United States
Environmental Protection
kl m m Agency
May 2011
EPA/600/R-10/075A
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 first 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 iv
Authors, Contributors, and Reviewers vi
Clean Air Scientific Advisory Committee Lead NAAQS Review Panel xi
Acronyms and Abbreviations xiii
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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
Dr. Stephen McDow—National Center for Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
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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. 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
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
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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
Dr. Stephen McDow—National Center for Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
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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. 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. MaryJane 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
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
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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—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. 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
Mr. Josh Drukenbrod, Office of Air Quality Planning and Standards, Office of Air and Radiation,
U.S. Environmental Protection Agency, RTP, NC
Ms. Katelyn Hausman—Department of Epidmiology, Gillings School of Global Public Health,
University of North Carolina, Chapel Hill, 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. 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. Brianna Young—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Reviewers
Dr. David Buchwalter—Department of Toxicology, North Carolina State University, Raleigh, 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
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
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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, Research Triangle Park, NC
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
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. 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
Dr. Jay Turner—Environmental and Chemical Engineering Department, Washington University, St.
Louis, MO
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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. 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
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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
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
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Dr. Michael Weitzman, New York University School of Medicine, New York, NY
* Members of the statutory Clean Air Scientific Advisory Committee (CASAC) appointed by the
EPA Administrator
Science Advisory Board Staff
Mr. Aaron Yeow, Designated Federal Officer, 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
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AST aspartate aminotransferase
ASV anode stripping voltammetry
ATLD ataxia-telangiectasia-like disorder
ATOFMS aerosol 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
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
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Ca
CAA
CaBP
CaCl2
CaC03
CaEDTA
CaMKII
cAMP
CASAC
CASM
CaS04
CaS04.2H20
CAT
CBLI
CBS A
CD
Cd
Cd(II)
Cd2+
CD3+
CD4+
CDC
CEA
CEC
cent
cert.
cf
CFL
CFR
cGMP
ChAT
CHD
CHL
CHO
C-horizon
CHV79
CI
Cir.
CKD
CKD-EPI
CL
CI
cr
Cl2
CLACE 5
CLS
CO
co2
CO,2"
calcium ion
Clean Air Act
calcium binding protein
calcium chloride
calcium carbonate; calcite
calcium ethylenediaminetetraacetic acid
calmodulin-dependent protein kinase II
cyclic adenosine monophosphate
Clean Air Scientific Advisory Committee
Comprehensive Aquatic Systems Model
calcium sulfate
gypsum
catalase
cumulative blood lead index
core based statistical area
cluster of differentiation
cadmium
cadmium (II)
cadmium ion
T lymphocyte
T helper cell
Centers for Disease Control
carcinoembryonic antigen
cation exchange capacity
central
certiorari
correction factor; latin abbreviation for conferre (used as "compared
with)
contanst flux layer
Code of Federal Regulations
cyclic guanosine monophosphate
chlorine acetyltranferase
coronary heart disease
Chinese hamster lung
Chinese hamster ovary cell line
soil horizon under A- and B- horizons, may contain lumps or shelves
of rock and parent material
Chinese hamster lung cell line
confidence interval
circuit
chronic kidney disease
Chronic Kidney Disease Epidemiology Collaboration
confidence limit
chlorine
chlorine ion
molecular chlorine
fifth Cloud and Aerosol Characterization Experiment in the Free
Troposphere campaign
Cincinnati Lead Study
carbon monoxide
carbon dioxide
carbonate ion
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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, CyplA2 cytochrome P450 family 1 member A2
CYP P450 cytochrome P450
A delta, difference, change
A5-3P-HSD delta-5-3-beta-hydroxysteroid dehydrogenase
8-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
difif differentiation
DIT developmental immunotoxicity
DMPS 2,3-dimercaptopropane-l-sulfonic acid
DMSA dimercaptosuccinic acid
DMSO dimethyl sulfoxide
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DNA
DoAD
DOC
DOM
DP-109
DP-460
DR
DRD4
DRD4.7
DRUM
D-serine
DSM-IV
DTH
DTPA
E
E2
e
EC
EC10
EC20
ec50
ECG
ECOD
Eco-SSLs
EDjo
EDTA
EFS
EOF
EGFR
eGFR
Eh
E-horizon
EI-MS
eNOS
EOG
EPA
EPT
ER
Erg-1
ERG
ERK
ERK1/2
EROD
ESCA
ESI-MS
ESRD
ET
ET-1
ETA-type receptors
EU
deoxyribonucleic acid
developmental origins of adult disease
dissolved organic carbon
dissolved organic matter
metal chelator
metal chelator
diet-restricted
dopamine 4 receptor
dopamine 4 receptor repeat alleles
Davis Rotating Unit for Monitoring
neuronal signal
Diagnostic Statistical Manual-IV
delayed-type hypersensitivity
diethylene triamine pentaacetic acid; techetium-diethylenetriamine-
pentaacetic acid
east; expression for exposure
estradiol
exponential function
endothelial cell
effect concentration for 10% of test population
effect concentration for 20% of test population
effect concentration for 50% of test population
electrocardiography; electrocardiogram
7-ethoxycoumarin-o-deethylase
ecological soil screening levels
effect dose for 10% of population
ethylenediaminetetraacetic acid
electrical field stimulus
epidermal growth factor
epidermal growth factor receptor
estimated glomerular filtration rate
electrochemical potential
soil horizon with eluviated or leached of mineral and/or organic
content
electron impact ionization mass spectrometry
endothelial nitric oxide synthase
end-of-grade
U.S. Environmental Protection Agency
ephemeroptera, plecoptera, trichoptera
endoplasmic reticulum
ether-a-go-go related gene
electroretinogram
extracellular signal regulated kinase
extracellular signal-regulated kinases 1 and 2
7-ethoxyresorufin-o-deethylase
electron spectroscopy for chemical analysis
electrospray ionization mass spectrometry
end stage renal disease
endothelin
vasoconstrictor endothelin-1
endothelin type A receptors
European Union
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EURO
eV
EXAFS
F0
Fi
F2
FAA
FAI
FAS
Fas-L
Fe
Fe(III)
FEM
FEV1
FI
FI-Ext
Fl
Fokl
FR
FrA
FR-FI
FRM
FSH
FSIQ
FT3
FT4
G
G2
g, mg, kg, |ig, ng, pg
G93A
GABA
GABAergic
GAD
GC
G-CSF
GD
GEE
GFAAS
GFAP
GFR
GGT
GH
GI
GIS
G+L
GLE
GM
GMR
GnRH
G6PD
GPEI
European emission standard
electronvolts
X-ray absorption fine structure spectroscopy
filial 0 generation
first offspring generation
second offspring generation
Federal Aviation Agency
free androgen index
apoptosis stimulating fragment
apoptosis stimulating fragment ligand
iron
iron III
Federal equivalence method
forced expiratory volume in 1 second
fixed interval
fixed interval with extinction
fluoride
polymorphism of the VDR in humans
Federal Register (Notice)
fractional anisotropy
fixed ratio-fixed interval
Federal reference method
follicle-stimulating hormone
full scale intelligence quotient (IQ)
free triodothyronine
free thyroxine
pregnancy; guanine
gap 2 Phase
gram(s), milligram(s), microgram(s), kilogram(s), nanogram(s),
picogram(s)
mouse model
y-aminobutyric acid; gamma aminobutyric acid
gamma aminobutyric acid-ergic
generalized anxiety disorder
gas chromatography
granulocyte colony-stimulating factor
gestational day
generalized estimating equations
graphite furnace atomic absorption spectrometry
glial fibrillary acidic protein
glomerular filtration rate
gamma-glutamyl transpeptidase
growth hormone
gastrointestinal
Geographic Information System
pregnancy plus lactation
gestationally-lead exposed
geometric mean
geometric mean blood lead ratio
gonadotropin-releasing hormone
glucose-6-phosphate dehydrogenase
glutathione transferase P (GST-P) enhancer I
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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
HNO3 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
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5HT serotonin
5-HT 5-hydroxytryptamine
5-HT2B 5-hydroxytryptamine 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-AES inductively coupled plasma atomic emission spectroscopy
ICPMS, ICP-MS inductively coupled plasma mass spectrometry
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
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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-PGFl a 6-keto-prostaglandin F1 a (vasodilatory prostaglandin)
ke V kiloelectron volt
Ki-67 antigen, cell cycle and tumor growth marker
Kim-1 kidney injury molecule-1
Kinder-KITAP Kinder-Testbatterie zur Aufmerksamkeitsprufung 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
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
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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
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NADP nicotinamide adenine dinucleotide phosphate
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
NAWQA National 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
NDMAR N-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 + N02)
NP nanoparticle
NPSH nonprotein sulfhydryl
NQOl NAD(P)H-quinone oxidoreductase (genotype)
NRC National Research Council
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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
OBS observations
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
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)
polycyclic aromatic hydro
Pb
lead
2°3pb
lead-203 radionuclide
2°4pb
stable isotope of lead-204
2°6pb
stable isotope of lead-206
2°7pb
stable isotope of lead-207
2°8pb
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
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PbCl2 lead chloride
PbCl3 lead (III) chloride; lead trichloride
PbCl4 lead (IV) chloride; lead tetrachloride
PbC03 cerrusite; lead carbonate
Pb(C03)2 lead (IV) carbonate
Pb(C03)2(0H)2 hydrocerussite
PbCr04 lead (II) chromate
PbD floor dust lead
PbFe6(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
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PMv
PMi(
PM2
PMii
PMii
p38MAPK
PMN
P5N
PND
POC
PP
PPb
ppm
PRP
PS
PSA
PSA-NCAM
PI
PTFE
PTHrP
PUFA
PVC
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.
In general terms, particulate matter with an aerodynamic diameter less
than or equal to a nominal 10 (im; a measurement of thoracic particles
(i.e., that subset of inhalable particles thought small enough to
penetrate beyond the larynx into the thoracic region of the respiratory
tract) in regulatory terms, particles with an upper 50%) cut-point of
10± 0.5 |im aerodynamic diameter (the 50% cut point diameter is the
diameter at which the sampler collects 50% of the particles and rejects
50%o of the particles) and a penetration curve as measured by a
reference method based on Appendix J of 40 CFR Part 50 and
designated in accordance with 40 CFR Part 53 or by an equivalent
method designated in accordance with 40 CFR Part 53.
In general terms, particulate matter with an aerodynamic diameter less
than or equal to a nominal 2.5 (im; a measurement of fine particles in
regulatory terms, particles with an upper 50% cut-point of 2.5 |im
aerodynamic diameter (the 50% cut point diameter is the diameter at
which the sampler collects 50% of the particles and rejects 50% of the
particles) and a penetration curve as measured by a reference method
based on Appendix L of 40 CFR Part 50 and designated in accordance
with 40 CFR Part 53, by an equivalent method designated in
accordance with 40 CFR Part 53, or by an approved regional method
designated in accordance with Appendix C of 40 CFR Part 58.
In general terms, particulate matter with an aerodynamic diameter less
than or equal to a nominal 10 |im and greater than a nominal 2.5 (im;
a measurement of thoracic coarse particulate matter or the coarse
fraction of PM10 in regulatory terms, particles with an upper 50%
cut-point of 10 |im aerodynamic diameter and a lower 50% cut-point
of 2.5 |im aerodynamic diameter (the 50% cut point diameter is the
diameter at which the sampler collects 50% of the particles and rejects
50%o of the particles) as measured by a reference method based on
Appendix O of 40 CFR Part 50 and designated in accordance with 40
CFR Part 53 or by an equivalent method designated in accordance
with 40 CFR Part 53.
The PM10_2 5 concentration of PM10_2 5 measured by the 40 CFR Part
50 Appendix O reference method which consists of currently
operated, collocated low-volume (16.7 Lpm) PM10 and PM2 5
reference method samplers.
p38 mitogen-activated protein kinase(s)
polymorphonuclear leukocyte
pyrimidine 5'-nucleotidase
post natal day
particulate organic carbon
polypropylene; pulse pressure
parts per billion
parts per million
post-reinforcement pause
dam stress; prenatal stress; phosphatidylserine
prostate specific antigen
polysialylated-neural cell adhesion molecule
proximal tubule
polytetrafluoroethylene
parathyroid hormone-related protein
polyunsaturated fatty acid
polyvinyl chloride
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220-
222-
PVD
Q
QRS
QT
QTc
P
pS
R
r
R2
r2
RAAS
RAC2
RBA
RBC
RBP
RD
Ref
RI-RI
RL
'Rn
Rn
RNA
ROI
ROS
RR
RSL
rtPCR
a
S
SAB
SATs
SBP
SCE
Sena
SD
SDN
SE
Se
sec
SEM
SES
Sess.
SGA
sGC
sGC-pi
SGOT
SGPT
SHBG
SHM
peripheral vascular disease
quantile; quartile; quintile
QRS complex in ECG
QT interval in ECG
corrected QT Interval
rho; bulk density; correlation
Pearson's r correlation coefficient
net drainage loss out of soil depth of concern; Spearman correlation
coefficient; upward resuspension flux; correlation
Pearson correlation coefficient
multiple regression correlation coefficient
correlation coefficient
renin-angiotensin-aldosterone system
gene encoding for Rac2
relative bioavailability
red blood cell
retinol binding protein
radial diffusivity
reference (group)
concurrent random interval
repeated learning
radon isotope
stable isotope of radon-222
ribonucleic acid
reactive oxygen intermediate/superoxide anion; regions of interest
reactive oxygen species
relative risk; risk ratio
roughness sublayer (transition layer, wake layer, interfacial layer)
reverse transcription polymerase chain reaction
sigma, standard deviation
south; sulfur; synthesis phase
U.S. EPA Science Advisory Board
Standard Assessment Tests
systolic blood pressure
sister chromatid exchange
a-synuclein
standard deviation
sexually dimorphic nucleus
standard error
selenium
second(s)
scanning electron microscopy; simultaneously extracted metal;
standard error of the mean
socioeconomic status
session
small for gestational age
soluble guanylate cyclase
soluble guanylate cyclase-beta 1
serum glutamic oxaloacetic transaminase
serum glutamic pyruvic transaminase
sex hormone binding globulin
Stockholm humic model
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siRNA small interfering RNA
SJW silver j ewelry 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
ti/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
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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
TNT 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
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 United States Environmental Protection Agency
USF uptake slope factor
USGS United States 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
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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
Xjj 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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Chapter 1 Contents
Chapter 1. Introduction 1-1
1.1. Legislative Requirements 1 -4
1.2. History of Reviews of the NAAQS for Lead 1-6
1.3. ISA Development 1-9
Figure 1-1. Identification of studies for inclusion in the ISA. 1-10
1.4. Document Organization 1-14
1.5. Document Scope 1-15
1.6. EPA Framework for Causal Determination 1-15
1.6.1. Scientific Evidence Used in Establishing Causality 1-16
1.6.2. Association and Causation 1-17
1.6.3. Evaluating Evidence for Inferring Causation 1-17
1.6.4. Application of Framework for Causal Determination 1-21
Table 1-1. Aspects to aid in judging causality 1-23
1.6.5. Determination of Causality 1-23
Table 1-2. Weight of evidence for causal determination 1-25
1.6.5.1. Effects on Human Populations 1-26
1.6.5.2. Effects on Ecosystems or Public Welfare 1-28
1.6.6. Concepts in Evaluating Adversity of Health Effects 1-28
1.7. Summary 1-29
Chapter 1 References 1 -30
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Chapter 1. Introduction
This first external review draft Integrated Science Assessment (ISA) is a concise review,
synthesis, and evaluation of the most policy-relevant evidence, and communicates critical science
judgments relevant to the National Ambient Air Quality Standards (NAAQS) review. As such, the
ISA forms the scientific foundation for the review of the primary (health-based) and secondary
(welfare-based) NAAQS for lead (Pb). The ISA accurately reflects "the latest scientific knowledge
useful in indicating the kind and extent of identifiable effects on public health which may be
expected from the presence of [a] pollutant in ambient air" (42 U.S.C. 7408). Key information and
judgments formerly contained in an Air Quality Criteria Document (AQCD) for Pb are incorporated
in this assessment. This ISA thus serves to update and revise the evaluation of the scientific evidence
available at the time of the previous review of the NAAQS for Pb that was concluded in 2008.
The draft Integrated Review Plan for the National Ambient Air Quality Standards for Lead
(U.S. EPA. 2011) identifies a series of policy-relevant questions that provide a framework for this
assessment of the scientific evidence. These questions frame the entire review of the NAAQS for Pb,
and thus are informed by both science and policy considerations. The ISA organizes, presents, and
integrates the scientific evidence which is considered along with findings from any risk analyses and
policy considerations to help the U.S. Environmental Protection Agency (EPA) address these
questions during the NAAQS review for Pb. In evaluating the health evidence, the focus of this
assessment is on scientific evidence that is most relevant to the following questions taken directly
from the Integrated Review Plan:
¦ What new evidence is available on exposure to Pb through air-related pathways? Can air-
related pathways be disentangled from water- and soil-related pathways using available
data?
¦ What new evidence is available regarding observational studies of Pb exposure? How do
these studies inform the assessment of exposure to air-related pathways?
¦ What new evidence is available on biological and other factors that could affect the
distribution and accumulation of Pb into blood and bone (e.g., age, diet, gender, race)?
Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and
Environmental Research Online) at http://epa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of
developing science assessments such as the Integrated Science Assessments (ISA) and the Integrated Risk Information System (IRIS).
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How and to what extent does previous or concurrent Pb exposure, including duration
(e.g., acute, subchronic, chronic) and pattern (e.g., continuous low, extreme peak) impact
blood and bone Pb?
What new evidence is available on the relationship between air Pb and blood Pb levels
and uncertainties in that relationship? What new knowledge exists regarding the
characterization of changes in this relationship when accounting for the multiple
pathways of Pb exposure and body burden associated with Pb exposure? What does the
current evidence indicate regarding variation in the relationship with variation in blood
Pb levels or air Pb levels?
To what extent does new scientific evidence increase our understanding of the
contributions of Pb from different sources and exposure pathways to blood Pb levels or
to other indicators of Pb body burden (e.g., contributions from various air-related
pathways, including diet and indoor dust pathways)?
How do results of recent epidemiologic studies and current or new interpretations of
previous findings expand our understanding of the relationship between body burdens of
Pb and neurological effects in children and adults, including deficits in IQ, behavior,
learning, and motor skills, as well as risk of neurodegenerative diseases? What new
evidence is available on the potential clinical relevance of these effects? Do recent
studies expand the current understanding of concentration-response relationships
pertinent to the range of Pb exposures currently experienced by the U.S. population?
For what Pb-induced health effects, is there sufficient evidence in multiple species to
support a quantitative comparison of exposures that induce the effects?
To what extent are the health effects observed in epidemiological studies attributable to
exposure to Pb rather than co-exposures to other toxic metals or environmental
contaminants?
In epidemiologic studies, what are the uncertainties in Pb effect estimates due to
potential confounding factors (e.g., demographic and lifestyle attributes, socioeconomic
status [SES], genetic susceptibility factors, occupational exposure, and access to medical
care)?
Based on the new body of evidence, what uncertainties remain regarding the nature and
shape of concentration-response relationships (e.g., threshold, linear, nonlinear)? What
evidence is newly available on the uncertainties related to other aspects of statistical
model specification and how can it be used to assess the influence of these uncertainties
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on the results of epidemiologic studies? What evidence is available from toxicological
studies of dose-response relationships?
¦ To what extent is key evidence now available regarding mechanisms for neurological
effects associated with "lower" (<10 (ig/dL or <5 (ig/dL) blood Pb levels (e.g., oxidative
stress)? What toxicological evidence is available on mechanisms and dose-response
relationships for other health outcomes (e.g., cardiovascular, renal, or immunological
effects), and is there coherence between this and epidemiologic findings for these
endpoints?
¦ To what extent is key new evidence available that could inform the understanding of
populations that are particularly susceptible to Pb exposures? What is known about
genetic traits, pre-existing conditions (obesity), or other factors that affect susceptibility
(sex)? To what extent is the strength of epidemiologic or toxicological evidence driven
by effects observed in populations with increased susceptibility?
¦ To what extent is key evidence now available to inform our understanding of
developmental lifestages that are particularly susceptible to Pb exposures? What is
known about critical windows of exposure for Pb with regard to their impact on
concentration-response relationships and/or effects elicited?
¦ What do the currently available studies indicate regarding the relationship between
exposures to Pb and health effects in those with preexisting diseases (e.g., renal diseases)
compared to healthy individuals? What medical conditions are identified as increasing
susceptibility to Pb effects? What is the nature and time-course of the development of
effects in previously healthy persons and in persons with pre-existing disease (e.g.,
cardiovascular disease)? What are the pathways and mechanisms through which Pb may
be acting for these groups?
In evaluating evidence on welfare effects of Pb, the focus will be on evidence that can help
inform these questions from the Integrated Review Plan:
¦ What new information is available about the nature of the effects of Pb on terrestrial
ecosystems, especially Pb that is relevant to air-related pathways? Is there new evidence
of effects at current ecosystem loads? Is there new evidence that, in combination with the
previously existing evidence, supports the development of critical loads for terrestrial
ecosystems?
¦ Is there new information available for establishing specific exposure levels at which
terrestrial ecological receptors are expected to experience effects?
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¦ Are there new empirical data or modeling results that would improve our understanding
of the movement of Pb in or through terrestrial systems, or would improve our
understanding of Pb bioavailability and pathways of exposure for terrestrial organisms?
¦ Is there new evidence that contributes to a better understanding of the nature and
magnitude of the potential effects of Pb on terrestrial ecosystem services?
¦ What new information is available about the nature of the effects of Pb on aquatic
ecosystems, especially Pb that is relevant to air-related pathways? Is there new evidence
of effects at current ecosystem loads? Is there new evidence that, in combination with the
previously existing evidence, supports the definition of critical loads for aquatic
ecosystems?
¦ Is there new information available for establishing specific exposure levels at which
aquatic ecological receptors are expected to experience effects?
¦ Are there new empirical data or modeling results that would improve our understanding
of the movement of Pb in or through aquatic systems or would improve our
understanding of Pb bioavailability and pathways of exposure for aquatic organisms?
¦ Is there new evidence that contributes to a better understanding of the nature and
magnitude of the potential effects of Pb on aquatic ecosystem services?
This introductory chapter (Chapter 1) of the Pb ISA presents: (1) background information on
pertinent Clean Air Act legislative requirements, the air quality criteria and NAAQS review process,
and the history of previous Pb NAAQS reviews; (2) an overview of the ISA development process
and an orientation to the general organizational structure and content of this Pb ISA; and (3) the
framework for causal determination used to evaluate the causal nature of air pollution-induced health
and environmental effects in NAAQS reviews.
1.1. Legislative Requirements
Two sections of the Clean Air Act (CAA) govern the establishment and revision of the
NAAQS. Section 108 (42 USC §7408) directs the Administrator to identify and list certain air
pollutants and then 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 and welfare;" "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
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scientific knowledge useful in indicating the kind and extent of identifiable effects on public health
or welfare which may be expected from the presence of [a] pollutant in ambient air...." 42 USC
§7408(b). Section 109 (42 USC §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."1 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."2
The requirement that primary standards include 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. See Lead Industries Association v. EPA, 647 F.2d 1130,
1154 (D.C. Cir 1980), cert, denied, 449 U.S. 1042 (1980); American Petroleum Institute v. Costle,
665 F.2d 1176, 1186 (D.C. Cir. 1981), cert, denied, 455 U.S. 1034 (1982); American Farm Bureau v.
EPA, 559 F. 3d 512, 533 (D.C. Cir. 2009); Coalition of Battery Recyclers Association v. EPA, 604 F.
3d 613, 617-18 (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 include 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, see 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
1 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).
2 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."
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Administrator's judgment. Sqq Lead Industries Association v. EPA, 647 F.2d at 1161-62; Whitman v.
American Trucking Associations, 531 U.S. 457, 495 (2001).
In setting primary and secondary standards that are "requisite" to protect public health and
welfare, respectively, 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, 475-76 (2001). Likewise, "Attainability and technological feasibility are not
relevant considerations in the promulgation of national ambient air quality standards." American
Petroleum Institute v. Costle, 665 F. 2d at 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..." Since the early 1980s, this independent review function has been
performed by the Clean Air Scientific Advisory Committee (CASAC).1
1.2. History of Reviews of the NAAQS for Lead
On October 5, 1978, EPA promulgated primary and secondary NAAQS for Pb under section
109 of the Act (43 FR 46246). Both primary and secondary standards were set at a level of 1.5 (ig
micrograms per cubic meter ((xg/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. This
standard was based on the 1977 AQCD for Pb (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 Lead AQCD (U.S. EPA. 1986a). the associated Addendum
(U.S. EPA. 1986b). 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
1 Lists of CASAC members and of members of the CASAC Pb Review Panel are available at:
http://vosemite.epa.gov/sab/sabproduct.nsfAVebCASAC/CommitteesandMembership7QpenDocument
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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 AQCD and conduct of the
NAAQS review were contained in two documents: Project Work Plan for Revised Air Quality
Criteria for Lead (U.S. EPA. 2005b) and Plan for Review of the National Ambient Air Quality
Standards for Lead (U.S. EPA. 2006c).1 The schedule for completion of this review was governed by
ajudicial order in Missouri Coalition for the Environment v. EPA (No. 4:04CV00660 ERW, Sept. 14,
2005; amended April 29, 2008 and July 1, 2008), which specified a schedule for the review of
duration substantially shorter than five years.
The scientific assessment for the review is described in the 2006 AQCD for Pb(U.S. EPA.
2006a). 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.
2006b). 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. 2007). 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. 2006d). 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 AQCD and presented and
interpreted results from the quantitative risk/exposure analyses conducted for this review. Based on
this evaluation, the Staff Paper presented OAQPS staff recommendations that the Administrator give
consideration to substantially revising the primary and secondary standards to a range of levels at or
below 0.2 (ig/m3.
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
1 In the current review, these two documents have been combined into an integrated plan (this document).
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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 Report. 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). In addition to public
comments on the proposal received during the public comment period, both written and oral at two
public hearings, the CASAC Pb Panel 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 revisions to the primary and secondary NAAQS
for Pb. In consideration of the much-expanded health effects evidence on neurocognitive effects of
Pb in children, EPA substantially revised the primary standard from a level of 1.5 (ig/m3 to a level of
0.15 (ig/m3. EPA's decision on the level for the standard was based on the weight of the scientific
evidence and guided by an evidence-based framework that integrates evidence for relationships
between Pb in air and Pb in children's blood and Pb in children's blood and IQ loss. The level of
0.15 (xg/m3 was estimated to protect against air Pb-related 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 estimates. The averaging time was revised to a
rolling 3-month period with a maximum (not-to-be-exceeded) form, evaluated over a 3-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 in total suspended particles (Pb-TSP) was retained,
reflecting the evidence that Pb particles of all sizes pose health risks. The secondary standard was
revised to be identical in all respects to the revised primary standards.2
Revisions to the NAAQS were accompanied by revisions to the data handling procedures, the
treatment of exceptional events and the ambient air monitoring and reporting requirements, as well
as emissions inventory reporting requirements.3 One aspect of the new data handling requirements is
1 The ANPR was one of the features of the revised NAAQS review process that EPA instituted in 2006. :In 2009, this component of the
process was replaced by reinstatement of the OAQPS policy assessment (previously termed the Staff Paper).
2 The 2008 NAAQS for Pb are specified at 40 CFR 50.16.
3 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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the allowance for the use of Pb-PMi0 monitoring for Pb NAAQS attainment purposes in certain
limited circumstances at non-source oriented sites. Subsequent to the 2008 rulemaking, additional
revisions were made to the monitoring network requirements.
1.3. ISA Development
EPA announced the initiation of the current periodic review of the air quality criteria for Pb
and the Pb NAAQS in April 2010 and issued a call for information from the public (75 FR 20843).
In addition to the call for information, publications were identified through an ongoing literature
search process that includes extensive computer database mining on specific topics in a variety of
disciplines. Literature searches were conducted to identify studies published since the last review,
focusing on publications from January 2006 to March 2011. Search strategies were iteratively
modified in an effort to optimize the identification of pertinent publications. Additional papers were
identified for inclusion in several ways: review of pre-publication tables of contents for journals in
which relevant papers may be published and independent identification of relevant literature by
expert authors and peer reviewers. It is anticipated that further identification of studies will occur
during the external review process by the public and CASAC. Publications considered for inclusion
in the ISA were added to the Health and Environmental Research Online (HERO) database recently
developed by EPA (http://hero.epa.gov); note that all citations in the ISA are electronically linked to
the database. Typically, only information that underwent scientific peer review and was published or
accepted for publication was considered. All relevant epidemiologic, animal toxicological, and
ecological and welfare effects studies published since the last review were considered, including
those related to exposure-response relationships, mode(s) of action (MOA), and susceptible
populations. Additionally, air quality and emissions data, studies on atmospheric chemistry,
environmental fate and transport, as well as issues related to Pb toxicokinetics and exposure were
considered for inclusion in the document. The process for identifying studies for consideration and
inclusion in the Pb ISA is provided in Figure 1-1. All references that were considered for inclusion
can be found within the HERO website (http://hero.epa.gov/lead). This site contains HERO links to
lists of references that are cited in the ISA, as well as those that were considered for inclusion, but
not cited in the ISA, with bibliographic information and abstracts.
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KEY DEFINITIONS
INFORMATIVE studies are well-designed,
properly implemented, thoroughly described.
HIGHLY INFORMATIVE studies reduce
uncertainty on critical issues, may include
analyses of confounding or effect
modification by copollutants or other
variables, analyses of concentration-
response or dose-response relationships,
analyses related to time between exposure
and response, and offer innovation in
method or design.
POLICY-RELEVANT studies may include
those conducted at or near ambient
concentrations and studies conducted in
U.S. and Canadian airsheds.
Informative
studies are
identified.
Yes
No
Studies are
evaluated for
inclusion in the
ISA.
Continuous,
comprehensive
literature review of
peer-reviewed
journal articles
Studies added to
the docket during
public comment
period.
Studies identified
during EPA
sponsored kickoff
meeting (including
studies in
preparation).
Studies that do not
address exposure
and/or effects of
air pollutant(s)
under review are
excluded.
Selection of
studies discussed
and additional
studies identified
during CASAC
peer review of
draft document.
Policy relevant and highly informative studies discussed in the ISA text include
those that provide a basis for or describe the association between the criteria
pollutant and effects. Studies summarized in tables and figures are included
becaues they are sufficiently comparable to be displayed together. A study
highlighted in the ISA text does not necessarily appear in a summary table or
figure.
ISA
Figure 1-1. Identification of studies for inclusion in the ISA.
1 The ISA builds upon the conclusions of the previous review of the air quality criteria for Pb,
2 presented in the 2006 Pb AQCD (U.S. EPA. 2006a). and focuses on peer reviewed literature
3 published thereafter and on any new interpretations of previous literature. The 2006 Pb AQCD (U.S.
4 EPA. 2006a) evaluated literature published through December 2005. In subsequent chapters, the
5 results of recent scientific studies are integrated with previous findings. Important older studies may
6 be discussed in detail to reinforce key concepts and conclusions or if they are open to
7 reinterpretation in light of newer data. Older studies also are the primary focus in some areas of the
8 document where research efforts have subsided, and these older studies remain the definitive works
9 available in the literature. Emphasis is placed on studies that examine effects associated with Pb
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concentrations relevant to current population and ecosystem exposures, and particularly those
pertaining to Pb 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.
Discussions in the ISA primarily focus on scientific evaluations that can inform the key policy
questions described in the Integrated Review Plan (U.S. EPA. 2011). Although emphasis is placed on
discussion of health and welfare effects information, other scientific information is also presented
and evaluated in order to provide a better understanding of the sources of Pb to ambient air,
measurement and concentrations of Pb in ambient air, its subsequent fate and transport in the
environment, pathways of human and ecological exposure, and toxicokinetic characteristics of Pb in
the human body, as well as the measurement of population exposure to Pb.
In general, in assessing the scientific quality and relevance of health and environmental effects
studies, the following considerations were taken into account when selecting studies for inclusion in
the ISA.
¦ Are the study populations, subjects, or animal models adequately selected and are they
sufficiently well defined to allow for meaningful comparisons between study or exposure
groups?
¦ Are the statistical analyses appropriate, properly performed, and properly interpreted?
Are likely covariates adequately controlled or taken into account in the study design and
statistical analysis?
¦ Are the air quality data, exposure, or dose metrics of adequate quality?
¦ Are the health or welfare effect measurements meaningful and reliable?
Studies published since the 2006 Pb AQCD are emphasized; however, evidence from studies
described in previous assessment that are needed to characterize the current state of the science as
well as new interpretations of older evidence is also considered. Among the studies EPA included in
the ISA, particular focus is for the following areas:
¦ New studies with adequate blood Pb data at the low end of the distribution (e.g.,
<10 (ig/dL);
¦ New studies that provide quantitative effect estimates for populations or lifestages and
concentrations of interest;
¦ Lead exposure or effects in susceptible populations and lifestages;
¦ Issues related to the potential for confounding of study effects/responses by non-Pb
exposure-related factors or variables, and to the modification of Pb-related effects;
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¦ The timing (e.g., across/within specific lifestages) and duration of exposure associated
with specific responses;
¦ Concentration-response relationships for specific Pb-related effects;
¦ The interpretation of Pb biomarkers in epidemiological studies; and air-to-blood Pb or
air-to-bone Pb relationships;
¦ Studies that evaluate Pb as a component of a complex mixtures of pollutants.
In selecting epidemiologic studies for inclusion in the present assessment, EPA has considered
studies containing information on: (1) recent or cumulative exposures relevant to current population
exposure levels of Pb; (2) health endpoints that repeat or extend findings from earlier assessments as
well as those not previously extensively researched; (3) populations and lifestages that are
susceptible to Pb exposures; (4) issues related to potential confounding, and modification of effects;
and/or (5) important methodological issues (e.g., timing and duration of exposure, concentration-
response relationships, interpretation of biomarkers in epidemiological studies, and air-to-blood/bone
relationships) related to Pb exposure effects. In selecting the most informative and policy-relevant
epidemiologic studies on which to give particular focus in the Pb ISA, emphasis is placed on those
most relevant to standard setting in the United States. Informative studies conducted in other
countries are discussed, as appropriate (e.g., studies for which the mean blood Pb level in the
population studied is comparable to the current mean blood Pb level in the corresponding U.S.
population).
In reviewing new studies that evaluated the response of laboratory animals to Pb exposure,
focus in on studies that reveal the effects of Pb exposure within the previously identified target
biological systems (e.g., neurological, cardiovascular, renal, immune). Additionally, particular focus
is on those studies that involve doses or blood Pb/bone Pb levels that approximate human doses or
blood Pb/bone Pb levels relevant to current U.S. populations. Studies at higher exposure doses that
result in body burdens above what is found in the current U.S. population are included when the
study can provide information relevant to potential MOA, information on exposure-response
relationships, or otherwise improve our understanding of susceptible populations. Studies of the
efficacy of chelation as a treatment for Pb poisoning in humans and laboratory animals are excluded,
except where they provide evidence for the reversibility of a given health effect.
In selecting informative studies of welfare effects, emphasis is placed on recent studies that:
(1) evaluate the occurrence of effects associated with Pb exposure at current ambient concentrations,
with a particular focus on ambient concentrations resulting from ambient air Pb; and/or (2)
investigate the effects of Pb on ecosystems at any scale. Studies conducted in geographical areas
outside the U.S. are included in the assessment if they contribute to the general knowledge of the
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effects of Pb irrespective of species or locality. As in the selection of health-related scientific studies,
welfare-related studies were selected that advance our understanding of MOA by which Pb directly
affects terrestrial and aquatic biota. These MOAs, as they pertain to Pb exposures of short or longer
duration, informs our understanding of indirect effects that Pb may exert more broadly on ecosystem
structure, function and services. Key studies identified for welfare effects are integrated into the
discussion to inform our interpretation of the ecological literature and our characterization of
uncertainties.
The criteria described here provide generalized benchmarks to guide the inclusion in the ISA
of the highest quality and most policy-relevant studies. Of most relevance for evaluation of studies is
whether they provide useful qualitative or quantitative information on exposure-effect or
exposure-response relationships for effects associated with current blood Pb or bone Pb levels likely
to be encountered in the U.S. population. Detailed critical analysis of all studies of the effects of Pb
on health and welfare, especially in relation to the above considerations, is beyond the scope of this
document. Since the last AQCD was completed in 2006, a considerable portion of the current ISA is
devoted to summarizing previously available evidence that contributed to the basis for the last
rulemaking.
As discussed previously, studies included in the text of the ISA are those deemed informative
to the NAAQS review process (e.g., policy relevant) and of adequate quality. The ISA text, tables
and figures highlight and summarize key study details that are needed to understand and interpret the
results of a study. This information, which was described in the text as well as reiterated in the annex
tables of previous documents, includes the air quality system (AQS) data; studies of fate and
transport in air, water, and soil; human exposure and dosimetry studies; blood or tissue Pb levels
corresponding to adverse health effects and dose and duration of exposure in toxicological studies;
and, effect estimates, study location, population, exposure metric and time window, as well as the
characteristics of the exposure/dose distribution for epidemiologic studies. In addition,
supplementary materials are provided in the form of output from the HERO database. A key function
of the HERO output is to document the base of evidence containing publications evaluated for the Pb
review, including any publications considered but not included in the ISA. This information is
presented as links to lists of references in the HERO database, which include bibliographic
information and abstracts and can be found at http://hero.epa.gov/lead. In addition, certain study
characteristics of epidemiologic studies, including location, ages investigated, outcomes, and health
endpoints, are summarized in tables included in Chapter 5.
In developing the Pb ISA, EPA began by reviewing and summarizing the evidence on: (1)
atmospheric sciences and exposure; (2) the health effects evidence from in vivo and in vitro animal
toxicological and epidemiologic studies; and (3) the welfare effects of Pb, including ecological
effects. In December 2010, EPA held a peer-review input workshop to obtain review of the scientific
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content of initial draft materials or sections for the ISA. The purpose of the peer-review input
workshop was to ensure that the ISA was up to date and focused on the most policy-relevant
findings, and to assist EPA with integration of evidence within and across disciplines. Subsequently,
EPA addressed comments from the peer-review input workshop and completed the initial integration
and synthesis of the evidence.
The integration of evidence on health or welfare effects involves collaboration between
scientists from various disciplines. As described in the section below, the ISA organization is based
on health and welfare effect categories. As an example, an evaluation of health effects evidence
would include summaries of findings from epidemiologic and toxicological studies, and integration
of the results to draw conclusions based on the causal framework described below. Using the causal
framework described in Section 1.6, EPA scientists consider aspects such as strength, consistency,
coherence and biological plausibility of the evidence, and develop draft causality judgments on the
nature of the relationships. The draft integrative synthesis sections and conclusions are reviewed by
EPA internal experts and, as appropriate, by outside expert authors. In practice, causality
determinations often entail an iterative process of review and evaluation of the evidence. The draft
ISA is released for review by the CASAC and the public. Comments on the characterization of the
science as well as the implementation of the causal framework are carefully considered in revising
and completing the ISA.
1.4. Document Organization
This ISA is composed of seven chapters. This introductory chapter presents background
information, and provides an overview of EPA's framework for making causal judgments. Chapter 2
is an integrated summary of key findings and conclusions regarding the source-to-dose paradigm,
toxicokinetics, MOA, important health effects of Pb, including neurological, cardiovascular, renal,
immunological, reproductive and developmental, and cancer outcomes, and welfare effects of Pb,
including terrestrial and aquatic ecosystem impacts. More detailed summaries, evaluations, and
integration of the evidence are included in Chapters 3 through 7. Chapter 3 highlights key concepts
or issues relevant to understanding the sources, ambient concentrations, atmospheric behavior, and
fate of Pb in the environment. Chapter 4 summarizes key concepts and recent findings on exposures
to Pb using a conceptual model that reflects the multimedia nature of Pb exposure, toxicokinetics,
biomarkers of Pb exposure and body burden, and models. 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 neurological, cardiovascular, renal, immunological,
reproductive and developmental, and cancer outcomes. Chapter 6 summarizes the evidence on
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potentially susceptible populations for health effects of Pb exposure. Chapter 7 evaluates welfare
effects evidence that is relevant to the review of the secondary NAAQS for Pb, including ecological
effects. The chapter also presents key conclusions and scientific judgments regarding causality for
welfare effects of Pb.
1.5. Document Scope
For the current NAAQS review of the primary Pb standard, relevant scientific information on
human exposures and health effects associated with exposure to ambient Pb has been assessed.
Previous reviews have included an extensive body of evidence from the major health disciplines of
toxicology and epidemiology on the health effects of Pb exposure (U.S. EPA. 1986a. 2006a). In this
review, the conclusions from previous reviews are summarized at the beginning of each health
outcome discussion to provide the foundation for consideration of evidence from recent studies.
In some cases where no new information is available, the summary of key findings and conclusions
from the previous Pb AQCDs serve as the basis for current key conclusions. Results of key studies
from previous reviews are included in ISA 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 review also assesses scientific information associated with known or anticipated
ecological and public welfare effects that is relevant to the review of the secondary Pb standard.
Research on the ecological effects of Pb, including impacts on vegetation, has been discussed
extensively in previous AQCDs. This review incorporates and discusses findings of recent studies,
building upon previous evaluations and conclusions.
1.6. EPA Framework for Causal Determination
The EPA has developed a consistent and transparent basis to evaluate the causal nature of air
pollution-induced health or environmental effects. The framework described below establishes
uniform language concerning causality and brings more specificity to the findings. This standardized
language was drawn from across the federal government and wider scientific community, especially
from the National Academy of Sciences (NAS) Institute of Medicine (IOM) document, Improving
the Presumptive Disability Decision-Making Process for Veterans, (2008) the most recent
comprehensive work on evaluating causality.
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This introductory section focuses on the evaluation of health effects evidence. While focusing
on human health outcomes, the concepts are also generally relevant to causality determination for
welfare effects. This section:
¦ describes the kinds of scientific evidence used in establishing a general causal
relationship between exposure and health effects;
¦ defines causation, in contrast to statistical association;
¦ discusses the sources of evidence necessary to reach a conclusion about the existence of
a causal relationship;
¦ highlights the issue of multifactorial causation;
¦ 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 (2006). EPA Guidelines for Carcinogen Risk Assessment (2005a). Centers for
Disease Control and Prevention (2004), and National Acid Precipitation Assessment Program (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. Moreover, these frameworks have supported
decision-making under conditions of uncertainty.
1.6.1. Scientific Evidence Used in Establishing Causality
Causality determinations are based on the evaluation and synthesis of evidence from across
scientific disciplines; the type of evidence that is most important for such determinations will vary
by assessment. The most direct evidence of a causal relationship between pollutant exposures and
human health effects comes from controlled human exposure studies. This type of study
experimentally evaluates the health effects of administered exposures in human volunteers under
highly-controlled laboratory conditions. Controlled human exposure studies are not done for Pb, and
thus, are unavailable for consideration.
In most epidemiologic or observational studies of humans, the investigator does not control
exposures or intervene with the study population. Broadly, observational studies can describe
associations between exposures and effects. In the case of Pb, most observational studies use
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biomarkers of Pb (i.e., blood or bone Pb) rather than exposures to relate to effects. These studies fall
into several categories: cross-sectional and longitudinal studies. "Natural experiments" offer the
opportunity to investigate changes in health with a change in exposure; these include comparisons of
health effects before and after a change in population exposures, such as the closure of a pollution
source.
Experimental animal data can help characterize effects of concern, exposure-response
relationships, susceptible populations, MO As and enhance understanding of biological plausibility of
observed effects. In the absence of human data, animal data alone may be sufficient to support a
likely causal determination, assuming that similar responses are expected in humans.
1.6.2. Association and Causation
"Causation" is a significant, effectual relationship between an agent and an effect on health or
welfare. "Association" is the statistical dependence among events, characteristics, or other variables.
An association is prima facie evidence for causation; alone, however, it is insufficient proof of a
causal relationship between exposure and disease or health effect. Determining whether an observed
association is causal rather than spurious involves consideration of a number of factors, as described
below. Much of the newly available health information evaluated in this ISA comes from
epidemiologic studies that report a statistical association between ambient exposure and health
outcomes.
Many of the health and environmental outcomes reported in these studies have complex
etiologies. Diseases such as asthma, cardiovascular disease, Parkinson's disease or cancer are
typically initiated by a web of multiple agents. Outcomes depend on a variety of factors, such as age,
genetic susceptibility, nutritional status, immune competence, and social factors (Gee & Pavne-
Sturges. 2004; Samet & C. C. Bodurow. 2008). Effects on ecosystems are also multifactorial with a
complex web of causation. Further, exposure to a combination of agents could cause synergistic or
antagonistic effects. Thus, the observed risk represents the net effect of many actions and
counteractions.
1.6.3. Evaluating Evidence for Inferring Causation
Moving from association to causation involves the elimination of alternative explanations for
the association. In estimating the causal influence of an exposure on health or environmental effects,
it is recognized that scientific findings incorporate uncertainty. "Uncertainty" can be defined as a
state of having limited knowledge where it is impossible to exactly describe an existing state or
future outcome, e.g., the lack of knowledge about the correct value for a specific measure or
estimate. Uncertainty characterization and uncertainty assessment are two activities that lead to
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different degrees of sophistication in describing uncertainty. Uncertainty characterization generally
involves a qualitative discussion of the thought processes that lead to the selection and rejection of
specific data, estimates, scenarios, etc. Uncertainty assessment is more quantitative. The process
begins with simpler measures (e.g., ranges) and simpler analytical techniques and progresses, to the
extent needed to support the decision for which the assessment is conducted, to more complex
measures and techniques. Data will not be available for all aspects of an assessment and those data
that are available may be of questionable or unknown quality. In these situations, evaluation of
uncertainty can include professional judgment or inferences based on analogy with similar situations.
The net result is that the assessment will be based on a number of assumptions with varying degrees
of uncertainty. Uncertainties commonly encountered in evaluating health evidence for the criteria air
pollutants are outlined below for epidemiologic and experimental studies. Various approaches to
evaluating uncertainty include classical statistical methods, sensitivity analysis, or probabilistic
uncertainty analysis, in order of increasing complexity and data requirements. The ISA generally
evaluates uncertainties qualitatively in assessing the evidence from across studies; in some
situations, quantitative analysis approaches, such as meta-regression, may be used.
Meta-analysis may be a valuable tool for evaluating evidence by combining results from a
body of studies. Blair et al. (1995) observe that meta-analysis can enhance understanding of
associations between exposures and effects that are not readily apparent in examination of individual
study results and can be particularly useful for formally examining sources of heterogeneity.
However, these authors note that meta-analysis may not be useful when the relationship between the
exposure and outcome is obvious, when only a few studies are available for a particular exposure-
outcome relationship, where there is limited access to data of sufficient quality, or where there is
substantial variation in study design or population. In addition, important differences in effect
estimates, exposure metrics, or other factors may limit or even preclude quantitative statistical
combination of multiple studies.
Epidemiologic studies provide important information on the associations between health
effects and exposure of human populations to ambient air pollution. 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 & Greenland. 1998). One approach to remove spurious associations from
possible confounders is to control for characteristics that may differ between exposed and unexposed
persons; this is frequently termed "adjustment." Appropriate statistical adjustment for confounders
requires identifying and measuring all reasonably expected confounders. Deciding which variables
to control for in a statistical analysis of the association between exposure and disease or health
outcome depends on knowledge about possible mechanisms and the distributions of these factors in
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the population under study. In addition, scientific judgment is needed regarding likely sources and
magnitude of confounding, together with consideration of how well the existing constellation of
study designs, results, and analyses address this potential threat to inferential validity.
Another important consideration in the evaluation of epidemiologic evidence is effect
modification. "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 & Greenland. 1998). Examples of effect modifiers in some of the
studies evaluated in this ISA include environmental variables, such as temperature or humidity,
individual risk factors, such as education, cigarette smoking status, age in a prospective cohort study,
and community factors, such as percent of population >65 years old. It is often possible to stratify
the relationship between health outcome and exposure or biomarker by one or more of these risk
factor variables. For variables that modify the association, effect estimates in each stratum will be
different from one another and different from the overall estimate, indicating a different exposure-
response relationship may exist in populations represented by these variables. Effect modifiers may
be encountered (1) within single-city time-series studies; or (2) across cities in a two-stage
hierarchical model or meta-analysis.
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 pollutant likely responsible for the health outcome 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. Further, the correlation between the air pollutant of
interest and various copollutants may make it difficult to discern associations between different
pollutant exposures and health effects. Thus, results of models that attempt to distinguish gaseous
and particle effects must be interpreted with caution. 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.
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Another way to adjust for potential confounding is through stratified analysis, i.e., examining
the association within homogeneous groups with respect to the confounding variable. The use of
stratified analyses has an additional benefit: it allows examination of effect modification through
comparison of the effect estimates across different groups. If investigators successfully measured
characteristics that distort the results, adjustment of these factors help separate a spurious from a true
causal association. Appropriate statistical adjustment for confounders requires identifying and
measuring all reasonably expected confounders. Deciding which variables to control for in a
statistical analysis of the association between exposure and disease or health outcome depends on
knowledge about possible mechanisms and the distributions of these factors in the population under
study. Identifying these mechanisms makes it possible to control for potential sources that may result
in a spurious association.
Adjustment for potential confounders can be influenced by differential exposure measurement
error. There are several components that contribute to exposure measurement error in 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. Previous AQCDs have examined the role of measurement error in time-series
epidemiologic studies using simulated data and mathematical analyses and suggested that "transfer
of effects" would only occur under unusual circumstances (i.e., "true" predictors having high
positive or negative correlation; substantial measurement error; or extremely negatively correlated
measurement errors) (U.S. EPA. 2004).
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. Thus, multicity studies
which use a consistent method to analyze data from across locations with different levels of
covariates can provide insight on potential confounding in associations. Intervention studies, because
of their quasi-experimental nature, can be particularly useful in characterizing causation.
In addition to controlled human exposure and epidemiologic studies, the tools of experimental
biology have been valuable for developing insights into human physiology and pathology.
Laboratory tools have been extended to explore the effects of putative toxicants on human health,
especially through the study of model systems in other species. These studies evaluate the effects of
exposures to a variety of pollutants in a highly controlled laboratory setting and allow exploration of
MOAs or mechanisms by which a pollutant may cause effects. Background knowledge of the
biological mechanisms by which an exposure might or might not cause disease can prove crucial in
establishing or negating a causal claim. There are, however, uncertainties associated with
quantitative extrapolations between laboratory animals and humans on the pathophysiological effects
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of any pollutant. Animal species can differ from each other in fundamental aspects of physiology and
anatomy (e.g., metabolism, airway branching, hormonal regulation) that may limit extrapolation.
Interpretations of experimental studies of pollutant effects in laboratory animals, as in the case
of environmental comparative toxicology studies, are affected by limitations associated with
extrapolation models. The differences between humans and rodents with regard to pollutant
absorption and distribution profiles based on metabolism, hormonal regulation, exposure dose, and
differences in target organ structure and anatomy, all have to be taken into consideration. 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
molecular differences, at this time there are uncertainties associated with quantitative extrapolations
between laboratory animals and humans of observed pollutant-induced pathophysiological
alterations under the control of widely varying biochemical, endocrine, and neuronal factors.
1.6.4. Application of Framework for Causal Determination
EPA uses a two-step approach to evaluate the scientific evidence on health or environmental
effects of criteria pollutants. The first step determines the weight of evidence in support of causation
and characterizes the strength of any resulting causal classification. The second step includes further
evaluation of the quantitative evidence regarding the concentration-response relationships and the
loads or levels, 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 most widely cited aspects of causality in epidemiology, and public health, in
general, were articulated by Sir Austin Bradford Hill (1965) and have been widely used (CDC. 2004;
I ARC. 2006; NRC. 2004; Samet & C. C. Bodurow. 2008; U.S. EPA. 2005a). Several adaptations of
the Hill aspects have been used in aiding causality judgments in the ecological sciences (Adams.
2003; Collier. 2003; Fox. 1991; Gerritsen et al. 1998). These aspects (Hill. 1965) have been
modified (Table 1-1) for use in causal determinations specific to health and welfare effects or
pollutant exposures.2 Some aspects are more likely than others to be relevant for evaluating evidence
on the health or environmental effects of criteria air pollutants. For example, the analogy aspect does
not always apply and specificity would not be expected for multi-etiologic health outcomes such as
asthma or cardiovascular disease, or ecological effects related to acidification. Aspects that usually
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.
2 The Hill aspects were developed for interpretation of epidemiologic results. They have been modified here for use with a broader array of
data, i.e., epidemiologic, controlled human exposure, 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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play a larger role in determination of causality are consistency of results across studies, coherence of
effects observed in different study types or disciplines, biological plausibility, exposure-response
relationship, and evidence from "natural" experiments.
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 important considerations 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 1-1 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 (CDC. 2004).
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Table 1-1. Aspects to aid in judging causality
Aspect
Decription
Consistency of the
observed association
Coherence
Biological plausibility
Biological gradient
(exposure-response
relationship)
Strength of the
observed association
Experimental evidence
Temporal relationship
of the observed
association
An inference of causality is strengthened when a pattern of elevated risks is observed across several
independent studies, conducted in multiple locations by multiple investigators. 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.
An inference of causality from epidemiologic associations may be strengthened by other lines of evidence
(e.g., controlled human exposure and animal toxicological studies) that support a cause-and-effect
interpretation of the association. Causality is also supported when epidemiologic associations are reported
across study designs and across related health outcomes. 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 paleological/ historical reconstructions). The
coherence of evidence from various fields greatly adds to the strength of an inference of causality. The
absence of other lines of evidence, however, is not a reason to reject causality
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. A lack of biological
understanding, however, is not a reason to reject causality.
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). There are,
however, many possible reasons that a study may fail to detect an exposure-response relationship. Thus,
although the presence of a biological gradient may support causality, the absence of an exposure-response
relationship does not exclude a causal relationship.
The finding of large, precise risks increases confidence that the association is not likely due to chance, bias,
or other factors. However, given a truly causal agent, a small magnitude in the effect could follow from a
lower level of exposure, a lower potency, or the prevalence of other agents causing similar effects. While
large effects support causality, modest effects therefore do not preclude it.
The strongest evidence for causality can be provided when a change in exposure brings about a change in
occurrence or frequency of health or welfare effects.
Evidence of a temporal sequence between the introduction of an agent and appearance of the effect
constitutes another argument in favor of causality.
Specificity of the
observed association
Analogy
As originally intended, this refers to increased inference of causality if one cause is associated with a single
effect or disease (Hill. 1965). Based on the current understanding this is now considered one of the weaker
guidelines for causality; for example, many agents cause respiratory disease and respiratory disease has
multiple causes. At the scale of ecosystems, as in epidemiology, complexity is such that single agents
causing single effects, and single effects following single causes, are extremely unlikely. The ability to
demonstrate specificity under certain conditions remains, however, a powerful attribute of experimental
studies. Thus, although the presence of specificity may support causality, its absence does not exclude it.
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.
1.6.5. Determination of Causality
In the ISA, EPA assesses the results of recent relevant publications, building upon evidence
available during the previous NAAQS review, to draw conclusions on the causal relationships
between relevant exposures or body burden, as measured by blood or tissue Pb levels, and health
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effects and relevant Pb concentrations and environmental effects. This ISA uses a five-level
hierarchy that classifies the weight of evidence for causation, not just association1; that is, whether
the weight of scientific evidence makes causation at least as likely as not, in the judgment of the
reviewing group. 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 (2008). EPA's Guidelines for Carcinogen Risk Assessment (2005a), and the U.S. Surgeon
General's smoking reports (CDC. 2004). In the ISA, EPA uses a series of five descriptors to
characterize the weight of evidence for causality. 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 1-2.
For the Pb ISA, determination of causality involved the evaluation of evidence for different
types of health effects associated with Pb biomarkers of exposure and body burden (i.e., blood and
tissue). In making determinations of causality for Pb, evidence was evaluated for health outcome
categories, such as neurological effects, and then conclusions were drawn based upon the integration
of evidence from across disciplines (e.g., epidemiology and toxicology) and also across the suite of
related individual health outcomes. To accomplish this integration, evidence from multiple and
various types of studies was considered. Response was evaluated over a range of observations which
was determined by the type of study and methods of exposure or dose and response measurements.
Results from different protocols were compared and contrasted.
1 It should be noted that the 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 (U.S. EPA. 2005a) and to provide a more
nuanced set of categories.
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Table 1-2. Weight of evidence for causal determination
Determination
Health Effects
Ecological and Welfare Effects
Causal
relationship
Evidence is sufficient to conclude that there is a causal
relationship with relevant blood or tissue Pb levels. That is,
blood or tissue Pb levels have been shown to result in
health effects in studies in which chance, bias, and
confounding could be ruled out with reasonable
confidence. For example: a) controlled human exposure
studies that demonstrate consistent effects; or b)
observational studies that cannot be explained by
plausible alternatives or are supported by other lines of
evidence (e.g., animal studies or mode of action
information). Evidence includes replicated and consistent
high-quality studies by multiple investigators.
Evidence is sufficient to conclude that there is a causal
relationship with relevant pollutant exposures. That is, the
pollutant has been shown to result in effects in studies in
which chance, bias, and confounding could be ruled out
with reasonable confidence. Controlled exposure studies
(laboratory or small- to medium-scale field studies) provide
the strongest evidence for causality, but the scope of
inference may be limited. Generally, determination is
based on multiple studies conducted by multiple research
groups, and evidence that is considered sufficient to infer
a causal relationship is usually obtained from the joint
consideration of many lines of evidence that reinforce
each other.
Likely to be a
causal
relationship
Evidence is sufficient to conclude that a causal
relationship is likely to exist with relevant blood or tissue
Pb levels, but important uncertainties remain. That is,
blood or tissue Pb levels have been shown to result in
health effects in studies in which chance and bias can be
ruled out with reasonable confidence but potential issues
remain. For example: a) observational studies show an
association, but confounding factors are difficult to
address and/or other lines of evidence (controlled human
exposure, animal, or mode of action information) are
limited or inconsistent; or b) animal toxicological evidence
from multiple studies from different laboratories that
demonstrate effects, but limited or no human data are
available. Evidence generally includes 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.
Suggestive
of a causal
relationship
Evidence is suggestive of a causal relationship with
relevant blood or tissue Pb levels, but is limited because
chance, bias and confounding cannot be ruled out. For
example, at least one high-quality epidemiologic study
shows an association with a given health outcome but the
results of other studies are inconsistent.
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 blood or tissue Pb levels.
The available studies are of insufficient quantity, quality,
consistency or statistical power to permit a conclusion
regarding the presence or absence of an effect.
The available studies are of insufficient quality,
consistency or statistical power to permit a conclusion
regarding the presence or absence of an effect.
Not likely to
be a causal
relationship
Evidence is suggestive of no causal relationship with
relevant blood or tissue Pb levels. Several adequate
studies, covering the full range of levels of exposure that
human beings are known to encounter and considering
susceptible populations, are mutually consistent in not
showing an effect at any level of exposure.
Several adequate studies, examining relationships with
relevant exposures, are consistent in failing to show an
effect at any level of exposure.
1 In drawing judgments regarding causality for the criteria air pollutants, EPA typically focuses
2 on evidence of effects at relevant pollutant exposures. For making causality judgments for Pb health
3 effects, the focus is on evidence of exposure or body burden as indicated by relevant (within one
4 order of magnitude) blood or tissue Pb levels of the current U.S. population (median blood Pb level
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= 1.3; 95th percentile = 4.1; 99th percentile = 7.2). Studies of the efficacy of chelation therapy in Pb-
poisoned children and toxicological studies in which Pb levels were sufficiently high to induce an
overtly toxic response in the animals (e.g., mortality) were specifically excluded. Studies of workers
exposed to Pb in occupational settings were generally considered in the causal determinations.
Building upon the determination of causality are questions relevant to quantifying health or
environmental risks based on our understanding of the quantitative relationships between pollutant
exposures or biomarkers and health or welfare effects. While the causality determination is based
primarily on evaluation of health or environmental effects evidence, EPA also evaluates evidence
related to the doses or biomarker levels at which effects are observed. Considerations relevant to
evaluation of quantitative relationships for health and environmental effects are summarized below.
1.6.5.1. Effects on Human Populations
Once a determination is made regarding the causal relationship between the pollutant and
outcome category, important questions regarding quantitative relationships include:
¦ What is the concentration-response, exposure-response, or dose-response relationship in
the human population?
¦ What is the interrelationship between incidence and severity of effect?
¦ What exposure conditions (dose or exposure, duration and pattern) are important?
¦ What populations appear to be differentially affected (i.e., more susceptible to effects)?
To address these questions, the entirety of policy-relevant quantitative evidence is evaluated to
best quantify those concentration-response relationships that exist. For Pb, evaluation of blood or
tissue Pb concentrations at which effects were observed for exposed populations, including
potentially susceptible populations, has been an important element of this process. The integration of
evidence resulted in identification of a study or set of studies that best approximated the
concentration-response relationships between blood Pb and various health outcomes, given the
current state of knowledge and the uncertainties that surrounded these estimates. To accomplish this,
evidence is considered from multiple and diverse types of studies. To the extent available, the ISA
evaluates results from across epidemiologic studies that use various methods to evaluate the form of
relationships between blood or tissue Pb concentrations and health outcomes and draws conclusions
on the most well-supported shape of these relationships. Animal data may also inform evaluation of
concentration-response relationships, particularly relative to MOAs and characteristics of susceptible
populations. Chapter 2 presents the integrated findings informative for evaluation of population
risks.
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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 full
concentration range encountered or if nonlinear relationships exist along any part of this range. In
general, the shape of the concentration-response curve varies, depending on the type of health
outcome, underlying MOA and dose. At the human population level, however, various sources of
variability and uncertainty, such as the low data density at the lowest blood Pb levels, possible
influence of exposure measurement error, and individual differences in susceptibility to Pb health
effects, tend to smooth and "linearize" the concentration-response function. In addition, many
chemicals and agents may act by perturbing naturally occurring background processes that lead to
disease, which also linearizes population concentration-response relationships ( Clewell & Crump.
2005; Crump et al. 1976; Hoel. 1980). These attributes of population dose-response may explain
why the available human data at ambient concentrations for some environmental pollutants (e.g., Pb,
PM, 03, environmental tobacco smoke [ETS], radiation) do not exhibit evident 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 (Roth man & Greenland. 1998). Of particular
interest for Pb is the shape of the concentration- response curve at the low end (<10 (xg/dL) of
current blood Pb concentrations observed in the U.S. population.
Publication bias is a source of uncertainty regarding the magnitude of health risk estimates. It
is well understood that studies reporting non-null findings are more likely to be published than
reports of null findings, and publication bias can also result in overestimation of effect estimate sizes
(loannidis. 2008). For example, effect estimates from single-city epidemiologic studies have been
found to be generally larger than those from multicity studies (Anderson et al. 2005).
Finally, identification of the susceptible population groups contributes to an understanding of
the public health impact of pollutant exposures. In this ISA, the term "susceptible population" will
be used as an overarching concept to encompass populations variously described as susceptible,
vulnerable, or sensitive. "Susceptible populations" is defined here as those populations that have a
greater likelihood of experiencing health effects related to exposure to an air pollutant (e.g., Pb) due
to a variety of factors including but not limited to: genetic or developmental factors, race, gender,
lifestage, lifestyle (e.g., smoking status and nutrition) or preexisting disease; as well as population-
level factors that can increase an individual's exposure to an air pollutant (e.g., Pb) such as
socioeconomic status [SES], which encompasses reduced access to health care, low educational
attainment, residential location, and other factors. Epidemiologic studies can help identify
susceptible populations by evaluating health responses in the study population. Examples include
stratified analyses for subsets of the population under study or testing for interactions or effect
modification by factors such as gender, age group, or health status. Experimental studies using
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animal models of susceptibility or disease can also inform the extent to which health risks are likely
greater in specific population groups. Further discussion of these groups is presented in Chapter 6.
1.6.5.2. 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?
¦ Under what exposure conditions (amount deposited or concentration, duration and
pattern) are effects observed?
¦ 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 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. For example, an ecological response to deposition of a given pollutant can
differ greatly between ecosystems. Where results from greenhouse or animal ecotoxicological
studies are available, they may be used to aid in characterizing exposure-response relations,
particularly relative to mechanisms of action, and characteristics of sensitive biota.
1.6.6. Concepts in Evaluating Adversity of Health Effects
In evaluating the health evidence, a number of factors can be considered in determining the
extent to which health effects are "adverse" for health outcomes such as changes in lung function or
in cardiovascular health measures. Some health outcome events, such as hospitalization for
respiratory or cardiovascular diseases, are clearly considered adverse; what is more difficult is
determining the extent of change in the more subtle health measures that is adverse. What constitutes
an adverse health effect may vary between populations. Some changes in healthy individuals may
not be considered adverse while those of a similar type and magnitude are 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). This statement updated the guidance for defining adverse
respiratory health effects that had been published 15 years earlier (ATS. 1985). taking into account
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new investigative approaches used to identify the effects of air pollution and reflecting concern for
impacts of air pollution on specific susceptible groups. In the 2000 update, there was an increased
focus on quality of life measures as indicators of adversity and a more specific consideration of
population risk. Exposure to air pollution that increases the risk of an adverse effect to the entire
population is viewed as adverse, even though it may not increase the risk of any identifiable
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.
1.7. Summary
This draft ISA is a concise evaluation and synthesis of the most policy-relevant science for
reviewing the NAAQS for Pb, and it is the chief means for communicating the critical science
judgments relevant to that NAAQS review. It reviews the most policy-relevant evidence from
atmospheric science, exposure, and health and environmental effects studies and includes
mechanistic evidence from basic biological science. A framework for making critical judgments
concerning causality was presented in this chapter. It relies on a widely accepted set of principles and
standardized language to express evaluation of the evidence. This approach can bring rigor and
clarity to current and future assessments. Once complete, the ISA should assist EPA and others, now
and in the future, to accurately represent what is presently known and what remains unknown
concerning the effects of Pb on human health and public welfare.
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Chapter 1 References
Adams. S. M. (2003). Establishing causality between environmental stressors and effects on aquatic ecosystems. Human
and Ecological Risk Assessment, 9, 17-35. http://dx.doi.org/10.1080/71360985Q
Anderson. H. R.. Atkinson. R. W.. Peacock. J. L.. Sweeting. M. J.. & Marston. L. (2005). Ambient particulate matter and
health effects: Publication bias in studies of short-term associations. Epidemiology, 16(2), 155-163.
http://dx.doi.org/10.1097/01.ede.000Q152528.22746.0f
ATS. (American Thoracic Society). (1985). Guidelines as to what constitutes an adverse respiratory health effect, with
special reference to epidemiologic studies of air pollution. American Review of Respiratory Disease, 131, 666-
668. http://www.ncbi.nlm.nih.gov/pubmed/3994164
ATS. (American Thoracic Society). (2000). What constitutes an adverse health effect of air pollution? This official
statement of the American Thoracic Society was adopted by the ATS Board of Directors, July 1999. American
Journal of Respiratory and Critical Care Medicine, 161, 665-673.
Blair. A.. Burg. J.. Foran. J.. Gibb. H.. Greenland. S.. Morris. R Zimmerman. R. (1995). Guidelines for application of
meta-analysis in environmental epidemiology. Regulatory Toxicology and Pharmacology, 22, 189-197.
http://dx.doi.org/10.1006/rtph.1995.1084
CDC. (Centers for Disease Control and Prevention). (2004). The health consequences of smoking: A report of the Surgeon
General. Washington, DC: U.S. Department of Health and Human Services.
Clewell. H. J.. & Crump. K. S. (2005). Quantitative estimates of risk for noncancer endpoints. Risk Analysis, 25(2), 285-
289. http://dx.d0i.0rg/l0.llll/j. 1539-6924.2005.00589.x
Collier. T. K. (2003). Forensic ecotoxicology: Establishing causality between contaminants and biological effects in field
studies. Human and Ecological Risk Assessment, 9, 259-266. http://dx.doi.org/10.1080/713609862
Crump. K. S.. Hoel. D. G.. Langlev. C. H.. & Peto. R. (1976). Fundamental carcinogenic processes and their implications
for low dose risk assessment. Cancer Research, 36(9 pt. 1), 2973-2979.
http://www.ncbi.nlm.nih.gov/pubmed/975067
Fox. G. A. (1991). Practical causal inference for ecoepidemiologists. Journal of Toxicology and Environmental Health,
Part A: Current Issues, 55(4), 359-373. http://dx.doi.org/10.1080/15287399109531535
Gee. G. C.. & Pavne-Sturges. D. C. (2004). Environmental health disparities: A framework integrating psychosocial and
environmental concepts. Environmental Health Perspectives, 112(11), 1645-1653.
http://dx.doi.org/10.1289/ehp.7074
Gerritsen. J.. Carlson. R. E.. Dvcus. D. L.. Faulkner. C.. Gibson. G. R.. Harcum. J.. & Markowitz. S. A. (1998). Lake and
reservoir bioassessment and biocriteria: Technical guidance document. (Report No. EPA 841-B-98-007).
Washington, DC: U.S. Environmental Protection Agency, Office of Water. Retrieved from
http://www.epa.gov/owow/monitoring/tech/lakes.html.
Hill. A. B. (1965). The environment and disease: Association or causation? Proceedings of the Royal Society of Medicine,
58(5), 295-300. http://www.ncbi.nlm.nih.gov/pubmed/14283879
Hoel. D. G. (1980). Incorporation of background in dose-response models. Federation Proceedings, 59(1), 73-75.
http://www.ncbi.nlm.nih. gov/pubmed/7351247
I ARC. (International Agency for Research on Cancer). (2006). IARC monographs on the evaluation of carcinogenic risks
to humans: Preamble. Lyon, France: World Health Organization. Retrieved from
http://monographs.iarc.fr/ENG/Preamble/CurrentPreamble.pdf.
ICF. (ICF International). (2006). Lead human exposure and health risk assessments and ecological risk assessment for
selected areas: Pilot phase: External review draft technical report. Research Triangle Park, NC: U.S.
Environmental Protection Agency, Office of Air Quality Planning and Standards.
Ioannidis. J. P. A. (2008). Why most discovered true associations are inflated. Epidemiology, 19(5), 640-648.
http://dx.doi.org/10.1097/EDE.0b013e31818131e7
NAPAP. (National Acid Precipitation Assessment Program). (1991). The experience and legacy of NAPAP: Report of the
Oversight Review Board of the National Acid Precipitation Assessment Program. Washington, DC: Author.
NRC. (National Research Council). (2004). Research priorities for airborne particulate matter: IV: Continuing research
progress. Washington, DC: National Academy Press.
Rothman. K. J.. & Greenland. S. (1998). Modern epidemiology (2nd ed.). Philadelphia, PA: Lippincott-Raven Publishers.
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Samet. J. M.. & C. C. Bodurow (Eds.). (Institute of Medicine). (2008). Improving the presumptive disability decision-
making process for veterans. Washington, DC: National Academies Press.
U.S. EPA. (U.S. Environmental Protection Agency). (1977). Air quality criteria for lead. (Report No. EPA-600/8-77-017).
Washington, DC: U.S. Environmental Protection Agency, Office of Research and Development. Retrieved from
http://www.ntis.gov/search/product.aspx?ABBR=PB280411.
U.S. EPA. fUS. Environmental Protection Agency). (1986a). Air quality criteria for lead. (Report No. EPA/600/8-83/028
aF-dF). Washington, DC: Author.
U.S. EPA. (U.S. Environmental Protection Agency). (1986b). Lead effects on cardiovascular function, early development,
and stature: An addendum to U.S. EPA Air Quality Criteria for Lead (1986). (Report No. EPA-600/8-83/028aF).
Washington, DC: U.S. Environmental Protection Agency, Office of Research and Development.
U.S. EPA. (U.S. Environmental Protection Agency). (1989). Review of the national ambient air quality standards for lead:
Exposure analysis methodology and validation: OAQPS staff report. (Report No. EPA-450/2-89-011). Research
Triangle Park, NC: U.S. Environmental Protection Agency, Office of Air Quality Planning and Standards.
U.S. EPA. (U.S. Environmental Protection Agency). (1990a). Air quality criteria for lead: Supplement to the 1986
addendum. (Report No. EPA/600/8-89/049F). Washington, DC: U.S. Environmental Protection Agency, Office of
Research and Development.
U.S. EPA. (U.S. Environmental Protection Agency). (1990b). Review of the national ambient air quality standards for
lead: Assessment of scientific and technical information: OAQPS staffpaper. (Report No. EPA-450/2-89-022).
Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Air Quality Planning and
Standards.
U.S. EPA. (U.S. Environmental Protection Agency). (1991). Strategy for reducing lead exposures. Washington, DC:
Author. Retrieved from http: //www, epa. gov/ttn/naaq s/standards/pb/data/leadstrategv 1991 .pdf.
U.S. EPA. (U.S. Environmental Protection Agency). (2004). Air quality criteria for particulate matter. (Report No.
EPA/600/P-99/002aF-bF). Research Triangle Park, NC: U.S. Environmental Protection Agency, National Center
for Environmental Assessment. Retrieved from http://cfpub.epa.gov/ncea/cfm/recordisplav.cfm?deid=87903.
U.S. EPA. fUS. Environmental Protection Agency). (2005a). Guidelines for carcinogen risk assessment. (Report No.
EPA/630/P-03/001F). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum.
Retrieved from http: //www, epa. gov/cancerguidelines/.
U.S. EPA. (U.S. Environmental Protection Agency). (2005b). Project work plan for revised air quality criteria for lead.
(Report No. NCEA-R-1465). Research Triangle Park, NC: U. S. Environmental Protection Agency. Retrieved
from http://cfpub.epa.gov/ncea/cfm/recordisplav.cfm?deid=l 13963.
U.S. EPA. (U.S. Environmental Protection Agency). (2006a). Air quality criteria for lead. (Report No. EPA/600/R-
05/144aF-bF). Research Triangle Park, NC: U.S. Environmental Protection Agency, National Center for
Environmental Assessment. Retrieved from http://cfpub.epa.gov/ncea/CFM/recordisplav.cfm?deid=l 58823.
U.S. EPA. (U.S. Environmental Protection Agency). (2006b). Analysis plan for human health and ecological risk
assessment for the review of the lead national ambient air quality standards (draft). Research Triangle Park, NC:
Author. Retrieved from http://www.epa.gOv/ttn/naaas/standards/pb/s pb cr pd.html.
U.S. EPA. (U.S. Environmental Protection Agency). (2006c). Plan for review of the national ambient air quality standards
for lead. Research Triangle Park, NC: Author. Retrieved from
http://www.epa.gOv/ttn/naaas/standards/pb/s pb cr pd.html.
U.S. EPA. (U.S. Environmental Protection Agency). (2006d). Review of the national ambient air quality standards for
lead: Policy assessment of scientific and technical information: OAQPS staffpaper - first draft. (Report No. EPA-
452/P-06-002). Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Air Quality
Planning and Standards.
U.S. EPA. (U.S. Environmental Protection Agency). (2007). Lead: Human exposure and health risk assessments for
selected case studies: Volume 1: Human exposure and health risk assessments - full-scale. (Report No. EPA-
452/R-07-014a). Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Air Quality
Planning and Standards. Retrieved from http ://www.ntis.gov/search/product.aspx?ABBR=PB2008102573.
U.S. EPA. (U.S. Environmental Protection Agency). (2011). Integrated review plan for the national ambient air quality
standards for lead: External review draft. (Report No. EPA-452/D-11-001). Research Triangle Park,NC: U.S.
Environmental Protection Agency, Office of Research and Development.
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Chapter 2 Contents
Chapter 2. Integrative Health and Ecological Effects Overview 2-1
2.1. Ambient Lead: Source to Concentration 2-2
2.1.1. Sources, Fate and Transport of Ambient Lead 2-2
2.1.2. Monitoring and Concentrations of Ambient Air Lead 2-3
2.1.3. Ambient Lead Concentrations in Non-Air Media and Biota 2-5
2.2. Exposure to Ambient Lead 2-5
2.3. Toxicokinetics 2-7
2.4. Lead Biomarkers 2-8
2.5. Health Effects 2-9
Table 2-1. Summary of causal determinations between exposure to Pb and health
outcomes 2-10
2.5.1. Neurological Effects 2-10
Figure 2-1. Snapshot of evidence for the spectrum of effects to the nervous system
associated with Pb exposure. Green=animal toxicological studies (left
side); purple=epidemiological studies (right side). 2-13
2.5.2. Cardiovascular Effects 2-13
2.5.3. Renal Effects 2-15
2.5.4. Immune System Effects 2-17
2.5.5. Heme Synthesis and RBC Function 2-18
2.5.6. Reproductive Effects and Birth Outcomes 2-20
2.5.7. Effects on Other Organ Systems 2-21
2.5.8. Cancer 2-23
2.5.9. Human Health Effects and Corresponding Blood Pb Levels 2-24
Table 2-2. Summary of Pb-induced health effects in children and the lowest mean
blood Pb level in the population(s) studied 2-25
Table 2-3. Summary of Pb-induced health effects in adults and the lowest mean blood
Pb level in the population(s) studied 2-26
2.6. Ecological Effects 2-26
Table 2-4. Summary of causal determinations for Pb in plants, vertebrates and
invertebrates 2-27
2.6.1. Summary of Terrestrial Ecosystem Effects 2-28
2.6.2. Summary of Aquatic Ecosystem Effects 2-30
2.6.3. Bioaccumulation of Lead in Terrestrial and Aquatic Biota as it Affects Ecosystem Services 2-33
2.6.4. Mortality 2-34
2.6.5. Growth Effects 2-35
2.6.6. Physiological Stress 2-35
2.6.7. Hematological Effects 2-36
2.6.8. Developmental and Reproductive Effects 2-37
2.6.9. Neurobehavioral Effects 2-38
2.6.10. Community and Ecosystem Level Effects 2-39
2.6.11. Ecological Effects and Corresponding Pb Concentrations 2-40
2.7. Integration of Health and Ecological Effects Overview 2-41
Table 2-5. Summary of Causal Determinations for Health and Ecological Effects 2-42
2.7.1. Modes of Action Relevant to Downstream Health and Ecological Effects 2-42
Table 2-6. Related human health effects resulting from the MO As ofPb and the lowest
level eliciting theMOA 2-43
2.8. Policy Relevant Considerations and Human Health 2-46
2.8.1. Air-to-Blood Relationships 2-46
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Table 2-7. Summary ofEstimated Slopes for Blood Pb toAirPb Relationships in
Children 2-48
2.8.2. Concentration-Response Functions 2-48
Figure 2-2. Comparison of associations between bloodPb and cognitive function among
various bloodPb strata. 2-49
2.8.3. Timing and Duration of Exposure 2-52
Figure 2-3. Associations of cognitive function in children with different degrees of
changes in bloodPb levels over time. 2-53
2.8.3.1. Persistence of Effects 2-54
2.8.4. Susceptible Populations and Lifestages 2-54
2.8.4.1. Children 2-55
2.8.4.2. Adults 2-55
2.8.4.3. Sex 2-55
2.8.4.4. Race and Ethnicity 2-56
2.8.4.5. Socioeconomic Status 2-56
2.8.4.6. Genes 2-57
2.8.4.7. Pre-existing Conditions 2-57
2.8.4.8. Nutrition and Lifestyle Factors 2-57
2.8.4.9. Stress and Cognitive Reserve 2-58
2.8.4.10. Co-exposure of Lead with Metals or Other Chemicals 2-58
2.9. Summary 2-59
Table 2-8. Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb 2-60
Chapter 2 References 2-65
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Chapter 2. Integrative Health and
Ecological Effects Overview
The subsequent chapters of this ISA will present the most policy relevant information related to this
review of the science supporting the NAAQS for Pb. This chapter integrates the key findings from the
disciplines evaluated in this current assessment of the Pb scientific literature, which includes studies of Pb
sources, fate and transport of Pb, ambient air concentrations, exposure assessments, toxicokinetics,
biomarkers and models of Pb burden, health (e.g., both toxicology and epidemiology), and ecological
effects of Pb. The EPA framework for causal determinations described in Chapter 1 has been applied to
the body of scientific evidence in order to collectively examine the health and ecological effects attributed
to Pb exposure in a two-step process. The first step is to establish causal relationships followed by
identification of concentration-response relationships
As described in Chapter 1, EPA assesses the results of recent relevant publications, building upon
evidence available during the previous NAAQS reviews, to draw conclusions on the causal relationships
between relevant pollutant exposures and health or environmental effects. 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:
¦ What is the concentration-response, exposure-response, or dose-response relationship in the
human population?
Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and Environmental
Research Online) at http://eDa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of developing science
assessments such as the Integrated Science Assessments (ISA) and the Integrated Risk Information System (IRIS).
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¦ What exposure conditions (dose or exposure, exposure pathways, duration and pattern) are
important?
¦ What populations and lifestages appear to be differentially affected i.e., more susceptible to
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?
To address these questions, in the second step of the EPA framework, the entirety of quantitative
evidence is evaluated to identify and characterize potential concentration-response relationships. This
requires the evaluation of the levels of the pollutant and the exposure durations at which effects were
observed for exposed populations including potentially susceptible populations.
This chapter summarizes and integrates the newly available scientific evidence that best informs
consideration of the policy-relevant questions that frame this assessment, presented in Chapter 1.
Section 2.1 discusses the trends in ambient Pb sources and concentrations, including the fate and transport
of Pb in the environment, and provides a brief summary of the topics covered. Section 2.2 presents the
evidence regarding exposure to ambient Pb and describes air-related Pb exposure pathways. Section 2.3
provides a discussion of the toxicokinetics of Pb and Pb biomarkers are discussed in Section 2.4.
Section 2.5 summarizes effects of Pb on specific health endpoints and Section 2.6 summarizes the
evidence relating to the ecological effects of Pb. Section 2.7 integrates the scientific evidence across
various health and ecological endpoints highlighting common modes of action where applicable. Finally,
Section 2.8 provides a discussion of policy relevant considerations including air-to-blood relationships,
concentration-response relationships, timing and duration of exposure and susceptible populations.
Section 2.9 is a summary of the health and ecological effects of Pb.
2.1. Ambient Lead: Source to Concentration
2.1.1. 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. 2006). 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 1990 and
then by an additional 77% from 1990 to 2008, at which time emissions were 1,200 tons per year. Data
from the 2008 National Emissions Inventory (NEI) (U.S. EPA. 2011) illustrate that piston engine aircraft
emissions now comprise the largest share (-49%) of total atmospheric Pb emissions; the 2008 NEI
estimated that 590 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
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related sources, and historic Pb. Chemical speciation of Pb had also been fairly well characterized in the
2006 Pb AQCD (2006). Estimates from the 1986 Pb AQCD (U.S. EPA. 1986. 2006) for organic
automotive Pb emissions provides an upper bound for organic vapor emissions of 20% of total Pb
dibromide and Pb bromide emissions from piston engine aircraft. Recent speciation studies of smelting
and battery-recycling operations have shown that Pb sulfide and Pb sulfates are abundant within the
emissions mixture for such industrial operations.
The atmosphere is the main environmental transport pathway for Pb, and on a global scale
atmospheric Pb is primarily associated with fine particulate matter (PM). Global atmospheric Pb
deposition peaked in the 1970s, followed by a more recent decline. On a local scale, Pb concentrations in
soils (including urban areas where historic use was widespread) can be substantial, and coarse Pb-bearing
PM experiences cycles of deposition and resuspension that serve to distribute it. 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. 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. Resuspension by wind and traffic can be an important
source of airborne Pb near sources where Pb occurs in substantial amounts in surface dust.
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, leading
to high soil concentrations, while fine Pb-bearing PM can be transported long distances, leading to
contamination of remote areas. Surface waters act as an important reservoir, with Pb lifetimes 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 without becoming airborne. Pb
containing sediment particles can be remobilized into the water column, and sediment concentrations tend
to follow those in overlying waters.
2.1.2. Monitoring and Concentrations of Ambient Air Lead
In recognition of the role of all PM sizes in ambient air Pb exposures, including the ingestion of
particles deposited onto surfaces, the indicator for the Pb NAAQS is Pb in total suspended particulate
(Pb-TSP). Although there is a lower rate of error in estimating ambient Pb from Pb-PM10 monitoring than
from Pb-TSP monitoring, the Pb-TSP indicator was retained in 2008 because ingestion after deposition in
the upper respiratory tract was considered an important component of Pb exposure. A new federal
reference method (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. Several FEMs
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have also been approved. The FRM is based on flame AAS. ICPMS is under consideration as a new FRM
for Pb-TSP.
Monitoring for ambient Pb levels is required for all areas where Pb levels have been shown or are
expected to contribute to maximum concentrations of 0.10 |ig/m3 or greater over a 3-year time period. Pb
is monitored routinely at state and local air monitoring stations (SLAMS) that report data used for
NAAQS compliance to the air quality system (AQS) database. Pb monitoring requirements have
experienced several changes since publication of the 2006 Pb AQCD (U.S. EPA. 2006). 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, and is planned for the national core multipollutant monitoring
network (NCore) network. While monitoring in multiple networks provides extensive geographic
coverage, measurements between networks are not directly comparable in all cases because different
particle size ranges are sampled in different networks. Depending on monitoring network, Pb is monitored
in TSP, PMio, or PM2 5. Monitors reporting to the AQS were considered for the purpose of this ISA 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. 2008). Non-source oriented
monitors were those monitors not considered to be source oriented.
Ambient air Pb concentrations have declined drastically over the period 1980-2009. The median
annual concentrations have dropped by 97% from 0.87 (ig/m3 in 1980 to 0.025 (ig/m3 in 2009. While the
sharpest drop in Pb concentration occurred during 1980-1990, a declining trend was observed between
1990 and 2009. Compared to 1980-1990, a smaller reduction was observable among source oriented Pb
concentration (56%) and non-source oriented Pb data (51%) for 2000-2009.
AQS data for source oriented and non-source oriented monitoring were analyzed for 2007-2009.
For source oriented monitoring, the three-month rolling average was measured to be above the level of
the NAAQS in fourteen 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/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. PM size
distribution also influenced how far the particle will travel upon initial emission or resuspension before
being deposited. Additionally, resuspension and urban background levels of Pb were uncertain influential
factors of ambient Pb concentrations. Given variability in these conditions, it was very difficult to predict
how Pb concentration varies over time and space. This was consistent with field studies to characterize Pb
concentrations that were described in the literature.
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Size distribution of Pb-bearing PM was demonstrated to vary substantially for several studies
presented, 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
explained by differences in sources across land use types. Additionally, Pb concentrations exhibited
varying degrees of association with other criteria pollutant concentrations. Overall, Pb was moderately
associated with PM2 5, PMi0 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. Br, Cu, and K concentrations also exhibited moderate associations with Pb concentrations. Such
correlations may suggest some common sources affecting the concentrations of various pollutants.
2.1.3. 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 have also been observed. 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.
2.2. Exposure to Ambient Lead
Exposure data considered in this assessment build upon the conclusions of the 2006 Pb AQCD
(2006). which found air Pb concentrations in the U.S. and associated biomarkers of exposure to have
decreased substantially following the ban on Pb in gasoline as well as earlier bans on Pb in house-hold
paints and solder. Pb exposure is difficult to assess because Pb has multiple sources in the environment
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and passes through various media. The atmosphere is the main environmental transport pathway for Pb,
and 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, and a fraction of that suspended Pb may originate from waters used
to irrigate the soil. Air-related pathways of Pb exposure are the focus of this assessment. In general, air-
related pathways include those pathways where Pb passes through ambient air on its path from a source to
human exposure. 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 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 non-air exposures are reviewed
in this section because they can also contribute to Pb body burden.
Section 4.1 presents data illustrating potential exposure mechanisms. Several studies suggested that
soil can act as a reservoir for historically deposited and contemporaneous Pb emissions from industrial or
other activities. Exposure to soil contaminated with deposited Pb can occur through resuspended PM as
well as shoe tracking and hand-to-mouth contact. In general, soil Pb concentrations tended to be higher
within inner-city communities compared with suburban neighborhoods. Infiltration of Pb dust 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 to Pb is typically higher than indoor or outdoor ambient Pb concentrations. These
findings regarding personal exposure may be related resuspension of Pb that occurs with body movement.
Observational studies using biomarkers of Pb as exposure metrics are also included in Section 4.1.
The median blood Pb level for the entire U.S. population is 1.2 (ig/dL and the 95th percentile blood Pb
level is 3.7 (ig/dL, based on 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.
Overall, trends in blood Pb levels have been decreasing among U.S. children and adults over the past 20
years. Concurrent changes in isotopic ratios of blood Pb samples reflect changes in source composition
over the past several decades. Several studies have regressed blood Pb as a function of environmental Pb
samples such as air Pb or Pb dust fall. Recent studies have observed a relationship between blood Pb and
soil Pb concentration. Studies have suggested that blood Pb is associated with exposure to Pb paints in
older homes, Pb released into drinking water, and occupational work with materials containing Pb.
Studies that examine blood Pb as a function of ambient air Pb measurements are discussed in Section
2.8.1 that follows.
Sequential extraction has been used to estimate the gastric bioavailability of particle bound Pb after
exposure occurs. Findings from these studies have been mixed, ranging from 13 to 86%, but such
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variation is likely a function of the particle sizes from which the Pb was extracted as well as the acid
mixture used to simulate gastric juices. Estimates of bioavailability of inhaled organic Pb to the lungs are
available only from older studies in the literature and suggest a that it is possible for all inhaled organic
Pb to enter the blood stream (Chamberlain et al.. 1975).
2.3. 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. Saturable binding to RBC proteins contributes to an
increase in the plasma/blood Pb ratio with increasing blood Pb level concentration and curvature to the
blood Pb-plasma Pb relationship. As blood Pb level increases and the higher affinity binding sites for Pb
in RBCs become saturated at approximately 40 (ig/dL blood, 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 compartment
(bone) and a smaller fast compartment (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 & Wittmers. 1992). 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.
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 to 20. 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 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). Due to the more rapid turnover of bone mineral
in children, changes in blood Pb concentration are thought to more closely parallel changes in total body
burden. However, some Pb accumulated in bone during childhood does persist into later life. Potential
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mobilization of Pb from the skeleton could occur in adults at times of physiological stress associated with
enhanced bone remodeling such as during pregnancy and lactation, menopause or in older adulthood,
extended bed rest, hyperparathyroidism, and weightlessness. Regardless of age, however, similar blood
Pb concentrations in two individuals (or populations) do not necessarily translate to similar body burdens
or similar exposure histories.
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 an elimination 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 may become evident with longer observation periods
following a decrease in exposure due to the gradual redistribution of Pb among other compartments via
the blood. Therefore, a single blood Pb concentration may reflect the near-term or longer-term history of
the individual to varying degrees, depending on the relative contributions of internal (e.g., bone) and
external sources of Pb to blood Pb, which in turn will depend on the exposure history and possibly age-
related and individual-specific (e.g., pregnancy, lactation) characteristics of bone turnover. In general,
higher blood Pb concentrations can be interpreted as indicating higher exposures (or Pb uptakes);
however, they do not necessarily predict higher body burdens, especially in adults.
2.4. Lead Biomarkers
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
variables that may affect bone remodeling (e.g., nutritional status, pregnancy, and menopause). Blood Pb
in adults is typically more an index of recent exposures than body burden, whereas bone Pb is an index of
cumulative exposure and body burden. In children, 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 some physiological
circumstances (e.g., osteoporosis), bone Pb may contribute to blood Pb in adults. The disparity between
blood Pb and body burden may have important implications for the interpretation of blood Pb
measurements in some epidemiology studies. Conceptually, measurement of long-term Pb body burden
(i.e., based on tibia Pb) may be the appropriate metric if the effects of Pb on a particular outcome are
lasting and cumulative. However, if the effects of Pb on the outcome represent the acute effects of current
exposure, then blood Pb may be the preferred metric. In the absence of clear evidence as to whether a
particular outcome is an acute effect of recent Pb dose or a chronic effect of cumulative Pb exposure, both
blood and bone metrics should be considered.
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Cross-sectional studies that sample blood Pb once generally provide an index of recent exposures.
In contrast, cross-sectional studies of bone Pb and longitudinal samples of blood Pb concentrations over
time provide an index of cumulative exposure and are more reflective of average Pb body burdens over
time. 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.
The concentration of Pb in urine is a function of the urinary Pb excretion and the urine flow rate.
Urine flow rate requires collection of a timed urine sample, which is often problematic in epidemiologic
studies. Collection of un-timed ("spot") urine samples, a common alternative to timed samples, requires
adjustment of the Pb measurement in urine to account for variation in urine flow (Diamond. 1988V
Urinary Pb concentration reflects, mainly, the exposure history of the previous few months; thus, a single
urinary Pb measurement cannot distinguish between a long-term low level of exposure or a higher acute
exposure. 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
better reflect body burden.
2.5. Health Effects
This section evaluates the evidence from toxicological and epidemiologic studies that examined the
health effects associated with exposure to Pb. The results from the health studies evaluated in
combination with the evidence from other disciplines (e.g., fate and transport, exposure sciences,
toxicokinetics) contribute to the causal determinations (Section 1.6.4) 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 2-1). Although not presented in depth in this chapter, a 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 2-1. Summary of causal determinations between exposure to Pb and health
outcomes
Outcome
Causality Determination
Neurological Effects
Causal Relationship
Cardiovascular Effects
Causal Relationship
Renal Effects
Causal Relationship
Immune System Effects
Causal Relationship
Effects on Heme Synthesis and Red Blood Cell Function
Causal Relationship
Reproductive Effects and Birth Outcomes
Causal Relationship
Cancer
Likely Causal Relationship
2.5.1. Neurological Effects
The 2006 Pb AQCD concluded that the collective body of epidemiologic studies provides clear and
consistent evidence for the effects of Pb exposure on neurocognitive function in children. This conclusion
was substantiated by findings in diverse populations that blood Pb levels were associated with a broad
spectrum of cognitive and behavioral endpoints, including IQ, higher-order processes such as language
and memory, academic achievement, behavior and conduct, sensory acuities, and changes in brain
structure and activity as assessed by magnetic resonance imaging (MRI) or magnetic resonance
spectroscopy (MRS). Toxicological studies not only provided coherence with similarly consistent findings
for Pb-induced impairments in learning, behavior, and sensory acuities, but also provided biological
plausibility by characterizing mechanisms for Pb-induced neurotoxicity. These mechanisms included Pb-
induced inhibition of neurotransmitter release and decreases in synaptic plasticity, neuronal
differentiation, and blood-brain-barrier integrity. Both epidemiologic studies (in children) and
toxicological studies, demonstrated neurocognitive deficits in association with blood Pb levels at and
below 10 (ig/dL, and evidence from both disciplines supported a nonlinear exposure-response
relationship, with greater effects estimated for lower blood Pb levels. Among environmentally-exposed
adults, the most consistent findings were associations between cumulative Pb exposure, as assessed by
serial blood Pb or bone Pb measurements, and cognitive deficits.
Building on this strong body of extant evidence, recent studies continue to demonstrate
associations between Pb exposure and neurological effects. While recent epidemiologic studies in
children continued to demonstrate associations with IQ, most evidence emphasized associations of blood
Pb levels (as low as 2 (ig/dL) with specific indices of neurocognitive function such as reading and verbal
skills, memory, visuospatial processing, and academic achievement. Nonetheless, these newer findings
are concordant with the previous body of evidence given that IQ is a global measure of cognitive function
that reflects the integration of several neurocognitive domains. Additional coherence for findings in
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children is provided by evidence in animals that blood Pb levels of 1.8 (ig/dL and higher are associated
with decrements in learning and memory. New findings in animals emphasized the role of stress in
potentiating the low dose effects of Pb on behavior and memory. In animals, the developmental period is
the most sensitive window for Pb-dependent neurotoxicity, whereas in children, concurrent blood Pb was
generally found to be the best predictor of cognitive decrements.
Recent studies in children continue to support associations of Pb exposure (blood Pb levels 3-11
(ig/dL) with a range of behavioral problems from anxiety and distractibility to conduct disorder and
delinquent behavior. Whereas previous evidence was not compelling, new evidence indicates associations
comparing the lowest quartiles of blood Pb level (0.8-1 (ig/dL versus <0.8 (ig/dL) and ADHD. These
findings for ADHD are well supported by observations in animals of Pb-induced increased response rates
and impulsivity. Both epidemiologic studies in children and toxicological studies demonstrate
associations of Pb exposure with deficits in visual acuity and hearing and auditory processing. New
evidence from toxicological studies demonstrates these effects at lower exposure levels (blood Pb levels
<15 (ig/dL). Combined evidence for Pb-associated neurocognitive deficits (e.g., inattention, conduct
disorder, and effects on sensory function) provides plausible mechanisms by which Pb exposure may
contribute to academic underachievement and to more serious problems of delinquent behavior.
Studies of adults without occupational Pb exposure have not provided consistent evidence for
associations between blood Pb and the range of neurological effects. One explanation for the weaker
evidence may be that cognitive reserve may compensate for the effects of Pb exposure on learning new
information. Compensatory mechanisms may become less effective with increasing age, explaining the
consistent associations between measures of cumulative Pb exposure and neurocognitive deficits. Among
recent studies of adults, blood Pb and bone Pb are associated with essential tremor and Parkinson's
disease, respectively. Consistent with these findings, toxicological studies demonstrate Pb-induced
decreased dopaminergic cell activity in the substantia nigra, which contributes to the primary symptoms
of Parkinson's disease. Biological plausibility also is provided by observations of developmental Pb
exposures of monkeys and rats inducing neurodegeneration in the aged brain. A recent epidemiological
study indicated that early-life ALAD activity, a biomarker of Pb exposure, may be associated with
schizophrenia later in adulthood. Consistent with these findings, toxicological studies have observed Pb-
induced emotional changes in males and depression in females. It is not surprising that Pb exposure may
increase the risk of different neurological endpoints in children and adults given the predominance of age-
dependent neurological processes, in particular, neurogenesis and brain development in children and
neurodegeneration in adults.
Extensive evidence from toxicological studies, as well as evidence in some aquatic and terrestrial
animal taxa (Section 2.6.9) clearly substantiates the biological plausibility for epidemiologic findings by
characterizing mechanisms underlying neurological effects. Pb induces complex neurochemical changes
in the brain that differ by region of the brain, neurotransmitter type, age, and sex of the organism. These
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changes remain aberrant over time but are dynamic in nature. Pb exposure of animals induces changes in
the transmission of dopamine, which plays a key role in cognitive functions mediated by the prefrontal
cortex and in motor functions mediated by the substantia nigra. Current toxicological research has been
expanded to document that early-life Pb exposure can contribute to neurodegeneration and neurofibrillary
tangle formation in the aged brain. Pb exposure can affect NMDA receptors, which can contribute to
mood disorders. Synapse formation, adhesion molecules, and nitrosative stress continue to be areas in
which research is being conducted related to Pb-associated neurotoxicity. Finally, the new area of
epigenetics shows that Pb exposure affects methylation patterns in rodent brains. These toxicological data
provide coherence with epidemiologic observations, in particular, associations of Pb exposure with
cognitive deficits, Parkinson's disease, and mood disorders.
In summary, recent evidence substantiates and expands upon the established epidemiologic and
toxicological literature demonstrating the neurological effects of Pb exposure. Both the consistency of
evidence across toxicological and epidemiologic studies and the coherence of findings across the full
spectrum of neurological endpoints, from mechanistic changes to impairments in cognitive function and
behavior and to poorer academic achievement and delinquency, are illustrated in Figure 2-1. In
epidemiologic studies of children, consistently positive associations of blood Pb levels with deficits in
neurocognitive function, attention, and sensory acuities support observed associations with academic
underachievement, which in turn, may explain associations with delinquent and criminal behavior. In
particular, observations of cognitive and behavioral deficits in association with blood Pb levels in the
range of 1-2 (ig/dL indicate that a threshold may not exist for the neurological effects of Pb in children.
Epidemiologic findings are strengthened by their biological plausibility in light of toxicological study
findings and their coherence with toxicological findings for similar or parallel endpoints and for the
mechanisms underlying the neurological effects. The collective body of evidence integrated across
epidemiologic and toxicological studies and across the spectrum of neurological endpoints is sufficient to
conclude that there is a causal relationship between Pb exposures and neurological effects.
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Rodents: Decrements in Seaming
(blood Ph 11.8 pg/dL and higher)
Rodents; Decrements in memory
(blood Ph 12.6 pg/dl and higher)
> Rodents: Depressed
females, emotional males (residual
blood Pb 5-7 pg/dL and higher)
1 urn i o n noon i d
res
Monkey hi i P H I t J
Mice; b
ii^jiRoclents; Increased apoptosis in
brain (0.78 pg/g Pb) i
Rodents: Retinal ERG
supernormality (males only.
blood Pb 12 pg/dL)
Neurotransmitter
aberrations in auditory
area of brainstem
(blood Pb 8.0 pg/dL)
n creased i
neurogi
(bloocf! )
s^^Rodents: Decreased
neurite outgrowth m
(biood Pb
^^S>Rodents; Changes
In neurotransmitter levels,
potentiated with stress
(biood Pb 11 Mg/dl)
Learning
Memory
Behavior
Mood Disorders
Motor Function
Neurodegeneration
Alzheimer's-like
Pathology
Visual Acuity
Auditory Acuity
Neurogenesis
Long Term Potentiation
Synaptic Changes
Blood Brain Barrier
Brain Architecture
Brain Activity
Affected Cell Types:
Neurons
Dendritic Cells
Astrocytes
Glial Cells
Neurotransmitter and
Receptor Changes
Children; Lower academic achievement
(blood Pb quantiies 2 |jg/dl_ and higher);
Lower IQ (mean blood Pb 5-10 jjg/dL):
decrements in higher-order processes
(mean blood Pb 2 pg/ciL and higher)
. Adults: Decrements in cognitive function
(mean blood Pb 3 pcj'dL and higher:
mean tibia Pb 11 pg/g and higher)
Children: Inattention imeari blood Pb 3.7
pg/dL); ADHD (blood Pb quantiies 0.8
pg/dL and higher):
Adults: Criminal arrests (mean blood
Pb = 8.3 pg/dL)
Adults: Mood and psychiatric effects
(blood Pb quantiies 3 pg/dL and
higher, mean tibia Pb 22 pg/g, mean
patella Pb 31-32 pg/g)
Adults: Essential tremor (mean blood
• Pb 2.7 pg/dL); Parkinson's Disease
itiDia pd quantile 19 pg/g and higher)
Increased hearing thresholds
Children: mean blood Pb 7 pg/dL
and higher
Adults: mean patella Pb 32 pg/g
Adults: Less gray matter
volume (mean blood Pb 13.3
pg/dL); White matter diffusion
abnormalities
(mean blood Pb 12.9 pg/dL)
Boys: Greater neurocognitive
deficits with lower affinity
dopamine receptor
(mean blood Pb 6.1 pg/dL)
Rodents and Monkeys:
20- 200 pM Pb
Physiological Cations
Displacement
Oxidative Stress
Figure 2-1. Snapshot of evidence for the spectrum of effects to the
nervous system associated with Pb exposure. Green=animal
toxicological studies (left side); purple=epidemiological
studies (right side).
2.5.2. Cardiovascular Effects
The 2006 Pb AQCD concluded that there was a relationship between increased Pb exposure and
increased adverse cardiovascular outcomes, including increased blood pressure (BP) and increased
incidence of hypertension (U.S. EPA. 2006). Meta-analysis of these studies found that each doubling of
blood Pb level (between 1 and >40 (ig/dL) was associated with a 1 mmHg increase in systolic BP and a
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0.6 mmHg increase in diastolic BP. In addition, most of the reviewed studies using cumulative Pb
exposure measured by adult bone Pb levels showed increased BP. Toxicological studies provided
evidence for exposure to low levels of Pb (e.g., 2 (ig/dL) resulting in increased BP in experimental
animals that persists long after the cessation of Pb exposure and also provided mechanistic evidence to
support the biological plausibility of Pb-induced hypertension, including oxidative stress, altered
sympathetic activity, and vasomodulator imbalance. Finally, limited evidence suggested a connection
between Pb exposure and the development of IHD, cerebrovascular disease, peripheral vascular disease
(PVD) and mortality.
Building on the strong body of evidence presented in the 2006 Pb AQCD, recent studies continue
to support associations between Pb exposure and exposure biomarkers and cardiovascular effects with
recent epidemiologic studies informing past uncertainties (e.g., confounding, low Pb exposures). A recent
study suggested that Pb has an acute effect on BP as a function of recent dose measured by blood Pb and a
chronic effect on hypertension risk as a function of cumulative exposure measured by tibia Pb (Martin et
al.. 2006). This study also verified the magnitude of change in BP observed in the past meta-analysis.
Additionally, recent epidemiologic studies provided evidence for associations between blood Pb and
hypertension in adults with relatively low blood Pb levels; a positive relationship was found in the
NHANES (1999-2002) data set at a geometric mean blood Pb level of 1.64 (ig/dL (Muntner et al.. 2005).
Animal toxicological studies also provide support for effects of low blood Pb level on increased BP with
statistically significant increases shown in animals with blood Pb levels as low as 2 (ig/dL. New studies
also demonstrate a partial reversibility of Pb-induced increased BP following Pb exposure cessation or
chelation.
Epidemiologic studies continue to investigate the relationship between bone Pb and increased BP.
Recent epidemiologic studies also emphasize the interaction between cumulative Pb exposure and factors
that moderate or modify the Pb effect on BP and hypertension (e.g., chronic stress and metabolic
syndrome). Further, recent epidemiologic studies found that the effects of Pb on cardiovascular endpoints
(including BP, pulse pressure [PP], and QT interval) were modified by genotypes (including ALAD and
genes involved in hemochromatosis or Fe metabolism). Epidemiologic and toxicological studies also
provided evidence for Pb exposure to contribute to increased development of atherosclerosis, thrombosis,
ischemic heart disease, peripheral artery disease, arrhythmia, and cardiac contractility. Animal
toxicological evidence continues to build on the evidence supporting the biological plausibility leading to
these cardiovascular alterations. New evidence extends the potential continuum of Pb-related
cardiovascular effects in adults by demonstrating associations of Pb concentrations in blood and bone
with both cardiovascular and all-cause mortality.
In summary, new studies evaluated in the current review support or expand upon the strong body of
evidence presented in the 2006 Pb AQCD that Pb exposure is causally associated with cardiovascular
health effects. Both epidemiologic and toxicological studies continue to demonstrate a consistently
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positive relationship between Pb exposure and increased BP or hypertension development in adults and
this relationship is observed in more recent studies of adults with blood Pb levels (mean: 2 (ig/dL) lower
than those reported in the 2006 Pb AQCD. While some studies evaluate concentration-response
relationships of blood Pb with BP or mortality, the information is inconclusive (Section 2.8.2). Recent
studies investigating measures of cumulative Pb exposure measures and suggest that bone Pb related
strongly to hypertension risk. Evidence of Pb increasing the risk of development of other cardiovascular
diseases also is shown. 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 mode of action of Pb and provided biological
plausibility for the consistently positive associations observed in epidemiologic studies between blood
and bone Pb and cardiovascular effects. These observed associations between Pb exposure and
cardiovascular morbidity are supported by recent reports of increased cardiovascular mortality.
Collectively, the evidence integrated across epidemiologic and toxicological studies as well as across the
spectrum of cardiovascular health endpoints is sufficient to conclude that there is a causal relationship
between Pb exposures and cardiovascular health effects.
2.5.3. Renal Effects
The 2006 Pb AQCD stated 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 (U.S. EPA. 2006). These findings were substantiated by the
coherence of effects observed across epidemiologic and toxicological studies. Toxicological studies
provided mechanistic evidence to support the biological plausibility of Pb-induced renal effects, including
oxidative stress leading to NO inactivation. Uncertainty remained on the implications of effects in
children, confounding, hyperfiltration, and reverse causality.
Recent epidemiologic studies in general and patient populations of adults have, with few
exceptions, been consistent in observing associations between bone or blood Pb levels and worse kidney
function; and provide important evidence that nephrotoxicity occurs at current population Pb biomarker
levels. Further, current evidence does not allow for the identification of a threshold for Pb-related
nephrotoxicity. The odds of reduced eGFR increased by 36% (95% CI: 0.99, 1.85) at blood Pb levels as
low as 1.6-2.4 (ig/dL and by 56% (95% CI: 1.17, 2.08) at blood Pb >2.4 (ig/dL. These studies benefit from
a number of strengths that vary by study but include: comprehensive assessment of Pb dose (using bone
Pb [as a measure of cumulative body burden], and chelatable Pb [as a measure of bioavailable Pb]);
prospective study design; and statistical approaches that utilize a range of exposure and outcome
measures, while adjusting for numerous kidney and Pb risk factors. General population studies also
benefit from large populations in both Europe and the U.S. At blood Pb levels that are common in the
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general U.S. population, Pb increases the risk for clinically relevant effects particularly in susceptible
populations such as those with underlying chronic medical diseases that increase chronic kidney disease
(CKD) risk such as diabetes mellitus and hypertension and co-exposure to other environmental
nephrotoxicants. The uncertainty around the role of reverse causality, which attributes increases in blood
Pb levels to compromised kidney excretion rather than as a causative factor for CKD, was reduced by
evidence that the association between blood Pb and serum creatinine occurred over the entire serum
creatinine range, including the normal range where reverse causality would not be expected. Further,
recent studies have extended the limited body of evidence for effects of Pb on the kidney in children.
Toxicological studies contribute support to effects of Pb in the early life window of exposure adding to
the strength of the association between Pb and altered renal function in children.
CKD results in substantial morbidity and mortality and 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 the renal and cardiovascular systems may result in a cycle
of increased severity of disease. Pb exposure has been causally linked to both increased BP and other
cardiovascular effects (Section 5.4) and renal dysfunction and, 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
mechanisms leading to these renal alterations including oxidative stress, which is also related to CVD,
infiltration of lymphocytes and macrophages associated with increased expression of NF-kB in proximal
tubules and infiltrating cells, mitochondrial dysfunction, renal cell apoptosis, and glomerular hypertrophy.
In summary, new studies evaluated in the current review support or expand upon the strong body of
evidence presented in the 2006 Pb AQCD that Pb exposure is associated with renal health effects.
Epidemiologic studies continue to demonstrate a consistently positive relationship between blood Pb level
and kidney dysfunction at blood Pb levels comparable to those occurring in the current U.S. population
with no evidence for a threshold across the range of levels studied. Uncertainty regarding effects in
children, confounding, hyperfiltration, and reverse causality have been reduced through consideration of
the recent evidence. By demonstrating Pb-induced oxidative stress and describing mechanisms of acute
changes following Pb exposure, toxicological studies provide biological plausibility for the associations
observed in epidemiologic studies between Pb and kidney dysfunction. Collectively, the evidence
integrated across epidemiologic and toxicological studies as well as across the spectrum of kidney health
endpoints is sufficient to conclude that there is a causal relationship between Pb exposures and renal
health effects.
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2.5.4. Immune System Effects
The collective body of evidence integrated across epidemiologic and toxicological study findings
consistently demonstrates that Pb exposure is associated with changes in a spectrum of immune mediators
and functions. The majority of results from animal studies indicates that immune changes are observable
at blood Pb levels in the range of 2 to 8 (ig/dL. Likewise, in the newly expanded body of epidemiologic
studies in environmentally-exposed children and adults, changes in immune function are demonstrated in
association with mean blood Pb levels in the range of 1 to 10 (ig/dL.
The strength of evidence for Pb-associated immune effects is derived not only from the consistency
of associations but also from the coherence of findings between toxicological and epidemiologic studies
and coherence of findings across the spectrum of related immune changes. Toxicological and
epidemiologic evidence links higher Pb exposures with decreases in various T cell subtypes. These
changes can affect cell-to-cell interactions that mediate acquired immunity required in subsequent
memory responses to antigen exposures; however, it is unclear what effect the observed magnitudes of
changes may have in attenuating acquired immunity.
The key immunomodulatory effect of Pb exposure, in terms of coherence across immune endpoints
and implications for developing immune-based diseases, is the skewing of immune function away from a
Thl phenotype towards aTh2 phenotype. In toxicological studies and epidemiologic studies, this shift is
well demonstrated by suppressed production of Thl cytokines (e.g., IFN-y) and increased production of
Th2 cytokines (e.g., IL-4). A recent in vitro study indicates that Pb may promote Th2 responses by acting
directly on dendritic cells, the major effector in antigen response. An increase in IL-4 from activated Th2
cells induces differentiation of B cells into Ab producing cells, thereby promoting the secretion of IgE,
IgA, and IgG. In support of this well-established mechanism, toxicological studies describe Pb-induced
changes in IgA, IgG, and IgM. Additionally, epidemiologic studies in children consistently link higher Pb
exposures with increases in B cell abundance and increases in IgE. Observations of Pb-associated
increases in Th2 responses and circulating IgE levels provide biological plausibility for epidemiologic
observations in children of associations of blood Pb with asthma and allergic conditions. Such
epidemiologic data are sparse, and additional studies with more rigorous methodology (e.g., longitudinal
design and adjustment for potential confounders such as smoking, SES, and exposures to other metals)
are needed to substantiate the findings.
Further evidence of Pb-associated suppressed Thl activity is provided by toxicological and
epidemiologic observations that Pb exposure is associated with impaired killing capacity of macrophages
and neutrophils. There is toxicological and epidemiologic evidence of suppressed Thl activity and effects
on macrophage and neutrophil functional activities. This evidence provides biological plausibility for
observations in animals of the Pb-induced suppression of the DTH response and the observations in both
animals and humans that Pb exposure and increases the risk of infection.
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Toxicological studies and a limited set of epidemiologic studies demonstrate that Pb induces
macrophages into a hyperinflammatory state as characterized by suppressed production of NO and
enhanced production of ROS, TNF-a, and the immunosuppressive PGE2. Specialized macrophages
residing in airways, reproductive organs, and in the nervous system indicate that immunomodulation may
underlie the documented associations of Pb exposure with effects in these organ systems. Although
limited mostly to toxicological studies, Pb has been shown to induce the generation of auto-antibodies,
suggesting that Pb exposure may increase the risk of autoimmune conditions.
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 to promote a Th2 phenotype and inflammation. The consistency and
coherence of findings among these related immune effects, in turn, establish the biological plausibility for
Pb exposure being associated with increased susceptibility to infection, autoimmunity, allergy, and effects
in other organ systems. Animal studies and to a limited extent, epidemiologic studies, demonstrate
increased susceptibility of prenatal exposures and enhanced responses with co-exposures to other metals.
The consistency of findings and the coherence between toxicological and epidemiologic findings across
the continuum of related immune responses are sufficient to conclude that there is a causal relationship
between Pb exposures and immune effects.
2.5.5. Heme Synthesis and RBC Function
Consistent with conclusions of the 2006 Pb AQCD as well as previous assessments, recent
evidence in the toxicological and epidemiologic literature supports the longstanding relationship between
Pb exposure and effects on hematological endpoints, including altered heme synthesis, decreased RBC
survival and function, and increased RBC oxidative stress.
Multiple occupational epidemiologic studies have shown that Pb affects several hematological
parameters such as Hb, PCV, MCV, MCH, and MCHC. Although the majority of occupationally- exposed
adults had blood Pb levels in excess of 20 (ig/dL, decreases in Hb and PCV were also observed in an
occupational cohort with a mean blood Pb level of 7 (ig/dL. In addition, Pb exposure was shown to reduce
Ca-ATPase and Ca-Mg-ATPase activity in RBC membranes at cord blood Pb levels of 3.54 (ig/dL.
Decreases in Ca-ATPase and Ca-Mg-ATPase activity leads to an increase in RBC [Ca2+]/, increased
membrane fragility, and abnormal morphological changes. Studies in children are less consistent than
those investigating occupationally-exposed adults; this may due to the comparatively shorter duration of
and magnitude of exposure experienced by children. Toxicological studies have also observed decreases
in hematocrit and hemoglobin and increases in hemolysis and reticulocyte density in rats and mice with
blood Pb levels as low as 6.6-7.1 (ig/dL. Pb exposure has also been observed to increase PS expression on
RBC membranes, leading to cell shrinkage, eryptosis, and destruction of the RBCs by macrophages.
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Suggestive evidence of disrupted hematopoiesis evidenced by decreased serum erythropoietin was
observed in occupationally exposed adults with blood Pb levels of 6.4 (ig/dL; toxicological studies in rats
also indicate that Pb is cytotoxic to RBC-progenitor cells after chronic exposure. Taken together, these
studies provide consistent evidence that exposure to Pb adversely effects 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.
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 6th step in heme biosynthesis converting coporphyrinogen III into
protoporphyrinogen IX), and ferrochelatase (catalyses the terminal step in heme synthesis converting
protoporphyrin IX into heme). Recently, numerous epidemiologic studies have confirmed that decreases
in RBC ALAD levels and activity are strongly associated with blood Pb levels in as low as 7.1 (ig/dL in
children and blood Pb levels as low as 6.4 (ig/dL in adults. Decreases in blood ALAD activity were also
seen in rats with blood Pb levels of 6.5 (ig/dL. There is also a considerable body of evidence for a
negative correlation between ALAD activity and Pb concentration in various invertebrate and vertebrate
taxa (Section 2.6.7). In addition to ALAD, recent studies have shown that Pb exposure inhibits the
activity of ferrochelatase, leading to increased RBC ZPP in humans and animals. Pb has also been shown
to inhibit the activities of other enzymes in RBCs, including those involved in nucleotide scavenging,
energy metabolism, and acid-base homeostasis.
Lastly, Pb exposure induces lipid peroxidation and oxidative stress in RBCs. Epidemiologic studies
have observed increases in MDA in occupationally-exposed adults with blood Pb levels as low as 7.9
(ig/dL. Other measures of oxidative stress observed included lowered activities of SOD, GR, and CAT,
and increased CRP. Indices of RBC oxidative stress were also seen in adolescents and children exposed to
Pb. In vitro and in vivo studies have also demonstrated that prior, con-current, or subsequent treatment
with various antioxidants has been shown to at least partially ameliorate Pb-induced oxidative stress in
RBCs.
In conclusion, the recent epidemiologic and toxicological literature provides strong evidence that
exposure to Pb is associated with numerous deleterious effects on the hematological system, including
altered heme synthesis mediated through decreased ALAD and ferrochelatase activities, decreased RBC
survival and function, decreased hematopoiesis, and increased oxidative stress and lipid peroxidation. The
consistency of findings in the epidemiologic and toxicological literature and coherence across the
disciplines is sufficient to conclude that there is a causal relationship between Pb exposure and heme
synthesis and RBC function.
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2.5.6. Reproductive Effects and Birth Outcomes
Epidemiologic and toxicological studies of the effects of Pb on reproductive outcomes have
covered outcomes such as female and male reproductive function, birth defects, spontaneous abortions,
infant mortality, preterm birth, low birth weight, and developmental effects.
Many of the Pb-induced effects in toxicological studies have been observed at maternal blood Pb
levels that do not result in overt clinical toxicity in the dams. Recent toxicological studies have shown the
effects of Pb exposure during early development to include disruption of endocrine function; delay in the
onset of puberty and alteration in reproductive function later in life; and changes in morphology or
histology in sex organs and placenta. Additionally, epidemiologic studies of reproductive factors among
males and females investigated whether Pb 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 adult males exposed to Pb in occupational settings with
blood Pb as low as 20-45 (ig/dL and sperm count and quality. Decreased sperm viability and altered
morphology in sperm is also observed in invertebrate species (Sections 7.2.4 and 7.3.4) Multiple studies
of Pb and puberty have shown inverse associations between blood Pb levels and delayed pubertal
development for girls and boys. These associations are consistently observed in multiple epidemiologic
studies and demonstrate effects on pubertal development at blood Pb levels < 1O^ig/dL.
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 suggested that oxidative stress is a major contributor to the toxic effects
of Pb on male and female reproductive systems. 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 demonstrated increased oxidative
stress also reported 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 evidence was available to
study the associations with spontaneous abortions. The previous Pb AQCD included a few studies that
reported possible associations between Pb and neural tube defects, but the recent epidemiologic studies
found no association. Possible 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
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consistent when using maternal blood Pb or umbilical cord and placenta Pb. Effects of Pb exposure
during early development in toxicological studies included reduction in litter size, implantation, birth
weight and postnatal growth.
Additional evidence for Pb exposure negatively affecting development is provided by toxicological
studies demonstrating developmental Pb exposures leading to impaired development of the retina, skin,
teeth and altered development of the hematopoietic and hepatic systems. In summary, the recent
toxicological and epidemiologic literature provides strong evidence that Pb exposure is related to delayed
onset of puberty in both males and females. Additionally, Pb exposure has been shown to have
detrimental effects on sperm (at higher blood Pb levels in epidemiologic studies and lower doses in the
toxicological literature). Furthermore, evidence from invertebrate and vertebrate taxa in both terrestrial
and aquatic ecosystems provide additional support for reproductive and developmental effects associated
with Pb exposure (Sections 2.6.8, 7.2.4, and 7.3.4). 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. Despite some inconsistencies for particular endpoints, the
evidence for Pb-related reproductive effects is strengthened by the coherence with similar findings in
invertebrate species. The collective body of evidence integrated across epidemiologic and toxicological
studies, with a focus on the strong relationship observed with negative effects on sperm and delayed
pubertal onset, is sufficient to conclude that there is a causal relationship between Pb exposures and
reproductive effects and birth outcomes.
2.5.7. Effects on Other Organ Systems
In the 2006 Pb AQCD, exposure to Pb was shown to exert effects in organ systems not yet
explicitly covered in the preceding sections of this document. These organ systems included the liver,
gastrointestinal tract, endocrine system, bone and teeth. In the current document, effects on these organ
systems, as well as effects on the respiratory system, have been organized in one section because the
amount of new evidence appearing since the 2006 Pb AQCD is limited. The few recent studies, however,
find that Pb may negatively affect the function of these systems at lower blood Pb levels than previously
described.
There is evidence from recent epidemiologic and toxicological studies that exposure to Pb results
in altered liver function and hepatic toxicity, including the observation of altered serum protein levels,
increased serum enzyme activities, and altered hepatic lipid metabolism. 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.
Effects observed in occupational cohorts of painters, battery- and jewelry-workers, as well as animal
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toxicological studies (applying a wide range of exposure regimens), occurred at blood Pb levels >20
(.ig/dL.
Relatively few human studies have been conducted on gastrointestinal toxicity of Pb since the
completion of the 2006 Pb AQCD. GI symptoms were observed in battery workers and painters exposed
to Pb in India (mean blood Pb level = 42.40 ± 25.53 and 8.04 ± 5.04 (ig/dL, respectively). Toxicological
evidence for Pb-induced GI health effects in rats includes altered muscle relaxations and markers of
oxidative stress in the gastric fundus and mucosa. The observation of oxidative stress was accompanied
gastric mucosal damage following short-term, sub-chronic and chronic exposures. The anterior intestine
of fish has also been identified as a target of Pb (Section 2.6.2).
The endocrine processes most impacted by exposure to Pb include changes in thyroid function, as
well as alteration in sex and stress hormone profiles. TSH was negatively correlated with blood Pb in
women that ate fish contaminated with Pb as well as other chemicals (median blood Pb level =1.7 (ig/dL,
less than the detection limit for the study), and FT4, but not FT3, was decreased in adolescent male auto
repair workers (blood Pb level = 7.3 ± 2.92 j^ig/dL). Significant differences in the levels of sex hormones,
including total and free testosterone, estradiol, aromatase, and luteinizing hormone, were observed in
Belgian adolescents residing in areas with different levels of industrial pollution including Pb (mean
blood Pb levels of 2.2 |_ig/dL.) Toxicological evidence for similar effects was observed in adults cow
reared in an environment containing Pb and other contaminants: positive correlations were reported
between blood Pb and plasma T3, T4, and estradiol levels. In study of children (mean age 9.5 years)
challenged with an acute stressor, increasing blood Pb was associated with significant increases in
salivary Cortisol responses comparing blood Pb levels of 1.1-1.4 (ig/dLto blood Pb levels <1 (ig/dL.
Multiple epidemiologic studies investigated the association between Pb exposure and bone and
tooth health in adults. High blood Pb has been observed to be associated with decreased BMD in non-
Hispanic white males (blood Pb level = 4.9 j^ig/dL). In elderly women, blood Pb levels (> 8 j^ig/dL) were
associated with an increased risk of falls and fractures, including osteoporosis-related falls. Linear
skeletal growth in children (7-17 years of age [mean blood Pb level = 7.7 (ig/dL)]), was negatively
correlated with increasing blood Pb levels. Epidemiologic studies (investigating Pb exposure and tooth
loss) reported 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 with oxidative stress/damage in
individuals exposed in an occupational setting.
New toxicology studies have reported ocular effects (i.e., retinal progenitor cell proliferation) at
blood Pb levels as low as <10 (ig/dL (Section 5.3.4.3), and one human study reported an association
between heavy smoking, increased blood Pb, and cataracts. Investigation of the respiratory effects of Pb
exposure has been limited; however, cross-sectional studies have indicated an association of increasing
blood Pb with increased prevalence of respiratory tract illnesses (Section 5.6.4.1) and asthma in children
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(Section 5.6.4.2). Additionally, Pb-induced production of ROS is implicated in increased BR and
decrements in lung function (Section 5.6.4.3).
2.5.8. Cancer
Toxicological literature on the genotoxic, mutagenic, and carcinogenic potential of Pb includes
strong evidence of effects in laboratory animals. Both the International Agency for Research on Cancer
(IARC) and the National Toxicology Program (NTP) have examined the role of Pb in cancer. IARC
classified inorganic Pb compounds as probable human carcinogens and organic Pb compounds as not
classifiable (IARC. 2006; Rousseau et al. 2005). The NTP reported Pb and Pb compounds to be
'reasonably anticipated to be human carcinogens'(NTP. 2004).
In laboratory studies, high-dose Pb has been demonstrated to be an animal carcinogen. Pb is likely
to be a human carcinogen based on strong evidence from animal toxicology data (IARC. 2006; Waalkes et
al.. 1995)and less definitive epidemiological data. 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. Pb is clastogenic and mutagenic in some but not all models. Clastogenicity and
mutagenicity may be possible mechanisms contributing to cancer, but are not necessarily associated with
the induction of cancer. Due to the disruption of metal cofactors binding to Zn-finger proteins, Pb 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, and with histones. These indirect effects may act at a post-translational level to alter protein
structure and DNA repair. In addition, some evidence of epigenetic changes associated with Pb exposure
is available in the recent literature. Epigenetic changes may further alter DNA repair or change the
expression of a tumor suppressor gene or oncogene. Thus, the animal toxicology literature provides a
strong base for understanding the potential contribution of Pb exposure to cancer in laboratory animals.
Multiple epidemiologic studies have been performed examining the association with cancer
incidence and mortality with Pb exposure assessed using biological measures and exposure databases.
Mixed results have been reported for cancer mortality studies; one strong epidemiologic study of US
adults (Schober et al.. 2006) demonstrated a positive association between blood Pb and cancer mortality,
but the other studies reported null results (Menke et al.. 2006). Although the previous Pb AQCD reported
that some studies were suggestive of an association between Pb exposure and lung cancer, current studies,
which mostly examined occupational exposure 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 genotypes. A limited amount of research on other types of cancer has been
performed. The previous Pb AQCD reported evidence that suggested an association between Pb exposure
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and stomach cancer, but recent studies of this association are lacking, with only one study published since
the last Pb AQCD (U.S. EPA. 2006). which reported mixed results.
Among epidemiologic studies on genotoxicity, positive associations were observed between high
Pb blood levels and sister chromatid exchange (SCE) 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 inducing SCE, MN and DNA strand breaks. Only PbCr04
produces chromosomal aberrations but this effect is likely due to chromate. Pb does not appear to be very
mutagenic unless a cell signaling pathway was disturbed.
Epigenetic effects, particularly with respect to methylation and effects on DNA repair were
observed consistently. In humans, epigenetic studies examining Pb and LINE-1 and Alu consistently
demonstrated an inverse association between patella Pb and global DNA LINE-1 methylation (Pilsner et
al.. 2009; R. O. Wright et al. 2010). Toxicological studies show that Pb can activate or interfere with a
number of signaling and repair pathways, though it is unclear whether these are in response to epigenetics
or genotoxicity. Thus, an underlying mechanism is still unknown, but likely involves either genomic
instability or epigenetic modifications or both.
Overall, there is some epidemiologic evidence supporting associations between Pb and cancer.
Strong evidence from toxicological studies demonstrates an association between Pb and cancer,
genotoxicity/clastogenicity or epigenetic modification. The collective body of evidence integrated across
epidemiologic and toxicological studies is sufficient to conclude that there is a likely causal relationship
between Pb exposures and cancer.
2.5.9. Human Health Effects and Corresponding Blood Pb
Levels
Tables 2-2 and 2-3 summarize the health effects in children (and adults and the lowest blood Pb
level at which the weight of the evidence substantiates a causal relationship. The 2006 Pb AQCD did not
identify a safe level of exposure for Pb and concluded that any threshold for Pb neurotoxicity would have
to exist at levels distinctly lower than the lowest exposures examined in the epidemiologic studies
included in the assessment. Recent studies continue to find associations between a wide range of health
endpoints and increasingly lower levels of blood Pb. The lack of a reference population with blood Pb
levels reflecting pre-industrial Pb exposures continues to limit the ability to identify a threshold.
Estimates of "background" blood Pb levels have been measured in ancient bones from pre-industrialized
societies. These studies suggest that the level of lead in blood in pre-industrial humans was approximately
0.016 (ig/dL (Flegal & Smith. 1992). approximately 65-fold lower than that currently measured in U.S.
populations. In this context, a blood Pb level of 1 (ig/dL is not relatively low. Further, if a threshold did
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1 exist, in order to demonstrate it, the scale at which blood Pb level is measured will likely have to be
2 adjusted to parts per million ((.ig/L) instead of parts per hundred thousand ((.ig/dL).
Table 2-2. Summary of Pb-induced health effects in children and the lowest mean blood
Pb level in the population(s) studied
Blood
Pb
Level
Neurological Effects Renal Effects
Immune Effects
Effects on Heme
Synthesis and RBC
Function
20 |jg/dL
• Macrophage hyper-
inflammation1
15 |jg/dL
• Lymphocyte activation'
• Increased Zn protoporphyrinp
• Lipid peroxidation9
10 |jg/dL
• Delinquent behavior8
• Increased hearing threshold15
• Altered antioxidant enzyme
activities'
• Decreased ALAD activities5
• Altered hematological
parameters (i.e., decreased
hemoglobin)
5 |jg/dL
• Inattention0
• Decrements in full scale IQd
• Decrements in specific
neurocognitive domains6 Decreased eGFRh
• Poorer school performance Decreased ebi-K
• ADHD9
• Increased B cell abundance*
• Increased IgE'
• Increased risk of infection"1
• Decreased T cell abundance"
• Allergic sensitization"
• Decreased Ca-Mg ATPase
activity"
1 |jg/dL
Note: Endpoints where the weight of the evidence, overall, substantiates the causal association with blood Pb levels in the range noted on the
figure. Since no evident threshold has yet been clearly established for most effects, the existence of such effects at still lower blood Pb levels
cannot be ruled out based on available information.
Supporting references: Wright et al. (2008)a; Hwang et al. (2009) and Schwartz and Otto (1991)b; Nicolescu et al. (2010)°: Kim et al. (2009)d; Krieg
et al. (2010)e; Miranda et al. (2009)'; Braun et al. (2006) and Braun et al. (2008)9; Fadrowski et al. (2010)h: Pineda-Zavaleta et al. (2004)': Lutz et al.
(1999V: Sarasua et al. (2000); Karmaus et al. (2005); Karmaus et al. (2005) ; Karmaus et al. (2005) ; Jedrychowski et al. (2011)°: Wang et al.
(2010)": Ahamed et al. (2006)q: Ahamed et al. (2005)'; Wang et al.(2010)TRiddell et al. (2007)': Huel et al. (2008)"
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Table 2-3. Summary of Pb-induced health effects in adults and the lowest mean blood
Pb level in the population(s) studied
Blood
Pb Level
Neurological
Effects
Cardiovascular
Effects
Renal
Effects
Immune
Effects
Reproductive
Effects and
Birth
Outcomes
Effects on Heme
Synthesis and
RBC Function
30 |jg/dL
• Arrhythmia9
• Decreased
neutrophil functionp
• Sperm
abnormalities"
• Decreased
hematocrit™
• Decreased Ca-Mg
ATPase activity"
20 |jg/dL
• Impaired renal
tubular function"
• Increased auto-
antibodies9
15 |jg/dL
• Brain MRI
changes8
• Increased hearing
threshold15
10 |jg/dL
• Criminal arrest0
• Lipid peroxidation"
• Decreased
antioxidant enzyme
activities"
• Altered
hematological
parameters
(e.g., decreased
hemoglobin, packed
cell volume)2
• Decreased serum
erythropoietin"
• Increased Zn
protoporphyrin1*
• Altered ALAD
activitybb
5 |jg/dL
1 |jg/dL
• Decrements in
cognitive function11
• Essential tremor0
• Mood disorders'
• Mortality"
• Decreased HRV'
• Ischemic heart
disease1
• Peripheral Artery
Diseasek
• Hypertension1
• Increased BPm
• Elevated serum
creatinine"
(| creatinine
clearance, GFR)
• Shift to Th2
cytokines'
• Increased
bronchial
reactivity5
• Increased
inflammation'
• Delayed puberty"
Note: Endpoints where the weight of the evidence, overall, substantiates the causal association with blood Pb levels in the range noted on the figure.
Since no evident threshold has yet been clearly established for most effects, the existence of such effects at still lower blood Pb levels cannot be
ruled out based on available information.
Supporting references: Brubaker et al. (2010: 2009)': Hwang et al. (2009) and Chuang et al. (2007)b; Wright et al. (20081°: Krieg et al. (2009)d; Dogu
et al. (2007)e; Bouchard et al. (2009)'; Reza et al. (2008)9; Menke et al. (2006)h; Park et al. (20097: Jain et al. (20070/iuntner et al. (2005T
Scinicariello et al. (2010) and Park et al. (2009)': Scinicariello et al. (2010) and Martin et al. (2006)m: Sun et al. (2008). Lin and Tai-vi (2007) and
Wang et al. (2010)yislih et al. (2004). Akesson et al. (2005) and Yu et al. (2004)°: Valentino et al. (1991)": El-Fawal et al. (1999)": Kim et al. (2007)r:
Min et al. (2008)5: Sonadei et al. (2010)': Telisman et al. (2007) and Hsu et al. (2009)": Hauser et al. (2008). Williams et al. (2010). Denham et al.
(2005). Selevan et al. (2003) and Wu et al. (2003)": Karita et al. (2005)": Abam et al. (2008)": Ergurhan-llhan et al. (2008)y: Ukaejiofo et al. (2009)z;
Sakata et al. (2007)aa: Wang (2010)""
2.6. Ecological Effects
1 This section evaluates the evidence from studies of ecological effects associated with exposure to
2 Pb. The results from the studies evaluated in combination with the evidence from other disciplines (e.g.,
3 fate and transport) contribute to the causal determinations for the ecological outcomes discussed in this
4 assessment. In the following sub-sections, a discussion of the causal determinations is presented for the
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ecological effects. Effects determined to be causal at the species level contribute to the body of evidence
for causal effects at the community and ecosystem scale. Where the causal determination varies
substantially between types of organisms (typically between plants and other organisms), the divergence
is noted.
The evidence used to formulate each causal determination is summarized here, and the
corresponding detailed discussion can be found in Chapter 7 of this document. In Chapter 7, the effects on
terrestrial (Section 7.2) and aquatic (Section 7.3) ecosystems are presented separately, and each of the
sections first discusses effects at the species level, followed by community and ecosystem levels. In each
of the two main sections, biogeochemistry and chemical effects of Pb that influence bioavailability are
considered first, as Pb must first move from the environmental media (soil, water, sediment etc.) into
biota. Next, uptake of Pb from soil and water are discussed, then new information on biological effects of
Pb on plants, invertebrates and vertebrates, followed by data on exposure-response relationships.
Ecosystem-scale responses to Pb exposure are considered along with critical loads, characterization of
sensitivity and vulnerability, and the effect of Pb on ecosystem services. Finally, in Section 7.4, a
synthesis of the effects of Pb observed across terrestrial and aquatic habitats is presented along with
causal determinations for those effects, which are also summarized in Table 2-4 below. In this chapter, Pb
effects on terrestrial and aquatic systems from Chapter 7 are summarized (Sections 2.6.1 and 2.6.2);
followed by a summary of the evidence for the causal determinations (Sections 2.6.3 to 2.6.10) and
consideration of atmospheric deposition of Pb as related to ecological effects (Section 2.6.11).
Table 2-4. Summary of causal determinations for Pb in plants, vertebrates and
invertebrates
Effect
Causality Determination
Bioaccumulation - All Organisms
Causal Relationship
Mortality - Plants
Inadequate to Infer Causal Relationship
Mortality - Vertebrates and Invertebrates
Causal Relationship
Growth - Plants
Causal Relationship
Growth - Invertebrates
Causal Relationship
Growth - Vertebrates
Suggestive of a Causal Relationship
Physiological Stress-All Organisms
Causal Relationship
Hematological Effects - Invertebrates and Vertebrates
Causal Relationship
Development and Reproduction- Invertebrates and Vertebrates
Causal Relationship
Development and Reproduction-Plants
Inadequate to Infer Causal Relationship
Neurobehavior - Invertebrates and Vertebrates
Causal Relationship
Community and Ecosystem Level Effects
Causal Relationship
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2.6.1. Summary of Terrestrial Ecosystem Effects
Section 7.2 focuses on the effects of Pb in terrestrial systems. 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) CEC; and (6) aging. Since
2006, further details have been contributed to the understanding of the role of pH, cation exchange
capacity (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 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,
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 bioaccumulates in plants, invertebrates and vertebrates in terrestrial
systems. 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. Multiple new studies showed that in trees,
the latter 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. Studies with trees found that soil Pb is generally translocated from the roots.
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 arthropods collected at contaminated sites, recent
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studies found gradients in accumulated Pb that corresponded to gradients in soil with increasing distance
from point sources.
There were a few new studies of Pb bioavailability and uptake in birds since the 2006 Pb AQCD.
Several found tissue levels in birds that indicated exposure to Pb, but none of the locations for these
studies was in proximity to 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 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, reduced growth was observed in some experiments, as well as genotoxicity, decreased
germination, and pollen sterility.
In terrestrial invertebrates, evidence for Pb effects have 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 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.
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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.
Evidence reviewed in Sections 7.2.3 and 7.2.4 demonstrates clearly that increased 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.
New evidence of effects of Pb at the community and ecosystem scale 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. Most recently published research
on community and ecosystem scale 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 over time. 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.
2.6.2. Summary of Aquatic Ecosystem Effects
Section 7.3 focuses on the effects of Pb in aquatic systems. Once atmospherically-derived Pb enters
surface waters, its fate and bioavailability are influenced by Ca2+ concentration, pH, alkalinity, total
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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.
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 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.
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 and 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 bioaccumulated 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.
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 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 higher trophic-level consumers, indicating that Pb
was bioconcentrated but not biomagnified. Some studies published since the 2006 Pb AQCD support the
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potential for transfer of Pb in aquatic food webs, while other studies indicate that Pb concentration
decreases with increasing trophic level (biodilution).
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 in this review. They include decreased growth,
deformation and disintegration of cells, and blocking of the pathways that lead to pigment synthesis, thus
affecting photosynthesis. Effects on plants supported by additional evidence in this review 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.
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. Growth effects are observed at lower concentrations in
some aquatic invertebrates since in the 2006 Pb AQCD, including juveniles of the freshwater snail
Lymnaea stagnalis where growth is affected at <4 (.ig Pb/L (Grosell et al.. 2006). 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 vertebrates. Additional mechanisms of Pb
toxicity have been elucidated in the gill and the renal system of fish since the 2006 Pb AQCD.
Furthermore, the mitogen-activated protein kinases, ERK1/2 and p3SVIAI>K were identified for the first
time as possible molecular targets for Pb neurotoxicity in ateleost (Leal et al.. 2006). 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.
Concentration-response data from plants, invertebrates and vertebrates is consistent with findings
in previous AQCDs of species differences in sensitivity to Pb in aquatic systems. In this ISA (and
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/L. Several studies in this review have provided evidence of effects at lower
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concentrations. Among the most sensitive species, growth of juvenile freshwater snails (L. stagnalis) was
inhibited at an EC;,, of <4 j.ig Pb/L. (Grosell & Brix. 2009; Grosell et al.. 2006V 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 ever reported for Pb (N. Wang et al.. 2010). 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 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-scale 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.
2.6.3. Bioaccumulation of Lead in Terrestrial and Aquatic
Biota as it Affects Ecosystem Services
Pb deposited on the surface of, or taken up by organisms has the potential to alter the services
provided by terrestrial and aquatic biota to humans. Ecosystem services are the benefits people obtain
from ecosystems. They include supporting, provisioning, regulating and cultural services that are vital for
the functioning of the biosphere and provide the basis for the delivery of tangible benefits to human
society. There is compelling evidence over several decades of research and in recent studies reviewed in
this ISA (Sections 7.2.3 and 7.3.3.) that Pb bioaccumulates in plants, invertebrates and vertebrates in
terrestrial and aquatic systems. Generally, there are considerable differences between species in the
amount of Pb taken up from the environment, and in the amounts of Pb retained in the organism. In order
for Pb to reach a biological receptor, the metal must first cross the membranes of organisms to reach the
target organ or site of storage. This process varies between plants, invertebrates and vertebrates, and
furthermore, uptake and sequestration are at times similar in unrelated species, while substantially
different between related ones. Uptake of Pb from environmental media is dependent upon the
bioaccessibility of Pb (reviewed in Chapter 3) which is influenced by many factors including, but not
limited to, temperature, pH, presence of humic acid and dissolved organic matter, presence of other
metals, and speciation of Pb.
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Pb is bioaccumulated in plants, invertebrates and vertebrates inhabiting terrestrial and aquatic
systems that receive Pb from atmospheric deposition. This represents a potential route for Pb mobilization
into the food web or into food products. For example, Pb bioaccumulation in leaves and roots of an edible
plant may represent an adverse impact to the provisioning of food, an essential ecosystem service. 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).
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. Overall, based on the
consistency of findings across taxa, the evidence is sufficient to conclude that there is a causal
relationship between Pb exposures and bioaccumulation of Pb that affects ecosystem services
associated with terrestrial and aquatic biota.
2.6.4. Mortality
The relationship between Pb exposure and mortality has been well demonstrated in terrestrial and
aquatic species as presented in Sections 7.2.5 and 7.3.5 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. However, the LC50 is usually much higher than current environmental levels
of Pb, even though physiological dysfunction that adversely impacts the fitness of an organism often
occurs well below concentrations that result in mortality.
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. 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
candidate or endangered species. The freshwater mussel Lampsilis rafinesqueana (Neosho mucket), is a
candidate for the endangered species list. The EC50 for foot movement (a measure of viability) for newly
transformed juveniles of this species was 188 (.ig Pb/L. Other invertebrates such as odonates may be
tolerant of Pb concentrations that greatly exceed environmental levels.
Thirty day LC50 values for larval fathead minnows ranged from 39 to 1,903 (ig/L in varying
concentrations of DOC, CaS04 and pH. (Grosell et al.. 2006). 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).
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The evidence is inadequate to conclude that there is a causal relationship between Pb and
mortality in plants.
The evidence is sufficient to conclude that there is a causal relationship between Pb exposures
and mortality in sensitive terrestrial and aquatic animal taxa.
2.6.5. Growth Effects
Evidence for Pb effects on growth is strongest in plants, with limited information in invertebrates
and vertebrates. There is evidence over several decades of research that Pb inhibits photosynthesis and
respiration in plants, both of which reduce growth (U.S. EPA. 1977. 2006). Many laboratory and
greenhouse toxicity studies have reported effects on plants, and there are few field toxicity studies. 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. The lowest 72-hour EC50 for growth inhibition reported for algae was 105 (.ig Pb/L
in Chaetoceros sp. Most primary producers experience EC50 values for growth in the range of 1,000 to
100,000 (.ig Pb/L (U.S. EPA. 2006).
In previous Pb AQCDs, growth effects of Pb have been reported in fish (growth inhibition), birds
(changes in juvenile weight gain), and frogs (delayed metamorphosis, smaller larvae). Growth effects
observed in invertebrates and vertebrates underscore the importance of lifestage to overall Pb
susceptibility. In general, juvenile organisms are more sensitive than adults. Several studies since the last
review have demonstrated effects of Pb on growth at lower concentrations than in previous literature.
Among the animal taxa tested, aquatic invertebrates were the most sensitive to the effect of Pb, with
adverse effects being reported as low as 4 (.ig Pb/L. Growth of juvenile freshwater snails L. stagnalis was
inhibited at EC2o <4 (.ig Pb/L (Grosell & Brix. 2009; Grosell et al.. 2006). In the freshwater mussel,
fatmucket (L. siliquoidea) juveniles were the most sensitive lifestage (N. Wang et al.. 2010). 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.
Evidence is sufficient to conclude that there is a causal relationship between Pb exposures and
growth effects in plants and invertebrates. Evidence is suggestive of a causal relationship between
Pb exposures and growth effects in vertebrates.
2.6.6. Physiological Stress
In this ISA and previous Pb AQCDs, there is consistent and coherent 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
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in algae, aquatic mosses, floating and rooted aquatic macrophytes, and terrestrial species. There is
considerable evidence for antioxidant activity in invertebrates, including gastropods, mussels, and
crustaceans, in response to Pb exposure. Upregulation of antioxidant enzymes are also observed in fish.
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 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.
Evidence is sufficient to conclude that there is a causal relationship between Pb exposures and
physiological stress in plants, invertebrates and vertebrates.
2.6.7. Hematological Effects
Hematological responses are commonly reported effects of Pb exposure in invertebrates and
vertebrates in both aquatic and terrestrial systems. In environmental assessments of metal-impacted
habitats, ALAD is a recognized biomarker of Pb exposure ("U.S. EPA. 2006). 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. Further evidence
from the 2006 Pb AQCD and this review for Pb effects on ALAD enzymatic activity including effects in
bacteria, amphibians and additional field and laboratory studies on fish suggest this enzyme is an
indicator for Pb exposure across a wide range of taxa. Limited evidence of Pb effects on other blood
parameters including altered serum profiles and changes in white blood cell counts in fish and amphibians
support the finding of the hematological system as a target for Pb in natural systems. 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 (Sections 2.5.5 and 5.7). 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 invertebrates and vertebrates.
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2.6.8. Developmental and Reproductive Effects
Evaluation of the literature on Pb effects in aquatic and terrestrial species indicates that exposure to
Pb is associated with adverse effects on development and reproduction. Evidence in this review and the
previous Pb AQCDs from invertebrate and vertebrate studies indicate that Pb is adversely affecting
reproductive performance in multiple species. In plants, few studies are available that specifically address
reproductive effects of Pb exposure.
Several new studies of snails, clams and rotifers support previous findings of adverse impacts to
embryonic development. In addition to affecting the embryo, Pb can alter developmental timing, 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.
Reproductive effects have also been observed in multi-generational studies. Larval settlement rate
and rate of population increase was adversely impacted in rotifers. Rotifers have decreased fertilization
rate associated with Pb exposure that appeared to be due to decreased viability of sperm and eggs.
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.
In 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 subtle
skeletal malformations. Previous Pb AQCDs have reported developmental effects in fish, specifically
spinal deformities in larvae. 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
reproductive parameters in fish observed to be impacted by Pb 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 observed adverse reproductive effects and Pb exposure. Decreased testis weight was
observed in lizards. In mammals, few studies in the field have addressed Pb specifically, due to most
available data in wild or grazing animals being from near smelters, where animals are co-exposed to other
metals. 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
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finding that is also an area of emerging interest in the 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. The strongest and most consistent evidence, which was coherent across
epidemiologic and toxicological studies, was for effects of Pb on sperm and the onset of puberty in males
and females.
Adverse effects of Pb on reproduction in invertebrates and vertebrates indicate that Pb is likely
affecting fecundity of Pb-exposed 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. In plants, the evidence is inadequate
to conclude a causal relationship between Pb exposures and plant reproductive effects.
2.6.9. Neurobehavioral Effects
Evidence from laboratory studies and limited data from field studies reviewed in Chapter 7 have
shown adverse effects of Pb on neurological endpoints in both aquatic and terrestrial animal taxa. These
include changes in behaviors that may decrease the overall fitness of the organism. There is also evidence
from both invertebrate and vertebrate studies that Pb adversely affects behaviors that may decrease the
ability of an organism to escape predators or capture prey.
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 p38VIAI>K
as possible molecular targets for Pb neurotoxicity in catfish (Leal et al.. 2006). 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 & Wang. 2009).
Decreased food consumption of Pb-contaminated diet has been demonstrated in some invertebrates
(snails) and vertebrates (lizards, pigs). 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 & Xing. 2008). 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 & Gochfeld. 2005). 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 are coherent with findings from studies in laboratory animals described in Sections
2.5.1 and 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). These data from laboratory
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toxicology studies, especially neurobehavioral findings and structural changes are highly coherent with
data from ecological studies. Overall, the evidence from aquatic and terrestrial systems is sufficient to
conclude that there is a causal relationship between Pb exposures and neurobehavioral effects in
invertebrates and vertebrates.
2.6.10. Community and Ecosystem Level Effects
Uptake of Pb into aquatic and terrestrial organisms and subsequent effects on survival,
reproduction, growth, behavior and other physiological variables at the species scale are likely to result in
effects at the population, community and ecosystem scale. The effects may include alteration of predator-
prey dynamics, species richness, species composition, and biodiversity. There are few field studies that
directly consider effects of Pb on these measures of ecosystem health. Ecosystem-level studies are
complicated by the confounding of Pb exposure with other factors such as 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.
Most direct evidence of community and ecosystem level effects is from near point sources. 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 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. Effects of Pb on mortality, growth, physiological stress, blood,
neurobehavioral and developmental and reproductive endpoints at the individual level are expected to
have ecosystem level consequences, and thus provide consistency and plausibility for causality in
ecosystem level effects.
Avoidance response to Pb exposure varies widely in different species and this could affect
community composition. 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).
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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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 (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 & 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.
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 the
lowest number of resident aquatic plant species when compared to sites with lower Pb concentrations (78-
92 |ag Pb/L) (Mishra et al.. 2008). This shift toward more Pb-tolerant species is also observed in terrestrial
plant communities near smelter sites (U.S. EPA. 2006). 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.
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 in natural systems. However, the evidence for Pb effects at higher levels of ecological
organization is sufficient to conclude that there is a causal relationship between Pb exposures and the
alteration of species richness, species composition and biodiversity in terrestrial and aquatic
ecosystems.
2.6.11. Ecological Effects and Corresponding Pb
Concentrations
There is limited evidence to relate ambient air concentrations of Pb to levels of deposition onto
terrestrial and aquatic ecosystems and subsequent movement of atmospherically-deposited Pb though
environmental compartments (e.g., soil, sediment, water, biota). Current evidence indicates that Pb is
bioaccumulated in biota; however, the sources of Pb in biota have only been identified in a few studies,
and the relative contribution of Pb from all sources is usually not known. There are large differences in
species sensitivity to Pb, and many environmental variables (e.g., pH, organic matter) determine the
bioavailability and toxicity of Pb. However, the proportion of observed effects of Pb attributable to Pb
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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.
Threshold levels for Pb in terrestrial and aquatic systems may serve as a tool for interpreting the
effects of atmospherically deposited Pb as a component of total Pb loading. For soils, ecological soil
screening levels (Eco-SSLs) have been developed by the EPA for Pb. Eco-SSLs are maximum
contaminant concentrations in soils that are predicted to result in little or no quantifiable effect on
terrestrial receptors. The Pb Eco-SSL values for terrestrial birds, mammals, plants and invertebrates are
11, 56, 120 and 1,700 mg Pb/kg soil (dry weight), respectively. In aquatic systems, national recommended
ambient water quality criteria have been developed by the EPA Office of Water to protect aquatic life and
human health in surface waters. The ambient water quality criteria for Pb are expressed as a criteria
maximum concentration (CMC) for acute toxicity and criteria continuous concentration (CCC) for
chronic toxicity. In freshwater, the CMC is 65 (.ig Pb/L and the CCC is 2.5 |_ig Pb/L at a hardness of 100
mg/L. In saltwater, these values are 210 (.ig Pb/L CMC and 8.1 (.ig Pb/L CCC, respectively. These U.S.
EPA Office of Water criteria were 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.
2.7. Integration of Health and Ecological Effects
Overview
The health and ecological effects considered for causal determination are summarized in the Table
2-5. The health endpoints include neurological, cardiovascular, renal, immune, hematological and
reproductive effects as well as cancer. The ecological endpoints considered for causal determination are
bioaccumulation, mortality, growth, physiological stress, hematological effects, developmental and
reproductive effects, neurobehavioral effects, and community and ecosystem level effects. The substantial
overlap between the ecological and health endpoints considered in the causal determinations allowed for
the integration of the evidence across these disciplines.
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Table 2-5. Summary of Causal Determinations for Health and Ecological Effects
Outcome
Human Health
Causal Determination
Ecological Receptors
Causal Determination
Neurological Effects3
Causal Relationship
Causal Relationship: Invertebrates and Vertebrates
Cardiovascular Effects
Causal Relationship
N/A
Renal Effects
Causal Relationship
N/A
Immune System Effects
Causal Relationship
N/A
Heme Synthesis and
RBC Function13
Causal Relationship
Causal Relationship: Invertebrates and Vertebrates
Reproductive Effects
and Birth Outcome0
Causal Relationship
Causal Relationship: Invertebrates and Vertebrates
Inadequate to Infer Causal Relationship: Plants
Cancer
Likely to be a causal relationship
N/A
Bioaccumulation
The causal determination for bioaccumulation
was developed respective to the impact of
bioaccumulation on ecosystem services. Thus,
although Pb bioaccumulates in all organisms
including humans, causality was not applicable
to bioaccumulation in humans.
Causal Relationship
Mortality
The strongest evidence of Pb-induced mortality
in humans was observed for cardiovascular
disease related mortality.
Causal Relationship: Invertebrates and Vertebrates
Inadequate to Infer Causal Relationship: Plants
Growth
N/A
Causal Relationship: Plants and Invertebrates
Suggestive of a Causal Relationship: Vertebrates
Physiological Stress
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/A
Causal Relationship
a In 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.
bThe 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.
c 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, were considered.
2.7.1. Modes of Action Relevant to Downstream Health and
Ecological Effects
1 The diverse health and ecological effects of Pb are mediated through multiple, interconnected
2 modes of action. This section summarizes the principle modes of action of human health endpoints
3 associated with Pb exposure and the concentrations at which those effects are observed. Then, effects of
4 Pb observed in organisms in aquatic and terrestrial ecosystems (Section 2.6) are evaluated along with
5 evidence from human and laboratory animals to determine the extent to which common modes of action
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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 2-6). 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. Dose captures Pb exposure
concentrations in in vitro systems or in animal models when no blood Pb level was reported. The
observable effect levels in humans reported in Table 2-6 are limited by the data and methods available 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.
Table 2-6. Related human health effects resulting from the MOAs of Pb and the lowest
level eliciting the MOA
Mode of Action
Related Health Effects
Lowest Level at which MOA
References
(ISA Section)
Observed
Blood Pb
a Dose3
Altered Ion Status
All Heath Effects of Pb
3.5 |jg/dL
50 pM, acute
Huel et al. (2008): Kern et al. (2000)
Protein Binding
Renal (5.5), Heme Synthesis and RBC
Function (5.7)
6.4 |jg/dL
50 |jM, acute
Chen et al. (2010): Klann and
Shelton (1989)
Oxidative Stress
All Heath Effects of Pb
5-10 |jg/dL
10 |jM, acute
Quinlan etal. (1988): Ahamed etal.
(2006): Yiin and Lin (1995)
Inflammation
Neurological (5.3), Cardiovascular (5.4),
Renal (5.5), Immune (5.6), Respiratory
(5.6.4), Hepatic (5.9.1)
3 |jg/dL
0.01 |jM, acute
Kim et al. (2007): Chetty et al. (2005)
Endocrine Disruption
Reproductive Effects and Birth Outcomes
(5.9), Bone and Teeth (5.9.4), Endocrine
System (5.9.3)
2 |jg/dL
20 ppm, acute
Kriea (2007): Wiebe and Barr (1988)
Cell
Death/Genotoxicity
Cancer (5.10), Reproductive Effects and Birth
Outcomes (5.8), Bone and Teeth (5.9.4)
3.1 |jg/dL
50 nM, acute
Van et al. (2004): Bonacker et al.
(2005)
aReported as blood Pb level and dose delivered (human, laboratory animal, and in vitro data).
Ecological studies have presented evidence for the occurrence of many of these modes of action in
animals, and to some degree in plants, however the connection to ecological outcomes must usually be
inferred because ecological studies are typically not designed to address mode of action directly. The level
at which Pb elicits a specific effect is more difficult to establish in terrestrial and aquatic systems due to
the influence of environmental variables on Pb bioavailability and toxicity and substantial species
differences in Pb susceptibility.
The alteration of cellular ion status (including disruption of Ca2+ homeostasis, altered ion transport
mechanisms, and perturbed protein function through displacement of metal cofactors) appears to be the
major unifying mode of action underlying all subsequent modes of action in plants, animals, and humans
(Figure 5-1). Pb will interfere with endogenous cation homeostasis, necessary as a cell signal carrier
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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. Pb-exposure decreases evoked release of neurotransmitters, while
simultaneously increasing spontaneous release of neurotransmitters through Ca2+ mimicry. Evidence for
these effects in Pb-exposed experimental animals and cell cultures have 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 is also responsible for 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. Intranuclear and intracytosolic inclusion body
formation has been observed in the kidney, liver, lung, and brain following Pb exposure to 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. Additionally, metallothionein is an important protein in the
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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 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
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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 clatogenicity 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.
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.
2.8. Policy Relevant Considerations and Human
Health
2.8.1. Air-to-Blood 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). In addition to the meta-analysis of Brunekreef (1984). seven 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 seven studies is 1-9 (ig/dL per (ig/m3, which encompasses
the estimate from the Brunekreef (1984) meta-analysis of (3-6 (ig/dL per (.ig/ni'). The Schnaas (2004) had
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a particularly strong experimental design in that is the only longitudinal study in which blood Pb
concentration was monitored repeatedly in individual children from age 6 months to 10 years. For
children who experienced the largest declines in air Pb (i.e., from 2.8 to <0.1 |_ig/m3). the predicted blood
Pb-air Pb slope (adjusted for age, year of birth, SES, and use of glazed pottery) was 0.213 1 n | j^ig/dL
blood] per ln| |_ig/m3 air]. The cross-sectional study done by Ranft (2008) attempted to account for
potential covariates that influence blood Pb (e.g., soil Pb concentration, gender, environmental tobacco
smoke, fossil heating system and parental education). It is the only study that reported a logarithmic blood
Pb-linear air Pb relationship, which results in an upward curvature of the blood Pb-air Pb relationship
(i.e., the blood Pb-air Pb slope increases with increasing air Pb concentration). In other studies (or based
on other studies), the blood Pb-air Pb relationship was either log-log (Brunekreef. 1984; Haves et al..
1994; Schnaas et al.. 2004). which predicts an increase in the blood Pb-air Pb slope with decreasing air Pb
concentration or linear (Hilts. 2003; Schwartz & Pitcher. 1989; Tripathi et al.. 2001). 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. Because air Pb
contributes to Pb in soil and indoor dusts, adjustment for the correlated covariates such as soil Pb would
introduce a downward bias in the slope estimate.
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Table 2-7. Summary of Estimated Slopes for Blood Pb to Air Pb Relationships in
Children
Reference
Study Methods
Model Description
Blood Pb-Air Pb Slope
(|jg/dL per |jg/m3)
Brunekreef et al. (1984)
Location: Various countries
Years: 1974-1983
Subjects: Children (varying age
ranges)
Analysis: Meta analysis of 18 studies
Model: Log-Log
Blood Pb: 5-41 |jg/dL (mean range for
studies)
Air Pb: 0.2-10 ^g/m3 (mean range for
studies)
All children: 18a, 6.1°
Children <20 ^ig/dL: 13a,3.0b
Hayes et al. (1994)
Location: Chicago, IL
Years: 1974-1988
Subjects: 0.5-6 yr (9,604 blood Pb
measurements)
Analysis: Regression of blood Pb
screening and quarterly average air Pb
Model: Log-Log 24a, 5.7°
Blood Pb: 12-30 |jg/dL (annual GM range)
Air Pb: 0.5-1.2 ^g/m3 (annual GM range)
Hilts et al. (2003) Location: Trail, BC
Years: 1989-2001
Subjects: 0.5-6 yr (292-536 blood Pb
measurements/yr)
Analysis: Regression of blood Pb
screening and community air Pb
following upgrading of a local smelter
Model: Linear
Blood Pb: 4.7-11.5 |jg/dL (annual median
range)
Air Pb: 0.03-1.1 [iglm3 (annual median
range)
6.5
Ranft etal. (2008)
Location: Germany
Years: 1983-2000
Subjects: 6-11 yr (n = 843)
Analysis: Pooled regression 5 cross-
sectional studies
Model: Log-Linear
Blood Pb: 2.2-13.6 ^ig/dL (5th-95th
percentile)
Air Pb: 0.03-0.47 ^ig/m3 (5th-95th
percentile)
3.2°
Schnaas etal. (2004)
Location: Mexico City
Years: 1987-2002
Subjects: 0.5-1 Oyr (n = 321)
Analysis: Regression of longitudinal
blood Pb measurements and annual
average air Pb data
Model: Log-Log
Blood Pb: 5-12 ^ig/dL (annual GM range) 4.8a,1.1b
Air Pb: 0.7-2.8 ^g/m3 (annual mean range)
Schwartz and Pitcher
(1989). U.S. EPA (1986)
Location: U.S.
Years: 1976-1980
Subjects: 0.5-7 yr (n = 7,000)
Analysis: NHANES blood Pb, gasoline
consumption data and Pb
concentrations in gasoline
Model: Linear
Blood Pb: 11-18 |jg/dL (mean range)
Air Pb: 0.36-1.22 ^jg/m (annual maximum
quarterly mean)
9.3
Schwartz and Pitcher
(1989). U.S. EPA (1986)
Location: Chicago, IL
Years: 1976-1980
Subjects: 0-5 yr (n = 7,000)
Analysis: Chicago blood Pb screening,
gasoline consumption data, and Pb
concentrations in gasoline
Model: Linear
Blood Pb: 18-27 |jg/dL (mean range)
Air Pb: 0.36-1.22 ^g/m3 (annual maximum
quarterly mean)
7.7
Tripathi etal. (2001)
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 |jg/dL (regional GM
range)
Air Pb: 0.11-1.18 \iglm3 (regional GM
range)
3.6
aAt an air concentration of 0.15 m/m3
bAt an air concentration of 1 [jg/m
c For a change in air Pb concentration from 0.025 to 0.465 ^g/m3
GM, geometric mean
2.8.2. Concentration-Response Functions
With each successive assessment to-date, the epidemiologic and toxicological study findings show
that progressively lower blood Pb levels are associated with cognitive deficits and behavioral impairments
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as well as other outcomes (U.S. EPA. 2006) (Tables 2-2 and 2-3). Furthermore, in the 2006 Pb AQCD,
compelling evidence for a steeper slope for the relationship between blood Pb level and children's IQ at
lower blood Pb levels was presented 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 & Rothenberg. 2005). and several individual studies.
This body of with the addition of more recent studies is presented 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 slopes at lower blood Pb levels. Relatively few studies examined the
concentration-response relationship between Pb in blood or bone and neurocognitive effects in adults. Of
the studies that did examine this relationship, findings were mixed with some studies reporting a linear
relationship with cognition and others reporting non-linear relationships.
Reference Exposure Period Outcome Blood Pb strata
_anphear et al. (2000) Concurrent Reading score All
<10
<7.5
<5
<2.5
~
~
~ ¦
~
1
Bellinger and Needleman (2003) Early childhood FSIQ >10
<10
~
~
Canfield et al. (2003) Lifetime avg FSIQ All
<10
*
Tellez-Rojo et al. (2006) Concurrent Bayley MDI £10
<10
5-10
<5
~
#
~
~
Hu et al. (2006) Prenatal Bayley MDI/10 >1.226a
<1.226a
~
~
7.5
<7.5
~-
•
~-
* 1
Schwartz (1994) Early childhood FSIQ >15
<15
~
-4.0 -3.0 -2.0 -1.0 0.0 1.0
Change in Cognitive Score (95% CI)
Note: a = Pb levels measured in plasma of maternal blood during 1st trimester of pregnancy. 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. Red symbols represent effect
estimates in lower blood Pb strata. Effect estimates without error bars are from studies that did not provide sufficient information in order to calculate 95% CIs.
Figure 2-2. Comparison of associations between blood Pb and cognitive
function among various blood Pb strata.
Concentration-response relationships were examined in several epidemiologic studies of blood
pressure and mortality. In a study of Korean workers, the Pb-induced increase in systolic blood pressure
was better described by a log linear function of blood Pb level than by a linear function (Weaver. 2010).
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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 and do not conflict with the epidemiologic findings.
Studies investigating both all-cause and cardiovascular mortality report both linear and non-linear
relationships. 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 non4inear relationships between blood Pb level and the hazard ratio for all cause, MI, stroke,
and cancer mortality. This latter analysis provided evidence for an association between blood Pb and
cardiovascular mortality in the NHANES population, where the mean blood Pb level was 2.58 (ig/dL, and
the hazard ratio of Pb with cardiovascular mortality reached its maximum at blood Pb levels between of 6
and 7 (ig/dL. Non4inear relationships between patella bone Pb and log HR for all-cause, cardiovascular,
and IHD mortality were reported by Weisskopf et al. (2009).
Concentration response information was provided in a small number of studies of Pb-related
nephrotoxicity in the occupational setting (Ehrlich et al.. 1998; Weaver et al.. 2003). 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.. 2003).
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 exposures. 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). Non-linear concentration response functions are reported in the air pollution
literature (Smith & Peel. 2010). 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.
In addition, the 2006 Pb AQCD 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 log-normally 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.
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The current body of evidence on the effects of Pb allows critical evaluation of several of the
aforementioned explanations for the supralinear curve concentration response function in epidemiological
studies. For example, 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 (Lanphear et al.. 2000; Lanphear et al.. 2005). 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 overwhelming effects of these other risk factors (Schwartz.
1994). Several studies have 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 5-10). Few studies have compared Pb effect estimates among groups in
different sociodemographic strata, and the limited data are mixed. Some have found greater Pb-associated
neurocognitive deficits in low-SES groups (Bellinger et al.. 1990; Schwartz. 1994). 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
nondisadvantaged populations (-4.5 points [95% CI: -5.6, -2.8]). It is important to note that blood Pb 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 dose-response relationship.
Although, the 2006 Pb AQCD did not identify a biological mechanism for a steeper slope at lower
than at higher blood Pb levels such a mechanism was not ruled out. In fact, several lines of evidence
support the possibility of low-dose and high dose-Pb acting through different mechanisms. For example,
in mice, lower-dose Pb is associated with differential responses of the neurotransmitters dopamine and
norepinephrine compared with control treatment and higher doses (Lcasure et al. 2008; Virgolini et al..
2005). These differential responses of neurotransmitter systems to low-dose Pb versus a higher-dose Pb
may provide mechanistic understanding of the nonlinearity of Pb-induced behavioral changes in animals
and may also explain the nonlinear blood Pb-neurocognitive and neurobehavioral associations reported
widely among children. Additional evidence points to differences in hormonal homeostasis by Pb
exposure level. In male mice with chronic Pb exposure (PND21 to 9 months of age), basal corticosterone
levels are significantly lower in the 50 ppm exposure group versus control or 150 ppm Pb.
Additional mechanistic understanding comes from differences in histological changes found in Pb-
exposed animals. Compared with high-dose Pb, low-dose Pb stimulates greater induction of c-fos, a
marker of neuronal activation and action potential firing (Lewis & Pitts. 2004). These findings may
underlie the nonlinear association between Pb exposure and learning and the U-shaped behavioral dose-
responsiveness seen with amphetamine-induced motor activity in males after GLE (Lcasure et al.. 2008).
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Sensory organ findings also show vastly different outcomes with low versus higher doses of Pb.
High-dose Pb produces subnormal retinal ERGs and low-dose Pb produces supernormal ERGs in both
children (Rothenberg et al.. 2002) and rodents (Fox & Chu. 1988; Fox & Farber. 1988; Foxetal.. 1991).
Inverted U-shaped dose-response curves have been seen for rod photoreceptor numbers or neurogenesis
(Giddabasappa et al.. 2011). retinal thickness (Fox et al.. 2010). and rod cell proliferation (Giddabasappa
et al.. 2011). Thus, these dichotomous histological findings are coherent with the functional retinal test or
the ERG where high-dose Pb produces subnormal ERGs and low-dose Pb produces supernormal ERGs.
Hierarchical enzyme activity also may explain nonlinear Pb dose-response relationships. The
phosphatase enzyme calcineurin has been shown to be inhibited by high dose Pb exposure and stimulated
by low dose Pb exposure (Kern & Audesirk. 2000). At low doses of Pb, Pb displaces calcium at its
binding sites on calmodulin and by acting as a calmodulin agonist at calcineurin's catalytic A subunit,
stimulates calcineurin activity. At high Pb doses, Pb can bind directly to a separate calcium-binding B
subunit, overriding the calmodulin-dependent effect and turning off the activity of calcineurin.
Interestingly, mice with modulated calcineurin expression exhibit aberrant behavior related to
schizophrenia (Mivakawa et al. 2003) or impaired synaptic plasticity and memory (Zeng et al.. 2001).
This example of the stimulatory effects of Pb at low doses and inhibitory effects of Pb at high doses gives
another example of biological plausibility for the non-monotonic dose response of Pb reported in multiple
studies.
2.8.3. Timing and Duration of Exposure
Epidemiologic studies reviewed in the 2006 Pb AQCD observed neurocognitive deficits in
association with prenatal, peak childhood, cumulative childhood, and concurrent blood Pb levels. Among
longitudinal studies that examined blood Pb level at multiple time points, several found that concurrent
blood Pb was associated with the largest decrement in IQ. A common limitation of epidemiologic studies
was the high correlation among Pb exposure metrics at different ages, making it difficult to distinguish
among effects of Pb exposure at different ages and to ascertain which developmental time periods of Pb
exposure were associated with the greatest risk of neurodevelopmental morbidity. Although prospective
cohort studies have provided valuable information on the effects of Pb exposure at different periods of
development, including the prenatal period and early childhood, the limitations noted in the previous 2006
Pb AQCD remain. Collectively, the epidemiologic evidence has not identified one unique time window of
exposure that poses the greatest risk to cognitive function in children (Figure 2-3). However, toxicological
studies included in the 2006 Pb AQCD demonstrated that developmental exposure to Pb was the most
sensitive window for Pb-dependent neurotoxicity and recent toxicological studies continue to support this
finding.
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Reference Outcome Blood Pb Variable Examined
Tong et al. (1998) FSIQ, 11-13 yr. < 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 4
4
-~
Bellinger et al. (1990) GCI, 24 and 57 mo Concurrent, Prenatal < 3a ~
Concurrent, Prenatal 3-10a ~
Concurrent, Prenatal > 10a ~
Chen et al. (2005) FSIQ, 7 yr. Low 2 yr, (<24.9), Low 7 yr, (<7.2) <
Low 2 yr, (<24.9), High 7 yr, (> 7.2) +
High 2yr, (>24.9), Low7yr, (<7.2) 1
High 2 yr, (> 24.9), High 7 yr, (> 7.2) +
>
Hornung et al. (2009) FSIQ, 6 yr. 0.5 ratio age 6 to 2 yrb 4
? n rptin pgp fi tn ? yr b ^
>
-1.0 -0.5 0.0 0.5
Change in Cognitive Score (95% CI)
Note: aEffect estimates represent associations between concurrent blood Pb level and cognitive
function (standardized to standard deviation) in children categorized by prenatal blood Pb level.
bValues represent the ratio of blood Pb level at age 6 years to that at age 2 years. FSIQ = Full-
scale IQ, GCI = General Cognitive Index. Cognitive function scores were standardized to their
standard deviation. Effect estimates in red represent blood Pb level variables associated with the
greater decrease in cognitive function.
Figure 2-3. Associations of cognitive function in children with different
degrees of changes in blood Pb levels overtime.
The 2006 Pb AQCD noted the importance of the duration of exposure in animal studies of a wide
array of effects including neurological, cardiovascular, renal, immune and reproductive effects of Pb.
Generally, epidemiologic studies of the effect of Pb exposure on human health have not been designed to
assess duration of exposure needed to induce those effects. Some exceptions are cohort studies with
repeated blood Pb measurements and cognitive assessments within short intervals of time during
pregnancy and in the first year of life (every 3 months). Studies have reported associations between
prenatal blood Pb levels and decrements in mental development (Bayley MDI) in children assessed
between 3 and 6 months (Bellinger et al.. 1984; Dietrich et al.. 1986; Dietrich et al.. 1987; Shcn ct al..
1998). In particular, Rothenberg et al. (1989) observed that maternal blood Pb levels from week 36 of
pregnancy to delivery was associated with less ability of infants to self-quiet in the first 30 days of life.
Additionally, in the Cincinnati cohort, blood Pb levels measured at neonatal day 10 also were associated
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with MDI decrements at age 3 and 6 months (Dietrich et al.. 1986; Dietrich et al.. 1987). Consistent with
the neurotoxicology literature, findings from these studies indicate that short-duration Pb exposure during
critical periods of in utero or neonatal development are associated with cognitive impairments in young
infants.
2.8.3.1. Persistence of Effects
The issue of persistence of the neurodevelopmental effects of 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. However, in these
studies, blood Pb levels remained relatively stable over time. Thus, the effect of concurrent exposures was
not ruled out. In addition, the persistence of effect appears to depend on the duration of exposure as well
as other factors that may affect an individual's ability to recover from an insult. Toxicological studies
from the 2006 Pb AQCD highlighted the importance of Pb exposure in early life in promoting
Alzheimer's like pathologies in the adult brain, demonstrating Pb-induced neurodegeneration and
formation of neurofibrillary tangles. 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.
Blood Pb generally is not associated with Alzheimer's disease in epidemiological studies of adults.
However, recent evidence indicates associations between early life ALAD activity, a biomarker of Pb
exposure, and schizophrenia later in adulthood. Consistent with these findings, toxicological studies have
observed Pb-induced emotional changes in males and depression changes in females.
2.8.4. Susceptible Populations and Lifestages
Potential susceptibility factors examined in Chapter 6 of this document include lifestage, sex,
genes, pre-existing diseases/conditions, race and ethnicity, SES, BMI, nutrition, stress, cognitive reserve
(e.g., the resilience of the mind), and co-exposure to other metals/toxicants. Studies are included in
Chapter 6 if the findings for the susceptible population or lifestage were compared across strata with and
without the potential susceptibility factor. By virtue of their design some cohort studies, including cohort
studies of pregnant women or other populations or lifestages with no comparison group, are discussed in
the endpoint-specific sections rather than in Chapter 6. This integrative summary, however, draws on
evidence relating to potentially susceptible populations and lifestages appearing throughout this
document. Also in Chapter 6, separate discussions of studies that evaluate factors that influence Pb
exposure and uptake, and studies that examine the modification of the association of Pb with health
endpoints by a possible susceptibility factor are presented. In this integrative summary both types of
studies are considered together.
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2.8.4.1. Children
Children may be more highly exposed to Pb compared to adults without occupational exposure to
Pb, through their behaviors (e.g., hand-to-mouth contact). Blood Pb levels are highest among the
youngest children and decrease with increasing age of the child (Table 6-1). Biokinetic factors that vary
by age, including bone turnover and absorption, also affect blood Pb levels. Childhood, as a susceptibility
factor related to Pb exposure and dose, is discussed in more detail in sectin 6.1.1.1. The kinetics of Pb,
and how absorption, distribution, and elimination may vary depending on lifestage, is discussed in
Section 4.2.
It is recognized that Pb can cross the placenta to affect the developing nervous system of the fetus
(Sections 4.2.2.4, 5.3.2.1) and there is evidence of increased susceptibility to the neurocognitive effects of
Pb exposure during several lifestages throughout childhood and into adolescence (for more detail, see
Section 5.3.2.1). Further, Pb exposure is associated with effects on the renal (Section 5.5.2.3), immune
(Section 5.6) and heme synthesis and RBC function (Section 5.7) of children. A limited number of studies
of immune parameters, transferring saturation, and iron-deficiency anemia that stratified children by age
report stronger associations among the youngest children. Childhood, as a susceptibility factor related to
Pb-induced health effects, is discussed in more detail in Section 6.2.1.1.
2.8.4.2. Adults
There is evidence that both recent and/or cumulative exposure to Pb may result in health outcomes
during adulthood, as indicated by consistent associations of blood Pb or bone Pb with cardiovascular
diseases 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. Mobilization
(Section 4.2.2.2) of Pb from the bone stores may occur during periods of physiological stress, including
older adulthood (Section 6.1.1.2). In recent studies, age was specifically examined as an effect modifier of
the association of Pb with mortality, cognition and blood pressure in adults; findings for mortality were
mixed while little evidence of modification by age was reported for the other specific outcomes.
Toxicological studies support the plausibility of differences in susceptibility to health effects depending
on lifestage.
2.8.4.3. Sex
In a recent NHANES analysis and several other studies, gender-based differences in blood Pb level
were observed among the adolescent and adult age groups, with higher blood Pb levels among males. The
gender-based differences were not substantial among the youngest age groups (1-5 years old and 6-11
years old). Studies of effect measure modification of the association of Pb and various health endpoints
including neurological effects such as cognition, kidney function blood pressure immune effects and
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cancer by sex were conducted. Overall, findings were mixed with the most consistent evidence reported
for neurological endpoints. Recent epidemiologic evidence increased the consistency of collective
evidence base and pointed to males having increased susceptibility for Pb-associated neurotoxicity.
Toxicological studies continue to demonstrate increased susceptibility of males for endpoints such as
sensory function, balance, stress hormone homeostasis, and brain membrane composition. Sex differences
were also observed in toxicological studies across a wide array of endpoints including of behavior,
memory, gross motor skills, obesity, and retinal decrements. See Section 6.1.2 and 6.2.2 for more details.
Hormone levels may affect susceptibility to Pb-related health effects and associations among
females may vary based on hormonal status, such as menopause or ovary removal. Toxicological and
epidemiologic evidence supports the potential susceptibility to Pb effects based on hormonal status with
findings on delayed onset of puberty and changes to the female reproductive tract.
2.8.4.4. Race and Ethnicity
Higher blood Pb and bone Pb levels among African Americans have been documented and recent
studies continue support previous findings. Further, larger proportions of both non-Hispanic blacks and
Mexican American children have blood Pb levels exceeding 5 (ig/dL of blood. The 2006 Pb AQCD noted
that a clear downward temporal trend was apparent in NHANES data during the previous two decades but
that blood Pb level was declining at different rates for groups within the population, which were defined
by race, as well as income and demographic factors include age of housing. Recent data suggest that the
gap in Pb exposure between African American and White subjects is decreasing, but African Americans
still tend to have higher blood Pb levels. Evidence of increased association of Pb with cardiovascular
outcomes among non Hispanic blacks and Mexican Americans was reported for the NHANES population.
Results of other recent epidemiological studies suggest that there may be race related susceptibility for
additional outcomes but the evidence is limited and confounding or modification by other factors, such as
SES, may be present. See Sections 6.1.3 and 6.2.7 for additional details.
2.8.4.5. Socioeconomic Status
The 2006 Pb AQCD noted that the geometric mean blood Pb concentration varied with SES and
other demographic characteristics that have been linked to Pb exposure. The gap between SES groups
with respect to Pb level appears to be diminishing, with Pb level being higher but not significantly higher
among lower income subjects. Lower SES individuals appear to represent a susceptible population. For
example, a study of Pb and IQ reported greater inverse associations among those in the lowest SES
groups. 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 contrast, there is
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some evidence that Pb-associated neurocognitive effects may be larger in magnitude among higher SES
populations. 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. See Sections 6.1.4 and
6.2.8 for additional details.
2.8.4.6. Genes
Several genes were examined as potential modifiers the associations between Pb and health effects.
Epidemiologic and toxicological studies reported ALAD and VDR variants may be health-related
susceptibility factors. Although limited, evidence suggests that risk of Pb-associated neurocognitive
deficits in children also may be modified by variants in genes for APOE, MTHFR, and dopamine
receptors. Other genes examined that may also affect susceptibility to Pb-related health effects were
DRD4, GSTM1, TNF-alpha, eNOS, and HFE. See Section 6.2.4 for additional details.
2.8.4.7. Pre-existing Conditions
Pre-existing diseases/conditions also have the potential to affect the association of Pb exposure
with health endpoints have been studied in relation to autism, diabetes, and hypertension. Recent
epidemiologic studies did not support modification of Pb and health endpoints by diabetes; however, past
studies have found diabetics to be a more susceptible population with regard to effects on renal function.
Recent epidemiological studies support finding from the 2006 Pb AQCD that hypertension is observed to
be a susceptibility factor i with both renal effects and heart rate variability demonstrating stronger
associations among hypertensive individuals compared to those that are normotensive. Although the
evidence is limited, current research has shown that in autistic children, blood Pb level is differentially
correlated with expression of immune-related genes. See Section 6.2.5 for additional details.
2.8.4.8. Nutrition and Lifestyle Factors
Body mass index (BMI), alcohol consumption, and nutritional factors were examined in recent
epidemiologic and toxicological studies. Modification of associations between Pb and various health
effects (mortality and heart rate variability) was not observed by BMI or 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 possibility of alcohol as a susceptibility factor. Among nutritional
factors, those with iron deficiencies were observed to be a susceptible population for Pb-related health
effects in both epidemiologic and toxicological studies. Other nutritional factors, such as calcium, zinc,
and protein, 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 modification. Folate was also examined in a recent epidemiologic study of birth size but no interaction
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was reported between Pb and folate. Further study of these and other nutritional factors will be useful in
determining susceptibility among individuals with various nutritional levels/deficiencies. See Sections
4.2, 6.2.9, 6.2.10, and 6.2.11 for additional details.
2.8.4.9. Stress and Cognitive Reserve
Animal toxicology findings described in the 2006 Pb AQCD demonstrated interactions between Pb
exposure and stress. Namely, 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.
In comparison, epidemiologic evidence for such interactions has been sparse. However, consistent with
historical animal studies, a recent epidemiological study indicated that positive social environment of
children as characterized by maternal self-esteem, may lessen the impact of Pb exposure on
neurodevelopment. Self-perceived stress was shown to modify the association of bone Pb with
hypertension. In addition, a greater association of Pb with cognitive function was found in workers with
lower cognitive reserve (Sections 6.2.12 and 6.2.13).
2.8.4.10. Co-exposure of Lead with Metals or Other Chemicals
The 2006 Pb AQCD reported that the majority of studies examined other chemicals as confounders
and not effect measure modifiers (U.S. EPA. 2006). Although the body of evidence remains limited,
recent epidemiologic studies have begun to explore the possible interaction between Pb and other metals
or chemical agents. These studies report some stronger associations between Pb and various health
endpoints with co-exposure to Cd, As, Mn, fluoride, tobacco smoke and urban pollutants.
Epidemiologic and toxicological studies have reported increased susceptibility to Pb-related health
effects among those with high Cd levels. Modification of associations of Pb with levels of reproductive
hormones and renal dysfunction were reported in epidemiologic studies. Toxicological evidence of a
synergism between Pb and Cd with regard to renal toxicity supported the epidemiological evidence. In
addition, exposure to Pb and As was associated with effects on immune function in children living near a
smelter and a statistical interaction between the metals was observed. Studies suggest that co-exposure to
As may increase the bioavailability of Pb establishing the plausibility of increased susceptibility of Pb-
related health effects when co-exposed to As. An interaction was also reported between Pb and Mn (Y
Kim et al.. 2009) in a study of IQ. F1 has been identified as a potential susceptibility factor in a
toxicological study but has not yet been explored in epidemiologic studies. The toxicological reported that
co-exposure of Pb and F1 increased Pb deposition in calcified tissues. Since Pb is acid soluble,
fluoridation may increase Pb concentration in water through leaching from pipes and Pb solder.
Additional details of the studies summarized, are found in Sections 6.2.14 and 6.2.15.
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Recent studies suggested that Pb-associated neurotoxicity in children are exacerbated with
co-exposures to environmental tobacco smoke but findings from recent epidemiologic studies examining
modification by smoking for other outcomes are mixed. Exposure to urban areas with larger industrial
sources and higher traffic has been suggested to increase Pb body burden and risk of Pb exposure. An
ecological study reports an association of accumulated metals in the soil (e.g., Pb, Zn, Cd, Ni, Mn, Cu,
Co, Cr and V) with reduced learning achievement among the students at the school (Mielke et al.. 2005).
2.9. Summary
This section summarizes the main conclusions of this assessment regarding the health and
ecological effects of Pb and, for health outcomes, the concentrations at which those effects are observed.
The conclusions from the 2006 Pb AQCD and the causality determinations for the health and ecological
effects of Pb exposure from this review are summarized in Table 2-8.
The 2006 Pb AQCD reviewed a strong body of evidence clearly substantiating the health effects of
Pb at contemporary exposure levels. Neurological effects in children and cardiovascular effects in adults
were the effects that were best substantiated as occurring at blood Pb concentrations as low as 5 to 10
(ig/dL. Other newly demonstrated immune and renal system effects among general population groups
were also emerging as low-level Pb-exposure effects of potential public health concern. New
epidemiologic and toxicological studies support the findings of the previous assessement and provide
additional evidence for these effects at increasingly lower levels.
The major conclusions reached in the 2006 Pb AQCD for terrestrial ecosystems focused on
evidence from smelter sites or other industrial point sources with elevated levels of Pb where death of
vegetation was found to cause a near-complete collapse of the detrital food web, creating a terrestrial
ecosystem in which energy and nutrient flows were minimal. In aquatic ecosystems, the best documented
links between Pb and effects on the environment were with effects on individual organisms.
Bioaccumulation of Pb in aquatic organisms was shown to alter the aquatic environment. Further, it was
noted that alteration of ecological interactions (e.g., species competitive behaviors, predator-prey
interactions, and contaminant avoidance behaviors) may have negative effects on species abundance and
community structure. However, the fact that both terrestrial and aquatic systems frequently contain
multiple metals and other stressors made it difficult to attribute observed adverse effects to Pb.
The current document presents detailed reviews of the scientific literature on the effects of Pb in
aquatic and terrestrial ecosystems, but, also integrates the evidence of effects across aquatic and terrestrial
habitats. The ecological endpoints considered for causal determination are bioaccumulation, mortality,
growth, physiological stress, hematological effects, developmental and reproductive effects, and
neurobehavioral effects. The substantial overlap between the ecological and health endpoints considered
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1 in the causal determinations allowed the integration of the evidence across these disciplines. This
2 organizational scheme represents an evolution from the structure of the 2006 Pb AQCD and lends
3 additional weight to the ecological evidence by reducing the uncertainty associated with attributing
4 ecosystem effects to Pb rather than other metals or toxicants that co-occur with Pb since much of the
5 evidence documenting the effects and modes of action of Pb comes from controlled animal experiments.
Table 2-8. Summary of evidence from epidemiologic, animal toxicological and ecological
studies on the effects associated with exposure to Pb
Health Outcome Conclusions from the 2006 Pb Conclusions from the 2011 1st Draft Pb
AQCD ISA
Neurological
Effects
Causal relationship
Neurocognitive The collective body of epidemiologic
Function and studies provided clear and consistent
Learning evidence for the effects of Pb exposure
on decreasing neurocognitive function in
children.
Recent epidemiologic studies in children continue
to demonstrate associations with IQ; most
evidence emphasizes associations of blood Pb
levels as low as 2 pg/dL with specific indices of
neurocognitive function (e.g., verbal skills,
memory, learning visuospatial processing). Among
environmentally exposed adults, the most
consistent findings were associations of
cumulative Pb exposure metrics with cognitive
deficits.
Neurobehavioral
Effects
Several epidemiologic studies reported
associations between Pb exposure and
that ranged from inattentiveness to self-
reported delinquent behaviors to criminal
activities. Uncertainty remained
regarding the critical time period of Pb
exposure. In addition, uncertainties
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 depression and
anxiety symptoms.
Recent studies in children continue to support
associations of Pb exposure (blood Pb levels 3-11
pg/dL) with a range of effects from anxiety and
distractibility to conduct disorder and delinquent
behavior. New evidence indicates associations
between blood Pb levels as low as 1 pg/dL and
ADHD diagnosis and contributing diagnostic
indices.
Sensory Organ The selective action of Pb on retinal rod No new epidemiologic studies on sensory organ
Function cells and bipolar cells is well function. Recent toxicological studies find retinal
documented in earlier AQCDs. There effects in male rodents at lower blood Pb levels
was coherence between the animal and (~12 pg/dL)
the human literature on the effects of
chronic Pb exposure on auditory
function.
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Neurodegenerative Several epidemiologic studies of Pb Recent epidemiological studies report associations
Diseases exposure and Alzheimer's disease, or with PD and essential tremor. Emerging
dementia did not report associations toxicological evidence suggests that early life
association, but each study had sufficient exposure may be associated with
limitations. neurodegeneration in adult animals.
Cardiovascular Epidemiologic studies consistently Causal relationship
Effects demonstrated associations between Pb
exposure and increased risk of CVD
outcomes. Experimental toxicology
confirmed Pb effects on the
cardiovascular system observed in
epidemiologic studies.
Blood Pressure A doubling of blood Pb level was Recent studies confirm these findings.
associated with a 1 mmHg increase in Associations of increased BP with Pb exposure
systolic blood pressure and a 0.6 mmHg observed at blood Pb levels < 2 pg/dL.
increase in diastolic blood pressure.
Hypertension Suggestive evidence that cumulative Pb Recent studies, including those using bone Pb as
exposure may be associated with a metric of cumulative exposure, confirm and add
hypertension. Animal studies weight to previous findings. Associations of
demonstrated that long-term exposure to hypertension with Pb exposure observed at blood
Pb resulted in hypertension that Pb levels < 2 pg/dL. Recent studies have
persisted after cessation of exposure. emphasized the interaction of cumulative
exposure to Pb with other factors including stress.
Cardiovascular Limited evidence in support of
Mortality cardiovascular mortality.
Recent studies including an NHANES analysis of
the association of blood Pb with cause-specific
mortality and study of older adults, which used
bone Pb as an exposure metric, addressed
limitations of previous studies and provide
additional evidence for an association of Pb with
cardiovascular mortality.
Renal Effects Circulating and cumulative Pb exposure Causal relationship
was associated with longitudinal decline Recent studies expand upon the strong body of
in renal function. Experimental studies evidence that Pb exposure is associated with
demonstrated that initial accumulation of k^ney dysfunction including increased serum
absorbed Pb occurred primarily in the creatinine and decreased creatinine clearance at
kidneys and hyperfiltration phenomenon |3|00Cj levels < 5 pg/dL
during the first 3 months of exposure
was noted.
Immune System
Effects
Epidemiologic studies suggested that Pb
exposure may be associated with effects
on cellular and humoral immunity
including changes in serum
immunoglobulin E levels in children.
Toxicological evidence supported these
findings and provided evidence for
effects on downstream events such as
inflammation and decreased host
resistance.
Causal relationship
Recent studies support the strong body of
evidence that Pb exposure is associated with both
cell-mediated and humoral immunity. The
consistency and coherence of findings among
related immune effects establishes the biological
plausibility for epidemiologic observations of
associations with infection, allergy and effects in
other organ systems.
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Heme Synthesis
and RBC Function
Pb exposure was associated with Causal relationship
disruption in heme synthesis in both Recent epidemiologic and toxicological studies
children and adults. Toxicological studies provide strong evidence that exposure to Pb is
demonstrated that Pb interferes with red associated with numerous deleterious effects on
blood cell survival and mobility. hematological system including altered heme
synthesis mediated through decreased ALD and
ferrochelatase activities, decreased RBC survival
and function, decreased hematopoiesis and
increased oxidative stress and lipid peroxidation.
Reproductive
Effects and Birth
Outcomes
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.
Causal relationship
Recent toxicological and epidemiologic studies
provide strong evidence for delayed onset of
puberty in males and females as well as 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
workers suggested a relationship
between Pb and cancers of the lung and
the stomach; the evidence was limited by
confounding by metal co-exposures
(e.g., to As,Cd), smoking, and dietary
habits. The 2003 NTP and 2004 IARC
reviews concluded that Pb and Pb
compounds were probable carcinogens,
based on limited evidence in humans
and sufficient evidence in animals.
Based on animal data and inadequate
human data Pb and Pb compounds
would be classified as likely carcinogens
according to the EPA Cancer
Assessment Guidelines for Carcinogen
Risk Assessment.
Likely causal relationship
Some epidemiologic evidence supporting
associations between Pb and cancer with the
strongest evidence from animal toxicology
between Pb and cancer,
genotoxicity/clastogenicity, or epigenetic
modification.
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Ecological/ Welfare Findings from 2006 Pb AQCD
Effect
Conclusions in the 2011 1st Draft ISA
Bioaccumulation
as it affects
Ecosystem
Services
Atmospheric Pb pollution has resulted in
accumulation of Pb in terrestrial and
aquatic systems throughout the world. In
field studies in aquatic systems, Pb has
been shown to significantly alter the
aquatic environment through
bioaccumulation and alterations of
community structure and function. Due to
low solubility of Pb in water, dietary Pb
(i.e., Pb absorbed to sediment,
particulate matter and food) may
contribute substantially to exposure and
toxicity.
Causal relationship
There is compelling evidence from several
decades of research and in recent studies that Pb
bioaccumulates in plants, invertebrates and
vertebrates in terrestrial and aquatic ecosystems.
Pb deposited on the surface of, or taken up by
organisms has the potential to contribute to
exposure and toxicity in primary and secondary
order consumers and alter the services provided
by ecosystems (i.e., provisioning services).
Mortality
No information on mortality in plants.
Effects of Pb on invertebrates and
vertebrates include decreased survival.
Inadequate to infer a causal relationship for plants
Insufficient evidence for mortality in plants.
Causal relationship in vertebrates and
invertebrates
Recent studies provide additional evidence for Pb
effects on mortality in invertebrates and
vertebrates.
Growth
Evidence of growth effects in algae,
aquatic plants, soil invertebrates and
aquatic invertebrates. Limited evidence
in avian and mammalian consumers.
Causal relationship in plants and invertebrates
Recent studies on growth in algae and
invertebrates find effects of Pb at lower
concentrations than previously, and additional
evidence for growth effects in plants.
Suggestive causal relationship in vertebrates
Limited studies considered effects on growth in
vertebrates.
Physiological Pb exposure may cause lipid
Stress peroxidation and changes in glutathione
concentrations. There are species
differences in resistance to oxidative
stress.
Causal relationship
Recent studies support the strong body of
evidence for upregulation of antioxidant enzymes
and increased lipid peroxidation associated with
Pb exposure in many species of plants,
invertebrates and vertebrates.
Hematological
Effects
Pb effects on heme synthesis were
documented in the 1986 AQCD and
continue to be studied in aquatic and
terrestrial biota. Changes in ALAD are
not always related to adverse effects but
may simply indicate exposure.
Numerous studies have reported the
effects of Pb exposure on blood
chemistry in aquatic and terrestrial biota.
Causal relationship
Recent studies expand the evidence for Pb effects
on ALAD enzyme activity in bacteria,
invertebrates, and vertebrates and altered serum
profiles and blood cell counts in vertebrates.
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Development and
Reproduction
No information on reproduction in plants.
Limited new evidence in invertebrates
and vertebrates.
Inadequate to infer a causal relationship for plants
There are an insufficient number of studies that
consider Pb effects on plant reproduction.
Causal relationship in invertebrates and
vertebrates
Recent studies expand the evidence for Pb effects
on embryonic development as well as for multi-
generational effects in invertebrates. In
vertebrates, there is new evidence for delayed
metamorphosis and altered steroid profiles in the
few species studied.
Neurobehavior
Exposure to Pb has been shown to affect Causal relationship
brain receptors in fish. Exposure to Pb in
laboratory studies and simulated
ecosystems may alter species
competitive behaviors, predator-prey
interactions and contaminant avoidance
behaviors.
Recent studies identify possible molecular targets
for Pb neurotoxicity in invertebrates and fish.
There is new evidence in a few invertebrate and
vertebrate species for behavioral effects
associated with Pb exposure.
Community and
Ecosystem Level
Effects of Pb difficult to interpret because Causal Relationship
of the presence of other stressors
including metals.
Uptake of Pb into aquatic and terrestrial organisms
and subsequent effects on survival, reproduction,
growth, behavior and other physiological variables
at the species scale is likely to lead to effects at
the population, community and ecosystem scale.
There is additional evidence for effects of Pb in
soil microbial communities, and in sediment-
associated and aquatic plant communities.
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Chapter 3 Contents
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 inPb 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. 3-3
Figure 3-3. Nationwide stationary source Pb emissions (tons/yr) in the U.S. by source
sector in 2008. 3-4
Figure 3-4. County-level Pb emissions (tons) from stationary sources in the U.S. in
2008. 3-5
Figure 3-5. Pb emissions (tons) at airports in the U.S. in 2008. The size of the symbol
indicates the magnitude ofPb emissions at each airport. 3-6
3.2.2. Anthropogenic Sources 3-6
3.2.2.1. Lead Emissions from Piston Engine Aircraft Operating on Leaded Aviation Gasoline
and Other Non-Road Sources 3-6
3.2.2.2. Fugitive Emissions from Metals Processing and Mining 3-7
3.2.2.3. Fossil Fuel Combustion 3-8
3.2.2.4. Other Industrial Sources 3-9
3.2.2.5. Roadway-Related Sources 3-10
3.2.2.6. Deposited Lead 3-11
Figure 3-6. Total U.S. Pb additives in on-roadgasoline used in on-road vehicles, 1927-
1995. 3-13
Figure 3-7. EstimatedPb aerosol inputs from on-road gasoline into 90 U.S. urbanized
areas (UAs), from 1950 through 1982. 3-15
3.2.3. Source Attribution 3-15
3.2.3.1. Lead Speciation and Source Apportionment 3-15
Table 3-1. Pb compounds observed in the environment 3-16
3.2.3.2. Lead Isotope Ratio Analysis 3-19
3.3. Fate and Transport of Lead 3-20
Figure 3-8. Fate of atmospheric lead. 3-21
3.3.1. Air 3-21
3.3.1.1. Transport 3-22
Figure 3-9. Scales of turbulence within an urban environment. 3-24
3.3.1.2. Deposition 3-25
3.3.1.3. Resuspension of Lead from Soil to Air after Lead Deposition 3-27
3.3.2. Water 3-28
3.3.2.1. Lead Transport in Water and Sediment 3-29
3.3.2.2. Deposition of Lead within Bodies of Water and in Sediment 3-30
Table 3-2. Surface sediment Pb concentrations for various continental shelves; see
Fang et al. (2009) and references therein. 3-31
3.3.2.3. Flux of Lead from Sediments 3-32
3.3.2.4. Lead in Runoff 3-33
3.3.3. Soil 3-40
3.3.3.1. Deposition of Lead onto Soil from Air 3-41
3.3.3.2. Sequestration of Lead from Water to Soil 3-43
3.3.3.3. Movement of Lead within the Soil Column 3-45
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Figure 3-10. Schematic model summarizing the estimated flux ofPb within a typical
podzol profile from northern Sweden using data from Klaminder et al.
(2006). 3-47
Figure 3-11. Eh-pH diagram for Pb in shooting range soils, Jefferson National Forest,
VA. 3-51
3.4. Monitoring of Ambient Lead 3-52
3.4.1. Ambient Measurement Techniques 3-52
3.4.1.1. Size-Selective PM Monitoring for Lead Concentrations 3-52
3.4.1.2. Federal Reference Method and Federal Equivalence Method Evaluation 3-53
Table 3-3. Specifications for Pb monitoring 3-54
3.4.1.3. Chemical Speciation Network, IMPROVE, and National Air Toxics Trends Network
Monitors 3-55
3.4.1.4. Other Measurement Methods for Total Lead 3-56
3.4.1.5. Sequential Extraction 3-57
3.4.1.6. Speciation Techniques 3-58
3.4.1.7. Continuous Lead Monitoring 3-60
3.4.2. Ambient Network Design 3-60
3.4.2.1. Monitor Siting Requirements 3 -62
3.4.2.2. Spatial and Temporal Coverage 3-63
Figure 3-12. Pb monitoring sites for SLAMS, CSN, NATTS, and IMPROVE networks,
2007-2009. 3-63
3.5. Ambient Air Lead Concentrations 3 -64
3.5.1. Spatial Distribution of Air Lead 3-64
3.5.1.1. Variability across the U.S. 3-64
Table 3-4. Summary data for source orientedPb monitors across the U.S. 3-65
Table 3-5. Summary data for sites at which source oriented statistics are at a maxima 3-65
Figure 3-13. Highest county-level source orientedPb-TSP concentrations (f.ig/m3),
maximum 3-month average, 2007-2009. 3-66
Figure 3-14. Highest county-levelPb-PMI0 concentrations (uy//n I, maximum 3-month
average, 2007-2009. 3-67
Figure 3-15. Highest county-levelPb-PMzsConcentrations (fig/m3), maximum 3-month
average, 2007-2009. 3-68
3.5.1.2. Intra-urban Variability 3-68
3.5.2. Temporal Variability 3-70
3.5.2.1. Multi-year Trends 3-70
Figure 3-16. National trends in Pb concentration (fig/m3), all FRM monitors, 1980-2009. 3-71
Figure 3-17. National trends inPb concentration (u.g in ), source oriented FRM
monitors, 1990-2009. 3-72
Figure 3-18. National trends inPb concentration (u.g-ln ). non-source oriented FRM
monitors, 1990-2009. 3-72
3.5.2.2. Seasonal Variations 3-73
Figure 3-19. Monthly source oriented Pb-TSP average (u.gin) over 12 months of the
year, 2007-2009. 3-74
Figure 3-20. Monthly non-source oriented lead-TSP average (u.gin ) over 12 months of
the year, 2007-2009. 3-75
Figure 3-21. Monthly lead-PMw average (fig/m3) over 12 months of the year, 2007-2009. 3-75
Figure 3-22. Monthly lead-PM2.s average (fig/m3) over 12 months of the year, 2007-2009. 3-76
3.5.3. Size Distribution of Lead-Bearing PM 3 -77
3.5.3.1. AQS Data Analysis 3-77
Table 3-6. Summary of comparison data for co-located lead-TSP and lead-PMI0
monitors. 3-77
Table 3-7. Summary of comparison data for co-located Pb-TSP andPb-PM2.5 monitors 3-78
Table 3-8. Summary of comparison data for co-located Pb-PMw andPb-PM2.5
monitors. 3-78
3.5.3.2. Size Distribution Studies in the Literature 3-79
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Figure 3-23. Comparison of urban background and near-road size fractions of lead-
bearing PM. 3-81
3.5.4. Lead Concentrations in aMultipollutant Context 3-82
Figure 3-24. Correlations of monitored Pb-TSP concentration with copollutant
concentrations, 2007-2008. 3-83
Figure 3-25. Correlations of monitored Pb-TSP concentration with co-pollutant
concentrations, 2009. 3-84
Figure 3-26. Correlations of monitored lead-PM2.5 concentration with copollutant
concentrations, 2007-2009. 3-86
3.6. Ambient Lead Concentrations in Non-Air Media and Biota 3-87
3.6.1. Soils 3-88
Table 3-9. Outdoor soilPb levels in various cities within the U.S. 3-88
Figure 3-27. Map of median Pb content in soil in New Orleans. 3-90
3.6.2. Sediments 3-92
Table 3-10. Sediment concentrations in various cities, prior to 2005 3-93
Figure 3-28. 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. 3-94
Figure 3-29. 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. 3-95
3.6.3. Rain 3-96
Figure 3-30. Trends inPb concentration in precipitation from various sites in Norway
over the period 1980-2005. 3-97
3.6.4. Snowpack 3-97
3.6.5. Natural Waters 3-99
3.6.6. Moss 3-100
3.6.7. Grass, Foliage, and Tree Rings 3-100
Figure 3-31. Trends in regional pollution near a copper smelter in Canada andPb
concentrations at the boundary ofheartwood trees within roughly 75 km of
the smelter. 3-101
3.6.8. Aquatic Bivalves 3-101
3.7. Summary 3-102
3.7.1. Sources of Atmospheric Lead 3-102
3.7.2. Fate and Transport of Lead 3-102
3.7.3. Ambient Lead Monitoring 3-104
3.7.4. Ambient Air Lead Concentrations 3-104
3.7.5. Ambient Lead Concentrations in Non-Air Media and Biota 3-105
Chapter 3 References 3-107
Chapter 3 Appendix 3-129
3.8. Variability across the U.S. 3-129
Table 3A-1. Distribution of 1-month average Pb-TSP concentrations (jig/m3) nationwide,
source-oriented monitors, 2007-9. 3-129
Table 3A-2. Distribution of 3-month moving average Pb-TSP concentrations (jig/m3)
nationwide, source-oriented monitors, 2007-9. Sites listed in the bottom six
rows of the table fall in the upper 90th percentile of the data pooled by site. 3-130
Table 3A-3. Distribution of annual 1-month site maxima TSP Pb concentrations (jig/m3)
nationwide, source-oriented monitors, 2007-2009. 3-131
Table 3A-4. Distribution of annual 3-month site maxima Pb-TSP concentrations (ng/m3)
nationwide, source-oriented monitors, 2007-2009. 3-131
Table 3A-5. One-month average Pb-TSP for individual county concentrations
nationwide (fig/m3), non-source-oriented monitors, 2007-2009 3-132
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Table 3A-6. Three-month moving average Pb-TSP for individual county concentrations
({ig/m3) nationwide, non-source-oriented monitors, 2007-2009 3-133
3.8.1. Intra-urban Variability 3-133
Figure 3A-1. Pb TSP monitor and source locations within Los Angeles County, CA (06-
037), 2007-2009. 3-137
Figure 3A-2. Wind roses for Los Angeles County, CA, from meteorological data at the
Los Angeles International Airport, 1961-1990. 3-138
Figure 3A-3. Box plots of annual and seasonal Pb TSP concentrations (ng/m3) from
source-oriented and non-source-oriented monitors within Los Angeles
County, CA (06-037), 2007-2009. 3-139
Table 3A-7. Correlations between Pb TSP concentrations from source-oriented and non-
source-oriented monitors within Los Angeles County, CA (06-037), 2007-
2009 3-140
Figure 3A-4. Pb TSP monitor locations within Hillsborough and Pinellas Counties, FL
(12-057 and 12-103), 2007-2009. 3-142
Figure 3A-5. Wind roses for Hillsborough/Pinellas Counties, FL, obtained from
meteorological data at Tampa International Airport, 1961-1990. 3-143
Figure 3A-6. Box plots of annual and seasonal Pb TSP concentrations (jig/m3) from
source-oriented and non-source-oriented monitors within Hillsborough and
Pinellas Counties, FL (12-057 and 12-103), 2007-2009. 3-144
Table 3A-8. 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. 3-145
Figure 3A-7. Pb TSP Monitor locations within Cook County, 1L (17-031), 2007-2009. 3-148
Figure 3A-8. Wind roses for Cook County, 1L, obtained from meteorological data at
O'Hare International Airport, 1961-1990. 3-149
Figure 3A-9. Box plots of annual and seasonal Pb TSP concentrations (jig/m3) from
source-oriented and non-source-oriented monitors within Cook County, 1L
(17-031), 2007-2009. 3-150
Table 3A-9. Correlations between Pb TSP concentrations from source-oriented and non-
source-oriented monitors within Cook County, 1L (17-031), 2007-2009. 3-151
Figure 3A-10. Pb TSP Monitor locations within Jefferson County, MO (29-099), 2007-
2009. 3-153
Figure 3A-11. Wind roses for Jefferson County, MO, obtained from meteorological
data at St. Louis/Lambert International Airport, 1961-1990. 3-154
Figure 3A-12. Box plots of annual and seasonal Pb TSP concentrations (jig/m3) from
source-oriented and non-source-oriented monitors within Jefferson County,
MO (29-099), 2007-2009. 3-155
Table 3A-10. Correlations between Pb TSP concentrations from source-oriented and non-
source-oriented monitors within Jefferson County, MO (29-099), 2007-2009 3-156
Figure 3A-13. Pb TSP Monitor locations within Cuyahoga County, OH (39-035),
2007-2009. 3-158
Figure 3A-14. Wind roses for Cuyahoga County, OH, obtained from meteorological
data at Cleveland/Hopkins International Airport, 1961-90. 3-159
Figure 3A-15. Box plots of annual and seasonal Pb TSP concentrations (ng/m3) from
source-oriented and non-source-oriented monitors within Cuyahoga
County, OH (39-035), 2007-2009. 3-160
Table 3A-11. Correlations between Pb TSP concentrations from source-oriented and non-
source-oriented monitors within Cuyahoga County, OH (39-035), 2007-
2009 3-161
Figure 3A-16. Pb TSP Monitor locations within Sullivan County, TN (47-163), 2007-
2009. 3-163
Figure 3A-17. Wind roses for Sullivan County, TN, obtained from meteorological data
at Bristol/Tri City Airport, 1961-90. 3-164
Figure 3A-18. Box plots of annual and seasonal Pb TSP concentrations (jj.g/m3) from
source-oriented and non-source-oriented monitors within Sullivan County,
TN (47-163), 2007-2009. 3-165
Table 3A-12. Correlations between Pb TSP concentrations from source-oriented and non-
source-oriented monitors within Sullivan County, TN (47-163), 2007-2009 3-166
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3.8.2. Size Distribution of Pb-Bearing PM 3-166
Table 3A-13. Correlations and average of the concentration ratios for co-located
monitors, TSP versus PMjo, TSP versus PMz5, andPMjo versus PM25. 3-166
3.8.3. Lead Concentration in a Multipollutant Context 3-169
Figure 3A-19a. Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2007-2008. Top: winter; bottom: spring. 3-169
Figure 3A-19b. Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2007-2008. Top: summer; bottom: fall. 3-170
Figure 3A-20a. Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2009. Top: winter; bottom: spring. 3-171
Figure 3A-20b. Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2009. Top: summer; bottom: fall. 3-172
Figure 3A-21a. Seasonal correlations of monitored Pb-PM2 5 concentration with
copollutant concentrations, 2007-2009. Left: winter; right: spring. 3-173
Figure 3A-21b. Seasonal correlations of monitored Pb-PM2 5 concentration with
copollutant concentrations, 2007-2009. Left: summer; right: fall. 3-174
Table 3A-14. Copollutant exposures for various trace metal studies 3-175
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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
methodologies, and recent data describing Pb concentrations as a function of size range. 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. 2006).
Section 3.2 provides an overview of the primary and secondary sources of 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), and associations with copollutants 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. 2011). which is the most recently available quality-assured Pb emissions data. This section also
reviews updated information from the peer-reviewed literature regarding sources of ambient Pb. Detailed
information about processes for primary and secondary anthropogenic emissions and naturally-occurring
emissions can be found in the 2006 Pb AQCD ("U.S. EPA. 2006). The papers cited herein generally
utilized PM sampling data, because ambient airborne Pb readily condenses to PM. The 2006 Pb AQCD
(U.S. EPA. 2006) employed the 2002 NEI (U.S. EPA. 2008a) or source analysis and 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
Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and Environmental
Research Online) at http://eDa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of developing science
assessments such as the Integrated Science Assessments (ISA) and the Integrated Risk Information System (IRIS).
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from industrial processes, incineration, and utilities, each contributing less than 5% (53%). Since that
time, states have additional information to the inventory. The mobile source category included
combustion products from organic Pb antiknock additives used in piston-engine aircraft (hereafter
referred to piston aircraft emissions).
3.2.1. National Emissions Inventory
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. 2006). This action
resulted in a 98% reduction in Pb emissions from 1970-1990. Total Pb emissions over the period 1990-
2008 decreased an additional 77%, from 5,200 tons in 1990 to 1,200 tons in 2008. Subsequent emissions
reductions are related to enhanced control of the metals processing industry. In 1990, metals processing
accounted for 42% (2,200 tons) of total Pb emissions. By 2008, metals processing accounted for 12%
(150 tons) of total emissions. This represented more than an order of magnitude decrease in Pb emissions
from metals processing. At the same time, emissions from piston engine aircraft varied only slightly over
this time period. In 1990, off-highway Pb emissions were 990 tons and represented 19% of total Pb
emissions. In 2008, off-highway Pb emissions from piston engine aircraft were slightly lower at 590 tons,
which comprised 49% of all Pb emissions. "Miscellaneous" emissions from other industrial processes,
solvent utilization, agriculture, and construction comprised 20% of emissions (240 tons) in 2008.
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¦ Highway Vehicles
¦ Metals Processing
¦ Fuel Combustion
¦ Piston Engine Aircraft
¦ Miscellaneous
1970 1975 19&0 1985 1990 1995 1999 2002 2005 2008
Source: U.S. EPA (2011)
Figure 3-1. Trends in Pb emissions (thousand tons) from stationary and
mobile sources in the U.S., 1970-2008.
Highway Vehicles
Metals Processing
Fuel Combustion
Piston Engine Aircraft
Miscellaneous
2002
2005
2008
Source: U.S. EPA (2011)
Figure 3-2. Trends in Pb emissions (thousand tons) from stationary and
mobile sources in the U.S., 1990-2008.
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Direct emissions of Pb into the atmosphere primarily come from piston engine aircraft, fuel
combustion, and industrial activities. Direct Pb point source emissions estimated by the 2008 NEI are
shown in Figure 3-3. Emissions from piston engine aircraft emissions comprised nearly half of all
emissions (590 tons). Industrial fuel combustion contributed 220 tons (18%) of Pb emissions in 2008,
followed by metal working and mining (12%), other industry (9%), dusts from construction (7%) and
miscellaneous contributions from agriculture, solvent utilization, and operation of commercial marine
vessels and locomotives (4%) (U.S. EPA. 2011). 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.
1 590
1 224
0 ^'ning
1 149
'n«"stry
1 107
C°nStr»ctl0n
1 SS
1 49
100 200 300 400 500 600 700
Emissions (tons)
Source: U. S. EPA (2011)
Figure 3-3. Nationwide stationary source Pb emissions (tons/yr) in the
U.S. by source sector in 2008.
There is substantial variability in Pb emissions from stationary sources 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. The upper 0.1% of stationary emissions came
from thirty-three counties. This category included all counties emitting more than 3.8 tons of Pb in 2008.
Jefferson County, MO was the highest emitting county, with over 21 tons of airborne Pb emissions in
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1 2008. Jefferson County is home to the Doe Run primary Pb smelting facility, which is the only remaining
2 operational primary Pb smelter in the U.S.
2008 NEI Pb Emissions by County
Tons Pb Emitted Annually
I 0.00 - 0.02
| 0.03 - 0.04
|. | 0.05 - 0.08
~~I 0.09 - 1.60 f<3
I 1 H
Source: U.S. EPA (2011)
Figure 3-4. County-level Pb emissions (tons) from stationary sources in
the U.S. in 2008.
3 Pb emissions from piston engine aircraft operating on leaded fuel occur at approximately 20,000
4 airports across the U.S. Figure 3-5 displays Pb emissions (tons) at airports around the continental U.S.
5 The map illustrates that airport emissions tend to be elevated around highly populated metropolitan
6 regions, which typically have multiple airports with varying activity levels. Clusters of airports around
7 metropolitan areas and megapolitan regions (multiple contiguous metropolitan areas) are notable from
8 Figure 3-5. Among these sites, piston aircraft emissions at airports within twenty-eight counties
9 cumulatively emitted greater than one ton of Pb in 2008 U.S. EPA (2011). Additionally, within the 2008
10 NEI, there were estimates of Pb emissions during flight from piston aircraft. These estimates are provided
11 by state and cumulatively account for 296 tons in 2008.
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Counties
Airports
Tons of Lead Emitted
• 0.00-0.06
• 0.07-0.16
• 0.17-0.32
• 0.33-0.60
• 0.61-1.19
United States
.. ¦ • . m •ilV
\ * • w " • * f*. >.
° • °^r °6 ° = °°«go ¦ ~ kB~ m
»»
Source: U.S. EPA (2011)
Figure 3-5. Pb emissions (tons) at airports in the U.S. in 2008. The size of
the symbol indicates the magnitude of Pb emissions at each
airport.
3.2.2. Anthropogenic Sources
1 Anthropogenic Pb source categories are organized below in order of magnitude reported on the
2 2008 NEI (U.S. EPA. 2011). with emissions from piston engine aircraft being the highest and resuspended
3 dust from previously deposited Pb being substantially lower.
3.2.2.1. Lead Emissions from Piston Engine Aircraft Operating on Leaded
Aviation Gasoline and Other Non-Road Sources
4 The largest source of Pb in the NEI is from piston engine aircraft operating on leaded aviation
5 gasoline emissions (U.S. EPA. 2011). Murphy et al. (2007) cited fuel consumption estimates provided by
6 the U.S. Department of Energy indicating that 575,000 kg/yr of Pb is used in piston engine aircraft fuel,
7 and based on the 1999 NEI, that 490,000 kg/yr (85%) become airborne upon combustion. Levin et al.
8 (2008) point out that emissions from piston engine aircraft are exempt from reporting to the EPA Toxic
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Release Inventory. 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. The
Canadian report noted average and maximum air Pb levels were 0.030 and 0.30 (ig/m3, respectively,
compared with background levels of 0.007 and 0.018 (.ig/irf (Conor Pacific Environmental Technologies
Inc. 2000). The Illinois report noted that air Pb concentrations were elevated downwind of O'Hare airport
compared with upwind levels (Illinois Environmental Protection Agency. 2002). 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. Fuel consumption rates can be obtained for some engine/aircraft
combinations by running FAA's Emissions and Dispersion Modeling System (FA A. 2011V The ASTM
specification for the maximum Pb content is "100 Low Lead", the most commonly used leaded avgas, is
2.12 g of elemental Pb/gallon (ASTM. 2007).
Dynamometer testing has indicated that Pb emissions from piston engine aircraft fuel combustion
can occur in the particulate and gaseous form. For example, Gidney et al. (2010) performed dynamometer
testing on automobiles operating on standard gasoline and on gasoline with low levels of organometallic
additives. Tetraethyl Pb was included since it is still used in avgas. The additives had trace levels of Pb
that exist in gas phase when temperatures are higher than 650 C, below which they condense to
particulate phase. 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 to act as a Pb "scavenging agent." When
ethylene dibromide reacts with Pb, it forms Pb bromide and Pb oxybromides, which are more volatile.
3.2.2.2. Fugitive Emissions from Metals Processing and Mining
Fugitive emissions from secondary Pb processing can be substantial over the course of a year, but
they are difficult to estimate. 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"6to 4.4 x 10"5 g
Pb/m2 sec at the Florida site, 9.4 x 10"7to 2.0 x 10"6 g/m2sec for the Texas site, and 8.8 x 10"7to 1.1 x 10"
6 g/m2sec 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"3to 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.
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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 PMio_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
Fossil fuel combustion accounts for roughly 12% of Pb emissions in the U.S. 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. There were no emission factors for crude oil or residual oil but these represent
potentially large sources (up to 100-500 tons/yr and up to 25-700 tons/yr, respectively). 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.
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 (J. M. Pacyna & Pacvna. 2001; Witt et al.. 2006). The Pb isotope ratio values
(mainly 206Pb/207Pb) for coals 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.
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/yr worldwide. Using a separate global model, Niisoe et
al. (2010) calculated emissions of Pb from coal combustion in Japan during 2000 to be 900 tons/yr, based
on 9 x 107 tons/yr coal combustion and an emission factor of 10 g Pb/ton. The equivalent value for Pb
emissions from China was 56,000 tons Pb/yr. Calculated Pb concentrations in surface air for China agreed
with this value within a factor of two, although there was a systematic underestimation suggesting an
incomplete knowledge of the Pb emissions (Niisoe et al.. 2010). It was notable, however, that the
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calculated emissions from Chinese coal combustion make up a substantial proportion (-66%) of the total
global Pb emissions from fossil fuels detailed in Rauch and Pacyna (2009).
Tan et al. (2006) compared several emissions sources in Shanghai, China. They estimated emission
values for on-road exhaust from use of Pb-free gasoline (238 ± 5 mg/kg), vehicle exhaust from leaded on-
road gasoline (7,804 ± 160 mg/kg), coal combustion (1,788 ± 37 mg/kg), metallurgic dust (6,140 ±
130 mg/kg), soil (11.7 ± 0.3 mg/kg), and cement (103 ± 2 mg/kg). Pb-free automobile gasoline has been
in use in Shanghai since 1997. The isotope ratios for each of these emission sources were determined.
Based on the 4.4 x 107 tons of coal combusted annually in Shanghai, an average coal Pb concentration of
13.6 ± 6.6 mg/kg, and an emission factor of 0.5, approximately 300 tons Pb was being emitted annually in
association with fine PM. They concluded that a major priority should be to reduce Pb emissions from
coal combustion now that the contribution from vehicle exhaust emissions has decreased.
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 (W. Wang et al.. 2006). This seasonal relationship contrasts with
observations for the U.S. described in the 2006 Pb AQCD (U.S. EPA. 2006) 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. Other Industrial Sources
Several Pb isotope studies have been used to distinguish contributions from industrial activities.
For example, in northern China, Wang et al. (2006) noted that, in the response to decreasing atmospheric
Pb concentrations in total suspended particles (TSP) samples collected from 1994 to 1998, the 206Pb/207Pb
isotope ratio showed a related trend with values of-1.149 in 1994 increasing to -1.161 in 1998. The Pb
concentration and isotope ratio values then remained approximately constant from 1998 through to 2001.
Although this was consistent with a decreasing contribution of Pb from on-road gasoline, the ratio values
were still lower than those for Chinese coal [206Pb/207Pb -1.18 in Mukai et al. (2001)1. suggesting that
local Pb ore sources typically have lower 206Pb/207Pb ratios. The range for Chinese ores was 1.081 to
1.176, with lower values corresponding to ores from northern China (Mukai et al.. 2001).
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. Deconvolution of sources was attempted using Pb
isotopes, but one major area of uncertainty was the amount and the isotope composition of Pb emanating
from incineration plants, particularly in the U.K. The isotopic signature of Pb recycled into the
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atmosphere by incineration of various industrial wastes could have shifted from relatively high 206Pb/207Pb
ratios consistent with local Variscan ores to lower values reflecting imported Precambrian ores. There
have, however, been other environmental studies concerning incineration, and these give a highly
consistent value 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 -1.155. 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 1.1427-1.1576 and 2.4260-2.4346, respectively, i.e., quite a narrow range (de la Cruz et
al.. 2009 and references therein).
3.2.2.5. Roadway-Related Sources
Contemporaneous Emissions from Vehicle Parts
Contemporaneous 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 every automobile in the U.S.
in order to balance the tires, and may become loose and fall off. 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 fine PM by the pounding forces of traffic (Root. 2000). For
example, Schauer et al. (2006) measured Pb emissions in two traffic tunnels and found that the fraction of
Pb in PM2 5 was no more than 17% of Pb measured in PMi0. 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 low but substantial quantities of
Pb in PM10 (0.02 ± 0.01 mg/g) and PM2 5 (0.01 ± 0.00 mg/g) for semi-metallic brake pads and in PMi0
(0.01 ± 0.00 mg/g) for low-metallic brake pads. 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 are 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.
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Unleaded Fuel
Unleaded fuel contains Pb as an impurity within crude oil (E. G. Pacvna et al.. 2007). Schauer et al.
(2006) measured Pb in PM2 5 from tailpipe emissions and observed significant quantities in on-road
gasoline emissions (83.5 ± 12.80 mg/kg) and higher but non-significant quantities in diesel emissions
(137 ± 133 mg/kg). Hu et al. (2009) investigated the heavy metal content of diesel fuel and lubricating oil.
They found <1-3 ppm Pb in samples of lubricating oil. Hu et al. (2009) 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. The Tan et al. (2006) study cited in Section 3.2.2.3 for Shanghai, China, where unleaded fuel has
been in use since 1997, illustrated that Pb emissions from unleaded on-road gasoline were substantially
lower than Pb emissions from coal combustion. Assuming that the natural Pb content of unleaded fuels in
the U.S. is similar, it is unlikely that road vehicle combustion of unleaded on-road gasoline is currently a
major contributor to total Pb emissions in the U.S.
3.2.2.6. Deposited Lead
Soil Pb can serve as a reservoir for deposited Pb. The following subsections describe studies of
previously deposited Pb from industrial, historical leaded on-road emissions, and urban sources such as
paint and building materials. The 2006 Pb AQCD (U.S. EPA. 2006) 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. Since publication of the 2006 Pb AQCD (U.S. EPA. 2006). further analysis
of the Harris and Davidson (2005) paper has revealed that the contributions of Pb from piston engine
aircraft were underestimated compared with the 2002 NEI. Assumptions of spatial uniformity incurred by
the "continuously stirred reactor" mass balance model and for mixing layer height used by Harris and
Davidson (2005) were also not valid because Pb concentrations are spatially heterogeneous at the urban
scale; see Section 3.5. Therefore, the estimate of 90% of airborne Pb from resuspension is not employed
in the current assessment. Currently, data are not available with sufficient spatial resolution to discern the
specific contribution of soil Pb resuspension to air Pb concentration, but resuspended soil Pb cannot be
eliminated as a potential source of airborne Pb.
Lead from Industrial Sites
Several studies have indicated elevated levels of Pb in soil exposed to industrial emissions,
including brownfield sites (Deng & Jennings. 2006; Dermont et al.. 2010; Hofcr et al.. 2010; Jennings &
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Ma. 2007; Sriskandan et al.. 2007; van Herwiinen et al.. 2007; Verstraete & Van Meirvenne. 2008). It is
possible that Pb in soil serves as a source of airborne Pb. Laidlaw and Filipelli (2008) reviewed the
literature on Pb resuspension from soil and then analyzed IMPROVE data to explore conditions under
which Pb may become resuspended. They observed a seasonal pattern in concentration of soil
resuspended in the atmosphere, and they also found that 83% of the variability in concentrations of soil in
the atmosphere was predicted by the 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.
Lead from Paint and Building Materials
Exterior painted structures have long been known to be a source of ambient Pb ("U.S. EPA. 2006).
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 under the steel
structures being between 2.5 to 11 times higher than median Pb concentrations not near a structure.
The studies described above considered paint as a source of Pb dust through gradual abrasion of the
painted surfaces. However, ambient 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 elevated levels
of 03 (11.3 ± 0.8 ppm or 150 times the level of the 8-h NAAQS) and N02 (11.6 ± 0.9 ppm, 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). Edwards et al. (2009) state that the high 03 and N02 concentrations simulated in the
chamber were equivalent of 4.3 and 3.7 years of exposure at 50 and 60 ppb, respectively.
Building demolition was listed as a source of urban Pb dust in the 2006 Pb AQCD (U.S. EPA.
2006). In a follow-up study to previous work cited therein, Farfel et al. (2005) observed that Pb dust
surface loadings 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 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.
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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 still may be persistent in the environment as a potential
source of airborne Pb. The historical combustion of leaded on-road gasoline has been estimated from
documents submitted by Ethyl Corporation to the U.S. Senate ("Airborne Lead Reduction Act of 1984."
1984) and a report by the U.S. Geological Survey (USGS. 2005); see Mielke et al. (2010b). These
estimates are presented in Figure 3-6. 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 non-
highway applications, including piston engine aircraft fuel, racing fuels, farm tractors, snowmobiles, and
boats.
Lead Additives in U.S. Gasoline
250,000
200,000
= 150,000
O
I-
u
100,000
50,000
Year
Source: Used with permission from Pergamon Press, Mielke et al. (2010b).
Figure 3-6. Total U.S. Pb additives in on-road gasoline used in on-road
vehicles, 1927-1995. Estimates were derived from the
proceedings of the U.S. Senate hearings on the Airborne Pb
Reduction Act of 1984, S. 2609 ("Airborne Lead Reduction Act
of 1984," 1984) and the U.S. Geological Survey Pb end use
statistics (USGS. 2005).
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The particle sizes of Pb emissions from on-road sources were estimated by the U.S. EPA (1986).
which indicated that 75% of Pb additives were emitted as exhaust. The tonnages of relatively large
>10 |am mass median aerodynamic diameter (MMAD) Pb-PM probably settled locally, especially in high
traffic urbanized areas where soil Pb, from historic emissions as well as contemporaneous sources, are
elevated adjacent to roadways and decrease with distance away from the roadway (Laidlaw & Filippelli.
2008). EPA (1986) indicated that 35% of the PM were < 0.25 |im in MMAD. The majority of PM, by
number, emitted from automobiles was in the ultrafine size range (Londahl et al.. 2009). However,
Londahl et al. (2009) did not include ethylene dibromide to the fuel in these experiments. As described in
the Gidney et al. (2010) study referenced in Section 3.2.2.1, it was found that ethylene dibromide acted to
scavenge Pb in PM to form more volatile Pb bromide and Pb oxybromide to produce gaseous Pb
emissions.
The use of Pb additives also resulted in a national scale of influence. For example, various sized
urbanized areas of the U.S. have different amounts of vehicle traffic associated with Pb (Mielke et al..
2010a). Figure 3-7 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 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 cities illustrated in Figure 3-7, depending on the amount of highway
construction in those cities. Hence, the map may overestimate potential amounts of Pb in soil, and
consequently of airborne Pb from resuspended soil, in some of the cities illustrated.
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86
81
15W 71 - « Ao
« , -if
-• • • a ;;; 45
6#E • «. M *\* •» J**5
% V i 2 s «*
83
%
•57
m*v
Vis
•*
¦a
#31
• " *34
®K •
• 3'
«•
,«i
•«
169
#60
56
%
039
ftj°i 38 V
U.S. Urbanized Areas #52 # A
Pb 1950-1982 (metric tons) #28 r® •
652 - 4,570
33l
79*
# 4,662 - 8,108 • *77 87 •
# 8,403 -16,614
89
# 16,623 - 91,878 •
•10
149,938
Source: Used with permission from Pergamon Press, Mielke et al. (201 Ob)
Figure 3-7. Estimated Pb aerosol inputs from on-road gasoline into 90
U.S. urbanized areas (UAs), from 1950 through 1982.The
numbers on the map are rankings of each UA. The size of each
dot refers to the magnitude of motor vehicle gasoline-related
emissions for each group of UAs. The extremes are, Los
Angeles UA (ranked #1) and Laredo, Texas (ranked #90). Some
of the UAs have been used as sites in soil Pb studies, as
indicated in Table 3-9.
3.2.3. Source Attribution
3.2.3.1. Lead Speciation and Source Apportionment
1 The following section describes new findings with respect to speciation of Pb content in aerosols.
2 Analytic techniques for speciation are explicated in Section 3.4. Forms of Pb commonly observed in the
3 environment are presented in Table 3-1 to serve as a reference for the categories of Pb sources described
4 in Sections 3.2.1 and 3.2.2. Detailed descriptions of the related chemistry were presented in the 2006 Pb
5 AQCD (U.S. EPA. 2006).
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Table 3-1. Pb compounds observed in the environment
Emission Source
Observed Pb Compounds
Minerals
PbS (Galena)
PbO (Litharge, Massicot)
Pb304 (Minimum or "Red Pb")
PbS04 (Anglesite)
Smelting Aerosols
Pb", PbS
PbS04, PbO
PbC03
Pb silicates
Coal Combustion Aerosols
PbS
PbSe
Coal Combustion Flue Gases
Pb", PbO, Pb02 (Above 1150 K)
PbCI2 (Low rank coals, above 1150 K)
PbS04 (Below 1150 K)
Wood Combustion
PbC03
Waste Incineration Aerosols
PbCI2, PbO
Soils Near Mining Operations
PbC03
PbS04
[PbFe6(S04)4(0H)12l
[Pb5(P04)3Cll
[Pb4S04(C03)2(0H)3l
PbS-Bi2S3
Pb oxides, silicates
Motor Vehicle Exhaust (Combustion of Leaded Fuel)1
PbBrCl
PbBrCI-2NH4CI
PbBrCl-NH4CI
Roadside Dust
PbS04, Pb"' PbS04(NH4)S04, Pb304, PbO-PbS04 and 2PbC03-Pb(0H)2
Other mobile sources:
Brake wear, wheel weights
Pb"
Racing vehicle emissions
Pb halides
Aircraft Engine Wear
Pb"
Lawn Mowers
Pb halides (battery leakage)
Source: Biggins and Harrison (1979. 1980): U.S. EPA (2006).
Chemical speciation of Pb by source was reviewed in the 2006 Pb AQCD (U.S. EPA. 2006). and is
fairly well understood and varies considerably between sources. Pb components from Pb smelters are
mainly elemental Pb (Pb°), Pb sulfide (PbS), Pb sulfates (PbS04, PbO-PbS04), and Pb oxide (PbO), with
Pb carbonate (PbC03) and Pb silicates. Other smelters also produce Pb, mainly as Pb oxide (PbO). Pb in
coal combustion emissions is mainly in the form of Pb sulfide (PbS) and Pb sulfate (PbS04) with some Pb
selenide (PbSe) as well as sulfates, oxides and chlorides. In wood combustion emissions PbC03 and Pb
oxides are important. Waste incineration produces mainly Pb chloride (PbCl2) and PbO. Resuspended
mining soils were reported to be abundant in PbC03 and PbS04, as well as a variety of complex salts
containing phosphates, hydroxides, chlorides, oxides, and silicates. On-road engine exhaust from leaded
gasoline use contained mostly Pb bromide and Pb chloride salts, including PbBrCl, PbBrCl-NH4Cl. and
PbBrCl-2NH4Cl. Road dust is rich in PbS04, Pb°, oxides, carbonates, and hydroxides. Motor vehicles also
contribute Pb° from brake wear and wheel weights.
The 2006 Pb AQCD (U.S. EPA. 2006) described the atmosphere as the major environmental
transport pathway for Pb, with Pb primarily present in submicron aerosols. Although not directly
addressed in the 2006 Pb AQCD (U.S. EPA. 2006). oiganolead vapor emissions were extensively
discussed in the 1986 Pb AQCD (U.S. EPA. 1986) which concluded that they were primarily emitted
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during manufacture, transport, and handling of leaded on-road gasoline. Organolead vapors contributed
less than 10% of vehicular Pb tailpipe emissions when leaded on-road gasoline was still in use. Studies of
Pb emissions within enclosed microenvironments where automobiles were the dominant Pb source cited
within the 1986 Pb AQCD (U.S. EPA. 1986). reported that organic Pb vapors contributed less than 20%
of total vehicular Pb emissions. More recent studies support this (Shotvk et al.. 2002). The 20% estimate
of organic Pb vapors from the previous studies of on-road Pb emissions may potentially provide an upper
bound for organic Pb vapors from current piston engine aircraft emissions.
Several recent studies have used speciation techniques either simply to determine the chemical
composition of emissions or for source attribution. In urban environments, airborne Pb concentrations are
likely a mix of various sources. Ogulei et al. (2006) performed source apportionment of PM25 samples in
Baltimore, MD and found multiple industrial, urban, and background influences. Sixty-three percent of
the Pb was associated with incineration, while 20% was associated with a wildfire episode from which
PM was transported from Quebec, 8% was associated with secondary nitrate, 6% was associated with
operations at a steel plant, and 3% was associated with local gasoline traffic. Dillner et al. (2006)
analyzed the composition of PM25 and TSP samples in Beijing, China and found that Pb comprised
roughly 0.2% of TSP and 1.4% of PM25 during industrial pollution events, 0.1% of TSP and 0.5% of
PM25 during urban pollution events, and 0.05% of TSP and 0.4% of PM25 during dust storms. During
industrial pollution events, the authors note that the amount of Pb in PM2 5 can be substantial.
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). Pb speciation
emissions from a sintering plant, a major component of the steel making process, were reported for the
first time, with cerussite, a Pb carbonate (PbC03-2H20), emerging as the most abundant species (Sammut
et al.. 2010). The predominance of carbonates as major Pb species in industrial emissions is unusual.
Choel et al. (2006) confirmed that Pb was strongly associated with sulfur in smelter emission PM, and
that Pb sulfates and Pb oxy-sulfates were the most abundant species, with important contributions from
Pb oxides. Zhang et al. (2009) used single particle aerosol mass spectrometry (ATOFMS) to speciate Pb-
bearing PM in Shanghai, China in 2007. PM containing Pb along with OC and/or EC was attributed to
coal combustion processes; this accounted for roughly 45% of Pb-bearing PM. PM producing high
correlations between CI and Pb were ascribed to waste incineration, while Pb-bearing PM with a strong
phosphate signal was attributed to the phosphate industry.
A few recent studies have used speciation techniques to characterize Pb and other components of
PMio, PM2 5, and PMi. 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.
Approximately two-thirds of PMi consisted of secondary aerosols, e.g. mainly sulfate, nitrate and
primary/secondary organics. The remaining third included PM from biomass burning, fossil fuel
combustion and various industrial sources. For the latter group, strong Pb-Zn-K-Na associations were
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observed. 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. Wojas and Almquist (2007)
used ICPMS to characterize trace metals in PM2 5, PMi0, and TSP samples obtained in Oxford, OH. They
observed that Pb was highly correlated with several other elements (Ca, Co, Cu, Fe, K, Mg, Mn, Mo, Ni,
Pb, Sb, Si, and Zn), suggesting that Pb and copollutants emanated from a variety of sources including
road dust and fuel combustion. Similarly, a study by Moffet et al. (2008) found that Pb-Zn-Cl 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. The unique single particle chemical associations closely matched
signatures indicative of waste incineration. A study of PMi0 and PM2 5 collected in Shanghai, China and
analyzed using extended X-ray absorption fine structure spectroscopy (EXAFS), found that the main
chemical forms of Pb were PbCl2 (41 ± 4%), PbS04 (37 ± 2%) and PbO (22 ± 3%) (Tan et al.. 2006).
There was no significant difference in the chemical forms of Pb between the PMi0 and PM2 5 samples.
The main sources of these forms of Pb, based on Pb isotopic composition, were coal combustion,
metallurgic dust and vehicle exhaust emissions (none though from leaded on-road gasoline).
Approximately 83% Pb was in the <2.5 |_im size range. Murphy et al. (2007) found that the volatility of
Pb and its compounds such as PbO results in its presence at high concentration in the submicron fraction
of PM emitted from coal emissions. PbS04, also derived from coal combustion, has low solubility
(Barrett et al.. 2010). Variations in the relative proportions of Pb-containing compounds may account for
the difference in Pb solubility in aerosols (Fernandez Espinosa & Ternero-Rodriguez. 2004; Tan et al..
2006; von Schneidemesser et al.. 2010).
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. For urban aerosols collected in
the U.S., two types of Pb-PM were found; in the main class, Pb was found together with K and usually
also Zn. There were also minor amounts of Na, EC and organic carbon (OC) including amines. The
second, minor class contained Pb together with Na, K, Zn, smaller amounts of Fe, EC and OC. The size
distribution of the first group of Pb-PM usually peaked around 200 nm.
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Sample solubility can inform speciation efforts. For example, a study involving weak acid leaching
was carried out by Erel et al. (2006). The transport of anthropogenic pollution by desert dust in the eastern
Mediterranean region was studied by determining major and trace element concentrations, organic
pollutants, and Pb isotope ratios. PMi0 samples were collected during 10 dust storms in 2001-2003. Most
samples were polluted to some extent with pollutants released by weak acid (0.5 M HN03) extraction
(including carbonates, oxides and surface-bound fractions). From the Pb isotope data, most of the Pb
came from recently north African emissions and from past Israeli emissions (Pb now residing in Israeli
soils).
3.2.3.2. 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 & Hu. 2010). Isotope ratio analysis was first applied to airborne PM in 1965 to identify the
impact of motor vehicle exhaust on marine and terrestrial Pb deposition in the Los Angeles area (Chow &
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 & 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 (Hsu et al.. 2006; Knowlton & Moran. 2010; Noble et al.. 2008; Widorv. 2006).
Shotyk and Krachler (2010) also used Pb isotopes to demonstrate that the fate of Pb from runoff can be
different from Pb with different origins. They observed that humus PM impacted by leaded on-road
gasoline that are derived from soil surfaces are likely to be more easily transferred to sediments than Pb of
other origins, with substantial amounts retained by lakes.
Recent studies have examined the use of Pb isotope ratios as a tool for source apportionment.
Duzgoren-Aydin and Weiss (2008) provide caveats for using isotope ratio analyses. They point out that Pb
isotope ratios may vary when Pb from several sources of different geological origins are introduced to the
same location. Duzgoren-Aydin (2007) warned that the presence of a complex mixture of contaminants
containing common Pb isotopes can lead to an overestimation of the contribution of one source (e.g., soil
contaminated by Pb emissions from on-road gasoline) and an underestimate of another source, such as
that from industry. For this reason, Cheng and Hu (2010) suggest that Pb isotope analysis only be used
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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 208Pb/206Pb 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 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.
3.3. 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 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 the contribution of atmospheric Pb to total Pb exposure. Figure 3-8
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
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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
Figure 3-8. Fate of atmospheric lead. Media through which Pb is
transported and deposited are shown in bold.
3.3.1. Air
The 2006 Pb AQCD (U.S. EPA. 2006) concluded that Pb was primarily present in submicron
aerosols, but that bimodal size distributions were frequently observed. Pb-PM in the fine fraction is
transported long distances, found in remote areas, and can be modeled using Gaussian plume models and
Lagrangian or Eulerian continental transport models as reported by several studies. Good agreement
between measurements and these models have been reported. Historical records of atmospheric
deposition to soil, sediments, peat, plants, snowpacks, and ice cores have provided valuable information
on trends and characteristics of atmospheric Pb transport. Numerous studies using a variety of
environmental media indicated a consistent pattern of Pb deposition peaking in the 1970s, followed by a
more recent decline. These findings indicated that the elimination of leaded gasoline for motor vehicles
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. 2006) 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
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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/m2yr. 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 United States ranged from 0.3 to 1.0 mg/m2yr. 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/m2yr 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. Pb in resuspended road dust exhibited a bimodal size distribution, but mass
was predominantly associated with coarse PM. The Pb fraction in resuspended dust ranged from 0.002 to
0.3%, with the highest fractions observed for paved road dust and lowest for agricultural soil.
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. 2006). much of the recent
research on Pb transport has focused on interactions between anthropogenic and coarser geogenic PM that
leads to incorporation of Pb into coarse PM as well as subsequent transformation on exposure to mineral
components of coarse PM. Using scanning electron microscopy (SEM), Schleicher et al. (2010) observed
interactions of anthropogenic soot and fly ash particles on the surfaces of coarse geogenic mineral
particles 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
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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)2H20 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.
Transport and Dispersion Mechanisms in Urban Environments
Several major U.S. sources of Pb emissions are located in urban areas. The urban environment can
be considered unique 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. This
section focuses on special features of urban environments and upon processes that influence the
distribution and redistribution of Pb-bearing PM.
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ABL
Urban can*
length L
Street
Canyon
Recirculation
ABL -
UBL -
RSL -
UCl -
USL -
ML -
CFL -
+/*) =
= Uh?MH-V
Rural
Atmospheric boundary layer
Urban boundary layer
Roughness sublayer
(transition layer, wake layer, interface I layer)
Urban canopy layer
Urban surface layer
Mixed layer
Constant flux layer
(i.e., inertial sublayer - ISL)
I *ua
Street
canyon
WIND
Source: Used with permission from Annual Reviews, Fernando (2010)
Figure 3-9. Scales of turbulence within an urban environment.Top:
multiple scales within the atmosphenc boundary layer.
Bottom: illustration of airflow recirculation within a single
street canyon located in the urban canopy layer.
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As shown in Figure 3-9, 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 (see Figure 3-9), which result in
lower mean components of air movement, higher turbulence components, and higher sheer stress within
the canyon compared with open field conditions (Britter & Hanna. 2003; Kastner-Klein & Rotach. 2004).
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 (Kastner-Klein et al.. 2003;
Kastner-Klein et al.. 2001; Soulhac et al.. 2009). All of these factors have the potential to influence human
exposure to atmospheric Pb in urban areas with substantial Pb emissions.
3.3.1.2. Deposition
Wet Deposition
The 2006 Pb AQCD (U.S. EPA. 2006) 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. Recent research confirms the general trend described in the 2006 Pb
AQCD (U.S. EPA. 2006) that Pb associated with fine PM is usually more soluble in rain water than Pb
associated with coarse PM, leading to a relatively greater importance of wet deposition for fine Pb and of
dry deposition for coarse Pb. 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.
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Dry Deposition
New measurements of Pb dry deposition fluxes are similar to those reported in the 2006 Pb AQCD
(U.S. EPA. 2006), except in industrialized urban areas, where it is considerably greater. 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 an urban industrial site than for New Brunswick. 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. Elements from anthropogenic
sources, including Pb, were generally associated with fine PM. In a study of Tokyo Bay (Sakata &
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 AQCD (U.S. EPA. 2006). They also reported that dry
deposition fluxes were greater in industrially impacted urban areas, ranging from 12-17 (ig/m2yr, more
than 10 times the upper bound of the range reported in the 2006 Pb AQCD (U.S. EPA. 2006).
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. 2006). 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/m2yr; this was lower than the value of 1.30-1.90 mg/m2yr 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.
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3.3.1.3. Resuspension of Lead from Soil to Air after Lead Deposition
As described in Section 3.2, the greatest Pb emissions in the United States occur in locations near
major specific point sources, including airports, secondary smelters, and other industrial operations
involving large scale metal processing or fuel combustion. However, in the absence of such sources and in
the vicinity of previous major sources, the 2006 Pb AQCD ("U.S. EPA. 2006) concluded that resuspension
by wind and traffic can be a substantial source of airborne Pb above background levels near sources, with
resuspended dust accounting for between 0.002 to 0.3% of PM mass. Since then, results from several
studies have provided support for a substantial contribution from resuspension by indicating a smoothed
soil Pb concentration profile that decreases with distance from various sources, including city centers
(Laid law & Filippelli. 2008). major freeways (Sabin. Lim. Venezia. et al.. 2006). and steel structures with
abrading paint (Weiss et al.. 2006). The smoothed profile is consistent with continual Pb resuspension and
deposition due to atmospheric turbulence. Recent Pb speciation results also indicate a substantial
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 (Piimitorc et al.. 2009).
Recent research on urban PM transport is also highly relevant to Pb transport and dispersion
because Pb is most prevalently particle-bound. Relevant results for Pb exposure in these areas include
observations that PM concentration peaks dissipate more rapidly on wider streets than in narrow street
canyons (Buonanno et al.. 2011); concentrations are typically low next to a building because either less
source material is available or less material penetrates the boundary layer of the building (Buonanno et
al.. 2011); and there are stronger inverse 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 was applied to a
road and then the grit's dispersion by traffic was measured over time to simulate resuspension and
transport of a trace metal such as Pb. 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.
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-Cazenave 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
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factors determining resuspension. Together these results demonstrate that under in the vicinity of previous
major emission sources and the absence of current major sources, resuspension can make a substantial
contribution to atmospheric Pb concentrations.
3.3.2. Water
As described in the 2006 Pb AQCD ("U.S. EPA. 2006), 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 by flushing, evaporation and sedimentation. 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. 2006), 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 is 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. 2006), runoff from storms was also identified as an important source. A
substantial portion of Pb susceptible to runoff is originates from atmospheric deposition. The 2006 Pb
AQCD (U.S. EPA. 2006) concluded that important contributors to Pb in dust on roadways included
vehicle wear, vehicle emissions, road wear, fluid leakage, and 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-
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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 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
& von der Kammer. 2008). which is consistent with previous research using cross flow filtration
(Pokrovskv & Schott. 2002; Ross & Sherrell. 1999) and SEM examination of single particles (Taillcfcrt et
al.. 2000).
Two distinct colloidal phases, one organic-rich (0.5-3 nm in diameter) and the other Fe-rich (>3 nm
in diameter), have been observed to coexist in both soil isolates and river water (Stolpe & 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 & 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 & von
der Kammer. 2008).
Observations in boreal rivers and soil pore waters in permafrost dominated areas of Central Siberia
indicated that Pb was transported with colloids in Fe-rich waters. Trace elements that normally exhibited
limited mobility (including Pb) had 40-80% of their annual flux in the nominal dissolved phase,
operationally defined as material that passes through a 0.45 |_im pore-size filter, and that these metals had
a higher affinity for organo-mineral Fe-Al colloids (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 (.un pore-size filter (Pokrovskv et al.. 2010). This was likely due to the presence of organic-bound
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colloids smaller than 0.45 |_im rather than true Pb dissolution (Pokrovskv et al.. 2010V Pb adsorbed on
colloidal surfaces rather than incorporated into particle structure is likely to be more readily dissolved
because dissolution of the entire particle is not required.
Recent research on retention of Pb in water bodies and sediments has focused on the estuarine and
marine environment, where considerable retention of Pb was observed in estuarine sediments. For a large
riparian system, the Trinity River, Texas, Warnken and Santschi (2009) found that 80% of riverine Pb was
retained in Lake Livingston, an estuarine region, while an additional 16% was removed to estuarine
sediments, and only about 4% eventually reached the ocean. Geochemical (sorption by Fe
oxyhydroxides), biological (seasonal uptake by sinking algae in Lake Livingston) and hydrological
(dilution effects by increasing flow rates) processes were mainly responsible for controlling dissolved
trace metal concentrations rather than pollution sources.
Overall, recent research on Pb transport in aquatic systems has provided a large body of
observations confirming that Pb transport is dominated by iron and organic rich colloids. In addition, new
results indicated that although the 2006 Pb AQCD (U.S. EPA. 2006) 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 & 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 (Chalmers et al.. 2007; Horow itz et al.. 2008).
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 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 (Roth well et al..
2008). At the other end of the pH scale, Pb tends to remain or become complexed, precipitated or sorbed
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to TSP, 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 (.un 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. (2010) reported that PM Pb accounted for
85 ± 15% and 50 ± 22% in Boston Harbor and Massachusetts Bay, respectively, while Lai et al. (2008)
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 & Berry. 2010). Several new advances in the study of Pb cycling in these
complex environments have been described in recent publications. Li et al. (2010) 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 PM-Pb 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 (L. Li et al.. 2010). Surface sediment Pb
concentrations for various continental shelves were collated and compared by Fang et al. (2009); see
Table 3-2.
Table 3-2. Surface sediment Pb concentrations for various continental shelves; see
Fang et al. (2009) and references therein.
Location
Digestion solution
Pb (mg/kg)
East China Sea
HCI/HNO3/HF
10-49 (27)"
Mediterranean, Israel coast
HNO3
9.9-20
Aegean Sea
HCI/HNO3/HF
21-44(34)
Banc d'Arguin, Mauritania
HCI/HNO3/HF
2.8-8.9
Campeche shelf, Gulf of Mexico
HCI/HNO3
0.22-20 (4.3)
Laptev Sea, Siberia
HCI/HNO3/HF
12-22
Pechora Sea, Russia
Not reported
9.0-22 (14)
aValues in parentheses are the average, where calculable
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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. (J. L. 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 (J. L. 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 (G. Lee et al.. 2008). 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 sub-oxic and anoxic zones of the sediment act as a Pb source to the overlying water of the lake
(Sennor et al.. 2007).
Sediment resuspension from marine environments is similarly important, with disturbance of bed
sediments by tidal action in estuarine areas resulting in a general 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 (L. Li et al.. 2010). 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).
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
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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. 2006). 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 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 surfaces waters, 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 deposited from the atmosphere. The 2006 Pb AQCD (U.S. EPA. 2006) identified
important contributors to Pb pollution in dust associated with roadways, such as vehicle wear, vehicle
emissions, road wear, fluid leakage, and atmospheric deposition That review identified contributors to
runoff from buildings, such as paint, gutters, roofing materials and other housing materials. The 2006 Pb
AQCD (U.S. EPA. 2006) also 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 roadway and urban runoff and snow melt.
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 & 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
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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 (Al-Malack. 2001; Lasheen et
al.. 2008; Weiss et al. 2006).
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 & McCreadv. 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 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 & 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 whilst 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 using , 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. 2006). 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. (J. 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
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pattern. Pb was associated with Cincinnati PM mainly in the <75 (.un 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.
Guo et al. (2006) 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. (2006) investigated a catchment that drained towards 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. (2006) 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 j^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
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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. 2006). The relocation of snow means that the area 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 (J. J. Sansalone & Buchberger. 1996). The
melt process usually occurs in a sequence: pavement melt, followed by roadside (impervious) and finally
pervious area melt. As part of this sequence, rain-on-snow can transport high loads of PM-associated
pollutants (Oberts. 2000). Westerlund and Viklander (2006) investigated differences in PM and Pb
concentrations between rainfall events occurring during snowmelt and rain periods. Runoff events
occurring during the snowmelt period (i.e. rain-on-snow) had about five times higher numbers of particles
(in the size range 4 to 120 |_im)/liter of runoff. The first rain-on-snow event was characterized by an
increase in the number of particles in the 4 to 25 (.un 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 & 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 & 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.
2006). including several recent publications on a forested watershed (Lake Plastic) in central Ontario
(Landre et al.. 2009. 2010; Watmough & Dillon. 2007) and nearby Kawagama Lake, Canada (Shotvk &
Krachler. 2010). Results indicated that bulk deposition substantially decreased to 0.49 mg/m2 in 2002
from 1.30-1.90 mg/m2 in 1989-91. The upland soils retained >95% of the Pb in bulk deposition, i.e.
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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 (.un size range, but after several hours Pb was mainly transported with
colloidal or DOM (<0.45 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 ofWatmough 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)).
In another study Rothwell et al. (2007) observed stormflow Pb concentrations almost three times
higher than those reported by Graham et al. (2006) for northeastern Scotland. The generally high
dissolved Pb stores and high stream water DOC concentrations (Rothwell. Evans. Daniels, et al.. 2007).
In a separate study, Rothwell et al. (2007) 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.
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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.
Several investigators considered a Pb isotope study by Dunlap et al. (2008) of a large
(>160,000 km2) riparian system (the Sacramento River, CA), which 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).
Rothwell et al. (2007) commented that although there have been substantial reductions in sulfur
deposition to U.K. uplands over the last few decades (Fow ler et al.. 2005). anthropogenic acidification of
upland waters is likely to continue due to nitrogen leaching from the surrounding catchment and this may
increase with nitrogen saturation (Curtis et al.. 2005). Rothwell et al. (2007) 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-
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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 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).
In heavily contaminated catchments (e.g., that of the Litavka River, Czech Republic (Zak et al..
2009)). 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-
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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, and that Pb is frequently associated with chromate near
roads, probably from paint used to mark road lines. 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 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. 2006).
3.3.3. Soil
The 2006 Pb AQCD ("U.S. EPA. 2006) 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,
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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 was associated with Mn and Fe oxides or carbonate.
3.3.3.1. Deposition of Lead onto Soil from Air
As described in the 2006 Pb AQCD (U.S. EPA. 2006). 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 can reach several thousand mg/kg. Major sources in urban soil were identified as
automotive traffic (prior to when leaded on-road gasoline was phased out), and deteriorating Pb-based
paint, with the highest Pb concentrations observed where traffic and population density were the greatest.
Highest concentrations were found in city centers and near roadways, and several studies reported
concentrations falling off rapidly with distance and depth of soil layers near roads. High Pb soil
concentrations were also observed near stationary sources such as smelters and battery disposal
operations, also decreasing 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/m2yr to about 3 mg/m2yr with fair agreement between locations in Canada,
Scandinavia, and Scotland and showed a substantial decrease compared to when leaded on-road gasoline
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was in widespread use (Fow ler et al.. 2006; Graham et al. 2006; Shotbolt et al.. 2008; Watmough &
Dillon. 2007).
Differences between throughfall and litterfall in forested areas have also 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).
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. Higher Pb concentrations at
greater altitudes were also found in Japan, especially above 600 m (Takamatsu et al.. 2010).
There is wide agreement that atmospheric deposition due to long-range transport from industrial
areas has been the major source of Pb to remote surface soils over the past decades, e.g. (Steinnes et al..
2005). However, another hypothesis proposes that the gradual increase in Pb content in O-horizon soil
with changing latitude is attributable to "plant pumping and organic binding" rather than to atmospheric
deposition (Rasmussen. 1997; Reimnnn et al.. 2008; Reimnnn et al.. 2001).
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, and that productive plants were able to accumulate more
particles from the air and that rates of net Pb accumulation by the mosses were in excellent agreement
with the fluxes obtained by direct atmospheric measurements at nearby monitoring stations. In addition,
Bindler et al. (2008) used Pb isotopes to compare the distribution of Pb in the forest soils with that of lake
sediments where no "plant pumping" processes could be invoked, and used Pb isotope ratios to
demonstrate that observations were consistent with anthropogenic Pb deposition to the soils rather than
intermixing of natural Pb from underlying mineral soil horizons.
Overall, recent studies provided deposition data that was consistent with deposition fluxes reported
in the 2006 Pb AQCD (U.S. EPA. 2006). and demonstrated consistently that Pb deposition to soils has
decreased since the phase-out of leaded on-road gasoline. Additional research highlighted the importance
of taking forest cycling and litter throughput account in estimating input by deposition. 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. Finally, although there has been
considerable discussion of plant pumping as an alternative hypothesis for explaining the increases in soil
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concentrations, much evidence is more consistent with atmospheric deposition as the explanation for
observed increases in soil Pb concentration.
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. 2006). 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).
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
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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. 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 O soil horizons in 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
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sequestration of Pb via formation of oxycarbonate minerals is only effective at near-neutral to alkaline pH
values (Figure 3-10 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 immobilize Pb in situ through the formation of
Pb phosphate minerals such as chloropyromorphite. Recent research investigated factors affecting the
long-term stability of such products, which depends on the equilibrium solubility and the dissolution rate
of the mineral, trace impurities, such as Pb(OH)2, the presence of complexing agents, and pH (Xie &
Giammar. 2007). Overall, in agreement with the 2006 Pb AQCD (U.S. EPA. 2006). 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. 2006). 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 & 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
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forested areas of the Northeast U.S. that had been re-sampled after a 21-year time period had elapsed, but
that measured soil Pb concentrations were lower than predicted concentrations from total deposition,
strongly suggesting that the O-horizon had not retained all of the atmospheric Pb, and that a proportion of
the atmospheric deposition must have leached into the underlying mineral layers. At some sites, mainly
those at the southern latitudes and lower altitude sites, the proportion of Pb that had been leached
downward from the O-horizon was quite considerable. Relative retention of Pb was influenced by the rate
of OM decomposition, depth of soil O-horizon, and pH. For soils where Pb was strongly retained by the
O-horizon, a relationship between Pb and Fe-rich phase was observed, but Pb was also significantly
correlated with other metals. XANES data suggested a possible interaction with an amorphous Fe oxide,
but spectra were not entirely explained by Fe and oxygen and an additional spectral feature suggested the
presence of a 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. (2006) 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. (2006) 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.
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Klaminder et al. (2008) considered the biogeochemical behavior of atmospherically derived Pb in
boreal forest soils in Sweden (Figure 3-10). The estimated annual losses via percolating soil water were
-2 mg/'m yr (Klaminder. Bindler. & Renberg. 2008) and so the annual loss, assumed to be from the mor
layer, was greater than the atmospheric input of -0 .5 mg/m2yr. 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. Bindler. Emtervd. et al.. 2006). These values were greater
than the values of 2-150 years determined for U.S. forest soils, e.g. (J. Kaste et al.. 2003; Watmough et
al.. 2004) 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.
0.5
0.05
. V7 /r —
0.05
Bs
B/C
* 0 02
Loss of lead
Buildup of
lead
¦—> Present atmospheric deposition (mg rrr2 yr1)
—* Plant uptake (mg rrr2 yr1)
* Soil water flux (mg rrr2 yr1)
Figure 3-10. Schematic model summarizing the estimated flux of Pb within
a typical podzol profile from northern Sweden using data from
Klaminder et al. (2006).The atmospheric deposition rate is
from (Klaminder. Bindler. Emtervd. et al.. 2006). the plant
uptake rates from (Klaminder et al., 2005) and estimated soil-
water fluxes from (Klaminder, Bindler. Laudon, et al.. 2006).
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Klaminder et al. (2008) 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.. 2007V 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 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
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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.
(2008V
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). 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 & Krueger. 2009). In an agricultural soil, fate of Pb in soils is related to agricultural
management. Although Pb was found to be strongly sorbed to the soil, downward migration was observed
and the movement of Pb to deeper soils was due to the soil mixing activities of earthworms (Fernandez et
al.. 2007). Thus in relatively unpolluted non-forested soils, as in forested soils, colloidal Fe and OM, pH,
and biophysical transport all enhance Pb mobility in soil. Pb transport in more highly contaminated soils
has also been the subject of recent research. In a vegetated roadside soil, Pb was leached from the upper
50 cm of the soil even though the pH was 7.2. Pb was transported on mobile particles and colloids in the
soil solution. Some of the colloids may have formed from OM produced by roots and decaying shoots.
The transport process was enhanced by preferential flow triggered by intense rainfall events. This study
suggested that the value of the effective sorption coefficient estimated under dynamic conditions was
unrelated to values measured in conventional batch studies. This indicates that the use of batch studies to
derive input values for sorption coefficients in transport models requires caution. It was concluded that
the primary control of Pb transport in the long term was the degree of preferential flow in the system (S^
Roulier et al.. 2008).
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
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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-horizonof 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 A1 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 3-11, 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 & Rimstidt. 2009).
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1.0
Pb02
Plattnerite
,2+
0.5
PbCO,
>
Cerussite
.c
UJ
Pb(OH)3"
-0.5
2
5
0
3
6
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9 10 11 12 13 14
1
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7
pH
Source: Used with permission from Elsevier Publishing, Scheetz and Rimstidt (2009)
Figure 3-11. 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; J. M. Kaste et al.. 2006). As was concluded in the 2006 Pb AQCD (U.S. EPA. 2006).
soils continue to act as a predominant sink for Pb.
While in some studies the flux of Pb, from the soil through aquatic ecosystems to lakes has peaked
and declined. In other studies, no recovery of lake sediments in response to emission reductions was
observed (Norton. 2007). For example, Klaminder et al. (2010) has shown that the Pb concentrations in
sub-Arctic lake sediments remain unchanged in recent years, with the lack of recovery linked to the
effects of soil warming, which affect Pb-OM transport from soil to the receiving lake systems. Shotyk and
Krachler (2010) also reported a disconnect between atmospheric deposition and recent changes in Pb
concentration and isotope ratios in the lake sediments. Simulations of future metal behavior suggest that
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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. Size-Selective PM Monitoring for Lead Concentrations
Ambient Pb is present in the atmosphere as PM and distributed over a wide range of PM sizes. In
recognition of the role of all PM sizes in ambient air Pb exposures, including the ingestion of PM
deposited onto surfaces, the indicator for the Pb NAAQS is Pb in TSP As described in Chapter 4,
ingestion of deposited Pb can be a substantial contributor to total Pb exposure. Additionally, a substantial
fraction of atmospheric Pb may be associated with PM larger than 10 |_im (ultracoarse PM). However, the
variability of capture efficiency for TSP using current TSP samplers is considerably greater than the
capture efficiency for PM10 using current PM10 samplers. For example, the symmetrical design of Federal
Reference Method (FRM) samplers for PM10 makes their collection efficiency independent of wind
direction, and collection efficiency is independent of wind speed under typical sampling conditions.
While the collection efficiency of TSP samplers is nearly 100% for fine PM up to 5 (.un diameter, there is
much greater variability associated with collection of larger PM (Wedding et al.. 1977). For example,
using the FRM for TSP, a directional difference of 45 degrees can result in a nearly two-fold difference in
15 |_im particle collection efficiency and a nearly five-fold difference in 50 |_im particle collection
efficiency (Rodes & Evans. 1985). 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 (Wedding et al.. 1977).
Recognizing the variability in capture efficiency associated with TSP samplers and the potential
benefit of an indicator with lower measurement variability, the last NAAQS review considered whether
the indicator for the Pb NAAQS should be revised from one based on Pb-TSP to one based on Pb-PM10.
The final decision in the review was to retain the Pb-TSP indicator. The rationale for this decision
included recognition of exposure due to Pb-TSP that would not be captured by PM10 sampling, the
paucity of information documenting the relationship between Pb-PM10 and Pb-TSP at the broad range of
Pb sources in the U.S., and uncertainty regarding the effectiveness of a Pb-PM10-based NAAQS in
controlling ultracoarse Pb-PM near sources where Pb concentrations are highest (73 FR 66991). Changes
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were made to the monitoring and data handling provisions, however, to allow for the siting of Pb-PMi0
monitors for compliance purposes in locations remote from sources, where the evidence indicates that
airborne Pb is predominantly in the PMi0 size fraction (73 FR 66964). For Pb-PMi0 monitoring to be
allowed under these regulations, the maximum three-month average Pb concentration at a site must not
exceed 0.10 |ig/m3 over a three-year time period. Additionally, if a Pb-PMi0 monitor is sited near a source,
the majority of the particle-bound Pb must be smaller than 10 |im (40 CFR Part 58).
3.4.1.2. Federal Reference Method and Federal Equivalence Method Evaluation
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 federal equivalence methods (FEM). Measurements for determinations of NAAQS compliance must be
made with FRMs or FEMs. As of August 2010, 1 manual reference method and 24 manual equivalent
methods had been approved for Pb (http://www.epa.gov/ttn/amtic/criteria.html). The FRM for Pb was
promulgated in 1979 and is based on flame atomic absorption spectroscopy (AAS) (40 CFR Part 50). The
FRM provides for collection of PM by high volume sampling and analysis of the PM for Pb by atomic
absorption spectrometry. Ambient air suspended in PM is collected on a glass fiber filter for 24 hours
using a high volume air sampler. The analysis of the 24-hour samples may be performed for either
individual samples or composites of the samples collected over a calendar month or quarter. Pb in PM is
then solubilized by extraction with nitric acid (HN03), 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 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. Inductively-coupled plasma
mass spectrometry (ICPMS) is under consideration as the new FRM for Pb-TSP.
PM10 monitoring can be used in limited circumstances to measure Pb concentration. (40 CFR 58).
The proposed method is based on sampling requirements for an existing Federal Reference Method for
the Determination of Coarse PM as PM10 - PM2 5 ("Reference Method for the Determination of Coarse
Particulate Matter as PM 10-2.5 in the Atmosphere." 2010). which requires a specially approved PM10C
sampler that meets more demanding performance requirements than conventional PM10 samplers.
Ambient air is drawn through an inertial particle size separator for collection on a polytetrafluoroethylene
(PTFE) filter. The analysis method for the FRM is based on x-ray fluorescence spectrometry. In addition,
several FEM have been approved based on a variety of principles of operation have been approved,
including: inductively coupled plasma optical emission spectrometry, or ICPMS. Specifications for Pb
monitoring are designed to help states demonstrate whether they have met compliance criteria.
Operational parameters required under Appendix G of 40 CFR Part 50 are listed in Table 3-3.
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Table 3-3. Specifications for Pb monitoring
Parameter
Specification"
Range
0.7-7.5 |jg Pb/m°
Sensitivity
0.2-0.5 |jq Pb/mL for 1 % change in absorbance
Lower Detectable Limit
0.07 yg Pb/m°
aAssumes sample volume of 2,400 m3.
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 a suite of elements simultaneously. An additional
advantage is low detection limits of 50-100 parts/trillion for Pb.
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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. Chemical Speciation Network, IMPROVE, and National Air Toxics
Trends Network Monitors
In addition to being monitored for regulatory purposes in the SLAMS network, Pb is also
monitored in three other sampling networks. Pb is monitored at 53 monitoring sites as a part of the
Chemical Speciation Network. Participating monitoring agencies responsible for site operation are given
flexibility in sampler design, with filter collection media best-suited for the analysis of specific
components (U.S. EPA. 1999a). Several samplers are approved for CSN monitoring, all of which collect
bulk PM species with multiple channels containing different types of filters appropriate for speciation
sampling. Pb is one of 33 elements in PM2 5 collected on Teflon filters every third day and analyzed by
energy dispersive X-ray fluorescence spectrometry.
Pb is also monitored at 110 aerosol visibility-monitoring sites as a part of the Interagency
Monitoring of Protected Visual Environments (IMPROVE) program. An additional 59 aerosol samplers
that are not directly operated through the program are operated following IMPROVE protocols.
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IMPROVE is a cooperative effort by Federal and state organizations to protect visibility in 156 national
parks and wilderness areas as described in the 1977 amendments to the Clean Air Act. Objectives are: (1)
to establish current visibility and aerosol conditions in high priority (class I) areas for visibility protection;
(2) to identify chemical species and emission sources responsible for existing man made visibility
impairment (3) to document long-term trends for assessing progress towards visibility goals; and (4) to
provide regional haze monitoring representing protected federal areas in accordance with the regional
haze rule. The IMPROVE sampler operates with four sampling modules, three for PM2 5 and one for
PM10. Pb is not measured in PM10, but one of the three PM2 5 modules contains a Teflon filter used for
determination of gravimetric mass, absorbance, and elemental analysis by Particle Induced X-Ray
Emission (PIXE) and XRF. A total of 9 elements are determined by XRF, including Pb (University of
California Davis. 1995).
Pb in PM10 is also monitored in the National Air Toxics Trends Station (NATTS) network (ERG.
2009). PM 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. 1999b). Pb is one of seven
core metals collected on Teflon filters and analyzed by ICPMS. The NATTS network was developed to
fulfill the need for consistent data quality of long-term monitoring data on hazardous air pollutants of
consistent data quality, for use in assessing trends and emission reduction program effectiveness,
assessing and verifying air quality models, exposure assessments, emission control strategies, and as
direct input for receptor modeling. 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.
Typically more than 100 pollutants are measured at each site, and monitoring is required for nineteen
species, including Pb. Pb monitoring is also required for PM2 5 samples at NCore monitoring sites
beginning no later than January 1, 2011, and monitoring of Pb in coarse PM (PM10-PM2 5) is also likely to
be included (U.S. EPA. 2006).
3.4.1.4. Other Measurement 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
PIXE
Proton-induced X-ray emission (PIXE) spectroscopy has been widely used to measure Pb in
atmospheric PM. It is the method used for Pb analysis in the IMPROVE network. 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
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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 (Ariola et al.. 2006; Chan et al.. 2008; Cong et al.. 2007; Johnson et al.. 2006;
Johnson et al.. 2008; Sanchez-Ccovllo et al.. 2009; Wahlin et al.. 2006) (Cohen et al.. 2010; Waheed et
al.. 2010).
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 & 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 in of Pb sulfate (Qi et al.. 2006). XPS
analysis is a surface technique that is suitable only to a depth of 20-5 OA.
Other Total Lead Methods
Anodic stripping voltammetry, atomic emission spectroscopy, and colorimetry have also been used
for measurement of atmospheric Pb (Finlavson-Pitts & 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 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.
3.4.1.5. 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
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more mobile in soil and is more bioavailable to organisms. This approach has also been used widely in
characterization of airborne PM. In its original application (Tessier et al.. 1979) metals extraction solvents
were selected to correspond to common species present in soil, and metals were classified as
exchangeable, bound to carbonates, bound to iron 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 & 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 (Canepari et al.. 2006;
Falta et al.. 2008; Voutsa & 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 & Iwatsuki. 2002; Wang et al.. 2002). 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 (Birmili et al.. 2006; Dos Santos et al.. 2009). 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; Povkio et al.. 2007; Sillanpaa et al.. 2005;
Smichowski et al.. 2008) and locations (Al-Masri et al.. 2006; Annihnldi et al.. 2007; Canepari et al..
2006; Cizmecioglu & Muezzinoglu. 2008; Dahl et al.. 2008; Dos Santos et al.. 2009; Fuiiwara et al..
2006; Gutierrez-Castillo et al.. 2005; Heal et al.. 2005; Perrino et al.. 2010; Richter et al.. 2007; Sato et
al.. 2008; W. Wang et al.. 2006; Yadav & Raiamani. 2006). leading to a better understanding of mobility
characteristics of Pb in airborne PM.
3.4.1.6. Speciation Techniques
XAFS
There have been few attempts to speciate Pb in atmospheric PM into individual species. 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
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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 (X. Y. Guo et al..
2006). 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 (Oiievauviller. 2000). These have
been thoroughly reviewed (Pvrzviiska. 1996) and method intercomparison studies have been conducted
(Oiievauviller. 2000). More 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 systems capable of analyzing a wide range of elements along with Pb, and these
approaches have also been recently reviewed (Hirner. 2006). Chromatographic systems in common use
are gas chromatography 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) (Hirner. 2006). Using these techniques, organometallic species are separated from
each other based on differences in retention times on chromatographic columns, and elemental Pb is
determined by the ICPMS used as a detector downstream of the column to measure elemental Pb in the
pure compounds after chromatographic separation. Pb speciation analysis has benefited from the
development of HPLC-ICPMS in particular (Oiievauviller. 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 & Hcumann.
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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 (Jitaru et al.. 2004;
Leal-Granadillo et al.. 2000; Poperechna & Hcnmnnn. 2005V
3.4.1.7. 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 (T. A. Cahill et al.. 1987; Raabe et al.. 1988) 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 (Bench et al.. 2002; C. F. Cahill. 2003). The
University of Maryland Semi-continuous Elements in Aerosol Sampler (Ki dwell & Ondov. 2001. 2004)
uses direct steam injection into 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 (Pratheret al.. 1994). This technique
can also be considered as an effective method for real time Pb measurement in PM (Silva & 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 (Bein et al.. 2007;
Johnson et al.. 2008; Peknev et al.. 2006; Reinard et al.. 2007; Snvder et al.. 2009) or specifically targeted
(Moffet. de Foy. et al.. 2008; Murphy et al.. 2007; Salcedo et al.. 2010) Pb measurements.
3.4.2. Ambient Network Design
Ambient air Pb concentration is detected by FRM monitors at state and local monitoring stations
(SLAMS) reporting data used for NAAQS compliance to the Air Quality System (AQS). Monitoring
requirements for Pb have evolved over the past ten years. Monitoring for ambient Pb levels has been
required for all major urban areas where ambient air Pb measurements have been elevated near or beyond
the level of the NAAQS. Alternately, state and local agencies have located Pb monitoring stations in
proximity to Pb point source emissions. Prior to 2006, monitoring sites were established where sources
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emitted 5 or more tons/yr or where smaller stationary sources were located in proximity to populated
neighborhoods.
Pb monitoring requirements have experienced several changes since publication of the last Pb
AQCD (U.S. EPA. 2006). In 2008 revisions for the Pb NAAQS were announced, and new monitoring
requirements to support NAAQS revision called for expanded monitoring at sources that emit Pb at a rate
of 1.0 or more tons/yr and non-source oriented monitoring at each Core Based Statistical Area (CBSA)
with a population of 500,000 or more ("National Ambient Air Quality Standards for Lead (final rule)."
2008). This corresponded to approximately 100 non-source oriented monitors. Some of the new monitors
were required to become operational by January 1, 2010, with the remainder operational by January 1,
2011. Subsequent revisions to these requirements have been promulgated, including reduction of the
source oriented monitoring threshold from 1.0 tons/yr to 0.5 tons/yr and replacing the requirement for
population-based CBSA monitoring with a requirement for non-source oriented Pb monitoring at National
Core multipollutant monitoring network (NCore) sites in CBSA's with a population of 500,000 or more
(75 FR 81126). 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.
Data used in this chapter are from the period 2007-2009. The number of source oriented and non-
source oriented monitors changed each analysis year because the monitoring requirements changed over
this time. Monitors were designated to be source-oriented if they were designated in AQS as source
oriented, or they were located within one mile of a 0.5 ton/yr or greater source, as noted in the 2005 NEI
(U.S. EPA. 2008a). Non-source oriented monitors were those monitors not considered to be source
oriented. This loosening of the restrictions was intended to accommodate the 2007-2008 data that were
obtained before the latest monitor designation requirements were implemented.
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. While monitoring in multiple networks improves geographic
coverage, measurements between networks are not directly comparable in all cases because different PM
size ranges are sampled in different networks. Depending on the monitoring network, Pb is monitored
either in TSP, PMi0, or PM25.
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3.4.2.1. Monitor Siting Requirements
Spatial scales defined for Pb monitoring range from microscale to regional scale (40 CFR Part 58):
¦ Microscale: Defines the concentrations in air volumes associated with area dimensions
ranging from several meters up to about 100 m. This scale applies to areas in close proximity
to Pb point sources.
¦ Middle Scale: Defines the concentration typical of areas up to several city blocks in size with
dimensions ranging from about 100 m to 0.5 km. This scale generally represents Pb air
quality levels in areas up to several city blocks.
¦ Neighborhood Scale: Defines concentrations within some extended area of the city that has
relatively uniform land use 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.
¦ Urban Scale: Defines concentrations within area of city-like dimensions on the order of 4 to
50 km. Such stations would be used to present ambient Pb concentrations over an entire
metropolitan area with dimensions in the 4 to 50 km range. An urban scale station would be
useful for assessing trends in citywide air quality and the effectiveness of larger scale air
pollution control strategies.
¦ Regional Scale: Defines usually an area of reasonably homogeneous geography without
larges sources, and extends from tens to hundreds of kilometers. This large scale of
representativeness has not been used widely for Pb monitoring, but provides important
information on background air quality and inter-regional pollutant transport.
Since the majority of Pb emissions mass comes from point sources, such as metals processing
facilities, waste disposal and recycling, and fuel combustion, the SLAMS network is primarily used to
assess the air quality impacts of Pb point sources. A second purpose of the SLAMS network is to
determine the broad population exposure from any Pb source. The most important spatial scales to
characterize the emissions from point sources are the micro, middle, and neighborhood scales.
Background information such as point source emissions inventories, climatological summaries, and
local geographical characteristics are used to identify areas where monitoring is necessary. After siting
each Pb station, specific siting criteria must be fulfilled for Pb monitoring in the SLAMS network. Micro
and middle scale monitors must be 2-7 m above ground. All other scale monitors must be 2-15 m above
ground. Monitors must be more than 2 m from supporting structures. Monitors must be more than 10 m
from trees. Microscale monitors designed for monitoring traffic related Pb must be 2-10 m from
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roadways. Distance from roadw ays for other scales depends on the purpose and scale of the monitor and
the level of traffic on the roadway.
3.4.2.2. Spatial and Temporal Coverage
Pb Monitor Locations in the United States in 2007-2009
* TSP nonsource Monitors * 4
- TSP source Monitors
- PM,a nonsource Monitors ^ * A * PM? 6 Monitors
Counties | 1 Counties
United States United States
* NATTS Monitors A * IMPROVE Monitors
I I Counties 1 i Counties
United States United States
Figure 3-12. Pb monitoring sites for SLAMS, CSN, NATTS, and IMPROVE
networks, 2007-2009.
Figure 3-12 shows Pb monitoring sites for all networks. The top left map shows the SLAMS
monitors reporting to the AQS system from 2007 to 2009. Monitors are indicated as Pb-TSP source
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oriented, Pb-TSP non-source oriented, and PMi0 non-source oriented. Source and non-source designations
used in the data analysis in Section 3.5 are indicated on this map. The top right map shows PM2 5 monitors
from the CSN. The bottom left map shows National Air Toxics Trend Networks sites, and the bottom right
map shows IMPROVE monitoring sites. There is a high density of FRM monitors in some cities
containing Pb sources, including Los Angeles, St. Louis, Pittsburgh, and Minneapolis. As a result,
population coverage varies across cities, with those cities with major Pb sources having greater coverage.
Coverage for PM2 5 in the CSN and IMPROVE network is geographically comprehensive. In comparison,
the FRM sites are more representative of source effects. NATTS sites have less extensive national
coverage.
3.5. Ambient Air Lead Concentrations
The following section summarizes data on ambient air Pb concentrations during the years 2007-
2009. The section begins with a description of Pb concentrations observed in TSP, PM10, and PM25 at
source oriented and non-source oriented monitors across the U.S. 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 AQS
data and the peer-reviewed literature is provided to evaluate the size distribution of Pb-bearing airborne
PM under varied ambient conditions. Finally, the relationship between Pb concentration and
concentrations of copollutants are presented. Summary information is presented within this section, and
detailed data are included in an Appendix 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 TSP FRM monitors, PM10 monitors, and PM2 5 monitors from the CSN for 2007-2009.
This information is useful to develop a sense of variability in Pb concentrations at a national scale. For
this analysis, source oriented monitors encompassed all those listed as "source oriented" in the AQS,
based on state agency reporting, plus those within one mile of a facility emitting 0.5 ton/yr or more. Non-
source oriented monitors were those monitors in the system not designated to be source oriented (Figures
3-13 though 3-16), the majority of U.S. counties do not have a Pb monitor.
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Concentrations of Pb Measured using TSP Monitors
Source oriented maximum 3-month average Pb data were obtained for 22 counties across the U.S.
during the period 2007-2009. Figure 3-13 illustrates that the level of the NAAQS was exceeded in
fourteen counties where source oriented monitoring was performed. Summary data are presented below in
Table 3-4, and detailed data for the one-month and three-month average and maxima source oriented Pb-
TSP concentrations are provided in Tables 3A-1 through 3A-4 in the Appendix. The mean was skewed
towards the 75th percentile of 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 national maxima were reached for 2007-2009 are
presented in Table 3-5. The highest monthly and three-month average concentrations occurred at the same
sites: Herculaneum, MO followed by Los Angeles, CA. The highest annual site max 1-month value
occurred in Cook County, IL in 2008, followed by Iron County, MO in 2008 and Hillsborough County, FL
in 2007. The Herculaneum and Los Angeles sites were also above the 90th percentile annual 3-month site
max Pb concentrations. The highest annual site max 3-month concentrations occurred in Herculaneum in
2008, Los Angeles in 2008, and Iron County, MO in 2008. The majority of data were below the level of
the NAAQS over the 3-year period, but high values at a subset of source oriented monitoring locations
tended to skew the nationwide distribution of Pb concentration data upwards.
Table 3-4.
Summary data for source oriented Pb monitors across the U.S.
Mean, pg/m3
Median, pg/m3
95th%, pg/m3 99th%, jjg/m3 Max, \iglm*
Monthly
0.24
0.070
0.98 2.1
4.4
3-mo rolling avg
0.24
0.080
0.98 1.9
2.9
Table 3-5.
Summary data for sites at which source oriented statistics are at a maxima
County
Ano Highest Monthly Highest 3-mo
Mean, pg/m Mean, pg/m
Highest Monthly
Annual Site Max,
[jg/m3 (Year)
Highest 3-mo
Annual Site Max,
[jg/m3 (Year)
Jefferson, MO
290990015 1.3
1.3
2.9 (2008)
Jefferson, MO
290990004 1.1
1.1
Los Angeles, CA
060371405 0.86
0.93
2.5 (2008)
Chicago, IL
180350009
4.4 (2008)
Iron, MO
290930016
4.2 (2008)
2.5 (2008)
Hillsborough, FL
120571066
3.6 (2007)
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2007-2009 Pb-TSPSource-Oriented County Maximum 3-Month Mean
Concentration:
" >= 1.55 ng/m3(5counties)
¦ 0.76-1.54 ng/m3 (2 counties)
" 0.16 - 0,75 ng/rrr (7 counties)
¦ 0.06 - 0.15 ng/m3 (4 counties)
<= .05 [ig/m5 (4 counties)
~ no data
Figure 3-13. Highest county-level source oriented Pb-TSP concentrations
(|jg/m3), maximum 3-month average, 2007-2009.
Non-source oriented maximum 3-month average Pb concentration data were obtained with TSP
monitors for 36 counties across the U.S. during the period 2007-2009; non-source PMi0 monitoring data
are not included here because they were obtained at select sites in 2009 only. The median for monthly and
3-month rolling average data was 0.010 (.ig/m ". Detailed data for site-specific Pb-TSP concentrations are
provided in Tables 3A-5 and 3 A-6 in the Appendix to this chapter. Nationwide, the mean monthly average
non-source onented Pb-TSP concentration for 2007-2009 was more than an order of magnitude lower
than the source-oriented data. Collectively, these data indicate that non-source oriented monitors tend to
measure concentrations an order of magnitude lower than the level of the NAAQS.
Concentrations of Pb Measured using PM10 Monitors
Figure 3-14 displays maximum 3-month average county-level data for Pb in PM10 concentrations
for 36 counties in which measurements were obtained. The data presented here are not compared to the
level of the NAAQS because PMi0 monitors were not incorporated into the non-source oriented
monitoring network until 2009. Among the 36 counties in which PM]0 monitoring was conducted, only
one county, Gila County, AZ, reported concentrations above 0.076 jxg/m3. Three other counties reported
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1 concentrations greater than 0.016 (.ig/nr: Wayne County, MI, Boyd County. KY, and the county of St.
2 Louis City, MO.
2007-2009 Pb-PMjo County Maximum 3-Month Mean
«
L,rr
i
!< \ •
_ i
. v: •
u/Tn,
m
m
fFJk
SreSM gss-
"i
V
m,
lyl
Concentration:
" >= 0.076 ng/m3 (1 county)
" 0.016-0.075 Mg/m: (3 counties)
" 0.006 - 0.015 ng/m- (17 counties)
<= .005 ng/m3 (15 counties)
~ no data
i E
: n-i TO5
Sfi
m~r
H4J4
TUJitl
A]
;
Figure 3-14. Highest county-level Pb-PM10 concentrations (|jg/m3),
maximum 3-month average, 2007-2009.
3 Figure 3-15 displays maximum 3-month average county-level data for Pb in PM2 5 concentrations
4 for 323 counties in which PM2 5 measurements were obtained for speciation in the CSN and IMPROVE
5 networks. The data presented here are not compared to the level of the NAAQS because PM2 5 monitors
6 are not deployed for that purpose. Among the 323 counties in which PM2 5 monitoring was conducted,
7 only eleven counties reported concentrations greater than 0.016 ug/'ni : Jefferson, AL, San Bernardino,
8 CA, Imperial, CA, Wayne, MI, Jefferson, MO, Erie, NY, Lorain, OH, Allegheny, PA, Berks, PA,
9 Davidson, TN. and El Paso, TX.
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Concentration:
Hg/m3(ll counties)
¦ 0.006 -0.015 pg/m3(71 counties)
<= .005 \xg/\r\"' (241 counties)
~ no data
2007-2009 Pb-PM2 5County Maximum 3-Month Mean
Figure 3-15. Highest county-level Pb-PM2.5 concentrations (ng/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, spatial
variability of Pb concentrations in urban areas may also be high. Such variability may not be detected if
one or a small number of central site monitors is in use, so cities with multiple monitors can be used to
characterize intra-urban variability. Intra-urban variability in Pb concentrations reported to AQS at the
individual county level was described in detail in the Appendix in Section 3A.2. County-level data were
used because PM-bound airborne Pb tends to settle over short distances to produce large spatiotemporal
variability, as described in Section 3.3.1 and revisited in Section 3.5.3. 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 combinations of source and urban contributions of Pb. 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.
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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. 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. Additionally, resuspension of wheel weights and
soils, emission of trace Pb during on-road gasoline combustion, and urban background levels of Pb are
uncertain influential factors in ambient Pb concentrations. This is consistent with field studies to
characterize Pb concentrations that are described in the literature.
A number of studies have characterized how Pb-bearing PM is distributed over the neighborhood
scale in the air. Martuzevicius et al. (2004) examined the spatial variability of PM25 samples obtained in
the greater Cincinnati, OH area at 6 urban, 4 suburban, and 1 rural site and found that Pb concentration in
PM2 5 had a coefficient of variation (CV, defined as the standard deviation of site measurements divided
by the average) of33.8%, compared withaCV for PM25 of 11.3% over all sites. Average Pb-PM25
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. (2006) measured Pb-PM with an upper cutpoint of 29 |im 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. (2006) 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.
Han et al. (2007) found that Pb concentrations within re suspended road dust were higher at an inner-city
ring road and an industrial site compared with residential or construction sites. Li et al. (2009) observed
that Pb concentration in PM2 5 samples was 2.3-3.0 times higher near a bus depot than next to a rural-
suburban road. Ondov et al. (2006) measured Pb-PM2 5 concentration at three Baltimore sites, one of
which was industrial and the other two of which were considered "receptor" sites. Average Pb-PM2 5
concentrations at the different sites were 8.3 ng/m3 at the industrial site and 1.9 ng/m3 and 7.2 ng/m3 at the
receptor sites. As a group, these studies support the analysis of intra-urban AQS data by illustrating that
intra-urban variability is most strongly related to type, strength, and location of sources.
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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
Figure 3-16 illustrates the trend in ambient air TSP-Pb concentrations during the years 1980-2009.
Over this time period, average air Pb concentrations have declined by 91% from 1.3 |ig/m3 (in 1980) to
0.12 (ig/m3 (in 2009). The median concentrations have declined by 97% from 0.87 (ig/m3 (in 1980) to
0.025 |ig/m3 (in 2009). 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 2009 following reductions in industrial use and processing of Pb, as described in Section 3.2.1.
In 1990, the median Pb concentration was 0.19 |ig/m3 and the average Pb concentration was 0.55 |ig/m3 to
yield 87% and 78% reductions, respectively, from 1990 to 2009. Average concentrations in these
calculations and in Figure 3-16, 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.
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4
3.5 -
3
£ 2.5
1
c
o
E 2 -
c
-------
3
2.5
\
2
1978-2008 Level of the NAAQS
1.5
/\
1
0.5
2009^vdoltheNAAQ^
0
% X X % % % %
\ % X
% X
X
Figure 3-17. National trends in Pb concentration (pg/m3), source oriented
FRM monitors, 1990-2009.Average concentration is shown by
the solid line, and the 10th and 90th percentiles are shown by
the dashed lines.
0.6
0.5
0.4
c
o
E
^ 0.3
OJ
c
o
u
-Q
Q.
Q.
1/1
0.2
2009Levej^theNAAQ^
0.1
0
\ \ % \ \ \ \ % \ \ % % % % \ \ \ % % ^
Figure 3-18. National trends in Pb concentration (pg/m3), non-source
oriented FRM monitors, 1990-2009. Average concentration is
shown by the solid line, and the 10th and 90th percentiles are
shown by the dashed lines.
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For source oriented monitors, average concentrations decreased from 0.93 (.ig/ni3 to 0.23 |_ig/m3
(75% decline) and upper 90th percentile concentrations decreased from 2.5 (ig/m3 to 0.54 (ig/m3 (78%
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. 2010). An abrupt decrease in
average concentrations between these years is evident in Figure 3-17. Note that the number of monitors
contributing to these statistics increased from 29 during 1990-1999 to 47 for 2000-2009.
For non-source oriented monitors, average concentrations decreased from 0.18 (ig/m3 to
0.020 (ig/m3 (88% decline) and upper 90th percentile concentrations decreased from approximately 0.49
(ig/m3 to approximately 0.05 (ig/m3 (87% decline) over the 20-year period. Average concentrations near
stationary sources were 5 to 12 times typical concentrations from non-source oriented monitoring
locations between 1990 and 1999; during the subsequent decade, average source oriented Pb
concentrations were 8 to 22 times higher than non-source concentrations (U.S. EPA. 2010). This
differential likely reflects the absence of Pb emissions from automobiles during 2000-2009. The number
of monitors contributing to these statistics increased from 29 during 1990-1999 to 53 for 2000-2009.
When both source oriented and non-source oriented monitoring sites are considered, average Pb
concentrations decreased by 73% between 2001 and 2008 for maximum 3-month average concentrations
at 24 sites that are near large stationary sources and 101 sites that are not near stationary sources (U.S.
EPA. 2010).
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, PM10, and PM2 5 data.
The average of Pb concentrations over all monitoring sites is higher in fall and lower in winter than
other seasons. Monthly average Pb concentrations averaged over multiple sites and over 3 years from
2007-2009 are shown in for Pb-TSP from source oriented monitors (Figure 3-19), Pb-TSP from non-
source oriented monitors (Figure 3-20), Pb-PM10 (Figure 3-21), and Pb-PM2 5 (Figure 3-22). For source
oriented Pb-TSP (Figure 3-19), monthly average concentrations were determined from between 146 and
154 samples in each month. The highest monthly average concentrations were observed in March, April,
and November, and exceeded 0.26 (ig/m3. For non-source oriented TSP (Figure 3-20), 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
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concentrations, each of which were below 0.015 |ig/m3; but concentrations were similar between spring,
summer and fall months.
1.5-
1.4-
1.3-
1.2-
1.1 -
1.0-
0.9-
0.8-
0.7-
0.6-
0.5-
0.4-
0.3-
0.2-
0.1 -
0.0-
—I—
Jan
—I—
Feb
—I—
Mar
Apr
May
—I—
Jun
—r~
Jul
—I—
Aug
Sep
Oct
Nov
—I—
Dec
Month
Figure 3-19. Monthly source oriented Pb-TSP average (pg/m3) over 12
months of the year, 2007-2009.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.
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—I—
Jan
—I—
Mar
—I—
Apr
-1—
Jul
—I—
Aug
—I—
Nov
I
~~r~
Dec
Feb
May
Jun
Sep
Oct
Month
Figure 3-20. Monthly non-source oriented lead-TSP average (pg/m3) over 12
months of the year, 2007-2009.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.
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 -
—I—
Jan
—I—
Mar
—I—
May
—r~
Jul
—I—
Aug
—I—
Oct
—I—
Dec
Apr
Sep
Month
Figure 3-21. Monthly lead-PM10 average (pg/m3) over 12 months of the year,
2007-2009.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.
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0.010 -
0.009 -
0.008 -
0.007 -
E
§ 0.006-
| 0.005 -
I 0.004 -
o
c
° 0.003 -
0.002 -
0.001 -
0.000 -
Figure 3-22. Monthly lead-PM2.5 average (|jg/m3) over 12 months of the year,
2007-2009.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.
For both Pb-PM10 (Figure 3-21) and Pb-PM2 5, (Figure 3-22) monthly average concentrations are
considerably higher in the fall than in other seasons, with lowest the three highest monthly average
concentrations observed in September, October, and November, and the average September concentration
more than double the average December concentration. Pb-PM10 monthly average concentrations were
determined from between 100 and 109 samples and Pb-PM2 5 from between 866 and 1,034 samples each
month. Some of the Pb-PM2 5 concentrations were below Method Detection Limits and the absolute
difference in monthly average concentrations between fall and other seasons for Pb-PM2 5 of 0.001 (.ig/nr1
is extremely small. In spite of this the observed trends for PM2 5 are consistent with the PM10
observations. Whether the seasonal trend for Pb-PM10 and Pb-PM2 5 differs from trends for both source
oriented and non-source oriented Pb-TSP because of the difference in source proximity, sampling
locations or difference in size range sampled is not clear.
Although details of the seasonal trends varied with PM size range and source proximity, the data as
a whole indicate that average monthly concentrations in the fall were consistently higher than the lowest
average monthly concentrations, and that average monthly concentrations in the winter were consistently
lower than the highest average monthly concentration regardless of PM size range or source proximity.
Overall, there this indicates a clear tendency toward higher fall concentrations and lower winter
concentrations. These results are consistent with observations at a single location by Melaku et al. (2008)
*
*
1
3
c r>ri p
c x b
t
- *
r l-r
T 3
C i—
r 3
L _
c 4
L _L
¥
* 3 C
-ill
~~1 i i i i i i i 1 i i i—
Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec
Month
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that atmospheric Pb concentrations in fall were higher than in winter in an intensive study of urban
Washington DC, but not consistent with their observations of higher summer concentrations than fall.
3.5.3. Size Distribution of Lead-Bearing PM
Size-selective monitoring data from AQS and the literature is examined in this section to improve
understanding of the size distribution of Pb-bearing PM. This information informs monitoring strategies
because high content or correlation of PM10 with TSP may allow for expanded usage of PM10 within the
FRM monitoring network. Additionally, size distribution data enhances understanding of the relationship
between sources and characteristics of airborne Pb-bearing PM.
3.5.3.1. AQS Data Analysis
This section employs AQS data for Pb concentrations from co-located TSP, PMi0, and/or PM2.5
FRM monitors to analyze correlations and ratios of concentrations obtained from the different monitors.
These data were used because relationships among the monitors provide information about the nature of
Pb-bearing PM at different locations (e.g., whether the mode is in the fine or coarse fraction). Correlations
indicate the extent to which the size fractions vary together in time, and the ratios signify the average
proportion of the smaller fraction to the larger fraction (e.g., the ratio of PM25 to PMi0 concentrations).
Estimation of the size distribution of Pb-bearing PM is possible at a limited number of monitoring
sites where monitors having different size-selective cut-points are co-located. Data for correlations
between monitors and average concentration ratios are available per collocation site in Table 3A-13 in the
Appendix. For the comparison between Pb-TSP and Pb-PMi0, 27 sites were available for analysis. A
summary of these data are provided in Table 3-6. Overall, the average correlation, p, was moderate,
although the wide range across monitors indicates suggests site-to-site variability. The average ratio of
Pb-PMio to Pb-TSP concentrations was relatively high at 0.88. When broken down by land type, the
average p went up slightly for urban and city center land use, with a lower average ratio of concentrations.
For suburban sites, the average p was reduced, but the average ratio of concentrations was near unity.
Average ratio of concentrations greater than one suggest that some portion of the TSP was not collected
due to instrument biases, as discussed in Section 3.4.
Table 3-6.
Summary of comparison data for co-located lead-TSP and lead-PM
10 monitors.
Monitors
Correlation
Average Ratio
PMi0:TSP
Average
Standard
Deviation
Range
Average
Standard
Deviation
Range
Overall
0.65
0.30
0.00-0.99
0.88
0.51
0.33-2.60
Urban and City
Center
0.73
0.25
0.29-0.99
0.73
0.19
0.54-1.10
Suburban
0.55
0.33
0.00-0.98
0.99
0.74
0.33-2.60
Rural
-
-
-
-
-
-
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Forty-five sites were available for analysis to compare Pb-TSP with Pb-PM2 5. A summary of these
data are provided in Table 3-7. Overall, the average p was moderate. As for the TSP to PMi0 comparison,
a wide range of correlations suggests that the tracking of the TSP and PM2 5 time series was quite variable
between sites and source influences. The average ratio of Pb-PM25 to Pb-TSP concentrations was
somewhat high, but with several monitors having ratios above one, sampler bias may have influenced
these statistics. When broken down by land type, the average p was somewhat reduced for urban and city
center land use. For suburban sites the average ratio of concentrations was lower, with some of the same
sampling bias issues. For rural sites, the average p was much higher. Correlation between PM2 5 and TSP
may suggest commonality of sources or processes influencing both size fractions at the same time.
Table 3-7.
Summary of comparison data for co-located Pb-TSP and Pb-PM2.s
monitors
Correlation
Average Ratio
PM2.5:TSP
Monitors
Average
Standard
Deviation
Range
Average
Standard
Deviation
Range
Overall
0.50
0.25
0.03-0.95
0.79
0.49
0.11-2.34
Urban and City
Center
0.43
0.26
0.03-0.90
0.75
0.54
0.11-2.34
Suburban
0.55
0.23
0.16-0.95
0.66
0.29
0.36-1.57
Rural
0.71
0.24
0.37-0.90
1.36
0.39
0.83-1.76
Pb-PM10 and Pb-PM2 5 data were compared for 50 sites. A summary of these data are provided in
Table 3-8. Overall, the average p was moderately high, with perfect correlation at a Providence, RI site
(AQS ID: 440070022) and very high correlation at several other sites. The average ratio of Pb-PM2 5 to
Pb-PM10 concentrations was 1.01, suggesting that some bias existed in the data so that PM2 5 was often
higher than PM10 concentration. Data for urban and city center and suburban sites were similar for
correlations and average ratios of concentration. For rural sites, the average p was slightly lower, and the
range of ratios of concentrations showed that PM2 5 concentrations were almost always higher than PM10
concentrations, suggesting bias among the rural monitors.
Table 3-8.
Summary of comparison data for co-located Pb-PM
10 and Pb-PM2.5
monitors.
Monitors
Correlation
Average Ratio
PM2.5: PM10
Average
Standard
Deviation
Range
Average
Standard
Deviation
Range
Overall
0.76
0.22
0.15-1.00
1.01
0.37
0.45-2.15
Urban and City
Center
0.72
0.24
0.15-1.00
0.85
0.38
0.45-2.15
Suburban
0.81
0.17
0.34-0.98
1.08
0.35
0.61-1.89
Rural
0.66
0.28
0.22-0.99
1.23
0.29
0.97-1.63
In a few cases, Pb-TSP, Pb-PM10, and Pb-PM2 5 monitors were co-located simultaneously, so that
more information regarding size distribution can be discerned. For example, in urban Jefferson County,
AL (AQS: 010730023), the average ratio of PM2 5 to TSP was 0.84, while the average ratio of PM10 to
TSP was 0.80 for the years 2005-2006. These reported values suggested that most of the PM was in the
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fine mode. But, the data also indicated instrumentation bias with either the PM25 or PMi0 monitors or
both, because the average concentration ratio was greater than one. In a suburban portion of Cook County,
IL (AQS: 170314201), the PM2 5 to TSP average ratio was reported to be 0.36, and the PMi0 to TSP
average ratio was reported to be 0.39. This suggested that the majority of the sample mass was from
particles larger than 10 |im. but that the distribution might have been bimodal since there was little
difference between PM2 5 and PMi0. Note that only some of the years reported for all three monitors
overlapped. In suburban Wayne, MI (AQS: 261630033), the PM25 to TSP average ratio was reported to
be 0.49, and the reported PMi0 to TSP ratio was reported to be 0.84. This suggests a smoother distribution
with one mode likely in the large fine or small coarse region of the distribution. Note that the PM2 5 to
TSP collocation includes one year of data more than the PMi0 to TSP collocation. Taken together, these
findings imply that the Pb-bearing PM size distribution varies substantially by location.
3.5.3.2. Size Distribution Studies in the Literature
Several studies in the literature have been designed to characterize the size distribution of Pb
concentrations in the vicinity of sources. The following section describes studies that have measured more
than one size fraction of Pb-bearing PM in the vicinity of industrial, urban, and/or traffic-related sources.
Some discussion of the variability of size distribution over space and time is also provided.
Size distributions of Pb-bearing PM have been measured near several industrial sites. For example,
Bein et al. (2006) measured the size distribution of Pb in PM from the Pittsburgh superfund site using
rapid single particle mass spectrometry and a Multi-Orifice Uniform Deposit Impactor (MOUDI). The
Pittsburgh, PA, superfund site had seventeen major PM sources within a 24-km radius. Bein et al.'s
(2006) measurements yielded different results on different days, with a bimodal distribution with modes
around 140 nm and 750 nm during an October, 2001 measurement and a single dominant mode around
800 nm during a March, 2002 measurement. Differences in the size distributions could have been related
to differences among wind speed, wind direction, and source contributions on the respective dates. Singh
et al. (2002) measured the mass distribution of Pb-PM in the coarse and fine PM size ranges 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-PM size distribution was unimodal with
a concentration peak in the 100-350 nm size range with 34% of the particles in this size bin. At the
downwind site, a bimodal distribution was observed with peaks in the 2.5-10 (j,m bin and the 350 nm-1
(jm bin, comprising 42% and 26% of the mass measured as PM10, respectively. Pb in the fine range only
comprised 13% of the particles in the 100-350 nm bin. 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.
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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 concentrations in the coarse to
fine PM size range. The size distribution was multimodal with a primary mode around 1 |im. a secondary
mode around 300 nm, and a very small additional mode around 5 |im. 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, 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 PMi0 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 PMi0 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.
Traffic can be a source of resuspension of Pb from deposited contemporaneous wheel weights or
industrial emissions or historic sources via traffic-induced turbulence. Sabin et al. (2006) compared the
size distribution of coarse Pb-PM captured at an urban background site and at a location 10 m from the I-
405 Freeway in the southern California air basin; data from Sabin et al. (2006) are displayed in Figure 3-
23. For both the urban background and near-road sites, the largest fraction was from PM sampled below
the 6 (mi cut point, but the near-road Pb-PM distribution appeared bimodal with a mode in the largest size
fraction. Over all 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. (2006) 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. 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. The Pb 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.
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 wheel weights rather than historic Pb on-road
gasoline emissions. Pb dust was shown by Bukowiecki et al. (2010) to be significantly higher at roadside
samples compared with urban background when the PM was in the coarse mode, measured as PMi0.2.5,
but not for fine modes measured as PM2 5_i and PMi_0.i. Chen et al. (2010) measured Pb in PMi0, PM2 5,
and PM0.i at a roadside location and in a tunnel in Taipei, Taiwan in 2008. While roadside and tunnel
concentrations of PMi0 and PM2 5 were roughly equivalent, Pb in PM01 was approximately 15 times
higher than by the roadside. The authors suggest that particle-bound Pb was emitted from on-road
gasoline and diesel engines. This could possibly be attributed to residual Pb in unleaded gasoline.
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¦ Urban Background ¦ Near-Road
Source: Adapted, with permission from Elsevier Publishing, from Sabin et al. (2006).
Figure 3-23. 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 non-road
sources. Harrison et al. (2003) measured the distribution of Pb in PMi0 at a roadside sampler in
Birmingham, U.K.. The size distribution was unimodal with approximately 2% of the Pb mass
(0.5 ng/irf) above the 10 |im cut point, 12% (3 ng/irf) in the 2-10 |im bin, 9% (2 ng/irf) in the 1-2 |im
bin, 53% of the Pb mass (14 ng/irf) in the 0.2-1.0 |im bin, and 24% (7 ng/irf) collected below the 0.2 |im
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 traffic (|3 = 0.067, R2 = 0.38) as well as PMi0 (|3 = 0.26,
R2 = 0.35). Briiggemann et al. (2009) observed a unimodal Pb size distribution with 51% of the mass in
the 0.42-1.2 (.un size bin. During winter, Pb concentrations were twice as high as during the summer, and
they were also higher when winds blew from the east. Briiggemann et al. (2009) suggested that this
finding reflected coal burning sources rather than road dust resuspension. Wang et al. (2006) observed a
bimodal Pb distribution in a heavily trafficked area of Kanazawa, Japan with incineration and generation
facilities also nearby. They observed a bimodal distribution with modes at the 0.65-1.1 |im and the 3.3-4.7
|im size bins. Wang et al.'s (2006) source apportionment work in this study suggested that the fine mode
derives from incineration and combustion of oil and coal.
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 |im
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and a slightly smaller secondary peak at 5.6 |im. 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-PM2 5 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-
PM10 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, PM25, 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/PMi0 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. 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 (jm cutpoint and on the 1.5-3.0 |im stage for near-road, in a traffic tunnel, and remote
and urban background sites. The highest concentrations were measured in the tunnel, at 3.3 ng/m3 for Pb-
PM0 5 and 10 ng/m3 for Pb-PMi 5.3 0. 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-PM) 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. The distribution of data along the curves did not change substantially
under the different conditions examined. 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.
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
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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 Pb 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 TSP-Pb concentrations with concentrations of other
criteria pollutants are summarized in Figure 3-24 for 2007-2008 data from 129 monitoring sites, and in
Figure 3-25 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
PMio. In Figure 3-24, and fewer for each copollutant in Figure 3-25. Each of these figures illustrates co-
pollutant correlations across the U.S. Additionally, seasonal correlations between Pb and co-pollutants are
provided in Figures 3A-19 through 3A-21 in the Chapter 3 Appendix, with seasonal co-pollutant
measurement data from the literature (Table 3A-14). As evident in each figure, there were considerably
fewer source-oriented sites available for co-located comparisons.
Source, SO2 -
US Overall
00 0
Non-Source, SO2 -
0
OdXIIHMOuum OOCXPO O O (3D
(
Source, PM2.5 -
O O
Non-Source, PM2.5 -
0
O aXSXXmHHBOK
wdoo) o
(
Source, PM10 -
8
o
o
Non-Source, PM10 -
0
qbmbbdbd^bob
KOOOGD
Source, O3 -
o o o o
Non-Source, O3 -
0 00 (do as a
IO (HIMl> OHDOOflD OO
Source, NO2 -
OO
Non-Source, NO2 -
ooood caapawnn
ODD O
Source, CO -
o o
Non-Source, CO -
0 00
oood oaucoDUD ano o
CD o
c
t 1 r
-1.0 -0.5 0.0 0.5 1.0
Spearman Correlation Coefficient
Figure 3-24. Correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2007-2008.
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Non-Source, SO2 -
US Overall
00 00
Source, PM2.5 -
0
Non-Source, PM2.5 -
CCD OOO
Non-Source, PM10 -
coo
0
0
0
Non-Source, O3 -
0
000 0
Non-Source, NO2 -
0 0
Non-Source, CO -
0
0
0
0 0
t 1 r
-1.0 -0.5 0.0 0.5 1.0
Spearman Correlation Coefficient
Figure 3-25. Correlations of monitored Pb-TSP concentration with co-
pollutant concentrations, 2009.
Overall, Pb was most strongly associated with PM2 5, PMi0 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 R = 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 R =
0.29 for CO, 0.17 for S02). The poorest associations were observed between Pb and O, (median R =
0.00). Although the overall associations of Pb concentration with PMi0 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 3-26, which describes
data from 3 years of CSN results. 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). Br, Cu, and K 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. Other species more useful for as diagnostic indicators of crustal, general combustion,
industrial emission, and coal combustion processes exhibited weaker, but still remarkable associations
with Pb, including crustal elements (median R = 0.32), EC (median R = 0.32), Mn and Fe (median R =
0.32 and 0.34, respectively), S (median R = 0.27), and Se (median R = 0.27). Except for S, in summer
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1 associations with each of these species were much weaker than in other seasons, with a Spearman
2 Correlation Coefficient greater than R = 0.3 observed only for Zn (median R = 0.37). The weakest
3 associations were with As, CI, Hg, Ni, N03, and Na (median R = -0.03 to 0.10).
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US Overall
Zinc -
Volatile Nitrate ~
Vanadium ~
Titanium
Sulfur "
Sulfate
Sodium Ion
Sodium
Silicon
Selenium
Potassium Ion
Potassium
Organic Carbon, Blank Adjusted
Non-Volatile Nitrate
Nitrate
Nickel
Mercury
Manganese
Magnesium
Iron
Elemental Carbon
Crustal -
Copper'
Chromium"
Chlorine
Carbonate Carbon ~
Calcium -
Cadmium
Bromine -
Arsenic -
Ammonium -
Aluminum ~
O <3© O
' '•)«>>!»
O O
O GD
O O
OOO
-1.0 -0.5 0.0 0.5 1.0
Spearman Correlation Coefficient
Figure 3-26. Correlations of monitored lead-PM2.5 concentration with
copollutant concentrations, 2007-2009.
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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 in 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. 2006) 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. Pb accumulation occurs more readily
for Pb salts applied to soils than for 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. 2006) 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. 2006) reports data from the
U.S. Geologic Service National Water-Quality Assessment (NAWQA), which are updated every ten years.
In the NAWQA 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.
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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. (2010b) 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 262 mg/kg, with maximum levels ranging from 461 to 348,000 mg/kg
(see Table 3-9). Soil Pb was thought to originate from present-day sources, such as industry, debrided
paint, and piston engine aircraft fuel, as well as historic sources, such as on-road gasoline emissions, as
described in Section 3.2.
Emissions trends have shown that industrial activities are now one of the largest sources of 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. In soil samples
obtained near a defunct smelter in El Paso, TX, in 1999 and 2005, Pingatore et al. (2009) found that TSP
concentration was predicted strongly by concentrations of Pb-humate, which is created by sorption of Pb
onto humic substances in soil. 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.
Table 3-9. Outdoor soil Pb levels in various cities within the U.S.
Study Year n iVMn" Meda Max"
Background soil Pb, U.S. 2001 1,319 16j>
City-State
Los Angeles, California 2010 550 9 216,174
Los Angeles, California 1995 343
Chicago Illinois 2008 57
Chicago Illinois 1987 667
Chicago-Urban Parks 1986 255 12 262 1,312
Chicago-Suburban Parks 1986 245 12 87 1,637
Illinois, Rural Parks 1986 177 12 37 937
Detroit, Michigan 2003 59 13 189 1,345
Detroit-Suburbs, Michigan 2003 76 4 16 810
Pontiac, Michigan 2003 38 15 86 495
Oakland, California 1995 358 7 347,900
Alameda, California 1993 138 22 3,187
Boston, Massachusetts 1988 195 7 13,240
Miami, Florida 2004 240 2 1,060
Seattle, Washington 1991 51 150 74,000
Washington, D.C. 1995 240 12 6,015
Minneapolis/St. Paul, Minnesota 1984 90 5 7,650
Minneapolis, Minnesota 1988 898 1 230 20,136
St. Pau, Minnesota 1988 832 1 170 7,994
Duluth, Minnesota 1988 229 1 144 11,110
St. Cloud, Minnesota 1988 124 1 4] 1,952
Rochester, Minnesota 1988 165 1 25 1,930
Outstate farms Minnesota 1988 781 1 31 7,111
Cleveland, Ohio 2006 50 19 811
Baltimore, Maryland 2008 122 OX)] 5,620
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Study Year
Min
Med
Max*
Baltimore, Maryland
1983
422
100
10,900
Indianapolis, Indiana
2005
116
46
565
Milwaukee, Wisconsin summary
data
1994
372
160
880
Milwaukee, Wisconsin cent city +
North and South Side
1994
256
240
7,220
Milwaukee, Wisconsin suburbs
1994
122
50
1,780
Cincinnati, Ohio
1990
60
3,166
Cincinnati, Ohio-Childcare centers 2008
69
17
4,636
Cincinnati, Ohio
1990
60
Sacramento, California
1995
232
57
229
320,834
Tampa, Florida
1994
146
<1
100
9,160
New Orleans, Louisiana Survey 1 2005
4,026
18
134
183,588
New Orleans, Louisiana Survey 2 2005
5,467
100
52,798
Louisiana Orleans Parish
1997
~1,540
<25
>200
<1,000
Louisiana Lafourche Parish
1997
-190
<25
Albuquerque, New Mexico
1981
43
5,280
Omaha, Nebraska
1979
176
Dayton, Ohio
1983
22
22
461
El Paso, Texas
2009
94
<20
8700
Honolulu, Hawaii
1988
18
Charleston, South Carolina
1975
164
7,890
New Haven, Connecticut
1982
487
30
7,000
Corpus Christi, Texas
1987
485
21
2,969
Pueblo, Colorado
2006
33
Connecticut
2008
174
<10
2,200
Gainesville, Florida
2004
202
2.13
1,091
Champaign Illinois
1976
116
20
1,061
Louisiana and Minnesota
1993
6,342
Urban-rural comparisons
Southeastern. Michigan
2004
171
7,400
Mt. Pleasant, Michigan
1992
189
100
16,839
Syracuse, New York
2009
2,998
45
Syracuse, New York
2002
162
Syracuse, New York
2002
194
80 (GM)
Lubbock, Texas
2008
52
35
Maine urban soils
1989
100
aMinimum, median, and maximum values are reported in units of mg/kg for individual cities cited in the review paper.
Source: Used with permission from Elsevier Publishing, Mielke et al. (Mielke et al.. 2010b').
Several studies explore the relationship between soil Pb concentration and land use. For example,
the Mielke et al. (2010b) review also found that soil Pb concentrations tended to be higher within inner-
city communities compared with neighborhoods surrounding city outskirts. 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 overtime. 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 & 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 over time from resuspension and deposition with contributions from historic
sources of on-road gasoline and Pb paint. 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,
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LA in Figure 3-27. 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 (mgfkg)
Source: Used with permission from Elsevier Publishing, Mielke et al. (2007)
Figure 3-27. Map of median Pb content in soil in New Orleans.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.
The amount of Pb within the inner-city is likely not from a single source but instead composed of
all modern and historic sources of Pb dust that have been released in the city including Pb from several
local industries, Pb dust from pulverized wheel weights, deteriorated Pb-based paint, Pb additives to on-
road gasoline, and defunct incinerators that once dotted New Orleans prior to being shut down by EPA in
the early 1970's. Similarly, 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. (2007) observed median and maxima deposition values of roughly 25,000 and
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265,000 |ag/m2. respectively. Median and maxima surface soil Pb concentrations were observed to be
1,000 and 20,000 mg/kg, respectively. Clark et al. (2006) performed isotopic analysis on urban garden
soils in Boston, MA 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% could be attributed to historic Pb on-road
gasoline emissions in a suburban area. The remainder of the Pb was attributed to paint degradation.
Several studies have examined the effects of roadways on Pb content in roadside dust. 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 age of the land parcel (square-root transformed), 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. Amato et al. (2009) observed that
deposited PMi0 onto roadways, measured as dust samples, in Barcelona, Spain was differentially enriched
with Pb. Pb concentration in PMi0 was highest at ring roads (229 ppm) and in the city center (225 ppm),
followed by demolition and construction sites (177 ppm) and near a harbor (100 ppm). Joshi et al. (2008)
also observed Pb dust concentrations to be highest at industrial sites followed by commercial and
residential sites 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.. 2006;
Weiss et al.. 2006). Caravanos et al. (2006) 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. (2003. 2005)
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
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increase by a factor of 200% for streets (87,000 (ig/m2), 138% for alleys (65,000 (.ig/ni2). and 26% for
sidewalks (23,000 (.ig/ni2) compared with pre-demolition Pb dust levels. At another demolition site,
smaller increases were observed: 29% for streets (29,000 (ig/m2), 18% for alleys (19,000 (.ig/ni2) and 18%
for sidewalks (22,000 (.ig/nr) (Farfel et al.. 2005V
Pb can be present 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.
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 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. These results suggest that soil Pb
concentration tends to be spatially heterogeneous in the absence of a source. 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.
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. Although, Pb was measured in snow, water, sediment, lichen
and fish during the multiyear project, this metal was not quantified in air samples.
In the WACAP study, bioaccumulation of airborne contaminants was demonstrated on a regional
scale in remote ecosystems in the Western U.S. Contaminants were shown to accumulate geographically
based on proximity to individual sources or source areas, primarily agriculture and industry. This finding
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was counter to the original working hypothesis that most of the contaminants found in western parks
would originate from eastern Europe and Asia.
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. A clear decline in Pb concentrations in sediments after the discontinued use of leaded on-road
gasoline was observed at almost all WACAP locations in the for nearly all WACAP sites in the
conterminous 48 states. 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. Pb was also consistently observed in WACAP
fish and wildlife samples.
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 3-10 presents data from
seven metropolitan areas (Cobb et al.. 2006). Differences among the intraurban concentration ranges
illustrate a high level of spatial variability within individual cities as well as high interurban variability.
The rural New Orleans site reported relatively low Pb sediment concentrations, and the highest Pb
sediment concentrations were reported for the city of New Orleans. Figure 3-28 and Figure 3-29 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. The data also illustrated that Pb
concentrations in sediment have declined in the U.S. since 1975 (Figure 3-29), prior to the phase-out of
on-road leaded gasoline.
Table 3-10. Sediment concentrations in various cities, prior to 2005
City
Avg Pb Concentration (nig/kg)"
Pb Concentration Range (nig/kg)"
Baltimore, MD
1-10,900
Miami, FL
275
25-1612
Mt. Pleasant, Ml
320
100-840
New Orleans, LA
784
31.7-5195
New Orleans, LA (rural outskirts)
11
4.8-17.3
St. Louis, MO
427
35-1860
Syracuse, NY
80
20-800
aDry weight basis.
Source: Used with permission from the American Chemical Society, Cobb et al. (2006).
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X X
Downstream
100 200 BOO 400 500 600
River km above Apalachicola Bay, FL
~ Pb in stream bed-sediment and reservoir-core samples
X Pb background in streambed-sediment and baseline reservoir-core samples
Source: Used with permission from the American Chemical Society, Callender and Rice (2000).
Figure 3-28. 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.Note that 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.
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160
140
120
"So
Is 100
E.
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0
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u
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0
0
Source: Used with permission from the American Chemical Society, Callender and Rice (2000).
Figure 3-29. 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.Note that 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.
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 atypical 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,
1980-1985
A -1985-1990
—X ¦ 1990-1995
Q)
O
>
Downstream
100 ZOO 300 400 500 600 700 S00
River km above Apalachicola Bav, FL
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several studies noted no appreciable increase in Pb concentration within Lake Pontchartrain soils and
sediments (Abel et al.. 2010; Abel et al. 2007; Cobb et al.. 2006; Preslev et al.. 2006; K. J. Schwab et al..
2007). 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
Recent results from locations outside the United States 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. 2006"). volume weighted Pb concentrations in
precipitation collected in 1993-94 from Lake Superior, Lake Michigan and Lake Erie ranged from -0.7 to
-1.1 (ig/L (Sweet et al.. 1998). These values fit well with the temporal trend reported in Watmough and
Dillon (2007). who calculated annual volume-weighted Pb concentrations to be 2.12, 1.17 and 0.58 (ig/L
for 1989-90, 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 & Krachler. 2010).
There have also been a few recently published, long-term European studies of Pb concentration in
precipitation including Berg et al. (2008) and Farmer et al. (2010). Berg 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 3-30).
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
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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; see Figure 3-30.
O)
3
n
D.
—•— Birkenes
K3rvatn
Jergul/Karasjok
— ¦ Osen
- Lista
—
\ \
\
^ VA*-
\ /*\ \ V * -
X ' * V * •
10
9
8
7
6
5
4
3
2
1
0
1980 1985 1990 1995 2000 2005
Source: Used with permission from Pergamon Press, Berg et al. (2008)
Figure 3-30. Trends in Pb concentration in precipitation from various sites
in Norway over the period 1980-2005.
Farmer et al. (2010) showed the trends in concentration of Pb in precipitation collected in a remote
part of northeastern Scotland over the period 1989-2007. The 2.6- and 3.0-fold decline in mean
concentration from 4.92 (ig/L (1989-1991) to 1.88 (ig/L (1999) and then to 0.63 (ig/L (2006-2007) is
qualitatively but not quantitatively in line with the sixfold 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. 2006). 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. 2006) 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
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undergo further transport more easily when snow melts. Deposition to snow was investigated in several
studies. Seasonal patterns of heavy metal deposition to snow on Lambert Glacier basin, east Antarctica,
were determined by Hur et al. (2007). The snow pit samples covered the period from austral spring 1998
to summer 2002 and Pb concentrations ranged from 1.29-9.6 pg/g with a mean value of 4.0 pg/g. This
was similar to a mean value of 4.7 pg/g (1965-1986) obtained by Planchon et al. (2003) for Coats Land,
northwest Antarctica. Estimated contributions to the Pb in Lambert Glacier basin snow were -1% from
rock and soil dust (based on Al concentrations) and -4.6% from volcanoes (based on the 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.
Lee et al. (2008) collected 42 snow samples during the period autumn 2004-summer 2005 from a
2.1m snow pit at a high-altitude site on the northeast slope of Mount Everest, Himalayas. Pb
concentrations ranged from 5-530 pg/g with a mean value of 77 pg/g. The mean value is clearly higher
than the Hur et al. (2007) value for Antarctica but is substantially lower than a mean concentration of 573
pg/g for snow from Mont Blanc, France (1990-1991) (collated in 2008). The mean Pb concentration for
Mount Everest snow was lower during the monsoon (28 pg/g) compared with the non-monsoon periods
(137 pg/g). From calculated enrichment factors (Pb/Alsnow:Pb/Alcrust), anthropogenic inputs of Pb were
partly important but soil and rock dust also contributed. The low Pb concentrations during monsoon
periods are thought to be attributable to low levels of atmospheric loadings of crustal dusts. K. Lee et al.
(2008) noted that their conclusions 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. (2008) suggested that these Pb concentrations were much higher than expected and
that their snow samples may have suffered from contamination during sampling and analysis.
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3.6.5. Natural Waters
Shotyk and Krachler (2007) measured 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 & 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). 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 and that for
contemporary snow was 672 (Johnson, n = 6; Parnell, n = 3) ng/L. Shotyk et al. (2010) gave 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. 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 that the removal process is very effective. Shotyk and Krachler
(2010) speculate that even at these very low Pb concentrations, much if not most of the Pb is likely to be
colloidal, as suggested by the 2006 Pb AQCD (U.S. EPA. 2006). Finally, Shotyk et al. (2010) suggest that
the pristine groundwaters from Simcoe County, Canada, provide a useful reference level against which
other water samples can be compared.
Although Pb concentrations in Kawagama Lake water were approaching "natural values," the
206Pb/207Pb ratios for the samples that had the lowest dissolved Pb concentrations of 10, 10 and 6 ng/L
were 1.16, 1.15 and 1.16, respectively. These values are far from those expected for natural Pb (the clay
fraction from the lake sediments dating from the pre-industrial period had values of 1.19-1.21) and it was
concluded that most of the dissolved Pb in the lake water was of industrial origin. Shotyk and Krachler
(2010) found that the full range of isotope ratios for Kawagama Lake water samples (Ontario, Canada)
was 1.09 to 1.15; this was not only much lower than the stream water values entering the lake but also
lower than the values attributed to leaded on-road gasoline in Canada (-1.15). The streamwater ratio
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values were -1.16 to 1.17 whilst 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.
3.6.6. Moss
Mosses can be used effectively for monitoring trends in Pb deposition as demonstrated in many
studies (Harmons et al.. 2008; Harmens et al.. 2010). For example, Harmens et al. (2008) showed that a
52% decrease in deposited Pb concentrations corresponded to a 57% decrease in Pb concentrations in
moss. Farmer et al. (2010) showed that there was good agreement between the 206Pb/207Pb ratio for
precipitation and mosses collected in northeast Scotland. A study in the Vosges mountains also found a
ratio value of 1.158 for a moss sample and a surface soil litter value of 1.167 and concluded that 1.158 to
1.167 represented the current atmospheric baseline (Geagea et al.. 2008). For rural northeast Scotland, a
combination of sources is giving rise to a 206Pb/207Pb ratio of-1.15 in recent precipitation and mosses
(Farmer et al.. 2010). Clearly, sources with a lower ratio than coal (-1.20) must be contributing
substantially to the overall emissions. Pb from waste incineration has been implicated as a possible
current source (cf. typical 206Pb/207Pb ratios for Pb from European incineration plants are -1.14 to 1.15 (de
la Cruz et al. (2009) and references therein).
3.6.7. Grass, Foliage, and Tree Rings
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 over time, 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. (2008) 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; see
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1 Figure 3-31. Both Pb concentrations and Pb isotope ratios declined with distance from the smelter (Aznar.
2 Richer-Lafleche. et al.. 2008; Aznar. Richer-Lafleche. et al.. 2008).
Trends
800
700
600
0)
I 500
cd 400
300
200
100
Humus
(11)
~ (41)
A (29)
q (21)
2002
0.8
0.6
04
0.2
Regional
pollution
Smelter
1950 1960 1970 1980 1990 2000
Tree rings
Distance class (km)
10-20
20-40
1 1 1
1950 1960 1970
—I
1980
1990 2000
Sapwood-h ea rtwood
boundary
Source: Used with permission from Elsevier Publishing, Aznar et al. (2008)
Figure 3-31. 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.
3.6.8. Aquatic Bivalves
3 Data from invertebrate waterborne populations can serve as in indicator of Pb contamination
4 because animals such as mussels and oysters take in contaminants during filter feeding. Kimbrough et al.
5 (2008) surveyed Pb concentrations in mussels, zebra mussels, and oysters along the coastlines of the
6 continental U.S. In general, they observed the highest concentrations of Pb in the vicinity of urban and
7 industrial areas. Company et al. (2008) measured Pb concentrations and Pb isotope ratios in bivalves
8 along the Guadiana River separating Spain and Portugal. Analysis of Pb isotope ratio data suggested that
9 high Pb concentrations were related to historical mining activities in the region. Elevated Pb
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concentrations were also observed by Company et al. (2008) in the vicinity of more populated areas.
Couture et al. (2010) report data from a survey of the isotopic ratios of Pb 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 resuspension of contaminated sediments enriched with Pb runoff from wastewater
treatment plants, municipal waste incinerators, smelters and refineries rather than from atmospheric
deposition (Couture et al.. 2010).
3.7. Summary
3.7.1. Sources of Atmospheric Lead
The 2006 Pb AQCD (U.S. EPA. 2006) 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
1990 and then by an additional 77% from 1990 to 2008, at which time emissions were 1,200 tons/yr.
Industrial processes, including metals processing and industrial fuel combustion, had replaced mobile
sources as the primary source of Pb to the atmosphere by the 2006 Pb AQCD (U.S. EPA. 2006). More
recent data from the 2008 NEI (U.S. EPA. 2011) illustrate that piston engine aircraft emissions now
comprise the largest share (-49%) of total atmospheric Pb emissions; the 2008 NEI (U.S. EPA. 2011)
estimated that 590 tons of Pb were emitted from aircraft point sources. Other sources of ambient air Pb, in
approximate order of importance, include metals processing, fossil fuel combustion, other industrial
sources, roadway related sources, and historic Pb.
Chemical speciation of Pb had been fairly well characterized in the 2006 Pb AQCD (U.S. EPA.
2006). Estimates from the 1986 Pb AQCD (U.S. EPA. 1986) for organic on-road Pb emissions provides
an upper bound for organic vapor emissions of 20% of total Pb dibromide and Pb bromide emissions from
piston engine aircraft. Recent speciation studies of smelting and battery recycling operations have shown
that PbS and Pb sulfates are abundant within the emissions mixture for such industrial operations.
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. Global atmospheric Pb deposition peaked in the
1970s, followed by a more recent decline. On a local scale, Pb concentrations in soils (including urban
areas where historic use was widespread) can be substantial, and coarse Pb-bearing PM experiences
cycles of deposition and resuspension that serve to distribute it. Both wet and dry deposition are important
removal mechanisms for atmospheric Pb. Because Pb in fine PM is typically fairly soluble, wet
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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. Resuspension by wind and traffic can be an important source of airborne Pb near sources
where Pb occurs in substantial amounts in surface dust.
In water, Pb is transported as free ions, soluble chelates, or on surfaces of iron and organic rich
colloids, and water columns behave as important reservoirs of Pb. 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 removal by sulfides is particularly important. 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, and sediment concentrations tend to follow those in overlying
waters. Resuspended Pb is largely 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 than atmospheric deposition. Resuspension and release from sediments largely
occurs during discrete events related to storms.
A complex variety of factors that 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, leading to high soil concentrations in those locations, while fine Pb-PM can be transported long
distances, leading to contamination of remote areas. Surface waters act as an important reservoir, with Pb
lifetimes largely controlled by deposition and resuspension of Pb in sediments. Pb retention in soil
depends on Pb speciation and a variety of factors intrinsic to the soil.
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3.7.3. Ambient Lead Monitoring
In recognition of the role of all PM sizes in ambient air Pb exposures, including the ingestion of
particles deposited onto surfaces, the indicator for the Pb NAAQS is Pb in Pb-TSP. Although there is a
lower rate of error in estimating ambient Pb from Pb-PM10 monitoring than from Pb-TSP monitoring, the
Pb-TSP indicator was retained in 2008 because ingestion after deposition in the upper respiratory tract
was considered an important component of Pb exposure. A new FRM for Pb-PM10 has been implemented
in which ambient air is drawn through an inertial PM size separator for collection on a PTFE filter.
Several FEMs have also been approved. The FRM is based on flame AAS. ICPMS is under consideration
as a new FRM for Pb-TSP.
Monitoring for ambient Pb levels is required for all areas where Pb levels have been shown or are
expected to contribute to maximum concentrations of 0.10 |ig/m3 or greater over a three-year time period.
Pb is monitored routinely at SLAMS that report data used for NAAQS compliance to the AQS database.
Pb monitoring requirements have experienced several changes since publication of the 2006 Pb AQCD
("U.S. EPA. 2006). In addition to FRM monitoring, Pb is also routinely measured in smaller PM fractions
in the CSN, IMPROVE, and the NATTS networks, and is planned for the NCore network. 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. Monitors
reporting to the AQS were considered for the purpose of this ISA 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/yr or greater source,
as noted in the 2005 NEI (U.S. EPA. 2008a'). Non-source oriented monitors were those monitors not
considered to be source oriented based on these two criteria.
3.7.4. Ambient Air Lead Concentrations
Ambient air Pb concentrations have declined drastically over the period 1980-2009. The median
annual concentrations for all monitors have dropped by 97% from 0.87 |ig/m3 in 1980 to 0.025 |ig/m3 in
2009. While the sharpest drop in Pb concentration occurred during 1980-1990, a declining trend was
observed between 1990 and 2009. A smaller reduction was observable among source oriented Pb
concentration (56%) and non-source oriented Pb data (51%) for 2000-2009.
AQS data for source oriented and non-source oriented monitoring were analyzed for 2007-2009.
For source oriented monitoring, the 3-month rolling average was measured to be above the level of the
NAAQS in 14 counties across the U.S. Fourteen monitoring sites had maximum 3-month rolling average
values that exceeded the level of the NAAQS. The maximum 3-month rolling average concentrations
ranged from 0.17-2.9 |ig/nr\
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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. PM size distribution also influenced how far the particle
will travel upon initial emission or resuspension before being deposited.
Size distribution of Pb-bearing PM was demonstrated to vary substantially for several studies
presented, depending on the nature of Pb sources and proximity of the monitors to the Pb sources. AQS
data were also used to estimate the size distribution of Pb-bearing PM at several sites with co-located Pb
monitors. On average, Pb-TSP and Pb-PMi0 were moderately correlated, but the correlation improved for
urban and city center land use types compared with all data. When comparing Pb-TSP with Pb-PM2 5,
correlations were lower in urban and city center areas compared with suburban and rural sites. A
relationship between land use type and correlation was less obvious when comparing Pb-PMi0 with Pb-
PM2 5. For urban and city center types, average p was fairly high. Average p increased for suburban sites
but decreased for rural sites. Variation in correlation of size-fractionated Pb samples among different land
use types may reflect differences among sources among land use types.
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.
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. After the phase-out of Pb from on-road gasoline,
concentrations in these media decreased to varying degrees. In rain, snowpack, and surface waters, Pb
concentrations have decreased considerably following elimination of leaded on-road gasoline. 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
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1 concentrations. In remote lakes, sediment profiles indicate higher Pb concentrations in near surface
2 sediment as compared to pre-industrial era sediment from greater depth and indicate peak concentrations
3 between 1960 and 1980, when leaded on-road gasoline was at peak use. Concentrations of moss, lichens,
4 peat, and aquatic bivalves have been used to understand spatial and temporal distribution patterns of air
5 Pb concentrations. Ingestion and water intake are the major routes of Pb exposure for aquatic organisms,
6 and food, drinking water, and inhalation are major routes of exposure for livestock and terrestrial wildlife.
7 Overall, Pb concentrations have decreased substantially in media through which Pb is rapidly transported,
8 such as air and water. Substantial Pb remains in soil and sediment sinks. Although in areas less affected
9 by major local sources, the highest concentrations are below the surface layers and reflect the phase-out
10 of Pb from on-road gasoline and emissions reductions from other sources.
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Chapter 3 Appendix
3.8. Variability across the U.S.
Table 3A-1.Distribution of 1-month average Pb-TSP concentrations (pg/m3) nationwide,
source-oriented monitors, 2007-9.Sites listed in the bottom six rows of the
table fall in the upper 90th percentile of the data pooled by site.
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
2007-
2009
i
i
1,802
0.2353
o Jo
0.007
0.012
0.028
0.070
0.254|0.670
0.976
2.061
4.440
2007
554
0.2500
0
0
0.008
0.012
0.028
0.070
0.290
0.706
1.085
1.982
3.620
2008
622
0.2803
0
0
0.008
0.013
0.034
0.080
0.300
0.741
1.164
2.501
4.440
2009
626
0.1774
0
0
0.006
0.010
0.023
0.063
0.194
0.535
0.787
1.280
2.438
Winter
443
0.2324
0
0
0.006
0.010
0.025
0.064
0.224
0.575
1.111
2.440
3.103
Spring
447
0.2695
0
0
0.008
0.013
0.031
0.077
0.333
0.794
1.08512.035
3.123
Summer
458
0.2142
0
0
0.008
0.013
0.032
0.074
0.265
0.652
0.85211.384
4.440
Fall
I 454
0.2260
0
0
0.007
0.012
0.026
0.064
0.226
0.670
1.00611.565
4.225
Nationwide statistics, pooled by site
2007-
2009
56
0.2384
0.001
0.001
0.010
0.014
0.031
0.098
0.350
0.730
0.862
1.306
1.306
2007
48
0.2547
0.000
0.000
0.010
0.014
0.031
0.090
0.323
0.783
1.048
1.520
1.520
2008
53
0.2848
0.001
0.001
0.011
0.021
0.032
0.121
0.344
0.818
1.226
1.542
1.542
2009
53
0.1779
0.000
0.000
0.007
0.011
0.029
0.075
0.219
0.453
0.718
0.917
0.917
Winter
56
0.2353
0.001
0.001
0.008
0.010
0.033
0.088
0.290
0.654
1.205
1.389
1.389
Spring
56
0.2759
0.000
0.000
0.008
0.017
0.035
0.093
0.425
0.863
1.119
1.605
1.605
Summer
56
0.2136
0.000
0.000
0.008
0.016
0.034
0.095
0.289
0.654
0.802
1.170
1.170
Fall
55
0.2286
0.001
0.001
0.012
0.017
0.029
0.071
0.389
0.651
1.031
1.243
1.243
Statistics for individual counties (2007-2009)
I 101109 j AL I Pike I I 441 210.511010.07010.07010.08310.09410.19010.35510.71011.27711.35611.43411.434
j 106037 |CA |Los Anqeles | ! 108! 510.242510.00210.00210.00810.01210.03710.07310.18810.69810.94212.50112.880
12057
FL
Hillsborough
i 62
2
0.2939
0.014
0.014
0.020
0.040,0.080
0.13910.333
0.600
0.84013.620
3.620
17031
IL
Cook
! 144
4
0.0241
0.010
0.010
0.010
0.010l0.014
0.022) 0.030
0.040
0.04610.070
0.084
17119
IL
Madison
36
1
0.1088
0.022
0.022
0.024
0.028
0.037
0.058
0.138
0.300
0.325
0.454
0.454
17163
IL
Saint Clair
36
1
0.0246
0.010
0.010
0.010
0.013
0.015
0.023
0.030
0.045
0.054
0.058
0.058
18035
IN
Delaware
72
2
0.3318
0.034
0.034
0.045
0.056
0.094
0.202
0.341
0.639
0.980
4.440
4.440
| 118089 lIN |Lake | | 361 110.0297|0.004|0.004|0.006|0.007|0.017|0.025|0.045|0.060|0.063|0.065|0.065
! j 18097 |lN iMarion ! ! 71! 210.019810.00310.00310.00510.00710.01010.01610.02810.04010.04610.05710.057
I 127037 |MN
Dakota I
35
110.2872
0.062
0.06210.072
0.098
0.13410.232
0.412
0.62010.670
0.73010.730
27163
MN
Washington
31
1
0.0006
0.000
0.000
0.000
0.000
0.000
0.000
0.001
0.003
0.003
0.004
0.004
29093
MO
Iron
144
4
0.4060
0.007
0.010
0.021
0.029
0.045
0.094
0.658
0.971
1.437
2.557
4.225
29099
MO
Jefferson
296
10
0.6096
0.011
0.019
0.071
0.124
0.212
0.454
0.798
1.353
1.764
2.440
3.123
29189
MO
Saint Louis
36
1
0.0361
0.005
0.005
0.005
0.005
0.008
0.050
0.050
0.050
0.050
0.066
0.066
34023
NJ
Middlesex
12
1
0.0101
0.007
0.007
0.007
0.007
0.008
0.008
0.008
0.010
0.034
0.034
0.034
36071
NY
Orange
106
3
0.0271
0.003
0.003
0.004
0.005
0.007
0.020
0.037
0.058
0.079
0.103
0.134
39035
OH
Cuyahoga
108
3
0.0473
0.004
0.004
0.007
0.008
0.013
0.024
0.060
0.130
0.190
0.210
0.220
39051
OH
Fulton
33
1
0.2434
0.015
0.015
0.055
0.066
0.093
0.190
0.360
0.510
0.620
0.690
0.690
39091
OH
Logan
68
2
0.0763
0.020
0.020
0.020
0.030
0.040
0.070
0.090
0.120
0.190
0.290
0.290
42007
PA
Beaver
32
1
0.1090
0.045
0.045
0.054
0.063
0.070
0.100
0.132
0.173
0.195
0.284
0.284
42011
PA
Berks
108
3
0.1053
0.030
0.033
0.037
0.040
0.050
0.073
0.131
0.242
0.302
0.360
0.542
47163
TN
Sullivan
108
3
0.0754
0.030
0.030
0.032
0.036
0.042
0.062
0.087
0.145
0.178
0.254
0.341
48085
TX
Collin
108
3
0.2910
0.007
0.009
0.030
0.052
0.102
0.186
0.389
0.673
0.904
1.121
1.564
Statistics for individual sites where overall average monthly mean > national 90th percentile (2007-2009)
1060371405! 24] j0.8623 j 0.242 j0.24210.283 j0.28410.321 j 0.60610.923 j 2.27712.50112.880|2.880
12909300161 36] |0.730210.16610.16610.18610.23510.34710.54510.837 |l-29512.43514.22514.225
1290930021! 36| I0.7970l0.084l0.084l0.093l0.099l0.409l0.696l0.967ll.453l2.438l2.557l2.557
i i i | i290990004
36
11.117710.242
0.242 (0.285
0.518
0.76411.005
1.519
1.905(2.101
2.416 (2.416
I I I I 1290990015
34
11.306010.155
0.155l0.339
0.421
0.83411.185
1.577
2.31913.103
3.12313.123
! ! ! I 1290990021
33
10.7498 i 0.084
0.08410.141
0.423
0.550! 0.681
0.901
1.16411.553
2.162i2.162
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Table 3A-2.Distribution of 3-month moving average Pb-TSP concentrations (jjg/m3)
nationwide, source-oriented monitors, 2007-9. Sites listed in the bottom six
rows of the table fall in the upper 90th percentile of the data pooled by site.
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
20079
1,737
0.236610.000
0.000
0.010
0.010
0.030
0.080
0.270
0.680
0.980
1.870
2.890
2007
507
0.2421
0.000
0.000
0.010
0.020
0.030
0.080
0.270
0.720
1.070
1.640
1.740
2008
612
0.2834
0.000
0.000
0.010
0.010
0.040
0.090
0.320
0.800
1.150
2.290
2.890
2009
618
0.1859
0.000
0.000
0.010
0.010
0.030
0.070
0.240
0.550
0.760
1.150
2.070
Winter
417
0.2434
0.000
0.000
0.010
0.010
0.030
0.070
0.270
0.660
1.120
2.080
2.620
Spring | |
436
I0.2638
0.000
0.000!0.010
0.010
0.030 i 0.080
0.330
0.740! 1.080
2.140 !2.890
jSummeri j j j j 448! |0.2269|0.000|0.000|0.010|0.020|0.030|0.080|0.290|0.670|0.900|1.730I2.140
Fall I ! ! ! ! 436! io.2131! 0.00010.00010.01010.02010.03010.080! 0.240! 0.650! 0.940! 1.32012.460
Nationwide statistics, pooled by site
200791 III I I 56| |0.2386|0.000|0.00010.01110.01710.03210.09910.33310.73310.93011.33211.332
2007 i
i 45i
0.2435
0.003
0.003
0.016
0.019,0.033
0.084|0.277 iO.687
0.83111.524
1.524
2008 |
! 53!
0.2903
0.000
0.000
0.010
0.022 !o.034
0.119 !0.383!0.806
1.301 h. 641
1.641
2009
53
0.1848
0.000
0.000
0.009
0.014
0.029
0.075
0.229
0.617
0.855
0.863
0.863
Winter
56
0.2482
0.000
0.000
0.010
0.015
0.041
0.104
0.316
0.683
1.150
1.308
1.308
Spring
56
0.2736
0.000
0.000
0.010
0.014
0.032
0.096
0.369
0.820
1.320
1.679
1.679
|Summer| | | | | 56| |0.224310.00010.00010.01010.01210.034|0.09310.28710.70710.90011.23211.232
Fall ! ! ! ! ! 55| |0.2132|0.000|0.000|0.011|0.017|0.032|0.085|0.341|0.592|0.960|1.164|1.164
Statistics for individual counties (2007-2009)
I 101109 | AL | Pike | | 44| 210.508010.10010.10010.12010.23010.27510.32510.68511.05011.15011.35011.350
06037
CA | Los Angeles
I 1081 5
0.2594 !0.010|0.010
0.010!0.01010.045
0.075 i 0.185 i 0.710
1.26012.450
2.490
12057
FL
Hillsborough
48
2
0.2248
0.040
0.040
0.040
0.040
0.095
0.135
0.250
0.490
0.580
1.770
1.770
17031
IL
Cook
144
4
0.0244
0.010
0.010
0.010
0.010
0.020
0.030
0.030
0.040
0.040
0.050
0.050
17119
IL
Madison
36
1
0.1144
0.040
0.040
0.040
0.040
0.060
0.105
0.155
0.210
0.230
0.280
0.280
17163
IL
Saint Clair
36
1
0.0250
0.010
0.010
0.010
0.020
0.020
0.020
0.030
0.040
0.040
0.050
0.050
18035
IN
Delaware
72
2
0.3422
0.050
0.050
0.070
0.080
0.120
0.210
0.430
0.710
0.930
2.140
2.140
18089
IN
Lake
36
1
0.0314
0.010
0.010
0.010
0.010
0.020
0.030
0.040
0.050
0.050
0.060
0.060
18097
IN
Marion
69
2
0.0199
0.000
0.000
0.010
0.010
0.010
0.020
0.030
0.030
0.040
0.050
0.050
27037
MN
Dakota
33
1
0.2742
0.100
0.100
0.120
0.140
0.210
0.240
0.350
0.400
0.470
0.570
0.570
27163
MN
Washington
24
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
29093
MO
Iron
144
4
0.4142
0.020
0.020
0.020
0.030
0.050
0.170
0.735
1.010
1.210
2.080
2.460
29099
MO
Jefferson
280
10
0.6211
0.040
0.070
0.120
0.155
0.280
0.475
0.835
1.250
1.655
2.620
2.890
29189
MO
Saint Louis
36
1
0.0383
0.010
0.010
0.010
0.010
0.010
0.050
0.050
0.050
0.050
0.060
0.060
34023
NJ
Middlesex
12
1
0.0192
0.010
0.010
0.010
0.010
0.010
0.010
0.020
0.050
0.060
0.060
0.060
136071 | NY lOrange |
102
3i0.0273i0.000
O.OOOiO.OOO
0.010
0.010 i 0.030
0.030
0.060 j 0.070
0.080 !0.090
| 139035 IOH jCuyahoqa I | 102| 3iO.048310.010]0.01010.01010.01010.020i0.03010.070]0.120i0.14010.160iO. 170
! ! 39051 iOH ! Fulton ! ! 25! 1 !o.2420! 0.050 !o.050 !0.050 !o.080 !o. 130! 0.240! 0.320! 0.450! 0.520 !o.570 !o.570
39091
OH
Logan
i 68
2
0.0809
0.030
0.030
0.030 i0.03010.050
0.07010.090
0.160
0.19010.260
0.260
42007
PA
Beaver
! 32
1
0.1115
0.060
0.060
0.070|o.080!o.090
0.110 S0.120
0.160
0.17010.170
0.170
42011
PA
Berks
108
3
0.1099
0.040
0.040
0.040
0.050
0.060
0.085
0.150
0.210
0.250
0.350
0.360
47163
TN
Sullivan
102
3
0.0769
0.030
0.030
0.040
0.040
0.050
0.060
0.090
0.130
0.180
0.210
0.260
48085
TX
Collin
108
3
0.2873
0.010
0.030
0.060
0.070
0.110
0.205
0.430
0.590
0.650
1.190
1.260
Statistics for individual sites where overall average monthly mean > national 90th percentile (2007-2009)
I I I I 10603714051 24| 10.930410.28010.28010.31010.33010.47010.66511.04512.29012.45012.49012.490
III! |2909300161 361 |0.732810.21010.21010.26010.38010.44510.57010.770 |l ¦ 17012.08012.46012.460
! ! ! ! J290930021! 36! !o.8258 ! 0.220 !o.220 !o.220 !o.300 !o.545 ! 0.880! 1.005! 1.210! 1.270! 1.940! 1.940
I | | 1290990004
36
11.138310.630
0.63010.680
0.700
0.81011.085
1.340
1.71012.130
2.140 i2.140
1290990015
34
11.3318
0.610
0.610 !0.640
0.730
0.94011.195
1.710
2.02012.630
2.89012.890
i i
S290990021
33
iO.7682
0.430
0.430! 0.440
0.520
0.61010.750
0.900
1.090) 1.310
1.32011.320
May 2011
3-130
DRAFT - DO NOT CITE OR QUOTE
-------
Table 3A-3.Distribution of annual 1-month site maxima TSP Pb concentrations (jjg/m3)
nationwide, source-oriented monitors, 2007-2009.Sites listed in the bottom
eight rows of the table fall in the upper 90th percentile of the data
pooled by site.
Year
Site ID-year
N (sites)
Mean
Min
1
5
10
25
50
75
90
95
99
max
Nationwide statistics
2007-2009
56
0.7961
0.004
0.004
0.025
0.040
0.076
0.289
0.747
2.557
3.620
4.440
4.440
2007
48
0.5517
0.003
0.003
0.025
0.035
0.054
0.192
0.735
1.565
2.162
3.620
3.620
2008
53
0.7157
0.004
0.004
0.014
0.039
0.059
0.247
0.754
2.416
3.123
4.440
4.440
2009
53
0.3758
0.003
0.003
0.016
0.019
0.065
0.141
0.536
1.124
1.357
2.438
2.438
Annual site max 3-month means >= national 90th percentile (2007-2009)
| 180350009-2008
4.4400
290930016-2008
4.2252
120571066-2007
3.6200
290990015-2008
3.1228
! 060371405-2008
2.8800
!
!
290930021-2008
i 2.5566
I I I I
180350009-2008
4.4400
290930016-2008
4.2252
Table 3A-4.Distribution of annual 3-month site maxima Pb-TSP concentrations (jjg/m3)
nationwide, source-oriented monitors, 2007-2009.Sites listed in the bottom
nine rows of the table fall in the upper 90th percentile of the data
pooled by site.
Year
Site ID-year
N (sites)
Mean
Min
1
5
10
25
50
75
90
95
99
max
Nationwide statistics
2007-2009
56
0.5409
0.000
0.000
0.020
0.030
0.060
0.215
0.590
1.940
2.460
2.890
2.890
2007
45
0.3616
0.010
0.010
0.030
0.030
0.050
0.130
0.520
1.210
1.350
1.740
1.740
2008
53
0.5177
0.000
0.000
0.010
0.030
0.050
0.170
0.530
1.770
2.460
2.890
2.890
2009
j53
0.3123
0.000
0.000
0.010
0.020
0.040
0.110
0.390
0.940
1.240
2.070
2.070
Annual site max 3-month means >= national 90th percentile (2007-2009)
060371405-2008
2.4900
290930016-2008
2.4600
290930016-2009
2.0700
290930021-2009
1.9400
1290990015-2008
2.8900
!
060371405-2008
2.4900
i
May 2011
3-131
DRAFT - DO NOT CITE OR QUOTE
-------
Table 3A-5.One-month average Pb-TSPfor individual county concentrations nationwide
(jjg/m3), non-source-oriented monitors, 2007-2009
Stcou
State
I N
County name I monthly
I means
N
sites
Mean
Min
1
5
10
25
50
75
90
95 ! 99
I
i
max
Statistics for individual counties (2007-2009)
06025
CA
Imperial
33
1
0.0218
0.009
0.009
0.010
0.011
0.013
0.019
0.029
0.036
0.041
0.041
0.041
06037
CA
Los Angeles
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
06065
CA
Riverside
48
2
0.0090
0.000
0.000
0.002
0.003
0.008
0.010
0.010
0.016
0.018
0.022
0.022
06071
CA
San Bernardino
47
2
0.0112
0.000
0.000
0.002
0.003
0.008
0.010
0.014
0.020
0.022
0.036
0.036
12103
FL
Pinellas
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
17031
IL
Cook
143
4
0.0144
0.010
0.010
0.010
0.010
0.010
0.014
0.017
0.020
0.024
0.028
0.032
17117
IL
Macoupin
36
1
0.0102
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.013
0.014
0.014
17119
IL
Madison
35
1
0.0187
0.010
0.010
0.010
0.010
0.010
0.015
0.022
0.030
0.038
0.066
0.066
17143
IL
Peoria
36
1
0.0113
0.010
0.010
0.010
0.010
0.010
0.010
0.012
0.014
0.015
0.018
0.018
18097
IN
Marion
35
1
0.0071
0.002
0.002
0.002
0.003
0.004
0.006
0.009
0.012
0.012
0.033
0.033
18163
IN
Vanderburgh
21
2
0.0048
0.001
0.001
0.002
0.002
0.004
0.005
0.005
0.006
0.010
0.010
0.010
25025
MA
Suffolk
24
1
0.0089
0.000
0.000
0.000
0.003
0.007
0.009
0.011
0.015
0.016
0.020
0.020
26163
Ml
Wayne
24
1
0.0187
0.007
0.007
0.009
0.012
0.013
0.018
0.022
0.027
0.032
0.045
0.045
27037
MN
Dakota
129
4
0.0039
0.000
0.000
0.000
0.000
0.000
0.003
0.006
0.008
0.010
0.018
0.036
27053
MN
Hennepin
128
4
0.0034
0.000
0.000
0.000
0.000
0.002
0.003
0.005
0.008
0.008
0.010
0.010
27123
MN
Ramsey
71
3
0.0071
0.000
0.000
0.000
0.000
0.004
0.006
0.010
0.015
0.020
0.028
0.028
27137
MN
Saint Louis
70
2
0.0017
0.000
0.000
0.000
0.000
0.000
0.002
0.003
0.005
0.006
0.010
0.010
27163
MN
Washington
36
1
0.0035
0.000
0.000
0.000
0.000
0.003
0.004
0.004
0.006
0.006
0.008
0.008
36047
NY
Kings
36
1
0.0142
0.010
0.010
0.010
0.010
0.011
0.013
0.016
0.020
0.021
0.026
0.026
39017
OH
Butler
34
1
0.0052
0.000
0.000
0.000
0.003
0.004
0.005
0.007
0.008
0.009
0.009
0.009
39029
OH
Columbiana
107
3
0.0145
0.000
0.000
0.004
0.006
0.008
0.012
0.019
0.031
0.034
0.045
0.050
39035
OH
Cuyahoga
70
2
0.0138
0.004
0.004
0.006
0.006
0.009
0.013
0.016
0.022
0.028
0.041
0.041
39049
OH
Franklin
36
1
0.0088
0.004
0.004
0.004
0.005
0.007
0.008
0.011
0.014
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
Washington
43
2
0.0041
0.002
0.002
0.002
0.002
0.003
0.004
0.005
0.006
0.007
0.010
0.010
42003
PA
Allegheny
36
1
0.0067
0.000
0.000
0.000
0.000
0.000
0.004
0.012
0.016
0.021
0.053
0.053
42045
PA
Delaware
20
1
0.0400
0.040
0.040
0.040
0.040
0.040
0.040
0.040
0.040
0.040
0.040
0.040
42129
PA
Westmoreland
36
1
0.0410
0.034
0.034
0.037
0.040
0.040
0.040
0.040
0.044
0.050
0.053
0.053
48061
TX
Cameron
35
1
0.0038
0.002
0.002
0.002
0.002
0.003
0.003
0.005
0.006
0.007
0.009
0.009
48141
TX
El Paso
22
2
0.0229
0.014
0.014
0.015
0.016
0.018
0.019
0.019
0.041
0.056
0.057
0.057
48201
TX
Harris
31
1
0.0054
0.003
0.003
0.004
0.004
0.005
0.005
0.006
0.007
0.009
0.010
0.010
48479
TX
Webb
31
1
0.0103
0.004
0.004
0.005
0.006
0.007
0.008
0.010
0.024
0.025
0.026
0.026
May 2011
3-132
DRAFT - DO NOT CITE OR QUOTE
-------
Table 3A-6.Three-month moving average Pb-TSP for individual county concentrations
(jjg/m3) nationwide, non-source-oriented monitors, 2007-2009
Stcou
State
County name
N
monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
Statistics for individual counties (2007-2009)
06025
CA
Imperial
33
1
0.0216
0.010
0.010
0.010
0.020
0.020
0.020
0.030
0.030
0.030
0.030
0.030
06037
CA
Los Angeles
117
5
0.0113
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.020
0.020
0.030
0.030
06065
CA
Riverside
48
2
0.0102
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.020
0.020
06071
CA
San Bernardino
47
2
0.0115
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.020
0.020
0.020
0.020
12103
FL
Pinellas
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
17031
IL
Cook
143
4
0.0139
0.010
0.010
0.010
0.010
0.010
0.010
0.020
0.020
0.020
0.020
0.020
17117
Macoupin
36
1
0.0100
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010|0.010
17119
Madison
35
1
0.0182
0.010
0.010
0.010
0.010
0.010
0.020
0.020
0.030
0.040
0.040
0.040
17143
Peoria
36
1
0.0100
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
18097
IN
Marion
35
1
0.0076
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.020
0.020
0.020
18163 {IN
Vanderburgh
21
2
0.0037
0.000
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.010
0.010
25025
MA
Suffolk
24
1
0.0092
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.010
0.020
0.020
0.020
26163
Ml
Wayne
24
1
0.0196
0.010
0.010
0.010
0.010
0.020
0.020
0.020
0.020
0.030
0.030
0.030
27037
MN
Dakota
129
4
0.0029
0.000
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.010
0.010
27053
MN
Hennepin
128
4
0.0024
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010j0.010
27123
MN
Ramsey
71
3
0.0071
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.020
0.020
0.020
27137
MN
Saint Louis
70
2
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
36
1
0.0011
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.010
36047
NY
Kings
36
1
0.0128
0.010
0.010
0.010
0.010
0.010
0.010
0.020
0.020
0.020
0.020
0.020
39017
OH
Butler
34
1
0.0061
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.010
0.010
0.010
39029
OH
Columbiana
107
3
0.0151
0.010
0.010
0.010
0.010
0.010
0.010
0.020
0.020
0.030
0.030
0.030
39035
OH
Cuyahoga
70
2
0.0142
0.000
0.000
0.010
0.010
0.010
0.010
0.020
0.020
0.020
0.030
0.030
39049
OH
Franklin
36
1
0.0094
0.000
0.000
0.000
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
39143
OH
Sandusky
12
1
0.0060
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.010
0.010
0.010
39167
OH
Washington
43
2
0.0008
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
42003
PA
Allegheny
36
1
0.0058
0.000
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.020
0.030
0.030
42045 | PA
Delaware
20
1
0.0400
0.040
0.040
0.040
0.040
0.040
0.040
0.040
0.040
0.040
0.040
0.040
42129
PA
Westmoreland
36
1
0.0414
0.040
0.040
0.040
0.040
0.040
0.040
0.040
0.050
0.050
0.050
0.050
48061
TX
Cameron
35
1
0.0024
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.010
48141
TX
El Paso
22
2
0.0256
0.020
0.020
0.020
0.020
0.020
0.020
0.030
0.040
0.040
0.040
0.040
48201
TX
Harris
31
1
CO
CD
O
O
O
0.000
0.000
0.000
0.000
0.000
0.010
0.010
0.010
0.010
0.010
0.010
48479
TX
Webb
31
1
0.0100
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
3.8.1. 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.
May 2011
3-133
DRAFT - DO NOT CITE OR QUOTE
-------
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
Data from 2007-2009 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), the 90th
percentile of the absolute difference in concentrations (P90) in ppm, 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:
where X,:l and X,/.: represent the observed hourly concentrations for time period i at sites j and k, andp 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 2007-2009 are near or below the analysis
method's stated method detection limit (MDL). The MDL is generally taken as 0.01 because it is the
upper value of the range of MDLs reported for 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 3A-1 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. 2008b'). 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. 2008b). 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. 2008b). 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
\^(x„-xA1
Equation 3A-1
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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
four-lane road, and monitor J was located approximately 130 m south of a 44ane highway. Figure 3A-2
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 3A-1 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 3A-3. The average annual concentration at monitor A was 0.074 (.ig/nr3. 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/m3. Given
the position of this monitor with respect to the Exide facility, there is the potential for recirculation of
fugitive Pb emissions in the air sampled by that monitor. The average annual Pb concentration at monitor
D was 0.12 |ig/m3. and the 75th percentile of year-round data exceeded the level of the NAAQS; in
spring, the 70th percentile exceeded 0.15 |ig/nr\ Monitor B reported the lowest values among the source-
oriented monitors with an average annual concentration of 0.013 |ig/nr\ 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/m3.
The non-source-oriented monitors located at sites F-J all recorded low concentrations, with average
values ranging from 0.004 to 0.018 |ig/m3 (Figure 3A-3). 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 3A-7), illustrate that Pb has high intra-urban spatial variability. For
the source-oriented monitors, the highest correlation 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 correlation between source-oriented
monitor B and non-source-oriented monitor F, with p = 0.74. 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.
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1 Comparisons between the non-source-oriented monitors revealed moderate correlation between sites (G
2 to J [p = 0.37 to 0.65]). Sites G, H, I and J are all located in the southwestern quadrant of Los Angeles. It
3 is possible that they are also exposed to a ubiquitous source that produces a common signal at these four
4 sites.
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¦Los/Angeled
Legend
• City-based Population Center
• County-based Population Center
Major Highways
Los Angeles County, CA
Figure 3A-1. Pb TSP monitor and source locations within Los Angeles
County, CA (06-037), 2007-2009.Note that 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.
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Source: NRCS (2011V
Figure 3A-2. Wind roses for Los Angeles County, CA, from meteorological
data at the Los Angeles International Airport, 1961-
1990.Clockwise from top left: January, April, July, and October.
Note that the wind percentages vary from month to month.
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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
A
B
D
H
*itt +
Hi
ti
Y1234 Y1234
Y1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y 1 2 3 4 Y1 2 3 4 Y1 2 34 Y1234
season
Figure 3A-3. Box plots of annual and seasonal Pb TSP concentrations
(pg/m3) from source-oriented and non-source-oriented
monitors within Los Angeles County, CA (06-037), 2007-2009.
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Table 3A-7. Correlations 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
1.00
0.16
0.10
0.08
0.27
-0.15
0.00
0.14
-0.02
-0.09
0.00
0.08
0.49
0.10
0.10
0.08
0.06
0.08
0.08
0.08
0.00
0.63
0.64
0.31
0.34
0.57
0.57
0.79
0.77
0.85
B
Source
1.00
0.05
0.05
0.07
0.74
0.12
0.28
0.11
0.10
0.00
3.59
0.25
0.10
0.02
0.02
0.01
0.02
0.02
0.00
0.96
0.84
0.71
0.46
0.48
0.61
0.60
0.81
C
Source
1.00
0.57
0.03
-0.08
0.26
0.28
0.20
0.13
0.00
1.76
2.14
3.59
4.22
3.59
3.59
3.92
0.00
0.68
0.77
0.95
0.96
0.98
0.98
0.99
D
Source
1.00
0.12
0.17
0.11
0.24
0.21
0.07
0.00
0.17
0.24
0.25
0.25
0.25
0.25
0.00
0.42
0.78
0.80
0.89
0.89
0.95
E
Source
1.00
0.13
0.06
0.24
0.07
0.18
0.00
0.10
0.10
0.11
0.11
0.11
Legend
P
P90
COD
0.00
0.61
0.64
0.78
0.79
0.90
F
Non-Source
1.00
0.02
0.19
0.09
0.09
0.00
0.02
0.02
0.02
0.02
0.00
0.39
0.61
0.58
0.82
G
Non-Source
1.00
0.65
0.39
0.38
0.00
0.01
0.02
0.02
0.00
0.54
0.61
0.85
H
Non-Source
1.00
0.51
0.40
0.00
0.01
0.01
0.00
0.55
0.77
I
Non-Source
1.00
0.37
0.00
0.01
0.00
0.78
J
Non-Source
1.00
0.00
0.00
1 Figure 3A-4 illustrates Pb monitor locations within Hillsborough and Pinellas Counties in FL,
2 which comprise the greater Tampa-St. Petersburg metropolitan area. Two source-oriented monitors (A and
3 B) were located within Hillsborough County, and one non-source-oriented monitor (C) was located in
4 Pinellas County. Monitor A was located 360 m north-northeast of the EnviroFocus Technologies battery
5 recycling facility, which produced 1.3 tons/yr ("U.S. EPA. 2008c'). and monitor B was located 320 m
6 southwest of the same facility. Monitor C was located next to a two-lane road in Pinellas Park, FL.
7 Figure 3A-5 displays seasonal wind roses for the Tampa-St. Petersburg metropolitan area. These
8 wind roses suggest shifting wind directions throughout the winter, spring, and summer. During the winter,
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
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 3A-6. The average annual concentration at monitor A was 0.15 |ig/m3. and the 95th percentile was
0.70 (ig/m3. During winter, the 60th percentile of the data met the level of the NAAQS. At this site, the
highest concentrations occurred during summer, which corresponded to the time when westerly winds
were stronger. Concentration data at monitor B were much higher, with an annual average of 0.45 |ig/m3
and a 95th percentile of 1.9 (.ig/m3. Annually, the 55th percentile exceeded the level of the NAAQS, and in
autumn the 45th percentile exceeded the NAAQS. The highest concentrations occurred in autumn,
coinciding with the time when winds blew from the northeast, when monitor B was most often downwind
of the battery recycling facility. The non-source-oriented monitor C always reported concentrations of 0.0
|ig/m3. This is likely related to its location next to a quiet road in a small city.
Intersampler correlations, shown in Table 3A-8, illustrate that Pb has high intra-urban spatial
variability. The source oriented monitors were anticorrelated (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.
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lampa—St. jjetersburg
Legend
a TSPSouteo Mwwet#
Figure 3A-4. Pb TSP monitor locations within Hillsborough and Pinellas
Counties, FL (12-057 and 12-103), 2007-2009.Top: view of all
Pb FRM monitors in Hillsborough and Pinellas Counties.
Bottom: Close up of industrial site around monitors A and B.
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Source: NRCS (2011V
Figure 3A-5. Wind roses for Hillsborough/Pinellas Counties, FL, obtained
from meteorological data at Tampa International Airport, 1961-
1990.Clockwise from top left: January, April, July, and October.
Note that wind percentages vary from month to month.
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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
3.0 "
2.9 "
2.8 "
2.7 "
2.6 "
2.5 "
2.4 "
2.3 "
2.2 "
2.1 "
2.0 "
1.9 "
ro£ 1.8"
1-7"
3 1.6 "
S 1.5-
% 1.4"
£ 13 "
5 12 -
u
c 11-
o '¦1
o i0 -
0.9 "
0.8 -
0.7 "
0.6 "
0.5 "
0.4 "
0.3 "
0.2 "
0.1 -
o.o -
/
l
i
B
i
c
I I I I I
Y 1 2 3 4 Y 1 2 3 4 Y1234
season
Figure 3A-6. Box plots of annual and seasonal Pb TSP concentrations
(pg/m3) from source-oriented and non-source-oriented
monitors within Hillsborough and Pinellas Counties, FL (12-
057 and 12-103), 2007-2009.
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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
Table 3A-8. 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
1.00
-0.08
0.00
1.20
0.50
0.00
0.71
1.00
B
Source
1.00
0.00
2.20
Legend
0.00
1.00
C
Non-source
P
1.00
P90
0.00
COD
0.00
Figure 3A-7 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/yr (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 and was within 10 km of 2 Pb sources (0.41 and 1.08 tons/yr) (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/yr 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 I-
294 but not located on an arterial road; it was 9 km southeast of a 0.56 tons/yr 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/yr
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 monitor, monitor G was located 2 km southwest
of facilities emitting 0.30 and 0.41 tons/yr (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 3A-8 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
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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
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 3A-9 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 of the NAAQS.
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. However,
the average concentration at this site was 0.012 |ig/m3. and 85% of data were below the MDL. In contrast,
non-source monitors, E, F, G, and H had average concentrations of 0.011-0.017 |ig/m3. 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 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 3A-9).
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 [p = 0.26]). This might have reflected
more substantial differences related to the additional influence of industrial sources nearby monitor A.
Monitors (C and D) were uncorrelated with each other and with monitors (A and B), likely because their
exposure to sources was substantially different. The source-oriented and non-source-oriented monitors
were generally not well correlated. The highest correlation occurred between monitors (D and H [p =
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2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
0.53]). 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 [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 (p = 0.36 to 0.45), and
the correlation between monitors (E and G) was 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.
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Chicago
_
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
Cook County, IL
Figure 3A-7. Pb TSP Monitor locations within Cook County, IL (17-031),
20Q7-2009.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.
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Source: NRCS (2011^
Figure 3A-8. Wind roses for Cook County, IL, obtained from meteorological
data at O'Hare International Airport, 1961-1990.Clockwise from
the top left: January, April, July, and October. Note that the
wind percentages vary from month to month.
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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
0.15 -
0.14
0.13 -
0.12 -
0.11 -
0.10 -
ST 0-09 -
_E
^ 0.08 -
.2 0.07 -
4—1
c 0.06 -
U
c
° 0.05 -
0.04 -
0.03 -
0.02 -
0.01 -
0.00 -
A
I1
B
c
r
D
E
III
F
Mill
G
Mill
H
Y 1 2 3 4 Y 1 2 3 4 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234
season
Figure 3A-9. Box plots of annual and seasonal Pb TSP concentrations
(pg/m3) from source-oriented and non-source-oriented
monitors within Cook County, IL (17-031), 2007-2009.
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1
2
3
4
5
6
7
8
9
10
11
12
Table 3A-9. 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
1.00
0.26
-0.01
0.08
0.06
0.32
0.18
0.06
0.00
0.03
0.06
0.04
0.04
0.03
0.03
0.04
0.00
0.29
0.38
0.43
0.41
0.36
0.36
0.45
B
Source
1.00
0.05
0.10
0.32
0.46
0.35
-0.01
0.00
0.04
0.03
0.03
0.02
0.02
0.03
0.00
0.33
0.36
0.34
0.29
0.30
0.40
C
Source
1.00
0.04
0.16
0.10
0.17
0.06
0.00
0.05
0.05
0.04
0.05
0.05
0.00
0.40
0.39
0.35
0.35
0.42
D
Source
1.00
0.21
0.37
0.07
0.53
0.00
0.01
0.01
0.02
0.01
0.00
0.19
0.24
0.28
0.15
E
Non-Source
1.00
0.36
0.40
0.07
0.00
0.02
0.01
0.01
Legend
0.00
0.24
0.24
0.20
F
Non-Source
P
1.00
0.41
0.45
P90
0.00
0.01
0.02
COD
0.00
0.24
0.26
G
Non-Source
1.00
0.05
0.00
0.02
0.00
0.27
H
Non-Source
1.00
0.00
0.00
Figure 3A-10 illustrates Pb monitor locations with Jefferson County, MO. Ten source-oriented
monitors surrounded the Doe Run primary Pb smelter in Herculaneum, MO on the west and northwestern
sides. The largest distance between these monitors was approximately 1.5 km. Monitor E located on the
Doe Run facility roughly 20 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 (2007-2009), and the facility was estimated to have emitted
41.1 tons Pb/yr (U.S. EPA. 2008d). Figure 3A-11 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
May 2011
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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
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 3A-12 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/m3 across the monitors. The
maximum concentration was measured at monitor C to be 21.6 |ig/nr\ 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 3A-10). Correlations range
from p = -0.04 to 0.96. High correlations (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 (p = 0.96, for
monitors (E and F), monitor F is 250 m directly east of monitor E. Low correlation (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 winds did not carry pollutant to these sites concurrently with the signal recorded by the
other monitors.
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1
\ J
\ TV
\ H° °
F °G
D ooc
°o,E
A° C
o
B
N
\ 1 -¦ J, t
,
J
I ' \i i I i i i 1
0 0.5 1 2 Kilometers s
Legend
TSP Source Monitors
City-based Population Center
County-based Population Center
— Interstates
— Major Highways
Bodies of Water
Urban Areas
Jefferson County. MO
Icrcti ani'tim
Figure 3A-10. Pb TSP Monitor locations within Jefferson County, MO (29-
099), 2007-2009.Note that 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.
May 2011
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Source: NRCS (2011^
Figure 3A-11. Wind roses for Jefferson County, MO, obtained from
meteorological data at St. Louis/Lambert International Airport,
1961-1990.Clockwise from top left: January, April, July, and
October. Note wind percentages vary from month to month.
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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
7.0 -
6.5 -
6.0 -
5.5 -
5.0
4.5 -
cn
F
4.0 -
on
=L
C
o
3.5 -
+-»
C
3.0 -
ni
u
c
o
2.5 -
u
2.0
1.5 -
1.0 -
0.5 -
A
B
m
D
II
H
h!!i
mil
!.i
Y1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y 1 2 3 4 Y1 2 3 4 Y1234
season
Figure 3A-12. Box plots of annual and seasonal Pb TSP concentrations
(pg/m3) from source-oriented and non-source-oriented
monitors within Jefferson County, MO (29-099), 2007-2009.
May 2011
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Table 3A-10. 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
1.00
0.59
0.80
0.83
0.57
0.64
0.33
0.35
0.07
0.05
0.00
0.71
1.55
0.42
1.93
1.14
1.41
0.74
0.92
0.78
0.00
0.46
0.48
0.30
0.55
0.45
0.57
0.64
0.67
0.69
B
Source
1.00
0.53
0.43
0.10
0.14
0.07
0.22
0.10
0.09
0.00
1.86
0.87
2.77
1.96
2.08
0.94
1.04
0.91
0.00
0.58
0.51
0.69
0.62
0.68
0.68
0.65
0.65
C
Source
1.00
0.86
0.59
0.72
0.26
0.27
-0.04
0.04
0.00
1.56
2.26
1.26
2.94
2.65
3.18
2.60
0.00
0.50
0.50
0.46
0.60
0.74
0.73
0.73
D
Source
1.00
0.71
0.80
0.41
0.56
0.14
0.18
0.00
1.83
1.02
1.38
0.76
0.88
0.70
0.00
0.50
0.36
0.53
0.61
0.63
0.66
E
Source
1.00
0.96
0.54
0.46
0.06
0.16
0.00
0.86
2.16
2.50
3.09
2.57
Legend
0.00
0.35
0.49
0.66
0.70
0.72
F
Source
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
1.00
0.87
0.28
0.38
0.00
1.53
2.10
2.08
0.00
0.61
0.63
0.66
H
Source
1.00
0.20
0.30
0.00
0.89
0.56
0.00
0.67
0.65
1
Source
1.00
0.79
0.00
0.62
0.00
0.48
J
Source
1.00
0.00
0.00
1
2
3
4
5
6
7
8
May 2011 3-156 DRAFT - DO NOT CITE OR QUOTE
Figure 3A-13 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/yr source facilities and one 0.2 tons/yr source ("U.S. EPA. 20086).
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. 20086). 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 ton/yr Victory White Metal
-------
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
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 3A-14
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 3A-15 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/nr\ Concentrations at monitor C averaged 0.017
|ig/m3. and those at monitors D and E averaged 0.014 |ig/m3 and 0.013 (ig/m3, 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 3A-
11. Monitors A and B appear to be anticorrelated (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 well with each other (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.
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Cleveland
Legend
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
Figure 3A-13. Pb TSP Monitor locations within Cuyahoga County, OH (39-
035), 2007-2009.Top: view of all Pb FRM monitors in Cuyahoga
County. Bottom left: Close up of industrial site around
monitors A and B. Bottom right: Close up of monitor D north
of I-490.
May 2011
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Source: NRCS (2011^
Figure 3A-14. Wind roses for Cuyahoga County, OH, obtained from
meteorological data at Cleveland/Hopkins International
Airport, 1961-90.Clockwise from top left: Jan, April, July, and
October. Note wind percentages vary from month to month.
May 2011
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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
0.25 -
0.24 "
0.23 -
0.22 "
0.21 "
0.20 "
0.19 -
0.18 "
0.17 -
0.16 "
mE 0.15 :
"m 0.14 -
~ 0.13 "
~ 0.12 "
<0
£ 0.11 -
§ 0.10 -
o 0.09 "
u
0.08 "
0.07 "
0.06 -
0.05 -
0.04 -
0.03 -
0.02 -
0.01 -
o.oo -
<
B
i i i i i
c
Jll
I I I I I
D
M'i
i i i i i
E
IJll
i i i i i
Y 1 2 3 4 Y 1 2 3 4 Y1234 Y1234 Y1234
season
Figure 3A-15. Box plots of annual and seasonal Pb TSP concentrations
(pg/m3) from source-oriented and non-source-oriented
monitors within Cuyahoga County, OH (39-035), 2007-2009.
May 2011
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
Table 3A-11. 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
1.00
-0.13
0.24
0.19
0.21
0.00
0.18
0.05
0.04
0.05
0.00
0.64
0.33
0.35
0.37
g Source
1.00
0.31
0.24
0.34
0.00
0.18
0.19
0.19
0.00
0.69
0.71
0.73
q Source
1.00
0.77
0.67
0.00
0.01
0.01
Legend
0.00
0.17
0.18
— Non-Source
P
1.00
0.67
P90
0.00
0.01
COD
0.00
0.17
— Non-Source
1.00
0.00
0.00
Figure 3A-16 illustrates Pb monitor locations within Sullivan County, TN. Three source-oriented
monitors were situated around an Exide Pb recycling facility emitting 0.78 tons/yr (U.S. EPA. 2008f).
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 3A-17
displays seasonal wind roses for Sullivan County. During winter and spring, the predominant winds come
from the southwest and west. In the summer, the percentage of wind coming from the west and southwest
is roughly equal to that for wind coming from the east and northeast, although the easterly winds are
calmer. During autumn, winds come predominantly from the northeast and east, although these winds
tend to be calmer than those originating from the southwest and west.
The data presented in Figure 3A-18 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
May 2011
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1
2
3
4
5
6
7
8
9
10
11
12
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/nr\ 0.26 |ig/nr'. 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 were quite low (Table 3 A-
12). The correlation between monitors B and C was 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.
May 2011
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Kingsport
Legend
TSP Source Monitors
City-based Population Center
County-based Population Center
— Interstates
|| Major Highways
| Bodies of Water
Urban Areas
Sullivan County, TN
Figure 3A-16. Pb TSP Monitor locations within Sullivan County, TN (47-163),
2007-2009.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.
May 2011
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Source: NRCS (2011^
Figure 3A-17. Wind roses for Sullivan County, TN, obtained from
meteorological data at Bristol/Tri City Airport, 1961-
90.Clockwise from top left: January, April, July, and October.
Note that the wind percentages vary from month to month.
May 2011
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Site
A
B
c
SITE ID
47-163-3001
47-163-3002
47-163-3003
MEAN
0.11
0.051
0.059
SD
0.11
0.036
0.047
OBS
334
362
345
% BELOW MDL
0
0
0
Source orientation
Source
Source
Source
0.44 "
0.42 "
0.40 -
0.38 "
0.36 "
0.34 "
0.32 "
0.30 "
0.28 -
£ 0.26 -
M o.24 "
T 0.22 "
| 0.20 -
c 0.18 "
a)
c 0.16 "
0
u 0.14 -
0.12 "
0.10 -
0.08 "
0.06 "
0.04 -
0.02 -
0.00 -
>
CO
C
I I I I I
Y 1 2 3 4 Y 1 2 3 4 Y1234
season
Figure 3A-18. Box plots of annual and seasonal Pb TSP concentrations
(pg/m3) from source-oriented and non-source-oriented
monitors within Sullivan County, TN (47-163), 2007-2009.
May 2011
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Table 3A-12. Correlations between Pb TSP concentrations from source-oriented and
non-source-oriented monitors within Sullivan County, TN (47-163), 2007-2009
A
B
c
Source
Source
Source
A
Source
1.00
-0.04
0.13
0.00
0.21
0.19
0.00
0.47
0.43
B
Source
1.00
0.45
0.00
0.06
Legend
0.00
0.23
C
Source
P
1.00
P90
0.00
COD
0.00
3.8.2. Size Distribution of Pb-Bearing PM
Table 3A-13. Correlations and average of the concentration ratios for co-located
monitors, TSP versus PM10, TSP versus PM2.5, and PM10 versus PM2.5.Data are
bolded for sites where the TSP, PM10, and PM2.5 monitors were co-
located for at least one sampling year.
Site County | Land Type | Years j ^orr j Ratio I Years j Corr j jYears Corr j
PMi0:TSP
PM2.5:TSP j PM2.5: PM10
010730023
Jefferson, AL
Urban and
center city
2005-2006
0.87
0.80
2005-2006
0.74
0.84
2005-2006
0.81
1.02
040139997
Maricopa, AZ
Urban and
center city
2006-2009
0.40
0.59
0600110
01
Alameda, CA
Suburban
1994-1998
0.00
2.31
060130002
Contra Costa, CA
Suburban
1994-1998
0.27
2.60
060130003
Contra Costa, CA
Suburban
1994-1997
0.31
1.04
060190008
Fresno, CA
Suburban
1995-2001
0.92
0.92
2000-2009
0.58
0.66
2000-2001
0.83
1.16
060250005
Imperial, CA
Suburban
1996-2001
0.77
0.91
2002-2009
0.95
0.74
060290014
Kern, CA
Urban and
center city
1995-2000
0.92
0.76
2001-2009
0.22
0.72
060371103
Los Angeles, CA
Urban and
center city
2002-2009
0.22
0.28
060374002
Los Angeles, CA
Suburban
1995-2000
0.72
0.39
CO
o
o
o
LO
LO
O
CO
O
Napa, CA
Urban and
center city
1994-1994
0.42
0.93
060658001
Riverside, CA jSuburban
1995-1997
0.27
0.39
2001-2009
0.59 |0.40 i
060730003
San Diego, CA jSuburban
2001-2009
0.35 jo.48 !
060750005
San Francisco, CA
Urban and
center city
1994-1998
0.29
0.98
060771002
San Joaguin, CA
Urban and
center city
1995-2000
0.73
0.82
o
o
o
LO
00
o
CO
o
Santa Clara, CA
Urban and
center city
1994-2000
0.32
0.95
2000-2002
0.63
0.63
060850004
Santa Clara, CA
Urban and
center city
2002-2009
0.11
0.78
2008-2009
0.76
1.41
060853001
Santa Clara, CA
Suburban
1994-1998
0.16
1.00
060990002
060990005
061112002
Stanislaus, CA
Stanislaus, CA
Ventura, CA
Urban and
center city
Urban and
center city
Suburban
1995-1998
1998-2000
0.94
0.79
0.82
1.02
2001-2009
0.47
0.74
May 2011 3-166 DRAFT - DO NOT CITE OR QUOTE
-------
Site
County
Land Type
Years
Corr
Avg
Ratio
Years
Corr
Avg
Ratio
Years
Corr
Avg
Ratio
PMi0:TSP
PM2.5:TSP
PM2.5: PM10
080010006
Adams, CO
Suburban
I '
2001-2006
0.16
0.43
| |
080410011
El Paso, CO
Urban and
center city
2001-2006
0.03
0.39
080770017
Mesa, CO
Urban and
center city
2004-2008
0.42
0.87
100032004
New Castle, DE
Urban and
center city
2003-2007
0.47
0.44
110010043
District of Columbia, DC
Urban and
center city
2004-2009
0.47
0.73
120571075
Hillsborough, FL
Urban and
center city
2001-2003
0.83
1.86
120573002
Hillsborough, FL
Rural
2004-2009
0.66
1.55
121030026|Pinellas, FL
Suburban | III II
2004-2009
0.56
1.89
130690002jCoffee, GA
Rural
2003-2009
0.82
1.46
130890002|DeKalb, GA
Suburban
2003-2009
0.92
1.69
150032004
Honolulu, HI
Urban and
center city
2003-2009
0.77
2.34
170314201
Cook, IL
Suburban
2006-2009
0.53
0.39
2002-2009
0.25
0.36
2005-2009
0.65
1.01
180970078
Marion, IN
Suburban
2002-2009
0.75
0.78
201731012
Sedgwick, KS
Suburban
1993-1997
0.63
0.33
202090015
Wyandotte, KS
Urban and
center city
1993-1997
0.66
0.55
202090020jwyandotte, KS
Urban and
center city
1993-1997
0.99
0.56
210430500
Carter, KY
Rural
2008-2009
0.22
1.39
211930003
Perry, KY
Suburban
2003-2008
0.76
1.47
220511001
Jefferson, LA
Suburban
2005-2006
0.96
0.79
220710010
Orleans, LA
Urban and
center city
2005-2006
0.94
0.64
220710012
Orleans, LA
Urban and
center city
2005-2006
0.91
0.53
220718105
Orleans, LA
Urban and
center city
2005-2006
0.98
0.80
220718401
Orleans, LA
Urban and
center city
2005-2006
0.89
0.45
220758400 Plaquemines, LA
Urban and
center city
2005-2005
0.98
0.83
220870004
St. Bernard, LA
Suburban
2005-2006
0.94
0.78
220878103
St. Bernard, LA
Urban and
center city
2005-2006
0.90
0.66
250250042
Suffolk, MA
Urban and
center city
2009-2009
0.43
0.37
2003-2009
0.38
0.78
260770905
Kalamazoo, Ml
Urban and
center city
1993-1996
0.98
0.77
260810020
Kent, Ml
Urban and
center city
2005-2007
0.58
0.70
261130001
Missaukee, Ml
Rural
2002-2007
0.73
1.38
261250010
Oakland, Ml
Urban and
center city
2001-2002
0.40
1.60
261390009
Ottawa, Ml
Urban and
center city
2000-2001
0.90
0.67
261610008
Washtenaw, Ml
Urban and
center city
2003-2007
0.58
0.71
261630001
Wayne, Ml
Suburban
2001-2007
0.68
0.58
261630019
Wayne, Ml
Suburban
2001-2002
0.45
0.51
261630033
Wayne, Ml
Suburban
2003-2009
0.87
0.84
2002-2009
0.84
0.49
2003-2009
0.86
0.61
270530053
Hennepin, MN
Urban and
center city
1996-2001
0.47
0.55
270530963
Hennepin, MN
Urban and
center city
2002-2009
0.64
0.42
280458104
Hancock, MS
Suburban
2005-2006
0.93
1.51
280458105|Hancock, MS
Rural
2005-2006
0.99
1.02
280458201 IHancock, MS
Suburban
2005-2006
0.94
1.05
280470008|Harrison, MS
Rural
2005-2006
0.87
0.99
280478101
Harrison, MS
Suburban
2005-2006
0.98
0.99
280478102
Harrison, MS
Suburban
2005-2006
0.80
1.10
280478103
Harrison, MS
Suburban
2005-2006
0.93
1.09
280590006
Jackson, MS
Urban and
center city
2005-2006
0.97
0.78
295100085
St. Louis City, MO
Urban and
center city
2004-2004
0.96
1.10
2004-2004
0.26
1.08
2003-2009
0.76
0.76
330110020jlHiMsborough, NH
Urban and
center city
2002-2005
0.42
0.81
May 2011
3-167
DRAFT - DO NOT CITE OR QUOTE
-------
Site
County
Land Type
Years
Corr
Avg
Ratio
Years
Corr
Avg
Ratio
Years
Corr
Avg
Ratio
PMi0:TSP
PM2.5:TSP
PM2.5: PM10
330150014
Rockingham, NH
Urban and
center city
2003-2005
0.80
0.72
350010023
Bernalillo, NM
Urban and
center city
2004-2009
0.04
0.73
360050110
Bronx, NY
Urban and
center city
2007-2009
0.80
0.60
360551007
Monroe, NY
Urban and
center city
2007-2009
0.77
1.30
360610062
New York, NY
Urban and
center city
2005-2005
0.96
0.57
360632008
Niagara, NY
Suburban
2005-2005
0.74
0.84
401091037|0klahoma, OK
Suburban j
2009-2009
0.25
1.57
401431127 Tulsa, OK
Urban and
center city
2009-2009
0.34
0.74
410390060
Lane, OR
Urban and
center city
2002-2004
0.33
1.65
410510246
Multnomah, OR
Urban and
center city
2002-2003
0.90
0.90
410290133
Jackson, OR
Urban and
center city
2009-2009
0.54
1.10
410390060
Lane, OR
Urban and
center city
2004-2009
0.76
2.15
410510246
Multnomah, OR
Urban and
center city
2003-2006
0.98
0.99
410610119
Union, OR
Urban and
center city
2004-2007
0.15
1.60
420450002
Delaware, PA
Urban and
center city
2002-2008
0.04
0.11
420710007|Lancaster, PA
Suburban i
2004-2007
0.78
0.80
421010004
Philadelphia, PA
Urban and
center city
2000-2007
0.37
0.38
421010055
Philadelphia, PA
Urban and
center city
2005-2007
0.36
0.32
421010136
Philadelphia, PA
Urban and
center city
2004-2005
0.75
0.38
440070022
Providence, Rl
Urban and
center city
2001-2002
0.78
0.54
2002-2009
1.00
0.70
440071010
Providence, Rl
Suburban
2001-2002
0.98
0.71
450250001
Chesterfield, SC
Rural
2002-2009
0.37
1.76
2004-2009
0.41
1.63
450790019
Richland, SC
Urban and
center city
2001-2006
0.74
0.58
470370023jDavidson, TN
Urban and j
center city
2003-2004
0.39
0.83
482011034
Harris, TX
Urban and
center city
2002-2005
0.41
0.40
482011039
Harris, TX
Suburban
2000-2009
0.34
0.63
490110004
Davis, UT
Suburban
2003-2009
0.68
0.91
500070007
Chittenden, VT
Rural
2004-2009
0.89
0.97
510870014
Henrico, VA
Suburban
2004-2008
0.54
0.79
2008-2009
0.91
1.08
530330038
King, WA
Suburban
2001-2002
0.64
0.53
o
00
o
o
CO
CO
o
CO
LO
King, WA
Urban and
center city
2003-2009
0.91
0.81
530630016
Spokane, WA
Suburban
I I
2005-2005
0.77
0.91
550270007
Dodge, Wl
Rural
I |2004-2005
0.90
0.83
2005-2009
0.59
1.02
May 2011
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DRAFT - DO NOT CITE OR QUOTE
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3.8.3. Lead Concentration in a Multipollutant Context
Source, S02
Non-Source, S02
Source, PM2.5
Non-Source, PM25
Source, PM10
Non-Source, PM10
Source, 03
Non-Source, 03
Source, N02
Non-Source, N02
Source, CO
Non-Source, CO
US Wint«y
o o
o (DB (D mnnm oce o CD o ocoo 0
o 000
o o 00 cd ocnii,> qp i
o
O O
00 o cd cd © ©oaaoaowooo© © o oca?
-1.0
¦0.5 0.0 0.5
Spearman Correlation Coefficient
1.0
Source, S02
Non-Source, S02
Source, PM25
Non-Source, PM25
Source, PM10
Non-Source, PM10
Source, 03
Non-Source, 03
Source, N02
Non-Source, N02
Source, CO
Non-Source, CO
US Spring
00 0
cd cbd m ootaotaDSD asooooo 00
00 000
O OO 1 H IIIIHBIIIIHin Offl o ®
0 00
O O OP Q CP QMMMDDOMD ODD QDfl) CP 0
O O O O
o oanDanoam cud ooo o o oo $
o o
o ooo oasansBD amnmsDoaso
o o
© o O OOCDOOTDDOO OOOGO O O O O
-1.0
-0.5 0.0 0.5
Spearman Correlation Coefficient
1.0
Figure 3A-19a. Seasonal correlations of monitored Pb-TSP concentration
with copollutant concentrations, 2007-2008. Top: winter;
bottom: spring.
May 2011
3-169
DRAFT - DO NOT CITE OR QUOTE
-------
Source, S02
Non-Source, S02
Source, PM25
Non-Source, PM25
Source, PM10
Non-Source, PM10
Source, 03
Non-Source, 03
Source, N02
Non-Source, N02
Source, CO
Non-Source, CO
US Summer
o 00
O QD
a> o
000 ooamxB anooo o
0 o
o oaDcooamoo ao
00
O OaXDHHC m> GCBDGDOCD o
o 000
-1.0
-0.5 0.0 0.5
Spearman Correlation Coefficient
1.0
Source, S02
Non-Source, S02
Source, PM25
Non-Source, PM25
Source, PM,0
Non-Source, PM-0
Source, 03
Non-Source, 03
Source, N02
Non-Source, N02
Source, CO
Non-Source, CO
US Fall 00 o
O OO GGODO© O CD O GO O CO O 0
OOO o
o o o o wrnmMummamo oo <3m>
o o ®
o o o o
1.0
-0.5 0.0 0,5
Spearman Correlation Coefficient
1.0
Figure 3A-19b. Seasonal correlations of monitored Pb-TSP concentration
with copollutant concentrations, 2007-2008. Top: summer;
bottom: fall.
May 2011
3-170
DRAFT - DO NOT CITE OR QUOTE
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Non-Source, S02 1 "
US Winter
00 0
0
Source, PM2.5 * -
0
Non-Source, PM25 '¦ -
0
0 000
0
Non-Source, PM10 : -
0 000
0 <
Non-Source, 03 i -
0 0
0
0
Non-Source, N02 i -
0
0
Non-Source, CO . -
0 0 00
0
-1.0 -0.5 0.0 0.5 1.0
Spearman Correlation Coefficient
Non-Source, S02 -
US Spring
O OO O
Source, PM2.5 -
O
Non-Source, PM25 -
O
8
O
Non-Source, PM10 -
0 0
00 0 0
Non-Source, 03 .
GD O
0
0
0
Non-Source, N02 .
0 0
Non-Source, CO _
0
0
0
0
i 1 r
-1.0 -0.5 0.0 0.5 1.0
Spearman Correlation Coefficient
Figure 3A-20a. Seasonal correlations of monitored Pb-TSP concentration
with copollutant concentrations, 2009. Top: winter; bottom:
spring.
May 2011
3-171
DRAFT - DO NOT CITE OR QUOTE
-------
Non-Source, S02 ~
US Summer
0 0
O O
Source, PM2.5 "
O
Non-Source, PM2.5 -
0
O O O
Non-Source, PM10 -
0
O O
Non-Source, 03 -
0
O (BED
O
Non-Source, N02 -
O O
Non-Source, CO -
0
OO O
O
-1.0 -0.5 0.0 0.5 1.0
Spearman Correlation Coefficient
Non-Source, S02
US Fall
8
0
0
Source, PM2.5
0
8
0
0
Non-Source, PM2.5
-
0
0
0
0
0
0
Non-Source, PM10
Non-Source, 03
0 0 OO 0 0 0
Non-Source, N02
0 0
Non-Source, CO -
0
0 0 0
1 1 r
-1.0 -0.5 0.0 0.5 1.0
Spearman Correlation Coefficient
Figure 3A-20b. Seasonal correlations of monitored Pb-TSP concentration
with copollutant concentrations, 2009. Top: summer;
bottom: fall.
May 2011
3-172
DRAFT - DO NOT CITE OR QUOTE
-------
US Winter
Zinc
Volatile Nitrate
Vanadium
Titanium
Sulfur
Sulfate
Sodium Ion
Sodium
Silicon
Selenium
Potassium Ion
Potassium
Organic Carbon, Blank Adjusted
Non-Volatile Nitrate
Nitrate
Nickel
Mercury
Manganese
Magnesium
Iron
Elemental Carbon
Crustal
Copper
Chromium
Chlorine
Carbonate Carbon
Calcium
Cadmium
Bromine
Arsenic
Ammonium
Aluminum
o o o
O octal
a
COflDO
GO 3S
o oop m
o
O 00 o
MP Q
-1.0 -0,5 0.0 0.5 1.0
Spearman Corrgisrtion Coefficient
US Spring
Zinc
Volatile Nitrate
Vanadium
Titanium
Sulfur ¦
Sulfate
Sodium Ion
Sodium
Silicon
Selenium
Potassium Ion
Potassium
Organic Carbon, Blank Adjusted
Non-Volatile Nitrate
Nitrate
Nickel
Mercury
Manganese
Magnesium
Iron
Elemental Carbon
Crustal
Copper ¦
Chromium
Chlorine
Carbonate Carbon
Calcium
Cadmium
Bromine
Arsenic
Ammonium
Aluminum
COSSDQBV
o 0>
QDOE
oo m
m oo
CO
-1.0 -0.5 0.0 0,5 1.0
Spearman Correlation Coefficient
Figure 3A-21a. Seasonal correlations of monitored Pb-PM2.5 concentration
with copollutant concentrations, 2007-2009. Left: winter;
right: spring.
May 2011
3-173
DRAFT - DO NOT CITE OR QUOTE
-------
US Summer
Zinc
Volatile Nitrate
Vanadium
Titanium
Sulfur •
Sulfate
Sodium Ion
Sodium
Silicon
Selenium
Potassium Ion
Potassium
Organic Carbon, Blank Adjusted
Non-Volatile Nitrate
Nitrate
Nickel
Mercury •
Manganese
Magnesium
Iron
Elemental Carbon
Crustal
Copper •
Chromium
Chlorine
Carbonate Carbon
Calcium
Cadmium
Bromine
Arsenic
Ammonium
Aluminum
o o o
o
qd
BOO
DOT
o oo at
oca
o o a
ip < >
oo o oc
®oo
O O
» <)
-1.0 -0.5 0.0 0.5 1.0
Spearman Correlation Coefficient
Figure 3A-21b. Seasonal correlations of monitored Pb-PM2.5 concentration
with copollutant concentrations, 2007-2009. Left: summer;
right: fall.
May 2011
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Table 3A-14. Copollutant exposures for various trace metal studies
Adgate et al. (2007)
Riediker et al. (2003)
Pekey et al.
Molnar et al. (2007)
l-R(medfb
Personal
Vehicle
Roadside
I -near Industry
l-R
I-School
l-Pre-School
(median)0
(range)0
(range)0
(range)5
(medianf
(median)"
(median)3
Location
Minnesota
New Jersey
Kocaell, 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
Ag
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 j |
1
1
1-2
Mo
Br
2.1
1.3
1.3
al: Indoor; Units: ng/m3
bR: Residential; Units: ng/m3
cUnits: ng/m3
May 2011
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Chapter 4 Contents
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 ofPb measurements for EPA Region 5 from the NHEXAS
study. 4-4
4.1.1.1. Airborne Lead Exposure 4-4
Table 4-2. Estimates of fixed effects multivariate modeling ofPb levels measured
during the NHEXAS-MD study 4-5
Table 4-3. Comparison of personal, indoor, and outdoor Pb-bearing PM
measurements from several studies. 4-7
4.1.1.2. Exposure to Lead in Soil and Dust 4-7
Table 4-4. Measurements of indoor Pb dust from various studies. 4-8
4.1.1.3. Dietary Lead Exposure 4-10
Table 4-5. Pb bioaccumulation data for various plants. 4-15
4.1.1.4. Occupational 4-16
4.1.1.5. Exposure to Lead from Consumer Products 4-17
Table 4-6. Pb content in various consumer products 4-18
4.2. Kinetics 4-19
4.2.1. Absorption 4-19
4.2.1.1. Inhalation 4-19
4.2.1.2. Ingestion 4-21
Figure 4-2. Estimated relative bioavailability (RBA, compared to Pb-acetate) of
ingested Pb in mineral groups, based on results from juvenile swine
assays. 4-24
4.2.2. Distribution 4-26
4.2.2.1. Blood 4-26
Figure 4-3. Plot of blood and plasmaPb concentrations in measured in adults and
children. 4-27
Figure 4-4. Relationship between Pb intake and blood Pb concentration in infants (n
= 105, age 13 weeks, formula-fed). 4-28
Figure 4-5. Simulation of quasi-steady state blood and plasmaPb concentrations in
a child (age 4 years) associated with varying Pb ingestion rates.
Simulation based on ICRP Pb biokinetics model (Leggett, 1993). 4-29
4.2.2.2. Bone 4-30
4.2.2.3. Soft Tissues 4-31
4.2.2.4. Fetus 4-32
4.2.2.5. Organic Lead 4-32
4.2.3. Elimination 4-32
4.3. Lead Biomarkers 4-34
4.3.1. Bone Lead Measurements 4-34
4.3.2. Blood Lead Measurements 4-35
Figure 4-6. Simulation of temporal relationships between Pb exposure and blood Pb
concentration in children. 4-38
4.3.3. Urine Lead Measurements 4-39
Figure 4-7. Simulation of relationship between urinary Pb excretion and body
burden in adults. 4-40
4.3.4. Lead in Other Potential Biomarkers 4-41
4.3.4.1. Teeth 4-41
4.3.4.2. Hair 4-41
May 2011 i DRAFT - DO NOT CITE OR QUOTE
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4.3.4.3. Saliva 4-42
4.3.4.4. Serum 5-ALA and ALAD 4-43
4.3.5. Relationship between Lead in Blood and Lead in Bone 4-43
4.3.5.1. Children 4-44
Figure 4-8. Simulation of relationship between blood Pb concentration and body
burden in children, with a constant Pb intake from age 2 to 5. 4-46
Figure 4-9. Simulation of relationship between time-integrated blood Pb
concentration and cumulative Pb absorption in children. 4-47
4.3.5.2. Adults 4-47
Figure 4-10. Simulation of relationship between blood Pb concentration, bone Pb
and body burden in adults. 4-49
4.3.6. Relationship Between Lead in Blood and Lead in Soft Tissues 4-52
Figure 4-11. Simulation of blood and soft tissue (including brain) Pb in children and
adults who experience a period of increased Pb intake. Simulation based
on 1CRP Pb biokinetics model (Leggett, 1993). 4-53
Figure 4-12. Simulation of blood and brain Pb in children and adults who experience
a period of increased Pb intake. 4-55
4.3.7. Relationship Between Lead in Blood and Lead in Urine 4-55
Figure 4-13. Top panel: Predicted relationship between plasma Pb concentration and
urinary Pb excretion in an adult (age 40 years). Lower panel:
Simulation of blood Pb, bone Pb and urinary excretion ofPb in an adult
who experiences a period of increased Pb intake. Simulation based on
1CRP Pb biokinetics model (Leggett, 1993). 4-57
4.4. Observational Studies of Lead Exposure 4-58
4.4.1. Lead in Blood 4-58
Figure 4-14. Temporal trend in blood Pb concentration. 4-59
Table 4-7. Blood Pb concentrations in the U.S. population 4-60
Figure 4-15. Box plots ofbloodPb levels among U.S. children (1-5 years old) from
the NHANES survey, 1988-2008. 4-61
Figure 4-16. Percent distribution ofbloodPb levels by race/ethnicity among U.S.
children (1-5 years) from the NHANES survey, 1988-1991 (top) and
1999-2004 (bottom). 4-62
Figure 4-17. a) Trends in 206Pb/204Pb isotope ratio in blood Pb and b) trends in blood
Pb levels among Australian study populations during the period 1990-
2000. 4-64
4.4.2. Lead in Bone 4-64
Table 4-8. Epidemiologic studies that provide bone Pb measurements for non-
occupational exposed populations 4-65
Table 4-9. Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations 4-73
4.4.3. Lead in Urine 4-76
Table 4-10. Urine Pb concentrations in the U.S. population 4-77
4.4.4. Lead in Teeth 4-77
Figure 4-18. Comparison of tooth enamel 4-78
4.5. Empirical Models of Lead Exposure-Blood Lead Relationships 4-78
4.5.1. Air Lead-Blood Lead Relationships 4-79
Table 4-11. Summary of estimated slopes for blood Pb to air Pb relationships in
humans 4-80
Figure 4-19. Predicted relationship between air Pb and blood Pb based on a meta
analysis of 18 studies. 4-81
4.5.1.1. Children 4-81
Table 4-12. Environmental Pb levels and blood Pb levels in children in Trail, British
Columbia 4-82
Table 4-13. U.S. gasoline Pb consumption and air Pb levels 4-84
Table 4-14. Air Pb levels and bloodPb levels in children in Mumbai, India 4-85
Figure 4-20. Predicted relationship between air Pb and blood Pb based on data from
Chicago, 1L (1974-1988). 4-86
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4.5.1.2. Adults 4-86
Table 4-15. Significant predictors of blood Pb concentration in bridge painters 4-87
1. 4.5.2. Environmental Lead-Blood Lead Relationships4-87
Table 4-16. Linear model relating environmental Pb exposure and blood Pb
concentration in children11 4-89
Figure 4-21. Predicted relationship between soil Pb concentration and blood Pb
concentration in children based data collected in the New Orleans child
Pb screening program (2000-2005) (Mielke et al., 2007). 4-90
Table 4-17. General linear model relating blood Pb concentration in children and
environmental Pb levels—Bunker Hill Superfund Site 4-92
4.6. Biokinetic Models of Lead Exposure-Blood Lead Relationships 4-92
4.7. Summary 4-93
4.7.1. Exposure 4-93
4.7.2. Kinetics 4-94
4.7.3. Lead Biomarkers 4-96
4.7.4. Air Lead-Blood Lead Relationships 4-97
Chapter 4. References 4-99
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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. The recent information provided here builds upon the conclusions of the
2006 Pb AQCD (2006). which found that air Pb concentrations and blood Pb levels have decreased
substantially following the 1996 ban on Pb in on-road vehicle gasoline, the 1978 ban on Pb in household
paints, and the 1986 and 1995 restrictions on uses of Pb solder. At the same time, 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. 2006) 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/ni1 for source-oriented monitors during that time period, while blood Pb levels reached a median of
1.70 (ig/dL among children ages 1-5 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. For the current review, Section 4.1. contains a
description of studies related to pathways for human exposure to Pb.
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 or water. 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, and a fraction of that resuspended Pb may even originate from waters used to
irrigate the soil. Additionally, Pb-bearing PM can be deposited from the air to soil or water through wet
and dry deposition. In general, air-related pathways include those pathways where Pb passes through
ambient air on its path from a source to human exposure. Air-related Pb exposures include inhalation and
ingestion of Pb-contaminated food, water or other materials including dust and soil. Non-air-related
exposures include ingestion of indoor Pb paint, Pb in diet as a result of inadvertent additions during food
Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and Environmental
Research Online) at http://eDa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of developing science
assessments such as the Integrated Science Assessments (ISA) and the Integrated Risk Information System (IRIS).
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processing, and Pb in drinking water attributable to Pb in distribution systems, as well as other generally
less prevalent pathways. The complicated nature of Pb exposure is illustrated 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 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 difficulty 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.
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AIR
Newly Emitted Pb
Historically Emitted Pb
'DOOR SOIL^
\ND DUST i
NATURAL WATERS
.AND SEDIMENTS
Non-air Pb
Releases,
%
PaintN
AIR
AIR
AIR
SOIL
SOIL
SOIL
WATER
WATER
WATER
PLANT
EXPOSURE
-
ANIMAL 1
exposure/
Diet
HUMAN
nnking Water
PIPES
WELR
COSMETICS
OYS etc
OCCUPATION
Figure 4-1. Conceptual model of multimedia Pb exposure.The Venn
diagram is used to illustrate the passage of Pb through
multiple environmental media compartments through which
exposure can occur.
Hie relative importance of different sources or pathways of potential exposure to Pb in the
environment is often difficult to discern. Individual factors such as home environment, location, and
susceptibility 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 m EPA Region 5 (Illinois, Indiana,
Michigan, Minnesota, Ohio, and Wisconsin) (Clayton et al.. 1999) as well as in Arizona (O'Rourke et al..
1999) and Maryland (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). It is plausible
that local resuspension of Pb-containing dust due to human activity increased personal exposure
concentrations of airborne Pb relative to indoor or outdoor air Pb concentrations. 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
measurable (CLs)c
Mean (CLs)c
50th (CLs)c
90th (CLs)c
Personal air (ng/m0)"
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)
Indoor air (ng/mJ)°
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/mJ)a
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/cmz)
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/cmz)
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)
Window sill dust (mg/kg)
239
95.8 (92.5; 99.0)
954.07(506.70; 1,401.4)
191.43 (140.48; 256.65)
1,842.8 (1,151.3; 2,782.5)
Standing tap water (|jg/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 (|jg/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 (|jg/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 (|jg/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 (|jg/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 (|jg/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 (|jg/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
(|jg/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 (|jg/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).
The lower and upper bounds of the 95% confidence limits (CL) are provided.
dPM50.
Source: Used with permission from Nature Publishing Group, Clayton et al. (1999)
4.1.1.1. Airborne Lead Exposure
1 Limited personal exposure monitoring data for airborne Pb were available for the 2006 AQCD
2 (U.S. EPA. 2006V As described above, the NHEXAS study showed personal air Pb concentrations to be
3 significantly higher than indoor or outdoor air Pb concentrations (Clavton et al.. 1999). Indoor air Pb
4 concentration was shown to be moderately correlated with floor dust and residential yard soil Pb
5 concentration (M. Rabinowitz et al.. 1985). Egeghy et al. (2005) performed multivariate fixed effects
6 analysis of the NHEXAS-Maryland data and found that Pb levels measured in indoor air were
7 significantly associated with log-transformed outdoor air Pb levels, ambient temperature, number of hours
8 in which windows were open, homes built before 1950, and frequency 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 P
p-value pa
p-value
P3
p-value
P3
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 soil"
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
Work 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
aEstimates 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: Used with permission from Nature Publishing Group, Egeghy et al. (2005).
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 samples of PM25 and found that average personal Pb-PM25 concentration was roughly three
times higher than outdoor Pb-PM2 5 concentration and two times higher than indoor Pb-PM2 5
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 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. In a pilot study in
Windsor, Ontario, Rasmussen et al. (2007) observed that personal exposure to Pb measured using a PM2 5
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monitor was roughly 40% higher than outdoor Pb concentration and 150% higher than indoor Pb
concentration. Pekey et al. (2010) measured indoor and outdoor trace element composition of PM25 and
PMio in Kocaeli, an industrial region of Turkey, and found that average airborne Pb concentrations were
higher outdoors than indoors for both PM2 5 and PMi0 during summer and for PMi0 during winter, but that
indoor Pb concentration was higher than outdoor Pb concentration for PM2 5 during winter. The indoor-to-
outdoor ratio of Pb in PM varied by environment; it tended to be less than one, but the ratio varied from
one microenvironment to another. The three studies that included personal samples recorded
measurements that were consistently higher than indoor or outdoor levels. It is likely that a number of
factors influenced the indoor-to-outdoor ratio of Pb in PM for these studies. These factors may have
included seasonal air exchange, which can vary as a function of window opening or air conditioning
usage, prevalence and strength of outdoor and indoor Pb sources, and size distribution of airborne Pb-
bearing PM.
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
PM2 5 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 3A-14 in the Appendix illustrates that Pb
exposures are typically well below the level of the NAAQS. The higher exposures occurred in a heavily
industrialized area with an incinerator and several industrial facilities including metal processing, cement,
petroleum refining, agriculture processing. Otherwise, exposures 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-bearing PM measurements
from several studies.
Study
Location
Pb Metric
Sampling
Period
Personal Pb
Indoor Pb
Outdoor Pb
Clayton etal.
(1999)
IL, IN, Ml, MN,
OH, Wl
Med. Pb-PM50
(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
Molnaretal.
(2007)
Stockholm,
Sweden
Avg. Pb-PM25
(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
etal. (2007)
Mexico City,
Mexico
Med. Pb-PM25
(ng/m3)
April-May,
2002
26
56
Puebla, Mexico
Med. Pb-PM25
(ng/m3)
April-May,
2002
4
4
Rasmussen et al.
(2007)
Windsor, Ontario,
Canada
Med. Pb-PM25
(mg/kg)
April, 2004
311
124
221
Pekey et al. (2010) Kocaeli, Turkey
Avg. Pb-PM25
(ng/m3)
May-June,
2006,
December,
2006-January
2007
Summer: 34
Winter: 85
Summer: 47
Winter: 72
Avg. Pb-PM10
(ng/m3)
May-June,
2006,
December,
2006-January
2007
Summer: 57
Winter: 125
Summer: 78
Winter: 159
4.1.1.2. Exposure to Lead in Soil and Dust
1 The 2006 AQCD (U.S. EPA. 2006) lists indoor Pb dust infiltrated from outdoors as a potential
2 source of exposure to Pb soil and dust. Outdoor soil Pb concentration may present a direct inhalation
3 exposure, or it can be tracked into homes to result in exposure to resuspended Pb PM or to Pb dust during
4 hand-to-mouth contact. A detailed description of studies of outdoor Pb concentration is provided in
5 section 3.6.1. Indoor measurements can reflect infiltrated Pb as well as Pb dust derived from debrided
6 paint, consumer products, or soil that has been transported into the home via foot traffic. Table 4-4
7 presents indoor Pb levels for several studies.
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Table 4-4. Measurements of indoor Pb dust from various studies.
Reference
Study Location
Metric (units)
Sample Site
Indoor Pb
Concentration
Caravanos et al.
(2006)
New York City, New York
Weekly dust loading
(ug/m*
Glass plate
Median: 52
Khoder et al. (2010) Giza, Egypt
Weeklv dust loading
(ug/m*
Glass plate
Median: 408
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
Gaitens et al. (2009) U.S. (nationwide)
Dust loading (^ig/m2) Smooth floor
Median: 1.7
Avg.: 4.4
Rough floor
Median: 5.6
Avg.: 16
Smooth windowsill
Median: 2.5
Avg.: 190
Rough windowsill
Median: 55
Avg.: 480
Mielke et al. (2001) New Orleans, Louisiana
Dust concentration
(mg/kg)
Multiple locations within homes
prepared for painting; sanded house
Range: <3-28,000
Multiple locations within homes
prepared for painting; scraped
house
Range: 7-1,200
Spalinger et al.
(2007)
Rural towns, Idaho
Dust concentration
(mg/kg)
Vacuum
Median: 120
Max: 830
Floor
Median: 95
Max: 1,300
Bunker Hill, Idaho
Superfund site
Vacuum
Median: 470
Max: 2,000
Floor
Median: 290
Max: 4,600
Several studies have demonstrated the infiltration of Pb dust into buildings. For example,
Caravanos et al. (2006) collected dust on glass plates at an interior location near an open window, a
sheltered exterior location, and an open exterior location for a two-year period in Manhattan, NY. Median
weekly dust loading was reported to be 52 (ig/m2 for the indoor site, 153 (ig/m2 for the unsheltered
outdoor site, and 347 |_ig/m2 for the sheltered outdoor site. This paper demonstrated the likely role of
outdoor Pb in influencing indoor dust Pb loading and indicated that under quiescent conditions (e.g., no
cleaning), the indoor dust Pb level might exceed EPA's hazard level for interior floor dust of 430 (ig/m2
(40 jxg/ft2). Khoder et al. (2010) used the same methodology to study Pb dust deposition in Giza, Egypt
and found a median weekly deposition rate of 408 |_ig/m2 and an exterior median deposition rate of 2,600
(ig/m2. In the latter study, Pb deposition rate correlated with total dust deposition rate (R=0.92), Cd
deposition rate (R=0.95), and Ni deposition rate (R=0.90). Statistically significant differences in Pb
deposition rates were observed between old and new homes (p<0.01) in the Khoder et al. (2010) study,
although the only quantitative information provided regarding home age stated that the oldest home was
22 years old when the study was performed in 2007. Khoder et al. (2010) found no statistically significant
difference between Pb loadings when segregating the data by proximity to roadways.
Residual Pb dust contamination following cleanup has been documented. For instance, Hunt et al.
(2008) performed tests where moderately elevated soil Pb 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
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no surface discoloration. The authors then used wet wipe samples to collect residual soil and estimate Pb
deposition and concentration. Elevated Pb loadings and concentrations were observed even after multiple
cleanings. 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 indicates that Pb-bearing fine
particles are not well captured by home cleaning. Johnson et al. (2009) surveyed the floors of 488 homes
in Syracuse, NY census tracts and found that the variability in Pb dust within homes was greater than the
variability between homes. A correlation between Pb dust loading on floors and the fraction of homes in
census tracts that were renter-occupied (partial R2 = 0.48; where total number of homes is the sum of
owner-occupied, renter-occupied, and vacant homes) was also observed in this study. Yu et al. (2006)
dissolved Pb dust, obtained from vacuuming carpet samples and found that 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 dust in homes
of children ages 12-60 months. The median value of Pb dust 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 value of 5.6
(ig/m2 (mean: 16 (.ig/m2). Floor Pb dust level 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 on smooth windowsills was 25 (ig/m2 (mean: 190 (.ig/ni2). When windowsills were not
smooth, the median Pb dust level was 55 (ig/m2 (mean: 480 |_ig/m2). Windowsill Pb dust level was also
found to be 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. It
was nearly significantly associated (p = 0.076) with deterioration of outdoor paint when homes were built
prior to 1950. Dust Pb levels were found by Egeghy et al. (2005) to be significantly associated with the
log-transform of soil Pb levels, 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).
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
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.
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Dust Pb concentrations have also been reported for homes in the vicinity of historical metals
mining and smelting sources. For example, Spalinger et al. (2007) measured Pb 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. 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
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 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 an industrial site are at much greater risk of exposure to Pb dust compared to the general
population.
4.1.1.3. 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, pesticides via vegetables, urban gardening, dietary supplements,
tobacco, cultural food sources, and breastfeeding.
Drinking Water
Differences in sources and transport of drinking water may cause variation in 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 that have higher Pb concentrations due to
leaching from the glass, the median Pb concentration was 8.5 ng/L (range < 1 to 761 ng/L). This was
significantly less than the EU, Health Canada and WHO drinking water maximum admissible
concentration of 10 (ig/L. 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
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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 lead to Pb exposures 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 would have implications for V release into drinking water. The stability of vanadinite in the
presence of monochloramine is also not known, and this might have implications for both Pb and V
release into drinking water.
In recent years, drinking water treatment plants in many municipalities have switched from using
chlorine to other disinfecting agents because their disinfection byproducts may be less carcinogenic.
However, chloramines are more acidic than chlorine and can increase Pb solubility (Raab et al.. 1991) and
increase Pb concentrations in tap water. For example, after observing elevated Pb concentrations in
drinking water samples, Kim and Herrera (2010) observed Pb oxide corrosion scales potentially 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 & 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 fluorosilic 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. (2007) modeled blood Pb levels among children living in Wayne County, NC as a
function of household age, use of chloramines and other covariates. It was found that blood Pb level was
significantly associated with the year the home was built (p <0.001), use of chloramines (p <0.001), and
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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. (2007) found
that significance increased with the age of the home, based on the assumption that older homes will have
more Pb pipes and Pb solder connecting the pipes. However, the study did not control for the presence of
Pb paint in the dwellings, so it is difficult to distinguish the effect of Pb pipes from the presence of paint
from that variable.
Several chemical mechanisms may contribute to release of Pb during use of chloramine
disinfection agents. Edwards and Dudi (2004) hypothesized that Pb leaching through chloramines
exposure through the breakdown of brass alloys and solder containing Pb. 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. (2009) 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 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
(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
Dietary Pb has the potential to emanate from soil Pb used for agricultural purposes. 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. Fernandez et al. (2007; 2010; 2008) 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 Pb were distributed bimodally with peaks around 2-20
|im and 50-100 |im (Fernandez et al.. 2010). For the agricultural plot, Pb concentration was constant
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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 contamination of crops can originate with treatment of crops. For
example, compost produced from wastewater sludge has the potential to add Pb to crops. Cai et al. (2007)
demonstrated that production of compost from sludge enriched the Pb content by 15-43% prior to its
application. Chen et al. (2008) observed that the median concentration of Pb in California crop soil
samples was 16.2 mg/kg (range: 6.0-62.2 mg/kg). Factor analysis suggested that the soil was enriched
with Pb in crop soils in the Oxnard/Ventura region. Chen et al. (2008) further observed that in three of the
seven California agricultural regions sampled, concentrations of Pb increased following addition of
fertilizer, but sub-proportionally to the increase in P and Zn indicators of fertilizer. In four regions, there
was no increase of Pb at all. Furthermore, Tu et al. (2000) observed a decrease in Pb fraction with
increasing P application. Nziguheba and Smolders (2008) also surveyed phosphate-based fertilizers sold
in European markets to determine the contribution of these fertilizers to heavy metal concentrations in
agricultural products. They observed a median 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) observed that the amount of Pb applied via fertilizers was only 2.6% of that from atmospheric
deposition. Although Pb in on-road vehicle gasoline has been phased out in the U.S., this remains a
relevant issue in the U.S. because some imported crops that 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, where leaded gasoline is legal. Rankin et al. (2005) observed that
the ratios of 207Pb to 206Pb and 208Pb to 207Pb were found to be 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 was observed to have 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.
Uptake of Pb by plants growing in contaminated soil has been 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
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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 were
demonstrated to have 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; these references also suggested high transfer from
roots to shoots among root crops, with shoots having roughly four times higher Pb bioaccumulation than
roots.
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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
GM
GSD
AM
SD
Min
Max
All
210
2.0%
14
63%
290%
0.015% 2500%
Cereals
Grain
All
1.0%
3.6
1.6%
0.19%
Straw
All
2.3%
3.5
4.0%
0.51%
9.6%
Maize
Grain
All
0.12%
2.3
0.17%
0.14%
0.052% 0.38%
Straw
All
0.28%
6.6
0.85%
1.3%
0.060% 2.3%
Rice
Grain
All
2.2%
1.4%
1.2%
3.2%
Leafy Vegetables
All
31
S.0%
13
210%
610%
0.32%
2500%
Sand
7.3%
1.5
3.3%
4.9%
11%
Loam
82%
1.0
82%
3.5%
79%
86%
Clay
4.1
5.1%
0.41%
12%
Non-Leafy Vegetables Fruits
All
1.5%
26
78%
170%
0.15%
390%
Shoots
All
0.42%
0.58%
1.17%
Legumes
Pods
All
17
0.53%
12
34%
120%
0.046% 490%
Sand
0.27%
3.2
0.42%
0.34%
0.065% 0.89%
Loam
0.14%
4.4
0.42%
0.34%
0.065% 0.89%
Clay
0.080% 1.0
0.33%
0.47%
0.046% 1.0%
Shoots
All
0.080%
Root Crops
Roots
All
27
1.5%
16
41%
0.024% 330%
Sand
6.4%
1.6
7.0%
3.4%
4.2%
12%
Loam
2.3%
4.7
0.50%
0.68%
0.024% 1.7%
Shoots
All
12
6.3%
15
250%
570%
0.30%
16%
Tubers
Tubers
All
30
0.15%
7.4
9.1%
48%
0.015% 260%
Sand
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
0.77%
2.6
1.0%
0.60%
0.15%
1.7%
Leaves
All
25%
Grasses
All
17
31%
36%
22%
11%
100%
Natural Pastures
All
34
92%
4.8
23%
29%
0.22%
100%
Leguminous Fodder
All
1.6%
All Cereals
All
20
0.43%
4.7
1.1%
1.4%
0.052%
Sand
0.61%
5.3
1.3%
1.3%
0.052% 3.2%
Loam
0.17%
3.9
0.53%
1.1%
0.059% 3.2%
Clay
0.90%
4.0
1,
0.22%
Pastures/Grasses
All
51
14%
4.2
27%
27%
0.22%
100%
Fodder
All
24
2.5%
12
130%
420%
0.060% 1600%
Sand
4.5%
2.3
5.6%
4.0%
1.6%
11%
Clay
0.82%
5.7
2.7%
4.6%
0.16%
9.6%
Source: Used with permission from Elsevier Publishers, Vandenhove et al. (2009").
Findings from Pb uptake studies have implications for urban gardening if urban soils may be
contaminated with Pb, as described in Section 4.1.1.2. For instance, Clark et al. (2006) tested the soil in
103 urban gardens in two Boston neighborhoods. They found that Pb-based paint contributed 40-80% of
measured Pb in the urban garden soil samples, with the rest coming from historical gasoline emissions.
Furthermore, Clark et al. (2006) estimated that Pb consumption from urban gardens can be responsible for
up to 25% of exposure to Pb in 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.
In addition to uptake of Pb through the roots of a plant, deposition of airborne Pb-bearing PM can
also contribute to human dietary Pb exposures, as described in the 2006 Pb AQCD (2006). In a recent
study, 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.
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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.
Fish
Accumulation in fish could also lead to human exposure to Pb. 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 Pb concentrations ranged from 580 to 6,500 mg/kg. In
total, 65% of those surveyed 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 content varied with respect to species of fish
but were found to be elevated in some species. Pb concentrations in fish were higher 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) compared with fish located further from the mines (Schmitt et al.. 2007). 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. There has
also been evidence of elevated Pb concentration within large game from mining areas (Reglero et al..
2009).
4.1.1.4. 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. It was not stated explicitly, but it is likely that Miller et
al. (2010) measured the PM as TSP because the Occupational Safety and Health Administration (OSHA)
permissible exposure limit (PEL) for Pb is based on TSP rather than a smaller size cut, and the OSHA
PEL was used for comparison. Concentrations measured by personal samples ranged from 2 to 6 (ig/m3,
and concentrations from area samples ranged from 4 to 14 (ig/m3. The OSHA PEL is 5 (ig/m3. In steel
production, sintering was found to be the largest source of airborne Pb exposure in a survey of operations
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(Sammut et al.. 2010). with Pb enrichment in PM reported to be 20,000 mg/kg, although total PM
concentration, reported to have 75% below 2.5 |_im diameter, was 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.
4.1.1.5. 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 (Levin
et al.. 2008; Miodovnik & Landrigan. 2009).
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Table 4-6. Pb content in various consumer products
Product
Category
Product
Location of
Purchase
Pb Content (units)
Reference
Cissus 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 etal (2010)
Maiva syivestris
Turkey
1.1-2.0 mg/kg
Hifsomnez et al
(2009)
Yugmiiihwang-tang, Boiungigki-tang, Sibieondaebo-
tang, Kuibi-tang, Ojeogsan
Korea
7.9x10 " to 2.5x10"J mg/kg
body weight/day
Kim et al (2009)
Alternative
and
Traditional
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)
Medicines
B-Success 28, Operation Sweep, Aloe Vera Plus
Bitter Aloes, Zarausmacine, Virgy-Virgy Computer
Worm-Expeller, Dorasine Powder, Sexual Energy,
U&DEE Infection Cleansing Powder, U&DEE Sweet
Bitter, Natural Power Stone, Charma 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
Nigeria
925-27,000 |jg
Obi et al (2006)
Shell of Hen's Egg
India
14 mg/kg
Sharma etal (2009)
Berberis (B. aristata, B. chitria, B. asiatica, B.
iyceum), 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 a I
(2006)
Greta powder
U.S. (California)
770,000 ppm
CDC (2002)
Candy
Tamarind Candy
U.S.
(Oklahoma)
Product: 0.15-3.61 mg/kg
Stems: 0.36-2.5 mg/kg
Wrappers: 459-27,125
mg/kg
Lynch et al (2000)
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 (2009)
Cosmetics
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
United Kingdom
0.15-1.56 mg/kg
McNeill et al (2006)
Tobacco
Cigarette Tobacco (21GPb concentrations)
Pakistan
Activity conc.: 7-20 Bq/kg
TahirandAlaamer
(2008)
Red and yellow painted toy vehicles and tracks
Brazil
500-6,000 mg/kg
Godoi et al (2009)
Toys
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
Non-yellow: 94 mg/kg
Greenway and
Gerstenberger (2010)
Soft plastic toys
India
Average (by city): 21-280
mg/kg
Kumar and Pastore
(2007)
Toy necklace
U.S.
388,000 mq/kq
Meyer et al (2008)
Soft plastic toys
Nigeria
2.5-1,445 mg/kg
Omolaoye et al
(2010b)
Vitamins
Vitamins for young children, older children, and
pregnant or lactating women
U.S.
Average:
Young children: 2.9 ^g/day
Older children: 1.8 ^g/day
Pregnant and lactating
women: 4.9 yg/day
Mindaket al. (2008)
aHoodia gordonii, from Eastern Cape, South Africa Euterpe oleracea from Ninole Orchard, Ninole, Hawaii
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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. 2006). 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, 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 AQCD also presented dermal absorption of
inorganic and organic Pb compounds, which is generally considered to be much less than by inhalation or
ingestion. No recent literature has advanced our knowledge of dermal absorption of Pb beyond that which
was included in the 2006 AQCD. 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+).
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 & Rov. 1994). Larger
particles (>2.5 |_un) 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 (Chamberlain et al.. 1978; Morrow et al.. 1980; M. B. Rabinowitz et al.. 1977).
Absorption half-times (ti/2) have been estimated for radon decay progeny in adults who inhaled aerosols
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220 222
of Pb and bismuth isotopes generated from decay of Rn or Rn. 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 &
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 attempted to quantify the bioavailability of Pb in atmospheric PM, although
different laboratory techniques are used throughout the literature, as described in Section 3.4. Unlike the
bioavailability methods described in Section 4.2.1.2, many of these in vitro assays have not been
validated with in vivo models. Therefore, any in vitro method can only be a simplistic surrogate of the
complex uptake of Pb from the gastrointestinal (GI) tract. Despite this limitation, the studies mentioned
below are the only ones to evaluate atmospheric Pb.
In a study of PMi0 and PM25 samples from downtown Vienna, Austria, Falta et al. (2008) used
synthetic gastric juice to investigate the bioavailability of heavy metals including Pb. The rationale was
that inhaled PM in the 2.5-10 (.un size range are mostly deposited in the tracheal and bronchial regions of
the lung from where they are transported within hours by mucociliary clearance, i.e., they are mainly
swallowed. In contrast, the <2.5 |_im particles are deposited in the pulmonary alveoli where they can stay
for months to years. The study aimed to determine the bioavailability of the 2.5-10 (.im PM. It is important
to note that they do not isolate the 2.5-10 |_im size range; instead, they infer the characteristics from the
difference between the PM2 5 and PMi0 fractions. The Pb concentrations associated with the two fractions
were almost identical, as was the percentage extracted by synthetic gastric juice (86% and 83% Pb for
PM2 5 and PMi0 fractions, respectively). The mean daily bioavailable mass was calculated to be 16 ng for
the PM2 5.10 size range. Since the quantitative clearance of these particles to the stomach was assumed,
this value represents an upper estimate for the amount of bioavailable Pb. Niu et al. (2010) determined the
bioavailability of Pb 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 PM smaller
than 10 (jm were 8,800 and 7,800 mg/kg for the two different locations. The bioavailability was based on
ammonium acetate extractability and it was found that, within the fine and nano-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 PM, with
percent bioavailability decreasing with increasing PM size. This indicated that Pb was potentially most
bioavailable in the nano-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
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|_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 bioavailable if ingested compared with PbO or Pb-acetate because they are less soluble.
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 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
203 203
experiment conducted with Pb tetramethyl Pb, 51% of the inhaled Pb 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 (i.e., piston-
engine aircraft). Mahaffey (1977) estimated that 40% of inhaled Pb 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 at a concentration of 0.001 |ig/m3 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 (Pocock et al.. 1983; J. Sherlock et al.. 1982; J. C. Sherlock et al.. 1984; J. C.
Sherlock & Ouinn. 1986). 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 (Heard & Chamberlain. 1982; James et al.. 1985;
Maddaloni etal.. 1998; M. B. Rabinowitz et al.. 1980; Watson et al.. 1986). The absence of food in the GI
tract increases absorption of water-soluble Pb in adults. Reported estimates of soluble Pb absorption range
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from 26 to 70% in fasted adults (Blake et al.. 1983; Heard & Chamberlain. 1982; James et al. 1985;
Maddaloni etal.. 1998; M. B. Rabinowitz et al.. 1980). Reported fed:fasted ratios for soluble Pb
absorption in adults range from 0.04 to 0.2 (Blake et al.. 1983; Heard & Chamberlain. 1982; James et al..
1985; M. B. Rabinow itz 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 (Alexander etal.. 1974; Ziegler et al.. 1978).
Experimental studies provide further evidence for greater absorption of Pb from the gut in young animals
compared to adult animals (Aungst et al.. 1981; Forbes & Reina. 1972; kostial et al.. 1978; Pounds et al..
1978). 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).
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 (Mahaffev & Annest. 1986;
Marcus & Schwartz. 1987; Schell et al.. 2004). 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; Barton. Conrad. Nub v. et al..
1978; Morrison & Ouarterman. 1987).
The effects of iron nutritional status on blood Pb include changes in blood Pb concentrations in
association with genetic variation in genes involved in iron metabolism. For example, genetic variants in
the hemochromatosis (HFE) and transferrin genes are associated with higher blood Pb concentrations in
children (Hopkins et al.. 2008). In contrast, HFE gene variants are associated with lower bone and blood
Pb levels in elderly men (Wright et al.. 2004).
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 & Mann. 1983; Heard & Chamberlain. 1982). A
similar effect of calcium occurs in rats (Barton. Conrad. Harrison, et al.. 1978). 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; Mahaffev et al.. 1986; Schell et al.. 2004; Ziegler et 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 (Bronner et al..
1986; Fullmer & Rosen. 1990). However, animal studies have also shown that multiple aspects of Pb
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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 & Fung. 1985V 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.. 2004V 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. Relative bioavailability (RBA) of Pb in
soils and dust has been most extensively studied in animal models. Relative bioavailability is the ratio of
the absorbed fraction (AF) of ingested dose of soil Pb to that of a water-soluble form of Pb (e.g., Pb-
acetate) that is considered to be completely bioaccessible (e.g., RBA = AFSoiiPb/AFpb.acetate)- In typical
studies, the absorbed fraction of the Pb dose is estimated based measurements of Pb measured in blood
and/or other tissues (e.g., kidney, liver, bone) after dosing . Gastric function of swine is thought to be
sufficiently similar to that of humans to justify use of swine as a model for assessing RBA of Pb in soils
(Casteel et al.. 1997; Casteel etal.. 2006; Juhasz et al.. 2009; U.S. EPA. 2007a; Weis & Lavelle. 1991).
Other practical advantages of the swine model over rodent models have been described, and include:
absence of coprophagia; ease with which Pb dosing can be administered and controlled; and higher
bioavailability of soluble Pb (e.g., Pb-acetate) in swine, which is more similar to humans than rats (D. M.
Smith et al.. 2009). Relative bioavailability 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 |im) tends to be lower than
smaller particles (Barltrop & Meek. 1979; Healv et al.. 1992).
Collectively, published studies conducted in swine have provided 39 estimates of Pb RBA for 37
different soil or "soil-like" test materials (Bannon et al. 2009; Casteel et al.. 2006; Marschner et al.. 2006;
D. M. Smith et al.. 2009). 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 aNIST 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
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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 4-2 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 overtime. 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%
(Freeman et al.. 1996; Freeman et al.. 1992; Freeman et al.. 1994; D. M. Smith et al.. 2008. 2009).
HIGH
0
Group
Source: Casteel et al. (2006).
Figure 4-2. Estimated relative bioavailability (RBA, compared to Pb-
acetate) of ingested Pb in mineral groups, based on results
from juvenile swine assays.
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):
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RBA = (0.878 • 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. 2007b'). 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 = 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 of in vitro assays applied to the same soil test materials have
found considerable variability in IVBA estimates (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). Given the dependence of IVBA outcomes on assay conditions, in
vitro assays used to predict in vivo RBA should be evaluated against in vivo RBA estimates to
quantitatively assess uncertainty in RBA predictions (U.S. EPA. 2007b).
Absorption of Pb in house dusts has not been rigorously evaluated quantitatively in humans or in
experimental animal models. The RBA for paint Pb mixed with soil has been reported to be
approximately 72% (95% CI: 44, 98) in juvenile swine, suggesting that paint Pb dust maybe highly
bioavailable (Casteel et al.. 2006). The same material yielded a bioaccessibility value (based on IVBA
assay) of 75% (Drexler & 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 would be that a validated IVBA
methodology for soil would perform well for predicting RBA of interior dust, this has not actually been
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experimentally confirmed. Factors that may affect in vitro predictions of RBA of interior dust Pb could
include particle size distribution of interior dust Pb and the composition of the dust matrix, which may be
quite different from that of soil.
Other estimates of "bioavailability" of Pb exposure samples are derived from less validated in vitro
methods. Roussel et al. (2010) estimated that 63% of soil Pb is bioavailable in the stomach, while 39% is
bioavailable in the intestines, using different acidities of solutions to simulate acids in the digestive
system. Yu et al. (2006) dissolved Pb dust, obtained from vacuuming carpet samples, into simulated
gastric and intestinal acids. 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% bioavailability 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 (M. B. Rabinowitz et
al.. 1976). The highest soft tissue concentrations in adults occur in liver and kidney cortex (Barry. 1975;
Gerhardsson et al.. 1986; Gerhardsson et al.. 1995; Gross et al.. 1975; Qldereid et al.. 1993). 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
(Bergdahl. Grubb. et al.. 1997; Bergdahl et al.. 1998; Bergdahl etal.. 1999; Hernandez-Avila et al.. 1998;
Manton et al.. 2001; Schutz et al.. 1996; D. Smith et al.. 2002). S-aminolevulinic acid dehydratase
(ALAD) is the primary binding ligand for Pb in erythrocytes (Bergdahl. Grubb. et al.. 1997; Bergdahl et
al.. 1998; Sakai et al.. 1982; Xie et al.. 1998). 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. Grubb. et al.. 1997; Bergdahl etal.. 1996; Bergdahl et al.. 1998). Of the three
principal Pb-binding proteins identified in RBCs, ALAD has the strongest affinity for Pb (Bergdahl et al..
1998) and appears to dominate the ligand distribution of Pb (35 to 84% of total erythrocyte Pb) at blood
Pb levels below 40 (.ig/dL (Bergdahl et al.. 1996; Bergdahl et al.. 1998; Sakai et al. 1982). Pb binding to
ALAD is saturable; the binding capacity has been estimated to be -850 |_ig/dL RBCs (or -40 j^ig/dL whole
blood) and the apparent dissociation constant has been estimated to be -1.5 (ig/L (Bergdahl et al.. 1998).
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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 (Barbosa.
Ramires. et al.. 2006; Bergdahl. Schutz. et al.. 1997; Bergdahl et al.. 1998; Bergdahl et al.. 1999; DeSilva.
1981; C. Jin et al.. 2008; Kang et al.. 2009; Manton et al.. 2001; D. Smith et al.. 2002). An example of this
is shown in Figure 4-3. Saturable binding of Pb to RBC proteins has several important consequences. As
blood Pb increases and the higher affinity binding sites for Pb in RBCs become saturated at
approximately 40 (ig/dL blood, 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.
Typically, at blood Pb concentrations <100 (ig/dL, only a small fraction (<1%) of blood Pb is found
in plasma (DeSilva. 1981; Manton & Cook. 1984; Marcus. 1985). 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; Ong & Lee. 1980). Pb in serum that is not bound to protein
exists largely as complexes with low molecular weight sulfhydryl compounds (e.g., cysteine,
homocysteine) and other ligands (Al-Modhefer et al.. 1991).
2.0
1.5
1.0
T3
O)
=L
.Q
CL
CO
E
CO
CO
Q- 0.5
0.0
oAdults •Children
*
0
20
40
60
80
100
Blood Pb (|jg/dL)
Source: Adapted, with permission from Elsevier Publishing and the Finland Institute of Occupational Health, from Bergdahl et al. (1997: 1999).
Figure 4-3. Plot of blood and plasma Pb concentrations in measured in
adults and children.
As shown in Figure 4-3, the limited binding capacity of Pb binding proteins in RBCs produces a
curvilinear relationship between blood and plasma Pb concentration. The limited binding capacity of RBC
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binding proteins also confers, or at least contributes, to a curvilinear relationship between Pb intake and
blood Pb concentration. A curvilinear relationship between Pb intake and blood Pb concentration has been
observed in children (Lacevetal.. 1985; Ryu et al.. 1983; J. C. Sherlock & Ouinn. 1986). As shown in
Figure 4-4, the relationship becomes pseudo4inear at relatively low daily Pb intakes (i.e., <10 (ig/day/kg)
and at blood Pb concentrations <25 (ig/dL.
45
40
30
_i
T3
CD
=L
0
100
200
300
400
Pb Intake (|jg/day)
Source: Adapted, with permission from Taylor & Francis Publishing, from Sherlock and Quinn (1986).
Figure 4-4. Relationship between Pb intake and blood Pb concentration in
infants (n = 105, age 13 weeks, formula-fed). Data represent
mean and standard errors for intake; the line is the regression
model (blood Pb = 3.9 + 2.43 (Pb intake [pg/week]1'3).
Figure 4-5 shows the predicted relationship between quasi-steady state blood and plasma Pb
concentrations in a 4-year old child using the ICRP model (ICRP. 1994; Leggett. 1993; Pounds & Leggett.
1998). 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
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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.
0.60
0.50 -
1) 0.40 -
¦Q
0.30 -
£
ra 0.20 -
CL
0.10
0.00
10 20 30 40
Blood Pb (iig/dL)
50
Blood
Plasma
o) 40
0.6 3
0-4 E
m 20
~i 1 r
100 200 300
Intake (tig/day)
400
Figure 4-5. Simulation of quasi-steady state blood and plasma Pb
concentrations in a child (age 4 years) associated with varying
Pb ingestion rates. Simulation based on ICRP Pb biokinetics
model (Leqqett. 1993).
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Studies conducted in swine provide additional evidence in support of RBC binding kinetics
influencing distribution of Pb to tissues. In these studies, the relationship between the ingested dose of Pb
and tissue Pb concentrations (e.g., liver, kidney, bone) was linear, whereas, the relationship between dose
and blood Pb was curvilinear with the slope decreasing as the dose increased (Casteel et al.. 2006).
Saturable binding of Pb to RBC proteins also contributes to a curvilinear relationship between urinary Pb
excretion and plasma Pb concentration (Section 4.2.3) (Bemdahl. Schutz. etal.. 1997; Besser et al.. 2008).
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
calcellous). The proportion of cortical to trabecular bone in the human body varies by age, but on average
is about 80 to 20 (ICRP. 1973; Leggett. 1993; O'Flahertv. 1998).
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 (Bros et al.. 1986; Miyake. 1986;
Verbeeck et al.. 1981). Several intracellular kinetic pools of Pb have been described in isolated cultures of
osteoblasts and osteoclasts which appear to reflect physiological compartmentalization within the cell,
including membranes, mitochondria, soluble intracellular binding proteins, mineralized Pb (i.e.,
hydroxyapatite) and inclusion bodies (Long et al.. 1990; Pounds & 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 & 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 (Barry. 1975; Barry & Connolly. 1981; Gross et al..
1975; Park. Mukherjee. et al.. 2009; Schroeder & Tipton. 1968). The cortical and trabecular bones have
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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 (M. B. 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%) (ICRP. 1973; Leggett. 1993; O'Flahertv. 1995). 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 (Cristv. 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 (Barry. 1975; Gerhardsson et al.. 1986;
Gerhardsson et al.. 1995; Gross et al.. 1975; Oldereid et al. 1993). 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 & Fow ler. 1989; Fow ler. 1989). The Pb-binding proteins in
rat are cleavage products of a2\x globulin, a member of the protein superfamily known as retinol-binding
proteins that are generally observed only in male rats (Fowler & 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 (D. R. 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 (Eaton et al. 1980; Waalkes & klaassen. 1985).
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 (Azar et al.. 1973; Casteel et al.. 1997; Casteel et al.. 2006; D. M. Smith et al.. 2008). In
contrast to Pb in bone, which accumulates Pb with continued exposure in adulthood, concentrations in
soft tissues (e.g., liver and kidney) are relatively constant in adults (Barry. 1975; Treble & Thompson.
1997). reflecting a faster turnover of Pb in soft tissue relative to bone.
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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. These ratios range from about 0.6 to 1.0 at the time of delivery (Carbone et al.. 1998;
Gover. 1990; Graziano et al.. 1990; B. Gulson. Jameson, et al.. 1998; kordas et al.. 2009; Manton. 1985).
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 (B. Gulson et al.. 1999).
Transplacental transfer of Pb may be facilitated by an increase in the plasma/blood Pb
concentration ratio during pregnancy (Lamadrid-Figueroa et al.. 2006; Montenegro et al.. 2008).
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 (Franklin et al.. 1997; O'Flahertv. 1998).
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 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 et al.. 1979). The kinetics of 203Pb in blood
showed an initial declining phase during the first 4 hours (tetramethyl Pb) or 10 hours (tetraethyl Pb) after
the exposure, followed by a reappearance of radioactivity back into the blood after -20 hours. The high
level of radioactivity initially in the plasma indicates the presence of tetraalkyl/trialkyl Pb. The
subsequent rise in blood radioactivity, however, probably represents water-soluble inorganic Pb and
trialkyl and dialkyl Pb compounds that were formed from the metabolic conversion of the volatile parent
compounds (Heard et al.. 1979).
Alkyl Pb compounds undergo oxidative dealkylation catalyzed by cytochrome P450 in 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; Kehoe. 1961a. 1961b. 1961c; M. B. Rabinowitz et al.. 1976). Absorbed Pb is
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excreted primarily in urine and feces, with sweat, saliva, hair, nails, and breast milk being minor routes of
excretion (Chamberlain et al.. 1978; Griffin et al.. 1975; Hursh et al.. 1969; Hursh & Suomela. 1968;
Kehoe. 1987; M. B. Rabinowitz et al.. 1976).
Approximately 30% of Pb excretion during the first 20 days after exposure is due to urinary and
fecal losses (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. Schutz. et
al.. 1997) (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 has been shown to be
linear, the plasma Pb relationship to blood Pb concentration is curvilinear (as described in Section 4.2.2.1
and demonstrated in Figure 4-5). This 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 4.2.7).
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; Chamberlain et al.. 1978; Manton & Cook. 1984; Manton & M alio v.
1983), which corresponds to half-time for transfer of Pb from plasma to urine of 0.1-0.16 days for a 70-kg
adult who has a plasma volume of ~3 L. The rate of urinary excretion of Pb was less than the rate of
glomerular filtration of ultrafilterable Pb, suggesting that urinary Pb is the result of incomplete renal
tubular reabsorption of Pb in the glomerular filtrate (Araki et al.. 1986); although, net tubular secretion of
Pb has been demonstrated in animals (Vander et al.. 1977; Victory et al.. 1979). 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 (Araki et al..
1990; Berger et al.. 1990; Chamberlain et al.. 1978; Roster et al.. 1989; Manton & M alio v. 1983; M. B.
Rabinowitz et al.. 1973; Ryu et al.. 1983) (Diamond. 1992). 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 fccal:urinary excretion ratio (-0.5) was observed during days 2-14 following
intravenous injection of Pb in humans (Booker et al.. 1969; Chamberlain et al.. 1978; 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 ty2 =10 days), and accounts for 70% of the total plasma clearance (O'Flahertv.
1995).
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Organic Lead
Pb absorbed after inhalation of tetraethyl and tetramethyl Pb is excreted in exhaled air, urine, and
feces (Heard et al.. 1979). Fecal:urinary excretion ratios were 1.8 following exposure to tetraethyl Pb and
1.0 following exposure to tetramethyl Pb (Heard et al.. 1979). Occupational monitoring studies of
workers who were exposed to tetraethyl Pb have shown that tetraethyl Pb is excreted in the urine as
diethyl Pb, ethyl Pb, and inorganic Pb (Turlakiewicz & Chmielnicka. 1985; Vural & Diivdu. 1995; W.
Zhang et al.. 1994).
4.3. Lead Biomarkers
This section describes the biological measurements of Pb and their interpretation as indicators of
exposure or body burden. The focus is on blood Pb and bone Pb and the interplay between them, as these
are the most commonly measured biomarkers in recent epidemiologic and toxicological studies.
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.
Numerous mechanistic models of Pb biokinetics in humans have been proposed, and these are
described in the 2006 Pb AQCD (U.S. EPA. 2006) 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 (ICRP. 1994; Leggett. 1993; Pounds & Leggett. 1998).
This model was originally developed for the purpose of supporting radiation dosimetry predictions;
however, it has also been applied in Pb risk assessment (Abrahams et al.. 2006; Khourv & Diamond.
2003; Lorenzana et al.. 2005). Portions of the model have been incorporated into an All Ages Lead Model
(AALM) that is being developed by EPA (2005).
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
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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,
because it has more bone mass and may afford better measurement precision than the calcaneus. Bone
analysis methods have included 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). 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. 2006V
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 and a high correlation with results
obtained from K-shell methods (intraclass correlation = 0.65).
Bone Pb measurements are typically expressed in units of Pb/g bone mineral. This 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 (Hu et al.. 2007; Theppeang et al.. 2008).
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 (Tlegal & Smith. 1995). Limits of detection for Pb using AAS are on the order of
5-10 (ig/dL for flame AAS measurements and approximately 0.1 (ig/dL for flameless AAS measurements
(Flegal & Smith. 1995; NIOSH. 1994). A detection limit of 0.005 (.ig/dL has been achieved for Pb in
blood samples analyzed by GFAAS.
For measurement of Pb in plasma, ICP-MS provides sufficient sensitivity (Schutz et al.. 1996).
While the technique has been applied to assessing Pb exposures in adults, it has not received widespread
use in epidemiologic studies.
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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) (Scinicnriello et al.. 2007). Since the ALAD 1 and
ALAD2 alleles can be codominantly 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, noninstitutionalized 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 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 (Astrin et al..
1987; Bergdahl. Schutz. etal.. 1997; H.-S. Kim et al.. 2004; Perez-Bravo et al.. 2004; C. M. Smith et al..
1995; Wetmur. 1994; Wetmur. Lehnert. et al.. 1991). 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 would be consistent with the structural differences that result in greater
electronegativity of ALAD 1 compared to ALAD2 (Wetmur. 1994; Wetmur. Kava. et al.. 1991).
measurements of Pb binding affinity to ALAD 1 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).
Both analyses reported the greatest differences for ALAD2 compared to ALAD1 in highly exposed adults
and little difference among environmentally-exposed adults; large differences were also observed for
children at low exposures. However, there were 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, with blood Pb levels being higher in individuals with ALAD2
alleles (Mivaki et al.. 2009; Shaik & 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 ALAD 1 and ALAD2 carriers (Montenegro et al.. 2006; Rabstein et al.. 2008;
Wananukul et al.. 2006) or lower blood Pb levels for individuals with ALAD 1-2/2-2 (Chia et al.. 2006).
Analyses of serial blood Pb concentrations measured in longitudinal epidemiologic studies have
found relatively strong correlations (e.g., r = 0.5-0.8) between individual blood Pb concentrations
measured after 6-12 months of age (Dietrich et al.. 1993; McMichael et al.. 1988; Otto et al.. 1985; M.
Rabinowitz et al.. 1984; Schnaas et al.. 2000). These observations suggest that, in general, exposure
characteristics of an individual child (e.g., exposure levels and/or exposure behaviors) tend to be
relatively constant across age. However, a single blood Pb measurement may not distinguish between a
history of long-term lower-level Pb exposure from a history that includes higher acute exposures
(Mushak. 1998). This is illustrated in Figure 4-6. Two hypothetical children are simulated. Child A has a
relatively constant Pb intake from birth, whereas Child B has the same long-term Pb intake as Child A,
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1 but with a 1-year elevated intake beginning at age 24 months (Figure 4-6, upper panel). The absorption
2 fraction is assumed to be the same for both children. Blood Pb samples 1 and 5 for Child A and B, or 2
3 and 4 for Child B, will yield similar blood Pb concentrations (~3 or 10 (ig/dL, respectively), yet the
4 exposure contexts for these samples are very different. Two samples (e.g., 1 and 2, or 4 and 5), at a
5 minimum, are needed to ascertain if the blood Pb concentration is changing over time. The rate of change
6 can provide information about the magnitude of change in exposure, but not necessarily about the time
7 history of the change (Figure 4-6, lower panel). Time-integrated measurements of Pb concentration may
8 provide a means for accounting for some of these factors and, thereby, provide a better measure of long-
9 term Pb exposure.
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Child B
15-
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Child D
TJ
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O 10-
CQ
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Age (months)
Figure 4-6. Simulation of temporal relationships between Pb exposure
and blood Pb concentration in children.Child A and Child B
have a relatively constant basal Pb intake (pg/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 pg/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 (Leqqett. 1993).
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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 (Araki et al.. 1990; Fukiii et al.. 1999). 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 exposure history of the previous few months; thus, a
single urinary Pb measurement cannot distinguish between a long-term low level of exposure or a higher
acute exposure. Urinary Pb measurements would be expected to correlate with concurrent blood Pb.
Chiang et al. (2008) reported a significant, but relatively weak correlation between urinary Pb levels
(|ig/dg creatinine) and individual Pb intakes (|_ig/day) estimated in a group of 10- to 12-year-old children
((3: 0.053, R = 0.320, p = 0.02, n = 57). A contributing factor to the relatively weak correlation may have
been the temporal displacement between the urine sampling and measurements used to estimate intake,
which may have been as long as six 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 4-7).
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1 1 1 1 1 1 1 1
25 30 35 40 45 50 55 60 65 70
Age (year)
¦o
o
O)
n g>
5
o
E
Figure 4-7. Simulation of relationship between urinary Pb excretion and
body burden in adults.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 (Leqqett. 1993).
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4.3.4. Lead in Other Potential 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 as biomarkers of exposure or body burden, 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 (B. L. Gulson et
al.. 1997; M. B. Rabinowitz et al.. 1993). Limited studies have demonstrated moderate-to-high
correlations between tooth Pb levels and blood Pb levels (M. B. Rabinowitz. 1995; M. B. Rabinowitz et
al.. 1989V
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 (M. B. Rabinowitz et al.. 1993). This 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) was able to discern 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., (Fosse et al.. 1995; Tvinnereim et al.. 1997). For example, deciduous
tooth Pb in the enamel provides information about in utero exposure whereas that in dentine from the
same tooth provides information about postnatal exposure until the tooth exfoliates at about 6-7 years of
age.
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
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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; M. B. Rabinowitz et al.. 1976)
and has been explored as a possibly noninvasive approach for estimating Pb body burden (Gerhardsson et
al.. 1995; Wilhelm et al.. 1989; Wilhelm et al.. 2002). Hair Pb measurements are subject to error from
contamination of the surface with environmental Pb and contaminants in artificial hair treatments (i.e.,
dyeing, bleaching, permanents) and are a relatively poor predictor of blood Pb concentrations, particularly
at low levels (<10-12 (.ig/dL) (Campbell & Toribara. 2001; G. Drasch et al. 1997; Esteban et al.. 1999; I
L. Rodrigues et al.. 2008). 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, 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 AQCD (U.S. EPA. 2006). Earlier reports suggested a relatively strong correlation
between salivary Pb concentration and blood Pb concentration (Brodeur et al.. 1983; Omokhodion &
Crockford. 1991; P'an. 1981); however, more recent assessments have shown relatively weak or
inconsistent associations (Barbosa. Heloisa. et al.. 2006; Costa de Almeida et al.. 2009; Nriagu et al..
2006). The differences in these outcomes may reflect differences in blood Pb concentrations, exposure
history and/or dental health (i.e., transfer of Pb between dentin and saliva) and possibly methods for
determining Pb in saliva. Barbosa et al. (2006) 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 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.
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4.3.4.4. Serum 5-ALA and ALAD
The association between blood Pb and blood ALAD activity and serum 5-aminolevulinc acid (5-
ALA) levels was recognized decades ago as having potential use as a biomarker of Pb exposure
(Hernberg et al.. 1970; Mitchell et al.. 1977). 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 (Opler et al..
2004; Qpler et al.. 2008).
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). A slower phase becomes evident with longer
observation periods following a decrease in exposure. 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.
Bone Pb stores can contribute 40-70% to blood Pb (B. Gulson et al.. 1995; Manton. 1985; D. R.
Smith et al.. 1996). 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 is dependent on age and
intensity of exposure. Slow bone volume compartments are much more labile in infants and children than
in adults as reflected by half-times for movement into the plasma (e.g., cortical ti/2 = 0.23 years at birth,
1.2 years at 5 years of age, 3.7 years at 15 years of age, and 23 years in adults; trabecular Xm = 0.23 years
at birth, 1.0 years at 5 years of age, 2.0 years at 15 years of age, and 3.9 years in adults) (Lcggett. 1993).
Children who have been removed from a relatively brief exposure to elevated environmental Pb exhibit
faster slow-phase kinetics than children removed from exposures that lasted several years, with half-times
of 10 and 20-38 months, respectively (Manton et al.. 2000). The longer half-times measured under the
latter conditions reflect the contribution of bone Pb stores to blood Pb following a change in exposure.
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 et al.. 1997; Inskip et al.. 1996; Kehoe. 1987; O'Flahertv et al.. 1982; D. R. Smith et al.. 1996).
even for exposures that occurred during childhood (F. E. 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.
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There is a stronger relationship between patella Pb and blood Pb than tibia Pb and blood Pb
(Hernandez-Avila et al.. 1996; Hu et al.. 1996; Hu et al.. 1998; Park. Mukheriee. et al.. 2009). Hu et al.
(1998) suggest that trabecular bone is the predominant bone type providing Pb back into circulation under
steady-state and pathologic conditions. The stronger relationships between blood Pb and trabecular Pb
compared with cortical bone is probably associated with the larger surface area of trabecular bone
allowing for more Pb to bind via ion exchange mechanisms and more rapid turnover making it more
sensitive to changing patterns of exposure.
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 4-8
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 (|_ig Pb/day) followed by an abrupt decline in intake. The
figure illustrates several important general concepts about the relationship between Pb in blood and bone.
While blood Pb approaches a quasi-steady state after a period of a few months with a constant rate of Pb
intake (as demonstrated by the vertical dashed line), Pb continues to accumulate in bone with continued
Pb intake after the quasi-steady state is achieved in blood. The model also predicts that the rate of release
of Pb from bone after 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 in to the bone mineral matrix.
Empirical evidence in support of this comes from longitudinal studies in which relatively high
correlations (r = 0.85) were found between concurrent (r = 0.75) or lifetime average blood Pb
concentrations (r = 0.85) and tibia bone Pb concentrations (measured by XRF) in a sample of children in
which average blood Pb concentrations exceeded 20 j^ig/dL; the correlations was much weaker (r <0.15)
among children who had average blood Pb concentration <10 (ig/dL (Wasserman et al.. 1994).
Two alternative blood Pb metrics depicted in Figure 4-8 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 (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
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1 event than prior to or following the event. Following cessation of exposure, the time-integrated blood Pb
2 and body burden diverge. This is expected, as the time-integrated blood Pb curve is a cumulative function
3 which cannot decrease over time and bone Pb levels will decrease with cessation of exposure.
4 The time-integrated blood Pb concentration will be a better reflection of the total amount of Pb that
5 has been absorbed, than the body burden at any given time. The time-integrated blood Pb concentration
6 will also reflect cumulative Pb absorption, and cumulative exposure if the absorption fraction is constant.
7 This is illustrated in the hypothetical simulations of an exposure event experienced by a child (Figure 4-
8 9). This pattern is similar for adults.
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8 -
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D)
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Q_
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'' /\ — -Bone
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/ ' \
/ N\
// « xV.
1/ \ —:
N ' ^ .
'/
£ •
-- 2.5
3.0
-- 2.0
-- 1.5
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Age (year)
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oa
o
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—Absorption
O
c
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T3
O
o
30
20
c
i 10
E
0
0
0 2 4 6 8 10
Age (year)
Figure 4-9. Simulation of relationship between time-integrated blood Pb
concentration and cumulative Pb absorption in children. 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 (Leggett,
1993).
4.3.5.2. Adults
In adults, where a relatively large fraction of the body burden residing in bone has a slower
turnover compared to blood, a constant Pb uptake (or constant intake and fractional absorption) gives rise
to a quasi-steady state blood Pb concentration, while the body burden continues to increase over a much
longer period, largely as a consequence of continued accumulation of Pb in bone. This pattern is
illustrated in Figure 4-10 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 4-10 reveals that the accumulation and elimination phases of blood Pb kinetics are not
symmetrical; elimination is slower than accumulation as a result of the gradual release of bone Pb stores
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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 4-10, 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 (Fleming et al.. 1997; Gerhardsson et al.. 1995V 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. This illustrates 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, bone Pb will increase with increasing duration of exposure and,
therefore, with age. In contrast, blood Pb 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).
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10
_ 8 -
6
= 4
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/ \ — • Bone
i /
' /
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i ' sV
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i *
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'
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300 -
250 -
200 -
150 -
B ood
Bone
DO
o
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(D
O
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Age (year)
Figure 4-10. Simulation of relationship between blood Pb concentration,
bone Pb and body burden in adults.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 (Leggett, 1993).
1 Tibia bone Pb has been shown to be correlated with time-integrated blood Pb concentration (i.e.,
2 CBLI). McNeill et al. (2000) compared tibia Pb levels and cumulative blood Pb indices in a population of
3 19- to 29-year-olds who had been highly exposed to Pb in childhood from the Bunker Hill, Idaho smelter.
4 They concluded that Pb from exposure in early childhood had persisted in the bone matrix until
5 adulthood. The bone Pb/CBLI slopes from various studies range from 0.022 to 0.067 |ig/g bone mineral
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per (.ig-year/dL (Healev et al.. 2008; Hu et al.. 2007). 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 has been 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.. 1993V
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 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. Mukheriee. et al..
2009V
Mobilization of Lead from Bone in Adulthood
Potential mobilization of Pb from the skeleton increases this contribution of bone Pb to blood Pb,
which occurs at times of physiological stress associated with enhanced bone remodeling such as during
pregnancy and lactation (Hertz-Picciotto et al.. 2000; Manton. 1985; Silbergeld. 1991). menopause or in
the elderly (Silbergeld et al.. 1988). extended bed rest (Markowitz & 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 (Franklin et al.. 1997; B. Gulson et al.. 1999; B. Gulson et al.. 2003; B. L.
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 (B. Gulson et al.. 2004; B. L. Gulson
etal.. 1997; Hertz-Picciotto et al.. 2000; Lagerkvist et al.. 1996; Lamadrid-Figueroa et al.. 2006;
Rothenberg et al.. 1994; Schuhmacher et al.. 1996). 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. (1998)
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. The contribution of
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skeletal Pb to blood Pb during the post-pregnancy period remained essentially constant at the increased
level of Pb mobilization.
Gulson et al. (2004) observed 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 (1200 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 (Ettinaer 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 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 (B. 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, women >30 years old had significant odds of having higher blood Pb levels (aged
30-34: OR = 2.39, p <0.001; aged 35-39: OR = 2.98, p <0.001; aged 40-44: OR = 7.69, p <0.001).
Similarly, younger women had less chance of having higher blood Pb levels compared with the reference
group (aged 18-19: OR = 0.60, p = 0.179; aged 20-24: OR = 0.54, p = 0.015). These findings indicate that
maternal blood Pb levels are more likely the result of Pb remobilization from bone stores from historic
exposures as opposed to contemporaneous exposures.
Lactation can affect the endogenous bone Pb release rate. After adjusting for patella Pb
concentration, an increase in blood Pb levels of 12.7% (95% CI: 6.2, 19.6) was observed for women who
practiced partial lactation and an increase of 18.6% (95% CI: 7.1, 31.4) for women who practiced
exclusive lactation compared to those who stopped lactation (Tellez-Roio et al.. 2002). In another Mexico
City study, Ettinger et al. (2004: 2006) 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
(Ettinger et al.. 2006: B. Gulson. Jameson, et al.. 1998: Kovashiki et al.. 2010). 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 (G. A. Drasch et al.. 1987). This decrease is most pronounced in females and may
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be due to osteoporosis and release of Pb from resorbed bone to blood (B. 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
(.ig/g bone) (Garrido Latorre et al.. 2003; Popovic et al.. 2005).
Studies of the effect of hormone replacement therapy on bone Pb mobilization have yielded
conflicting results (Berkowitz et al.. 2004; Garrido Latorre et al.. 2003; Korrick et al.. 2002; Popovic et
al.. 2005; 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 4-11 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
Soft Tissue
1.0
0.6 c/>
4 6
Age (year)
10
10
5
"3>
•= 6 -
.a
a.
¦a
o
o
CO
4 -
- -
Blood
/ jr
Soft Tissue --
t /
1 /
1 /
1 f
1 /
-¦
-
\
2.4
2.0
o
r-f
SJ_
W
1.2 w
c
o
TJ
08 ~
3
(Q
- 0.4
0.0
25 30 35 40 45 50 55 60 65 70
Age (year)
Figure 4-11. Simulation of blood and soft tissue (including brain) Pb in
children and adults who experience a period of increased Pb
intake. Simulation based on ICRP Pb biokinetics model
(Leqqett. 1993).
1 Information on Pb levels in human brain are limited to autopsy data and the simulation of brain Pb
2 shown in Figure 4-12 reflects general concepts derived from observations made in non-human primates,
3 dogs and rodents. These observations suggest that peak Pb levels in the brain are reached 6 months
4 following a bolus exposure and within two months approximately 80% of steady state brain Pb levels are
5 reached (Leggett. 1993). There is a relatively slow elimination of Pb from brain (ti/2 ~ 2 years) compared
6 to other soft tissues (Leggett. 1993). This slow elimination rate is reflected in the slower elimination
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1 phase kinetics in shown Figure 4-12. Although in this model, brain Pb to blood Pb transfer half-times are
2 assumed to be the same in children and adults, uptake kinetics are assumed to be faster during infancy and
3 childhood, which achieves a higher fraction of the soft tissue burden in brain, consistent with higher
4 brain/body mass relationships. This is reflected in the simulation as slower brain Pb accumulation in
5 children. The uptake half times predicted by Leggett (1993) vary from 0.9 to 3.7 days, depending on age.
6 Brain Pb kinetics represented in the simulations are simple outcomes of modeling assumptions and cannot
7 currently be verified with available observations in humans.
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-I 1
4 6
Age (year)
10
10
5
u>
~ 6 -
¦Q
Q.
¦a
o
o
m
4 -
2 -
Blood
Brain
n 1 1 1 1 1 1 r~
25 30 35 40 45 50 55 60 65 70
Age (year)
o
ff
oo
E)
5'
-o
¦c
(Q
Figure 4-12. Simulation of blood and brain Pb in children and adults who
experience a period of increased Pb intake.Simulation based
on ICRP Pb biokinetics model (Leqqett. 1993).
4.3.7. Relationship Between Lead in Blood and Lead in
Urine
Urinary filtering and excretion of Pb is associated with plasma Pb concentrations. Given the
curvilinear relationship between blood Pb and plasma Pb, a secondary expectation is for a curvilinear
relationship between blood Pb and urinary Pb excretion that may become evident only at relatively high
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1 blood Pb concentrations (e.g., >25 (ig/dL). Figure 4-13 shows these relationships predicted from the
2 model. In this case, the exposure scenario shown is for an adult (age 40 years) at a quasi-steady state
3 blood Pb concentration; the same relationships hold for children. At low blood Pb concentrations (<25
4 (ig/dL), urinary Pb excretion is predicted to closely parallel plasma Pb concentration for any given blood
5 Pb level (Figure 4-13, top panel). It follows from this that, similar to blood Pb, urinary Pb will respond
6 much more rapidly to an abrupt change in Pb exposure than will bone Pb. One important implication of
7 this relationship is that, as described previously for blood Pb, the relationships between urinary Pb and
8 bone Pb will diverge with increasing exposure duration and age, even if exposure remains constant.
9 Furthermore, following an abrupt cessation of exposure, urine Pb (i.e., not provoked by administration of
10 chelating agents) will quickly decrease while bone Pb will remain elevated (Figure 4-13, lower panel).
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80
> 60
re
5
O)
.Q
o.
a>
c
40
20
Urine
— Plasma
0.20 »
t/>
0.15 5
10 20 30 40
Blood Pb (|ig/dL)
50
16
12 -
>- J*
fD "O
"D ^
buo
«!
c O
.E o
8 -
m 4
— Bone Urine Blood
\
\
V
£
«%
Jb /
S
s
y
-
y
/
f
/
/
. - -
/
1 111II Hllw _
l 1 1 r
25 30 35 40 45 50 55 60 65 70
Age (year)
Figure 4-13. Top panel: Predicted relationship between plasma Pb
concentration and urinary Pb excretion in an adult (age 40
years). 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 (Leggett, 1993).
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23
4.4. Observational Studies of Lead Exposure
4.4.1. Lead in Blood
Overall, trends in blood Pb levels have been decreasing among U.S. residents over the past twenty
years. Blood Pb concentrations in the U.S. general population have been monitored in the NHANES.
Analyses of these data have shown 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
(Brodv et al.. 1994; Pirkle et al.. 1994; Pirkle et al.. 1998; J. Schwartz & Pitcher. 1989). The temporal
trend for the period 1988-2008 is shown in Figure 4-14. 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 4-15 uses NHANES data to illustrate 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 (NC'HS. 2010). For 2005-2008, 95% of childhood blood Pb
levels were less than 5 (ig/dL.
When data were aggregated for years 1999-2004, Pb concentrations in children were highest in the
ethnicity category non-Hispanic black (GM 2.8, 95% CI: 2.5, 3.0) compared to the categories Mexican-
American (GM 1.9, 95% CI: 1.7, 2.0) and non-Hispanic white (GM 1.7, 95% CI: 1.6, 1.8) (Jones et al..
2009). Figure 4-16 demonstrates the change in percent of children with various blood Pb levels by
race/ethnicity from 1988-1991 and 1999-2004. When these data were aggregated for the years 1988-2004,
residence in older housing, poverty, age, and being non-Hispanic black were significant risk factors for
higher Pb levels (Jones et al.. 2009). The geometric mean blood Pb concentration among adults > 20 years
of age was 1.38 (ig/dL (95% CI: 1.31, 1.46) based on the sample collected during the period 2007-2008
(CDC. 2011). Based on these same data, the geometric mean for all males was 1.47 (ig/dL (95% CI: 1.39,
1.56), and for females was 1.11 (ig/dL (95% CI: 1.06, 1.16).
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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
Figure 4-14. Temporal trend in blood Pb concentration.Shown are
geometric means and 95% CIs based on data from NHANES III
Phase 1 (Brodvet al„ 1994: Pirkleetal.. 1994): NHANES III
Phase 2 (Pirkleetal.. 1998): and NHANES IV (CDC. 2011). Data
for adults during the period 1988-1994 are for ages 20-49
years, and > 20 years for the period 1999-2008.
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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
All
1999-2000
1.66
1.60, 1.72
7970
2001-2002
1.45
1.39, 1.51
8945
2003-2004
1.43
1.36, 1.50
8373
2005-2006
1.29
1.23, 1.36
8407
2007-2008
1.27
1.21, 1.34
8266
1-5 yr
1999-2000
2.23
1.96,2.53
723
2001-2002
1.70
1.55, 1.87
898
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
6-11 yr
1999-2000
1.51
1.36, 1.66
905
2001-2002
1.25
1.14,1.36
1044
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
1011
12-19 yr
1999-2000
1.10
1.04, 1.17
2135
2001-2002
0.94
0.90, 0.99
2231
2003-2004
0.95
0.88, 1.02
2081
2005-2006
0.80
0.75, 0.85
1996
2007-2008
0.80
0.74, 0.86
1074
>20 yr
1999-2000
1.75
1.68, 1.81
4207
2001-2002
1.56
1.49, 1.62
4772
2003-2004
1.52
1.45, 1.60
4525
2005-2006
1.41
1.34, 1.48
4509
2007-2008
1.38
1.31, 1.46
5364
Males
1999-2000
2.01
1.93,2.09
3913
2001-2002
1.78
1.71, 1.86
4339
2003-2004
1.69
1.62, 1.75
4132
2005-2006
1.52
1.42, 1.62
4092
2007-2008
1.47
1.39, 1.56
4147
Females
1999-2000
1.37
1.32, 1.43
4057
2001-2002
1.19
1.14, 1.25
4606
2003-2004
1.22
1.14, 1.31
4241
2005-2006
1.11
1.05, 1.17
4315
2007-2008
1.11
1.06, 1.16
4119
Mexican - Americans
1999-2000
1.83
1.75,1.91
2742
2001-2002
1.46
1.34, 1.60
2268
2003-2004
1.55
1.43, 1.69
2085
2005-2006
1.29
1.21, 1.38
2236
2007-2008
1.25
1.15, 1.36
1712
Non-Hispanic blacks
1999-2000
1.87
1.75,2.00
1842
2001-2002
1.65
1.52, 1.80
2219
2003-2004
1.69
1.52, 1.89
2293
2005-2006
1.39
1.26, 1.53
2193
2007-2008
1.39
1.30, 1.48
1746
Non-Hispanic whites
1999-2000
1.62
1.55, 1.69
2716
2001-2002
1.43
1.37, 1.48
3806
2003-2004
1.37
1.32, 1.43
3478
2005-2006
1.28
1.19, 1.37
3310
2007-2008
1.24
1.16, 1.33
3461
Source: Adapted from data from the NHANES (CDC. 2011")
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10
Source: Adapted from data from the NHANES (NCHS, 2010)
Figure 4-15. Box plots of blood Pb levels among U.S. children (1-5 years
old) from the NHANES survey, 1988-2008.Top: all data.
Bottom: data for subjects having blood Pb levels less than 15
pg/dL.
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70
60
= 50
a>
5 40
M-
O
c 30
a>
u
ai
a. 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 from the American Academy of Pediatrics, Jones et al. (2009)
Figure 4-16. 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 (B. Gulson et
al.. 2008; D. L. Johnson et al.. 1996; Laidlaw et al.. 2005). Seasonal variation in blood Pb concentration
was also evident in individual children when repeated blood Pb measurements were made over a 5-year
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21
22
23
24
25
26
27
28
29
30
period (B. Gulson et al.. 2008). 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; Syracuse, NY; New Orleans, LA). Blood
Pb data was obtained from public health screening programs conducted in the three cities during
1998-2003. 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 (D. Johnson & Bretsch. 2002; D. L. Johnson et al.. 1996); New
Orleans, 2,295 (Mielke et al.. 2007). 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).
This analysis provides a possible explanation for the seasonal patterns of blood Pb concentrations in
children that involves weather-dependent entrainment and air transport of surface dusts.
Trends in blood Pb levels have been accompanied by changes in Pb isotope ratios within 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 (B. Gulson et al.. 2006a; B. Gulson et al.. 2008; Ranft et al.. 2006.
2008). 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 various study populations in Australia during the period
1990-2000 (Figure 4-17). 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.
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source
O Adelaide
BK Adult
^ females
~ BK children
A Broken Hill
V Hobart
<| PbinCa
t> Port Pirie
i Sydney
' children
17.60-
n 17.20
16.80-
R Sq Linear = 0.53
16.40-
12.5—
10.0-
7.5-
m
-Q
CL
5.0-
2.5-
0.0-
"i i i i i r
1990 1992 1994 1996 1998 2000
i 1 1 r
1990 1992 1994 1996 1998 2000
(a)
Year
(b) Year
Source: Used with permission from Academic Press, Gulson et al. (2006a)
Figure 4-17. a) Trends in 206Pb/204Pb isotope ratio in blood Pb and b) trends
in blood Pb levels among Australian study populations during
the period 1990-2000.
4.4.2. Lead in Bone
An extensive national database (i.e., NHANES) is available for blood Pb concentrations in children
and adults, as described in Section 4.4.1. Bone Pb concentrations are less well characterized. Tables 4-8
and 4-9 are compilations of data from epidemiologic studies that provided bone Pb concentrations and/or
variability in concentrations among individuals without reported occupational exposure and those with
occupational exposures, respectively. In non-occupationally exposed individuals, typical group mean tibia
bone Pb concentrations ranged from 10 to 30 |ig/g. Patella bone Pb levels are typically higher than tibia
bone Pb levels in the studies considered (Table 4-8). For example, in the Normative Aging Study, patella
bone Pb concentrations were approximately 32 |ig/g. whereas tibia bone Pb concentrations were about 22
|ig/g. Occupationally exposed individuals generally had greater bone Pb concentrations than seen in
control groups (i.e., unexposed). Bone Pb data in Table 4-9 for occupationally exposed individuals were
also generally higher compared to non-occupationally exposed individuals (Table 4-8).
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Table 4-8. Epidemiologic studies that provide bone Pb measurements for non-
occupational^ exposed populations
Reference Cohort Age(yrs) N Location Study Period bSer Co^fSg) of®°"e,Pb
Bandeen-
Baltimore
50-70
1140
Baltimore, MD 2001-2005
Cumulative
Tibia
Mean±SD
Not reported
Roche etal.
Memory
Tibia: 18.8 ±
(2009)
Study cohort
11.6
Bellinger et
Not reported
5-8
79
Boston, MA 1989-1990
Cumulative
Tibia
Mean
High
al. (1994)
(recruited)
Patella
(Range):
exposure:
19-20 (follow-
Tibia: 5.4 (3-
>24
up)
16)
Low
exposure:
Patella: 9.2
<8.7
(4-18)
Cheng et al.
Normative
Mean±SD:
833
Boston, MA 8/1/1991-
Cumulative
Tibia
Mean±SD
Lowest
(2001)
Aging Study
Normotensive
males
12/31/1997
Patella
Tibia:
quintile: Tibia:
cohort
: 65.49 ±7.17
Normotensive
8.5
Borderline
: 20.27 ±
Patella: 12.0
hypertension:
11.55
68.3 ± 7.79
Borderline
Highest
Definite
hypertension:
quintile:
hypertension:
23.46 ±15.02
Tibia: 36.0
67.93 ± 6.79
Definite
Patella: 53.0
hypertension:
22.69 ±14.71
Patella:
Normotensive
: 28.95 ±
18.01
Borderline
hypertension:
33.73 ±21.76
Definite
hypertension:
32.72 ±19.55
Coon etal.
Participants
>50
121
Southeastern 1995-1999
Cumulative
Tibia
Mean±SD:
Tibia
(2006)
from Henry
Mean: 69.9
cases
Michigan (participants
Calcaneu
Tibia: 12.5 ±
Q1:0-5.91
Ford Health
414
received
s
7.8
Q2: 5.92-
System
controls
primary health
Calcaneus:
10.40
(HFHS)
care services)
20.5 ±10.2
Q3: 10.41-
15.50
Q4:> 15.51
Calcaneus
Q1: 0-11.70
Q2:11.71-
19.07
Q3: 19.08-
25.28
Q4: > 25.29
Elmarsafaw
Normative
Not reported
471
Greater 6/1991-
Not reported
Tibia
Mean±SD:
Not reported
yet al.
Aging Study
elderly
Bostonarea, 12/1994
Patella
Tibia: 21.6 ±1
(2006)
males
MA
2.0
Patella: 31.7
±18.3
Glass et
Baltimore
Mean: 59.4
1,001
Baltimore, MD 2001-2005
Cumulative
Tibia
Mean±SD:
NPH Scale:
al.,(2009)
Memory
Range: 50-70
(lifetime)
Tibia: 18.8 ±
Lowest
Study
11.1
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
May 2011
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Reference Cohort Age (yrs)
Location Study Period
Prior Pb Bone Pb Bone Pb
Exposure biomarker Cone. (|jg/g)
Distribution
of Bone Pb
(Mg/g)
Hsiehetal. Not reported Mean:
(2009) Control:
46.06
18 Not reported
controls
Not reported Control group for Tibia
occupational Patella
exposure group
Mean±SD
Tibia Control:
18.51 ±22.40
Patella
Control: 7.14
±9.81
Not reported
Hu et al.
(1996)
(As
reported in
Navas-
Acien et al.,
(2008))
Normative
Aging Study
48-92
Mean ±SD:
66.6 ± 7.2
590
males
Boston, MA
8/1991- Cumulative
12/1994
Tibia Mean±SD:
Patella 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
Jainet al.
(2007)
VA-
Normative
Aging Study
Not reported
837
males
Greater
Boston, MA
9/1/1991- Not reported
12/31/2001
Tibia
Patella
Mean ± SD
Tibia:
Non-Cases:
21.4 ± 13.6
Cases: 24.2 ±
15.9
Patella:
Non-cases:
30.6±19.7
Cases: 36.8 ±
20.8
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)
Range:
Tibia:
Noncases:
-3-126
Cases: -5-75
Patella:
Noncases: -
10-165
Cases: 5-101
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)
Cases:
Tertile 1:
15.3±4.3 (5-
19)
Tertile 2: 25.7
±3.8(21-33)
Tertile 3: 53.3
±17.3(35-
101)
Kamel et al.
Not reported
30-80
256
New England
1993-1996
Cumulative
Tibia
Mean±SE
Controls
(2002):
controls
(Boston, MA)
Control group for
Patella
Tibia
Tibia: N (%)
Kamel et al.
(Bone
occupational
Controls: 11.1
-7-7:14(34)
(2005);
samples
exposure group
±1.6
8-14:12 (29)
Kamel et al.
collecte
Patella
15-61: 15
(2008)
d from
Controls:
(37)
41
16.7 ±2.0
controls
Patella: N
)
(%)
-4-9:14(34)
10-20: 14
(34)
21-107: 13
(32)
Khalil etal.
1982 Pb
Control
51
Eastern
1982-2004
Control group for
Tibia
Median (IQR)
Not reported
(2009)
Occupational
mean: 55
controls
Pennsylvania
occupational
Tibia Control:
Study
exposure group
12 (-8-32)
May 2011
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Reference Cohort Age (yrs)
Location Study Period
Prior Pb Bone Pb Bone Pb
Exposure biomarker Cone. (|jg/g)
Distribution
of Bone Pb
(Mg/g)
Korrick et
al. Q999)
(As
reported in
from Navas-
Acien et al.,
(2008)
Nurses'
Health Study
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
284
females
(89
cases;
195
controls
)
Boston, MA 7/1993-
7/1995
Nonoccupational!
y exposed
Tibia
Patella
Mean ± SD
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.
Not reported
22.0-60.2
135
Republic of
10/24/1997-
Control group for
Tibia
Mean ± SD
Not reported
(2001) (As
Mean ±SD:
controls
Korea
8/19/1999
occupational
Tibia
reported in
Controls:
exposure group
Controls: 5.8
Navas-
34.5 ±9.1
±7.0
Acien et al.,
(2008))
Range
Tibia
Controls:
-11-27
Martin et al.
Baltimore
50-70
964
Baltimore, MD 5/2001 -
Cumulative
Tibia
Mean ± SD
Tibia IQR:
(2006)
Memory
Study
Mean: 59.4
9/2002 (1st
study visit)
8/2002-
3/2004 (2nd
study visit -
tibia Pb
measured)
(lifetime)
Tibia: 18.8 ±
12.4
11.9-24.8
Needleman
Not reported
12-18
194
Allegheny
4/1996-
Not reported
Tibia
Mean ± SD
Table 4
etal. (2002)
Mean age ±
male
County, PA
8/1998
Tibia Cases
distributes
SD:
African
youth
cases
(cases);
Pittsburgh, PA
(ppm):
All subjects:
bone Pb by >
25 or <25 for
American
146
(controls)
11.0 ±32.7
race, two
cases: 15.8 ±
male
African
parental
1.4
youth
American: 9.0
figures, and
African
controls
±33.6
parent
American
White: 20 ±
occupation
controls: 15.5
27.5
±1.1
Tibia Controls
White cases:
(ppm):
15.7 ±1.3
All subjects:
White
1.5 ±32.1
controls: 15.8
African
±1.1
American: -
1.4 ± 31.9
White: 3.5 ±
32.6
Osterberg
Not reported
Median: 41.5
19 male
Not reported
Not reported
Control group for
Finger
Median
Not reported
etal. (1997)
controls
occupational
bone
(range)
(As
exposure group
Finger Bone
reported in
Controls:
Shih etal.,
4 (-19-18)
(2007))
May 2011
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Reference Cohort Age (yrs)
Location Study Period
Prior Pb Bone Pb Bone Pb
Exposure biomarker Cone. (|jg/g)
Distribution
of Bone Pb
(Mg/g)
Parketal.
(2006)
Normative
Aging Study
Mean: 72.9 ±
6.5
413
males
Greater
Boston, MA
11/14/2000-
12/22/2004
(HRV
measurement
s taken)
1991-2002
(bone Pb
measurement
s 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)
Median (IQR)
for No. of
metabolic
abnormalities
Tibia:
0:18.5 (10.5-
23)
1:19(11-28)
2:19(12-26)
Patella:
0:22 (13.5-
32)
1:25 (16-36)
2:20 (15-32)
Estimated
Patella:
0:16.3 (10.8-
24.8)
1:17.1 (11-
29.3)
2:15.1 (9.4-
22.1)
Parketal.
Normative
Mean: 67.3 ±
613
Greater
8/1991 -
Not reported
Tibia
Median (IQR)
Table 1
(2009)
Aging Study
7.2
males
Boston, MA
12/1995
Patella
distributes
Tibia: 19(14-
27)
Patella: 26
tibia and
patella Pb by
genotype;
(18-37)
Table 2
distributes
tibia and
patella Pb by
number of
qene variants
Parketal.
VA Normative
Mean: 64.9
448
Eastern
1991-1996
Cumulative
Tibia
Mean±SD
Tibia IQR: 15
(2010)
Aging Study
(at bone Pb
males
Massachusett
(chronic
Patella
Tibia: 22.5 ±
Patella IQR:
cohort
measurement
s
exposure)
14.2
21
)
Patella: 32.5
±20.4
Table 2
provides age-
adjusted
mean bone
Pb levels
(age, race,
education,
smoking
[pack-yr],
occupational
noise, noise
notch, BMI,
hypertension,
diabetes)
Payton et
VA Normative
Mean: 66.8
141
Boston, MA
4/1993-
Not reported
Tibia
Mean ± SD
Not reported
al. (1998)
Aging Study
males
3/1994
Patella
Tibia: 22.5 ±
cohort
12.2
Patella: 31.7
±19.2
Peters etal.
Normative
Mean: 66.9
513
Boston, MA
1991-1996
Cumulative
Tibia
Mean ± SD
Not reported
(2007)
Aging Study
cohort
male
cases
Patella
Tibia: 21.5 ±
13.4
Patella: 31.5
±19.3
Rajan etal.
VA Normative
Mean: 67.5
1075
Boston, MA
1991-2002
Not reported
Tibia
Mean ± SD
Not reported
(2007)
Aging Study
(at bone
males
Patella
Tibia: 22.1 ±
Cohort
scan)
13.8
Patella: 31.4
±19.6
May 2011
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Reference Cohort Age (yrs)
Location Study Period
Prior Pb Bone Pb Bone Pb
Exposure biomarker Cone. (|jg/g)
Distribution
of Bone Pb
(Mg/g)
Rajanetal. VA Normative >45 720
(2008) Aging Study males
Cohort
Boston, MA 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 VA Normative Mean: 67.1 526
al. (2003) Aging Study males
Cohort
Boston, MA
1/1/1991-
12/31/1995
Not reported
Tibia
Patella
Mean ± SD
Tibia: 21.9 ±
13.5
Patella: 32.1
±19.8
Roels etal.
(1994)
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
J33]_
Not reported 30-60
68
males
Belgium
Not reported
Control group for
occupational
exposure group
Rothenberg
etal. (2002)
(As
reported in
Navas-
Acien et al.,
(2008))
Tibia
Geometric
Mean
(Range)
Tibia
Controls:
Normotensive
: 21.7 (<15.2-
69.3)
Hypertensive:
20.2 (<15.2-
52.9)
Total: 21.4
(<15.2-69.3)
Tibia
Mean ± SD
Calcaneu
Tibia: 8.0 ±
s
11.4
Calcaneus:
10.7 ± 11.9
Not reported
Not reported
15-44
Mean ±SD:
31.0 ±7.7
720 Los Angeles, 6/1995-
females CA 5/2001
Not reported
Tibia
quartiles:
Q1: -33.7-0.9
Q2: 1.0-8.0
Q3: 8.1-16.1
Q4:16.2-42.5
Calcaneus
quartiles:
Q1: -30.6-3.0
Q2: 3.1-10.0
Q3:10.1-18.7
Q4: 18.8-49.0
Shih et al.,
(2006)
Baltimore
Memory
Study cohort
Mean: 59.39 985 Baltimore, MD Not reported Not reported
Tibia
Mean ± SD:
Tibia: 18.7 ±
11.2
Not reported
May 2011
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Reference Cohort Age (yrs)
N
Location Study Period
Prior Pb Bone Pb Bone Pb
Exposure biomarker Cone. (|jg/g)
Distribution
of Bone Pb
(Mg/g)
Stokes et
al. Q998),
as reported
in Shih et
al., (2007)
Not reported
19-29 (in
1994)
Mean±SD: 276
Cases: 24.3 ± controls
3.18
Control: 24.2
±3.02
Cases: 9
months-9 yr
(during
1/1/1974-
12/31/1975)
257 Silver Valley, 7/10/1994-
cases ID; Spokane, 8/7/1994
WA
Cumulative
(lifelong)
Environmental
(resided near Pb
smelter during
childhood)
Tibia
Mean
(Range):
Tibia Cases:
4.6 (-28.9-37)
Tibia
Controls: 0.6
(-46.4-17.4)
Tibia
No. of Cases:
<1 Mg/g:
31.5%
1-5 Mg/g:
24.4%
5-10 |jg/g:
22.3%
>10 |jg/g:
21.8%
No. of
Controls:
<1 Mg/g:
50.4%
1-5 Mg/g:
25.6%
5-10 Mg/g:
19.4%
>10 Mg/g:
4.7%
Mean ± SD
Tibia
concentration
by age group:
Cases:
19-21:1.47 ±
8.35
22-24: 4.48 ±
7.45
25-27: 4.82 ±
8.92
28-30: 6.64 ±
9.53
Controls:
19-21:1.27 ±
6.60
22-24: -0.61
±6.19
25-27: 0.60 ±
8.60
28-30:1.74 ±
6.42
Van
Not reported
Mean: 61.5 47
Rochester,
Not reported
Cumulative
Tibia Mean ± SD
Not reported
Wijngaarde
NY
Calcaneu Tibia: 2.0 ±
n etal.
s 5.2
(2009)
Calcaneus:
6.1 ±8.5
Wasserman
Yugoslavia
10-12 167
Kosovska,
5/1985-
Cumulative
Tibia Mean ± SD:
Tibia
etal. (2003)
Prospective
children
Mitrovica,
12/1986
(lifetime)
Tibia
quartiles:
Study of
Kosovo,
(mother's
Environmental
Pristina: 1.36
Q1: -14.4-
Environment
Yugoslavia;
enrollment)
(Pb smelter,
±6.5
1.85
al Pb
Pristina,
1986-1999
refinery, battery
Mitrovica:
Q2: 1.85-10.5
Exposure
Kosovo,
(follow-up
plant)
39.09 ± 24.55
Q3: 10.5-35
Yugoslavia
through age
12 yr)
Tibia Pb
measured 11-
Q4: 35-193.5
Table 3
distributes
13 yr old
tibia Pb by
sex, ethnicity,
address at
birth relative
to factory,
and maternal
education
May 2011
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Reference Cohort Age (yrs)
Location Study Period
Prior Pb Bone Pb Bone Pb
Exposure biomarker Cone. (|jg/g)
Distribution
of Bone Pb
(Mg/g)
Weisskopf
etal.
(2004), as
reported in
Shihet al.
194225
(2007)
Normative
Aging Study
Mean ± SD:
67.4 ± 6.6
466
males
Boston, MA 1991-2002
Environmental
Tibia
Patella
Median (IQR)
Tibia: 19
(12,26)
Patella: 23
(15, 35)
Tibia IQR: 14
Patella IQR:
20
Table 3
shows mean
Pb levels
across cate-
gorical
variables (yr
of education,
smoking
status,
computer
experience,
first language
English)
31 Boston, MA Bone Pb Not reported Tibia Median (IQR) Not reported
males measured: Patella Tibia
1994-1999 Lowest quin-
Scans tile: 13(9-17)
performed: Highest quin-
2002-2004 tile: 41 (38-
59)
Patella
Lowest quin-
tile: 9 (5-15)
Highest quin-
tile: 63 (43-
86)
Weisskopf
et al. (2007) Aging Study
cohort
VA Normative Mean:
Lowest
Patella
quintile: 73.2
Highest
Patella
quintile: 80.7
Weisskopf VA Normative Mean: 68.7
et al. (2007) Aging Study
cohort
1,089 Boston, MA
males
1993-2001
Concurrent and
cumulative
Tibia
Patella
Median (IQR)
Tibia: 20 (13-
28)
Patella: 25
(17-37)
Table 1
shows the
distribution of
Pb
biomarkers
by categories
of covariates
(age, edu-
cation, smo-
king status,
alcohol
intake, physi-
cal activity,
computer
experience,
first language
Weisskopf
Normative
Mean ± SD
868
Greater
1991-1999
Cumulative
Tibia
Mean ± SD
Patella
et al. (2009) Aging Study
(at Patella
males
Boston area,
Patella
Tibia: 21.8 ±
tertiles:
(95% white)
baseline)
Tertile 1:65.2
±7.1
Tertile 2: 66.5
±6.5
Tertile 3: 70.2
±7.2
MA
13.6
Patella: 31.2
±19.4
1: <22
2: 22-35
3: >35
Weisskopf
BUMC, BWH,
Mean:
330
Boston, MA
2003-2007
Cumulative
Tibia
Mean ± SD:
Tibia
etal. (2010)
BIDMC,
Cases: 66.5
cases
1991-1999
Patella
Tibia: 10.7 ±
quartiles:
HVMA,
Controls:
308
(NAS patients
12.1
Q1: <3.1
Normative
69.4
controls
bone Pb
Patella: 13.6
Q2: 3.5-9.6
Aging Study
measured)
±15.9
Q3: 10.0-17.0
(NAS),
Q4: >17.3
Harvard
Cooperative
Patella
Program on
quartiles:
Aging
Q1: <2.7
(HCPOA)
Q2: 3.5-11.0
Q3:11.3-20.9
Q4: >20.9
May 2011
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Reference Cohort Age (yrs)
Location Study Period
Prior Pb Bone Pb Bone Pb
Exposure biomarker Cone. (|jg/g)
Distribution
of Bone Pb
(Mg/g)
Weuve et
al. (2006)
VA Normative >45
Aging Study
cohort
720
males
Boston, MA
1991
(measuring
bone Pb
levels)
End date not
reported
Cumulative
Tibia
Patella
Median (1st-
3rd quartile):
Tibia: 19(13-
28)
Patella: 27
(18-39)
Weuve et Nurses'
al. (2009) Health Study
cohort
Table 1
shows
distribution of
mean Pb bio-
marker levels
by
characteristic
s of
participants
(age,
education,
computer
experience,
smoking
status,
alcohol
consumption,
fertile of cal-
cium intake,
fertile of
physical
activity,
diabetes)
47-74
587
females
Boston, MA 1995-2005 Recent and
cumulative
Tibia
Patella
Mean ± SD:
Tibia: 10.5 ±
9.7
Patella: 12.6
±11.6
Not reported
Wright etal.
Normative
Mean ± SD:
736 Boston, MA 1991-1997
Environmental Tibia
Mean ± SD:
Tibia:
(2003), as
Aging Study
68.2 ± 6.9
males
Patella
Difference in
reported in
Tibia: 22.4 ±
mean from
Shih etal.
15.3
Lowest-
194225
Patella: 29.5
highest
(2007)
±21.2
quartile: 34.2
Patella:
Difference in
mean from
lowest-
highest
quartile: 47
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Table 4-9. Epidemiologic studies that provide bone Pb measurements for occupationally
exposed populations
Reference Cohort
Age
(years)
Number
of Location
Subjects
Study
Period
Prior Pb
Exposure
Bone Pb
biomarker
Bone Pb
Concentration
(ng/g)
Distribution
of Bone Pb
(ng/g)
Bleecker et
Canada
Cumulative:
80 males
Canada
Not
Occupational
Tibia
Mean ± SD
Not reported
al. (1997)
Lead Study
24-64
reported
(Pb smelter
(Tibia):
(As reported
Younger:
workers)
Cumulative: 41.0
inShihetal.,
24-43
± 24.44
(2007))
Older: 44-64
Younger: 35 ±
Mean ± SD:
Cumulative:
44.1 ±8.36
Younger:
37.2 ± 4.57
Older: 50.9
±4.86
24.11
Older: 46.9 ±
23.59
Range (Tibia):
Cumulative: -12-
90
Younger: -12-80
Older: 3-90
Bleecker et
Not reported
Mean: 39.7
61
Northern
Not
Occupational
Tibia
Mean:
Not reported
al. (2007)
Canada
reported
(primary Pb
Tibia: 38.6
smelter
workers)
Caffo et al.
Not reported
Mean: 60.39
513 males
Delaware
1994-1997
Cumulative
Tibia
Mean ± SD:
Not reported
(2008)
and New
(Phase 1
Occupational
Peak Tibia:
Jersey, US
recruitment
(Former
23.99 ±18.46
)
2001-2003
(Phase 2
recruitment
)
organolead
manufacturin
g workers)
Dorsey et al.
Not reported
Mean: 43.4
652
Korea
10/24/1997 Occupational
Tibia
Mean ± SD:
Not reported
(2006)
-8/19/1999
(Pb workers)
Patella
Tibia: 33.5 ±
(enrolled)
43.4
Patella: 75.1 ±
101.1
Glenn et al.
Not reported 40-70
496 males
Eastern U.S.
6/1994-
Occupational
Tibia
Mean ± SD:
Not reported
(2003). as
Mean: 55.8
6/1996
(Chemical
Tibia: 14.7 ±9.4
reported in
(baseline)
(enrolled)
manufacturin
(at yr 3)
Navas-Acien
6/1998
g facility;
Peak Tibia: 24.3
et al., (2008)
(follow-up
period
ended)
inorganic and
organic Pb)
±18.1
Range:
Tibia: -1.6-52 (at
year 3)
Peak Tibia:
-2.2-118.8
Glenn et al.
(2006)
Not reported
0-36.2
(baseline)
Mean ± SD:
41.4 ±9.5
(baseline)
575
(76%
male; 24%
female)
South Korea
10/1997-
6/2001
Cumulative Tibia
and recent
Occupational
(Pb-using
facilities)
Mean ± SD:
Tibia: 38.4 ±
42.9
Tibia-Women:
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
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Reference Cohort
Age
(years)
Number
of Location
Subjects
Study
Period
Prior Pb
Exposure
Bone Pb
biomarker
Bone Pb
Concentration
(ng/g)
Distribution
of Bone Pb
(ng/g)
Hanninen et
Not reported Mean±SD:
54
Helsinki,
Not
Occupational
Tibia
Mean±SD:
Not reported
al. (1998). as
Male: 43
(43 males,
Finland
reported
(Pb acid
Calcaneus
reported in
Female: 48
11
battery
Tibia:
Shih etal.,
BPb(max)
females)
factory
BPb (max)
(2007)
<2.4
workers)
<2.4 |jmol/L:
|jmol/L: 41.7
19.8 ± 13.7
±9.3
BPb (max)
BPb (max)
>2.4 |jmol/L:
>2.4 |jmol/L:
35.3 ±16.6
46.6 ±6.2
Calcaneus:
BPb (max)
<2.4 |jmol/L:
78.6 ±62.4
BPb (max)
>2.4 |jmol/L:
100.4 ±43.1
Hsieh etal.
Not reported Mean:
22 cases
Location NR
Not
Occupational
Tibia
Mean ± SD
Not reported
2009 (2009)
Cases:
18 controls
reported
(Pb paint
Patella
Tibia
45.71
factory
Case: 61.55 ±
Controls:
workers)
30.21
46.06
Control: 18.51 ±
22.40
Patella
Case: 66.29 ±
19.48
Control: 7.14 ±
9.81
Kamel et al.
Not reported 30-80
109 cases
New
1993-1996
Cumulative
Tibia
Mean ± SE
Cases
(2002):
256
England
Occupational
Patella
Tibia
Tibia Pb: N (%)
Kamel et al.
controls
(Boston,
(Pb fumes,
Cases: 14.9 ±
-7-7: 21 (20)
(2005);
(Bone
MA)
dust, or
1.6
8-14:35(34)
Kamel et al.
samples
particles)
Controls: 11.1 ±
15-61:48(46)
(2008)
collected
1.6
from 104
Patella Pb: N
cases and
Patella
(%)
41
Cases: 20.5 ±
-4-9: 27 (26)
controls)
2.1
10-20: 40 (38)
Controls: 16.7 ±
21-107:37(36)
2.0
Khaliletal. 1982 Pb Mean: 83 cases Eastern
(2009) Occupationa Cases: 54 51 controls Pennsylvani
I Study Controls: 55 a
cohort
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)
1982-2004
Occupational
(Pb battery
plant
workers)
Tibia
Osterberg et
al. (1997). as
reported in
Shih etal.
(2007)
Median (IQR)
Tibia
Cases: 57 (20-
86)
Controls: 12
(-8-32)
Not reported
Not reported Median:
41.5
38 male
cases
19 male
controls
Not reported
Not
reported
Occupational
(secondary
Pb smelter -
inorganic Pb)
Finger bone
Median
Finger Bone:
High Cases: 32
Low cases: 16
Control: 4
Range
Finger Bone:
High Cases:
17-101
Low cases: -7-
49
Control:-19-18
Not reported
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Reference Cohort
Age
(years)
Number
of
Subjects
Location
Study
Period
Prior Pb
Exposure
Bone Pb
biomarker
Bone Pb
Concentration
(ng/g)
Distribution
of Bone Pb
(ng/g)
Roelsetal. Not reported 30-60
76 male Belgium
Not
Occupational Tibia
Geometric Mean Not reported
(1994)
cases
reported
(Pb smelter
(Range)
68 male
workers)
controls
Mean case
Tibia Cases:
exposure: 18
Normotensive:
yr (range: 6 to
64.0(19.6-
36 yr)
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)
Schwartz
U.S. Mean±SD:
535 male Eastern U.S.
6/1994-
Occupational
Tibia
Mean ± SD
Not reported
(2000) et al.,
Organolead Cases: 55.6
cases
10/1997
(tetraethyl
Current Tibia:
as reported
Study ± 7.4
118 male
(enrolled)
and
Cases: 14.4 ±
in Shih et al.,
Controls:
controls
Completed
tetramethyl
9.3
(2007)
58.6 ±7.0
2-4 annual
Pb
follow-up
manufacturin
Peak Tibia:
visits
g facility)
Cases: 22.6 ±
Tibia Pb
16.5
taken in
3rd year
Schwartz et
Not reported 41.7-73.7
543 males Eastern U.S.
1995
Occupational
Tibia
Mean ± SD
Not reported
al. (2000). as
(Combined)
(recruited)
(former
Tibia:
reported in
Mean ± SD:
1996-1997
organolead
Combined: 14.4
Navas-Acien
Combined:
(Tibia Pb
manufacturin
±9.3
et al.(2008)
57.6 ±7.6
taken
g workers)
Hypertensive:
Hypertensiv
during the
15.4 ±9.1
e: 60.2 ±6.9
3rd yr)
Nonhypertensive
Nonhyperte
: 14.0 ±9.3
nsive: 56.6
±7.5
Range Tibia:
Combined: -1.6-
52
Schwartz et
Not reported
Mean:
803 cases South Korea
10/24/1997
Occupational
Tibia
Mean ± SD
Not reported
al. (2001):
Exposed:
135
-8/19/1999
(battery
Tibia
Lee et al.
40.4
controls
manufacturin
Cases: 37.1 ±
(2001)
Control:
34.5
g, secondary
smelting, Pb
oxide
manufacturin
g, car radiator
manufacturin
g)
40.3
Control: 5.8 ±
7.0
Range:
Tibia
Cases: -7-338
Controls: -11-27
Schwartz et
Not reported
Mean at T
576 South Korea
10/1997-
Occupational
Tibia
Mean ± SD
Tibia:
al. (2005)
visit: 41.4
6/2001
(current and
Tibia: 38.4 ± 43
25th percentile
former Pb
workers)
atV1:14.4
75th percentile
at V1: 47.1
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Reference Cohort
Age
(years)
Number
of
Subjects
Location
Study
Period
Prior Pb
Exposure
Bone Pb
biomarker
Bone Pb
Concentration
(ng/g)
Distribution
of Bone Pb
(ng/g)
Stewart etal. U.S.
40-70 (in
534 males Eastern U.S. Not
Occupational Tibia
Mean ± SD
Current Tibia
(1999). as Organolead
1995)
reported
(tetraethyl
Tibia:
Pb: N (%)
reported in Study
38% > 60
and
Current: 14.4 ±
<5: 77 (14.2)
Shihet al.,
yrs
tetramethyl
9.3
5-9.99:113
(2007)
Mean: 58
Pb
Peak: 23.7 ±
(20.8)
manufacturin
17.4
10-14.99: 119
g facility)
(21.9)
Range: Tibia
15-19.99: 117
Current: -1.6-52
(21.5)
Peak: -2.2-105.9
>20: 118
(21.7)
Peak Tibia Pb:
N (%)
<5: 49 (9.1)
5-9.99: 64
(11.8)
10-14.99: 70
(12.9)
15-19.99: 87
(16.1)
20-24.99: 79
(14.6)
25-29.99: 55
(10.2)
>30: 137
(26.1)
Stewart etal. Not reported Mean: 56.1
532 males Eastern U.S. 1994-
Cumulative Tibia
Mean ± SD Not reported
(2006)
1997;
Occupational
Current Tibia:
2001-2003
(Organolead
14.5 ±9.6
workers - not
Peak Tibia: 23.9
occupational!
±18.3
y exposed to
Pb at time of
enrollment)
Weaver etal.
(2008)
Not reported Mean ± SD:
43.3 ± 9.8
652
South Korea 12/1999-
6/2001
Occupational Patella
(Current and
former Pb
workers;
plants
produced Pb
batteries, Pb
oxide, Pb
crystal, or
radiators, or
were
secondary Pb
smelters)
Mean±SD
Patella: 37.5 ±
41.8
Not reported
4.4.3. Lead in Urine
1 Urine Pb concentrations in the U.S. general population have been monitored in the NHANES. Data
2 from the most recent survey (CDC. 2011) are shown in Table 4-10. The geometric mean for the entire
3 sample for the period 2007-2008 (n = 2,627) was 0.52 jj.g/g creatinine (95% CI: 0.48, 0.55). The
4 geometric means for males (n = 1,327) and females (n = 1,300) were 0.50 jj.g/g creatinine (95% CI: 0.47,
5 0.53) and 0.53 jj.g/g creatinine (95% CI: 0.49, 0.57), respectively.
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Table 4-10. Urine Pb concentrations in the U.S. population
Survey Stratum
Period
Geometric Mean (|jg/g CR)
95% Confidence Interval
Number of Subjects
All
1999-2000
0.721
0.700, 0.742
2465
2001-2002
0.639
0.603, 0.677
2689
2003-2004
0.632
0.603, 0.662
2558
2005-2006
0.546
0.502, 0.573
2576
2007-2008
0.515
0.483, 0.549
2627
6-11 yr
1999-2000
1.170
0.975,1.41
340
2001-2002
0.918
0.841,1.00
368
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
12-19 yr
1999-2000
0.496
0.460, 0.535
719
2001-2002
0.404
0.380, 0.428
762
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
>20 yr
1999-2000
0.720
0.683, 0.758
1406
2001-2002
0.658
0.617,0.703
1559
2003-2004
0.641
0.606, 0.679
1543
2005-2006
0.573
0.548, 0.600
1520
2007-2008
0.546
0.513,0.580
1857
Males
1999-2000
0.720
0.679, 0.763
1227
2001-2002
0.639
0.607, 0.673
1334
2003-2004
0.615
0.588, 0.644
1281
2005-2006
0.551
0.522, 0.582
1271
2007-2008
0.502
0.471,0.534
1327
Females
1999-2000
0.722
0.681,0.765
1238
2001-2002
0.639
0.594, 0.688
1355
2003-2004
0.648
0.601,0.698
1277
2005-2006
0.541
0.507, 0.577
1305
2007-2008
0.527
0.489, 0.568
1300
Mexican - Americans
1999-2000
0.940
0.876,1.01
884
2001-2002
0.810
0.731,0.898
682
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
Non-Hispanic blacks
1999-2000
0.722
0.659, 0.790
568
2001-2002
0.644
0.559, 0.742
667
2003-2004
0.609
0.529, 0.701
723
2005-2006
0.483
0.459, 0.508
692
2007-2008
0.452
0.414,0.492
589
Non-Hispanic whites
1999-2000
0.696
0.668, 0.725
822
2001-2002
0.615
0.579, 0.654
1132
2003-2004
0.623
0.592, 0.655
1074
2005-2006
0.541
0.500, 0.585
1041
2007-2008
0.506
0.466, 0.550
1095
Values are [jg Pb/g creatinine
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. (2010). Tooth
enamel samples from 127 subjects born between 1936 and 1993 were analyzed for Pb concentration and
Pb isotope ratios of the tooth enamel and compared with those parameters for sediment cores and
estimates of Pb emissions from gasoline during the years when 50% enamel formation was estimated to
occur. They found that the log-transform of tooth enamel concentration was significantly predicted by the
log-transform of Lake Erie sediment core data obtained by Graney et al. (1995) (p <0.00001) and by the
log-transform of U.S. consumption of Pb in gasoline (p <0.00001); Figure 4-18. Additionally, Robbins et
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7
8
9
10
11
12
13
al. (2010) found that 2u7Pb/2u6Pb was significantly predicted by the 2u7Pb/2u6Pb observed in the Lake Erie
sediment cores obtained by Graney et al. (1995) (p <0.0001) and for this study (p <0.0002).
100i
<1)
i 75-
>
"c§
O 50-
c
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
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 2006 AQCD and its 1986 predecessor (U.S. EPA. 1986. 2006) described epidemiological
studies of relationships between air Pb and blood Pb. Much of the pertinent earlier literature (e.g., prior to
1984) was summarized by Brunekreef (1984). 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[jj.g/dl blood Pb] per ln| |_ig/m3 air Pb] (R2= 0.69; Figure 4-19). This corresponds to an increase of 6.3
(ig/dL blood Pb for an increase in air Pb concentration from 0.15 to 1.5 (ig/m3. When the analysis was
limited to children whose blood Pb concentrations were <20 (ig/dL, the regression coefficient was 0.2159
(R2=0.33), which corresponds to an increase of 3.2 (ig/dL blood Pb for an increase in air Pb from 0.15
tol.5 |_ig/m3. Blood Pb-air Pb slopes are presented for recent studies in the following paragraphs. These
data are summarized in Table 4-11.
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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
Slope (|jg/dL/|jg/m )
Children Populations
Brunekreef et al. (1984)
Location: Various countries
Years: 1974-1983
Subjects: Children (varying age ranges)
Analysis: Meta analysis of 18 studies
Model: Log-Log
Blood Pb: 5-41 |jg/dL (mean
range for studies)
Air Pb: 0.2-10 ^ig/m3 (mean
range for studies)
All children: 18a,6.1"
Children <20 ^ig/dl: 13a,
3.0b
Haves et al. (1994)
Location: Chicago, IL
Years: 1974-1988
Subjects: 0.5-6 yr (9,604 blood Pb measurements)
Analysis: Regression of blood Pb screening and quarterly
average air Pb
Model: Log-Log
Blood Pb: 12-30 ^ig/dL
(annual GM range)
Air Pb: 0.5-1.2 ^g/m3 (annual
GM range)
24a, 5.7°
Hilts et al. (2003)
Location: Trail, BC Model: Linear
Years: 1989-2001 Blood Pb: 4.7-11.5 ^ig/dL
Subjects: 0.5-6 yr (292-536 blood Pb measurements/yr) (annual median range)
Analysis: Regression of blood Pb screening and community air Air Pb: 0.03-1.1 ^g/m
Pb following upgrading of a local smelter (annual median range)
6.5
Ranft etal. (2008)
Location: Germany
Years: 1983-2000
Subjects: 6-11 yr (n = 843)
Analysis: Pooled regression 5 cross-sectional studies
Model: Log-Linear
Blood Pb: 2.2-13.6 ^ig/dL (5th-
95th percentile)
Air Pb: 0.03-0.47 ^ig/m3 (5th-
95th percentile)
3.2°
Schnaas etal. (2004)
Location: Mexico City
Years: 1987-2002
Subjects: 0.5-10 yr (n = 321)
Analysis: Regression of longitudinal blood Pb measurements
and annual average air Pb data
Model: Log-Log
Blood Pb: 5-12 |jg/dL (annual
GM range)
Air Pb: 0.7-2.8 [iglm3 (annual
mean range)
4.8a, 1.1"
Schwartz and Pitcher
(1989). U.S. EPA Q986)
Location: US
Years: 1976-1980
Subjects: 0.5-7 yr (n = 7,000)
Analysis: NHANES blood Pb, gasoline consumption data and
Pb concentrations in gasoline
Model: Linear 9.3
Blood Pb: 11-18 |jg/dL (mean
range)
Air Pb: 0.36-1.22 ^ig/m3
(annual maximum quarterly
mean)
Schwartz and Pitcher
(1989). U.S. EPA (1986)
Location: Chicago, IL
Years: 1976-1980
Subjects: 0-5 yr (n = 7,000)
Analysis: Chicago blood Pb screening, gasoline consumption
data, and Pb concentrations in gasoline
Model: Linear 7.7
Blood Pb: 18-27 |jg/dL (mean
range)
Air Pb: 0.36-1.22 ^ig/m3
(annual maximum quarterly
mean)
Tripathi et al. (2001) 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 ^ig/dL
(regional GM range)
Air Pb: 0.11-1.18 ^g/m3
(regional GM range)
3.6
Adult Populations
Rodrigues et al. (2010) Location: New England, U.S. Model: Log-log 5.4a, 4.2°
Years: 1994-1995 Blood Pb: 16.1 ^ig/dL (GM,
Subjects: Adult bridge painters (n-84,1 female) 1,7 GSD)
Analysis: Regression analysis of blood Pb and air Pb data Air Pb: 58 ^g/m3 (GM, 2.8
(personal monitors) collected during work performing various GSD)
job-related tasks
aAt an air concentration of 0.15 m/m3
bAt an air concentration of 1 [jg/m
cFor a change in air Pb concentration from 0.025 to 0.465 ^g/m3
GM, geometric mean; GSD, geometric standard deviation
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9
10
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12
13
60
All children
< 20 ug/dL
50
T3 40
O)
=L
-? 30
Q_
"D
O
-9 20
m
0
5
10
15
20
25
AirPb (|jg/m3)
Data provided from Brunekreef (1984).
Figure 4-19. Predicted relationship between air Pb and blood Pb based on
a meta analysis of 18 studies. The regression model is:
ln[pg/dl_ blood Pb] = 0.3485ln[pg/m3 air Pb] + 2.85 for all
children and ln[pg/dl_ blood Pb] = 0.2159ln[pg/m3 air Pb] +
2.62 when the sample was restricted to populations that had
blood Pb concentrations <20 pg/dL.
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 (n: 292-536 per year) who lived within 4 km from the smelter. 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
(.ig/rn1 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 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
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21
22
23
24
25
26
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 include
aggressive public education and exposure intervention programs (Hilts. 1996; Hilts et al.. 1998).
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 (ug/dL)
11.5
5.9
4.7
NA
Air Pb (Lig/m")
1.1
0.3
0.03
6.5 ± 0.52 (FT=0.99, p=0.050)
Soil Pb (mg/kg)
844
756
750
0.068 ± 0.008 (R"=0.99, p=0.069)
Interior Dust Pb (mg/kg)
758
583
580
0.035 ± 0.005 (FT=0.98, p=0.097)
A new smelter process began operation in 1997. Values for air, soil and dust Pb are annual averages; 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 |_ig Pb/m3 for air and 72-877 mg Pb/kg for soil. The results of multivariate regression
analyses were reported in terms of the relative increase (the geometric mean blood Pb ratio, GMR) for an
increase in air or soil Pb from the 5th to 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 Pb varied depending on the sampling data and ranged from 1.41 to
2.89, with most recent data (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)/5th-95th percentile air
or soil Pb) for calculating equivalent airblood Pb ratios. The model predicts an increase of 3.2 (ig/dL
blood Pb per 1 (ig/m3 increase in air Pb. 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 0.6 (ig/dL per mg/kg soil. The
degree of confounding of the GMR and estimates resulting from the air and soil Pb correlation was not
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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. 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 (i.e., in linear scale, the blood Pb-air Pb slope increases with
increasing air Pb concentration). 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.
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 PMi0 (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 appropriate air
monitoring zones. Associations between blood Pb concentration, air Pb concentration and other variables
(e.g., age, year of birth, family use of glazed pottery) were evaluated using multivariate regression
models. The regression model (r2 = 0.96) predicted blood Pb-air Pb slopes that decreased with year of
birth. 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 predicted slope for this group of children was 0.213 (95% CI: 0.114-0.312)
In [jj.g/dL blood] per ln| (.ig/ni1 air]. This slope corresponds to an increase of 2.1 (ig/dL blood Pb for an
increase in air Pb from 0.15 to 1.5 (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. 1986). 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 7 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 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
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gasoline Pb/day. U.S. EPA (1986) 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 7.8 (ig/dL per (ig/m3, based on the NHANES and
Chicago data, respectively (e.g., 2.14/0.23 = 9.3).
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)3
Air Pb
(M9/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/m3 air per 100 metric tons/day (SE= 0.020, R2= 0.95, p<0.0001).
Conversion factor is 10/365 days/year.
Source: U.S. EPA (1986).
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 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.
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Table 4-14. Air Pb levels and blood Pb levels in children in Mumbai, India
Blood Pb (|jg/dL)
Air Pb (jjg/m°
)
GSD
Location
N
GM
GSD
N
GM
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
14.4
1.64
7
1.12
1.12
The linear regression coefficient is 3.62 |jg/dl_ blood per |jg/m3 air (SE= 0.61, R2= 0.76, p<0.001).
GM, geometric mean; GSD, geometric standard deviation; N, number of subjects.
Source: Data are from Tripathi et al. (2001").
1 Hayes et al. (1994) analyzed data collected as part of the Chicago, IL blood Pb screening program
2 for the period 1974-1988, following the phase-out of leaded gasoline. The data included 9,604 blood Pb
3 measurements in children (age: 6 months to 6 years) and quarterly average air Pb concentrations
4 measured at 12 monitoring stations in Cook County, IL. Quarterly median blood Pb levels declined in
5 association with quarterly mean air Pb concentrations. The regression model predicted a slope of 0.24 In
6 [fj.g/dL blood] per ln| (.ig/nr1 air], as illustrated in Figure 4-20. This corresponds to an increase of 11.1
7 (.ig/dL blood Pb for an increase in air Pb from 0.15 to 1.5 (ig/m3.
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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. (1994).
Figure 4-20. Predicted relationship between air Pb and blood Pb based on
data from Chicago, IL (1974-1988).The regression model is:
ln[pg/dl_ blood Pb] = 0.24ln[Hg/m3 air Pb] + 3.17.
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/nr1
(GSD 2.8), with a maximum daily value of 210 |ag/nr\ The Pb concentrations reported were 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 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/nr1) | was 0.11 (SE = 0.05, p = 0.03). This
corresponds to a 1.3-fold increase in blood Pb concentration for a 10-fold increase in air Pb concentration.
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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.
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 (kig/m°)
0.11 (0.05)
0.03
0.05 (0.07)
0.45
Hand wipe at break (^g 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).
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 (Abadin
& Wheeler. 1997; Bowers et al. 1994; Carlisle & Wade. 1992; Maddaloni et al.. 2005; Stern. 1994. 1996;
U.S. EPA. 2003). 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:
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PbB = E • ISF
Equation 4-2
PbB = E-AF-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 tu2 for elimination of
Pb in blood).
The U.S. EPA Adult Lead Methodology (ALM) is a example of a slope factor model that has had
extensive regulatory use in the EPA Superfund program for assessing health risks to adults associated with
non-residential exposures to Pb in contaminated soils (Maddaloni et al.. 2005; U.S. EPA. 1996a). The
model was developed to predict maternal and fetal blood Pb concentrations that might occur in relation to
maternal exposures to contaminated soils. The model assumes an uptake slope factor of 0.4 (ig/dL blood
per !_ig/day Pb uptake. Additional discussion of slope factor models that have been used or proposed for
regulatory use can be found in the 2006 AQCD (U.S. EPA. 2006).
Previous studies included in the 2006 AQCD (U.S. EPA. 2006) 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
etal.. 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 (Bomschein et al.. 1985; Lanphear & Roghmann. 1997; U.S. EPA. 1996b).
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)|
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1 and ln[floor dust Pb concentration ((.ig/ft2)| was 0.386 (SE 0.089) for "not smooth and cleanable" surfaces
2 (e.g., high-pile carpets) and 0.205 (SE 0.032) for "smooth and cleanable" surfaces (e.g., uncarpeted or
3 low-pile carpets). These coefficients correspond to a 2.4-fold or 1.6-fold increase in blood Pb
4 concentration, respectively, for a 10-fold increase in floor dust Pb concentration.
Table 4-16. Linear model relating environmental Pb exposure and blood Pb concentration
in children3
Variables
Overall p-value
Levels
Estimate (SE)
p-Value
Intercept
0.172
-0.517(0.373
0.172
Age (in yr)
< 0.001
Age
Age2
Age3
Age4
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
1950-1959
1940-1949
Before 1940
-0.121 (0.052)
-0.198 (0.058)
-0.196 (0.060)
-0.174 (0.056)
-0.207 (0.065)
-0.012 (0.072)
0.000
0.024
0.001
0.002
0.003
0.003
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
United States'1
Mexico
Elsewhere
-0.077 (0.219)
0.000
0.353 (0.097)
0.154(0.121
0.728
< 0.001
0.209
Floor surface/condition x log
floor PbD
< 0.001
Intercept for missing
Not smooth and cleanable
Smooth and cleanable or
carpeted
0.178 (0.094)
0.386 (0.089)
0.205 (0.032)
0.065
< 0.001
< 0.001
Floor surface/condition x log
floor PbD)2
Not smooth and cleanable
Smooth and cleanable or
carpeted
0.023(0.015)
0.027 (0.008)
0.124
0.001
Floor surface/condition x (log
floor PbD)3
Uncarpeted not smooth and
cleanable
Smooth and cleanable or
carpeted
-0.020 (0.014)
-0.009 (0.004)
0.159
0.012
Log windowsill PbD
0.002
Intercept for missing
Slope
0.053 (0.040
0.041 (0.011
0.186
< 0.001
Home-apartment type
< 0.001
Intercept for missing
Mobile home or trailer
One family house, detached
One family house, attached
Apartment (1-9 units)
Apartment (> 10 units)
-0.064 (0.097
0.127(0.067)
-0.025 (0.046)
0.000
0.069 (0.060)
-0.133 (0.056)
0.511
0.066
0.596
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
an = 2,155 (age 10-60 mo); R2 = 40%
includes the 50 states and the District of Columbia
Source: Dixon etal. (2009).
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Mielke et al. (2007) analyzed data on blood Pb and soil Pb concentration collected as part of a
universal blood Pb screening program in New Orleans (2000-2005). The data set included 55,551 blood
Pb measurements for children 0-6 years of age and 5,467 soil Pb measurements. Blood Pb and soil Pb
concentrations were matched at the level of census tracts. The association between blood Pb
concentration and soil Pb concentration was evaluated using non-parametric permutation methods. The
resulting best-fit model (R2=0.528) was:
PbB = 2.038 + (0.172 • PbS° 5)
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 1000 ppm (Figure 4-21).
10
8
6
Blood=2.038+0.172 x Soil0-5
4
2
0
0
400
800
1200
1600
2000
Soil Pb (ppm)
Figure 4-21. Predicted relationship between soil Pb concentration and
blood Pb concentration in children based data collected in the
New Orleans child Pb screening program (2000-2005) (Mielke
et al.. 2007). 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.
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. (2007).
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Regression coefficients for soil Pb (random effects generalized least squares regression) ranged from
0.217 to 0.214 (per soil Pb0 5), which is equivalent to approximately a 2-fold increase in blood Pb
concentration for an increase in soil Pb concentration from 100 to 1000 ppm.
Several studies have linked elevated blood Pb levels to residential soil exposures for populations
living nearby industrial or mining facilities. Gulson et al. (2009) studied the blood Pb and isotopic Pb
ratios of children younger than 5 years old and adults older than 18 years old living in the vicinity of a
mine producing Magellan Pb ore in western Australia. They observed a median blood Pb level of 6.6
(ig/dL for the children, with isotopic ratios indicating contributions from the mine ranging from 27 to
93%. A weak but significant linear association between blood Pb level and percent Magellan Pb was
observed (R2= 0.12, p = 0.018). Among children with blood Pb levels over 9 (ig/dL and among adults, the
isotopic ratios revealed Pb exposures from a variety of sources. 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 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.
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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
ln(yard 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 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
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analogous mathematical representation of relationships between Pb intakes and Pb levels in body tissues
(e.g., blood Pb concentration). They include parameters that represent processes of Pb transfer (a) from
portals of entry into the body and (b) from blood to tissues and excreta. Linked together, exposure-intake
and intake-biokinetics models (i.e., integrated exposure-intake-biokinetics models) provide an approach
for predicting blood Pb concentrations (or Pb concentrations in other tissues) that corresponds to a
specified exposure (medium, 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
(Leggett. 1993; O'Flahertv. 1993. 1995). 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. 2006).
4.7. Summary
4.7.1. Exposure
Exposure data considered in this assessment build upon the conclusions of the 2006 AQCD for Pb
(2006). 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
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difficult to assess because Pb has multiple sources in the environment and passes through various media.
Air-related pathways of Pb exposure are the focus of this assessment. Pb can be emitted to air, water, or
soil. In addition to primary emission of particle-bound or gaseous Pb to the atmosphere, Pb can be
suspended to the air from disturbance of soil or dust, and a fraction of that suspended Pb may even
originate from waters used to irrigate the soil. Pb-bearing PM can be deposited from the air to soil or
water through wet and dry deposition. In general, air-related pathways include those pathways where Pb
passes through ambient air on its path from a source to human exposure. Air-related Pb exposures include
inhalation and ingestion of Pb-contaminated food, water or other materials including dust and soil. Non-
air-related Pb exposures may 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. 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.
Section 4.1 presents data illustrating potential exposure mechanisms. Several studies suggested that
soil can act as a reservoir for Pb emissions from industrial or and other activities. Exposure to soil
contaminated with deposited Pb can occur through hand-to-mouth contact as well as inhalation of
resuspended Pb-bearing PM. In general, soil Pb concentrations tended to be higher within inner-city
communities compared with neighborhoods surrounding city outskirts. Infiltration of Pb dust 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 concentrations for Pb are typically higher than indoor or outdoor ambient Pb
concentrations. These findings may be related to local resuspension with body movement.
4.7.2. Kinetics
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 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. 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. As blood Pb increases and the higher affinity binding sites for Pb in RBCs
become saturated at approximately 40 (ig/dL blood, a larger fraction of the blood Pb is available in
plasma to distribute to brain and other Pb-responsive tissues.
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The burden of Pb in the body may be viewed as divided between a dominant slow compartment
(bone) and a smaller fast compartment (soft tissues). 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 & Wittmers. 1992V 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. 1995V Thus, most
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 to 20. 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 to the plasma (e.g., cortical ti/2 = 0.23 years at birth, 3.7
years at 15 years of age, and 23 years in adults; trabecular ti/2 = 0.23 years at birth, 2.0 years at 15 years of
age, and 3.8 years in adults) (Leggett. 1993V Due to the more rapid turnover of bone mineral in children,
changes in blood Pb concentration are thought to more closely parallel changes in total body burden.
However, some Pb accumulated in bone during childhood does persist into later life. Potential
mobilization of Pb from the skeleton could occur in adults at times of physiological stress associated with
enhanced bone remodeling such as during pregnancy and lactation, menopause or in the elderly, extended
bed rest, hyperparathyroidism, and weightlessness. Regardless of age, however, similar blood Pb
concentrations in two individuals (or populations) do not necessarily translate to similar body burdens or
similar exposure histories.
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 an elimination 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 may become evident with longer observation periods
following a decrease in exposure due to the gradual redistribution of Pb among other compartments via
the blood. Therefore, a single blood Pb concentration may reflect the near-term or longer-term history of
the individual to varying degrees, depending on the relative contributions of internal (e.g., bone) and
external sources of Pb to blood Pb, which in turn will depend on the exposure history and possibly age-
related and individual-specific (e.g., pregnancy, lactation) characteristics of bone turnover. In general,
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higher blood Pb concentrations can be interpreted as indicating higher exposures (or Pb uptakes);
however, they do not necessarily predict higher body burdens, especially in adults.
4.7.3. Lead Biomarkers
Observational studies using biomarkers of Pb are included in 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 (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. Overall, trends in blood Pb
levels have been decreasing among U.S. children and adults over the past twenty years. Concurrent
changes in isotopic ratios of blood Pb samples reflect changes in source composition over the past several
decades. Recent studies have observed a relationship between blood Pb and soil Pb concentration.
Additionally, studies have shown that blood Pb is also associated with non-air-related, non-policy-
relevant exposure to Pb paints in older homes, Pb released into drinking water, and occupational work
with materials containing Pb.
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
variables that may affect bone remodeling (e.g., nutritional status, pregnancy, menopause). Blood Pb in
adults is typically more an index of recent exposures than body burden, whereas bone Pb is an index of
cumulative exposure and body burden. In children, 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 some physiological
circumstances (e.g., osteoporosis), bone Pb may contribute to blood Pb in adults. The disparity between
blood Pb and body burden may have important implications for the interpretation of blood Pb
measurements in some epidemiology studies. Conceptually, measurement of long-term Pb body burden
(i.e., based on tibia Pb) may be the appropriate metric if the effects of Pb on a particular outcome are
lasting and cumulative. However, if the effects of Pb on the outcome represent the acute effects of current
exposure, then blood Pb may be the preferred metric. In the absence of clear evidence as to whether a
particular outcome is an acute effect of recent Pb dose or a chronic effect of cumulative Pb exposure, both
blood and bone metrics should be considered.
Cross-sectional studies that sample blood Pb once generally provide an index of recent exposures.
In contrast, cross-sectional studies of bone Pb and longitudinal samples of blood Pb concentrations over
time provide an index of cumulative exposure and are more reflective of average Pb body burdens over
time. 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
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variability in Pb exposures can produce episodic (or periodic) oscillations in blood Pb concentration that
may not be captured with low sampling frequencies.
The concentration of Pb in urine is a function of the urinary Pb excretion and the urine flow rate.
Urine flow rate requires collection of a timed urine sample, which is often problematic in epidemiologic
studies. Collection of un-timed ("spot") urine samples, a common alternative to timed samples, requires
adjustment of the Pb measurement in urine to account for variation in urine flow (Diamond. 1988).
Urinary Pb concentration reflects, mainly, the exposure history of the previous few months; thus, a single
urinary Pb measurement cannot distinguish between a long-term low level of exposure or a higher acute
exposure. 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
better reflect body burden.
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 AQCD was drawn from a meta-analysis by
Brunekreef (1984). In addition to the meta-analysis of Brunekreef (1984). seven more recent studies have
provided data from which estimates of the blood Pb-air Pb slope can be derived for children (Table 4-11).
The range of estimates from these seven studies is 1-9 (ig/dL per (ig/m3, which encompasses the estimate
from the Brunekreef (1984) meta-analysis of (3-6 (ig/dL per (.ig/ni3). The Schnaas et al. (2004) had a
particularly strong experimental design in that is the only longitudinal study in which blood Pb
concentration was monitored repeatedly in individual children from age 6 months to 10 years. For
children who experienced the largest declines in air Pb (i.e., from 2.8 to <0.1 (.ig/ni3), the predicted blood
Pb-air Pb slope (adjusted for age, year of birth, SES, and use of glazed pottery) was 0.213 ln|(.ig/dL
blood] per ln| (.ig/ni3 air]. The cross-sectional study done by Ranft et al. (2008) attempted to account for
potential co-variates that influence blood Pb (e.g., soil Pb concentration, gender, environmental tobacco
smoke, fossil heating system and parental education). It is the only study that reported a logarithmic blood
Pb-linear air Pb relationship, which results in an upward curvature of the blood Pb-air Pb relationship
(i.e., the blood Pb-air Pb slope increases with increasing air Pb concentration). In other studies (or based
on other studies), the blood Pb-air Pb relationship was either log-log (Brunekreef. 1984; Haves et al..
1994; Schnaas et al.. 2004). which predicts an increase in the blood Pb-air Pb slope with decreasing air Pb
concentration or linear (Hilts. 2003; J. Schwartz & Pitcher. 1989; Tripathi et al.. 2001). 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. Because air Pb
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contributes to Pb in soil and indoor dusts, adjustment for the correlated co-variates such as soil Pb would
introduce a downward bias in the slope estimate.
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Chapter 5 Contents
Chapter 5. Integrated Health Effects of LeadExposure 5-1
5.1. Introduction 5-1
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 toxic effects. 5-2
5.2.2. Altered Ion Status 5-3
5.2.2.1. Disruption of Ca2+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 5-8
Table 5-1. Enzymes and proteins potentially affected by exposure toPb and the metal
cation cofactors necessary for their proper physiological activity 5-12
5.2.2.4. Mitochondrial Abnormality 5-12
5.2.3. Protein Binding 5-15
5.2.3.1. Intranuclear and Cytoplasmic Inclusion Bodies 5-15
5.2.3.2. Cytosolic Lead Binding Proteins 5-16
5.2.3.3. Erythrocytic Lead Binding Proteins 5-17
5.2.3.4. Metallothionein 5-18
5.2.4. Oxidative Stress 5-19
5.2.4.1. 5-ALA Oxidation 5-20
5.2.4.2. Membrane and Lipid Peroxidation 5 -20
5.2.4.3. NAD(P)H Oxidase Activation 5-22
5.2.4.4. Antioxidant Enzyme Disruption 5-22
5.2.4.5. Nitric Oxide Signaling 5-24
5.2.5. Inflammation 5-25
5.2.5.1. Cytokine Production 5-26
5.2.6. Endocrine Disruption 5-28
5.2.6.1. Hypothalamic-Pituitary-Gonadal Axis 5-28
5.2.6.2. Hypothalamic-Pituitary-Thyroid Axis 5-29
5.2.7. Cell Death and Genotoxicity 5-30
5.2.7.1. DNA Damage 5-30
5.2.7.2. Mutagenicity 5-32
5.2.7.3. Clastogenicity 5-33
5.2.7.4. Epigenetic Effects 5-38
5.2.7.5. Gene Expression 5-39
5.2.7.6. Apoptosis 5-40
5.2.8. Summary 5-40
Table 5-2. Related health effects resulting from the MOAs ofPb and the lowest level
eliciting theMOA reported as bloodPb level and dose delivered 5-41
5.3. Neurological Effects 5-43
5.3.1. Introduction 5-43
5.3.2. Neurocognitive Function and Learning 5-46
5.3.2.1. Epidemiologic Studies of Cognitive Function in Children 5-46
Figure 5-2. Associations ofbloodPb levels with full-scale IQ (FSIQ) among children. 5-48
Table 5-3. Additional characteristics and quantitative results for studies presented in
Figure 5-2 5-49
Figure 5-3. Effect modification of association between concurrent blood Pb level andlQ
by blood Mn level. 5-51
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Figure 5-4. Associations of blood Pb levels with standardized scores for specific indices
of cognitive function in children. 5-54
Table 5-4. Additional characteristics and quantitative results for studies presented in
Figure 5-4 5-54
Figure 5-5. Comparing model results for 4th-grade EOG mathematics scores. 5-59
Figure 5-6. Reduction in EOG achievement test scores at each percentile of the test
distribution. 5-60
Figure 5-7. Associations ofbloodPb measures at various lifestages with cognitive
function in children. 5-63
Figure 5-8. Associations ofPb biomarkers at various lifestages with behavioral indices
in children. 5-64
Table 5-5. Additional characteristics and quantitative results for studies presented in
Figures 5-7and5-8 5-65
Figure 5-9. Regression of fitted MDI score at 36 months on log-transformed
concentration of cord blood Pb level by sex. 5-68
Figure 5-10. Associations of cognitive function in children with different degrees of
changes in blood Pb levels over time. 5-70
Table 5-6. Additional characteristics and quantitative results for studies presented in
Figure 5-10 5-71
Figure 5-11. Estimated 1Q in combined Cincinnati and Rochester cohorts for 3 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
level (white squares). 5-73
Figure 5-12. Neurological summary array of toxicological outcomes after Pb exposure.
Dosimetric representation reported by blood Pb level. (ID corresponds to
Table 5-7.) 5-74
Table 5-7. Summary of findings from neurotoxicological exposure-response array
presented in Figure 5-12. 5-75
5.3.2.2. Toxicological Studies of Neurocognition, Memory and Learning 5-75
Figure 5-13. Mean basal corticosterone levels of female and male offspring exposed to
lifetime Pb (0, 50, 150 ppm) and/or stress (PS (dam stress) or OS (offspring
stress)). 5-79
Figure 5-14. Changes in F1 performance (F1 overall performance, run rate, PRP) in
female offspring with maternal Pb exposure with various stressors in
adulthood (restraint, cold, novelty). 5-81
Figure 5-15. Changes in F1 performance (F1 overall performance, run rate, PRP) in male
offspring with maternal Pb exposure with various stressors in adulthood
(restraint, cold, novelty). 5-81
Table 5-8. Summary of effects of maternal and lifetime Pb exposure on F1 performance
watera. 5-83
5.3.2.3. Toxicological Studies on the Effects of Chelation 5-84
5.3.2.4. Epidemiologic Studies of Cognitive Function in Adults 5-86
Figure 5-16. Associations of blood and bonePb levels with cognitive function among
adults without occupational exposures to Pb. 5-88
Table 5-9. Additional characteristics and quantitative results for studies presented in
Figure 5-16 5-89
Figure 5-17. Exploration of nonlinear association of tibia Pb concentration with annual
rate of cognitive decline, by class ofHFE genotype. 5-94
Figure 5-18. Nonlinear association between patella bone Pb concentration and the
relative change in response latency over time on the pattern comparison test
(reference = 0 at mean of patella Pb concentration). 5-95
5.3.3. Neurobehavioral Effects 5-98
5.3.3.1. Epidemiologic Studies of Behavioral Effects in Children 5-98
Figure 5-19. Associations ofbloodPb levels with behavioral indices in children. 5-100
Table 5-10. Additional characteristics and quantitative results for studies presented in
Figure 5-19. 5-100
Figure 5-20. Estimates and 95% CI of direct and indirect effects of concurrent blood Pb
concentrations at age 7 on BASC-TRS. 5-103
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Figure 5-21. Estimates and 95% CI of direct and indirect effects of concurrent blood Pb
concentrations at age 7 on BASC-PRS. 5-103
Figure 5-22. Adjusted odds ratio for ADHD among U.S. children (ages 4-15 years) from
NHANES 1999-2002 by quintile of blood Pb level. 5-104
5.3.3.2. Epidemiologic Studies of Behavior, Mood, and Psychiatric Effects in Adults 5-107
5.3.3.3. Toxicological Studies of Neurobehavioral Outcomes 5-108
5.3.3.4. Toxicological Studies of Mood Alterations 5-110
Figure 5-23. Animal toxicology evidence of possible Pb-dependent contributors to the
development of mood disorders. 5-111
Figure 5-24. Schematic representation of the contribution ofPb exposure to the
development of a phenotype consistent with schizophrenia. 5-112
Figure 5-25. Neurogenesis (production of new cells) in the rat dentate gyrus after
postnatal Pb exposure. 5-113
5.3.4. Sensory Acuity 5-114
5.3.4.1. Epidemiologic Studies of Children 5-114
5.3.4.2. Epidemiologic Studies of Adults 5-114
5.3.4.3. Toxicological Studies of Sensory Organ Function 5-116
Table 5-11. Summary of toxicological studies ofPb on the retina. 5-117
Figure 5-26. Retinal a-wave and b-wave ERG amplitude in GLE adult males. 5-118
Figure 5-27. Retinal dopamine metabolism in adult control and GLE rats. 5-118
5.3.5. Neurodegenerative Diseases 5-119
5.3.5.1. Epidemiologic Studies of Adults 5-119
5.3.5.2. Toxicological Studies 5-122
5.3.6. Studies of Mechanisms of the Neurological Effects of Lead 5-126
5.3.6.1. Effects on Brain Physiology and Activity 5-126
5.3.6.2. Cholesterol and Lipid Homeostasis 5-129
5.3.6.3. Oxidative Stress 5-129
5.3.6.4. Nitrosative Signaling andNitrosative Stress 5-131
5.3.6.5. Synaptic Changes 5-131
5.3.6.6. Blood Brain Barrier 5-133
5.3.6.7. Cell Adhesion Molecules 5-134
5.3.6.8. Glial Effects 5-134
5.3.6.9. Neurotransmitters 5-135
5.3.6.10. Neurite Outgrowth 5-138
5.3.6.11. Epigenetics 5-139
5.3.7. Examination of the Lead Concentration-Response Relationship 5-139
Figure 5-28. Comparison of associations between blood Pb and cognitive function among
various bloodPb strata. 5-141
Table 5-12. Additional characteristics and quantitative results for studies presented
in Figure 5-28 5-141
5.3.8. Summary and Causal Determination 5-146
Figure 5-29. Snapshot of evidence for the spectrum of effects to the nervous system
associated with Pb exposure. 5-151
5.4. Cardiovascular Effects 5-152
5.4.1. Introduction 5-152
5.4.2. Blood Pressure and Hypertension 5-153
5.4.2.1. Epidemiology 5-153
Figure 5-30. Slope of BP (mmHg) per ug-d!. bloodPb level at 1 ug;di. or per 10 ugg
bonePb (95% CI) for associations of bloodPb (closed circles) and bone Pb
(open circles) with systolic BP (SBP; blue), diastolic BP (DBP; red), and
pulse pressure (PP; purple). 5-154
Table 5-13. Additional characteristics and quantitative data for associations of blood
and bone Pb with BP measures for results presented in Figure 5-30 5-155
Figure 5-31. Odds ratio (95%> CI) for associations of blood and bonePb with
hypertension measures. 5-158
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Table 5-14. Additional characteristics and quantitative data for associations of blood
and bone Pb with hypertension measures for results presented in Figure 5-
31 5-159
Figure 5-32. The relationship between tibia Pb and estimated SEP for those with high
self-reported stress versus those with low self-reported stress. 5-162
5.4.2.2. Toxicology 5-165
Figure 5-33. Rat blood Pb levels reported to be associated with changes in SEP from the
current ISA and 2006 Pb AQCD. 5-167
Figure 5-34. Changes in BP after Pb exposure (as bloodPb level) in unanesthetized adult
rats across studies. 5-168
5.4.2.3. Hypertension Mechanisms 5-168
5.4.2.4. Summary 5-174
Figure 5-35. Change in SEP, association size in mmHg with 95% CI, associated with a
doubling in the blood Pb concentration. 5-176
Figure 5-36. Prospective and cross-sectional increase in SEP andDEP and relative risk
of hypertension per 10 jug/g increase in bonePb levels. 5-178
5.4.3. Vascular Effects and Cardiotoxicity 5-179
5.4.3.1. Effects on Vascular Cell Types 5-179
5.4.3.2. Cholesterol 5-181
5.4.3.3. Heart Rate Variability 5-181
5.4.3.4. Peripheral Artery Disease 5-182
5.4.3.5. Ischemic Heart Disease 5-183
5.4.3.6. Atherosclerosis 5-184
Table 5-15. Characteristics and quantitative data for associations of blood and bone Pb
with other CVD measures. 5-185
5.4.3.7. Summary 5-186
5.4.4. Mortality Associated with Long-Term Lead Exposure 5-186
Figure 5-37. Multivariate adjusted relative hazards of all-cause and cardiovascular
mortality. 5-188
Figure 5-38. Multivariate adjusted relative hazard (left axis) of mortality associated with
blood Pb level between 1 jug/dL and 10 jug/dL. 5-189
Figure 5-39. Relative risk of all cause mortality for different bloodPb levels compared
with referent level of 1.5 jug/dL (12.5th percentile). 5-190
Figure 5-40. Nonlinear association between patella bone Pb concentration and the log of
HR (logHR) for all-cause, cardiovascular, and ischemic heart disease
adjusted for age, education, smoking status, and pack-years of smoking
among participants without ischemic heart disease at baseline. 5-191
Figure 5-41. Multivariate adjusted relative hazard (left axis) of mortality associated with
bloodPb levels between 1 jug/dL and 15 jug/dL. 5-192
5.4.4.1. Summary 5-192
Figure 5-42. Hazard ratios between bloodPb (closed markers), bonePb (open markers),
all-cause mortality (diamonds), and cardiovascular mortality (circles). 5-194
Table 5-16. Additional characteristics and quantitative data for associations of blood
and bone Pb with CVD mortality for results presented in Figure 5-42. 5-195
5.4.5. Air Lead-PM Studies 5-196
5.4.5.1. Hospital Admissions 5-196
5.4.5.2. Mortality 5-197
5.4.6. Summary and Causal Determination 5-197
5.5. Renal Effects 5-200
5.5.1. Introduction 5-200
5.5.1.1. Kidney Outcome Measures 5-201
5.5.2. Nephrotoxicity and Renal Pathology Related to Lead Effects 5-202
5.5.2.1. Toxicology 5-202
Table 5-17. Indicators of renal damage in male rats exposed to 50 ppm Pbfor 40 and 65
days, starting at parturition 5-204
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Table 5-18. Acute and chronic effects ofPb on the kidney/renal system — evidence from
animal toxicology studies. 5-209
5.5.2.2. Epidemiology in Adults 5-209
Figure 5-43. Kidney metric slopes on bloodPb or bonePb. 5-212
Figure 5-44. Percent change for kidney outcomes associated with blood Pb. 5-214
Table 5-19. Additional characteristics and quantitative data for associations of blood
and bone Pb with kidney outcomes for results presented in Figures 5-43 and
5-44 5-215
Figure 5-45. Added variable plot of association between serum creatinine and bloodPb
in 267 Korean Pb workers in the oldest age tertile. 5-218
Table 5-20. Patient population studies: kidney function decline 5-220
Table 5-21. Clinical randomized chelation trials in chronic kidney disease patients 5-223
5.5.2.3. Epidemiology in Children 5-224
5.5.2.4. Associations between Lead Dose and New Kidney Outcome Measures 5-225
5.5.3. Mechanisms of Lead Nephrotoxicity 5-226
5.5.3.1. Altered Uric Acid 5-226
5.5.3.2. Oxidative Damage 5-227
5.5.3.3. Lead Effect on Renal Ganglio sides 5-229
5.5.3.4. Role of Metallothionein 5-230
Figure 5-46. Dose-responsive representation of the effect ofPb on renal outcomes in
animal toxicology studies. 5-231
Table 5-22. Additional characteristics for results of toxicological studies presented in
Figure 5-46 5-232
5.5.4. Effects of Exposure to Lead Mixtures 5-232
5.5.4.1. Lead and Cadmium 5-233
5.5.4.2. Lead, Cadmium, and Arsenic 5-234
5.5.4.3. Lead and Zinc 5-235
5.5.4.4. Lead and Mercury 5-235
5.5.5. Impact of Treatment with Antioxidants on Renal Lead Accumulation and Pathology 5-236
5.5.5.1. Treatment with Antioxidants 5-236
5.5.5.2. Treatment with Antioxidants plus Chelators 5-239
5.5.6. Summary and Causal Determination 5-240
5.6. Immune System Effects 5-242
5.6.1. Introduction 5-242
Figure 5-4 7. Immunological pathways by which Pb exposure may increase risk of
immune-related diseases. 5-243
5.6.2. Cell-Mediated Immunity 5-245
5.6.2.1. T Cells 5-245
Figure 5-48. Comparisons of the relative abundance ofCD3+ T cells among groups with
increasing bloodPb level (fjg/dL). 5-247
Table 5-23. Comparison of serum abundance ofT-cell subtypesa among various blood
Pb groups. 5-24 7
5.6.2.2. Lymphocyte Activation 5-248
5.6.2.3. Delayed-type Hypersensitivity 5-249
5.6.2.4. Macrophages and Monocytes 5-250
5.6.2.5. Neutrophils 5-251
5.6.2.6. Dendritic Cells 5-252
5.6.2.7. Natural Killer (NK) Cells 5-253
5.6.3. Humoral Immunity 5-253
Figure 5-49. Comparison oflgE levels among groups with increasing bloodPb level
(Hg/dL). 5-256
Figure 5-50. Comparison of the relative abundance ofB cells among groups with
increasing bloodPb level (fjg/dL). 5-256
Table 5-24. Additional characteristics and quantitative date? for results presented in
Figures 5-49 and 5-50 5-257
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5.6.4. Immune-based Diseases 5-258
5.6.4.1. Host Resistance 5-258
5.6.4.2. Asthma and Allergy 5-259
Figure 5-51. Associations of blood Pb levels with immune-based conditions. 5-260
Table 5-25. Additional characteristics and quantitative data for results presented in
Figure 5-51. 5-261
5.6.4.3. Other Respiratory Effects 5-262
Table 5-26. Associations of air-Pb with respiratory effects 5-264
5.6.4.4. Autoimmunity 5-264
5.6.4.5. Specialized Cells in Other Tissues 5-265
Figure 5-52. Specialized macrophages in nonlymphoid tissue may serve as a significant
link between Pb and disease in multiple organ systems. 5-266
5.6.4.6. Tumors 5-267
5.6.5. Mechanisms of Lead-Induced Immunomodulation 5-267
5.6.5.1. Inflammation 5-267
5.6.5.2. Increased Prostaglandin E2 and Decreased Nitric Oxide 5-268
5.6.5.3. Cellular Death (Apoptosis, Necrosis) 5-270
5.6.5.4. Cytokine Production 5-270
5.6.6. Immune Effects of Lead within Mixtures 5-272
5.6.7. Summary and Causal Determination 5-273
5.7. Effects on Heme Synthesis and Red Blood Cell Function 5-276
5.7.1. Summary of Findings from 2006 PbAQCD 5-276
5.7.2. Effects on Red Blood Cell Functions 5-278
5.7.2.1. Lead Uptake, Binding, and Transport into Red Blood Cells 5-278
5.7.2.2. Red Blood Cell Survival, Mobility, and Membrane Integrity 5-278
5.7.2.3. Red Blood Cell Hematopoiesis 5-283
5.7.2.4. Membrane Proteins 5-283
5.7.3. Effects on Red Blood Cell Heme Metabolism 5-284
Figure 5-53. Schematic representation of the enzymatic steps involved in the heme
synthetic pathway. 5-285
5.7.3.1. Red Blood Cell 5-Aminolevulinic Acid Dehydratase 5-285
5.7.4. Other Heme Metabolism Enzymes 5-287
5.7.5. Effects on Other Hematological Parameters 5-287
5.7.5.1. Energy Metabolism 5-287
5.7.5.2. Other Enzymes 5-288
5.7.6. Red Blood Cell Oxidative Stress 5-288
5.7.6.1. Oxidative Stress, Lipid Peroxidation, and Antioxidant Enzymes 5-289
5.7.6.2. Antioxidant Defense 5-291
5.7.7. Summary and Causal Determination 5-291
5.8. Reproductive Effects and Birth Outcomes 5-293
5.8.1. Effects on Female Reproductive Function 5-294
5.8.1.1. Effects on Female Sex Endocrine System and Estrus Cycle 5-294
5.8.1.2. Effects on Fertility 5-296
Table 5-27. Summary of recent epidemiologic studies of effects on fertility for females 5-296
5.8.1.3. Effects on Puberty 5-298
Table 5-28. Summary of recent epidemiologic studies of effects on puberty for females 5-298
Figure 5-54. Toxicological Exposure-Response Array for Reproductive Effects ofPb. 5-301
Table 5-29. Toxicological Exposure-Response Array Summaries for Reproductive
Effects ofPb presented in Figure 5-54 5-302
5.8.1.4. Summary of Effects on Female Reproductive Function 5-303
5.8.2. Effects on Male Reproductive Function 5-304
5.8.2.1. Effects on Sperm/Semen Production, Quality, and Function 5-304
Table 5-30. Summary of recent epidemiologic studies of effects on sperm and semen 5-304
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5.8.2.2. Hormone Levels 5-309
Table 5-31. Summary of recent epidemiologic studies of effects on hormones for males 5-310
5.8.2.3. Fertility 5-312
5.8.2.4. Puberty 5-312
Table 5-32. Summary of recent epidemiologic studies of effects on puberty for males. 5-313
5.8.2.5. Effects on Morphology and Histology of Male Sex Organs 5-314
5.8.2.6. Summary of Effects on Male Reproductive Function 5-314
5.8.3. Effects on Ovaries, Embryo Development, Placental function, and Spontaneous Abortions 5-315
Table 5-33. Summary of recent epidemiologic studies of effects on spontaneous
abortions. 5-315
5.8.4. Infant Mortality and Embryogenesis 5-318
5.8.5. BirthDefects 5-318
Table 5-34. Summary of recent epidemiologic studies of effects on neural tube defects. 5-318
5.8.6. PretermBirth 5-319
Table 5-35. Summary of recent epidemiologic studies of effects on preterm birth. 5-320
5.8.7. Low Birth Weight/Fetal Growth 5-321
Table 5-36. Summary of recent epidemiologic studies of effects on low birth weight and
fetal growth. 5-322
5.8.8. Toxicological Studies of Developmental Effects 5-326
5.8.8.1. Developmental Effects on Blood and Liver 5-326
5.8.8.2. Developmental Effects on Skin 5-328
5.8.8.3. Developmental Effects on the Retina 5-328
5.8.8.4. Developmental Effects on Teeth 5-328
5.8.9. Summary and Causal Determination 5-328
5.9. Effects on Other Organ Systems 5-330
5.9.1. Effects on the Hepatic System 5-330
5.9.1.1. Summary of Key Findings of the Effects on the Hepatic System from the 2006 Lead
AQCD 5-331
5.9.1.2. New Epidemiologic Studies 5-331
5.9.1.3. New Toxicological Studies 5-333
5.9.2. Effects on the Gastrointestinal System 5-337
5.9.2.1. Summary of Key Findings on the Effects on the Gastrointestinal System from the 2006
Lead AQCD 5-337
5.9.2.2. New Epidemiologic Studies 5-337
5.9.2.3. New Toxicological Studies 5-338
5.9.3. Effects on the Endocrine System 5-339
5.9.3.1. Summary of Key Findings of the Effects on the Endocrine System from the 2006 Lead
AQCD 5-339
5.9.3.2. New Epidemiologic Studies 5-339
5.9.3.3. New Toxicological Studies 5-341
5.9.4. Effects on Bone and Teeth 5-342
5.9.4.1. Summary ofKey Findings of the Effects on Bone and Teeth from the 2006 Lead
AQCD 5-342
5.9.4.2. New Toxicological and Epidemiologic Studies 5-343
5.9.5. Effects on Ocular Health 5-346
5.9.5.1. Summary ofKey Findings of the Effects on Ocular Health from the 2006 Lead AQCD 5-346
5.9.5.2. New Toxicology and Epidemiology Studies 5-346
5.9.6. Effects on the Respiratory System 5-346
5.9.7. Summary 5-347
5.10. Cancer 5-349
5.10.1. Cancer Incidence and Mortality 5-349
Table 5-37. Summary of epidemiologic studies of cancer incidence and mortality 5-350
5.10.1.1. Overall Cancer Mortality 5-353
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5.10.1.2. Lung Cancer 5-354
5.10.1.3. Brain Cancer 5-354
5.10.1.4. Breast Cancer 5-355
5.10.1.5. OtherCancers 5-356
5.10.1.6. Toxicological Models of Carcinogenicity 5-357
5.10.2. Cancer Biomarkers 5-358
5.10.3. DNA and Cellular Damage 5-358
5.10.3.1. Epidemiologic Evidence for DNA and Cellular Damage 5-358
5.10.3.2. Toxicological Evidence for DNA and Cellular Damage 5-359
5.10.3.3. Mechanisms of Action 5-365
5.10.3.4. Effects of Lead within Mixtures 5-366
5.10.4. Epigenetics 5-367
5.10.5. Summary and Causal Determination 5-368
5.11. Overall Summary 5-369
Table 5-38. Summary of causal determinations for Pb. 5-369
Chapter 5. References 5-370
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Chapter 5. Integrated Health Effects
of LeadExposure
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 common to all health effects
evaluated in the ISA and those specific to particular endpoints. Subsequent sections consist of
assessments of the epidemiologic and toxicological evidence for the effects of Pb exposure on major
health effect categories such as neurological 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 effects and birth outcomes (Section 5.8). Section 5.9 provides
reviews of the evidence for Pb effects on health outcomes for which a fewer number of studies are
available, 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 health effect categories (e.g., neurological, cardiovascular, renal)
include a brief summary of conclusions from the 2006 Pb AQCD and 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). Each major section (e.g., neurological,
cardiovascular, renal effects) concludes with an integrated summary of the findings and a conclusion
regarding causality. Based upon the framework described in Chapter 1, a determination of causality is
made for a broad health effect category, such as neurological effects, with coherence and biological
plausibility being based on evidence available across disciplines and also across the suite of related health
endpoints.
Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and Environmental
Research Online) at http://eDa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of developing science
assessments such as the Integrated Science Assessments (ISA) and the Integrated Risk Information System (IRIS).
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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
targeted. A mode of action is the sequence of key events (i.e., empirically observable precursor steps) that
cumulatively result in the formation of negative health outcomes. Although the toxic 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). The following section draws information from all of the health effects
sections in the current document and identifies the major modes of action operating at the molecular,
cellular, and tissue/organ level. In turn, individual health effect sections bridge these effects to toxicities
observed on the organismal level. 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 that information with older studies that represent the current state of the science on
the possible modes of action of Pb.
Altered Ion
Status
(5.2.2}
Inflammation
(5.2.5)
Protein
Binding
(5.2,3)
Oxidative
Stress
(5.2,4)
Endocrine
Disruption
(5.2.6)
Note: The sections 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 toxic effects.
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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 toxic action,
ultimately adversely affecting 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 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.
Ca2+ homeostasis is particularly important in bone cells, as the skeletal system serves as the major
dynamic reservoir of Ca2 in the body (Long et al.. 1992; Wiemann et al.. 1999). Bone cells also are
unique in that they 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.. 1989; Schanne et al.. 1997). 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 bone cell's ability 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) and Pb versus Pb alone (Long & 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 (CRAC) influx of Ca2+ independent of
any inhibitory effect on Ca2+- ATPases.
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Ca2+ homeostasis has also been shown be to disturbed in erythrocytes due to Pb exposure
(Ouintanar-Escorza et al.. 2010; Ouintanar-Escorza et al.. 2007; Shin et al.. 2007). In blood samples taken
from Pb-exposed workers (blood Pb level = 74.4 ± 21.9 (ig/dL), the | Ca2 | was approximately 2.5-fold
higher than that seen in nonexposed workers (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. When erythrocytes from 10 healthy volunteers were exposed to Pb at concentrations of
0.2 to 6.0 (.iM for 24 or 120 hours, dose-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)) for 12-120 hours demonstrated duration-related increases at 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, 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, Ca2+ homeostasis has been observed in platelets and white
blood cells. Dowd and Gupta (1991) observed statistically significant increases in |Ca2 | in human
platelets exposed to as little as 1 (.iM Pb for 3.5 hours. The observed increase in Ca2+ levels was attributed
to increased influx of external Ca2+, possibly through receptor-operated Ca2+ channels. In mouse splenic
lymphocytes exposed to Pb, | Ca2 | was increased at exposure levels as low as 1 (.iM when incubated for
10 minutes or greater (S. Li et al.. 2008). 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, possible 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 100 nM Pb for 1-48 hours, although the observed decreases were not time-
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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, cytosolic concentrations of free Ca2+ are maintained at low levels (0.1
(.iM) by extrusion and internal compartmentalization processes (Hue! et al.. 2008). An important
component of the maintenance of Ca2+ homeostasis is transmembrane transport of Ca ions via Ca2+-
ATPase and voltage-sensitive gates (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 (blood Pb = 55.6 ± 6.3
(.ig/dL) (Kharoubi. Slimani. Aoues. et al.. 2008) and in erythrocytes of rats exposed to 0.2% Pb in
drinking water for 5 weeks (blood Pb = 97.56 ±11.8 (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 mg/L Pb in drinking water for 3 months (blood Pb = 378 (ig/dL) (Rafalowska et al..
1996) or 15 mg/kg Pb injected intraperitoneally for 7 days (blood Pb = 112.5 (ig/dL) (Struzynska.
Dabrowska-Bouta. et al.. 1997). The activity of Ca2 -ATPase in the sarcoplasmic reticulum of rabbits
exposed to 0.01 (.unol/L Pb was similarly decreased (Hechtenberg & 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 mg/L Pb acetate gestationally, and
in drinking water after weaning to the age of 6, 8, 10, or 12 weeks (H. T. Liu et al.. 2008). In 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 P0 generation (Eder et al.. 1990).
However, the F, generation of Pb-depleted rats displayed decreased activities in both enzymes (meaning
that animals with higher exposure to Pb, i.e., "control" animals, had higher enzymatic activities). A
similar increase in the activity Na+-K+-ATPase was observed in rats exposed to 20 mg/kg Pb
intraperitoneally for 14 consecutive days (Jehan & Motlag. 1995). Co-exposure of Pb with zinc and
copper greatly reduced the increase in ATPase activity observed. Although the precise mechanism was not
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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 (blood Pb = 43.0 ±
7.6 (ig/dL).
In vitro studies of ATPase activities in human erythrocyte ghosts have also shown that Pb affects
the transport of metal ions across membranes. Calderon-Salinas et al. (1999) observed that 10-50 mM Pb
and Ca2+ were capable of inhibiting the passive transport of each other in human erythrocyte ghosts
incubated with both cations. Subsequent inhibition experiments indicated that both cations share the same
electrogenic transport pathway (Sakuma et al.. 1984). Further study by this group (Calderon-Salinas.
Ouintanar-Escorza. et al.. 1999) demonstrated that Pb can noncompetitively block the transport of Ca2 by
inhibiting the activity of Ca2+-Mg2+-ATPase at concentrations of 1-5 mM. 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 the activity ofNa+-K+-ATPase was
decreased in erythrocyte ghosts exposed to concentrations as low as 10 (ig/dL Pb; 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 (blood Pb = 12.34 ±
2.44 (males) and 16.85 ± 6.01 (ig/dL (females)). Increased 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. Huel et al. (2008) found that
newborn hair and cord blood Pb levels (1.22 ± 1.41 jj.g/g and 3.54 (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 cord blood Pb
Pb has also been shown to disrupt cation transport mechanisms through direct action on voltage-
sensitive cation channels. Audesirk and Audesirk (1991. 1993) 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 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) confirmed 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,
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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 (Uiihara 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 0.2 nM, possibly through blocking the intracellular Ca2+ binding domain, or through Ca2+
dependent dephosphorylation of the channel (L. R. Sun & Suszkiw. 1995).
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 dose-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 the channel's 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 dose-
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 & Manalis. 1984; J. Suszkiw et al..
1984). However, concentrations of Pb as low as 5 (.iM were also observed to actually increase the
spontaneous release of neurotransmitters in these same experiments. Subsequent research by other groups
confirmed that Pb demonstrates Ca2+-mimetic properties in enhancing neurotransmitter release from cells
in the absence of Ca2+ and Ca2+-induced depolarization. Tomsig and Suszkiw (1993. 1995) reported that
Pb exposure induced the release of norepinephrine (NE) from bovine adrenal chromaffin cells, and was
considerably more potent at doing so than Ca2+ (K0 5 of 4.6 nM for Pb versus 2.4 (.iM for Ca2+). Activation
of protein kinase C (PKC) was observed to enhance the Pb-induced release of NE. Westrink and Vijverber
(2002) observed that Pb acted as a high affinity substitute for Ca2+, and triggered enhanced catecholamine
release from PC12 cells at 10 (.iM in intact cells and 30 nM 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 GD15-16 and
continuing to 120 days postnatal, decreases in total potassium-stimulated hippocampal GABA release
were seen at exposure levels of 0.1-0.5% (blood Pb = 26.8 ± 1.3 - 61.8 ± 2.9 (ig/dL) (Laslev & Gilbert.
2002). Maximal effects were observed at 0.2% Pb in drinking water, but effects were less evident at 0.5%,
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and were absent at 1.0%. In the absence of Ca2+, potassium-induced release was increased in 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 & 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 & 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 (Garza et al.. 2006;
Kirberger & Yang. 2008). The coordination chemistry that normally regulates metal-protein binding
makes many proteins particularly susceptible to perturbation from Pb, as it is able to function with
flexible coordination numbers and can bind multiple ligands (Garza et al.. 2006; Kirberger & Yang.
2008). However, due to differences in its physical properties, Pb induces abnormal conformational
changes when it binds to proteins (Bitto et al.. 2006; Garza et al.. 2006; Kirberger & Yang. 2008; 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 protein kinase C (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 et al.. 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 Ca2+-dependent activation by five orders of
magnitude. Long et al. (1994) confirmed these findings, reporting that Pb had a Kact 4800 times smaller
than Ca2+ (55 pM versus 25 |_iM. following a 3 minute exposure). However, Ca2+ had a higher maximal
activation of PKC than 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 Ca2+, but not subsequent sites. Tomsig
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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 100 nM (Belloni-Olivi et al.. 1996). The observation that
neither Ca2+ nor diacylgycerol was increased due to exposure indicates that Pb-induced activation of PKC
is due to direct interaction with the protein. Pb-induced alterations in PKC have also been observed in
other tissues, including increased activity in rabbit mesenteric arteries at picomolar concentrations of Pb
(Chai & Webb. 1988; Watts et al.. 1995) and human erythrocytes from Pb-exposed workers (blood Pb =
5.4 to 69.3 (ig/dL) (K.-Y. Hwang et al.. 2002). and decreased activity in mouse macrophages and the rat
brain cortex at micromolar concentrations (Lison et al.. 1990; Murakami et al.. 1993).
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 (Saimi & Kung. 2002; Vetter & Leclerc. 2003).
Haberman 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 100 nM, EC50 = 0.5-1.0 stimulation of brain membrane phosphorylation at Pb
concentrations greater than 400 nM after 1 minute incubation, and increased binding of calmodulin to
brain membranes at Pb concentrations greater than 1 (.iM after 10 minutes incubation. Haberman et al.
(1983) reported the affinity of Pb for calmodulin's Ca2+-binding sites was approximate to that of Ca2+
itself (Kd ~ 20 (iM), whereas Richardt et al. (1986) observed that Pb was slightly more potent than Ca2+ at
binding calmodulin (IC50 =11 and 26 (.iM, respectively). Both studies indicated that Pb was much more
effective at binding calmodulin than 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 400-550 pM (threshold =100 pM) and 450-500 nM
(threshold =100 nM), 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 430 pM [threshold = 300 pM versus EC50 1200 nM (threshold = 200 nM; 50-minute
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incubation]). When incubated with physiological concentrations of Ca2+, Pb induced phosphodiesterase
activity at concentrations as low as 50 pM. Pb activated calcineurin, a phosphatase with widespread
distribution in the brain and immune system, at threshold concentrations as low as 20 pM in the presence
of Ca2+ (incubation time = 30 minutes), but inhibited its activity at concentrations greater than 200 pM
(Kern & 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-dose (<1 |_iM. 20-minute incubation) stimulatory effects
of Pb exposure on the activity of Ca2+-Mg2+-ATPase was due to Pb binding to calmodulin and subsequent
activation of the ion pore. Ferguson et al. (2000) observed that exposure of rat hippocampal neurons to Pb
for 1 to 48 hours resulted in increased activation of a calmodulin-dependent Ca2+ extrusion mechanism.
Pb has also been observed to alter the activity of other proteins that rely on Ca2+ binding for normal
cellular function. Osteocalcin is a matrix protein important in bone resorption, osteoclast differentiation,
and bone growth and has three Ca2+-binding sites (Dowd et al.. 2001). Incubation of osteocalcin in
solution with Ca2+ and Pb resulted in the competitive displacement of Ca2+ by Pb (Dowd et al.. 1994). Pb
was found to bind to osteocalcin more than 1000-times more tightly than Ca2+ (Kd = 1.6 ± 0.42 nM versus
0.007 mM, respectively), and analysis with NMR indicated 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 confirmed that Pb bound osteocalcin approximately 10,000-times more tightly than Ca2+ (Kd =
0.085 (.iM versus 1.25 mM, 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 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 syntapotagmin I, at concentrations as low as the
nanomolar range (Bouton et al.. 2001; Qsterode & Ulberth. 2000; Richardt et al.. 1986).
Pb can displace metal cations other than Ca2+ that are requisite for protein function. One of the most
researched targets for molecular toxicity of Pb is the second enzyme in the heme synthetic pathway,
aminolevulinic acid dehydratase (ALAD). ALAD contains four 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. Grubb. et al.. 1997). and exposure to Pb results in
inhibition of the enzyme in the erythrocytes of Pb-exposed workers and adolescents (blood Pb level >10
(.ig/dL) (Ademuviwa. Ugbaia. Oio. et al.. 2005; Ahamed et al.. 2006). in human erythrocytes exposed to
Pb for 60 minutes (K = 0.07 pM) (Simons. 1995). and in rats exposed to 25 mg/kg Pb once a week for 4
weeks (blood Pb level = 6.56 ± 0.98 (ig/dL) (M. K. Lee et al.. 2005). Further experiments indicated that
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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. 1995V
Zinc-binding domains are also found in transcription factors and proteins necessary for gene
expression, including GATA proteins and transcription factors TFIIIA, Spl, and Erg-1 (Ghering et al..
2005; Hanas et al.. 1999; M. Huang et al.. 2004; Zawia et al.. 1998) (G. R. Reddv & Zawia. 2000). Pb
was found to form tight complexes with the cysteine residues in GATA proteins (|3Pb = 6.4 / 10 ' M"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 = 1 M). Pb also binds to the zinc domain of TFIIIA, inhibiting its ability to bind DNA at
concentrations as low as 10 (.iM (Hanas et al.. 1999; M. Huang et al.. 2004). Huang et al. (2004) also
reported that exposure to Pb caused the dissociation of TFIIIA-DNA adducts and that NMR spectroscopy
indicated that altered TFIIIA activity is the result of a Pb-induced abnormal protein conformation.
Pb exposure modulated the DNA-binding profiles of the transcription factors Spl and Erg-1 in rat
pups exposed to 0.2% Pb acetate via lactation, resulting in a shift in DNA-binding towards early
development (i.e., the first week following birth) (G. R. Reddv & Zawia. 2000; Zawia 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 & Zawia. 2000). However, in DNA-binding
assays utilizing recombinant Sp 1 (which has three zinc finger motifs, opposed to only one in the synthetic
peptide), incubation with as little as 37 (.iM Pb resulted in the abolishment of Spl's DNA-binding
capabilities (Razmiafshari & 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 (blood Pb = 7.01 ± 1.64 (ig/dL) and in human erythrocyte lysates
exposed to 100 nM Pb for as little as 5 minutes (Baranowska-Bosiacka et al.. 2009). and cGMP
phosphodiesterase at picomolar concentrations in homogenized bovine retinas (D. 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. These alterations in protein function have
implications for numerous cellular and physiological processes, including cell signaling, growth and
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differentiation, gene expression, energy metabolism, and biosynthetic pathways. Table 5-1 provides a list
of enzymes and proteins whose function may be 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 Action
Metal Cation; Reference
Aminolevulinic acid dehydratase
1
Zn: Simons (1995)
Ferrochelatase
1
Fe (2Fe-2S Cluster); Crooks (2010)
Superoxide dismutase
It
Mn, Cu, Zn, Fe; Antonyuk et al. (2009).
Borqstahl et al. (1992)
Catalase
It
Fe (Heme): Putnam etal. (2000)
Glutathione peroxidase
It
Se: Rotruck et al. (1973)
Enzymes
Guanylate cyclase
1
Fe (Heme); Boerrigter and Burnett (2009)
cGMP phosphodiesterase
1
Ma. Zn: Ke (2004)
NAD synthase
1
Mg; Haraetal. (2003)
NAD(P)H oxidase
t
Ca: Lesenev (1999)
Pyrimidine 5'-nucleotidase
1
Mg, Ca; Bitto etal. (2006),Amici (1997),
Paglia and Valentine (1975)
Erythrocyte
phosphoribosyltransferase
1
Mg (Mn, Ca, Co, Ni, Zn); Deng et al. (2010).
Arnold and Kelley (1978)
ATPase
It
Ca, Mg, Na-K; Technische Universitat
Braunschweiq (2011)
Ion Channels/
Transport
Mitochondrial transmembrane
pore
t
Ca; He etal. (2000)
Calcium-dependent potassium
channel
t
Ca: Silkin et al. (2001). Alvarez et al. (1986)
Signal
Protein kinase C
It
Ca; Garza et al. (2006)
Transduction
Calmodulin
t
Ca; Garza et al. (2006)
Pb Binding
Metallothionein
t
Zn, Cu; Yu et al. (2009)
DNA Binding
GATA transcriptional factors
1
Zn; Hanas et al. (1999). Huang et al. (2004)
| indicates increased activity; J, indicates decreased activity; indicates activity can be alternatively increased or decreased.
5.2.2.4. Mitochondrial Abnormality
3 Alterations in mitochondrial function, including disruptions in ion transport, ultrastructural
4 changes, altered energy metabolism, and perturbed enzyme activities due to Pb intoxication are well
5 documented in the scientific literature. Exposure of rats to Pb in feed (1% Pb for 4, 6, 8, 10, 12, or 20
6 weeks) or drinking water (300 ppm for 8 weeks, 500 ppm for 7 months, or 1% for 9 months) resulted in
7 gross ultrastructural changes in renal tubule and epididymal mitochondria characterized as a general
8 swollen appearance with frequent rupture of the outer membrane, distorted cristae, loss of cristae,
9 frequent inner compartment vacuolization, observation of small inclusion bodies, and fusion with adjacent
10 mitochondria (Gover. 1968; Goveretal.. 1968; Marchlewicz et al.. 2009; Navarro-Moreno et al.. 2009; L.
1 1 Wan a et al.. 2010V
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Transmembrane mitochondrial ion transport mechanisms are perturbed by exposure to Pb. Pb
inhibits the uptake of Ca2 into mitochondria (Parr & 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
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 (L. Wang. H. Wang, et al.. 2009). and retinal rod photoreceptor cells
incubated with 10 nM to 10 (.iM for 15 minutes (L. H. He et al.. 2000). Further research indicated that
observed 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; L. H. He et al.. 2000). Opening of the
MTP is the first step of the mitochondrial-regulated apoptotic cascade pathway in many cells (Lidskv &
Schneider. 2003; Rana. 2008). 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 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.. 1980). Statistically significant Pb-induced decreases in ATP and adenylate energy charge
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) and cellular ATP levels were decreased in
differentiated PC-12 cells incubated with as little as 1 (.iM Pb for 48 hours (Prins et al.. 2010). The
observed decrease in cellular ATP levels in Prins et al. (2010) were correlated with a Pb-induced decrease
in the expression of the voltage-dependent anion channel (VDAC), which maintains cellular ATP levels in
neurons. 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
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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. 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 mg/L 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. (1997), 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 (Dowd et al.. 1990; Grabowska & Guminska. 1996). 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 level = 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 & Hlvnczak. 2003). These decreases in nucleotide pools were accompanied by an
increase in purine degradation products (adenosine, etc.). Similar decreases in cellular nucleotide pools
were observed when rats were exposed to 1% Pb in drinking water for four weeks (Baranowska-Bosiacka
& Hlvnczak. 2004). In erythrocytes, nucleotides are synthesized via salvage pathways such as the adenine
pathway, which requires adenine phosphoribosyltransferase (APRT). The activity of this enzyme is
inhibited by exposure to Pb in human and rat erythrocytes (see above for dose 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 (blood Pb = 34.6
± 20.7 (ig/dL) and observed an apparent dose-dependent decrease in NAD synthetase activity with
increased blood Pb. The blood Pb 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., 5-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
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heme. Farant and Wigfield (1987. 1990) 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. Taketami 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 statistically significantly associated with increased erythrocyte ZPP levels
in humans (average blood Pb ranging from 21.92 to 53.63 (.ig/dL) (Ademuviwa. Ugbaia. Oio. et al.. 2005;
Counter et al.. 2007; Mohammad et al.. 2008; Patil. Bhagwat. Patil. Dongre. Ambekar. Jailkhani. et al..
2006) and animals (blood Pb = 24.7 (.ig/dL) (Rendon-Ramirez et al.. 2007).
5.2.3. Protein Binding
Pb is able to bind 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 toxic effects through sequestration. The ability of
Pb to bind proteins was first reported by Blackmon (1936): Pb intoxication was observed to induce the
formation of intranuclear inclusion bodies in the liver and kidney. Further research into the composition
of intranuclear inclusion bodies and the identification of specific Pb-binding proteins has been conducted
since that time.
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 (Choie & Richter. 1972; Fow ler et al.. 1980; Gover.
Leonard, et al.. 1970; Gover. May, et al.. 1970; McLachlin etal.. 1980; Navarro-Moreno et al.. 2009;
Oskarsson & Fowler. 1985). 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 (Deveci. 2006; Navarro-Moreno et al.. 2009). Intranuclear and cytoplasmic inclusions have
also been found in organs other than the kidney, including liver, lung, and glial cells (Gover & Rhvne.
1973; J. Singh et al.. 1999). Pb found within nuclei has also been shown to bind to the nuclear membrane
and histone fractions (Sabbioni & Marafante. 1976).
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Upon denaturing intranuclear inclusion bodies with strong denaturing agents, Moore et al. (1973)
observed that proteins included in the bodies were rich in aspartic and glutamic acid, glycine, and
cysteine. Further work by Moore and Goyer (1974) characterized the protein as a 27.5 kDa protein that
migrates as a single band on acrylamide gel electrophoresis. In contrast with Moore and Goyer's findings,
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
(Choie et al.. 1975; McLachlin etal.. 1980). 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 & 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 & 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 postnatal exposure of rats
to 0.2-1.0% Pb in drinking water increased the levels of another brain protein, glial fibrillary acidic
protein (GFAP), 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) observed three proteins (MW = 11.5, 63, and >200
kDa) in rat kidney cytosol, and that 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 & Fowler. 1984. 1985). Cadmium and zinc, but not Ca2 or
Fe, prevent 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 intranuclear inclusion bodies
(Mahaffev et al.. 1981; Mahaffev & Fowler. 1977; Mistrv et al.. 1986). 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)
(Fow ler & DuVal. 1991; D. R. Smith et al.. 1998).
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 (DuVal & Fowler. 1989; Goering et al..
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1986; Ouintanilla-Vega et al.. 1995). 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 (DuVal & Fowler. 1989; Fowler et al.. 1993V Some evidence exists that cytosolic Pb-binding
proteins directly target Pb and compartmentalize intracellular Pb as protective measure against toxicity
(Y. Oian et al.. 2000; Y. Oian et al.. 2005).
5.2.3.3. Erythrocytic Lead Binding Proteins
The majority (94%) of Pb in whole blood is found in erythrocytes (Ong & Lee. 1980a). Originally,
the major Pb-binding protein in erythrocytes was identified as hemoglobin (B. S. Cohen et al.. 2000;
Lolin & O'Gorman. 1988; Ong & Lee. 1980a. 1980b; Raghavan & Gonick. 1977). However, Bergdahl et
al. (1997) 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 |_ig/dL) and rats (Pb exposure = 25 mM in drinking
water) (Boudene etal.. 1984; Fujita et al.. 1981; Fujita et al.. 1982).
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 (Astrin et al.. 1987; H.-S. Kim et al.. 2004; Perez-Bravo et al.. 2004; Scinicariello et al..
2007; C. M. Smith. Hu. et al.. 1995; Wetmur. 1994; Wetmur et al.. 1991; Y. Zhao et al.. 2007). 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 (Y. Chen et al.. 2008; Chia et al.. 2007; Chia et
al.. 2006; E. F. Krieg. Jr. et al.. 2009; Scinicariello et al.. 2010; Wananukul et al.. 2006).
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. Grubb. et al.. 1997). 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 (B. S. Schwartz. Lee. Stewart. Sithisarankul. etal.. 1997).
decreased levels of zinc protoporphyrin (H.-S. Kim et al.. 2004; Scinicariello et al.. 2007). lower cortical
bone Pb (C. M. Smith. Wang, et al.. 1995). and lower amounts of DMSA-chelatable Pb (B. S. Schwartz et
al. 2000; B. S. Schwartz. Lee. Stewart. Ahn. et al.. 1997; Scinicariello et al.. 2007). However, the
findings that ALAD-2 polymorphisms reduced the bioavailability of Pb are somewhat equivocal. Wu et
al. (2003) 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
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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. ("2003) 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 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 suggests that the increased plasma levels of Pb in subjects
with ALAD 1-2/2-2 genotypes increases the probability of adverse health effects in these individuals.
ALAD's capacity for binding Pb has been estimated at 85 (ig/dL in erythrocytes and 40 |_ig/dL in
whole blood (Bergdahl et al. 1998). The 45 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.. 1980.
1981; Raghavan & 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.. 1980. 1981). 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. 1993b). 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
ALAD's Pb-binding capacity identified by Bergdahl et al. (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. 1993b) 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 metal-binding protein, most often zinc or copper, that is rich in
cysteine residues and plays an important role in the protection against heavy metal toxicity, trace element
homeostasis, and scavenging free radicals (J. Yu et al.. 2009). Exposure to Pb acetate induces the
production of Pb- and Zn-metallothionein in mice exposed via i.p. or i.v. injection at 30 mg/kg (Maitani et
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al.. 1986). in mice exposed via i.p. injection at 300 |_imol/kg (J. Yu et al.. 2009). or in rats exposed via i.p.
injection at 24 (imol/100g (Ikebuchi etal.. 1986). The induced Pb-metallothionein consisted of 28% half-
cysteine and reacted with an antibody for Zn-metallothionein II (Ikebuchi etal.. 1986). In contrast,
exposure of rats to Pb via drinking water (200 or 300 mg/L) failed to induce metallothionein in the
kidneys or intestines (Jamieson et al.. 2007; L. Wang. D. W. Chen, et al.. 2009). Goering and Fowler
(1987a. 1987b) 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 &
Fowler. 1987a. 1987b). 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 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 (Ou et al.. 2002). Interestingly,
metallothionein-null mice were unable to form intranuclear inclusion bodies and accumulated less renal
Pb than the wild-type mice (Ou 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 (Tokar et al. 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. (2010) observed that Pb-exposed workers with a mutant metallothionein allele had
higher blood Pb levels than carriers of the normal allele (24.17 and 21.27 versus 17.03 (ig/dL), and were
more susceptible to Pb toxicity.
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), that 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
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erythrocytes, which may be responsible for a number of Pb-induced pathologies (Gonick et al.. 1997;
Khalil-Manesh. Gonick. Cohen. Bergamaschi. et al.. 1992; Khalil-Manesh et al.. 1994; Salawu et al..
2009; Sandhir & Gill. 1995; Shan et al.. 2009; Vaziri. 2008b). 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).
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-ALA dehydratase (S-ALAD)
both identified as targets (Bergdahl. Grubb. et al.. 1997). Through its sulfhydryl and metal ion disrupting
properties, Pb incorporates with and inhibits a number of enzymes in the heme biosynthetic process,
including 8-ALA synthetase, S-ALAD, and ferrochelatase. Pb is able to disrupt the Zn ions requisite for
the activity of 8-ALAD, the rate limiting step in heme synthesis, leading to enzyme inhibition at
picomolar concentrations (Simons. 1995). Additionally, low blood Pb levels (7 (ig/dL) have been found to
inhibit the activity of 8-ALAD in humans with a threshold value around 5 (ig/dL (Ahamed et al.. 2005;
Sakai & Morita. 1996). A significant negative correlation (r = -0.6) was found between blood Pb levels in
adolescents (4-20 (ig/dL) and blood 8-ALA activity (Ahamed et al.. 2006). This inhibition of 8-ALAD
results in the accumulation of S-ALA in blood and urine, where S-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 S-ALA produces 02, OH', H202, and an ALA radical (Monteiro et al..
1989; Monteiro etal.. 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 S-ALA oxidation, as
described above, interacts with and disrupts membrane lipids (Bechara et al.. 1993; Oteiza et al.. 1995).
Additionally, Pb has the capacity to stimulate ferrous ion initiated membrane lipid peroxidation serving as
a catalyst for these events (Adonaylo & Oteiza. 1999; Quinlan et al.. 1988). Increased lipid peroxidation
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measured as TBARS from liposomes, microsomes, and erythrocytes was shown after Pb (Pb(N03)2)
exposure for < 2 hours at concentrations as low as 10 (.iM in vitro (Aruoma et al.. 1989; Ouinlan 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 & Lin. 1995). After Pb exposure (4-12 (ig/dL, 24 hours, in vitro), 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 does not promote lipid
peroxidation, however it may accelerate peroxidation by H202 (Ouinlan et al.. 1988). This could be due to
altering membrane structure, restricting phospholipid movement, and facilitating the propagation of
peroxidation.
Pb induces changes in the fatty acid composition of a membrane, which could lead to oxidative
damage. Exposure to Pb (>62.5 ppm in drinking water, 3 weeks) in chicks promoted an increase in
arachidonic acid (AA, 20:4) as a percentage of total fatty acids, and decreased the relative proportion of
shorter chain fatty acids (linoleic acid, 18:2) (Lawton & 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 & Knowles. 1993). Changes in membrane fluidity will disturb
the conformation of the active sites of membrane associated enzymes, disrupt metabolic regulation, and
alter membrane permeability and function.
A number of recent studies report increased measures of lipid peroxidation in various organs,
tissues, and species. Occupational Pb exposure resulting in elevated blood Pb levels (>8 (ig/dL) in various
countries provides evidence of lipid peroxidation, including increased plasma malondialdehyde levels
(Dogru et al.. 2008: Ergurhan-Ilhan et al.. 2008: D. A. Khan et al.. 2008: Mohammad et al.. 2008: Patil.
Bhagwat. Patil. Dongre. Ambekar. & Das. 2006: Patil. Bhagwat. Patil. Dongre. Ambekar. Jailkhani. et al..
2006: Ouintanar-Escorza et al.. 2007). One study found a correlation between the MDA levels and blood
Pb levels even in the unexposed workers who had lower (i.e., <12 jj.g/dl) blood Pb levels (Ouintanar-
Escorza et al.. 2007). Other studies found increased evidence of lipid peroxidation among the general
population, including children, with elevated blood Pb levels (>10 (ig/dL) (Ahamed et al.. 2008: Ahamed
et al.. 2006: Y. P. Jin et al.. 2006). Ahamed et al. (2006) found a significant positive correlation (r = 0.7)
between blood Pb levels between 4-20 (ig/dL in adolescents and blood MDA level. Similar results have
been shown after Pb exposure in animal studies (Adegbesan & Adenuga. 2007: M. K. Lee et al.. 2005:
Pandva et al.. 2010: D. Y. Yu et al.. 2008). An enhanced rate of lipid peroxidation has been found in Pb
treated (50 jj.g, 1-4 hours) rat brain homogenates (Rehman et al.. 1995) and in specific brain regions,
hippocampus and cerebellum, after Pb exposure (500 ppm, 8 weeks) to rats (Bennet et al.. 2007). Overall,
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there was a correlation between the blood Pb level and measures of lipid peroxidation often measured by
MDA levels.
Interestingly, studies in many species of plants, invertebrates, and vertebrates exhibit increased
lipid peroxidation (Chapter 7.4.4). 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 (Lcsenev et al.. 1999). Two studies provide
evidence for increased activation of NAD(P)H oxidase contributing 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 gp91 phox in the brain, heart, and renal cortex of Pb treated rats (100 ppm in
drinking water, 12 weeks). This upregulation was present in Pb-treated (1-10 ppm) human coronary artery
endothelial cells, but not vascular smooth muscle cells (VSMC), which do not express the protein (Ni et
al.. 2004). It is possible thatNAD(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 will 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 it 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
metalloproteins 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 et al.. 1992). A number of studies have found decreased activity
of these enzymes (Ergurhan-Ilhan et al.. 2008; Mohammad et al.. 2008; Pandva et al.. 2010; Patil.
Bhagwat. Patil. Dongre. Ambekar. & Das. 2006; Patil. Bhagwat. Patil. Dongre. Ambekar. Jailkhani. et al..
2006; D. Y. Yu et al.. 2008). whereas others observe increased activity (Ahamed et al.. 2008; M. K. Lee et
al.. 2005). Pb exposure (500 ppm, 1, 4, and 8 weeks) in rats showed that organ SOD and CAT responded
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differently depending on the dose and tissue investigated. Activity of SOD and CAT varied based on the
brain region analyzed and time of exposure (Bonnet 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 mg/1,
blood Pb level 20.4 and 24.5 j^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 decrease levels of GSH. Acute administration of Pb (0.1 (.iM in vitro, 18
(ig/dL human) results in decreased blood and organ GSH and cysteine content, which may be due to
increased GSH efflux from tissues (Ahamed et al.. 2008; Ahamed et al.. 2009; Ahamed et al.. 2005;
C'hettv et al.. 2005; Flora et al.. 2007; Nakagawa. 1989. 1991; Pandva et al.. 2010). Chronic 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 (Corongiu & Milia. 1982; Daggett et al.. 1998;
J. M. Hsu. 1981). However, other studies have found that chronic Pb exposure resulting in blood Pb level
between 6.6 and 22 (.ig/dL. causes the depletion of GSH (Ercal et al.. 1996; M. K. Lee et al.. 2005;
Mohammad et al.. 2008). Thus, the time of exposure is important to consider when measuring GSH
levels.
Glutathione reductase (GR) is able to reduce GSSG back to GSH. Therefore, an increased
GSSG/GSH ratio is considered indicative of oxidative stress. Pb exposure has been shown to increase the
GSSG/GSH ratio (Ercal et al.. 1996; Mohammad et al.. 2008; Sandhir & Gill. 1995). even at blood Pb
level below 10 (ig/dL in children (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 have found
decreased (Bokara et al.. 2009; Sandhir & Gill. 1995; Sandhir et al.. 1994). increased (Howard. 1974;
Sobekova et al.. 2009). and no change (J. M. 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, Pb is able to decrease the activity of GPx and GST (M. K. Lee et al.. 2005;
Nakagawa. 1991; Schrauzer. 1987; D. Y. Yu et al.. 2008). Similar to other antioxidant enzymes,
compensatory upregulation of these enzymes is described after Pb treatment (Bokara et al.. 2009;
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Conterato et al.. 2007; Daggett et al.. 1998; Ergurhan-Ilhan et al.. 2008). However, other studies have
shown that these enzymes may not be able to compensate for the increased Pb-induced ROS, further
contributing to the oxidative environment (Farmand et al.. 2005).
Recently, y-glutamyltranferase (GGT) within its normal range has been regarded as an early and
sensitive marker of oxidative stress. This may be because cellular GGT metabolizes extracellular GSH to
be used in intracellular GSH synthesis. Thus, cellular GGT acts as an antioxidant enzyme by increasing
the intracellular GSH pool. However, the reasons for the association between GGT and oxidative stress
have not been fully realized (D. H. Lee et al.. 2004). Occupational Pb exposure (blood Pb level, 29.1
(ig/dL) results in increased serum GGT levels (D. A. Khan et al.. 2008). Interestingly, low blood Pb levels
found in a sample of the U.S. population (NHANES III) were positively associated with serum GGT
levels showing a dose-response relationship at levels <7 (ig/dL (D. H. Lee et al.. 2006).
5.2.4.5. Nitric Oxide Signaling
NO, also known as endothelium-derived relaxing factor, is a potent endogenous signaling
molecule involved in vasodilation. Acute and chronic Pb exposure decreases the biologically active NO,
not through reduction in NO-production capacity (Vaziri & Ding. 2001; Vaziri. Ding, et al.. 1999). but as
a result of inactivation and sequestration of NO by ROS (Malvezzi et al.. 2001; Vaziri. Liang, et al..
1999). 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. Liang, et al..
1999). NO is also produced by macrophages in the defense against certain infectious agents, including
bacteria. Studies have found that Pb exposure can significantly reduce production of NO in immune cells
(J.-E. Lee et al.. 2001; Pineda-Zavaleta et al.. 2004; L. Tian & Lawrence. 1995). possibly leading to
reduced host resistance (L. Tian & 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), that 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 significant upregulation in NOS isotypes (Gonick et al.. 1997; Vaziri & Ding. 2001; Vaziri. Ding, et
al.. 1999). A direct inhibitory action of Pb on NOS enzymatic activity has been rejected (Vaziri. Ding, et
al.. 1999). Instead, the upregulation of NOS occurs as compensation for the decreased NO resulting from
ROS inactivation (Vaziri & Ding. 2001; Vaziri et al.. 2005; Vaziri & 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 (GTP). Soluble guanylate cyclase is a heterodimer requiring one molecule of
heme for enzymatic activity (Boerrigter & Burnett. 2009). In vascular smooth muscle cells (VSMC), sGC
serves as the NO receptor. Marked reduction in plasma concentrations and urinary excretion of cGMP is
observed after Pb exposure (5 ppm for 30 days and 0.01% for 12 months) (khalil-Manesh. Gonick.
Weiler. et al.. 1993; M. Marques et al.. 2001). In addition, Pb exposure downregulated the protein
abundance of sGC in vascular tissue (Courtois et al.. 2003; Farmand et al.. 2005; M. Marques et al..
2001). This downregulation in sGC was prevented by antioxidant therapy suggesting that oxidative stress
also plays a role in Pb-induced downregulation of sGC (M. Marques et al.. 2001).
5.2.5. Inflammation
Misregulated inflammation represents one of the major effects of Pb-induced immunotoxicity. 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 (C'hettv et al.. 2005).
Traditional immune mediated inflammation can be seen with bronchial hyper-responsiveness,
asthma, and respiratory infections after 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 (Carmignani et al.. 2000; Mudipalli.
2007). 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 (18 (.ig/dL or 100 ppm for 14 weeks) (Ramesh et al.. 2001; Rodriguez-ltiirbe et al.. 2005).
which has been coupled with activation of the redox sensitive nuclear transcription factor kappa B
(NFkB) and lymphocyte and macrophage infiltration (23-27 (ig/dL, 100 ppm for 14 weeks) (Bravo et al..
2007). These events could be in response to the oxidative environment developed from Pb, 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 humans (48 (ig/dL) and cell models (0.01 (.iM, 48 h)
(Cardenas et al.. 1993; C'hettv et al.. 2005; Flohe et al.. 2002; Knowles & Donaldson. 1997; J. J. Lee &
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Battles. 1994). Dietary Pb supplementation (500 ppm, 19 days) can increase the percentage of cell
membrane AA, the precursor of cyclooxygenase and lipoxygenase metabolism to PGs and leukotrienes
(Knowles & Donaldson. 1990). Additionally, Pb (1 (.iM. 20 (.ig/dL) may promote the release of AA via
activation of phospholipase A:, as shown in isolated VSMC (Dorman & Freeman. 2002V
Inflammation may be the result of increased pro-inflammatory signaling or may exacerbate these
signaling pathways. Pb can elevate the expression of NFkB, as well as expression of activator protein-1
(AP-1) and cJun (Bravo et al.. 2007; korashv & Ei-kadi. 2008; korashv & El-Kadi. 2008; Pvatt et al..
1996; Ramesh et al.. 1999). Pb exposure (25 (.iM) to dendritic cells stimulates phosphorylation of the
Erk/MAPK pathway, but not p38, STAT3 or 5, or CREB (D. Gao et al.. 2007)
5.2.5.1. Cytokine Production
There are three modes of major effects of Pb on immune cytokine production. First, Pb can act on
macrophages to elevate the production of pro-inflammatory cytokines such as TNF-a and IL-6 (S. Chen
etal.. 1999; Y.-J. Cheng et al.. 2006; 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 Thl responses are suppressed and Th2 responses are
promoted (S. Chen et al.. 2004; Miller etal.. 1998). Third, when acquired immune responses occur
following exposure to Pb, Thl lymphocyte production of cytokines is suppressed (e.g. IFN-y) (Heo et al..
1996; Lvnes et al.. 2006). In contrast, Th2 cytokines such as IL-4, IL-5, and IL-6 are elevated (D. Gao et
al.. 2007; D. Kim & 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.
Iavicoli et al. (2006) reported below-background blood Pb concentrations produced significant
changes in cytokine levels. At the lowest dietary Pb concentration (0.8 (ig/dL), IL-2 and IFN-y were
elevated over the controls, indicating an enhanced Thl response. However, as the dietary and blood Pb
concentrations increased (resulting in blood Pb level 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 extremely 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 relative 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 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 and increased IL-6 was found in the forebrain (Struzvnska et al.. 2007).
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Consistent with animal studies, epidemiologic studies have also demonstrated Pb-associated
decreases in Thl-type cytokines and increases in Th2-type cytokines. Among adults without occupational
Pb exposures in Incheon, Korea with blood Pb levels ranging from 0.337 to 10.47 |_ig/dL. Kim et al.
(2007) found associations of blood Pb with serum levels of TNF-a and IL-6 that were larger among men
with blood Pb levels > 2.51 (ig/dL (median). In models that adjusted for age, sex, BMI, and smoking
status, a 1 (ig/dL increase in blood Pb 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).
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 levels among different occupational groups with different mean blood
Pb levels (Pi Lorenzo et al.. 2007; Valentino et al.. 2007; Yucesov et al.. 1997a). The exception was the
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 effect estimates were not provided, statistically significant associations were
observed between blood Pb and IL-10 and TNF-a, with R2 values of 0.249 and 0.235, respectively.
Exposed workers also had higher levels of most Th2 cytokines (IL-2, IL-6, IL-10, and TNF-a) and lower
levels of the Thl cytokine IFN-y. Levels of IL-2, IL-6, and IL-10 showed an increasing trend 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 j^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). Excluded from
this study were exposed workers from the highest end of the distribution of blood Pb levels in DiLorenzo
et al. (2006). Mean TNF-a levels showed a monotonic increase from the low to high blood Pb group,
which was suggestive of a concentration-dependent relationship. Levels of granulocyte colony-
stimulating factor (G-CSF) did not differ between the intermediate and high blood Pb groups among the
Pb recyclers; however, G-CSF levels were higher in the Pb recyclers than in the unexposed controls.
Furthermore, among all subjects, blood Pb showed a strong, positive correlation with G-CSF. Yucesoy et
al. (1997a) found statistically significant lower serum levels of the Thl cytokines, IL-1(3 and IFN-y, in
workers compared with controls; however levels of the Th2 cytokines, IL-2 and TNF-a levels, were
similar between groups. Overall, 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.
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5.2.6. Endocrine Disruption
5.2.6.1. Hypothalamic-Pituitary-Gonadal Axis
Pb is a potent endocrine disrupting chemical that causes reproductive and developmental effects at
moderate levels of exposure 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. Chronic Pb exposure has been shown to reduce serum levels of follicle-
stimulating hormone (FSH), luteinizing hormone (LH), testosterone, and estradiol (Biswas & Ghosh.
2006; Dearth et al.. 2002; E. F. Krieg. Jr.. 2007; Ron is et al.. 1998; Rubio et al.. 2006; Sokol & Berman.
1991). This is likely through the inhibition of LH secretion and the reduction in the expression of the
steroidogenic acute regulatory protein (StAR) (B. M. Huang et al.. 2002; B. M. Huang & Liu. 2004;
Ron is et al.. 1996; V. Srivastava et al.. 2004). StAR is the rate-limiting step essential in maintaining
gonadotropin-stimulated steroidogenesis, which results in the formation of testosterone and estradiol. Pb
(prenatal exposure resulting in blood Pb level 3 (ig/dL) decreases basal StAR synthesis, but not
gonadotropin-stimulated StAR synthesis, suggesting that Pb may not directly affect ovarian
responsiveness to gonadotropin stimulation (V. Srivastava et al.. 2004). Instead, Pb may act on the
hypothalamic-pituitary level to alter LH secretion, which is necessary to drive StAR production and
subsequent sex hormone synthesis. Release of LH and FSH from the pituitary is controlled by
gonadotropin-releasing hormone (GnRH). Pb exposure (10 (.iM, 90 min) in rat brain median eminence
cells can block GnRH release (Bratton et al.. 1994). Pb may also interfere with release of pituitary
hormones through interference with cation-dependent secondary messenger systems, 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) is able to decrease the number of estrogen receptors found in
the uterus and receptor binding affinity (Wiebe & 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 (P. C. Hsu. Liu, et al. 1998; Madhavi et
al.. 2007; Rubio et al.. 2006; Salawu et al.. 2009; Shan et al.. 2009; C. H. Wang et al.. 2006). Direct
generation of ROS in epididymal spermatozoa was observed after Pb exposure (i.p. 20 or 50 mg/kg, 6
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wk) (P. C. Hsu. Hsu, et al.. 1998). In addition, lipid peroxidation in the seminal plasma was significantly
increased in a group of Pb-exposed workers with blood Pb >40 (ig/dL (A. 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). Pb exposed humans (>4 j^ig/dL). animals (14 j^ig/dL). and gonadal cells (0.05 mg/mL)
show a decrease in plasma IGF-1 (Dearth et al.. 2002; Huseman et al.. 1992; kolesarova et al.. 2010; Pine
et al.. 2006). which may be the result of decreased translation or secretion of IGF-1 (Dearth et al.. 2002V
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 etal.. 1993; Huseman et al.. 1987; Huseman et al.. 1992). 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 52
pM, 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 have shown a negative
association between elevated blood Pb and thyroxine (T4) and free T4 levels without alteration in
triiodothyronine (T3), suggesting that long-term Pb exposure may depress thyroid function (Dundar et al..
2006; Robins et al.. 1983; Tuppurainen et al. 1988). However, animal studies on thyroid hormones have
shown mixed results. Pb exposed cows were reported to have an increase in plasma T3 and T4 levels
(Swarup et al.. 2007). whereas mice and chickens manifested decreased serum T3 concentrations after Pb
exposure accompanied by increased lipid peroxidation (Chaurasia et al.. 1998; Chaurasia & 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 (Chaurasia & 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 subjects
(Abdelouahab et al.. 2008; Gustafson et al.. 1989; Lopez et al. 2000; B. Singh et al.. 2000). which would
result in increased T4 levels. Overall, results on the effects of Pb on the HPT axis are inconclusive.
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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 whether observed genotoxic effects are due to chromate or Pb in studies using this form of
inorganic Pb, only studies utilizing other forms of inorganic Pb (e.g., Pb nitrate, acetate) are discussed
below.
5.2.7.1. DNA Damage
A number of studies in human populations have observed positive associations between Pb
exposure and DNA damage, as measured as DNA strand breaks. Most of these associations have been
observed in occupationally-cxposed populations (Danadevi et al.. 2003; de Restrepo et al.. 2000; Fracasso
et al.. 2002; G rover et al.. 2010; Hengstler et al.. 2003; Minozzo et al.. 2010; Pal us et al.. 2003; Shaik &
Jamil. 2009). It is important to note that occupationally-exposed workers have very high blood Pb levels,
and in one study (de Restrepo et al.. 2000) the association between blood Pb and DNA damage was only
observed in workers with blood Pb 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 Paulus
et al. ("2003) and Minozzo et al. (2010) observed no correlation (mean blood Pb = 50.4 (range = 28.2-
65.5) and 59.43 ± 28.34 (ig/dL, respectively). Lastly, workers occupationally-exposed to Pb are also
potentially exposed to other genotoxic materials, making it difficult to rule out confounding co-exposures.
For example, Hengstler et al. (2003) examined workers exposed to Pb, cadmium, and cobalt and observed
that neither blood or air Pb (4.4 (IQR: 2.84-13.6) (ig/dL; 3.0 (IQR: 1.6-50.0) |_ig/m3) was 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. 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 blood Pb levels of 19.5 (11.3-49.2) or 28.6 (11.4-47.5) (ig/dL, compared to blood Pb level
of 4.6 (0.1-8.7) (ig/dL for children living distant to the smelter. DNA damage was significantly increased
in children living nearest the smelter, compared to the intermediate distant children, but was not
significantly different from children living farthest away from the smelter. Multivariate analysis (which
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considered the children's urinary As levels, highest in children farthest from the smelter), revealed no
significant associations between DNA damage and blood Pb. Further, DNA repair ability was also
observed to be nonrelated to blood Pb. Alternatively, Yanez et al. ("2003) observed that children living
close to a mining complex (blood Pb level= 11.6, range = 3.0 to 19.5 j^ig/dL) did have increased levels of
DNA damage compared to control children that lived further away from the mining facility (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 dose
(K. D. Devi et al.. 2000). The three highest doses had responses which were similar in magnitude and
were actually lower than the responses to lower doses tested. In mice exposed to Pb (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, and 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
dose-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 level = 19.5 and 21.9 (ig/dL,
respectively) compared to control rats (Nava-Hemandez et al.. 2009). The level of DNA damage was not
dose-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, 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 HeLa cells when exposed to 500 |_iM for 20-25 hours or 100
(.iM for 0.5-4 hours (Hartwig et al.. 1990; R. D. Snvder & Lachmann. 1989). Pb nitrate, administered to
lymphoma cells at 1-10 mM for 6 hours, did not result in any DNA-protein crosslinks (M. Costa et al..
1996). Pb acetate was observed by Wozniack 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 (.iM. DNA-protein crosslinks were only observed in the 100 (.iM exposure group,
suggesting that the decreased strand breaks observed in the high dose group may be a result of increased
crosslinking in this group. Shaik et al. (2006) also observed DNA damage in human lymphocytes exposed
to 2.1-3.3 mM Pb nitrate for 2 hours. Although there was a dose-dependent increase in DNA damage from
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2.1 to 3.3 mM, no statistics were reported and no unexposed control group was included making it
difficult to interpret these results. Gastaldo et al. (2009) observed that exposure of human endothelial cells
to 1-1000 |iM Pb nitrate for 24 hours resulted in a dose-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 mM,
respectively) did not induce single or double DNA strand breaks or DNA-protein or DNA-DNA
crosslinks in CHV79 cells exposed to Pb acetate. However, both Xu et al. (2006) and Kermani et al.
(2008) reported Pb acetate induced DNA damage in PC 12 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 dose-dependent manner in hepatoma
cells exposed to 50-5000 (.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 HeLa 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 (jM for 24 hours.
Exposure to Pb inhibited nonhomologous end joining (NHEJ) repair, but increased two other repair
pathways, MRE11-dependent and Rad51-related repair (Gastaldo et al.. 2007). Interestingly, 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 (J. P. Li et al.. 2008). though this result is difficult to interpret. Roy and
Rossman (1992) observed an increase in UV-induced mutagenicity when CHV79 cells were co-exposed
to 400 |iM Pb acetate (a nonmutagenic dose 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 dose-dependent
accumulation of apurinic/apyrimidinic site incision activity, indicating that DNA repair was adversely
affected (McNeill et al.. 2007).
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 oxxupational Pb exposures or a number of other heavy metals and
organic contaminants. Blood Pb (range 78.2-251.0 nM) was statistically significantly positively
associated with HPRT mutation frequency in the total population. Also, women with high blood Pb (i.e.,
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greater than the population median, not reported) demonstrated a greater mutation frequency compared to
women with lower blood Pb.
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 500-2000 (.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) (J. P. Li et al.. 2008;
C. Y. Wang et al.. 2008). 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 the findings from human cell line studies, although differences in mutagenicity 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 (Hart wig et al.. 1990). but elicited a
mutagenic response in CHV79 cells (gpt assay) exposed to 1700 (.iM for 5 days (N. K. Rov & 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 dose. Pb nitrate was
alternatively found to be nonmutagenic in CHV79 cells (gpt assay) exposed to 0.5-2000 (.iM for 5 days
(N. K. Rov & Rossman. 1992). or mutagenic in the same cell line (HPRT assay) exposed to 50-5000 (.iM
for 5 days (Zelikoff et al.. 1988). However, mutagenicity was only observed at 500 (.iM. and was higher
than that observed at higher doses. Pb sulfate was also observed to be mutagenic in CHV79 cells (HPRT
assay) exposed to 100-1000 (.iM for 24 hours, but as with Pb nitrate, it was not dose-dependent (Zelikoff
etal.. 1988). Pb chloride was the only Pb compound tested in animal cell lines that was consistently
mutagenic: three studies from the same laboratory observed dose-dependent mutagenicity in the gpt assay
in Chinese hamster ovary cells exposed to 0.1-1 (.iM Pb chloride for one hour (Ariza et al.. 1998; Ariza &
Williams. 1996. 1999).
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
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(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 (mean blood Pb = 10.48 - 86.9 (ig/dL) and sister chromatid
exchange (SCE) was observed in a number of occupational studies (Anwar & Kamal. 1988; Bilban. 1998;
Diivdu et al.. 2005; Diivdu et al.. 2001; X. P. Huang et al.. 1988; Pal us et al.. 2003; Pinto et al.. 2000;
Wiwanitkit et al.. 2008). However, there are numerous methodological issues that limit firm conclusions
from being drawn, most notably that occupational co-exposures to other genotoxic materials were
possible, although some studies excluded workers with exposures to known mutagens (X. P. Huang et al..
1988; Pinto et al.. 2000). In most studies that attempted to investigate the dose-response relationship in
workers, no association was observed between increasing blood Pb levels and the incidence of SCE
(Diivdu et al.. 2001; Pal us et al. 2003; 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), although
the association was only statistically significant 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 incidence of SCE (T. Rajah & Ahuia. 1995; T. T.
Rajah & Ahuia. 1996). However, these two studies may have suffered from limited statistical power to
observe an effect as they only included very small Pb exposed populations. Mielzynska et al. (2006)
investigated the incidence of SCE in children exposed to Pb and PAHs in Poland. Children had an average
blood Pb concentration of 7.69 (ig/dL and 7.87 SCEs/cell. Male children had higher blood Pb
concentrations compared to females, but lower numbers of SCEs. No control population was included in
this study, and thus interpretation of these findings is difficult.
Pb exposure has been observed to induce SCE in multiple laboratory animal studies. In mice
exposed to up to 100 mg/kg Pb acetate intraperitoneally, Pb induced SCE at 50 and 100 mg/kg (Fahmv.
1999). Pb nitrate, also administered i.p., induced the formation of SCE in a dose-dependent manner (10-
40 mg/kg) in the bone marrow of exposed mice (Dhir et al.. 1993). Nayak et al. (1989) exposed pregnant
mice to 100-200 mg/kg Pb nitrate via i.v. injection and observed an increase in SCE 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 (1/10th the LD50) via intraperitoneal injection on alternating days for 11 or 21 days, for a total of 5
or 10 exposures. Induction of SCE in the bone marrow of exposed rats was increased over controls in a
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 exposures encountered by the general population
(e.g., drinking water exposure).
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Only two studies were found that investigated SCE formation in human cell lines due to Pb
exposure. Statistically significant, dose-dependent increases in SCE were observed in human lymphocytes
obtained from a single donor when incubated with 1, 5, 10, or 50 (.iM Pb nitrate (Ustundag & Duvdii.
2007). Melatonin and N-acetylcysteine were reported to ameliorate these effects, indicating Pb may
induce SCE 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.).
Studies investigating SCE in rodent cells were more equivocal than those in human cells. Pb
sulfate, acetate, and nitrate were found to not 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 ability to detect Pb-induced SCE. 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, dose-
dependent increase in SCE in Chinese hamster ovary cells. Lin et al. (1994) also observed a dose-
dependent increase in SCE in Chinese hamster cells exposed to 3-30 (.iM Pb nitrate for 2 hours.
Micronucleus Formation
Micronucleus formation in Pb-exposed workers was investigated in numerous occupational studies
(Bilban. 1998; Grover et al.. 2010; M. I. Khan et al.. 2010; Minozzo et al.. 2004; Minozzo et al.. 2010;
Palus et al.. 2003; Pinto et al.. 2000; Shaik & Jamil. 2009; Vaglenov et al.. 1998; Vaglenov et al.. 2001).
The workers in the occupational studies generally had high blood Pb levels (>20 (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 concentration of only 10.48 (ig/dL. Studies that
analyzed workers according to the magnitude of their blood Pb levels reported no correlation between Pb
exposure and the observation of micronuclei (Minozzo et al.. 2004; Minozzo et al.. 2010; Pal us et al..
2003; Pinto et al.. 2000). although Pinto et al. (2000). Grover et al. (2010). and Minozzo et al. (2010) did
report an association between micronuclei formation and duration of exposure. Only one study was found
that investigated micronucleus formation in a nonworker population; Mielzynska et al. (2006) examined
associations of blood Pb and urinary PAH metabolite levels with micronuclei formation in children in
Poland. Children, with an average blood Pb concentration of 7.69 (ig/dL, were observed to have 4.44
micronucleated cells per 1000 cells analyzed. Although no control group was included in this study, a
statistically significant positive correlation was observed between blood Pb concentrations and
micronuclei frequency, and children with blood Pb greater than 10 (ig/dL had significantly more
micronucleated cells than children with blood Pb 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
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polychromatic erythrocytes (PCEs) compared to controls. Similarly, Alghazal et al. (2008) exposed rats to
100 mg/L 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 atotal of 5 or 10 exposures. Formation of micronuclei
in the bone marrow of exposed rats was increased over 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 dose-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 dose-dependent. Mice exposed to 1 g/L Pb
acetate via a more environmentally relevant route of exposure, drinking water, for 90 days had
statistically significant increases in micronucleated PCEs (C. C. Marques et al.. 2006).
Three studies were found that reported increased micronucleus formation in human cell lines
treated with Pb. Dose-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 (Turkez et al; Ustundag &
Duvdii. 2007). Gastaldo et al. (2007) also observed a dose-dependent increase in micronuclei in human
endothelial cells exposed to 1-1000 (.iM Pb nitrate for 24 hours. Only two animal cell culture studies
investigating micronuclei formation were found. One study observed that micronuclei were not induced in
Chinese hamster cells exposed to 3-30 (.iM Pb nitrate for 2 hours (R. H. Lin et al.. 1994). whereas the
other observed that Pb acetate induced a dose-dependent increase in Chinese hamster cells when
administered at 0.05-10 (.iM for 18 hours (Bonacker et al.. 2005).
Chromosomal Aberrations
Chromosomal aberrations (e.g., chromosome breaks, nucleoplasmic bridges, di- and acentric
chromosomes, and rings) were examined in a number of occupational studies (Bilban. 1998; De et al..
1995; Grover et al.. 2010; X. P. Huang et al.. 1988; Pinto et al.. 2000; Shaik & Jamil. 2009).
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 are difficult to interpret in
the context of the general population, also pertain to the present findings. No correlation was observed
between increasing blood Pb levels and incidence of chromosomal aberrations, although an association
was observed between duration of exposure and chromosomal damage (Grover et al.. 2010; Pinto et al..
2000). Two studies reported no association between occupational exposure to Pb and chromosomal
aberrations (Andreae. 1983; Anwar &Kamal. 1988). Smejkalova (1990) observed chromosomal damage
and aberrations in children living in a heavily Pb-contaminated area of Czechoslovakia and an area with
less contamination, although the difference between the two was not statistically significant. Although
blood Pb levels were statistically significantly higher in the Pb-contaminated children, they were
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generally comparable (low 30s versus high 20s (ig/dL, respectively), indicating there may not be enough
of an exposure difference to detect a significant difference in aberration rates.
The majority of animal studies investigating Pb-induced genotoxicity focused on the ability of Pb
to produce chromosomal damage. Fahmy (1999) exposed mice to 25-400 mg/kg Pb acetate i.p., either as a
single dose or repeatedly for 3, 5, or 7 days. Chromosomal damage was observed to increase in bone
marrow cells (100-400 mg/kg) and spermatocytes (50-400 mg/kg) in a dose-dependent manner after both
exposure regimens. Pb nitrate was also observed to produce dose-dependent chromosomal damage in
mice exposed i.p. to a single dose of 5, 10, or 20 mg/kg (Dhir. Sharma. etal.. 1992). 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 dose-response as the response was similar in all doses
tested. Nayak et al. (1989) 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 dose. 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 dose-response reported and few cells were analyzed (Navak. Rav. & Persaud. 1989). 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.. 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 doses. Dhir et al. (1992) also observed a dose-dependent increase in chromosomal
damage in mice exposed via gavage, albeit at much lower doses: 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.
Only three studies were found that investigated the ability of Pb to induce chromosomal damage in
human cell lines and all three reported negative findings. Wise et al. (2004; 2005) observed that Pb
glutamate was not mutagenic in human lung cells exposed to 250-2,000 (.iM for 24 hours. Shaik et al.
(2006) observed that Pb nitrate did not increase chromosomal aberrations in primary lymphocytes
(obtained from healthy volunteers) when incubated with 1.2 or 2 mM for 2 hours. Four 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
500-2000 (.iM for 24 hours (J. P. Wise. Sr. et al.. 1994). 3-30 (.iM for 2 hours (R. H. Lin et al.. 1994). or
0.05-1 (.iM for 3-12 hours (Cai & Arenaz. 1998). Wise et al. (1994) did observe increased chromosomal
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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 etal.. 1991V The observed hypomethylation in the liver was associated with an increase
in cell proliferation. Two additional studies in humans observed that DNA methylation patterns in adults
and cord blood were inversely correlated with bone Pb levels (Pilsner et al.. 2009; R. Q. Wright et al..
2010). Changes in DNA methylation patterns could potentially lead to dysregulation of gene expression
and altered tissue differentiation.
Mitogenesis
Only a few studies have investigated the potential epigenetic effects of Pb exposure in human
populations. One such epigenetic effect investigated was mitogenesis that induces cells to proliferate
when they should not. Three studies (Minozzo et al.. 2004; Minozzo et al.. 2010; T. Rajah & Ahuja. 1995)
were found that reported that Pb reduced mitogenesis in Pb-exposed workers (blood Pb = 35.4 (ig/dL,
59.4 (ig/dL, and not reported, respectively). The observation of decreased cell division in exposed
workers may indicate that cells suffered DNA damage and died during division, or that division was
delayed to allow for DNA repair to occur. It is also possible that Pb exerts an aneugenic effect and arrests
the cell cycle.
Many studies have investigated the ability of Pb to induce mitogenesis in animal models, and have
consistently shown that Pb nitrate can stimulate DNA synthesis and cell proliferation in the liver of
animals exposed to 100 (.iM/kg via i.v. injection (Columbano et al.. 1987; Columbano et al.. 1990; Coni et
al.. 1992; Lcdda-Columbano et al.. 1992; Nakaiima et al.. 1995). 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 (Y.-J. Cheng et al.. 2002; Menegazzi et al..
1997). Only one study was found that observed a mitogenic effect after inhalation: exposure to 0.01M Pb
acetate for 4 weeks resulted in increased cellular proliferation in the lungs (Fortoul et al.. 2005).
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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) (Heiman & Tonner. 1995) and primary oligodendrocyte
progenitor cells (1 (.iM for 24 hours) (W. Deng & Poretz. 2002). and had no observable effects on growth
in glioma cells (0.01-10 (.iM for 12-72 hours) (M. Y. 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) (S. S. Wise et
al.. 2005V 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 dose-dependent
inhibition of DNA synthesis reported in hepatoma cells (1-100 (.iM for 72 hours) (Heiman & Tonncr.
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 |_iM -1 mM for 1-7 days and 0.078-320 (.iM for 48 hours, respectively) (Apostoli et
al.. 2000; lavicoli et al.. 2001). 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 0-100 (j,M for 48 hours (Kermani et al.. 2008). Pb
nitrate was alternatively reported to increase (R. H. Lin et al.. 1994) and decrease (Cai & Arenaz. 1998)
the mitotic index in Chinese hamster ovary cells exposed to 1 |iM 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 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 glutathione transferase P (GST-P) with both Pb compounds. The
induction of GST-P by Pb was observed to operate 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 to not inhibit CYP1A2
by direct enzyme inhibition, but rather to decrease the amount of CyplA2 mRNA. In contrast, Korashy
and El Kadi (2004) observed that exposure of murine hepatoma cells to 10-100 |iM Pb nitrate for 24
hours increased the amount of CyplAl mRNA while not influencing the activity of the enzyme.
NAD(P)H:quinone oxidoreductase and glutathione S-transferase Ya activities and mRNA levels were
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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 metabolism, the pentose phosphate pathway, and amino acid metabolism (Glahn et
al.. 2008).
One additional animal study provides 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 PND1-20, the expression of proteolipid protein (PLP), 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 (MBP), 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.
5.2.7.6. Apoptosis
Occupational exposure to Pb and induction of apoptosis was investigated in three studies. One
study directly reported that exposure to Pb increased apoptosis compared to nonexposed controls
(Minozzo et al.. 2010). whereas the other two reported that two early indicators of apoptosis, karyorrhexis
and karyolysis, were elevated in exposed workers (Grover et al.. 2010; M. I. Khan et al.. 2010). Pb nitrate
was also observed to induce apoptosis in the liver of exposed animals (Columbano et al.. 1996; Nakaiima
etal.. 1995). Apoptosis was observed in rat fibroblasts exposed to Pb acetate and rat alveolar
macrophages exposed to Pb nitrate (lavicoli et al.. 2001; Shabani & Rabbani. 2000). Observation of Pb-
induced apoptosis may represent the dysregulation of genetically-controlled cell processes and tissue
homeostasis.
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). Evidence for the majority of these modes of action is observed at
low blood Pb level in humans and animals, between 2 and 5 (ig/dL, and at doses as low as the picomolar
range in animals and cells. Dose captures Pb exposure concentration in invitro systems or in animal
studies when no blood Pb level was reported. These observable effect levels are limited by the data and
methods available 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.
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Table 5-2. Related health effects resulting from the MOAs of Pb and the lowest level
eliciting the MOA reported as blood Pb level and dose delivered
MOA
Related Health Effects
(Section) ISA
Lowest Level at which
MOA Observed
Studies
Blood Pb
Dose
Altered Ion Status
All Heath Effects of Pb
3.5 |jg/dL
50 pM
Huel et al. (2008) Kern et al. (2000)
Protein Binding
Renal (5.5), Effects on Heme Synthesis and Red
Blood Cell Funtion (5.7)
6.4 |jg/dL
50 |jM
Chen et al. (2010) Klann and Shelton
(1989)
Oxidative Stress
All Heath Effects of Pb
5-10 |jg/dL
10 |JM
Quinlan et al. (1988) Ahamed et al.
(2006) Yiin and Lin (1995)
Inflammation
Neurological (5.3), Cardiovascular (5.4), Renal
(5.5), Immune (5.6), Respiratory (5.6.4), Cancer
(5.10), Hepatic (5.9.1)
3 |jg/dL
0.01 [)M
Kim et al. <2007) Chettv et al. (2005)
Endocrine
Disruption
Reproductive Effects and Birth Outcomes (5.8),
Bone and Teeth (5.9.4), Endocrine System (5.9.3)
2 |jg/dL
20 ppm
Krieg (2007) Wiebe and Barr (1988)
Cell Cancer (5.10), Reproductive Effects and Birth
Death/Genotoxicity Outcomes (5.8), Bone and Teeth (5.9.4)
3.1 |jg/dL
50 nM
Van et al. (2004) Bonacker et al. (2005)
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 | have been shown to increase 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. Disruption
of ion transport not only leads to altered Ca2+ homeostasis, it can also result in perturbed neurotransmitter
function. Pb-exposure decreases evoked release of neurotransmitters, while simultaneously increasing
spontaneous release of neurotransmitters through Ca2+ mimicry. Pb is 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 pertubation 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
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binding protein in blood is ALAD with carriers of the ALAD-2 allele potentially exhibiting higher Pb
binding affinity. Additionally, metallothionein is a 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 5-ALA, NAD(P)H oxidase
activation, membrane and lipid peroxidation, and antioxidant enzyme depletion. Through the inhibition of
5-ALAD due to displacement of Zn, 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 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
production of related cytokines.
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 are
mixed and require further investigation.
Genotoxicity and cell death has been investigated after Pb exposure in humans, animals, 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 evidence of mutagenesis and clastogenicity in highly exposed humans, however weak
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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. 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. 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.
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. Neurological Effects
5.3.1. Introduction
The central nervous system undergoes rapid differentiation in utero and in early life, which makes
the developing fetus and infant particularly susceptible to neurological effects associated with
environmental exposures (Landrigan et al.. 1999; Rice & Barone. 2000). Based on evidence from diverse
prospective and cross-sectional studies focusing primarily on effects associated with Pb exposure levels
less than 10 (ig/dL, 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. 2006). Several individual studies observed inverse associations
between blood Pb levels and IQ that persisted at blood Pb levels in the range of 2-8 (ig/dL (Lanphear et
al.. 2000; J. Schwartz. 1994). 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). The 2006 AQCD described associations of blood Pb with a broad range of additional, related
neurodevelopmental endpoints, including academic achievement and performance, motor skills, mood,
and antisocial and delinquent behavior (U.S. EPA. 2006).
Toxicological studies not only provided coherence with similarly consistent findings for Pb-
induced impairments in learning, behavior and attention, and sensory acuities, but also provided
biological plausibility by characterizing mechanisms for Pb-induced neurological 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
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levels of children being associated with deficits in multiple functional domains such as cognitive function,
motor function, memory, mood, and behavior. Additional biological plausibility was provided by
observations of associations of childhood blood Pb levels with decreased neuronal density and neuronal
loss measured in adulthood, as assessed by magnetic resonance imaging techniques (Cecil et al.. 2005;
Meng et al.. 2005; Trope et al.. 2001; Trope et al.. 1998; Yuan et al.. 2006).
A key finding across several epidemiologic studies of children was a larger estimated effect for a
given incremental increase in blood Pb levels on neurocognitive deficits in children with lower blood Pb
levels compared with children with higher blood-Pb levels (Bellinger & Needleman. 2003; Can field.
Henderson, et al.. 2003; kordas et al.. 2006; Lanphear et al.. 2005; Rothenberg & Rothenberg. 2005;
Tellez-Roio et al. 2006). Among these studies were two analyses of the pooled cohort data, both of which
demonstrated a supralinear relationship between blood Pb level and IQ with a steeper negative slope
observed at blood Pb levels <10 (.ig/dL (Lanphear et al.. 2005; Rothenberg & Rothenberg. 2005).
Consistent with epidemiologic findings, toxicological studies also observed nonlinear Pb exposure-
response relationships for outcomes such as behavioral responses, neuronal activation, and dopamine
release.
Another area of focus included the comparison of various lifestages of Pb exposure in terms of risk
of neurodevelopmental deficits. Toxicological studies demonstrated that in utero with or without early
postnatal exposure to Pb was the most sensitive window for Pb-dependent neurological effects.
Epidemiologic studies observed neurocognitive deficits in association with prenatal, peak childhood,
cumulative childhood, and concurrent blood Pb levels. Among studies of children that examined multiple
lifestages of exposure, several found that concurrent blood Pb was associated with the largest decrement
in IQ (Baghurst et al.. 1992; Cnn field. Henderson, et al.. 2003; A. Chen et al.. 2005; Lanphear et al..
2005). with some studies finding that the magnitude of association increased with age (Dietrich. Berger.
Succop. et al.. 1993; Factor-Litvak et al.. 1999; Ris et al.. 2004; G. A. Wasserman et al.. 1994). Several
studies of children compared effect estimates for Pb levels measured in blood, deciduous tooth, or tibial
bone and found that compared with blood Pb levels, tooth or bone Pb levels were associated with an equal
or larger magnitude of neurodevelopmental deficits (Bellinger et al.. 1991; Greene & Emhart. 1993;
Needleman et al. 1979; G. A. Wasserman et al.. 2003). These findings pointed to an effect of cumulative
childhood exposure. A common limitation of epidemiologic studies of children was the high correlation
among Pb exposure metrics at different ages, making it difficult to distinguish among effects of Pb
exposure at different ages (Lanphear et al.. 2005) and to ascertain which developmental periods of Pb
exposure were associated with the greatest risk of neurodevelopmental decrements. The issue of
persistence of the neurological effects of Pb exposure also was considered, with some evidence
suggesting that the associations of blood Pb levels with neurodevelopmental outcomes persisted into
adolescence and young adulthood in the absence of marked reductions in blood Pb level (Needleman et
al.. 1990; Ris et al.. 2004; Tong et al.. 1996).
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In epidemiologic studies of adults, neurological effects (e.g., impaired memory, attention, reaction
time, visuomotor tasks and reasoning, alterations in visual or brainstem evoked potentials, postural sway,
and nerve conduction) were mostly clearly indicated in occupationally-exposed workers in association
with blood Pb levels in the range of 14 to 40 (.ig/dL (Baker et al.. 1979; Cantarow & Trumper. 1944).
Studies of environmentally-exposed adults produced mixed findings; however, bone Pb levels (Weisskopf
et al.. 2004; R. Q. Wright et al.. 2003) were associated with decrements in neurocognitive function more
so than were blood Pb levels (E. F. Krieg. Jr et al.. 2005; Nordberg et al.. 2000). These findings suggested
that cumulative Pb exposures, including higher past exposures, may be important contributors to
neurological effects in adults.In addition to cognitive function, blood and bone Pb levels also were
associated with greater self-reported anxiety and depression scores among environmentally-exposed
adults (Rhodes et al.. 2003). These findings in adults were strongly supported by parallel findings in
female and male rodents for Pb-induced depression and emotional changes, respectively. Studies of adults
also reported associations of blood Pb with risk of amyotrophic lateral sclerosis (ALS) and essential
tremor, although the body of literature was smaller and evidence was less consistent than that for
cognitive function. Whereas toxicological studies demonstrated Pb-induced neurodegeneration and
formation of neurofibrilary tangles commonly associated with Alzheimer's disease pathophysiology,
blood Pb level generally was not associated with Alzheimer's disease in adults.
As discussed throughout this section, recent epidemiologic and toxicological studies continue to
demonstrate associations between exposure to or biomarkers of Pb and neurological effects and expand
upon the previous body of evidence by demonstrating similar effects (e.g., cognitive function,
impairments in behavior) at lower blood Pb levels (1-5 |_ig/dL). Whereas previous evidence was
inconsistent, new studies in children report positive associations between blood Pb levels and attention
deficit hyperactivity disorder (ADHD). New toxicological studies expand evidence for prenatal and
postnatal Pb exposure effects on learning, memory, and attention and provide insight into the contribution
of stress to this paradigm. New or expanded areas of toxicological research related to Pb exposure include
mood disorders, neurofibrilary tangle formation, and adult dementia after early life Pb exposures.
Historically important areas of toxicological research are further expanded with recent publications of Pb-
dependent 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
neurological and neurobehavioral outcomes associated with Pb exposure.
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5.3.2. Neurocognitive Function and Learning
5.3.2.1. Epidemiologic Studies of Cognitive Function in Children
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 exposure and outcome,
examining persistence of neurocognitive deficits to older ages, and comparing risk estimates among blood
Pb levels 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 blood Pb levels, race/ethnicity, and SES. Individual cohort studies in diverse populations were
consistent in demonstrating that blood Pb levels in the range of 5 to 10 (ig/dL (Bellinger et al.. 1987;
Bellinger & Needleman. 2003; Bellinger et al.. 1991; Can field. Henderson, et al.. 2003; J. Schwartz.
1994; Tellez-Roio et al.. 2006), were associated with losses in full-scale IQ (FSIQ) pointsin children.
These findings were substantiated in a pooled analysis of seven prospective studies (Boston, MA;
Cincinnati, OH; Rochester, NY; Cleveland, OH; Mexico City, Mexico; Port Pirie, Australia; and Kosovo,
Yugoslavia) by Lanphear et al. (2005) as well as multiple meta-analyses that included both cross-sectional
and prospective studies (Needleman & Gatsonis. 1990; Pocock et al.. 1994; J. Schwartz. 1994) (Figure 5-
2 and Table 5-3).
The analysis pooling data from seven prospective studies included 1333 children with mean
(5th-95th percentile) blood Pb levels of 12.4 (ig/dL (4.1-34.8 (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, birth weight, and maternal education, IQ measured at schoolage (mean
6.9 years) was inversely associated with concurrent, peak, average lifetime, and early childhood blood Pb
levels, with the largest decrement in IQ estimated for concurrent blood Pb levels (-0.23 points [95% CI: -
0.32, -0.14] at a blood Pb level of 1 (ig/dL).
In Lanphear et al. (2005). various models were investigated to characterize the shape of the blood
Pb dose-response relationship. Consistent with findings from individual studies, Lanphear et al. (2005)
found that a log-linear model best fit the data, with a greater decrease in IQ estimated for an increase in
concurrent blood Pb from <1-10 (ig/dL (6.2 points [95% CI: 3.8, 8.6]) than an increase from 10 to 20
(ig/dL (1.9 points [95% CI: 1.2, 2.6]). Sensitivity analyses, in which one study was successively excluded,
revealed that no single study was responsible for driving the results. Although HOME score was not
available in the Rochester study, exclusion of that cohort's data resulted in a less negative effect estimate.
However, the change was less than 3%.
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Studies published since the 2006 Pb AQCD continue to demonstrate associations between
increasing blood Pb level and decrements in FSIQ (Figure 5-2 and Table 5-3). Whereas most studies
demonstrated decrements in FSIQ in association with blood Pb levels ranging from 5 to 10 (ig/dL, Kim et
al. (2009) was particularly noteworthy for demonstrating an association in a population with lower blood
Pb levels (mean 1.73 |_ig/dL. range 0.42-4.91 |_ig/dL). Children ages 8 to 11 years of age in Korea were
tested using the Korean Educational Development Institute-WISC, which assesses vocabulary, arithmetic,
picture arrangement, and block design. In a 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 increase in blood Pb level was associated with a -1.03
point decrease (95% CI: -1.71, -0.36) in FSIQ. Increasing blood Pb levels also was associated in
performance and verbal IQ (PIQ and VIQ, respectively). Although several important confounders were
considered, there was no direct assessment of the home environment and the primary caregiver's IQ in
this study which are notable limitations.
Kim et al. (2009) also examined effect modification of the blood Pb-FSIQ relationship by blood
manganese (Mn) levels. The mean (range) blood Mn level was 14.3 (ig/dL (5.3-29.02), respectively.
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 j^ig/dL).
Multivariate linear regression models predicting FSIQ, VIQ, and PIQ used concurrently measured blood
Pb level as the predictor variable in the low and high Mn groups. The associations for blood Pb level with
FSIQ and VIQ were larger in magnitude for children in the high Mn group (e.g., -5.3 FSIQ points [95%
CI: -10.1, -5.3] per 1 (ig/dL increase in blood Pb level), compared with children in the low Mn group
(e.g., -4.0 FSIQ points [95% CI: -9.9, 1.80] per 1 (ig/dL increase in blood Pb level) (Figure 5-3).
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Reference
Exposure
Period
Mean (SD)
Blood Pb
(MQ/dL)
FSIQ Age
(Years)
1.73 (0.80)
8-11
4
4.7 (2.3)
6
*
5.4 (3.3)
7.5
«
6.5 (4.9)
10
•
7.0
4
~
9
*
*
11
*
7.4 (4.3)
3 or 5
•
7.8
6-10
NR
6.5
NR
15-17
Kim et al. (2009)
Walkowiak et al. (1998)
Chiodo et al. (2004)
Bellinger et al. (1992)
Min et al. (2009)
Concurrent
Concurrent
Concurrent
Early childhood
Concurrent
Canfield et al. (2003)
Schnaas et al. (2006)
Dietrich et al. (1993)
Ris et al. (2004)
Baghurst et al. (1992)
Tong et al. (1996)
Factor-Litvak et al. (1999)
Wasserman et al. (2003)
Surkan et al. (2007)
Cumulative
Prenatal
Concurrent
Early childhood
Early childhood
Cumulative
Cumulative
Cumulative
Concurrent
21.7 (25 -50 ) 7-8
17.8 (5.8)
30 and 8a
31.6 and 6.3a
5-10 vs. 1-2b
11-3
7
10-12
6-10
Pooled/Meta-Analyses
Lanphearet al. (2005)
Schwartz et al. (1994)
Pocock et al. (1994)
Concurrent 6.9(1.2) 4.8-10
Early childhood 6.5-23 School age
Early childhood 6.8-21.2 5-14
Concurrent 7.4-23.7 5-14
Change in FSIQ (95CI)
Note: In general, studies are presented in ascending order of mean blood Pb level, followed by a
study analyzing blood Pb level as a categorical variable and then by pooled/meta-analyses. Effect
estimates are standardized to a 1 [jg/dL increase in blood Pb. The various tests used to measure
IQ are scored on a similar scale (approximately 40-160). aEffect estimate represents the loss in
FSIQ points in children with blood Pb levels 5-10 [jg/dL, with children with blood Pb levels 1-2
[jg/dL serving as the reference group. bThese values represent the mean blood Pb levels in the two
groups from different cities.
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 presented in
Figure 5-2
Study
Population/
Location
Blood Pb
Levels (|jg/dL)
Statistical Analysis
FSIQ Assessment
Effect
Estimate
(95% Cl)a
Kimetal. 279 children in Seoul, Concurrent mean
(2009) Seongnam, Ulsan, and (SD): 1.73 (0.80)
Yeoncheon, Korea, ages
8-11 yr in April-December
2007
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
-1.03 (-1.71, -
0.36)
Walkowiak et 384 children in East
a I. (1998) Germany, age 6 yr in
1994
Concurrent (age 6
yr) mean (SD): 4.7
(2.3)
Log linear regression model
adjusted for city, visual acuity,
contrast sensitivity, parental
education, sex, breastfeeding,
height, nationality
WISC verbal and block
design summed index
-0.51 (-1.03, 0.01)
Chiodo et al. 246 African-American
(2004) children Detroit, Ml
Concurrent (age 7.5
yr) mean (SD): 5.4
(3.3)
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
-0.20 (-0.35, ¦
0.05)b
Bellinger et
al. (1992)
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)
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
-0.58 (-0.99, -
0.18)
Min et al.
267 primarily African-
Concurrent mean
Linear regression model
WISC-R at age 4 yr
-0.50 (-0.89, -
(2009)
American children in the
(range): 7.0 (1.3-
adjusted for HOME score,
WISC-R at age 9 yr
0.11)
Cleveland, OH area
23.8)
cargiver's vocabulary test, sex,
WISC-R at age 11 yr
-0.41 (-0.78, -
followed from birth
parity, maternal marital status,
0.04)
(1994-1996) to age 11 yr.
head circumference at birth
-0.54 (-0.91, -
Children were exposed
0.17)
prenatally to multiple
drugs.
Canfield et
al. (2003)
172 children born 1994-
1995 in Rochester, NY
followed from infancy to
age 3-5 yr
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 3 or 5
yr.
-1.37 (-2.56, -
0.17)
Schnaas et
al. (2006)
175 children in Mexico
City, Mexico followed
from birth (1987-1992)
followed until age 10-15
yr.
Prenatal (3rd
trimester) geometric
mean (5-95th): 7.8
(2.5-24.5)
Linear mixed effects regression
model adjusted for sex, SES,
maternal IQ, HOME score, birth
weight, postnatal blood Pb,
random slope for subject
WISC-R at ages 6-10 yr
-3.4 (-5.6, -1.3)
Dietrich et al.
(1993)
253 children in
Cincinnati, OH followed
from birth (1979-1985) to
age 20-23 yr
Concurrent
Linear regression model
WISC-R at age 6.5 yr
-0.33 (-0.60, -
0.06)
Ris etal.
(2004)
195 children in
Cincinnati, OH followed
from birth (1979-1985) to
age 20-23 yr
Early childhood (age
6.5 yr)
Linear regression model
adjusted for maternal IQ, sex,
and average total HOME score
WISC-III indices at age 15-
17 yr factored into
Learning/IQ
-0.08 (-0.16,
0.004)
Baghurst et
al. (1992)
494 children in Port Pirie,
Australia followed from
birth (1979-1982) to age
11-13 yr.
Early childhood (avg
of age 0-3 yr) 25-
50th: 17.4, 50-75th:
21.7
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
WISC-Ratage 7-8yr
-3.3 (-6.5, -0.2)
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Study
Population/
Location
Blood Pb
Levels (|jg/dL)
Statistical Analysis
FSIQ Assessment
Effect
Estimate
(95% Cl)a
Tong et al.
375 children in Port Pirie, Cumulative 0-7 yr
Regression model adjusted for WISC-R at age 11 -13 yr
-4.3 (-8.5, -0.14)
(1996)
Australia followed from mean (SD): 17.8
birth (1979-1982) to age (5.8)
sex, age, schoolgrade, parental
occupational prestige, HOME
11-13 yr.
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
Factor-Litvak 577 children in Kosovo,
et al. (1999) Yugoslavia followed from
birth (1985-1986) to age
10-12 yr
Cumulative (4-7 yr) Log linear regression model
mean0: 30 (K. adjusted for HOME score,
Mitrovica, 8 (Pristina) ethnic group, maternal age, birth
weight, maternal Raven's score,
maternal education, birth order,
sibshipsize, sex
WISC-R at age 7 yr
-8.3 (-11.4,-5.05)
Wasserman
etal. (2003)
290 children in Kosovo,
Yugoslavia followed from
birth (1985-1986) to age
10-12 yr
Lifetime avg mean:
31.6 (K. Mitrovica,
6.3 (Pristina)
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
-2.3 (-4.0, -0.58)
Surkan et al.
(2007)
389 children ages 6-10
yr. from Boston, MA and
Farmington, ME
Concurrent mean
(range): 2.2 (1-10)
Linear regression model
adjusted for caregiver IQ, child
age, SES, race, birth weight.
WISC-III at ages 6-10 yr
-6.07 (-10.7, -
1.36), blood Pb =
5-10 vs. 1-2
Pooled/Meta-analyses
Lanphear et
al. (2005)
1333 children pooled
from Boston, Cincinnati,
Cleveland, Mexico City,
Port Pirie, Rochester,
and Yugoslavia cohorts
Concurrent mean
(SD): 6.9 (1.2)
Log linear regression model
adjusted for HOME score, birth
weight, maternal IQ, maternal
education
FSIQ measured at ages 4.8-
10 yr
-0.23 (-0.32, -
0.14)
Schwartz et
al. (1994)
Meta-analysis of 7 Early childhood (2-3
studies with sample sizes yr) range in study
75-579 children means: 6.5-23
Meta-analysis of combining
effect estimates from individual
studies
FSIQ measured at
schoolage
-3.75 (-4.91, -
2.60)
Pocock etal.
(1994)
Meta-analysis of 5
prospective (over 1,100
children and 14 cross-
sectional studies (3,499
children)
Prospective studies
mean at age 2 yr:
6.8-21.2
Cross-sectional
means: 7.4-23.7
Meta-analysis of combining
effect estimates from individual
studies
FSIQ measured at ages 5-14
Prospective studies
Cross-sectional studies
-2.7 (-4.1,-1.2)
-3.7 (-4.8, -2.5)
aEffect estimates are standardized to a 1 |jg/dl_ increase in blood Pb level.
b95% 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.
Quantitative data not presented. Means estimated from a figure.
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High Mn
Jll ¦ "" Ifll lill tllllll r II llll II Mill l 1. — i . L
7"" i '
I llll J1111UL111 Llll - ] I 1
I III! Ill III '» MMj H1HI "If I II >11111
-0.2 0.0 0.2 0.4 0.6
Blood Lead Concentration
-0.2 o.o
Blood Lead Concentration
Source: Used with permission from Elsevier Science, Kim et al. (2009).
Figure 5-3. Effect modification of association between concurrent blood
Pb level and IQ by blood Mn level.High and low Mn refers to
levels above and below the median of 14 |jg/dl_, respectively.
Min and colleagues (2009) examined the relationship between Pb exposure assessed at age 4 years
and children's IQ and academic achievement at 4, 9, and 11 years of age in a sample of 278 urban
children originally enrolled in a prospective study on the effects of prenatal poly-drug exposure
(determined by assay of infant meconium or urine, maternal urine, or maternal self-report). The study
population was primarily African-American (86%) and low SES (98%); 39% of mothers had not finished
high school, and 14% were married at the time of enrollment. The mean blood Pb level at age 4 years was
7.0 (ig/dL (SD 4.1, range 1.3-23.8); 36% had blood Pb levels <5 (ig/dL, and 19% had levels >10 (ig/dL.
The researchers utilized restricted cubic spline functions for blood Pb level to test for a nonlinear
relationship between blood Pb levels and FSIQ. Although the cubic spline term did not attain statistical
significance, analyses suggested a steeper slope at lower Pb levels (up to 7 j^ig/dL) These findings were
consistent with those from the pooled analysis (Lanphear et al.. 2005) and other studies (Tellez-Roio et
al.. 2006).
Also similar to previous studies with repeated assessments of cognitive function overtime, Min et
al. (2009) found that the association between concurrent blood Pb level and FSIQ persisted with
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increasing age. The magnitude of the inverse association between blood Pb level and FSIQ was consistent
at ages 4, 9 and 11 years. A 1 (ig/dL increase in blood Pb level was associated with a loss in IQ points of
0.50, 0.41, and 0.54 at ages 4, 9, and 11 years, respectively (Figure 5-2 and Table 5-3). The findings of
this study also indicated that specific cognitive domains may be more sensitive to Pb exposure at different
stages of development. Non-verbal reasoning decrements assessed using the Wechsler Preschool and
Primary Scales of Intelligence-Revised (WPPSI-R) Performance IQ (PIQ) and Wechsler Intelligence
Scales (WISC)-IV Perceptual Reasoning Index were consistently associated with increasing blood Pb
level even at younger ages while verbal decrements did not become apparent until assessments at 11 years
of age. Lower reading scores were associated with increased Pb exposure at 9 and 11 years while math
scores were not affected until age 11 years. An important consideration that may limit the generalizability
of these findings is the high prevalence of prenatal exposure to cocaine (51% of subjects) and alcohol
(77% of subjects). However, accounting for prenatal drug exposure did not attenuate or modify (i.e., no
interaction effects) the negative associations between blood Pb level and cognitive outcomes.
Surkan et al. (2007) used data originally collected for the New England Children's Amalgam Trial
(NECAT), a study of 6 to 10 year old English-speaking children from urban Boston, Massachusetts and
rural Farmington, Maine designed to assess the effect of amalgam dental fillings on children's
neurodevelopment. At baseline (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. In analyses that excluded 3 children with
blood Pb level >10 (ig/dL, children with blood Pb levels of 5 to 10 (ig/dL had significantly lower WISC-
III Full FSIQ scores (-6.07 points (95% CI: -10.7, -1.36]) compared with children who had levels of 1 to 2
(ig/dL (referent group), adjusting for age, race/ethnicity (Black or Hispanic vs. non-Hispanic white), birth
weight, SES, and primary caregiver IQ.
Specific Indices of Cognitive Function in Children
In addition to FSIQ, an index of global cognitive function, blood Pb levels also were associated
with specific cognitive abilities, including attention, executive function, language, memory and learning,
and visuospatial processing in previous studies of children and adolescents (Bellinger et al.. 1991;
Bellinger & Stiles. 1993; Can field. Kreher. et al.. 2003; Chiodo et al. 2004; Dietrich et al.. 1991; Dietrich
etal.. 1992; kordas et al.. 2006; Lanphear et al.. 2000; Needleman et al. 1979; Ris et al.. 2004; Tellez-
Rojo et al.. 2006). Studies often find associations with several endpoints, and because several tests of
neurocognitive function are interrelated, it is difficult to ascribe the effects of Pb exposure to a specific
domain of neurocognitive function. For example, in U.S. representative analysis of NHANES III (1988-
1994) data, which included 4853 children ages 6-16 years with a geometric mean blood Pb level of 1.9
(ig/dL, Lanphear et al. (2000) found that a 1 (ig/dL increase in blood Pb levels was associated with
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decreases in arithmetic (-0.70 points [95% CI: -1.0, -0.37]), reading (-0.99 points [95% CI: -1.4, -0.62]),
block design (-0.10 points [95% CI: -0.18, -0.02]), and digit span (-0.05 [95% CI: -0.09, -0.01]) subtests.
Finding that blood Pb levels are associated with a spectrum of neurocognitive indices provides biological
plausibility for associations observed between blood Pb levels and IQ. Furthermore, these tests of
attention, learning, and memory in humans have parallel tests in animals, and compared with evidence for
IQ, evidence for these specific tests may improve understanding of the coherence between findings in
humans and animals (Rice. 1996).
Studies published since the 2006 Pb AQCD continue to observe associations between increasing
blood Pb level and decrements in these specific indices of cognitive function. Compared with studies of
IQ, studies of specific cognitive indices consistently find associations at lower blood Pb levels
(population means: 1.2 to 7 (ig/dL and quantities of blood Pb levels as low as 2 (ig/dL) (Figure 5-4 and
Table 5-4). Recent studies of cognitive function also expanded on previous evidence by providing
information on effect modification by genetic, nutritional, and caregiving quality (E. F. Krieg. Jr. et al..
2010; Pilsner et al.. 2010; Solon et al.. 2008; Surkan et al.. 2008).
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Reference
Exposure
Period
Blood PbLevel
(|jg/dL)a
Outcome
Jedrychowski et al. (2009b)
Cho et al. (2010)
Krieg et al. (2010)
Miranda et al. (2007)
Miranda et al. (2009)
Froehlich et al. (2007)
Surkan et al. (2008)
Pilsner et al. (2010)
Hu et al. (2006)
Prenatal
Concurrent
Concurrent
1.23 (0.44 to 5)
1.9 (0.67)
1.95 (0.16)
Early childhood 2 vs. 1
3 vs. V
Early childhood 2 vs. 1d
3 vs.1d
Concurrent
6.1 (4.9)
Concurrent 6.4 (4.3)
Prenatal 6.7 (3.6)
Prenatal 7.1 (5.1)
Bayley MDI
No omission errors'3 *
Block design
Digit Span
Reading score
Math score
End of grade scorec
End of grade score
Spatial memory
Rule learning
Spatial span
Problem solving
Bayley MDI, all subjects
Bayley MDI, high maternal self-esteem
Bayley MDI, low maternal self-esteem
Bayley MDI
Bayley MDI *
~
~
~
-0.8 -0.4 -0.2 0 0.2
Change in test score (standardized) per unit increase in blood Pb (95% CI)
Note: Test scores were standardized to their standard deviation to facilitate comparisons among
tests with different scales. Studies are presented in ascending order of blood Pb level. Effect
estimates are standardized to a 1 [jg/dL increase in blood Pb. MDI = Mental Developmental Index.
aBlood Pb level refers to the study mean (SD), unless otherwise specified. bStandard error was
estimated from p-value. Effect estimate represents association of blood Pb level with number of
correct answers on test. cStandard error was estimated from p-value. dEffect estimate compares
test scores of children in different categories of blood Pb concentration.
Figure 5-4. Associations of blood Pb levels with standardized scores for
specific indices of cognitive function in children.
Table 5-4. Additional characteristics and quantitative results for studies presented in
Figure 5-4
Study
Population/
Location
Blood Pb
Levels (|jg/dL)
Statistical Analysis
Cognitive Index
Effect Estimate
(95% Cl)a
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
-0.26 (-0.49, -0.02)
Cho et al. (2010)
667 children ages 8-
11 yr in 2008
Five Korean cities
Concurrent mean
(range): 1.9 (0.53-
6.16)
Log linear regression model
adjusted for age, sex, parental
education, maternal IQ, child
IQ, birth weight, urinary
cotinine
No omission errors"
No comission errors'1
Word reading score0
using KEDI-WISC at
ages 8-11 yr
-0.45 (-3.02, 2.06)
-1.80 (-4.33, 0.26)
-1.37 (-3.77, 1.03)
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Study
Population/
Location
Blood Pb
Levels (|jg/dL)
Statistical Analysis Cognitive Index
Effect Estimate
(95% Cl)a
Krieg et al.
(2010)
842 children ages
12-16 yr
U.S. NHANES III
(1991-1994)
Concurrent mean
(95% CI): 1.95
(1.63,2.27)
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
-0.05 (-0.08, -0.009)
-0.04 (-0.06, -0.02)
-0.05 (-0.07, -0.02)
-0.03 (-0.05, -0.004)
Miranda et al. 8603 4th grade Early childhood
(2007) children, 2000-2004 (ages 0-5 yr.)
7 counties in central range: 1 ->10
North Carolina
Linear regression model
adjusted for sex, race,
free/reduced-price lunch,
parental education, daily
computer use, charter school,
age of blood Pb screening,
school system
4th grade end-of-
grade score0
-0.08 (-0.16,0), blood Pb2
|jg/dL vs. 1 |jg/dLd
-0.13 (-0.21,-0.05), blood
Pb 3 |jg/dL vs. 1 ^ig/dLd
Miranda et al. 57,678 4th grade Early childhood
(2009) children, 2001-2005 (ages 9-36 mos.)
All 100 North mean (range): 4.8
Carolina counties (1-16)
Linear regression model
adjusted for race, sex, parental
education, free/reduced-price
lunch, charter school, school
system
4th grade end-of-
grade score
-0.04 (-0.07, -0.001), blood
Pb 2 |jg/dL vs. 1 ^ig/dLd
-0.05 (-0.09, -0.02), blood
Pb 3 |jg/dL vs. 1 ^ig/dLd
Froelich et al.
(2007)
174 children age 5 yr Concurrent mean
Rochester, NY (SD): 6.1 (4.9)
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
5yr
-0.02 (-0.06. -0.008)
-0.03 (-0.06, -0.001)
-0.007 (-0.01,0)
-0.04 (-0.09, 0.01)
Surkan et al.
(2008)
309 children ages
12-36 mo during
1996-2001 or 2004-
2005
Mexico City, Mexico
Concurrent mean
(SD): 6.4 (4.3)
Linear mixed effects
regression model adjusted for
sex, maternal age, maternal
IQ, maternal education, parity,
alcohol consumption, smoking,
cohort, maternal self-esteem
Bayley MDI, all
subjects
Bayley MDI, high
maternal self-esteem
Bayley MDI, low
maternal self-esteem
assessed at ages 12-
36 mo
-0.013 (-0.033, 0.0007)
0.027 (-0.037, 0.09)
-0.02 (-0.04, -0.001)
Pilsner etal.
(2010)
255 children age 24
mo born 1994-1995
Mexico City, Mexico
Prenatal (cord
blood) mean (SD):
6.7 (3.6)
Linear regression model
adjusted for maternal age,
maternal IQ, marital status,
parity, gestational age,
inadequate folate intake,
MTHFR genotype
Bayley MDI assessed -0.051 (-0.087,-0.016)
at age 24 mo
Hu et al. (2006)
146 children born
1997-1999 followed
prenatally to age 24
mo
Mexico City, Mexico
Prenatal (maternal
blood Pb) in 1st
trimester mean
(range):
7.1 (1.5-43.6)
Log linear regression model
adjusted for concurrent blood
Pb, sex, maternal age, current
weight, height-for-age Z score,
maternal IQ
Bayley MDI assessed -0.36 (-0.70, -0.01)
at age 24 mo
Studies not included in figure due to lack of sufficient data to calculate z-scores
Surkan et al.
(2007)
534 children ages 6-
10 yr
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
Reading score -5.20 (-9.45, -0.95)
Math score -4.02 (-7.6, -0.43),
Assessed using WIAT blood Pb level 5-10 |jg/dL
at age 6-10 yr vs. 1 -2 |jg/dLd
Chandramouli et
al. (2009)
488 children born
1991-1992 followed
from birth to age 7-8
yr
Avon, U.K.
Early childhood
(age 30 mos.)
mean (SD): 4.22
(3.12)
Log linear regression model
adjusted for sex, child IQ,
maternal education, home
ownership, maternal smoking,
HOME score, paternal SES,
family adversity index,
parenting attitudes at 6 mos.
Standardized
Assessment Test
assessed at age 7-8
yr
-0.61 (-0.82, -0.46),
continuous blood Pb
-1.08 (-1.71,-0.69), blood
Pb 2-5 |jg/dL vs. 0-2 |jg/dLd
-0.49 (-0.79, -0.31), blood
Pb 5-10 |jg/dL vs. 0-2 |jg/dLd
-0.44 (-0.93, -0.21), blood
Pb >10 |jg/dLvs. 0-2 |jg/dLd
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Study
Population/
Location
Blood Pb
Levels (|jg/dL)
Statistical Analysis Cognitive Index
Effect Estimate
(95% Cl)a
Solon etal.
(2008)
502 children ages 6-
35 mo
377 children 3-5 yr,
2003-2004
Visayas, Philippines
Concurrent mean
(range): 7.1
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)
BSID-II (ages 6-35
mo)
WPPSI-III (ages 3-5
yr)
-3.32 (-5.02,-1.60)
-2.47 (-4.58, -0.35)
MDI = Mental Development Index, MTHFR = methylenetetrahydrofolate reductase, WIAT = Weschler Individual Achievement Test
aEffect estimates are transformed to a z-score and standardized to a 1 |jg/dl_ increase in blood Pb level.
b95% 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. Effect estimate represents association of blood Pb
with correct answers on test.
°Standard error was estimated from p-value.
dEffect estimates compare test performance of children in higher blood Pb groups to children in lowest blood Pb group.
Krieg et al. (2010) examined an older subset of children (ages 12-16 years) from NHANES III
(1988-1994) that were previously examined in Lanphear et al. (2000). Similar to Lanphear et al. (2000).
Krieg et al. (2010) found associations of increasing concurrent blood Pb level with decrements in block
design, digit span, reading score, and arithmetic score in these older children (Figure 5-4 and Table 5-4).
In another study of Korean children (ages 8-11 years) with similar blood Pb levels (mean: 1.9 (ig/dL [SD:
0.67]), Cho and colleagues (2010) found concurrent blood Pb levels to be associated with decreases in in
tests of attention (errors in responding to targets); however associations were not statistically significant
in models that adjusted for urinary cotinine levels (Figure 5-4 and Table 5-4). Although associations of
blood Pb levels with word and color naming were negative, effect estimates were associated with wide
95% CIs.
Krieg et al. (2010) additionally 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 performance for some tests and a greater Pb-associated decrement in performance for other
tests. 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 slopes for digit
span score (95% CI) per 1 (ig/dL increase in blood Pb level were -1.5 (-2.2, -0.71) for the CC genotype
and -0.26 (-0.99, 0.46) for the TT genotype. Conversely, the TT genotype was associated with the greatest
Pb-associated decrease in arithmetic score. The slopes (95% CIs) for math score per 1 (ig/dL increase in
blood Pb level were -8.4 (-11.5, -5.2) for the TT genotype and -2.7 (-10.1, 4.6) forthe CC genotype.
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Effect modification by VDR rs731236 was more consistent across cognitive tests, with larger decrements
in blood Pb-associated cognitive performance among children with the CC genotype.
Multiple studies in different Mexico City mother-child dyads recently reported on associations
between blood Pb levels (e.g., maternal, cord blood, or child postnatal Pb levels) and mental development
in children at age 24 or 36 months (H. Hu et al.. 2006; Pilsner et al.. 2010; Surkan et al.. 2008). Hu et al.
(2006) assessed the impact of timing of exposure prenatally among 146 mother-child dyads 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 child's blood Pb level at 24 months when
the 24-month Bayley Mental Development Index (MDI) was ascertained. Comparing whole blood and
plasma Pb levels collected at each of 3 trimesters, this group found that 1st trimester blood Pb (or plasma
Pb) was the best predictor of subsequent 24-month Bayley Scale MDI scores. Another study excluding
the first trimester demonstrated an inverse association between IQ assessed at age 6-10 years and third
trimester maternal blood Pb level, but not with maternal blood Pb levels measured at other times during
pregnancy or within child Pb blood levels averaged over 6 to 10 years (Schnaas et al.. 2006V
Surkan et al. (2008) found negative associations (statistically nonsignificant) of concurrent blood
Pb levels with Bayley MDI and Psychomotor Development Index overall in a population of 379 Mexico
City children between ages 12 and 36 months. However, when data were stratified by maternal self-
esteem, negative associations were observed among children with mothers with in the lowest three
quartiles of self-esteem but not among children with mothers in the highest quartile of self-esteem (Figure
5-4 and Table 5-4). These findings indicate that higher maternal psychosocial functioning (e.g., stress,
anxiety, depression, self-esteem) may contribute to better caregiving, which in turn may improve
neuropsychological functioning of the child. These limited data in humans are well-supported by findings
from animal studies that have shown that environmental enrichment can reverse the effects of early stress
experiences on reactions such as depressed behavior, HPA activation, and immunosuppression (Francis et
al.. 2002; Laviola etal.. 2008; Laviola et al.. 2004; Morelev-Fletcher et al.. 2003). With specific regards
to Pb exposure, Schneider et al. (2001) and Guilarte et al. (2003) demonstrated that the social isolation or
enrichment can exacerbate or protect against, respectively, the neurological effects from Pb exposure. 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
& Jirtle. 2008). Pb exposure of animals and blood Pb levels in humans have been associated with altered
DNA methylation patterns which in turn, may be associated with altered gene expression patterns
(Section 5.3.6.11 and Section 5.10.4).
In a recent study in Mexico City, Mexico, investigators found increasing cord blood Pb levels to be
associated with lower MDI scores among children at age 24 months (-0.051 points [95% CI: -0.087, -
0.016] in standardized score per 1 (ig/dL increase in blood Pb level) (Pilsner et al.. 2010). Investigators
additionally examined effect modification by variants in the methylenetetrahydrofolate reductase
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(MTHFR) gene. The MTHFR enzyme is involved in folate metabolism, specifically, catalyzing the
conversion of 5,10-methylenetetrahydrofolate to 5-methylenehydrofolate, 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). Although Pilsner et al. (2010) found that both cord blood Pb
levels and the MTHFR 677T 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.
Instead of analyzing MTHFR genetic variants to represent folate metabolism, Solon et al. (2008)
measured red blood cell folate levels among children in the Philippines. Not only did investigators find
an association between increasing blood Pb level and lower cognitive test performance (Table 5-4), but
they 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 negative marginal effect on MDI (-0.80 to -2.44 points),
whereas among children with higher folate levels, blood Pb level did not have a negative marginal impact.
Thus, in contrast with those from Pilsner et al. (2010). findings from Solon et al. (2008) indicate that
children with folate deficiencies may be at increased susceptibility to Pb effects on cognitive function.
Academic Performance in Children
Although the preponderance of evidence for Pb-associated neurodevelopmental deficits is for IQ
and specific indices of cognitive function, academic achievement and school performance are corollaries
to aptitude that may be more objective measure of one's 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 an individual's actual
knowledge 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's 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 outcomes at
older ages (ages 8-18 years), suggesting early exposure to Pb may have persistent effects (Bellinger et al..
1992; Leviton et al.. 1993; Needleman et al. 1990). Results from the longitudinal study by Bellinger et al.
(1992) was particularly noteworthy for examining associations of blood Pb at several ages with scores on
the Kaufman Test of Educational Achievement at age 10 years. Only blood Pb level at age 2 years showed
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a statistically significant association with lower predicted academic achievement. Additionally, the
association was robust to adjustment for IQ, indicating that blood Pb levels may be associated with
reduced performance on academic tasks not reflected in indices of IQ. Several studies also found
associations between concurrent blood Pb levels and academic achievement (Al-Saleh et al.. 2001;
Kordas et al.. 2006; Lanphear et al.. 2000; C.-L. Wang et al.. 2002). Among recent studies, academic
performance was examined less frequently; however, findings are consistent with the extant body of
evidence.
Miranda et al. (2007) linked blood Pb surveillance data collected between 0 and 5 years with end-
of-grade (EOG) testing data at 4th grade for 7 of the largest counties in North Carolina. Approximately
22-30% of children in these counties were screened for Pb poisoning, and the total sample size in the
analysis was approximately 8,600 children for both math and reading achievement tests. For both reading
and math, achievement test scores were inversely associated with early childhood blood Pb screening data
(Figures 5-4 and 5-5 and Table 5-4).
£ -0 5
u
o -i.o
-1.5
-2.0
-2.5
-3.0
-3.5
~ Model 1: linear BLL
~ Model 3: dummy for each BLL (1 to 210)
-4.0
2
3
6
4
5
7
8
9
Blood lead levels (ug/dL)
Source: Used with permission from Elsevier Science, Miranda et al. (2007).
Figure 5-5. Comparing model results for 4th-grade EOG mathematics
scores.Based on a referent individual who was screened at 2
years of age and is a white female, living in Wake County, NC,
parents with a high school education, not enrolled in the
school lunch program, and who does not use a computer
every day. Baseline score is 262.6.
Similar results were obtained in a subsequent study expanded to the entire state of North Carolina
(Miranda et al.. 2009) (Figure 5-4 and Table 5-4). Investigators additionally used quantile regression to
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estimate effects for conditional percentiles of EOG (e.g., what is the 10th percentile of EOG test scores
conditioned on early childhood blood Pb levels) rather than conditional means. Compared with linear
regression, quantile regression is more robust in response to outliers and predicts outcomes at the top and
bottom tails of the distribution of the outcome rather than at the mean. The distributions m EOG scores
for children with blood Pb levels greater than or equal to 10 (.ig/dL were more spread out than those for
children with lower blood Pb levels. With increasing blood Pb levels, the lower tail of the EOG
distribution was stretched out more so than the middle or upper tail of the distribution. For example, in
comparisons of children with blood Pb levels of 5 (ig/dL versus children with blood Pb levels of 1 (.ig/dl..
children in the 5th percentile of EOG have a greater decrease in EOG score compared with children in the
95 th percentile of EOG (Figure 5-6). These findings indicate that children residing at the lowest
performance regions of the EOG distribution may be differentially affected by blood Pb levels as
compared with children in the middle or higher regions. Similarly, using quantile regression, Miranda et
al. (2009) showed that, in addition to elevated blood Pb levels, cumulative social risk (lower parental
education, being enrolled in a school lunch program) further enhanced the negative effects on academic
achievement in these children.
a, 0
c
Q
w
ro
CO
E
&
2
o
o
«
0
O
w
0)
(A
re
£
o
0J
Q
Cummulative Deficit: Decrease in EOG scores by multiple risk factors
5% 10% 15* 20* 25% 30% 35% *0% 45% 50% 55% 60% 65% 70% 75% 80% KX 90% 95%
-2 ¦
-4
-6 ¦
-8
-10 ¦
-12
I I I
re-Baseline (BLL=1pg/dL, no school
lunch program, parents completed college)
Effect of reduction in parental education
~from completed college
to only completed hign school
I Income effect as indicated by
f enrollment in school lunch program
> Effect of increased BLL from 1 to 5 pg/dL
Quantile
Source: Used with permission from Elsevier Science, Miranda et al. (2009)
Figure 5-6. Reduction in EOG achievement test scores at each percentile
of the test distribution.Note greater effect of Pb at low end of
the distribution.
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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 at
age 7 years) among participants of the Avon Longitudinal Study of Parents and Children conducted in the
U.K (Figure 5-4 and Table 5-4). While these aforementioned recent studies found negative associations
for early childhood blood Pb levels, unlike the longitudinal assessment by Bellinger et al. (1992), they did
not have available blood Pb measurements at other lifestages to compare associations with blood Pb
levels at other lifestages. Therefore, results from these recent studies do not preclude associations with
blood Pb levels at other lifestages. Among children (ages 6-10 years) participating in NECAT, increasing
concurrent blood Pb level was associated with poorer performance on the Wechsler Individual
Achievement Test, even when adjusted for IQ (Surkan et al.. 2007). In analyses adjusted for child IQ,
caregiver IQ, age, SES, race, and birth weight, children with concurrent blood Pb levels 5-10 (ig/dL
scored 5.2 (95% CI: 0.95, 9.45) points and 4.0 (95% CI: 0.43, 7.6) points lower on reading and math
composite scores on the respectively, compared to children with levels of 1-2 (ig/dL. Blood Pb levels
were similarly associated with other tests of cognitive function (e.g., FSIQ, working memory, cognitive
flexibility and a number of executive functioning domains [i.e., ability to formulate, test, and adapt
hypotheses]). Children with blood Pb levels 3-4 (ig/dL had lower scores compared with children with
blood Pb levels 1-2 (ig/dL; however, differences were not statistically significant.
Age-based Susceptibility to Lead-associated Neurodevelopmental Deficits
Plasticity is a consequence of environmental exposures during critical life periods affecting key
physiological systems that orchestrate underlying developmental processes (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 (Landrigan et al.. 1999; Rice & Barone. 2000). This hypothesis is well-
supported by findings in animals that prenatal Pb exposure alters brain development via changes in
synaptic architecture (Section 5.3.6.5) and neuronal outgrowth (Section 5.3.6.10) and leads to
impairments in memory and learning (Section 5.3.2.2) and emotional and depressive changes postnatally
(Section 5.3.3.4). Unlike other organ systems, the unidirectional nature of CNS development limits the
capacity 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 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 5.3.5.2).
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With repeated assessment of children prenatally to later childhood and early adulthood, the
prospective cohort studies have aimed to distinguish among the effects of blood Pb levels at different
periods of development. In the collective body of evidence, neurocognitive decrements have been
associated with prenatal, early childhood, childhood average, and concurrent blood Pb levels. In these
studies, the identification of developmental periods when children are most sensitive to Pb-associated
neurocognitive decrements has been complicated by the high degree of correlation in children's blood Pb
levels over time and the confounding of age and peak blood Pb levels (Dietrich. Berger. & Succop. 1993;
Lanphear et al.. 2005; Needleman et al.. 1990V
As described in detail in the 2006 Pb AQCD, several studies with varying lengths of follow-up
demonstrated associations of prenatal blood Pb levels with neurodevelopmental deficits thoughout
childhood and into early adulthood (U.S. EPA. 2006). The prenatal period may be susceptible life stage of
Pb exposure not only because of the nervous system developmental process occurring 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 even related to remote exposures (Gulson et al.. 2003; H.
Hu & Hemandez-Avila. 2002). Pb can cross the placenta to affect the developing fetal nervous system
(Rabinowitz. 1988). Maternal and umbilical cord blood Pb levels generally are highly correlated,
indicating that a newborn infant's blood Pb levels reflects that of the mother (Schell et al.. 2003).
Associations with prenatal blood Pb levels were demonstrated most consistently for cognitive
function and behavior assessed between infancy and age 3 years (Figures 5-7 and 5-8 and Table 5-5).
Among studies examining associations of early-life blood Pb measures (maternal, cord, or neonatal
blood), results were mixed as to whether prenatal (Bellinger et al.. 1984; H. Hu et al.. 2006) or concurrent
blood Pb levels (G. Wasserman et al.. 1992; G. A. Wasserman et al.. 1998) were associated with a greater
decrement in cognitive function. Several studies found that prenatal or neonatal blood Pb levels were
associated with neurodevelopmental decrements assessed in neonates (within 30 days) or early in infancy
(within 3 months), which indicated that relatively short-durations of Pb exposure may be associated with
negative neurological effects (Dietrich et al.. 1987; Emhart et al.. 1986; Rothenberg et al.. 1989; Shenet
al.. 1998).
A recent analysis of 444 children participating in the Krakow Prospective Cohort Study
corroborated previous findings for prenatal exposure and found associations at lower umbilical cord blood
Pb levels than those in previous studies. Children in Jedrychowski et al. (2009b) had a median (range) of
cord blood Pb levels of 1.23 (0.44-6.9 (ig/dL), and increasing umbilical cord blood Pb levels was
associated with lower 36-month Bayley MDI (Figure 5-4 and Table 5-4). Investigators also observed a
larger magnitude of effect in males compared with females (Figure 5-9). These findings are consistent
with the hypothesis that the developing male central nervous system may be more vulnerable than
females' to environmental insults resulting in later behavioral problems (Moffitt et al.. 2001).
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Mean (SD)
blood Pb (ng/dL)
Exposure
Period
Rothenberg et al. (1989)
Dietrich etal. (1987)
Dietrich etal. (1986)
Hu et al. (2006)
Gomaa et al. (2002)
Jedrychowski et al. (2009b)
Wasserman et al. (1992)
Bellinger et al. (1987)
Assessment at School Age
Wasserman et al. (1994)
Bellinger et al. (1991)
Dietrich et al. (1992)e
Dietrich et al. (1993)
Dietrich et al. (1993)
Baghurst et al. (1992)
Lanphear et al. (2005)
Pocock et al. (1994)
Ris et al. (2004)
6.3 (4.5)
5.9 (3.4)
8.0 (3.8)
4.5(2.9)
7.1 (5.1)
NR
5.2 (3.4)
4.8 (3.7)
6.7 (3.4)
1.29(1.24, 1.34)a
14.4(10.4)
35.4, 8.5b
14.4(10.5)
39.9, 9.6b
6.8 (6.3)
2.8 (1.7)d
6.4 (4.1)
8.2 (3.8)
4.8 (3.3)
11.9(6.4)
8.4 (3.8)
4.8 (3.1)
10.1 (5.6)
NR
8.3 (3.7)
5.0 (3.4)
11.8(6.3)
Self-quieting, 30 days
MDI. 3 mo
MDI, 6 mo
MDI, 24 mo
MDI, 24 mo
MDI, 24 mo
MDI, 24 mo
MDI, 24 mo
GCI, 4 yr
GCI, 57 mo
FWS, 5 yr
Coordination, 6 yr
FSIQ, 6.5 yr
Prenatal
Concurrent
Prenatal
Neonatal
Prenatal (1st trimester)
Prenatal (average)
12 month
Concurrent
Prenatal
Prenatal
Prenatal
Concurrent
Prenatal
Concurrent
Early childhood, 2 yr
Early childhood, tooth
Concurrent
Prenatal
Neonatal
Concurrent
Lifetime avg
Prenatal
Neonatal
Concurrent
Lifetime avg
Prenatal
Neonatal
Concurrent
Lifetime avg
7.4 (2n quartile)
16.6 (2nd quartile)
15.7 (2nd quartile)
FSIQ, 7 yr
Prenatal
Early childhood, 2 yr
Lifetime avg
12.7 (4.0-34.5)
18.0 (6.2-47.0)a
12.4 (4.1-34.8)a
9.7 (2.5-33.2)a
NR
6.8-21.2
NR
NR
NR
Learning, 15-17 yr
Early childhood
Peak
Lifetime average
Concurrent
Around birth
Early childhood, 2 yr
Postnatal mean
Prenatal
Early childhood, 6.5 yr
Early childhood avg
~
~
f
Schnaas et al. (2006)
NR
FSIQ, 6-10 yr
9.8 (2.8-36.4)'
6.2 (2.2-18.6)f
Early childhood avg
Later childhood avg
~ ~
Bellinger et al. (1992)
>10 vs. <3C
FSIQ, 10 yr
Prenatal
6.5 (4.9)
2.9 (2.4)
Early childhood, 2 yr
Concurrent, 10 yr
~
-7.0
-6.0
-5.0 -4.0 -3.0 -2.0 -1.0
Change in Cognitive Score (95% CI)
0.0
1.0
Note: Effect estimates are standardized to a blood Pb levels of 1 [jg/dL. Studies are arranged in
order of ascending age of cognitive function assessment. Cognitive function test scores are not
standardized to a similar scale because not all studies provided sufficient data.Red = prenatal or
neonatal blood Pb levels, Blue = Early childhood levels, Black = concurrent or lifetime average
levels. MDI = Mental Development Index, Bayley Scales; NR = Not reported; GCI = General
Cognitive Index, McCarthy Scales; FWS = Filtered Word Test, Kaufman Assessment Battery for
Children, FSIQ = Full-scale IQ. a = 95% CI for blood Pb levels, b = values represent mean blood
Pb levels in the two towns studied, c = Effect estimate compares children in different categories of
blood Pb levels, d = mean (SD) and effect estimate for tooth Pb levels (pg/g). e = Sufficient data
were not provided in order to calculate 95% CIs. f = range of blood Pb levels.
Figure 5-7. Associations of blood Pb measures at various lifestages with
cognitive function in children.
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Study
Mean (SD)
Blood Pb (ug/dL)
Wasserman etal. (1998) 16.1 (2.6)
25.8 (19.1)
Anxiety/ Prenatal
Depression, Concurrent
3 yrs
Leviton et al. (1993) 4.8-6.3 (2nd quartile) Hyperactivity, Prenatal
2.0-2.9 (2nd quartile)3 Girls, 8 yrs Early childhood, 6 yr
Bellinger etal. (1994) 1.98(0.38)
3.4 (2.4)b
Behavioral Prenatal <-
Problems, 8 yrs Early childhood, 6.5 yr
Risetal. (2004) NR
NR
NR
Inattention,
15-17 yr
Prenatal
Early childhood, 6.5 yr
Early childhood avg
Dietrich et al. (2001) 8.9 (3.9)
NR
NR
Delinquent
Behavior,
15-17 yrs
Prenatal
Early childhood, 6.5 yr
Early childhood avg
-1.0 -0.5 0.0 0.5 1.0 1.5 2.0
Change in Behavioral Index per 1 unit increase in Pb biomarker (95% CI)
Note: Positive effect estimates represent an increase in behavioral index. Effect estimates are
standardized to a blood Pb levels of 1 [jg/dL. Studies are arranged in order of ascending age of
behavioral assessment. Behavioral asssessment scores are not standardized to a similar scale
because not all studies provided sufficient data. Red = prenatal blood Pb levels; Black = blood Pb
levels at other lifestages. a = second quartile levels (pg/g) and effect estimate for tooth Pb levels b
= mean (SD) in ppm and effect estimate for tooth Pb levels.
Figure 5-8. Associations of Pb biomarkers at various lifestages with
behavioral indices in children.
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Table 5-5. Additional characteristics and quantitative results for studies presented in
Figures 5-7 and 5-8
Study
Population/Location
Blood Pb Levels
(pg/dL)
Statistical Analysis
„ . Effect Estimate
Outcome (95o/o
Rothenberg etal.
42 children followed
Maternal week 36
Regression model
Self-quieting Prenatal:-0.091 (-0.18,0)
(1989)
prenatally to child age 30
gestation mean (SD):
adjusted for smoking,
ability (regulation
days
15.0(6.4)
single mother, problems in
of state) at age 30
Mexico City, Mexico
Maternal at birth
pregnancy, alcohol use in
days
mean (SD): 15.5(5.7)
previous month, use of
Assessed using
spinal block, gravidity,
Newborn
income
Brazelton
Assessment
System
Dietrich etal.
(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)
Huetal. (2006)
146 children born 1997-
1999 followed prenatally to
age 24 mo
Mexico City, Mexico
Prenatal (maternal
blood Pb) in 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 1a 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)
Jedrychowski et 444 children born 2001- Prenatal (cord blood)
al. (2009b) 2004 followed prenatally to geometric mean
age 36 mo (range): 1.29 (0.44-5)
Krakow, Poland
Linear regression model Bayley MDI
adjusted for maternal assessed at age
education, birth order, 36 mo
prenatal ETS, sex
Prenatal: -2.9 (-5.0, -0.75)
Wasserman etal.
(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. Mistrovica: 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
Prenatal: -3.2 (-7.2, 0.86)
Concurrent: -4.1 (-6.2, -
2.0)
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 wg/dL
vs. blood Pb levels <3
Cognitive function assessments at school age
Wasserman etal.
(1994)
332 children followed
prenatally to age 3-4 yr
Kosovo, Yugoslavia (i<.
Mitrovica, Pristina)
Prenatal (cord blood)
mean (SD): 14.4
(10.4)
Concurrent means:
K. Mistrovica: 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
Prenatal: -3.2 (-5.1, -1.2)
Concurrent: -4.1 (-6.2, -2.0)
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0 , Blood Pb Levels ~ t. t. , A , ~ t Effect Estimate
Study Population/Location Statistical Analysis Outcome (95% Cl)a
Bellinger et al. 170 children followed from Early childhood (24
(1991') birth (1979-1981) to age mo) mean (SD): 6.8
57 mo (6.3)
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
(-5.7, -0.2)
Early childhood tooth: -2.5
(-10.2, 5.2)
Concurrent blood: -2.3 (-
6.0,1.4)
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
Total FWS Prenatal:-0.25, p< 0.01c
assessed using Neonatal: -0.38, p < 0.01°
KABC at age 5 yr Concurrent: -0.19, p < 0.01c
Lifetime avg: -0.16, p <
0.01°
Dietrich et al.
(1993)
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 obstretric
complications, perinatal
status, sex, social class,
maternal intelligence,
quality of rearing
environment, earlier
measures of
neurobehavioral status
Bruininks- Prenatal: -0.04 (-0.20, 0.12)
Oseretsky Test of Neonatal: -0.15 (-0.33,
Motor Proficiency 0.03)
assessed at age Concurrent: -0.18 (-0.26, -
6yr 0.10)
Lifetime avg: -0.11 (-0.19, -
0.03)
Dietrich et al.
(1993)
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
Prenatal: 0.15 (-0.26, 0.56)
Neonatal: -0.03 (-0.42,
0.36)
Concurrent: -0.33 (-0.60, -
0.06)
Lifetime avg: -0.13 (-0.35,
0.09)
Prenatal (maternal
28-36 wk gestation):
NR
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)
Schnaas etal.
(2006)
150 children followed from
prenatally (1987-1992) to
age 6-10 yr
Mexico City, Mexico
Log linear mixed effects
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
Prenatal (28-36 weeks
gestation): -3.9 (-6.5,1.4)
Early childhood avg: 0.10 (-
3.9,4.1)
Later childhood avg: 0.17 (-
1.4,1.8)
Baghurst et al.
(1992)
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
Prenatal: 0.26 (-0.67,1.5)
Early childhood: -2.0 (-3.8,
-0.21)
Lifetime avg: -1.6 (-3.7,
0.52)
Bellinger etal.
(1992)
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 FSIQ assessed
adjusted for HOME score using WISC-R at
(age 10 and 5), child agelOyr
stress, race, maternal IQ,
SES, sex, birth order,
marital status
Prenatal: -0.48 (-5.7, 4.7),
blood Pb >10 ug/dL vs. <3
|jg/dLb
Early childhood: -0.58 (-
0.99,-0.17)
Concurrent: -0.46 (-1.5,
0.56)
Lanphearetal.
(2005)
1333 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
Early childhood: -0.14 (-
0.23, -0.06)
Peak: -0.20 (-0.29,-0.11)
Lifetime avg: -0.15 (-0.22, -
0.09)
Concurrent: -0.23 (-0.32, -
0.14)
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0 , Blood Pb Levels ~ t. t. , A , ~ t Effect Estimate
Study Population/Location Statistical Analysis Outcome (95% Cl)a
Pocock et al.
(1994)
Meta-analysis of 5
prospective (over 1100
children and 14 cross-
sectional studies (3499
children)
Early childhood (2 yr) Meta-analysis of FSIQ assessed Around birth: 0.26 (-1.5,
range in means: combining effect using various 2.0)
6.8-21.2 estimates from individual tests at ages 5-10 Early childhood:-2.7 (-4.1,
studies yr 1.2)
Postnatal mean: -1.3 (-2.9,
0.37)
Ris etal. (2004) 195 children in followed NR
from birth (1979-1985)
to age
15-17 yr
Cincinnati, OH
Linear regression model
adjusted for maternal IQ,
sex, and average total
HOME score
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
Wasserman etal.
(1998)
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 score
Anxiety/depression Prenatal: 1.16 (0.02, 2.3)
assessed using Concurrent: 1.45 (0.04,
Child Behavior 2.86)
Checklist at 3 yr
Leviton etal.
(1993)
1923 children followed
from birth (1979-1980)
to age 8 yr
Boston area, MA
Prenatal blood 2na
quartile: 4.8-6.3
Early childhood
(tooth) second
quartile: 2.0-2.9 |jg/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.
(1994)
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
Prenatal: -0.31 (-1.7,1.07)
Early childhood: 1.8 (0.49,
3.1)
Risetal. (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 score
Inattention
composite
assessed using
Continuous
Performance Test
Prenatal: 0.16 (0.04, 0.27)
Early childhood, 6.5 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 (bornl 979-
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
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 = Weschler Intelligence Scale for Children, NR = Not reported
aEffect estimates are standardized to a 1 |jg/dl_ increase in blood Pb level in analyses of blood Pb as a continuous variable.
bEffect 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.
cSufficent data were not provided in order to calculate 95% CI.
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7
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10
11
12
13
14
15
16
17
18
19
m
o
CD
•>—
o
o
CO
CD
CO
in
o>
o
o>
Boys
Girls
T—
-.5
Graphs by gender
Source: Used with permission from Elsevier Science, Jedrychowski et al. (2009a).
Figure 5-9. Regression of fitted MDI score at 36 months on log-
transformed concentration of cord blood Pb level by sex.
In studies that examined cognitive and behavioral indices measured in schoolaged children (ages 4-
17 years), statistically significant, negative associations with prenatal or neonatal (10 days after birth)
blood Pb measures were observed less frequently (Figures 5-7 and 5-8 and Table 5-5). Multiple studies
conducted in the Cincinnati cohort found that blood Pb levels measured 10 days after birth but not
maternal blood Pb during pregnancy were associated with impairments in cognitive function, auditory
processing, and motor function as well as increased behavioral problems in children between ages 4 and 6
years (Dietrich. Berger. & Succop. 1993; Dietrich. Berger. Succop. et al.. 1993; Dietrich et al.. 1991;
Dietrich et al.. 1992). In these Cincinnati studies, concurrent blood Pb levels generally were estimated to
have similar magnitudes of effect. In most of the studies examining associations of prenatal or neonatal
blood Pb levels with neurodevelopmental outcomes in later childhood, early or cumulative childhood
blood Pb levels were associated with the greater decrements in function. A larger effect estimate for peak
blood Pb levels was corroborated in meta-analysis of results from five cohort studies (Pocock et al.. 1994)
(Figure 5-7 and Table 5-5). These findings may indicate the lack of persistence of early Pb exposures.
Further, associations of cord blood Pb levels with neurodevelopmental measures in infancy may reflect
associations with blood Pb levels in infancy, which are expected to be similar to those during the prenatal
period. Cord blood Pb level may be a good surrogate of early postnatal blood Pb levels.
Early childhood blood Pb levels also were associated with diverse neurodevelopmental effects
assessed later in childhood and into early adulthood in both recent (reviewed earlier in Section 5.3.2.1,
Figure 5-4, and Table 5-4) and previous studies that did not compare various lifestages of Pb exposure
95% CI Fitted values
Beta coeff. = -4.9,
P = 0.244
-i 1 r- -i 1 1 r
0 .5 1 -.5 0 .5 1
Pb level {log transformed)
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9
10
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16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
(Cecil et al.. 2005; Tong et al.. 2000; Yuan et al.. 2006V 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, Pb exposure may affect higher-order neurodevelopmental processes that are more
reliably assessed at later ages when children's processes modalities 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. In some studies that have reported associations between early-
childhood blood Pb levels and neurodevelopmental decrements later in life, children's blood Pb levels had
not markedly changed over time Thus, the early-childhood blood Pb levels may have been serving as
surrogates of concurrent or cumulative blood Pb levels.
As depicted in Figures 5-7 and 5-8 and Table 5-5, several studies estimated larger decreases in
neurodevelopmental endpoints for concurrent or lifetime average blood Pb levels than 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, and early childhood blood Pb levels were all
negatively associated with IQ, with the largest magnitude of decrease associated with concurrent blood Pb
levels (Lanphear et al. 2005). Childhood average blood Pb levels (Dietrich. Berger. & Succop. 1993;
Dietrich. Berger. Succop. etal.. 1993; Lanphear et al.. 2005) and tooth Pb levels (Bellinger et al.. 1994)
have been associated with neurodevelopmental effects, indicating that biomarkers of cumulative
childhood Pb exposure also may contribute to neurodevelopmental effects in children. Associations with
concurrent blood Pb level were also demonstrated consistently in studies without comparisons to
exposures at other lifestages (Figures 5-2 and 5-4 and Tables 5-3 and 5-4).
Some studies have aimed to improve assessment of age-based susceptibility by examining children
with different degrees of changes in blood Pb levels over time (i.e., children whose blood Pb level ranking
changed over time) (Bellinger et al.. 1990; A. Chen et al.. 2005; Hornung et al.. 2009; Tong et al.. 1998).
Except for Tong et al. (1998), these studies have demonstrated stronger effects of concurrent blood Pb
levels (Figure 5-10 and Table 5-6). Tong et al. (1998) found that early-life blood Pb level was associated
with a larger deficit in IQ. 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 intellectual attainment scores 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 lower
decline (<10 (ig/dL). These findings indicated a stronger effect of higher blood Pb levels early in life even
among groups with lower concurrent levels.
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Reference Outcome Blood Pb Variable Examined
Tong et al. (1998) FSIQ, 11-13 yr. < 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
~
Bellinger et al. (1990) GCI, 24 and 57 mo Concurrent, Prenatal < 3 ^
Concurrent, Prenatal 3-10a ~
Concurrent, Prenatal > 10 ~
Chen et al. (2005) FSIQ, 7 yr. Low 2 yr, (<24.9), Low 7 yr, (<7.2) <
Low 2 yr, (<24.9), High 7 yr, (> 7.2) ~
High 2yr, (>24.9), Low7yr, (<7.2) <
High 2 yr, (> 24.9), High 7 yr, (> 7.2) ~
>
«
Hornung et al. (2009) FSIQ, 6 yr. 0.5 ratio age 6 to 2 yrb <
2.0 ratio age 6 to 2 yrb +
~
-1.0 -0.5 0.0 0.5
Change in Cognitive Score (95% CI)
Note: Effect estimates represent associations between concurrent blood Pb level and cognitive
function (standardized to standard deviation) in children categorized by prenatal blood Pb level.
bValues represent the ratio of blood Pb level at age 6 years to that at age 2 years. FSIQ = Full-
scale IQ, GCI = General Cognitive Index. Cognitive function scores were standardized to their
standard deviation. Effect estimates in red represent blood Pb level variables associated with the
greater decrease in cognitive function.
Figure 5-10. Associations of cognitive function in children with different
degrees of changes in blood Pb levels overtime.
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2
3
4
5
6
7
8
9
10
Table 5-6. Additional characteristics and quantitative results for studies presented in
Figure 5-10
. Population/
study Location
Blood Pb
Levels
(pg/dL)
Statistical Analysis
Outcome
Effect Estimate
(95% Cl)a
Tongetal. 375 children Means: 21.2 Log linear regression model
(1998) followed from birth (age 2 yr), 7.9 adjusted for sex, birth weight, birth
(1979-1982) to age (age 11-13 yr) rank, feeding style, breastfeeding
11-13 yr duration, maternal IQ, maternal
Port Pirie, Australia age, SES, HOME score, 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
Bayley MDI at age 2 yr,
McCarthy GCI at age 4
yr, WISC-R at ages 7,
and 11 -13 yr
<10.2 |jg/dL decline: 0.03
(-0.15, 0.21 )b
10.2-16.2 m/dL decline: 0.04
(-0.15, 0.23?
>16.2 |jg/dL decline: -0.01
(-0.20, 0.18)b
Bellinger et 170 children
al. (1990) followed 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 level0
Prenatal <3 |jg/dL: -0.16
(-0.43, 0.11)
Prenatal 3-10 |jg/dL: -0.14
(-0.57, 0.29)
Prenatal > 10 |jg/dL: -0.46
(-0.81,-0.11)
Chen etal. 780 children
Mean (SD):
Linear regression model adjusted WISC-III at age 7 yr
Low age 2 (<24.9 |jg/dL, Low
(2005) participating in the
Age 2 yr: 26.2
for city, race, sex, language,
age 7 (<7.2 ^ig/dL): 0d
TLC trial from age
(5.1)
parental education, parental
Low age 2, High age 7: -0.27
12-33 mo to age 7
Age 5yr: 12.0
employment, single parent, age at
(-0.48, -0.05)
yr
(5.2)
blood Pb measurement, caregiver
High age 2, Low age 7: 0
Baltimore, MD;
Age 7 yr: 8.0
IQ
(-0.21,0.20)
Cincinnati, OH;
(4.0)
High age 2, High age 7: -0.28
Newark, NJ;
(-0.47, -0.10)
Philadelphia, PA
Hornung et
al. (2009)
462 children
followed from birth
(1979-1984) to age
6yr
Rochester, NY and
Cincinnati, OH
Geometric mean
(5th-95th):
Peak: 13.6(4.6-
34.4)
Early childhood:
8.9 (3.0-23.8)
Lifetime mean:
8.5 (3.0-22.1)
Concurrent: 6.0
(1.9-17.9)
Linear regression model adjusted
for city, HOME score, birth weight,
maternal IQ, maternal education
FSIQ assessed using
WISC-R at age 6 yr
0.5 ratio of blood Pb level at
age 6 to age 2: 0 (reference)
2.0 ratio of blood Pb level at
age 6 to age 2 yr: -0.70 (-1.0,
-0.40)
Effect estimates represent the cognitive function score or change in score over time standardized to its standard deviation.
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.
cEffects 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 s 10 [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 ng/dL), low
levels at age 2 (<11.4 ng/dL) and high levels at age 7 (>7.2 ng/dL), high levels at age 2 (>11.4 |jg/dl_) and low levels at age 7 (<7.2 ng/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.
In several different U.S. cohorts of children, larger decrements on neurocognitive function were
estimated for concurrent blood Pb levels (Bellinger etal.. 1990; A. Chen et al.. 2005; Hornung et al..
2009) (Figure 5-10 and Table 5-6). In the Boston cohort, Bellinger et al. (1990) found that at age 57
months, cognitive performance, as assessed by McCarthy GCI, was similar between children with higher
(>10 (ig/dL) and lower (<3 (ig/dL) prenatal blood Pb levels. Additionally, increasing concurrent blood Pb
levels (age 57 months) were associated with the largest decline in GCI scores overtime (score at age 57
months - score at age 24 months) among children with high prenatal blood Pb levels (>10 (ig/dL), which
indicated an effect among children with both high early and concurrent blood Pb levels (Figure 5-10 and
Table 5-6). These findings indicated that by age 5 years, children with higher prenatal blood Pb levels
appear to recover the Pb-associated decrements in cognitive function unless concurrent blood Pb levels
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21
remain 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.
As part of the multicenter Treatment of Lead-Exposed Children (TLC) trial, Chen et al. ("2005)
evaluated how change in blood Pb over time was related to IQ at later ages. The TLC was a clinical trial
designed to examine the effect of chelation using succimer to prevent cognitive impairment in 780 urban
children enrolled at 12 to 33 months of age with elevated blood Pb concentrations (20-44 (ig/dL). Blood
Pb and IQ were assessed at ages 24 and 36 months and 3, 5, and 7 years. Concurrent blood Pb level above
the median level (>7.2 (ig/dL) was associated with a larger decrease in IQ, regardless of whether prenatal
blood Pb levels were low or high (less than or greater than the median of 11.4 (ig/dL, respectively).
Pooling the Cincinnati and Rochester cohorts (n = 397), Hornung et al. (2009) also created a new
indicator of Pb exposure: the ratio of blood Pb level at 6 years of age to that at 2 years of age. The greatest
decrease in cognitive and behavioral development was observed for children with blood Pb ratios greater
than 1 (indicating an increase in blood Pb level from 2 to 6 years of age) (Figures 5-10 and 5-11 and Table
5-6). Presumably areas under the curve would be similar among children with blood Pb level ratios of 1,
greater than 1, and less than 1, indicating that cumulative blood Pb levels would not be predictive. It is
important to note that in these aforementioned studies, blood Pb levels were higher than those currently
measured in among children in the U.S. Additionally, children in these study populations experienced
larger decreases in blood Pb levels over time. It is unclear whether these findings would apply to children
Blood Pb levels 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.
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10
11
12
13
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15
16
18
6-yea r:2-year
ratio = 0.5
IQ = 89.0
6-year:2-year
ratio = 1.25
IQ = 83.7
16
14
12
OJ 10
~
- - -a- -
6-year:2-year
ratio = 1.0
IQ = 85.5
8
5 6
4
2
0
2
3
6
4
5
Age (years)
Source: Hornung et al. (Hornuna et al.. 2009)
Note: All three patterns have an identical mean blood Pb level level of 10 [jg/dL.
Figure 5-11. Estimated IQ in combined Cincinnati and Rochester cohorts
for 3 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 level (white squares).
In the collective body of epidemiologic evidence of children, it is difficult to ascertain which
lifestage of Pb exposure is associated with the greatest susceptibility to Pb-associated neurodevelopmental
effects. Associations have been observed with prenatal, early-childhood, lifetime average, and concurrent
blood Pb levels as well as childhood tooth Pb levels. The assessment of age-based susceptibility is
complicated further by the fact that blood Pb levels in children, although highly affected by recent dose,
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 cumulative dose (Section 4.3.5).
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 schoolage and into adolescence. Additional results from Hornung et al. (2009) and recent
studies described earlier in the section support the conclusion from the 2006 AQCD that concurrent blood
Pb level appears to be the best predictor of neurodevelopmental effects in children. These findings are
consistent with the understanding that the nervous system continues to develop throughout childhood.
Thus, the course of cognitive development may be modified in children, depending on concurrent blood
Pb levels or positive caregiving environment.
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Blood Lead (|Jg/dl)
Behavioral; Neonate; Rat; Female (1
Behavioral; Neonate; Rat; Male (2
Cognition; Neonate; Rat; Both (3
Cognition; Adult; Monkey; Both (4
Cognition; Neonate; Rat; Both (5
Cognition; Adult; Rat; Male (6
Cognition; Neonate; Mouse; Both (7
Cognition; Neonate; Rat; Both (8
Morphology; Neonate; Rat; Both (9
Morphology; Adult; Monkey; Female (10
Morphology; Adult; Rat; Male (11
Morphology; Neonate; Mouse; Both (7
Morphology; Neonate; Rat; Both (8
Morphology; Adult; Rat; Male (6
Motor function; Neonate; Mouse; Male (12
Neurotransmitter; Neonate; Rat; Both (13
Neurotransmitter; Neonate; Mouse; Both (14
Neurotransmitter; Neonate; Mouse; Both (12
Neurotransmitter; Adult; Rat; Male (15
Neurotransmitter; Adult; Rat; Female (15
Neurotransmitter; Neonate; Rat; Female (16
Oxidative Stress; Adult; Monkey; Female (10
Physical Development; Adult; Rat; Female (9
Physical Development; Neonate; Rat; Female (1
Physical Development; Neonate; Rat; Male (2
Physical Development; Adult; Rat; Male (6
Stress; Neonate; Rat; Female (17
Stress; Neonate; Rat; Female (16
Stress; Adult; Rat; Male (18
Stress; Neonate; Rat; Male (16
Stress; Adult; Rat; Male (15
Figure 5-12. Neurological summary array of toxicological outcomes after
Pb exposure. Dosimetric representation reported by blood Pb
level. (ID corresponds to Table 5-7.)
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Table 5-7. Summary of findings from neurotoxicological exposure-response array
presented in Figure 5-12.
Study ID
Reference
Blood Pb Level
(ug/dL)
Outcome
1
Beaudin et al. (2007)
13 & 31
Behavior, neonate: Lactational Pb exposure, offspring deficient in Reward Omission testing.
2
Kishi et al. (1983)
59 & 186
Behavior, neonate: Pb exposure (oral gavage of pups) during lactational period, Changed emotional
behavior, males.
3
Stangle et al. (2007)
13 & 31
Cognition; Developmental Pb exposure (PND1-30): Impaired learning with visual discrimination task,
heightened response to errors.
6
Gong & Evans (1997)
38 & 85
Cognition-Adult male 21 day Pb exposure: Hyperactivity with Habituation to new cage environment.
4
Rice (1990)
32 & 36
Cognition-Chronic Pb exposure from birth: Spatial discrimination reversal task impairment.
7
Li et al. (2009)
80S 100
Cognition-Gestational & lactational Pb exposure: Morris water maze performance impaired.
8
Li et al. (2010)
80S 102
Cognition- Gestational & lactational Pb exposure: Morris water maze performance impaired.
5
Overmann (1977)
33, 174 & 226
Cognition-Pb exposure (oral gavage of pups) during lactation: Response inhibition impaired.
17
Cory-Slechta et al. (2010)
10 & 13
Stress: Corticosterone-Lifetime Pb plus stress: Affects Fl performance, dopamine and serotonin levels in
female offspring.
16
Virgolini, Rossi-George,
Weston et al. (2008)
25
Stress: Corticosterone-Maternal Pb plus stress: Affects Fl performance.
16
Virgolini, Rossi-George,
Weston et al. (2008)
19 & 35
Stress.
18
Virgolini et al. (2005)
15 & 27
Stress: Corticosterone-Chronic Pb plus stress: Affects neurotransmitters & Fl performance.
15
Virgolini, Rossi-George,
Lisek et al. (2008)
11 &/or 31
Stress: Corticosterone-Maternal Pb plus stress: Affects Fl performance, dopamine, serotonin, and NE
levels.
9
Huetal. (2008)
4 & 12
Morphology; Gestational Pb exposure: Neurite outgrowth marker PSA-NCAM decreased in rat pups.
10
Wu et al. (2008)
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).
11
Tavakoli-Nezhad etal.
(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.
7
Li et al. (2009)
40 & 100
Morphology; Gestational & lactational Pb exposure: Increased levels of inflammatory cytokines &
exocytosis related proteins in brains of pups at weaning.
8
Li et al. (2010)
80S 102
Morphology: Increased levels of Alzheimer disease-associated proteins in mice with gestational and
lactational Pb exposure.
6
Gong & Evans (1997)
85
Morphology; 21 day Pb exposure to adult males: Marker of neuronal injury-elevated hippocampal glial
fibrillary acidic protein (GFAP).
12
Leasure et al. (2008)
10 & 42
Motor function; Mouse maternal (dam) Pb exposure: Induced decreased rotarod performance in offspring
(1 year-old male offspring).
13
Bielarczyk et al. (1996)
1.8,3.8, 22
Neurotransmitter; Perinatal (GD16-PND28) Pb exposure: Decreased hippocampal ChAT activity and
increased hippocampal tyrosine hydroxylase activity.
14
Fortune & Lurie (2009)
8 & 43
Neurotransmitter; Mouse maternal (dam) Pb exposure: Affects offspring superior olivary complex
(auditory) neurotransmitters.
12
Leasure et al. (2008)
10 & 42
Neurotransmitter; Mouse maternal (dam) Pb exposure: Affects 1 year old male offspring dopamine
homeostasis.
15
Virgolini, Rossi-George,
Lisek et al. (2008)
31
Neurotransmitter; Gestational and lactational Pb exposure: Induced NE aberrations in adult rat offspring
(both sexes).
16
Virgolini, Rossi-George,
Weston et al. (2008)
19 & 30
Neurotransmitter; Gestational and lactational Pb exposure: Induced DA and 5HT changes in rat offspring.
10
Wu et al. (2008)
19 & 26
Oxidative stress: Elevated oxidative DNA damage in aged brains of female monkeys (exposed to Pb as
infants).
9
Hu et al. (2008)
15
Physical developmen; t-Gestational Pb exposure: Early brain synapse development impaired
(hippocampal PSA-NCAM and sialytransferase).
1
Beaudin et al. (2007)
13 & 31
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
Kishi et al. (1983)
59 & 186
Physical development; Pb exposure during lactation (oral gavage): Delayed development of righting reflex
in male rats.
6
Gong & Evans (1997)
85
Physical development; Adult male rats (21 day Pb exposure): Neurotoxicity measured with brain glial
fibrillary acidic protein (GFAP).
16
Virgolini, Rossi-George,
Weston et al. (2008)
19 & 35
Stress: Corticosterone levels affected.
17
Cory-Slechta et al. (2010)
10 & 13
Stress: Corticosterone-neurotransmitter-Lifetime Pb exposure in female rats plus stress: Dopamine
homeostasis affected.
5.3.2.2. Toxicological Studies of Neurocognition, Memory and Learning
The 2006 AQCD reported deficits in the Morris water maze with Pb exposure. The Morris water
maze tests memory and learning by having a mouse swim and locate or remember the location of a
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platform submerged in opaque water. New research since the 2006 AQCD continues to show Pb-induced
impaired Morris water maze performance. Data on neurocognition and learning as well as other
neurotoxicological endpoints with dose responsive data are shown in Figure 5-12 and accompanying
Table 5-7. Dams received Pb acetate dissolved in drinking water (0.1%, 0.5%, and 1% with corresponding
blood Pb level of 4, 8 and 10 (ig/dL at postnatal day [PND] 21) throughout gestation and lactation.
Beginning at weaning, Pb exposed pups were subjected to Morris water maze performance testing. Pb
exposed pups had 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 (ig/dL, respectively, indicating impaired
memory and learning (Li et al.. 2009). The pups in Li et al. (2009) were not separated by sex. Another
study found that dietary supplementation with various supplements listed below or with
methioninecholine concomitant with Pb exposure in weanling males shortened the escape latency of Pb-
exposed pups to more closely resemble the escape latency of control pups (G. Fan et al.. 2009; Fan et al..
2010). Zinc or methionine were effective dietary supplements in the G. Fan et al. 2009 study (2009);
glycine, taurine, vitamin C, vitamin Bl, tyrosine had no effect on the Morris water maze results. These
data on the effect of Pb on learning and memory in the Morris water maze confirm findings by two other
labs. Jett et al. (1997) showed increased escape latency in adult rats exposed to Pb via direct injections
into the dorsal hippocampus. Kuhlmann et al. (1997) 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. This new study
confirms the findings of earlier studies that learning and memory are significantly impaired in rodents
who are exposed to Pb early in life.
Working Memory
Working memory is the ability to temporarily keep information in mind while using the
information to perform a related or unrelated task. The Morris water maze is able to measure working
memory in addition to learning. Using this test, the 2006 AQCD found working memory was significantly
affected in chronic developmentally exposed (Pb Acetate in feed 10 days prior to mating through PND
21) female offspring at PND 21 (Jett. Kuhlmann. Farmer, et al.. 1997). Delayed spatial alternation (DSA)
is another test used to measure working memory. With DSA, an animal receives rewards based on
responses at two separate levers. Work from the 2006 AQCD showed that Pb-exposed animals had
deficits under DSA testing. These deficits included increased response errors, decreased percent of correct
responses, and perseverance at one lever (repeatedly pressing the same lever without moving between the
two locations). These results have been consistently shown with non-human primate studies (continuous
Pb exposure or juvenile to adult exposure) and less consistently shown with rats (juvenile only or juvenile
to adult exposure). Working memory is a subcategory of executive function or goal-oriented problem
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solving. Deficits in working memory are thought to underlie associations between blood Pb levels and
ADHD in humans.
Response Inhibition
Response inhibition is another measure of executive function and is measured with multiple tests
that measure premature responses, decreased pause time between two scheduled events and increased
perseverance. These tests include Differential Reinforcement of Low Rates of Responding (DRL), DSA,
Fixed Interval testing (FI), FI with Extinction (FI-Ext), Fixed Ratio-FI (FR-FI), and Signal Detection with
Distraction. Multiple studies from the 2006 AQCD and earlier literature have shown that early life Pb
exposure contribute to response inhibition across the spectrum of these aforementioned tests. Monkeys
with moderate blood Pb levels (11-13 (ig/dL) learned the DRL task more slowly but eventually acquired
reinforcement rates equal to 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.
Learning Ability, Schedule-Controlled Behavior
The 2006 AQCD discussed learning or cognition as measured with schedule controlled-behaviors
including fixed interval (FI) and fixed ratio (FR) operant conditioning and found that FI response rate was
affected differentially with low level and high level Pb exposures increasing and decreasing FI response
rate in females, respectively. This curvilinear response has since been further explored in more recent
work, much of which also includes the effect of psychological stress on Pb exposure.
Learning Ability with Stress
The combined paradigm of Pb exposure and stress experienced by a person or 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. Data on stress and Pb with dose responsive data
endpoints are shown in Figure 5-12 and accompanying Table 5-7. Cory-Slechta and colleagues have
conducted multiple investigations in this area. Most recently, they have shown enhanced learning deficits
in female rats (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 to mating through lactation and pups from a mixed sex litter received drinking water Pb (50
ppm) exposure through the remainder of their lifetime resulting in blood Pb levels of dams and pups
ranging from 5-13 (ig/dL.
Pb plus stress-related outcomes were followed 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:
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50 or 150 ppb Pb acetate drinking water solutions) (Vimolini. Rossi-George. Lisek. et al.. 2008). Dams
underwent restraint stress at gestational 16-17. Marked increases in response rates on FI performance in
the Pb-stress female offspring versus control was found in animals 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, this effect was potentiated in the animals exposed to Pb plus additional stress.
Similarly, lifetime Pb exposure (50 or 150 ppm, blood Pb level 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, another potentiated effect (Rossi-George et al.. 2011) (Table 5-8).
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 start responding too early due to a decreased pause or
PRP interval. Abberant FI performance in infants and children has been used as a marker for impulsivity.
Separately, overall FI response rate was significantly increased in Pb exposure alone and with maternal or
offspring stress at the 50 ppm exposure dose. At 150 ppm, 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 striatial 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 confirm learning deficits and provide possible mechanistic explanations.
Pb exposure over various exposure windows has been shown to affect corticosterone levels in
rodents. 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 seen in this model (Corv-Slechta et al..
2004). By 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) that were enhanced with maternal stress. Postnatal exposure of male rodents to Pb
(PND 21-5 months of age) showed 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 permanent changes in the HPA axis
(corticosterone levels) in both sexes which are dynamic as the animal ages.
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Female
Male
900-1
600-
O
c
300-
Basal
Final
PbxS
900-1
Basal
Final
600-
300-
S:PS
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Specifically Pb and Pb plus maternal stress initially reduced the ability of DEX to suppress
corticosterone. With time, the effect of this DEX test in males induced prolonged corticosterone
suppression or failure to return to baseline as seen in control animals. In summary, dam Pb exposure
induced negative feedback hypofiinction in both sexes with an inverse U dose response function. A
similar recent study explored the effect of lifetime Pb exposure on the HPA axis, looking at basal
corticosterone levels in male and female offspring at two time periods (2 months old and 10 months old;
before and after behavioral testing, respectively) in adulthood (Rossi-George et al.. 2011). Pb and stress
have no effect on basal corticosterone levels in males at either time period (Figure 5-13). At the first time
period, Pb exposure elevated basal corticosterone levels in a dose-dependent fashion in females, and Pb
plus stress attenuated the PB-dependent elevations in corticosterone to baseline levels (Figure 5-13). At
the second time period, Pb and stress accelerated the age-dependent decrease in basal corticosterone
levels in females (Figure 5-13).
These two studies of lifetime exposure (Rossi-George et al. 2009) reported different basal stress
hormone levels with Pb exposure. Males with lifetime Pb exposure had no significant corticosterone
response to Pb exposure; whereas males with dam Pb exposure had significant decreases in corticosterone
at 5 months of age in the 50 ppm exposure group only (not seen in 150 ppm Pb exposure group). On the
other hand, females had dose-dependent corticosterone responses to Pb exposure in both exposure models
(lifetime Pb exposure and dam Pb exposure.
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Restraint
Overall Rale
j diri
0 $0 ISO 0 50 150 0 SO ISO
!u» Rate +
#-| *" IM-i 120-1 •
0 SO ISO
PoKtretnforcement Pause Time
0 -i 3001 128-1
L L i
0 50 150 0 SO 150 0 SO 150
(Pb (ppm)] [Pb (ppm)] [Pb (ppm)]
Source: Used with permission from Elsevier Science, Virgolini, Rossi-George, Weston, et al. (2008).
Figure 5-14. Changes in Fl performance (Fl overall performance, run rate,
PRP) in female offspring with maternal Pb exposure with
various stressors in adulthood (restraint, cold, novelty).
trail Rate
-j 1S0- .120-
Iri M ll
0 SO 160 0 50 150 0 50 ISO
iRate
ISO-. +¦ 120-
y m n
0 SO 150 0 50 150 0 60 150
0 60 150
Postreinforcement Pause Time
»-| 160-1 160 -j *
ii li t ¦
0 50 150 0 SO 150 0 SO 150
[Pb (ppm)) [Pb (ppm)] |Pb (ppm)]
Source: Used with permission from Elsevier Science, Virgolini, Rossi-George, Weston, et al. (2008).
Figure 5-15. Changes in Fl performance (Fl overall performance, run rate,
PRP) in male offspring with maternal Pb exposure with various
stressors in adulthood (restraint, cold, novelty).
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Maternal stress alone also led to HPA axis negative feedback hypofiinction. 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 supression in males. These data
together show that HPA axis alterations could be an underlying mechanism linking commonalities
between the contribution of Pb and stress to adverse health outcomes.
Schedule-control 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 who 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. (2008) 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 who were not behaviorally trained 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 (Figure 5-13). 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 demonstrated a U shaped dose response corticosterone curve with 50Pb significantly less
than control or 150Pb. 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). When drawing conclusions about the effects of Pb exposure, these data demonstrate that
behaviorally trained animals have an altered HPA axis and response to Pb exposure versus animals who
are housed under conditions without daily handling by caregivers.
Another study looked at female rats with lifetime Pb exposure combined with prenatal stress and
found enhanced learning deficits(drinking water 50 ppm Pb acetate, offspring blood Pb 7-13 jj.g/dl)
(Corv-Slechta et al.. 2010). Learning was evaluated with multiple schedule of repeated learning (RL) and
performance testing. Repeated learing was impaired but performance was not affected with Pb exposure.
The impaired RL was further enhanced with prenatal stress. There were significant associations between
Pb/stress and coticosterone 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 significant decreases in hippocampal nerve
growth factor (NGF) versus control. Thus, this study demonstrates that lifetime Pb exposure with or
without prenatal stress induced learning deficits in female mice. In a similar study, the authors proposed
that associations of Pb and stress with learning deficits (Fi testing in females) may be related to
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aberrations in corticosterone and dopamine (Rossi-George et al. 2011). Earlier work has shown that dam
or prenatal stress (PS) affects the HPA axis of the offspring. A newer study was conducted to determine
the influence of low level dam Pb exposure and prenatal stress on offspring stress challenge responsivity
(intermittent stress as an adult) (Rossi-George et al.. 2011). In a similar study, the authors proposed that
associations of Pb and stress with learning deficits (F1 testing in females) may be related to aberrations in
corticosterone and dopamine (Rossi-George et al.. 2011). Dam Pb exposure (50 or 150 ppm Pb acetate)
followed by intermittent stressors (cold, novelty or restraint) to offspring as adults induced significant
changes in FI response rate. Females were more sensitive to the adult intermittent stressors at the higher
dose of Pb (150 ppm) with significant increases in FI response rate and decreased PRP, i.e. increased
impulsivity (Figure 5-14). 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 5-15).
Corticosterone levels were followed in this study and showed dose dependent correlations with FI
outcomes in females but were independent of dose in males.
Table 5-8. Summary of effects of maternal and lifetime Pb exposure on FI performance
water3.
Pb (ppm) Maternal Pb
Lifetime Pbu
Overall rate
PRP
Overall rate
PRP
0 ppm
0-PS
NO SIGNIFICANT EFFECT" NO SIGNIFICANT EFFECT NO SIGNIFICANT EFFECT NO SIGNIFICANT EFFECT
0-OS
NO SIGNIFICANT EFFECT
*|-23%
NO SIGNIFICANT EFFECT NO SIGNIFICANT EFFECT
50 ppm
50-NS
NO SIGNIFICANT EFFECT NO SIGNIFICANT EFFECT
* j* 95%
NO SIGNIFICANT EFFECT
50-PS
NO SIGNIFICANT EFFECT NO SIGNIFICANT EFFECT
*|79.2%
*|-42%
50-OS
* j* 64.9%
NO SIGNIFICANT EFFECT
*|74.7%
* J,-39.3%
150 ppm
150-NS
* j* 42.4%
* J,-30.3%
NO SIGNIFICANT EFFECT NO SIGNIFICANT EFFECT
150-PS
NO SIGNIFICANT EFFECT
*|-25.7%
* j* 90.7%
*|-44.7%
150-OS
*|59.2%
NO SIGNIFICANT EFFECT
*|78.5%
NO SIGNIFICANT EFFECT
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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 life time
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; T, represents increase; I, represents decrease.
bData from Virgolini et al. (2005"). *Denotes significant effect versus Oppm control (p<0.05).
°Data from current study. *Denotes significant effect versus Oppm control (p<0.05).
Source: Used with permission from Elsevier Science, Rossi-George et al. (2011") (Table 1).
Cognitive Flexibility
Cognitive flexibility is the ability to reallocate mental resources when situations
change (Monsell. 2003). Discrimination reversal learning is used to measure cognitive
flexibility, which is a subclass of executive function. The 2006 AQCD reported reversal
learning deficits in monkeys with blood Pb of 11-20 |J.g/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 & Strupp. 1997). Interestingly, recent
work has shown that NMDA receptors and D2-like receptors, two known targets of Pb, are
involved in discrimination reversal learning (Herold. 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 AQCD reported
monkeys with cognitive flexibility impairment under RI-RI (Newland et al.. 1994).
Selective Attention
Few animal toxicology studies measure selective attention. Those that do employ signal detection
with distraction, a test looking for increased omissions after exposure to an external distraction. The
newest publication in this area showed no effect with this test after juvenile through adolescence exposure
in female rats (Staimlc et al.. 2007). The two dose groups yielded blood Pb levels of 13 jj.g/dl and 3 lug/dl
(Stannic et al.. 2007).
5.3.2.3. Toxicological Studies on the Effects of Chelation
Earlier work in the animal toxicology 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 & Evans. 1997) and forced-swim immobility (P. W. Stewart et al.. 1996). A more
recent study looked at the effect of succimer treatment on various neurobehaviroal and neurocognitive
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-
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Pb, m-Pb+succimer, and h-Pb+succimer are 1.5, 12.6, 31, 2.8, and 8.5 (ig/dL, respectively; brain Pb levels
at the same time for the same groups are 41, 1040, 3690, 196, and 1370 ng/g dry weight, respectively.
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 and did not improve the
impaired inhibitory control (Stannic et al.. 2007). These are important findings because they provide
evidence that certain adverse neurobehavioral or cognitive outcomes associated with Pb exposure appear
to be reversible with chelation therapy.
A 3-week course of Pb acetate (PND 1-17, dam drinking water) plus or minus succimer/chelator
(PND 31-52) treatment was performed 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 therapy had normalized reactivity to reward
omission and errors (Beaudin et al.. 2007). Adverse Pb-induced behavioral outcomes were attenuated with
chelation therapy in Pb-exposed animals.
Another study (G. Fan et al.. 2009) looked at methionine choline supplementation in Pb exposed
animals to understand its effect on Pb disposition in various tissues (blood, bone, brain) and how this
might contribute 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 & Blusztain. 1994). In this model, methionine choline attenuated Pb-dependent memory
and learning defecits (Section 5.3.2.2). Exposure of weanling male rats to Pb acetate in drinking water
(300 mg/L) out to PND60 produced a blood Pb level of 60 (ig/dL, bone Pb of 165 jj.g/g and brain Pb of
0.63 jj.g/g. Methionine choline supplementation significantly attenuated blood Pb and bone Pb but
produced a non-significant attenuation of brain Pb (0.51 jj.g/g) in mice 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 are neuroprotective in the ALS
mouse neurodegenerative model or Tg(SODl-G93A) (Petri et al.. 2007).
In summary, succimer or chelation therapy appears to be able to restore Pb-dependent impairments
of learning and arousal as well as being neuroprotective in a dose dependent fashion. In these studies
succimer use was most efficacious at lower doses of Pb exposure. Chelation does not restore Pb-
dependent impaired inhibitory control. Chelation with the supplement methionine choline affected the
disposition of Pb in various tissues, significantly attenuating blood and bone Pb levels and non-
significantly attenuating brain Pb.
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5.3.2.4. Epidemiologic Studies of Cognitive Function in Adults
Adults without Occupational Lead Exposures
The 2006 AQCD cited more consistent associations of bone Pb levels but not concurrent blood Pb
levels with cognitive performance in environmentally-exposed adults (U.S. EPA. 2006). Studies published
since the 2006 Pb AQCD continue to support the previous findings. Several studies analyzed data from
NHANES III (1991-1994) and investigated effect modification by age and genetic variants. Krieg et al.
(2009) overall did not find blood Pb level to be associated consistently with poorer performance on
cognitive testing among 2,090 adults 20-59 years of age nor among 1,796 adults 60 years of age and
older. There were also few statistically significant differences in the association with blood Pb between
ALAD genotypes (E. F. Krieg. Jr. et al.. 2009). Among the 20-59 year-old adults, a borderline significant
difference was observed for mean reaction time, however, among subjects in the CC and CG ALAD
genotype groups combined (ALAD2 carriers in the terminology above) (n=161), reaction time improved
(i.e., faster reaction time) with increasing blood Pb level. In contrast, ALAD2 subjects had a greater
increase in the number of errors on a symbol-digit substitution task (p = 0.07) in association with
increasing blood Pb level. In the same study, contrasting observations were made in adolescent NHANES
participants. Effect modification by ALAD2 is not entirely clear as it may increase susceptibility to Pb-
associated health effects by increasing blood Pb levels or diminish Pb bioavailability by maintaining it in
a sequestered state in the bloodstream. Krieg et al. (2010) also found differences in the association
between blood Pb level and scores on a symbol-digit substitution test by the VDR variants, rs731236 and
VDRrs2239185, as well as VDRhaplotype. Similar to observations in children (Section 5.3.2.1), results
were not consistent 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.
Shih et al. (2006) studied participants in the Baltimore Memory Study (BMS), a longitudinal study
of men and women 50-70 years of age residing in Baltimore, MD with a mean (SD) blood Pb level of
3.46 (2.23) (ig/dL. A total 1,140 out of 2,351 (48.5%) subjects participated from neighborhoods that
represented a diversity of race and SES. Of these, 991 had complete data for blood and tibia bone Pb,
cognitive testing, and covariates. After excluding the three participants with the highest bone Pb values,
the two participants with the two most negative bone Pb values, and one person with a blood Pb level 10
standard deviations greater than the mean, the analytic sample was 985. Scores on individual cognitive
tests from a battery of 20 in-person administered tests were grouped into different cognitive domains
(language, processing speed, eye-hand coordination, executive function, verbal memory and learning,
visual memory, and visuoconstruction) by transforming individual test scores into z-scores and averaging.
Negative associations were observed more consistently for tibia Pb levels than for blood Pb levels (Figure
5-16 and Table 5-9). Tibia bone Pb levels were associated with worse performance on tests in all domains
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1
2
3
4
5
6
7
8
9
10
11
12
in models adjusted for age, sex, testing technician, and presence of the apolipoprotein (APO)E-s4 allele.
The magnitudes and statistical significance of associations were attenuated with additional adjustment for
education, race, and wealth. In thesefully-adjusted models, all domains except language and processing
speed were negatively associated with tibia bone Pb, but only the association with visuoconstruction was
borderline statistically significant. Analysis of a quadratic term for tibia Pb indicated no evidence of
nonlinearity, thus results for linear models were presented. In linear models, visuoconstruction scores
decreased by 0.0044 SDs (95% CI: -0.0091, 0.0003) per 1 jxg/g bone increase in tibia Pb level. The mean
(SD) tibia Pb in this group was 18.7 (11.2) jxg/g bone. Of particular note in this study is that it was the
first such study with a large proportion of African-Americans (n=395). In a subsequent analysis,
increasing tibia Pb levels were associated with a greater decrease in cognitive performance among
subjects living in neighborhoods with a greater number psychosocial hazards (Glass et al.. 2009) (Figure
5-16 and Table 5-9).
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Data (Mg/dL)
Outcome
Subgroup/Bone type
Blood Pb
Krieg and Butler (2009")
Krieg et al. (2009*1
Krieg etal. (2010)
Shihetal. (2006)a
Rajan et al. (2008")
Weuve et al. (2009")
Gao et al. (2008")
2.88 (6.91)
2.85 (7.31)
4.02 (3.56)
2.85 (7.32)
Symbol digit substitution
Serial digit learning
Symbol digit substitution
Serial digit learning
Word recall
Story recall
Symbol digit substitution ^
Ages 20-39 yrs
Ages 40-59 yrs
Ages 20-39 yrs
Ages 40-59 yrs
Ages 20-59 yrs, ALAD GG
Ages 20-59 yrs, ALAD CC/CG
Ages 20-59 yrs, ALAD GG
Ages 20-59 yrs, ALAD CC/CG
Ages £60 yrs, ALAD GG
Ages 260 yrs, ALAD CC/CG
Ages 260 yrs, ALAD GG
Ages 260 yrs, ALAD CC/CG
Ages 20-59 yrs, VDR hap CC
Ages 20-59 yrs, VDR hap CT
Ages 20-59 yrs, VDR hap TC
Ages 20-59 yrs, VDR hap TT
3.5 (2.2) Language
Eye-hand coordination
Executive functioning
Visuoconstruction
5.3 (2.9), 4.8 (2.7)b Visuospatial construction
Executive functioning
Verbal memory
Perceptual speed
2.9 (1.9)
Composite cognitive score
Composite, except letter
0.39 (0.63) (plasma) Composite cognitive score
Bone Pb
Shihetal. (2006)a
18.7 (11.2)
Language
Eye-hand coordination
Executive functioning
Visuoconstruction
Bandeen-Roche et al. (2009) 18.8 (11.(
Glass et al. (20091
Rajan et al. (2008")
Weuve et al. (2009")
18.8 (11.1)
21.9 (13.8), 21.2
29.3 (19.1), 27.9
10.5 (9.7)
12.6 (11.6)
Eye-hand coordination
African-Americans, s 15
African-Americans, >15
White, <15
White, > 15
Verbal memory
African-Americans, s 15
African-Americans, >15
White, <15
White, > 15
Language
Middle NPHd
High NPH"
Eye-hand coordination
Middle NPH11
High NPHd
Visuospatial construction
Tibia
Executive functioning
Tibia
Verbal memory
Tibia
Perceptual speed
Tibia
Visuospatial construction
Patella
Executive functioning
Patella
Verbal memory
Patella
Perceptual speed
Patella
Composite cognitive score
Tibia
Composite, except letter
Tibia
Composite cognitive score
Patella
Composite, except letter
Patella
-O
-0.6 -0.5 -0.4 -0.3 -0,2 -0.1 0
Change in Cognitive Score (95% CI)
0.1
0.2
Note: Black diamonds = blood Pb, blue circles = tibia Pb, white circles = patella Pb. ALAD = aminolevulinate dehydratase, VDR = vitamin D receptor, NPH = neighborhood
psychosocial hazard. sEffect estimates for Model B are presented. bBlood Pb levels represent levels in ALAD wildtype and ALAD2 carriers, respectively. cEffect estimates for the
following strata: African-Americans with tibia Pb levels s 15 |jg/dL, African-Americans with tibia Pb levels >15 |jg/dL, whites with tibia Pb levels s 15 |jg/dL, and whites with tibia
Pb levels >15 [jg/dL. dEffect estimates for the interaction between tibia Pb levels and NPH fertile, with the lowest NPH fertile serving as the reference group. ®Tibia Pb levels
represent levels in ALAD wildtypes and ALAD2 carriers, respectively. 'Patella Pb levels represent levels in ALAD wildtypes and ALAD2 carriers, respectively.
Figure 5-16. Associations of blood and bone Pb levels with cognitive
function among adults without occupational exposures to Pb.
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Table 5-9. Additional characteristics and quantitative results for studies presented in
Figure 5-16
. Population/
Study Location
Pb
Biomarker Statistical Analysis Cognitive Test
Data
Subgroup/
Model
Effect Estimate
(95% Cl)a
Blood Pb Studies
Krieg and 2823 adults, Blood mean
Butler ages 20-59 yr, (SD): 2.88
(2009) U.S. NHANES (6.91) ^ig/dL
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
Symbol Digit Substitution
(mean total latency, sec)
Serial digit learning score
Assessed using the
Neurobehavioral
Evaluation System 2
Ages 20-39 yr
Ages 40-59 yr
Ages 20-39 yr
Ages 40-59 yr
-0.014 (-0.061, 0.032)b
-0.042 (-0.087,
0.0009)b
-0.017 (-0.067, 0.034)b
0.058 (-0.028, 0.143)b
Krieg et al. 2090 adults,
(2009) ages 20-59 yr
1976 adults,
ages >60 yr
U.S. NHANES
III (1991-1994)
Blood mean
(SD):
20-59 yr: 2.85
(7.31) |jg/dL
>60 yr: 4.02
(3.56) |jg/dL
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)
Symbol Digit Substitution
(mean total latency, sec)
Serial digit learning score
Word recall
Story recall
Assessed using the
Neurobehavioral
Evaluation System 2
Ages 20-59 yr
ALADGG
ALAD CC/CG
Ages 20-59 yr
ALADGG
ALAD CC/CG
Ages >60 yr
ALADGG
ALAD CC/CG
Ages >60 yr
ALADGG
ALAD CC/CG
-0.018 (-0.049, 0.013)b
-0.072 (-0.153, 0.008)b
-0.003 (-0.072, 0.067)b
0.056 (-0.088, 0.201 )b
-0.021 (-0.08, 0.034)
0.007 (-0.114, 0.129)
0.024 (-0.028, 0.077)
-0.131 (-0.301,0.039)
Krieg et al.
(2010)
2093 adults,
ages 20-59 yr
1799 adults,
ages >60 yr
U.S. NHANES
III (1991-1994)
Blood mean
(SD):
20-59 yr: 2.85
(7.32) |jg/dL
>60 yr: 4.02
(3.39) |jg/dL
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)
Symbol Digit Substitution
(mean total latency, sec)
Serial digit learning score
(# errors, pos is bad)
Word recall
Story recall
Assessed using the
Neurobehavioral
Evaluation System 2
Ages 20-59 yr
VDR haplotype CC
VDR haplotype CT
VDR haplotype TC
VDR haplotype TT
Ages 20-59 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
Ages >60 yr
VDR haplotype CC
VDR haplotype CT
VDR haplotype TC
VDR haplotype TT
-0.535 (-1.18, 0.107)b
0.019 (-0.038, 0.077)b
-0.069 (-0.140, 0.002)b
-0.095 (-0.192, 0.001)b
-0.346 (-0.665, 0.026)b
-0.044 (-0.126, 0.039)b
0.061 (0.074, 0.135)b
0.006 (-0.107, 0.119)b
-0.226 (-0.536, 0.087)
-0.028 (-0.118, 0.067)
-0.010 (-0.139, 0.115)
-0.028 (-0.264, 0.209)
0.043 (-0.494, 0.581)
0.001 (-0.057, 0.060)
0.010 (-0.095, 0.116)
-0.049 (-0.128, 0.064)
Shihetal. 985 adults, Blood mean
(2006) mean age: 59 (SD): 3.5
yr (2.2) [igldl
Baltimore
Memory Study,
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
Language
Eye-hand coordination
Executive functioning
Visuoconstruction
Assessed using Raven's
Colored Progressive
Matrices
Model A
Model B
Model A
Model B
Model A
Model B
Model A
Model B
-0.0060
-0.0019
-0.0110
-0.0076
-0.0143
-0.0101
-0.0191
-0.0143
(-0.029,
(-0.020,
(-0.032,
(-0.026,
(-0.034,
(-0.027,
(-0.046,
(-0.038,
0.017)
0.016)
0.01)
0.011)
0.005)
0.007)
0.008)
0.009)
Rajan et
al. (2008)
720 males,
ages >45 yr
Normative
Aging Study,
Boston, MA
Blood mean
(SD):
5.3 (2.9)
|jg/dL (ALAD
wildtype)
4.8 (2.7)
|jg/dL
(ALAD2
carriers)
Linear regression adjusted
for blood Pb main effect,
ALAD genotype, age at
cognitive test, education,
alcohol consumption,
cumulative smoking, English
as first language
Visuospatial,
constructional praxis
Executive function verbal
fluency
Verbal memory, word
recall
Perceptual speed, mean
latency
Assessed using CERAD,
Neurobehavioral
Evaluation Sysem, WIAS-
R
-0.017 (-0.077, 0.043)c
-0.01 (-0.073, 0.053)c
0.001 (-0.06, 0.062)c
-0.06 (-0.14, 0.02)c
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Pb
study Location
K" Biomarker Statistical Analysis Cognitive Test
Data
Subgroup/
Model
Effect Estimate
(95% Cl)a
Weuve et 587 females, Blood mean Generalized estimating
al. (2009) ages 47-74 yr (SD): 2.9 equations adjusted for age,
Nurses' Health (1.9) |jg/dL age-squared at Pb
Study, Boston, assessment, age at cognitive
MA assessment, education,
husband's education, alcohol
consumption, smoking
status, physical activity,
aspirin use, ibuprofen use,
use of Vitamin E
supplements, menopausal
status and postmenopausal
hormone use
Composite cognitive score
Composite except letter
fluency
Assessed using
Telephone Interview for
Cognitive Status and East
Boston Memory Test
-0.008 (-0.036, 0.020)
0.008 (-0.037, 0.054)
Gao etal.
(2008)
188 adults,
mean age 69.2
yr
Sichuan and
Shandong
Provinces,
China
Plasma mean ANCOVA adjusted for age,
(SD): 0.39 sex, education, BMI, APOE
(0.63) |jg/dL e4
Composite cognitive score
Assessed using CERAD,
CSID, III story recall,
Animal fluency test, III
token test
-0.009 (-0.025, 0.007)
Excluded From Figure Due Insufficient Data To Standardize Test Scores
Weuve et
al. (2006)
720 males,
ages >45 yr
Normative
Aging Study,
Boston, MA
Blood mean
(range): 5.2
(£1-28) |jg/dL
Linear mixed effects
regression adjusted for
smoking status, alcohol
consumption, calorie
adjusted calcium intake,
regular energy expenditure
on leisure time physical
activity, diabetes
MMSE score
ALAD wildtypes
ALAD2 carriers
-0.013 (-0.053, 0.027)
-0.087 (-0.180, 0.007)
Bone Pb Studies
Shih et al.,
(Shih et
ajj
985 adults,
mean age: 59
yr
Baltimore
Memory Study,
Baltimore, MD
Tibia mean Linear regression adjusted Language
(SD): 18.7 for:
(11.2) |jg/g Model A: age, sex,
technician, presence of
APOE-e4 allele
Model B: Model I, years of
education, race/ethnicity,
wealth
Eye-hand coordination
Executive functioning
Visuoconstruction
Assessed using Raven's
Colored Progressive
Matrices
Model A -0.08 (-0.13,-0.04)
Model B 0.006 (-0.03, 0.04)
Model A -0.08 (-0.12,-0.04)
Model B -0.03 (-0.06, 0.01)
Model A -0.08 (-0.11,-0.04)
Model B -0.014 (-0.05, 0.02)
Model A -0.022 (-0.17,-0.07)
Model B -0.044 (-0.09, 0.003)
Bandeen- 1140 adults, Tibia mean Marginal longitudinal linear
Roche et ages 50-70 yr (SD): 18.8 regression models adjusted
a I. (2009) Baltimore (11.6) ^jg/g forage, household wealth,
Memory Study education, race/ethnicity
cohort Demographic characteristics,
Baltimore, MD socioeconomic status,
race/ethnicity
Eye hand coordination
African-Americans
White
Verbal memory
African-Americans
White
Visual memory
African-Americans
White
Pb< 15 |jg/g
Pb >15 |jg/g
Pb< 15 |jg/g
Pb >15 |jg/g
Pb< 15 |jg/g
Pb >15 |jg/g
Pb< 15 |jg/g
Pb >15 |jg/g
Pb< 15 |jg/g
Pb >15 |jg/g
Pb< 15 |jg/g
Pb >15 |jg/g
-0.04 (-0.112, 0.03)
0.02 (-0.05, 0.10)
-0.02 (-0.06, 0.01)
0.04 (-0.002, 0.08)
-0.008 (-0.08, 0.06)
0.002 (-0.09, 0.09)
0.04 (-0.003, 0.09)
-0.07 (-0.13, -0.005)
0.001 (-0.007, 0.09)
-0.0009 (-0.011, 0.09)
0.003 (-0.003, 0.08)
-0.004 (-0.011, 0.03)
Glass et
al. (2009)
1001 adults,
mean age 59
yr
Baltimore
Memory Study,
Baltimore, MD
Tibia Pb Multilevel hierarchical
mean (SD): regression model adjusted
18.8 (11.1) forage, sex, race/ethnicity,
ug/g education, testing technician,
time of day
Language
Eye-hand coordination
Executive functioning
Visuoconstruction
Assessed using Raven's
Colored Progressive
Matrices
Middle NPH
High NPH
Middle NPH
High NPH
Middle NPH
High NPH
Middle NPH
High NPH
0.01 (-0.08, 0.09)d
-0.09 (-0.17, -0.001 )d
-0.04 (-0.12, 0.05)d
-0.06 (-0.15, 0.02)d
-0.02 (-0.10, 0.06)d
-0.10 (-0.18, -0.02)d
-0.003 (-0.14, 0.08)d
-0.006 (-0.17, 0.05)d
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Pb
study Location
K" Biomarker Statistical Analysis Cognitive Test
Data
Subgroup/
Model
Effect Estimate
(95% Cl)a
Rajan et
720 males,
Mean (SD):
Linear regression adjusted
Visuospatial, constructional
Tibia
-0.17 (-0.33, -0.007)
al. (2008)
ages >45 yr
Tibia: 21.9
for bone Pb main effect,
praxis
Patella
0.01 (-0.10,0.12)
Normative
(13.8) Mg/g
ALAD genotype, age at
Tibia
-0.07 (-0.23, 0.08)
Aging Study,
(ALAD
cognitive test, education,
Executive function verbal
Patella
-0.01 (-0.12, 0.10)
Boston, MA
wildtype),
alcohol consumption,
fluency
Tibia
0.05 (-0.10, 0.21)
21.2 (11.6)
cumulative smoking, English
Patella
0.07 (-0.40, 0.18)
Mg/g (ALAD2
as first language
Verbal memory, word recall
Tibia
-0.12 (-0.39, 0.06)
carriers)
Patella
-0.08 (-0.22, 0.06)
Patella: 29.3
Perceptual speed, mean
(19.1) Mg/g
latency
(ALAD
wildtype),
Assessed using CERAD,
27.9(17.3)
Neurobehavioral Evaluation
Mg/g (ALAD2
System, WIAS-R
carriers)
Weuve et Nurses' Health Tibia mean Generalized estimating
a I. (2009) Study cohort (SD): 10.5 equations adjusted forage,
587 subjects (9.7) |jg/g age-squared at Pb
Age range: 47- Patella mean assessment, age at cognitive
74 yr (SD):12.6 assessment, education,
All females (11.6) Mg/g husband's education, alcohol
Boston, MA consumption, smoking
status, physical activity,
aspirin use, ibuprofen use,
use of Vitamin E
supplements, menopausal
status and postmenopausal
hormone use
Composite cognitive score
Composite except letter
fluency
Assessed using Telephone
Interview for Cognitive
Status and East Boston
Memory Test
Tibia
Patella
Tibia
Patella
-0.04 (-0.09, 0.005)
-0.01 (-0.05, 0.03)
-0.05 (-0.10, -0.003)
-0.03 (-0.07, 0.01)
Excluded From Figure Due Insufficient Data To Standardize Test Scores Or Calculate Effect Estimates
Weisskopf
etal. (2007)
1089 males,
mean age
68.7 yr
Normative
Aging Study,
Boston, MA
Mean (IQR):
20 (13-28)
Mg/g (Tibia)
25 (17-37)
|jg/g (Patella)
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
Visuospatial, pattern
comparison
(pos is bad, latency)
Executive function verbal
fluency
Short-term memory, word
list
Assessed using CERAD,
Neurobehavioral Evaluation
System, WIAS-R, MMSE,
VMI
Tibia
Patella
Tibia
Patella
Tibia
Patella
0.79 (0.40,1,2)e
0.73 (0.40,1,2)e
-0.40 (-1.6, 0.80)e
-0.86 (-2.00, 0.30)e
-0.28 (-1.2, 0.60)e
-0.81 (-1.7, 0.05)e
Weuve etal.
(2006)
720 males,
ages >45 yr
Normative
Aging Study,
Boston, MA
Median (1st-
3rd quartile):
Tibia: 19(13-
28) Mg/g
Patella: 27
(18-39) |jg/g
Linear mixed effects
regression adjusted for
smoking status, alcohol
consumption, calorie
adjusted calcium intake,
regular energy expenditure
on leisure time physical
activity, diabetes
MMSE score
Tibia
ALAD wildtype
ALAD2 carrier
Patella
ALAD wildtype
ALAD2 carrier
-0.03 (-0.14, 0.07)
-0.11 (-0.30,0.06)
-0.03 (-0.11, 0.04)
-0.12 (-0.30, 0.06)
Wang et al.
358 males,
Median: 19
Linear regression adjusted MMSE
HFE wildtype
-0.20 (-1.0, 0.70)e
(2007)
median ages:
Mg/g (Tibia),
for age, years of education,
One HFE variant
-1.40 (-3.3, 0.40)e
67.2 yr (HFE
23 Mg/g
nonsmoker, former smoker,
Two HFE variants
-6.3 (-10.4, -2.1)e
wild-type)
(Patella)
pack-years, nondrinker,
67.7 yr (HFE
alcohol consumption, English
variant)
as first language, computer
Normative
experience, diabetes
Aging Study,
Boston, MA
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
. Population/
study Location
Pb
Biomarker Statistical Analysis
Data
Cognitive Test
Subgroup/
Model
Effect Estimate
(95% Cl)a
Van 47 adults,
Mean (SD):
Linear regression adjusted
Delayed matching, %
Wijngaarden mean age
2.0 (5.2) |jg/g
for age, gender, educational
correct
Calcaneus
87.56f
etal. (2009) 61.5 yr
(Tibia)
level, history of hypertension
Lowest fertile
86.67
Rochester,
6.1 (8.5) |jg/g
Medium fertile
80.67, p = 0.027
NY
(Calcaneus)
Highest fertile
Tibia
85.42f
Lowest fertile
87.08
Medium fertile
82.44, p = 0.25
Highest fertile
Calcaneus
Total trials
2.54f
Lowest fertile
2.61
Medium fertile
2.72, p = 0.21
Highest fertile
Tibia
2.62f
Lowest fertile
2.59
Medium fertile
2.66
Highest fertile
Assessed using
CANTAB and Montreal
Cognitive Assessment
aEffect estimates have been standardized to the standard deviation of the cognitive test scores and standardized to a 1 |jg/dL increase in blood Pb
and 10 |jg/g increase in bone Pb.
bThe directions of effect estimates were changed to indicate a negative slope as a decrease in cognitive performance.
°Effect estimates indicate interactions between Pb and ALAD genotype.
dEffect estimates indicate interactions between Pb and category of neighborhood psychosocial hazard (NPH), with the lowest fertile of NPH serving
as the reference group.
eEffect estimate refers to the change in cognitive function score overtime.
'Results refer to mean cognitive function scores among tertiles of bone Pb.
Weuve et al. (2009) studied the association of blood and bone Pb levels with cognitive function
among a subset of 587 women from the Nurses Health Study from whom blood and bone Pb
measurements were taken between the ages of 47 and 74 years. The mean (SD) blood Pb level in this
group was 2.9 (1.9) (ig/dL. The women had had Pb measured as part of their participation in two separate
sub-studies, one of osteoporosis and the other of hypertension. The inclusion criteria included residence in
the Boston area and absent of major chronic diseases at the time of Pb measurement. Cognitive function
was assessed via a battery of telephone administered tests (Telephone Interview of Cognitive Status, digit
span backwards, alphabetizing span, animal naming [category fluency], "f' naming [letter fluency], and a
composite verbal memory score) an average of 5 years after Pb measurement. As in the aforementioned
studies of adults, negative associations were observed more consistently for tibia and patella Pb levels
than for blood Pb levels (Figure 5-16 and Table 5-9). Contrary to expectation, scores on the "f' naming
test were positively associated with patella and tibia bone Pb levels. 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) increase in
tibia Pb level was associated with 0.051-point decrease (95% CI: -0.010, -0.003) in the standardized
composite score. This effect size of tibia bone Pb level was approximately equivalent to the effect size for
3 years of age among these women. The magnitude of effect was smaller for patella Pb levels (-0.033
[95% CI: -0.080, 0.014) per 1 SD unit increase in patella Pb level)
Several analyses of Normative Aging Study (NAS) have contributed to greater understanding of the
relationship between biomarkers of Pb dose and neurocognitive effects in adults. The NAS is conducted
at the VA Outpatient Clinic in Boston, MA and is a multidisciplinary longitudinal investigation of the
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aging process that was established in 1961 and originally involved 2,280 men residing in the Greater
Boston area aged 21 to 80 years with no current or past chronic medical conditions. Participants are
evaluated every 3-5 years with in-person clinical assessments, self-administered questionnaires. In-person
administered cognitive tests included those in the general domains of attention/executive function, short-
term memory, visuospatial, and verbal/language. About 60-70% of participants, depending on the specific
cognitive test, took individual tests twice approximately 3.5 years apart. The average (SD) age of the men
at baseline was 69 (± 7) years, and the median blood Pb level was 5 (ig/dL. Beginning in 1991,
measurement of bone Pb levels began, with 68% of active NAS subjects participating. All analyses were
adjusted for age, age squared, education, smoking, and alcohol intake. In one study of Mini-Mental State
Examination (MMSE) tests scores among 720 men 45 years of age and older, blood Pb levels were
negatively associated with MMSE score among ALAD2 carriers (Weuve et al.. 2006). A 3 (ig/dL
increment in 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 in the same cohort considered performance on a battery of oher
cognitive tests and did not find a consistent pattern of modification of the association between blood or
bone Pb levels and cognitive function by ALAD genotype (Raian et al.. 2008). Nonetheless, Raj an et al.
(2008) found tibia Pb levels to be associated more consistently with lower cognitive performance
compared with blood or patella Pb levels (Figure 5-16 and Table 5-9).
The NAS also examined effect modification by hemochromatosis (HFE) gene variants. 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 increase
in tibia bone Pb level (15jxg/g) was associated with a 0.22 point steeper annual decline (95% CI: -0.39, -
0.05) in MMSE score among men with any variant HFE allele (either H63D or C282Y). The magnitude
of association was less negative among those with only HFE wildtype alleles (F. T. Wang et al.. 2007)
(Figure 5-17). As indicated in Figure 5-17, a nonlinear association was observed between tibia Pb levels
and change in MMSE score, with a steeper decline per unit increase in tibia Pb level at higher tibia Pb
levels. Effect modification by nongenetic factors also were examined in the NAS cohort. Increasing bone
Pb levels were associated with greater decreases in cognitive function among individuals with higher
individual-level perceived stress (Peters et al.. 2007).
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-0.2
-0.4
-0.6
- - HFE wildtype
— HFE variant allele
-0.8
0
10
20
30
40
50
Tibia lead biomarker (ug/g)
Source: Wang et al. (2007).
Note: The lines indicate curvilinear trends estimated from the penalized spline method. Among HFE
wild-types, the optimal degree of smoothing was 1, meaning that 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), with an optimal 1.68 degree of smoothing. The model
was adjusted forage, years of education, nonsmoker, former smoker, pack-years, nondrinker,
alcohol consumption, English as first language, computer experience, and diabetes.
Figure 5-17. Exploration of nonlinear association of tibia Pb concentration
with annual rate of cognitive decline, by class of HFE
genotype.
Two large NAS studies examined the relationship of blood and bone Pb levels with the change in
cognitive function overtime. Weisskopf et al. (2007) expanded on an earlier study of Pb biomarkers and
cognitive function (Pavton et al.. 1998) The average (SD) age of the men at baseline was 69 (±7) years,
and the median blood Pb concentration was 5 (ig/dL. Two measurements of cognitive function, collected
approximately 3.5 years apart, were available 60-70% of participants. All analyses were adjusted for age,
age squared, education, smoking, and alcohol intake. There was little association between blood Pb levels
and cognitive test scores, except possibly for vocabulary scores, although this association was greatly
weakened when the five men with the highest blood Pb levels (>15 (ig/dL) were excluded. For bone Pb
analyses, the authors used repeated measures analysis with an interaction term between bone Pb level and
time in order to estimate the association between bone Pb level and decline in cognitive test score over
time. There were no main effect associations with bone Pb levels; however, increases in patella and tibia
bone Pb levels were associated with decreases in cognitive performance overtime. The only test for
which this reached statistical significance (p<0.05), however, was increased response latency on a pattern
comparison test. Contrary to expectation, bone Pb levels were associated with fewer errors on the same
pattern comparison test. The authors proposed that this may be related to slowing reaction time to
improve accuracy. When the 9 men with the highest bone Pb levels were removed, the association was no
longer statistically significant.
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In an examination of the patella Pb concentration-reaction time relationship, Weisskopf et al.
(2007) found a statistically significant nonlinear association, with latency times on the pattern comparison
test becoming worse over time (i.e., larger values or slower response latencies) up to approximately 60
(xg/g bone mineral, but the change over time leveling off at higher concentrations (Figure 5-18). Below 60
|ig/g. a 20 |ig/g difference in patella Pb level was associated with a decrease in pattern comparison test
score of approximately 0.15 ms. Intriguingly, however, the cognitive tests where the association with
bone Pb was significantly worse among ALAD-2 carriers (constructional praxis and pattern comparison)
were the same tests that were significantly associated with bone Pb in an earlier study (Weisskopf.
Proctor, et al.. 2007).
E,
c
J2
c
o
<1)
CD
C
n
JO
a>
tr
CM
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.
Figure 5-18. Nonlinear association between patella bone Pb concentration
and the relative change in response latency overtime on the
pattern comparison test (reference = 0 at mean of patella Pb
concentration).
In a somewhat similar approach to that taken in the NAS, BMS investigators took advantage of
repeat cognitive testing of study subjects (91% of the original cohort returned for a second round of
testing and 83% for a third round each at approximately 14-month intervals) to analyze associations of
blood and bone Pb levels with changes in cognitive performance over time cohort (Bandeen-Roche et al..
2009). An interquartile range increase in tibia Pb level (12.7 (xg/g) was associated with a 0.019 units per
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year decrease in eye-hand coordination z-score. The association was somewhat stronger among African-
Americans than it was among whites (Figure 5-16 and Table 5-9). Of the other tests, only change in
verbal memory and learning suggested some association with tibia Pb levels. In analyses of what they
term persistent effects (analogous to the main effects analyses of longitudinal data in the NAS), a similar
pattern was found as that reported by Shih et al. (2006) in the original BMS cross-sectional analyses.
Increasing tibia Pb levels were associated with decreases in cognitive performance; however, the effect
sizes decreased as more covariates were added to models. In models adjusted for age, sex, race, and
education, performance on executive function, verbal memory and learning, and visual memory test
decreased with increasing tibia Pb levels, with p-values ranging from 0.16 to 0.26. In contrast to the
results of the longitudinal analyses described above, race-stratified analyses of persistent effects 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 compared with
African-Americans.
Other studies generally confirm the same pattern as described in the larger studies above. One
study of 188 rural Chinese men and women found a weakly negative association between plasma Pb
levels and a composite cognitive score based on a battery of in-person administered tests (S. J. 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; Hemandez-Avila
etal.. 1998). In another study of 47 men and women in Rochester, NY (55-67 years of age), subjects in
the higher two tertiles of calcaneal bone Pb level had lower scores on a delayed matching-to-sample task
(van Wiingaarden et al.. 2009) (Table 5-9). The pattern was similar for a paired associates learning task,
although the results did not reach statistical significance. In analyses of tibia Pb levels, subjects in the
highest tertile of tibia Pb level performed worse on cognitive tests (Table 5-9).
In summary, among adults without occupational exposures to Pb, there is weak evidence for an
association between blood Pb levels and cognitive function. The strongest evidence of association
between blood Pb levels and cognitive function in adults was provided by the various NHANES analyses
of various age and genetic variant subgroups (E. F. Krieg. Jr. & Butler. 2009; E. F. Krieg. Jr. et al.. 2009;
E. F. Krieg. Jr. et al.. 2010). These NHANES analyses did not have bone Pb measurements for
comparison. Consistent with the conclusion of the 2006 Pb AQCD, recent studies continued to
demonstrate associations between bone Pb levels and cognitive deficits in adults (Figure 5-16 and Table
5-9). Recent studies also demonstrated that bone Pb levels may be associated specifically with cognitive
decline over time. Among recent studies that analyzed both blood and bone Pb levels, bone Pb levels, in
particular tibia Pb levels, were associated with greater decreases in cognitive performance than were
blood Pb levels across of the various cognitive tests that were performed. The discrepant findings for
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blood and bone Pb levels indicate that biomarkers of cumulative Pb exposure, including higher levels in
the past, may the best predictors of neurocognitive function in adults. The effects associated with
cumulative Pb exposure are also demonstrated by observations that tibia Pb levels were associated with
larger decreases in cognitive performance than were patella Pb levels. Tibia bone has a slower rate of
turnover compared with patella bone and is an indicator of longer-term Pb exposure.
Adults with Occupational Lead Exposures
The 2006 Pb AQCD concluded that the evidence for an association between blood Pb levels and
cognitive function in adults was most consistent among adults occupationally exposed to Pb. Results from
a few recent studies of occupationally-exposed adults support the previous conclusions. 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 bone 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 associated with poorer performance on a battery of neurocognitive tests (B. S.
Schwartz et al.. 2005; B. S. Schwartz et al.. 2001). Dorsey et al. (2006) found patella Pb levels to be
associated with poorer performance in the domains of manual dexterity, sensory PNS function, and
depression symptoms. In this occupational cohort, however, the associations between patella Pb levels
and cognitive function were not as strong as the associations with either blood or tibia Pb levels.
A follow-up study of the original 1982 Lead Occupational Study was conducted in 2001-2004
among 83 of the original 288 Pb-exposed workers and 51 of the original 181 controls (Khali 1. Morrow, et
al.. 2009). 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 appear to 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 significantly (p <0.05) worse than the controls in total
cognitive score and in the spatial and general intelligence domains. They also performed worse in all
other domains (e.g., motor, executive, and memory) although not significantly so. A similar pattern was
observed in analyses using tibia Pb levels measured at the follow-up visit as the exposure variable.
Associations also were seen with blood Pb levels (median among the exposed: 12 (ig/dL), although these
generally did not reach statistical significance. Among the former Pb workers, 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 other domains as well, although
they were not statistically significant.
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Two additional studies aimed to characterize factors that either mediate or modify the association
between Pb biomarkers and cognitive function. A study among 61 current Pb smelter workers with an
average age of 40 years and blood Pb of 29.1 (ig/dL found that both a time-weighted integrated blood Pb
measure (p = 0.09) and tibia Pb level (p = 0.08) were associated with longer times to complete the
grooved pegboard test (Bleecker. Ford. Vaughan. et al.. 2007). In another study to examine the modifying
effects of cognitive reserve (assessed by performance on the Wide Range Achievement Test-R for
reading) on the Pb-cognitive function association among 112 Pb smelter workers, a time-weighted
integrated blood Pb measure (an index of cumulative exposure) was associated with decrements in motor
performance (p<0.05), and among those with low cognitive reserve, tests of attention tasks and a digit
symbol task as well (Bleecker. Ford. Celio. et al.. 2007V
Iwata et al. (2005) examined the 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, increasing
blood Pb level was associated with increases in 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 dose-response model of their data.
Although the data were slightly better fit with a supralinear dose-response function, the linear function
was also statistically significant.
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.
Carriers of the E4 variant are at twofold increased risk of developing Alzheimer's disease, although the
majority of such individuals still do not develop the disease. A study of occupationally-exposed adults
found that among individuals with at least one ApoE-s4 allele, Pb was associated with greater decrements
in tests such as digit symbol, pegboard assembly, and complex reaction time (W. F. Stewart et al.. 2002).
5.3.3. Neurobehavioral Effects
5.3.3.1. Epidemiologic Studies of Behavioral Effects in Children
Several epidemiologic studies reviewed in the 2006 Pb AQCD reported associations between blood
Pb levels and problems with behavior and social conduct that ranged from inattentiveness to self-reported
delinquent behaviors to criminal activities (Bellinger et al. 1994; Bellinger & Rappaport. 2002; Burns et
al.. 1999; Dietrich et al.. 2001; Needleman et al.. 2002; Needleman et al.. 1996; G. A. Wasserman et al..
1998). Recent studies continue to demonstrate similar associations and provide new evidence for
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associations of blood Pb levels with ADHD diagnosis and diagnostic indices (Figure 5-19 and Table 5-
10). Noncognitive effects of Pb are more complex to study relative to IQ tests. 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) (R,
F. White et al.. 2009). Most of these 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
(7-22 years of age). Most of these studies examined associations with blood Pb levels assessed at one
time point; however, even the prospective studies with serial measurements of blood Pb levels, found
associations with both prenatal and early childhood blood Pb levels (Dietrich et al.. 2001; J. P. Wright et
al.. 2008) and lifetime average blood Pb levels (Burns et al.. 1999). Therefore, uncertainty remained over
what was the critical time period of Pb exposure for increasing risk of behavioral problems and
misconduct. In many of the studies noted above, blood Pb level also was associated with IQ and other
endpoints of cognitive function, thus it was unclear whether blood Pb-associated neurocognitive deficits,
poor school performance, and attention problems, in turn, progressed to antisocial and delinquent
behavior later in life.
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Reference
Blood Pb (ug/dL)
Exposure Period
Outcome
Chen et al. (2007)
8
8
8
Concurrent
Concurrent
Concurrent
Behavioral symptoms
Externalizing problems
School problems
Nicolescu et al. ("20101
3.7 (IQR: 2.6)
3.7 (IQR: 2.6)
Concurrent
Concurrent
Distractibility
Inattention
Roy et al. (2009")
11.4 (5.3)
11.4 (5.3)
11.4 (5.3)
Concurrent
Concurrent
Concurrent
Anxiety
Inattention
Hyperactivity
Nicolescu et al. (2010)
3.7 (IQR: 2.6)
Concurrent
ADHD score
Roy et al. (2009")
11.4 (5.3)
Concurrent
ADHD index
Braun et al. (200613
1.1-1.3 (3* quintile)
0.8-1.0 (2nd quintile)
1.4-2.0 (4th quintile)0
>2.0 (5th quintile)3
Concurrent
Concurrent
Concurrent
Concurrent
ADHD Dx
ADHD Dx
ADHD Dx
ADHD Dx
Froehlich et al. (20091
>2.0a
Concurrent
ADHD Dx, prenatal smoke exposure
Braun et al. (20081
0.8-1.0 (2ncl quartile)3
1.0-1.4(3™ quartile)3
1.5-10 (4th quartile)3
Concurrent
Concurrent
Concurrent
Conduct disorder
Conduct disorder
Conduct disorder
Dietrich et al. (20011
NR
Early childhood
Delinquent behavior
Wright et al. (20081
8.3 (3.8)
8.3 (4.8)
Prenatal
Early childhood avg
Criminal arrests
Criminal arrests
Odds ratio (95% CI)
Note: Studies are presented in order of increasing severity of behavioral outcome. Odds ratios are
standardized to a 1 [jg/dL increase in blood Pb level in analyses of blood Pb level as a continuous
variable. IQR = Interquartile range, ADHD = attention deficit hyperactivity disorder, NR = Not
reported. aEffect estimate compares children in higher quantiles of of blood Pb level, with children
in the lowest blood Pb quantile serving as the reference group.
Figure 5-19. Associations of blood Pb levels with behavioral indices in
children.
Table 5-10. Additional characteristics and quantitative results for studies presented in
Figure 5-19.
Study Population/Location
Blood Pb
Levels
(pg/dL)
Statistical Analysis Outcome
Odds ratio
(95% Cl)a
Chenetal. 780 children participating Concurrent
(2007) in TLC trial followed mean (range):
between ages 2-7 yr (0-26)
Baltimore, MD; Cincinnati,
OH; Newark, NJ;
Philadelphia, PA
Linear regression model
adjusted for city, race, sex,
language, parental eduction,
parental employment, single
parent, age at blood Pb
measurement, caregiver IQ
Behavioral problems
Externalizing problems
School problems
At age 5-7 yr assessed
using CPRS-R, BASC-TRS,
and BASC-PRS
1.02 (0.99, 1.06)
1.04(1.00,1.07)
1.03(1.00,1.06)
Nicolescu 83 children ages 8-12 yr Concurrent
et al. tested in 2007 mean (IQR): 3.7
(2010) Bucharest and Pantelimon, (2.6)
Romania
Log-linear regression model
adjusted for city, sex, age,
computer experience,
handedness, eye problems,
number of siblings, parental
education, prenatal smoking,
family psychopathology
Distractibility
Inattention
At age 8-12 yr assessed
using German version of
Teacher's Conner's scales
1.35 (0.99,1.85)
1.14(0.95,1.37)
Roy et al. 756 children ages 3-7 yr Concurrent
(2009) tested 2005-2006 mean (SD): 11.4
Chennai, India (5.3)
Linear regression model
adjusted for age, sex,
hemoglobin, average monthly
income, parental education,
number of other children,
clustering in school and
classroom
Anxiety
Inattention
Hyperactivity
At age 3-7 yr assessed
using Conners'ADHD/DSM-
IV Scales and Behavior
Rating Inventory of
Executive Function
1.19(1.03,1.36)
5.31 (1.46,19.3)
2.39 (0.77, 7.39)
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Blood Pb
Study Population/Location Levels
(M9/dL)
Statistical Analysis Outcome
Odds ratio
(95% Cl)a
Nicolescu 83 children ages 8-12 yr Concurrent
et al. tested in 2007 mean (IQR): 3.7
(2010) Bucharest and Pantelimon, (2.6)
Romania
Log-linear regression model
adjusted for city, sex, age,
computer experience,
handedness, eye problems,
number of siblings, parental
education, prenatal smoking,
family psychopathology
ADHD score
At age 8-12 yr assessed
using German version of
Conner's scales
1.16(0.98,1.37)
Roy et al. 756 children ages 3-7 yr Concurrent
(2009) tested 2005-2006 mean (SD): 11.4
Chennai, India (5.3)
Linear regression model
adjusted for age, sex,
hemoglobin, average monthly
income, parental education,
number of other children,
clustering in school and
classroom
ADHD index
At age 3-7 yr assessed
using Conners'ADHD/DSM-
IV Scales and Behavior
Rating Inventory of
Executive Function
4.26 (0.90, 20.08)
Braun et 4,704 children ages 4-15 Concurrent
al. (2006) yr 3rd quintile: 1.1-
U.S. NHANES 1999-2002 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 |jg/dL vs. blood
Pb <0.8 |jg/dLb
2.1 (0.7, 6.8), blood Pb
1.1-1.3 ^jg/dL vs. blood
Pb <0.8 |jg/dLb
2.7 (0.9, 8.4), blood Pb
1.4-2.0 |jg/dL vs. blood
Pb <0.8 |jg/dLb
4.1 (1.2,14.0), blood Pb
> 2.0 |jg/dL vs. blood Pb
<0.8 |jg/dLb
Froehlich 2,588 children, ages 2nd quartile: 0.9-
etal. 8-15 yr 1.3
(2007) U.S. NHANES 2001-2004
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 |jg/dL plus
prenatal ETS exposure
vs. blood Pb level <0.8
|jg/dL and no prenatal
ETS exposure11
Braun et
al. (2008)
Children ages 8-15 yr
U.S. NHANES 2001-2004
2nd quartile: 0.8- Logistic regression with
1.0 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 at age 8-
15 yr
7.24(1.06,49.47), blood
Pb level 0.8-1.0 ^ig/dL
vs. blood Pb <0.8 |jg/dLb
12.37(2.37,64.56),
blood Pb level 1.1-1.4
ug/dL vs. blood Pb <0.8
|jg/dLb
8.64(1.87,40.04), blood
Pb level 1.5-10 |jg/dL vs.
blood Pb <0.8 |jg/dLb
Dietrich et 195 children followed from Early childhood Linear regression model
al.(2001) birth (1979-1985) to age (0-6yravg): NR adjusted for HOME score,
15-17 yr parental IQ, current SES
Cincinnati, OH
Delinquent behavior
assessed at ages 15-17 yr
using the Self-Report of
Delinquent Behavior
1.21 (1.08,1.37)
Wright et 250 adults followed from Early childhood
al. (2008) birth (1979-1985) to age mean (SD): 8.3
19-24 yr (4.8)
Cincinnati, OH
Negative binomial regression
models adjusted for maternal
IQ, sec, SES, maternal
education
Criminal arrests at ages 20-
23 yr
1.05(1.00,1.09)
aEffect estimates are standardized to a 1 |jg/dl_ increase in blood Pb level.
bOdds in higher quantile of blood Pb level compared to that in lowest quantile of blood Pb level.
Burns et al. (1999) and Silva et al. (1988) aimed to characterize the direct association of blood Pb
level with behavior by adjusting for child IQ in their models. They found positive associations, suggesting
that blood Pb level may have an independent effect on behavior. However, because a decrement in IQ
may on the causal pathway to behavioral problems, including both IQ and behavioral problems may result
in an underestimate of the effect on behavior. Hence, to account for the relationship between IQ and
behavior, Chen and colleagues (2007) used structural equation modeling (specifically path analysis) to
estimate the direct effects of blood Pb level on behavioral problems as well as indirect effects mediated
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through child's IQ (assessed using the WPPSI-R at age 5 years and the WISC-III) at age 7 years). Among
5- and 7-year-old participants from the TLC trial (described in Section 5.3.2.1), conduct problems were
assessed using the Conners' Parent Rating Scale-Revised, and other behaviors were assessed with the
Behavior Assessment System for Children-Teacher Rating Scale (BASC-TRS) and Behavior Assessment
System for Children-Parent Ratings Scale (BASC-PRS). In models that adjusted for sex, race, clinical
site, language (English or Spanish), parent's education, parent's employment, age at blood Pb testing,
caregiver's IQ, and treatment group, peak blood Pb levels at age 2 years were not associated with conduct
problems at age 5 years or any of the BASC scores at age 7 years. Concurrent blood Pb levels had
statistically significant direct effects on behavioral symptoms index, externalizing problems and school
problems as assessed by BASC-TRS and externalizing problems (outbursts of behavior) as assessed by
BASC-PRS (Figures 5-20 and 5-21).
Chen et al. (2007) also found significant indirect effects (mediated through child's IQ) of blood Pb
levels at age 7 years with all measurements except the BASC-TRS externalizing problems and BASC-
PRS internalizing problems (repressing problems). However, because blood Pb level was estimated to
have direct effects on several behavioral endpoints and, in for some endpoints, a larger magnitude of
direct effect, the authors inferred IQ to only partially mediate associations between blood Pb levels and
behavioral outcomes (Figures 5-20 and 5-21). The study also had limitations, including the lack of
adjustment for HOME score and lack of information on other covarying family and neighborhood
characteristics that may be relevant (i.e., social stressors). Also, these findings may not be generalizable 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. In this study, only concurrent blood Pb levels were
examined, 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.
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Adaptive skills
Behavioral Symptoms Index
Externalizing problems
Internalizing problems
School problems
-T"
-2
Oirect effecis
Indirect effects
0 2 4
Point estimates and 95% CIs
Source: Used with permission from the American Academy of Pediatrics, Chen et al. (2007)
Note: Values represent change in behavioral score per 10 pg/dL increase in blood Pb level.
Figure 5-20. Estimates and 95% CI of direct arid indirect effects of
concurrent blood Pb concentrations at age 7 on BASC-TRS.
Adaptive sKI lis
Behavioral Symptoms Index
Externalizing problems
Internalizing problems
Direct effects
Indirect effects
-4 -2 0 2 4 6 3
Point estimates and 95% CIs
Source: Used with permission from the American Academy of Pediatrics, Chen Chen et al. (2007)
Note: Values represent change in behavioral score per 10 pg/dL increase in blood Pb level.
Figure 5-21. Estimates and 95% CI of direct and indirect effects of
concurrent blood Pb concentrations at age 7 on BASC-PRS.
The 2006 Pb AQCD summarized indirect evidence for associations of blood Pb level with
behavioral features of ADHD including distractibility, poor organization, lacking persistence in
completing tasks, and daydreaming (Bellinger et al.. 1994; Fergusson et al.. 1993; Needleman et al.. 1979;
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G. A. Wasserman et al.. 2001; G. A. Wasserman et al.. 1998). Prior studies examining the association of
blood Pb level with a diagnosis of ADHD were limited by small sample size, and results were
inconclusive (David et al.. 1972; Gittleman & Eskenazi. 1983). In addition, many earlier studies of
inattention and impulsivity included children with higher blood Pb levels than those observed in
contemporary children.
A recent analysis of data from NHANES (1999-2002) found a positive relationship between blood
Pb level and ADHD (parent-report of a diagnosis of ADHD or use of stimulant medication) (Braun et al..
2006). This analysis included 4,704 children aged 4-15 years. These authors reported a monotonic
increase in odds of ADHD from the lowest to highest quintile of blood Pb level (Figure 5-22). Children in
the fifth quintile of blood Pb level (>2.0 (ig/dL) had the highest odds of ADHD compared with those in
the lowest quintile (<0.8 (ig/dL) (OR: 4.1 [95% CI: 1.2, 14]). Children in the other three higher quintiles
of blood Pb level also had increased odds of ADHD relative to the reference group; however, the
associations were not statistically significant (Tables 5-19 and 5-10).
In the same NHANES dataset, however, restricted to children ages 8-15 years, Froehlich and
colleagues (2009) demonstrated the joint effects of prenatal tobacco smoke (maternal report) and blood
Pb levels. They found independent effects on ADHD for 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. A statistically significant
interaction also was found. Compared to children in the lowest fertile of blood Pb levels with no exposure
to prenatal tobacco smoke, children in the highest fertile of blood Pb level with exposure to prenatal
tobacco smoke had the greatest odds of ADHD (OR: 8.1 [95% CI: 3.5, 18.7]).
5
4
DC 3
O
< 2
1
0
Source: Braun et al. (2006)
Note: Adjusted for child's age, gender, race/ethnicity, preschool attendance, serum ferritin, prental
tobacco smoke exposure, smoker in the household, and insurance status (p for trend = 0.012).
Figure 5-22. Adjusted odds ratio for ADHD among U.S. children (ages 4-15
years) from NHANES 1999-2002 by quintile of blood Pb level.
<0.8 0.8-1.0 1.1-1.3 1.4-2.0 >2,0
Quintiles of blood lead concentration (ug/dL)
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Pb has long been known to impact dopaminergic neurons. Pb will inhibit depolarization-evoked
neurotransmitter release and stimulate spontaneous neurotransmitter release, including effects specific to
dopaminergic neurons (Section 5.3.6.9). This background would make dopamine related genes strong
candidates as effect modifiers of Pb-associated neurodevelopmental effects. There are several known
functional variants in dopamine related genes, including the dopamine 2 receptor, dopamine 4 receptor
(DRD4), and the dopamine transporter. The DRD4 gene has a 48 base pair long repeat in exon 3. The
longer repeat alleles (DRD4.7) appear to have less binding affinity for dopamine and have been
associated with ADHD. However, DRD4.7 also has been associated with sustained attention, response
inhibition, and quicker response time. Thus, it is not clear whether DRD4.7 would increase or decrease
susceptibility to Pb-associated ADHD. Froehlich et al. (2007) studied 174 children from the Rochester
prospective birth cohort and measured DRD4 genotype, blood Pb level, and several functional domains
using the Cambridge Neuropsychological Testing Automated Battery. Increasing blood Pb levels were
associated with poorer rule learning and reversal, spatial span, and planning. These negative associations
were larger in magnitude among boys and those lacking DRD4.7.
In a population of children in a similar age range (8-11 years) and similar blood Pb levels (mean
1.9 (ig/dL) as in the NHANES studies, Cho et al. (2010) found a statistically significant relationship of
blood Pb levels with parent- and teacher-rated ADHD symptoms (i.e., inattentiveness, hyperactivity, and
total score).
Nicolescu et al. (2010) examined the relationship between blood Pb level and ADHD-related
behaviors among 83 children, ages 8-12 years, living near a metal-processing plant in Romania. These
authors examined associations of Pb, mercury, and aluminum biomarkers with performance on four
different attention tasks on the computerized German Kinder-KITAP as well as behavior ratings using the
parent and teacher reports on the Connors scale. Only concurrent blood Pb levels, and not other metals,
were consistently associated with increased odds for ADHD-related behaviors (OR: 1.16 [95% CI: 0.98,
1.37] per 1 (ig/dL increase in blood Pb level. Notably, blood Pb levels ranged from 1.1 to 14.3 (ig/dL, and
when analyses were repeated taking out the 5 children with blood Pb levels at or above 10 (ig/dL,
associations with both questionnaire-based ADHD ratings and KITAP-performance remained statistically
significant.
Roy and colleagues (2009) examined associations between blood Pb level and a range of
behavioral problems in 756 children, ages 3-7 years, in Chennai, India. In this population, mean (SD)
blood Pb level was 11.4 (5.3) (ig/dL. Anxiety, social problems, inattention, hyperactivity, and ADHD as
well as executive function were assessed based on the teacher's report of the Conners' Teacher Rating
Scales-39, Conners' ADHD/Diagnostic and Statistical Manual for Mental Disorders, 4th Edition Scales,
and the Behavior Rating Inventory of Executive Function questionnaire. In generalized estimating
equations, increasing blood Pb level was associated with higher anxiety, social problems, and ADHD
index scores (Figure 5-19 and Table 5-10).
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Recent studies add to the collective body of evidence demonstrating associations between blood Pb
level and a variety of conduct problems including conduct disorder (Braun et al.. 2008). oppositional
defiant disorder (Nigg et al.. 2008) and more serious behaviors such as criminal behavior (J. P. Wright et
al.. 2008). These findings were corroborated in a recent meta-analysis on Pb and conduct problems
(Marcus et al.. 2010) that included 19 studies with atotal of 8,561 children and adolescents (mean ages
ranging from 3.5 years to 18.4 years). An overall medium effect size was estimated (r across studies was
0.19, p<0.001). Moreover, the meta-analysis demonstrated a consistent relationship between increasing
blood Pb levels and conduct problems despite considerable heterogeneity across studies (e.g., ways in
which conduct problems were defined and assessed, participant ages, participant blood Pb levels). In the
meta-analysis, covariates such as SES, birth weight, parental IQ, and family environment did not
attenuate the relationship between blood Pb level and conduct problems. Interestingly, a larger magnitude
of effect was estimated for hair Pb levels compared with bone or blood 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 addition to examining ADHD in the NHANES population, Braun and colleagues (2008) also
examined conduct disorder defined using DSM-IV criteria. The prevalence of conduct disorder was
higher among males, older children (13-15 years, relative to those less than 13 years), and children with
higher blood Pb levels. In analyses adjusted for child's age, gender race/ethnicity, SES, mother's age at
child's birth, prenatal tobacco smoke exposure (based on maternal report), and child's cotinine levels,
children in the highest quartile of blood Pb level (>1.5 (ig/dL) had increased odds of conduct disorder
relative to children in the lowest quartile (0.2-0.7 (ig/dL) (OR: 8.7 [95% CI: 1.9, 40]). Poisson regression
models showed that compared with children in the lowest quartile of blood Pb level, children in the
highest quartile had 1.73 (95% CI: 1.2, 2.4) times as many conduct disorder symptoms.
In their longitudinal study of children in the U.K., Chandramouli et al. (2009) found associations of
blood Pb level at age 30 months not only with hyperreactivity but also with antisocial behavior at ages 7
and 8 years. Child behavior was assessed using three methods: parent- and teacher-report on the Strengths
and Difficulties Questionnaire at age 7 years, the Development and Well-being Assessment at age 8 years,
and the Anti-social Behaviour Interview at age 8 years. Attention was measured using the Test of
Everyday Attention for Children at age 8 years. Blood Pb levels showed statistically significant positive
association with antisocial behavior. Similar to Chen et al. (2007) and Burns et al. (1999). increasing
blood Pb level was associated with antisocial behavior, independently of IQ.
Wright et al. (2008) recently examined the relationship between prenatal and postnatal blood Pb
levels and arrests for criminal offenses at ages 19-24 years in the CLS. In this birth cohort, prenatal and
postnatal blood Pb levels previously have been reported to be associated with self- and parent-reported
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delinquent and social acts at ages 16-17 (Dietrich et al.. 2001). Data on criminal arrests for participants
and their mothers were obtained from a computer search of Hamilton County, Ohio criminal justice
records. They also examined the absolute change in arrest rates between participants with higher levels of
blood Pb compared to those with lower blood Pb levels defining attributable risk as the average difference
in annual arrest rates between participants at the 95th percentile of blood Pb and those at the 5th
percentile. Mean blood Pb concentrations were 8.3 (ig/dL (range 1 to 26) for the prenatal period (maternal
blood), 13.4 |_ig/dL (range 4 to 37) for the average between birth and age 6 years, and 8.3 (ig/dL (range 2
to 33) at age 6 yr. 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. Relative risks for violent criminal arrests were
also larger for prenatal blood Pb levels and age 6-year blood Pb levels. The RRs for arrests involving
violent crimes per 1 jxg/dl increment in blood Pb level were 1.06 (95% CI: 0.97, 1.15) for prenatal blood
Pb level, 1.05 (95% CI: 1.01, 1.10) for average childhood blood Pb level, and 1.08 (95% CI: 1.03, 1.14)
for blood Pb level at age 6 years. Although interactions terms of 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]). Similar to findings from Dietrich et al. (2001).
prenatal and early childhood blood Pb levels were associated positively with risk of criminial behavior in
early adulthood. Results from the two studies combined suggest that in addition to the prenatal period,
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.
5.3.3.2. Epidemiologic Studies of Behavior, Mood, and Psychiatric Effects in
Adults
Effects of blood Pb levels on emotional regulation in adults have received far less attention than
that in children and cognitive function in adults. Nonetheless, evaluation of mood states is an integral part
of the World Health Organization's (WHO) neurocognitive test battery and, indeed, it has been suggested
that the assessment of mood with the Profile Of Mood States may be particularly sensitive to toxicant
exposures (1987). With respect to Pb exposure, 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 of mood disorders and
anxiety among Pb-exposed workers than unexposed controls (mean blood Pb levels ranging from 15-38
(.ig/dL) (Baker et al.. 1984; Baker et al.. 1985; Litis et al.. 1977; Maizlish et al.. 1995; Parkinson et al..
1986; B. S. Schwartz et al.. 2005). In one of the few previous studies of adults without occupational
exposure to Pb and more relevant to blood and bone Pb levels measured currently in the U.S. among
adults without occupational exposures, an association was observed between both blood (mean: 6.3 (ig/dL
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[SD: 4.16] and bone (tibia mean: 21.9 jj.g/g [SD: 13.5]) Pb levels and depression and anxiety symptoms
(Brief Symptom Inventory, BSI) among men in the NAS (Rhodes et al.. 2003).
More recently, subsequent assessments of the NAS men have provided understanding of effect
modification of the associations of blood, patella, and tibia Pb levels with psychiatric symptom
dimensions by ALAD genotype (Raian et al.. 2007). In addition to corroborating associations of Pb with
BSI symptoms reorted by Rhodes et al. ("2003). Rajan et al. (2007) found that of eight symptom
dimensions considered, associations of tibia bone Pb levels with phobic anxiety, positive symptom total,
and anxiety scale were modified by ALAD genotype, with interaction terms attaining statistical
significance for phobic anxiety and positive symptom total. For all psychiatric symptoms, the association
with tibia Pb was worse among ALAD 1-1 carriers, which was the opposite genotype observed to have
larger Pb-associated decrements in cognitive performance (Section 5.3.2.4).
A study of 1,987 adults age 20-39 years participating in NHANES 1999-2004 was the largest study
of psychiatric disorders and the study with the lowest blood Pb levels (mean 1.61 |_ig/dL [SD: 1.72])
(Bouchard et al.. 2009). Investigators examined cases of major depressive disorder (MDD), panic disorder
and generalized anxiety disorder (GAD) as determined using the WHO Composite International
Diagnostic Interview, which follows criteria defined in DSM. Compared with those in the lowest quintile
of blood Pb level (<0.7 (ig/dL), adults in the highest quintile (>2.11 |_ig/dL) had increased odds of MDD
(OR: 2.32 [95% CI: 1.13, 4.75]) and panic disorder (OR: 4.94 [95% CI: 1.32, 18.48]). Odds ratios were
even larger in analyses excluding current smokers. While those in higher quintiles of blood Pb level (>0.7
(ig/dL) had increased odds to GAD, the associations were not statistically significant. Although studies in
adults without occupation Pb exposure are sparse, consistent with studies of occupationally-exposed
adults and experimental evidence, they demonstrate associations of blood (as low as 2.11 (ig/dL) and
bone Pb level (population mean tibia Pb level 22 jj.g/g) with psychiatric outcomes.
5.3.3.3. Toxicological Studies of Neurobehavioral Outcomes
Pb is a known risk factor for neurobehavioral changes with preferentially targeted sites including
the prefrontal cerebral cortex, cerebellum, and hippocampus; affected functions include cognition,
execution of motor skills, and memory/behavior, respectively. As discussed in earlier AQCDs, young
animals are especially susceptible to the effects of Pb due to differences in structure of the nervous system
and to the ongoing development 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 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 ability of Pb to mimic calcium. Very early research on neurobehavioral endpoints
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failed to capture the disposition of Pb and its resulting body burden or blood Pb and is thus difficult to use
in risk assessment.
Since then, the 1986 AQCD reported the literature on rodent and nonhuman primate Pb-induced
aberrant operant conditioning tasks in rodents and non-human primates (blood Pb 11 (ig/dL to 15 (ig/dL)
with other studies yielding hyperactive or inappropriate Pb-mediated responses (U.S. EPA. 1986) possibly
of hippocampal origin with a curvilinear response, decreasing at higher doses possibly due to impairment
of motor function at the high doses (Crofton et al.. 1980; Maetal.. 1999). Pb exposure in lab animals
contributed to distractibility, reduced adaption capacity to changes in behavior, impaired ability to inhibit
inappropriate responses and perservation (U.S. EPA. 2006). Pb is known to impair learning (blood Pb 11
(ig/dL) measured using Fixed Interval tasks (FI) defined as scheduled reinforcement delivered after a
fixed period since the last reward, with premature responses not rewarded. The 2006 AQCD found FI
response rates (blood Pb 58 to 94 (ig/dL) were sensitive to Pb exposure, which was primarily accounted
for by decreased interresponse times. Inter-response rates and overall run rate are the two subcomponents
of FI response rate. Spatial and non-spatial discrimination reversal or reversal of a previously learned
habit is significantly affected after developmental Pb exposure and is exacerbated with distracting stimuli;
discrimination reversal has been shown to be especially sensitive to Pb exposure. Repeat-acquisition
testing revealed that these deficits are likely not due to sensory or motor impairment at this dose. The
results from different studies testing the effect of Pb on memory are mixed with impaired memory shown
at blood Pb level of 10 (ig/dL and improved memory in other studies. Low dose Pb does not appear to
affect short-term memory. Memory tests may give incorrect results when opportunities exist for impaired
attention to contribute to test results (U.S. EPA. 2006). Together, the data from the 2006 AQCD showed
that social behavior and learning in rodents and nonhuman primates is significantly affected by Pb
exposure (blood Pb 15-40 (ig/dL).
In the new literature, gestationally-Pb exposed (GLE) male mice (low and high dose Pb, 10 (ig/dL
and 42 (ig/dL blood Pb at PND10) were significantly less active than control mice and low dose GLE
mice were significantly less active than high dose GLE mice, demonstrating a non-linearity of GLE dose
responsiveness (Leasure et al.. 2008). A similar Pb dose response non-linearity (baseline corticosterone)
was seen in male mice exposed post-weaning to Pb (Virgolini et al.. 2005). Activity level of GLE female
mice versus control was unaffected (Leasure et al.. 2008). Amphetamine induced motor activity was
monitored in male and female GLE mice at 1 year of age. Amphetamine-induced activity of male low and
high dose GLE offspring was significantly elevated versus control; GLE females had no change in
sensitivity to amphetamine-induced motor activity (Leasure et al.. 2008).
Rotarod performance measures endurance, balance and coordination in mice. GLE male mice had
significantly shorter mean latencies to fall from the rotarod compared with controls; females were
unaffected. Further, low dose GLE male mice had significantly poorer rotarod performance than high
dose GLE male mice (fell off quicker, 10 (ig/dL and 42 (ig/dL blood Pb at PND10), showing non-linearity
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of dose-responsiveness to GLE (Lcasure et al.. 2008). Other rotarod experiments at higher doses of Pb
exposure and at various speeds of rotarod rotation yielded mixed results (Grant et al.. 1980; kishi et al..
1983; Overmnnn. 1977).
Herring gull chicks exposed to a single IP bolus dose (100 mg/kg Pb acetate) of Pb on post-natal
day two at a dose created to be similar to that which wild herring gulls are exposed to in the wild were
found to have neurobehavioral deficits, including 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 parent bird's mouth, decreased time spent in the shade
(behavioral thermoregulation), decreased learning in food location testing, decreased recognition of
familiar individuals (caretaker or sibling), and slower development of motor skills (treadmill test) versus
control birds (Burger & Gochfeld. 2005). The impaired thermoregulation with Pb exposure agrees with
earlier work in rat pups who also showed impaired thermoregulatory behavior, i.e., impaired ultrasonic
vocalization (Davis. 1982). These studies in herring gull chicks demonstrate that a single early life
exposure to Pb can induce neurobehavioral deficits that affect survival skills.
5.3.3.4. Toxicological Studies of Mood Alterations
Neurotoxicological studies often focus on motor, sensory, behavioral or cognitive outcomes and
often fail to evaluate psychological pathologies. Recent epidemiologic studies have reported that prenatal
or early life blood Pb levels or ALAD changes, a biomarker of Pb exposure, may be a risk factor for
development of mood disorders in adulthood (i.e. schizophrenia, major depression or panic disorders)
(Bouchard et al.. 2009; Qpler et al.. 2004). With this in mind, animal studies of Pb exposure during
pregnancy and lactation and outcomes in offspring look to address the role of developmental Pb exposure
on emotional state and mood disorder-like behavior in adult offspring. Wistar rats exposed to lOmg of Pb
acetate daily by gavage during pregnancy (G) or pregnancy plus lactation (G+L) produced pups that were
then tested in the open field test or the forced swimming test also known as Porsolt's test. Blood Pb levels
in the pups at PND70 was 5-7 (ig/dL. The open field test can measure emotion and exploratory behavior;
Pb (G+L) treated male rats had increased emotionality with the open field test. Pb exposed (G+L) female
offspring had a significant increased depressive phenotype in the forced swim test (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.
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; T. W. W. Pace & Miller. 2009). Some researcher use
sickness behavior and its associated malaise as a model for depression. Examples from animal models
include the study by Dyatlov and Lawrence (2002) with Pb exposure in mice. In this study, sickness
behavior, which is due to an interaction of the immune system and the CNS in Pb-exposed mice, was
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potentiated by Pb-exposure (blood Pb level 17 (ig/dL) and this correlated with depletions in specific
thymic T-cell populations. Pb-exposure also potentiated the infection-dependent elevation in IL-1|3, a
cytokine which has been shown to inhibit hippocampal glutamate release in young and not aged animals.
Sickness behavior was induced with Listeria monocytogenes infection; Pb exposure was from birth
through lactation and continued for a brief period after weaning until the experiment was terminated.
Sickness behavior is characterized by overall malaise, decreased food intake, immobility and changes in
core body temperature. Pb potentiated the sickness behavior in exposed animals.
Neuro-
transmitters
NMDA Receptor
Hypofunction
Humans
i
Epidemilogic
associations
Immune
Elevated Cytokines
Sickness Behavior
Behavioral
Structural
Decreased
hippocampal
neurogenesis
Rodents
More
Depression
emotiona
Figure 5-23. Animal toxicology evidence of possible Pb-dependent
contributors to the development of mood 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) but the specific link between prenatal
Pb exposure, using ALAD as a biomarker, and schizophrenia is just beginning to appear in
the literature and has been described in two publications (Opler et al., 2004; Opler et al.,
2008). Because of this, the animal toxicology literature 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 (Figures 5-23, 5-24, and 5-25). Pb may bind
a divalent cation site in the NMDAR and allosterically inhibit glycine binding
(Hashemzadeh-Gargari & Guilarte, 1999); human studies of patients with schizophrenia
have shown aberrations at this site (Coyle & Tsai, 2004). These findings are consistent with
the glutamatergic hypothesis of schizophrenia which shows that NMDAR noncompetitive
antagonist use in patients with schizophrenia exacerbates their psychotic symptoms and that
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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 (Kempermann et al.. 2008; Reif et al.. 2006). in animal models of
schizophrenia {M, 2007, 670321} and in animal models with developmental Pb exposure
(Jaako-Movits et al.. 2005; Verina et al.. 2007) (Figures 5-23, 5-24, and 5-25). 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 in
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. Thus, a Pb-
dependent contribution to mood disorders exists in the toxicology literature with support
from behavioral, neurochemical and ultrastructural data.
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 5-24. Schematic representation of the contribution of Pb exposure
to the development of a phenotype consistent with
schizophrenia.
As recently reviewed by Wright (2009). social stress and physical environmental toxins impact
overlapping biological processes which determine adaptive plasticity in early neurodevelopment.
Development of CNS organization into functional neuronal and synaptic networks can be determined by
environmental signals which modify neuronalgenesis, synaptic formation and synaptic pruning (LeDoux.
2003). Environmental factors can promote or disrupt this process depending on whether they are positive
(social supports, good nutrition, etc.) or negative (psychosocial stress, chemical toxicants, malnutrition,
trauma, etc.). While plasticity allows recovery from short-term toxic exposures, the neural mechanisms
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1 underlying the plasticity of the developing brain exposed to chronic toxic exposures could induce
2 permanent structural or organizational changes via altered neuronal growth and/or
3 synaptogenesis/pruning. While historically research has focused on how social and physical
4 environmental factors independently affect children's health, evolving theory and methodologies
5 underscore the importance of studying integrated effects (L. D. White et al.. 2007; R. J. Wright. 2009).
6 Recent studies highlight how social conditions influence susceptibility to future environmental exposures
7 and, when contemporaneously exposed, how social-physical environmental interactions may account for
8 more variance in explaining risk than main effects of either factor alone (R. J. Wright. 2009).
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5.3.4. Sensory Acuity
5.3.4.1. Epidemiologic Studies of Children
Although not as widely examined as cognitive and behavioral outcomes, several studies
demonstrated associations of blood Pb level with increased hearing thresholds and decrements in auditory
processing in children (U.S. EPA. 2006V Such evidence has been limited largely to studies described in
the 2006 Pb AQCD, including large U.S. studies, including NHANES II (J. Schwartz & Otto. 1987) and
the Hispanic Health and Nutrition Examination Survey (HHANES) (J. Schwartz & 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). Blood Pb level also was associated with increases in hearing threshold 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]) (Osman et al.. 1999). In the HHANES and Polish studies, associations persisted in analyses
restricted to subjects with blood Pb levels below 10 (ig/dL. In the CLS, blood Pb level was associated
with impaired auditory processing, albeit at higher concentrations. In this cohort, the mean (SD) lifetime
average blood Pb level was 17.4 (ig/dL (8.8), and a 1 (ig/dL increase in lifetime average blood Pb was
associated with a 0.07-point (p <0.05) lower score on the Filtered Word test, indicative of incorrect
identification of filtered or muffled words (Dietrich et al.. 1992). Despite the higher blood Pb levels in the
CLS, the observed associations with auditory function were consistent with those with related endpoints,
including cognitive deficits (Section 5.3.2.1) and behavioral problems (Section 5.3.3.1).
5.3.4.2. Epidemiologic Studies of Adults
Rather than evidence for effect on hearing thresholds, among adults, evidence of association
between blood Pb levels and auditory function comprised changes in auditory evoked brainstem
potentials (U.S. EPA. 2006). Two studies of hearing thresholds came to somewhat different conclusions.
One study examined 183 Pb-workers with blood Pb levels from 1 to 18 (ig/dL and among multiple
frequencies examined, found correlations between increasing blood Pb level and increased hearing
threshold at 4 kHz (Forst et al.. 1997). A second study included 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,
gender, and duration of employment. These findings may indicate that in an occupational setting with
high Pb exposure, any one blood Pb measure may not be the best biomarker for cumulative exposure over
the duration of work.
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Studies published since the 2006 Pb AQCD have produced findings consistent with previous
studies. While studies of Pb-exposed workers continued to dominate, an NAS study provided new
evidence for increasing bone Pb levels being associated with hearing loss in adult males without
occupational Pb exposures (Park et al.. 2010). 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. In models that adjusted for age, smoking,
alcohol consumption, years of noise exposure, as well as Mn, As, and Se levels in blood, blood Pb levels
was associated with a statistically significant higher average hearing threshold (0.5-6 kHz).
Similar findings were reported in a study of 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 (Y.-H. Hwang et al.. 2009). The participants had an
average blood Pb level of 54.3 (ig/dL (SD: 34.6). Average noise levels also were measured in work areas
and dichotomized at 80dB. In analyses adjusted for age and work area noise (dichotomized at 80dB),
workers with blood Pb > 7 (ig/dL had a statistically significant increased odds (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
1-ig/dL.
Park et al. (2010) analyzed data from 448 men in the NAS with an audiometric hearing test within
5 years of bone Pb measurements (all but 5 of these men had audiometric testing before the bone Pb
measurement) and who did not have unilateral hearing loss. 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 was associated with a statistically significant higher
hearing thresholds for frequencies greater than 1 kHz. In analyses of pure tone average hearing loss, a 21
jxg/g increase in patella bone Pb level (interquartile range) was associated with an OR of 1.48 (95% CI:
1.14, 1.91) after controlling for all covariates. Similar, but slightly weaker associations were seen with
tibia bone Pb levels. Audiometric data collected from the same men approximatly 20 years earlier
(median observations/participant: 5; median follow-up: 23 years) were used to assess the association
between tibia bone Pb levels and the change in hearing thresholds over that time. Increasing tibia Pb level
was associated with faster rate of increase in hearing threshold for frequencies of 1, 2, and 8 kHz and a
pure tone average. For the pure tone average, a 15 jxg/g increase in tibia bone Pb level (interquartile
range) was associated with an increased in hearing threshold of 0.05 dB per year. Blood Pb was not
examined in this study. Together with those from studies of Pb-exposed workers, findings from the NAS
study primarily indicate that biomarkers of cumulative Pb exposure are associated with increased hearing
thresholds in adults. Because only bone Pb levels were examined in the NAS study, further investigation
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is required to characterize potential differences between measures of cumulative and recent Pb exposures
in the effects on hearing in adults without occupational exposures.
5.3.4.3. Toxicological Studies of Sensory Organ Function
Pb affects multiple aspects of the nervous system including the sensory organs. The 1986 and 2006
Pb AQCDs detailed the effects of Pb on the retina, CNS visual processing areas, and the auditory system
and described possible or known mechanisms of action where available. The new research examines
effects on these systems at even lower Pb exposures and blood Pb levels.
The Effect of Lead on Vision
The selective action of Pb on retinal rod cells and bipolar cells is well documented in earlier
AQCDs and research in this area continues to date (Fox et al.. 1997; Fox & Sillman. 1979). Pb exposure
during perinatal development and adulthood has also been shown to affect the visual cortex (L. G. Costa
& Fox. 1983) and subcortical neurons (Cline et al.. 1996V The 1986 AQCD mentioned that neonatal rats
exposed to Pb out to PND21, through gestational and lactational exposure, had significant alterations in
visual evoked responses and impaired visual acuity; it was hypothesized that a decreased number of
cholinergic receptors and alterations in the ratio of inhibitory to excitatory systems in the cerebrospinal
axis may be the underlying mechanisms leading to these retinal changes (U.S. EPA. 1986). A 1996
publication detailing environmentally relevant doses of Pb administered to tadpoles showed that Pb
inhibited the growth of developing neurons in the subcortical retinotectal pathway, the main efferent from
the retina (Cline et al.. 1996). The 2006 AQCD evolved to detail Pb-induced impairment of retinal
function in non-human primates as well as focusing on mechanisms of action for specific physiological
retinal changes using both in vitro and in vivo evidence, where available. With Pb-induced retinal effects,
decreased maximal ERG amplitude or sensitivity and increased mean ERG latency was linked to
increased retinal cGMP both in vitro and in vivo. Delayed dark adaptation and increased response
thresholds at scotopic backgrounds were linked to in vivo apoptotic endpoints including rod bipolar cell
death, increased Bax (apoptosis protein) translocation, increased cytochrome c release (apoptosis trigger),
and decreased rhodopsin; in vitro evidence also included retinal apoptosis from the calcium/Pb signal
localized to the mitochondrial permeability transition pore. Other endpoints seen in Pb-induced impaired
retinal function included competitive inhibition of cGMP phosphodiesterase (PDE) in vitro and decreased
stimulated cGMP PDE in vivo; also decreased retinal Na+/K+ ATPase activity has been reported both in
vivo and in vitro. The effects of early life Pb exposure on the retina in monkeys was detailed in work by
Reuhl et al. (1989) in the 2006 AQCD. Chronic Pb exposure from birth to age 6 produced
cytoarchitectural changes in visual projection areas of the brain of rodents; maximum blood Pb level in
the low and high dose group reached 20 (ig/dL and 220|_ig/dL. respectively. Liliental et al. (1988) found
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decreased amplitudes and increased latencies in visual evoked potentials from electroretinograms. Pb
affected amplitude under dark conditions (dark adaptation, B waves affected) and latencies under bright
conditions; blood Pb levels were 40 and 50 (ig/dL in the 350 or 600 ppm Pb groups, respectively. Earlier
work in rodents found that a moderate to high postnatal Pb exposure induced ERG subnormality (Fox &
C'hu. 1988; Fox & Farber. 1988; Otto & Fox. 1993). Thus, the historical animal toxicology literature
shows multiple effects on vision from Pb exposure (Table 5-11, high dose).
Table 5-11. Summary of toxicological studies of Pb on the retina.
High Dose Pb
Low Dose Pb
Gestational (GLE) or Postnatal Pb exposure
Persistent subnormal scotopic ERG
Delayed Dark Adaptation
Decreased thickness of ONL INL (GLE)
Dose dependent decreased DA homeostasis (GLE)
Increased retinal cell apoptosis (postnatal Pb)
No increased rod neurogenesis (GLE)
No increased progenitor cell proliferation (GLE)
Gestational Pb exposure
Persistent scotopic ERG supeanormality
Increased thickness of the ONL
Dose dependent decreased DA homeostasis
Males affected, females not affected
No increase in retinal apoptosis
Increased progenitor cell proliferation
Increased neurogenesis of rods
Larger size of retina
The current literature for this ISA has more work by the Fox lab showing retinal effects in rodents
after human equivalent gestational Pb exposure (GLE, gestationally out to PND10); this is
developmentally equivalent to the in utero retinal development period in humans. Pb exposure during
various developmental windows and at specific doses has been shown to significantly affect
electroretinographs (ERG) both in Pb-exposed humans and rodents (Fox et al.. 2008; Rothenberg et al..
2002) (Table 5-11 and Figure 5-26). Consistent with low dose Pb exposure associated ERG
supernormality in children (Rothenberg et al.. 2002). Fox et al. (2008) found low and moderate ose
gestational Pb-exposure (GLE) induced persistent supernormal scotopic ERGs in rodents. Low and
moderate GLE also induced increased rod and rod bipolar cell neurogenesis (proliferation) and increased
thickness and cell number of the outer and inner neuroblastic layers of the retina (ONL and INL) (Fox et
al.. 2008; Giddabasappa et al.. 2011). Rodents with moderate dose GLE (blood Pb level 25 (.ig/dL) had
27-fold increased retinal progenitor cell proliferation (Giddabasappa et al.. 2011). Extension of the in vivo
studies to isolated cultured cells showed GLE increased and prolonged proliferation of retinal progenitor
cells (Giddabasappa et al.. 2011). Nitric oxide has been shown to regulate retinal progenitor cell
proliferation in chick embryos (Magalhaes et al.. 2006).
Because Pb exposure has been shown to impair NOS activity in other organs (Section 5.2.4.5),
these authors postulate that impaired NO production may be linked to aberrant retinal cell proliferation
(Giddabasappa et al.. 2011). GLE did not significantly affect apoptosis during retinal development
(Giddabasappa et al.. 2011) but it did contribute to increased proliferation of retinal cells. GLE induced
decreased DA synthesis and use in a dose-dependent fashion (Fox et al.. 2008) (Figure 5-27).
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IS 305
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o 200
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A | ¦ Control •
ZH LowPb "jL
¦ Moderate Pb
¦ High Pb |
ill II
low mtensitv High intensity
JtBSt stsmulus >) (test stimulus 21
f3: «0O
5®
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Note: *p <0.05
High intensity
(test stimulus 21
1,00
low intensity
HMl fit Ml MM
(test stimulus I)
Itert stimulus 21
Source: Fox et al. (2008)
Figure 5-26. Retinal a-wave and b-wave ERG amplitude in GLE adult males.
Dark-adapted
Light-adapted
Control
Low Moderate
GLE
Low Moderate
GLE
0.35 _
0.30
0.25
0.20 «
... 4/5
-0.15 o
DOPAC/DA
HvA/DA
£ 0.9-
« 0.3
Control
Low Moderate
GLE
Control
Low Moderate
GLE
Note: *p <0.05
Source: Fox et al. (2008)
Figure 5-27. Retinal dopamine metabolism in adult control and GLE rats.
1 Many outcomes in this study showed an inverse U Pb dose-response curve as is shown with the
2 high dose exposures having vastly different effects from low dose GLE. GLE exposure at high doses
3 produced ERG subnormality, rod cells loss, and decreased rod neurogenesis (Fox et al.. 2008) (Table 5-
4 11). The high dose GLE rodents showed dose-dependent decreased DA synthesis and use (Figure 5-27).
5 These new animal toxicology data confirm the epidemiologic data showing ERG supernormality at low
6 dose GLE. They provide further insight into retinal changes showing increased proliferation of Pb-
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exposed retinal progenitor cells without changes in apoptosis. GLE induced dose-dependent decrements
in retinal DA use and synthesis.
Lead-Induced Auditory Effects
The 2006 Pb AQCD mentioned auditory effects on non-human primates who were exposed to Pb
throughout gestation and out to age 8-9 (blood Pb levels 33-56 (ig/dL during Pb exposure period).
Auditory evoked potentials, 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 level had
returned to baseline levels. In Pb-exposed animals, half of the pure tone detection thresholds were outside
of the control range at certain frequencies (Rice. 1997). which is consistent with data from humans
developmentally exposed to Pb. Thus, these authors found that early life Pb exposure impaired auditory
function. The cochlear nerve in both developing and mature humans appears to be especially sensitive to
the Pb insult. At low to moderate Pb exposures, elevated thresholds and increased latencies are seen in
brainstem auditory evoked potentials. There is coherence between the animal and the human literature on
the effects of chronic Pb exposure on auditory function.
5.3.5. Neurodegenerative Diseases
5.3.5.1. Epidemiologic Studies of 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. Two NAS studies found
associations between increasing bone Pb levels and decreasing MMSE scores ("Weisskopf et al. 2004; R.
Q. Wright et al.. 2003). which pointed to an association with dementia, given that the MMSE is widely
used as a screening tool for dementia. Overall, studies had sufficient limitations, and findings were
inconclusive ("U.S. EPA. 2006). New studies on dementia are not available to assess further the
association of Pb biomarkers with dementia. Similarly, new studies examining Alzheimer's disease are
not available, and as in 2006, the evidence is inconclusive regarding the association between Pb
biomarkers and Alzheimer's disease. 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. The 2006 Pb AQCD reviewed two case-
control studies that measured blood Pb levels. One study found no difference in mean blood Pb levels
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between the 16 ALS cases (mean: 12.7 (ig/dL) and 39 controls (mean: 10.8 (ig/dL) (Vinceti et al.. 1997).
Another study that examined blood and bone Pb levels in a New England-area population found increased
odds of ALS among subjects with blood Pb levels > 3 (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
elevated odds of ALS; however, associations were not statistically significant. In analyses of Pb
biomarkers as continuous variables, odds ratios for all three biomarkers were similar; however, only
associations for blood Pb level were statistically significant. Kamel et al. (2002) also found that an
estimate of cumulative Pb exposure based on occupational history to be significantly associated with ALS
(Kamel et al.. 2002). Thus, the stronger findings for blood Pb level were surprising given that bone Pb
level is a better biomarker of cumulative Pb exposure than is blood Pb level. 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, additional studies have been conducted in the New England-area case-
control study. One study indicated that the association between blood Pb level and ALS was not modified
by the ALAD genotype (Kamel et al.. 2005). Another study examined survival with 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, however, 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. However, because the strongest
findings for survival were found for tibia Pb, it might be expected a bias would be most apparent in a
study examining associations of tibia Pb levels and ALS incidence. However, this was not observed in the
one study that had bone and blood Pb biomarkers (Kamel et al.. 2002).
Another case-control study examined blood Pb levels and odds of ALS among 184 cases (33 were
either progressive muscular atrophy or primary lateral sclerosis, mean blood Pb level: 2.41 (ig/dL) and
194 controls (mean blood Pb level: 1.76 (ig/dL) (Fang et al.. 2010). The cases were recruited from the
National Registry of U.S. Veterans with ALS, and controls were recruited from among U.S. Veterans
without ALS 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 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
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turnover suggests that reverse causation is not likely driving the association between blood Pb level and
ALS. Whether other types of reverse causality are occurring, 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, several studies have found associations of blood and bone Pb levels with ALS;
however, the issues of reverse causality and bias due to survival time make it difficult to draw firm
conclusions.
Parkinson's Disease
A few studies, some ecological (Aquilonius & Hartvig. 1986; Rvbicki et al.. 1993) and some case-
control relying on questionnaire data or occupational history (Gore 11 et al.. 1997; Gulson et al.. 1999;
Tanner et al.. 1989) have indicated associations between exposure to heavy metals, particularly Pb, and
risk of PD, although the evidence is limited and far from conclusive. Coon et al. (2006) expanded on
earlier studies that had found more than a twofold increased risk of PD among adults occupationally-
exposed to Pb for more than 20 years (Gore 11 et al.. 1997; Gulson et al.. 1999) by examining associations
with bone and whole-body Pb levels. Coon et al. (2006) included 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 jj.g/g) and
calcaneus (OR: 1.50 [95% CI: 0.75, 3.00] for levels > 25.29 jj.g/g) bone Pb levels were at elevated odds of
PD compared to those in the lowest quartile. The highest OR for PD was estimated for subjects in the
highest quartile of whole-body lifetime exposure to Pb, compared to the lowest quartile of exposure (OR:
2.27 [95% CI: 1.13, 4.55] for levels > 80.81 jag/g).
The second study to explore the association between biomarkers of Pb and PD was published
recently (Weisskopf et al. 2010). This study was based in the Boston, MA area and had more than twice
as many cases as Coon et al. (2006): 330 cases and 308 controls. Subjects in the highest quartile of tibia
Pb level (>16.0 jj.g/g) had elevated odds of PD compared to those in the lowest quartile (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 towards
the null as these groups are likely to share many exposures.
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Essential Tremor
The 2006 Pb AQCD described two studies that found associations between blood Pb levels and
odds of PD in New York City-based populations (Louis et al.. 2005; Louis et al.. 2003). Since then, 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). Blood Pb
levels in this study were comparable to those found in the earlier New York based studies: the median
blood Pb was 2.7 (ig/dL among cases and 1.5 (ig/dL among controls. After adjusting for age, sex,
education, cigarette smoking (yes versus no), cigarette pack-years, and ethanol use (yes vs. no), a 1 (ig/dL
increase in blood Pb level was associated with more than a fourfold increase in odds of essential tremor
(OR: 4.19 [95% CI: 2.59, 6.78]). This OR was much larger than that obtained in the New York 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 occupation that could have contributed to Pb exposure (possibly stored in bone from where it
could have contributed to blood Pb levels at the time of the study) was not assessed. One of the earlier
New York studies also found a very high OR for essential tremor among ALAD2 carriers (per 1 (ig/dL
increase in blood Pb level, OR: 71.8 [95% CI: 1.08, 4789.68); however, the very wide 95% CI indicated
lack of precision in the effect estimate.
Several studies of essential tremor have reported a very strong association between blood Pb levels
and essential tremor, although studies employing biomarkers of Pb have had relatively small sample sizes
and have produced imprecise effect estimates.
5.3.5.2. Toxicological Studies
In epidemiological studies, Pb level (bone and blood) is associated with increased odds of ALS
(kamel et al.. 2005) and paradoxically with longer survival time in patients diagnosed with ALS (Kamel
et al.. 2008). Chronic Pb exposure (Pb acetate in drinking water at 200 ppm from weaning onward, 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 does not significantly increase the onset of the ALS disease in SOD1 Tg mice, but Pb exposure
was associated with longer survival time in SOD1 Tg mice (Barbeito et al.. 2010). Baseline levels of
VEGF are elevated in astrocytes from the ventral spinal cord of untreated SOD1 Tg control mice versus
control, non-transgenic animals. VEGF was not induced in Pb treated non-transgenic astrocytes. Further,
Pb-exposed SOD1 Tg mice had significant elevations of astrocyte VEGF versus vehicle treated SOD1
animals (Barbeito et al.. 2010). Control animals exposed to Pb showed no elevation in VEGF expression
above control vehicle-treated animals (Barbeito et al.. 2010). Other research has suggested that ALS
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initiation is dependent on motor neuron function and ALS progression is dependent on astrocyte and
microglia function (Boil lee et al.. 2006; Yamanaka et al.. 2008).
Consistent these findings, other 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 is 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 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 are 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 (L. Oian & Flood. 2008; D. Zhang et al.. 2010). 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 could induce cell death or apoptosis in various
models including rat brain (Tavakoli-Nezhad et al.. 2001). retinal rod cells (L. H. He et al.. 2000; 2003).
cerebellar neurons (Oberto et al.. 1996). and PC12 cells (Sharifi & Mousavi. 2008). This study addresses
chronic (40 days) Pb exposure-induced hippocampal apoptosis in young (exposure starting at 2-4 weeks
of age) and old (exposure starting at 12-14 weeks of age) male rats exposed to 500 ppm Pb by drinking
water (blood Pb, 98 (ig/dL); apoptosis was verified by light and electron microscopy, and increased pro-
apoptotic Bax protein levels (Sharifi 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 at the end of lactation or PND20 was 80 (ig/dL.
These data showed 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.. 2007). The cerebellum did not have elevated apoptotic mRNA levels in this model. This
study shows temporal and regional changes in activation of death protein message levels in male
offspring. Thus, the new data continue to confirm that Pb exposure induced apoptosis in brains of exposed
animals.
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Lead-Induced Neuronal Plaque Formation
Studies from the 2006 Pb AQCD highlighted the importance of Pb exposure in early life in
promoting Alzheimer's like pathologies in the adult brain. Pb has been recognized for decades as having a
more profound effect on children than adults who receive the same exposure. In the last decade, the
developmental origins of adult disease (DoAD) paradigm and the similar Barker hypothesis have reported
that early life exposures can result in aberrant adult outcomes. Epidemiology data show that blood Pb
level effects on neurological outcomes fit this paradigm. Recent evidence in the toxicological literature
points to latent effects in rodent and non-human primate models of gestational and/or early life exposure
to Pb including neurodegeneration similar to Alzheimer's like pathologies, obesity (in males only), retinal
aberrations (male only), and neurobehavioral aberrations. Immune outcomes fitting the DoAD paradigm
involve tissue inflammation and loss of organ function. 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
(2008) 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 xenobiotics
rather than relying on adult exposure information to predict probable health risks from prenatal, neonatal
or juvenile exposure (Dietert & Piepenbrink. 2006). Mechanistically, some of these pathologies have been
associated with changes in the epigenome. Multiple Pb studies point to sensitive windows of exposure;
early life or developmental exposures are far more sensitive than adult exposures. Alzheimer's disease is
characterized by amyloid-beta peptide (Ab) accumulation, hyper-phosphorlyation of the tau protein,
neuronal death and synaptic loss. The toxicological evidence for Pb in contributing to the AD pathology
follows.
The fetal basis of amyloidogenesis has been explored extensively by Zawia's laboratory in both
rodents and non-human primates. Amyloid deposits in the brain are seen in patients with Alzheimer's
disease and in the aged brain of individuals with Down Syndrome who display an Alzheimer's-like
pathology. Mechanistically, amyloid plaques originate from the cleavage of the amyloid precursor protein
(APP) to Ab, which comprises the plaque. By exposing rodents to Pb as neonates or as aged animals, it
was determined that neonatal Pb exposure is a sensitive window for induction of the amyloidogenesis in
the aged animal brains and that exposure to Pb in old animals did not contribute to plaque formation.
Following cortical APP gene expression over the lifetime of male rodents exposed neonataly via lactation
to Pb (PND1-PND20 exposure, dam drinking water Pb acetate 200 ppm, pup PND20 blood Pb level 46
(ig/dL and cortex 0.41 jj.g/g wet weight of tissue), one sees a biphasic significant increase above control
animals in APP expression with the first increased phase manifesting neonatally and the second phase
manifesting in old age (82 weeks of age) (M.R. Basha et al.. 2005). A concomitant biphasic response is
seen in specificity protein 1 (Spl), a transcription factor known to be related to APP expression. Ab, the
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amyolid plaque constituent, was also significantly elevated in these aged animals developmentally
exposed to Pb. A separate subset of rodents exposed to Pb only as aged adults (18-20 weeks of age) were
unresponsive in APP or Spl expression or Ab production after Pb exposure, indicating the developmental
window and not adult exposure as the susceptible period for Pb-dependent amyloidogenesis. The Zawia
lab (J. Wu et al.. 2008) has confirmed similar amyloid findings in brains of monkeys that were exposed to
Pb as infants (PND1-PND400), i.e., significantly higher message levels 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 (Rice. 1990. 1992). After weaning and
when still being dosed with Pb, the monkeys had blood Pb level of 19-26 (ig/dL. As a caveat, by the time
neonatally Pb-exposed animals become aged and manifest with amyloid plaques, blood Pb level and brain
cortex Pb levels have returned to control or baseline levels. Thus, the rodent and non-human primate
toxicology studies concur and show that developmental Pb exposure induced elevations in neuronal
plaque proteins in the aged animals.
Transcription factors are essential in the regulation of the developing brain. Pb exposure is known
to perturb DNA binding of transcription factors including SP1 at essential sites like zinc finger proteins.
In Long-Evans hooded rat pups exposed during lactation, these Pb-induced developmental perturbations
of SP1 DNA binding can be ameliorated by exogenous zinc supplementation (M. R. Basha et al.. 2003).
The mechanism by which Ab peptide formation is affected by Pb exposure has been explored by
Behl et al. It has been shown that the choroid plexus is able to remove beta-amyloid peptides from the
brain extracellular matrix and that Pb impairs this function and may be mediated by the
metalloendopeptidase, insulin-degrading enzyme (IDE), which metabolizes Ab (Behl et al.. 2009).
Further studies with developmental Pb exposure (gestational plus lactational, dam drinking water
solutions of 0.1%, 0.5% or 1%, blood Pb level 400, 800 and 1,000 |_ig/L) showed that the hippocampus
contained neurofibriallary 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.. 2010). The multiple new studies on Pb-dependent changes in the neurofibrillary proteins show
that developmental Pb exposure induced significant increases in neuronal plaque associated proteins,
indicating that early life Pb exposures may contribute to dementia in adulthood.
Data from the animal toxicology literature point to an early life window in which Pb exposure can
contribute to pathological brain changes consistent with those seen in Alzheimer's disease including Ab
peptide accumulation and activation of Ab supporting transcription factors as well as tau
hyperphosphorylation.
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5.3.6. Studies of Mechanisms of the Neurological Effects of
Lead
5.3.6.1. Effects on Brain Physiology and Activity
A growing body of epidemiologic evidence demonstrates associations of Pb biomarkers with
electrophysiologic changes in the brain. By providing insight into the underlying mechanisms by which
Pb exposure may disrupt brain function, findings from these studies have provided biologically plausible
evidence for the effects of Pb exposure on cognitive, psychological and behavioral consequences
observed in children and adults. Much of the early work was conducted by Otto and colleagues (Otto et
al.. 1985; Otto & Fox. 1993). which found associations of blood Pb level with auditory and visual evoked
potentials. Rothenberg et al. (1994) 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 a decrease in interpeak intervals in auditory evoked
potentials at age 5-7 years. Prenatal blood Pb level showed a biphasic relationship, with a negative
association at blood Pb levels of 1-8 (ig/dL and a positive association at blood Pb levels of 8-30 (ig/dL.
These findings provide mechanistic support for observations of Pb-associated changes in sensory acuities
(Section 5.3.4.1).
Studies using magnetic resonance imaging (MRI) or spectroscopy (MRS) as clinical outcome
measures have been limited in number and sample size, but have shown associations of blood Pb level
with alterations in brain physiology such as reduced levels of N-acetylaspartate, creatine, or choline in
young adults (Cecil et al.. 2005; Meng et al.. 2005; Trope et al.. 2001; Trope et al.. 1998; Yuan et al..
2006). These changes have been linked to decreased neuronal density or loss and alteration in myeline.
Notably, Trope et al. (2001; 1998) 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 is unclear. Additionally, these studies compared subjects with relatively high childhood blood
levels (23-65 (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 mean childhood 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 associations of blood Pb level with other indices of
language in childhood.
Since the 2006 Pb AQCD, studies examining MRI data have been limited to CLS cohort
participants as adults (ages 19-24), and recent results continue to support associations of childhood blood
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Pb levels with physiological changes in the brain in adulthood. These recent studies expanded on previous
studies by including larger sample sizes and aiming to characterize Pb effects more precisely by
examining blood Pb levels at different periods in childhood and aiming to link changes in brain activity to
neurodevelopmental deficits. Brubaker et al. (2009) used diffusion tensor imaging to examine
associations between mean childhood blood Pb and white matter integrity in 91 young adults,
hypothesizing that childhood Pb exposure may alter adult white matter architecture via deficits in axonal
integrity and myelin organization. Fractional anisotropy (FrA), mean diffusivity (MD), axial diffusivity
(AD), and radial diffusivity (RD) were measured on an exploratory voxel-wise basis. In adjusted
analyses, mean childhood blood Pb levels were associated with decreased FrA throughout white matter.
Regions of the corona radiata demonstrated highly significant Pb-associated decreases in FrA and AD and
increases in MD and RD. The genu, body, and splenium of the corpus callosum demonstrated highly
significant Pb-associated decreases in RD, smaller and less significant decreases in MD, and small areas
with increases in AD. The results of this analysis suggest multiple insults appear as distinct patterns of
white matter diffusion abnormalities in the adult brain which may be indicative of altered myelination and
axonal integrity. Additionally, childhood blood Pb levels appear to differentially affect neural elements,
which 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 and by examining associations between
MRI-assessed brain volume and historic neuropsychological assessments, provided insight into the
potential clinical significance of changes in brain physiology associated with blood Pb levels (Cecil et al..
2008). Using conservative, minimum contiguous cluster size and statistical criteria (700 voxels,
unadjusted p <0.001), approximately 1.2% of the total gray matter was significantly and inversely
associated with mean childhood blood Pb level. 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). Comparing neuropsychological
factor scores with gray matter volume, 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 are notable in light of other studies linking brain
volume changes with altered function and suggest that MRI changes association with blood Pb levels may
be indicative of decrements in neurocognitive and neurobehavioral function. Schwartz and colleagues
(2007) showed that larger RO1 volumes were associated with better cognitive function in 5 or 6 cognitive
domains (visuoconstruction, processing, speed, visual memory, executive functioning, and eye-hand
coordination). More recent studies by Raine and colleagues suggest that deficits in cortical volume or
activity found in select brain regions, including the prefrontal gray matter, may predispose individuals to
impulsive, aggressive, or violent behavior.
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In a subsequent analysis, Brubaker et al. (2010) investigated the developmental trajectory of
childhood blood Pb levels on adult gray matter. Adjusted voxel-wise regression analyses were performed
for associations between adult gray matter volume loss and yearly mean blood Pb levels from 1 to 6 years
of age in the entire cohort and by sex. Investigators observed significant inverse associations between
gray matter volume loss and yearly mean blood Pb levels from 3 to 6 years of age. The extent of
prefrontal gray matter loss associated with yearly mean blood Pb levels increased with advancing age of
the subjects. These results indicate that blood Pb levels later in childhood are associated with greater
losses in gray matter volume than are childhood mean or maximum blood Pb levels. This study
demonstrates that maximum blood Pb levels do not fully account for gray matter changes, particularly in
the frontal lobes of young men. Notably, although they did not consider Pb, Yang et al. (2005) reported
volume reduction in gray matter in psychopaths, adding additional evidence that these physiological
changes may be related to overt deleterious outcomes. Consistent with Wright et al. (2008). Cecil et al.
(2008) found that gray matter volume loss associated with childhood blood Pb levels was much larger in
CLS male adults than female adults. In an expanded analysis of the developmental trajectory of childhood
blood Pb levels on adult gray matter, Brubaker et al. (2010) found that the inverse associations between
gray matter volume loss and yearly mean blood Pb levels were more pronounced in the frontal lobes of
men than women for blood Pb levels measured at all ages.
Whereas the aforementioned CLS studies examined associations of childhood blood Pb levels, a
recent analysis of the NAS participants indicated that biomarkers of cumulative, long-term Pb exposure
also may be associated with changes in brain structure and function in older adults. Weisskopf et al.
(2007) found increasing bone Pb level to be associated with increased myoinositol (mI)/Cr ratio with
increasing bone Pb concentration among 31 elderly men from the NAS. A higher ml/Cr ratio may be
indicative of glial activation and is a signal reportedly seen in the early stages of HIV-related dementia
and Alzheimer's disease.
Studies of Pb-workers also found associations of blood and bone Pb levels with changes in brain
structure and physiology. Stewart et al. (2006) studied 532 former organolead workers with a mean age of
56 years and found that an estimate of past peak tibia bone Pb (mean: 23.9 jj.g/g) was significantly
associated with more white matter lesions (WML). Higher estimated peak tibia Pb also was associated
with smaller total brain volume and volumes of frontal and total gray matter, parietal white matter, the
cingulate gyrus, and insula. In this same group, Caffo et al. (2008) found evidence that the association
between tibia Pb level and cognitive function, in particular, visuo-construction domain tasks, and to a
lesser degree, executive function and eye-hand coordination, were mediated by the association between
tibia Pb levels and brain region volume changes. In a similar study of 61 current Pb smelter workers with
an average age of 40 years, higher estimates of cumulative Pb exposure were also associated with WML,
and there was evidence that this association accounted for, in part, the association between higher
cumulative Pb exposure and worse performance on the grooved pegboard test (Bleecker. Ford. Vaughan.
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et al.. 2007). Another small study of 15 current occupationally Pb-exposed workers (mean blood Pb level:
63.5 (ig/dL) and 19 non-Pb exposed controls (mean blood Pb level: 8.74 jj.g/dl) found smaller
hippocampal volumes in MR I scans among the Pb workers (Jiang et al.. 2008). A lower ratio of the brain
metabolites N-acetyl-aspartate (NAA) and creatine (Cr), indicative of neuronal density, was also found
among the Pb workers, as well as an increased lipid to creatine ratio, indicative of cell membrane or white
matter damage. Similarly, a study of 22 Pb paint factory workers and 18 controls found lower NAA/Cr
ratios among the Pb exposed workers as well as lower choline/Cr, possibly indicative of reduced cell
membrane turnover (Hsieh et al.. 2009).
5.3.6.2. 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 manifested with Pb-induced cholesterogenesis and phospholipidosis in brain tissue (Ademuviwa et
al.. 2009). Pb-dependent changes in brain cholesterol produced an inverse U dose response curve with the
highest brain cholesterol at 200 ppm 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 (Kojima 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 non-significant 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 dose-dependent fashion; 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 at 0, 200, 300, and 400 ppm were 7, 41, 61, and 39 (ig/dL, respectively. In summary, 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
this Pb-dependent phospholipidosis and cholesterogenesis.
5.3.6.3. 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 biomarker 8-hydroxy-2'-
deoxyguanosine (8-oxo-dG). The contribution of reactive oxygen or nitrogen species to these Pb induced
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changes was assayed by looking at the ratio of 8-oxo-dG to 2-deoxyguanosine (2-dG). 2-dG is a DNA
nucleoside enzyme that can generate 8-oxo-dG from a parent compound forming a DNA adduct or
biomarker during conditions of nitrosative or oxidative stress. The 8-oxo-dG to 2-dG ratio data from
rodent male offspring recapitulated 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, non-Pb 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, the monkey data were the same as the rodent data. The
ratio of 8-oxo-dG to 2-dG in the brains of aged monkeys (23 years) after being exposed as infants, was
significantly elevated above controls (X Wu et al.. 2008). Similar to the amyloid data, the oxidative stress
markers showed no significant changes above baseline when animals were exposed to Pb as aged adults
(Bolin et al.. 2006; J. Wu et al.. 2008). 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.
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 (Regunathan & Sundaresan. 1984; Sterling et al.. 1982; Verma et al.. 2005;
Yun & Hover. 2000). 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 neuotransmission. The
mitochondrial Na/K ATPase is important in maintaining the inner mitochondrial membrane potential
(delta omega sub m) and the health of the mitochondria. To address the effect of Pb exposure on these
mitochondrial parameters, mice were mated, produced offspring and nursed the offspring until PND8 at
which time the brains were collected from the pups (Baranowska-Bosiacka et al.. 2011). Cerebellar
granular cells were harvested from these PND 8 control and Pb-exposed animals (0.1% Pb acetate in dam
drinking water, blood Pb level 4 (ig/dL and cerebella Pb 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 control. Neuronal Na/K ATPase activity was significantly
lower in cortical granule cells from Pb-exposed mice versus controls. Mitochondrial mass was unaffected
with Pb treatment, but mitochondrial membrane potential was significantly decreased with Pb exposure.
Pb-exposed crayfish who are placed under hypoxic conditions adapt to the situation by decreasing their
metabolism (Morris et al.. 2005). manifesting with whole organism findings consistent with these cell
data. These data show impaired mitochondrial function and energy production in neuronal cells from mice
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with gestational and lactational Pb exposure with concomitant increases in mitochondrial and cellular
ROS production.
5.3.6.4. 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 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 nNOS with perinatal Pb exposure. Namely, with repeated swim
tests, control animals more quickly find a submerged platform, i.e. escape, than do Pb-exposed animals.
After either 4 or 8 weeks of Pb exposure to weanling male rats (blood Pb level 0.3 umol/L), hippocampal
NOS and NO are significantly decreased. Dietary supplementation concomitant with 8 weeks of Pb
exposure induced significant increases in hippocampal NOS (taurine or glycine) and decreases in
hippocampal NOS (vitamin C, methionine, tyrosine, or vitamin Bl). These animals also had significant
changes in hippocampal NO with supplementation. NO increased with taurine and decreased with vitamin
C, tyrosine or glycine co-exposure with Pb. Dietary supplementation after 4 weeks of Pb exposure in
weanling males (4 week blood Pb level 2.3 umol/L & 8 week Blood Pb level 0.39 umol/L), induced
significant increases in NO with the supplements tyrosine, methionine, or ascorbic acid. Zinc
supplementation in this model had no effect on the NO system. The conclusions of this study are that
various combinations of nutrients significantly attenuate Pb-dependent 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-dependent decrements in NO/NOS concentrations (G. Fan et al.. 2009).
5.3.6.5. Synaptic Changes
Work in earlier criteria documents as well as earlier publications in the scientific literature point to
an effect of developmental Pb exposure on synapse development, which mechanistically may contribute
to multiple Pb related aberrant outcomes, including changes in long-term potentiation (LTP) and
facilitation. Earlier work has shown that developmental Pb exposure is responsible for altered density of
dendritic hippocampal spines (Kiraly & Jones. 1982; Petit & LcBoutillier. 1979). aberrant synapse
elimination (Lohmann & Bonhoeffer. 2008) and abnormal long-term and short-term plasticity
(Mac Donald et al.. 2006). Newer research using the Drosophila larval neuromuscular junction model has
shown that stimulation with multiple action potentials (also called AP trains) induced significant increases
in intracellular calcium and significant delays in calcium decays back to baseline levels at the pre-
synaptic neuronal bouton in developmentally Pb exposed larvae versus control. 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 (T. He et al.. 2009). Intracellular calcium in Pb exposed larvae was no
different from 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 ability 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 (T. 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. (2009) focused on inflammatory endpoints and synaptic changes after
gestational plus lactational dam drinking water Pb exposure (solutions of 0.1%, 0.5% or 1%, offspring
blood Pb level 40, 80 and 100 (ig/dL, respectively at PND 21). Hippocampal TNF-alpha 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.. 2009). Thus, Li et al. (2009)
found significant difference in hippocampal synaptic protein composition and increased pro-inflammatory
cytokine levels in the brains of Pb-exposed offspring.
Neurotransmission is an energy-dependent process with calcium-dependent ATP releases found at
the synaptic cleft. At the synapse, ATP is metabolized by ecto-nucleotidases. In heme synthesis, Pb is
known to substitute for the cation zinc in another nucleotidase, pyrimidine 5'-nucleotidase, and is thus
used as a biomarker of Pb exposure. Acute exposure (96h) of male and female zebrafish to Pb acetate (20
(ig/L) in their water induced significant decreases in ATP hydrolysis in brain tissue. This dose is deemed
to be an environmentaly relevant dose. With a chronic exposure (30 days), Pb acetate promoted the
inhibition of ATP, ADP and AMP hydrolysis; these data are consistent with findings in rodents
(Baranowska-Bosiacka et al.. 2011). The authors hypothesize that at 30 days, this change in nucelotide
hydrolysis was likely due to post-translational modification because message level of enzymes
responsible for the hydrolysis, NTPDasel and 5'-nucleotidase, were unchanged (Senger et al.. 2006).
Thus, Pb is shown to affect nucleotidase 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
environmentally relevant concentrations of PbAcetate (20 ug/L water), AChE activity was significantly
inhibited in zebrafish brain tissue. In Pb-exposed fish, AChE activity returned to baseline by 96 hours and
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maintained baseline activity after chronic exposure of 30 days. Thus, Pb is also able to affect synaptic
homeostasis of AChE in the brain of exposed zebrafish (Richetti et al.. 2010).
Pb is known to act as an antagonist of the NMDA receptor. The NMDAR is essential for proper
pre-synaptic 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.. 2010). 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 is
seen in both GABAergic and glutamatergic neurons. Pb also induced an increase in number of
presynaptic contact sites. But, these sites may be non-functional as they lack the protein receptor
complexes necessary for proper vesicular exocytosis. Another factor involved in growth and signaling of
pre-synaptic neurons is BDNF, which is synthesized and released by post-synaptic neurons. BDNF is
regulated by the NMDAR and acts in a retrograde fashion, participating in pre-synaptic maturation. In
this model, both pro-BDNF and BDNF release were significantly attenuated with Pb exposure. Further,
exogenous BDNF administration rescued the aforementioned Pb-dependent pre-synaptic effects. Thus,
this cell culture model shows that Pb-dependent pre-synaptic aberrations are controlled by NMDAR-
dependent BDNF effects on synaptic transmission.
5.3.6.6. 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 cerebrospinal fluid barrier
(BCB). The blood brain barrier, 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 is known to affect the blood brain barrier even at low Pb
concentrations resulting in increased permeability (Dvatlov etal.. 1998; Moorhouse et al.. 1988;
Struzvnska. Walski. etal.. 1997; Sundstrom et al.. 1985). Because the BBB is under-developed early in
life, prenatal and perinatal Pb exposure results in higher brain Pb accumulation than does similar
exposures later in life (Moorhouse et al.. 1988). Earlier AQCDs have shown that the chemical form of Pb,
and its ability to interact with proteins and other blood components affects its ability to penetrate the BBB
(U.S. EPA. 2006). Pb compromises the function of the BCB, and decreased the CSF level of transthyretin,
a thryoid binding protein made in the choroid plexus. The choroid plexus and cerebral endothelial cells
that form these BBB and BCS tight junctions are known to accumulate Pb more than other cell types and
regions of the CNS. Hypothyroid status can contribute to impaired learning and IQ deficits. More recent
research with weanling rats fed Pb through Pb acetate drinking water exposure manifested histologically
with leaky cerebral vasculature as detected with lanthum nitrate staining of the brain parenchyma, an
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indication of BBB impairment, that was ameliorated or that 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 (0. Wang et al.. 2007). These data demonstrate that Pb induced a leaky BBB in
weanling rats with associated decreases in the junctional protein occludin; dietary supplementation with
iron was able to ameliorate these Pb-induced impairments of the BBB in male rats.
5.3.6.7. 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 contribute to
nervous system barrier function disruption, tissue development dysregulation, immune dysfunction, and
affect learning and memory (Prozialeck et al.. 2002).
5.3.6.8. 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,
strucutral support to neurons, and myelination, among other effects. Glial cells are known 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 a dose of 0.25 ± 1.0 |_iM
Pb acetate via Pb-dependent reduction in glutamine synthestase activity in the astroglia; astroglia take up
glutamate after its release and convert it to glutamine. Pb causes hypomyelination and demyelination (F.
Coria et al.. 1984) mediated through the oligodendrocytes with younger animals being more susceptible to
the effects of (Tiffanv-Castiglioni et al.. 1989). Unfortunately, Pb accumulation in young glial cells may
contribute to a lifelong exposure to this Pb sink in the brain as it is released over time where it can
damage 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) performed in vitro patch clamp monitoring of rat hippocampal CA1
section LTPs collected from pups exposed to Pb Acetate in utero, lactationally and through drinking water
out to PND28. D-serine supplementation relieved the chronic Pb exposure dependent impaired magnitude
of hippocampal LTP (H. Sun et al.. 2007). which is known to be regulated by the NMDAR (Bear &
Malenka. 1994). The use of 7-chlorokynurenic acid, an antagonist of the glycine binding site of the
NMDAR-the binding site of D-serine, effectively abolished D-serine's rescue of the LTP. NMDAR-
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independent LTP hippocampal neurotransmission with slices from Pb-exposed mossy-CA3 synapses was
not rescued by exogenous D-serine supplementation. These data indicate that glial transmission may
provide promising therapeutic targets for intervention after Pb exposure or with other affective or
cognitive disorders known to manifest with aberrant NMDAR-dependent neurotransmission.
5.3.6.9. 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 signals
for its homeostasis. Pb-dependent alterations in neurotransmission are discussed in further detail below.
Monoamine Neurotransmitters and Stress
Combined exposures of maternal stress and Pb exposure can synergistically enhance behavioral
and neurotoxic outcomes in exposed offspring, sometimes even potentiating an effect that would
otherwise be sub-threshold. Virgolini et al. (2008) found that effects on the central nervous system by
developmental Pb exposure (2 months prior to mating through lactation, 50 or 150 ppm Pb acetate
drinking water exposure, blood Pb level 11 (ig/dL and 35 (ig/dL, respectively) are enhanced by combined
maternal and offspring stress. Offspring neurotransmitter concentrations were significantly affected with
Pb exposure, but the most interesting findings were those of potentiated effects or effects that were not
seen with Pb exposure alone or stress alone. These potentiated effects were only seen 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 no stressors and Pb exposure. Thus, Pb alone did not significantly
affect 5HT levels. 5HIAA concentration was significantly increased with Pb exposure alone in the
striatum of male offspring at 150 ppb Pb exposure; with the remaining exposures, 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.
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Monoamine Neurotransmitters and Auditory Function
The monoamine neurotransmitters include DA, 5HT, and NE. Earlier work has shown that perinatal
rat Pb exposure induced increased tyrosine hydroxylase, increased DA and increased cerebral cortex
catecholamine neurotransmission (Bielarczvk et al.. 1996; C. B. Devi et al.. 2005; Leret et al.. 2002).
Earlier publications detailing the window of exposure, duration of exposure and dose of Pb used have
varying effects on monoamine transmitters. In more recent work, these neurotransmitters, among others,
have been implicated in auditory function in the brainstem in various integration centers there including
the lateral superior olive (LSO), and the superior olivary complex (SOC). The SOC is vital for sound
detection in noisy settings among other functions. Low level Pb exposure has been associated with altered
processing of auditory temporal signals in animal studies (Finkelstein etal.. 1998; Lurie et al.. 2006).
Because Pb alters auditory processing, the monoamine system is a potential target for Pb-mediated
interactions. Blood Pb levels for control, very low Pb (VLPb) and low Pb (LPb) exposure groups are 1.4,
8.0, and 42.2 (ig/dL, respectively. Developmental Pb 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-hyrdoxylase (DbH), a marker for NE. This immunochemistry was significant
for both VLPb and LPb exposure for VMAT2 and DbH but 5HT only had significant decrements 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 relating to auditory integration with interaction from the monoamine neurotransmitter system
are affected by developmental Pb exposure (Fortune & Lurie. 2009).
Dopamine
The 2006 AQCD detailed low dose Pb-dependent 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 dopamine (DA) metabolism, DA and DOPAC, a DA metabolite.
Thus, these were measured in various brain regions of year old males to determine if GLE affected DA
metabolism. Low and high dose GLE in male rodents induced significant elevations in the DOPAC to DA
ratio, and DOPAC concentration in the forebrain. In the forebrain, DA was significantly decreased in low
dose GLE males and significantly elevated in GLE high dose males compared to controls. In the striatum,
DOPAC was significantly elevated in both low and high dose GLE exposed 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 control. These new data expand upon the monoamine literature base which
reports perinatal rat low concentration Pb exposure induced increased sensitivity of the dopamine
receptors (D2 and D3) (Corv-Slechta et al.. 1992; Gedeon et al.. 2001). produced higher DA levels (C. B.
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Devi et al.. 2005; Leret et al.. 2002). and enhanced catecholamine neurotransmission in the cerebral
cortex, cerebellum, and hippocampus (C. B. Devi et al.. 2005).
The interaction of dopamine and the nitric oxide system in the striatum was studied after prenatal
Pb exposure. Blood Pb was not reported in this study, but similarly treated Wistar rat pups in other studies
report blood Pb levels at parturition in range of 50-100 (ig/dL (Grant et al.. 1980V 7-nitroinidazole (7-NI),
a selective inhibitor of nNOS, enhanced amphetamine-evoked dopamine release in the rat striatum
(Nowak et al.. 2008). Prenatal Pb exposure attenuated 7-NI's facilitatory effect on dopamine 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-dependent dopaminergic changes.
Dopamine and Vision
In various experimental animal models, the loss of retinal dopamine or zinc is associated with
supernormal rod-mediated scotopic electroretinograms (ERGs), pointing to the retina as a sensitive target
of low dose Pb exposure (19 (ig/dL). In the human and non-human primate literature, GLE is associated
with increased amplitude (supernormality) of ERGs (Lilienthal et al.. 1994; Lilienthal etal.. 1988;
Rothenberg et al.. 2002) and in the animal toxicology literature postnatal Pb exposure induces
subnormality of the ERGs. New research in the animal toxicology literature recapitulated the low dose
human literature, showing that low-and moderate (LPb or MPb) level gestational Pb exposure in the rat
produced supernormal ERGs with associated significant increases in retinal neurogenesis and significant
decreases in retinal dopamine use and dopamine turnover, DA and DOPAC:DA ratio, respectively. High
gestational Pb exposure (HPb) produced significant subnormal ERGs, similar to the findings with
postnatal human Pb exposures. Rats (dams) were exposed to Pb acetate in drinking water starting 2 weeks
prior to mating and throughout gestation and lactation until PND 10, a period of developmental exposure
that is equivalent to gestational exposure in humans with peak blood Pb PND 1-10 of 12, 24, and 46 (ig/dL
in LPb, MPb, and HPb, respectively. LPb and MPb gestational exposure induced increased cellularity or
retinal thickness in the outer nuclear layer, inner nuclear layer and total retina (Leasure et al.. 2008). In
conclusion, the retina is a sensitive site to low-dose Pb exposure and gestational Pb exposure produced
dose-dependent decreases in DA use and turnover; inverted U shaped Pb dose response curves 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 is known to contribute to aberrations in LTP or LTD in the hippocampus
via reduced NMDA current, among other mechanisms (L. Liu et al.. 2004). The 2006 AQCD detailed that
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Pb induced decreases in stimulated glutamate release that affected LTP. Further, it detailed that the Pb-
dependent decreased magnitude and increased threshold of the LTP in the hippocampus is biphasic or
non-linear. NMDAR subtypes have been shown to be significantly decreased with developmental Pb
exposure (Guilarte & McGlothan. 1998V Recent work looking at supplement use, found Pb-dependent
decreases in message and protein level of NMDAR subunit NR1 was rescued with methioninecholine
coexposure in these weanling male rats (Fan et al.. 2010). Fan et al. (Fan et al.. 2010) found that Pb-
dependent suppression of the NMDAR subunits NR2A and NR2B were not rescued with
methioninecholine treatment. Other recent mechanistic studies have found that pretreatment of primary
fetal brain neuronal rat cultures with glutamic acid, a NMDAR agonist, reversed Pb-dependent reductions
in NMDAR subunits (S.-Z. Xu & Raianna. 2006) whereas pretreatment with the NMDA antagonist MK-
801 exacerbated Pb-induced NMDAR defecits (S.-Z. Xu & Raianna. 2006). Thus, glutamic acid or
methioninecholine may offer therapeutic possibilities for Pb-induced neuronal NMDAR decrements.
The Guilarte lab has made extensive contributions to the Pb animal toxicology literature and a
recent publication details the effect of low dose developmental/lifetime Pb exposure on changes in
hippocampal neurogenesis in adulthood (Verina et al.. 2007). an emerging area of research affecting long-
term potentiation, spatial learning, neuronal outgrowth, and possibly mood disorders like schizophrenia.
NMDAR mediates the integration of new neurons into existing neuronal pathways in the adult
hippocampal DG, which is important to learning and memory. Lifetime Pb chow exposure (dam Pb
acetate exposure 10 days prior to mating through pregnancy out 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. A marker for
immature or newly formed neurons showed a significant decrease in the length-density of these cells in
the outer portion of the DG in Pb-exposed animals. 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 reversal
of learning impairment, increased expression of hippocampal NMDA receptor subunit 1, and increased
induction of brain derived neurotrophic factor mRNA (Guilarte et al.. 2003).
5.3.6.10. Neurite Outgrowth
The 2006 AQCD reported Pb decreased neurite outgrowth at 20|ag/dL noting that Pb interfered
with neurite outgrowth via protein kinase mediated pathways (MAPK/ERK); earlier work has
documented decreased primary DA neuron outgrowth with 0.001 (.iM Pb exposure (Lidskv & Schneider.
2004). More recent studies have shown that dam exposure to low dose Pb (blood Pb level 4 (ig/dL) of
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dams 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; PSA-NCAM is transiently expressed in newly formed neurons (0. S. Hu et
al.. 2008) during the period of neurite outgrowth from embyrogenesis until the early postnatal period and
is down-regulated in the adults except in areas known to exhibit synaptic plasticity (Seki & Arai. 1993).
NCAM is important for memory formation, plasticity and synapse formation and early life Pb exposure in
laboratory rodents affects its expression.
5.3.6.11. Epigenetics
DNA methyltransferase activity was significantly decreased in cortical neurons from both monkey
(aged animals) and mouse brains (fetal cells exposed to Pb in culture, 0.1 (.iM Pb) after Pb exposure (J. F.
Wu et al.. 2008). 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.
5.3.7. 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 decrements 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
AQCD (U.S. EPA. 2006). Furthermore, in the 2006 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 5-28
and Table 5-12). While the majority of epidemiologic evidence indicated differences in effect estimates
above and below 10 (ig/dL, several studies of children with mean blood Pb levels less than 5 (ig/dL
estimated larger effects for children with <5 (ig/dL (compared with children with blood Pb levels 5-10
(ig/dL, and >0 (ig/dL) (Tellez-Roio et al.. 2006). <2.5 (ig/dL (compared with children with blood Pb
levels <5 (ig/dL, <.5 (ig/dL, <10 (ig/dL, and all subjects) (Lanphear et al.. 2000). and <1.2 (ig/dL
(maternal plasma Pb, compared with plasma Pb levels >1.2 (ig/dL) (H. Hu et al.. 2006). Using data from
NHANES 1999-1994, Lanphear et al. (2000) examined differences in effect estimates among multiple
strata of blood Pb levels and found the largest deficit in reading score per 1 (ig/dL increment in blood Pb
among children with blood Pb levels less than 1 (ig/dL. As lower concentrations of Pb exposure are being
used experimentally, the toxicological literature reports nonlinear concentration-response relationships for
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1 some endpoints and similar to the epidemiologic literature, shows larger effects in lower Pb exposure
2 groups.
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Reference
Exposure Period Outcome Blood Pb strata
Lanphear et al. (2000)
Concurrent
Bellinger and Needleman (2003) Early childhood
Canfield et al. (2003) Lifetime avg
Tellez-Rojo et al. (2006)
Hu et al. (2006)
Kordas et al. (2006)
Lanphear et al. (2005)
Schwartz (1994)
Concurrent
Early childhood
Reading score All
<10
<7.5
<5
<2.5
FSIQ
FSIQ
>10
<10
All
<10
Bayley MDI £10
<10
5-10
<5
Bayley MDI/10 >1.226®
<1.226®
Math score All
<10
FSIQ >10
<10
>7.5
<7.5
FSIQ >15
<15
-4.0
-3.0 -2.0 -1.0 0.0
Change in Cognitive Score (95% CI)
1.0
Note: a = Pb levels measured in plasma of maternal blood during 1st trimester of
pregnancy. 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. Red symbols represent
effect estimates in lower blood Pb strata. Effect estimates without error bars are from
studies that did not provide sufficient information in order to calculate 95% CIs.
Figure 5-28. Comparison of associations between blood Pb and cognitive
function among various blood Pb strata.
Table 5-12. Additional characteristics and quantitative results for studies presented
in Figure 5-28
Blood Pb
Study Population/Location Levels
(M9/dL)
Statistical Analysis Outcome
Blood Pb
stratum
(pg/dL)
Effect
Estimate
(95% Cl)a
Lanphear et
al. (2000)
4853 children ages 6-16 yr Concurrent mean
NHANES 1988-1994 (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)
Bellinger et
al. Q992)
Bellinger and
Needleman
(2003)
148 children followed from Early childhood
birth (1979-1981) to age (age 2 yr) mean
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°
-1.56"
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Blood Pb
Study Population/Location Levels
(M9/dL)
Statistical Analysis Outcome
Blood Pb
stratum
(pg/dL)
Effect
Estimate
(95% Cl)a
Canfield et
al. (2003)
172 children born 1994- Lifetime avg (3 or Mixed effects models
1995 followed from infancy 5 yr) mean (SD):
to age 3-5 yr 7.4(4.3)
Rochester, NY
adjusted for sex, maternal
race, parental smoking,
child iron status, maternal
income, maternal IQ,
HOME score
Stanford-Binet at
age 3 or 5 yr
All
<10
-0.87 (-1.19, -0.55)
-1.37 (-2.56, -0.17)
Tellez-Rojo et 294 children followed from
al. (2006) birth (1994-1995, 1997-
1999) to age 2 yr
Mexico City, Mexico
Concurrent (age Linear regression model Bayley MDI at age
2 yr) mean (SD): adjusted for sex, birth 2 yr
4.28 (2.25) weight, maternal IQ
>10
<10
5-10
<5
0.07 (p = 0.84)D
-1.04 (p <0.01)b
-0.94 (p = 0.12)b
-1.71 (p = 0.01)b
Huetal. 146 children followed
(2006) prenatally (1997-1999) to
age 24 mo
Mexico City, Mexico
Maternal 1st Linear regression model
trimester plasma adjusted for plasma Pb in
Pb median: 1.226 2nd trimester, plasma Pb in
3rd trimester, child 24 mo
blood Pb, sex, maternal
age, height-for-age Z
score, maternal IQ
Bayley MDI at age
24 mo
>1.226
<1.226
-4.0°
-15.0b
Kordas et al. 602 children in 1a grade
(2006) Torreon, Mexico
Concurrent mean Linear regression model
(SD): 11.4(6.1)
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)
Lanphear et
al. (2005)
1333 children pooled from
Boston, Cincinnati,
Cleveland, Mexico City,
Port Pirie, Rochester, and
Yugoslavia cohorts
Median (5th- Linear regression model FSIQ measured at >10 -0.13 (-2.3,-0.03)
95th) adjusted for HOME score, ages4.8-10yr <10 -0.80 (-1.74,-0.14)
Peak: 18.0 (6.2- birth weight, maternal IQ, >7.5 -0.16 (-2.4,-0.08)
47.0) maternal education <7.5 -2.94 (-5.16,-0.71)
Schwartz Meta-analysis of 7 studies
(1994) 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
schoolage
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,
investigators did not provide sufficient information in order to calculate 95% CIs.
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 lower blood Pb levels. A linear model
was found to be inadequate as the polynomial terms for concurrent blood Pb were statistically significant.
These findings were corroborated by a separate analysis by Rothenberg and Rothenberg (2005) which
found that the log-linear model fit the relationship between blood Pb level and IQ better than did a linear
model.
Studies of adults have not widely examined the shape of the relationship between blood or bone Pb
level and cognitive performance. In the various NHANES analyses, only log-linear models were used to
fit the data (E. F. Krieg. Jr. & Butler. 2009; E. F. Krieg. Jr. et al.. 2009; E. F. Krieg. Jr. et al.. 2010). Other
studies examined nonlinearity with the use of quadratic terms, penalized splines, or visual inspection of
bivariate plots (Bandeen-Roche et al.. 2009; Shih et al.. 2006; Weisskopf. Proctor, et al.. 2007). While
there was some evidence for nonlinearity for some cognitive tests (Figures 5-17 and 5-18), the majority of
results suggested linear associations. Shih et al. (2006) found that a quadratic term for tibia Pb was not
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statistically significant and found a linear model fit adequately the relationship between tibia Pb level and
various tests of cognitive performance. In contrast to most studies, Wang et al. (2007) found that among
HFE variant carriers, there was a steeper decline in MMSE score at higher tibia Pb levels (20-25 jxg/g,
Figure 5-17).
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, saturation of biological mechanisms at higher levels, larger proportion of
susceptible populations at lower exposure levels, and variations in other risk factors among exposure
levels (Stavner et al.. 2003). Other explanations for nonlinearity include different mechanisms operating
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
neurocognitive function has not been examined widely in epidemiologic 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,
which argues against a larger proportion of susceptible populations at lower blood Pb levels (Lanphear et
al.. 2000; Lanphear et al.. 2005). 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
overwhelming effects of these other risk factors (J. Schwartz. 1994). Several studies have found
significant associations of these sociodemographic risk factors with neurocognitive deficits, and Miranda
et al. (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 5-6). Few studies have compared Pb effect estimates among groups in different sociodemographic
strata, and the limited data are mixed. Greater Pb-associated neurocognitive deficits in low-SES groups
were reported 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 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 dose-
response relationship.
Rothenberg and Rothenberg (2005) formally assesed the influence of residual confounding on the
nonlinear blood Pb 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
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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.
Consistent with the epidemiologic literaure, toxicological studies also find nonlinear relationships
between Pb exposure and neurological effects in animals. In particular, multiple studies have shown U- or
inverse U-shaped curves with lower exposures of Pb having different or often the opposite effect from
higher doses. U-shaped Pb exposure-responses include rotoarod performance, adult forebrain dopamine
levels, amphetamine-induced motor activity, and latency to fall from rotarod (Lcasure et al.. 2008).
Inverted U-shaped Pb dose-responses include histological findings such as the numbers of rod
photoreceptors and bipolar cells, forebrain dopamine use, activity level, and adult body weight (Lcasure
et al.. 2008) as well as ERG wave amplitudes (Fox et al.. 2008) and hippocampal neurogenesis (Fox et al..
2008; Gilbert et al.. 2005).
Because toxicological studies typically do not have confounding, exposure measurement error or
other epidemiologic influences, i.e., susceptibility, they have permitted assessment of a mechanistic basis
for nonlinear Pb exposure- or dose-response relationships. Several lines of evidence support the
possibility of low-dose and high dose-Pb acting through differential activation of mechanisms. For
example, in mice, lower Pb exposure (50 ppm) is associated with differential responses of the
neurotransmitters dopamine and norepinephrine compared with control treatment and higher doses (150
ppm) (Lcasure et al.. 2008; Virgolini et al.. 2005). These differential responses of neurotransmitter
systems to lower versus higher Pb exposures may provide mechanistic understanding of the nonlinearity
of Pb-induced behavioral changes in animals and may also explain the nonlinear blood Pb-neurocognitive
and neurobehavioral associations reported widely among children. 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 are significantly lower in the 50 ppm exposure
group than in the control or 150 ppm Pb exposure group.
Additional mechanistic understanding comes from differences in histological changes found in Pb-
exposed animals. Compared with higher Pb exposure, lower Pb exposure stimulates greater induction of
c-fos, a marker of neuronal activation and action potential firing (Lewis & Pitts. 2004). These findings
may underlie the nonlinear association between Pb exposure and learning and the U-shaped behavioral
dose-responsiveness seen with amphetamine-induced motor activity in males after GLE (Lcasure et al..
2008).
Sensory organ findings in animals also show vastly different outcomes with low versus higher Pb
exposure. Higher Pb exposure produces subnormal retinal ERGs and lower Pb exposure produces
supernormal ERGs in both children (Rothenberg et al.. 2002) and rodents (Fox & Chu. 1988; Fox &
Farber. 1988; Fox et al.. 1991). Inverted U-shaped dose-response curves have been seen for rod
photoreceptor numbers or neurogenesis (Giddabasappa et al.. 2011) and retinal thickness (Fox et al..
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2010). Thus, these dichotomous histological findings are coherent with the functional retinal test or the
ERG where higher Pb exposure produces subnormal ERGs and lower exposure Pb produces supernormal
ERGs.
Hierarchical enzyme activity also may explain nonlinear Pb concentrations-response relationships.
The phosphatase enzyme calcineurin has been shown to be inhibited by higher Pb exposure and
stimulated by lower Pb exposure (Kern & Audesirk. 2000). At lower Pb exposure, Pb displaces calcium at
its binding sites on calmodulin and by acting as a calmodulin agonist at calcineurin's catalytic A subunit,
stimulates calcineurin activity. At higher Pb exposure, Pb can bind directly to a separate calcium-binding
B subunit, overriding the calmodulin-dependent effect and turning off the activity of calcineurin.
Interestingly, mice with modulated calcineurin expression exhibit aberrant behavior related to
schizophrenia or impaired synaptic plasticity and memory (Zeng et al.. 2001). This example of the
stimulatory effects of Pb at lower exposure and inhibitory effects at higher exposure gives another
example of biological plausibility for the nonlinear concentration-response relationship reported for Pb in
multiple studies.
The supralinear concentration-response relationship widely documented for Pb is consistent with
the lack of a threshold for Pb-associated neurological 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 the Boston prospective cohort data by regressing IQ and blood Pb level on potential
confounders including age, race, maternal IQ, SES, and HOME score and fitting a nonparametric
smoothed curve to the residuals of both regression models (variation in IQ 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 j^ig/dL). Thus, in the Boston study, the
association between blood Pb level and IQ 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, has been 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 (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 to 0.025
(ig/dL in NHANES. This has allowed categorization of children in mulitple 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
neurological effects. Several recent studies reported associations between blood Pb levels and deficits in
neurocognitive and neurobehavioral endpoints in populations with mean blood Pb levels <2 (ig/dL (Braun
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et al.. 2008; Braun et al.. 2006; C'ho et al.. 2010; E. F. Krieg. Jr. et al.. 2010). In comparisons of various
quantiles of blood Pb, Chandramouli et al. (2009) observed a lower SAT score among children in the U.K.
with blood Pb levels 2-5 (ig/dL compared with children with blood Pb levels 0-2 (ig/dL. Likewise,
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. 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). Collectively, these new findings in children, as summarized in this document, do not provide
evidence for a threshold for the neurological effects of Pb in the ranges 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 of Pb in blood in preindustrial humans was approximately 0.016
(ig/dL (Flegal & Smith. 1992). approximately 65-fold lower than that currently measured in U.S.
populations and lower than the levels at which neurological effects have been observed (1 j^ig/dL). Thus,
the current evidence does not preclude the possibility of a threshold existing in the large range of blood
levels between 1 (ig/dL and preindustrial "background" levels.
5.3.8. Summary and Causal Determination
The 2006 Pb AQCD concluded that the collective body of epidemiologic studies provides clear and
consistent evidence for the effects of Pb exposure on neurocognitive function in children. This conclusion
was substantiated by the coherence of findings across studies of diverse design and populations (varying
distributions of blood Pb levels, SES, parental intelligence, and quality of caregiving) that blood Pb
levelswere associated with a broad spectrum of neurocognitive and neurobehavioral indices, including
cognitive function (IQ), higher-order processes such as language and memory, academic achievement,
behavior and conduct, delinquent and criminal activity, sensory acuities, and changes in brain structure
and activity as assessed by MRI or MRS (Figure 5-29). Toxicological studies not only provided coherence
with similarly consistent findings for Pb-induced impairments in parallel tests of learning, behavior and
attention, and sensory acuities, but also provided biological plausibility by characterizing mechanisms for
Pb-induced neurological effects (Figure 5-29). These mechanisms included Pb-induced inhibition of
neurotransmitter release, decline in synaptic plasticity, decreases in neuronal differentiation, and
decreases in the integrity of the blood-brain-barrier. Both epidemiologic studies in children and
toxicological studies reviewed in the 2006 Pb AQCD demonstrated neurocognitive deficits in association
with blood Pb levels at or below 10 (ig/dL, and evidence from both disciplines also supported a nonlinear
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concentration-response relationship, with greater cognitive or behavioral decrements per unit increase in
blood Pb level estimated for lower blood Pb levels or estimated for lower Pb exposures. Among adults,
although associations of blood Pb level with the spectrum of neurological effects (e.g., impairments in
memory, attention, mood, balance, and motor function) were most consistently observed in
occupationally-exposed adults with blood Pb levels > 14 (ig/dL, studies of adults without occupational Pb
exposures indicated associations between biomarkers of cumulative Pb exposure, serial blood Pb or bone
Pb measurements, and decrements in cognitive function.
Building on the strong body of evidence presented in the 2006 AQCD, recent studies continue to
support associations of Pb biomarkers or exposures with neurological effects (Figure 5-29). Although
fewer in number, recent studies in children corroborate findings from the several previous longitudinal
and cross-sectional studies in demonstrating associations between blood Pb levels and FSIQ. In the
cumulative body of evidence, negative associations between blood Pb level and IQ are best substantiated
at mean blood Pb levels in the range of 5-10 (ig/dL; however, an association was observed in a recent
study with a mean blood Pb level of 1.73 (ig/dL. A majority of recent epidemiologic studies in children
has focused on examining specific indices of neurocognitive function such as reading and verbal skills,
memory, learning, and visuospatial processing and has demonstrated associations with blood Pb levels as
low as 2 (ig/dL (population mean or quantiles). The consistently positive associations observed between
blood Pb levels and this diverse set of neurocognitive indices provides coherence with findings for IQ, a
global measure of cognitive function that reflects the integration of these individual domains. Additional
coherence for findings in children is derived from evidence in animals that blood Pb levels of 11.6 (ig/dL
and higher are associated with changes in learning and memory (Figure 5-29). Recent toxicological
studies continue to demonstrate that in utero and early postnatal exposure to Pb is the most sensitive
window for Pb-dependent neurological effects. In the 2006 Pb AQCD, uncertainty was noted among
studies in children regarding the relative importance of prenatal, early life, concurrent, and lifetime
measures of Pb exposures. It also was noted that distinguishing among the effects of Pb exposures at
different lifestages is difficult in epidemiologic studies due to the high correlations among children's
blood Pb levels over time. Although blood Pb levels at all of these lifestages are associated with
neurocognitive deficits in children, stronger effects generally are estimated for concurrent blood Pb
levels, and recent evidence indicates that among both children with relatively lower or higher early
childhood blood Pb levels, concurrent blood Pb levels are more strongly associated with neurocognitive
deficits. In addition to performance on various neurocognitive tests, recent epidemiologic studies in
children link blood Pb levels (in quantiles as low as 2 (ig/dL) with factors that may be indicators of
children's life success, including the level of educational attainment and end-of-grade score. In particular,
observations of lower 4th grade end-of-grade score among children with blood Pb levels of 2 (ig/dL
compared with children with blood Pb levels of 1 (ig/dL indicate that a threshold may not exist for the
neurodevelopment effects of Pb in children.
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Recent studies in children continue to support associations of blood Pb levels (population means or
quantiles of 3-11 (ig/dL) with a range of behavioral problems from anxiety and distractability to conduct
disorder and delinquent behavior (Figure 5-29). Whereas previous evidence was not compelling, new
evidence indicates associations between blood Pb levels and ADHD diagnosis and contributing diagnostic
indices. In particular, a recent NHANES analysis demonstrated associations at blood Pb levels between 1
and 2 (ig/dL). These findings for ADHD are well-supported by observations in animals of Pb-induced
increased response rates and impulsivity. Additional coherence is provided by evidence in aquatic and
terrestrial species for Pb affecting behaviors that decrease the ability of organisms to escape predators or
capture prey (Chapter 7.1, 7.2). Both epidemiologic studies in children and adults as well as toxicological
studies demonstrate associations of Pb biomarkers or exposure with deficits in visual acuity and hearing
and auditory processing. New evidence from toxicological studies demonstrates the effects of lower blood
Pb levels (<15 (ig/dL) with retinal changes in male offspring. Combined evidence for Pb-associated
neurocognitive deficits, inattention, conduct disorder, and effects on sensory function provides plausible
mechanisms by which Pb exposure may contribute to academic underachievement and to more serious
problems of delinquent behavior.
Studies in adults without occupational exposure to Pb have not provided consistent evidence for
associations or blood or bone Pb levels with the range of neurological effects. Levels of Pb in bone,
particularly in tibia, which is an indicator of cumulative Pb exposure, including higher exposures in the
past, are better predictors of cognitive performance rather than a single blood Pb measurement. One
explanation for the overall weaker body of evidence 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 between biomarkers of
cumulative Pb exposure (serial blood measurements, tibia Pb levels) and neurocognitive deficits. Among
recent studies of adults, blood Pb levels and bone Pb levels have been associated with essential tremor
and Parkinson's Disease, respectively. These findings are well-supported by toxicological evidence for
Pb-induced decreased dopaminergic cell activity in the substantia nigra, which contributes to the primary
symptoms of Parkinson's disease. Biological plausibility also is provided by observations of
developmental Pb exposures of monkeys and rats resulting in neurodegeneration in aged brains. Recent
evidence also indicates associations between early-life ALAD activity, a biomarker of Pb exposure, and
schizophrenia later in adulthood. Consistent with these findings, toxicological studies have observed Pb-
induced emotional changes in males and depression changes in females. It is not surprising that Pb
exposure may increase the risk of different neurological endpoints in children and adults given the
predominance of different neurological processes operating at different ages, in particular, neurogenesis
and brain development in children and neurodegeneration in adults.
Several host and environmental factors may modify the association between Pb exposure and
neurological effects in children. Interactions of blood Pb levels with race/ethnicity and SES continue to be
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poorly characterized. Although the 2006 Pb AQCD cited mixed epidemiologic evidence for effect
modification by sex, recent epidemiologic evidence points to males having increased susceptibility for
Pb-associated neurological effects. Toxicological studies continue to demonstrate increased susceptibility
of males for endpoints such as sensory function, balance, stress hormone homeostasis, and brain
membrane composition. Although limited, evidence suggests that risk of Pb-associated neurocognitive
deficits in children also may be modified by variants in genes for apolipoprotein E and dopamine
receptors. In addition to host factors, recent studies suggested that associations between blood Pb levels
and neurological effects in children are greater with coexposures to environmental tobacco smoke and
manganese. Historical animal toxicology findings demonstrate interactions between Pb exposure and
stress. Namely, 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 with lower Pb exposures. In comparison,
epidemiologic evidence for such interactions has been sparse. However, consistent with historical animal
studies, a recent study indicated that positive social environment of children as characterized by maternal
self-esteem, attenuates the negative association between blood Pb level and cognitive function. While
effect modification by these host and environmental factors has not been examined widely in
epidemiologic studies, new studies provide information on potentially susceptible populations that may
benefit from early intervention to reduce the risk of neurological effects. Furthermore, the robust evidence
for varying susceptibilities to Pb-induced neurological effects provides a basis to integrate
mechanistically the findings of toxicology and epidemiology.
Extensive evidence from toxicological studies clearly substantiates the biological plausibility for
epidemiologic findings by characterizing mechanisms underlying neurological effects. Pb exposure of
animals induces dopamine changes in animals (Figure 5-29). Dopamine plays a key role in cognitive
functions mediated by the prefrontal cortex and also motor functions mediated by the substantia nigra,
and animal toxicological findings provide mechanistic support for associations in humans between blood
Pb levels and neurocognitive 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 brain. Pb induces 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 over time but are dynamic in nature. The effect of Pb on
NMDA receptors and the contribution of this paradigm to mood disorders is detailed. Synapse formation,
adhesion molecules, and nitrosive stress continue to be areas of research with known Pb-associated
adverse outcomes. Finally, the new area of epigenetics details that Pb exposure affects methylation
patterns in rodent brains. These toxicological data complement the expanding epidemiologic data and
often provide coherence between the two fields.
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In summary, recent evidence substantiates and expands upon the established epidemiologic and
toxicological literature of neurologic effects associated with Pb exposure. In epidemiologic studies of
children, consistently positive associations of blood Pb levels with deficits in neurocognitive function,
attention, and sensory acuities support observed associations with school performance as assessed by end-
of-grade scores and level of educational attainment, which in turn, may explain associations with
delinquent and criminal behavior. In particular, observations of lower academic achievement and ADHD
among children in quantiles of blood Pb levels in the range of 1 to 2 (ig/dL have not indicated that a
threshold exists for the neurodevelopment effects of Pb in children. Epidemiologic findings are
strengthened by the coherence and biological plausibility provided by toxicological findings for similar or
parallel endpoints and for the mechanisms underlying the neurological effects (Figure 5-29). The
collective body of evidence integrated across epidemiologic and toxicological studies and across the
spectrum of neurological endpoints is sufficient to conclude that there is a causal relationship between
Pb exposures and neurological effects.
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5.4. Cardiovascular Effects
5.4.1. Introduction
Both human and animal studies provide consistent evidence for an association of increased BP and
arterial hypertension with chronic exposure to Pb resulting in adult blood Pb levels below 5 (ig/dL. In
addition, studies have suggested a connection between measures of Pb exposure and other cardiovascular
diseases in adults such as ischemic heart disease, cerebrovascular disease, peripheral vascular disease, and
cardiovascular disease related mortality. Toxicological studies explore the underlying mechanisms by
which Pb exposure can lead to human cardiovascular health outcomes. Such studies have demonstrated
that the Pb content in heart tissue reflects the increases in blood Pb levels (Lai et al.. 1991). In general,
associations between blood Pb and bone Pb (particularly in the tibia) with health outcomes in adults
indicate acute effects of recent dose and chronic effects of cumulative dose, respectively. In some
physiological circumstances of increased bone remodeling or loss (e.g., osteoporosis and pregnancy), Pb
from bone of adults may also contribute substantially to blood Pb concentrations. Additional details on
the interpretation of Pb in blood and bone are provided in Section 4.3.5. Additionally, as the
cardiovascular and renal systems are intimately linked, cardiovascular effects can arise secondarily to Pb-
induced renal injury (Section 5.5).
The previous Pb AQCD (U.S. EPA. 2006) concluded that both epidemiologic and animal
toxicological studies support the relationship between increased Pb exposure and increased cardiovascular
outcome, including increased BP, increased incidence of hypertension, and cardiovascular morbidity and
mortality. Meta-analysis of these human studies found that each doubling of 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 (Nawrot et al.. 2002). On a population-wide basis, the measured effect size translates into a
large number of events for a moderate population size and thus has important health consequences for the
occurrence of stroke, myocardial infarction, and sudden death. It was also noted that most of the reviewed
studies using cumulative Pb exposure measured by bone Pb also showed increased BP (Y. Cheng et al..
2001; H. Hu et al.. 1996) or increased hypertension with increasing bone Pb (B.-K. Lee et al.. 2001). 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 ratio of hypertension between 1.28 and 1.86, depending upon the study. Two studies
observed averaged increased 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. Since bone Pb measures Pb accumulation over time, duration
of past exposure to Pb plays a role in increased BP.
The previous Pb AQCD also provided compelling evidence for a number of mechanisms leading to
increased BP, and the development of hypertension and other cardiovascular diseases observed after Pb
exposure. The strongest evidence supported the role of oxidative stress in the pathogenesis of Pb-induced
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19
20
21
22
23
24
25
26
27
28
29
30
31
hypertension. Additionally, several studies focused on other pathways or cellular, molecular, and tissue
events promoting the Pb-induced increase in BP. These mechanisms include inflammation, adrenergic and
sympathetic activation, renin-angiotensin-aldosterone system (RAAS) activation, vasomodulator
imbalance, and vascular cell dysfunction. Studies continue to support the observed increase in BP and
hypertension development following Pb exposure, as well as build on the evidence for the biological
pathways of these effects. This section reviews the published studies pertaining to the cardiovascular
effects of Pb exposure in experimental animals, isolated vascular tissues, cultured vascular cells, and
humans. Emphasis has been placed on studies published since the 2006 AQCD (U.S. EPA. 2006).
however the large body of evidence that existed prior to that review has been summarized and
incorporated into the current review.
5.4.2. Blood Pressure and Hypertension
5.4.2.1. Epidemiology
The most commonly used indicator of cardiovascular morbidity is increased BP and its derived
index, hypertension. Hypertension diagnoses in these studies require that the patient or subject have
diastolic and/or systolic BP above certain cut-points or be taking anti-hypertensive medicines. These BP
cut-points have historically been established by reference to informed medical opinion and as medical
knowledge improves BP cut-points defining hypertension have been lowered overtime. Consequently,
different studies using "hypertension" as a cardiovascular outcome may assign 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 at or
above 90 or taking anti-hypertensive medications). Studies in the medical literature show that increasing
BP is associated with increased rates of cardiovascular disease 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 (Chobanian et al.. 2003; Ingelsson et al..
2008; Kannel. 2000a. 2000b; Neaton et al.. 1995; Pastor-Barriuso et al.. 2003; Prospective Studies
Collaboration. 2002).
Several recent general population and occupational cohort studies examined the associations of
blood Pb and/or bone Pb with BP (Figure 5-30 and Table 5-13) as well as the associations of these Pb
exposure metrics with hypertension (Figure 5-31 and Table 5-14). In a cross-sectional analysis, Martin et
al. (2006) examined the association of blood and tibia Pb with BP and hypertension in a community-based
cohort study of older adults (n = 964). Four models evaluated associations for BP and hypertension
considering SES and race/ethnicity. Blood Pb but not tibia Pb was a strong and significant predictor of BP
in all models with an approximately 1 mmHg increase in systolic BP with each 1 (ig/dL increase in blood
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1 Pb level and an approximately 0.5 mmHg increase in diastolic BP per 1 (ig/dL increase in blood Pb level.
2 Tibia Pb but not blood Pb was associated with hypertension in logistic regression models. The authors
3 applied propensity analysis to their models to better account for the effect of risk factors such as
4 race/ethnicity, age and SES that were strongly associated with tibia Pb level. The propensity score
5 analysis and model adjustment did not substantially change the numerical findings and conclusions (e.g.
6 tibia Pb and hypertension were positively associated independent of race/ethnicity and socioeconomic
7 status). No evidence for effect modification by race/ethnicity was found. Overall, the results suggest that
8 Pb has an acute effect on BP as a function of recent dose and a chronic effect on hypertension risk as a
9 function of cumulative exposure.
Reference
Strata
Pb Distribution3
Exposure Metric0
SBP
Martin et al. (2006)
2.9 (2.0, 4.4)
Blood Pb
1 ^
1 *
Scinicariello et al. (2010)
NHANES III-Whites
1.6(0.8, 3.3)
Blood Pb
|—•—
NHANES III-Blacks
1.4(0.6, 3.6)
Blood Pb
#
NHANES III - Mexicans 2.0 (1.0, 3.9)
Blood Pb
$
Weaver et al. (2008)
Korean Workers
27.2 (19.3, 28.3)
Blood Pb
•
Glenn et al. (2006)
Korean Workers
Blood Pb (concurrent)^
Blood Pb (longitudinalp
Martin et al. (2006)
15.7 (10.5, 23.5)
Tibia Pb
Peters et al. (2007)
High Stress
18.1 (12.2, 26.9)
Tibia Pb
: o
Low Stress
18.1 (12.2, 26.9)
Tibia Pb
o,
Glenn et al. (2006)
Korean Workers
Tibia Pb (historical)
Q
Weaver et al. (2008)
Korean Workers
74.3 (67.3, 82.0)
Patella Pb
5
Peters et al. (2007)
High Stress
26.9(18.4, 39.3)
Patella Pb
o
DBP
Scinicariello (2010)
NHANES III-Whites
1.6(0.8, 3.3)
Blood Pb
•!
NHANES III-Blacks
1.4(0.6, 3.6)
Blood Pb
•
NHANES III - Mexicans 2.0 (1.0, 3.9)
Blood Pb
*—9
Martin et al. (2006)
2.9 (2.0, 4.4)
Blood Pb
15.7 (10.5, 23.5)
Tibia Pb
o
pp
Zhang etal. (2010)
HFE Wild-type
18 (12, 27)
Tibia Pb
t _ • ¦
H63D variant
19 (14, 26)
Tibia Pb
i o
C282Y variant
20 (14, 27)
Tibia Pb
Any HFE variant
19 (14, 27)
Tibia Pb
1 c
HFE Wild-type
26 (17, 34)
Patella Pb
H63D variant
27 (19, 37)
Patella Pb
C282Y variant
25 (17, 37)
Patella Pb
Any HFE variant
26 (18, 37)
Patella Pb
-2 0 2 4 6 8
Slope of BP (mmHg) per pg/dL BLL at 1 pg/dL or per pg/g bone Pb
Note: aPb distribution is the median (IQR) estimated to make comparable. bEffect estimates were
standardized to 1 [jg/dL blood Pb or 10 |jg/g bone Pb.
Figure 5-30. Slope of BP (mmHg) per pg/dL blood Pb level at 1 pg/dL or per
10 pg/g bone Pb (95% CI) for associations of blood Pb (closed
circles) and bone Pb (open circles) with systolic BP (SBP;
blue), diastolic BP (DBP; red), and pulse pressure (PP; purple).
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Table 5-13. Additional characteristics and quantitative data for associations of blood and
bone Pb with BP measures for results presented in Figure 5-30
Study
Population
/Location
Parameter Pb Data
Statistical Analysis
Effect Estimate
(95% CI)
Martin et al.
(2006)
964 men and women,
50-70 yr, 40% African
American, 55% White,
5% other, in Baltimore,
MD
BP Concurrent Mean
Blood Pb:
Mean (SD): 3.5 (2.3)
|jg/dL
African American: 3.4
(2.3)
White: 3.5 (2.4)
Tibia Pb:
Mean (SD): 18.8 (12.4)
Mg/g
African American: 21.5
(12.6)
White: 16.7(11.9)
Multiple linear regression base
model adjusted for age, sex, BMI,
antihypertensive medication use,
dietary sodium intake, dietary
potassium intake, time of day,
testing technician, serum total
cholesterol. SES, race/ ethnicity
also included in select models that
are presented in Figure 5-30 and
tabulated here.)
Blood Pb
SBP: 13=1.05(0.53,1.58)
DBP; |3=0.53 (0.25, 0.81)
Tibia Pb:
SBP: p=0.07 (-0.05, 0.14)
DBP: p=0.05 (-0.02, 0.08)
mmHg per |jg Pb/dL
blood
mmHg per |jg Pb/g bone
Glenn et al.
(2006)
575 Pb exposed
workers, age 18-65 yr,
in South Korea
(10/1997-6/2001)
BP Blood Pb mean (SD):
Visit 1:20.3 (9.6), Women
Visit 2: 20.8 (10.8),
Women
Visit 3:19.8 (10.7),
Women
Visit 1:35.0 (13.5), Men
Visit 2: 36.5 (14.2), Men
Visit 3: 35.4 (15.9), Men
Tibia Pb, mean (SD):
Visit 1:28.2 (19.7),
Women
Visit 2: 22.8 (20.9),
Women
Visit 1:41.7 (47.6), Men
Visit 2: 37.1 (48.1), Men
Patella Pb, mean (SD):
Visit 3 49.5 (38.5) Women
Visit 3 87.7 (117.0)
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.
Model 1 (short-term)
Blood Pb concurrent
|B=0.08 (-0.01,0.16)
Blood Pb (longitudinal)
|B=0.09 (0.01,0.16)
Model 4: short and
longer-term)
Blood Pb concurrent
p=0.10 (0.01,0.19)
Blood Pb longitudinal:
P=0.09 (0.01,0.16)
Per 10 |jg/dL blood Pb
Weaver etal.
(2008)
652 current and former
Pb workers in South
Korea (12/1999-
6/2001)
BP Blood Pb:
Mean (SD):30.9 (16.7)
|jg/dL
Patella Pb:
Mean (SD): 75.1 (101.1)
Mg/g
Linear regression model adjusted
forage, gender, BMI, diabetes,
antihypertensive and analgesic
medication use, Pb job duration,
work status, tobacco and alcohol
SBP
Patella Pb
p=0.0059 (p=0.41)
Blood Pb
P=0.1007 (p=0.01)
Interaction between blood
Pb/patella Pb with ALAD
and vitamin D receptor
polymorphisms not
significant.
Peters etal. 513 elderly men (mean BP
(2007) 67 y) from Normative
Aging Study in Greater
Boston, MA area
Tibia Pb:
mean (SD):21.5 (13.4)
Mg/g
Patella Pb:
Mean (SD):31.5 (19.3)
Mg/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:
P=3.57 (0.39, 6.75)
Low Stress:
P=-0.21 (-1.70,1.29)
per SD increase in tibia
Pb
Patella Pb/ High Stress:
P=2.98 (-0.12, 6.08)
perSD increase in tibia
Pb
Patella Pb/ Low Stress:
NR
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Study
Population
/Location
Parameter Pb Data
Statistical Analysis
Effect Estimate
(95% CI)
Scinicariello et 6,016 NHANES III BP Blood Pb:
al. (2010) participants > 17 yr Overall Mean (SE): 2.99
(0.09) pg/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
Ln blood Pb
SBP
Non-Hispanic whites:
P±SE=1.05±0.37 (p=0.01)
Non-Hispanic blacks:
p±SE=2.55±0.49
(p=0.001)
Mexican Americans:
p±SE=0.84±0.46 (p=0.08)
DBP
Non-Hispanic whites:
p±SE= -0.14±0.49
(p=0.77)
Non-Hispanic blacks:
P±SE=1.99±0.44
(p=0.0002)
Mexican Americans:
p±SE=0.74±0.38 (p=0.06)
Significant interactions
with blood Pb and ALAD
observed in relation to
SBP for non-Hispanic
whites and non-Hispanic
blacks
Zhang etal. 619 older adult males PP
(2010) (mean 67 yr) enrolled
in the VA-NAS in
Greater Boston, MA
Wild type
Tibia Pb: Med (IQR):8(12-
27) Mg/g
Patella Pb:
Med(IQR):26(17-37) jjg/g
C282Y
Tibia Pb: Med(IQR):20
(14-27) ^g/g
Patella Pb:
Med(IQR):25(17-37) jjg/g
H63D
Tibia Pb:
Med(IQR):19(14-26) |jg/g
Patella Pb:
Med(IQR):27(19-37)|jg/g
Linear mixed effects regression
models with repeated
measurements adjusted for age;
education; alcohol intake;
smoking; daily intakes of calcium,
sodium, and potassium; total
calories; family history of
hypertension; diabetes; height;
heart rate; high-density lipoprotein
(HDL); total cholesterol:HDL ratio;
and waist circumference
PP
Tibia Pb per 13 |jg/g:
Wild Types: p=0.38
(0,1.96)
H63D: p=3.30 (0.16,
6.46)
C282Y: p=0.89 (0, 5.24)
Any HFE: p=2.90 (0.31,
5.51)
Patella Pb per 19 |jg/g:
Wild Type: p=0.26 (0,
1.78)
H63D: p=2.95 (0, 5.92)
C282Y: p=0.55 (0,1.66)
Any HFE: p=2.83
(0.32,5.37)
Perlstein etal.
(2007)
593 predominantly
white men from VA-
NAS in Greater
Boston, MA area
(1991-1997)
PP Blood Pb:
Overall mean (SD): 6.12
(4.03) |jg/dL
Mean (SD) quintiles:
Q1
Q2
Q3
Q4
Q5
2.3 (0.8) |jg/dL
3.9 (0.3) |jg/dL
5.4 (0.5) |jg/dL
7.4 (0.6) |jg/dL
12.4 (4.4) |jg/dL
Tibia Pb:
Median: 19 |jg/g
Mean (SD) quintiles:
Q1
Q2
Q3
Q4
Q5
7.4(3.2) |jg/g
14.1 (1.4) |jg/g
18.9 (1.4) |jg/g
24.9 (2.2) |jg/g
40.9 (14) |jg/g
BP association assessed using
spearman correlation coefficients.
PP association(adjusted mean
difference) assessed using
multiple linear regression model
adjusted for age, height, race,
heart rate, waist circumference,
diabetes, family history of
hypertension, education level
achieved, smoking, alcohol intake,
fasting plasma glucose, and ratio
of total cholesterol to HDL
cholesterol
Tibia Pb:
SBP r=0.06 (p=0.15)
DBP r=-0.02 (p=0.63)
Blood Pb:
SBP r=0.05 (p=0.28)
DBP r=0.12(p=0.01)
Pulse Pressure
Tibia Pb:
>Median: 4.2 (1.9,6.5)
mmHg (mean higher than
men below the median)
-------
Study
Population
/Location
Parameter Pb Data
Statistical Analysis
Effect Estimate
(95% 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
Pooled Estimates per 10
|jg/g increase in Tibia Pb
Prospective/SBP
|B=0.33 (-0.44,1.11)
X-sectional SBP
|3=0.26 (0.02, 0.50)
X-sectional DBP
|3=0.02 (-0.15,0.19
x-Sectional hypertension
OR=1.04 (1.01,1.07)
Pooled Estimates per 10
|jg/g increase in patella
Pb
hypertension
OR=1.04 (0.96,1.12)
Gumpetal. 122 children age 9.5 yr BP, TPR
(2007) in Oswego, NY (total
peripheral
vascular
resistance)
Blood Pb:
Mean (SD): 4.6 (2.5)
|jg/dL
Pb is a mediator/modifier
and moderator in the
analysis, no effects
presented
Yazbecketal. 971 pregnant women, BP
(2009) age 18-45 yr, in France
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. (2006) 67 yr) from Normative
Aging Study in Greater
Boston, MA area
Blood Pb:
Mean (SD): 6.6 (4.3)
|jg/dL
Tibia Pb:
Mean (SD):21.6 (12.0)
Mg/g
Patella Pb:
Mean (SD): 31.7 (18.3)
Mg/g
Linear regression models
adjusted for age, BMI, family
history of hypertension, history of
smoking, dietary sodium intake,
and cumulative alcohol ingestion
Tibia Pb
High calcium group (>800
mg/d):
SBP: |3=0.40.(0.11,0.70)
Low calcium group (<800
mg/d):
SBP: 13=0.19(0.01,0.37)
mmHg per |jg/g tibia Pb
References not included in Figure 5-30 are included in this table.
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Reference Study Strata Blood Pb (|Jg/dL) Comparison
Scinicariello et al. (2010)
NHANESIII Non-Hispanic Whites
0.7-1.4
Q1 Reference
1988-1994
1.5-2.3
Q2 v Q1
age ^17 y
2.4-3.7
Q3 v Q1
Mean blood Pb =
3.8-52.9
Q4 v Q1
2.99 |jg/dL
2.4-3.7
ALAD2 v 1b
Non-Hispanic Blacks
0.7-1.4
Q1 Reference
1.5-2.3
Q2 v Q1
2.4-3.7
Q3 v Q1
3.8-52.9
Q4 v Q1
2.4-3.7
ALAD2 v 1b
Mexican Americans
0.7-1.4
Q1 Reference
1.5-2.3
Q2 v Q1
2.4-3.7
Q3 v Q1
3.8-52.9
Q4 v Q1
2.4-3.7
ALAD2 v 1b
Park et al. (2009)
Vlean blood Pb =
3.52 |jg/dL
NHANESIII
Overall
White Men
Black Men
White Women
Black Women
Men <50
Men > 50
Women < 50
Women >50
3.52 (0.10)
j-#-
Muntner et al. (2005)
Vlean blood Pb =
1.64 (jg/dL
NHANESIII
1999-2002
Non-Hispanic Whites
Non-Hispanic Blacks
<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
Q1 Reference
Q2 v Q1
Q3 v Q1
Q4 v Q1
Q1 Reference
Q2 v Q1
Q3 v Q1
Q4 v Q1
Q1 Reference
Q2 v Q1
Q3 v Q1
Q4 v Q1
^
Mexican Americans
Martin et al. (2006)c
Baltimore, MD
3.5 (2.3)
4
~
Yazbeck etal. (2009)
France
Pregnant Women
1.2-1.7
1.71-2.30
>2.30
Q1 Reference
Q2 v Q1
Q3 v Q1
Q4 v Q1
Tibia Pb (ug/g)
Martin et al. (2006)c
Baltimore, MD
18.8 (12.4)
¦o
^eters et al. (2007)d
Boston, MA
High Stress
High Stress
21.5 (13.4)
Patella Pb (ug/g)
31.5 (19.3)
0
1 1 1 1
0.0 1.0 2.0 3.0 4.0 5.0 6.0
Odds Ratio (95% CI)
Note: aThe outcomes plotted are hypertension with the exception of Yazbeck et al. (2009) which
measured pregnancy induced hypertension and Peters et al. (2007) which measured hypertension
incidence. ALAD2 v 1 indicates comparison between ALAD 2 carriers (e.g. ALAD1-2 and ALAD2-
2) and ALAD 1 homozygotes (e.g., ALAD1-1). cEffect estimates were standardized to 1 [jg/dL
blood Pb. dEffect estimates were standardized to 1 |jg/g bone Pb.
Figure 5-31. Odds ratio (95% CI) for associations of blood and bone Pb
with hypertension measures.
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Table 5-14. Additional characteristics and quantitative data for associations of blood and
bone Pb with hypertension measures for results presented in Figure 5-31
Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
(95% CI)
Scinicariello et
al. (2010)
6,016 NHANES
III participants >
17 y
Hypertension
(current use of
antihypertensive
medication, SBP
> 140 mmHg, or
DBP > 90 mmHg)
Blood Pb:
Mean (SE): 2.99 (0.09) ^ig/dL
Q1 0.7-1.4 |jg/dL,
Q2 1.5-2.3 |jg/dL,
Q3 2.4-3.7 |jg/dL,
Q4 3.8-52.9 ^ig/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
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
Non-Hispanic blacks:
Q1 Reference
Q2 POR=1.83 (1.08,
3.09)
Q3 POR=2.38 (1.40,
4.06)
Q4 POR=2.92 (1.58,
5.41)
ALAD1-2/2-2: POR= 3.40
(0.05,219.03)
ALAD-1 reference
Mexican Americans:
Q1 Reference
Q2 POR=0.74 (0.24,
2.23)
Q3 POR=1.43 (0.61,
3.38)
Q4 POR=1.27 (0.59,
2.75)
ALAD1-2/2-2: POR= 0.49
(0.08, 3.20)
ALAD-1 reference
POR for hypertension
with ALAD2 carriers
across quartiles of blood
Pb level also reported.
ALAD2 carriers
associated with
hypertension in non-
Hispanic whites.
Parketal.
12,500 NHANES Hypertension
NHANES III Blood Pb
Logistic regression models
OR's per SD (0.75 |jg/dL)
(2009)
III participants
3.52 (0.10)
adjusted for age, education,
in log blood Pb:
smoking status, cigarette
Overall
White men
smoking, BMI, hematocrit,
1.12(1.03,1.23).
<50 yr 4.02 (0.16)
alcohol consumption, physical
White men: 1.06 (0.92,
>50 yr 4.92 (0.18)
activity, antihypertensive
1.22)
Black men
medication use, and diagnosis
Black men: 1.17 (0.98,
<50 yr 4.55 (0.15)
of type-2 diabetes
1.38)
>50 yr 7.57 (0.22)
White women:1.16 (1.04,
White women
1.29)
<50 yr 2.09 (0.07)
Black women: 1.19 (1.04,
>50 yr 3.53 (0.12)
1.38)
Black women
Men <50 yr
<50 yr 2.52 (0.09)
0.98 (0.80,1.22)
>50 yr 4.49 (0.16)
Men >50 yr
1.20 (1.02,1.41),
Women <50 yr
1.23 (1.04,1.46),
Women >50 yr
1.09(1.94,1.26).
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Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
(95% CI)
Muntneret al.
(2005)
9,961 NHANES
(1999-2002)
participants
Hypertension
(current use of
antihypertensive
medication, SBP
> 140 mmHg, or
DBP > 90 mmHg)
Blood Pb:
Overall Mean (CI): 1.64 (1.59-
1.68) |jg/dL
quartile 1: <1.06 ^ig/dL,
quartile 2:1.06-1.63 |jg/dL,
quartile 3:1.63-2.47 |jg/dL,
and quartile 4: >2.47 |jg/dL
Multivariable logistic
regression models adjusted for
age, sex, diabetes mellitus,
BMI, cigarette smoking,
alcohol consumption, high
school education, and health
insurance status
Adjusted OR of having a
blood Pb level of 10 |jg/dL
Non-Hispanic white:
Q1 reference
Q2 OR=1.12 (0.83, 1.50)
Q3 OR=1.03 (0.78, 1.37)
Q4 0R=1.10(0.87, 1.41)
Non-Hispanic black
Q1 reference
Q2 OR=1.03 (0.63, 1.67)
Q3 OR=1.12 (0.77, 1.64)
Q4 0R=1.44 (0.89, 2.32)
Mexican American
Q1 reference
Q2 OR=1.42 (0.75,2.71)
Q2 0R=1.48 (0.89, 2.48)
Q3 0R=1.54 (0.99, 2.39)
Significant trend (p=0.04)
Martin et al.
(2006)
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) ^ig/dL
Tibia Pb:
Mean (SD): 18.i
(12.4) Mg/g
Logistic regression models
adjusted for age, sex, BMI,
antihypertensive medication
use, dietary sodium intake,
dietary potassium intake, time
of day, testing technician, and
serum homocysteine
Blood Pb level
OR=1.02 (0.87,1.19)
Tibia Pb
OR=1.24 (1.05, 1.47)
mmHg per |jg Pb/dL
blood
mmHg per |jg Pb/g bone
Peters etal.
(2007)
513 elderly men
(mean 67 y)
from Normative
Aging Study in
Greater Boston,
MA area
Hypertension
(mean SBP >140
mmHg, DBP >90
mmHg; or
physician
diagnosis)
Tibia Pb:
mean (SD): 21.5 (13.4) |jg/g
Patella Pb:
Mean (SD): 31.5 (19.3) |jg/g
Cox proportional hazards
models 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
Hypertension Incidence
High Stress
RR=2.66 (1.43, 4.95) per
SD increase in tibia Pb
RR=2.64 (1.42, 4.92) per
SD increase in patella Pb
Yazbeck etal.
(2009)
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) |jg/dL
No PIH group mean (SD): 1.9
(1.2) |jg/dL
Q1: <1.20 |jg/dL
Q2:1.20-1.70 ^ig/dL
Q3: 1.71-2.30 ^ig/dL
Q4: >2.30 ^ig/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
Log Blood Pb
OR=3.29 (1.11, 9.74) per
1 |jg/dL maternal blood
Pb level
Q1: reference
Q2: OR 1.84 (0.77, 4.41)
Q3:OR=2.07 (0.83, 5.13)
Q4: OR=2.56 (1.05, 6.22)
Weaver etal.
(2008)
652 current and
former Pb
workers in South
Korea (12/1999-
6/2001)
Hypertension
(mean SBP > 140
mmHg, DBP > 90
mmHg; and/or
use of
antihypertensive
medications; or
physician
diagnosis)
Blood Pb:
Mean (SD): 31.9 (14.8) |jg/dL
Patella Pb:
Mean (SD): 37.5 (41.8) |jg/g
Logistic regression models
adjusted for age, gender, BMI,
diabetes, antihypertensive and
analgesic medication use, Pb
job duration, work status,
tobacco and alcohol use
None of the Pb exposure
metrics examined were
(blood, patella, and In
patella) were significantly
associated with
hypertension (results not
reported)
Chen etal.
(2006)
2,994,072
pregnant women
in United States
(1998)
PIH
(gestational
hypertension as
increased SBP of
> 30 mmHg or
DBP of> 15
mmHg after 20th
wk of gestation)
Pb in TSP
Seasonal mean at conception:
0.0940 ijg/m3
Seasonal mean at birth:
0.0950 Ljg/m3
Generalized estimating
equations (GEEs) adjusted for
maternal age, race, education,
OR at conception
Q1 Referent
Q2 1.07 (1.05,1.08)
marital status, parity, adequacy Q3 1.22 (1.20,1.25)
of care, and tobacco use
Q4 1.16 (1.15,1.18)
0.05 LJg/m3 increased .04
(1.03.1.04)
OR at birth
Q1 Referent
Q2 1.07 (1.05, 1.09)
Q31.21 (1.19, 1.23)
Q41.15(1.13, 1.17)
0.05 LJg/m3 increased .04
(1.04.1.05)
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Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
(95% CI)
Elmarsafawy et
al. (2006)
471 elderly men
(mean 67 y)
from Normative
Aging Study 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) ^ig/dL
Tibia Pb:
Mean (SD): 21.6 (12.0) |jg/g
Patella Pb:
Mean (SD): 31.7 (18.3) |jg/g
Logistic regression models
adjusted for age, 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)
References not included in Figure 5-31 are included in this table.
In another cross-sectional analysis, Perlstein et al. (2007) examined the association of BP and pulse
pressure (PP) among predominantly white older adults in the greater Boston area. 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. A statistically significant association between blood Pb and
DBP was observed in adjusted models but the correlations of BP with tibia Pb were not significant. Men
with tibia Pb above the median had a significantly higher mean PP compared to men with tibia Pb below
the median (4.2 mmHg [95%CI: 1.9, 6.5]). The trend toward increasing PP with increasing tibia Pb was
significant although none of the confidence intervals for PP referenced to the lowest quintile of tibia Pb
excluded the null value.
Peters et al. (2007) examined the modification by self-reported stress of the associations of Pb
exposure (tibia and patella Pb) with BP and hypertension in a cohort of subjects enrolled in the VA-NAS.
Cross-sectional analyses of the effect of bone Pb and stress on BP and hypertension were conducted.
Increased but nonsignificant associations between hypertension status and SBP with bone Pb were
observed. Interaction of stress with tibia Pb (tibia Pb (3=3.77 [CI: 0.46, 7.09]) and stress with patella Pb
was significant (patella Pb |3=2.60 [CI: -0.95, 6.15]) in systolic BP models (neither bone, self-reported
stress, nor the interaction predicted DBP). Figure 5-32 shows the association between SBP and tibia Pb,
comparing those with high and low self-reported stress. Peters et al. (2007) also used Cox proportional
hazards models to assess the interaction of stress and bone Pb level on the association of bone Pb with the
development of hypertension among those free of hypertension at baseline. The results of this analysis
showed interactions between both tibia and patella Pb and the incidence of hypertension (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 provide information regarding factors that
moderate or modify Pb effects on cardiovascular health. Gump et al. (2005) described significantly
greater total peripheral vascular resistance (TPR) associated with increased blood Pb among 122 children
(mean age 9.5 years) under acute stress, assessed by minor tracing or reaction time tasks which are
consistent with a and [3 adrenergic activation. In a new analysis (Gump et al.. 2007) significant effects of
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SES on stress-induced reactivity of BP and TPRare reported. A significant SES by blood Pb level
interaction is also reported suggesting possibly heightened effects of blood Pb level on stress changes in
TPR and BP in low SES groups.
140 -i
E
E 125 -
®> 130 -
115-
135 -
4 O High perceived
• Low perceived:
•
, ~ f • O High perceived stress
« • Low perceived stress
Trend (high stress)
• Trend (low stress)
110 1 1 1 1 1
-10 0 10 30 50 70 90
Tibia lead (^g/g)
Source: Peters et al. (2007)
Figure 5-32. The relationship between tibia Pb and estimated SBP for those
with high self-reported stress versus those with low self-
reported stress.
Elmarsafawy et al. (2006) examined the modification of Pb effect by dietary calcium, with 467
subjects from the VA-NAS. Responses on a semi-quantitative dietary frequency questionnaire with one-
year recall were converted to estimated calcium intake. Hypertension was modeled using logistic
regression and included interaction terms between Pb (tibia, patella and blood Pb) and a dichotomized
calcium intake variable (split at 800 mg/day). They also constructed alternative models stratified on the
calcium variable. Nonsignificant increases in hypertension associated with elevated blood, tibia, and
patella Pb were observed in both low and high calcium groups. The only significant interaction reported
was between BMI and calcium in a tibia Pb model. The authors report that in linear regression models of
BP stratified by calcium status, SBP increased 0.40 (95% CI: 0.11, 0.70) mmHg for every 1 (.ig/g increase
in tibia Pb concentration in the high calcium group, and 0.19 (95% CI: 0.01, 0.37) mmHg in the low
calcium group.
Glenn et al. (2006) simultaneously modeled the multiple Pb dose measures of individuals over
repeated time periods, assessing cross-sectional as well as longitudinal relationships. The initial blood Pb
level was used as a baseline covariate and the difference in blood Pb level between visits were computed
for subsequent visits. The bone Pb measures were used to indicate historical exposure and cumulative
exposure. Four models were specified: Model 1 was conceptualized to reflect short-term exposure; Model
2 to reflect longer-term exposure controlling for recent dose; Model 3 to reflect longer-term exposure
controlling for cross-sectional influence of cumulative dose; and Model 4 to reflect both short-term
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change with recent dose and longer-term change with cumulative dose. Concurrent and longitudinal
measures of blood Pb were associated with SBP in Model 1 (short-term exposure) and Model 4 (short-
and longer-term exposures). No associations with tibia Pb at baseline were observed while historical tibia
Pb metric was negatively associated with SBP in each of the models. This study suggests that Pb exposure
may act continuously on systolic BP and reduction in exposure may contribute to reductions in BP, while
cumulative Pb burden may contribute to hypertension incidence by other mechanisms over longer time
periods. Elevated BP may reflect an immediate response to Pb at a biochemical site of action as a
consequence of recent dose or a persistent effect of cumulative doses over a lifetime.
In a separate analysis of the third year cross-sectional results of the same occupationally-exposed
group, Weaver et al. (2008) examined associations between patella Pb and blood Pb level and SBP, DBP,
and hypertension to determine interactions of the patella Pb effects with ALAD and vitamin D receptor
(VDR) polymorphisms. None of the Pb exposure metrics were associated with DBP. Patella Pb alone was
not significantly associated with SBP, while blood Pb, either alone or with patella Pb was positively and
significantly associated with SBP. The patella Pb-age and blood Pb-age interactions were not significant.
There were no significant effects of blood Pb or patella Pb on 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 SBP were not significant. Mean blood Pb level was high (30.9 j^ig/dL) compared to non-
occupational groups.
Weaver et al. (2010) provided the results of further analysis of the Korean worker cohort (V. M.
Weaver et al.. 2008). with a focus on determining functional form of the concentration-response
relationships. The coefficient indicates that every doubling of blood Pb level is 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 & 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.
Yazbeck et al. (2009) conducted a cross-sectional study examining a community-based group of
pregnant women to determine the association of Pregnancy Induced Hypertension (PIH) with blood Pb
level and unlike most other studies adjusted their model for metal blood concentrations of cadmium,
manganese, and selenium. 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. An association between blood Pb and PIH was observed (OR 3.29 [95% CI: 1.11, 9.74])
between PIH cases (2.2 ±1.4 (ig/dL blood Pb) and normotensive patients (1.9 ±1.2 (ig/dL blood Pb).
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
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PIH. They observed no significant interactions among blood Pb level, any of the other elements, and
maternal characteristics in predicting the risk of PIH. Interaction between selenium and Pb concentrations
was not significant, and the putative protection effects of selenium through antioxidative properties were
not confirmed in this study.
Chen et al. (2006) reported an ecological study of air Pb concentration and PIH, aggregated by all
50 U.S. states and the District of Columbia. The PIH data were taken from the live births and infant
deaths up to 1 year of age compiled by the National Center for Health Statistics (CDC. 2000) for the year
1998. Associations between state level air Pb and state level PIH were reported (OR 1.04 [95% CI: 1.03,
1.04] per 0.05 |_ig/m3 Pb). No individual level data were used in the analysis. Wells et al. (2011) measured
the influence of cord blood Pb on BP in 285 women at admission to the Johns Hopkins Hospital in
Baltimore, MD, during labor and delivery. Women in the fourth quartile of blood Pb elevations (>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 benchmark dose (BMD) and the associated lower confidence
limit for benchmark dose (BMDL) for one standard deviation (SD) increase in BP, which is
approximately equivalent to a 10% increase above the mean for the first quartile blood Pb "controls". The
BMD approach is used here only as a means of characterizing the exposure level where effects might be
found. These BMDL results indicate that the 95% lower bound confidence limit on the venous blood Pb
level that is associated with a 1 SD increase is about 1.85 (ig/dL for all BP outcomes. While these reported
results are similar to those in the 2006 Pb AQCD as well as those found 25 years ago but with blood Pb
levels an order of magnitude lower, the authors did not provide enough information to allow for
verification of the BMD analysis.
Zhang et al. (2010) examined the effect of polymorphisms of the hemochromatosis gene (HFE) on
the bone Pb effect on PP among older adult men participating in the VA-NAS. 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 the cross-sectional PP study of many of the same subjects
of the VA-NAS group, without genotyping, reviewed above (Perl stein et al.. 2007). The effect of bone Pb
(tibia and patella) among those with the H63D variant was greater compared to those with the wild-type
or the C282Y variant. In another gene-environment interaction analysis, Scinicariello et al. (2010) used
NHANES III (1988-1994) data to examine the interaction between ALAD genotype and blood Pb in
relation to BP in a cross-sectional analysis. A significant interaction between log blood Pb level and
ALAD 1-2/2-2b among non-Hispanic whites and non-Hispanic blacks was observed. In addition,
associations of blood Pb with SBP and DBP across race/ethnicity strata were presented. The strongest
associations were observed among non-Hispanic blacks. Scinicariello et al. (2010) also examined the
association of blood Pb level with hypertension. Significant associations between blood Pb level and
hypertension were observed among non-Hispanic blacks and nonsignificant increases were observed
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among non-Hispanic whites and Mexican Americans (with the exception of Q2 association for Mexican
Americans.) In addition, non-Hispanic white ALAD2 carriers in the highest blood Pb level quartile had a
significantly higher association with hypertension compared with ALAD1 homozygous individuals.
Muntner et al. (2005) also used the NHANES data (1999-2002) to examine the cross-sectional
effect of blood Pb on 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 among those with blood Pb measurements (1.28 (ig/dL
[95% CI: 1.23, 1.33] in the 18-39 age group to 2.32 (ig/dL [95% CI: 2.20, 2.44] in the 75 and older age
group.) Associations were observed between blood Pb level and hypertension across race/ethnicity groups
with significant trends observed for non-Hispanic blacks and Mexican Americans. Park et al. (2009)
examined the association of blood Pb as well as bone Pb, which was predicted from blood Pb, with
hypertension. The predicted bone Pb metrics were derived from models using VA-NAS data and applied
to the NHANES (1988-1994) population. Blood Pb was associated with hypertension overall in the
NHANES part of this study, with larger associations among black men and women as well as older adults.
Associations with estimated bone Pb were also observed.
5.4.2.2. Toxicology
An array of studies have provided evidence that extended exposure to low levels of Pb (<5 (ig/dL)
can result in delayed onset of hypertension in experimental animals that persists long after the cessation of
Pb exposure (U.S. EPA. 2006). Tsao et al. (2000) found significantly increased systolic and diastolic BP
in rats with blood Pb levels relevant to human exposure (2.15 (ig/dL). As this was the lowest Pb level
tested, no evidence of a threshold was evident. After Pb exposure is removed, blood, heart, aorta, and
kidney Pb levels decreased quickly within the first three months (H.-R. 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. Between 4 and 7 months after Pb exposure, the still-
elevated BP did not decrease further, thus never returning to control BP levels. Decreases in BP were
closely correlated with decreases in blood Pb level after exposure cessation. Prenatal Pb exposure in rats
given a low calcium diet also resulted in increased arterial pressure (Bogden et al.. 1995).
Experimental animal studies continue to provide evidence to conclude that Pb exposure results in
delayed, yet sustained arterial hypertension. Increased systolic BP developed in rats after exposure to 90 -
10,000 ppm Pb (as Pb acetate in drinking water) for various time periods resulting in blood Pb level
between 19.3-240 (.ig/dL (Badavi et al.. 2008; Bagchi & Preuss. 2005; Bravo et al. 2007; Grizzo &
Cordellini. 2008; Hevdari et al.. 2006; Reza et al.. 2008; Vargas-Robles et al.. 2007; L.-F. Zhang et al..
2009). However, past studies have shown statistically significant elevations in BP in rats with lower blood
Pb level (Figure 5-33). Consistent with measurements of systolic BP, Pb exposure (100 ppm for 14
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1 weeks; blood Pb level 24 (ig/dL) also caused an increase in intra-aortic mean arterial pressure (Bravo et
2 al.. 2007). One study tested low levels of Pb exposure (30 ppm; blood Pb level 7.6 (ig/dL) and did not
3 find a statistically significant increase in systolic BP despite elevated blood Pb level after 8 weeks of
4 treatment, however the data do represent a trend of increasing BP (Rizzi et al.. 2009). Additionally, pups
5 of Pb exposed dams (1,000 ppm through pregnancy and lactation) exhibited increased blood Pb level
6 (58.7 (ig/dL) and increased arterial systolic BP after weaning (Grizzo & Cordcllini. 2008) suggesting a
7 role for childhood Pb exposure leading to adult disease.
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Chang et al. (2005")
Tsao et al. (2000")
Rizzi et al. (2009")
Chang et al. (1997")
Carmignani et al. (2000)
Heydari et al. (2006")
Reza et al. (2008")
Bravo et al. (2007")
Zhang et al. (2009")
® Dose with Statistically Significant Effect
¦ Dose with No Statistically Significant Effect
10 15 20 25 30
Blood Pb (ijg/dL)
35
Note: Red square = blood Pb level where no statistically significant
change in BP was observed; Blue circles = lowest blood Pb level reported
with statistically significant changes in BP; Arrow line = higher blood Pb
level reported in the same study with significant changes in SBP; A SBP =
the change in SBP from control to first statistically significant blood Pb
level in mmHg; n = number of animals in treatment group.
Figure 5-33. Rat blood Pb levels reported to be associated with changes in
SBP from the current ISA and 2006 Pb AQCD.
1 Pb induced hypertension persists long after cessation of Pb exposure. Bagchi and Preuss (2005)
2 found that elevated systolic BP was maintained for 210 days after Pb exposure cessation. Chang et al.
3 (2005) reported a partial reversibility of effect after cessation of Pb exposure, where Pb-induced elevated
4 BP decreased but did not return to control levels 7 months post Pb exposure. However, chelation therapy
5 using Na2CaEDTA was able to return systolic BP to levels comparable to untreated rats (Bagchi & Preuss.
6 2005). Studies reporting the effect of Pb (as blood Pb level) on systolic BP in unanesthetized adult rats
7 since 1992 report a positive increase in BP with increasing blood Pb level (Figure 5-34).
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(Malvezzi et al.. 2001; Vaziri. Liang, et al.. 1999). Oxidative stress from Pb exposure in animals may be
due to upregulation ofNAD(P)H oxidase (Ni et al.. 2004; Vaziri et al.. 2003). induction of Fenton and
Haber-Weiss reactions (Ding et al.. 2000; Ding et al.. 2001). 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 plays a role in Pb-induced downregulation of sGC (Courtois
et al.. 2003; Farmand et al.. 2005; M. Marques et al.. 2001). The reduction of the vasodilator NO leads to
increased vasoconstriction and BP.
Pb-induced oxidative stress also induces renal tubulointerstitial inflammation which plays a crucial
role in models of hypertension (Rodriguez-lturbe et al.. 2005; Rodriguez-lturbe et al.. 2004).
Tubulointerstitial inflammation from treatment with Pb has been coupled with activation of the redox
sensitive NFkB (Ramesh et al.. 2001). Pb-induced hypertension, inflammation, and NFkB activation can
be ameliorated by antioxidant therapy (Rodriguez-lturbe 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 continue to provide evidence for the role of ROS and NO metabolism in Pb-
induced hypertension and vascular disease. Increased SBP after Pb exposure has been accompanied by
increased superoxide (02) and 02 positive cells (Bravo et al.. 2007; Vargas-Robles et al.. 2007), elevated
urinary malondialdehyde (MDA) (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-HT (L. F. Zhang et al.. 2005). Increased SBP, intra-
aortic mean arterial pressure, and MDA after Pb exposure (100 ppm; blood Pb level 23.7-27 j^ig/dL) were
also prevented by treatment with the immunosuppressive, mycophenolate mofetil (MMF) (Bravo et al..
2007). MMF has been shown to inhibit endothelial NAD(P)H oxidase, which could explain the decrease
in oxidative stress and BP. Red grape seed extract was 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).
Exposure to Pb can also affect the activity and levels of antioxidant enzymes. Male and female rats
exposed to Pb for 18 weeks (100-1,000 ppm) had altered responses in antioxidant enzyme found in heart
tissue (Alghazal. Lenartova. et al.. 2008; Sobekova et al.. 2009). Pb exposure (>100ppm) in female rats
increased the activity of cardiac SOD, GST, GR, and GPx and increased cardiac TBARS (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.
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Oxidative stress can trigger a cascade of events that promote cellular stress, renal inflammation,
and hypertension. As was shown previously (Rodriguez-Iturbe et al.. 2004). Pb exposure can increase
renal NFkB, 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 is also able to induce inflammation in human endothelial
cells as a model for vessel intima hyperplasia (Zelleretal.. 2010). The proinflammatory cytokine,
interleukin-8 (IL-8) protein and mRNA were increased, dose and time dependency, after in vitro Pb
exposure (5-50 (.iM). Enhanced IL-8 production was mediated through activation of the transcription
factor Nrf2 (but not NFkB, 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).
Oxidative stress affects vascular reactivity and tone through inactivation and sequestration of NO,
causing a reduction in biologically active NO. Recent studies confirm these past conclusions on the
interplay of ROS and NO metabolism in the cardiovascular effects of Pb. Elevated SBP and altered
vasorelaxation after Pb exposure is accompanied by a decrease in total nitrates and nitrites (NOx)
(Hevdari et al.. 2006; L. F. Zhang et al.. 2007). Serum NOx levels in Pb-treated rats remained depressed
for 8 weeks and then reversed after 12 weeks, despite continued elevation in SBP (Hevdari et al.. 2006).
This return of serum NOx levels could be a result of compensatory increases in endothelial NOS (eNOS)
attempting to replenish an over-sequestered NO supply. With this in mind, studies have shown increased
eNOS protein expression after chronic Pb exposure in kidney (L. F. Zhang et al.. 2007) 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 (L. F. Zhang et al.. 2007).
NO, also known as endothelium-derived relaxing factor, is a potent endogenous vasodilator.
Studies continue to investigate the effects of Pb on NO dependent vascular reactivity. Perinatal Pb
exposure (1000 ppm through pregnancy and lactation, blood Pb level 58.7 (ig/dL) resulted in a greater
increase in maximal contraction to L-NAME, which decreases NO production, with a greater effect in
the endothelium than the smooth muscle cells (Grizzo & Cordellini. 2008). Additionally, blocking NOS
with L-NAME abolished the relaxant response evoked by ACh, which triggers the release of NO from
the endothelial cell, in aortic rings of perinatally exposed rats. Cyclooxygenase (COX) inhibition
decreased the EC50 of the ACh response in Pb treated animals. This study suggests that pups exposed to
Pb through pregnancy and lactation have an altered vascular reactivity that is endothelium dependent and
occurs due to the altered release of NO and a COX-derived vasoconstrictor (Grizzo & Cordellini. 2008).
Similarly, a recent study provides evidence that acute Pb exposure increases rat tail artery reactivity
in an endothelium dependent manner due to a COX-derived vasoconstrictor and in part free radicals and
NO (Silveira et al.. 2010). Acute exposure of rat tail artery to Pb (100 (.iM. 1 h) increased reactivity to
phenylephrine. Pb exposure decreased ACh induced relaxation, suggesting damage to the endothelium. Pb
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did not affect smooth muscle integrity since sodium nitroprusside (SNP)-induced vasorelaxation was
unchanged. Inhibition of NOS increased the Pb pressor response whereas COX inhibition eliminated the
response to PHE. Treatment with the SOD mimetic Tempol decreased, but did not eliminate, the Pb
pressor response (Silveira et al. 2010). A second study showed that Pb (90 ppm) exposure did not change
the rat thoracic aortic ring relaxation response curves to the NO donor, SNP (Rizzi et al.. 2009V
Conversely, Skoczynska and Stojek ("2005) found that Pb exposure (50 ppm; blood Pb level 11.2
(ig/dL) enhanced NO-mediated vasodilation by ACh in rat mesenteric arteries and NOS inhibition
enhanced the ACh relaxant response. In rat renal interlobar arteries, Pb exposure blunted the increase in
Angll mediated contraction from NOS inhibition by L-NAME (Varaas-Robles et al.. 2007).
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
norepinephrine and plasma catecholamines (Carmignani et al.. 2000; H.-R. Chang et al.. 1997). and
decreased (3 adrenergic receptor density and (3 agonist-stimulated cAMP production in the aorta and heart
(H.-R. Chang et al.. 1997; Tsao et al.. 2000). These stimulatory effects on the SNS paralleled effects on
BP, cardiac contractility, and carotid blood flow. Increased Pb induced arterial pressure and heart rate
were abrogated by ganglionic blockade (C.-C. Lai et al.. 2002) and gradually decreased 7 months after Pb
exposure cessation along with Pb-induced SNS alterations (H.-R. Chang et al.. 2005).
Increased BP can be caused by vascular narrowing leading to increased total peripheral resistance,
resulting from activation of the SNS. 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, norepinephrine (NE), and dopamine), hormones of
the SNS. Indeed, the isolated mesenteric vessel bed from Pb treated rats (50 ppm blood Pb level: 11.2
(ig/dL, but not 100 ppm blood Pb level: 17.3 (ig/dL) exhibited increased reactivity to NE (Skoczvnska &
Stojek. 2005). Similarly, 100 ppm Pb did not affect the NE induced contractile response after 10 months
of exposure (L.-F. Zhang et al.. 2009). suggesting a small range of doses affecting pressor response to NE.
Catecholamines act primarily through the adrenergic and dopaminergic receptors. Antagonists of al-
adrenergic, a2-adrenergic, (3-adrenergic, and dopamine D1 receptors abolish Pb-induced aortic contraction
(Fazli-Tabaei et al.. 2006; Hevdari et al.. 2006). Phenylephrine-induced aortic contractions were enhanced
by treatment with Pb (100 ppm; blood Pb level: 26.8 (ig/dL), indicating a specific role for the al-
adrenergic receptor. Additionally, Pb blunted the isoproterenol-induced relaxation, supporting a role for
the [^-adrenoceptors (Hevdari et al.. 2006; Vassallo et al.. 2008).
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Recently, there has been mixed evidence for Pb disrupting vascular reactivity to other pressor
agents. One study found that Pb (50 ppm; 12 weeks; blood Pb level: 11.2 (ig/dL) increased acetylcholine
(ACh) induced relaxation in rats (Skoczvnska & Stoiek. 2005). Additionally, studies have shown no
change in ACh induced vasorelaxation after Pb exposure (Grizzo & Cordellini. 2008; Rizzi et al.. 2009).
However, Zhang et al. (2007) and Silveira et al. (2010) found that Pb (1 ppm and 100 (.iM. 1 h) blunted
ACh induced relaxation in isolated rat thoracic aorta and tail artery, respectively. Another study
investigated the influence of Pb on vasoconstriction from 5-hydroxytryptamine (5-HT). 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 (L. F. 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 (L.-F. Zhang et al.. 2009).
Studies continue to investigate the role of NO, also known as endothelium-derived relaxing factor,
in Pb induced changes in vascular reactivity. A recent study provides evidence that acute Pb exposure
increases rat tail artery reactivity in an endothelium dependent manner due in part to free radicals and
NO (Silveira et al.. 2010). Acute exposure of rat tail artery to Pb (100 (.iM. 1 hour) increased reactivity to
phenylephrine. Pb exposure decreased ACh induced relaxation, suggesting damage to the endothelium.
However, Pb did not affect smooth muscle integrity since SNP-induced vasorelaxation was unchanged.
Similarly, Pb (90 ppm) exposure did not change the rat thoracic aortic ring relaxation response curves to
the NO donor, SNP (Rizzi et al.. 2009). Inhibition of NOS increased the Pb pressor response to PHE
(Silveira et al.. 2010). Another study showed that blocking NOS with L-NAME, abolished the relaxant
response evoked by ACh, which triggers the release of NO from the endothelial cell, in aortic rings of
perinatally exposed rats (1,000 ppm through pregnancy and lactation, blood Pb level 58.7 (ig/dL) (Grizzo
& Cordellini. 2008). Additionally, perinatal Pb exposure resulted in a greater increase in maximal
contraction to L-NAME, which decreases NO production, with a greater effect in the endothelium than
the smooth muscle cells. This study suggests that pups exposed to Pb through pregnancy and lactation
have an altered vascular reactivity that is endothelium dependent and occurs due in part to the altered
release of NO (Grizzo & Cordellini. 2008). 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). NOS inhibition by L-NAME increased Angll-induced vasoconstriction in control but not Pb-
exposed arteries, suggesting impaired NO availability (Vargas-Robles et al.. 2007). Conversely,
Skoczynska and Stojek (2005) found that Pb exposure (50 ppm; blood Pb level 11.2 (ig/dL) enhanced
NO-mediated vasodilation by ACh in rat mesenteric arteries and NOS inhibition enhanced the ACh
relaxant response.
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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 blood pressure regulation, and has been shown to be affected by Pb
exposure. Meta-analysis found that Pb exposure (30-40 (ig/dL) increases plasma renin activity and renal
tissue renin in young rats, but not old (Vandcr. 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 killikrein activities; and renal angiotensin II (Angll) positive cells
(Carmignani etal.. 1999; Rodriguez-Iturbe et al.. 2005; Sharif! et al.. 2004). 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 continue to implicate the RAAS in the development of Pb-induced hypertension,
especially during early exposure in young animals. Low level Pb (100 ppm, 14 weeks; blood Pb level
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 SBP and was ablated by the anti-inflammatory antioxidant, MMF.
Similarly, early high level Pb (1% Pb, 40 days; blood Pb level >240 (ig/dL after exposure, 12-13 (ig/dL
after chelation after 1 year) accumulation resulted in sustained hypertension (Bagchi & Preuss. 2005).
Treatment with the Angll receptor blocker, Losartan, resulted in a greater decrease in SBP in Pb exposed
rats than control rats that continued into later periods of follow-up (day 283). Increased SBP 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. 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). These studies point to the activation of the RAAS in the course of Pb-induced
hypertension.
Vasomodulators
The balance between production of vasodilators and vasoconstrictors is important in the regulation
of blood pressure and cardiovascular function. The previous AQCD reported that the effects of Pb on
vasomodulators are contradictory. Urinary excretion of the vasoconstrictor, thromboxane (TXB2), and the
vasodilatory prostaglandin, 6-keto-PGFla, were unchanged in rats with Pb-induced hypertension (Gonick
etal.. 1998). However, in vitro Pb exposure promoted the release of the prostaglandin substrate,
arachidonic acid, in vascular smooth muscle cells (VSMC) via activation of phospholipase A2 (Dorman &
Freeman. 2002). Plasma concentration and urinary excretion of the vasoconstrictive peptide, endothelin
(ET) 3 was increased after low (100 ppm), but not high level (5000 ppm) Pb exposure in rats (Gonick et
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al.. 1997; Khalil-Manesh et al.. 1994; Khalil-Manesh. Gonick. Wcilcr. et al.. 1993). 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.. 2003V Additionally, Pb-exposed animals exhibited fluid retention and a dose dependent
decline in the vasodilator, atrial natriuretic factor (ANF) (Giridhar & Isom. 1990). These studies suggest
that Pb may interfere with the balance between vasodilators and vasoconstrictors forming the hormonal
regulation of vascular contraction and blood pressure.
The imbalance in vasomodulators is one explanation for the concentration-dependent
vasoconstriction observed after Pb exposure (Piccinini et al.. 1977; Valencia et al.. 2001; Watts et al..
1995). Vasoconstriction after Pb exposure was not reported in all studies (Shelkovnikov & 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 attenuation of acetylcholine- and NO-mediated vasodilation
(M. 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
(Piccinini et al.. 1977; Valencia et al.. 2001; Watts et al. 1995).
One 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 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 decreases the Pb induced pressor response to ACh
(Grizzo & 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.
5.4.2.4. Summary
The 2006 Pb AQCD reported a clear positive association between blood Pb level and BP. The effect
was modest, but highly significant, as determined by a meta-analysis (Nawrot et al.. 2002) of over 30
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studies comprising over 40,000 subjects (Figure 5-35), reporting that each doubling of blood Pb was
associated with a 1 mmHg increase in systolic BP and a 0.6 mmHg increase in diastolic BP. Recent
studies support this conclusion at lower blood Pb levels (<2 (ig/dL) and add to the evidence base on
susceptibility factors and bone Pb associations with BP and hypertension at levels <20 jj.g/g. Associations
of bone and/or blood Pb with systolic BP and hypertension were higher among non-whites, those
reporting high stress, and those with the HFE H63D and ALAD genotypes.
A recent study in an ethnically diverse community-based cohort of women and men aged 50-70
years of age suggests that Pb has an acute effect on BP as a function of recent dose measured by blood Pb
and a chronic effect on hypertension risk as a function of cumulative exposure measured by tibia Pb
(Martin et al.. 2006). This study verified other studies by demonstrating that with each increase of 1
(ig/dL blood Pb level, systolic BP would increase 1 mmHg and diastolic BP would increase 0.5 mmHg.
Additionally, recent epidemiologic studies provided evidence for associations between blood Pb and BP
and hypertension at relatively low blood Pb level; a positive relationship was found in the NHANES data
(1999-2002) at a geometric mean blood Pb level of 1.64 (ig/dL (Muntner et al.. 2005). Animal
toxicological studies also provide support for effects of low blood Pb level on increased BP with
statistically significant increases shown as low as 2 (ig/dL (Tsao et al.. 2000). Collectively, all animal
toxicological studies providing blood Pb level and BP measurements report positive increases in BP with
increasing blood Pb level (Figure 5-34). New studies also demonstrate reversibility of Pb-induced
increased BP following Pb exposure cessation or chelation.
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POCOCK 81,
KRCMHOUT 85
ORSSAUD 85
WEISS 86
DE KORT 87
LOCKETT 87
PARKINSON 87
RABiNOWITZ 87
EL WOOD [CI 58
ELWOOQ [HP! 88
ELWOOD (HP) 88
GARTSIOE [WJ 88
GARTSIOE [81 B8
GARTSIOE [WJ 88
GARTSIOE IB) 88
NERI [FWI 88
NERI 88
GRANDJEAN 89
GRANDJEAN 89
REIMER 89
APOSTOLI 90
APOSTOll 90
MORRIS 90
MORRIS 90
SHARP Ml 90
SHARP IB! 90
STAESSEN 90
STAESSEN 90
M01LER 92
HEN5E 93
HENSE 93
MAHESWARAN 93
MENDITTO 94
PROCTOR 96
STAESSEN (P) 96
STAESSEN (PI 96
SOKAS IWI 97
BOST 99
BOST 99
CHll 99
CHU 99
ROTHENBERG (Nil 99
ROTHENBERG [I) 99
SCHWARTZ 00
DEN HOND (W) 01
DEN HOND (B) 01
DEN HOND IW) 01
DEN HOND IB] 01
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diastolic BP. A pooled estimate for an increase in systolic BP of 0.26 mmHg (95% CI: 0.02, 0.50) for the
cross-sectional studies was reported. The estimate for the longitudinal studies was 0.33 mmHg (95% CI: -
0.44, 1.11). With the exception of one study, positive associations of bone Pb with hypertension were also
reported. Pooled estimates of 1.04 (95% CI: 1.01, 1.07) per 10 jj.g/g increase in tibia Pb and 1.04 (95%
CI: 0.96, 1.12) per 10 jj.g/g increase in patella Pb were reported.
Recent epidemiological studies have also emphasized the interaction between long-term Pb
exposure and factors that moderate or modify the Pb effect, like chronic stress and metabolic syndrome,
on BP and hypertension. Bone Pb coupled with high stress was associated with a strong and reliable
increased risk of developing hypertension in an originally nonhypertensive group (Peters et al.. 2007).
Also, long duration Pb exposure interacted with components of the metabolic syndrome to drive HRV in
directions associated with increased cardiovascular events (Park et al.. 2006).
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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*6
Cheng 200113
Glenn 2003'z
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 2001« 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)
Schwartz20 2000 14.4 0.74 (-0.73 to 2.21)
Korrick16 1999 13.3 -
Overall: 0.26 (0.02 to 0.50)
Patella lead
Cheng13 2001 31.4 -
Hu 961B/Cheng 0113 32.1 0.29 (-0.36 to 0.95)
Korrick161999 17.3 -
-10 12
Increase in SBP (mmHg / year)
-10 12
Increase in DBP (mmHg / year)
005 I 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: Used with permission from Elsevier Publishers, Navas-Acien et al. (2008)
In the Normative Aging Study, Hu et al (1996) reported the cross-sectional association between
bone Pb levels and the prevalence of hypertension and Cheng et al (2001) reported the cross-
sectional association between bone Pb levels and SBP 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 5-36. Prospective and cross-sectional increase in SBP and DBP and
relative risk of hypertension per 10 pg/g increase in bone Pb
levels.
1 Recent epidemiologic studies investigated the interaction of genotypes with effects of Pb on the
2 cardiovascular system. Significant evidence was presented for modification of the effect of blood Pb level
3 on BP by ALAD genotype (Scinicariello et al.. 2010). Additionally, polymorphisms in the
4 hemochromatosis gene modified the pulse pressure response to bone Pb exposure, where pulse pressure
5 represents a good predictor of cardiovascular morbidity and mortality and an indicator of arterial stiffness
6 (A. Zhang et al.. 2010). Park et al. (2009) provided further evidence of gene variants, specifically those
7 related to iron metabolism, impacting the effect of long-term Pb exposure on the cardiovascular system,
8 evaluated by QT interval changes.
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Not only has Pb exposure been shown to increase BP and hypertension, but Pb exposure can
contribute to the development of other cardiovascular diseases. Recent epidemiologic and toxicological
studies provide evidence for increased atherosclerosis, thrombosis, ischemic heart disease, peripheral
artery disease, arrhythmia, and cardiac contractility.
Animal toxicological evidence continues to build on the evidence supporting the mechanisms
leading to these cardiovascular alterations. 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 consistently positive
associations observed between blood and bone Pb and cardiovascular effects.
5.4.3. Vascular Effects and Cardiotoxicity
Not only has Pb exposure 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 &
Kopp. 1985; Re vis et al.. 1981). Epidemiologic studies discussed in the previous AQCD provided limited
evidence to support the association of ischemic heart disease (IHD) and peripheral artery disease (PAD)
with increased blood Pb.
5.4.3.1. Effects on Vascular Cell Types
The endothelium layer is an important constituent of the blood vessel wall, which regulates
macromolecular permeability, vascular SMC tone, tissue perfusion, and blood fluidity. Damage to the
endothelium is an initiating step in development of atherosclerosis, thrombosis, and tissue injury. Given
that chronic Pb exposure promotes a number of these diseases, numerous studies have investigated the
role of Pb on endothelial dysfunction. The endothelial layer makes up only a small part of the vascular
anatomy; the majority of the vessel wall is composed of vascular SMC, which work in concert with the
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, studies
continue to investigate the effects of Pb on SMC.
Pb exposure (50 (.iM. 2 weeks) stimulated SMC invasiveness in isolated human arteries leading to
the invasion of medial SMC into the vessel intima and development of intimal hyperplasia, a key step in
atherosclerotic progression (Zeller et al.. 2010). In addition, Pb exposure (50 (.iM. 12 hours) promoted
SMC elastin expression and increased arterial extracellular matrix in isolated human arteries. SMC
invasiveness was also increased in culture by treatment with supernatant of Pb-treated human EC (50
(iM), suggesting that Pb-exposed EC secrete an activating compound. This compound was confirmed to
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be interleukin-8 (IL-8). Pb exposure (5-50 |_iM) was able to dose-dependently 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 SMC invasion and arterial intima thickening (Zeller et al.. 2010). This
study provides evidence that Pb exposure stimulates EC to secrete IL-8 in an Nrf2-dependent manner that
stimulates SMC invasion from the vessel media to intima leading to a vascular thickening and possibly
atherogenesis.
A number of cardiovascular diseases, including atherosclerosis, are characterized by increased
inflammatory processes. Numerous studies have shown that Pb exposure is able to induce an
inflammatory environment in humans and animals by increasing inflammatory mediators like
prostaglandin E2 (PGE2). Human aortic vascular SMC treated with Pb (1 (.iM, 1-12 hours) exhibited
increased secretion of PGE2 time-dependently through enhanced gene transcription fW. C. Chang et al.).
This was preceded by a Pb-induced increase in 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) attenuated Pb-induced cPLA2 and COX activation and PGE2 secretion. Overall this
study suggests that Pb can induce proinflammatory events in vascular SMC in the form of increased PGE2
secretion and cPLA2 and COX-2 expression through activation of EGFR via ERK1/2 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 (Fuiiwara et al.. 1997; kaii et al.. 1995; kishimoto etal.. 1995;
Ueda et al.. 1997). However, Pb exposure does not 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 results in specific cytotoxicity (caspase3/7 activation) through
endoplasmic reticulum (ER) stress that can be 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 EC induces 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.
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5.4.3.2. Cholesterol
As blood cholesterol rises so does the risk of coronary heart disease. Early occupational studies
(Ademuviwa. Ugbaia. Idumebor. et al.. 2005; Bener et al.. 2001; kristal-Boneh et al.. 1999) at higher than
current blood Pb levels reported higher total cholesterol levels related to Pb exposure, and mixed results
for HDL, LDL, and triglycerides. More recently, Poreba et al. (2010). in an occupational study, reports no
significant differences in parameters of lipid metabolism between Pb exposed and unexposed
hypertensive patients. Other Pb studies (Menke et al.. 2006) adjust models for total cholesterol to control
for this coronary heart disease risk factor and note that mean total cholesterol was higher at higher blood
Pb levels. In developing models to predict bone Pb levels, Park et al. (2009) noted that total and high-
density lipoprotein cholesterol were selected as 2 of 18 predictors for the bone Pb level model. Their
findings suggested that total and HDL cholesterols, as part of a larger groups of predictors, may be critical
contributors to such prediction models. The major risk factor that lipids represent for heart disease make
relating lipid levels to lead exposures an interesting but challenging hypothesis to test.
5.4.3.3. Heart Rate Variability
Pb has been shown to not only affect vascular contractility but also cardiac contractility. Park et al.
(2006) investigated the interaction of key markers of the metabolic syndrome and bone Pb effect on heart
rate variability (HRV) in a group of 413 older adults with patella Pb measurements from the VA-NAS.
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 blood pressure
>130/85 mmHg, and high fasting glucose (>110 mg/dL). Those using antihypertensive medication or
diabetes medications were counted as high BP or high fasting glucose, respectively. The strongest
relationships with bone Pb were among those with three or more metabolic abnormalities. Tests for trend
by number of metabolic abnormalities was significant at p<0.05 for patella Pb. These results suggest
multiplicative effects of long duration Pb exposure 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 (nonsignificant) bone Pb main effects on HRV measures were linear. However,
the relationship with LF/HF was linear in log(LF/HF).
Park et al. (2009) followed up a previous report (Y. Cheng et al.. 1998). which found prolongation
of corrected QT interval (QTc) with increasing bone Pb in men <65 years, but not in men > 65. In the
recent work, authors stratified multiple regression models on polymorphisms in three genes known to
alter iron metabolism, hemochromatosis (HFE), transferrin (TF), and heme oxygenase-1 (HMOX-1), and
related QTc intervals to blood, tibia and patella Pb level. They also used interaction models with cross-
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product terms between genotype and the Pb biomarker. The distributions of all genotypes but the HFE
mutant, H63D, were in Hardy-Weinberg equilibrium. Subjects homozygous for the other HFE mutant,
C282Y, had higher bone Pb concentrations and those homozygous for H63D and heterozygous for both
C282Y and H63D had lower bone Pb. The HMOX-1 variant alone, compared to the wild type, showed a
significant interaction with tibia Pb (increased QTc [11.35 msec] for each 13 jj.g/g increase in bone Pb in
L-allele variants). No other gene variant alone showed significant QTc differences from wild types, either
for tibia and patella Pb or for (linear) blood Pb. Lengthening of QTc with increased 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. Tests for linear trend in QTc by increasing number of gene
variants from 0 to 3 were significant for both tibia and blood Pb. This study provided further suggestive
evidence of gene variants impact on long-term Pb exposure on the cardiovascular system.
Increased incidence of arrhythmia and atrioventricular conduction block increased in rats after 12
weeks of Pb exposure (100 ppm; blood Pb level 26.8 (ig/dL) (Reza et al.. 2008). Also, Pb exposure after 8
weeks increased heart rate and systolic BP. These 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 dose-dependently reduced 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 could change the response to inotropic agents and blunted the
force produced during contraction. Conversely, past studies have found that Pb exposure increases
intracellular Ca2 content (Favalli et al.. 1977; Lai et al.. 1991; 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 were
discussed previously, also examined the association of blood Pb with PAD. They observed an increasing
trend in the prevalence of PAD with increasing 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 a previous NHANES analysis of the association of blood Pb with PAD
conducted by Navas-Acien et al. (2004).
Navas-Acien et al. (2004) reported a significant trend of increasing OR with increasing quartile of
blood Pb or Cd. However, the authors reported a significant association of blood Cd with PAD and a non-
significant association with blood Pb. These authors tested both Pb and Cd in separate models, tested the
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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). When blood Pb and blood Cd were in the same
model together, the ORs were diminished slightly but both had significant trends of increasing OR with
increasing quartile of the metal. Their interaction was not significant. In a later analysis, Navas-Acien et
al. ("2005) used the same 1999-2000 NHANES data set that they used in their 2004 paper, 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. Urine Cd, but not urine
Pb, was reliably elevated in all models in PAD subjects, while there were indications of elevations in
antimony and tungsten. Spot urine Pb measurements are less reliable compared to blood Pb
measurements. In another NHANES analysis Navas-Acien et al. ("2005') reported associations between
PAD and urinary Pb level that were sensitive to adjustment for urinary creatinine.
5.4.3.5. Ischemic Heart Disease
A few studies discussed in the previous Pb AQCD reported association between Pb exposure and
increased risk of cardiovascular outcomes associated with IHD, including left ventricular hypertrophy (J.
Schwartz. 1991) and myocardial infarction (Gustavsson et al.. 2001). Recently, Jain et al. (2007) reported
the incidence of IHD (physician confirmed MI, angina pectoris) among older adult males enrolled in the
VA-NAS 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. Blood,
tibia, and patella Pb levels were log-transformed. Significant hazard ratios were reported for the
association of IHD with blood Pb level and patella Pb level. When blood Pb and patella Pb were included
simultaneously in the model, HR were 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 the 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). This suggests that both contribute independently to IHD. In an ecological study,
Marchwinska-Wyrwal et al. (2010) report simple associations between 15-year air Pb averages in 15
Silesian cities in Poland with similar health and socioeconomic conditions but widely varying air
contamination, including air Pb and air Cd. They report associations between air Pb and CVD by city
when each city was weighted equally.
IHD, characterized by reduced blood supply to the heart, may result from increased thrombosis. A
recent study suggests that Pb exposure promotes a procoagulant state that would contribute to thrombus
formation (Shin et al.. 2007). Pb exposure (i.v. 25 mg/kg, 1 hour) in a rat model of venous thrombosis
resulted in increased thrombus formation. Additionally, increased blood coagulation (5 (.iM) and thrombin
generation (2-5|iM) was observed in a dose-dependent manner after Pb exposure to human erythrocytes.
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This enhanced procoagulant activity in Pb-treated erthyrocytes was the result of increased outer cell
membrane phosphatidylserine (PS) exposure (human RBC: 2-5 (.iM: rat RBC: 5 (iM). Similar to these in
vitro results, PS externalization on erythrocytes was increased in Pb exposed rats (50-100 mg/kg).
Increased PS exposure was likely the result of increased intracellular calcium (5 (iM). enhanced
scramblase activity (5-10 (iM). inhibited flippase activity (5-10 (iM). and ATP depletion (1-5 (.iM) after Pb
exposure (Shin et al.. 2007).
5.4.3.6. Atherosclerosis
Studies provide evidence for increased atherosclerosis and intimal medial thickening (IMT) after
Pb exposure. The association between stroke subtypes and severity of cerebral atherosclerosis was
examined in relation to a single blood Pb level and total 72-hour urine Pb amount (body Pb store - EDTA
mobilization test) in a cross-sectional study of 153 patients (mean age 63.7) receiving digital subtraction
angiography in Chang Gung Memorial Hospital in Taiwan from 2002 to 2005 (T.-H. Lee et al.. 2009).
Body Pb stores were positively associated with the severity of artery stenosis in the intracranial carotid
system but not the extracranial and vertebrobasilar systems. As development of atherosclerosis is a
lifelong process, cumulative body Pb stores may be more sensitive than single blood Pb in the prediction
of atherosclerosis.
Zeller et al. (2010) examined human radial and internal mammary arteries exposed to Pb in culture
and reported a dose-dependent increase in arterial intimal thickness (nonsignificant at 5 |_iM. significant at
50 (.iM. 2 weeks) and intimal extracellular matrix accumulation (50 (iM). Also, Pb exposure promoted
endothelial cell (EC) proliferation (5 and 50 (.iM, 72 hours) and SMC elastin expression (50 (iM, 12
hours), as discussed above (Section 5.4.3.1) (Zeller et al.. 2010). A second study showed that Pb exposure
(100 ppm, 10 months; blood Pb level 28.4 (ig/dL) to rats also increased the aortic media thickness, media-
lumen ratio, and medial collagen content (L.-F. Zhang et al.. 2009). These morphological changes to the
vessel do not only imply initiation of arteriosclerosis. These vascular changes could be to blame for
decreased contractile response of the vessel due to altered visco-dynamic vessel properties or could be an
effect of Pb-induced hypertension.
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Table 5-15. Characteristics and quantitative data for associations of blood and bone Pb
with other CVD measures.
Study
Population/
Location
Parameter Pb Data
Statistical Analysis
Effect Estimate (95%
Cl)a
Ischemic Heart Disease
Jain etal. (2007) 837 men from VA- IHD Blood Pb: Non-cases 6.2 (4.3)
NAS in Greater (Ml or angina |jg/dL; Cases 7.0 (3.8) |jg/dL
Boston, MA area pectoris) Patella Pb: Non-cases 30.6
(1991-2001) (19.7) |jg/dL; Cases 36.8
(20.8) |jg/dL
Tibia Pb: Non-Cases 21.4
(13.6) |jg/g; Cases 24.2 (15.9)
|jg/g
Cases:
Blood Pb range: 1.0 to 20.0
|jg/dL
Patella Pb range: 5.0 to 101
Mg/g
Tibia Pb range: -5 to 75 |jg/g
Cox's 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 |jg/dL
OR= 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
metric
Marchwinska-
Residents from 13 CVD
Air Pb (avg yearly rate):
Linear regression models
13= 1.52 (0.76)
Wyrwal et al.
Silesian cities,
127.8-359.9 ng/m3
(2010)
Poland
Peripheral artery disease
Muntneret al.
9,961 NHANES PAD
Blood Pb:
Logistic regression models
OR (95% CI):
(2005)
(1999-2002)
Q1: <1.06 |jg/dL,
adjusted for age,
1.00 (Reference),
participants
Q2:1.06-1.63 ^ig/dL
race/ethnicity, sex, diabetes
1.00 (0.45,2.22),
Q3:1.63-2.47 ^ig/dL
mellitus, BMI, cigarette
1.21 (0.66,2.23),
Q4: >2.47 ^ig/dL
smoking, alcohol
1.92(1.02,3.61)
consumption, high school
education, health insurance
status
Navas-Acien et al.
(2005)
790 participants, age
> 40 y, from
NHANES (1999-
2000)
PAD
Urine Pb:
Mean (10th-90th %):
(0.2-2.3)
Logistic regression adjusted
0.79 |jg/L 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:OR=1.17 (0.81,
1.69)
Model 2: OR=1.17 (0.78,
1.76)
Model 3: OR=0.89 (0.45,
1.78)
Per IQR increase of urinary
Pb
Array of metals in urine also
evaluated.
Heart rate variability
Park etal. (2006)
413 men from
Normative Aging
Study in Greater
Boston, MA area
(11/14/2000 -
12/22/2004)
HRV Patella Pb:
(SDNN, HF, Median (IQR): 23.0 (15-34)
HFnorm, LF, |jg/g
LF„™, LF/HF) Estimated8: Median (IQR):
16.3 (10.4-25.8) |jg/g
Tibia Pb:
Median (IQR): 19.0(11-28)
|jg/g
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-
blockers, calcium channel
blockers, and/or ACE
inhibitors
Tibia Pb: Model 2
Change (95%CI)
HF:-0.9 (-3.8, 2.1)
LF: 0.9 (-2.0, 3.9)
Log LF/HF: 3.3 (-10.7,19.5)
Per 17 |jg/g tibia Pb
Patella Pb: Model 2 Change
(95%CI)
HF:-0.6 (-3.1, 1.9)
LF: 0.6 (-1.9,3.1)
Log LF/HF: 3.0 (-8.7, 16.2)
Per 15.4 |jg/g patella Pb
Among those with metabolic
syndrome the strongest
effects were observed.
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Study
Population/
Location
Parameter Pb Data
Statistical Analysis
Effect Estimate (95%
Cl)a
Parketal. (2009)
613 men from
QTc" interval Blood Pb:
Linear regression models
Blood Pb:
Normative Aging
Median (IQR): 5 (4-7) ^ig/dL
adjusted for age, BMI,
p:1.3 (-0.76, 3.36)
Study in Greater
smoking status, serum
IQR: 3 |jg/dL
Boston, MA area
Patella Pb:
calcium, and diabetes.
(8/1991 -12/1995)
Median (IQR): 26 (18-37) |jg/g
Patella Pb:
13:2.64(0.13, 5.15)
Tibia Pb:
IQR: 19 |jg/g
Tibia Pb:
p: 2.85 (0.29, 5.40)
IQR: 13 |jg/g
Median (IQR): 19 (14-27) |jg/g
a Estimated patella Pb accounts for declining trend in patella Pb levels between analysis of bone Pb and HRV.
b Heart-rate-corrected QT interval calculated by Bazett's formula
5.4.3.7. Summary
There are a limited number of studies that investigate the association between Pb exposure and
cardiovascular effects other than hypertension (Table 5-15). These studies have presented associations
between various measures of Pb, representing distinct exposure time periods, and atherosclerosis, IHD,
PAD, and HRV. Also, limited, mixed evidence of occupational exposure to Pb and altered cholesterol
have been reported. Additionally, studies in isolated vascular tissues and cells provided mechanistic
evidence to support the biological plausibility of these other vascular effects and cardiotoxicity. A recent
study provided evidence for the interaction between Pb exposure and gene variants for iron metabolism
on the prolonged QT interval (Park. Hu. et al.. 2009). Blood Pb (>2.5 (ig/dL) was associated with greater
risk for PAD in two NHANES analyses (Muntner et al.. 2005; Navas-Acien et al.. 2004). In addition, in
the VA-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 the 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 EC, which was responsible for stimulating SMC invasion into the
vessel intimal layer. Pb exposure also increased extracellular matrix and elastin, primary sites for lipid
deposition in the vessel wall. Overall, the relatively few studies that investigate associations between
biomarkers of Pb exposure and other cardiovascular events provide supportive evidence for the role of Pb
in the development of these diseases, yet further research is needed to understand these relationships.
5.4.4. Mortality Associated with Long-Term Lead Exposure
The previous Pb AQCD (U.S. EPA. 2006) stated that collectively the then available analyses of
NHANES II and III data suggest a significant 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. This involved two
NHANES analyses that examined the association of blood Pb with all cause and cause-specific mortality
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(Lustberg & Silbergeld. 2002; Schober et al.. 2006). As blood Pb levels in adults are comprised of both
recent Pb exposure and Pb mobilization from bone, it is unclear whether the mortality associated with
blood Pb levels are due to current or cumulative Pb exposures. 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 what they are experiencing currently. Using NHANES II data, Lustberg and
Silbergeld ("2002) found significant increases in all-cause, circulatory and cancer mortality, comparing
those with blood Pb levels from 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 associations for all cause mortality,
cardiovascular, and cancer mortality comparing those with blood Pb levels from 5-9 (ig/dL and above 10
(ig/dL to those with blood Pb levels less than 5 (ig/dL.
Three new studies make substantial additions to the evidence base. A further analysis of the
NHANES III database by a different research group using different methods provides further information
addressing uncertainties from other earlier analyses. Two longitudinal prospective studies in different
cohorts conducted by different researchers with different methods in different parts of the U.S. provide
coherence to the evidence base. Menke et al. (2006) examined the associations of all-cause and cause-
specific mortality using NHANES III data. Subjects at least 18 years of age were followed up to 12 years
after they were surveyed and 1,661 deaths were identified. Those with blood Pb levels from 3.63 to 10
(ig/dL had significantly higher risks of all-cause (1.25 [95% CI: 1.04, 1.51]), cardiovascular (1.55 [95%
CI: 1.08, 2.24]), MI (1.89 [95% CI: 1.04, 3.43]), and stroke (2.51 [95% CI: 1.20, 2.26]) mortality
compared to those with blood Pb levels less than 1.93 (ig/dL and non-significantly increased risk of
cancer mortality. Hazard ratios were not higher comparing those with blood Pb levels from 1.94 to 3.62
(ig/dL to those with blood Pb levels <1.93 (ig/dL. However, trends of increasing hazard with increasing
blood Pb tertile were significant (p <0.017) for all models of all CVD presented. Menke et al. (2006)
evaluated several of the model covariates (e.g. diabetes, hypertension, and GFR) in a subgroup analysis
(Figure 5-37). The authors reported that there were no interactions between blood Pb and other adjusted
variables and found a high consistency of HRs across models with a varying number of control variables
(indicating little residual confounding). 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 = 9,686) and
adjusted for covariates including age, sex, ethnicity and smoking rather than the full suite of covariates
evaluated by Menke et al. (2006). Schober et al. (2006) reported significant associations comparing those
with blood Pb levels > 10 j^ig/dL to those with blood Pb levels <5 (ig/dL for all-cause (1.59 [95% CI: 1.28,
1.98]), CVD (1.55 [95% CI: 1.16, 2.07]), and cancer (1.69 [95% CI: 1.14, 2.52]) mortality and generally
non-significant increases comparing those with blood Pb levels from 5-9 (ig/dL to those with blood Pb
levels <5 (ig/dL.
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Subgroup
Hazard ratio of all-cause mortality (95% CI)
Hazard ratio of cardiovascular mortality (95% CI)
Age (years)
<60
1.75(1.25-2.44)
V
ii
CT>
O
1.31 (1.08-1.58)
Race-ethnicity
Non-Hispanic white
1.32(1.09-1.60)
Non-Hispanic black
1.23(0.99-1.52)
Mexican-American
1.17(0.86-1.60)
Sex and menopausal status
Male
1.41 (1.11 -1.78)
Female
1.24(1.00-1.54)
Pre-menopausal
1.02 (0.54- 1.95^
Post-menopausal
1.24(1.00-1.54)
Residence
Rural
1.28(1.05-1.54)
Urban
1.42(1.18-1.72)
Smoking
Never
1.21 (0.93-1.58)
Former
1.61 (1.33-1.94)
Current
1.34 (0.96-1.87)
Body mass index (kg/m2)
<25
1.51 (1.16-1.96)
>= 25
1.28(1.03-1.58)
Hypertension
No
1.31 (1.08-1.58)
Yes
1.32(1.09-1.60)
Diabetes
No
1.37(1.19-1.58)
Yes
1.12(0.73 - 1.71)
Estimated glomerular filtration
rate (ml/mi rV1.73m2)
<60
1.44(1.01 -2.06)
>= 60
1.32(1.12 • 1.56)
Overall
1.34(1.16-1.54)
-4-
¦
2.00 (1.24-3.22)
1.49(1.12-1.99)
1.49 (1.12-1.99)
1.13(0.79-1.61)
1.55 (0.90 - 2.68)
1.35 (0.84 -2.18)
1.63(1.25-2.11)
2.71 (0.93-7.91)
1.46(1.04-2.03)
1.41 (1.01 -1.96)
1.75(1.19-2.56)
1.57 (1.10-2.24)
2.07 (1.49-2.89)
1.05 (0.54 - 2.04)-
2.02(1.32-3.11)
1.34 (0.94-1.91)
1.48 (0.96-2.26)
1.49(1.15-1.94)
1.59(1.31 -1.92)
1.16(0.67-2.00)
1.75 (1.06-2.88)
1.49 (1.18- 1.89)
1.53 (1.21 -1.94)
0.5
0.5
Source: Used with permission from Lippincott Wiliams & Wilkins, Menke et al. (2006)
Note: Hazard ratios were calculated for a 3.4-fold increase in blood Pb level with log-blood Pb as a
continuous variable. This increase corresponds to the difference between the 80th and 20th
percentiles of the blood Pb distribution (4.92 pg/dL versus 1.46 pg/dL, respectively).
Figure 5-37. Multivariate adjusted relative hazards of all-cause and
cardiovascular mortality.
1 Both Menke et al. (2006) and Schober et al. (2006) present mortality curves that plot the hazard
2 ratios against blood Pb level. Figure 5-38 shows the mortality curves for both stroke and MI reported by
3 Menke et al. (2006). which reach a peak around 6-7 ug/dL. The curves were fitted by restricted quadratic
4 splines with knots at the 10th percentile (1.00 (.ig/dL; 0.048 |.iM/L), the 50th percentile (2.67 (.ig/dL; 0.287
5 fiM/L). and the 80th percentile (5.98 ug/dL) blood Pb levels.
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2.0
1.8
1.6
g> 1.0
J 0.8
a;
ce 0.6
0.4
0.2
0.0
Source: Schoberet 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 5-39. Relative risk of all cause mortality for different blood Pb levels
compared with referent level of 1.5 pg/dl_ (12.5th percentile).
In addition to the NHANES analyses described above, two studies of older adult males (Weisskopf
et al.. 2009) and older adult females (KhaliI. 2010: Khalil. Wilson, et al.. 2009) were conducted.
Weisskopf et al. (2009) used data from the VA-NAS to determine the association 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) in study subjects. The strongest associations were observed between mortality
and patella Pb concentration. Non-significant increases in CVD mortality with tibia Pb and no association
between blood Pb and mortality were observed. Tibia bone Pb concentration is thought to reflect a longer
cumulative exposure period than patella bone Pb because the residence time of Pb in trabecular bone is
shorter than in cortical bone. Ischemic heart disease contributed most to the relationship between patella
Pb and all CVD death with HR of 2.69 (95% CI: 1.42, 5.08). Although there was high correlation between
tibia and patella Pb (Pearson r = 0.77), trabecular bone Pb may have more influence on circulating blood
Pb level, and thus organ concentration of Pb, than cortical bone Pb because of its shorter residence time in
bone. In contrast to the NHANES analyses, blood Pb was not significantly related to cardiovascular
mortality in this 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 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,
which were 17 and 40 years, respectively. Further the blood Pb tertile analysis of the Weisskopf study
\
\
\
- - - - '
" /
/
/
1
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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could have been affected if the majority of a hypothesized non-linear effect was contained largely in the
lowest (reference) blood Pb tertile.
Weisskopf et al. (2009) also conducted an exposure-response analysis. The test for linear trend of
HR by bone Pb tertile was significant in both tibia and patella Pb models. The linear relationship using
tertile patella Pb was confirmed in other models using continuous patella Pb and non-linear penalized
spline terms, where higher order components were non-significant. The number of knots and their
placement within the Pb variable, which can influence these results though the number and placement of
knots of the penalized spline, were determined by an iterative best fit procedure to the data.
Concentration-response relationships shown in Figure 5-40 are approximately linear for patella Pb on the
log hazard ratio scale for all CVD, but appear non-linear for ischemic heart disease (p <0.10). Peak HR is
shown around 60 (ig/g, beyond which HR tends to decrease, though the confidence limits are wide.
All-cause
l-J I I
Mill III I
All cardiovascular
lllllllllllll III I I ¦
0 50 100
Patella lead, pg/g
0 50 100
Patella lead, pg/g
Ischemic Heart Disease
minhimin mm11
0 50 100
Patella lead, pg/g
Source: Used with permission from Lippincott Wiliams & 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 absicissa.
Figure 5-40. Nonlinear association between patella bone Pb concentration
and the log of HR (logHR) for all-cause, cardiovascular, and
ischemic heart disease adjusted for age, education, smoking
status, and pack-years of smoking among participants without
ischemic heart disease at baseline.
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The association of Pb with mortality has also been examined among women enrolled in the Study
of Osteoporotic Fractures (SOF) (Khalil. Wilson, et al.. 2009). This prospective cohort (N = 533) enrolled
female volunteers from two locations across the U.S. (Baltimore, MD and Monongahela Valley, PA). All-
cause mortality comparing those with blood Pb levels >8 (ig/dL to those with blood Pb levels <8 (ig/dL
was significantly increased (1.59 [95% CI: 1.02, 2.49]). Combined cardiovascular disease mortality (1.78
[95% CI: 0.92, 3.45]), coronary heart disease mortality (3.08 [95% CI: 1.23, 7.70]), but not stroke
mortality (1.13 [95% CI: 0.34, 3.81]) HR was increased among the women enrolled in this study. In
addition, Khalil et al. (2010) provided both tertile and quintiles analyses, as well as exposure-response
results shown in Figure 5-41.
-
LEAD VALUE
• relative hazard
Median spline
Source: Khalil et al. (2010)
Figure 5-41. Multivariate adjusted relative hazard (left axis) of mortality
associated with blood Pb levels between 1 ygldL and 15 |jg/dL.
5.4.4.1. Summary
The mortality results in this review support and expand upon findings from the previous Pb AQCD
(U.S. EPA. 2006). which included two NHANES mortality studies (Lustberg & Silbergeld. 2002; Schober
et al.. 2006). The recent NHANES mortality study discussed above (Menke et al.. 2006) addressed many
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of the limitations of the these earlier studies, including control for a wider range of potential confounders,
testing for interactions with Pb, consideration of exposure-response relationships, extensive model
evaluations, and including sub-categories of CVD. Further, an association with increased mortality was
observed at lower population blood Pb concentrations. The mean blood Pb level of the NHANES III
population was 2.58 (ig/dL. The Pb attributable risk of increased cardiovascular mortality in the
NHANES III analysis of Menke et al. (2006) reached its maximum at blood Pb levels between of 6 and 7
(ig/dL. In addition, the first evidence that bone Pb, a metric of cumulative Pb exposure, is associated with
increased mortality was reported. Several studies report associations between the accumulated Pb in bone
and higher CVD morbidity, which are consistent with the mortality findings.
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 exposure measure, assumptions of linearity of the Pb exposure term, and choice of
exposure metric. Quantitative differences in Pb-associated mortality across NHANES II and NHANES III
studies or between different NHANES III may be explained by the use of continuous or ordered blood Pb
exposure variables 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
national effects of Pb on cardiovascular disease-related mortality. The study uses the nationally
representative NHANES III (1988 - 1994) sample allowing results to be generalized to the segment of the
US population included in the sample. The results provide confirmation of earlier published studies and
address some of the key weaknesses noted in those studies. 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 with
trabecular bone Pb. The Khalil et al. (2010; 2009) Study of Osteoporotic Fractures provides supporting
results in a different study cohort consisting of white females aged 65-87 years. Further, a number of prior
studies have already found association between accumulated Pb reflected in bone Pb measurements and
higher CVD morbidity (see CVD morbidity section), to which is added new findings of increased
mortality due to CVD from long-term Pb exposure (bone Pb). Despite the differences in design and
methods across studies, effects on mortality were consistently observed (Figure 5-42 and Table 5-16). In
studies that broke out CVD-related mortality into sub-categories, MI, stroke, and IHD mortality, death-
causes related to higher BP and hypertension were all significantly elevated as Pb exposure increased.
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Reference
Outcome
Study
N
Blood Pb
Comparison
i
Menke et al. (2006)
All Cause
NHANES III
13,946
; >3.63
Q3 v Q1
|-#-
1988-1994
1.94-3.62
Q2 v Q1
-f
Mean age = 44
<1.93
Q1 Reference
Mean blood Pb = 2.58
CVD
>3.63
Q3 v Q1
S—•—
1.94-3.62
Q2 v Q1
—
<1.93
Q1 Reference
Ml
>3.63
Q3 v Q1
! •
1.94-3.62
Q2 v Q1
—f
<1.93
Q1 Reference
Stroke
>3.63
Q3 v Q1
! •
1.94-3.62
Q2 v Q1
<4 •
<1.93
Q1 Reference
1
Schober et al. (2006)
All Cause
9757
>10
Q3 v Q1
i #
5-9
Q2 v Q1
age > 40 yr
<5
Q1 Reference
i
CVD
>10
Q3 v Q1
! #
5-9
Q2 v Q1
*•—
<5
Q1 Reference
Lustbert & Silbergeld (2002)
All Cause
NHANES II
4190
20-29
Q3 v Q1
1976-1980
10-19
Q2 v Q1
Mean age = 54 yr
< 10
Q1 Reference
Mean blood Pb = 14.0
CVD
20-29
Q3 v Q1
10-19
Q2 v Q1 Ig
< 10
Q1 Reference
Khalil et al. (2009)
All Cause
SOF
533
>8
Dichotomous
—~
women
CVD
1986-1988
>8
Dichotomous
—#
Mean age = 70 yr
CHD
>8
Dichotomous
—m
Mean blood Pb = 5.3
Stroke
>8
Dichotomous
9
<8
Q1 Reference
Weisskopf et al. (2009)
All Cause
VA-NAS
868
>6
Q3vQ1 a.
1991 -1999
4-6
Q2vQ1 a
Men
<4
Q1 Reference
Mean age = 67.3 yr
CVD
>6
Q3 v Q1 g ¦
Mean blood Pb = 5.7
4-6
Q2 v Q1
<4
Q1 Reference
IHD
>6
Q3 v Q1
k
4-6
Q2 v Q1
<4
Q1 Reference
r
Tibia Pb (ng/g)
All Cause
Mean = 30.6
Q3 v Q1
A
CVD
Mean = 30.6
Q3 v Q1
V
IHD
Mean = 30.6
Q3 v Q1
Mean = 13.9
Q1 Reference
u
Patella Pb (pg/g)
All Cause
Mean = 49.6
Q3 v Q1
Mean = 28.8
Q2 v Q1
Mean = 14.5
Q1 Reference
y
CVD
Mean = 49.6
Q3 v Q1
Mean = 28.8
Q2 v Q1
O
Mean = 14.5
Q1 Reference
IHD
Mean = 49.6
Q3 v Q1
Mean = 28.8
Q2 v Q1
~
Mean = 14.5
Q1 Reference
o
-1 1 3 5 7 9
Hazard Ratio (95% CI)
Figure 5-42. Hazard ratios between blood Pb (closed markers), bone Pb
(open markers), all-cause mortality (diamonds), and
cardiovascular mortality (circles).
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Table 5-16. Additional characteristics and quantitative data for associations of blood and
bone Pb with CVD mortality for results presented in Figure 5-42.
Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
(95% CI)
Menke et al.
(2006)
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
I00-I99), Ml (ICD-9 410-414
and 429.2; ICD-10 I20-I25),
stroke (ICD-9 430-434 and
436-438; ICD-10 I60-I69).
Blood Pb:
Mean: 2.58 |jg/dL
Tertiles:
<1.93 |jg/dL,
1.94-3.62 |jg/dL,
> 3.63 |jg/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)
Schoberetal.
2006 (2006)
9686 adult
participant of
NHANES III,
> 40 yr
All cause and cause-specific
mortality
Ordered
categorical blood
Pb level:
<5 |jg/dL
5-9 |jg/dL
>10 |jg/dL
Also ln(blood Pb
level)
Survey-design adjusted Cox
proportional hazard adjusted for
sex, age, race/ethnicity, smoking,
education level, median follow-up
time = 8.55 y
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)
Weisskopf etal.
(2009)
868 men, >55 yr,
95% white, from
Normative Aging
Study in Greater
Boston area, MA
All cause and cause-specific
mortality
Blood Pb:
Mean (SD): 5.6
(3.4) |jg/dL
Patella Pb:
Mean (SD):31.2
(19.4) pg/g
Tertiles:
<22 |jg/g,
22-35 |jg/g,
>35 |jg/g
Tibia Pb:
Mean (SD):21.8
(13.6) Mg/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 etal.
(2009)
533 women, 65- All cause and cause-specific Blood Pb:
87 y, from Study mortality
of Osteoporotic
Fractures cohort
in Baltimore, MD
and Monongahela
Valley, PA
Mean (SD;
range): 5.3 (2.3;
1-21) |jg/dL
Cox proportional hazards
regression analysis adjusted for
age, clinic, BMI, education,
smoking, alcohol intake, estrogen
use, hypertension, total hip BMD,
walking for exercise, and diabetes,
, mean follow-up time = 12 ± 3 y.
All cause (> 8 |jg/dL vs.
<8 |jg/dL):
1.59(1.02,2.49)
CVD: 1.78(0.92,3.45)
Coronary Heart
Disease:
3.08(1.23, 7.70)
Stroke: 1.13(0.34,3.81)
Cancer: 1.64 (0.73,
3.71)
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Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
(95% CI)
Neubergeretal.
(2009)
Residents at or
near Tar Creek
Superfund site,
Ottawa County,
OK (exposed pop.
5,852, unexposed
pop. 16,210)
Cause-specific mortality
Not reported
Standardized mortality ratio based
on 2000 US 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)
Cocco et al.
(2007)
933 male Pb
smelter workers
from Sardinia,
Italy (1973-2003)
All cause and cause-specific
mortality
Not reported
Standardized mortality ratio
All cause: 56 (46, 68)
CVD: 37 (25, 55)
References not included in Figure 5-42 are included in this table.
5.4.5. Air Lead-PM Studies
5.4.5.1. Hospital Admissions
In addition to blood Pb, some recent epidemiologic studies have used Pb measured in PM10 and
PM2 5 air samples to represent Pb exposures. Some studies have analyzed Pb individually, whereas others
have applied source apportionment techniques to analyze Pb as part of a group of correlated components.
A common limitation of air-Pb studies is the variable size distribution of Pb-bearing PM (Section 3.5.3)
and its relationship with blood Pb levels. Relative to studies of Pb biomarkers, time-series studies provide
weak evidence for association between PM2 5-Pb concentrations and cardiovascular hospital admissions
and mortality in older adults. In a time-series study of 106 U.S. counties, Bell et al. (2009) found that an
increase in lag 0 PM25-Pb was associated with an increased risk of cardiovascular hospital admissions.
Quantitative results were not presented; however, the 95% CI was wide and included the null value. Ostro
et al. (2007) found that a 5 ng/m3 (interquartile range) increase in lag 3 PM2 5-Pb was associated with a
1.89% increase (95% CI: -0.57, 4.40%) in cardiovascular mortality in six California counties in the cool
season. In both of these studies, statistically significant positive associations were observed for other PM
components such as nickel, vanadium, and zinc also. In the absence of detailed data on correlations
among components or results adjusted for copollutants, it is difficult to exclude confounding by these
other components.
To address correlations among PM chemical components, some studies have applied source
apportionment techniques to group components into common source categories. In these source-factor
studies, it is difficult to attribute findings to a particular component in a group. In a study of 20 counties
near Atlanta, GA, Sarnat et al. (2008) found that PM2 5-Pb mass was explained by a "woodsmoke" factor,
which was associated with a 2.4% (95% CI: 1.5, 3.3%) increased risk of cardiovascular hospital
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admission (lag 0, per interquartile range increase). Less than 10% of variation in PM2 5-Pb mass was
explained by "woodsmoke," thus the association may not be attributable specifically to Pb. Andersen et
al. (2007) found that in Copenhagen, Denmark, variation in PMi0-Pb was explained by a "vehicle" factor
that also included copper and iron. A 0.6 (ig/m3 increase in the 3-day lagged sum of "vehicle" factor
pollutants was not associated with increased risk of cardiovascular hospital admissions among adults age
65 years and older (RR: 0.999, [95% CI: 0.993, 1.004]).
5.4.5.2. Mortality
Time-series studies of PM25-Pb have reported positive associations with all-cause mortality. 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-9th percentile) increase in PM25-Pb. In six California counties,
Ostro et al. (2007) found that a 5 ng/m3 (interquartile range) increase in PM2 5-Pb was associated with a
1.74% (95% CI: 0.24, 3.26%) increased risk in all-cause mortality during the cool season. The limitations
of air-Pb studies have been described previously (Section 5.5.5.1), and are relevant to the interpretation of
these findings for all-cause mortality.
5.4.6. Summary and Causal Determination
The 2006 Pb AQCD concluded that there was a relationship between increased blood Pb and bone
Pb and increased adverse cardiovascular outcomes in adults, including increased BP and increased
incidence of hypertension (U.S. EPA. 2006). This was substantiated by the coherence of effects observed
across epidemiologic and toxicological studies. The large evidence base of epidemiologic studies
conducted by many researchers in many locations using different designs found a clear positive
association between blood Pb level and BP. Meta-analysis of these studies found that each doubling of
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 (Nawrot et al.. 2002). In addition, most of the reviewed studies using
cumulative Pb exposure measured by bone Pb levels also showed increased BP. Similarly, toxicological
studies provided evidence for exposure to low levels of Pb resulting 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
oxidative stress, altered sympathetic activity, and vasomodulator imbalance. Studies in the 2006 Pb
AQCD also suggested a connection between Pb exposure and other cardiovascular diseases such as
ischemic heart disease, cerebrovascular disease, peripheral vascular disease, and cardiovascular disease
related mortality, however this evidence was limited.
Building on the strong body of evidence presented in the 2006 Pb AQCD, recent studies continue
to support associations between long-term Pb exposure and cardiovascular effects with recent
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epidemiologic studies informing past uncertainties (e.g. confounding, low Pb exposures). A recent study
in an ethnically diverse community-based cohort of women and men aged 50-70 years of age suggests
that Pb has an acute effect on BP as a function of recent dose measured by blood Pb and a chronic effect
on hypertension risk as a function of cumulative exposure measured by tibia Pb (Martin et al.. 2006). This
study verified other studies by demonstrating that with each increase of 1 (ig/dL blood Pb level, systolic
BP would increase 1 mmHg and diastolic BP would increase 0.5 mmHg. Additionally, recent
epidemiologic studies provided evidence for associations between blood Pb and hypertension at relatively
low blood Pb levels; a positive relationship was found in the NHANES data set at a geometric mean
blood Pb level of 1.64 (ig/dL (Muntner et al.. 2005). Animal toxicological studies also provide support for
effects of low blood Pb level on increased BP with statistically significant increases shown as low as 2
(ig/dL (Tsao et al.. 2000). Collectively, all animal toxicological studies providing blood Pb level and BP
measurements report positive increases in BP with increasing blood Pb level (Figure 5-34). New studies
also demonstrate reversibility of Pb-induced increased BP following Pb exposure cessation or chelation.
Epidemiologic studies continue to investigate the relationship between bone Pb, representing
cumulative Pb exposure, and increased BP. Navas-Acien et al. (2008) published a meta-analysis of
epidemiological studies examining this association. Studies passing the detailed inclusion criteria all
showed positive relationships between bone Pb measures and BP and all but one that characterized
hypertension showed positive risk or odds ratios associated with bone Pb. Recent epidemiological studies
have also emphasized the interaction between long-term Pb exposure and factors that moderate or modify
the Pb effect, like chronic stress and metabolic syndrome, on BP and hypertension. Bone Pb coupled with
high stress was associated with a strong and reliable increased risk of developing hypertension in an
originally nonhypertensive group (Peters et al. 2007). Also, long duration Pb exposure interacted with
components of the metabolic syndrome to drive HRV in directions associated with increased
cardiovascular events (Park et al.. 2006).
Recent epidemiologic studies investigated the interaction of genotypes with effects of Pb on the
cardiovascular system. Significant evidence was presented for modification of the effect of blood Pb level
on BP by ALAD genotype (Scinicariello et al.. 2010). Additionally, polymorphisms in the
hemochromatosis gene modified the pulse pressure response to bone Pb exposure, where pulse pressure
represents as a good predictor of cardiovascular morbidity and mortality and an indicator of arterial
stiffness (A. Zhang et al.. 2010). Park et al. (2009) provided further evidence of gene variants, specifically
those related to iron metabolism, impacting the effect of long-term Pb exposure on the cardiovascular
system, evaluated by QT interval changes.
Not only has Pb exposure been shown to increase BP and hypertension, but Pb exposure can
contribute to the development of other cardiovascular diseases. Recent epidemiologic and toxicological
studies provide evidence for increased atherosclerosis, thrombosis, ischemic heart disease, peripheral
artery disease, arrhythmia, and cardiac contractility (blood Pb levels >2.5 (ig/dL).
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Animal toxicological evidence continues to build on the evidence supporting the mechanisms
leading to these cardiovascular alterations. 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 consistently positive
associations observed between blood and bone Pb and cardiovascular effects.
New evidence extends the potential continuum of Pb-related cardiovascular effects by
demonstrating associations between Pb exposure and both cardiovascular and all-cause mortality. All-
cause mortality was positively associated with increased blood Pb level. The recent analysis of the
nationally representative NHANES III (1988-1994) sample reported positive associations with
cardiovascular mortality, with stronger associations with myocardial infarction and stroke mortality
(Menke et al.. 2006). These findings were supported by a community-based cohort of women age 65-87
years, in which effect estimates were increased for mortality from cardiovascular disease and coronary
heart disease (Khalil. Wilson, et al.. 2009). Weisskopf et al. (2009) published the first mortality study
using bone Pb as an exposure index. This prospective study found that trabecular bone Pb levels were
associated with increased mortality from cardiovascular disease and ischemic heart disease with hazard
ratios of 5.6 and 8.4, respectively.
In summary, new studies evaluated in the current review support or expand upon the strong body of
evidence presented in the 2006 AQCD that Pb exposure is causally associated with cardiovascular health
effects. Both epidemiologic and toxicological studies continue to demonstrate a consistently positive
relationship between Pb exposure and increased BP or hypertension development in adults and this
relationship is observed at adult blood Pb levels (mean 2 (ig/dL) lower than that reported in the 2006
AQCD. While some studies evaluate exposure-response relationships, the information is inconclusive.
Recent studies investigate cumulative Pb exposure measures and suggest that bone Pb related strongly to
hypertension risk. Evidence of Pb increasing the risk of developing other cardiovascular diseases has also
been shown. 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 mode of action of Pb and provided biological plausibility for
the consistently positive associations observed in epidemiologic studies between blood and bone Pb and
cardiovascular effects. These observed associations between Pb exposure and cardiovascular morbidity
are supported by recent reports of increased cardiovascular mortality. Collectively, the evidence integrated
across epidemiologic and toxicological studies as well as across the spectrum of cardiovascular health
endpoints is sufficient to conclude that there is a causal relationship between Pb exposures and
cardiovascular health effects.
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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. After chronic Pb exposure, pathological changes in the renal
system include 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.
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 interrelated
nature of 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 blood Pb and bone Pb (particularly in the tibia) with health
outcomes in adults indicate acute effects of recent dose and chronic effects of cumulative dose,
respectively. In some physiological circumstances of increased bone remodeling or loss (e.g., osteoporosis
and pregnancy), Pb from bone of adults may also contribute substantially to blood Pb concentrations.
Blood Pb in children, although highly affected by recent dose, is also influenced by Pb stored in bone due
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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.
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 epidemiology studies and clinical settings. Creatinine is the most commonly used
endogenous compound; blood urea nitrogen (BUN) has also been used. 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, which 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 (C'ockcroft & 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 etal.. 1999; Levey
et al.. 2000). which estimates GFR, has become the standard in the kidney epidemiology and clinical
communities. With widespread use of the MDRD equation, it became clear that it 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 and the inability to adjust for
muscle mass has led to evaluation of alternative serum biomarkers such as cystatin C, a cysteine protease
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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 with serum creatinine (Fried.
2009). Despite this, recent research indicates that serum cystatin C varies by age, sex, and race (Kottgen
et al.. 2008) and a cystatin C-based eGFR equation that includes age, sex, and race was recently
developed (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 et al..
1993). Elevated levels may indicate an increased risk for subsequent kidney dysfunction. However, most
of these markers are research tools only, and their prognostic value remains uncertain since prospective
studies of most of these markers in nephrotoxicant-exposed populations are quite limited to date.
Recently, microalbuminuria has been identified as a PT marker, not just glomerular as previously thought
(Comper & 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 (Devaraian. 2007; M. A. Ferguson et al.. 2008).
5.5.2. Nephrotoxicity and Renal Pathology Related to Lead
Effects
5.5.2.1. Toxicology
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 0.01% [low dose] or 0.5% [high 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. Gonick. Cohen. Bergamaschi. et al.. 1992;
Khalil-Manesh. Gonick. & Cohen. 1993; Khalil-Manesh. Gonick. Cohen. Alinovi. et al.. 1992). At 6
months of exposure, GFR decreased, albuminuria was present, and pathology ensued with focal tubular
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atrophy and interstitial fibrosis formation. This pathology remained consistent out to 12 months, and at 12
months glomeruli developed focal and segmental sclerosis. Similarly, GFR remained decreased after 12
months of exposure. The evidence provided by toxicological studies that showed a difference in GFR
with acute Pb exposure (hyperfiltration) versus chronic exposure (decreased GFR) provided biological
plausibility to epidemiological studies that observed a similar phenomenon by age in humans with Pb
exposure. These dichotomous changes in GFR (acute versus chronic Pb exposure) are consistent between
the toxicological and epidemiology 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
biomarker since renal ALAD is protected by the kidney antioxidant GSH. Urinary NAG and GST levels
increase after 3 months of high dose Pb exposure (blood Pb level 125 (ig/dL) (Dehpour et al.. 1999;
Khalil-Manesh. Gonick. Cohen. Alinovi. et al.. 1992). whereas only urinary NAG was increased
following low dose Pb exposure (blood Pb level 29 (.ig/dL) (Khalil-Manesh. Gonick. & Cohen. 1993).
Occupational studies found that urinary NAG correlated best with recent blood Pb changes.
The adverse effects of chronic Pb exposure as detailed above are partially rescued with chelation
therapy such as DMSA (Khalil-Manesh. Gonick. Cohen. Bergamaschi. et al.. 1992). 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 further
detail in the antioxidant section (Section 5.5.5), produced similar findings. There was a function 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 confirmed the previously observed increase in serum creatinine following
chronic Pb exposure in rats. Annabi 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-dependent renal dysfunction are listed in Table 5-17
below. Other investigators have also shown that chronic Pb exposure has adverse 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).
Qiao et al. (2006) measured the effect of Pb on the expression of the renal fibrosis-related nuclear
factor-kappa B (NFkB), 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 NFkB increased
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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, followed by
remnant kidney surgery (left kidney mass reduced by 2/3 and right kidney removed) and then
continuation of Pb exposure for an additional 12 weeks. Pb-treated rats had higher systolic BP, lower
creatinine clearance, and higher proteinuria than controls. Most striking was development of worse
arteriolar disease, peritubular capillary loss, tubulointerstitial damage, and macrophage infiltration. Pb
treatment was associated with significant worsening of pre-glomerular vascular disease, as characterized
by an increase in the media-to-lumen ratio. There was also a higher percentage of segmental sclerosis
within glomeruli and a tendency for a higher number of sclerotic glomeruli. Additionally, a loss of
peritubular capillaries, as reflected by a reduction in thrombomodulin staining, was observed. This was
associated with worse tubular injury (osteopontin staining) due to more interstitial fibrosis (type III
collagen staining) and a greater macrophage infiltration in the interstitium. The increase in macrophages
was associated with higher renal MCP-1 mRNA. Low level 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 et al. (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 blood Pb level may exacerbate pre-existing underlying
kidney disease.
Table 5-17. Indicators of renal damage in male rats exposed to 50 ppm Pb for 40 and 65
days, starting at parturition
Biomarker
PND40
Control
PND40
Pb
PND65
Control
PND65
Pb
Blood Pb level
(pg/dL)
1.8
12.7
2.1
7.5
Plasma Creatinine
(mg/L)
4.5+0.21
5.35+0.25a
4.55+0.27
6.04+0.29a
Plasma Urea
(mg/L)
0.37+0.019
0.47+0.021a
0.29+0.009
0.29+0.009
Plasma Uric Acid
(mg/L)
7.51+0.44
7.65+0.32
9.39+0.82
5.91+0.53a
"p <0.001
Source: Used with permission from John Wiley & Sons, Berrahal et al. (2011)
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Histological Changes
Historical studies discussed in previous Pb AQCDs have identified Pb-dependent renal damage by
the presence of dense intranuclear inclusion bodies, which are capable of sequestering Pb (Gover. May, et
al.. 1970). Chelators like CaNa2EDTA removed these inclusion bodies from affected nuclei (Gover et al..
1978). 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 seen in the kidney with
acute 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
(Fow ler et al.. 1980; Gover. Leonard, et al.. 1970). Symptoms similar to the PT transport-associated
Fanconi syndrome appear with Pb exposure, albeit often at high doses of Pb. These symptoms, which
include increased urinary electrolyte excretion (zinc), decreased Na-K-ATPase activity, mitochondrial
aberrations, and aminoaciduria, have also been reported in Pb exposed 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 in histological preservation. 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. (1992; 1992); in particular, the enlarged
glomeruli are consistent with the early hyperfiltration caused by Pb.
A recent study has also reported inclusion body formation in the nuclei, cytoplasm, and
mitochondria of PT cells of Pb exposed rats (50 mg Pb/kg bw i.p., every 48 h 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 junctions.
Vogetseder et al. (2008) examined the proliferative capacity of the renal PT (particularly the S3
segment) following i.v. administration of Pb to juvenile and adult male Wistar rats. Proliferation induction
was examined by detection of Bromo-2'-deoxyuridine (BrdU), Ki-67 (labels S, G2, and M phase cells),
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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 caused by Pb administration may be a result of
reduced cell cycle inhibition by p27kip l. Acute Pb treatment increased the incidence of cyclin D1 in the
BrdU-positive cells suggesting Pb was able to accelerate reentry into the cell cycle and cause proliferation
in the PT. Pb-dependent proliferation has also been reported in the retina with gestational 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 of Pb-mediating nephrotoxic
manifestations. Drug-induced PT phospholipidosis is seen clinically with use of the potentially
nephrotoxic aminoglycoside drugs, including gentamicin (Baronas et al.. 2007).
Various antioxidants have been shown to attenuate histopathological changes to the
kidney. Ozsoy et al. (In Press ) 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 of Pb and extract produced histology indiscernible from control. Post treatment
with extract partially rescued the Pb induced histopathology. El-Neweshy and El-Sayed
studied the influence of vitamin C supplementation (20 mg/kg pretreatment every other day)
on Pb-induced histopathological alterations in 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 seen. Rats treated with Pb acetate plus vitamin C exhibited
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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 antioxidants are capable of preventing or rescuing Pb-dependent 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 will 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 affected by Pb exposure.
Antioxidants attenuated Pb-dependent oxidative/nitrosative stress in the kidney and abrogated the Pb-
induced increased BP (Vaziri. Ding, et al.. 1999). Administration of the vasoconstrictor endothelin-1 (ET-
1) affected mean arterial pressure (MAP) and decreased GFR (Novak & Banks. 1995). Acute high dose
Pb exposure completely blocked the ET-1-dependent 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 L-NAME in both Pb-treated and control
animals. Pb exposure significantly increased BP (134 ± 3 versus 100 ± 6 mm Hg), 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 conclude that Pb exposure increases vascular tone through nitric
oxide-dependent and independent mechanisms, increasing renal vascular sensitivity to vasoconstrictors.
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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 adverse outcomes. All of these
outcomes, in conjunction with Pb-dependent depletion of antioxidants (e.g. GSH) and elevation of lipid
peroxidation point to possible susceptibility of the kidney to apoptosis or necrosis. Literature in this area
is emerging.
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, NFkB activation and
overexpression of tubulointerstitial angiotensin II.
Navarro-Moreno et al. (2009) examined the effect of 500 ppm Pb in drinking water over 7 months
on the structure (including intercellular junctions), function, and biochemical properties of PT cells of
Wistar rats. Pb effects in epithelial cells consisted of an early loss of the apical microvillae, 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
speculate that Pb may compete with calcium in the establishment and maintenance of intercellular
junctions.
Tubular necrosis was also observed in rats exposed to 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
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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 conclude 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-Neweshv & El-Saved).
Wang et al. (2009) 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, were seen 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 were seen during Pb exposure.
Pb-dependent 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 5-18 presents the acute and chronic renal effects of Pb exposure observed in recent and past
animal toxicology studies.
Table 5-18. Acute and chronic effects of Pb on the kidney/renal system - evidence from
animal toxicology studies.
Acute
Chronic
Mitochondrial dysfunction
Renal cell apoptosis
Nuclear Inclusion Body Formation
3roximal Tubule Cytomegaly
Glomerular Hypertrophy
ncreased GFR
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 significant advances in research on the impact of Pb on the kidney in the 20 years
following the 1986 Pb AQCD (U.S. EPA. 1986) were noted in the 2006 Pb AQCD (U.S. EPA. 2006).
These included research in general and CKD patient populations at much lower levels (5-10 (ig/dL) of Pb
exposure than previously studied. Furthermore, chelation therapy in CKD patients was evaluated, also at
levels of exposure not previously thought to be amenable to chelation. Through these lines of research, it
became clear that chronic Pb nephropathy, characterized by tubulointerstitial nephritis due to chronic,
high-level Pb exposure (Bonucchi et al.. 2007). is a small portion of the contribution that Pb makes to
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kidney dysfunction overall in the population. Pb, at much lower doses than those causing chronic Pb
nephropathy, acts as a cofactor with other more established kidney risks to increase the risk for CKD and
disease progression in susceptible patients. The populations at risk for kidney dysfunction (diabetics and
hypertensives) are increasing worldwide, particularly in countries where obesity is epidemic. Pb exposure
continues to decline in many industrialized countries, although less so among minority populations,
which, notably, are also higher risk groups for CKD. Thus, the extent of the public health impact of Pb on
the kidney depends on the balance of these two factors.
In the 2006 Pb AQCD (U.S. EPA. 2006). several key issues could not be completely resolved based
on the Pb-kidney literature published to date. These included the lowest Pb dose at which adverse kidney
effects occur, whether associations at current Pb levels are due to higher past exposures, 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 for several of these challenges have been published.
General Population Studies
The 2006 Pb AQCD reported studies that examined the effect of Pb exposure on 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. 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. 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 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.
Four studies assessing the kidney impact of Pb exposure have been published in the Normative
Aging Study (NAS) population (R. Kim et al.. 1996; Pavton et al.. 1994; Tsaih et al.. 2004; M. T. Wu et
al.. 2003). Participants in this study were originally recruited in the 1960s in the Greater Boston area.
Inclusion criteria included male gender, age 21 to 80 years, and absence of chronic medical conditions.
Longitudinal data contained in two 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 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 (R. Kim et al.. 1996).
Participants were randomly selected to be representative of the entire NAS population in terms of age and
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follow-up. Kidney status was assessed with serum creatinine. Data from four evaluations were available
for the majority of participants. Mean (SD) age, blood Pb level, and serum creatinine, at baseline, were
56.9 (8.3) years, 9.9 (6.1) (ig/dL, and 1.2 (0.2) mg/dL, respectively. In the longitudinal analysis, using
random-effects modeling, ln-transformed blood Pb was associated with change in serum creatinine over
the subsequent follow-up period in the 428 participants whose highest blood Pb level was < 25 (ig/dL ((3=
0.027 [95% CI: 0.0, 0.054]); associations in the entire group and subsets with different peak blood Pb
levels (< 10 or 40 (ig/dL) 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 significant positive association between blood Pb and concurrent serum
creatinine remained. 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, 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 Normative Aging Study 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. Pb dose 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. For In blood Pb, the positive association with serum creatinine
was substantially stronger in diabetics ((3 = 0.076 [95% CI: 0.031, 0.121]) compared to non-diabetics ([3 =
0.006 [95% CI: -0.004, 0.016]). A similar relationship was observed for tibia Pb. An interaction was also
observed between tibia Pb and hypertension, although it is possible that many of the 26 diabetics were
also included in the hypertensive group and were influential there as well. Reverse causality was
addressed in a sensitivity analysis of participants whose serum creatinine was <1.5 mg/dL; the authors
reported that longitudinal associations did not materially change. These studies are depicted in either
Figure 5-43 and Table 5-19.
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Study/Age (yr) Pb Level3
CROSS-SECTIONAL RESULTS
Serum Creatinine = Scr, mg/L
Kim et al. (1996) 8.4 (5.7, 12.4)
Avgage (SD) 58.9 (12.4)
Tsiah et al. (2004)
Baseline blood Pb
Avg age (SD) 66 (6.6)
Follow-up Blood Pb
Avg age (SD): 72 (6.5)
5.5(3.7, 8.1)
3.9 (2.8, 5.6)
Creatinine Clearance, mL/min divided by 100
Akesson et al. (2005) 2.2 (1.7, 3.0)
Median age (95th percentile) 58 (54,63)
Staessen et al. (1992) 7.5 (5.2, 10.9)
Avg age (SD) 48 (16)
IogeCreatinine Clearance, mL/min
Payton et al. (1994) 7.3 (5.4, 9.9)
Avg age (SD) 64 (7.4)
Glomerular filtration rate, mL/min divided by 100
Akesson et al. (2005) 2.2 (1.7, 3.0)
GFR mL/min/1.73 m2
Fadrowski et al. (2010) 1.4 (0.7, 2.9)
Age range (12 to 20)
Navas-Acien et al. (2009) geomean 1.58
Age range >20
LONGITUDINAL RESULTS
Change in Scr, mg/L per yrs of follow-up
Tsiah et al. (2004)
Baseline Blood Pb 3.9 (2.8, 5.6)
Baseline Tibia Pb
18.2 (12.3, 26.9)
Change in Scr, mg/L-adj. for time between visits
Kim et al. (1996) 8.4 (5.7, 12.4)
Change in eGFR, mL/min over 4 yrs divided by 10
Yu et al. (2004) 3.2(2.5,4.1)
Age range (25 to 82)
Populations
Blood Pb <40 ng/dL from NAS
Blood Pb <25 |jg/dL from NAS
Blood Pb <10 (jg/dL from NAS
Diabetic from NAS
Nondiabetic from NAS
Hypertensive from NAS
Nonhypertensivefrom NAS
Diabetic from NAS
Nondiabetic from NAS
Hypertensive from NAS
Nonhypertensivefrom NAS
Swedish women
Belgian women
Men in NAS
Swedish women
Adolescents from NHANES III
NHANES 1999-2006
Diabetic from NAS
Nondiabetic from NAS
Hypertensive from NAS
Nonhypertensivefrom NAS
Diabetic from NAS
Nondiabetic from NAS
Hypertensive from NAS
Nonhypertensivefrom NAS
Blood Pb <40 ng/dL from NAS
Blood Pb <25 |jg/dL from NAS
Blood Pb <10 |jg/dL from NAS
CKD patients
-0.6 -0.5 -0.4 -0.3 -0.2 -0.1 0.0 0.1 0.2 0.3 0.4 0.5 0.6
Change in kidney metric per pg/dL blood Pb, at 1 |jg/dl_
aPb level presented as median blood Pb level and (IQR) in |jg/dL or bone
Pb level (IQR) in |jg/g unless otherwise noted.
Note: The kidney function/blood Pb curves fit by a log-linear model are
depicted by their slopes at a blood Pb level of 1 pg/dL. Comparisons of the
magnitude of the effect should not be made between effects having
different kidney metrics. For uniform presentation, blood Pb level
distributional statistics were converted to median and IQR by assuming
that blood Pb is normally distributed. The white shaded areas include
kidney function tests where an increase is considered impaired function.
The gray shaded areas include kidney function tests where a decrease is
considered impaired function.
Figure 5-43. Kidney metric slopes on blood Pb or bone Pb.
2 NHANES data analyses benefit from a number of strengths including large sample size, ability to
3 adjust for numerous Pb risk factors, and the fact that the study population is representative of the U.S.
4 non-institutionalized, civilian population. As a result, the impact of Pb on the kidney has been examined
5 in multiple NHANES datasets obtained over the last few decades (Figure 5-44 and Table 5-19). The
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results of these publications, covering different time frames, have been consistent in providing support for
Pb as a CKD risk factor, including NHANES III, conducted from 1988-1994, where hypertensives and
diabetics were observed to be susceptible populations (Muntner et al.. 2003) and NHANES 1999-2002
(Muntner et al.. 2005V
A recent publication examined this relationship in NHANES data collected from 1999 through
2006 (Navas-Acien et al.. 2009). The geometric mean 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 to the lowest blood Pb quartiles. Thus, in
the subset of the population with the most severe kidney disease (both reduced eGFR and albuminuria),
the risk from Pb was greater. Cadmium was included as a covariate and 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 below).
An important contribution of all three 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 Pb levels during the time periods covered.
Additional studies in this category have also reported worse kidney function related to Pb dose (Gosw ami
et al.. 2005; Hernandez-Serrato et al.. 2006; L. H. Lai et al.. 2008).
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Study
Quartiles of Blood Pb Distribution Used
Muntner et al. (2003)
Elevated Serum Creatinine
Hypertensive
Normotensive
Navas-Acien et al. (2009)
Albuminuria >30 mg/g creatinine
eGFR <60 mL/min/1,73m
4
a*
i
•
i
¦#—
n
ji
a
|jg/dL Blood Pb
-40 -10 10 50
% Change per |jg/dL Blood Pb
Note: 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). The blood Pb distribution of the reference
group is shaded gray and the other group is shaded black. These articles
reported ORs of kidney function measures by grouping by quartiles of
blood Pb and then comparing each group to the quartile with the lowest
blood Pb (reference group).
Figure 5-44. Percent change for kidney outcomes associated with blood
Pb.
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Table 5-19. Additional characteristics and quantitative data for associations of blood and
bone Pb with kidney outcomes for results presented in Figures 5-43 and 5-44
Reference Population
Study
Location;
Time n
Period
Pb Level
Outcome
Model
Change or %
Change
in Kidney Metric
(95% CI)
Cross-Sectional
Muntner et
al. (2003)
NHANESI
adults
US; 1988-
1994
4813
Mean (SD) blood Pb
Hypertensives: 4.2 (
0.14) |jg/dL
Q1: 0.7 to 2.4
Q2: 2.5 to 3.8
Q3: 3.9 to 5.9
Q4: 6.0 to 56.0
Normotensives: 3.3
(0.10) |jg/dL
Q1: 0.7 to 1.6
Q2: 1.7 to 2.8
Q3: 2.9 to 4.6
Q4: 4.7 to 52.9
Elevated Serum
Creatinine Logistic
(99th percentile of each regression
race-sex specific
distribution for healthy
young adults)
Hypertensives
Q1: Referent
Q2: 28% (2, 54)
Q3:8%(4,12)
Q4: 5% (3, 6)
Normotensives
Q2:10% (-57, 78)
Q3: 3% (-9,15)
Q4:1 % (-4, 5)
Akesson et WHILA, adult Sweden; 820 Median (5-95%) =
al. (2005) women 6/1999-1/2000 2.2 (1.1-4.6) |jg/dL
Cystatin C-based
eGFR (Larsson et al.
2004)
Creatinine clearance
Multiple
Linear
regression
-2.0 (-3.2, -0.9)
Linear
regression
-1.8 (-3.0, 0.7)
Navas-Acien NHANES III, US; 1999-
et al. (2009) adults 2006
14,778 Geometric mean =
1.58 |jg/dL
Q1: < 1.1
Q2: 1.2 to 1.6
Q3: 1.7 to 2.4
Q4: >2.4
eGFR <60
mL/minute/1.73 m2
Logistic
regression
Q1: Referent
Q2: 19% (-44, 83)
Q3: 28% (0.0, 56)
Q4: 19% (7, 31)
Albuminuria and eGFR
<60 mL/minute/1.73 m2
Logistic
regression
Q1: Referent
Q2: -37% (-83, 8)
Q3: -8% (-25,10)
Q4: 7% (-2,16)
Fadrowskiet NHANES, US; 1988-
aI. (2010) adolescents 1994
769
Median = 1.5 |jg/dL
Q1: <1.0
Q2:1.0 to 1.5
Q3:1.6 to 2.9
Q4: >2.9
Cystatin C-based
eGFR (mL/min/1.73 m2;
calculated using the
Filler and Lepage
equation)
Linear
regression
Q4: -0.42 (-0.73,-0.11)
Kimetal. Adult males Boston, MA;
(1996) 1979-1994
459 Median = 8.6 |jg/dL
Serum creatinine
concentrations
Random-
effects model
< 40 |jg/dL blood Pb:
0.016(0.004,0.028)
< 25 |jg/dL blood Pb:
0.019(0.006,0.032)
< 10 |jg/dL blood Pb:
0.030(0.011,0.049)
Payton et al.
(1994)
Adult males
Boston, MA;
1988-1991
744
Mean (SD) = 8.1
(3.9) ug/dL
Ln creatinine clearance
Multiple linear
regression
Ln blood Pb
¦4.0 (-8.0, -0.1)
Tsaih et al.
(2004)
Adult males
Boston, MA;
8/1991-1995
with mean 6 yr
follow-up
448
Mean (SD)
Blood Pb = 6.5 (4.2)
|jg/dL
Tibia Pb = 21.5
(13.5) Mg/g
Patella Pb = 32.4
(20.5) |jg/g
Serum creatinine
Multiple linear
regression
Baseline blood Pb
Diabetic: -0.05 (-0.23,
0.12)
Nondiabetic: -0.02 (-0.06,
0.02)
Hypertensive: -0.01 (-0.09,
0.07)
Nonhypertensive: -0.03 (-
0.07,0.01)
Follow-up blood Pb
Diabetic: 0.22 (-0.14, 0.58)
Nondiabetic: 0.14 (0.03,
0.26)
Hypertensive: 0.35 (0.16,
0.54)
Nonhypertensive: 0.06 (-
0.07,0.19)
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Reference Population
Study
Location;
Time
n
Pb Level
Outcome
Model
Change or %
Change
in Kidney Metric
Period
(95% CI)
Staesson et
al. (1992)
Adults
Belgium;
1985-1989
1,981
Blood Pb Mean (SD)
Males: 11.4 |jg/dL
Females: 7.5 yg/dL
Creatinine clearance
Multiple linear
regression
-52.9 (-84, 21)
Longitudinal
Kim etal.
Adult males
Boston, MA;
459
Median = 8.6 |jg/dL
Change in serum
Random-
< 40 |jg/dL blood Pb: 0.01
(1996)
1979-1994
creatinine
concentrations
effects model
(-0.0, 0.02)
< 25 |jg/dL blood Pb: 0.01
(-0.0, 0.03)
< 10 |jg/dL blood Pb: 0.02
(-0.0, 0.04)
Tsaih et al.
(2004)
Adult males
Boston, MA;
8/1991-1995
with mean 6 yr
follow-up
448
Mean (SD)
Blood Pb = 6.5 (4.2)
kig/dL
Tibia Pb = 21.5
Change in serum
creatinine
Multiple linear Blood Pb
regression Diabetics (n=26): 0.076
(0.03, 0.12)
Nondiabetic (n=422):
(13.5) Mg/g
Patella Pb = 32.4
(20.5) |jg/g
0.006 (-0.004, 0.02)
Hypertensive (n= 115):
0.008 (-0.01,0.03)
Normotensive (n=333):
0.009 (-0.003, 0.021)
Tibia Pb
Diabetics (n=26): 0.082
(0.03, 0.14)
Nondiabetic (n=422):
0.005 (-0.01,0.02)
Hypertensive (n= 115):
0.023 (0.003, 0.04)
Normotensive (n=333):
0.0004 (-0.01, 0.01)
Yu etal.
Adult CKD
Taipei, Taiwan; 121
Mean (SD) = 4.2
Change in MDRD
Generalized
-4.01 ml/min/1.73 m2 body
(2004)
patients
48 month
(2.2) |jg/dL
eGFR
estimating
surface area (p=0.0148) in
longitudinal
equations
the GFR over the follow-
study period
up period for each 1 |jg/dL
increment of blood Pb
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Occupational Studies
The vast majority of studies in the literature on the impact of Pb on the kidney have been
conducted in the occupational setting. In general, study size and extent of statistical analysis are much
more limited than for general population studies. Publications in only three 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 (V. M. Weaver et al.. 2009). Tibia Pb was evaluated at the beginning of each
follow-up period and Pb measures were adjusted for baseline Pb dose 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. 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 men which may
have been important factors in the effect modification observed by sex.
Chia and colleagues observed a significant, positive association between blood Pb and urine NAG
in linear regression models after adjustment for age, gender, race, exposure duration, ALAD G177C
polymorphism and the interaction between ALAD and blood Pb (Chia et al.. 2006). Similar positive
associations were observed between blood Pb and a wider range of EBE markers in models that adjusted
for age, gender, race, exposure duration, and the HpyCH4 ALAD SNP (discussed below) (Chia et al..
2005). A study of 155 male workers reported significant, positive correlations between blood and urine Pb
and urine NAG and albumin after controlling for age and job duration (Y. Sun. Sun. Zhou. Zhu. Lei, et al..
2008). An important additional study that analyzed occupational Pb exposure is discussed below under
patient population studies (Evans et al.. 2010).
Two studies 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 (Y. Sun. Sun. Zhou. Zhu. Lei, et al.. 2008). and 25.3
(.ig/dL (T. A. Lin & Tai-vi. 2007).
A number of other publications in the five years since the 2006 Pb AQCD have reported
significantly worse kidney outcomes in unadjusted analyses in occupationally exposed workers compared
to unexposed controls (Patil et al.. 2007) and/or significant correlations between higher Pb dose and
worse kidney function (Alinovi et al.. 2005; Garcon et al.. 2007; D. A. Khan et al.. 2008; T. A. Lin & Tai-
vi. 2007; Y. Sun. Sun. Zhou. Zhu. Lei, et al.. 2008). One small study found no significant differences
(Qrisakwe et al.. 2007).
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Overall, the occupational literature published in the last five years on the kidney impact of Pb
exposure has been more consistent in reporting significant associations than 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.
Publications that include dose-response information provide evidence of Pb-related nephrotoxicity
in the occupational setting across the Pb dose ranges analyzed (Ehrlich et al.. 1998; V. M. Weaver. Lee, et
al.. 2003). Data in 267 Korean Pb workers in the oldest age tertile (mean age = 52 years) reveal no
threshold for a Pb effect (beta = 0.0011, p = <0.05; regression and lowess lines shown) (V. M. Weaver.
Lee, et al.. 2003) (added variable plot shown in Figure 5-45).
_i
T3
O)
E
CD
C
c
o
CC
CD
o
E
=3
CD
CO
"O
CD
-»—•
CD
o
CO
D
TJ
<
Adjusted blood Pb level (|jg/dL)
Source: Used with permission from the BMJ Publishing Group, Weaver et al. (2003)
Note: Both the adjusted regression line and the line estimated by the smoothing method of the S-
PLUS statistical software function lowess 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 5-45. 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, increased risk for subsequent CKD occurs. 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 three occupational populations (Hsiao et al.. 2001;
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Roels et al.. 1994; V. M. Weaver. Lee, et al.. 2003). a study of adults who were Pb poisoned as children
(H. Hu. 1991). and a study in European children (De Burbiire 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. Gonick. Cohen. Alinovi. et al.. 1992). Pb could induce glomerular hypertension resulting in
hyperfiltration by several mechanisms including increased ROS, changes in eicosanoid levels, and/or an
impact on the renin-angiotensin system (Roels et al.. 1994; Vaziri. 2008b). 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 (V. M. Weaver. Lee, et al.. 2003; V. M. Weaver. Schwartz, et al.. 2003). significant
associations in opposite directions were observed only when relevant effect modifiers were included in
the model. This is a valid concern for risk assessment, since the factors involved in these inverse
associations in Pb-exposed populations are not well defined at present.
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 criteria
and markers of kidney damage are not required for the second criteria.
Several key studies in CKD patient populations have been published in the last five years (Table 5-
20). One CKD patient study, discussed in the 2006 Pb AQCD, remains the hallmark for prospective
evaluation of susceptible patient populations to determine if CKD progression (kidney function decline) is
greater in participants with higher baseline Pb dose. Yu et al. (2004) 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 in the lower EDTA chelatable Pb levels (<80 jug Pb/72 hours) group (2/58). Associations
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between baseline chelatable or blood Pb level and change in eGFR (estimated by the MDRD equation
(Lcvev et al.. 1999) were modeled separately using GEE. 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. Two recent studies expanded the CKD patient populations in
which this effect was observed to those with diabetic nephropathy (J.-L. Lin. Lin-Tan. Yu. et al.. 2006)
and with the lowest Pb body burdens studied to date (J.-L. Lin. Lin-Tan. Li. et al.. 2006). Results of these
observational studies have been summarized (V. Weaver & Jaar. 2010).
Table 5-20. Patient population studies: kidney function decline
Study
n
Baseline
mean
(SD)
blood Pb
(ng/dL)
Baseline mean
(SD) chelatable
Pb (jo.g/72 hours)
Baseline mean
(SD) eGFR
(njl/min/1.73
m )
Years of
follow-up
Decline in
eGFR per 1 SD
higher Pb dose
at baseline per
year
Comments
Lin etal.
(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 etal.
(2006)
87
6.5 (3.4)
108.5(53.8)
35.1 (9.0)
1
3.87
Type II diabetics with
nephropathy
Lin etal.
(2006)
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: Used with permission from UpToDate.com, Weaver et al. (2010)
A recent population-based case-control study examined occupational Pb exposure as a risk factor
for severe CKD (Evans et al.. 2010). The study included 926 cases with first time elevations of serum
creatinine >3.4 mg/dL for men and >2.8 mg/dL for women and 998 population-based controls.
Occupational Pb exposure was assessed using an expert rating method based on job histories. 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/m3 (the current US OSHA limit is 50
|ig/m3). Using multivariable logistic regression modeling, the adjusted OR for CKD was 0.97 (95% CI:
0.68-1.38) in Pb-exposed compared to non-exposed participants. No significant increased odds were
observed for low, medium or high exposed groups using either average or cumulative exposure metrics.
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
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4.55 mL/min/1.73 m2/y in patients without occupational Pb exposure. Thus, in this study, no adverse
kidney effect of occupational Pb exposure was evident.
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. Two case-control studies
examining occupational risk factors for CKD have been published; one found Pb exposure to be a risk
factor (Nuvts et al.. 1995). the other did not although moonshine alcohol consumption was a risk
presumably due to Pb level (Steenland et al.. 1990). The prospective observational aspect of this study is
similar in design to the work of Lin and colleagues but differs in several important respects. In this study,
only occupational Pb exposure was considered whereas the work in Taiwan excludes occupational
exposure and uses Pb dose 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 |ag/g) ("V. M.
Weaver. Lee, et al.. 2003) 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 (Elinderet al.. 1986; Stromberg et al.. 1995). although blood
Pb levels were lower than those noted during the U.S. phase-out. Finally, the severe degree of CKD in this
population 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, which is characterized
by the use of dialysis to perform the normal functions of the kidney. Incidence and prevalence in the US
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. One study reported much higher blood Pb levels than had been
appreciated by the treating clinicians (Davenport et al.. 2009). Of 271 adult patients on regular thrice
weekly dialysis, 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's r = 0.38, p-value <0.001); negatively correlated
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with urine output (r = -0.44, p-value <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 (B. Chen et al.. 2009). 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 Pb dose 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 (B. Chen et al.. 2009; 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.. 2008V 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 significantly 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 examined
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 (J.-L.
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 negatively correlated
with logarithmic-transformed blood Pb levels after adjustment (J.-L. 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 and highest compared to the lowest tertiles (hazard ratio= 2.1 [95% CI: 2.0-2.2] and 3.3 [95% CI:
1.3-13.5], respectively). Given other recent publications in hemodialysis patients by this group, it would
be valuable to examine this risk after adjustment for serum ferritin (Jenq et al.. 2009). hemoglobin A1C
(Lin-Tan. Lin. Wang, et al.. 2007). and blood cadmium (C. W. Hsu et al.. 2009).
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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 5-21 below.
Table 5-21. Clinical randomized chelation trials in chronic kidney disease patients
Reference
Group
n
Baseline
mean
(SD)
blood Pb
(ng/dL)
Baseline
mean (SD)
chelatable Pb
(ng/72 hr)
Baseline mean
(SD) eGFR
(njl/min/1.73
m )
Months of
treatment/
follow-up
Change in
eGFR per yr
(njl/min/1.73
m )
Comments
Lin etal.
(2003)
Chelated
32
6.1 (2.5)
150.9(62.4)
32.0 (12.1)
27
+ 1.07
Largest study to
date
Control
32
5.9 (3.0)
144.5(87.9)
31.5(9.0)
-2.7
Lin etal.
(2006)
Chelated
15
7.5 (4.6)
148.0(88.6)
22.4 (4.4)
15
-3.5
Type II diabetics
with nephropathy
Control
15
5.9 (2.2)
131.4(77.4)
26.3 (6.2)
-10.6
Lin etal.
(2006)
Chelated
16
2.6 (1,0)a
43.1 (13.7)
41.2 (11.2)
27
+3.0
Lowest Pb
exposed and
treated range BLB
> 20- <80 |ig)
Control
16
3.0 (1.1)
47.1 (15.8)
42.6 (9.7)
-2.0
Lin-Tan et al.
(2007)
Chelated
58
5.0 (2.2)
164.1 (111.1)
36.8 (12.7)
51
-0.3
Non-diabetic
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 (Jacobsen et
al.. 2001; Saxena & Flora. 2004). EDTA benefits in a Pb rodent model appeared to occur via reduced
oxidation (Saxena & Flora. 2004). 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 non-Pb exposed rat model
(Gonick et al.. 1996). Benefits from chelation reported in rodent models of Pb-related nephrotoxicity
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(Khalil-Manesh. Gonick. Cohen. Bergamaschi. et al.. 1992; Sanchez-Fructuoso. Blanco, et al.. 2002;
Sanchez-F ructuoso. Cano. et al.. 2002) did not appear to occur via reversal of structural damage (Khalil-
Manesh. Gonick. Cohen. Bergamaschi. et al.. 1992); again suggesting that improved hemodynamics from
reduction of reactive oxidant species, which could be due to reduced Pb and/or directly from the chelating
agent, may be a mechanism (Gonick et al.. 1996). However, the most parsimonious explanation for the
combination of Lin's observational and experimental chelation work is that Pb is the underlying reason.
Moreover, if the benefit can be replicated, this could be a valid treatment regardless of the mechanism or
whether Pb is involved.
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 dose
assessment that includes bioavailable body burden, longitudinal statistical analysis, and control for
multiple kidney risk factors. However, the generalizability of the results to broader populations is
unknown. In addition, the observed effect of Pb on decline in GFRhas been variable; the annual decline
in eGFRper standard deviation (SD) higher Pb dose at baseline was much lower in the 2003 study than in
subsequent publications (see Table 5-21 above). Small sample sizes and differences in renal diagnoses
between groups may be factors in this variability. However, if confirmed in large populations at multiple
centers and shown not to worsen cognition or other effects through Pb mobilization, chelation could yield
important public health benefits.
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 (Tnglis 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 (C. 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 at much lower
environmental exposure 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).
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Three studies that included analysis of clinical kidney outcomes were discussed in the 2006 Pb AQCD.
One found no difference in mean serum creatinine between 62 exposed and 50 control children (Fels et
al.. 1998). Two larger studies observed significant associations that were in opposite directions; blood Pb
was positively associated with serum cystatin-C in 200 17-year-old Belgian adolescents but negatively
associated with serum creatinine and cystatin C in 300-600 European children (n varied by outcome)
(Dc Burbiire et al.. 2006; Staessen et al.. 2001).
Therefore, one of the key gaps identified in the 2006 Pb AQCD was limited data in children and
adolescents particularly with respect to GFR measures. A recently published NHANES analysis in
adolescents begins to fill this gap (Fadrowski et al.. 2010V Associations between blood Pb and kidney
function were investigated in 769 adolescents aged 12-20 years in the US NHANES III, conducted from
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 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 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 to -5.0) lower eGFR. In contrast, the association between blood Pb and
creatinine-based eGFR, although in the same direction, was not statistically significant. Additional
research in children, including with longitudinal follow-up, a range of outcome assessment methods, and
with exposure only after Pb was banned from gasoline, is warranted.
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 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 between cadmium dose measures were stronger for the cystatin C based outcome.
Staessen et al. (2001) found a significant association between 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 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
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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 3941 adults who participated in the 1999-2002 NHANES cystatin C
subsample (Spector et al.. 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.
5.5.3. Mechanisms of Lead Nephrotoxicity
5.5.3.1. Altered Uric Acid
Individuals who have been heavily exposed to Pb are at increased risk for both gout and kidney
disease (Batuman. 1993; Shadick et al.. 2000). Pb is thought to increase serum uric acid by decreasing its
kidney excretion (Ball & Sorensen. 1969; Emmerson. 1965; Emmerson & Ravenscroft. 1975). Research
during the past decade indicates that uric acid is nephrotoxic at lower levels than previously recognized
(R. J. Johnson et al.. 2003). Therefore, the 2006 Pb AQCD reviewed literature implicating increased uric
acid as one mechanism for Pb-related nephrotoxicity (Shadick et al.. 2000; V. M. Weaver et al.. 2005).
However, this is not the only mechanism, since associations between blood Pb and serum creatinine
remained significant even after adjustment for uric acid (V. M. Weaver et al.. 2005). 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 (J.-L. Lin et al..
2001).
Conterato et al. (2007) followed various parameters of kidney function after acute or chronic Pb
exposure in rats. Acute exposure to Pb acetate consisted of a single i.p. injection of 25 or 50 mg/kg Pb
acetate, while chronic exposure was one daily i.p. injection of either vehicle or Pb acetate (5 or 25 mg/kg
) for 30 days. Acute and chronic exposure at both dose levels increased plasma uric acid levels.
Conversely, Annabi Berrahal et al. (2011) found that plasma uric acid levels decreased after 65 days of Pb
exposure (post-puberty; blood Pb 7.5 (ig/dL) (Table 5-17). Plasma urea levels increased after 40 days of
exposure (puberty; blood Pb 12.7 (.ig/dL). Changes in plasma urea are used as an acute renal marker of
injury.
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5.5.3.2. Oxidative Damage
A role for ROS in the pathogenesis of experimental Pb-induced hypertension and renal disease has
been well established (Vaziri. 2008a. 2008b; Vaziri & 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 acute Pb exposure can elevate kidney GST levels, affecting glutathione metabolism (Daggett
etal.. 1998; Moser et al.. 1995; Oberlev et al.. 1995).
Annabi Berrahal et al. (2011) reported on the effects of age-dependent exposure to Pb on
nephrotoxicity in male rats (Table 5-17). Pups were exposed to Pb lactationally (as a result of dams
consuming water containing 50 ppm Pb acetate) until weaning. Thereafter the male pups were exposed to
the same solution from weaning (day 21) until sacrificed at age 40 days (puberty; blood Pb 12.7 (ig/dL)
and at age 65 days (post-puberty; blood Pb 7.5 (ig/dL). MDA concentration in kidney was significantly
increased relative to controls to the same degree at both 40 and 65 days, while total sulfhydryl groups
were significantly decreased only at 65 days. These changes reflect an increase in oxidative stress after
exposure to Pb.
Conterato et al. (2007) examined the effect of Pb acetate on the cytosolic thioredoxin reductase
activity and oxidative stress parameters in rat kidneys. Acute exposure to Pb acetate consisted of a single
i.p. injection of 25 or 50 mg/kg Pb acetate, while chronic exposure 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, 8-ALAD, GST, GPx, non protein thiol groups (NPSH), CAT, as
well as plasma creatinine, uric acid, and inorganic phosphate levels. Acute exposure at the 25 mg Pb dose
level resulted in increased SOD and thioredoxin reductase-1 activity, while exposure to the 50 mg dose
level increased CAT activity and inhibited 8-ALAD activity in the kidney. Chronic exposure at the 5 mg
dose level of Pb inhibited 8-ALAD and increased GST, NPSH, CAT, and thioredoxin reductase-1.
Chronic exposure to the 25-mg dose level reduced 8-ALAD, but increased GST, NPSH, and plasma uric
acid levels. No changes were observed in TBARS, GPx, creatinine or inorganic phosphate levels after
either acute or chronic exposure. As both acute and chronic exposure to Pb increased thioredoxin
reductase-1 activity, the authors suggest that this enzyme may be a sensitive indicator to exposure at low
Pb dosage.
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 well as the blood activity and urinary concentration of S-ALA. The concentrations of
GSH and NPSH were decreased by Pb administration, while MT concentration was unchanged. 8-ALAD
in blood was decreased, whereas urinary S-ALA was increased by Pb administration. Negative
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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
El-Neweshy and El-Sayed 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. Rats treated with Pb
acetate plus vitamin C exhibited relatively mild karyomegaly and eosinophilic intranuclear inclusion
bodies of proximal tubules in 5 rats, while an additional 5 rats were normal. Normal glomeruli were noted
in all. Thus vitamin C could be shown to ameliorate the renal histopathological effects of Pb intoxication.
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 analyzed. Pb-exposed pups had
decreased body weights. Blood Pb level in the control group was 1.43 (ig/dL, in the Pb group it was 22.8
(ig/dL, in the Pb plus zinc plus vitamins it was 21.2 (ig/dL, and in the zinc plus vitamin group blood Pb
was 0.98 (ig/dL. The kidney TBARS were significantly elevated in Pb exposed pups, while treatment with
vitamins and zinc returned TBARS to control levels. Kidney catalase activity was significantly increased
above control with Pb treatment; however supplement with zinc and vitamins reduced catalase activity
towards 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 is effective in
attenuating the redox imbalance induced by developmental, chronic low-level Pb exposure.
Bravo et al. (2007) reported further that mycophenolate mofetil (an immunosuppressive agent used
in renal transplantation which inhibits T and B cell proliferation) administration 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.
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Although the majority of studies of the effects of Pb exposure have been on male rats, two studies
have appeared which compare the response of male rats with female rats (Alghazal. Lenartova. et al..
2008; Sobekova et al.. 2009). 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 1000 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 1000 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. (2008) 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 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 response of male and female rats to Pb exposure. Differences could 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 females, so that the biological half4ife
of xenobiotics in the females is longer.
5.5.3.3. Lead Effect on 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. Perez Aguilar et al. (2008) studied
changes in renal gangliosides following Pb exposure (600 ppm Pb acetate in their 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 S-ALA.
The following renal gangliosides were measured by immunohistochemistry and by thin layer
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chromatography: GM1, GM2, GM4, and 9-O-acetylated modified form of the GD3 ganglioside (9-O-Ac-
GD3). The ganglioside pattern was mainly characterized by a decrease in the GM1 ganglioside as well as
by a mild increase in GM4 and GM2 gangliosides, while the strongest alteration was observed in the 9-0-
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
exerts a dual effect on MT expression in the kidney: enhancement of MT gene transcription but
suppression of MT mRNA translation.
Zuo et al. (2009) explored 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.
Figure 5-46 (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. Dysfunction in
kidney function measures, including urinary flow, ALP, microalbumin, and NAG, was observed at blood
Pb concentrations above 19.7 (.ig/dL (L. Wang et al.. 2010).
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® Highest Concentration ~Lowest Cone, with Response
i Highest Cone, with No Response O LowestConcentration
Biomarker (L. Wang et al.. 2010)
Inflammation (Roncal et al.. 2007)
Oxidative Stress - Redox (Navarro-
Moreno et al.. 2009)
Oxidative Stress - Redox (L. Wang
et al.. 2010)
Oxidative Stress - Redox (Masso-
Gonzalez & Antonio-Garcia. 2009)
Morphology (Roncal et al.. 2007)
Morphology (Navarro-Moreno et al..
2009)
Morphology (L. Wang et al.. 2010)
Morphology (Masso-Gonzalez &
Antonio-Garcia. 2009)
Kidney function (Navarro-Moreno et
al.. 2009)
Kidney function (Roncal et al.. 2007)
Kidney function (L. Wang et al..
2010)
Kidney function (Ademuviwa et al..
2009)
~
~
~
~
~
~
~
~
~
~
~
~
~—
1 10 100
Blood Pb Level (ng/dL)
Figure 5-46. Dose-responsive representation of the effect of Pb on renal
outcomes in animal toxicology studies.
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Table 5-22. Additional characteristics for results of toxicological studies presented in
Figure 5-46
Reference
Blood Pb Level
with Response
(Mg/dL)
Outcome
Wang et al. (2010)
20
Biomarker - Aberrant NAG, GGT, |32-microgobulin expression in adult female mice
with chronic Pb exposure.
Roncal et al. (2007)
26
Inflammation - Elevation in number of macrophages & marker MCP-1 in Pb exposed
kidneys with remnant kidney surgery.
Navarro-Moreno et al. (2009)
43
Oxidative stress - Chronic Pb exposure in males increased kidney lipid peroxidation
(i.e. TBARS)
Wang et al. (2010)
20
Oxidative Stress - Pb caused increased lipid peroxidation (i.e. MDA production),
elevated kidney antioxidant enzymes (SOD, GPx, CAT), and depleted GSH in
immature female rats.
Masso-Gonzalez et al. (2009)
23
Oxidative stress - Elevated TBARS and catalase activity in weanling pups exposed to
Pb during gestation and lactation
Roncal et al. (2007)
26
Morphology - Pb induced pre-glomerular vascular disease of kidney (i.e. sclerosis,
fibrosis, peritubular capillary loss)
Navarro-Moreno et al. (2009)
43
Morphology - Electron micrography of chronic Pb exposure in male rats showed
lumen reduction, microvilli loss, brush border loss, and mitochondrial damage
Wang et al. (2010)
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)
23
Morphology - Pb elevated relative kidney weight at PND21 in animals with neonatal
Pb exposure.
Navarro-Moreno et al. (2009)
43
Kidney function - Pb exposed males had elevated urinary pH and protein, and
glucose and blood in the urine.
Roncal et al. (2007)
26
Kidney function - Remnant kidney surgery and Pb exposure induced decreased
creatinine clearance and proteinuria.
Wang et al. (2010)
20
Kidney function - Elevated urinary total protein, urinary albumin, and serum urea
nitrogen in immature female rats exposed to Pb.
Ademuviwa et al. (2009)
39 and 61
Kidney Function - Renal phospholipidosis and depletion of renal cholesterol in male
pups after gestational Pb exposure.
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 are 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 (U. S. 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-dependent lipid peroxidation and abrogates the Pb-dependent
attenuation of GR and SOD. Concomitant exposure to the cations aluminum and Pb protects Pb-exposed
animals from ensuing nephropathy (Shakoor et al.. 2000). In summary, Pb is known to affect processes
mediated by endogenous divalent cations. In addition, exposure to other metals or divalent cations can
modulate Pb disposition and its effects in the body.
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5.5.4.1. Lead and Cadmium
Cd shares many similarities with Pb; it is a ubiquitous PT nephrotoxicant at high exposure levels
and accumulates in the body. Despite this, few studies have evaluated associations between low4evel Cd
exposure and CKD or the impact of joint exposure of these or other metals on CKD. As discussed in the
2006 Pb AQCD, Cd, at the lower exposure levels common in the U.S. and other developed countries, has
a substantial impact on associations between Pb exposure and the kidney EBE marker, NAG, even in the
presence of occupational level Pb exposure. In one report, mean NAG, although higher in the Pb-exposed
group compared to controls, was correlated with urine Cd and not blood or tibia Pb (RoeIs 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 a 66.9 jxg/g bone mineral increase in tibia Pb (V.
M. Weaver. Lee, et al.. 2003).
The 2006 Pb AQCD noted that data examining the dose-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-Acien
et al.. 2009). Geometric mean 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 odds ratios 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 although effect modification was not observed.
However, the odds ratios comparing participants in the highest with the lowest quartiles of both metals
were 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 CKD risk factors in the general population, and provide novel evidence of increased risk in
those with higher environmental exposure to both metals.
However, a very recent study suggests that interpretation of low4evel 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 low4evel Cd exposure in workers who are occupationally exposed to other
nephrotoxicants such as Pb, Cd dose was assessed 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) ju.g/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 (V. M. Weaver et al.). After adjustment for age, sex, BMI,
urine creatinine, smoking, alcohol, education, annual income, diastolic BP, current or former Pb worker
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job status, new or returning study participant, and blood and tibia Pb, higher ln-urine Cd was associated
with higher calculated creatinine clearance, eGFR(P = 8.7 ml/min/1.73 m2 [95% CI: 5.4, 12.1]) and ln-
NAG, but lower serum creatinine. These unexpected paradoxical associations have been reported in two
other publications (De Burbiire et al.. 2006; Hotz et al.. 1999) and have been observed in two 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. (2009) 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 rtPCR to 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, consisting of 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 our knowledge of the synergistic
effects of Pb and Cd on kidney mitochondria.
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 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 addtion 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 & Fowler. 1977).
Stress proteins were examined after exposure to mixtures of Pb and other metals. Induction of MT
was strongest in groups with Cd treatment. However, co-exposure to Pb and As induced higher levels of
MT protein than either Pb or As exposure alone in kidney tubule cells. Heat shock proteins (Hsps) are
commonly altered under the situation of exposure to metal mixtures. Both in vitro low dose studies and in
vivo studies showed that Hsps were induced in a metal/metalloid, dose and time-specific manner (G.
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Wang et al.. 2005). Additive or more than additive interactions occur 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. Zinc (10 mg/kg i.p.) and ascorbic acid (10, 20 and 30 mg/kg i.p.) treatment
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 (DR, 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 marginal zinc rats had higher renal Pb (35%) and lower
renal zinc (16%) concentrations than control rats. On the other hand, supplemental zinc was more
protective than the control diet against renal Pb accumulation (33% lower). Standard procedures for
indirect immunoperoxidase staining were used to determine MT localization in the kidney. Pb had no
effect on MT staining intensity, distribution, or relative protein amounts. Western blot analysis confirmed
that MT levels were responsive to dietary 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 heat shock proteins,
glucose-regulating proteins and metallothioneins in a rat proximal tubular cell line (NRK-52E) exposed to
subcytotoxic doses of inorganic mercury (HgCl2, 1-40 (j,M) and Pb (PbCl2, 2-500 (j,M) were analyzed.
Reactive oxygen and nitrogen species 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 heat shock proteins and metallothionein.
Reverse transcription PCR was used for quantification of metallothionein. The higher doses of mercury
(20 (j,M and 40 (jM) were shown to markedly inhibit growth of the cell line while the higher doses of Pb
(60 (j,M to 500 (j,M) 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
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hours. Mercury induced a significant increase in both and reactive nitrogen species, the reactive nitrogen
species maximal at 24 hours, and the ROS at 48 hours. Pb (60 or 300 (.iM) did not cause an increase in
reactive oxygen or reactive nitrogen species beyond the levels seen in control cells. Total GSH
significantly increased in cells grown in the presence of Pb; the effect was dose-dependent and GSH
reached its maximal value at a dose of 300 |_iM Pb. The effect of mercury was biphasic, 10 (jM
significantly enhancing GSH by 600%, while the amount of GSH detected after 20 (.iM mercury only
increased by 50% compared to control. GST activity was enhanced by both Pb and mercury. Heat shock
proteins 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 metallothionein 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
dose- 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 metallothionein synthesis clearly differ from the
study of Jamieson et al. (2008) that found no increase in metallothionein following Pb exposure. This
discrepancy remains to be clarified.
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) 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 towards 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 towards normal by Pb plus NAC treatment (MDA increased and the
remainder decreased). Thus NAC can be shown to have both an anti-oxidative and a chelator effect on Pb
intoxication.
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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 as to their capacity to reduce Pb concentration in blood and soft tissues and to recover Pb-
induced oxidative stress in male Wistar rats who were exposed to Pb acetate (0.1% in drinking water) for
20 weeks. Rats were then treated orally with five days of DMSA or its two analogues at a dose of up to
100 mg/kg once daily. 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 8-aminolevulinic acid synthatase (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.
Abdallah et al. (2010) explored the effect of Pb toxicity 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
in a dose of 5 mg/kg daily for 6 weeks. No blood Pb levels were reported. TBARS were elevated above
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 (48.6 ±5.6 versus 95.5 ± 10.1 nmol/g
tissue, oxidized, and 35.4 ±3.0 versus 61.4 ± 5.1 nmol/g tissue, reduced). On the other hand, levels of
oxidized and reduced coenzyme Q10 were unchanged. Thus the reduced levels of coenzyme Q
attributable to Pb intoxication may participate 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 plus 4-hydroxyalkenals (4-HAD). Pb increased kidney lipid peroxidation products, but these were
reduced towards normal by melatonin. 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 conclude 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.
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Ozsoy et al. 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 and L-carnitine group serum creatinine was reduced to
control values and the histopathological changes were reversed. Immunological staining for Cu/Zn-SOD
was elicited by Pb feeding and reduced by L-carnitine. The authors attribute the beneficial effects of L-
carnitine to its antioxidant effect.
Reddy et al. (2010) used Sprague-Dawley rats that were treated 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.
Kharoubi et al. (2008) described the prophylactic effects of Wormwood (Artemisia absinthium L.)
plant extracts on kidney function on 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 (blood Pb 55.6 (ig/dL compared to 22.3 (ig/dL). 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 towards 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.
The effectiveness of various plant or bacterial extracts as antioxidants in the kidney was explored
in two separate publications. 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 Pb acetate, 20 ppm, and/or an extract of A vulgaris, 100
ppm, 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
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glomeruli while rats treated with Pb and then the extract showed improvement in tubular structure, but
interstitial fibrosis was still present. This experiment confirmed 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
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 exposed to 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 thus increased renal blood flow by enhancing the delivery of
NO. This question arises because of the observations of Lin et al. (2006) that repeated injections of
CaNa2EDTA leads to improvement in kidney function in patients with chronic renal failure, even in
individuals with very low body Pb stores. 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 suggest 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. (2006) 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 8-ALAD activity, both in vivo and ex vivo. Ex vivo, human blood was
pre-incubated with BAL or DMSA (10 (.iM) 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) 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 8-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 conjecture that the
chelators may deplete the cells of zinc, an essential element for 8-ALAD activity. Supporting the
chelation effect seen is the Santos study is work by Bradberry and Vale (2009). Hamidinia et al. (2006).
and Aslani et al. (Asian i 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 associated with blood and bone Pb levels (U.S.
EPA. 2006). Data in general and patient populations provided consistent evidence of low4evel Pb
nephrotoxicity (Akesson et al.. 2005; R. Kim et al.. 1996; Tsaih et al.. 2004; C.-C. Yu et al.. 2004); effects
on eGFR were observed in human hypertensives at mean blood Pb level of 4.2 (ig/dL (Muntner et al..
2003). These findings were substantiated by the coherence of effects observed across epidemiologic and
toxicological studies. Both human and animal studies have observed 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, chronic exposure 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 public health significance of such effects in
the general population, the implications of hyperfiltration, and reverse causality.
Recent epidemiologic studies in adult general and patient populations, with few exceptions,
continue to be consistent in observing associations between blood and bone Pb levels and worse kidney
function and provide important evidence that nephrotoxicity occurs at current population levels of
biomarkers of Pb exposure. These studies benefit from a number of strengths that vary by study but
include comprehensive assessment of Pb dose with bone Pb as a measure of cumulative body burden and
chelatable Pb as a measure of bioavailable Pb; prospective study design; and statistical approaches that
utilize a range of exposure and outcome measures, while adjusting for numerous kidney and Pb risk
factors. Large sample sizes provide strength to the general population studies. Reexamination of a study
from the 2006 Pb AQCD provided data to conclude that a 10-fold increase in blood Pb (e.g., from 1 to 10
(ig/dL) would result in 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
adverse impact on eGFR as an increase of 4.7 years in age or 7 kg/m2 in body mass index (Akesson et al..
2005). In populations with lower blood Pb levels, 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. At blood Pb levels that are common in the general U.S. population, Pb increases
the risk for clinically relevant effects particularly in susceptible populations such as those with underlying
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chronic medical diseases that increase CKD risk such as diabetes mellitus and hypertension and co-
exposure to other environmental nephrotoxicants.
Absence of impact in the occupational setting cannot be used as a rationale for discounting Pb-
related nephrotoxicity at lower environmental levels. Research in the occupational setting has
traditionally been far less consistent than in environmentally exposed populations. 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 cannot be resolved solely by utilizing stronger research techniques. 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.
Important data on the kidney effects of Pb on children were reported in a recent NHANES analysis
in adolescents that observed an association between higher blood Pb and lower cystatin C-based eGFR
(Fadrowski et al.. 2010). These findings are consistent with a rodent model 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
(Berrahal et al.. 2011V These limited studies add to the strength of the association between blood Pb and
altered renal function in children despite the need for additional research.
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 associated with
increased glomerular sclerosis, tubulointerstitial injury, increased collagen staining, and an increase in
macrophages associated with higher levels of MCP-1 mRNA (Roncal et al.. 2007). 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 mechanisms
leading to these renal alterations, especially, similar to the cardiovascular system, the influence of Pb-
induced oxidative stress. The mode of action of Pb in the kidneys has been extended to the field of
immunology, where it was shown that low Pb exposure results in infiltration of lymphocytes and
macrophages associated with increased expression of NFkB in proximal tubules and infiltrating cells.
Additionally, recent evidence expands on the evidence of acute effects of Pb, including mitochondrial
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dysfunction, renal cell apoptosis, and glomerular hypertrophy. These mechanisms are useful in
understanding the occurrence of acute hyperfiltration followed by chronic kidney dysfunction.
Current evidence does not allow for the identification of a threshold for Pb-related nephrotoxicity;
increased odds of CKD (characterized by eGFR <60 mL/min/1.73 m2) were apparent in NHANES
analyses that included data collected as recently as 2006 (Navas-Acien et al.. 2009). The odds of reduced
eGFR increased by 36% (95% CI: 0.99, 1.85) at blood Pb levels as low as 1.6-2.4 (ig/dL and by 56%
(95% CI: 1.17, 2.08) at blood Pb >.4 ^ig/dL.
In summary, new studies evaluated in the current review support or expand upon the strong body of
evidence presented in the 2006 Pb AQCD that biomarkers of Pb exposure are associated with renal health
effects. Epidemiologic studies continue to demonstrate a consistently positive relationship between blood
Pb level and kidney dysfunction at blood Pb levels (mean <2 (ig/dL) comparable to those occurring in the
current U.S. population with no evidence for a threshold across the range of levels studied. By
demonstrating Pb-induced oxidative stress and describing mechanisms of acute changes following Pb
exposure, toxicological studies provide biological plausibility for the associations observed in
epidemiologic studies between Pb and kidney dysfunction. Collectively, the evidence integrated across
epidemiologic and toxicological studies as well as across the spectrum of kidney health endpoints 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, Pb is one of the most
extensively researched and studied immunotoxicants. Experimental 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 (DIT), were first conducted during the 1970s.
Despite the long history of Pb-associated immunotoxicity research, the immune-based 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, are a relatively recent finding from within the last 10-15 years (Dictcrt & McCabe. 2007). Over the
same time period, similar advances in the understanding of Pb-associated changes in immunological
parameters in humans without occupation Pb exposures have substantiated the immunomodulatory effects
of Pb.
In the 2006 Pb AQCD (U.S. EPA. 2006). both toxicological studies in animals and epidemiologic
studies of humans provided strong evidence that the immune system was one of the more sensitive
systems affected by Pb exposure. However, rather than producing overt cytotoxicity or pathology, Pb
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exposure of experimental models and blood Pb levels in humans were found to be associated with
alterations in the abundance and function of a variety of immune cells (Figure 5-47). In both toxicological
and epidemiologic studies, macrophages and T lymphocytes were observed to be particularly sensitive to
Pb, but Pb-associated changes were also reported in B lymphocytes and neutrophils. Several of these
changes were observed at blood Pb levels <10 (ig/dL or the equivalent in humans and experimental
models, levels at which neurological effects also were observed.
Alterations in these aforementioned immune cells can lead to changes changes in cell-to-cell
interactions, multiple signaling pathways, and inflammation that affect both innate and acquired
immunity, that in turn, influence risk of developing infectious, allergic and autoimmune diseases as well
as exacerbating inflammatory responses in other organ systems (Figure 5-47).
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
Increased risk of atopy
and allergic disease
Increased risk of tissue
inflammatory diseases
Increased risk of
autoimmunity
Reduced host resistance
to bacterial infection
Figure 5-47. Immunological pathways by which Pb exposure may increase
risk of immune-related diseases.
Studies conducted in animal and in vitro models have provided consistent evidence for Pb inducing
effects on the range of immune effects presented in this continuum. Among the hallmarks reported for Pb-
induced changes in functional pathways are: (1) a suppression ofT-derived lymphocyte helper (Th)l-
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driven cell-mediated immunity (as measured by a delayed-type hypersensitivity [DTH] response); (2) an
increase in Th2-driven immunoglobulin E (IgE) antibody production; and (3) a proinflammatory shift in
macrophage function. The latter was characterized by increased production of reactive oxygen
intermediates reactive oxygen species (ROS), prostaglandin E2 (PGE2), and inflammatory cytokines such
as tumor necrosis factor-a (TNF-a) and interleukin (IL)-6 and decreased production of IL-12 and nitric
oxide (NO). In animal studies, these effects of Pb were more prominent with in utero Pb exposures and in
males. In the 2006 Pb AQCD, epidemiologic evidence was available for relatively fewer immune
endpoints (e.g., proinflammatory shifts in immune cell function); however, coherence was observed
between toxicological and epidemiologic findings for Pb-associated changes in circulating IgE levels and
T and B cell abundance and activation (U.S. EPA. 2006V Although many epidemiologic studies indicated
associations between blood Pb levels and immune system effects, several limitations of study design and
analytic methods were noted, including cross-sectional analyses; small sample size; inconsistent
adjustment for potential confounders such as age, sex, smoking, and comorbid conditions; and limited
assessment of the magnitude of association between blood Pb levels and changes in immune function.
Consistent with inhibition of Thl activity, toxicological evidence presented in the 2006 Pb AQCD
linked Pb exposure of animals to impaired host resistance and increased risk of certain infections (U.S.
EPA. 2006). Consistent with inducing a hyperinflammatory state and local tissue damage, Pb exposure
was found to induce generation of autoantibodies, indicating an elevated risk of autoimmune reactions.
Additionally, the demonstrated shift toward a Th2 response suggested that Pb could elevate the risk of
atopy and allergic responses. While toxicological studies provided the evidence for biological plausibility,
the epidemiologic evidence was too sparse to draw conclusions regarding associations between blood Pb
levels and these broader indicators of immune dysfunction in humans.
Studies published since the 2006 AQCD support the previous findings of Pb-induced immune
effects and demonstrate similar effects at lower blood Pb levels (<2-5 (ig/dL). Recent studies also expand
on the array of immunological parameters affected by Pb exposure as presented in Figure 5-47. For
example, new toxicological evidence indicates that Pb modulates function of dendritic cells. Results from
new toxicological and epidemiologic studies strengthen 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 5-47). New epidemiologic studies examine signaling
molecules such as proinflammatory cytokines and NO to produce findings parallel with toxicological
studies. Another important advance is the increasing knowledge of the broader role of Pb-associated
immune modulation in mediating Pb effects in nonlymphoid tissues (e.g., in the neurological,
reproductive, and respiratory systems). Although primarily cross-sectional in design, recent epidemiologic
studies address many limitations of earlier studies through greater examination of children and adults
without occupation Pb exposures with blood Pb levels more comparable to those currently measured in
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the U.S. population and greater consideration of confounding by age, sex, smoking, and comorbid
conditions.
5.6.2. Cell-Mediated Immunity
5.6.2.1. T Cells
Toxicological and epidemiologic evidence reviewed in the 2006 Pb AQCD consistently
demonstrated Pb effects on decreasing T cell populations (U.S. EPA. 2006). which mediate responses to
antigens and infectious agents. Consistent with previous findings, in a recent study of Wistar rats
administered 200 ppm Pb acetate, the percentages of CD4+ (T helper) and CD8+ (cytotoxic T) T cells
were decreased (with CD4-CD8- cells elevated) in the submaxillary lymph nodes (p <0.05), with
intraperitoneal (i.p.) exposure but not oral exposure (Teiion et al.. 2010). The 2006 Pb AQCD also
described numerous toxicological studies in which Pb exposure shifted the development and/or activation
of CD4+ T cell populations such that production of Th2 cytokines was favored and production of Thl
cytokines was suppressed (U.S. EPA. 2006). Recent studies expand information on the potential
mechanisms underlying T cell activation. In cultures of human CD4+ T cells, Pb (1 (iM, 30 minutes) has
been shown to activate transcription factor NFkB (regulates T cell activation) (Pvattetal.. 1996) and to
increase, in a dose-dependent manner (10 and 50 (.iM PbCl2, 24 hours), the expression of MHC class II
surface antigens (HLA-DR), which mediates the CD4+ response to exogenous antigens (Guo et al..
1996b). Heo et al. (2007) provide evidence for the direct effects of Pb on T cells by showing that Pb (25
(.iM) blocked production of the Thl cytokine interferon-y (IFN-y) in cultures of stimulated mouse T cells
by suppressing translation of the protein. This blockage was rescued with the addition of IL-12. Kasten-
Jolly et al. (2010) found that Pb may not necessarily skew towards a Th2 phenotype via a direct effect on
T cells. In this study, developmental Pb exposure of mice (O.lmM Pb acetate in drinking water of dams
from GD8 to PND21, resulting in pup blood Pb levels 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, the preferential signaling pathways for T cells and occurred with concomitant increases in
adenylate cyclase 8 and phosphatidylinositol 3-kinase, indicating that Pb may promote Th2 activity via T
cell-independent pathways.
Epidemiologic studies provide evidence for blood Pb levels being associated with a shift in
production from Thl to Th2 cytokines in humans (Section 5.6.5.4); however, the extant evidence for
effects on T cells in humans is derived largely from older studies describing changes in the abundance of
several T cell subtypes, including CD3+, CD4+, and CD8+, that may affect cell-to-cell interactions
required in acquired immunity responses to antigens. Among studies of subjects without occupational Pb
exposures, decreases in the abundance of CD3+ (Figure 5-48 and Table 5-23), CD4+, and CD8+ T cells
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often were observed among children with blood Pb levels in the range of 5 to 44 (ig/dL (Lutz et al.. 1999;
Sarasua et al.. 2000; Z. Y. Zhao et al.. 2004V However, Karmaus et al. (2005) observed decreases in
several T cell subtypes in children with blood Pb levels 2.2-2.8 (ig/dL compared with children with blood
Pb levels <2.2 (ig/dL, indicating that decreases in T cell abundance also may occur at blood Pb levels that
are more comparable to those of the current U.S. general population. In association with higher blood Pb
levels (> 10 (ig/dL), Zhao et al. (2004). Fischbein et al. (1993). and Mishra et al. (2010) observed a
decrease in the ratio of CD4+/CD8+ cells, indicative of a compromised response to viral infections. These
findings are consistent with the documented effects of Pb exposure on diminished host resistance and
associations with bacterial and viral infections (Section 5.6.4.1). Changes in the CD4+/CD8+ ratio have
not been examined in populations with lower blood Pb levels. In a large multicity U.S. study conducted in
subjects living near a Pb smelting operation plus demographically-matched controls, Sarasua et al. (2000)
indicated that the associations between blood Pb levels and T cells may vary with age. Investigators
analyzed blood Pb level as a continuous variable and found that among children 6-35 months in age, a 1
(ig/dL increase in blood Pb level was associated with decreases in the percentage of CD3+ (-0.18 [95%
CI: -0.34, -0.02]), CD4+ (-0.10 [95% CI: -0.24, 0.04]), and CD8+ (-0.04 [95% CI: -0.15, 0.07]) T cells;
however, in older children (36-71 months, 6-15 years) and adults (16-75 years), many effect estimates
were positive.
Despite the consistency of findings across studies, it is not clear what may be the impact of the
small magnitudes of change in T cell abundance on the cell-to-cell interactions that mediate downstream
acquired immune responses. For example, in populations without occupational Pb exposures, Pb-
associated decreases in the relative abundance of CD3+ cells range between 1 and 9% (Table 5-23).
Larger decreases (20-35%) are observed in studies of occupationally-exposed males with higher blood Pb
levels than those expected in the general population (>25 (.ig/dL) (Fischbein et al.. 1993; Undeger et al..
1996).
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Study Population
Karmaus et al. (2005) Children, ages 7-10yr.
Sarasua et al. (2000) Children, ages 6-35 mo.
Sarasua et al. (2000) Children, ages 36-71 mo.
Sarasua et al. (2000) Children, ages 6-15yr.
Blood Pb Group
Lutz et al. (1999)
Children, ages 9mo.-6 yr.
Zhao et a I. (2004) Children, ages 3-6 yr.
Sarasua et al. (2000) Subjects, ages 16-75 yr.
Pinkerton et al. (1998) Controls
Pb workers
Fischbein et al. (1993) Controls
Pb workers
Pb workers
Underger et al. (1996) Controls
Pb workers
<2.2
2.2-2.83
2.84-3.4
>3.4
0.6-4.9
5-9.9
10-14.9
>15
0.6-4.9
5-9.9
10-14.9
>15
0.6-4.9
5-9.9
10-14.9
>15
<10
10-14
15-19
20-44
<10
>10
0.6-4.9
5-9.9
10-14.9
>15
<25
>25
16.7
74.8
0.2 0.4 0.6 0.8 1
Relative CD3+cell abundance (normalized to lowest blood Pb group)
1.2
Note: Bars represent the abundance of CD3+ cells normalized to the level measured in the lowest
blood Pb group (depicted as black or dark blue). Bars in black or gray represent results in subjects
without occupational Pb exposures, and bars in dark or light blue represent results in subjects with
occupational Pb exposures.
Figure 5-48. Comparisons of the relative abundance of CD3+ T cells among
groups with increasing blood Pb level (|jg/dL).
Table 5-23. Comparison of serum abundance of T-cell subtypes3 among various blood Pb
groups.
Blood Pb
Study
Population
group
CD3+
CD4+
CD8+
CD4+:CD8+ CD45RO+ CD45RA+
(pg/dl)
Children
Karmaus et
331 children, ages 7-10
<2.2
2118
1214
712
358
al. (2005)"
yr
2.21-2.83
1919
1106
634
321
Hesse, Germany
2.84-3.41
1979
1128
661
348
>3.41
2184
1123
662
351
Sarasua et
382 children, ages 6-30
0.6-4.9
67.8
46.6
19.9
al. (2000)°
mo
5-9.9
66.6
45.5
20.2
Multiple U.S. locations
10-14.9
67.1
47.1
19.1
> 15
65.1
45.2
18.8
Sarasua et
562 children, ages 36-71
0.6-4.9
68.1
42.0
23.5
al. (2000)°
mo
5-9.9
69.4
44.0
23.1
Multiple U.S. locations
10-14.9
68.6
44.2
21.4
> 15
76.1
41.6
24.4
Sarasua et
675 children ages 5-16 vr 0.6-4.9
69.2
43.0
24.7
al. (2000)°
Multiple U.S. locations
5-9.9
69.5
43.0
25.4
10-14.9
71.1
44.5
24.5
> 15
66.3
44.0
20.1
Lutz etal.
279 children, ages 9 mo-
<10
68.4
(1999)d
6yr
10-14
69.7
Springfield, MO
15-19
64.6
20-44
65.3
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Study
Population
Blood Pb
group
(pg/dl)
CD3+
CD4+
CD8+
o
D
+
o
D
00
+
CD45RO+
CD45RA+
Zhao et al.
(2004)
75 children, ages 3-6 yr <10
Zhejiang Province, China > 10
55.2
54.6
27.1
23.7e
20.6
23.2e
1.41
1.09e
Adults without Occupational Pb Exposures
Sarasua et
al. (2000)°
433 children and adults,
ages 16-75 yr
Multiple U.S. locations
0.6-4.9
5-9.9
10-14.9
> 15
71.6
73.4
78.5
70.9
50.3
53.3
54.9
49.6
24.9
23.9
24.1
25.3
Adults with Occupational Pb Exposures
Fischbein et
al. (1993)
36 unexposed controls,
mean age 47 yr
51 firearms instructors,
mean age 48 yr
NR
<25
>15
72.6
57
46.9e
45.6
29.6
21.4e
25.7
23
25.4
1.95
1.38
0.95e
Pinkerton et
al. (1998)'
84 unexposed controls,
mean age 30 yr
145 Pb smelter workers,
mean age 33 yr
<2
39
73.1
74.9
44.7
44.6
23.0
23.3
1.9
1.9
45.6
44.1
Underger et
al. (1996V1
25 unexposed controls,
ages 22-56 yr
25 Pb battery plant
workers, ages 22-55 yr
16.7
74.8
2044.7
1644
1140.3
858.8e
977.6
829.1
1.3
1.1
Yucesoy et
al. (1997b)
10 unexposed controls,
ages 25-42 yr
47 Pb battery plant
workers, ages 19-49 yr
4.0
59.4, 58.4
30.8
30.1
Mishra et al.
(2010)
21 unexposed controls,
median age 27 yr
26 three-wheel drivers,
median age 31 yr
33 Pb battery workers,
median age 27 yr
4.5
6.7
132 (103)
55
37
31e
26
27
26
2.6
1.4
1.3e
44
43
40
61
73
70e
aData are presented as the percentage of T cell subtype among all T cells unless otherwise specified.
bData represent the number of cells/[jL serum. Means are adjusted for age, sex, ETS, number of infections in the previous 12 months, serum lipid concentration, and
organochlorine exposures.
cMeans are adjusted for age, sex, and location of study.
dMeans are adjusted for age.
ep<0.05 for difference among groups.
'Means adjusted for age, race, current smoking status, and workshift.
9Data represent the number of cells/[jL serum.
5.6.2.2. Lymphocyte Activation
Pb exposure produces an expansion of alloreactive T and B lymphocytes. This occurs as a result of
reversing the normal suppression that is mediated by a macrophage-like subpopulation. As discussed in
Section 5.6.5.2, changes in NO production appear to be involved in this process (Farrer et al.. 2008).
Additionally, Pb alters antigen processing that occurs in antigen presenting cells (APCs, e.g., primarily
dendritic cells and macrophages), which appear to shift signals to T cells skewing both the nature of the
subsequent response and the spectrum of activities among those expanded populations of lymphocytes
(Farrer et al.. 2005). Gao et al. (2007) reported that dendritic cells that matured in the presence of Pb
promoted enhanced alloreactive T cell proliferation compared to control dendritic cells. An additional
effect of Pb has been described relative to activation of T cells. Using the local lymph node assay
(LLNA), Carey et al. (2006) found that PbCl2 was able to provide a costimulatory signal to antigens that
could activate T cells. The exact mechanistic basis for this is not known.
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Epidemiologic findings are mostly limited to studies of occupationally-exposed adults and to
associations with blood Pb level >10 (ig/dL. In the only study of children in Missouri, the percentage of
activated T cells (as indicated by the cell surface marker HLA-DR) was higher in children with blood Pb
level 15-19 (ig/dL than in children with blood Pb level <15 (ig/dL; however, activated cells were not
elevated in children with blood Pb levels 20-44 (ig/dL (Lutz et al.. 1999). In contrast with toxicological
findings, the direction of lymphocyte proliferation (in response to mitogens and/or to specific antigens) in
association with occupational Pb exposure is unclear. Pb promotes the activation of Th2 cells and
suppresses Thl cells, therefore, the differential activation of specific subtypes may be not discernable in
studies that measure overall lymphocyte proliferation. Whereas some studies have reported similar levels
of lymphocyte proliferation (< 1% difference) between Pb-exposed workers and unexposed controls (N.
Cohen et al. 1989; Oueiroz. Perlingeiro. et al.. 1994). others have reported lower lymphocyte
proliferation among Pb-exposed workers (8-25%) (Alomran & Shleamoon. 1988; Fischbein et al.. 1993;
kimber et al.. 1986; Mishraet al.. 2003). With the exception of Fischbein et al. (1993). these latter studies
included subjects with high blood Pb levels (>60 j^ig/dL). Additionally, with much of the evidence limited
to comparisons of mean proliferative responses between Pb-exposed and -unexposed workers, it is
difficult to apply findings to populations with lower blood Pb levels. In one of the few analyses of
correlation between blood Pb levels and lymphocyte proliferation, Mishra et al. ("2003) found a lack of
correlation, suggesting the influence of another occupationally-related factor on lymphocyte proliferation.
5.6.2.3. Delayed-type Hypersensitivity
The DTH assay is commonly used as an indicator of 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 directed against the antigen (sensitizing phase)
and the recruitment of antigen-specific T cells and monocytes to the site of antigen deposition (elicitation
phase). In the 2006 Pb AQCD and several recent reviews, a suppressed DTH response was identified as
one of the most consistently observed and well-established immunomodulatory effects of Pb exposure in
animal models (Dietert & McCabe. 2007; Mishra. 2009; U.S. EPA. 2006). A majority of the evidence for
Pb suppressing the DTH response is provided by the historical literature in which effects were observed in
association with both gestational (Bunn. Ladies, et al.. 2001; Bunn. Parsons, et al. 2001a. 2001b; S. Chen
et al.. 2004; S. Chen et al.. 1999; Faith et al.. 1979; J.-E. Lee et al.. 2001; Miller et al.. 1998) and postnatal
(Laschi-Loquerie et al.. 1984; M. J. McCabe. Jr. et al.. 1999; Muller et al.. 1977) Pb exposures of animals.
In some studies, the suppressed DTH response was accompanied by a decreased production of IFN-y (S.
Chen et al.. 1999; J.-E. Lee et al.. 2001). which is the primary cytokine that stimulates recruitment of
macrophages, a key component of the DTH response. The concomitant decrease in IFN-y demonstrated
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further that Pb-induced suppression of the DTH response reflects the inhibition of Thl functional
activities.
One of the most salient findings collectively was that DTH was suppressed in animals with blood
Pb levels ranging from <2 to 5 (.ig/dL (Bunn. Ladies, et al.. 2001; Bunn. Parsons, et al.. 2001a. 2001b;
Miller et al.. 1998; Mulleretal.. 1977). In some studies that examined Pb exposures at different stages of
gestation, exposures later in gestation suppressed DTH (Bunn. Parsons, et al.. 2001b; J.-E. 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 mice. The Pb-exposed
dendritic cells inhibited the OVA-specific DTH footpad response in mice compared with mice exposed to
control dendritic cells.
The capacity 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). Further, coherence is provided by
associations observed between Pb exposure or blood Pb levels and other responses related to the
inhibition of Thl-driven adaptive immune responses, including increased susceptibility to developing
certain infections and tumors (Section 5.6.4.1).
5.6.2.4. Macrophages and Monocytes
Macrophages and monocytes, the blood form of tissue macrophages, are among the most sensitive
targets of Pb-induced immune effects. The 2006 Pb AQCD emphasized the large number of toxicological
studies showing the effects of Pb on a wide range of alterations in macrophage function to promote a
hyperinflammatory phenotype (U.S. EPA. 2006). These changes include enhanced production of ROS,
suppressed production of NO, enhanced production of TNF-a, excessive metabolism of arachidonic acid
into immunosuppressive metabolites (e.g., prostaglandin E2), impaired phagocytic activity and lysosomal
function, and potentially altered receptor expression [e.g., toll-like receptors]). These studies are
described in detail in Section 5.6.5.2. Because macrophages are major resident populations in most tissues
and organs and are also 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 and organs (Section 5.6.4.5). Recent studies by Bussolaro et al. (2008). Kasten-Jolly et al. (2010).
and Mishra et al. (2006) using mouse macrophages exposed in vitro to Pb reinforced the capacity of Pb to
induce a broad spectrum of functional alterations in macrophages such as lipopolysaccharide (LPS)-
induced production of NO. Although not examined widely, epidemiologic evidence suggests that Pb
exposure may be associated with changes in macrophage function in humans. Pineda-Zavaleta et al.
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(2004) examined children in Lagunera, Mexico, attending schools at varying distances from an active Pb
smelter (range of blood Pb levels: 3.5-47.5 j^ig/dL) and found that blood Pb levels were associated with a
hyperinflammatory state in macrophages as indicated by a decrease in NO production and increase in
superoxide anion production (Section 5.6.5.2). Studies of occupationally-exposed adults mostly observe
that Pb exposure is associated with impaired function of macrophages. Adjusting for age, race, smoking,
and workshift, Pinkerton et al. (1998) found lower abundance of monocytes among Pb smelter workers
(7.8%) than among unexposed controls (8.5%). Fischbein et al. (1993) found lower abundance of HLA-
DR+ cells in firearms instructors with blood Pb level <25 (ig/dL (8.8%) and > 25 (ig/dL (8.7%) than
among unexposed controls (15.2%). HLA-DR+ is an indicator of activated function state of APCs and is
upregulated in response to cell signaling.
5.6.2.5. Neutrophils
Although neutrophils were found not to be a significant direct target of Pb in the 2006 Pb AQCD
(U.S. EPA. 2006). the modulation of their activity by Pb may have important consequences on the
dysregulation of inflammation and ability to response to infectious agents. Studies of cultured human
polymorphonuclear cells (PMNs) (Govema et al.. 1987) and occupationally-exposed adults (Bergeret et
al.. 1990; Queiroz et al.. 1993; Queiroz. Costa, et al.. 1994; Valentino etal.. 1991) have found reduced
PMN functionality, as indicated by reduced chemotactic response, phagocytic activity, respiratory
oxidative burst activity, or reduced ability to kill ingested antigen, among Pb workers compared with
controls. Important limitations to applying these epidemiologic findings broadly include male-only study
populations, relatively high blood Pb levels of workers (range of mean levels: 33.1-71 (ig/dL) and the
lack of direct examination of associations between blood Pb level and neutrophil function.
In both studies of animals and occupationally-exposed adults, Pb exposure has been associated
with an increase in neutrophil counts, which has been interpreted as a compensatory response to Pb-
induced impairment in neutrophil chemotactic activity and a hyperinflammatory response. In a study by
Kibayashi et al. (2010) to investigate host responses to gunshot wounds in the brain, neutrophils were a
major responding cell. Implantation of Pb spheres (compared with glass spheres in the controls) in male
Wistar rats led to major neutrophil infiltration with inflammatory-related damage that included apoptosis
and indications of neurodegeneration.
In a group of 68 ceramic, Pb recycling, or bullet manufacturing workers and 50 controls selected
among food plant workers, DiLorenzo et al. (2006) observed that a 1 (ig/dL increase in blood Pb level
was associated with an increase in ANC of 21.8 cc 11 s/j^lL (95% CI: 11.2, 32.4 eelIs/j^lL) adjusted for age,
BMI, and smoking status. The geometric mean (range) of blood Pb levels was 20.5 (ig/dL (3.2-120)
among workers and 3.5 (ig/dL (1-11) among controls. Eight workers with medium to high Pb exposures
(exact blood Pb levels not reported) had neutrophilia (n >7,500 cells/mm3) versus no controls, suggesting
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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 j^ig/dL, and workers with blood Pb levels >30 (ig/dL, ANC was observed to increase
monotonically, supporting a blood Pb dose-dependent relationship. When the three blood Pb groups were
further stratified by current smoking, two-way ANOVA indicated positive interaction between blood Pb
level and current smoking. Higher blood Pb level was associated with higher ANC only when current
smokers were compared. Among nonsmokers, ANCs were similar across blood Pb groups.
Coherence 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 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
complement C3 protein among Pb-exposed workers compared with unexposed controls, with Ewers et al.
(1982) additionally observing an inverse association between blood Pb level and C3 among workers.
However, the implications of these findings are limited due to the high blood Pb levels in these
occupationally-exposed groups (range of blood Pb levels: 18.6-85.2 ng/dL and 38-100 ng/dL,
respectively) and by the lack of adjustment for potential confounding variables.
5.6.2.6. Dendritic Cells
Since the 2006 Pb AQCD (U.S. EPA. 2006). new evidence from both an ex vivo and in vitro model
suggests 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 vs. Th2 T-
cell populations and cytokine profiles. But new research techniques (D. Gao & Lawrence. 2010) have
provided 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. They found that Pb (25 (.iM. 10 days) altered 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 control
cells but the same amount of IL-10 (inhibits production of Thl cytokines). The effect of Pb in altering the
cytokine 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 towards a Th2 phenotype. Strengthening the role of
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dendritic cells in mediating Pb immune effects were ex vivo results from the same study that Pb-nai've
Balb/c mice implanted with Pb-treated dendritic cells were skewed towards Th2 activity as indicated by
inhibited DTH (Section 5.6.2.3) and IgG2a antibody (Section 5.6.3) responses (D. Gao et al.. 2007V
5.6.2.7. Natural Killer (NK) Cells
The collective body of toxicological and epidemiologic evidence indicates that the innate immune
NK cells are not affected to a large extent by Pb exposure. This conclusion is underscored by
epidemiologic observations that blood Pb levels are associated with T or B cell abudance but are not
associated with NK cell abundance or the level of functional activity (Karmaus et al.. 2005; Pinkerton et
al.. 1998; Sarasua et al.. 2000). Likewise, similar means ofNK cell abundance or functional activity have
been observed in Pb-exposed workers and unexposed controls (Fischbein et al.. 1993; Kimber et al..
1986; Mishra et al.. 2003; Pinkerton et al.. 1998; Undeger et al.. 1996; Yucesov et al.. 1997b). Consistent
with epidemiologic findings, in a recent in vitro study comparing the toxicity of metals for different
populations of immune cells, Fortier et al. (2008) found that PbCl2 (7.5-20.7 j^ig/dL), did not significantly
affect NK cytotoxicity compared with the DMSO vehicle; however, PbCl2 did not affect other immune
parameters (e.g., monocyte phagocytic activity or lymphocyte proliferation as well.
5.6.3. Humoral Immunity
The 2006 Pb AQCD indicated that Pb exposure was associated with enhanced humoral immune
responses as characterized by the proliferation of B cells and increased production of Ig antibodies (U.S.
EPA. 2006). Both toxicological and epidemiologic studies (Figure 5-49 and Table 5-24) have
demonstrated Pb-associated increases in IgE production, which is strongly implicated in mediating
allergic responses and inflammation in allergic asthma. In animal studies, Pb exposures induced
concomitant increases in IgE and IL-4 production by T cells (S. Chen et al.. 1999; J. E. Snvder et al..
2000). consistent with the hypothesis that Th2-mediated mechanisms can induce class switching of B
cells for the production of IgE. Additional coherence and biological plausibility for Pb effects on humoral
immunity have been provided by epidemiologic observations for increases in B cell abundance in
association with increasing blood Pb level or occupational Pb exposure (Figure 5-50 and Table 5-24).
Earlier toxicological and epidemiologic studies also found similar associations of Pb with increases in
other classes of Igs including IgG, IgM, and IgA.
Recent toxicological evidence continues to support the role ofT cell-mediated mechanisms in Pb-
induced activation of B cells and production of Ig antibodies. 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-ovalabumin (TNP-OVA)] hapten-protein conjugate with or without co-exposures to PbCl2. Seven
days later, they examined the effects of PbCl2 on the LLNA response to TNP-Ficoll or TNP-OVA. PbCl2
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exposure (25-50 jj.g, injected) increased the numbers of T and B cells in the lymph node against both
TNP-Ficoll and TNP-OVA. Further, in a dose-dependent manner, PbCl2 induced statistically significant
elevations of IgM, IgG2a, and IgGl antibody 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 seeing 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.
Other animal studies provided strong support for Pb exposure stimulating humoral responses
preferentially via Th2-mediated mechanisms. Gao et al. (2006) found that Pb exposure (50 jj.g, i.p.,
3 times/week, 3 weeks) of Balb/c mice elevated IgGl over that ofThl-driven IgG2a. Similar results were
reported when Pb-exposed dendritic cells were used to initiate antibody production (D. Gao & Lawrence.
2010). In a highly-specialized strain of IFN-y knockout mice (lacking the capacity to produce IFN-y), Pb
exposure had the reverse effect of increasing the IgG2a/IgGl ratio. These results were 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 (D. Gao et al. 2006). In a microarray study examining the DIT of Pb in Balb/c mice, Kasten-
Jolly et al. (2010) found that early-life Pb exposure (O.lmM 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, associations between increasing blood Pb level and increasing serum IgE
level are demonstrated primarily in children (Annesi-Maesano et al.. 2003; Karmaus et al.. 2005; Lutz et
al.. 1999; L. Sun et al.. 2003). Most studies indicate that the association between blood Pb level and IgE is
nonmonotonic (Figure 5-49 and Table 5-24). Whereas many studies were focused on examining
differences between children with blood Pb levels below and above 10 (ig/dL (Figure 5-49 and Table 5-
24), Karmaus et al. (2005) demonstrated 28% increases in IgE among children with blood Pb levels 2.84-
3.41 and >3.4 (ig/dL compared with children with blood Pb level <2.2 (ig/dL. Another strength of this
study was the adjustment for potential confounding by various organochlorine compounds, age, number
of infections in the previous 12 months, and serum lipids. A clear blood Pb dose-dependent relationship
was not observed as the mean IgE was lower among children in the second quartile of blood Pb levels
(2.2-2.8 (ig/dL) than among children in the lowest quartile (<2.2 j^ig/dL). Annesi-Maesano et al. (2003)
provided additional information on potential critical developmental windows of Pb exposure. Cord blood
IgE was associated with infant hair Pb level but not with cord or placental Pb level. From these findings,
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the authors inferred a stronger effect of Pb exposure integrated over the entire gestational period rather
than exposures closer to birth. Additionally, the magnitude of association was larger in the subgroup with
nonallergic mothers, pointing to the possible stronger effect of family history of allergy on IgE levels at
birth compared with blood Pb level. Sarasua et al. (2000) was unique among studies of children in
examining associations of blood Pb level with IgA, IgG, and IgM. Associations were consistently positive
(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 increase in IgA, IgG,
and IgM, respectively, per 1 (ig/dL increase in blood Pb level, adjusted for age, sex, and location) and
tended to be larger in magnitude in the youngest age group (6-35 months), suggesting an increased
susceptibility of exposure at younger ages. In this study, associations with IgE were not examined.
In studies of adults, one without (Pizent et al.. 2008) and one with occupational Pb exposure (Heo
et al.. 2004). investigators reported higher IgE levels in association with higher blood Pb levels. In a study
of urban adults in Zagreb, Croatia with blood Pb levels between 0.56 and 7.4 (ig/dL, a statistically
significant, positive association between blood Pb level and 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 increase in blood Pb level was associated with a 0.60 increase in log IgE (95%
CI: 0.58, 1.18). Investigators did not report an effect estimate in men because it did not attain statistical
significance. The study included 166 women and 50 men, thus, it was difficult to ascertain whether there
was suggestion of association in men but lack of power to indicate statistical significance. Although blood
Pb levels were lower in women (mean: 2.16 j^ig/dL, range 0.56-7.35 (ig/dL) than in men (mean: 3.17
(ig/dL, range 0.99-7.23), both groups had levels similar to those reported in studies of children. Among
battery manufacturing workers (mostly males), Heo et al. (2004) found a monotonic increase in IgE levels
among workers with blood Pb levels <0, 10-29, and > 30 (ig/dL.
A majority of the epidemiologic evidence for the effects of Pb on humoral immunity comprises
comparisons of serum IgA, IgG, and IgM levels between Pb-exposed and -unexposed workers (Alomran
& Shleamoon. 1988; Anetor & Adenivi. 1998; Ewers et al.. 1982; kimber et al.. 1986; Pinkerton et al .
1998; Oueiroz. Perlingeiro. et al.. 1994; Sarasua et al.. 2000; Undeger et al. 1996). In contrast with
toxicological findings, epidemiologic evidence reflects more mixed associations, with studies reporting
higher, lower, and similar Ig levels in Pb-exposed workers compared with -unexposed controls. Among
studies reporting lower Ig levels in Pb-exposed workers were a few with high mean blood Pb levels
among exposed workers (56.3 and 74.8 (.ig/dL) (Anetor & Adenivi. 1998; Undeger et al.. 1996).
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Study
Karmaus et al.
(2005)
Population
Children, ages 9 mo.-6yr.
Sun et al. (2003) Children, ages 3-6 yr.
Lutz et al. (1999) Children, ages 9 mo.-6 yr.
Heo et al (2004) Pb-exposed workers
Blood Pb Group
2.2-2.83
2.84-3.4
10-14
15-19
20-44
10-29
1 2 3 4 5
Relative IgE concentration (normalized to lowest blood Pb group)
Note: Bars represent the IgE normalized to the level measured in the lowest blood Pb group
(depicted as black).
Figure 5-49. Comparison of IgE levels among groups with increasing blood
Pb level (pg/dL).
Study Population
Karmaus et al. (2005) Children, ages 7-10 yr.
Sarasua et al. (2000) Children, ages 6-35 mo.
Sarasua et al. (2000) Children, ages 6-15yr.
Zhao et al. (2004)
Children, ages 3-6 yr.
Sarasua et al. (2000) Subjects, ages 16-75 yr.
Pinkerton et al. (1998) Controls
Pb workers
Fischbein et al. (1993) Controls
Pb workers
Pb workers
Underger et al. (1996) Controls
Pb workers
Blood Pb Group
<2.2
2.2-2.83
2.84-3.4
>3.4
0.6-4.9
5-9.9
10-14.9
>15
Sarasua et al. (2000) Children, ages 36-71 mo. 0.6-4.9
5-9.9
10-14.9
>15
0.6-4.9
5-9.9
10-14.9
>15
<10
>10
0.6-4.9
5-9.9
10-14.9
>15
<2
39
NR
<25
>25
16.7
74.8
0.6 0.7 0.8 0.9 1 1.1
Relative B cell abundance (normalized to lowest blood Pb group)
Note: Bars represent the abundance of B cells normalized to the level measured in the lowest blood Pb group
(depicted as black or dark blue). Bars in black or gray represent results in subjects without occupational Pb
exposures, and bars in dark or light blue represent results subjects with occupational Pb exposures.
Figure 5-50. Comparison of the relative abundance of B cells among
groups with increasing blood Pb level (pg/dL).
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Table 5-24. Additional characteristics and quantitative data3 for results presented in
Figures 5-49 and 5-50
Study
Population
Blood Pb
group (|jg/dL)
igE
igG
igM
igA
B cells
Children
Karmaus et al.
(2005)"
331 children, ages 7-1 Oyr
Hesse, Germany
<2.2
2.21-2.83
2.84-3.41
>3.41
46
30
59
59°
1210
1214
1241
1201
150
143
153
148
123
121
133
136
418
353
389
393
Sun et al.
(2003)
Zeijiang Province, China
217 children, ages 3-6 yr.
<9.9
>9.9
Lutz etal.
(1999)
279 children, ages 9 mo-6 yr
Springfield, MO
<10
10-14
15-19
20-44
51.8
74.0
210.7
63.7d
13.4
12.6
16.9
11.1
Sarasua etal.
(2000)e
382 children, ages 6-30 mo
Multiple U.S. locations
0.6-4.9
5-9.9
10-14.9
>15
609
666f
680f
630
103
108
105
124f
50.1
55.0
58.2
61,4f
19.1
20
20.4
22.2
Sarasua etal.
(2000)e
562 children, ages 36-71 mo
Multiple U.S. locations
0.6-4.9
5-9.9
10-14.9
>15
817
813
856
835
120
116
125
121
88.6
90.9
96.3
94.1
18.4
17.6
19.2
18.6
Sarasua etal.
(2000)e
675 children ages 5-16 yr
Multiple U.S. locations
0.6-4.9
5-9.9
10-14.9
>15
1031
1094f
1048
1221
128
131
136
106
140
143
140
108
16.1
15.8
15.3
20.1
Zhao et al.
(2004)
75 children, ages 3-6 yr
Zhejiang Province, China
<10
>10
16.58
16.82
Adults without Occupational Pb Exposures
Sarasua etal.
(2000)e
433 children and adults,
ages 16-75 yr
Multiple U.S. locations
0.6-4.9
5-9.9
10-14.9
>15
1099
1085
1231
1169
175
175
262f
139
252
242
283
193
13.9
13.0
12.4
14.8
Adults with Occupational Pb Exposures
Heo et al.
(2004)
Korea
606 Pb battery plant workers
<10
10-29
>30
112.5
223.3
535,89
Pinkerton etal.
(1998)h
84 unexposed controls,
mean age 30 yr
145 Pb smelter workers,
mean age 33 yr
<2
39
1090
1110
94.5
106.2
180
202
14.6
13.2
Fischbein et al.
(1993)
36 unexposed controls,
mean age 47 yr
51 firearms instructors,
mean age 48 yr
NR
<25
>15
8.6
10.5
11.2°
Undergeret al.
(1996)
25 unexposed controls,
ages 22-56 yr
25 Pb battery plant workers,
ages 22-55 yr
16.7
74.8
1202.1
854.69
140.4
93.39
210.3
168.1
635.9'
545.5
algE data are presented as lU/mL. Other Ig data are presented as mg/dL. B cell data are presented as the percentage of B cells among all lymphocytes unless otherwise
specified.
bAII means are adjusted for age, sex, ETS, number of infections in the previous 12 months, serum lipid concentration, and organochlorine exposures. B cell data represent the
number of cells/[jL serum.
cp<0.05 for difference among groups.
dp<0.05 for differences among groups adjusted for age.
eMeans are adjusted for age, sex, and location of study.
fps0.05 in comparison with mean in lowest blood Pb group.
9p<0.05 for differences among groups.
hMeans adjusted for age, race, current smoking status, and workshift.
'Data represent the number of cells/[jL serum.
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5.6.4. Immune-based Diseases
5.6.4.1. Host Resistance
The capacity of Pb to reduce host resistance to bacteria has been known for almost 40 years and
was supported by several toxicological studies described in the 2006 Pb AQCD (U.S. EPA. 2006).
Biological plausibility has been provided by observations of Pb affecting mechanisms underlying
diminished host resistance, e.g., decreased capacity for Thl-driven acquired immune antiviral responses
(indicated by the overproduction of PGE2) and increased inflammatory responses in target tissue resulting
in further damage to host protective barriers. Gupta et al. ("2002) demonstrated that elevated Pb exposure
of mice (>125 mg/kg Pb, 13.0 (ig/dL blood Pb) reduced host resistance to viral infections as indicated by
an increased viral titre and increased mortality. Host resistance to bacteria such as Listeria requires
effective Thl-driven responses including the production of IL-12 and IFN-y (Lara-Tei ero & Pamer. 2004).
but these are suppressed by Pb (D. Gao et al.. 2007). The lack of IFN-y can inhibit appropriate and timely
macrophage activation. However, beyond this, suppression of NO production and along with it, the
microbicidal metabolite peroxynitrite, can compromise host resistance to some bacteria (U.S. EPA. 2006).
Recently, mechanisms through which Pb impacts both innate immune cells and natural host defense
barriers to increase the likelihood of serious bacterial infections have been delineated further. Kasten-Jolly
et al. (2010) showed that developmental exposure of mice to Pb 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).
With limited investigation, the effect of Pb on host resistance to parasitic agents is uncertain. The
2006 Pb ACQD described one study in which Pb-exposed (30-100 mM) mouse macrophages had
diminished ability to kill Leishmania enrietti parasites (Mauel et al.. 1989); however, given the well-
established effect of Pb in promoting Th2 activity, it is plausible that Pb could enhance host resistance to
parasites that require robust Th2 responses (e.g., helminths) (U.S. EPA. 2006). In a recent study, Pb
enhanced host resistance to malaria (Koka et al.. 2007). However, this was related to the capacity of Pb to
induce eryptosis and the rapid removal of malaria-infected erythrocytes and not to Pb-induced alterations
in immune function.
The collective body of epidemiologic data is sparse; however, several studies have found an
association between blood Pb levels and respiratory infections. In a Boston-area study of 1978 children
(ages 4-8 years), Rabinowitz et al. (1990) reported that compared with children with blood Pb levels <10
(ig/dL, children with blood Pb levels >10 (ig/dL were more likely to have parental report "other
respiratory infections" (not defined by authors) (OR: 1.5 [95% CI: 1.0, 2.3]), severe ear infections (OR:
1.2 [95% CI: 1.0, 1.4]), and illnesses other than cold or influenza (OR: 1.3 [95% CI: 1.0, 1.5]) (Figure 5-
51 and Table 5-25). Analyses did not adjust for any potential confounders. Likewise, without considering
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confounders, Karmaus et al. (2005) found that children with more than 10 infections in the previous 12
months had higher blood Pb levels (mean: 3.3 (ig/dL) compared with children with 1 to <5 infections
(mean: 2.8 (ig/dL) or 5-10 infections (mean: 2.6 (ig/dL) in the same time period. A unique study was
conducted in children in Cordoba, Argentina, in which Pb exposure was assessed by measuring Pb in total
deposition samples and in lichen biomonitors from sites near 4 city health clinics (Carreras et al.. 2009V
Lichen has been recognized for the uptake, accumulation, and sequestration of environmental chemicals
from the air. Pb content in lichen and colocated total deposition samples were highly correlated,
indicating that the lichen was a suitable indicator of environmental concentrations. In an ecologic analysis
that did not consider potential confounders, clinics near sites with higher levels of Pb in total deposition
and lichen samples had higher frequency of visits by children for pharyngitis, tonsilitis, and laryngitis.
Because other metals, including manganese, iron, copper, and nickel were also associated with respiratory
illnesses, it was difficult to characterize the independent effects of Pb. Similar to studies in children,
among adults, frequency of self-reported colds or influenza was greater among Pb battery or smelter plant
workers (28.8%) than among unexposed controls (16.1%); however, a statistical analysis was not
performed on the data (Ewers et al.. 1982). Thus, while several studies in humans indicate associations
between indicators of Pb exposure and infectious illnesses, they are limited by weak analytic methods and
lack of consideration for potential confounders. Evaluation of the relationship between Pb exposure and
host resistance in humans would be improved by more longitudinal investigation and rather than group
comparisons, analyses of associations between blood Pb levels and viral or bacterial infections.
5.6.4.2. Asthma and Allergy
Toxicological studies and to varying degrees, epidemiologic studies, have demonstrated Pb effects
on multiple immunological pathways, including elevated production of Th2 cytokines such as IL-4,
increased IgE antibody production (Figure 5-49), and Pb-induced inflammatory cell dysfunction. These
are well-recognized pathways in the development of allergy and allergic disease, including asthma. It has
been suggested that low exposure to Pb exerts immunostimulating effects in contrast to higher exposure,
which has been implicated in suppressing immune function (Mishra et al.. 2006). Coherent with the
mechanistic evidence, several epidemiologic studies indicate associations between blood Pb levels and
asthma and allergy (Figure 5-51 and Table 5-25). In univariate analyses of children near Boston, MA,
Rabinowitz et al. (1990) reported a positive association between high blood Pb level (>10 (ig/dL) and
asthma (RR: 1.3 [95% CI: 0.8, 2.0]) but not eczema (RR: 1.0 [95% CI: 0.6, 1.6]).
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Reference
Population
Outcome
Blood Pb Level Mean or Group
Rabinowitz et al. (1990) Children
Rabinowitz et al. (1990") Children
Jedrychowski etal. (2011") Children
Pizent et al. (2008") Adult males
Rabinowitz et al. (1990) Children
Joseph et al. ("2005") Children, Caucasian
Other respiratory infection ^10 vs. <10
Severe ear infection
Children, African American medical care
Children, African American
Children, African American
Children, African American
Children, African American
Prenatal: 1.16
Concurrent: 2.02
Prevalent asthma slOvs. <10
Incident asthma requiring s5 vs. Caucasian<5
vs. African American<5
Min etal. (2004)
Adults
BR
£10 vs. African American<5
<5 vs. Caucasian<5
s5 vs. Caucasian<5
£10 vs. Caucasian<5
2.9
f
0.0 1.0 2.0 3.0
Odds ratio (95% CI)
Note: Odds ratios are are standardized to a 1 |jg/dl_ increase in blood Pb level or 1 unit in log of blood Pb. SPT = skin prick test, BR = bronchial responsiveness.
Figure 5-51. Associations of blood Pb levels with immune-based
conditions.
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Table 5-25. Additional characteristics and quantitative data for results presented in
Figure 5-51.
Reference
Population
Blood Pb level mean or group
(pg/dL)
Outcome
Odds
ratio
(95% Cl)a
Rabinowitz etal.
(1990)
Boston area, MA
1768 children
>10 vs. <10°
Other respiratory infection
Severe ear infection
1.5 (1.0,2.3)
1.2 (1.0,1.4)
Rabinowitz etal.
(1990)
Boston area, MA
1768 children.
>10 vs. <10°
Eczema
1.0 (0.6,1.6)
Jedrychowski et al.
(2011)
Krakow, Poland
224 children followed
prenatally
to age 5yr
Prenatal: 1.16
Concurrent: 2.02
Positive SPT
2.3 (1.1,4.6)
1.1 (0.7,1.6)
Pizent et al. (2008)
Zagreb, Croatia
50 males, ages 21 -67 yr
21.6
Positive SPT
0.92 (0.86,
0.98)
Rabinowitz etal.
(1990)
Boston area, MA
1768 children.
>10 vs. <10°
Prevalent asthma
1.3(0.8,2.0)
Joseoh etal. (2005)
Southeastern Ml
4634 children followed
prospectively from age 1 -3 yr
Caucasian >5 vs. Caucasian <5C
African American >5 vs. African American
<5d
African American >10 vs. African
American <5d
African American <5 vs. Caucasian <5°
African American >5 vs. Caucasian <5°
African American >10 vs. Caucasian <5°
Incident asthma requiring
medical care
2.7(0.9,8.1)
1.1 (0.8,1.7)
1.3(0.6,2.6)
1.8 (1.3,2.4)
1.5 (1.2,1.8)
3.0 (1.2,7.1)
Zhao et al. (2004)
75 children, ages 3-6 yr
Zhejiang Province, China
<10
> 10
BR
1.02 (1.00,
1.03)
SPT = skin prick test, BR = bronchial responsiveness.
aOdds ratios are standardized to a 1 |jg/dl_ increase in blood Pb level, except in studies in which Blood Pb is analyzed as a categorical variable.
bOdds ratio in children with blood Pb level s 10 [jg/dl_ with children with blood Pb level <10 |jg/dl_ as the reference group. No additional covariates included in model.
cRelative risk in each specified subgroup with Caucasian children with blood Pb level <5 |jg/dl_ as the reference group. Model covariates include sex, birth weight,
and annual income.
dRelative risk in each specified subgroup with African American children with blood Pb level <5 |jg/dl_ as the reference group. Model covariates include sex, birth weight, and
annual income.
Positive associations between blood Pb level and asthma- and allergy-related conditions also were
observed in large studies with multivariate analyses. An additional common strength of these studies was
the prospective follow-up of subjects that allowed investigators to establish temporality between the
measurement of blood Pb level and onset of disease. In a study of 4,634 children (ages 1-3 years) in
southeastern Michigan 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 most strongly among Caucasian children (relative risk [RR]:
2.7 [95% CI: 0.9, 8.1]) (Joseph et al.. 2005) (Figure 5-51 and Table 5-25). In analyses restricted to African
American children, blood Pb level was weakly associated with asthma requiring medical care (Figure 5-
51 and Table 5-25). In analyses that used Caucasian children with blood Pb level <5 (ig/dL as the
reference group, blood Pb level was associated with statistically significant increases in the risk of asthma
among African children in all blood Pb level categories, which indicated a racial/ethnic effect rather than
a Pb effect. However, among African American children, the RR was much higher in the > 10 (ig/dL
blood Pb category (RR: 3.0 [95% CI: 1.2, 7.1]) than in the > 5 (ig/dL (RR: 1.5 [95% CI: 1.2, 1.8]) or <5
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(ig/dL blood Pb categories (RR: 1.8 [95% CI: 1.3, 2.4]), which pointed to a possible race/ethnicity by
blood Pb level interaction. These findings should be interpreted with caution due to the small number of
asthmatics requiring medical care in the high blood Pb level categories (5 Caucasian children with blood
Pb > 5 (ig/dL and 9 African American children with blood Pb level >10 j^ig/dL).
While the aforementioned studies examined concurrent blood Pb levels, findings from a recent
prospective birth cohort study demonstrated that prenatal blood Pb levels (cord blood Pb level mean: 1.16
(ig/dL [95% CI: 0.12, 1.22]) was associated with an increased risk of allergic sensitization at age 5 years
(Jedrvchowski et al.. 2011). In models that adjusted for sex, parity, maternal age, maternal education,
maternal atopy, and environmental tobacco smoke exposure, measures of prenatal Pb exposure (cord or
maternal blood Pb) were associated with greater risk of positive skin prick test (SPT) to dust mite, dog, or
cat allergen compared with concurrent blood Pb levels (Figure 5-51 and Table 5-25). The greater risk
associated with prenatal blood Pb measures was substantiated by the weak correlation observed between
umbilical cord and age 5-year blood Pb levels (r = 0.29). Larger risks were estimated for Pb than for other
risk factors, including blood levels of mercury, polycyclic aromatic hydrocarbon DNA adducts, and
residential levels of dust mite or pet allergen.
Interestingly, in a study of 216 adults without occupational Pb exposures, Pizent et al. (2008) found
that among women, the association between blood Pb level and total IgE was statistically significant,
whereas the association with positive SPT to common inhaled allergens was not. An increase in IgE
mediates the acute inflammatory response to allergens. Among men, not only was the opposite observed,
but the association between blood Pb level and positive SPT was negative (OR: 0.92 [95% CI: 0.86, 0.98]
in stepwise regression models that considered age, smoking intensity, and alcohol consumption).
Interpretation of the findings is difficult because only statistically significant effect estimates were
reported; thus it is not possible to ascertain whether there were suggestions of association in women or
whether there were discrepant findings for the related outcomes of IgE and positive SPT.
5.6.4.3. Other Respiratory Effects
Increased bronchial responsiveness (BR) is a characteristic feature of asthma and other respiratory
diseases and can result from activation of innate immune responses and increased airway inflammation.
Compared with findings for IgE or asthma, evidence of association between blood Pb levels and BR is
weak (J. Y. Min et al.. 2008; Pizent et al.. 2008). In a study of 525 middle-aged adults in Seoul, Korea,
Min et al. (2008) found an association between blood Pb levels and BR. Study subjects had a mean (SD)
blood Pb level of 2.90 (1.59) (ig/dL. A 1 (ig/dL increase in blood Pb level was associated with an increase
in BR index (log [% decline in forced expiratory volume in 1 second (FEV,)/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 (J. Y. Min et al.. 2008). In contrast to Min et al. (2008). Pizent et al. (2008)
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observed a negative association between blood Pb level and nonspecific BR in men (2.4% decrease [95%
CI: -4.2, -0.52%] in percent change FEVi post-histamine challenge per 1 (ig/dL increase in blood Pb
level). Although this counterintuitive finding was not discussed by investigators, it was consistent with
the negative association found between blood Pb level and SPT among men in this study.
In less rigorous analyses that compared occupational groups with relatively low Pb exposures,
associations of blood Pb levels with lung function were less clearly indicated (A. Y. M. Jones et al.. 2008;
A. Y. M. Jones et al.. 2006). In a comparison of male drivers of buses with (mean blood Pb level: 5.0
(ig/dL) and without (mean blood Pb level: 3.7 (ig/dL) air conditioning in Hong Kong, China, drivers of
non-air conditioned buses had lower exposures to PMi0, lower blood Pb levels but lower forced vital
capacity and similar FEVi (A. Y. M. Jones et al.. 2006). In this study, the authors attributed the slightly
higher blood Pb levels among air conditioned bus drivers to the poor efficiency in the filters and higher
PMio 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 (A. Y. M. Jones et al.. 2008). Neither study directly examined associations between blood
Pb levels and lung function.
In addition to blood Pb, several recent epidemiologic studies have used Pb measured in PMi0 and
PM2 5 air samples to represent Pb exposures. Some studies have analyzed the Pb component individually,
whereas others have applied source apportionment techniques to analyze Pb as part of a group of
correlated components. A majority of air-Pb studies has found associations with asthma-related morbidity
in children and respiratory-related hospitalizations and mortality in older adults (Table 5-26). Despite the
concordance between the findings of air-Pb and blood Pb studies, a common limitation of air-Pb studies is
the variable size distribution of Pb-bearing PM (Section 3.5.3) and its relationship with blood Pb levels.
Additionally, in these air-Pb studies, other PM components such as elemental carbon, copper, and zinc
also were associated with respiratory effects. In the absence of detailed data on correlations among
components or results adjusted for copollutants, it is difficult to exclude confounding by these other
components.
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Table 5-26. Associations of air-Pb with respiratory effects
Study Population Air-Pb Data
Statistical analysis
Outcome
Effect Estimate
(95% Cl)a
Gent et al. (2009) CT. MA 2-dav ava Pb-PM^
149 asthmatic Mean: 5.1 ng/m3
children, ages 4-12 yr
1-yr follow-up 2000-
2003
Generalized estimating
equations with one-day lag
autoregressive structure,
adjusted for season, day of
week, and date
Wheeze
Shortness of breath
Fast-acting inhaler
use
OR per 5 ng/m°
increase11
1.07 (p = 0.13)
1.12 (p = 0.01)
1.04 (p = 0.10)
Hong etal.
(2007)
Dukjeok Island, Korea
43 children, grades 3-
6
6-wk follow up March-
May 2004
Lag 1 Pb-PMn
Mean (SD): 51 ng/mJ
(31)
Linear mixed effects models
with random effect for subject
adjusted for age, sex, height,
weight, asthma history,
passive smoking exposure at
home, GST genotype,
temperature, relative
humidity, air pressure, day of
week
Morning PEF
Daily average PEF
Change per 1 log
increase
-6.83 L/min (p <0.01)
-6.37 L/min (p <0.01)
Ostro et al. 6 California counties Lag 3 Pb-PM2s
(2007) Adults, ages > 65 yr Mean (IQR): 4 (4)
2000-2003 ng/m3
County-specific poisson
regression adjusted for day
of week, smoothing splines
for temperture and humidity
(3 df), smoothing spline for
time (4 df). County-specific
estimates combined using
random effects model
Respiratory mortality
All-year
Summer-only
RR per 4 ng/m° increase
1.011 (0.99,1.033)
Statistically significant,
quantitative results not
reported
Bell et al. (2009)
106 U.S. counties
Adults, ages > 65 yr
1999-2005
Lag 0 Pb-PM2.5
Mean (SD): 4.89 ng/m3
(31)
Andersen et al.
(2007)
Bayesian hierarchical
regression to combine
county-specific estimates
adjusted for day of week,
seasonality, smooth function
of time, daily temperature,
previous 3-day's temperature
and dew point temperature
Respiratory hospital
admissions
Results were reported in
a figure. RR per IQR
increase was negative
with wide 95% CI
RR per 0.6 ^ig/mJ
increase
0.95(0.89,1.02)
1.20(0.98, 1.47)
Copenhagen,
Denmark
Adults, ages > 65 yr
Children, ages 5-18 yr
1999-2004
Lag 3 "vehicle" PM10
factor0 comprising Pb,
copper, iron, and other
trace metals
Sarnat et al.
(2008)
Generalized additive Poisson
regression adjusted for
smoothed splines for
temperature and dew point
temperature (4-5 df),
smoothed spline for calendar
time (3-5 df), influenza, day
of week, public holidays,
school holidays, and pollen
(only for asthma models)
Respiratory hospital
admissions (adults
ages > 65 yr)
Asthma (children
ages 5-18 yr)
20 Atlanta, GA-area
counties
1998-2002
All ages
"Woodsmoke" PM2.s
factor0 comprising Pb
(minor contribution),
potassium, organic
carbon, ammonium
Mean: 1.6 mg/m3 (cool
season), 0.8 mg/m3
(warm season)
Generalized linear Poisson
regression adjusted for day
of week, holidays, hospital,
cubic splines for time
(monthly knots), temperature,
and mean dew point
temperature (knots at 25th
and 75th percentiles)
Respiratory hospital
admissions
Increment NR
0.999 (0.993, 1.004)
OR = odds ratio, PEF = peak expiratory flow, RR = relative risk, NR = not reported.
aEffect estimates are reported as given in the study and are not standardized because of variability in exposure metrics among studies,
investigators did not provide sufficient information to calculate 95% CIs.
°Souce apportionment techniques were applied to group correlated components into common source categories.
5.6.4.4. Autoimmunity
1 The 2006 AQCD described animal studies in which Pb exposure induced the generation of
2 autoantibodies (Bunn et al.. 2000; El-Fawal et al.. 1999; Hudson et al.. 2003; Waterman et al.. 1994).
3 Whereas some evidence linked this risk of autoimmunity to a shift towards Th2 responses, other evidence
4 pointed to a shift towards Thl responses. While recent studies did not examine Pb-induced production of
5 autoantibodies, some provided indirect evidence by indicating that the changes induced by Pb had broader
6 implications for increasing risk of autoimmunity. For example, Kasten-Jolly et al. (2010) examined the
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impact of developmental Pb exposure of mice on changes in gene expression in the spleen. They 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-
exposed mice also indicated the potential for autoantibody generation. Evidence of Pb-associated
autoimmune responses in humans is limited to a 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).
5.6.4.5. Specialized Cells in Other Tissues
Resident macrophages in tissues represent a significant target for Pb-induced immune effects, and
alteration in the function of these cells can contribute to organ/tissue dysfunction, cell death, tissue
pathology and tissue-specific autoimmune reactions. Among the specialized populations 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 effects of
Pb on these cells are important as they may contribute to disease in nonlymphoid tissues (Figure 5-52).
Because these specialized cells are not always recognized as macrophages, the resulting diseases and
conditions are not always recognized as being linked with Pb-induced immune dysfunction.
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Residentbone 4
macrophages
1
Osteoblasts and
Osteoclasts
> t
Periodontitis
anc' Altered liver metabolism
Chondrogenesis 0xidativedamage
Glutathionestatus
Figure 5-52. Specialized macrophages in nonlymphoid tissue may serve as
a significant link between Pb and disease in multiple organ
systems.
Fan et al. (2009) 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. In the central nervous system, subacute exposure of Wistar rats to Pb (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). Among the
cytokines detected in the brains of these Pb-treated rats were IL-1|3, TNF-a and IL-6. In bone, resident
tissue macrophages regulate osteoblast function (M. K. Chang et al.. 2008). and osteoblasts are known
targets of Pb. This can contribute to later life diseases such as arthritis [reviewed in Zoeger et al. (2006)1.
Pb-induced elevation of TGF-|3 production is also involved in chondrogenesis in bone.
Resident immune cells in reproductive organs are also potential targets of Pb. 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.
Pb
Kupffercells Microglia Alveolar
and Astrocytes macrophages
V
Neurobehavioral and
neurodegenerative
conditions
~
Bronchial hyper-reactivity
Respiratory allergy
Asthma
Susceptibilityto infections
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5.6.4.6. Tumors
While toxicological evidence supports high doses of Pb directly promoting tumor formation or
inducing 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
1300 ppm of Pb acetate in drinking water had enhanced moloney sarcoma virus-induced tumor growth
compared with control animals. These findings indicated that Pb-induced immunomodulation affecting
tumors likely results 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 immuntoxicology, particularly when early life exposures are involved
(Dietert).
5.6.5. Mechanisms of Lead-Induced Immunomodulation
5.6.5.1. Inflammation
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 (Section 5.2.5). It is important to note that
inflammation can manifest in any tissue where immune cell mobilization and tissue insult occurs (as with
an infection). Enhanced inflammation and tissue damage occurs through the modulation of inflammatory
cell function and production of proinflammatory cytokines and metabolites. 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). Chetty et al. (2005) reported that Pb-induced
inflammatory damage involves the depletion of glutathione. 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).
Traditional immune-mediated inflammation can be seen with asthma, respiratory infections, and
BR in association with Pb exposure. Nonetheless, inflammation also is a general response to tissue injury
whether mediated by infection, toxic insult, or other stresses. Thus, Pb-induced misregulation of
inflammation could exacerbate disease and damage in almost any organ given the distribution of immune
cells as both permanent residents and infiltrating cell populations. As described in Section 5.2.5,
misregulation of inflammation represents a potential mode of action for Pb induced effects on the liver,
kidney, and vasculature.
In epidemiologic studies, whereas Pb-associated changes in proinflammatory cell function (Section
5.6.2) and cytokine production (Section 5.6.5.4) have been demonstrated, less certain are the effects of Pb
exposure on nonspecific indicators of inflammation that may be related to multisystemic effects as have
been demonstrated in toxicological and in vitro studies. Using 1999-2004 NHANES data for adults 40
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years of age or older, Songdej et al. (2010) examined the relationship between blood Pb levels (mean:
1.89 (ig/dL) and the inflammation markers, C-reactive protein (CRP), fibrinogen, and white blood cell
(WBC) count. Adjusting for age, sex, race/ethnicity, education, income, body mass index (BMI), physical
activity, smoking status, diabetes status, inflammatory disease status, and cardiovascular disease status,
the researchers found that men appeared to be more susceptible than women to Pb-associated
inflammation. Among women, most odds ratios for associations between quintiles of blood Pb level and
tertiles of CRP, fibrinogen, and WBC count were less than 1.0 whereas corresponding odds ratios in men
tended to be greater than 1.0. Additionally, compared with men with blood Pb levels less than 1.16 j^ig/dL,
men with blood Pb levels of 1.16-1.63 (ig/dL, 2.17-3.09 (ig/dL, and > 3.09 |_ig/dL had statistically
significant increases in CRP (ORs: 2.22, 2.12, 2.85, respectively). For all inflammation markers, although
the highest quintile of blood Pb level was associated with the largest odds ratio, a blood Pb dose-
dependent association was not observed. Consistent with these findings, among men in Incheon, Korea
without occupational Pb exposures, Kim et al. (2007) reported positive associations of blood Pb level
with WBC as well as with IL-6. These findings of associations between blood Pb levels and inflammatory
mediators 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 (Fuller & Zhang. 2001; Hage & Szalai. 2007).
In a genome-wide association study that included 37 autistic and 15 nonautistic children (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, 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) (Y. Tian et al.. 2011). Although blood Pb levels
were similar between autistic and nonautistic children and correlations were observed in both groups, they
were in opposite directions (positive among autistic and negative among nonautistic). These results are
consistent with findings that Pb increases MHC molecule surface expression in mouse and human HLA
antigen presenting cells (Guo et al.. 1996a; M. J. McCabe & Lawrence. 1991) and also suggest that Pb-
associated changes in the expression of immune genes may be modified by underlying autism.
5.6.5.2. Increased Prostaglandin E2 and Decreased Nitric Oxide
Consistent with the findings presented in the 2006 Pb AQCD (U.S. EPA. 2006). recent studies
continue to show that Pb exposure alters the levels of signaling molecules. These signaling molecules are
involved in mediating inflammation and host resistance (Figure 5-47). For example, Pb exposure
increases arachidonic acid metabolism, elevating the production of prostaglandin E2 (PGE2) (Chettv et al..
2005). Additionally, production of nitric oxide (NO) by macrophages is decreased at low-moderate
exposure levels (Farrer et al.. 2008; Pineda-Zavaleta et al.. 2004). Decreases in NO can impact not only
innate host defenses, but also, acquired immunity. This is proposed to occur via the Pb-induced release of
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myeloid cell (CDllb+)-mediated suppression of CD4+ T cell proliferation (Farrer et al.. 2008). The result
is that Pb exposure leads to an increased production of alloreactive T and B cells. Farrer et al. (2008)
found that Pb decreases inducible nitric oxide synthease function in myeloid cells without decreasing its
abundance. The resulting loss of NO production also leads to a reduction in the production of the
microbicidal metabolite peroxynitrite, thus weakening host defenses against bacteria. When this is
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, 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 support associations of blood Pb level with
suppressed NO production (Barbosa et al.. 2006; Mishra et al.. 2006; Pineda-Zavaleta et al.. 2004;
Valentino et al.. 2007) and increased ROS production (Pineda-Zavaleta et al.. 2004). Among children, in
Pineda-Zavaleta et al. (2004). with increasing residential proximity to the Pb smelter, mean 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. NO release from macrophages (indirectly activated by
phytohemagglutinin, PHA) decreased with increasing blood Pb level. After adjusting for age and sex, a 1
(ig/dL increase in blood Pb level was associated with a decrease in NO of 0.00089 (95% CI: -0.0017, -
0.00005) nmol/(ig protein and an increase in superoxide anion of 0.00389 (95% CI: 0.00031, 0.00748)
(imol/mg protein. Because PHA activates macrophages indirectly through the activation of lymphocytes
and IFN-y directly activates macrophages, these results suggest that Pb suppressed T cell-mediated
macrophage activation and stimulated cytokine-induced macrophage activation. Results also
demonstrated a larger magnitude of association between blood Pb levels and superoxide anion release in
males. Although not described in detail, the association between blood Pb level and NO was reported to
be not negative in girls. Barbosa et al. (2006) also observed negative associations between blood Pb level
and plasma NO (quantitative results not provided) in a group of adults in Sao Paolo, Brazil residing near a
battery plant, although it was not possible to identify immune cells as the specific sources of NO.
In contrast to studies of populations without occupational Pb exposures, studies of occupationally-
exposed groups provided less clear indication of association of blood Pb level with NO and ROS. Among
30 male Pb recycling plant workers and 27 unexposed controls, despite large differences in blood Pb
levels, levels of ROS released from neutrophils (indicators of respiratory burst) were similar between
workers and controls (Mishra et al.. 2006). NO production of neutrophils after stimulation with zymosan-
A was lower in controls, but the difference was not statistically significant. In a study of male foundry
workers (mean blood Pb level: 21.7 |_ig/dL). pottery workers (mean blood Pb level: 9.7 (ig/dL), and
unexposed workers (mean blood Pb level: 3.9 (ig/dL), Valentino et al. (2007) also found lower plasma NO
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levels in controls compared with Pb-exposed workers. The means (ranges) of plasma NO levels in
controls, pottery workers, and foundry workers were 23.73 (.iM (11.21 to 55.71 (iM), 28.44 (.iM (15.23 to
57.65 (iM), and 25.30 (.iM (15.03 to 61.98 (iM), respectively. Further, although quantitative results were
not reported, blood Pb level was reported to be not correlated with NO.
5.6.5.3. Cellular Death (Apoptosis, Necrosis)
In a study in mice, Bishayi and Sengupta (2006) found that Pb exposure elevated DNA
fragmentation in splenic macrophages. Using mouse resident peritoneal macrophages, Gargioni et al.
(2006) found that inorganic Pb 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 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.
The Pb-induced DNA damage, determined by the comet assay, was accompanied by elevations in p53 and
Bax expression with no change in Bcl-2 expression (creating a Bax/Bcl-2 imbalance). The authors
propose that this is a likely route to Pb-induced apoptosis and tumorigenesis.
5.6.5.4. Cytokine Production
Toxicological studies have indicated that Pb affects immune cytokine production via action on T
cells, macrophages, and dendritic cells (Section 5.6.2). The combination of these three pathways of
cytokine changes induced by Pb creates a hyperinflammatory state among innate immune cells, and
acquired immunity is skewed toward Th2 responses. As illustrated in Figure 5-47, downstream effects
include altered IgE production, ROS production, and inflammation. Cheng et al. (2006) found that Pb
exposure affected TNF-a production in macrophages by affecting signaling pathways. This can result in
local tissue damage during the course of immune responses affecting such targets as the liver. In a study
involving macrophage-mediated liver injury in mice (A/J), substimulatory levels of Pb (10 (.iM)
coadministered with LPS stimulated the phosphorylation of p42/44 mitogen-activated protein kinase
(MAPK) and TNF-a expression (Y.-J. 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. These findings are consistent with those of Gao et al. (2007). which showed that treatment
of mouse dendritic cells with Pb produced an increased phosphorylation of the Erk/MAPK signaling
molecule.
The most consistent immunomodulatory effect of Pb is the skewing of immune responses away
from Thl and toward Th2. Pb was observed to skew toward Th2 cytokine production in both dendritic
cells (5.6.2.6) and T cells. Lynes et al. (2006) observed that Pb suppressed the production of Thl cytokine
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IFN-y. Gao et al. (2007) observed that Pb exposure elevated production of Th2 cytokines such as IL-4,
IL-5 and IL-6. This shift to a Th2 phenotype also was demonstrated in cultures of human blood
monocytes activated with Salmonella enteritidis or with monoclonal antibodies of CD3, CD28, and CD40
and exposed to inorganic Pb. Pb exposure 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).
In a recent study conducted across a lifetime (developmental through adulthood) using a broad
range of dietary Pb concentrations in Swiss mice (females and males), Iavicoli et al. (2006) suggested a
nonlinear hierarchical cytokine response. At the lowest dietary Pb concentration (0.8 (ig/dL), IL-2 and
IFN-y were elevated over the controls, indicating an enhanced Thl response. However, as dietary Pb
exposure increased (resulting in blood Pb levels 12-61 j^ig/d L). 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 low Pb exposures and that compensatory mechanisms may be
stimulated at low Pb exposures. Other studies have found variable Pb-induced changes in IL-2, with no
change or elevated production, depending upon the protocol used. Recently, Gao et al. (2007) found that
Pb-treated dendritic cells promoted a slight but statistically significant increase in IL-2 production among
lymphocytes. A recent study on bone (in vivo and in vitro) chondrogenesis found Pb-induced increases in
production of TGF-(3 (Zuscik et al.. 2007).
Consistent with toxicological studies, epidemiologic studies also find higher blood Pb levels in
humans to be associated with a shift towards production of Th2 cytokines relative to Thl cytokines. One
consequence of an increase in the Th2 cytokine IL-4 is the activation of B cells to induce B cell class
switching to IgE. In particular, Lutz et al. (1999) provided evidence for Pb affecting this pathway by
finding that children with blood Pb levels greater than 10 (ig/dL had both higher levels of IL-4 and IgE
(Section 5.6.3). Among adults in Incheon, Korea without occupational Pb exposures with blood Pb levels
ranging from 0.337 to 10.47 (ig/dL, Kim et al. (2007) found associations of blood Pb level with serum
levels of TNF-a and IL-6 that were larger among men with blood Pb levels > 2.51 (ig/dL (median). 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 associations between levels of blood Pb and plasma TNF-a were 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 metabolism of ROS, may confer increased susceptibility to 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.
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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 (Pi Lorenzo et al.. 2007; Valentino et al.. 2007; Yucesov etal.. 1997a). The exception was the
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 effect estimates were not reported, statistically significant associations were
observed between blood Pb level and IL-10 and TNF-a. Levels of IL-2, IL-6, and IL-10 showed an
increasing trend from the lowest to highest blood Pb group. In contrast with most other studies, both
exposed worker groups had lower IL-4 levels compared with controls. In a similar analysis, DiLorenzo et
al. (2007) separated exposed workers into intermediate (9.1-29.4 (ig/dL) and high (29.4-81.1 |_ig/dL)
blood Pb level groups, with unexposed workers comprising the low exposure group (blood Pb levels 1-11
(ig/dL). Excluded from this study were exposed workers from the highest end of the distribution of blood
Pb levels in DiLorenzo et al. (2006). Mean TNF-a levels showed a monotonic increase from the low to
high blood Pb level group. A synergistic effect was observed between blood Pb levels and smoking.
Among current smokers, a 12- to 16-fold difference in TNF-a levels was observed among blood Pb
groups, whereas a less than twofold difference was observed among nonsmokers. In Yucesoy et al.
(1997a). levels of the Thl cytokines, IL-ip and IFN-y. were lower in workers than in controls; however,
differences were not observed in levels of the Th2 cytokines, IL-2 and TNF-a.
5.6.6. Immune Effects of Lead within Mixtures
One of the most striking observations regarding Pb effects on the immune system since the 2006 Pb
AQCD concerns the effects of metal mixtures. Recent studies indicate that immune effects may be
observed with lower levels of Pb exposure when they occur in conjunction with other metals. Information
on interactions among metal mixtures may improve risk assessment methods because most human
exposures to chemicals involve mixtures at low environmental levels. In a study of mice exposed to Pb
acetate (10 mg/kg by weight), As (0.5 mg/kg by weight), or both, Bishayi and Senguta (2006) reported a
greater than additive effect of coadministered Pb and As on macrophages in producing a decrease in
bacterial resistance, myeloperoxidase (MPO) release, and NO production. Investigators assessed the Pb-
As interaction using the multivariate ANOVA for the experimental results of MPO release 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.
Institoris et al. (2006) reached a similar conclusion after observing that lymph node weight
decreased with exposure to 20 mg/kg Pb plus a second metal (Cd or Hg) but not with 20 mg/kg of Pb
alone. Another study conducted in rats (Jadhav et al.. 2007) found that mixtures of Pb, Hg, and other
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metals at 10 to 100 times the concentrations of the individual components typically measured in drinking
water in India suppressed lymphocyte counts and antibody titers and increased neutrophil counts. In
contrast with these aforementioned studies, Fortier et al. (2008) found that PbCl2-exposed (7.5-20.7
(ig/dL) human leukocytes did not have alterations in lymphocyte proliferation, monocytic phagocytic
activity, or NK cell activity. Although the combination of 20.7 (ig/dL PbCl2 plus 12.0 (ig/dL
methylmercuric chloride (MeHgCl) decreased lymphocyte proliferation, these effects were attributed to
MeHgCl, which had a stronger suppressive effect individually. The majority of evidence indicating
synergism between Pb and metals such as As, Cd, and Hg suggests that a threshold for producing Pb-
induced immune effects may be lower if additional metals are present.
In a study designed to mimic exposure to particles associated with urban traffic, Wei et al. exposed
human endothelial cells in culture (EA.hy926) to urban fine particles (PM2 5 enriched with Pb and other
metals). Investigators found that the particle mixture increased ROS production and mitochondrial -
mediated apoptosis; however, they did not test metals individually and could not attribute findings to Pb
individually or interactions between Pb and other metals within the mixture. Similarly, in two studies of
mice exposed to Pb and Cd in drinking water (1 ppm of each metal in drinking water for 28 days), Pb and
Cd altered antibody titres (Massadeh et al.. 2007). The main aims of these studies were to demonstrate
that the effects of Pb and Cd could be reversed with administration of Nigella sativa L (black cumin) or
garlic extract. However, investigators did not test each metal individually to assess interactions between
metals.
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)
reported interactions between Pb and As. In addition to Pb, As contamination of drinking water 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 negatively associated with NO (similar to Pb), and
a statistically significant negative interaction was observed between Pb and As, indicating that high
internal doses of both metals were associated with a larger decrease in NO than that associated with either
metal alone. Urinary As was negatively associated with superoxide anion (opposite direction of Pb), and
the Pb by As interaction was positive. Thus, although higher urinary As was associated with a lower
superoxide anion level, higher internal doses both Pb and As were associated with a larger increase in
superoxide anion than that associated with blood Pb level alone.
5.6.7. Summary and Causal Determination
The collective body of evidence integrated across epidemiologic and toxicological studies
consistently demonstrates that the immune system is a major target of Pb. Rather than resulting in overt
cytotoxicity to lymphoid tissues, Pb exposure has been associated predominantly with more subtle
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changes in a spectrum of immune mediators and functions. The importance of Pb as an immunomodulator
is particularly evident when one considers: (1) the relative sensitivity of the immune system to Pb-
induced modulation; (2) the lifetime health ramifications of exposure to Pb in early development,
particularly given the recognized sensitivity of the developing immune system to environmentally-
induced programming; and (3) the consequent broad spectrum of diseases and illnesses in multiple
physiological systems potentially related to Pb-associated immune dysfunction. The majority of results
from animal studies indicates that immune changes are observable at blood Pb levels of <8.0 (ig/dL, with
new evidence in mice demonstrating that nonmonotonic cytokine changes occur at blood Pb levels of 2-3
(ig/dL. Consistent with the animal studies, in the newly expanded body of epidemiologic studies in
children and adults without occupational Pb exposures, changes in immune parameters are commonly
demonstrated in association with mean blood Pb levels in the range of <2 (ig/dL to 10 (ig/dL.
The strength of evidence for Pb-associated immune effects is derived not only from the consistency
of associations but also from the coherence of findings between toxicological and epidemiologic studies
and coherence of findings among the spectrum of immune changes operating within particular pathways.
A large body of toxicological evidence demonstrates Pb-induced suppression of T cell proliferation, and
epidemiologic evidence in children consistently links elevated blood Pb levels (as low as 2.2-3.4 j^ig/dL)
with decreases in T cell abundance. These changes can affect cell-to-cell interactions that mediate
acquired immunity required in subsequent memory responses to antigen exposures. Despite the
consistency of evidence for some T-cell subtypes (e.g., CD3+, CD4+), it is unclear what effect the
observed magnitudes of decrease may have in attenuating acquired immunity.
The prominent effect of Pb exposure on T cells, in terms of coherence with effects on other
immune endpoints and implications for developing immune-based diseases, is the skewing of immune
function away from a Thl phenotype towards a Th2 phenotype. In toxicological studies, this shift is well-
established by suppressed production of Thl cytokines (e.g., IFN-y) and increased production of Th2
cytokines (e.g., IL-4). A recent toxicological study builds on this extant evidence by indicating that Pb
may promote Th2 responses by acting directly on dendritic cells, the major effector in antigen response.
Findings from recent epidemiologic studies strengthen the evidence with observations of similarly
skewed cytokine profiles in association with blood Pb levels in humans in the range of 2.5-10 (ig/dL.
Further, toxicological and epidemiologic studies link the Pb-associated predominance of Th2 cytokines
with downstream effects on humoral immunity by demonstrating Pb-associated changes in B cell
abundance and changes in circulating antibody 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. IgE is the primary mediator for type 1 hypersensitivity resulting in various
allergic conditions and asthma. In support of this well-established mechanism, toxicological studies
describe Pb-induced (blood Pb levels 10-30 (ig/dL) changes in IgA, IgG, and IgM. Additionally,
epidemiologic studies in children consistently demonstrate that blood Pb levels are positively associated
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with B cell abundance (blood Pb levels in the range of 5-10 j^ig/dL) and increases in IgE (blood Pb levels
as low as 2.8-3.4 (ig/dL). Observations of Pb-associated increases in Th2 responses and circulating IgE
levels provide biological plausibility for epidemiologic observations in children of associations of blood
Pb levels (in the range of 1.16-10 (ig/dL) with asthma and allergic conditions. Such epidemiologic data
are limited, and additional studies with more rigorous methodology (e.g., longitudinal design to establish
temporality, improved assessment of Pb exposures, adjustment for potential confounders such as
smoking, SES, and exposures to other metals) are needed to substantiate the findings.
Suppression of Thl function by Pb places individuals at greater risk of certain infectious diseases
and cancer. Compared with the relationships between Pb-suppressed Thl-dependent antitumor processes
and effects on tumor formation, Pb effects on decreasing host resistance are relatively well-established.
Pb exposure of animals (resulting in blood Pb levels <2-5 (ig/dL) suppresses the DTH response, and a
recent in vitro study indicates such effects may be mediated by dendritic cells. Further evidence of Pb-
associated suppressed Thl activity is provided by toxicological and epidemiologic observations that Pb
exposure and blood Pb levels, respectively, are associated impaired phagocytic and chemotactic activity
of macrophages and neutrophils (blood Pb levels >25 (ig/dL in humans). Thl-dependent IFN-y normally
enhances the killing capacity of macrophages. Epidemiologic studies additionally demonstrate that higher
blood Pb levels are associated with lower neutrophil respiratory burst (superoxide anion release),
indicative of diminished degradation of phagocytosed particles. Toxicological and epidemiologic
evidence for suppressed Thl activity and effects on macrophage and neutrophil functional activities
provide biological plausibility for observations in animals and humans for associations between blood Pb
levels (in the range of 3-10 (ig/dL in humans) and increased risk of infection.
Toxicological studies demonstrate that Pb induces 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).
Although examined in fewer studies, epidemiologic studies find higher blood Pb levels (in the range of 6-
20 (ig/dL) to be associated with higher serum levels of ROS, TNF- a and lower serum levels of NO.
These proinflammatory effects of Pb exposure on macrophages provide support for associations of Pb
with elevated risk of disease in multiple physiological systems. The strongest evidence comprises Pb-
induced changes in specialized macrophages such as alveolar macrophages, testicular macrophages, and
microglia, whose altered homeostasis or function may contribute to documented associations of blood Pb
levels with asthma and bronchial reactivity, poor reproductive performance, and neurodegeneration,
respectively. Although limited mostly to toxicological studies, Pb has been shown to induce the
generation of autoantibodies, suggesting that Pb exposure may increase the risk of developing
autoimmune conditions. These findings are supported by Pb-induced inflammation and tissue damage and
Pb-induced T cell activation in response to new antigens.
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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 changes in immune cell abundance
and function that subsequently lead to a broad spectrum of changes in both cell-mediated and humoral
immunity to promote a Th2 phenotype and hyperinflammatory state. Toxicological and epidemiologic
findings of Pb-associated decreases in T cells, inhibition of Thl-type responses, and impaired phagocytic
activity of macrophages and neutrophils provide biological plausibility for evidence linking Pb exposure
with increased risk of bacterial and viral infection. Additionally, toxicological and epidemiologic evidence
of a Pb-associated promotion of the Th2 phenotype, increased B cell abundance, increased synthesis of
IgE, and increased inflammation support observations in children for associations of blood Pb levels with
asthma and allergic conditions. Although not widely examined in humans, toxicological findings indicate
that Pb-induced immunomodulation may have broader implications for autoimmunity and mediating Pb
effects in other physiological systems such as the nervous, cardiovascular, and reproductive systems.
Animal studies and to a limited extent, epidemiologic studies, demonstrate increased susceptibility from
prenatal exposures and enhanced responses with co-exposures to other metals. The consistency of
findings among toxicological and epidemiologic studies and the coherence of findings between these
disciplines and across the continuum of related immune responses 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 RBCs (red blood cells [RBC]) 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 Pbinduced 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 decreases RBC survival, as well as alters 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
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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 results 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 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 (See Figure 5-53 for 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 Pb
exposure with a concentration-response change in the ratio of activated/nonactivated ALAD activity that
is 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. 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 Ca2+ or
magnesium (Mg2+) ATPases.
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The 2006 Pb AQCD document identified oxidative stress as an important potential mechanism of
action resulting from Pb exposure. Increased lipid peroxidation and inhibition of antioxidant enzymes
(e.g., superoxide dismutase [SOD], catalase [CAT]) was observed following exposure to Pb.
5.7.2. Effects on Red Blood Cell Functions
As stated in the 2006 Pb AQCD, Pb poisoning is 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
reduced in non- or mildly anemic, iron-deficient children in India fed an iron-fortified diet for 30 weeks
(p <0.02); blood Pb levels were not reduced in the children not receiving the iron-fortified diet. Although
a number of studies find decreases in RBCs and/or Hct levels associated with 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. Lead Uptake, Binding, and Transport into Red Blood Cells
The 2006 Pb AQCD reports 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 leaves the
cell after entry. Simons (1993 b) found that in vitro uptake of 203Pb (1-10 (.iM) occurred via an anion
exchanger while the efflux was 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. 1986V
However, Bergdahl et al. (1997) suggested that ALAD was the primary Pb binding protein and not Hb.
The 2006 Pb AQCD also reports that the majority (approximately 98%) of Pb accumulates in RBC
cytoplasm bound to protein and only about 2% is found in the membrane. This is related to the high ratio
of Pb in RBCs compared to plasma Pb. Further Information on Pb binding and transport in blood can be
found in the kinetics section of Chapter 4 (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.
5.7.2.2. Red Blood Cell Survival, Mobility, and Membrane Integrity
Although it is known that Pb exposure shortens the RBC life span and alters 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 in free Ca2+. There are also newer studies that
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examine the relationship between Pb and RBC survival, mobility, and membrane integrity. Pb-exposed
workers in a recycled automobile battery factory in Mexico had 7.5 times as much blood Pb as was found
in unexposed workers (74.4 ± 21.9 (ig/dL in 15 exposed workers versus. 9.9 ± 2 (ig/dL in 15 unexposed
workers, p <0.01) (Ouintanar-Escorza et al.. 2007). Intracellular RBC free calcium levels |Ca2 |, were
significantly higher in the Pb-exposed workers than in the unexposed workers (79 ± 13 versus. 30 ± 9 nM,
p <0.05). The level of |Ca2 | was highly correlated (R2 = 0.754) with blood Pb levels in exposed workers;
| Ca2 | was also elevated in the unexposed workers who had low blood Pb levels ranging from 7.9 (ig/dL
to 11.9 (ig/dL. The observed elevation in | Ca2 | was related to an increased uptake and a decreased efflux
due to reduced Ca2+-Mg2+-ATPase activity. In the RBCs of Pb-exposed workers, the activity of Ca2+-
Mg2+-ATPase was 28 ± 8 nmol Pi/mg protein/min versus. 40 ± 9 in unexposed workers (p <0.05). These
changes were associated with increased fragility of the RBCs and dramatic morphological alterations,
including the increased presence of ecinocytes (cells without normal biconcave shape) and crenocytes
(speculated cells) in Pb-exposed workers. Similar dose-dependent effects were observed when RBCs
from healthy subjects were incubated with Pb at concentrations from 0.2 to 6.0 (.iM (Ouintanar-Escorza et
al.. 2010). Abam et al. (2008) also observed a decrease in the activity of RBC membrane bound Ca2+-
Mg2+-ATPase in workers exposed to Pb in a number of occupations. While the authors did not observe an
increase in | Ca2 | or Mg in RBCs from exposed workers, the RBC membrane | Ca2 | and Mg2+
concentration were increased. Abam et al. (2008) did not report on morphological changes in erythrocytes
from exposed workers. Ciubar et al. (2007) found that RBC morphology was disrupted with 50% or more
of RBCs exposed to Pb concentrations of 0.5 (.iM or higher for 24 hours at 37°C having lost the typical
discocytic morphology and displaying moderate to severe echinocytosis. Exposure of RBCs to higher
concentrations of Pb nitrate resulted in cell shrinkage. Ademuyiwa et al. (2009) observed that the
cholesterol content of RBC plasma membranes, but not the phospholipid content, was statistically
significantly higher in rats exposed to 200 ppm Pb acetate (blood Pb = 40.63 ± 9.21 (ig/dL) through
drinking water compared to controls. Further, the cholesterol/phospholipid ratio was increased in the 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
hematological parameters. (Ukaeiiofo et al.. 2009) studied hematological effects of Pb in 81 male subjects
exposed to Pb at three different manufacturing companies in Nigeria. 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 an average 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). Exposed
subjects had significantly reduced Hb and packed cell volume (PCV) levels and increased percentage of
reticulocytes. Although the differences were significant between the exposed and control subjects, the
study authors state that the results in the exposed subjects were at the lower range of normal for
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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 control group II (p <0.01). Patil et al. (2006) examined hematology
effects in jewelry workers in India occupationally exposed to Pb. Blood Pb was significantly higher in the
jewelry workers (48.56 ± 7.39 (ig/dL) compared to individuals not occupationally exposed (12.52 ± 4.08
(ig/dL). There was no significant difference in the Hb, MCV, or MCH; however, jewelry workers had a
significant decrease in PCV and total RBCs, accompanied by a significant increase in MCHC. In addition,
the exposed workers had a significant increase in total leukocyte levels. In battery manufacturing workers
in India, significant decreases in Hb, PCV, MCV, MCH, MCHC, and total RBCs were observed (Patil.
Bhagwat. Patil. Dongre. Ambekar. Jailkhani. et al.. 2006) with similar blood Pb levels (53.63 ± 16.98
(ig/dL). Industrial workers in Pakistan occupationally exposed to Pb (blood Pb = 29.1 (range 9.0-61.1)
(ig/dL) had a significant increase (3.5-fold higher median) in blood Pb levels compared to age and gender
matched controls (blood Pb = 8.3 (range 1.0-21.7) (ig/dL) (D. A. Khan et al.. 2008V The industrial
workers had a significantly lower Hb, but not a significant difference in the number of RBCs. There were
no significant differences in the number of white blood cells (WBCs) or in platelet counts. Karita et al.
(2005) examined the relationship between hematological parameters in Pb-exposed workers (blood Pb =
26.9 (ig/dL) in a variety of occupational settings. Blood Pb was found to be negatively correlated with
Hb, Hct, and total RBC count (p <0.01). Fonte et al. (2007) described a case-report in which a 47-year old
male exposed to Pb fumes and vapors at a recycling plant (blood Pb = 148 (ig/dL) experienced
normocytic, normochromic anemia, along with reticulocytosis and RBC basophilic stipplings. Following
chelation therapy, the hematological symptoms improved. Taken together, these studies provide consistent
evidence that occupational exposure to Pb reduces the number 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 in Ukaejiofo et al. (2009) seems to represent compensation for
decreased RBC survival due to Pb exposure.
Studies in children were more equivocal than those in adults. Riddell et al. (2007) found that 21%
of children 6 months to 5 years of age living in rural Philippines had blood Pb levels greater than 10
(ig/dL (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. However, Huo et al. (2007) found that children living near
an area where electronic waste was recycled in China had significantly higher blood Pb levels than
children in the neighboring town with no waste recycling (15.3 versus 9.94 (ig/dL), but no difference in
the Hb levels in the children in the two towns was detected (127.55 g/L in children with the higher blood
Pb levels versus 123.46 g/L). Ahamed et al. (2006) studied male urban adolescents in India. The 39
adolescents were separated into groups according to their blood Pb level (group 1: <10 (ig/dL, group 2:
>10 (ig/dL). Although the groups were similar in their age, height, weight, and body mass index, group 2
had a significantly lower PCV compared to group 1. The equivocal findings in studies investigating
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hematological effects in children may be due to the comparatively shorter time period and magnitude of
exposure versus, those seen in occupational 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 100 nM to 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. Both studies demonstrated that Pb exposure resulted in increased hemolysis,
observed as a significant increase in extracellular Hb. Pb-induced hemolysis in these experiments may be
due to inhibition of RBC phosphoribosyltransferases (Section 5.7.5.1). The in vitro studies indicated a
concentration-dependent increase in the amount of hemolysis, with a significant (threefold) increase even
at the lowest concentration tested (i.e., 100 nM). Lee et al. ("2005) observed that rats orally administered
Pb (25 mg/kg) once a week for 4 weeks had an average plasma Pb level of 6.5 (ig/dL (9.6-fold higher than
controls, p <0.05), and had significant decreases in Hct, Hb, and RBCs (p <0.05) (M. K. Lee et al.. 2005).
Male mice administered 50 mg/kg Pb nitrate in distilled water via gavage for 40 days had significantly
reduced total RBC counts, total leukocyte counts, Hb, lymphocytes, and monocytes (Sharma et al.. 2010).
Rats exposed to 2 g/L Pb acetate in drinking water for 30 days had significantly decreased RBCs, Hb,
PCV, MCH, and MCHC compared to controls (p <0.05) (Simsek et al.. 2009). Microcytic and basophilic
erythrocytic granulations were commonly seen in Pb-exposed animals. An indication that the decrease in
RBC count was related to decreased survival, and not a disruption of hematopoiesis, was the observation
of significantly increased reticulocyte density and total count compared to controls (p <0.05). Mice
exposed to 1 g/L Pb acetate in drinking water for 90 days, but not those exposed for 15 or 45 days, had
significantly decreased RBC counts and Hct compared to controls (p <0.05) (C. C. Marques et al.. 2006).
Spleen weights were also observed to be increased relative to body weight in animals exposed to Pb for
45 days. Male rats administered Pb acetate in the drinking water for 4 weeks at concentrations ranging
from 100 to 1,000 ppm had a dose-dependent increase in blood Pb (range: 6.57 to 22.39 (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 (M. Y. Lee et al.. 2006). Slight, nonsignificant, increases in
PS expression on RBC membranes were also observed. 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/1 Pb acetate for three hours were much smoother than untreated RBC membranes when
examined by atomic force microscopy. The authors postulate that the observed smoothing in treated RBC
membranes may be due to clusterization of band 3 protein. Band 3 (anion exchanger 1 [AE1]), is a
chloride/bicarbonate (CT'/HCOO 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
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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 AEl's normal activity.
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 (Folleret 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 to 3 (.iM caused increased activation of K+ channels that lead to cell shrinkage and scramblase
activation. The activation of scramblase increased the exposure to PS on the cell membrane, which causes
an increase in 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 increase in expression of PS
was 26.8 ± 3.15% (compared to deionized water) 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 (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 dose-dependent manner following exposure to Pb. To
confirm these findings in vivo, Shin et al. (2007) exposed male rats i.p. to 25, 50, or 100 mg/kg Pb
acetate. 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. Ciubar et al. (2007) also found that exposure to Pb nitrate (0.5-2 (.iM)
resulted an increase in PS exposure and cell shrinkage, which they stated were indicators of cell
apoptosis. As reported above, Khairullina et al. (2008) observed 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 mice, PS exposure was much greater
than in wild type mice. Decreased RBCs and increased reticulocytes were also observed, an indication of
high cell turnover.
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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 tricycle taxi drivers (n = 27) working at Kathmandu who were
not anemic. 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 level of serum erythropoietin (12.7 versus 18.8 mU/mL)
compared to the nondrivers and there was a statistically significant inverse relationship between the level
of serum erythropoietin and blood Pb (r = -0.68, p <0.001). No other hematological effects were
observed. 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 this is generally considered a measure of kidney
toxicity, it can also be considered an indicator that Pb could possibly affect the level of RBCs through
decreased levels of serum erythropoietin.
Celik et al. (2005) observed that exposure of female rats to 140, 250, or 500 mg/kg Pb acetate once
per week for 10 weeks resulted in statistically significantly decreased numbers of polychromatic RBCs
(PCE) and increased numbers of micronucleated PCEs, compared to controls (p <0.001). Alghazal et al.
(2008) exposed male and female rats to 100 mg/L Pb acetate daily for 125 days and observed statistically
significant 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, a significant 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. (2008) indicate that Pb is cytotoxic in male rats only, but is genotoxic in both sexes, Celik
et al. (2005) indicates Pb is cytotoxic in female rats as well. Mice exposed to 1 g/L Pb acetate in drinking
water for 90 days had statistically significant increases in micronucleated PCEs; a small, but not
statistically significant decrease in the PCE/NCE ratio was also observed (C. C. Marques et al.. 2006).
Cyto- and genotoxicity in RBC precursor cells is a strong indication 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 report, Pb has been found to affect RBC membrane
polypeptides in exposed workers (Fukumoto et al.. 1983);(Apostoli et al.. 1988). 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 occurred at blood
Pb levels greater than 50 (ig/dL. Exposure to Pb acetate at concentrations above 100 nM for 60 minute
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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 (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
cord blood Pb (p <0.0001 versus, p <0.05). Maternal Pb levels were not correlated with Capump 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.
5.7.3. Effects on 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 5-53). The observation of decreased Hb (measured as total Hb,
MCH, or MCHC) in occupationally exposed adults (Karita et al.. 2005; D. A. Khan et al.. 2008; Patil.
Bhagwat. Patil. Dongre. Ambekar. Jailkhani. et al.. 2006; Ukaeiiofo et al.. 2009). children (Riddel 1 et al..
2007). and experimental animal models (M. K. Lee et al.. 2005; C. C. Marques et al.. 2006; Sharma et al..
2010; Simsek et al.. 2009) is a direct measure of decreased heme synthesis due to Pb intoxication.
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Mitochondria
Pb
Heme
Pb
Protoporphyrinogen IX
Coproporphyrinogen III
Cytosol
2x6-aminolevulinic acid
Pb
ALA dehydratase
(porphobilinogen
synthase
"T urinary 6-ALA
4, ALA-D
Porphobilinogen (PBG)
4 C02
^ Coproporphyrinogen III
^ Uroporphyrinogen III
hydroxymethylbilane
Uroporphyrin ogen Uroporphyrinogen
III synthetase decarboxylase
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 J, arrows.
Figure 5-53. Schematic representation of the enzymatic steps involved in
the heme synthetic pathway.
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.
Patil et al. (2006) examined jewelry workers (blood Pb = 48.46 ± 7.39 (ig/dL) in India
occupationally exposed to Pb. In the study both the activated and nonactivated ALAD activities were
measured. The study authors state that this is because decreases in ALAD activity reached a plateau and
anemia can result in increased ALAD levels. Therefore, they considered the ratio of
activated/nonactivated ALAD to be a good indicator of Pb toxicity. As described in the 2006 Pb AQCD,
Scheuhammer (1987) studied the usefulness of this ratio in avian RBCs and found it to be a sensitive,
dose-dependent measure of Pb exposure regardless of the route of exposure. They found that the
activated/nonactivated ALAD ratio could be used to determine the oral exposure due to the highly
positive correlation between Pb exposure in the 5-100 ppm range and ALAD activity ratio. Patil et al.
(2006) observed a statistically significant decrease in nonactivated ALAD and an increase in the ratio of
activated to nonactivated ALAD, compared to nonexposed controls (p <0.05 and 0.001, respectively) .
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The study authors state that this indicates the inhibition of heme synthesis. Urinary excretion of both ALA
and PBG was statistically significantly increased in jewelry workers, a further indication of RBC ALAD
inhibition. Similar results were seen in battery manufacturing workers in India (Patil. Bhagwat. Patil.
Dongre. Ambekar. Jailkhani. et al.. 2006). Pb-exposed workers (blood Pb = 74.4 ±21.9 (.ig/dL) in a
recycled automobile battery factory in Mexico had a significant decrease (-90%) in RBC ALAD activity
compared to unexposed workers (9.9 ± 2 (.ig/dL) (Ouintanar-Escorza et al.. 2007). Painters in India with
an average blood Pb level of 21.92 (ig/dL had a significant decrease in ALAD levels compared to controls
with an average blood Pb level of 3.06 (ig/dL (p <0.01) (Mohammad et al.. 2008). Stoleski et al. (2008)
observed that workers in a Pb smelter in Macedonia (blood Pb = 16.4 ± 8.5 |_ig/dL) had significantly
decreased ALAD activity (p <0.001) and increased ALA levels (p <0.0005) compared to workers with no
exposure to (blood Pb = 7.0 ±5.4 (ig/dL). Lastly, Ademuyiwa et al. (2005) observed that workers in
mechanic workshops (blood Pb = 27.0 ± 1.1 - 48.9 ± 19.1 (ig/dL j^ig/dL) had significant decreases in
ALAD activity compared to controls (15.8 ± 2.8 (ig/dL, p <0.001). Petrol station workers had the highest
degree of ALAD inhibition relative to controls (77%), whereas welders only had a 36% decrease in
ALAD activity compared to controls. A case report by Fonte et al. (2007) described a worker
occupationally exposed to Pb vapors (blood Pb = 148 (ig/dL) with ALAD levels substantially decreased
compared to normal levels (3 versus >25 U/L). Zinc protoporphyrin (ZPP) levels were also described to
be greatly increased. Following chelation therapy, the patient's clinical picture improved.
Wang et al. (2010) found that even with low to moderate blood Pb levels, there was a
concentration-dependent decrease in ALAD activity in both children and adults (blood Pb = 7.1 and 6.4
(ig/dL, respectively) in rural southwest China. Further, Wang et al. (2010) observed that the relationship
between blood Pb and ALAD activity was nonlinear and exponential, with more significant 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. (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, group 2: >10 (ig/dL). Although the groups were similar in their age, height, weight,
and body mass index, group 2 had a significantly lower ALAD activity than group 1 (p <0.001). When all
39 adolescents were examined together, an inverse relationship was found between blood Pb and ALAD
activity. Ahamed et al. (2005) also observed decreased ALAD activity in Indian children (aged 4-12) with
a mean blood Pb level ofll.39± 1.39 (ig/dL compared to children with mean blood Pb levels of 3.93 ±
0.61 (ig/dL. Similar decreases were also observed in children 3-6 years of age with >10 (ig/dL, compared
to children <10 (ig/dL (Y. P. Jin et al.. 2006).
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,
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which were associated with significant decreases (approximately 50% lower than control levels) in RBC
ALAD activity (M. K. Lee et al.. 2005).
5.7.4. Other Heme Metabolism Enzymes
The 2006 Pb AQCD report indicates Pb affects RBC PBG synthase (Farant & Wigfield. 1987);
(Farant & Wigfield. 1990; Simons. 1995). PBG deaminase (Tomokuni & Ichiba. 1990). and TF
endocytosis and iron transport across membranes (Z. M. Qian & Morgan. 1990). all of which are directly
or indirectly involved in heme synthesis. Although there are no new studies that measure the effect Pb has
on other heme metabolism enzymes' activities, a number of studies investigated the effect blood Pb had
on 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. However, Pb inhibits this insertion of
Fe2+ into the protoporphyrin ring and instead, Zn is inserted into the right creating ZPP A number of
recent studies have shown that blood Pb is statistically significantly associated with increased RBC ZPP
levels in humans (Ademuviwa. Ugbaia. Oio. et al.. 2005; Counter et al.. 2007; Mohammad et al.. 2008;
Patil. Bhagwat. Patil. Dong re. Ambekar. Jailkhani. et al.. 2006) and animals (Rendon-Ramirez et al..
2007). Patil et al. (2006) observed a small, but not statistically significant increase in RBC ZPP levels
among silver jewelry workers exposed to Pb, compared to nonexposed controls. Interestingly, Q. Wang et
al. (2010) found that in children and adults living in a rural area of Southwest China, ZPP levels were
decreased at the low blood levels of Pb and were only increased with higher blood Pb levels. The authors
suggest that this may be representative of ALAD activities at low Pb levels, which contributes to lower
ZPP levels. Scinicariello et al. (2007) performed a meta-analysis and observed that Pb-exposed
individuals that carried the ALAD2 allele had slightly, but not statistically significant, lower
concentrations of blood ZPP levels compared to carriers of the ALAD1 allele.
5.7.5. Effects on 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 reports 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 report, there have been few studies examining Pb effects on
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energy metabolism. Baranowska-Bosiacka et al. (2009) examined the effects of Pb on RBC APRT and
hypoxanthine-guanine 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
concentration range from 100 nM to 100 (.iM for 5-30 minutes of exposure. 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 100 nM concentration, but the
decrease was also dose-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 dose-dependent decrease in
intracellular ATP in human RBCs in vitro with significant decreases found even with the lowest
concentration (i.e., 1 (.iM).
5.7.5.2. Other Enzymes
The 2006 Pb AQCD reports 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 the inhibition of
Ca2+-Mg2+ ATPase. Only two studies since the 2006 report 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 metalloprotein that use Zn to
catalyze the equilibrium between carbon dioxide and bicarbonate in the cells of higher invertebrates.
Although they found Pb nitrate inhibited both carbonic anhydrase isozymes in a concentration-dependent
manner, the concentrations used (i.e., 2 x 10"4to 1 x 10"3 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 enzyme's active site 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) (Ahamed & Siddiqui. 2007; Nemsadze et al.. 2009). Many studies have
found an association between the level of blood Pb and lipid peroxidation, antioxidant levels, or
indicators of ROS production.
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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. Patil et al. (2006) found significantly higher blood Pb in the jewelry
workers (48.56 ± 7.39 (ig/dL) compared to individuals not occupationally exposed (12.52 ± 4.08 (ig/dL)
to Pb. These workers had significantly higher plasma MDA, along with significantly lower levels of RBC
SOD, catalyses, and plasma ceruloplasm, all of which are indicators of oxidative stress in the RBCs. Patil
et al. (2006) found similar effects in a group of battery manufacturing workers in India. Levels of MDA
were significantly positively correlated with blood Pb in workers occupationally exposed to Pb, but no
correlation was observed in controls. These effects were also demonstrated in vitro by Ciubar et al.
(2007). but only with concentrations of 2 (.iM (highest concentration tested). In this study, RBCs from
nine volunteers were incubated with Pb at concentrations ranging from 0.1-2 (.iM for 24 hours. Evidence
of lipid peroxidation was also observed in auto repair apprentices in Turkey with blood Pb levels as low
as 7.9 (.ig/dL (compared to 2.6 (.ig/dL in controls) (Ergurhan-Ilhan et al.. 2008). including increases in
glutathione peroxidase (GPx) and MDA, as well as decreases in a-tocopherol and (3-carotene. Decreases
were observed in SOD and CAT, but the results did not achieve statistical significance. Decreased
glutathione reductase (GR) activity was observed in human RBCs incubated with 5-18 (.iM Pb in vitro
(Coban et al.. 2007). Industrial workers in Pakistan occupationally exposed to Pb had a significant
increase in blood Pb levels (mean: 29.1 (ig/dL, range: 9.0 to 61.1 j^ig/dL) compared to age and gender
matched controls (mean: 8.3 (ig/dL, range: 1.0 to 21.7 (ig/dL) (D. A. Khan et al.. 2008). The industrial
workers also had increased oxidative stress as measured by increased levels of serum MDA and
C-reactive protein (CRP). In painters in India with an average 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). Quintanar-Escorza et al. (2007) found elevated RBC lipid peroxidation measured as increased
MDA levels in Pb-exposed workers in a recycled automobile battery factory in Mexico. There was a
correlation in the MDA levels and blood Pb even in the unexposed workers who had low (i.e., <12 (ig/dL)
blood Pb levels, although the observed correlation in exposed workers was greater. Similar effects were
seen when RBCs from healthy volunteers with no Pb exposure were incubated with 0.4 (jM Pb for 24
hours (Quintanar-Escorza et al.. 2010). MDA concentrations, SOD and GPx activities were observed to be
elevated in normotensive exposed workers compared to controls (S. Kasperczvk et al. 2009). The
concentration of MDA was statistically significantly greater in workers with hypertension compared to
both controls and Pb-exposed normotensive workers, whereas the activity of GR in hypertensive workers
decreased to levels comparable to those seen in the control group. Exposure related increases in lipid
peroxidation were also observed in occupationally exposed workers in Poland (S. Kasperczvk et al..
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2005). The concentrations of MDA and 7-ketocholesterol (oxidized cholesterol) were significantly greater
in exposed hypertensive workers compared to normotensive workers.
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 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: group I, 3.93 ± 0.61 jj.g/dl; group II, 7.11 ± 1.25 (ig/dL; and group
III, 11.39 ± 1.39 (ig/dL. Lipid peroxidation, measured as blood MDA, was statistically significant greater
in group III, compared to group II and I, whereas GSH was decreased 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, group 2: >10 (ig/dL). Although the groups
were similar in their age, 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, there was a correlation between the 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 with
neurological disorders. There was a significantly higher blood Pb level in the study population compared
to the control 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 and) 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 demonstrated 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 controls (Y. P.
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 (M. K. Lee et al.. 2005). Other
indications of 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, while simultaneously increasing the activity of SOD and GPx (Kharoubi.
Slimani. Krouf. et al.. 2008). CAT activity was also slightly elevated in Pb-exposed rats, but the increase
failed to reach statistical significance. SOD activity was significantly decreased in rats injected with 15
mg/kg Pb i.p. for seven days, but not rats exposed to 5 mg/kg Pb (Berrahal et al.. 2007). GPx activity and
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MDA concentrations were slightly elevated in the exposure groups, but 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 were 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. (2006). found that 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 2000 ppm Pb in drinking water for 5 weeks,
MDA levels were significantly increased, whereas vitamin E concentrations were significantly decreased
(Cavlak et al.. 2008). In the case of MDA, co-exposure to Pb and a number of sulfur-containing
antioxidants (e.g., L-methionine, N-acetylcysteine, and L-homocysteine) reduced concentrations to a level
not statistically significantly different from controls, but statistically smaller than concentrations observed
with Pb alone. Exposure to L-methionine an N-acetylcysteine also reduced Pb-induced depletion of
vitamin E.
5.7.7. Summary and Causal Determination
There is consistent toxicological and epidemiologic evidence that exposure to Pb induces adverse
effects on hematological endpoints, including altered heme synthesis, decreased RBC survival and
function, and increased RBC oxidative stress. Pb preferentially partitions into RBCs following exposure,
with RBC concentrations approximately 100-fold greater than those observed in the plasma (C. Jin et al..
2008; Timchalk et al.. 2006).
Multiple occupational epidemiologic studies have observed that Pb affects several hematological
parameters such as Hb, PCV, MCV. MCH, and MCHC (Karita et al.. 2005; D. A. Khan et al.. 2008; Patil.
Bhagwat. Patil. Dongre. Ambekar. & Das. 2006; Patil. Bhagwat. Patil. Dongre. Ambekar. Jailkhani. et al..
2006; Ukaeiiofo et al.. 2009). Although the majority of occupationally exposed adults had blood Pb levels
in excess of 20 (ig/dL, decreases in Hb and PCV were observed in adults with blood Pb levels of 7 (ig/dL.
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In addition, Pb exposure was shown to reduce Ca2+- and Ca2+-Mg2+-ATPase activity in RBC membranes,
which leads to an increase in RBC | Ca2 |,. increased membrane fragility, and abnormal morphological
changes (Ciubar et al.. 2007; Huel et al. 2008; Ouintanar-Escorza et al.. 2010; Ouintanar-Escorza et al..
2007). Heul observed a reduction of RBC Ca2+-Mg2+-ATPase activity at a cord blood Pb level of 3.54
(ig/dL. Studies in children are more equivocal than those investigating occupationally exposed adults; this
may due to the comparatively shorter duration of and magnitude of exposure experienced by children.
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 (Baranowska-Bosiacka
et al.. 2009; M. K. Lee et al.. 2005; Sharma et al.. 2010; Simsek et al.. 2009). Pb exposure has also been
observed to increase PS expression on RBC membranes, leading to cell shrinkage, eyrptosis, and
destruction of the RBCs by macrophages (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 blood Pb level of 6.4 (ig/dL; toxicological studies in rats also indicate that Pb is
cytotoxic to RBC progenitor cells. Taken together, these studies provide consistent evidence that exposure
to Pb adversely effects 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.
Recently, numerous epidemiologic studies have confirmed that decreases in RBC ALAD levels and
activity are strongly associated with Pb exposure in adults and children at blood Pb levels as low as 6.4
and 7.1 (.ig/dL. respectively (Ademuviwa. Ugbaia. Oio. et al.. 2005; Mohammad et al.. 2008; Patil.
Bhagwat. Patil. Dongre. Ambekar. & Das. 2006; Patil. Bhagwat. Patil. Dongre. Ambekar. Jailkhani. et al..
2006; Ouintanar-Escorza et al.. 2007). Decreases in blood ALAD activity were also seen in rats with
blood Pb levels of 6.5 (ig/dL (M. K. 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 humans
(Ademuviwa. Ugbaia. Oio. et al.. 2005; Counter et al.. 2007; Mohammad et al.. 2008; Patil. Bhagwat.
Patil. Dongre. Ambekar. Jailkhani. et al.. 2006) and animals (Rendon-Ramirez et al.. 2007). Pb has also
been shown to inhibit the 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 adults with blood Pb levels as low as 7.9
(ig/dL (Ergurhan-Ilhan et al.. 2008; D. A. Khan et al.. 2008; Mohammad et al.. 2008; Patil. Bhagwat.
Patil. Dongre. Ambekar. & Das. 2006; Patil. Bhagwat. Patil. Dongre. Ambekar. Jailkhani. et al.. 2006;
Ouintanar-Escorza et al. 2007). Other measures of oxidative stress observed included lowered activities
of SOD, GR, and CAT, and increased CRP. Indices of RBC oxidative stress were also seen in adolescents
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and children exposed to Pb (Ahamed et al.. 2008; Ahamed et al.. 2006; Y. P. 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 at least partially ameliorate Pb-induced oxidative stress in RBCs (Casadoa
et al.. 2006; Cecil et al.. 2008; 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
(Section 7.4.5). The most consistently observed effect in metal impacted environments is decreased RBC
ALAD activity. This effect has been observed across a wide range of taxa, including bivalves, fish,
amphibians, birds, and mammals. More limited evidence exists regarding deleterious effects of Pb on
serum enzyme levels and white blood cell counts in birds and mammals.
In conclusion, the recent epidemiologic and toxicological literature provides strong evidence that
exposure to Pb is associated with numerous deleterious effects on the hematological system, including
altered heme synthesis mediated through decreased ALAD and ferrochelatase activities, decreased RBC
survival and function, decreased hematopoiesis, and increased oxidative stress and lipid peroxidation. The
consistency of findings in the epidemiologic and toxicological literature and coherence across the
disciplines is sufficient to conclude that there is a causal relationship between Pb exposures and effects
on heme synthesis and red blood cell function.
5.8. Reproductive Effects and Birth Outcomes
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 people with environmentally relevant 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. 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 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). Briefly, the relatively small number of
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studies found consistently positive associations between recent Pb exposure and pregnancy-induced
hypertension.
Overall, the recent reproductive literature continues to support associations reported in earlier
AQCDs between Pb exposure and adverse outcomes on various parameters of sperm (function, motility,
count, integrity, histology). The toxicological and epidemiologic literature also support the finding that Pb
exposure is consistently associated with delayed onset of puberty in both males and females. The new
information from epidemiologic and toxicological studies and conclusions from previous AQCDs are
summarized below.
5.8.1. Effects on Female Reproductive Function
The epidemiologic studies presented on Pb and female reproductive function in the 2006 AQCD
("U.S. EPA. 2006) provided little evidence on the possible associations between Pb exposure and female
reproduction and fertility. However, the 1986 and 2006 Pb AQCDs (U.S. EPA. 1986. 2006) reported
toxicological findings that Pb exposure is 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 AQCD, many epidemiologic studies have been published regarding Pb levels in women
and their effects on reproduction. 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 studies have examined the association between Pb and its effects on hormones and the
estrus cyclce. Epidemiologic studies support the toxicological findings, which have the majority of the
evidence.
An epidemiologic study using the NHANES III database 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) (E. F. Krieg. Jr.. 2007). Deviation from normal FSH and
LH levels may indicate endocrine disruption related to ovary functioning. Researchers determined that as
blood Pb levels increased, serum FSH and LH increased among both post-menopausal women and
women with both ovaries removed. There was also an increasing trend for pre-menopausal women who
were not menstruating or pregnant, although the association was not statistically significant for LH.
Increasing blood Pb levels were associated with decreasing 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 found that the lowest level of blood Pb for which
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a statistically significant association could be observed between blood Pb and FSH was 1.7 (ig/dL among
women with their ovaries removed. A limitation 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.
Another epidemiologic study was performed in Kaohsiung City, Taiwan among two groups of women
aged 23-44: 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 (S. H. Chang et al.. 2006). The mean blood Pb in this study was 3.12 (.ig/dL (SD 0.19 (.ig/dL). The
study reported a positive association between increased blood Pb levels and serum estradiol
concentrations, which reflects ovary activity.
The effect of Pb exposure on the female endocrine system has been demonstrated in toxicological
studies in the 1986 and 2006 Pb AQCD (U.S. EPA. 1986. 2006). 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
previously, Pine et al. (2006) observed that maternal Pb exposure causes a decrease in basal LH levels in
pre-pubertal female Fisher 344 rat pups when compared to non-Pb exposed pups during gestation and
lactation. Dumitrescu et al. (2008) observed alteration of hormone levels in female Wister rats after
ingesting Pb acetate (50, 100, 150 ppb) in drinking water for six months. 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, 3J3-HSD
and 17J3-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.
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 that Pb acetate
concentrations of 0.046 mg/mL and 0.063 mg/mL statistically significantly inhibited 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 & Gupta.
2006).
No recent toxicological studies were found that addressed Pb-induced effects on the estrus cycle.
Overall, toxicological studies report alterations in hormone levels related to Pb concentration. This
was also observed in epidemiologic studies. Although these changes are observed, there are discrepancies
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about the direction of the hormone changes related to Pb. One explanation is that the direction of effect
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. 2006). Recent epidemiologic studies and studies in experimental animals support this
finding. The epidemiologic studies are summarized in Table 5-27.
Table 5-27. Summary of recent epidemiologic studies of effects on fertility for females
Reference
Study' Studv
,LSSSi OU,COme P°P^0n
Exposure Mean Pb (SD) Adjusted Effect
Measurement in |jg/dL Estimates
Al-Saleh et al.
(2008)
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 |jg/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,
099)
Chang et al. (2006) Kaohsiung Fertility
City, Taiwan
1999
2000-2001
Women receiving Blood Pb 3.12(0.19) OR (95% CI)
care at a infertility Fertility
clinic in 2000-2001 or <2.5 |jg/dl: 1.00 (Ref)
delivering a normal >2.5 |jg/dl: 2.94 (1.18,
infant at a nearby 7.34)
medical center in
1999
Silberstein et al.
(2006)
Providence,
NS
Rl Achieving
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
Epidemiologic studies examined women having difficulty conceiving by performing studies
among patients of fertility clinics or undergoing in vitro fertilization (IVF). The first of these was
performed in Kaohsiung City, Taiwan among women aged 23-44 (S. H. Chang et al.. 2006). 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
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records of a postpartum care unit at a medical center. Higher odds of infertility were observed when
comparing women with blood Pb levels >2.5 (ig/dL to those with blood Pb levels < 2.5 (ig/dL. Another
study examining fertility reported on women in Saudi Arabia aged 19-50 years who were undergoing IVF
treatment (Al-Saleh et al.. 2008). Women were categorized as having achieved a pregnancy versus not
achieved a pregnancy and achieved fertilization versus not achieving fertilization. The majority of women
had follicular Pb levels that were below the level 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 an inverse association was detected between
blood Pb and fertilization. Finally, a study that included nine women undergoing IVF treatment in Rhode
Island (Silberstein et al.. 2006) found that median follicular Pb levels in women who achieved pregnancy
were lower than the follicular Pb levels among non-pregnant 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 studies are not a sample of the general population and therefore cannot be
generalized to all women of childbearing age.
Several studies observed a decrease in litter size when females were exposed to Pb before mating
or during pregnancy (Dumitrescu. Alexandra, et al.. 2008; lavicoli. Carelli. Stanek. Castcllino. Li. et al..
2006; Teiion et al.. 2006). Pups in Teijon et al.'s study receiving 400 ppm Pb acetate in drinking water had
blood Pb of 97 (.ig Pb/dL blood at 1 wk post-weaning and 18.2 (ig Pb/dL blood at 2 wk 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. (1998). who did not observe differences in sex ratio if liters
from females 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 dose-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 g/L 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 0.05 mg/kg body weight daily before
mating and during pregnancy with a resulting blood Pb of 2.49 (ig/mL. Although reproductive
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1 performance was not affected in this study, the authors did report an alteration in implantation enzymes.
2 Cathepsin-D activity decreased and alkaline phosphatase activity increased after Pb exposure.
3 Epidemiologic and toxicological studies on the effect of Pb on fertility outcomes have generated
4 inconsistent results. However, there is some indication that increased Pb exposure may decrease fertility.
5.8.1.3. Effects on Puberty
5 Recent toxicological studies of rodents have examined the effects of Pb on pubertal and
6 reproductive organ development and on biomarker development. There have also been recent
7 epidemiologic studies examining Pb levels and onset of puberty, which are summarized in Table 5-28 and
8 in the text below.
Table 5-28. Summary of recent epidemiologic studies of effects on puberty for females
Reference
Study Location
and Years
Outcome
Study
Population
Exposure Mean Pb
Measurement (SD) in
|jg/dL
Adjusted Effect
Estimates
Denham et al.
(2005)
Akwesasne Mohawk
Nation (boundaries of
New York, Ontario,
and Quebec
NS
Age at 10- to 16.9-yr-old girls Blood Pb 0.49 (0.905)
menarche in the Akwesasne
community Median: 1.2
Coefficients for binary
logistic regression
predicting menarche with
Pb centered at the mean:
log blood Pb -1.29 (p-value
0.01)
log blood Pb -squared: -
1.01 (p-value 0.08)
Non-linear relationship
observed and Pb below the
mean did not appear to
affect the odds of
menarche. Increasing
blood Pb from 0.49 to 0.98
|jg/dL decreased the odds
of menarche attainment by
72%
Gollenberg et
al. (2010)
U.S.A.
1988-1994
Luteinizing
hormone (LH)
and inhibin B
Girls ages 6-11 from Blood Pb
the NHANES III study
Median 2.5
(range 0.07,
29.4)
blood Pb
>10 |jg/dL: 5%
OR (95% CI) for exceeding
pubertal inhibin B cutoff
(>35pg/mL)
<1 |jg/dl: 1.00 (Ref)
1-4.9 |jg/dl: 0.38 (0.12,
1.15)
> 5|jg/dl: 0.26 (0.11, 0.60)
OR (95% CI) for exceeding
pubertal LH cutoff (>0.4
mlU/mL)
<1 |jg/dl: 1.00 (Ref)
1-4.9 |jg/dl: 0.98 (0.48,
1.99)
> 5|jg/dl: 0.83 (0.37,1.87)
*a sensitivity analysis
including only those with
blood Pb <10 |jg/dl had
similar results butORs
were slightly attenuated
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Reference
Study Location
and Years
Outcome
Study
Population
Exposure Mean Pb
Measurement (SD) in
|jg/dL
Adjusted Effect
Estimates
Naickeretal.
(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 Pb at 13 yr 4.9(1.9)
of age blood Pb levels
> 10 |jg/dL: 1%
OR (95% CI)
Delay in breast
developmental age 13
<5 |jg/dl: 1.00 (Ref)
> 5|jg/dl: 2.34 (1.45,3.79)
Delay in pubic hair
development at age 13
<5 |jg/dl: 1.00 (Ref)
> 5|jg/dl: 1.81 (1.15,2.84)
Delay in attainment of
menarche at age 13
<5 |jg/dl: 1.00 (Ref)
> 5|jg/dl: 2.01 (1.38,2.94)
Selevan et al.
(2003)
U.S.A.
1988-1994
Tanner
staging and
age at
menarche
Girls ages 8-18 from Blood Pb
the NHANES III study
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%
Blood Pb
levels>10|jg/dL:
NHWhites:
0.3%
NHBIacks: 1.6%
Mexican-
Americans:
2.3%
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)
Pubic hair development
NH Whites:
1 |jg/dl: 1.00 (Ref)
3|jg/dl: 0.75 (0.37,1.51)
NH Blacks:
1 |jg/dl: 1.00 (Ref)
3|jg/dl: 0.62 (0.41, 0.96)
Mexican Americans:
1 |jg/dl: 1.00 (Ref)
3|jg/dl: 0.70 (0.54, 0.91)
HR (95% CI) Included only
girls 8-16
Age at menarche
NH Whites:
1 |jg/dl: 1.00 (Ref)
3|jg/dl: 0.74 (0.55,1.002)
NH Blacks:
1 |jg/dl: 1.00 (Ref)
3|jg/dl: 0.78 (0.63, 0.98)
Mexican Americans:
1 |jg/dl: 1.00 (Ref)
3|jg/dl: 0.90 (0.73,1.11)
Tomoumetal. Cairo, Egypt Hormones Healthy children aged Blood Pb NS for girls only
(2010) 2007 and pubertal 10-13 seeking (combined with
development treatment for minor boys in the
health problems and study the mean
living in one of two was 9.46 (3.08))
designated areas (one
with high-risk for Pb
contamination and
one with no Pb
source)
Breast Development
<10 |jg/dl:
Stage 2: 36.4%
Stage 3: 63.6%
> 10|jg/dl:
Stage 2:100%
Stage 3: 0%
Chi-square p-value<0.01
Pubic Hair Development
<10 |jg/dl:
Stage 2: 36.4%
Stage 3: 63.6%
> 10|jg/dl:
Stage 2: 77.8%
Stage 3: 22.2%
Chi-square p-value>0.05
•Quantitative results for
hormones not provided
Wolff etal. New York City, NY Pubertal 9-yr old girls from the Blood Pb
(2008): Wolf et 1996-1997 stages defined study hospital and
al.(2007) using nearby pediatric
standard offices
drawings
Median: 2.4
PR (95% CI) (unit not
given, assume results are
per 1 ua/dL)
Breast stage: 1.01 (0.79,
1.30)
Pubic hair stage: 1.25
(0.83,1.88)
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Reference Study Location Outcome Study Exposure Mean Pb Adjusted Effect
and Years Population Measurement (SD) in Estimates
|jg/dL
Wuetal. U.S.A. Tanner Girls ages 8-16 from Blood Pb 2.5(2.2) OR (95% CI)
(2003) 1988-1994 staging and the NHANES III study Breast development
age at Weighted 0.7-2.0 |jg/dl: 1.00 (Ref)
menarche proportion of the 2.1-4.9 |jg/dl: 1.51 (0.90,
sample with 2.53)
blood Pb 5.0- 5.0-21.7 jjg/dl: 1.20 (0.51,
21.7:5.9% 2.85)
Pubic hair development
0.7-2.0 |jg/dl: 1.00 (Ref)
2.1-4.9 |jg/dl: 0.48 (0.25,
0.92)
5.0-21.7 |jg/dl: 0.27 (0.08,
0.93)
Menarche
0.7-2.0 |jg/dl: 1.00 (Ref)
2.1-4.9 |jg/dl: 0.42 (0.18,
0.97)
5.0-21.7 |jg/dl: 0.19 (0.08,
043)
Several epidemiologic studies investigated the association between blood Pb and indicators of
puberty onset. A study performed in NYC among 9 year old girls reported no association between Pb
levels and pubertal development (Wolff et al.. 2008). However, a study among girls aged 10-13 (median
age 12) reported decreased 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 non-linear positive association between
blood Pb and age at menarche (Denham et al.. 2005). No association was observed below blood Pb of
0.49 (ig/dL. A study conducted in South Africa reported a positive association between blood Pb levels
and age at menarche and pubertal development (Naicker et al.. 2010). Blood Pb levels were associated
with delayed pubertal development and later age at menarche. This study illustrates not only the
association between Pb and pubertal development, but that delays can occur at low Pb levels. Multiple
studies have been performed examining blood Pb levels and puberty using the NHANES III database
(Gollenberg et al.. 2010; Sole van et al.. 2003; T. Wu et al.. 2003). One study included girls aged 8-16 and
reported an positive association for delayed attainment of menarche and pubic hair development, but not
for breast development (T. Wu et al.. 2003). The associations were observed even at low levels of blood
Pb. Another NHANES III study included girls 8-18 years of age and reported the results stratified by race
(Selevan et al.. 2003). Blood Pb levels were inversely associated with Tanner stage of breast and pubic
hair development and age at menarche among African Americans and with breast and pubic hair
development among Mexican Americans. The associations were in the same directions for whites, but
none of the associations reached statistical significance. A third study using the NHANES III database
examined the association between Pb and reproductive hormones among girls 6-11 years old (Gollenbem
et al.. 2010). Blood Pb levels were inversely associated with inhibin B, a protein that inhibits FSH
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production, but no association was observed for LH. 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 these hormones are, "believed to be relevant
for younger girls... near the onset of puberty and... serve as markers for hypothalamic-pituitary-gonadal
functioning."
Puberty; Neonate/adult; Mouse; Female
(lavicoli. Carelli. Stanek. Castellino. Li. et al..
2006)
Neurotransmitter; Adult; Mouse; Both
(Leasure et al.. 2008)
Physical development; Adult; Mouse; Male
(Leasure et al.. 2008)
o Highest Concentration
~ LowestCone, with Response
A Highest Cone, with No Response
o Lowest Concentration
Eye; Adult; Rat; Both
(Fox et al.. 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)
100
Blood Pb Level (|jg/dL)
Figure 5-54. Toxicological Exposure-Response Array for Reproductive
Effects of Pb.
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Table 5-29. Toxicological Exposure-Response Array Summaries for Reproductive Effects
of Pb presented in Figure 5-54
Reference
Blood Pb level
with Effect (pg/dl)
Altered Outcome
Iavicoli I. etal. (2006)
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 etal. (1998)
25-130
Semen quality affected
Teijon etal. (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 etal. (1998)
25-130
Semen quality affected
Teijon etal. (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 confirm these findings and show that puberty onset is one of
the more sensitive markers of Pb exposure as is demonstrated in the exposure response array (Table 5-29
and Figure 5-54). Dumitrescu et al. (2008) exposed adult Wister 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 dose-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 dose-dependent relationship.
In another recent study, Iavicoli et al. (2006) reported a statistically significant delay in several
biomarkers of sexual maturity in offspring (F, generation) born to dams that ingested 3.5-40 ppm in their
daily diet. Maternal ingestion of Pb at the various doses resulted in female pup blood Pb levels of 3.5-13
(ig/dL. For all diet groups, there was a delay in age at vaginal opening, age of first estrus, age of vaginal
plug formation, and age of first parturition. 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) induced statistically significant acceleration
of markers of sexual maturation in female offspring versus background Pb level animals (blood Pb of 2
(ig/dL) animals. There were statistically significant increases in time of vaginal opening (30% increased),
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 dose-dependent fashion with very low dose Pb having
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a statistically significant earlier onset of puberty than the background Pb animals. 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 animals continued to be exposed to same concentrations of Pb
over multiple generations through the diet. Data results of 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 female puberty. The authors reported that offspring
exposed to Pb during gestation and lactation (12 mg/mL; mean maternal blood Pb level 40 (ig/dL)
exhibited a marked increase in LH and luteinizing hormone releasing hormone (LHRH) secretion after
IGF-1 administration (200 ng7(.iL) resulting in restored timing of vaginal opening such that they were the
same as control. 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. EPA. 2006) 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 a positive association between 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 to be IGF-1-dependent.
5.8.1.4. Summary of Effects on Female Reproductive Function
In summary, Pb exposure affects female reproductive function as demonstrated by both
epidemiologic and toxicological studies. At low Pb levels, associations are observed with delayed puberty.
Some evidence is also available regarding Pb levels and altered hormone levels as well as decreased
fertility, although studies reported inconsistent findings for the later.
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5.8.2. Effects on Male Reproductive Function
The 2006 Pb AQCD ("U.S. EPA. 2006) reported on male Pb exposure/levels and reproductive
functions as measured by sperm count/motility/morphology, time to pregnancy, reproductive history, and
chromosomal aberrations. Despite limitations in many of the studies, most of the studies found slight
associations between high Pb levels (i.e. > 45 j^ig/dL) and reduced male fecundity or fertility (U.S. EPA.
2006). Evidence provided in the 1986 Pb AQCD (U.S. EPA. 1986) 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.
5.8.2.1. Effects on Sperm/Semen Production, Quality, and Function
Multiple epidemiologic and toxicological studies have examined the relationship between Pb and
sperm and semen production, quality, and function. These studies are summarized in the text below. In
addition, recent epidemiologic studies are included in Table 5-30.
Table 5-30. Summary of recent epidemiologic studies of effects on sperm and semen
Reference
Study
Location Study
and Population
Years
Exposure
Measurement
Mean Pb
(SD) in
|jg/dL
Adjusted Effect Estimates
Hsu et al. (2009) Taiwan
NS
Men working at a
battery plant
Blood Pb
Categorized into 3 groups:
<25 |jg/dl, 25-45 ^ig/dl, >45
|jg/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)
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Reference
Study
Location Study
and Population
Years
Exposure
Measurement
Mean Pb
(SD) in
|jg/dL
Adjusted Effect Estimates
Kasperczyketal. Poland
(2008) 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
|jg/dl), low exposed workers
(blood Pb 25-40 |jg/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
Meeker etal. Michigan Men aged 18-55 Blood Pb
(2008) NS going to infertility
clinics (distinction
not made
between clinic
visits for male or
female fertility
issues)
Median: 1.50 OR (95% CI) for having below reference-
(IQR 1.10, 2.00) 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)
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Reference
Study
Location Study
and Population
Years
Exposure
Measurement
Mean Pb
(SD) in
|jg/dL
Adjusted Effect Estimates
Naha and Kolkata,
Chowdhury (2006) India
NS
Men aged 31 -45
that were non-
occupationally
exposed 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 13.62
(2.45)
Low exposure
48.29 (4.91)
High exposure
77.22 (1.25)
Semen Pb
measurement
Controls 3.99
(1.36)
Low exposure
10.85(0.75)
High exposure
18.30(2.08)
p-values for difference across the three
groups for mean values of semen profiles
were <0.01 for: sperm count, sperm
protein, sperm DNA hyploidy, sperm DNA,
sperm RNA, sperm viability, sperm
membrane lipid peroxidation, seminal
plasma total ascorbate, seminal plasma
DHAA, sperm ATPase activity, sperm
motility, sperm velocity, seminal plasma
fructose
Naha and Manna Bangalore,
(2007) India
NS
Non-
occupationally
exposed 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)
p-values for difference across the three
groups for mean values of semen profiles
were <0.01 for: liquefaction time, seminal
volume, sperm count, sperm DNA
hyploidy, sperm morphological abnormality,
sperm motility, sperm ATPase activity,
seminal plasma fructose, seminal plasma
total protein, seminal plasma free amino
acid, seminal plasma cholesterol
Slivkova etal.
(2009)
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
Telisman et al. Croatia Men aged 19-55, Blood Pb
(2007) 2002-2005 never
occupationally
exposed to metals
and going to a
clinic for infertility
examination or for
semen donation
to be used for
artificial
insemination
Median: 4.92 Standardized regression coefficients for
(range 1.13- log blood Pb (units not given)
14.91) 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
1 International epidemiologic studies of men occupationally exposed to Pb have reported on
2 associations between Pb levels and semen/sperm count and quality. Most of these studies included
3 individuals occupationally exposed to Pb and have reported blood Pb levels over 40 (ig/dL. For example,
4 two studies performed in India (Naha & Chowdhury. 2006; Naha & Manna. 2007) reported that men in
5 the highest exposure group (men working in battery or paint manufacturing plants for 10-15 years for 8
6 hours/day) had mean blood Pb levels of 77.22 (ig/dL in one study (Naha & Chowdhury. 2006) and 68.26
7 in the other study (Naha & Manna. 2007). Control groups in these studies (those without occupational Pb
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exposure) had 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 increased hyploidy of sperm DNA
and morphologic abnormalities (Naha & Chowdhurv. 2006; Naha & Manna. 2007). Decreased viability
and increased lipid peroxidation were detected (Naha & Chowdhurv. 2006). A study performed in Taiwan
reported that men with greater 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
(P. C. Hsu et al.. 2009). 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 Poland
included employees exposed to Pb and compared them with a group of male office workers (A.
Kasperczvk 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 induces tissue damage in sperm via reactive oxygen species, was
greater in the highest exposure group compared to the controls.
One study performed in Croatia recruited men who had never been occupationally exposed to
metals (Telisman et al.. 2007). Increased blood Pb was associated with increased percent 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 biomarkers for Pb (erythrocyte protoporphyrin
and S-aminolevulinic acid dehydratase [ALAD]) were used instead.
Two studies examined Pb levels and semen quality of men at infertility clinics (Meeker et al.. 2008;
Slivkova et al.. 2009). Meeker et al. (2008) detected no associations between increases in blood Pb and
semen concentration, morphology, or motility (although a slight positive trend was observed between
increasing Pb levels and motility in unadjusted models). In models that include multiple metals, blood Pb
was associated with being below the WHO's 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.
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 induced statistically significant effects on semen quality and
testicular pathology at blood Pb of 16-24 (ig/dL (Moorman et al.. 1998). Recent studies confirm earlier
studies that Pb alters sperm parameters such as sperm count, viability, motility, and morphology. 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
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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 dose-
dependent fashion (8-24 mg/kg body weight), reduced the length of certain stages of the spermatogenic
cycle of rat seminiferous tubules and thus affected spermatogenesis. Reshma Anjum et al. 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). Pb induced morphological abnormalities in sperm in a
dose-dependent manner (Allouche et al.. 2009; Massanvi et al.. 2007; Oliveira et al.. 2009; Salawu et al..
2009; Shan et al.. 2009; Tapisso et al.. 2009; C. H. Wang et al.. 2006). Sperm abnormalities reported after
Pb exposures are 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 the above studies
observed changes in every parameter. This may be due to the use of different strains or species, chemical
form of the Pb compound administered, dosage schedule, duration of exposure, and age of animals at the
time of the study (Oliveira et al.. 2009).
Data from recent studies 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 catalse (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 ROS production and prevented the 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
reduction in sperm count, sperm motility, and sperm viability (Madhavi et al.. 2007; Rubio et al.. 2006;
Salawu et al.. 2009; Shan et al.. 2009; C. H. Wang et al.. 2006).
Recent studies also demonstrate that Pb may be directly toxic to mature spermatozoa (Hernandez-
Ochoa et al.. 2006; Tapisso et al.. 2009) as well as primary spermatocytes (Nava-Hernandez et al.. 2009;
Rafique et al.. 2009). Nava-Hernandez et al. had two Pb exposure groups via drinking water. 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 j^ig/dL); no statistically significant difference in blood Pb levels existed
between the two Pb exposure groups because the L2 group drank less water than the LI group. Piao et al.
(2007) report Pb exposure caused DNA damage to sperm; the Pb exposure group had blood Pb of 67 jj.g/1.
Piao et al. (2007) looked at 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
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significantly lower than in Pb alone group. Hernandez-Ochoa et al. (2006) report that Pb reaches the
sperm nucleus in the epididymis of mice chronically exposed to Pb (mean blood Pb 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 intraperitoneal injection. Nava-Hernandez (2009)
reported a dose-dependent increase in DNA damage in rat primary spermatocytes after a 13-week
exposure period to Pb acetate in drinking water (mean blood Pb levels between 19.5 and 21.9 (ig/dL).
Rafique et al. (2009) reported degenerative changes from pyknosis to apoptosis in primary spermatocytes.
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 (C. H. Wang et al..
2006). Dong et al. (2009) reported a dose-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. (2006)
observed a dose-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 (Shan et al.. 2009; C. H. Wang et al.. 2006) suggesting
ROS generation is a major contributing factor in decreased male fertility observed after chronic Pb
exposure.
Similar to the results summarized in previous AQCDs, recent epidemiologic and toxicological
studies report negative effects of high levels of Pb on sperm and semen. Future studies will aid in
determining whether this association is observed at lower Pb levels.
5.8.2.2. Hormone Levels
The 2006 Pb AQCD (U.S. EPA. 2006) provided evidence that Pb acts as an endocrine disruptor in
males at various points along the hypothalamic-pituitary-gonadal axis. The 2006 document also reported
inconsistencies in the reported effects of Pb exposure on circulating testosterone levels. Recent
epidemiologic and toxicological studies are reported below. Epidemiologic studies are summarized in
Table 5-31.
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Table 5-31. Summary of recent epidemiologic studies of effects on hormones for males
Study
Reference Location Outcome
and Years
stun,, Dnnnhtinn Exposure MeanPb Adjusted Effect
Study Population Measurement (SD) in pg/dL Estimates
Meeker etal. Michigan
(2008) NS
FSH, LH, Inhibin Men aged 18-55 going to Blood Pb Median: 1.50
B, testosterone, infertility clinics (distinction not (IQR 1.10, 2.00)
SHBG, FAI, made between clinic visits for
testosterone/LH male or female fertility issues)
Regression coefficients
(95% CI)
FSH
1st quartile: 1.00 (ref)
2nd quartile: 0.13 (-0.10,
0.37)
3rd quartile: 0.10 (-0.15,
0.35)
4th quartile: 0.07 (-0.18,
0.31)
LH
1st quartile: 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
Reference Location Outcome
and Years
stun,, Dnn..hHnn Exposure MeanPb Adjusted Effect
Study Population Measurement (SD) in pg/dL Estimates
Naha and Bangalore, FSH, LH,
Manna (2007) India testosterone
NS
Non-occupationally exposed
controls and occupationally
exposed workers
Categorized by Blood Pb
work history as measurement
controls, low Controls 10.25
exposure (7-10 yr (2.26)
of exposure for 8 Low exposure
h/day) and high 50.29 (3.45)
exposure (>10 yr High exposure
of exposure for 8 68.26 (2.49)
h/day)
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
Telisman et al. Croatia
(2007) 2002-2005
FSH, LH, Men aged 19-55, never
testosterone, occupationally exposed to
estradiol, metals and going to a clinic
prolactin 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)
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 was associated with increasing serum
testosterone and estradiol but decreasing serum prolactin (Telisman et al.. 2007). In addition, the
interaction term of 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 (Naha & Manna. 2007). Among men recruited from infertility clinics in Michigan, median
blood Pb levels were much lower than those observed in the other studies of Pb and hormone levels
among men (Meeker et al.. 2010). No association was detected between blood Pb and levels of FSH, LH,
InhibinB, 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.
In a recent toxicological study, Rubio et al. (2006) observed a decrease in testosterone levels in Pb
acetate-treated rats in a dose-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). Pandya et al. (2010) reported altered hepatic sterodogenic enzyme activity. 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
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steroidogenic enzyme, A5-3J3-HSD activity and serum corticosterone levels indicating disruption of the
adrenocortical process. Dose-dependent decreases in serum testosterone were reported in Pb-exposed
male rats (Reshma Anium et al.). 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. 2006).
Overall, recent epidemiologic and toxicological studies report mixed outcomes regarding hormone
aberrations associated with Pb exposure. These results are similar to those from the 2006 Pb AQCD on
the effects of Pb exposure on circulating testosterone levels.
5.8.2.3. Fertility
Epidemiologic studies have been performed comparing Pb and infertility in men. A study
conducted in Turkey reported blood and seminal plasma Pb levels were different in fertile and infertile
men [mean blood Pb 23.16 (ig/dL (SD 5.59 (ig/dL) for fertile men and 36.82 (ig/dL (SD 12.30 (ig/dL) for
infertile men (p-values for ANOVA <0.001)] (Kiziler et al.. 2007). 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 studies are not a sample of the general
population regarding fertility.
5.8.2.4. Puberty
Research has also been published examining the association between blood Pb and onset of puberty
in males. These studies are summarized in Table 5-32.
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Table 5-32. Summary of recent epidemiologic studies of effects on puberty for males.
Reference
Location
Study Exposure Mean Pb (SD) Adjusted Effec, Estimates
Population Measurement in |jg/dL
Hauseretal.
(2008)
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 |jg/dl:
3%
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 |jg/dl: 1.00 (Ref)
> 5 |jg/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 |jg/dl: 1.00 (Ref)
> 5 |jg/dl: 0.74 (0.34,1.60)
Tomoum et al.
(2010)
Cairo, Egypt
2007
Hormones and
pubertal
development
Healthy children
aged 10-13
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 boys only
(combined with girls
in the study the
mean was 9.46
(3.08))
Testicular size
<10 |jg/dl:
Stage 1:0%
Stage 2: 44.4%
Stage 3: 55.6%
> 10|jg/dl:
Stage 1:33.3%
Stage 2: 66.7%
Stage 3: 0%
Chi-square p-value<0.01
Pubic Hair Development
<10 |jg/dl:
Stage 1:0%
Stage 2: 55.6%
Stage 3: 44.4%
> 10|jg/dl:
Stage 1:33.3%
Stage 2: 66.7%
Stage 3: 0%
Chi-square p-value<0.05
Penile staging
<10 |jg/dl:
Stage 1:11.1%
Stage 2: 44.4%
Stage 3: 44.4%
> 10|jg/dl:
Stage 1:58.3%
Stage 2: 41.7%
Stage 3: 0%
Chi-square p-value<0.05
Mean testosterone level
<10 |jg/dl:
4.72 (SD 1.52)
> 10|jg/dl:
1.84 (SD 1.04)
•Quantitative results for LH and
FSH not provided
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Williams etal. Chapaevsk, Pubertal stages Healthy boys Blo<
(2010) Russia defined using aged 8-9 at 8-9
2003-2008 standard enrollment who
drawings had annual
follow-up
evaluations
Blood Pb at ages Median: 3 (IQR 2-5) HR (95% CI)
blood Pb level >10 Pubertal onset based on testicular
|jg/dl: 3%
volume
<5 |jg/dl: 1.00 (Ref)
> 5 |jg/dl: 0.73 (0.55, 0.97)
Genital development
<5 |jg/dl: 1.00 (Ref)
> 5 |jg/dl: 0.76 (0.59, 0.98)
Pubic hair development
<5 |jg/dl: 1.00 (Ref)
> 5 |jg/dl: 0.69 (0.44, 1.07)
Studies were performed among a cohort of Russian boys enrolled between ages 8-9 (Hauser et al..
2008; P. L. Williams et al.. 2010). Both the cross-sectional study (Hauser et al.. 2008) and the prospective
study with annual follow-ups (P. L. Williams et al.. 2010) demonstrated an association between increased
blood Pb levels and later onset of puberty. In addition, in a study of boys and girls 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 Pb group, those boys with higher blood Pb had lower
testosterone, FSH, and LH levels.
Thus, recent studies have demonstrated a negative effect of Pb on pubertal development among
boys that exists even at low 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. 2006)
supported 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 histological changes in the testes as well as germ cells.
Histological changes of testes in Pb-treated animals included seminiferous tubule atrophy,
Sertoli cell and Leydig cell shrinkage with pyknotic nuclei (Shan et al.. 2009; C. H. Wang et
al.. 2006). dilatation of blood capillaries in the interstitium, undulation of basal membrane,
and occurrence of empty spaces in seminiferous epithelium ( Massanyi et al.. 2007).
5.8.2.6. Summary of Effects on Male Reproductive Function
Associations between Pb exposure and male reproductive function vary by outcome. The strongest
evidence of an association is the relationship observed between high levels of Pb and negative effects on
sperm and semen in both recent epidemiologic and toxicological studies and previous AQCDs. Recent
toxicological studies also reported an association between Pb exposure and histological changes in the
testes and germ cells. In addition, recent epidemiologic studies found Pb exposure to be associated with
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delayed pubertal development at low blood Pb levels. This is supported by earlier toxicological studies.
Similar to the 2006 Pb AQCD, 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.
5.8.3. Effects on Ovaries, Embryo Development, Placental
function, and Spontaneous Abortions
The 2006 Pb AQCD ("U.S. EPA. 2006) included studies of Pb exposure among men and women and
their associations with spontaneous abortions. The 2006 AQCD concluded that overall there was little
evidence to support an association between Pb levels among women and spontaneous abortion (U.S. EPA.
2006). Most of the studies examined in the 2006 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 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 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.
Table 5-33. Summary of recent epidemiologic studies of effects on spontaneous
abortions.
study study Exposure Pb Adjusted Effect
Location outcome popu|atjon Measurement Estimates
Median (IQR): Median Placenta Pb:
25.8 (21.0, 36.8) Women who had not
previously miscarried: 27
|jg/kg
Women who had previously
miscarried: 39 |jg/kg
(p-value for difference:
0.039)
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)
Gundackeretal. Vienna, Previous Women recruited Whole placentas
(2010) Austria miscarriage during the second shortly after birth
2005 trimester of
pregnancy
Lamadrid-Figueroa
etal. (2007)
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)
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Yinetal. (2008)
Shanxi
Province,
China
2004-2006
Anembryonic Women age 25-35 Maternal blood Pb
pregnancy yr old and at 8-12 after miscarriage for
Cases: 5.3 (95%
CI: 5.2, 5.9)
Comparisons between log-
transformed blood Pb levels
of cases and controls
weeks of gestation cases and at study
at study entry; enrollment for
cases were controls
anembryonic
pregnancies and
controls were
normal pregnancies
that ended in a live
birth between 37-
42 weeks
Controls: 4.5
(95% CI: 3.7,
5.0)
performed via Student's t-
test had a p-value of 0.03
Table 5-33, above, provides a summary of the recent epidemiologic studies examining the
association between Pb 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 Pb was inversely correlated with folate and vitamin B12, but this correlation was not
observed among those who delivered at term; no models examining Pb levels were controlled for nutrient
status. 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 women in the lowest tertile. The authors state that the plasma/whole blood ratio represents the
availability of Pb 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 although the number of women included in the
study was small (only 8 women reported previously having a miscarriage) (Gundacker et al.. 2010).
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 a dose-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.
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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. 2006). Wang et al. (2009) observed that elevated maternal blood Pbs (0.6-1.74 (.iM)
compared to control (0.04 (.iM) 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
dose-dependent increasing pathology of cytoarchitecture and cytoplasmic organelles. The authors also
reported a positive expression of NF-kB, a transcription factor that controls the expression of genes
involved in immune responses, apoptosis, and cell cycle, in the cytotrophoblasts, decidual cells, and small
vascular endothelial cells in rat placenta under a low Pb level exposure condition which correlated to
blood 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. that had dose-dependent decreases in serum testosterone, decreased
male reproductive organ weight and decreased sperm were mated to untreated females. These untreated
dams had dose-dependent decreased implantation rate and higher pre- and post-implantation loss,
indicating paternally mediated fetal loss.
As observed in sperm cells, Pb stimulates changes in antioxidant enzyme activity in rat ovaries
indicating that a contributing factor in Pb-induced ovarian toxicity may be oxidative stress. 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 histological changes in ovarian cells of pigs (kolesarova et al.. 2010)
and rats (Nampoothiri et al.. 2007; Nampoothiri & 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.
2006).
Overall, the recent studies support the conclusions of the last Pb AQCD that there is insufficient
evidence among epidemiologic studies to suggest an association between Pb and spontaneous abortions.
In addition, 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
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increased risk of early spontaneous abortions. Toxicological data provide mechanistic understanding of
the contribution of Pb 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 up to the current date provide mixed effects on
the role of Pb in spontaneous abortions.
5.8.4. Infant Mortality and Embryogenesis
The 2006 AQCD ("U.S. EPA. 2006) 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. 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; Uzbekov et al.. 2007; Y.-Y. Wang et al.. 2009). No recent
epidemiologic studies have reported on the relationship between Pb levels and infant mortality.
5.8.5. Birth Defects
The 2006 Pb AQCD (U.S. EPA. 2006) reported the possibility of small associations between high
Pb exposure and birth defects, but many of the studies used occupational histories instead of actual
measures of blood Pb levels. Among the studies included in the 2006 AQCD, two studies reported
possible associations between parental exposure to Pb and neural tube defects (Bound et al.. 1997; I mens
etal.. 1998). Recent studies also examined Pb levels and neural tube defects (Table 5-34). No other recent
studies of Pb levels/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 AQCD
toxicological studies demonstrated associations between exposure to high doses of Pb and increased
incidences of teratogenic effect in experimental animals.
Table 5-34. Summary of recent epidemiologic studies of effects on neural tube defects.
Reference
Study
Location
Study Exposure Mean Pb (SD) in
Population Measurement M9^L
Adjusted Effect
Estimates
Brenderetal.
(2006)
Texas
1995-2000
Infants of Mexican- Maternal blood Pb taken Cases: 2.4 (1.9)
American women 5-6 wks post-partum Controls: 2.5 (1.6)
Blood Pb<6.0 |jg/dL: 1.0 (Ref)
Blood Pb>6.0 |jg/dL: 1.5 (95%
CI: 0.6, 4.3)
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Zeyreketal. (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
Among the recent epidemiologic studies (described in Table 5-34), 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. These measurements were taken 5-6
weeks postpartum, which is a limitation of this study because the levels may be different than during the
developmental period of gestation. The OR comparing those with at least 6 (ig/dL Pb to those with less
than 6 (ig/dL Pb was 1.5 (95% CI: 0.6, 4.3). This increased after adjusting for breast feeding, although
this variable was not a confounder due to its inability to be associated with neural tube defects. For these
women neither occupational exposure to Pb or proximity of residence to a facility with Pb air emissions at
the time of conception were associated with increased odds of neural tube defects.
Overall, in contrast to studies from the 2006 AQCD, recent studies of Pb and neural tube defects
observed no associations.
5.8.6. Preterm Birth
Research on preterm birth included in the 2006 Pb AQCD (U.S. EPA. 2006) reported inconsistent
findings regarding the relationship between Pb and gestational age. Recent studies have examined this
potential association and again mixed results were reported (Table 5-35). Of these studies, the ones that
categorized births as preterm or term all defined preterm birth as less than 37 weeks of gestation. One
limitation to note for these studies is that if Pb affects spontaneous abortion and length of gestation via a
similar pathway, then the studies that only collect data at delivery and not at earlier stages of pregnancy
would be biased towards the null.
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Table 5-35. Summary of recent epidemiologic studies of effects on preterm birth.
Reference
Study
Location
Outcome
Study
Population
Exposure
Measurement
Mean Pb
(SD) in
g/dL
Adjusted Effect
estimates
Berkowitz et al.
(2006)
Idaho
1970-1981
Preterm birth
(<37wk)
Singleton infants
with 28-45 wk
gestation
Three time periods of two NS
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"
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 >5 |jg/dL: mean gestational age
blood Pb: 4.7 38 wks
(12.1) <5 |jg/dL: mean gestational age
39 wks
Linear regression: gestational
age decreased 1 wk with every 1
|jg/dL increase in umbilical cord
blood Pb (exact values and 95%
CI not given)
Jelliffe-Pawlowski
etal. (2006)
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 >10 |jg/dl:
Pb during pregnancy 30.9%
Odd Ratios:
<5 |jg/dl: 1.00 (Ref)
6-9 |jg/dl: 0.8 (0.1, 6.4)
10-19 |jg/dl: 1.1 (0.2, 5.2)
20-39 |jg/dl: 4.5 (1.8,10.9)
> 40|jg/dl: 4.7 (1.1,19.9)
<10 |jg/dl: 1.00 (Ref)
> 10|jg/dl: 3.2 (1.2, 7.4)
Jones et al. Tennessee Gestational
(2010) 2006 Age: preterm
(<37wk), term
(37-40 wk),
post-term (>40
wk)
Singleton births >
27 wk 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:
Preterm birth: 1.4
Term birth: 1.2
Post-term birth: 1.3
p-value for difference: >0.10
Vigeh et al. (2011) Tehran, Iran Preterm birth
2006 (20-37 wk)
Singleton births Maternal blood Pb
from non-smoking,
non-obese
mothers aged 16-
35 and referred for
prenatal care
during the 8th-12th
week of gestation
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
MQ/dL)
1 One study of preterm birth included women living in two different residential areas over three
2 different time periods (Berkowitz et al.. 2006). One residential area was consistently unexposed but the
3 other had a period of high Pb emissions due to damage at a local factory. The three time periods examined
4 preterm birth rates before, during, and after the time of high exposure. No association was observed
5 between women living in the high exposure area compared to those in the low exposure area during any
6 of the time periods, but the number of preterm infants born during the period of high exposure was small.
7 In another study, measurements of umbilical cord blood were taken after birth at a hospital in Nagpur,
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India (Patel & 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 varied between
infants with >5 (ig/dL Pb and infants with < 5 (ig/dL 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 (E. A. Jones et al.. 2010). Mean levels of
umbilical cord blood Pb were slightly higher among infants born preterm but the difference was not
statistically significant. In a study taking place in California, women with information on blood Pb levels
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 i ffe -Paw 1 owsk i 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 increased 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 refined
into further categories, when compared to maximum blood Pb levels < 5 (ig/dL the positive association
between maximum blood Pb measurement and preterm birth was present at 20 (ig/dL and higher. Finally,
a study in Iran reported higher maternal blood Pb for preterm births than term births fVigeh et al.. 2011).
A positive association between maternal blood Pb levels and preterm birth was observed.
In sum, as in the 2006 Pb AQCD, recent epidemiologic studies report inconsistent findings for the
relationship between Pb and preterm birth.
5.8.7. Low Birth Weight/Fetal Growth
The 2006 Pb AQCD reported inconsistent study results examining 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. 2006). Since then, multiple 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 5-36
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 5-36. Summary of recent epidemiologic studies of effects on low birth weight and
fetal growth.
Reference
Location
Study
population
Exposure
Measurement
Mean Pb
(SD) in
|jg/dL
Adjusted Effect
Estimates
Berkowitz et al.
(2006)
Idaho
1970-1981
Low birth weight
(<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
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"
Not specified
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)
Gundackeretal.
(2010)
Vienna,
Austria
2005
Birth length, birth
weight, head
circumference
Infants of women
recruited during
their second
trimester
Maternal blood Pb
between wk 34-38 of
gestation, whole
placentas and umbilical
cord Pb shortly after
birth, meconium
samples in first five days
after birth
Median (IQR):
Maternal blood
Pb: 2.5 (1.8,
3.5)
Umbilical cord
blood Pb: 1.3
(0.8, 2.4)
Placenta Pb:
25.8 ug/kg
(21.0, 36.8
Mg/kg)
Meconium Pb:
15.5 |jg/kg (9.8,
27.9 |jg/kg)
Regression coefficients
(units not given, assume
results are per 10 |jg/dL or
1 Mg/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)
Iranpouretal.
(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 for t-tests:
Maternal blood Pb 0.07
Umbilical cord blood Pb:
0.20
P-values for correlations:
Maternal blood Pb and
Birth weight:
Low birth weight: 0.17
Normal birth weight: 0.3
P-values for correlations:
Umbilical cord blood Pb
and birth weight:
Low birth weight: 0.84
normal birth weight: 0.26
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 |jg/dl: 1.00 (Ref)
> 10|jg/dl: 0.82 (0.57,
1.17)
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Reference
Location
Study Exposure
population Measurement
Mean Pb
(SD) in
|jg/dL
Adjusted Effect
Estimates
Jelliffe-Pawlowski
etal. (2006)
California
1995-2002
Low birth weight
(<2,500g)
Small for gestational
age (birth weight for
Singleton births
to non-smoking
mothers with
blood Pb
gestational age <10th measures during
percentile of race-
and gender- specific
norms
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 >10 |jg/dl: Odd Ratios:
blood Pb during 30.9% Low birth weight <5 |jg/dl:
pregnancy 1.00 (Ref)
6-9 |jg/d|- -
10-19 |jg/dl: 2.7 (0.5, 14.8)
20-39 |jg/dl: 1.5 (0.3, 7.7)
> 40|jg/dl: -
<10 |jg/dl: 1.00 (Ref)
> 10|jg/dl: 3.6 (0.3, 40.0)
Small for gestational age
<5 |jg/dl: 1.00 (Ref)
6-9 |jg/dl" -
10-19 |jg/dl: 2.3 (0.6, 9.2)
20-39 |jg/dl: 2.1 (0.7, 6.7)
> 40|jg/dl: -
<10 |jg/dl: 1.00 (Ref)
> 10|jg/dl: 4.2 (1.3,13.9)
Jones et al. (2010) Tennessee
2006
Low birth weight Singleton births Umbilical cord blood Pb
(<2.500g) >27 wks
gestation from
mothers aged 16-
45 living in the
Shelby County
area for at least 5
mo during
pregnancy
2.4 (4.3) Geometric Mean:
Geometric Low birth weight: 1.2
mean: 1.3 Normal birth weight: 1.3
p-value for difference:
>0.10
Kordas et al. Mexico Head circumference,
(2009) City, birth weight, birth
Mexico length
1994-1995
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 |jg/g
(9-8 Mg/g)
Maternal blood
Pb > 10|jg/dl:
27%
Umbilical cord
blood Pb >
10|jg/dl: 13.7%
Regression coefficients
(SE) (adjusted for
maternal BMI, maternal
height, infant gestational
age, and other variables)
for each 1 ug/g increase in
tibia Pb:
Birth weight: -4.9 (1.8)
Birth length: -0.02 (0.01)
Head circumference: -0.01
(0.01; p-value<0.05)
Women with 4th quartile
tibia Pb (15.6-76.5 yglg)
delivered infants 140 g
less than women with tibia
Pb in the lowest quartile
Lamb et al. (2008) Mitrovica
Height and BMI at
Participants of
Mid-pregnancy blood Pb Mitrovica: 20.56
Regression Coefficients
and
birth
the Yugoslavia
(7.38)
(95% CI) for 1 Mg/dL
Pristina,
Study of
Pristina: 5.60
increase in Pb:
Yugoslavia
Environmental
(1.99)
BMI
1985-1986
Lead Exposure,
Pregnancy
Outcomes, and
Childhood
Mitrovica: -0.18 (-0.69,
0.33)
Pristina:-0.14 (-0.69, 0.42)
Height
Development
Mitrovica: 0.43 (-0.83,
1.69)
Pristina: 0.35 (-0.64,1.34)-
Llanos and Ronco Santiago,
(2009) 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
Mg/g (004 Mg/g)
Controls: 0.04
Mg/g (0.009
Mg/g)
P-value for Mann-Whitney
U-test <0.01
Williams etal.
(2007)
Tennessee Birth weight
2002
Infants from Air Pb levels during first
singleton births trimester of pregnancy
or the firstborn
infant in a set of
multiples
0.12 ^g/m3
(0.04 \iglm3)
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
Increase of Pb from 0 to
0.13 (maximum) relates to
a 124g decrease in birth
weight
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Reference
Location
Study
population
Exposure
Measurement
Mean Pb
(SD) in
|jg/dL
Adjusted Effect
Estimates
Zentneretal. Santo Birth weight and
(2006) Amaro, length
Brazil
2002
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 |jg/dL):
Length -0.46 (p-value
0.003) and Weight -0.275
(0.048)
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 blood Pb levels were greater in 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, no differences in the
associations were detected. Another study using maternal blood Pb was conducted in California (Jelliffe-
Pawlowski et al.. 2006). Women's blood Pb measurements during pregnancy were matched with the
corresponding birth certificates. The adjusted OR for low birth weight that compared women with blood
Pb levels >10 j^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 large 95% CI. An association was detected for high blood Pb and
having an infant who was small of his/her gestational age (SGA; defined as birth weight <10th percentile
of normal weight for population-based singleton race and gender specific infants of the same gestational
age). 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 (Iranpour et al.. 2007). Finally, 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).
One study examining the association between Pb levels and birth weight used tibia bone
measurements from the mothers living in Mexico City (Kordas et al.. 2009). Pb tibia 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 increased among women with certain genotypes
(statistical tests not reported).
Multiple studies examined the relationship using Pb measured from the placenta or umbilical cord.
First, the study by Iranpour et al. (2007) discussed above investigated the association between umbilical
cord blood Pb levels in addition to their examination of whole blood. They again report no difference in
levels between term infants of normal and low birth weight. Researchers in Chile collected the placentas
from term births and compared the Pb levels for those born with normal birth weights to those with low
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birth weights (Llanos & 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 Pb level was found to be inversely correlated with length and weight of the infants. A third
study recruited women in Pakistan (Janiua et al.. 2009). Umbilical cord blood Pb levels were not
associated with 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 among infants born to women
living primarily in urban areas of Memphis, TN (E. A. Jones et al.. 2010). 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.
Two 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 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. The
other 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 women living in high exposure areas during the period of high exposure
compared to those living in unexposed areas. 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
exposure, the odds of low birth weight among term births was increased among those living in the
exposed area compared to those in the unexposed 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 fifth percentile of birth weight for their
sex and age) was also associated with living in the exposed area during the time period of 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 unexposed 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. 2006). More recent studies
have reported conflicting results. Wang et al. (2009) 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
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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, female pups had a decreased birth weight when compared to male pups
of the same litter (only statistically significant at the 400 ppm dose). 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 (mean
blood Pb level 22.8 (ig/100mL), 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-and high-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 (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. Toxicological Studies of Developmental Effects
5.8.8.1. Developmental Effects on Blood and Liver
The 1986 and 2006 AQCD 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 report an increase in the number of
erythrocytes; however their size was diminished by 62%. Pb produced microcitic anemia as evidenced by
decreased hemoglobin content and hematocrit values without changes in mean corpuscular hemoglobin
(MCH) concentration. No changes were observed in alkaline phosphatase (ALP) activity, CAT activity, or
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thiobarbituric acid reactive substances (TBARS) production 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 PNDO or 21.
Masso-Gonzalez and Antonia-Garcia (2009) reported normochromic and microcytic anemia and a
significant decrease in hematocrit values and blood S-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 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 in 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
(ig/kg). In contrast, long-term exposure (5 months) to the same concentrations of Pb nitrate concentration
during gestation resulted in decreased hepatic SOD activity.
Effects on hepatic Phase I and Phase II enzymes after early developmental exposure of offspring to
Pb during gestation and lactation was evaluated by Pillai et al. (2009). In the study, pregnant Charles
Foster rats were administered 0.05 mg/kg body weight Pb subcutaneously throughout gestation until
PND21. Pups were evaluated on PND56. Results of the study show that Phase I xenobiotic-metabolizing
enzymes (NADPH- and NADH cytochrome c reductase) and Phase II xenobiotic- and steroid-
metabolizing enzymes (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). Pb-induced histological changes observed include massive fatty degeneration in
hepatocytes, large vacuoles in cytoplasm, appearance of pycnotic 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.
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5.8.8.2. Developmental Effects on Skin
The 2006 Pb AQCD (U.S. EPA. 2006) reported one study that demonstrated Pb-induced
abnormalities in skin development. No current studies were identified that addressed Pb-induced skin
alterations.
5.8.8.3. Developmental Effects on the Retina
The 2006 AQCD concluded that Pb exposure during early postnatal development (blood Pb -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 in pups on postnatal days 0-10 exposed to Pb during gestation 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 (ERG) in adult rats similar to ERG
findings in male and female children after gestational exposure to low- and moderate-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. High-level Pb exposure (109 ppm) statistically significantly decreased the number of
rods in central and peripheral retina Pb-exposure induced dose-dependent decreases in adult rat retinal
dopamine synthesis and utilization/release.
5.8.8.4. 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..
2005). Arora et al. (2005) demonstrated that Wistar rat pups exposed to Pb during gestation and lactation
(40 mg/L of Pb nitrate in drinking water of pregnant dams) had higher concentrations of Pb on the surface
of enamel and in the dentine immediately adjacent to the pulp. 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.
5.8.9. Summary and Causal Determination
Many epidemiologic and toxicological studies of the effects of Pb on reproductive outcomes have
been performed since the 2006 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
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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 dose-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 AQCD explored 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 AQCD on male fertility showed perturbed semen
quality. Recent studies have shown the effects of Pb exposure during early development to include
disruption of endocrine function; delay in the onset of puberty and alteration in reproductive function later
in life; and changes in morphology or histology in sex organs and placenta. Additionally, epidemiologic
studies of reproductive factors among males and females investigated whether Pb 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 (ig/dL and sperm count and quality. Multiple
studies of Pb and puberty have shown associations between blood Pb levels and delayed pubertal
development for girls and boys. These associations are consistently observed in multiple epidemiologic
studies and demonstrate effects on pubertal development at blood Pb levels <10ug/dL.
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 AQCD; inconsistent
evidence of a relationship with Pb was available for preterm birth and little evidence was available to
study the associations with spontaneous abortions. The previous AQCD included a few studies that
reported possible associations between Pb and neural tube defects, but the recent epidemiologic studies
found no association. Possible associations were observed between Pb and low birth weight when
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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. Effects of Pb exposure
during early development on toxicological studies included reduction in litter size, implantation, birth
weight and postnatal growth.
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 skin 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.
In conclusion, the recent toxicological and epidemiologic literature provides strong evidence that
Pb exposure is related to delayed onset of puberty in both males and females. Additionally, Pb exposure
has been shown to have detrimental effects on sperm (at high blood Pb levels in epidemiologic studies
and in low doses in the toxicological literature). The data on preterm birth, low birth weight, spontaneous
abortions, birth defects, hormonal influences, and fecundity are a bit more mixed and 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 negative
effects on sperm and delayed pubertal onset is sufficient to conclude that there is a causal relationship
between Pb exposures and reproductive effects and birth outcomes.
5.9. Effects on Other Organ Systems
5.9.1. Effects on the Hepatic System
Hepatotoxic effects of Pb in various animal and human models 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 various forms of Pb. Induction of
oxidative stress 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
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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 some of the frequent toxicological effects in the liver following
exposure to Pb included increased hepatic cell proliferation, cholesterol synthesis, DNA synthesis and
glucose-6-phosphotase dehydrogenase (G6DP) activity resulting in Pb-induced hyperplasia. The AQCD
concluded that cytochrome (CYP) P450 levels decreased following single exposures to Pb, primarily 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 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 a (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. More recent Pb exposure experiments
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.
In humans, the 2006 Pb AQCD stated that nonspecific liver injury generally observed as increases
in liver enzymes in the serum was reported in occupational studies. In addition, similar to effects noted in
animal studies, cytochrome P450 activity was also suppressed in humans following exposure to Pb under
various conditions. The 2006 Pb AQCD concluded that hepatic effects occurred only at high Pb exposure
levels.
5.9.1.2. New Epidemiologic Studies
A few studies examined liver biochemical parameters effects on antioxidant status and oxidative
stress resulting from occupational exposures. Patil et al. (2007) examined the effect of occupational Pb
exposure to liver and kidney function in silver jewelry workers (SJW), battery manufacturing workers
(BMW) and spray painters (SP) in western Maharashtra, India. Blood Pb was statistically significantly
increased in all three groups: 53.63 ± 16.98 (BMW), 48.56 ± 7.39 (SJW), and 22.32 ± 8.87 (ig/dL (SP),
compared to controls (12.52 ± 4.08). Liver function enzymes including serum glutamic oxaloacetic
transaminase (SGOT)/AST, and serum glutamic pyruvic transaminase (SGPT)/ALT levels were only
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increased in the SP group compared to the concurrent control group but not in the SJW and BMW groups.
Total serum protein levels were decreased in all three groups, while serum albumin levels were decreased
in SJW and SP groups and increased in the BMW group. Serum globulin levels were decreased and the
albumin/globulin levels were increased in the BMW and SJW groups compared to controls. In addition,
bilirubin levels were increased only in the BMW group. In a study examining the impact of Pb exposure
in an occupational setting, Khan et al. (2008) reported that workers in Pakistan occupationally exposed to
Pb (blood Pb = 29.1 [range 9.0 to 61.1] (ig/dL) had a significant increase (3.5-fold higher median) in
blood levels compared to age and gender matched controls (blood Pb = 8.3 [range 1.0 to 21.7] (ig/dL).
Oxidative stress markers such as MDA, and gamma-glutamyl transpeptidase (GGT) were significantly
increased in workers as were ALT levels. Serum albumin and total protein levels were significantly
decreased in the examined workers, compared to controls. Based on these results, study authors
concluded that Pb exposure causes oxidative stress and changes in liver enzymes that may lead to hepatic
toxicity in exposed workers. Can et al. (2008) also reported changes in liver function enzymes among
battery workers and muffler repair workers exposed to Pb in an occupational setting. Blood Pb was
elevated in both worker groups (36.83 ±8.13 and 26.99 ± 9.42 (ig/dL for battery workers and muffler
repair workers, respectively) versus, controls (14.81 ± 3.01 (ig/dL). The study authors reported that total
protein, globulin, and lactate dehydrogenase (LDH) levels were within or very close to the normal range,
but were statistically significantly higher in both worker groups compared to controls. Additionally,
increases in cholesterol and ALP were increased only in battery workers and muffler repair workers,
respectively. Though an increase in LDH levels among the workers was observed, the study authors stated
that this increase was not related to liver injury. Total protein, globulin, ALP, and LDH were also observed
to be significantly correlated to blood Pb levels in workers. Though liver enzyme function changes were
nominal, the study authors concluded that in an occupational setting, exposure to Pb may lead to liver
injury. While Can et al. (2008) did consider the impact of smoking in their analysis, it is not clear whether
Patil et al. (2007) and Khan et al. (2008) considered the impact on these factors in their analysis. In a
single case study report of a 40-year old Iranian male accustomed to using opium as a pain reliever,
Verheij et al. (2009) reported that a liver biopsy taken following elevated liver function enzymes exhibited
bile intracytoplasmic pigmentation in the hepatocytes. The study authors reported that the blood Pb levels
were highly elevated in the patient (86.0 (ig/dL), and attributed exposure to Pb from Pb-contaminated
opium consumption. The liver parenchyma also revealed disrupted architecture along with regenerated
nodules. Pathomorphological changes, rarely seen in humans, were also reported in the form of active
hepatitis along with micro vesicular and macro vesicular steatosis, hemosiderosis, and cholestatis as well as
lymphocytic cholangitis. The study authors stated that following chelation therapy, liver enzymes returned
to normal suggesting reversal of the histological findings. However, the reversibility was not confirmed
with another liver biopsy. A case report by Fonte et al. (2007) described a worker occupationally exposed
to Pb vapors (blood Pb = 148 (ig/dL) with hypersideremia, mixed bilirubinemia, and elevated levels of
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ALT and AST. Following chelation therapy, the patient's clinical symptoms resolved, indicating the
reversibility of the Pb-induced effects on the liver.
5.9.1.3. New Toxicological Studies
Hepatic Metabolism
As stated in the AQCD 2006, acute exposures to Pb nitrate and other Pb compounds causes a
decrease in Phase I enzymes accompanied by a simultaneous increase in Phase II enzymes. The
conclusions presented in the AQCD 2006 were also reviewed and corroborated by Mudipalli (2007).
Changes in biochemical parameters, suggestive of liver damage, in male Wistar rats 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 an increase in aspartate aminotransferase (AST), alanine aminotransferase
(ALT), serum alkaline phosphatase (ALP), and gamma glutamyl transpeptidase (GGT) levels (S. et al..
2009). In treated animals, the blood Pb levels steadily increased throughout the study period, reaching a
maximum of approximately 110 (ig/dL. The study authors reported that similar biochemical changes were
not observed in animals treated with Pb acetate as well a mineral rich diet and concluded that nutritional
management is important in managing Pb-related poisoning. 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 (blood Pb = 86 ± 6) compared to control cows (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. Upadhyay et
al. (2009) investigated the effects of Pb exposure on biochemical alterations in Sprague-Dawley rats
exposed to 35 mg/kg via i.p. injection for 3 days. The activities of ALT, AST, serum ALP, and acid
phosphatase were all significantly increased over control in exposed animals, whereas alkaline
phosphatase activity was decreased in exposed animals. Concomitant treatment with zinc and varying
levels of ascorbic acid were observed to ameliorate the toxic effects of Pb.
Pillai et al. (2009) investigated gestational and lactational exposure to Pb on hepatic phase I and II
enzymes in male and female rats. Pregnant rats were exposed to 50 j^ig/kg Pb acetate via subcutaneous
injection daily throughout gestation, and continuing until PND21. 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, compared to controls. In rats treated
with 25 (ig/kg Pb and Cd, the effect on phase I enzymes was increased. Pb exposure additionally
decreased the activities of phase II enzymes uridine diphosphate-glucoronyl transferase and GST in males
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and females, but no effect was observed on GGT or 17(3-hydroxysteroid oxidoreductase. Additionally, no
effect was observed in Pb-exposed rats regarding 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 following Pb
exposure. 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, but were similar between
the 2 dosing regimens at 3 months postweaning. 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 in a non-dose 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, exposing the mice to 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 increased LPS- induced TNF-a levels via the protein kinase C (PKC)/MAPK pathway.
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 blood Pb levels of 40.63 ± 9.21, 61.44 ±
4.63, and 39.00 ± 7.90 j^ig/dL, respectively. Animals exposed to 200 ppm had 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.
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, a decrease in
phospholipid levels was observed at 400 ppm. The authors also reported a positive correlation between
tissue cholesterol and phospholipids compared to Pb accumulation in liver across all doses. In contrast,
the association between tissue triglyceride levels and Pb accumulation was negative. In a related study,
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Khotimchenko and Kolenchenko (2007) reported that adult male albino rats treated with 100 mg/kg Pb
acetate for as little as 14 days exhibited disorders in lipid metabolism that were supported by increased
levels of total cholesterol and triglyceride levels in the liver tissue. 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 may cause subtle effects at the cellular level that may lead to
hepatotoxicity. Kojima and Degawa (2006) examined the gender-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. 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 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 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 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 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 wellnourished 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 exposed to 0.025 mg/kg Pb via i.p. injection 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 fraction (Pandya
et al.. 2010). Nonstatistically significant decreases were also observed in GSH levels and GPx and GR
activities in exposed animals. Yu et al. (2008) reported similar dose-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
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in 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 confirmed 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 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
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. Pups exhibited liver damage that was supported 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. In the fetuses from dams exposed for 1 month prior to
pregnancy, an dose-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. In a study examining the role of
low molecular weight thiols on peroxidative mechanism, 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. Metalothionine 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 exposed to 35 mg/kg Pb via i.p. injection, whereas hepatic
GSH was significantly decreased (Upadhvav et al.. 2009). 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 (Pillai et al.. 2009).
The studies presented above all confirm 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.
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5.9.2. Effects on the Gastrointestinal System
Gastrointestinal effects of Pb exposure primarily include abdominal pain, constipation, and internal
paralysis in humans. In animals, degeneration of the intestinal epithelial mucosa and a decrease in
duodenal motility has been reported.
5.9.2.1. Summary of Key Findings on the Effects on the Gastrointestinal System
from the 2006 Lead AQCD
The 2006 Pb AQCD states 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. 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, calcium, and vitamin D indicated competition in absorption between Pb and calcium.
Diet supplement with vitamin D Pb 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 (l,25-(OH)2D3) levels. Overall, the 2006 Pb AQCD states that studies in rat intestine have
shown that the largest amount of 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 AQCD also states that the 1986 Pb AQCD reported that common gastrointestinal effects in humans
following exposure to Pb include early symptoms of abdominal pain, constipation, and internal paralysis
with these symptoms generally observed at blood Pb range of 30-80 (ig/dL.
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. In a case study, Cabb et al. (2008) reported similar symptoms in
a 3-year-old child diagnosed with elevated blood Pb (19 (ig/dL). The child was reported to be
complaining of nonlocalized abdominal pain along with vomiting, nausea, constipation, lack of appetite,
joint pains, fatigue, irritability as well as headaches. Based on detailed nutritional information obtained
from the mother, the study authors inferred that the child was regularly ingesting candy contaminated
with Pb. Following this the child was treated by a folk healer with "greta" an orange powder that
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worsened the abdominal symptoms. When analyzed, it was determined that "greta" contained 14,000
mg/kg Pb monoxide. When the child was taken off "greta" his GI symptoms began to resolve. Based on
this study report, it may be concluded that Pb is capable of causing severe GI trauma and the GI effects
may be reversible once exposure to Pb is ceased. In a similar case study, Fonte et al. (2007) reported that
a 47-year-old male exposed to Pb fumes and vapors at a recycling plant experienced similar symptoms
reported by Cabb et al. (2008) in the 3-year-old child. The male patients' Pb levels were elevated with a
blood Pb level of 148 (ig/dL. Once exposure to Pb had ceased and the patient was treated with EDTA, his
condition improved and the blood Pb dropped to 16 (ig/dL. Kuruvilla et al. (2006) also reported
gastrointestinal effects including stomach pain and gastritis along with other Pb-related clinical
manifestations in battery workers and painters (blood Pb = 42.40 ± 25.53 and 8.04 ± 5.04 (ig/dL,
respectively) occupationally exposed to Pb in India.
5.9.2.3. New Toxicological Studies
Two studies pertaining to gastrointestinal effects of Pb exposure were identified. Santos et al.
(2006) examined the impact of Pb exposure on nonadrenergic noncholinergic (NANC) relaxations in rat
gastric fundus. Male Wistar rats treated with 0.008% Pb acetate via drinking water for 15, 30, and 120
days exhibited 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 at 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),
exhibited at dose-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. 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 the controls.
Additionally, there was a significant increase in gastric lipid peroxidation in 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 exposure to Pb may increase the formation of gastric ulcers
as a result of changes in the oxidative metabolism in the stomach.
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5.9.3. Effects on the Endocrine System
Endocrine processes that are most commonly impacted by Pb exposure include changes in the
thyroid, such as changes in the thyroid stimulating hormone (TSH), triiodothyronine (T3), and thyroxine
(T4). In addition changes in the male sex hormone levels have also been reported following exposure to
Pb.
5.9.3.1. Summary of Key Findings of the Effects on the Endocrine System from
the 2006 Lead AQCD
Endocrine processes that may be impacted by Pb 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 at blood Pb levels exceeding 30-40 (ig/dL. Summary of key findings
pertaining to reproductive hormones in females is presented in the section on reproductive effects and
birth outcomes (Section 5.8.1).
5.9.3.2. New Epidemiologic Studies
Thyroid hormone levels were reported to be impacted following exposure to various environmental
contaminates, including Pb, by Abdelouahab et al. (2008). Croes et al. (2009). and Dundar et al. (2006).
Abdelouahab et al. (2008) performed a cross-sectional study in a Canadian population exposed to various
environmental contaminants, including Pb, following consumption of freshwater fish. The median blood
Pb level of men included in this study was 3.1 (ig/dL, whereas for women the median was 1.7 (ig/dL. It is
important to note that the median blood Pb level for women is 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 that consumed fish contaminated
with Pb and other environmental pollutants. No impacts in T3 and T4 levels were 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. Dundar et al. (2006) examined the effects of blood Pb on thyroid function in 42 male
adolescent auto repair workers exposed long term to Pb. A control group comprised of 55 healthy subjects
was also used for comparison purpose. Mean blood Pb levels were reported to be higher in the auto repair
workers compared to the control subjects (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
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outcome, the study authors reported that long-term 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 and degree of sexual maturity in 1679 adolescents residing in
nine study areas with varying exposures to multiple industrial pollutants including Pb. The median 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 correction for confounding, significant interregional differences were observed in the
levels of sex hormones including total and free testosterone, estradiol, aromatase, and luteinizing hormone
as well as FT3 hormone levels. When individual neighborhoods were analyzed within the larger study
areas, altered levels of testosterone, aromatase, and FT3 levels were also observed. Altered sexual
maturation was also observed among boys and girls of individual study areas, compared to the sexual
maturation of the entire cohort. Though varying levels of sex hormones and rates of sexual maturation
was observed, the study authors reported that these changes are 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 contribute to the effects noted.
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 age 2.62 ±1.2 years, respectively. For prenatal blood Pb, the children were broken 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 was comprised of a gallon
of one part ice to one part water into which a child was asked to submerge his or her dominant arm for a
minute, 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 appeared to have 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 (ig/dL, can alter children's adrenocortical responses 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 levels in
142 young, urban African-American and Hispanic children in winter and summer to determine the
seasonal increase in blood Pb and its association with vitamin D, age and race. There was a
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winter/summer (W/S) increase in blood Pb levels in children aged between 1 and 3 years (winter blood Pb
= 4.94 ± 0.45 (SE) (ig/dL, summer blood Pb = 6.54 ± 0.82 (SE) (ig/dL), with a smaller W/S increase
observed in children aged between 4 and 8 years (winter blood Pb = 3.68 ± 0.31 (SE) |_ig/dL. summer
blood Pb = 4.16 ± 0.36 (SE) (ig/dL). 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. Summer and winter concentrations of 1,25-
(OH)2D3 were observed to differ in children aged 4-8 years and the correlation between the serum 1,25-
(OH)2D3. No difference in l,25-(OH)2D3 was observed in children 1-3 years old. 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 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
A single study examining the impact of Pb exposure on the endocrine system was identified. 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 (blood Pb = 86 ± 6 (ig/dL) and near closed Pb and operational zinc smelters (blood
Pb = 51 ± 9 (ig/dL) when compared to cows in unpolluted areas (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 exposure on adrenal and male gonadal
functions in Wistar rats exposed to 8.0 mg/kg Pb acetate via i.p. injection for 21 days. Exposure to Pb 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-exposed animals,
whereas adrenal weights were increased. Spermatogenesis was decreased and the percent of spermatid
degeneration was increased in animals exposed to Pb. Lastly, serum concentrations of testosterone, FSH,
and LH, were decreased in Pb-exposed animals. Supplementation with testosterone during the last 14
days of exposure to Pb was observed to ameliorate these effects.
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5.9.4. Effects on Bone and Teeth
Primary effects on bone as a result of Pb exposure include 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 is easily substituted by
Pb following Pb exposure. Exposure to high levels of Pb 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 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 following exposure to Pb.
Calcium in bone was easily substituted by Pb and taken up by the bone causing changes in bone cell
function. Exposure to Pb during gestation and immediate postnatal period was reported to significantly
depress early bone growth with the effects showing a dose-dependent trend, though this effect was not
manifested below certain exposure thresholds. Bone effects following short-term exposure to Pb were not
reported in mature animals; however, long-term exposures to Pb along with poor nutrition had an adverse
effect on bone growth as well bone density. Systemic effects of Pb exposure include 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 blood stream. Bone cell
cultures exposed to Pb were shown to impact vitamin D-stimulated production of osteocalcin
accompanied by inhibition of secretion of bone-related proteins such as osteonectin and collagen. In
addition, Pb exposure also 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).
Like the 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 during
growth, most of the Pb in the teeth remains in a state of permanent storage. High dose exposure of Pb to
animals 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
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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 AQCD also reported that a number of epidemiologic and animal studies have
both separately suggested that Pb is a caries-promoting element.
5.9.4.2. New Toxicological and Epidemiologic Studies
As reported in the 2006 Pb AQCD, Pb is capable of causing significant toxicological effects in
bones of humans and animals following short-term and long-term exposure. The association between Pb
exposure and osteoporosis was examined in three different epidemiologic studies. Campbell and Auigner
(2007) examined subjects >50 years of age using the Third National Health and Nutrition Examination
Survey (NHANES III) for association between Pb exposure and osteoporosis. The study authors used the
bone mineral density (BMD) in the hip as the primary variable to examine groups comprised of non-
Hispanic white men (mean blood Pb = 4.9, range: 0.7 to 48.1 jj.g/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 BMD in the non-Hispanic white males who had the lowest
blood Pb levels was statistically significantly higher than the males with higher blood Pb levels. Similar
associations, although not statistically significant, were reported among white females. 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 gender/race. Since the NHANES data
were comprised of 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. (2008)
examined the association between Pb exposure and osteoporosis in 192 (155 males; 37 females) Chinese
individuals occupationally-exposed to Pb (blood Pb = 20.22 and 15.5 (ig/dL, respectively). BMD 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 BMD in
groups that had high urinary Pb levels (> 5 jj.g/g creatinine) compared to groups with low urinary Pb in
both genders. In contrast a significant difference was only observed between blood Pb and BMD 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 Pb
exposure and osteoporosis in older women. The study authors conducted a prospective study using 533
women aged 65-78 years with a mean blood Pb of 5.3 ± 2.3 (ig/dL to determine the association between
blood Pb and recurring fractures. The BMD was 7% lower in the total hip (p <0.02) and 5% lower in the
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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 BMD. 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 nonspine 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
negative health consequences including osteoporosis-related fractures.
The effect of Pb exposure on bone development in younger children has been studied before and an
association has been found between Pb exposure and bone development. Ignasiak et al. (2006) conducted
a study on school children aged 7-15 years (463 males, 436 females) living close to copper smelters and
refineries in Poland to assess the impact of Pb exposure on their growth status. The mean blood Pb for all
participants was 7.7 ±3.5 (ig/dL (range: 2.0 to 33.9). Blood Pb levels were similar between males and
females except at age 14.8 years, when females had lower blood Pb than males. Study results indicated
that there was a statistically significant and linear relationship between blood Pb and reduced weight,
height, and trunk, leg and arm lengths. Regression analysis revealed that, per every 10 (ig/dL increase in
blood Pb levels, height decreased 2.1 and 2.9 cm for males and females, respectively. This decrease in
height was more influenced by decreases in leg length than trunk length. Based on these results, the study
authors concluded that linear skeletal growth was reduced with increases in blood Pb levels and these
effects were seen even at levels below 10 (ig/dL. These results also indicated that there was attenuation in
osteoblast activity as a result of Pb exposure.
To understand the significance 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 or 3 mM Pb, a dose-dependent impact on
the CRACI observed, as was a dose-dependent increase in the influx of Pb into human OLC. These
results suggest that Pb interferes with CARCI 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
was enhanced after CRACI had been induced. Since skeletal growth in Pb-exposed children is stunted,
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 assessed 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
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exhibited an increase in TGF-(3, but BMP-2 signaling was inhibited. Pb was also reported to induce NFkB
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 NFkB.
Effects of Pb exposure on teeth were examined in three epidemiologic studies. Since individuals
previously exposed to Pb may be impacted due to the endogenous 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 Normative Aging Study. Tooth loss in men was categorized as
0, 1-8 or > 9. Individuals with > 9 teeth missing had significantly higher bone Pb concentrations
compared to those with no tooth loss; no significant difference in blood Pb levels was observed between
the categories of teeth loss. Following adjustment for different variables (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. Odds of tooth loss were 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 United States (US), Saraiva et al. (2007) used data from NHANESIII by analyzing
data for 2,500 men and 2,399 women aged between 30 and 56 years. The analysis took into account
various covariates including age, NHANESIII phase, cotinine levels, poverty ration, race/ethnicity,
education, BMD, diabetes, calcium intake, dental visits, and menopause in women. After adjusting for
these covariates and comparing individuals with a 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 and statistical association between periodontitis and 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 (7.38 ± 4.41 versus. 2.27 ±
1.49 (ig/dL). No clinical periodontal or oxidative stress parameters were significantly difference 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
was observed. These results demonstrate that there is significant association between clinical periodontal
parameters and oxidative stress/damage indices in Pb-exposed apprentices. Following multiple regression
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analysis, a statistically significant association between gingival index and working status, family income
and either probing depth or clinical attachment level was noted.
5.9.5. Effects on Ocular Health
Ocular effects most commonly observed following 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 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 Toxicology and Epidemiology Studies
New animal studies pertaining to the ocular effects of Pb have investigated endpoints such as
retinal progenitor cell proliferation and neurogenesis, and have observed effects at exposures that resulted
in blood Pb levels as low as <10 (ig/dL (Section 5.3.4.3). One human study pertaining to ocular effects of
Pb was 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 (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 j^ig/dL). 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.
5.9.6. Effects on the Respiratory System
The collective body of toxicological and epidemiologic studies demonstrates Pb-induced 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
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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 respiratory tract illnesses (Section 5.6.4.1) and 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 5.6.4.3).
Collectively, panel and time-series studies demonstrate associations between Pb measured in PM2 5 or
PMio air samples and decreases in lung function and increases in respiratory symptoms, hospitalizations,
and mortality (Section 5.6.4.3). However, common limitations of these air-Pb studies are the variable size
distribution of Pb-bearing PM (Section 3.5.3) and its relationship with blood Pb levels as well as the lack
of adjustment for other correlated PM chemical components.
5.9.7. Summary
There is evidence from epidemiologic and toxicological studies that exposure to Pb results 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 humans (Can et al.. 2008; D. A. Khan et al.. 2008; Patil et al.. 2007) and animals (Y.-J.
Cheng et al.. 2006; S. et al.. 2009). Increased hepatic cholesterogenesis, altered triglyceride and
phospholipid levels, and disorders in lipid metabolism supported by increased levels of total cholesterol
and triglycerides has been reported in the animal literature (Ademuviwa et al.. 2009; Khotimchenko &
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 (Adegbesan & Adenuga. 2007; D. A. Khan et al.. 2008; Khotimchenko & Kolenchenko. 2007; D. Y.
Yu et al.. 2008). Indices of oxidative stress were additionally observed in the livers of fetuses exposed
throughout gestation (Uzbekov et al.. 2007).
Relatively few human studies have been conducted on the gastrointestinal toxicity of Pb since the
completion of the 2006 Pb AQCD. Two cases studies reporting on GI symptoms in a child and adult
report that elevated blood Pb was associated with nonlocalized abdominal pain, vomiting, nausea,
constipation, lack of appetite, fatigue, and headaches (Cabb et al.. 2008; Fonte et al.. 2007). The adult's
blood Pb level was reported as 148 (ig/dL. Both subjects' symptoms were reported to diminish following
cessation of exposure or chelation therapy. Similar GI symptoms (stomach pain and gastritis) were
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observed in battery works and painters 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 health 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; M. R. V. Santos
et al.. 2006). The observation of oxidative stress was accompanied gastric mucosal damage.
The endocrine processes most impacted by exposure to Pb include changes in thyroid function, as
well as alteration in sex and stress hormone profiles. TSH was negatively associated with blood Pb in
women that ate contaminated fish (median blood Pb = 1.7 j^ig/dL) but not men, and FT4, but not FT3, was
decreased in adolescent male auto repair workers (Abdelouahab et al.. 2008; Dundar et al.. 2006).
Alterations in the levels of sex hormones, including total and free testosterone, estradiol, aromatase, and
luteinizing hormone, were observed in Belgian adolescents environmentally exposed to Pb (C'roes et al..
2009). Toxicological evidence for similar effects was observed in adults cows reared in a contaminated
environment and exposed to Pb. A positive correlation was reported between blood Pb and plasma T3 and
T4 levels (Swarup et al.. 2007). Mean plasma estradiol levels were also significantly higher in Pb-exposed
cows. In children challenged with an acute stressor, increasing blood Pb was associated with significant
increases in salivary Cortisol responses, even at blood Pb levels less than 10 (ig/dL (Gump et al.. 2008).
Numerous epidemiologic studies investigated the association between Pb exposure and
osteoporosis in adults. High blood Pb has been observed to be associated with decreased BMD in non-
Hispanic white males (Campbell & Auinner. 2007). whereas urinary Pb, but not blood Pb, was seen to be
associated with decreased BMD in Chinese individuals occupationally exposed to Pb (Y. Sun. Sun. Zhou.
Zhu. Zhang, et al.. 2008). In elderly women, blood Pb levels were associated with an increased risk of
falls and fractures, including osteoporosis-related falls (khalil et al.. 2008). Linear skeletal growth in
children exposed to Pb was reduced with increased blood Pb; these effects were seen at blood levels <10
(ig/dL (Tgnasiak 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 resulted in
altered skeletal development (Jang et al.. 2008; Zuscik et al.. 2007). Epidemiologic studies investigating
Pb exposure and tooth loss report 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 with oxidative
stress/damage in individuals exposed in an occupational setting (Arora et al.. 2009; Saraiva et al.. 2007;
Yetkin-Ay et al.. 2007).
New toxicology studies have reported ocular effects (i.e., retinal progenitor cell proliferation) at
blood Pb levels as low as <10 (ig/dL (Section 5.3.4.3), and one human study reports and association
between heavy smoking, increased blood Pb levels, and cataracts (Mosad et al.. 2010). Investigation of
the respiratory effects of Pb exposure has been limited; however, cross-sectional studies have indicated an
association of increasing blood Pb with increased prevalence of respiratory tract illnesses (Section
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5.6.4.1) and asthma in children (Section 5.6.4.2). Additionally, Pb-induced production of ROS is
implicated in increased BR and decrements in lung function (Section 5.6.4.3).
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 confirmatory of those effects noted in the 2006 Pb AQCD. However, recent evidence of these
effects is relatively limited in number, 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) has 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. 2006; Rousseau et al.. 2005). Additionally, the National Toxicology
Program has listed Pb and Pb compounds as "reasonably anticipated to be human carcinogens" (NTP.
2004V
In the following sections, recent epidemiologic and toxicological studies published since the 2006
Pb AQCD regarding Pb and cancer mortality and incidence are examined. In addition, recent studies of Pb
and DNA and cellular damage, as well as epigenetics studies, are summarized. 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. 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 5-37 provides an
overview of the study characteristics and results for the epidemiologic studies that reported effect
estimates. This section also presents toxicological evidence on the carcinogenicity of Pb.
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Table 5-37. Summary of epidemiologic studies of cancer incidence and mortality
Reference
Study
Location
Cancer
Outcome
Study
Population
Exposure Mean Pb (SD1 Adjusted Effect Estimates
Measurement y ' '
(Menke et al..
2006)
U.S.A.
Overall
cancer
mortality
NHANES III
cohort with Blood
Pb measures in
1988-1994
- at least 12 years
of follow-up
-blood Pb <10
|jg/dL
Blood Pb
2.58 |jg/dL
(geometric mean)
Tertile 1: <1.93
uq/dL
Tertile 2:1.94-3.62
|jg/dL
Tertile 3: >=3.63
|jg/dL
Tertile 1:1.00
Tertile 2: 0.72 (95% CI: 0.46,1.12)
Tertile 3:1.10 (95% CI: 0.82,1.47)
Schoberetal. U.S.A. Overall NHANES III Blood Pb Blood Pb<5 ^ig/dL: Blood Pb<5 |jg/dL: 1.00
(2006) cancer cohort 67.7% Blood Pb 5-9 ^ig/dL: 1.44 (95% CI:
mortality - at least 40 years Blood Pb5-9 1.12,1.86)
of age ^ig/dL: 26.0% Blood Pb>10 |jg/dL: 1.69 (95% CI:
Blood Pb>=10 1.14,2.52)
|jg/dL: 6.3% Note:Modification by age assessed and
associations varied slightly
Weisskopf etal.
(2009)
Boston, MA
Overall
cancer
mortality
Normative Aging
Study
-included men
only
- mean follow-up
period for this
study: 8.9 yr
Blood Pb,
patella Pb
Blood Pb: 5.6
Mg/dl (3.4)
Tertile 1 of Blood
Pb: <4 |jg/dL
Tertile 2 of Blood
Pb: 4-6 |jg/dL
Tertile 3 of Blood
Pb: >6 |jg/dL
Tertile 1 of patella
Pb: <22|jg/g
Tertile 2 of patella
Pb: 22-35|jg/g
Tertile 3 of patella
Pb: >35 |jg/g
Blood Pb Tertile 1:1.00
Blood Pb Tertile 2:1.03 (95% CI: 0.42,
2.55)
Blood Pb Tertile 3: 0.53 (95% CI: 0.20,
1.39)
Patella Pb Tertile 1:1.00
Patella Pb Tertile 2: 0.82 (95% CI: 0.26,
2.59)
Patella Pb Tertile 3: 0.32 (95% CI: 0.08,
1.35)
Khalil etal.
Baltimore,
Overall
Subgroup of the
Blood Pb Level
Blood Pb<8 |jg/dL: 1.00
(2009)
MD, and
cancer
Study of
5.3 (2.3) |jg/dL
Blood Pb> 8 |jg/dL: 1.64 (95% CI: 0.73,
Monongahela
mortality
Osteoporotic
3.71)
Valley, PA
Fractures cohort
- included white
women aged 65-
87
12 yr (+/- 3 yr)
follow-up
Lundstrom etal.
(2006)
Sweden Lung cancer
(incidence
and
mortality)
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 |jmol/L,
controls 2.7
|jmol/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
|jmol/Pb, controls
11.9 ^mol/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^mol/L
Median cumulative Blood Pb index:
0.99 (95% CI: 0.96,1.02) per |jmol/L
Note: similar results were observed
when restricted to smokers only
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Reference
Study Cancer Study
Location Outcome Population
Measurement Mean P»
Adjusted Effect Estimates
Jones et al.
(2007)
Humberside, Lung cancer
UK mortality
Male tin smelter
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,
(2007) Canada
Lung cancer Men aged 35-79
and other
cancer
incidence
Interview of job
history and
exposure matrix
Ever exposed to:
Organic Pb 3.0%
Inorganic Pb
17.0%
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)
Lung 1.3 (95% CI: 0.5, 3.1)
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)
Lung 1.1 (95% CI: 0.7,1.7)
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)
Lung 0.8 (95% CI: 0.6,1.1)
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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_ . Study
Reference LoCation
Cancer Study
Outcome Population
Measurement Mean P»
Adjusted Effect Estimates
van Wijngaarden U.S.A.
and Dosemeci
(2006)
Brain cancer National
mortality
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,
(2006) Boston, MA,
and
Pittsburgh,
PA
Brain cancer NCI Brain Tumor
incidence 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
withALADI genotypes; these
associations were not observed for
glioma incidence
Bhattietal. Phoenix, AZ, Brain cancer NCI Brain Tumor
(2009) Boston, MA, incidence Study
and - included non-
Pittsburgh, Hispanic whites
PA 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
Overall Children living in Pb-related air
cancer this province at pollution
incidence least five years measures
Glioma: 70.5
|jg/m3y (193.8
|jg/m3y)
Glioblastoma
multiform: 97.5
^jg/m3y (233.9
|jg/m3y)
Meningioma:
101.1 LJg/m3y
(408.7 |jg/m y)
Controls: 69.7
Ljg/m3y (248.8
Mg/m'y)
Per 100 ^g/m'V increase in cumulative
Pb exposure
Glioma: 1.0 (95% CI: 0.9,1.1)
Glioblastoma multiform: 1.0 (95% CI
0.9,1.1)
Meningioma: 1.1 (95% CI: 1.0,1.2)
Note: modification by SNPs was
conducted and associations varied by
SNP
Absalon and Silesia
Slesak (2010) province,
Poland
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. Island of Krk,
(2007) 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)
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Reference
Study Cancer Study
Location Outcome Population
Measurement Mean Pb Adjusted Effect Estimates
Santibanez et al. Valencia and Esophageal PANESOES study Interviews to
(2008) Alicante, cancer included 30-80 yr determine
Spain incidence old men occupational
hospitalized in any history and a job
of the participating exposure matrix
study hospitals
NA
All esophageal cancers:
Unexposed: 1.00
Low workplace Pb exposure (<0.237
|jmol/L): 0.79 (95% CI: 0.43, 1.46)
High workplace Pb exposure (>0.237
|jmol/L): 1.69 (95% CI: 0.57, 5.03)
Esophageal squamous cell carcinoma:
Unexposed: 1.00
Low workplace Pb exposure (<0.237
|jmol/L): 0.70 (95% CI: 0.34, 1.43)
High workplace Pb exposure (>0.237
|jmol/L): 0.91 (95% CI: 0.22, 3.75)
Adenocarcinoma:
Unexposed: 1.00
Low workplace Pb exposure (<0.237
|jmol/L): 0.95 (95% CI: 0.32, 2.82)
High workplace Pb exposure (>0.237
|jmol/L): 5.30 (95% CI: 1.39, 20.22)
'associations not changed or slightly
increased when restricted to
occupational exposures >15yr
5.10.1.1. Overall Cancer Mortality
Recent studies have been performed examining the association between biologically measured Pb
levels and cancer mortality. The Third National Health and Nutrition Examination Survey (NHANES III)
included a nationally representative sample of U.S. adults who had blood Pb measurements taken and
were followed up for 12 years. Mean blood Pb levels were 2.58 (ig/dL (individuals with blood Pb levels
greater than 10 (ig/dL were excluded from the study). No association was observed between blood Pb and
cancer mortality (HR of highest tertile compared to lowest tertile: 1.10 [95% CI: 0.82, 1.47]) (Menke et
al.. 2006). Another analysis of the NHANES III population, restricted to individuals 40 years and older at
the time of blood Pb collection, included individuals with all blood Pb levels (including those greater than
10 (ig/dL) (Schober et al.. 2006). Overall, 68% of the study population had blood Pb levels less than 5
(ig/dL and 6% had blood Pb levels greater than 10 (ig/dL. Among individuals who died of cancer during
the study period, 52% had blood Pb levels less than 5 (ig/dL and 12% had blood Pb levels greater than 10
(ig/dL. In this study, median follow-up time was 8.6 years and a positive association was observed
between blood Pb and cancer mortality. The RRs were 1.69 (95% CI: 1.14, 2.52) for individuals with
blood Pb levels of at least 10 (ig/dL and 1.44 (95% CI: 1.12, 1.86) for blood Pb levels of 5-9 (ig/dL
compared to individuals with blood Pb levels less than 5 (ig/dL. When categorized by age groups, 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. A positive association for blood Pb levels of 10
(ig/dL and greater was present among those 40-74 years old and 85 years and older. A study of men from
the greater Boston area enrolled in the Department of Veterans Affairs Normative Aging Study reported
mean blood Pb levels of 5.6 (ig/dL (SD 3.4 (ig/dL) but this measure was poorly correlated with measured
bone Pb (Weisskopf et al.. 2009). No association was detected between either measure of Pb and cancer
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mortality in adjusted analyses. As part of the Study of Osteoporotic Fractures study, White women aged
65-87 were included in a sub-study of blood Pb levels and cancer mortality and were followed for
approximately 12 years (Khalil. Wilson, et al.. 2009). Mean blood Pb levels were 5.3 (.ig/dL (SD 2.3
(ig/dL) and no association was detected between blood Pb and cancer mortality in the study population.
5.10.1.2. Lung Cancer
A study was conducted among smelter workers to examine the relationship between Pb and lung
cancer (incidence and mortality combined) (Lundstrom et al.. 2006). Pb exposure was measured with
three different variables (peak blood Pb values, number of years Pb samples were obtained, and
cumulative blood Pb index) but none showed an association with lung cancer. Median follow-up in the
study was about 30 years and the peak blood Pb values during employment were 49.7 (ig/dL for lung
cancer cases and 55.9 (ig/dL for controls. A study in the UK of tin smelters reported no association
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 (S. R. Jones et al.. 2007).
Pb exposure was calculated in this study 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 and the smelters 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.
Other studies of Pb and lung cancer were performed by comparing the lung tissue of individuals
with lung cancer to those without lung cancer. The controls for these studies were individuals with
metastases in the lung from other primary cancers (De Palma et al.. 2008) and individuals with non-
cancerous lung diseases (De Palma et al.. 2008; kuo et al.. 2006). One study reported increased Pb
concentrations were observed 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) (De Palma et al.. 2008). but no statistical difference in Pb levels was reported
for lung tissue of individuals with lung cancer compared to controls in the other study (Kuo et al.. 2006).
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. 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 &
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Doscmcci. 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 with ALAD2 genotypes (Raiaraman et al.. 2006). The association was not
present among individuals with ALA/) I genotypes. A third study of the association between Pb exposure
and brain tumors reported none or slight overall associations with types of brain tumors; however,
positive associations were observed among individuals with certain single nucleotide polymorphisms
(SNPs) (Bhatti et al.. 2009). After control for multiple comparisons, individuals with GPX1 variants had
positive associations between cumulative Pb exposure and glioblastoma multiforme and meningioma,
whereas individuals without RAC2 variants showed a positive association between Pb and glioblastoma
multiforme and individuals without XDH variants displayed a positive association between Pb and
meningioma.
5.10.1.4. Breast Cancer
A few studies examined Pb levels and breast tumors among individuals with and without breast
tumor and/or cancer present. One study of newly diagnosed breast cancer patients and controls examined
Pb levels in blood and hair samples and reported higher levels in both for cancer cases, although the
difference in the hair samples was not statistically significant (Alatise & 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. Malik,
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
increased 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 Pb levels. Additionally, the sample sizes are
often small and the studies may be underpowered.
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5.10.1.5. Other Cancers
Studies of multiple 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 & Slesak. 2010). The highest Pb levels
were measured in 1990 when over 50% of the study area exceeded the limit of 1 |_ig/m2year. 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). No correlations were reported.
A study performed among men evaluated multiple cancer outcomes and determined exposures to
organic Pb, inorganic Pb, and Pb in gasoline emissions via interviews regarding job histories and then
subsequent exposure approximations by chemists and hygenists (Rousseau et al.. 2007). Organic Pb
exposure was positively associated with stomach cancer. 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
esophageous, 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 (association not present when
population controls were employed 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
adenocarcinmoa of the esophagus (Santibanez et al. 2008).
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, increased Pb levels were observed in the
blood and bladder tissue of individuals with bladder cancer (Golabek et al.. 2009). the kidney tissue of
individuals with renal cell carcinoma (with highest levels among those with the highest stage tumors)
(Calvo et al.. 2009). the tissue (but not serum) of individuals with laryngeal cancer (Olszewski et al..
2006). the blood of individuals with gastric cancer (Khorasani et al.. 2008). the plasma and hair of
individuals with gastrointestinal cancer (Pasha et al.. 2010). the blood and hair of individuals with non-
specified types of cancer (Pasha et al.. 2007; Pasha. Malik. & Shah. 2008). and the hair of individuals
with benign tumors (Pasha. Malik, 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 (C. N. 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
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detected in corpses without neoplastic tumors of the kidney (Cerulli et al.. 2006). All of these comparison
studies are limited by the inability to examine temporality; the presence 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.
5.10.1.6. Toxicological Models of Carcinogenicity
Carcinogenicity in Animal Models
Inorganic Pb has been shown to act as a carcinogen in animal toxicology models. Most commonly
the kidneys (Azar et al.. 1973; Kasprzak et al.. 1985; Roller et al.. 1985; Van Esch & Kroes. 1969) are
targeted but the testes, brain, adrenals, prostate, pituitary, and mammary gland have also been affected
(IARC. 2006). More recently it has been shown that early life transplacental and lactational exposure of
laboratory rodents to inorganic Pb induces carcinogenicity in adulthood (Waalkes et al.. 1995). Chronic,
lifetime exposure to Pb is also associated with carcinogenicity in laboratory rodents (Roller et al.. 1985).
One recent study considered Pb-dependent carcinogenesis in laboratory animals. Tokar et al. (2010)
considered tumorigenesis in rodents. Homozygous metallothionein I/II knockout mice and their
corresponding wild type controls (groups of ten mice each) were exposed by drinking water to 2,000 or
4,000 ppm Pb(Ac)2 and compared to untreated controls. Study animals were exposed in utero, through
birth and lactation, and then directly to drinking water until 8 weeks old. The 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
metallothionein can protect against Pb-induced tumorigenesis. Concerns with the study are that the doses
are at levels of Pb humans would not likely be exposed to 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 has any impact on Pb carcinogenesis.
Neoplastic Transformation Studies, Human Cell Cultures
Three studies considered Pb-dependent carcinogenesis in human cells. Xie et al. (2007) treated
BEP2D cells (human papilloma virus- immortalized human bronchial cells) with 0, 1, 5, or 10 |_ig/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-
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immortalized human skin fibroblasts deficient in Mrell) 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 Mrell deficient cells. Xie et
al., (2008) treated BJhTERT cells (hTERT-immortalized human skin fibroblasts) and ATLD-2 cells
(hTERT-immortalized human skin fibroblasts deficient in Mrell) with 0, 0.1, 0.5, and 1 (ig/cm2 PbCr04
for 24 or 120 h. Mrel 1 was required to prevent PbCr04-induced neoplastic transformation.
Immune Modulation of Tumorogenesis by Pb
As described in the prior AQCD for Pb ("U.S. EPA. 2006). Pb-induced immunotoxicity affecting
tumors results from a combination of suppressed Thl responses and misregulated inflammation. The
intersection of these two general Pb-induced alterations that elevate the risk of cancer. First, Pb-induced
misregulation of inflammation involving innate immune cells results in chronic insult to tissues. Decades
of excessive lipid and DNA oxidation production by overproduction of ROS and weakened anti-oxidant
defenses increase the likelihood of mutagenesis, cellular instability, and tumor cell formation. In support
of this, Xu et al. (J. Xu et al. 2008) found evidence that supports the association with Pb exposure and
DNA damage and concluded that it is a 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 never occupationally exposed to metals examined prostate
specific antigen (PSA), a biomarker for prostate cancer. This study reported a positive association
between blood Pb and PSA levels in adjusted analyses (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 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). A
study of male policeman reported mean blood Pb levels for the study population of 43.5 (ig/dL
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(Wiwanitkit et al.. 2008). When categorized 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) (Duvdii et al.. 2005V The exposed workers had
higher SCE levels and also a greater number of cells with the SCEs per cell higher than the 95th
percentile of the population. Finally, a study of children aged 5-14 years old (mean blood Pb levels of
7.69 (ig/dL [SD 4.29 |_ig/dL|) reported no correlation with SCE (Mielzviiska et al.. 2006). However, the
study did report a positive association between blood Pb and micronuclei levels.
Other studies of DNA damage have reported mixed results. A study of children ages 6-11 years old
and environmentally-exposed to Pb reported no association between blood Pb and DNA basal 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 or oxidative DNA damage (De Coster et al. 2008). A
study conducted among workers exposed to Pb (mean blood Pb levels of 30.3 j^ig/dL) and unexposed
controls (mean blood Pb levels of 3.2 (ig/dL) reported positive associations between the exposed group
and cytogenetic damage (measured by micronuclei frequency), chromosomal aberrations, and DNA
damage (although this was not statistically significant in linear regression models controlling for age)
(Grover et al.. 2010). A study of painters in India, where Pb-concentrations in paint are high, reported
mean blood Pb levels of 21.56 (ig/dL (SD 6.43 (ig/dL) among painters who reported painting houses for
8-9 hours/day for 5-10 years (M. I. Khan et al. 2010); the mean blood Pb levels were 2.84 (ig/dL (SD
0.96 (ig/dL) for healthy workers who had not been occupationally exposed to Pb. Cytogenetic damage
was increased among the painters compared to the healthy controls. Another study compared the blood Pb
of metal workers and office workers and reported greater blood Pb levels (both current and 2 year
average) among the metal workers (blood Pb levels at least 20 (ig/dL for metal workers compared to
blood Pb levels less than 10 (ig/dL for the office workers) (Olewiiiska 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 & Jamil. 2009). Higher chromosomal aberrations, micronuclei 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
One study, Tapisso et al. (2009). considered sister chromatid exchanges (SCE) in rodents. Algerian
mice (groups of six mice each) were exposed by intraperitoneal injection to 5 or 10 doses of 0.46 mg/kg
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Pb(Ac)2 and compared to untreated controls. 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.
Two studies considered SCE in cultured human cells. In one study, Ustundag and Duydu (2007).
considered the ability of N-acetylcysteine and melatonin to reduce led nitrate-induced SCE in a single
human donor. Cells were treated with 0, 1, 5, 10, or 50 (.iM Pb(N03)2. 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. Exposure times
were not provided. 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. In the other study, Turkez et al. 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. Exposure times
were not provided. The full interpretation of these data is limited by the limited number of donors and the
absence of an exposure time for the SCE assay.
Micronuclei Formation
Two studies considered MN in rodents. Alghazal et al. (2008). considered the ability of Pb(Ac)2
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 than unexposed controls. The second study, Tapisso et al., (2009). considered Pb
alone, Pb plus zinc and Pb plus cadmium-induced MN in rodents. Algerian mice were exposed by
intraperitoneal injection to 5 or 10 doses of 0.46 mg/kg Pb(Ac)2 and compared to untreated controls. The
MN in bone marrow were elevated after Pb exposure, which increased with time. Co-exposure with
cadmium or zinc did not further increase MN levels.
Three studies considered MN in cultured human cells. In one study, Ustundag and Duydu (2007)
considered the ability of N-acetylcysteine and melatonin to reduce led nitrate-induced MN in a single
human donor. Cells were treated with 0, 1, 5, 10, or 50 (.iM Pb(N03)2. MN 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. Exposure times
were not provided. 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. The second study, Turkez et al., considered the ability of
boron compounds to prevent Pb chloride-induced MN in human lymphocytes. Cells were obtained from 4
non-smoking donors. Both 3 and 5 ppm Pb chloride induced a statistically significant increase in MN
levels over controls. Boron induced a statistically significantly attenuation of these levels. Exposure times
were not provided. The full interpretation of these data is limited by the limited number of donors and the
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absence of an exposure time for the MN assay. The third study, Gastaldo et al. (2007). evaluated the
ability of Pb to induce MN. Human endothelial HMEC cell line was treated with 1-1000 (.iM Pb(N03)2
for 24 h. MN increased in a statistically significant, concentration-dependent manner.
HPRT Mutations
Two studies evaluated HPRT mutations in human cell cultures. Li et al., (2008). evaluated Pb(Ac)2-
induced HPRT in the non-small-cell lung carcinoma tumor cell line, CL3 and normal human diploid
fibroblasts (specific tissue source not provided). All cells were exposed to 0, 100, 300 or 500 (.iM Pb(Ac)2
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. A second study, Wang et al. (2008).
investigated Pb(Ac)2-induced HPRT mutations in CL3 cells. All cells were exposed to 0, 100, 300 or 500
(.iM Pb(Ac)2 for 24 hours in serum-free medium ± a 1-hour pretreatment with a PKC-alpha inhibitor or
siRNA fpr PKC-alpha. Pb alone did not induce HPRT mutations. Inhibiting the PKC-alpha via either
inhibitor statistically significantly increased Pb-induced mutagenesis.
One study considered HPRT in animal cell culture. McNeill et al. (2007) considered Pb(Ac)2
induced HPRT mutations in Chinese hamster ovary AA8 cells and AA8 cells overexpressing human
Apel. Cells were treated with 5 |_iM Pb(Ac)2 for 6 hours. No increases in HPRT mutations were observed
after Pb exposure in either cell line.
Chromosomal Aberrations
Only one study (El-Ashmawv et al.. 2006) considered Pb in laboratory rodents. The study focused
on dietary exposure to Pb(Ac)2 administered as a single dose of 0.5% w/w. Male Swiss albino mice, 30
per group, were studied. In addition, the authors considered the protective effects of turmeric and myrrh
powder. The study reported statistically significant levels of chromosomal aberrations in the Pb treatment
alone group, particularly with respect to fragments, deletions, ring chromosomes, gaps, and end-to-end
associations. The turmeric and myrrh powders were protective. Concerns with the study include the use of
only a single dose of Pb(Ac)2 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
common in this study raising questions about the quality of the metaphase preparations. One 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 or
why they only used one dose or analyzed fewer metaphases per dose.
Seven studies considered the ability of Pb to induce chromosomal aberrations in cultured human
cells. One study (Pasha Shaik et al.. 2006) considered the ability of Pb(N03)2 to induce chromosomal
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aberrations in primary human peripheral blood lymphocytes obtained from healthy, non-smoking donors.
Cells were treated with 0, 1.2 or 2 mM Pb(N03)2 for 2 h. 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. (2006). treated
WHTBF-6 cells (hTERT-immortalized human lung cells) with 0, 0.1, 0.5, or 1 (ig/cm2 PbCr04 for 24-120
hours or with 0, 0.1, 0.5, 1, 5 or 10 (ig/cm2 PbO for 24 or 120 hours. PbCr04 induced statistically
significant, concentration-dependent increases in centrosome abnormalities and aneuploidy. Wise et al.
(2006) treated BEP2D cells with 0, 0.5, 1, 5, or 10 |_ig/cm2 PbCr04 for 24 hours. PbCr04 induced
statistically significant concentration-dependent increases in chromosomal aberrations. Holmes et al.
(2006). treated WHTBF-6 cells with 0, 0.1, 0.5, or 1 (ig/cm2 PbCr04 for 24-72 hours. PbCr04 induced
statistically significant, concentration-dependent increases in chromosomal aberrations. The effects were
attributed to the chromate anion. Wise et al. (2006). treated WHTBF-6 cells with 0, 0.1, 0.5, or 1 (ig/cm2
PbCr04 for 24-120 hours. PbCr04 induced statistically significant, concentration-dependent increases in
spindle assembly checkpoint disruption, effects of mitosis and aneuploidy. By contrast, chromate-free
PbO did not induce centrosome amplification. The effects were attributed to the chromate anion. Xie et al.
(2007) treated BEP2D cells with 0, 1, 5, or 10 |_ig/cm2 PbCr04 for 24 hours. PbCr04 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 PbCr04 for 24 hours in a
study comparing 4 chromate compounds. PbCr04 induced statistically significant, concentration-
dependent increases in chromosomal aberrations, but was the least potent chromate based on administered
concentration.
Five studies considered the ability of PbCr04 to induce chromosome aberrations in rodent cell
cultures. All focused on PbCr04. Duzevik et al. (2006) treated Chinese hamster ovary (CHO) cells with 0,
0.1, 0.5, or 1 (ig/cm2 PbCr04 for 24 h. Specific CHO lines used included AA8 (wildtype) EM9 (XRCC1-
deficient), and H9T3 (EM9 complemented with human XRCC1 gene). PbCr04 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 PbCr04 (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 PbCr04 for 24 h. Specific CHO lines used included AA8 (wildtype) KO40 (Fancg-
deficient), and 40BP6 (Fancg complemented). PbCr04 induced statistically significant, concentration-
dependent increases in chromosomal aberrations that were increased by Fancg deficiency. Camrye et al.,
(2007) treated CHO cells with 0, 0.1, 0.5, 1, 5, or 10 (ig/cm2 PbCr04 for 24 hours. Specific CHO lines
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used included CHO-K1 (parental), xrs-6 (Ku80 deficient), and 2E (xrs-6 complemented with Ku80 gene).
PbCr04induced statistically significant, concentration-dependent increases in chromosomal aberrations
that were not affected by Ku80 deficiency. Stackpole et al. (2007) treated CHO and Chinese hamster lung
(CHL) cells with 0, 0.1, 0.5, or 1 (ig/cm2 PbCr04 for 24 hours. Specific CHO lines used included AA8
(wildtype), irslSF (XRCC3-deficient), and lSFwt8 (XRCC3 complemented). CHL lines used included
V79 (wildtype), irs3 (Rad51C deficient) and irs3#6 (Rad51C complemented). PbCr04 induced
statistically significant, concentration-dependent increases in chromosomal aberrations that were
statistically significantly increased by both XRC3 and Rad51C deficiency.
Three studies considered the ability of PbCr04 to induce chromosome aberrations in marine
mammal cell cultures. All focused on PbCr04. 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 PbCr04 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 PbCr04 for 24
hours. Wise et al. (2011) treated primary sperm whale skin fibroblasts with 0, 0.5, 1, 3, 5, and 10 (ig/cm2
PbCr04 for 24 hours. In all three studies, PbCr04 induced statistically significant, concentration-
dependent increases in chromosomal aberrations.
COMET Assay
Three studies considered the ability of Pb to induce DNA single strand breaks measured by the
comet assays in laboratory animals. Xu et al. (2008) considered the ability of Pb(Ac)2 to induce DNA
damage measured by the comet assay in lymphocytes in male ICR mice. Animals (5 per group) were
given Pb(Ac)2 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(Ac)2 to induce DNA damage measured by the
comet assay in primary spermatocyte DNA of male Wistar rats. Animals (3 per group) were treated for 13
weeks with 0, 250 or 500 mg/1 Pb in their drinking water. There was statistically significantly less DNA
damage in the controls compared to the two treatment groups. Narayana and Al-Bader (2011) considered
the ability of Pb(N03)2 to induce DNA damage measured by the comet assay 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.
Two studies considered the ability of Pb to induce DNA strand breaks measured by the comet assay
in cultured human cells. Only, one study (Pasha Shaik et al.. 2006) considered the ability of Pb to induce
DNA single strand breaks using the comet assay in primary human peripheral blood lymphocytes
obtained from healthy, non-smoking donors. Cells were treated with 0, 2.1, 2.4, 2.7, 3.0, 3.3 Pb(N03)2 for
2 hours. Increased comet tail length with increased dose was reported.
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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 Mrell) with 0,
0.1, 0.5, and 1 (ig/cm2 PbCr04 for 24 hours. PbCr04 induced a concentration-dependent increase in DNA
double strand breaks measured by the comet assay.
Two studies considered Pb-induced DNA single strand breaks in rodent cell cultures using the
comet assay. Xu et al. (2006). treated PC12 cells with 0, 0.1, 1 or 10 (.iM Pb(Ac)2. 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(Ac)2 for 48 hours. There was an
increase in several comet assay measurements including tail length.
Other Assays
Three studies considered the ability of Pb to induce DNA double strand breaks measured 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 PbCr04 for 24 hours. PbCr04 induced a
concentration-dependent increase in DNA double strand breaks measured by gamma-H2A.X foci
formation. Gastaldo et al. (2007) evaluated the ability of Pb to induce DNA double strand breaks measure
with gamma-H2A.X foci formation and by pulse-field gel electrophoresis in cultured human cells. Human
endothelial HMEC cell line was treated with 1 to 1,000 |_iM Pb(N03)2 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 lag/cm2 PbCr04 for 24 hours in a study comparing 4 chromate compounds. PbCr04 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 3 compounds.
Four studies considered Pb and DNA repair. All were done in cultured cells. Li et al., (2008).
evaluated Pb(Ac)2-induced effects on nucleotide excision repair efficiency in CL3 cells. All cells were
exposed to 0, 100, 300 or 500 (.iM Pb(Ac)2 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. Human endothelial HMEC cell line was treated with 100 (.iM Pb(N03)2 for 24
hours. Pb inhibited non-homologous end joining (NHEJ) repair, over activates 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 PbCr04 for 24 or 120 hours. Mrell was required to prevent PbCr04-induced
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DNA double strand breaks. McNeill et al, (2007) considered Pb(Ac)2 effects on Ape 1. Chinese hamster
ovary cells (AA8) were treated with 0, 0.5, 5, 50, or 500 (.iM Pb(Ac)2 and then whole cell extracts were
used to determine AP site incision activity. The data show that Pb reduced AP endonuclease function.
Finally, two studies considered Pb-induced cellular proliferation in laboratory animals. Fortoul et al.
(2005) exposed adult male CD1 mice (24 animals per group) to 0.01 M Pb(Ac)2, 0.006 M ca.dmium
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(Ac)2 for 48
hours. As measured by the MTT assay, Pb decrease cell proliferation at all concentrations tested.
5.10.3.3. Mechanisms of Action
The carcinogenic mechanism of action of Pb is poorly understood. Three well-accepted general
paradigms of carcinogenesis include multistage carcinogenesis (including initiation, promotion, and
progression), genomic instability, and epigenetic modification. Of the aforementioned paradigms, it is
unclear which of these best fit Pb. 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.
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. There are some data that suggest Pb
may interfere with some of these processes, but the data are few as these areas are rarely studied for Pb.
Furthermore, the bulk of the literature in this area involves PbCr04 and it is unclear if the effects are due
to Pb or chromate. Epigenetic modifications lead to cancer by altering cellular functions without altering
the genetic material. The most commonly studied epigenetic change is methylation alterations. Data show
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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, but it not likely to occur by a multistage carcinogenesis
paradigm. More work is needed to determine the mechanism.
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 protecting 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.
5.10.3.4. Effects of Lead within Mixtures
Three studies considered the impact of mixtures with Pb. All considered genotoxicity. Mendez-
Gomez et al., (2008). considered 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
arsenic or 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 by intraperitoneal to 5 or
10 doses of 0.46 mg/kg Pb(Ac)2 compared to untreated controls. The MN in bone marrow were elevated
after Pb alone exposure, which increased with time. Co-exposure with cadmium or zinc increased SCE
levels, but did not further increase MN levels. Glahn et al., (2008) performed a gene array study in
primary normal human bronchial epithelial cells from 4 donors treated with 550 jj.g/1 Pb chloride, 15 jj.g/1
cadmium sulphate, 25 jj.g/1 cobalt chloride or all three combined for 72 hours. There was a clear
interaction of all three metals impacting RNA expression.
One new publication details the interaction of Pb and selenium in virus-dependent mammary tumor
formation. 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 5.10.3.5, 5.10.3.7 and
5.10.3.10.).
One study considered Pb and selenium interactions in carcinogenesis in laboratory animals.
Schrauzer (2008) considered the impact of selenium on carcinogensis by studying 4 groups of weanling
virgin female C3H/St mice infected with murine mammary tumor virus (groups of 20-30 mice). 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(Ac)2 (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(Ac)2. The
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effects of selenium and Pb on the tumors caused by the virus were studied. The study is primarily focused
on the general effects of a low selenium diet. The data suggest that selenium is anticarcinogenic as in the
control groups the animals exposed to the higher selenium levels had fewer mammary tumors and these
tumors had a delayed onset. Pb exposure with low selenium caused the same delayed onset as 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.
5.10.4. Epigenetics
Epigenetic studies have been conducted to examine the associations between Pb levels and global
DNA methylation markers [Alu and long interspersed nuclear element-1 (LINE-1)] (Pilsner et al.. 2009;
R. Q. Wright et al. 2010). Wright et al. (2010) utilized a sample of participants from the Normative Aging
Study with mean Pb levels of 20.5 g/g (SD 14.8 g/g) for tibia measures, 27.4 g/g (SD 19.7 g/g) for patella
measures, and 4.1 (ig/dL (SD 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. When examining the
relationship between patella Pb and LINE-1 more closely, a non-linear trend was observed with leveling
off at higher Pb levels. No associations were observed for tibia or blood Pb and either LINE-1 or Alu. The
second study included maternal-infant pairs from the Early Life Exposures in Mexico to Environmental
Toxicants (ELEMENT) study and measured LINE-1 and Alu methylation in umbilical cord blood samples
(Pilsner et al.. 2009). In unadjusted models, maternal tibia Pb levels [mean 10.5 jj.g/g (SD 8.4 jj.g/g)] were
inversely associated with Alu methylation and maternal patella Pb levels [12.9 jj.g/g (SD 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 4 donors after in vitro treatment of the cells with 550 jj.g/1
Pb chloride, 15 jj.g/1 cadmium sulphate, 25 jj.g/1 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 epigentic
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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.
5.10.5. Summary and Causal Determination
In summary, the toxicology literature on the genotoxic, mutagenic, and carcinogenic potential of Pb
have strong evidence of effects in laboratory animals. In laboratory studies, high-dose Pb has been
demonstrated to be an animal carcinogen. There are data to suggest Pb is a human carcinogen among
toxicological studies, but they are not definitive. The three toxicological studies of neoplastic
transformation in cultured cells both were positive, but both focused on PbCr04 and attributed the
positive response to the Cr04 and not the Pb. 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. In separate studies, selenium and metallothionein are 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 zinc, Pb
replaces zinc at many zinc binding 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.
These indirect effects may act at a post-translational level to adversely alter protein structure and DNA
repair. Also, epigenetic changes associated with Pb exposure 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 in an adverse fashion. Thus, the animal toxicology literature provides a strong base for
understanding the potential contribution of Pb exposure to cancer in laboratory animals.
Multiple epidemiologic studies have been performed examining cancer incidence and mortality,
determined with biological measures and exposure databases. Mixed results have been reported for cancer
mortality studies; one strong epidemiologic study demonstrated a positive association between blood Pb
and cancer mortality, but the other studies reported null results. Although the previous Pb AQCD reported
some studies demonstrating an association between Pb exposure and lung cancer, current studies mostly
include studies of occupational exposure 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 genotypes. A limited amount of research on other types of cancer has been
performed. The previous AQCD reported evidence that suggested an association between Pb exposure
and stomach cancer, but recent studies of this association are lacking. One study examining Pb and
stomach cancer has been performed since the last AQCD and the results of the study are mixed.
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Among epidemiologic studies, positive associations were observed between high Pb levels and
SCE 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
inducing SCE, MN and DNA strand breaks, but continues to not produce chromosomal aberrations except
for PbCr04; 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.
Epigenetic effects, particularly with respect to methylation and effects on DNA repair were also
consistently seen. Epigenetic studies examining Pb and LINE-1 and Alu consistently demonstrated an
inverse association between patella Pb and LINE-1 methylation. Toxicological studies do show that Pb
can activate or interfere with a number of signaling and repair pathways, though it is unclear whether
these are in response to epigenetics or to genotoxicity. Thus, an underlying mechanism is still uncertain,
but likely involves either genomic instability or epigenetic modifications or both.
Overall, there is some epidemiologic evidence supporting associations between Pb and cancer.
Strong evidence from toxicological studies demonstrates an association between Pb and cancer,
genotoxicity/clastogenicity or epigenetic modification. The collective body of evidence integrated across
epidemiologic and toxicological studies is sufficient to conclude that there is a likely causal relationship
between Pb exposures and cancer.
The evidence reviewed in this chapter describes the recent findings regarding the health effects of
Pb. Table 5-38 provides an overview of the causal determinations for each of the health categories
evaluated.
Table 5-38. Summary of causal determinations for Pb.
5.11. Overall Summary
Health Category
Causal Determination
Neurological Effects
Causal relationship
Cardiovascular Effects
Causal relationship
Renal Effects
Causal relationship
Immune System Effects
Causal relationship
Effects on Heme Synthesis and Red Blood Cell Function
Causal relationship
Reproductive Effects and Birth Outcomes
Causal relationship
Cancer
Likely causal relationship
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Chapter 6 Contents
Chapter 6. Susceptible Populations and Lifestages 6-1
6.1. Susceptibility Factors and Lifestages Related to Lead Exposure and Dose 6-2
6.1.1. Lifestage 6-3
6.1.1.1. Young Children 6-3
Table 6-1. BloodPb levels broken down by age and gender, 2007-2008 NHANES 6-3
Table 6-2. Percentage of children within six brackets of blood Pb levels, from NHANES
2003-2004 6-4
6.1.1.2. Adults 6-5
6.1.2. Sex 6-6
6.1.3. Race and Ethnicity 6-6
Figure 6-1. 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). 6-7
Figure 6-2. SoilPb concentration exposure among the population of three parishes within
greater metropolitan New Orleans, by race and ethnicity. 6-8
6.1.4. Socioeconomic Status (SES) 6-8
6.1.5. Proximity to Lead Sources 6-9
6.2. Susceptibility Factors and Lifestages Related to Lead Induced Health Effects 6-10
Table 6-3. Summary of new evidence for characteristics examined as potential
susceptibility factors and lifestages for lead-related health effects 6-10
6.2.1. Lifestage 6-11
6.2.1.1. Children 6-11
6.2.1.2. Older Adults 6-12
6.2.2. Sex 6-14
6.2.3. Hormones 6-16
6.2.4. Genetics 6-17
6.2.4.1. Aminolevulinate Dehydratase 6-17
6.2.4.2. Vitamin D Receptor 6-18
6.2.4.3. Methylenetetrahydrofolate reductase 6-19
6.2.4.4. Apolipoprotein E 6-19
6.2.4.5. Hemochromatosis 6-19
6.2.4.6. Other Genetic Polymorphisms 6-20
6.2.5. Pre-existing Diseases/Conditions 6-20
6.2.5.1. Autism 6-20
6.2.5.2. Atopy 6-21
6.2.5.3. Diabetes 6-21
6.2.5.4. Hypertension 6-22
6.2.6. Smoking 6-22
6.2.7. Race/Ethnicity 6-23
6.2.8. Socioeconomic Status 6-24
6.2.9. Body Mass Index 6-24
6.2.10. Alcohol Consumption 6-25
6.2.11. Nutrition 6-25
6.2.11.1. Calcium 6-25
6.2.11.2. Iron 6-26
6.2.11.3. Zinc 6-26
6.2.11.4. Folate 6-26
6.2.11.5. Protein6-26
6.2.12. Stress 6-26
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6.2.13. Cognitive Reserve 6-27
6.2.14. Other Metal Exposure 6-27
6.2.14.1. Cadmium 6-28
6.2.14.2. Arsenic 6-28
6.2.14.3. Manganese 6-28
6.2.15. Fluoride 6-28
6.3. Summary 6-29
Chapter 6. References 6-32
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Chapter 6. Susceptible Populations
and Lifestages
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. The
NAAQS are intended to provide an adequate margin of safety for both the population as a whole and
those individuals potentially at increased risk for health effects in response to ambient air pollution
(Section 1.1). Several studies have evaluated factors to identify populations at greater risk for Pb-related
health effects. Many studies have termed such increased risk as the susceptibility and/or vulnerability of
an individual to Pb. The definition for both of these terms has been found to vary across studies, but in
most instances susceptibility refers to biological or intrinsic factors (e.g., lifestage, sex) while
vulnerability refers to non-biological or extrinsic factors (e.g., socioeconomic status [SES]) (U.S. EPA.
2009. 2010). Additionally, in some cases, the terms "at-risk" and sensitive populations have been used to
encompass these concepts more generally. However, in many cases, a factor identified that increases an
individual's risk for morbidity or mortality effects from exposure to an air pollutant cannot be easily
categorized as either a susceptibility or vulnerability factor.
As developed in previous ISAs and reviews (Sacks et al.; U.S. EPA. 2009. 2010). an all
encompassing definition for "susceptible population" is used to circumvent the need to distinguish
between susceptible and vulnerable, and to identify the populations at greater risk for Pb-induced health
effects. This definition identifies susceptible populations as the following:
Individual- and population-level characteristics that increase the risk of Pb-related health
effects in a population including, but not limited to: genetic background, birth outcomes (e.g., low
birth weight, birth defects), race, sex, lifestage, lifestyle (e.g., smoking status, nutrition), preexisting
disease, SES (e.g., educational attainment, reduced access to health care), and characteristics that
may modify exposure to Pb (e.g., time spent outdoors).
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 more susceptible
to Pb. Toxicological studies provide support and biological plausibility for factors that may lead to
increased susceptibility for Pb-related health effects through the study of animal models of disease.
Therefore, the results from these studies, combined with those results obtained through stratified analyses
Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and Environmental
Research Online) at http://eDa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of developing science
assessments such as the Integrated Science Assessments (ISA) and the Integrated Risk Information System (IRIS).
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in epidemiologic studies, comprise the overall weight of evidence for the increased susceptibility of
specific populations to Pb-related health effects.
The first section of this chapter summarizes susceptibility of population groups related to
differential Pb body burden. The studies presented in this section supplement the material provided in
Chapters 3 and 4 by examining how susceptibility factors, such as age, race, ethnicity, and SES may affect
Pb body burden, as measured by blood Pb or bone Pb. These biomarkers are influenced to varying
degrees by exposure, absorption, biokinetics, and diet. The second section of this chapter discusses the
epidemiologic and toxicological studies evaluated in Chapters 5 that provide information on potentially
susceptible factors related to Pb-induced health effects. Highlighted studies include only those where the
population was stratified into subgroups (e.g., males vs. females) 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.
Additionally, the section on susceptibility and Pb body burden explores how susceptibility factors may be
related to differential Pb exposures, where data are available. Numerous studies that focus on only one
potentially susceptible population are described in previous chapters, but these studies are not discussed
in detail in this chapter because they lack an adequate comparison group. For example, pregnancy is a
potentially susceptible lifestage for mothers and fetuses, but because there are no comparison groups for
stratified analyses, these studies are presented in Chapter 5 and are not included here. Included
toxicological studies may categorize the study population by age, sex, diet, genetics, etc. or are those with
animal models of disease.
Additionally, it is understood that some of the stratified variables may not be effect measure
modifiers but instead may be mediators of Pb-related health effects. Factors that are mediators are on the
causal pathway between Pb and health outcomes, while effect measure modifiers are factors that result in
changes in the measured association 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 susceptibility factors. Some of the factors included in this chapter could be mediators and/or
modifiers. These are noted in Table 6-3.
6.1. Susceptibility Factors and Lifestages
Related to Lead Exposure and Dose
Elevated Pb biomarkers have been shown to be statistically related to several population
characteristics, including age, gender, race and ethnicity, SES, and urbanization ("U.S. EPA. 2006). In most
cases, exposure, absorption, and biokinetics of Pb are all influenced to varying degrees by susceptibility
factors. The relative importance of susceptibility factors on exposure, absorption, and biokinetics varies
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on an individual basis and is difficult to quantify. The following section distinguishes studies
demonstrating a relationship between each susceptibility factor and exposure status from those that are
associated with increased biomarker levels without a clear attribution of the relative effects of those
factors on exposure, absorption, and bioavailability.
6.1.1. Lifestage
6.1.1.1. Young Children
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 (U.S.
EPA. 2006). Children can 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 gender. Among
children, highest blood Pb levels occurred in the 1-5 year age group, and within this subgroup, 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). Jones et al. (2009)
analyzed the NHANES dataset for the years 1988-2004 to study trends in blood Pb among different age
groups over time (see Table 6-2). They observed greater percentages of children aged 1-2 year having
blood Pb levels between 2.5 and 5 (ig/L compared with 3-5 year old children. Similarly, the 1-2 year old
group had larger percentages with blood Pb levels between 5 and 7.5 (ig/dL compared with 3-5 year old
children. These differences may be attributable to differences in exposure, age-dependent variability in
absorption and biokinetics, or diet (milk/formula versus child 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 broken down by age and gender, 2007-2008 NHANES
Age
Gender
Avg.
Std. Dev.
5%
25%
50%
75%
95%
99%
1 -5 yr
total
2.03
2.01
0.69
1.08
1.54
2.34
4.50
10.56
male
2.01
2.14
0.71
1.10
1.50
2.40
4.21
8.56
female
2.05
1.85
0.66
1.02
1.60
2.28
4.65
10.70
1 yr
total
2.62
3.26
0.76
1.24
1.80
2.88
6.23
16.94
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2.70
3.98
0.76
1.29
1.80
2.75
5.59
19.82
female
2.53
2.27
0.79
1.22
1.75
2.98
6.42
11.39
2 yr
total
2.09
1.52
0.76
1.28
1.77
2.40
4.34
8.40
male
1.99
1.11
0.72
1.30
1.76
2.40
3.82
6.04
female
2.20
1.84
0.82
1.27
1.77
2.35
4.52
10.75
3yr
total
1.99
1.80
0.65
1.04
1.50
2.32
4.92
8.55
male
1.99
1.55
0.68
1.11
1.50
2.35
5.09
8.46
female
1.99
2.08
0.65
0.93
1.40
2.16
4.64
9.81
4yr
total
1.75
1.31
0.69
1.00
1.40
1.90
4.03
6.31
male
1.80
1.47
0.71
1.05
1.30
2.07
4.09
6.89
female
1.66
1.03
0.64
0.98
1.41
1.89
3.78
5.31
5yr
total
1.62
1.22
0.64
0.87
1.29
1.90
3.37
7.29
male
1.58
0.97
0.72
0.95
1.29
1.88
3.38
4.96
female
1.68
1.50
0.62
0.80
1.30
2.04
3.33
8.52
6-11 yr
total
1.27
0.87
0.49
0.75
1.06
1.50
2.84
4.80
male
1.29
0.92
0.48
0.74
1.08
1.50
2.90
4.76
female
1.25
0.83
0.50
0.78
1.03
1.49
2.80
4.80
12-19 yr
total
0.99
0.73
0.40
0.59
0.81
1.13
2.11
4.00
male
1.13
0.82
0.44
0.69
0.94
1.30
2.41
4.11
female
0.83
0.57
0.37
0.52
0.69
0.91
1.70
3.36
20-64 yr
total
1.76
1.57
0.53
0.91
1.40
2.10
4.20
7.42
male
2.13
1.93
0.68
1.12
1.65
2.49
5.22
9.07
female
1.41
1.01
0.47
0.78
1.13
1.73
3.13
5.10
65+ yr
total
2.31
1.64
0.80
1.30
1.90
2.70
5.22
8.54
male
2.63
1.66
0.99
1.52
2.19
3.20
5.86
9.03
female
1.98
1.56
0.75
1.17
1.62
2.40
4.14
6.95
Source: CDC (2009a).
Table 6-2. Percentage of children within six brackets of blood Pb levels, from NHANES
2003-2004
n
Geometric
mean
<1 Mg/dL,
%
1 to <2.5
Hg/dL, %
2.5 to <5
Hg/dL, %
5 to <7.5
Hg/dL, %
7.5 to <10
Hg/dL, %
>10
Hg/dL, %
Overall
2532
1.9
14.0
55.0
23.6
4.5
1.5
1.4
Gender
Girl
1211
1.9
14.1
54.5
23.9
4.5
1.4
1.7
Boy
1321
1.9
14.0
55.5
23.2
4.6
1.5
1.3
Age
1-2 yr
1231
2.1
10.6
51.0
27.9
6.7
1.4
2.4
3-5 yr
1301
1.7
16.2
57.6
20.7
3.1
1.5
0.9
Race/Ethnicity
Non-Hispanic
Black
755
2.8
4.0
42.5
36.2
9.4
4.6
3.4
Mexican
American
812
1.9
10.9
61.0
22.1
3.4
1.3
1.2
Non-Hispanic
White
731
1.7
17.6
57.1
19.7
3.6
0.8
1.2
PIR
<1.3
1302
2.4
6.7
49.3
32.5
6.9
2.8
1.8
>1.3
1070
1.5
19.9
60.4
16.0
2.3
0.6
0.8
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6.1.1.2. Adults
Blood Pb levels tend to be higher in the elderly compared with the general adult population. 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 both blood Pb (p < 0.001) and saliva Pb (p < 0.001). Average
blood Pb levels among 14- to 24-year-old subjects was 2.6 (ig/dL compared with 4.3 (ig/dL among
subjects aged 55 years or older; similarly, average saliva Pb levels among 14- to 24-years-old subjects
was 2.5 |ig/L compared with 3.6 |ig/L 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 (see section 4.1) in conjunction with remobilization of stored Pb during bone
loss (Section 4.2).
Theppeang et al. (2008) 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, Yale energy index, and 2 diet variables;
(3 = 0.57, p <0.01 in a model including age, gender, and an interaction term for gender and age, which was
also statistically significant at p = 0.03). Theppeang et al. (2008) 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.
Fetal and child Pb biomarkers have been demonstrated to relate to maternal Pb biomarkers; several
older studies in the literature are presented in the 2006 AQCD (U.S. EPA. 2006). 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
and 35-39 had statistically significant higher odds of having greater blood Pb levels than women in the
25- to 29-years-old 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 decay during pregnancy, as described in Section
4.2. Elevated blood Pb levels among mothers also present a potential exposure to their children in utero or
through breast milk.
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6.1.2. 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 higher blood Pb levels (median: 1.50 (ig/dL) than females (median: 1.14 (ig/dL)
(see Table 6-1). Among adults aged 20-64 years, median blood Pb levels among males was 46% higher
than for females, and average levels were 51% higher for males compared with females. Among adults 65
years or older, median levels were 36% higher, and average levels were 33% higher. In their study of Pb
burden among Baltimore adults aged 50-70 years, Theppeang et al. (2008) 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. (2008) noted no difference (p = 0.12) between men
(18.0 (ig/dL) and women (19.4 (ig/dL).
Among U.S. children, the 2007-2008 NHANES data show that blood Pb levels are 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 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
gender-based differences in blood Pb levels are not substantial until adolescence.
6.1.3. 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. (2009) both analyzed NHANES survey data to examine trends in childhood blood Pb
levels. Data from the Jones et al. (2009) study, using 2003-2004 NHANES data (CDC. 2011). are shown
in Figure 6-1. They found that differences among racial and ethnic groups with regard to the percentage
with blood Pb levels greater than 2.5 (ig/dL have decreased since the period of 1976-1980 when
NHANES II was conducted. The non-Hispanic black group still had higher percentages with blood Pb
levels above 2.5 ng/dL compared with non-Hispanic whites and Mexican Americans, with the largest
observable differences 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. Theppeang et al.
(2008) also explored the effect of race and ethnicity on several Pb biomarkers in a study of older adults
living in Baltimore. 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 (ig/dL, C: 3.6 (ig/dL, p = 0.69).
Differences between bone Pb levels in African American and Caucasian subjects could potentially be
related to differential exposures in the home environment.
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70
60
1988-1991
50
40
30
20
<1 1 - < 2.5 2.5 - < 5 5 - < 7.5 7.5-<10 >10
70
60
1999 - 2004
50
30
<1
l-< 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 from the American Academy of Pediatrics, Jones et al. (2009)
Figure 6-1. 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).
1 Differences in exposure among ethnic and racial groups have also been noted. In a study of three
2 parishes in the greater metropolitan New Orleans area, Campanella and Mielke (2008) found that, where
3 soil Pb levels were less than 20 mg/kg, the population was 55% white, 36% black, 3.0% Asian, and 6.0%
4 Hispanic, based on the 2000 Census, with the percentage based on the total number living in Census
5 blocks with the same soil Pb levels. In contrast, they found that the population was 34% white, 62%
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black, 1% Asian, and 4% Hispanic on Census blocks in which soil Pb levels were between 1,000 and
5,000 mg/kg (Figure 6-2). As described in Section 6.14, the differences observed by Campanella and
Mielke (2008) may also be attributable to SES factors, or SES may be a confounding factor in the
relationship between Pb soil levels and race.
¦ White
Soil Pb Concentration (^g/g)
Source: Campanella and Mielke (2008).
Figure 6-2. Soil Pb concentration exposure among the population of three
parishes within greater metropolitan New Orleans, by race and
ethnicity.
6.1.4. Socioeconomic Status (SES)
Socioeconomic factors have sometimes been associated with Pb biomarkers, although these
relationships have not always been consistent. Nriagu et al. (2006) performed a multiple regression of
blood Pb and saliva Pb levels on various socioeconomic, demographic, and exposure variables among an
adult population in Detroit. 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
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<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) cite NHANES data analyzed in Pirkle et al. (1994) that the percentage of
children aged 1-5 years with blood Pb levels above 10 (ig/dL was 4.5% for the lowest income group
compared with 0.7% for the highest group. For the 1999-2002 NHANES, there was 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%), although 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) (Levin et al.. 2008). 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
exceeding 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
(R. L. Jones et al.. 2009). Likewise, Jones et al. (2009) analyzed blood Pb levels with respect to poverty-
income ratio (PIR). They found statistically significant differences in median blood Pb for PIR <1.3
compared with PIR >1.3. The percentage of 1- to 5-year0old children having blood Pb above 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 an average income of $40,000 per year, while soil Pb between 5,000 and 20,000 was associated with
an average income of $24,000 per year.
6.1.5. Proximity to Lead Sources
Airborne and soil Pb concentrations are higher in some urbanized areas, as described in Sections
3.2, 3.3, 3.5 and 4.1, as a result of historical and contemporaneous Pb sources. High concentrations of
ambient air Pb in PM tend to occur in the most urbanized areas and in close proximity to traffic in
metropolitan areas (Laid law & Filippelli. 2008; Mielke et al.. 2010; Weiss et al.. 2006). Moreover, air Pb
concentrations exhibit high spatial variability even at low concentrations (-0.01 ng/m3) (Martuzevicius et
al.. 2004). These conditions present the potential for additional risk of Pb exposure in urban areas.
Proximity to an industrial source likely contributes to higher Pb exposures, as described in the 2006
AQCD (U.S. EPA. 2006) 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 urban industrial site had statistically significantly higher cord
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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 current industrial Pb exposures 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 groups with and without industrial exposures
confounded these results.
6.2. Susceptibility Factors and Lifestages
Related to Lead Induced Health Effects
The following section evaluates potential susceptibility factors examined as effect measure
modifiers of various Pb-related health effects. Table 6-3 provides an overview of the factors examined and
populations identified as susceptible to Pb-related health effects based on the recent evidence integrated
across disciplines. Each characteristic is described in greater detail in the following sections.
Table 6-3. Summary of new evidence for characteristics examined as potential
susceptibility factors and lifestages for lead-related health effects
Factor Evaluated
Susceptible Population
Lifestage
Children
Sex
Males8 Females8
Genetics
ALADa, VDRa*, DRD4a*, GSTM1a, TNF-alphaa, eNOSa, APOEa, HFEa
Pre-existing Disease
Autismab, Atopyab, Hypertension"
Smoking
Smokers8
Race/Ethnicity
Non-Hispanic Blacks8, Hispanics8
Socioeconomic Status (SES)
Low SES8
Nutrition
Iron deficiency
Stress
High stress
Cognitive Reserve
Low cognitive reserve8"
Other Metals
Cd8, As8, Mn8
Additional evidence is needed to confirm whether the characteristic evaluated results in increased susceptibility to Pb-related health effects.
"Possible mediator
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6.2.1. Lifestage
The previous Pb AQCD reported on susceptibility by age ("U.S. EPA. 2006). The greatest ingestion
of Pb is often at the same time as development in children. Older adults are more likely to have age-
related degeneration of bones and organ systems and a possible redistribution of Pb. Therefore, increased
susceptibility of Pb-related health effects is a concern for these populations. Below is information from
epidemiologic and toxicological studies regarding studies of susceptibility for children and older adults.
6.2.1.1. 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. 2010a'). It is recognized that Pb can cross the placenta to affect the developing nervous system
of the fetus (Sections 4.2.2.4 and 5.3.2.1) and there is evidence of increased susceptibility to the
neurocognitive effects of Pb exposure during several lifestages throughout childhood and into
adolescence (for more detail, see Section 5.3.2.1). Epidemiologic studies have investigated susceptibility
among infants compared to adults or infants to young children.
A study including multiple U.S. locations examined associations of blood Pb levels with various
immune parameters among individuals living near Pb industries and matched controls (Sarasua et al..
2000). For several of the endpoints, the association in the youngest group (6-35 months) and the oldest
group (16-75 years) were in different 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 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 susceptibility for Pb-
associated inappropriate 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 (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 note that it is
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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.
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 susceptible to Pb
exposure (Fullmer et al.. 1985). Younger animals tend to attain a higher blood Pb level than older animals
exposed to the same dose (mg/kg body weight/day) of Pb (Berrahal et al.. 2011). Additionally, infants
may be a susceptible population because Pb easily crosses the placental barrier and accumulates in fetal
tissue during gestation (Pillai et al.. 2009; Uzbekov et al.. 2007; Y.-Y. Wang et al.. 2009).
Overall evidence indicates young age as a potential susceptibility factor 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.
6.2.1.2. 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. 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. 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). Although the hazard ratios
were greater for all-cause and cardiovascular mortality among those less than 60 years old, an increase in
the hazard ratios was also observed among those 60 years of age and older and 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 of at least 10 (ig/dL to blood Pb levels less than 5 (ig/dL and all-cause
mortality for all age groups (40-74, 75-84, and 85+ year olds), although the association for 75-84 year
olds did not reach statistical significance. The association was also present comparing blood Pb levels of
5-9 (ig/dL to blood Pb levels less than 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 comparing blood Pb levels of 10 (ig/dL
and greater to those less than 5 (ig/dL were elevated among all age groups. Finally, the association
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between blood Pb levels 10 (ig/dL and greater 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 of 5-9 (ig/dL to less than 5 (ig/dL. The other age groups
had similar point estimates but the associations were not statistically significant.
The NHANES III study cohort was also used to investigate the association between blood Pb and
cognitive test scores (Krieg et al.. 2009). The relationship was examined among adults aged 20-59 and 60
years and older, but no association was observed in either of the age groups. A study using the Normative
Aging Study cohort reported an interaction between Pb and age (R. Q. Wright. Tsaih. et al.. 2003). The
inverse association between age and cognitive function was greater among those with high blood or
patella Pb levels. Effect estimates were in the same direction for tibia Pb but the interaction was not
statistically significant.
Finally, a study of current and former Pb workers reported that an interaction term of Pb and age
(dichotomous cutpoint at 67th percentile but exact age not given) examined in a model of Pb 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.
Toxicological studies have demonstrated biological plausibility for increased susceptibility among
older populations. Demineralization associated with aging may increase the pool of available Pb to the
blood. Cory-Slechta (1990) administered various doses of Pb for a constant period to young animals,
adults, or aged animals and found increased susceptibility to Pb in the aged animals due to increased
exposure from elevated bone resorption.
Also the kidneys of older animals appear to be more susceptible to 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). Susceptibility 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). These pathologies include the development of neurofibrillary tangles and increased amyloid
precursor protein and its product beta-amyloid (Basha et al.. 2005; Zawia & Basha. 2005). Some of these
findings were seen in animals that no longer had elevated blood Pb levels.
Results for age-related modification of the association between Pb and mortality had mixed results
and no difference by age was observed for the associations between Pb and other health effects. However,
toxicological studies that inform on Pb-related health effects by age may be relevant in humans. Future
studies will be instrumental in understanding older age as a susceptibility factor.
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6.2.2. 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% were female. The distribution did vary by age with a greater prevalence of females >
65 years old compared to males (SSDAN. 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. 2006).
Multiple epidemiologic studies have examined Pb-related effects on cognition stratified by sex. In
previous studies using the Cincinnati Lead Study cohort, Dietrich et al. (1987) and Ris et al. (2004)
observed interactions between blood Pb and sex for both prenatal and postnatal exposures; associations of
prenatal and postnatal blood Pb decrements in memory, attention, and visuoconstruction were observed
only among male adolescents. More recently, Wright et al. (2008) examined early life Pb exposure and
criminal arrests in adulthood. The attributable risks were greater among males than females. Additionally,
the association between childhood blood Pb levels and gray matter volume loss was greater among males
than females (Cecil et al.. 2008). In an expanded analysis of the developmental trajectory of childhood
blood Pb levels on adult gray matter, researchers found that inverse associations between yearly mean
blood Pb levels and volume of gray matter loss were more pronounced in the frontal lobes of males than
females (Brubaker et al.. 2010). Multiple studies were also conducted in Port Pirie, Australia that
examined Pb exposures at various ages throughout childhood and adolescence (Baghurst et al.. 1992;
McMichael et al.. 1992; Tong et al.. 2000). These studies observed Pb effects on cognition were stronger
in girls throughout childhood and into early adolescence. A study in Poland also investigated the
association between cord Pb levels and cognitive deficits and reported an inverse association for boys at
36 months but not for girls (Jedrvchowski et al.. 2009). No association was detected for boys or girls at 24
months.
An epidemiologic study examined the association between 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 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 among biomarkers. A study comprised mostly of females reported positive
associations between 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 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 Pb levels and inflammatory markers (Songdei et
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al.. 2010). Although there was no pattern, a few of the associations were positive between 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). Additionally,
blood Pb was inversely associated with the release of NO among males but not females.
Epidemiologic investigations of cancer have also examined the associations by sex. A study of the
association between occupational exposure 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 inverse association was observed between ever exposure to Pb and
meningiomas for females. An ecologic analysis of Pb pollution levels and cancer incidence among
children reported weak correlations overall and the weak correlations were more apparent among males,
whereas no correlation was observed among females (Absalon & Slesak. 2010).
A study of all-cause and cardiovascular mortality using the NHANES III cohort reported no
modification of the association between 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 animals exposed to Pb. In a subsequent publication, Donald et al. (1987)
showed that non-social behavior 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 showed males experienced emotional changes and
females depression-like changes with Pb exposure (de Souza Lisboa et al.. 2005). In another study,
gestational exposure to Pb impaired memory retrieval in males at all 3 doses of Pb exposure; memory
retrieval was only impaired in low-dose females (Yang et al.. 2003).
Sex-specific differences were also observed for gross motor skills; at the lowest Pb dose, balance
and coordination were most affected among males (Lcasure et al.. 2008). In addition, obesity in adult
offspring exposed to low dose Pb in utero was reported for males but not females (Lcasure et al.. 2008).
Obesity was also found in male offspring exposed in utero to high doses of Pb that persisted to 5 weeks of
age/end of the study, but among females, 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
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 also
varied between males and females (Alghazal et al.. 2008; Sobekova et al.. 2009). 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).
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Pb and stress are co-occurring factors that act in a sex-divergent manner to affect behavior,
neurochemistry, and corticosterone. Pb and stress act synergistically to affect fixed interval operant
behavior and corticosterone in female offspring. Virgolini et al. (2008) found that effects on the central
nervous system by developmental Pb exposure are enhanced by combined maternal and offspring stress
and females are most susceptible. Behavioral related outcomes after gestational and lactational Pb
exposure with and without stress show sex-differences in exposed offspring (Virgolini. Rossi-George.
Weston, et al.. 2008). 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
exposure (pup blood Pb level 64 (ig/dL at birth) induced reductions in SDN volume in 35% of Pb-
exposed males {McGivern, 1991, 49264}. Interestingly, another chemical that is known to cause a
hypothalamic lesion in this area, monosodium glutamate, is associated with adult onset obesity; adult
onset obesity is seen in the Pb literature.
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 the
susceptible population, whereas studies from Australia pointed to females as the susceptible population. A
difference in sex is also supported by toxicological studies. Further research will confirm the presence or
absence of sex-specific associations between Pb and various health outcomes.
6.2.3. Hormones
It is possible that hormone levels may affect susceptibility to Pb-related health effects. Among
women, various hormone-related categories were examined for the relationship between blood Pb levels
and follicle stimulating hormone (FSH) and luteinizing hormone (LH) levels. A positive association was
observed between Pb levels and FSH levels among women who were post-menopausal, who were pre-
menopausal but not on birth control, menstruating, or pregnant, or who had both ovaries removed (Krieg.
2007). An inverse association was observed for women taking birth control pills. For Pb and LH, there
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was a positive association among women who were post-menopausal or had both ovaries removed. No
associations for either hormone were observed for women who were pregnant or menstruating.
Toxicological studies provide evidence that Pb affects hormones and support the possible
susceptibility to Pb-related health effects by hormonal status. Delayed onset of puberty (Pine et al.. 2006)
as well as changes in the female reproductive tract have been reported in the literature after Pb exposure.
Pb exposure can also alter estrous cyclicity (U.S. EPA. 2006). change cervical structure, and lead to
ovarian dysfunction with altered membrane composition (Kolesarova et al.. 2010; Nampoothiri et al..
2007; Nampoothiri & Gupta. 2006; U.S. EPA. 2006). Additionally, embryo development and implantation
enzymes are aberrant with Pb exposure (Nandi et al.. 2010). In female offspring, delays in F1 and F2
puberty onset, first estrous, and age of first parturition have been reported with in utero Pb exposure
(lavicoli et al.. 2006).
6.2.4. Genetics
The 2006 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. 2006). The majority
of discussion focused on aminolevulinate dehydratase (ALAD) and vitamin D receptor (VDR). These two
genes, as well as additional genes examined in recent studies, are discussed below.
6.2.4.1. Aminolevulinate Dehydratase
The aminolevulinate dehydratase 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. 2006). ALAD is a
polymorphic protein with three isoforms: ALAD 1-1, ALAD 1-2, and ALAD 2-2. Multiple studies have
examined the association between ALAD2 polymorphisms and blood Pb levels (Y. Chen et al. 2008;
Chi a et al.. 2007; Chi a et al.. 2006; Montenegro et al.. 2006; Scinicariello et al.. 2007; Scinicariello et al..
2010; Sob in et al.. 2009; Zhao et al.. 2007); ALAD polymorphisms may be biologically related to varying
Pb levels. In addition, studies have examined whether ALAD variants alters associations between Pb and
various health effects.
Associations between Pb and brain tumors observed in one study varied by ALAD genotype status
(Raiaraman et al.. 2006). Positive associations between Pb exposure and meningioma were reported
among ALAD2 individuals but this association was not found among individuals who had the ALAD 1
allele. No associations were observed between Pb and glioma regardless of ALAD genotype.
A study of current and former workers exposed to Pb examined the association between blood Pb
and blood pressure and reported no modification by ALAD genotype (Weaver et al.. 2008). However,
another study of blood Pb and blood pressure reported interactions between blood Pb and ALAD, but this
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varied by race (non-Hispanic white, non-Hispanic black, and Mexican-American) (Scinicariello et al..
2010V
Individuals with ALAD 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. 2003; Weaver et al.. 2005). A study of Chinese battery plant storage workers reported
workers with the ALAD2 allele demonstrated greater associations between Pb and renal injury (Gao et al..
2010). Another study of renal function among workers in Asia also reported greater associations between
Pb and renal function by ALAD, especially at high levels of Pb (Chi a et al.. 2006).
Studies investigating the association between Pb levels and cognitive function have also examined
modification by ALAD polymorphisms. In a study using a cohort from NHANES III, slight associations
were observed in some cognitive tests for 20-59 year olds with CC and CG ALAD genotypes (Krieg et
al.. 2009). However, other studies reported no difference in the association between blood Pb and
cognitive function by ALAD variant, although some difference was found when examining bone Pb
levels and cognitive function (Raian et al.. 2008; Weisskopf et al.. 2007; Weuve et al.. 2006).
6.2.4.2. 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).
Three genetic variants or polymorphisms of the vitamin D receptor in humans have been
characterized (BsmI, Fokl, and Apal) and have been reported to account for 75% of the differences in
bone density in humans (Morrison et al.. 1994; Morrison et al.. 1992). The BsmI polymorphisms are
denoted as bb (homozygote), BB (homozygote), and Bb (heterozygote). Bone measurements of Pb levels
in exposed workers found that bone Pb was highest in individuals with the BB genotype, intermediate in
the heterozygotes and lowest in the bb genotype group (Schwartz et al.. 2000; Theppeang et al.. 2004).
People with the bb genotype or ff (Fokl polymorphism) genotype have lower bone Pb than subjects with
other genotypes. Subjects with the aa (Apal polymorphism) or ff genotype have lower plasma Pb than
subjects with other genotypes (Rezende et al.. 2008). Plasma Pb is followed to look at bio-available Pb,
instead of blood Pb, which largely reports Pb bound to the red blood cell. Thus, subjects with the
haplotype combining a, b and f alleles for the aforementioned respective polymorphisms have lower
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plasma Pb and bone Pb levels (Rezende et al.. 2008). A study of pregnant women examining Pb in
maternal and umbilical cord serum and blood found no association between VDR polymorphisms and
these parameters. However there was lower Pb in women with the VDRhaplotype combining the alleles f
a and b, a group that is likely genetically less prone to Pb toxicity during pregnancy (Rezende et al..
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6.2.4.3. 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.
6.2.4.4. Apolipoprotein E
Apolipoprotein E (APOE) 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 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 Pb to be associated with greater decrements in 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 Pb and Bayley's MDI than children with
this allele, although the interaction term was not statistically significant (R. Q. Wright. Hu. et al.. 2003).
6.2.4.5. 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 (F. T. Wang et al.. 2007). No association between tibia Pb and cognitive
function was present for men with HFE wildtype, but a decline in function associated with 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.. 2010).
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6.2.4.6. Other Genetic Polymorphisms
Some genetic polymorphisms have only one study examining whether they modify Pb-related
health effects. These include dopamine 4 receptor (DRD4), glutathione S-transferase Mu 1 (GSTM1),
tumor necrosis factor-alpha (TNF-alpha), endothelial nitric oxide synthase (eNOS), and various SNPS.
A prospective birth cohort reported that increasing blood Pb levels were associated with poorer rule
learning and reversal, spatial span, and planning in their study population (Froehlich et al.. 2007). These
inverse associations were exacerbated among those lacking DRD4.7.
A study of university students in 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 (J. H. Kim et al.. 2007).
The relationship between blood Pb levels and inflammation was examined among individuals with
TNF-alpha GG, GA, or AA alleles. An association was present for those with TNF-alpha GG but not for
those with TNF-alpha GA or AA (J. H. Kim et al.. 2007).
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.. 2006). 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 Pb 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 glioblastma 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.
6.2.5. Pre-existing Diseases/Conditions
Studies have also been performed to examine whether certain morbidities make individuals more
susceptible to Pb-related effects on health. Recent studies have explored relationships for autism, atopy,
diabetes, and hypertension.
6.2.5.1. Autism
Rates of autism have increased in recent years. A study reported a prevalence rate in 2006 of 9.0
per 1,000 population (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. 2009b).
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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.. 2011V Blood Pb levels
among both autistic and non-autistic children were associated with expression of the inflammation genes
under study, however the associations observed were in opposite directions (positive association among
autistic children and inverse among non-autistic children).
6.2.5.2. Atopy
Atopy, a type of allergic hypersensitivity, was evaluated as a susceptibility factor in a study of Pb
and IgE (Annesi-Maesano et al.. 2003). The study examined Pb levels (measured via hair) 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 atopic mothers, the correlation was positive,
although smaller, and was not statistically significant.
6.2.5.3. 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 susceptibility 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 renal function among
diabetics 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).
The NHANES III study evaluated whether the association between 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 did not differ for individuals with and
without diabetes. However, results from the previous Pb AQCD found that individuals with diabetes are at
"increased risk of Pb-associated declines in renal function" (U.S. EPA. 2006). 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 this potential susceptibility factor.
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6.2.5.4. Hypertension
Hypertension affects approximately 24% of adults in the U.S. and the prevalence of hypertension
increases with age (61% of individuals 75 years of age and older have hypertension) (Pleis et al.. 2009).
The Normative Aging Study cohort mentioned above for 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 hypertensive but
not normotensive individuals (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 Pb and heart rate variability was not observed
among hypertensive or normotensive individuals, the estimates were different, with greater odds among
hypertensive individuals (Pb positively related to low frequency power and the ratio of low frequency to
high frequency power and inversely related to high frequency power).
A study using the NHANES III cohort reported a positive association between Pb and both all-
cause and cardiovascular mortality for hypertensive and normotensive individuals but the associations did
not differ based on hypertensive status (Menke et al.. 2006).
The 2006 Pb AQCD reported that individuals with hypertension had increased susceptibility to Pb-
related effects on renal function (U.S. EPA. 2006). 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 hypertensive individuals, but the difference between hypertensive and
normotensive adults is not observed for Pb-related mortality.
6.2.6. 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 Pb 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 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 (Di Lorenzo et al.. 2006). However, among
current smokers, higher Pb exposure was associated with higher ANC. Additionally, a positive
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relationship was observed between higher blood Pb levels and TNF-alpha 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 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 blood Pb levels and prevalence
of attention-deficit/hyperactivity disorder (ADHD). An interaction was observed between Pb and prenatal
tobacco smoke exposure; those children with high Pb levels and prenatal tobacco smoke exposure had the
highest odds of ADHD (Froehlich et al.. 2009).
Overall, the studies have mixed findings on whether smoking modifies the relationship between Pb
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 susceptibility by this factor.
6.2.7. 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. 2010b). Studies of multiple Pb-related health outcomes examined effect
measure modification by race.
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 was reported.
Another study using the NHANES III cohort reported on blood Pb levels and hypertension. While no
association was observed between blood Pb and non-Hispanic Whites or Hispanics, a positive association
was reported for non-Hispanic Blacks (Scinicariello et al.. 2010). Another study using NHANES datasets
examined the associations between blood Pb and hypertension (Muntner et al.. 2005). Although none of
the associations were statistically significant, increased odds were observed among non-Hispanic blacks
and Mexican-Americans but not for non-Hispanic whites.
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 African Americans 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 African Americans and Mexican Americans had blood Pb levels
greater than 5 (ig/dL, respectively.
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A study linking educational testing data for 4th grade students in North Carolina reported declines
in reading and mathematics scores with increasing levels of Pb (Miranda et al.. 2007). Although not
quantitatively reported, a figure depicts the association stratified by race; the slopes appear to be similar
for white and black children.
Blood Pb and asthma was examined for white and black children living in Michigan (Joseph et al..
2005). When utilizing separate referent groups for the two races, the only association is an increase
among whites (although not statistically significant), but when restricting to the highest blood Pb levels,
the association was no longer apparent. Whites with low blood Pb levels were used as the referent group
for both races in additional analysis. Although the estimates were elevated for black children compared to
white children (including at the lowest blood Pb levels), the confidence intervals for the associations
overlapped indicating a lack of a difference by race.
The results of these recent epidemiologic studies suggest that there may be race-related
susceptibility for some outcomes, although the overall understanding of potential effect measure
modification by race is limited by the small number of studies. Additionally, these results may be
cofounded by other factors, such as socioeconomic status.
6.2.8. 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. 2010c). Ris et al. (2004) examined modification of the associations between
early-life Pb exposure and Learning/IQ among adolescents in the Cincinnati Lead Study. In models
examining the association between Pb and Learning/IQ, prenatal and 78-month Pb concentrations were
associated with larger decrements in Learning/IQ in the lower two quintiles of socioeconomic status
(SES) (measured based on family SES levels).
6.2.9. 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 the two categories of body mass index
(BMI) (non-obese [<25 kg/m2] and obese [> 25 kg/m2]) but there was no difference between the
categories. Using the Normative Aging Study, 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).
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No modification by BMI/obesity was observed among recent epidemiologic studies. Future studies
of Pb-related health effects and BMI will aid in determining susceptibility by this factor.
6.2.10. Alcohol Consumption
There are a limited number of studies examining alcohol as a susceptibility factor. 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). Pb and ethanol also decreased other endogenous renal antioxidants (glutathione and non-
protein sulfhydryls) (Jurczuk et al.. 2006).
6.2.11. Nutrition
Different components of diet may affect the association between Pb and health outcomes. 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). 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 recent 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 & Antonio-Garcia. 2009). The previous Pb AQCD (U.S. EPA. 2006)
reported limited data available to assess modification by nutritional status; however, potential
modification by iron and calcium were noted. Recent epidemiologic and toxicological studies of specific
mineral intakes/dietary components are detailed below.
6.2.11.1. Calcium
Using the Normative Aging Study, researchers examined the association between Pb and
hypertension by calcium intake (Elmarsafawv et al.. 2006). The associations between Pb (measured and
modeled separately for blood, patella, and tibia) and hypertension did not differ based on dichotomized
calcium intake (800 mg/day). However, toxicological studies have shown that dietary calcium deficiency
induces increased Pb absorption and retention (Fullmer. 1992; Mvkkanen & Wasserman. 1981; Six &
Gover. 1970). Also, low calcium levels in the body stimulate the production of vitamin D and increased
synthesis of calcium-binding proteins to which Pb can bind (Richardt et al.. 1986). Increased calcium
intake reduces accumulation of Pb in bone and mobilization of Pb during pregnancy and lactation
(Bogden et al.. 1995).
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6.2.11.2. Iron
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. EPA. 2006). 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, with those girls with iron
deficiency having a stronger inverse association between Pb and inhibin B than those who were iron
sufficient (Gollenberg et al.. 2010). Toxicological studies also report 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 GD 6-14 had decreased litter size, pups with reduced fetal weight and reduced crown-
rump length, increased litter resorption, and a higher dam blood Pb level in the highest exposure groups
(Saxena et al.. 1991; Singh et al.. 1993). Thus, iron deficiency makes female rats of reproductive age
more susceptible to Pb-dependent embryo and feto-toxicity (Singh et al. 1993).
6.2.11.3. Zinc
No epidemiologic studies have been performed to examine the effect of zinc on Pb-related health
outcomes. Toxicological studies by Jamieson et al (2008; 2006) reported that a zinc deficient diet
increases bone and renal Pb content (deposition in kidney tissue) and impairs skeletal growth and
mineralization. A zinc-supplemented diet attenuated bone and renal Pb content.
6.2.11.4. Folate
A study by Kordas et al. (2009) examined Pb and birthsize among term births in Mexico City. The
authors reported no interaction between Pb and folate levels.
6.2.11.5. Protein
No recent epidemiologic studies have evaluated protein intake as a susceptibility factor for Pb-
related health effects. However, a toxicological study demonstrated that differences in maternal protein
levels could affect the extent of Pb-induced immunotoxicity among offspring (S. Chen et al.. 2004).
6.2.12. Stress
A study of bone Pb levels and hypertension reported modification of the association by perceived
stress levels (Peters et al.. 2007). Among individuals with greater stress levels, stronger associations of Pb
levels on hypertension was present. Among the same study population, higher stress was also noted to
affect the association between 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..
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2008). In another study, the 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 (Corv-Slechta et al.. 2008; Rossi-George et al.. 2009;
Virgolini. Rossi-George. Lisek. et al.. 2008; Virgolini. Rossi-George. Weston, et al. 2008). including
dysfunctional corticosterone responses (Rossi-George et al.. 2009; Virgolini. Rossi-George. Weston, et al..
2008). Additionally, toxicological studies have demonstrated that immune stress also affects associations
with Pb. Chicken with low Pb exposure in ovo and viral stressors had increased immune cell mobilization
and trafficking dysfunction (Lee et al.. 2002). Similarly, mice with neonatal Pb exposure and an immune
challenge had a sickness behavior phenotype, likely driven by IL-6 production (Dvatlov & Lawrence.
2002).
Similar to studies of stress in animals, maternal self-esteem has also been shown to modify
associations between Pb 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 observed of Pb on MDI and PDI scores; greater decreases
in MDI and PDI associated with Pb levels were observed among mothers 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 but also may 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 health effects by stress-level. Susceptibility to Pb-related
health effects by stress is supported by toxicological studies.
6.2.13. Cognitive Reserve
A study of Pb smelter workers reported that an inverse association between Pb levels and cognitive
function was present among workers with low cognitive reserve but no association was present in workers
with high cognitive reserve (Bleecker et al.. 2007). Associations between Pb and motor functions existed
among all workers regardless of cognitive reserve. No other recent epidemiologic studies were performed
examining cognitive reserve as a susceptibility factor.
6.2.14. Other Metal Exposure
The 2006 Pb AQCD reported that the majority of studies examined other toxicants as confounders
and not effect measure modifiers (U.S. EPA. 2006). Recent epidemiologic studies have, however, begun
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to explore the possible interaction between Pb and other metals. These studies, as well as toxicological
studies of these metals, are described below.
6.2.14.1. Cadmium
In a study of girls in the NHANES III cohort, inverse associations were observed between blood Pb
and inhibin B concentrations (Gollenberg et al.. 2010). These inverse 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 have reported that the addition of Cd to Pb treatment of rats 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 (G. S. Wang & Fow ler. 2008). In other studies, Cd
synergistically exacerbated Pb-dependent renal mitochondrial dysfunction (L. Wang et al.. 2009).
Overall, epidemiologic and toxicological studies have reported increased susceptibility to Pb-
related health effects among those with high Cd levels as well.
6.2.14.2. Arsenic
In a study of immune function among children living at varying distances from a Pb smelter in
Mexico, exposure to both metals were associated with greater decreases in NO and greater increases in
superoxide anion (Pineda-Zavaleta et al.. 2004). Recent toxicological studies that have examined the
addition of arsenic (As) to Pb and Cd mixtures report increases in bioavailability of Pb (G. S. Wang &
Fow ler. 2008). Thus, there is biologic plausibility of increased susceptibility of Pb-related health effects
when co-exposed to Pb and As.
6.2.14.3. Manganese
Among children in Korea taking part in a study of IQ, an interaction was reported between Pb and
manganese (Mn) (Y. Kim et al. 2009). Compared to children with low blood Mn levels, those with high
blood Mn levels had greater decreases in full scale IQ and verbal IQ associated with blood Pb levels. No
effect modification was observed for the association between Pb levels and performance IQ.
6.2.15. Fluoride
F1 has been identified as a potential susceptibility factor in a toxicological study but has not yet
been explored in epidemiologic studies. A recent toxicological study by Sawan et al. (2010) reported co-
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exposure with F1 increased Pb deposition in calcified tissues. Future investigation among humans will be
important for understanding whether fluoride present in water and other substances increases Pb
deposition in humans and modifies the association between Pb and various health effects.
6.3. Summary
Section 6.1 of this chapter provides a review of the literature regarding factors influencing Pb
exposure or biomarkers of Pb exposure. For most studies, relationships between factors and Pb biomarker
levels were presented without attribution to exposure, diet, absorption, or biokinetic factors because the
studies were not designed to make such conclusions. Where available, studies that shed light on the effect
of susceptibility factors on exposure were included. The factors examined in Section 6.1 included age,
gender, race and ethnicity, SES, and residential proximity to Pb sources.
In Section 6.2 of this chapter, epidemiologic and toxicological studies that contributed information
on potential susceptibility factors for Pb-related health effects were evaluated. Overall, this review
provided evidence that various factors may lead to increased susceptibility to Pb-related health effects
(see Table 6-3 for evidence from current studies). Section 6.2 included most of the factors from Section
6.1, plus various genes, pre-existing diseases/conditions, smoking, BMI, nutrition, stress, cognitive
reserve, and exposure to other metals.
Among children, the youngest age groups were observed to be most susceptible to having elevated
Pb body burden, with blood Pb levels decreasing with increasing age of the children. Recent
epidemiologic studies of infants/children detected susceptibility to Pb-related health effects, and this was
supported by toxicological studies.
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 (see Section 4.2). Studies
of older adults had inconsistent findings for effect 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 susceptibility to health effects.
Some studies suggest that males have higher blood Pb levels than females; this was supported by
stratifying the total sample of NHANES subjects. Gender-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 mixed, although it appears that there are some differences in associations for males and females.
This is also observed in toxicological studies. In addition, the associations among females may vary based
on hormonal status. Future research will be useful in determining which Pb-related health effects are
greater for males or females and whether hormones play a role in susceptibility.
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Regarding race and ethnicity, recent data suggest that the difference in blood Pb levels between
African American and White subjects is decreasing over time, but African Americans still tend to have
higher Pb body burden and exposures. Similarly, the gap between socioeconomic 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 as a susceptibility factor indicate that some modification of
associations between Pb and health effects may be present. Compared to whites, non-white populations
were observed to be more susceptible; 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, lower SES individuals appear to represent a susceptible population. A
study of Pb and IQ reported greater inverse associations among those in the 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 ALAD and VDR variants may be health-related
susceptibility factors. Other genes examined that may also affect susceptibility to Pb-related health effects
were MTHFR, DRD4, GSTM1, TNF-alpha, eNOS, APOE, and HFE.
Pre-existing diseases/conditions also have the potential to affect the association between Pb
exposure and various health endpoints. Recent epidemiologic studies did not support modification of Pb
and health endpoints by diabetes; however, past studies have found diabetics to be a susceptible
population with regard to renal function. More research on this population will be important for
determining which, if any, health effects related to Pb are different among diabetics. Hypertension was
observed to be a susceptibility factor in both past and recent epidemiologic studies. Studies of Pb and
both renal effects and heart rate variability demonstrated greater odds of the association among
hypertensive individuals compared to those that are normotensive. Recent epidemiologic studies also
examined autism and atopy as potential susceptibility factors. Future research will allow for a greater
understanding of potential modification by these conditions, but current research has shown that autistic
children and infants of mothers without atopy may have increased odds of Pb-induced health effects.
Recent epidemiologic studies examining smoking as a susceptibility factor reported mixed
findings. It is possible that smoking modifies the effects of only some Pb-related health effects. Further
studies of current, former, and prenatal smoking exposures related to Pb and health effects will provide
additional information on susceptibility.
BMI, alcohol consumption, and nutritional factors were examined in recent epidemiologic and
toxicological studies. Modification of associations between Pb and various health effects (mortality and
heart rate variability) was not observed by 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 possibility of alcohol as a susceptibility factor. Among nutritional factors, those with
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iron deficiencies were observed to be a susceptible population for Pb-related health effects in both
epidemiologic and toxicological studies. Other nutritional factors, such as calcium, zinc, and protein,
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 modification.
Folate was also examined in a recent epidemiologic study of birth size but no interaction was reported
between Pb and folate. Further study of these and other nutritional factors will be useful in determining
susceptibility among individuals with various nutritional levels/deficiencies.
Stress was also evaluated as a susceptibility factor and although there were a small number of
recent epidemiologic studies, increased stress was observed to negatively impact the association between
Pb and health endpoints. Toxicological studies supported this finding.
A recent epidemiologic study evaluated cognitive reserve as a modifier of the associations between
Pb and cognitive and motor functions. Cognitive reserve was an effect measure modifier for the
association between Pb and cognitive function but not motor function. Future studies evaluating Pb-
related health effects and cognitive reserve will provide more information on this possible susceptibility
factor.
Finally, interactions between Pb and other metals were evaluated in recent epidemiologic and
toxicological studies of health effects. High levels of other metals, such as Cd, As, and Mn, were observed
to negatively affect the associations between Pb and various health endpoints.
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Chapter 7 Contents
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 7-3
7.1.3. Critical Loads as an Organizing Principle for Ecological Effects of Atmospheric Deposition 7-4
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 7-7
7.2. Terrestrial Ecosystem Effects 7-8
7.2.1. Introduction to Terrestrial Ecosystem Effects 7-8
7.2.2. Soil Biogeochemistry and Chemical Effects 7-9
7.2.2.1. pHandCEC 7-9
7.2.2.2. Organic Matter 7-10
7.2.2.3. Aging 7-11
7.2.3. Bioavailability in Terrestrial Systems 7-12
7.2.3.1. Plants 7-13
7.2.3.2. Invertebrates 7-19
7.2.3.3. Terrestrial Vertebrates 7-22
7.2.3.4. Food Web 7-23
Table 7-2. Soil-to-tissue bioaccumulation factors for various terrestrial plant,
invertebrate, and vertebrate species 7-25
7.2.4. Biological Effects 7-26
7.2.4.1. Plants and Lichen 7-26
7.2.4.2. Invertebrates 7-30
7.2.4.3. Terrestrial Vertebrates 7-33
7.2.5. Exposure and Response of Terrestrial Species 7-35
7.2.6. Community and Ecosystem Effects 7-37
7.2.7. Critical Loads in Terrestrial Systems 7-41
7.2.8. Soil Screening Levels 7-42
7.2.9. Characterization of Sensitivity and Vulnerability 7-43
7.2.9.1. Species Sensitivity 7-43
7.2.9.2. Nutritional Factors 7-43
7.2.9.3. Soil Aging and Site-Specific Bioavailability 7-44
7.2.9.4. Ecosystem Vulnerability 7-44
7.2.10. Ecosystem Services 7-45
7.2.11. Summary of Effects in Terrestrial Systems 7-46
7.2.11.1. Biogeochemistry and Chemical Effects 7-46
7.2.11.2. Bioavailability and Uptake 7-47
7.2.11.3. Biological Effects 7-48
7.2.11.4. Exposure Response 7-49
7.2.11.5. Community and Ecosystem Effects 7-49
7.2.11.6. Critical Loads, Sensitivity and Vulnerability 7-50
7.3. Aquatic Ecosystem Effects 7-50
7.3.1. Introduction to Aquatic Ecosystem Effects 7-50
7.3.2. Biogeochemistry and Chemical Effects 7-52
7.3.2.1. Other Metals 7-54
7.3.2.2. Biofilm 7-55
7.3.2.3. Carbonate 7-55
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7.3.2.4. Dissolved Organic Matter (DOM) 7-56
7.3.3. Bioavailability in Aquatic Systems 7-57
7.3.3.1. Plants and Algae 7-61
Table 7-3. Bioconcentration factors for Pb in aquatic plants 7-65
7.3.3.2. Invertebrates 7-65
Table 7-4. Bioaccumulation factors for Pb in aquatic invertebrates 7-69
7.3.3.3. Vertebrates 7-69
Table 7-5. Bioaccumulation factors for Pb in fish 7-75
7.3.3.4. Food Web 7-76
7.3.4. Biological Effects 7-78
7.3.4.1. Plants and Algae 7-78
7.3.4.2. Invertebrates 7-81
7.3.4.3. Vertebrates 7-85
7.3.5. Exposure and Response of Aquatic Species 7-94
7.3.6. Community and Ecosystem Effects 7-98
7.3.7. Critical Loads in Aquatic Systems 7-101
7.3.8. Characterization of Sensitivity and Vulnerability 7-102
7.3.8.1. Seasonally-Affected Physiological Changes 7-102
7.3.8.2. Increased Nutrient Uptake 7-103
7.3.8.3. Temperature and pH 7-103
7.3.8.4. Life Stage 7-104
7.3.8.5. Species Sensitivity 7-104
7.3.8.6. Ecosystem Vulnerability 7-106
7.3.9. Ecosystem Services 7-106
7.3.10. Summary of Aquatic Effects 7-108
7.3.10.1. Biogeochemistry and Chemical Effects 7-108
7.3.10.2. Bioavailability 7-108
7.3.10.3. Biological Effects 7-110
7.3.10.4. Exposure and Response 7-111
7.3.10.5. Community and Ecosystem Effects 7-111
7.3.10.6. Critical Loads, Sensitivity and Vulnerability 7-112
7.4. Causality Determinations for Lead in Terrestrial and Aquatic Systems 7-113
Table 7-6. Summary of Pb causal determinations for plants, invertebrates and vertebrates 7-113
7.4.1. Bioaccumulation of Lead in Terrestrial and Aquatic Biota as it Affects Ecosystem Services 7-113
7.4.2. Mortality 7-115
7.4.3. Growth Effects 7-116
7.4.4. Physiological Stress 7-117
7.4.5. Hematological Effects 7-118
7.4.6. Developmental and Reproductive Effects 7-118
7.4.7. Neurobehavioral Effects 7-120
7.4.8. Other Physiological Effects 7-121
7.4.9. Community and Ecosystem Level Effects 7-121
Chapter 7. References 7-124
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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
basic concepts of metal ecotoxicology and ecosystem services. 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, and critical loads for soils. A similar
discussion of the effects of Pb on aquatic ecosystems is presented in Section 7.3, including water-only
exposures and sediment related effects. Both the terrestrial and aquatic system 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 Pb in both terrestrial and
aquatic systems are presented in Section 7.4. Areas not addressed here include literature related to 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 critical concepts for understanding how Pb from atmospheric deposition may affect
Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and Environmental
Research Online) at http://eDa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of developing science
assessments such as the Integrated Science Assessments (ISA) and the Integrated Risk Information System (IRIS).
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organisms, communities, and ecosystems. The sections that follow provide more detail for how aquatic
and terrestrial ecosystems respond to Pb and how environmental chemistry interacts with organisms to
affect exposure and uptake.
7.1.1. Ecosystem Scale, Function, and Structure
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. 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. 2008). 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.
Ecosystems are most often defined by their structure, and are based on the number and type of
species present. 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 et al.. 2005).
Individual plants or animals may exhibit changes in metabolism, enzyme activities, hormone function, or
overall growth rates or may suffer gross lesions, tumors, deformities, or other pathologies. Effects on the
nervous system of animals may cause behavioral changes that alter breeding behaviors or predator
avoidance. However, effects on individuals must result in changes to their survival or reproductive output
to have any effect on the population. Population level effects of pollutants include changes overtime 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
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removing one species from the assemblage and allowing another to emerge (Petraitis & Latham. 1999):
and (3) alterations in trophic structure.
Alternatively, ecosystems may be defined on a functional basis, 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 (Baitell. 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 (Baitell. 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.. 1989V 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 (Baitell. 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.
7.1.2. Ecosystem Services
Ecosystem structure and function may be translated into ecosystem services (Dailv. 1997).
Ecosystem services are the benefits people obtain from ecosystems (Millennium Ecosvstem Assessment.
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 (Dailv. 1997; Westman.
1977; WRI. 2000). These ecological benefits, in turn, provide economic benefits and values to society
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(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 &
Kennedy. 1997V
Particular concern has developed within the past decade regarding the consequences of decreasing
biological diversity (Avensu et al.. 1999; Chapin et al.. 1998; Hooper & Vitousek. 1997; Tilman. 2000;
Wall. 1999). 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 (Chapin et al.. 1998; Dailv & Ehrlich. 1999; Levi in. 1998; Peterson et al.. 1998; Pimm.
1984; Tilman. 1996; Tilman & Downing. 1994; Wall. 1999). Biodiversity is an important consideration at
all levels of biological organization, including species, individuals, 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 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 (Gitav et al.. 2001) i.e., food (including game meat,
roots, seeds, nuts, and other fruit, spices, fodder), 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
Critical loads were first defined for regulating emissions of sulfur and nitrogen oxides, but have
since been applied to exposure of other pollutants, including metals (Adams & Chapman. 2007). A critical
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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 & Grennfelt. 1988). Because critical loads for Pb differ by ecosystem
(aquatic-water; aquatic-sediment; terrestrial), differ by environmental chemistry properties that impact
bioavailability, differ by species, and differ by endpoint of concern, they can be used as an organizing
principle for linking atmospheric deposition with ecological impairment at multiple spatial scales. There
are two aspects to consider: (1) the critical load at a steady state in the environment (i.e., how much input
is required to balance the rate of output), and (2) the time required to reach the critical load (i.e., the lag
time between onset of exposure and induction of measurable effects). This is particularly true for
terrestrial ecosystems where changes in soil geochemistry, as a result of either changing land use or
ecological succession, may significantly alter the amount of sequestration of Pb, thus changing its
bioavailability and critical load if based on total metal. Ideally, therefore, critical loads for metals should
be defined on the basis of bioavailable metal rather than total metal. This approach is being used in
aquatic systems through the application of the biotic ligand model (BLM) (Di Toro et al.. 2001; Di Toro et
aL 2005). but is proving to be more difficult for modeling terrestrial systems.
For aquatic systems, a dynamic equilibrium exists between the surface water, the water column,
and the sediment compartments (which must be defined when determining the critical load for each
compartment). Although the sediment generally acts as a sink for pollutants in the water column,
especially metals in particle form or as insoluble metal complexes, there may be some re-entrainment
from the sediments into the water column. This sediment-water interface may change the solubility or
bioavailability of the metal, thereby altering the critical load for the water column, particularly if
expressed in the form of total metal. Note, however, that while sedimentation processes may change the
time to steady state, they will not affect the ultimate critical load once steady state is achieved.
The following pieces of information are required to calculate a critical load, each of which is
discussed in more detail in the subsequent sections of this chapter:
¦ Ecosystem at risk;
¦ Receptors of concern (plants, animals, etc.);
¦ Endpoints of concern (organism, population or community responses, changes in ecosystem
services or functions);
¦ Dose (concentration) - response relationships and threshold levels of effects;
¦ Bioavailability and bioaccumulation rates;
¦ Naturally occurring (background) Pb (or other metal) concentrations; and
¦ Biogeochemical modifiers of exposure.
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As stated in the 2008 ISA for Oxides of Nitrogen and Sulfur-Ecological Criteria 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. 2008). 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.
In the short term, metal emissions generally have greater adverse 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.
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 the modifying factors listed in the previous section. 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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0.69x^x10000
*1/2 ~ n
yxTF +
pKd
Equation 7-1
where:
d is the soil depth in meters (m)
y is the annual crop yield (tons/ha yr)
TF is the ratio of the metal concentration in plant to that in soil
R is the net drainage loss out of the soil depth of concern (m3/ha yr)
P is the bulk density of soil (kg(dry weight/L)
Kd is the ratio of the metal concentration in soil to that in soil pore solution (L/kg)
Metals removed by crops (or plants in general) comprise a very small fraction of the total soil metal
and can be ignored for the purpose of estimating time to steady state. Thus, equation 7-1 is
simplified to:
0.69 xdx 10000
hn-
pKd
Equation 7-2
and becomes a function of soil depth, the amount of rainfall, soil density, and soil properties that
affect Kd. Pb has a relatively long time to steady state compared to other metals, as shown in
Table 7-1.
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
O
o
_Q
Q_
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) (de Groot et al.
1998).
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, Fairborother et al. (2007)
In aquatic systems, ty2 for Pb in the water column depends on the ratio of the magnitudes of the
fluxes coming from and going into the sediment, the ratio of the depths of the water column and sediment,
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and the sediment half-life. Sediment Xm is dependent upon the particulate and dissolved fractions and is
calculated as for soils (Equation 7-2).
Re-entrainment of Pb particles via windblown dust from surface soils or dry sediments may occur.
Amount and distance of re-entrained particles and deposition rates are dependent upon wind velocity and
frequency; size, density, shape, and roughness of the particle; soil or sediment moisture; and terrain
features including openness (including amount of vegetation), aspect relative to wind direction, and
surface roughness. Resuspension is defined in terms of a resuspension factor, K, with units of m"1, or a
resuspension rate (A), with units of sec"1 (Equation 7-3). The resuspension rate, A, is the fraction of a
surface contaminant that is released per time and is defined by:
A = *
c
Equation 7-3
where:
R is the upward resuspension flux ((.ig/nr/sec)
C is the soil (or dry sediment) Pb concentration (|ig/m2)
Such emissions may have local impacts, but are not likely to have long-range effects, as particles
generally remain low to the ground and are not lifted into the upper atmosphere. Although re-
entrainment may alter the particle size distribution in a local area, it generally does not alter the
bioavailable fraction, and deposited particles will be subject to the same biogeochemical forces
affecting bioavailability. Therefore, exposure via re-entrainment should be considered additive to
exposure from atmospheric particulate deposition in terrestrial and aquatic ecosystems.
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
2006 Pb AQCD. The literature on terrestrial ecosystem effects of Pb, published since the 2006 Pb AQCD,
is considered with brief summaries from the 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 in terrestrial systems is reviewed in Section 7.2.2. Section 7.2.3. considers the
bioavailability and uptake of Pb by plants, invertebrates, and wildlife 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 scale are discussed in Section 7.2.6. Section 7.2 concludes with a
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discussion of critical loads in terrestrial systems (Section 7.2.7), characterization of sensitivity and
vulnerability of ecosystem components (Section 7.2.8), and effects on ecosystem services (Section 7.2.9).
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-desorption relationship of total Pb with
inorganic compounds (e.g., oxides of Al, Fe, Si, Mn; clay minerals); (3) adsorption-desorption reactions
of dissolved Pb phases on soil organic matter; (4) pH; (5) cation exchange capacity (CEC); and (6) aging.
Adsorption-desorption of Pb to soil solid phases is largely controlled by total metal loading. Therefore,
areas with high Pb deposition will exhibit a lower fraction of total Pb partitioned to inorganic and organic
matter. Decreasing soil pH, CEC, and organic matter have been strongly correlated to increases in the
concentration of dissolved Pb species. Aging of metals in soils results in decreased amounts of labile
metal as the Pb becomes incorporated into the soil solid phase (McLaughlin et al.. 2010). Data from
recent studies have further defined the impact of pH, CEC, organic matter (OM), and aging on Pb
mobilization and subsequent bioavailability in soils.
7.2.2.1. pH and CEC
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 about 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 both (1) Fe and Mn oxide
interactions; and (2) the formation of weak electrostatic bonds with charged soil surfaces. Similarly, the
mobility of smelter-produced metals in forest soils was found to be greater than in adjacent agricultural
lands (Douav et al. 2009). The higher solubility was caused by the decreased soil pH of the forest
environments. Further, decreasing the soil pH via simulated acid rain events increased naturally occurring
Pb bioavailability in field tests (X. Hu et al. 2009).
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)
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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.
7.2.2.2. Organic Matter
Organic matter (OM) 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, OM, total calcium carbonate equivalent, and
total free sesquioxides also exhibited higher total Pb concentration, indicating that less Pb had been
removed 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 OM binding affinity and uncertainties associated with metal speciation. They
found that natural variations in fulvic acid binding properties were the most important variable in
predicting Pb speciation. Guo et al. (2006) determined that the -COOH and -OH groups associated with
soil OM were important factors in Pb sequestration in soil, and Pb sorption was increased as pH was
raised from 2 to 8. Because organic content increased the Pb sequestration efficiency of soils, OM content
had an inhibitory effect on Pb uptake by woodlouse species Oniscus asellus and Porcellio scaber (Gal et
al.. 2008). Vermeulen et al. (2009) demonstrated that invertebrate bioaccumulation of Pb from
contaminated soils was dependent on pH and OM, but that other unidentified habitat-specific differences
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 atmosphere-deposited Pb
in soil matrices was strongly influenced by soil type, indicating that certain soil types may retain Pb for
longer periods of time. In humic forest soils, the highest Pb concentrations were measured in the humified
bottom layer, whereas 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
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 for 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
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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).
7.2.2.3. Aging
Smolders et al. (2007) reviewed the effects of aging of Pb in soils on the toxicity of Pb to plants
and soil invertebrates, with aging defined primarily as leaching following initial influx, but also as
binding and complexation. In nearly half of the Pb soil studies reviewed, observed dose-response curves
could not be established 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 soil 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 7.3.3), the authors
proposed a terrestrial BLM approach to estimate the risk of metals to terrestrial organisms. However,
Antunes et al. (2006) noted that there were several key challenges involved in development of a terrestrial
BLM applicable to plants, particularly the reliable measurement of free ion activities and ligand
concentration in the rhizosphere, the identification of the organisms' ligands associated with toxicity, and
the possible need to incorporate kinetic dissolution of metal-ligand complexes as sources of free ion.
Further, Pb in aged field soils has been observed to be less available for uptake into terrestrial organisms,
likely as a result of increased sequestration within the soil particles (Antunes et al.. 2006). Magrisso et al.
(2009) used a bioluminescent strain of the bacterium Cupriavidus metallidurans to detect and quantify Pb
bioavailability in soils collected adjacent to industrial and highway areas in Jerusalem, Israel, and in
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/L, whereas reproduction was always significantly affected at Pb
concentrations of 0.7 mg/L and higher, independent of the soil treatment. Leaching soils prior to use in
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bioassays had only a slight effect on Pb toxicity to resident springtails, suggesting that among the
processes that constitute aging of Pb in field soils, leaching is not particularly important with respect to
bioavailability.
Red-backed salamanders (Plethodon cinereus) exposed to Pb-amended soils (553, 1,700, 4,700,
and 9,167 mg Pb/kg) exhibited lowered appetite and decreased white blood cell counts at the two highest
concentrations, as compared to controls {B, 2010, 379076}. 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 for 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)." 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). Characteristics of the toxicant that affect bioavailability are: (1)
chemical form or species; (2) particle size; (3) lability; and (4) source. 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. Bioaccumulation factors (BAF's) (i.e., the ratio of Pb concentrations in tissues of terrestrial
biota to concentrations in soil or in food items) reported in this section are summarized in Table 7-2. The
2006 Pb AQCD extensively reviewed the methods available for quantitative determination of the
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mobility, distribution, uptake, and fluxes of atmospherically delivered Pb in ecosystems, and they are not
reviewed in this section.
7.2.3.1. Plants
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.
Leaf and Root Uptake
Field studies carried out in the vicinity of Pb smelters have determined the relative importance of
direct foliar uptake and root uptake of atmospheric Pb deposited in soils. Hu and Ding (2009) analyzed
ratios of Pb isotopes in the shoots of commonly grown vegetables and in soil at three distances from a
point 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 Rostellaria procumbens, ratio = 1.7). However, the two species
studied at Site II were not studied at Site I or Site III. In the control site (Site III), no plant was found with
a Pb 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,
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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.
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.
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 determined to be from direct adsorption from the
atmosphere. Further, Aznar et al. (2009) 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, which the authors attributed to "incidental processing" of atmospheric Pb. Foliar damage
and growth reduction were also observed in the trees (Aznar. Richer-Lafleche. et al.. 2009). They were
significantly correlated with Pb concentration in the litter layer, where Pb comes from atmospheric
deposition and closely reflects it. 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 adversely affecting trees (Aznar. Richer-Laflechea. et al.. 2009). 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). The advent of laser ablation inductively coupled
plasma mass spectrometry has made measurement of Pb concentration in individual tree rings possible
(Watmough. 1999; Witte et al.. 2004). 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. (2008) 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
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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 (Acerpseudoplatanus) 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 (L. 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 & 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 & 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 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
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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-m 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 & Al-Khashman. 2007). 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 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).
Translocation and Sequestration of Lead in Plants
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. 2006). 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. 2006).
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
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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/kg) by Probst et al. (2009) revealed dark ultrastructural
abnormalities that were demonstrated to be metal-rich particles located in or on root cell walls. It is
unclear whether the presence of these structures had any effect on overall plant health.
Clark et al. (2006) investigated Pb bioavailability in garden soils in Roxbury and Dorchester, MA.
The sources of Pb were considered to be Pb from paints and from leaded gasoline additives, with 40 to
80% coming from paint. The average Pb concentration in foliar tissue of bean plants was 14 ± 5 mg Pb/kg
while the concentration in the bean pod was only 20.6 mg Pb/kg. For mustard plants, there was a linear
relationship (R2=0.85) between Pb concentration in plant tissues and Pb concentration in the soil (both for
plants grown in situ and those grown under greenhouse conditions).
Murray et al. (2009) investigated the uptake and accumulation of Pb in several vegetable species
(carrot [Daucus carota\, radish [Raphanus sativus], lettuce [Lactuca sativa], soybean [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/kg dry mass. Similarly, (C'ho et al. 2009) showed that green onion (Allium
Jistulosum) 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). 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).
Recent research has shown that Pb translocation to stem and leaf tissues does occur at significant
rates in some species, including the legume Sesbania drummondii (Peralta-Videa et al.. 2009) and
buckwheat (Fagopyrum esculentum) (Tamura et al.. 2005). Wang et al. (2006) 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,
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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 & Tsadilas. 2009).
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. (2009) 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 shoot concentration of 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 Jinping. These results indicate significant interspecies differences in
Pb uptake, as well as potential soil-specific 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 (Y. Li et al.. 2009). 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.
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.
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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 (G. Wang et al.. 2006) and buckwheat has recently been shown to be an efficient translocator
of Pb from soil to above-ground shoots (Tarnura et al.. 2005). However, there is broad variability in Pb
uptake and translocation rates among plant species, and interspecies variability has been shown to interact
with other factors such as soil type. Field studies carried out in the vicinity of Pb smelters (X. Hu et al.
2009) show that Pb may accumulate in shoot tissue through direct stomatal uptake rather than by soil-
root-shoot translocation. 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 etal.. 1991).
7.2.3.2. 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 exposure
via soil, and that uptake and bioaccumulation of Pb by invertebrates is lower than that observed for other
metals.
Snails
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 & Richmond. 2010). While bioconcentration factors
(BCF's) for Cd were observed to increase over the 64-day study period, the rate of Pb assimilation
remained consistent over time and no evidence for a regulatory mechanism for Pb was observed. The
authors observed that, for additional Pb to be retained, snails would have to grow additional soft tissue.
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. De
Vaufleurv. et al.. 2006). There were no observed effects of Pb exposure or accumulation on survival or
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growth in C. asperses or H. aspersa. In another study (Ebenso & 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 1,200 mg Pb/kg at the three
main waste sites, while leaf tissues of radish (Raphanus sativus) grown at these sites averaged 7, 30, and
68 mg Pb/kg dry weight, respectively. Although plant concentrations were low, they were correlated with
elevated snail Pb tissue concentration. 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). 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 (1999 European 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/kg in
control organisms. Viscera also exhibited increased Pb levels at the two highest exposures, with measured
tissue concentration of 1.2 and 1.1 mg Pb/kg, respectively, as compared with control tissue Pb levels of
0.4 mg Pb/kg. However, there was no significant increase in snail-tissue Pb concentration when natural
soil was used in place of ISO medium, and there was no relationship between soil Pb concentration and
snail tissue concentration, strongly suggesting the presence of soil variables that modify bioavailability.
Notten et al. (2008) investigated the origin of Pb pollution in soil, plants, and snails by means of Pb
isotope ratios. They found that a substantial proportion of Pb in both plants and snails was from current
atmospheric exposure.
Earthworms
Soil characteristics that interact with bioavailability of Pb may include biogeochemistry associated
with different soil horizons, source of Pb, and proportion of soil: leaf litter. These have been studied
principally with various species of earthworms. 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/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.
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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 bioaccumulation in the endogeic earth wo rm Me iaph ire
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. Accumulation
studies conducted with Eisenia fetida earthworms documented the difficulty of extrapolating
accumulation kinetic constants from one soil type to another, and showed that many soil physiochemical
properties, including pH, OM, and CEC, among others, work in conjunction to affect metal bioavailability
(Nahmani et al.. 2009). However, once taken up from the environment, more than half of the
bioaccumulated Pb appears to be contained within earthworm tissue and cell membranes (Li et al.. 2008).
Despite significant Pb uptake by earthworms, Pb in earthworm tissue may not be bioavailable to
predators. Pb in the earthworm Aporrecioc/ea 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). From this, the authors concluded that only a
minority of earthworm-absorbed Pb would be toxicologically available to cause adverse effects in the
earthworms or in their predators. Earthworm activity can alter Pb bioavailability and subsequent uptake
by earthworms themselves and other organisms. The presence of earthworms may increase soil pH though
the secretion of cutaneous mucus, and worm activity is generally associated with increased bioavailability
of Pb. Sizmur and Hodson (2009) speculated that earthworms affect Pb mobility by modifying the
availability of cations or anions. 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.
Arthropods
Cicadas pupating in historically Pb-arsenate-treated soils accumulated Pb at concentrations similar
to those reported previously for earthworms (Robinson et al.. 2007). Likewise, tissue Pb levels measured
in Coleoptera specimens collected from areas containing average soil concentration of 45 and 71 mg
Pb/kg exhibited a positive relationship with soil Pb content, although abundance was unaffected (Schipper
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et al.. 2008). By contrast, the Pb sequestration rates that were observed in two woodlouse species, O.
asellus and/! scaber, were species-dependent (Gal et al.. 2008V Both species were field collected at Pb-
contaminated sites (average concentration, 245 mg Pb/kg dry weight; range, 21-638 mg Pb/kg dry
weight), with O. asellus Pb levels averaging 43 mg Pb/kg over all sites, while P. scaber contained no
detectable Pb residues. Pb concentration measured in granivorous rough harvester ants (Pogonomyrmex
rugosus), in the seeds of some plant species they consume, and in surface soil, were all shown to decline
with increasing distance from a former Pb smelter near El Paso, Texas, where soil leachable Pb at the
three sites of ant collection ranged from 0.003 to 0.117 mg Pb/kg (Del Toro et al.. 2010). Ants
accumulated approximately twice as much Pb as was measured in seeds, but the study did not separate the
effects of dietary exposure from those of direct contact with soil or respiratory intake.
7.2.3.3. Terrestrial Vertebrates
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 |ag Pb/dl, with levels between 20 and 59 jj.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 coramandelicus) following oral dosing of
0.5 mg/kg, 1.25 mg/kg or 2.5 mg/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 & 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
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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).
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).
Interestingly, 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/L Pb (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 individual 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 particulates 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
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bioavailability to mammals (Kaufman et al.. 2007). Earthworms were determined to contain an average of
727 mg Pb/kg, and the Pb content of unwashed leaf tissues averaged 2,945 mg Pb/kg. Canonical
correspondence analysis detected no relationship between earthworm and soil Pb concentration, but did
show correlation between unwashed vegetation and soil concentration. The authors noted that the
relatively high Pb concentration of unwashed as opposed to washed vegetation indicated the potential
importance of aerial deposition (or dust resuspension) in determining total vegetative Pb concentration.
Based on the mammalian gastric model, they noted that 50% of vegetation tissue Pb and 77% of
earthworm tissue Pb was expected to be bioavailable to consumers. The avian gizzard model indicated
that 53% of soil Pb and 73% of earthworm Pb was bioaccessible to birds, and, for both mammals and
birds, the bioaccessible fraction of Pb was a function of total Pb concentration.
The transfer of Pb from soils contaminated by a Pb-zinc mine was limited along a soil-plant-insect-
chicken food chain (Zhuang et al.. 2009). In soils averaging 991 mg Pb/kg, Rumex K-l plants sequestered
an average of 1.6 mg/kg wet weight Pb 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 holm oak (Quercus ilex), gum rockrose (Cistus ladanifer), elmleaf
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 concentration were not significantly greater than those of the non-mining areas.
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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/! 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/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 periodically returns to background level with each shedding. Granular
Pb is unavailable. Trophic attenuation is thus likely.
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.
Table 7-2. Soil-to-tissue bioaccumulation factors for various terrestrial plant,
invertebrate, and vertebrate species
Species
Soil Concentration
Tissue
Type
Tissue
Concentration
Bioaccumulation
Factor
Reference
Vegetable
379 mg/kg
Root
47.3 mg/kg
0.12
X. Hu & Ding (2009)
Vegetable
379 mg/kg
Shoot
36.2 mg/kg
0.1
X. Hu & Ding (2009)
P. angulata
4020 mg/kg
Root
527.1 mg/kg
0.13
Cui etal. (2007)
P. angulata
4020 mg/kg
Shoot
331.1 mg/kg
0.08
Cui etal. (2007)
A. theophrasti
4020 mg/kg
Root
38.7 mg/kg
0.01
Cui etal. (2007)
A. theophrasti
4020 mg/kg
Shoot
61.4 mg/kg
0.02
Cui etal. (2007)
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Species
Soil Concentration
Tissue
Type
Tissue
Concentration
Bioaccumulation
Factor
Reference
S. vestita
2409 mg/kg
Shoot
3142 mg/kg
1.3
Y.Li etal. (2009)
S. vestita
2409 mg/kg
Root
7457 mg/kg
3.1
Y.Li etal. (2009)
S. grevilloides 5330 mg/kg
Shoot
69 mg/kg
0.01
Y.Li etal. (2009)
D. carota
40.1 mg/kg
Root
12.2 mg/kg
0.3
Murray etal. (2009)
A. fistulosum
3560 mg/kg
Whole plant 23.7 mg/kg
0.007
Cho etal. (2009)
C. baccatum
0.6 mg/kg
Fruit
0.86 mg/kg
1.4
Antonious & Kochhar (2009)
A. achatina
1200 mg/kg
Whole body 468.5 mg/kg
0.39
Ebenson & Ologhobo (2009b)
A. achatina
200 mg/kg
Whole body 91.4 mg/kg
0.46
Ebenson & Ologhobo (2009b)
A. achatina
20 mg/kg
Whole body 12.2 mg/kg
0.61
Ebenson & Ologhobo (2009b)
H. aspersa
39 mg/kg
Viscera
1.1 mg/kg
0.03
De Vaufleury et al. (2006)
H. aspersa
52 mg/kg
Viscera
1.2 mg/kg
0.02
De Vaufleury et al. (2006)
L. terrestris
710 mg/kg
Whole body 20.6 mg/kg
0.03
Darling & Thomas (2005)
O. asellus
245 mg/kg
Whole body 43 mg/kg
0.18
Gal et al. (2008)
T. taipoides
111 mg/kg
Liver
1.63 mg/kg
0.002
Reynolds etal. (2006)
T. taipoides
111 mg/kg
Whole body 3.85 mg/kg
0.005
Reynolds etal. (2006)
L. limosa
336 mg/kg
Egg
0.17 mg/kg
0.0005
Roodbergen et al. (2008)
L. limosa
336 mg/kg
Feather
2.79 mg/kg
0.008
Roodbergen et al. (2008)
7.2.4. Biological Effects
Various adverse 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. 2004). Despite a few well-known examples of metal antagonism (e.g., Cu and Mo or Cd and
Zn), it is common practice to assume additivity of effects (Fairbrother et al.. 2007). Because this review is
focused on effects of Pb, studies reviewed for this section and the following include only those for which
Pb was the only, or primary, metal to which the organism was exposed.
7.2.4.1. 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,
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antioxidant capacity is exceeded, and metal exposure causes peroxidation of lipids and DNA damage,
eventually leading to accelerated senescence and potentially death (Stobrawa & Lorcnc-Plucinska. 2008).
Effects on Photosystem and Chlorophyll
The effect of Pb exposure on the structure and function of plant photosystem II was studied in giant
duckweed, Spirodela polyrrhiza (Ling & Hong. 2009). Although this is an aquatic plant, photosystem II is
present in all plants, and effects on photosystem II observed in any plant species are likely to occur in all
of them. 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. (2008) 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. Liu et al.
(2010) 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 & 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 &
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. As compared to other metals,
Pb caused less physiological damage, which the authors attributed to the metal's high affinity for binding
to and sequestration within cell walls.
The effect of Pb exposure on chlorophyll content of the moss and liverwort species Thuidium
delicatulum, T. sparsifolium, and Ptychanthus striatus was investigated following simulated atmospheric
exposures of 10"10 to 10"2 M Pb (Shakva et al.. 2008). Both chlorophyll a and total chlorophyll content of
the mosses delicatulum and T. sparsifolium) decreased with increasing Pb exposure, but the effect was
not statistically significant. For the liverwort, Pb exposure resulted in significant 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 have been demonstrated in broad bean (Vicia faba) (C.-R. Wang et
al.. 2010; C.-R. Wang. Wang. Tian. Yu. et al.. 2008; C. Wang. Y. Tian. et al.. 2010) and tomato
(Lycopersicon esculentum) (C.-R. Wang. Wang. Tian. Xue. et al.. 2008). where they were accompanied by
proportional increases in superoxide dismutase (SOD), glutathione, guaiacol peroxidase, as well as lipid
peroxidation, and decreases in catalase. 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 for M uniflorum. Antioxidant activities were also markedly greater in the root tissues than
the shoot tissues of the two plants, and were very likely related to the higher Pb accumulation rate of the
roots. The effectiveness of the up-regulation of antioxidant systems in preventing damage from Pb uptake
was evidenced by lower malondialdehyde (MDA) (a chemical marker of lipid peroxidation) concentration
inM uniflorum versus C. arietinum, indicating a lower rate of lipid peroxidation in response toM
uniflorum's higher antioxidant activity.
Gupta et al. (2010) contrasted responses of two ecotypes 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 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.. 2010) 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,
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carotenoids, photosystem II activity, C02 gas exchange (respiration), and MDA content. Lichen Pb levels
were significantly correlated only with MDA content.
Growth
There is evidence of effects of Pb on higher growth processes as well. 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. Exposure to 600 (.iM Pb and greater decreased carotenoid content, while
chlorophyll a was decreased at Pb exposures of 450 (.iM and higher. However, no effects were seen in
growth or chlorophyll production of maize (Zea mays) following growth in smelter ash-spiked soils
containing 1,466 mg Pb/kg (and 18.6 mg Cd/kg) (Komarek et al.. 2009). Pb concentrations of 7,331
mg/kg (98.0 mg/kg Cd) were required to elicit chlorosis and the expected decreased in growth.
Chinese cabbage (Brassica pekinensis) exposed to Pb-containing soils (0, 4, and 8 mM/kg dry
weight) 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 concentration of 0.2, 0.4, and 0.8 mg/kg dry weight
total Pb and collected leaf samples for 11 days. At exposures of 0.4 and 0.8 mg/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.
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 & 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.
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 & Tripathi. 2008) Plants
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.
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Interestingly, as in zebra finch (Section 7.2.3.3), Ca was found to counteract the toxic 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/L in the presence of Ca concentration of 1.2 mM and higher (Antosiewicz. 2005).
7.2.4.2. 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 & 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 response to Pb exposure (2.5 (.iM) (Xing. Guo. et al.. 2009). 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. Rui. et al.. 2009). 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 (D. Y. Wang & Xing. 2008). Chemotaxis towards NaCl, cAMP,
and biotin was also decreased in C. elegans nematodes exposed to Pb concentration greater than 2.5 (.iM
(Xing. Du. et al.. 2009). 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 (D. Y. Wang & Xing. 2008).
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 (D. Y. Wang & 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
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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.
E. andrei earthworms exposed to 21 different soils, each containing 2,000 mg/kg freshly added Pb,
for 28 days exhibited highly variable mortality, ranging from 0% to 100%, (Bradham et al.. 2006). Pb
body burden of exposed worms ranged from 29 to 782 mg Pb/kg. Internal Pb concentration was also
negatively correlated to reproductive output. CEC and pH were found to be the principal soil
characteristics determining the differences in those effects, although the apparent role of CEC may only
have been due to its correlation with other soil characteristics. Low soil Pb concentration (5 mg/kg) also
decreased the protein content of E. fetida earthworms during a 7-day exposure (M. Li et al.. 2009). 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), indicating that Pb exerted a detrimental effect on worm
metabolism. By contrast, Svendsen et al. (2007) found that L. 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.
As in plants, induction of antioxidant activity is affected by exposure to Pb in invertebrates. The
induction of antioxidant activity was correlated to standard toxicity measurements in Theba pisana snails
(Radw an et al.. 2010). Topical application of Pb solutions (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. 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. Further, decreased
food consumption, growth, and shell thickness were observed in juvenile A. achatina snails exposed to
Pb-contaminated (concentration greater than 134 mg/kg) diet for 12 weeks (Ebenso & 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/kg, a concentration-dependent relationship was
established for growth, with significant reduction observed at 12-week exposures to 20 mg Pb/kg (Ebenso
& Qloghobo. 2009b). However, consumption of field-collected Pb-polluted U. dioica leaves containing 3
mg/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).
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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.
Lock et al. (2006) compared the toxicity of both laboratory-spiked soils and field-collected Pb-
contaminated soils to springtails (F. Candida). The 28-day EC50 values derived for it 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 adverse effect on Sinella curviseta springtail survival or reproduction (Xu et al.. 2009).
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 & 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."
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 & 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
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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
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 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.
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
intraperitoneal 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
concentration 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 adverse
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. Coeurdassier. et al.. 2006). 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.
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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
(Bemlund et al.. 2010). Moss samples indicated that Pb deposition in study areas ranged between 100 and
2,000 mg Pb/kg dry weight during operations and 200 and 750 mg Pb/kg when operations ceased. A
simultaneous slight reduction was observed in pied flycatcher blood Pb levels, from 0.4 to 0.3 mg Pb/kg).
However, clutch size was decreased in pied flycatchers inhabiting the mining area both during and after
mining operations, and mean nestling mortality was 2.5 and 1.7 higher after mining operations in the
mining region than in reference areas, respectively. 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.
The level 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 (ig/L (±51), which was
higher than blood Pb levels from the mining area (44 ± 34 (.ig/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
intraperitoneal 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.
Again, dietary or other health deficiencies unrelated to Pb exposure are likely to exacerbate the
effects of Pb. Ca-deficient female zebra finches guttata) had a suppressed secondary humoral immune
response following 28-day exposures to 20 mg/L Pb 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 adversely affect several mammalian species.
Young adult rats reared on a diet containing 1,500 mg/kg Pb acetate for 50 days demonstrated less
plasticity in learning than non-exposed rats (McGlothan et al.. 2008). indicating that Pb exposure caused
significantly altered 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, lutenizing hormone, and pituitary growth hormone were
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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-zinc smelters in India, blood Pb levels were
positively correlated with plasma levels of the thyroid hormones thyroxine (T4) and tri-iodothyronine
(T3) and the hepatic biomarkers alanine transaminase and aspartate transaminase (Swarup et al.. 2007).
Total lipids, total protein and albumin levels were decreased in the same animals.
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 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.. 2009).
7.2.5. Exposure and Response of Terrestrial Species
Given that exposure to Pb may adversely affect organisms at the individual, population, or
community level, determining the rate and concentration at which these effects occur is essential in
predicting the overall risk to terrestrial organisms. However, data from controlled studies using a single
compound are scarce relative to field studies, which in turn often investigate effects of multiple metal
contaminants and afford too little control on interacting variables to be of use in establishing a general
dose-response function. 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 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.. 2009). 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
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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/kg and 1,000 mg/kg Pb, 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 soil, there
was no statistical relationship with earthworm survival during a 42-day exposure period (Nahmani et al..
2007). However, Pb concentration in soil leachate solution was significantly correlated with decreased
earthworm survival and growth (linear regression: R2= 0.64, p< 0.0001). The 42-day Pb EC50 for E. fetida
growth was 6,670 mg Pb/kg.
Langdon et al. (2005) exposed three earthworm species (E. andrei, L. rubellus, and A caliginosa)
to Pb nitrate-amended soils at concentrations of 1,000 to 10,000 mg Pb/kg to determine species variability
in uptake and sensitivity. Twenty-eight-day LC50 values for the three species were 5,824 mg Pb/kg, 2,867
mg Pb/kg, and 2,747 mg Pb/kg, respectively, indicating that L. 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 (2009) reported LC50 values for E. fetida that are similar to
those for L. rubellus and A. caliginosa. It is likely that these apparent species differences are a result of
differential bioavailability of the Pb in test soils. However, the Pb body burden of all three species in the
study by Langdon et al. (2005) increased with increasing environmental concentration, and there were no
species differences in Pb tissue content. When given a choice between treated and untreated soils, all
worm species exhibited significant avoidance of Pb-contaminated soils, and altering pH (and,
consequently, Pb bioavailability) had no impact on avoidance (Langdon et al. 2005). Field earthworms
may thus be able to reduce their exposure to Pb through behavior.
The individual and population-level responses of the springtail Paronychiurus kimi to Pb were
determined by Son et al. (2007) using artificial soils, following the 1999 European 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/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 I'
kimi springtails exposed to Pb, and proposed that more specific measures of population-level endpoints
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(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.
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 biotic ligand model proposed by Smolders et al. (2009).
have proven difficult to develop due to active physiological properties of soil organisms affecting either
uptake (such as root phytochelatins) or sequestration of Pb (such as granule formation in root tissues and
earthworms, or substitution of Pb for calcium in bones).
7.2.6. Community and Ecosystem Effects
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. 2006). 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 & 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-
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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.. 2010V However,
above certain Pb concentration, toxic effects on both plants and microbial communities may prevent these
ameliorating effects. R.Y. Yang et al. (2008) 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 et al.. 2006). Inoculation with both earthworms and fungi
increased plant growth at sites contaminated with mine tailings compared to that observed at sites with
75% less Pb contamination. Most likely, this was a result of the decrease in bioavailable (DTPA-
extractable and ammonium acetate-extractable) Pb to 17% to 25% of levels in areas without the
earthworm and arbuscular mycorrhizal fungi amendments. The presence of earthworms in metal-
contaminated soils decreased the amount of water-soluble Pb (Sizmur & 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. 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 (C. B. 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 (R. Y. 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 a better estimate of ecological impacts than microbial diversity or abundance
measurements. When the microbial properties of metal-contaminated urban soils were compared to those
of rural soils, significant differences were detected in basal community respiration rates and microbial
abundance (Y. Yang et al.. 2006). 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. This suggests that anthropogenic Pb contamination may adversely affect soil microbial
communities, and alter their ecological function. Most studies of metal-induced changes in microbial
communities have been conducted using mixtures of metals. Akerblom et al. (2007) tested the effects of
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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. Total phospholipid fatty acid content was determined to negatively correlate
with increasing Pb exposure, indicating alteration of the microbial community. (S. Khan et al.. 2010).
found that following a 2-week exposure to three levels of Pb (150, 300, and 500 mg Pb/kg), both acid
phosphatase and urease levels (measures of soil microbial activity) decreased significantly, although they
had recovered by the ninth week. In addition, the number of culturable bacteria was also decreased, but
only at the highest exposure concentration tested.
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.. 2006V Six soil types of differing pH were treated with
Pb nitrate concentrations of 0.5, 2, 8, and 32 mM. Basal respiration was adversely affected 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 exposure negatively affected the prey capture ability of certain fungal species. The densities of
traps constructed by nematophagous fungi decreased in soils treated with 0.15 mM Pb chloride (Mo et al..
2008). Nematophagous fungi are important predators of soil-dwelling nematodes, collecting their prey
with sticky nets, branches, and rings. This suppression caused a subsequent reduction in fungal nematode
capturing capacity, and could result in increased nematode abundance.
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 from the contributions of copper and zinc that were added with the
Pb.
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) (O. Hu et al.. 2007V 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.
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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. Thus,
it was concluded that even long-term exposure to minimally bioavailable Pb can alter the structure of soil
decomposer communities, which could in turn decrease decomposition rates.
Microbial communities associated with habitats other than soils are also affected by exposure to
atmospherically deposited Pb. Alder (Alnus nepalensis) leaf microorganism populations were greater in
number at non-affected sites than at sites adjacent to a major Indian highway with increased Pb pollution
(Joshi. 2008). The density, species richness, and biomass of testate amoebae communities grown on
Sphagnum fallax mosses were significantly decreased following moss incubation in Pb solutions of either
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 cyanobacteria biomass (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 adversely 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
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function of soil microorganisms. However, in an example of ecological mutualism, plant-associated
arbuscular mycorrhizal fungi protect the host plant from Pb uptake and fungal viability seems to be
protected by the plants. Similarly, soil microbial communities (bacterial species as well as fungi) in Pb-
contaminated soils are improved by re vegetation. 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. However, it is well known from historical
data that Pb can have a widespread and dramatic effect on populations of waterfowl exposed to spent shot
(Beveret al.. 1996) and may be negatively affecting loons in the northeastern U. S. through ingestion of
Pb fishing sinkers (Scheuhammer & Norris. 1996). Studies at shooting ranges and downwind of smelters
previously reported in the 2006 Pb AQCD demonstrate effects of dispersed Pb on terrestrial soil and plant
communities with resulting decreases in secondary consumer vertebrate species.
7.2.7. Critical Loads in Terrestrial Systems
The critical load is the environmental concentration predicted or estimated from available literature
data, above which adverse effects to organisms are likely to occur. It is based on the relative toxicity of
the compound to species or ecosystem processes of concern, and an estimate of residence time in the soil
environment. The concept of critical load is discussed in more detail in Section 7.1.3 of this chapter and
in Section 7.3 of the 2006 Pb AQCD (U.S. EPA. 2006).
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 & Groenenberg.
2009). Similarly, although Pb soil concentration 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).
In some cases, risk assessment for Pb predicts risks to terrestrial animals at environmental
concentrations that fall below natural background levels, and uncertainties associated with standard
toxicity testing can become magnified through the risk assessment process, degrading the reliability of
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estimates of hazardous environmental concentrations. Smolders et al. (2009) concluded from their
comparison of field soils, spiked soils, and artificially aged spiked soils that corrections for aging and
interacting soil properties in spiked soils will make predicted-no-effect concentrations rise above
background. Buekers et al. (2009) proposed the use of a Tissue Residue Approach as a risk estimation
method for terrestrial vertebrates that does not predict risks at background levels, and has smaller
uncertainty. This approach eliminates the need for quantitative estimation of food intake or Pb species
bioavailability. Blood Pb no-adverse-effect concentration (NAEC) and lowest adverse effects
concentration (LAEC) derived from 25 literature 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. For mammals, the hazardous concentration for 5% of species (HC5)
values obtained from the species sensitivity distributions ranged from 11 to 18 jj.g Pb/dL blood; HC5
values for birds ranged from 65 to 71 (ig/dL. The authors proposed the use of 18 and 71 (ig/dL as critical
threshold values for mammals and birds, respectively, which are below the lowest NAEC for both data
sets used, and are above typical background Pb values.
Short of conducting expensive in vivo toxicity studies, it is difficult to determine environmental Pb
toxicity given the variation of physiochemical and soil properties that alter bioavailability and toxicity.
Furman et al. (2006) proposed the use of a physiologically based extraction test to predict risks posed to
waterfowl from environmental Pb contamination. The extraction process was modeled after gastric and
intestinal conditions of waterfowl, and was used to gauge the bioavailability of Pb from freshly amended
and aged contaminated soils. The concentration of Pb extracted through the use of the physiologically
based extraction test was demonstrated to be significantly correlated to Pb tissue concentration in
waterfowl exposed via in vivo studies of the same soils.
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. 2006). Preference is given to studies using the most
bioavailable form of Pb, to derive conservative values. Soil concentration protective of avian and
mammalian diets are calculated by first converting dietary concentration to dose (mg/kg body weight per
day) for the critical study, then using food (and soil) ingestion rates and conservatively derived uptake
factors to calculate soil concentration that would result in unacceptable dietary doses. This frequently
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results in Eco-SSL values below the average background soil concentration, 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 F. Candida collembola (springtails) were between 2,060 and
3,210 mg/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/kg Pb-spiked soil (Xu et al. 2009). 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. Guo. et al.. 2009). and also demonstrated more pronounced neural
degeneration than older larvae and adults (Xing. Rui. et al.. 2009).
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 adverse 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
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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)). 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 et al. (1997). as cited in Dauwe et
al. (2006)).
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 ED10 values derived in aged soils to freshly spiked soils
(factors greater than one 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, beach, 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,
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nutrient poor soil. As agricultural areas are taken out of production and revert to old fields and eventually
forests, their vulnerability to Pb is likely to increase as a result of decreasing OM and acidification of soils
(from discontinuation of fertilizing and liming). On the other hand, increasing density of native or
invasive plants with associated arbuscular mycorrhizal fungi will likely act to ameliorate some of the
effects of Pb (see previous discussion of studies by Sudova and Vostka (2007) and Wong et al. (2007). It
is, however, difficult to categorically state that certain plant or soil invertebrate communities are more
vulnerable to Pb than others, as the available toxicity data have not yet been standardized for differences
in bioavailability (because of use of different Pb salts, different soil properties, and different lengths of
aging of soil prior to testing), nutritional state, or organism age, or other interacting factors. Data from
field studies are complicated by the co-occurrence of other metals and alterations of pH, such as
acidification from S02 in smelter emissions, that 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.
7.2.10. Ecosystem Services
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.
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A few studies since the 2006 Pb AQCD consider the impact of metals in general on ecosystem
services. 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 & 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 (Miaula &
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 copper (kwartirnikov 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 (Ka\ & 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 on
the maintenance of four major functions: carbon transformations, nutrient cycles, soil structure
maintenance, and the regulation of diseases and pests and these parameters may be altered by metal
deposition (Kibblewhite et al.. 2008). Pb impacts to terrestrial systems reviewed in the previous sections
provide evidence for impacts to supporting, provisioning, and regulating ecosystem services provided by
soils. For example, earthworms were shown to impact soil metal mobility and availability, which in turn
resulted in changes to microbial populations (biodiversity), pH, dissolved organic carbon, and metal
speciation (Sizmur & Hodson. 2009). all of which may directly affect soil fertility.
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: Causality determinations for Pb in
Terrestrial and Aquatic Systems for a synthesis of all evidence dating back to the 1977 AQCD considered
to determine causality.
7.2.11.1. Biogeochemistry and Chemical Effects
The amount of Pb dissolved in 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
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reactions of dissolved Pb phases on soil OM; (4) pH; (5) CEC; and (6) aging. Since 2006, further details
have been contributed to the understanding of the role of pH, CEC, OM, and aging. Smolders et al. (2009)
demonstrated that the two most important determinants of 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. Multiple new studies showed that in trees, this
proportion 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. Studies with trees found that soil Pb is generally translocated from the roots.
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 arthropods collected at contaminated sites, recent
studies found gradients in accumulated Pb that corresponded to gradients in soil with increasing distance
from point sources.
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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/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, 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 adverse effects
were found at lower exposure in younger nematodes, and effects on longevity and fecundity were shown
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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 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 scale include several studies of the
ameliorative effects of mycorrhizal fungi on plant growth, attributed to decreased uptake of Pb by plants,
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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 scale 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
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 organisms, or creating extraction methods that more closely mimic
uptake and accumulation.
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. 2006)
on the effects of Pb on aquatic systems. 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
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an aquatic life stage (e.g., amphibians). The current mean and range of Pb concentrations in surface
waters (mean 0.66 (.ig Pb/L, range 0.04 to 30 (.ig Pb/L), sediments (mean 120 (.ig Pb/g dry weight, range
0.5 to 12,000 (.ig Pb/g dry weight) and fish tissues (mean 1.03 |_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 was reported in the previous 2006 Pb AQCD (U.S. EPA. 2006). 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 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 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 currently expressed as a criteria maximum
concentration (CMC) for acute toxicity and criterion continuous concentration (CCC) for chronic toxicity
(U.S. EPA. 2010). In freshwater, the CMC is 65 (.ig Pb/L and the CCC is 2.5 |_ig Pb/L at a hardness of 100
mg/L. In saltwater, these values are 210 (.ig Pb/L CMC and 8.1 (ig Pb/L CCC, respectively.
A draft document intended to update the AWQC for Pb (Great Lakes Environmental Center. 2008)
was recently prepared for the EPA. This draft included significantly revised values for both acute and
chronic endpoints that were also based on a hardness equation; the revised chronic AWQC in particular is
much higher than in previous AWQC due to a substantial reduction in the acute to chronic ratio (ACR).
Recent studies have suggested that the ACRs used in the existing criteria documents were too high,
possibly because of the age and size of fish used in those studies (Mebane et al.. 2008).
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. 2005). 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
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with the metals extracted via acid-volatile sulfides (AVS) since metals that bind to AVS (such as Pb)
should not be toxic in sediments where AVS occurs in greater qualities than SEM.
Since the publication of the 2006 Pb AQCD both of these methods, for estimating sediment criteria
for metals, have continued to be used and refined. The SEM approach was further refined in the
development of the sediment BLM (Pi Toro et al.. 2005V The BLM is discussed further in Section 7.3.3.
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 am phi pod exposures (J. A. Besser et al.. 2009; Mac Donald et al.. 2009). However,
when the relationship between invertebrate habitat (epibenthic and benthic) and environmental Pb
bioaccumulation was investigated, De Jonge et al. (2010) determined that different environmental
fractions of Pb were responsible for invertebrate uptake and exposure. Pb uptake by benthic invertebrate
taxa was not significantly correlated to AVS Pb levels, but rather to total sediment concentrations
(De Jonge et al.. 2009). Conversely, epibenthic invertebrate Pb body burdens were better correlated to
AVS concentrations, rather than total Pb sediment concentrations (De Jonge et al.. 2010).
In the following sections, new information since the 2006 Pb AQCD on Pb in aquatic ecosystems
will be presented. Throughout the sections, brief summaries of conclusions from the 2006 Pb AQCD are
included where appropriate. Section 7.3 is organized to consider uptake of Pb and effects at the species
level, followed by community and ecosystem level effects. Section 7.3.2 considers the biogeochemistry
and chemical effects of Pb in aquatic systems. New research on the bioavailability and uptake of Pb into
aquatic organisms including plants, invertebrates and vertebrates is presented next (Section 7.3.3). Effects
of Pb on the physiology of aquatic fauna and biota (Section 7.3.4) are followed with data on exposure and
response of aquatic organisms (Section 7.3.5). Ecosystem-scale responses are reviewed in Section 7.3.6
followed by a brief consideration of critical loads in aquatic systems (Section 7.3.7), characterization of
sensitivity and vulnerability of ecosystem components (Section 7.3.8) and the effects of Pb on ecosystem
services associated with aquatic environments (Section 7.3.9).
7.3.2. Biogeochemistry and Chemical Effects
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 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 iron and manganese oxides. For instance, in neutral to
acidic aquatic environments, Pb is typically present as PbS04, PbCl4, Pb2+, cationic forms of Pb
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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 CI : 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 calcium and magnesium 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 While the BLM is not yet used in setting regulatory criteria for Pb in the U.S., its application in risk
analysis has become widely recognized (Fairbrother et al.. 2007). Sediment quality criteria have yet to be
adopted by EPA, although an equilibrium partitioning procedure is now available to predict which
sediments have metal concentrations that are not toxic to aquatic organisms ("U.S. EPA. 2005). 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
(2006). It is important to note that this method cannot accurately predict which sediments are toxic or
which metal is the primary risk driver.
A more detailed understanding of the biogeochemistry of Pb in aquatic systems (both the water
column and sediments) is critical to accurately predicting toxic effects of Pb to aquatic organisms. It
should be recognized, however, that in addition to exposure via sediment and water, chronic exposures to
Pb also include dietary uptake, even though the toxicokinetics of this exposure pathway are not yet well
understood in aquatic organisms and the influence of the bioavailability factors described above is
unknown. Furthermore, changes in environmental factors that reduce the bioaccessible Pb fraction can
result in either sequestration in sediments or subsequent release as mobile, bioaccessible forms. This
section provides updated information about the influence of chemical parameters that affect Pb
bioaccessibility in the aquatic environment (in sediments and the water column).
Several models are available for estimating the speciation of dissolved Pb. These models were
tested by Balistrieri and Blank (2008) by comparing the speciation of dissolved Pb in aquatic systems
affected by historical mining activities with that predicted by several models, including Windermere
humic aqueous model (WHAM VI), non-ideal competitive absorption Donnan-type model (NICA-
Donnan), and Stockholm humic model (SHM). Accurate prediction of labile Pb concentrations was
achieved only with SHM, although other metal concentrations were better described by the WHAM
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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 & 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 & 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/Lto 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 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
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
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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).
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. They concluded
that the fraction of Pb associated with Fe-oxides was most frequently transferred to biofilms and the other
biological components of the sampled systems (Farag et al.. 2007). Consequently, an increase in the Pb
Fe-oxide fraction could signify a potential increase in the bioaccessible pool of Pb. The authors also noted
that this fraction may promote downstream transport of Pb contamination. Ancion et al. (2010)
investigated whether urban runoff metal contaminants could modify biofilm bacterial community
structure and diversity and therefore potentially alter the function of biofilms in stream ecosystems. They
found that accumulation rates for metals in biofilm were maximal during the first day of exposure and
then decreased with time. Equilibrium between metal concentrations in the water and in the biofilm was
reached for all metals after 7-14 days of exposure. The affinity of the biofilm for Pb was, however, much
greater than for Cu and Zn. With respect to recovery, the release of metals was slow and after 14 days in
clean water 35% of Pb remained in the biofilm. By retaining 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
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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, (Pimephalespromelas) (Manor et al.. 2010). and by Pb toxicity to
fathead minnow and the cladoceran Ceriodaphnia dubia (Manor et al.. 2011).
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..
2010). 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 (Lamolas & Slavovkova. 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 (Lamolas 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, C. Kesslerii, 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 marine invertebrate Paracentrotus
lividus embryo-larval bioassay (Sanchez-Marin. Santos-Echoandia. et al.. 2010). 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. Slavovkova. et al.. 2010).
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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. Slavevkova. et al.. 2010). Formation of a
ternary complex that is better absorbed by biological membranes was another proposed mechanism that
could describe the increased bioaccessibility to marine invertebrates of Pb bound to humic acid (Sanchez-
Marin et al.. 2007).
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 Chlorella 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 & 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 bioaccessibility models was conducted at sites with distinct point-sources of Pb
contamination, and only one model (SHM) was found to adequately predict Pb bioaccessibility.
7.3.3. Bioavailability in Aquatic 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,
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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.. 2007V 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. 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, zinc
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.
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
adverse effects (Luoma & 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 (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, recent information on bioavailability and uptake in algae, plants, invertebrates and
vertebrates from marine and freshwater systems are reviewed with summary material from the 2006 Pb
AQCD where appropriate. An overview of the BLM is presented as the most widely used method for
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predicting both the bioaccessible and bioavailable fractions of Pb in the aquatic environment. This is
followed by a discussion of 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.
In addition to the biogeochemical effects that govern the environmental pool of accessible Pb,
reactions of Pb with biological surfaces and membranes determines the bioavailability and uptake of the
metal by aquatic organisms. The BLM predicts both the bioaccessible and bioavailable fraction of Pb in
the aquatic environment, and can be used to estimate the importance of environmental variables such as
DOC in limiting uptake by aquatic organisms (Alonso-Castro et al. 2009). The BLM integrates the
binding affinities of various natural ligands in surface waters and the biological uptake rates of aquatic
organisms to determine the site-specific toxicity of the bioavailable fraction. In the 2006 Pb AQCD,
limitations of the use of BLM in developing air quality criteria were recognized including the focus of
this model on acute endpoints and the absence of consideration of dietary uptake as a route of exposure.
Atmospheric deposition of Pb to aquatic systems and subsequent effects on ecosystem receptors is likely
characterized as a chronic, cumulative exposure rather than an acute exposure. Recommendations from
the 2006 Pb AQCD included developing both chronic toxicity BLMs and BLMs that consider the dietary
route of Pb uptake. 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
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determined that the model significantly underestimated Pb toxicity to marine invertebrates (Pi Toro et al..
2005). This may be because pore water metal concentrations were not modeled. Consequently, these
results may suggest that either 1) mobilization of Pb concentrations from sediments into pore water is
greater in marine environments, or 2) marine invertebrates are significantly more susceptible to Pb
exposures than are freshwater species.
A number of deviations from results predicted by Pb exposure models (such as the BLM) were
documented by Ahlf et al. (2009). They highlighted that uptake of metals by sediment-dwelling bivalves
was significantly greater than predicted, because bivalves accumulate Pb from multiple sources not
included in the model, such as ingestion of algae, bacteria, and colloidal matter. Species-specific dietary
assimilation of ingested particulate-bound metals is also likely to play a role in the toxicity of Pb to
aquatic organisms, yet insufficient data are available to permit modeling of this additional factor (Ahlf et
al.. 2009). The authors outlined the need for additional data in developing bioavailability models for
chronic metal exposures.
Similarly, although the presence of humic acid is considered to reduce the bioavailable fraction of
metals in surface water, green algae uptake and biosorption of metals, including Pb, was actually
increased by humic acid. The authors determined that humic acid bound to algal surfaces served to
increase the total number of metal binding sites over those afforded solely by the algal surface (Lamelas
& 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 & Wilkinson. 2005). Despite this, the authors noted
that, in most cases, the BLM could predict acute metals toxicity with a reasonable degree of accuracy.
Veltman et al. (2010) proposed an integration of BLM and bioaccumulation models in order to
more accurately predict metal uptake by fish and invertebrates. Although Pb was not the specific focus of
the paper, calculated metal absorption efficiencies for marine fish species from both BLM and
bioaccumulation models were determined to be highly comparable for Ag, Cd, Cu, and Zn. Authors also
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.
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7.3.3.1. Plants and Algae
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. Uptake and translocation studies of Pb in plants and algae reviewed in the
2006 Pb AQCD indicated that all 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 2006 Pb AQCD 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 the water column,
with 2.7, 6.9, 19, 45 and 145 (.ig Pb/mg accumulated at aqueous Pb concentrations of 5, 10, 30, 50, and
100 (.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 (known as "biodilution"). An aquatic moss, Fontinalis antipyretica,
accumulated up to an average of 3 (.iniol Pb/g dry weight over a 7-day exposure to 100 |_imol Pb, despite
saturation of intracellular Pb concentrations after 5 days of exposure (Rau 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 (.Z. Z. Chen et al.. 2010). 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.
Pb bioaccumulation studies conducted with five species of marine algae, (Tetraselmis chuii,
Rhodomonas salina, Chaetoceros sp., Isochrysis galbana and Nannochloropsis gaditana) 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 (Dcbelius et al.. 2009).
When exposed to water concentrations of up to 100 (.iniol Pb, 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 (S. Mishra
et al.. 2006). 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 (.iniol
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Pb) 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).
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 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 BCF's 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 wetlands 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
Alternanthera philoxeroides accumulated 1.4 and 10 mg Pb/kg dry weight. Fritioff and Greger (2006)
determined that anywhere from 24-59% of the total Pb taken up by Potamogeton natans aquatic plants
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 & 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
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tightly bound to sediments, and therefore the relationship between total sediment Pb and Pb in aquatic
plant tissues was weaker (Ebrahimpour & Mushrifah. 2009).
Lemna sp., a rooted 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 &
Sternberg. 2008V
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-Castro 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 & Lo Giudice. 2010).
Environmental Pb concentrations in water and sediment averaged 0.4 |_ig Pb/L and 2.7 mg Pb/kg. These
levels yielded root and rhizome concentrations of 17 and 15 mg Pb/kg, respectively, whereas stem and
leaf Pb concentrations were lower (9.9 and 13 mg Pb/kg). These tissue concentrations were significantly
correlated to both water and sediment concentrations (Bonanno & Lo Giudice. 2010). The roots of two
salt marsh species, Sacorconia fructicosa 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 sediments in vegetated zones indicating that Pb uptake by plants
reflects the input in sediments. Conversely, the semi-aquatic plant Ammonia 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 & 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-
Cvmerman & Kempers. 2007). A similar significant correlation was established between reed sweet grass
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root Pb concentration and sediment Pb concentrations, yielding BCFs ranging from 0.5 to 1.5, with an
average BCF of 0.9 (Skorbiowicz. 2006).
Pb tissue concentrations of aquatic plants P. australis and Ludwigia prostrata collected from
wetlands containing an average of 52 mg Pb/kg in surficial sediments were predominantly in root tissues,
indicating poor translocation of Pb from roots. In the former, Pb decreased from an average of 37 mg
Pb/kg in roots to 17, 14, and 12 mg Pb/kg in rhizome, stem and leaf tissues, respectively, while L.
prostrata Pb tissue concentrations decreased from 77 mg Pb/kg in fibrous root to 7 and 43 mg Pb/kg in
stem and leaf tissues (H. J. Yang et al. 2008). The authors proposed that this diminished transfer ability
explained the relatively low BCF's for Pb uptake in these two species, when compared with those of other
metals (Table 7-3).
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/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.
BCFs for Pb in root tissue from mangrove tree species range between 0.09 and 2.9, depending on
the species and the habitat, with an average BCF of 0.84. The average BCF for mangrove species leaf
tissue was considerably less (0.11), as these species are poor translocators of Pb (MacFarlane et al.. 2007).
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In contrast, 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.. 2009V The authors
proposed that the plant's active mobilization of nutrients from soil during growth also resulted in
increased Pb uptake and sequestration.
Given that sediments are a significant sink for Pb entering aquatic systems, it is not surprising that
rooted macrophytes bioaccumulate significant quantities of the metal. Although there are some
similarities to Pb accumulation observed in terrestrial plants (e.g., preferential sequestration of the metal
in root tissue), Pb appears to be more bioavailable in sediment than it is in soil. This may be a result of
differences in plant physiology between aquatic and terrestrial plants (e.g., more rapid growth or more
efficient assimilation of nutrients and ions from a water-saturated medium). While rooted macrophytes are
likely to be chronic accumulators of Pb sequestered in sediments, aerial deposition of Pb into aquatic
systems may result in pulsed inputs of labile Pb that would be available for uptake by floating
macrophytes and algae.
Reported values for BCF's in aquatic plants from the 2006 Pb AQCD range from 840 to 20,000 (.
Table AX7-2.3.1 U.S. EPA. 2006). Duckweed (Lemna minor) had BCF values ranging from 840 to 3,560
depending on the method of measurement. Additional BCF's established for aquatic plants since the 2006
Pb AQCD are summarized in Table 7-3 and include data on field-collected plants as well as BCF's
obtained from laboratory exposures.
Table 7-3. Bioconcentration factors for Pb in aquatic plants
Species
BCF
Test conditions
Reference
Typha latifolia
649
10 days, Pb nitrate-spiked water
Alonso-Castro et al. (2009)
Spirulina platensis
1500
10 days, Pb nitrate-spiked water
Arunakumara et al. (2008)
Ceratophyllum demersum
0.2
Field-collected plants
Bi et al. (2007)
Spirogyra communis
0.2
Field-collected plants
Bi et al. (2007)
Phragmites australis
6.4
Field-collected plants
Bonanno and Lo Giudice (2010)
Taxodium distichum
0.02
Field-collected tissue
Devall etal. (2006)
Hydrilla verticillata
26
Field-collected plants
Dwivedi et al. (2008)
Eleocharis acicularis
0.8
11 mo, field-collected sediment
Ha etal. (2009)
Lemna sp.
0.01
7 days, ash water extract
Horvatetal. (2007)
Mangrove species
0.8
Field-collected tissue
MacFarlane et al. (2007)
Glyceria aquatica
0.9
Field-collected tissue (roots)
Skorbiowicz (2006)
Phragmites australis
0.7
Field-collected plants
Yang et al. (2008)
Ludwigia prostata
1.5
Field-collected plants
Yang et al. (2008)
7.3.3.2. Invertebrates
Uptake and subsequent bioaccumulation of Pb in marine and 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 2006 Pb AQCD on bioaccumulation and sequestration of Pb in aquatic
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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, life
stage, age, and metabolism can alter the accumulation rate. In this section, new information on Pb uptake
from 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.
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 metal accumulation than dissolved metal concentrations in sediment pore
water based on bioaccumulation studies using contaminated sediments from six different sites (Roulier et
al.. 2008). 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. Additionally, Metian et al.
(2009) showed that king scallop (Pecten maximus) exhibited low bioaccumulation efficiency of Pb from
spiked sediment.
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
(J. M. 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. (komiarova & Blust. 2009a). I), 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 (.iniol Pb/kg dry
weight, and was similar to the amount absorbed by algae (Komiarova & Blust. 2009a).
Stable isotope analysis was to used measure uptake and elimination simultaneously in netspinning
caddisfly larvae (Hydropsyche sp.) exposed to aqueous Pb concentrations of 0.2 to 0.6 |_ig Pb/L (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 d"1 for Pb-exposed larvae was similar in both presence and absence of
the metal in the water. Caddisflies accumulated significant amounts of the metal over 18 days of
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exposure. Measured tissue concentrations ranged from approximately 15 to 35 (.ig Pb/g. Hydropsy chid Pb
BCF's 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). In larvae of the mosquito, Culex quinquefasciatus, exposed to 100 (.ig Pb/L for seven days the BCF
was 62 (kitvatanachai et al.. 2005).
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). Metian et al. (2009) investigated the uptake and bioaccumulation of 210Pb in Chlamys varia
(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 C. varia than P. maximus and soft tissue distribution patterns differed
between the species. C. varia accumulated Pb preferentially in the digestive gland (50%) while in P.
maximus, 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 with P. maximus showed low bioaccumulation efficiency of Pb from sediment.
With the exception of the above-mentioned study with scallops (Metian et al.. 2009) recent reports
on Pb distribution generally supports the findings of the 2006 Pb AQCD that Pb is primarily sequestered
in the gills, hepatopancreas, and muscle. Uptake of Pb by the crayfish (Cherax destructor) exposed to
5,000 |ag 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). The gills were the main sites
of Pb accumulation in Pinctada fucata (pearl oyster) followed by mantle, in 72-hour exposures to 103.5
|ag 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 of Mytilus edulis: after a 10 day depuration, Pb content was
decreased in the gills and digestive gland of these mussels (Einspom et al.. 2009). In blue crabs,
Callinectes sapidus, collected from a contaminated and a clean estuary in New Jersey, U.S., the
hepatopancreas was found to be the primary organ for Pb uptake (Reichmuth et al.. 2010). Body burden
analysis following 96 hour exposure to 50, 100 and 500 |ag 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).
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 & 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
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hemocytes (Einspom et al.. 2009). In the digestive gland of the variegated scallop, Pb was also mainly
bound to organelles, i.e., 66% of the total metal burden (Bustamante & 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).
Since the publication of the 2006 Pb AQCD, additional factors have been considered that may
affect Pb uptake in aquatic organisms. Pb tissue concentrations fluctuated seasonally in mussels (Mytilus
galloprovincialis) harvested near Istanbul, Turkey (Ozden. 2008). Tissue Pb concentrations were lowest
during the summer months (average of 0.9 mg Pb/kg), followed by spring, autumn and winter (1.3, 1.4,
and 1.6 mg Pb/kg, respectively). The authors speculated that the slight seasonal differences indicate that
bioavailability of the metal may be related to seasonal changes in surface water or sediment chemistry.
Additionally, alterations in growth over the year, as well as different rates of Ca uptake may have
impacted Pb bioaccumulation rates. 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).
Reported BAF values for Pb in aquatic invertebrates from the 2006 Pb AQCD ranged from 499 to
3,670 [See Table AX7-2.3.2 (U.S. EPA. 2006)1. Since the publication of 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 & Fisher. 1997) (Table 7-4). A
complicating factor in establishing BAF values is that laboratory studies usually assess uptake in water-
only or sediment only exposures while field studies take into account dietary sources of Pb as well as
waterborne Pb resulting in BAF values that are frequently 100-1,000 times larger than BCF values for the
same metal and species (DeForest et al.. 2007). Mean Pb levels in both predatory and grazing
zooplanktonic species in El Niagra reservoir, (in Aguascalientes, Mexico) were used to calculate BAF
values (Rubio-Franchini et al.. 2008) to assess biomagnification of Pb. The BAF of the predatory rotifer
Asplanchna brigthwellii (BAF 49,300) was up to four times higher than the grazing cladocerans Daphnia
similis (BAF 9,022) and Moina micrura (BAF 8,046). Limpet (Patella sp.) from the Lebanese Coast had
Pb BAF values ranging from 2,500 to 6,000 and in the same field study mussel (Brachidontes variabilis)
Pb BAF values ranged from 7,500-8,000 (Nakhle et al.. 2006).
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Table 7-4. Bioaccumulation factors for Pb in aquatic invertebrates
Species
BAF
Test conditions
Reference
Hydropsyche sp. (caddisfly
larvaej
54
18 days, Pb laboratory exposure
Evans etal. (2006)
Culex Quinquefasciatus
(mosquito)
62
7 days, Pb laboratory exposure
Kitvatanachai et al. (2005)
Daphnia similis (zooplankton)
9,022
Field-collected
Rubio-Franchini et al. (2008)
Moina micrura (zooplankton)
8,046
Field-collected
Rubio-Franchini et al. (2008)
Asplanchna brigthwellii (rotifer)
49,344
Field-collected
Rubio-Franchini et al. (2008)
Patella sp. (limpet)
6,000
Field-collected
Nakhle etal.(2006)
Brachidontes variabilis (mussel)
8,000
Field-collected
Nakhle etal.(2006)
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). A mean log of 4.3 of the metal concentration in mussels (ng Pb/kg wet flesh
weight)/metal concentration in water (ng Pb/L) was determined for Pb for M. galloprovincialis in this
study based on the rate constants of uptake and efflux in a series of transplantation experiments between
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.
7.3.3.3. 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 water 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. New data on uptake studies in marine fishes are
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presented followed by new evidence for additional Pb detoxification mechanisms in fish. Finally, new
data on uptake and tissue distribution of Pb in amphibians and aquatic mammals are presented.
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).
Since the publication of the 2006 Pb AQCD, multiple studies on uptake of Pb from water by
fathead minnow 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.. 2006a). In the
larval minnows, Pb concentration in the intestine exhibited the highest initial accumulation of all tissues
on day 3 but then decreased for the remainder of the experiment while concentrations in the other organs
continued to increase. By day 30, gill tissue exhibited the highest Pb concentration (approximately 120 |_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. (2006a) with highest concentration in gill, followed by kidney,
anterior intestine, and carcass. Addition of humic acid and carbonate both independently reduced uptake
of Pb in these fish over the exposure time period. Interestingly, fathead minnow eggs collected daily
during 21 day breeding assays that followed the chronic exposure described above accumulated similar
levels of Pb from the test solutions regardless of Pb concentration or water chemistry (e.g., addition of
humic acid and carbonate) (Manor 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 (Manor et al.. 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
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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 dry weight at 20 (.iM) during a 14-day
exposure (Atli & 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). 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 (.ig Pb/L for 24 days accumulated 6 (.ig Pb/g
dry weight in gill tissue and Pb concentrations in liver tissue reached 14 (.ig 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 (.iniol 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/kgh to 35 L/kgh by increasing pH from 6 to 7, and from 20 L/kgh to 35 L/kgh
by increasing Ca concentration from 0.1 mmol to 0.5 mmol (komiarova & 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 (Carrasius 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 (khozhina & 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 of 0.03 mg 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 jag 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).
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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 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.. 2006a) 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..
2006a). By day 30, gill tissue exhibited the highest Pb concentration followed by whole fish and carcass
(whole fish minus gill, liver, muscle and intestine). The most rapid rate of Pb accumulation in this species
occurs within the first 10 days of exposure (Mager et al.. 2008). African catfish (Clarias gariepinus)
exposed to aqueous Pb concentrations of 50 to 1,000 (.ig Pb/L (as Pb nitrate) for 4 weeks accumulated
significant amounts of Pb in heart (520-600 mg Pb/kg), liver (150-242 mg Pb/kg), and brain (120-230 mg
Pb/kg) tissues (Kudirat. 2008). Doubling the exposure time to 8 weeks increased sequestration of Pb in
these tissues as well as in skin (125-137.5 mg Pb/kg) and ovaries (30-60 mg Pb/kg).
Since the publication of the 2006 Pb AQCD, several studies have focused on dietary uptake of Pb
in teleosts. Alves et al. (2006) administered a diet of three concentrations of Pb (7, 77 and 520 (.ig Pb/g
dry weight) to rainbow trout for 21 days. Doses were calculated to be 0.02 (.ig Pb/day (control), 3.7
(.ig Pb/day (low concentration), 39.6 (.ig Pb/day (intermediate concentration) and 221.5 (.ig Pb/day (high
concentration). Concentrations in the study were selected to represent environmentally relevant
concentrations in prey. After 21 days exposure to the highest concentration, Pb accumulation was greatest
in the intestine, followed by carcass, kidney and liver leading the authors to hypothesize that the intestine
is the primary site of exposure in dietary uptake of Pb. All tissues, (gill, liver, kidney, intestine, carcass)
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 (.ig 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 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).
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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. Du. et al.. 2009).
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 & Wood. 2006). Pb accumulation in the gut was followed by bone, kidney, liver, spleen,
gill, carcass, brain and white muscle (Alves & 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 (Qio & 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 ficld-collcctcd 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 internally from dietary sources, toxicity does not correlate with dietary exposure, but does correlate
with gill accumulation from waterborne exposure (Alves et al.. 2006). Comparison of uptake rates across
the gut and gill have shown that transporter pathways in the gill have a much higher affinity for Pb than
do similar pathways in the gut (Qio & Wood. 2007).
Since the 2006 Pb AQCD, several field studies have considered Pb uptake and bioaccumulation in
fish as a tool for environmental assessment. Pb tissue concentrations were elevated in several species 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). The Western Airborne
Contaminants Assessment Project assessed concentrations of semi-volatile organic compounds and metals
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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. Results from this
study are considered in in Chapter 3 of this ISA.
Marine 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 gills and
intestine (W. X. Wang & Rainbow. 2008). Pb was significantly accumulated in gill, liver, plasma, kidney,
rectal gland, intestine, skin, muscle of a marine shark species, spotted dogfish (Scyliorhinus canicula)
exposed to 2,072 |_ig Pb/L for one week (Dc Boeck et al.. 2010). 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).
The 2006 Pb AQCD considered detoxification mechanisms in fish including mucus production and
Pb removal by scales through chelation 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 & 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.
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 |ag 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 diffuse, non-quantified sources of Pb (e.g., in sediments or in
dietary sources) may be contributing to Pb uptake by marine fish.
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
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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.
Reported BCF's in fish species from the 2006 Pb AQCD were 42 for brook trout (Salvelinus
fontinalis) and 45 for bluegill (Lepomis macrochirus). Since the 2006 Pb AQCD, additional BAF's have
been established in water-only exposures, from dietary exposures and from field-collected fish (Table 7-
5).
Table 7-5. Bioaccumulation factors for Pb in fish
Species
BAF
Test conditions
Reference
Clarias gariepinus
800
56 days, Pb nitrate
Kudirat (2008)
Ctenopharyngodon idella
44
Field-collected
Malik et al. (2010)
Pimephales promelas
100 to 100,000
30 days, Pb nitrate
Grosell et al. (2006a)
Carassius auratus
0.04
Field-collected
Khozhina and Sherriff (2008)
Danio rerio
1.4
63 days, dietary exposure
Bovle et al. (2010)
Amphibians
Since the 2006 Pb AQCD, there are a few new studies that consider uptake of Pb in amphibians. In
a chronic study with tadpoles of the Northern Leopard frog (Rana pipiens), Pb tissue concentrations were
evaluated following exposures to 3, 10, and 100 |_ig Pb/L from embryo to metamorphosis. The tadpole
tissue concentrations ranged from 0.1 to 224.5 mg Pb/kg dry mass and were positively correlated to Pb
concentrations in the water (T. H. Chen et al.. 2006). 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/g wet weight) for 12 days (Mouchct et al.. 2007). Pb
concentrations were measured in livers, bodies without liver and whole bodies in Southern leopard frog
{Rana spenocephala) 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.
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Mammals
Studies that consider uptake of Pb in aquatic mammals are limited. Kannan et al. (2006) in a
comparison of trace element concentrations in livers of free-ranging sea otters (Enhydra lutris nereis)
found dead along the California coast, detected Pb in all individuals sampled (N=80) in a range of 0.019
to 1.06 |ag Pb/g. The otters were classified by cause of death (infectious causes, non-infectious causes,
those that died in an 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.
7.3.3.4. Food Web
At the time of the publication of the 2006 Pb AQCD, trophic transfer of Pb through aquatic food
chains was considered to be negligible (U.S. EPA. 2006). 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 bioconcentrated but not biomagnified (Eisler. 2000; U.S.
EPA. 2006). New literature since the 2006 Pb AQCD provides evidence of the potential for Pb to be
transferred in aquatic food webs while other studies indicate Pb is decreased with increasing trophic level.
This section incorporates recent literature on transfer of Pb through 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 Niagra reservoir, Aguascalientes,
Mexico, an ecosystem that lacks fishes (Rubio-Franchini et al.. 2008). Pb was measured in sediment,
water, and zooplankton samples of this freshwater system. BCF's were calculated for predatory and
grazing zooplanktonic species. The BCF of the rotifer A brigthwellii (BCF 49,300) was up to four times
higher than the grazing cladocerans D. similis (BCF 9,022) and M. micrura (BCF 8,046). According to
the authors, since M. micrura are prey for A brigthwellii this may explain the biomagnifications of Pb
observed in the predatory rotifer and provides evidence that Pb biomagnifies at intermediate trophic
levels. Partial evidence for biomagnification was observed in a subtropical lagoon in Mexico with
increases of Pb concentration occurring in 14 of the 31 (45.2%) of trophic interactions considered
(Ruelas-lnzunza & Paez-Osuna. 2008). The highest rate of 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).
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
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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. 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 of N. diversicolor and there was a significant inverse linear
relationship between the trophic transfer coefficient and prey Pb concentration. In the same study, another
predator, the decapod Palaemonetes varians, did not significantly accumulate Pb from N. diversicolor
indicating that trophic transfer is dependent on species-specific differences in metal assimilation
efficiencies and accumulation patterns.
In a recent dietary metal study, field-collected invertebrates representing ecologically relevant
sources of Pb were fed to zebrafish, to assess bioavailability of this metal via food. The polychaete worm
N. diversicolor was collected from two sites; an estuary contaminated with Pb and a reference site with
low metal concentrations (Boyle et al.. 2010). Male zebrafish fed Pb-enriched N. diversicolor had
significant increases in whole-body Pb burden when compared to zebrafish fed prey from the reference
site, brine shrimp or flake food diets. There was a trend toward increased Pb levels in females under the
same dietary regimen. In this study, deposit feeding invertebrates were shown to mobilize sediment-
bound metals in the food chain since zebrafish were exposed only to biologically incorporated metal.
The concentration of Pb in the tissues of various aquatic organisms was measured during the
biomonitoring of mining-impacted stream systems in Missouri, U.S. Generally, Pb concentrations
decreased with increasing trophic level: detritus contained 20 to 60 |ag 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 (J. M. 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 (J. M. Besser et al. 2007).
Other studies have traced Pb in aquatic food webs and have found no evidence of biomagnification
of Pb with increasing trophic level. Pb exposure at the base of the food web did not biomagnify in a
simplified-four level marine food chain from Tetraselmis suecica (phytoplankton) to Artemia franciscana
(crustacean, brine shrimp) then L. vannamei (crustacean white shrimp) and finally to Haemulon scudderi
(fish, grunt) (Soto-Jimenez et al.). In the southeastern Gulf of California, Mexico, Pb was not positively
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transferred (biomagnification factor <1) through primary producers (seston, detritus) and 14 consumer
species in a lagoon food web (Jara-Marini et al.. 2009). No biomagnification of Pb was detected from
mesozooplankton to macrozooplankton in Bahia Blanca estuary, Argentina (Fernandez Severini et al.). In
a Brazilian coastal lagoon food chain, Pb was significantly higher in invertebrates than in fishes (Pereira
et al.. 2010). Watanabe et al. (2008) also observed decreasing Pb concentrations through a stream
macroinvertebrate food web in Japan from producers to primary and secondary consumers.
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 melanastomus) and the introduced zebra mussel
(Dreissena polymorpha) have created a new benthic pathway for transfer of Pb and other metals
(Southward Hogan et al.. 2007). The goby is a predator of the benthic zebra mussel, while the endemic
smallmouth bass (Micropterus dolomieui) feed on goby. Since the introduction of goby into the lake, total
Pb concentrations have decreased in bass. The authors attribute this decrease of Pb in bass to changes in
food web structure, changes in prey contaminant burden or declines in sediment Pb concentrations.
7.3.4. Biological Effects
This section focuses on the studies of biological effects of Pb on aquatic biota including algae,
aquatic plants, invertebrates, fish and other biota with an aquatic lifestage (e.g., amphibians) published
since the 2006 Pb AQCD. 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, adverse effects to blood chemistry, and decreases in enzyme levels. Functional growth
responses resulting from Pb exposure include changes in growth patterns, gill binding affinities, and
absorption rates. 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. Since this document focuses on atmospheric
sources of Pb to ecosystem receptors, areas of research not addressed here include literature related to
exposure to Pb from shot or pellets. Biological effects of Pb on algae and plants are considered below,
followed by information on effects on aquatic invertebrates and vertebrates.
7.3.4.1. Plants and Algae
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.
Pb exposure in microalgae species has been linked to several adverse effects, including disruption of
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thylakoid structure and inhibition of growth in both Scenedesmus quadricauda and Anabeana flos-aquae
fArunakiimara & Zhang. 2008). Arunakumara et al. (2008) determined the effect of aqueous Pb on the
algal species S. platensis using solutions of Pb-nitrate. While low Pb exposures (5 (.ig Pb/mL) stimulated
10-day algal growth, growth was inhibited at higher concentrations of 10, 30, 50, and 100 (.ig Pb/mL 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 |_ig Pb/mL. The effect of Pb exposure on the structure and function of
plant photosystem II was studied in giant duckweed, S. polyrrhiza (Ling & 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.
An increase in levels of antioxidant enzymes is commonly observed 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 1000
(imol. When exposure duration was increased to 7 days, only SOD activity remained significantly
increased by Pb exposure (Dazv 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, 10, and 100 mg Pb/L (Dogan et al.. 2009). This, along with the induction
of some antioxidant systems and the reduction of growth at the highest two exposures, indicated that
exposure to the metal caused significant stress, and that toxicity increased with exposure. In addition,
native Myriophyllum quitense exhibited elevated antioxidant enzyme activity (glutathione-S-transferase,
glutathione reductase, 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).
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, activities of these
antioxidant enzymes decreased (S. Mishra et al.. 2006).
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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 to 300 mg Pb/L); total
phenols also increased with increasing Pb exposure. Authors noted that this species exhibited similar
responses under extreme temperatures, drought, and disease (Sobrino et al.. 2010). 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 & Fasidi. 2006V 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 algae Scenedesmus vacuolatus were significantly increased over that of non-exposed
algal cultures {F, 2006, 358857}. 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.
Pb exposure (as Pb-nitrate) caused oxidative damage, growth inhibition, and decreased biochemical
parameters, including photosynthetic pigments, proteins, and monosaccharides, in Wolffia arrhiza plants.
Fresh weight of plants was reduced following both 7- and 14-day exposures to Pb concentrations greater
than 10 mmol, while chlorophyll a content was decreased at concentrations greater than 1 mmol Pb
(Piotrowska et al.. 2010).
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 mg 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 and 10 mg Pb/L, while exposure to 15 mg Pb/L was associated with notable plant mortality (Hurd &
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,
although the authors determined that this resulted from poor uptake of Pb by L. minor (Paczkow ska et al..
2007)
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7.3.4.2. Invertebrates
Effects of Pb on aquatic invertebrates recognized in the 2006 Pb AQCD include adverse impacts on
reproduction, growth, survival and metabolism. 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. As observed in terrestrial invertebrates, the antioxidant system,
survival, growth and reproduction are affected by Pb in aquatic organisms. In aquatic invertebrates, Pb
has also been shown to affect stress responses and osmoregulation. New evidence that supports previous
findings of Pb on reproduction and growth in invertebrates are reviewed here as well as limited studies on
behavioral effects.
Recent literature strengthens the evidence indicating that Pb affects enzymes and antioxidant
activity in aquatic invertebrates. Increased SOD activity was observed in mantles of pearl oyster but
decreased with time although always remaining higher than in the control animals during 72-hour
exposures to 0.5 (.iM Pb (Jing et al.. 2007). In contrast, activity of Se-dependent glutathione peroxidase
was not changed 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 (Y. Zhang et al.. 2010). In contrast, Einsporn et al. (2009)
observed no change in catalase activity in the digestive gland and gill of blue mussel M. edulis following
exposures to 2,500 (.ig Pb/L as Pb nitrate for 10 days and measured 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
(Vlahogianni & Valavanidis. 2007). Catalase activity was decreased in the mantle of these mussels but
fluctuated in their gills, as compared with the control group. In the bivalve C. farreri exposed to Pb, there
was induction of lipid peroxidation measured as MDA of 24% and a 37% reduction in 7-ethoxyresorufin-
o-deethylase (EROD) activity when compared to controls (Y. Zhang et al.. 2010). In black mussel
exposed for 10 days to sublethal concentrations of Pb, MDA levels were increased in mantle and gill
(Vlahogianni & Valavanidis. 2007).
Aminolevulinic acid dehydratase (ALAD) is a recognized biomarker of exposure across a wide
range oftaxa 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
from Spain, ALAD activity was negatively correlated with total Pb concentration in seven marine
bivalves (C. gallina, Mactra corallina, Donax trunculus, Cerastoderma edule,M. galloprovincialis,
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Scrobicularia plana and Crassotrea angulata), however, the authors of this study indicated the need to
consider species-dependent responses to Pb (Company et al.. 2011). Pb content varied significantly
among species and was related to habitat (sediment versus substrate) and feeding behavior. In red fingered
marsh crab, Parasesarma erythodactyla, 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).
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 |ag 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. 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 |ag Pb/L (Jing et al.. 2007). Activity
of AcPase, a lysosomal marker enzyme, was detected at 24 hours and decreased at subsequent time
points. Phenoloxidase activity was depressed compared with controls and remained significantly lower
than control at 72 hours of exposure to Pb.
The effect of Pb on the osmoregulatory response has been studied after 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, Dilocarcinuspagei (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.
Additional evidence of reproductive and developmental effects of Pb on aquatic invertebrates is
available since the 2006 Pb AQCD. 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 |ag Pb/L, while no embryos survived in the highest concentration.
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Formation of tentacles and eyes was significantly impaired in embryos of the freshwater ramshorn snail
Marisa cornuarietis at 15,000 (.ig Pb/L (Sawasdee & Kohler. 2010). Theegala et al. (2007) observed that
the rate of reproduction was significantly impaired in Daphnia pulex at >500 (.ig Pb/L in 21 day
exposures. 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, C. quinquefasciatus, exposed to 50 (ig Pb/L , 100 (ig Pb/L or 200 |ag
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 (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 |ag 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/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.
Since the publication of the 2006 Pb AQCD, limited studies on marine invertebrates have indicated
adverse effects of Pb on reproduction in saltwater environments. In a long term (approximately 60 days)
sediment bioassay with the marine amphipod Elasmopus laevi, onset to reproduction was significantly
delayed at 118 (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
was also reduced with increasing Pb concentration in sediment. 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 j.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 etal.. 2007).
The protective barrier against Pb toxicity formed by the egg structure in some invertebrates (e.g.,
Daphnia) 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
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the embryo and since the eggs are fixed on substrata in shallow coastal waters they may be subject to
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, however concentrations associated
with the eggshell increased throughout the 48 day exposure. These results are consistent with cuttlefish
eggs collected from the field in which Pb was only detected in the eggshell and indicate the protective
barrier provided by cuttlefish egg to Pb toxicity (Mi ram and et al.. 2006).
As noted in the 2006 Pb AQCD, Pb exposure negatively affects the growth of aquatic invertebrates.
Some studies reviewed in the previous document suggested that juveniles do not discriminate between the
uptake of essential and non-essential metals (Arai et al.. 2002). In new literature, the freshwater
pulomonate 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 & Brix. 2009; Grosell et al..
2006b). 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 & 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 susceptibility
of this species.Wang et al., {, 2009, 533439} 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/L (A. S. Sobrino-
Figueroa et al.. 2007). Rate of growth of the deposit feeding polychaete Capitella sp. decreased
significantly with increasing concentrations of Pb associated with sediment (Horng et al.. 2009).
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. Additionally, the carapace of hard shelled crabs have
lower concentrations of Pb than the exuvium of the soft shell crabs, leading the authors to speculate that
some of the metal might be excreted during the molting process. 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 (Uca pugnax) 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.
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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. Valve
closing speed was used as a measure of physiological alterations due to Pb exposure in the Catarina
scallop (A. Sobrino-Figueroa & 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. Feeding rate of the blackworm L. variegatus was significantly
suppressed by day 6 of a 10 day sublethal test in Pb-spiked sediments (Penttinen et al.. 2008) as compared
to feeding rates at the start of the experiment. However, this decrease of approximately 50% of the initial
feeding rate was also observed in the controls; therefore it is likely caused by some other factor other than
Pb exposure.
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 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
polychaete Capitella sp., exposure to varying concentrations of Pb associated with sediment up to 0.41
(imol/g had no effect on survival (Horng et al.. 2009). 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 Pb/L Pb (185,000 |ag Pb/L) (Tollett
et al.. 2009). Other species are more susceptible to Pb in the environment and these responses are
reviewed in Section 7.3.5.
7.3.4.3. Vertebrates
Biological effects of Pb on fish that have been studied since the 2006 Pb AQCD report are
reviewed here, and limited new evidence of Pb effects on amphibians, and marine mammals are
considered. As noted in the 2006 Pb AQCD, commonly observed effects of Pb on fish included inhibition
of heme formation, alterations in brain receptors in fish, adverse effects on blood chemistry, and decreases
in some enzyme activities. (U.S. EPA. 2006). Functional responses resulting from Pb exposure included
increased production of mucus, changes in growth patterns, and gill binding affinities. According to Eisler
(2000) and reviewed in the 2006 Pb AQCD, the general symptoms of Pb toxicity in fish include
production of excess mucus, lordosis, anemia, darkening of the dorsal tail region, degeneration of the
caudal fin, destruction of spinal neurons, ALAD inhibition, growth inhibition, renal pathology,
reproductive effects, growth inhibition and mortality. More recent experimental data presented here
expand and support these observations. As in terrestrial vertebrates, Pb has been shown to affect
antioxidant and enzymatic activity in aquatic vertebrates and new evidence of this since the 2006 Pb
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AQCD is reviewed in this section. This section also presents the limited 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 and growth 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.
Fish
In environmental assessments of metal-impacted habitats, ALAD is a recognized biomarker of Pb
exposure ("U.S. EPA. 2006). 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.. 2005; Schmitt et al.. 2007). 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 (Heieret al.. 2009). Costa et al. (2007) observed inhibition of
ALAD in hepatocytes of the neotropical traira (Hoplias malabaricus) following dietary dosing of 21 |ag
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).
Pb was shown to inhibit hepatic cytochrome P450 in carp (Cyprinus 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 |_ig 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.
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 (Carassius auratus
gibelio) injected with 10, 20 or 30 mg Pb/kg wet weight Pb-chloride showed a significant increase in the
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rate of production of superoxide ion and hydrogen peroxide in liver (Ling & Hong. 2010V 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, glucose-6-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 (Mageretal.. 2008). In the freshwater fish Nile tilapia, liver catalase, liver alkaline
phosphatase, sodium and potassium-ATPase (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 & Canli. 2007). Liver catalase activity significantly increased in the 5 and 20 (.iM concentrations
while 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. In another study
with (). niloticus, 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 & Canli. 2008V
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 |_ig 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, aspirate
transaminase, and lactate dehydrogenase activities significantly decreased in kidney in a concentration-
dependent manner (Dai. Fu. et al.. 2009) 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. Du. et al.. 2009).
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.
Data on the physiological effects of Pb on marine elasmobranchs are limited. De Boeck et al.
(2010) exposed the spotted dogfish to 2,072 |_ig Pb/L for one week and measured metallothionein
induction, and the electrolytes Na, K, Ca and CI. No effects were observed in Pb-exposed fish in any of
the physiological parameters measured in this study, however Pb was measured in all organs (De Boeck et
al.. 2010).
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Additional evidence of the neurotoxic effects of Pb on teleosts has become available since the 2006
Pb AQCD. The mitogen-activated protein kinases (MAPK), extracellular signal-regulated kinase
(ERK)l/2 and p38MAPK were identified for the first time as possible molecular targets for Pb neurotoxicity
in ateleost (Leal et al.. 2006). The phosphorylation of ERK1/2 and p38MAPK by Pb was determined in
vitro and in vivo in the catfish (Rhamdia queleri). R. quelen exposed to 1,000 |_ig Pb/L acetate for two
days showed a significant increase in phosphorylation of ERK1/2 and p38MAPK in the nervous system.
Incubation of cerebellar slices for 3 hours in 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
(Hoplias malabaricus) 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 (.ig Pb/L Pb for 4 weeks included perivascular edema, focal areas of malacia, and
diffuse areas of neuronal degeneration (Adevemo. 2008b').
Adverse effects of Pb on blood chemistry of fish were noted in the 2006 Pb AQCD and limited new
literature since the last Pb review has considered effects on blood. 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.
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 & Wood. 2004; Rogers & 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 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 & Blust. 2009b).
Long-term exposures of Pb can adversely 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
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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 reabsorption of Ca+, Mg2+, glucose, and water in
the teleost kidney (Patel et al.. 2006). Renal parameters including urine flow rate, glomerular filtration
rate, urine pH, and ammonia excretion were monitored in a 96-hour exposure of rainbow trout to 1.2 mg
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).
Limited new studies on reproductive effects of Pb in fish from oocyte formation to spawning are
available. Decreased oocyte diameter and density in the toadfish (Tetractenos glaber) were associated
with elevated levels of Pb in the gonad (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. The effects of
metals on embryonic stage of fish development in Cyprinus 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. 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
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controls and in low HC03" and DOC treatments with Pb. No significant differences were present between
treatments in egg hatchability.
Reproductive effects of Pb have also been observed at the cellular level, 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).
Reduction of growth was noted as an adverse effect of Pb on fish in the 2006 Pb AQCD. No new
evidence of growth effects in fish have been reported with the exception of Grosell et al. (2006a'). 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. (2006a') observed a significant increase in growth in some groups exposed to
higher concentrations, however, the increase in body mass was noted to have occurred in tanks with high
mortality earlier in the exposure (Grosell et al.. 2006a'). 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. Du. et al.. 2009). Growth and survival were not adversely 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). 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. (2011) did
not observe growth impairment in fathead minnows at environmentally relevant concentrations of Pb.
Since the publication of the 2006 Pb AQCD, several studies integrating behavioral and
physiological measures of Pb toxicity have been conducted on 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
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hypothalamus and mesencephalic areas of the brains of Pb-treated fish. No changes in feeding activity
were noted between non-treated and treated 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, 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 momoamines. 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 either 35 or 120 (ig Pb/L for 300 days, then subsequently tested in a breeding
assay for 21 days. The time interval between 1st and 5th ingestion of 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 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.
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 7.2.2.3 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. 2006). 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 (T. H. Chen et al.. 2006). In this chronic study,
the concentrations represent the range of Pb found in surface freshwaters across the U.S. The lowest
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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. 2002). 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) 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 (Rana
nigromaculata) treated with 100 |_ig Pb/L, 200 |_ig Pb/L, 400 |_ig Pb/L, 800 |_ig Pb/L or 1,600 |_ig Pb/L Pb-
nitrate by epidermal adsorption for 30 days (M. Z. Wang & 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 DNAtail 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 toad (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 |_ig Pb/L) did not exhibit genotoxic effects
while both 1,000 and 10,000 |_ig Pb/L significantly increased MNE to 14 and 202, respectively compared
to the control (6 MNE). In another chronic genotoxic study, erythrocytic micronuclei and erythrocytic
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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
(Y. M. Zhang et al.. 2007). 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.
In a study with 4-day-old X. laevis tadpoles exposed to a range of concentrations of Pb from 25,500
to 137,000 (.ig Pb/L for 24 hours, acetylcholinesterase was significantly inhibited in all treatments
(Gungordu et al.. 2010). The authors suggest that the 35-60% inhibition of acetylcholinesterase is
indicative of a neurotoxic effect. In the same study, glutathione-s-transferase activity significantly
increased in a concentration-dependent manner. Alanine aminotransferase and aspartate aminotransferase
activities were decreased, however, the degree of inhibition did not reflect Pb concentration. The
concentrations used in this study were selected based on the LC50 of Frog Embryo Teratogenesis Assay
results w ith X. laevis and are much higher than ambient levels of Pb.
Birds
In addition to effects on amphibians, 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 |_ig Pb/L for 30 minutes on day 0 of
development did not increase malformations or mortality of embryos (kertesz & 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.
Mammals
Although Pb continues to be detected in tissues of marine mammals in U.S. coastal waters (Bryan
et al.. 2007; Kannan et al.. 2006; Stavros et al.. 2007) few studies exist that consider biological effects
associated with Pb-exposure. Pb effects on immune parameters 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
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.
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7.3.5. Exposure and Response of Aquatic Species
To support the development of air quality criteria standards that are protective of aquatic
ecosystems, threshold levels for Pb effects on aquatic populations must be evaluated. The Annex of the
2006 Pb AQCD ("U.S. EPA. 2006) summarized data on exposure-response functions for freshwater and
marine invertebrates (Table AX7-2.4.1) and freshwater and marine 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).
A series of 72-hour Pb toxicity tests were conducted with five marine microalgae species chuii.
R. salina, Chaetoceros sp., I. galbana and N. 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 (Dcbelius 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. Additionally, the aquatic freshwater microalgae Scenedesmus obliquus was significantly more
susceptible 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 were derived (Atici et al..
2008). Experiments with the blue-green algae Spirulinaplatensis produced a LC50 value of 75.3 (.ig
Pb/mL (95% CI: 58.5, 97.0)( Arunakumara et al.. 2008).
In the 2006 Pb AQCD, adverse 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 publication of the 2006 Pb
AQCD, recent studies have identified the freshwater snail L. stagnalis as a species that is extremely
sensitive to Pb exposure (Grosell & Brix. 2009; Grosell et al.. 2006b). Growth of juvenile L. stagnalis
was inhibited at an EC2o of < 4 |_ig Pb/L. In contrast, freshwater juvenile ramshorn snails M cornuarietis
were less sensitive to Pb with the same LOEC for hatching rate and LC50, calculated to be about 10,000
(.ig Pb/L (Sawasdee & Kohler. 2010).
Additional studies on Pb effects in aquatic invertebrates published since the 2006 Pb AQCD have
indicated differences in susceptibility 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 adults in seawater
resulted in an LC50 of 1,270 |_ig Pb/L, an NOEC of 190 |_ig Pb/L and a LOEC of 390 |ag 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.
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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 (N. Wang et al.. 2010). 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 EC50 220 |_ig Pb/L) in contrast to older juveniles (2 months old) with an EC50 >426 |_ig/L. The
24-hour median effect concentration for glochidia (larvae) of L. siliquoidea in 48-hour acute toxicity tests
was >299 (ig/L. A 28 day exposure chronic value of 10 (ig Pb/L was obtained from 2-month-old L.
siliquoidea juveniles, and was the lowest genus mean chronic value ever reported for Pb (N. Wang et al..
2010). 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. In contrast, 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 & Valavanidis. 2007).
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 al. 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.. 2006b). The EC2o for reduced growth and
emergence of the midge Chironomus dilutus was 28 |ag 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 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
l_ig Pb/L was observed in water-only exposures of the blackworm Lumbriculus variegatus (Penttinen et
al.. 2008).
Cladocerans are commonly tested aquatic organisms, with data from three species: D. magna, D.
pulex and Cerodaphnia dubia, representing approximately 70% of available metal toxicological literature
on this group (L. C. Wong et al.. 2009). Since the publication of the 2006 Pb AQCD, additional studies
have generated acute toxicity values for other cladocerans. Median lethal concentrations for Moina
micrura (LC50 690 |ag Pb/L), Diaphanosoma birgei (LC50 3,160 |ag Pb/L), and A Ion a rectangular (LC50
7,000 (ig Pb/L) indicate differences in susceptibility to Pb in these freshwater species 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).
Exposure-response assays on other freshwater species have been conducted since the 2006 Pb
AQCD. In the mayfly Baetis tricaudatus, the 96-hour LC50 was 664 |ag Pb/L (Mebane et al.. 2008). An
EC2o value of 66 |ag 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). For rotifer Brachionuspatulus neonates, the
24-hour LC50 was 6,150 j.ig Pb/L (Garcia-Garcia et al.. 2007). In a 48-hour toxicity test with the rotifer
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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.. 2006b').
Since the publication of 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 |ag Pb/L. The 96-hour LC50 for
larvae of the same species was 353 |_ig Pb/L {Wang, 2009, 533439}. In comparison, juvenile Catarina
scallop (A. ventricosus) had a LC50 of 830 |_ig Pb/L in a 96-hour exposure (A. S. Sobrino-Figueroa et al..
2007). Morphological deformities were observed in 50% of veliger larvae of blacklip abalone (Haliotis
rubra) at 4,100 |_ig Pb/L following a 48-hour exposure to Pb, suggesting this species is not as sensitive to
Pb as other marine invertebrate larvae (Gorski &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 life stages 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 |ag 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 |ag Pb/L and a maximum effect at 3,000 |ag 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 a NOEL of 250 |_ig Pb/L.
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/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/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 |_ig Pb/L using the Frog Embryo
Teratogenesis Assay (Gungordu et al.. 2010).
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. 2006). 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
freshwater and marine fish. At the time of the 2006 Pb AQCD, there was a lack of exposure-response data
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in marine fish. No new exposure-response studies in marine fish have been conducted since the previous
Pb review.
A series of studies published since the 2006 Pb AQCD have been conducted and have further
elucidated the influence of water chemistry parameters on Pb uptake and toxicity in fathead minnow
resulting in additional dose-response data for this species. Grosell et al. (2006b) 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.. 2006a').
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. (2011)
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.. 2011).
In the 2006 Pb AQCD, fish size was recognized as an important variable in determining the adverse
effects of Pb. 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 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/L, maximum acceptable toxicant concentration=36 (ig/L, ECi0=26 (ig/L, EC2o=34 (ig/L, and LC50=55
(.ig/L. Results from the 62-day test, with fish length as the endpoint, were NOEC=8 (ig/L, MATC=12
(ig/L, ECio=7(ig/L, EC2o=102 (ig/L and LC50=120 j^ig/L. In acute tests run concurrently with the chronic
tests, 96-hour LC50 values were 120 and 150 (ig/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.
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7.3.6. Community and Ecosystem Effects
As discussed in the 2006 Pb AQCD, exposure to Pb is likely to have adverse impacts in aquatic
environments via effects at several levels of ecological organization (individuals, populations,
communities, or ecosystems). These adverse effects resulting from toxicity of Pb would be evidenced by
changes in species composition and richness, in ecosystem function, and in energy flow. The 2006 Pb
AQCD concluded that, in general, there was insufficient information available for single materials in
controlled studies to permit evaluation of specific impacts on higher levels of organization (beyond the
individual 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, Pb exposure has been demonstrated to alter predator-prey interactions, as
well as feeding and avoidance behaviors. 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.
Both plant species and type of habitat were determined to be factors affecting the rate of Pb
accumulation from contaminated sediments. While the rooted aquatic plant E. canadensis was observed
to accumulate the highest concentrations of Pb, the authors concluded that submerged macrophytes
(versus emergent plants) as a group were the most likely to accumulate Pb and other heavy metals
(Kurilenko & 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. A total of 11 species of
macrophytes were collected. 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 (V.
K. Mishra et al.. 2008). The site with the highest Pb concentrations (103 to 118 (.ig Pb/L) had the lowest
number of resident macrophyte species; these included E. crassipes, L. minor, Azolla pinnata and
Spirodela 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 (V. K. 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).
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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.
The faunal composition of seagrass beds in a Spanish coastal saltwater lagoon was found to be
impacted by Pb in sediment, plants, and biofilm (Marin-Guirao et al.. 2005). Sediment Pb concentrations
ranged from approximately 100 to 5,000 mg Pb/kg and corresponding biofilm concentrations were 500 to
1,600 mg Pb/kg, with leaf concentrations up to 300 mg Pb/kg. Although multiple community indices
(abundance, Shannon-Wiener diversity, Simpson dominance index) did not vary from site to site,
multivariate analysis and similarity analysis indicated significant differences in macroinvertebrate
communities between sites with different sediment, biofilm, and leaf Pb concentrations. Differences were
largely attributable to three amphipod species (Microdeutopus sp., Siphonoecetes sabatieri, Gammarus
sp.). This indicates that, although seagrass abundance and biomass were unaffected by Pb exposure,
organisms inhabiting these plants still may be adversely impacted.
In 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.
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 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).
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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 & 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. brigthwellii collected
from a Pb-impacted reservoir in Mexico, contained 384 ng Pb/mg and exhibited a water-to-tissue BCF of
49,344. The authors theorized that Pb biomagification 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 etal.. 2008). They showed thatM micrura had twice the Pb body burden of D. similis, another
grazing cladoceran species present in the reservoir. These two species exhibited average Pb tissue
concentrations of 57 and 98 ng Pb/mg, respectively, with respective water column BCFs of 9,022 and
8,046. Conversely, an examination of the simultaneous uptake of dissolved Pb by the algae P. subcapitata
and the cladoceran D. magna suggests that the dietary exposure route for the water column filter-feeder is
minor. Although Pb accumulated in the algal food source, uptake directly from the water column was
determined to be the primary route of exposure for D. magna (Komiarova & Blust. 2009c).
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 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 risk assessment conducted in southern Florida freshwater canals determined that the 90th
percentile of bulk Pb sediment concentrations in systems was 105 mg Pb/kg, which was predicted to
result in a sediment pore water concentration of 2.6 |_ig Pb/L. This estimated pore water concentration was
contrasted with acute 10th percentile toxic concentrations derived from a series of species sensitivity
distributions: 8.7 |_ig Pb/L for arthropods, 223 |_ig Pb/L for fish species, and 116 (ig Pb/L for all species
(Rand & Schuler. 2009). Although the predicted sediment pore water Pb concentration was below the
derived acute toxicity hazardous concentration for 10% of species (HC10) values, it was considered
possible that chronic exposure to such concentrations could impact some arthropod populations. A
chronic species sensitivity distribution constructed with Pb NOEC values for all aquatic species produced
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a chronic 10th percentile Pb NOEC value of 1.6 (ig/L, a further indication that some aquatic species
(likely arthropod species) may be impacted by the Pb contamination in southern Florida canals (Rand &
Schuler. 2009).
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.
fructicosa 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 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 & Gopal. 2009V 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.
7.3.7. Critical Loads in Aquatic Systems
Since the publication of the 2006 Pb AQCD there is no new significant information on critical
loads of Pb in aquatic systems. Refer to Section 7.3.6 of the 2006 Pb AQCD for a discussion of critical
loads of Pb in aquatic systems.
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7.3.8. Characterization of Sensitivity and Vulnerability
Data from the literature indicate that exposure to Pb may affect survival, reproduction, growth,
metabolism, and development in a wide range of 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 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 aquatic organisms. A review of the more recent literature corroborated these
findings, and identified seasonally-affected physiological changes and life-stage as other important
determinants of differential sensitivity to Pb.
7.3.8.1. 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.. 2007V
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 & Anbazhagan. 2009V
Couture et al. (2010) investigated seasonal and decadal variations in Pb sources to mussels (M
edulis) from the French Atlantic shoreline. Pb concentrations in the mussels were 5-66 times higher than
the natural background value for the north Atlantic. The 206Pb/207Pb signature indicated that the
bioaccumulated Pb was anthropogenic in origin. The signature was not, however, the same as that emitted
in western Europe as a result of leaded gasoline combustion, although that was a major emission source to
the atmosphere during a large part of the study period (1985-2005). Instead, it was most similar to that of
Pb released into the environment from wastewater treatment plants, municipal waste incinerators and
industries such as metal refineries and smelters. Thus continental runoff rather than atmospheric
deposition was identified as the main source of Pb to the French coastal area. The strong seasonal
variations in 206Pb/208Pb were used to conclude that resuspension of Pb triggered by high river runoff
events was a key factor affecting bioaccumulation of Pb inM edulis. In another biota monitoring study,
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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. 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.
7.3.8.2. 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.
7.3.8.3. Temperature and pH
Water temperature also appears to affect the toxicity of Pb to aquatic organisms, with higher
temperatures leading to greater responses. Pb toxicity to crayfish increased 7 to 10% when the water
temperature was increased by 4°C, and by 14% when the temperature increased by 7°C. The authors
determined that the increased toxicity was a result of the negative impact of Pb on crayfish respiration,
which was exacerbated by the lower dissolved oxygen concentrations at higher water temperatures (M. A.
Q. Khan et al.. 2006). 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 & Taner. 2009). Decreased pH was also demonstrated to
increase the uptake of environmental Pb in aquatic plants (Uvsal & Taner. 2009; C. Wang. X. Yan. et al..
2010). Additionally, Birceanu et al. (2008) determined that fish (specifically rainbow trout) were more
susceptible to Pb toxicity in acidic, soft waters characteristic of sensitive regions in Canada and
Scandinavia. Hence, fish species endemic to such systems may be more at risk from Pb contamination
than fish species in other habitats.
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7.3.8.4. Life Stage
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 LC50s 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.
(2010) demonstrated that the newly transformed juvenile mussels, L. siliquoidea and L. rafmesqueana, 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 (N. Wang et al.. 2010).
Pb exposures also differentially affected life-stages of the marine polychaete H. elegans. 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 off-set (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 |_ig Pb/L and
greater (Gopalakrishnan et al.. 2008).
7.3.8.5. Species Sensitivity
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 Hydropysche sp, caddisflies,
with only the mayfly exhibiting increased Pb tissue concentrations when collected from Pb-contaminated
sites; the caddisfly Pb tissue concentrations were similar between reference and Pb-contaminated areas.
The authors also examined the lifestage specific accumulation of Pb for E. virgo mayflies, and although
there was no statistical difference in Pb tissue concentrations between different lifestages, Pb
bioaccumulation did change as mayflies aged (Cid et al.. 2010). Data from 20 years of monitoring of
contaminant levels in filter-feeding mussels Mytilus sp. and oysters Crassotrea virginica 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).
Species-specific Ca requirements have also 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
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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 & Brix. 2009V Grosell and
colleagues also noted that reduced snail growth following chronic Pb exposure was likely a result of
reduced Ca uptake (Grosell et al.. 2006b').
There is some indication that molting may comprise an additional sequestration and excretion
pathway for aquatic animals exposed to Pb (Bergev & Weis. 2007; Mohapatra et al.. 2009; Soto-Jimenez
et al.; Tollett et al.. 2009). Crab species U. pugnax (Bergev & Weis. 2007) and Scylla serrata (Mohapatra
et al.. 2009). white shrimp L. vannamei (Soto-Jimenez et al.) as well as Libellulidae dragonfly nymphs
(Tollett et al.. 2009) have been shown to preferentially sequester Pb in exoskeleton tissue, where it is later
shed along with the hardened exterior tissue. Consequently, aquatic arthropod species and those species
that shed their exoskeleton more frequently may be able to tolerate higher environmental Pb
concentrations than non-arthropods or slow-growing molting species, as this pathway allows them to
effectively lower Pb body burdens.
Some tolerant species of fish (e.g., mummichog) have the ability to sequester accumulated Pb in
metal-rich granules or heat-stable proteins (Goto & Wallace. 2010). Fish with such abilities are more
likely to thrive in Pb-contaminated environments than other species. 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 & Gopal. 2009).
Tolerance to prolonged Pb exposure may develop in aquatic invertebrates and fish. Multi-
generational exposure to low levels of Pb appears to confer some degree of metal tolerance in
invertebrates such as Chironomusplumosus larvae; consequently, previous population Pb exposures may
decrease species' susceptibility to Pb contamination (Vedamanikam & Shazilli. 2008). However, the
authors noted that metal tolerant larvae were significantly smaller than larvae reared under clean
conditions, and that transference of Pb-tolerant C. plumosus larvae to clean systems resulted in a
subsequent loss of tolerance. Evidence of acclimation to elevated Pb in fathead minnow was suggested in
the variations in ionoregulatory parameters that were measured on day 10 and 30 in fish exposed to 115
(.ig Pb/L for 30 days. At the end of the experiment, whole body Ca2+ was elevated while Na+ and K+
recovered from elevated levels at 30 days (Grosell et al.. 2006a).
A series of species sensitivity distributions constructed by Brix et al. (2005) 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
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species sensitivity distributions for both cladoceran and copepod 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 (L. C. Wong et al.. 2009V
7.3.8.6. 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 susceptible to effects of Pb include habitats that are acidified due to
atmospheric deposition of pollutants, runoff from mining activities or lakes and streams with naturally
occurring organic acids. Hence, fish and invertebrate species endemic to such systems may be more at
risk from Pb contamination than corresponding species in other habitats.
7.3.9. Ecosystem Services
There are limited publications at this time that address Pb impacts to ecosystem services associated
with aquatic systems and most of the available literature is for 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
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A few recent studies consider the impact of Pb and other heavy metals on ecosystem services
provided by estuaries (Smith et al.. 2009) and salt marsh {Bromberg Gedan, 2009, 672706}. These
systems are natural sinks for metals and other contaminants. They provide habitat and breeding areas for
both terrestrial and marine wildlife and are locations for bird watching. In a Monte-Carlo 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 may pose an ecologically relevant risk to dunlin,
Calidris alpina (Smith et al.. 2009). However, the authors noted that a major source of uncertainty in this
study was the NOAEL values for Pb. Pb can be toxic to salt marsh plant species and decaying plant
detritus may result in resuspension of Pb into the aquatic food chain {Bromberg Gedan, 2009, 672706}.
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 &
Dillon. 2007). Several studies have addressed the response of natural wetlands to Pb (Gambrell. 1994;
Odiim. 2000). Recent reviews of pollution control by constructed wetlands (Mander & Mitsch. 2009).
removal of metals by constructed wetlands (Marchand et al.. 2010) and phytoremediation of metals by
wetland plants (Rai. 2008) indicate that these systems can remove Pb from the aquatic environment. 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. 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. The conclusions of a recent assessment report are that
the highest concentrations of Pb are found in mussels and oysters near urban and industrial centers and
the only region that exceeded the Food and Drug Administration action level for Pb in clams, oysters and
mussels of 1.7 mg Pb/kg wet weight was Lake Michigan, where maximum concentrations were 1.9 mg
Pb/kg wet weight in the years 2004-2005 (Kimbrough et al.. 2008). 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 an ecosystem service that 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
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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.
7.3.10. 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 AQCD considered to determine
causality.
7.3.10.1. Biogeochemistry and Chemical Effects
Once the atmospherically-derived 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.
7.3.10.2. 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
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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.
Recent studies on bioavailability of Pb in 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
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.
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 (Einspom et al.. 2009; Einspom & Koehler. 2008). Since the
2006 Pb AQCD, BCF's have been measured for several species in field studies and these BCF's tend to be
higher than calculated BCF's from laboratory exposures (Table 7-4).
Tissue accumulation of Pb via gill and dietary uptake has been further characterized in aquatic
vertebrates and stable isotope techniques have been applied to further elucidate bioaccumulation of Pb in
this ISA. 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 of 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 & Wallace. 2010).
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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 higher trophic-level consumers indicating that Pb was
bioconcentrated but not biomagnified (Eisler. 2000; U.S. EPA. 2006V 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).
7.3.10.3. 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 (Grose ll et al..
2006b). 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 p38VIAPK were 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
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Ca homeostasis as previously characterized in the 2006 Pb AQCD (Rogers & Wood. 2004; Rogers &
Wood. 2003V 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.
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. 2006). 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.
7.3.10.4. 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 an endpoint adversely affected by Pb exposure. The lowest EC50 for growth observed in marine
microalgae and freshwater microalgae was in the range of 100 (ig/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 & Brix. 2009; Grosell et al.. 2006b'). 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 (N. Wang et al.. 2010).
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 AQCD.
7.3.10.5. Community and Ecosystem Effects
Since the publication of 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 towards 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
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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 &
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
species sensitivity distributions constructed by Brix et al. ("2005) 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.
7.3.10.6. Critical Loads, Sensitivity and Vulnerability
Since the 2006 Pb AQCD there is no new significant information on critical loads of Pb in aquatic
systems. Recent studies have identified seasonally-affected physiological changes and life-stage 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 adverse effects.
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7.4. Causality Determinations for Lead in
Terrestrial and Aquatic Systems
This section presents key conclusions regarding causality determinations for welfare effects of Pb
(Table 7-6). Evidence considered in establishing causality was drawn from the 1977 (U.S. EPA. 1977).
1986 (U.S. EPA. 1986) and 2006 Pb AQCD (U.S. EPA. 2006) for Pb where appropriate as well as the
current ISA. EPA's framework for causality described in Chapter 1 was applied and the causal
determinations are highlighted. In this ISA, effects determined to be causal at the species level contribute
to the body of evidence for causal effects at the community and ecosystem scale. Some of the effects of
Pb observed in terrestrial and aquatic organisms are also considered in the chapters of the ISA that
evaluate evidence for human health effects associated with Pb exposure.
Table 7-6. Summary of Pb causal determinations for plants, invertebrates and vertebrates
Effect
Causality Determination
Bioaccumulation as it Affects Ecosystem Services-All organisms
Causal
Mortality-Plants
Inadequate
Mortality- Invertebrates and Vertebrates
Causal
Growth-Plants
Causal
G rowth-l nvertebrates
Causal
Growth-Vertebrates
Suggestive
Physiological Stress-All organisms
Causal
Hematological Effects-Invertebrates and Vertebrates
Causal
Development and Reproduction-Invertebrates and Vertebrates
Causal
Development and Reproduction-Plants
Inadequate
Neurobehavior-lnvertebrates and Vertebrates
Causal
Community and Ecosystem Level Effects
Causal
7.4.1. Bioaccumulation of Lead in Terrestrial and Aquatic
Biota as it Affects Ecosystem Services
Pb deposited on the surface of, or taken up by organisms has the potential to alter the services
provided by terrestrial and aquatic biota to humans. Ecosystem services are the benefits people obtain
from ecosystems. They include supporting, provisioning, regulating and cultural services that are vital for
the functioning of the biosphere and provide the basis for the delivery of tangible benefits to human
society. There is compelling evidence over several decades of research that Pb bioaccumulates in plants,
invertebrates and vertebrates in terrestrial and aquatic systems. Generally, there are considerable
differences between species in the amount of Pb taken up from the environment and in the amounts of Pb
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retained in the organism. In order for Pb to reach a biological receptor, the metal must first cross the
membranes of organisms to the target organ or site of storage. This process varies between plants,
invertebrates and vertebrates, and furthermore, uptake and sequestration are at times similar in unrelated
species, while substantially different between related ones. Uptake of Pb from environmental media is
dependent upon the bioaccessibility of Pb (reviewed in Chapter 3) which is influenced by many factors
including, but not limited to, temperature, pH, presence of humic acid and dissolved organic matter,
presence of other metals, and speciation of Pb.
Terrestrial plants accumulate Pb via direct stomatal uptake into foliage and incorporation of
atmospherically-deposited Pb from soil into root tissue, followed by variable translocation to other
tissues. Near smelters and other point-sources of Pb, leaf uptake may in some cases be greater than root
uptake, otherwise root uptake is predominant. Translocation of soil Pb to shoots and leaves is limited in
most species as plants sequester a large portion of Pb in root tissue. There are considerable species-
dependent differences in rate of uptake and translocation of Pb to other parts of the plant. Uptake and
sequestration of Pb primarily in roots of terrestrial plants is also observed in wetland species and algae.
Rooted aquatic plants take up Pb primary from sediments while floating aquatic plants take up Pb from
water.
Uptake of atmospherically deposited Pb from soil is the primary exposure route in terrestrial plants
and invertebrates. Bioaccessibility of Pb to soil-dwelling organisms is influenced by soil factors including
soil type, amount of OM, pH, and CEC and there are considerable differences in uptake among species.
Species exhibit different accumulation efficiencies and compartmentalize sequestered Pb differently.
There is limited evidence of contributions to total Pb body burden via dietary exposures in primary and
secondary consumers.
Aquatic organisms can uptake and bioaccumulate Pb from the water column, sediments or via
dietary exposure. However, as in terrestrial organisms, uptake and subsequent bioaccumulation of Pb in
aquatic plants, invertebrates and vertebrates varies greatly between species and across taxa. Invertebrates
may also sequester Pb in the exoskeleton, which is subsequently shed. Pb in aquatic invertebrates is
primarily sequestered in the gill and hepatopancreas. Uptake of Pb in fish is well characterized and occurs
primarily via direct uptake of dissolved Pb from the water column through gill surfaces and by ingestion
of Pb-contaminated diets. Pb in these organisms is primarily sequestered in the gill. In dietary exposures
in fish, Pb also targets the anterior intestine indicating the importance of water-only versus dietary uptake
exposures.
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. Recent
research has suggested that dietary Pb (i.e., Pb adsorbed to sediment, particulate matter, and food) may
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contribute to exposure and toxicity in primary and secondary order consumers (including humans).
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. Overall, based on the
consistency of findings across taxa, the evidence is sufficient to conclude that there is a causal
relationship between Pb exposures and bioaccumulation of Pb that affects ecosystem services
associated with terrestrial and aquatic biota.
7.4.2. Mortality
The relationship between Pb exposure and mortality has been well demonstrated in terrestrial and
aquatic species as presented in Sections 7.2.5 and 7.3.5 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. However, the LC50 is usually much higher than current environmental levels
of Pb, even though physiological dysfunction that adversely impacts the fitness of an organism often
occurs well below concentrations that result in mortality.
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.
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 that
approach ambient levels. These impacted species may include candidate or endangered species. The
freshwater mussel Lampsilis rafinesqueana (Neosho mucket), is a candidate for the endangered species
list. The EC50for foot movement (a measure of viability) for newly transformed juveniles of this species
was 188 (.ig Pb/L. (N. Wang et al.. 2010). However, tolerance to Pb varies substantially among
invertebrate species. As reviewed in the 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. 2006). 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. Toxicity of Pb to aquatic invertebrates is highly dependent on water quality
parameters such as pH, DOC and Ca2+. For example, 48 hour LC50 values for C. dubia range from 280 to
>2,700 |ag Pb/L when tested at varying pH levels (U.S. EPA. 2006). Other 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.
Early experiments from Pb mining areas indicated local extinction of fish from streams. Mortality
in fish is dependent on the sensitivity of the species tested and on water quality parameters. Higher
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toxicity tends to occur in acidic waters where more free-Pb ion is available for uptake. The interaction
between water quality parameters and Pb toxicity may result in a range of concentrations that cause
equivalent toxicity. For example, 96-hr LC50 values in fathead minnow range from 810->5,400 |_ig Pb/L in
varying pH and hardness (U.S. EPA. 2006). Increased mortality 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 (Grosell et al.. 2006b). 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).
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, 2005 #19390}. Mortality associated with Pb exposure in vertebrates is supported by the
consistently positive associations between Pb exposure and mortality observed in human epidemiologic
studies (Section 5.3.1).
The evidence is inadequate to conclude that there is a causal relationship between Pb
exposure and mortality in plants.
The evidence is sufficient to conclude that there is a causal relationship between Pb exposures
and mortality in sensitive terrestrial and aquatic animal taxa.
7.4.3. Growth Effects
Evidence for Pb effects on growth is strongest in plants with limited information on invertebrates
and vertebrates. There is evidence over several decades of research that Pb inhibits photosynthesis and
respiration in plants all of which reduce the growth of the plant (U.S. EPA. 1977. 1986. 2006). Many
laboratory toxicity studies report effects on plants; however, there are few field studies. 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. The lowest 72-hour EC50 for growth inhibition reported for algae was 105
|ag Pb/L in Chaetoceros sp. Most primary producers experience EC50 values for growth in the range of
1,000 to 100,000 (.ig Pb/L (U.S. EPA. 2006).
In previous AQCDs, growth effects of Pb have been reported in fish (growth inhibition), birds
(changes in juvenile weight gain), and frogs (delayed metamorphosis, smaller larvae). Growth effects
observed in invertebrates and vertebrates underscore the importance of lifestage to overall Pb
susceptibility. In general, juvenile organisms are more sensitive than adults. Several studies since the 2006
Pb AQCD have demonstrated adverse effects of Pb on growth at lower concentrations than in previous
literature. Among the animal taxa tested, aquatic invertebrates were the most sensitive to the effect of Pb,
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with adverse effects being reported as low as 4 |_ig Pb/L. Growth of juvenile freshwater snails L. stagnalis
was inhibited at EC:„ <4 j.ig Pb/L (Grose 11 & Brix. 2009; Grose 11 et al.. 2006b'). In the freshwater mussel,
fatmucket (L. siliquoidea) juveniles were the most sensitive lifestage (N. Wang et al.. 2010). 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. Evidence is sufficient to conclude that there is a causal relationship
between Pb exposures and growth effects in plants and invertebrates. Evidence is suggestive of a
causal relationship between Pb exposures and growth effects in vertebrates.
7.4.4. Physiological Stress
In this review and in previous Pb AQCDs there is consistent and coherent evidence that Pb induces
oxidative stress in plants, invertebrates, and vertebrates. This is consistent with evidence in humans and
experimental animal studies for oxidative stress development, due in many instances to the antagonism of
normal metal ion functions (Section 5.2.4). This oxidative stress is characterized by increased reactive
oxygen species and membrane and lipid peroxidation that can promote tissue damage, cytotoxicity, and
dysfunction. Increased reactive oxygen species are often followed by a compensatory and protective
upregulation in antioxidant enzymes, such that this observation is indicative of oxidative stress
conditions. Additionally, continuous reactive oxygen species production may overwhelm this defensive
process leading to further oxidative stress and injury.
Building on this strong body of evidence presented previously, recent studies provide consistent
and coherent evidence of upregulation of antioxidant enzymes and increased lipid peroxidation associated
with Pb exposure in many species of plants, invertebrates and vertebrates. In plants, increases of
antioxidant enzymes with Pb exposure occur in algae, aquatic mosses, floating and rooted aquatic
macrophytes, and terrestrial species. There is considerable evidence for antioxidant activity in
invertebrates, including gastropods, mussels, and crustaceans, in response to Pb exposure. Markers of
oxidative damage are also observed in fish and amphibians. Across all biota, there are differences in the
induction of antioxidant enzymes that appear to be species-dependent.
Additional stress responses to Pb exposure observed in terrestrial and aquatic invertebrates
including elevated heat shock proteins, osmotic stress, lowered metabolism and decreased glycogen levels
have been reported since the 2006 Pb AQCD. Heat shock protein induction by Pb exposure has been
observed in zebra mussels and mites. 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 exhibit a Pb-induced hypometabolism under conditions of environmental hypoxia in the
presence of the metal (Morris et al.. 2005). Tissue volume regulation is adversely affected in freshwater
crabs (Amado et al.. 2006). Glycogen levels in the freshwater snail B. glabrata were significantly
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decreased at near environmentally relevant concentrations of Pb (50 (ig/L and higher) (Ansaldo et al..
2006V
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. Evidence is sufficient to
conclude that there is a causal relationship between Pb exposures and physiological stress in plants,
invertebrates and vertebrates.
7.4.5. Hematological Effects
Hematological responses are commonly reported effects of Pb exposure in invertebrates and
vertebrates in both aquatic and terrestrial systems. In environmental assessments of metal-impacted
habitats, ALAD is a recognized biomarker of Pb exposure ("U.S. EPA. 2006). ALAD activity is negatively
correlated with total Pb concentration in bivalves. Lower ALAD activity has been correlated with
elevated blood Pb levels in fish and mammals as well. In the 1986 AQCD, decreases in red blood cell
ALAD activity following Pb exposure were well documented in birds and mammals. 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, suggest this enzyme is an
indicator for Pb exposure across a wide range of taxa. Limited evidence of Pb effects on other blood
parameters including altered serum profiles and changes in white blood cell counts in fish and amphibians
support the finding of the hematological system as a target for Pb in natural systems. This evidence is
strongly coherent with observations from human epidemiologic and animal toxicology studies (Section
5.7). There is consistent toxicological and epidemiologic 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 and laboratory animals (Section 5.7). 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 invertebrates and vertebrates.
7.4.6. Developmental and Reproductive Effects
Evaluation of the literature on Pb effects in aquatic and terrestrial species indicates that exposure to
Pb is associated with reproductive effects. Various endpoints have been measured in aquatic and terrestrial
organisms to assess the effect of Pb on development, fecundity and hormone homeostasis. Evidence in
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this review and the previous Pb AQCDs from invertebrate and vertebrate studies indicate that Pb is
adversely affecting reproductive performance in multiple species. However, there are typically only
limited studies available from different taxa.
In plants, few studies are available that specifically address reproductive effects of Pb exposure.
One study with grass pea showed Pb exposure increased pollen sterility.
Several studies with invertebrates provide evidence for adverse impacts to embryonic development,
specifically in snails, clams and rotifers. In addition to affecting the embryo, Pb can alter developmental
timing, 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.
Reproductive effects have also been observed in multi-generational studies. Larval settlement rate,
and rate of population increase was adversely impacted in rotifers. Rotifers have decreased fertilization
rate associated with Pb exposure that appeared to be due to decreased viability of sperm and eggs.
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.
In 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 subtle
skeletal malformations. Previous Pb AQCDs have reported developmental effects in fish, specifically
spinal deformities in larvae. 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 effects on 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-
contaminated prey. Additional reproductive parameters in fish observed to be impacted by Pb include
decreased oocyte diameter and density in toadfish associated with elevated Pb levels in gonad.
In terrestrial vertebrates, evidence from Chapter 7 of this ISA and in previous Pb AQCDs indicates
an association between observed adverse reproductive effects and Pb exposure. 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. Decreased testis weight was observed in lizards. In
mammals, few studies in the field have addressed Pb specifically, due to most available data in wild or
grazing animals being from near smelters, where animals are co-exposed to other metals.
In Chapter 5 evidence from mammals demonstrates a consistency of adverse effects of Pb on sperm
and the onset of puberty in males and females with strong evidence from both toxicology and
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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. 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.
Adverse effects of Pb on reproduction in invertebrates and vertebrates indicate that Pb is likely
adversely affecting fecundity of Pb-exposed organisms in both aquatic and terrestrial habitats, and
therefore 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. In plants, the
evidence is inadequate to conclude a causal relationship between Pb exposures and plant
reproductive effects.
7.4.7. Neurobehavioral Effects
Evidence from laboratory studies and limited data from field studies reviewed in Chapter 7 have
shown adverse effects of Pb on neurological endpoints in both aquatic and terrestrial animal taxa. These
include changes in behaviors that may decrease the overall fitness of the organism. There is also evidence
from both invertebrate and vertebrate studies that Pb adversely affects behaviors that may decrease the
ability of an organism to escape predators or capture prey.
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 p38VIAPK
as possible molecular targets for Pb neurotoxicity in catfish (Leal et al.. 2006). 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 & Wang. 2009).
Decreased food consumption of Pb-contaminated diet has been demonstrated in some invertebrates
(snails) and vertebrates (lizards, pigs). 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. In limited studies available on snails, tadpoles, scallops,
hatchling turtles and fish there is evidence that Pb may affect the ability to escape or avoid predation. The
motility of nematodes was adversely affected in Pb-contaminated soils (D. Y. Wang & Xing. 2008). 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
& Gochfeld. 2005). 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.
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These findings are coherent with findings from studies in laboratory animals as described in
Section 5.3 that show that Pb induces changes in learning, memory, attention and motor skills. Pb induced
new behaviors including hyperactivity and mood disorders. Within the sensory organs, the visual and
auditory systems are affected by Pb exposure. Changes in structure and function of neurons and
supporting cells in the brain are detailed including effects on the blood brain barrier. Mechanisms
including the displacement of physiological cations, oxidative stress and changes in neurotransmitters and
receptors are detailed. Based on evidence from several cohort and cross-sectional studies in diverse
populations, the overall weight of the available evidence provides clear and consistent evidence of
association between blood Pb concentrations and decrements in neurodevelopmental outcomes in young
children (Section 5.1). 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). These data from laboratory toxicology studies, especially neurobehavioral findings and structural
changes are highly coherent with data from ecological studies. Overall, the evidence from aquatic and
terrestrial systems is sufficient to conclude that there is a causal relationship between Pb exposures and
neurobehavioral effects in invertebrates and vertebrates.
7.4.8. Other Physiological Effects
In addition to the above mentioned physiological effects of Pb on organisms in terrestrial and
aquatic systems for which there is sufficient evidence to infer causality, there are a few recent studies that
can be linked to effects observed in humans for which there is insufficient evidence across taxa. Pb
exposure has been demonstrated to result in changes to DNA structure and chromosomal alterations in
plants, gastropods, mussels and fish. DNA damage, chromosomal damage and aberrations, and
micronucleus formation are also observed in humans and laboratory animals exposed to Pb (Section
5.10). Additional new evidence in this review indicates that Pb can interfere with renal function in fish,
specifically with ionoregulation of Na and CI and reabsorption Ca2+, Mg2+ glucose and water (Patel et al..
2006). In humans and laboratory animals, Pb is a recognized nephrotoxicant and is considered to be
causal of kidney damage (Section 5.5).
7.4.9. Community and Ecosystem Level Effects
Uptake of Pb into aquatic and terrestrial organisms and subsequent effects on survival,
reproduction, growth, behavior and other physiological variables at the species scale are likely to result in
effects at the population, community and ecosystem scale. The effects may include alteration of predator-
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prey dynamics, species richness, species composition, and biodiversity. There are few field studies
available that directly consider effects of Pb on these measures of ecosystem health. Ecosystem-level
studies are complicated by the confounding of Pb exposure with other factors such as 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.
Most direct evidence of community and ecosystem level effects is from near point sources. For
terrestrial systems there are several decades of research on impacts to natural ecosystems near smelters,
mines, and other industrial sources of Pb where Pb levels are elevated. 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 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. Effects of Pb on mortality, growth, physiological
stress, blood, neurobehavior and developmental and reproductive endpoints at the individual level are
expected to have ecosystem-level consequences, and thus provide consistency and plausibility for
causality in ecosystem-level effects.
Avoidance response to Pb exposure varies widely in different species and this could affect
community composition. For example, frogs and toads lack avoidance response while snails and fish
avoid higher concentrations of Pb (U.S. EPA. 2006). New evidence since the 2006 Pb AQCD indicates
that some species of worms avoid Pb-contaminated soils (Langdon et al.. 2005)
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 was shown to ameliorate
effects of Pb contamination on a microbial community.
In aquatic ecosystems there are numerous field studies on reductions of species abundance,
richness or diversity particularly in benthic macroinvertebrate communities coexisting with other metals.
For example, in the 2006 Pb AQCD, the Coeur d'Alene River watershed in Idaho, U.S. was used as a case
study for Pb effects at the population and community level. Significant negative correlations were
observed between Pb in water column and total taxa richness and EPT taxa richness in the river. In a
simulated aquatic microcosm a reduction in abundance and richness of protozoan species was observed
with increasing Pb concentration from 50 to 1,000 j.ig Pb/L (Fernandez-Leborans & Antonio-Garcia.
1988)
Since the last Pb AQCD, there is further evidence for effects of Pb in sediment-associated
communities. Exposure to three levels of sediment Pb contamination (322, 1,225, and 1,465 (.ig Pb/g dry
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weight) in a microcosm experiment significantly reduced nematode diversity and resulted in profound
restructuring of the community structure (Mahmoudi et al.. 2007V Sediment-bound Pb contamination
appears to differentially affect members of the benthic invertebrate community, potentially altering
ecosystems dynamics in small urban streams (Kominkova & Nabelkova. 2005). Although surface water
Pb concentrations in monitored streams were determined to be very low, concentrations of the metal in
sediment were high enough to pose a risk to the benthic community (e.g., 34 to 101 mg Pb/kg). These
risks were observed to 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.
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) (V
K. Mishra et al.. 2008). This shift toward more Pb-tolerant species is also observed in terrestrial plant
communities near smelter sites (U.S. EPA. 1986. 2006). 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.
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 in natural systems. However, the evidence for Pb effects at higher levels of ecological
organization is sufficient to conclude that there is a causal relationship between Pb exposures and the
alteration of species richness, species composition and biodiversity in terrestrial and aquatic
ecosystems.
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