United States Environmental Criteria and
Environmental Protection Assessment Office
Agency Research Triangle Park, NC 27711
EPA/600/8-83/0286F
June 1986
Research and Development
Air Quality
Criteria for Lead
Volume II of IV
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EPA/600/8-83/028bF
June 1986
Air Quality Criteria for Lead
Volume II of IV
U.S. ENVIRONMENTAL PROTECTION AGENCY
Office of Research and Development
Office of Health and Environmental Assessment
Environmental Criteria and Assessment Office
Research Triangle Park, NC 27711
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DISCLAIMER
This document has been reviewed in accordance with U.S. Environmental
Protection Agency policy and approved for publication. Mention of trade
names or commercial products does not constitute endorsement or
recommendation.
ii
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ABSTRACT
The document evaluates and assesses scientific information on the health
and welfare effects associated with exposure to various concentrations of lead
in ambient air. The literature through 1985 has been reviewed thoroughly for
information relevant to air quality criteria, although the document is not
intended as a complete and detailed review of all literature pertaining to
lead. An attempt has been made to identify the major discrepancies in our
current knowledge and understanding of the effects of these pollutants.
Although this document is principally concerned with the health and
welfare effects of lead, other scientific data are presented and evaluated in
order to provide a better understanding of this pollutant in the environment.
To this end, the document includes chapters that discuss the chemistry and
physics of the pollutant; analytical techniques; sources, and types of
emissions; environmental concentrations and exposure levels; atmospheric
chemistry and dispersion modeling; effects on vegetation; and respiratory,
physiological, toxicological, clinical, and epidemiological aspects of human
exposure.
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CONTENTS
VOLUME I
Chapter 1. Executive Summary and Conclusions i_i
VOLUME II
Chapter 2. Introduction 2-1
Chapter 3. Chemical and Physical Properties [' 3-1
Chapter 4. Sampling and Analytical Methods for Environmental Lead \\ 4-1
Chapter 5. Sources and Emissions '[ " 5-1
Chapter 6. Transport and Transformation '] 5-1
Chapter 7. Environmental Concentrations and Potential Pathways to Human Exposure 7-1
Chapter 8. Effects of Lead on Ecosystems " s-1
VOLUME III
Chapter 9. Quantitative Evaluation of Lead and Biochemical Indices of Lead
Exposure in Physiological Media g-1
Chapter 10. Metabolism of Lead 10-1
Chapter 11. Assessment of Lead Exposures and Absorption in Human Populations ....... li-l
Volume IV
Chapter 12. Biological Effects of Lead Exposure 12-i
Chapter 13. Evaluation of Human Health Risk Associated with Exposure to Lead
and Its Compounds 13-1
IV
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TABLE OF CONTENTS
.1ST OF FIGURES
,IST OF TABLES
2. INTRODUCTION
3. CHEMICAL AND PHYSICAL PROPERTIES
3. 1 INTRODUCTION
3. 2 ELEMENTAL LEAD
3. 3 GENERAL CHEMISTRY OF LEAD
3. 4 ORGANOMETALLIC CHEMISTRY OF LEAD
3. 5 FORMATION OF CHELATES AND OTHER COMPLEXES
3. 6 REFERENCES
3. A APPENDIX: PHYSICAL/CHEMICAL DATA FOR LEAD COMPOUNDS
3A. 1 Data Tables
3A. 2 The Chelate Effect
3A. 3 References
4. SAMPLING AND ANALYTICAL METHODS FOR ENVIRONMENTAL LEAD
4. 1 INTRODUCTION
4. 2 SAMPLING
4.2.1 Regulatory Siting Criteria for Ambient Aerosol Samplers
4.2.2 Ambient Sampling for Participate and Gaseous Lead
4.2.2.1 High Volume Sampler (hi-vol)
4.2.2.2 Dichotomous Sampler
4. 2. 2. 3 Impactor Samplers
4.2.2.4 Dry Deposition Sampling
4.2.2.5 Gas Collection
4.2.3 Source Sampl i ng
4. 2. 3. 1 Stationary Sources
4.2.3.2 Mobile Sources
4.2.4 Sampling for Lead in Water, Soil, Plants, and Food
4. 2. 4. 1 Precipitation
4. 2. 4. 2 Surface Water
4.2.4.3 Soils
4.2.4.4 Vegetation
424.5 Foodstuffs
4.2.5 Filter Selection and Sample Preparation
4. 3 ANALYSIS
4 3.1 Atomic Absorption Analysis (AAS)
4.3.2 Emission Spectroscopy
43.3 X-Ray Fl uorescence (XRF)
434 Isotope Dilution Mass Spectrometry (IDMS)
43.5 Colorimetric Analysis
436 Electrochemical Methods: Anodic Stripping Voltammetry
(ASV), and Differential Pulse Polarography (DPP)
437 Methods for Compound Analysi s
4 4 CONCLUSIONS
4. 5 REFERENCES
ix
xi
2-1
3-1
3-1
3-1
3-2
3-3
3-4
3-8
3A-1
3A-1
3A-3
3A-4
4-1
4-1
4-2
4-2
4-6
4-6
4-8
4-9
4-10
4-11
4-11
4-12
4-12
4-13
4-13
4-14
4-15
4-15
4-16
4-16
4-17
4-18
4-19
4-20
4-22
4-22
4-23
4-24
.... 4-24
4-9H
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TABLE OF CONTENTS (continued).
Page
SOURCES AND EMISSIONS 5-l
5-1
5-4
5-5
5-5
5-6
5-6
5-6
5-16
5-19
5-20
6-1
6-1
6-2
6-2
6-4
6-4
6-6
6-8
6-8
6-16
6-16
6-18
6-19
6-21
6-21
6-21
6-22
6-23
6-25
6-25
6-27
6-29
6-29
6-34
6-34
6-35
6-38
6-39
6-41
ENVIRONMENTAL CONCENTRATIONS AND POTENTIAL PATHWAYS TO HUMAN EXPOSURE 7-1
7.1 INTRODUCTION 7-1
7.2 ENVIRONMENTAL CONCENTRATIONS 7-1
7.2.1 Ambient Air 7-1
7.2.1.1 Total Airborne Lead Concentrations 7-3
7.2.1.2 Compliance with the 1978 Air Quality Standard 7-8
7.2.1.3 Changes in Air Lead Prior to Human Uptake 7-20
vi
5. 1 HISTORICAL PERSPECTIVE
5.2 NATURAL SOURCES
5. 3 MANMADE SOURCES
5. 3. 1 Production
5.3.2 Utilization
5.3.3 Emissions
5.3.3.1 Mobile Sources
5.3.3.2 Stationary Sources
5.4 SUMMARY
5.5 REFERENCES
6 TRANSPORT AND TRANSFORMATION
6 1 INTRODUCTION
6 2 TRANSPORT OF LEAD IN AIR BY DISPERSION
621 Fluid Mechanics of Dispersion ...............
6.2.2 Influence of Dispersion on Ambient Lead Conce
6221 Confined and Roadway Situations
6.2.2.2 Dispersion of Lead on an Urban Scale
6.2.2.3 Dispersion from Smelter and Refinery
6.2.2.4 Dispersion to Regional and Remote Lo
6 3 TRANSFORMATION OF LEAD IN AIR
63 1 Partirlp Size Distribution
6*} 9 firnanir fVannV* Pha<£^ 1 P^ri in Air
6.3.3 Chemical Transformations of Inorganic Lead in
6 4 RFMOVAI OF 1 FAD FROM THF ATMOSPHERE
6411 Mprhani <;m<; nf rirv rienosition
6412 Dry deposition models
6413 Calculation of drv deoosition
6.4.1.4 Field measurements of dry deposition
surrogate natural surfaces
6.4 2 Wet Deposition
6.5 TRANSFORMATION AND TRANSPORT IN OTHER ENVIRONMENTAL
651 Soil
6.5.2 Water
6.5.2 1 Inorganic
6.5.2.2 Organic . .
6. 5. 3 Vegetation Surfaces
6.6 SUMMARY
6.7 REFERENCES
ntrations
Locations
cations
Air
on
MEDIA
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TABLE OF CONTENTS (continued).
7.2.2 Lead in Soil 7-26
7.2.2.1 Typical Concentrations of Lead in Soil 7-28
7.2.2.2 Pathways of Soil Lead to Human Consumption 7-32
7.2.3 Lead in Surface and Ground Water 7-36
7.2.3.1 Typical Concentrations of Lead in Untreated Water 7-36
7.2.3.2 Human Consumption of Lead in Water 7-37
7.2.4 Summary of Environmental Concentrations of Lead 7-39
7.3 POTENTIAL PATHWAYS TO HUMAN EXPOSURE 7-40
7.3.1 Baseline Human Exposure 7-41
7.3.1.1 Lead in Inhaled Air 7-43
7.3.1.2 Lead in Food 7-44
7.3.1.3 Lead in Drinking Water 7-52
7.3.1.4 Lead in Dusts 7-54
7.3.1.5 Summary of Baseline Human Exposure to Lead 7-58
7.3.2 Additive Exposure Factors 7-58
7.3.2.1 Special Living and Working Environments 7-58
7.3.2.2 Additive Exposures Due to Age, Sex, or Socioeconomic
Status 7-68
7.3.2.3 Special Habits or Activities 7-68
7.3.3 Summary of Additive Exposure Factors 7-70
7.4 SUMMARY 7-71
7.5 REFERENCES 7-73
7A. APPENDIX: SUPPLEMENTAL AIR MONITORING INFORMATION 7A-1
7A.1 Airborne Lead Size Distribution 7A-1
7B. APPENDIX: SUPPLEMENTAL SOIL AND DUST INFORMATION 7B-1
7C. APPENDIX: STUDIES OF SPECIFIC POINT SOURCES OF LEAD 7C-1
7C.1 Smelters and Mines 7C-1
7C.1.1 Two Smelter Study 7C-1
7C.1.2 British Columbia, Canada 7C-2
7C.1.3 Netherlands 7C-2
7C.1.4 Belgium * 7C-2
7C.1.5 Meza River Valley, Yugoslavia 7C-5
7C.1.6 Kosova Province, Yugoslavia 7C-6
7C.1.7 Czechoslovakia 7C-6
7C.1.8 Australi a 7C-6
7C.2 BATTERY FACTORIES 7C-6
7C.2.1 Southern Vermont 7C-6
7C.2.2 North Carolina 7C-9
7C.2.3 Oklahoma 7C-9
7C.2.4 Oakland, CA 7C-10
7C.2.5 Manchester, England 7C-10
7D APPENDIX: SUPPLEMENTAL DIETARY INFORMATION FROM THE U.S. FDA TOTAL DIET STUDY .. 7D-1
7E. REFERENCES 7^1
vii
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TABLE OF CONTENTS (continued).
8. EFFECTS OF LEAD ON ECOSYSTEMS 8_-|
8.1 INTRODUCTION '.'.'.'.'.'.'.'.'.'.'.'. 8-1
8.1.1 Scope of Chapter 8 8-1
8.1.1.1 Plants '/.''. 8-3
8.1.1.2 Animals ' a-3
8.1.1.3 Microorganisms 8-4
8.1.1.4 Ecosystems 8-4
8.1.2 Ecosystem Functions 8-4
8.1.2.1 Types of Ecosystems 8-4
8.1.2.2 Energy Flow and Biogeochemical Cycles 8-5
8.1.2.3 Biogeochemistry of Lead 8-6
8.1.3 Criteria for Evaluating Ecosystem Effects 8-8
8.2 LEAD IN SOILS AND SEDIMENTS 8-12
8.2.1 Distribution of Lead in Soils 8-12
8.2.2 Origin and Availability of Lead in Aquatic Sediments 8-14
8.3 EFFECTS OF LEAD ON PLANTS 8-15
8.3.1 Effects on Vascular Plants and Algae 8-15
8.3.1.1 Uptake by Plants 8-15
8.3.1.2 Physiological Effects on Plants 8-19
8.3.1.3 Lead Tolerance in Vascular Plants 8-23
8. 3.1.4 Effects of Lead on Forage Crops 8-24
8.3.1.5 Effects on Algae 8-24
8.3.1.6 Summary of Plant Effects 8-25
8.3.2 Effects on Bacteria and Fungi 8-25
8.3.2.1 Effects on Decomposers 8-25
8.3.2.2 Effects on Nitrifying Bacteria 8-28
8.3.2.3 Methylation by Aquatic Microorganisms 8-29
8.3.2.4 Summary of Effects on Microorganisms 8-29
8.4 EFFECTS OF LEAD ON DOMESTIC AND WILD ANIMALS 8-29
8.4.1 Vertebrates 8-29
8.4.1.1 Terrestrial Vertebrates 8-29
8.4.1.2 Effects on Aquatic Vertebrates 8-35
8.4.2 Invertebrates 8-36
8.4.3 Summary of Effects on Animals 8-40
8.5 EFFECTS OF LEAD ON ECOSYSTEMS 8-40
8.5.1 Delayed Decomposition 8-41
8.5.2 Circumvention of Calcium Biopurification 8-42
8.5.3 Population Shifts Toward Lead Tolerant Populations 8-44
8.5.4 Biogeochemical Distribution of Lead in Ecosystems 8-44
8.6 SUMMARY . 8-46
8.7 REFERENCES "\[[ 8-48
vm
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LIST OF FIGURES
Figure Page
3-1 Metal complexes of lead 3-6
3-2 Softness parameters of metals 3-6
3-3 Structure of chelating agents 3-7
4-1 Acceptable zone for siting TSP monitors 4-5
5-1 Chronological record of the relative increase of lead in snow strata, pond
and lake sediments, marine sediments, and tree rings 5-2
5-2 The global lead production has changed historically 5-4
5-3 Location of major lead operations in the United States 5-9
5-4 Estimated lead-only emissions distribution per gallon of combusted fuel 5-14
5-5 Trend in lead content of U.S. gasolines, 1975-1984 5-15
5-6 Trend in U.S. gasoline sales, 1975-1984 5-17
5-7 Lead consumed in gasoline and ambient lead concentrations, 1975-1984 5-18
6-1 Horizontal and vertical distributions of lead 6-7
6-2 Spatial distribution of surface street and freeway traffic in
the Los Angeles Basin (103 VMT/day) for 1979 6-9
6-3 Annual average suspended lead concentrations for 1969 in the
Los Angeles Basin, calculated from the model of Cass (1975) 6-10
6-4 Profile of lead concentrations in the northeast Pacific 6-13
6-5 Lead concentration profiles in the oceans 6-13
6-6 Lead concentration profile in snow strata of northern Greenland 6-15
6-7 Airborne mass size distributions for ambient and vehicle aerosol lead 6-17
6-8 Predicted relationship between particle size and deposition velocity at
various conditions of atmospheric stability and roughness height 6-24
6-9 Variation of lead saturation capacity with cation exchange
capacity in soil at selected pH values 6-33
6-10 Lead distribution between filtrate and suspended solids in
stream water from urban and rural compartments 6-36
7-1 Principle pathways of lead from the environment to human consumption 7-2
7-2 Percent of urban stations reporting indicated concentration interval 7-6
7-3 Seasonal patterns and trends quarterly average urban lead concentrations 7-7
7-4 Comparison of trends in maximum quarterly average lead concentrations at
36 sites, 1975-1984 7-9
7-5 Airborne mass size distributions for lead taken from the literature 7-22
7-6 Decrease with distance in soil lead concentrations adjacent to a highway 7-30
7-7 Paint pigments and solder are two additional sources of potential lead
exposure which are not of atmospheric origin 7-42
7-8 Change in drinking water lead concentration in a house with lead
plumbing for the first use of water in the morning. Flushing rate
was 10 liters/minute 7-53
7C-1 Concentrations of lead in air, in dust, and on children's hands, measured
during the third population survey. Values obtained less than 1 km from the
smelter, at 2.5 km from the smelters, and in two control areas are shown 7C-4
7C-2 Schematic plan of lead mine and smelter from Mexa Valley, Yugoslavia study ... 7C-7
8-1 The major components of an ecosystem are the primary producers,
grazers, and decomposers 8-7
IX
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LIST OF FIGURES (continued).
Figure Page
8-2 The ecological success of a population depends in part on the
availability of all nutrients at some optimum concentration 8-10
8-3 This figure attempts to reconstruct the right portion of a
tolerance curve 8-11
8-4 Within the decomposer food chain, detritus is progressively
broken down i n a sequence of steps 8-28
8-5 The atomic ratios Sr/Ca, Ba/Ca and Pb/Ca (0) normally
decrease by several 8-43
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LIST OF TABLES
Table Page
3-1 Properties of elemental lead 3-2
3A-1 Physical properties of inorganic lead compounds 3A-1
3A-2 Temperature at which selected 1ead compounds 3A-3
4-1 Design of national air monitoring stations 4-3
4-2 TSP NAMS criteria 4-4
4-3 Description of spatial scales of representativeness 4-7
4-4 Relationship between monitoring objectives and
appropriate spatial scales 4-7
5-1 U.S. utilization of lead by product category 5-7
5-2 Estimated atmospheric lead emissions for the U.S., 1981, and the world 5-8
5-3 Light-duty vehicular particulate emissions 5-11
5-4 Heavy-duty vehicular particulate emissions 5-11
5-5 Recent and projected consumption of gasoline lead 5-12
6-1 Summary of microscale concentrations 6-5
6-2 Enrichment of atmospheric aerosols over crustal abundance 6-14
6-3 Distribution of lead in two size fractions at several sites
in the United States 6-18
6-4 Summary of surrogate and vegetation surface deposition of lead 6-26
6-5 Annual and seasonal deposition of lead at the Walker Branch Watershed,
1976-77 6-28
6-6 Estimated global deposition of atmospheric lead 6-28
7-1 Atmospheric lead in urban, rural and remote areas of the world 7-4
7-2 Air lead concentrations in major metropolitan areas 7-10
7-3 Stations with air lead concentrations greater than 1.0 ug/m3 7-13
7-4 Distribution of air lead concentrations by type of site 7-21
7-5 Vertical distribution of lead concentrations 7-24
7-6 Comparison of indoor and outdoor airborne lead concentrations 7-27
7-7 Summary of soil lead concentrations 7-32
7-8 Background lead in basic food crops and meats 7-33
7-9 Summary of lead in drinking water supplies 7-39
7-10 Summary of environmental concentrations of lead 7-39
7-11 Summary of inhaled ai r 1 ead exposure 7-43
7-12 Addition of lead to food products 7-46
7-13 Prehistoric and modern concentrations in human food from a marine food
chain 7-47
7-14 Recent trends of lead concentrations in food items 7-49
7-15 Total consumption, by age and sex, of food and beverages 7-50
7-16 Total consumption, by age and sex, of lead in food and beverages 7-51
7-17 Summary by source of lead consumed in food and beverages 7-52
7-18 Current baseline estimates of potential human exposure to dusts 7-57
7-19 Summary of baseline human exposures to lead 7-59
7-20 Summary of potential additive exposures to lead 7-62
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LIST OF TABLES
Table
7A-1 Information associated with the airborne lead size distributions of
Figure 7-5 lk-2
7B-1 Lead dust on and near heavily traveled roadways 7B-2
7B-2 Lead concentrations in street dust in Lancaster, England 7B-2
7B-3 Lead dust in residential areas 7B-3
7B-4 Airborne lead concentrations based on personal samplers 7B-3
7C-1 Lead concentrations in indoor and outdoor air 7C-3
7C-2 Airborne concentrations of lead during five population surveys 7C-5
7C-3 Atmospheric lead concentrations (24-hour) in the Meza Valley, Yugoslavia 7C-8
7C-4 Concentrations of total airborne dust ... Czechoslovakia 7C-8
7C-5 Lead concentrations in soil at ... Oakland, CA 7C-10
7D-1 Food list and preliminary lead concentrations 7D-2
7D-2 Scheme for condensation of 201 categories ... into 9 categories 70-8
8-1 Estimated natural levels of lead in ecosystem 8-12
8-2 Estimates of the degree of contamination of herbivores,
omnivores, and carnivores 8~33
xii
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LIST OF ABBREVIATIONS
AAS
Ach
ACTH
ADCC
ADP/0 ratio
AIDS
AIHA
All
ALA
ALA-D
ALA-S
ALA-U
APDC
APHA
ASTM
ASV
ATP
B-cells
Ba
BAL
BAP
BSA
BUN
BW
C.V.
CaBP
CaEDTA
CaNa,EDTA
CBO i
Cd
CDC
CEC
CEH
CFR
CMP
CNS
CO
COHb
CP-U
cBah
D.F.
DA
6-ALA
DCMU
DPP
DNA
DTH
EEC
EEG
EMC
EP
Atomic absorption spectrometry
Acetylcholine
Adrenocorticotrophic hormone
Antibody-dependent cell-mediated cytotoxicity
Adenosine diphosphate/oxygen ratio
Acquired immune deficiency syndrome
American Industrial Hygiene Association
Angiotensin II
Aminolevulinic acid
Aminolevulinic acid dehydrase
Aminolevulinic acid synthetase
Aminolevulinic acid in urine
Ammonium pyrrolidine-dithiocarbamate
American Public Health Association
Amercian Society for Testing and Materials
Anodic stripping voltammetry
Adenosine triphosphate
Bone marrow-derived lymphocytes
Barium
British anti-Lewi site (AKA dimercaprol)
benzo(a)pyrene
Bovine serum albumin
Blood serum urea nitrogen
Body weight
Coefficient of variation
Calcium binding protein
Calcium ethylenediaminetetraacetate
Calcium sodium ethylenediaminetetraacetate
Central business district
Cadmium
Centers for Disease Control
Cation exchange capacity
Center for Environmental Health
reference method
Cytidine monophosphate
Central nervous system
Carbon monoxide
Carboxyhemoglobin
Urinary coproporphyrin
plasma clearance of p-aminohippuric acid
Copper
Degrees of freedom
Dopamine
delta-aminolevulinic acid
[3-(3,4-dichlorophenyl)-l,l-dimethylurea
Differential pulse polarography
Deoxyribonucleic acid
Delayed-type hypersensitivity
European Economic Community
Electroencephalogram
Encephalomyocardi ti s
Erythrocyte protoporphyrin
xi i i
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LIST OF ABBREVIATIONS (continued).
EPA U.S. Environmental Protection Agency
FA Fulvic acid
FDA Food and Drug Administration
Fe Iron
FEP Free erythrocyte protoporphyrin
FY Fiscal year
G.M. Grand mean
G-6-PD Glucose-6-phosphate dehydrogenase
GABA Gamma-aminobutyric acid
GALT Gut-associated lymphoid tissue
GC Gas chromatography
GFR Glomerular filtration rate
HA Humic acid
Hg Mercury
hi-vol High-volume air sampler
HPLC High-performance liquid chromatography
i.m. Intramuscular (method of injection)
i.p. Intraperitoneally (method of injection)
i.v. Intravenously (method of injection)
IAA Indol-3-ylacetic acid
IARC International Agency for Research on Cancer
ICD International classification of diseases
ICP Inductively coupled plasma emission spectroscopy
IDMS Isotope dilution mass spectrometry
IF Interferon
ILE Isotopic Lead Experiment (Italy)
IRPC International Radiological Protection Commission
K Potassium
LDH-X Lactate dehydrogenase isoenzyme x
LCgQ Lethyl concentration (50 percent)
LD5Q Lethal dose (50 percent)
LH Luteinizing hormone
LIPO Laboratory Improvement Program Office
Natural logarithm
Lipopolysaccharide
Long range transport
n>RNA Messenger ribonucleic acid
ME Mercaptoethanol
MEPP Miniature end-plate potential
MES Maximal electroshock seizure
MeV Mega-electron volts
MLC Mixed lymphocyte culture
MMD Mass median diameter
MMAD Mass median aerodynamic diameter
Mn Manganese
MND Motor neuron disease
MSV Moloney sarcoma virus
MTD Maximum tolerated dose
n Number of subjects or observations
N/A Not Available
xiv
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LIST OF ABBREVIATIONS (continued).
NA
NAAQS
NAD
NADB
NAMS
NAS
NASN
NBS
NE
NFAN
NFR-82
NHANES II
Ni
NTA
OSHA
P
P
PAH
Pb
PBA
Pb(Ac)?
PbB
PbBrCl
PBG
PFC
pH
PHA
PHZ
PIXE
PMN
PND
PNS
P.O.
ppm
PRA
PRS
PWM
Py-5-N
RBC
RBF
RCR
redox
RES
RLV
RNA
S-HT
SA-7
S.C.
son
S.D.
SOS
S.E.M.
Not Applicable
National ambient air quality standards
Nicotinamide Adenine Dinucleotide
National Aerometric Data Bank
National Air Monitoring Station
National Academy of Sciences
National Air Surveillance Network
National Bureau of Standards
Norepinephrine
National Filter Analysis Network
Nutrition Foundation Report of 1982
National Health Assessment and Nutritional Evaluation Survey II
Nickel
Ni tri1otri acetoni tri1e
Occupational Safety and Health Administration
Phosphorus
Significance symbol
Para-aminohippuric acid
Lead
Air lead
Lead acetate
concentration of lead in blood
Lead (II) bromochloride
Porphobilinogen
Plaque-forming cells
Measure of acidity
Phytohemagglutinin
Polyacrylamide-hydrous-zirconia
Proton-induced X-ray emissions
Polymorphonuclear leukocytes
Post-natal day
Peripheral nervous system
Per os (orally)
Parts per million
Plasma renin activity
Plasma renin substrate
Pokeweed mitogen
Pyrimi de-5'-nucleoti dase
Red blood cell; erythrocyte
Renal blood flow
Respiratory control ratios/rates
Oxidation-reduction potential
Reticuloendothelial system
Rauscher leukemia virus
Ribonucleic acid
Serotonin
Simian adenovirus
Subcutaneously (method of injection)
Standard cubic meter
Standard deviation
Sodium dodecyl sulfate
Standard error of the mean
xv
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LIST OF ABBREVIATIONS (continued).
SES
SCOT
slg
SLAMS
SMR
Sr
SRBC
SRMs
STEL
SW voltage
T-cells
t-tests
TBL
TEA
TEL
TIBC
TML
TMLC
TSH
TSP
U.K.
UMP
USPHS
VA
WHO
XRF
T
Zn
ZPP
Socioeconomic status
Serum glutamic oxaloacetic transaminase
Surface immunoglobulin
State and local air monitoring stations
Standardized mortality ratio
Stronti urn
Sheep red blood cells
Standard reference materials
Short-term exposure limit
Slow-wave voltage
Thymus-derived lymphocytes
Tests of significance
Tri-n-butyl lead
Tetraethyl-ammonium
Tetraethyllead
Total iron binding capacity
Tetramethyllead
Tetramethyllead chloride
Thyroid-stimulating hormone
Total suspended particulate
United Kingdom
Uridine monophosphate
U.S. Public Health Service
Veterans Administration
Deposition velocity
Visual evoked response
World Health Organization
X-Ray fluorescence
Chi squared
Zinc
Erythrocyte zinc protoporphyrin
MEASUREMENT ABBREVIATIONS
dl
ft
g
g/gal
g/ha-mo
km/hr
1/nrin
ing/km
ug/m3
mm
pin
umol
ng/cm2
nm
nM
sec
t
deciliter
feet
gram
gram/gallon
gram/hectare-month
kilometer/hour
liter/minute
mi 11igram/kilometer
microgram/cubic meter
millimeter
micrometer
micromole
nanograms/square centimeter
nanometer
nanomole
second
tons
xv i
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GLOSSARY VOLUME II
A horizon of soils - the top layer of soil, immediately below the litter layer;
organically rich.
anorexia - loss of appetite.
anthropogenic - generated by the activities of man.
apoplast - extracellular portion of the root cross-section.
Brownian movement - the random movement of microscopic particles.
carnivore - meat-eating organism.
catenation - linkage between atoms of the same chemical element.
cation exchange capacity (CEC) - the ability of a matrix to selectively exchange
positively charged ions.
chemical mass balance - the input/output balance of a chemical within a defined
system.
coprophilic fungi - fungi which thrive on the biological waste products of
other organisms.
detritus - the organic remains of plants and animals.
dictyosome - a portion of the chloroplast structurally similar to a stack of
disks.
dry deposition - the transfer of atmospheric particles to surfaces by sedimen-
tation or impact!on.
ecosystem - one or more ecological communities linked by a common set of
environmental parameters.
electronegativity - a measure of the tendency of an atom to become negatively
charged.
enrichment factor - the degree to which the environmental concentration of an
element exceeds the expected (natural or crustal)
concentration.
galena - natural lead sulfide.
gravimetric - pertaining to a method of chemical analysis in which the
concentration of an element in a sample is determined by weight
(e.g., a precipitate).
herbivore - plant-eating organism.
humic substances - humic and fulvic acids in soil and surface water.
xvi i
-------
hydroponically grown plants - plants which are grown with their roots immersed
in a nutrient-containing solution instead of
soil.
Law of Tolerance - for every environmental factor there is both a minimum and
a maximum that can be tolerated by a population of plants
or animals.
leaf area index (LAI) - the effective leaf-surface (upfacing) area of a tree as
a function of the plane projected area of the tree canopy.
LCrn - concentration of an agent at which 50 percent of the exposed population
bu dies.
lithosphere - the portion of the earth's crust subject to interaction with the
atmosphere and hydrosphere.
mass median aerodynamic diameter (MMAD) - the aerodynamic diameter (in pm) at
which half the mass of particles in
an aerosol is associated with values
below and half above.
meristematic tissue - growth tissue in plants capable of differentiating into
any of several cell types.
microcosm - a small, artificially controlled ecosystem.
mycorrhizal fungi - fungi symbiotic with the root tissue of plants.
NAOP - National Atmospheric Deposition Program.
photolysis - decomposition of molecules into simpler units by the application
of light.
photosytem I light reaction - the light reaction of photosystem converts light
to chemical energy (ATP and reduced NADP).
Photosystem I of the light reaction receives ex-
cited electrons from photosystem II, increases
their energy by the absorption of light, and
passes these excited electrons to redox
substances that eventually produce reduced
NADP.
primary producers - plants and other organisms capable of transforming carbon
dioxide and light or chemical energy into organic compounds.
promotional energy - the energy required to move an atom from one valence
state to another.
saprotrophs - heterotrophic organisms that feed primarily on dead organic
material.
stoichiometry - calculation of the quantities of substances that enter into
and are produced by chemical reactions.
xviii
-------
stratospheric transfer - in the context of this document, transfer from the
troposphere to the stratosphere.
symplast - intracellular portion of the root cross-section.
troposphere - the lowest portion of the atmosphere, bounded on the upper level
by the stratosphere.
wet deposition - the transfer of atmospheric particles to surfaces by precipi-
tation, e.g., rain or snow.
xix
-------
AUTHORS, CONTRIBUTORS, AND REVIEWERS
Chapter 3: Physical and Chemical Properties of Lead
Principal Author
Dr. Derek Hodgson
Department of Chemistry
University of North Carolina
Chapel Hill, NC 27514
The following persons reviewed this chapter at EPA's request:
Dr. Clarence A. Hall
Air Conservation Division
Ethyl Corporation
1600 West 8-Mile Road
Ferndale, MI 48220
Dr. David E. Koeppe
Department of Plant and Soil Science
Texas Technical University
Lubbock, TX 79409
Dr. Samuel Lestz
Department of Mechanical Engineering
Pennsylvania State University
University Park, PA 16802
Dr. Ben Y. H. Liu
Department of Mechanical Engineering
University of Minnesota
Minneapolis, MN 55455
Dr. Michael Oppenheimer
Environmental Defense Fund
444 Park Avenue, S.
New York, NY 10016
Dr. William R. Pierson
Research Staff
Ford Motor Company
P.O. Box 2053
Dearborn, MI 48121
Dr. Gary Rolfe
Department of Forestry
University of Illinois
Urbana, IL 61801
Dr. Glen Sanderson
University of Illinois
Illinois Natural History Survey
Urbana, IL 61801
Dr. Rodney K. Skogerboe
Department of Chemistry
Colorado State University
Fort Collins, CO 80521
Dr. William H. Smith
Greeley Memorial Laboratory
and Environmental Studies
Yale University, School of
Forestry
New Haven, CT 06511
Dr. Gary Ter Haar
Toxicology and Industrial Hygiene
Ethyl Corporation
Baton Rouge, LA 70801
Dr. James Wedding
Engineering Research Center
Colorado State University
Fort Collins, CO 80523
XX
-------
Chapter 4: Sampling and Analytical Methods for Environmental Lead
Principal Authors
Dr. James Wedding
Engineering Research Center
Colorado State University
Fort Collins, CO 80521
Dr. Rodney K. Skogerboe
Department of Chemistry
Colorado State University
Fort Collins, CO 80521
Contributing Author
Dr. Robert Bruce
Environmental Criteria and Assessment Office
MD-52
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
The following persons reviewed this chapter at EPA's request:
Dr. John B. Clements
Environmental Monitoring Systems Laboratory
MD-78
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. Tom Dzubay
Inorganic Pollutant Analysis Branch
MD-47
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. Clarence A. Hall
Air Conservation Division
Ethyl Corporation
1600 West 8-Mile Road
Ferndale, MI 48220
Dr. Derek Hodgson
Department of Chemistry
University of North Carolina
Chapel Hill, NC 27514
Dr. Bill Hunt
Monitoring and Data Analysis Division
MD-14
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. David E. Koeppe
Department of Plant and Soil Science
Texas Technical University
Lubbock, TX 79409
Dr. Samuel Lestz
Department of Mechanical
Engineering
Pennsylvania State University
University Park, PA 16802
Dr. Ben Y. H. Liu
Department of Mechanical
Engineering
University of Minnesota
Minneapolis, MN 55455
Dr. Michael Oppenheimer
Environmental Defense Fund
444 Park Avenue, S.
New York, NY 10016
Dr. William R. Pierson
Research Staff
Ford Motor Company
P.O. Box 2053
Dearborn, MI 48121
Dr. Gary Rolfe
Department of Forestry
University of Illinois
Urbana, IL 61801
Dr. Glen Sanderson
University of Illinois
Illinois Natural History Survey
Urbana, IL 61801
xxi
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Mr. Stan Sleva
Office of Air Quality Planning and Standards
MD-14
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. William H. Smith
Greeley Memorial Laboratory
and Environmental Studies
Yale University, School of Forestry
New Haven, CT 06511
Dr. Robert Stevens
Inorganic Pollutant Analysis Branch
MD-47
U.S. Environmental Protection
Agency
Research Triangle Park, NC 27711
Dr. Gary Ter Haar
Toxicology and Industrial Hygiene
Ethyl Corporation
451 Florida Boulevard
Baton Rouge, LA 70801
Chapter 5: Sources and Emissions
Principal Author
Dr. James Braddock
Mobile Source Emissions Research Branch
MD-46
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Contributing Author
Dr. Tom McMullen
Environmental Criteria and Assessment Office
MD-52
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. Robert Elias
Environmental Criteria and
Assessment Office
MD-52
U.S. Environmental Protection
Agency
Research Triangle Park, NC 27711
The following persons reviewed this chapter at EPA's request:
Dr. Clarence A. Hall
Air Conservation Division
Ethyl Corporation
1600 West 8-Mile Road
Ferndale, MI 48220
Dr. Derek Hodgson
Department of Chemistry
University of North Carolina
Chapel Hill, NC 27514
Dr. David E. Koeppe
Department of Plant and Soil Science
Texas Technical University
Lubbock, TX 79409
Dr. Samuel Lestz
Department of Mechanical Engineering
Pennsylvania State University
University Park, PA 16802
xxii
Dr. William R. Pierson
Research Staff
Ford Motor Company
P.O. Box 2053
Dearborn, MI 48121
Dr. Gary Rolfe
Department of Forestry
University of Illinois
Urbana, IL 61801
Dr. Glen Sanderson
University of Illinois
Illinois Natural History Survey
Urbana, IL 61801
Dr. Rodney K. Skogerboe
Department of Chemistry
Colorado State University
Fort Collins, CO 80521
-------
Dr. Ben Y. H. Liu
Department of Mechanical Engineering
University of Minnesota
Minneapolis, MN 55455
Dr. Michael Oppenheimer
Environmental Defense Fund
444 Park Avenue, S.
New York, NY 10016
Dr. James Wedding
Engineering Research Center
Colorado State University
Fort Collins, CO 80523
Dr. William H. Smith
Greeley Memorial Laboratory
and Environmental Studies
Uale University, School of Forestry
New Haven, CT 06511
Dr. Gary Ter Haar
Toxicology and Industrial Hygiene
Ethyl Corporation
451 Florida Boulevard
Baton Rouge, LA 70801
Chapter 6: Transport and Transformation
Principal Author
Dr. Ron Bradow
Mobile Source Emissions Research Branch
MD-46
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Contributing Authors
Dr. Robert Eli as
Environmental Criteria and Assessment Office
MD-52
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. Rodney Skogerboe
Department of Chemistry
Colorado State University
Fort Collins, CO 80521
The following persons reviewed this chapter at EPA's request:
Dr. Clarence A. Hall
Air Conservation Division
Ethyl Corporation
1600 West 8-Mile Road
Ferndale, MI 48220
Dr. Derek Hodgson
Department of Chemistry
University of North Carolina
Chapel Hill, NC 27514
Dr. David E. Koeppe
Department of Plant and Soil Science
Texas Technical University
Lubbock, TX 79409
Dr. William R. Pierson
Research Staff
Ford Motor Company
P.O. Box 2053
Dearborn, MI 48121
Dr. Gary Rolfe
Department of Forestry
University of Illinois
Urbana, IL 61801
Dr. Glen Sanderson
Illinois Natural History Survey
University of Illinois
Urbana, IL 61801
xxm
-------
Or. Samuel Lestz
Department of Mechanical Engineering
Pennsylvania State University
University Park, PA 16802
Dr. Ben Y. H. Liu
Department of Mechanical Engineering
University of Minnesota
Minneapolis, MN 55455
Dr. Michael Oppenheimer
Environmental Defense Fund
444 Park Avenue, S.
New York, NY 10016
Dr. William H. Smith
Greeley Memorial Laboratory
and Environmental Studies
Yale University, School of
Forestry
New Haven, CT 06511
Dr. Gary Ter Haar
Toxicology and Industrial Hygiene
Ethyl Corporation
451 Florida Boulevard
Baton Rouge, LA 70801
Dr. James Wedding
Engineering Research Center
Colorado State University
Fort Collins, CO 80523
Chapter 7: Environmental Concentrations and Potential Pathways to Human
Exposure
Principal Authors
Dr. Cliff Davidson
Department of Civil Engineering
Carnegie-Mellon University
Schenley Park
Pittsburgh, PA 15213
Dr. Robert Eli as
Environmental Criteria and
Assessment Office
MD-52
U.S. Environmental Protection
Agency
Research Triangle Park, NC 27711
The following persons reviewed this chapter at EPA's request:
Dr. Carol Angle
Department of Pediatrics
University of Nebraska
College of Medicine
Omaha, NE 68105
Dr. Lee Annest
Division of Health Examin. Statistics
National Center for Health Statistics
3700 East-West Highway
Hyattsville, MD 20782
Dr. Donald Barltrop
Department of Child Health
Westminister Children's Hospital
London SW1P 2NS
England
Dr. A. C. Chamberlain
Environmental and Medical
Sciences Division
Atomic Energy Research
Establishment
Harwell 0X11
England
Dr. Neil Chernoff
Division of Developmental Biology
MD-67
U.S. Environmental Protection
Agency
Research Triangle Park, NC 27711
Dr. Julian Chisolm
Baltimore City Hospital
4940 Eastern Avenue
Baltimore, MD 21224
xxiv
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Dr. Irv Billick
Gas Research Institute
8600 West Bryn Mawr Avenue
Chicago, IL 60631
Dr. Joe Boone
Clinical Chemistry and
Toxicology Section
Centers for Disease Control
Atlanta, GA 30333
Dr. Robert Bornschein
University of Cincinnati
Kettering Laboratory
Cincinnati, OH 45267
Dr. Jack Dean
Immunobiology Program and
Immunotoxicology/Cel1 Biology program
CUT
P.O. Box 12137
Research Triangle Park, NC 27709
Dr. Fred deSerres
Associate Director for Genetics
NIEHS
P.O. Box 12233
Research Triangle Park, NC 27709
Dr. Robert Dixon
Laboratory of Reproductive and
Developmental Toxicology
NIEHS
P.O. Box 12233
Research Triangle Park, NC 27709
Dr. Claire Ernhart
Department of Psychiatry
Cleveland Metropolitan General Hospital
Cleveland, OH 44109
Dr. Sergio Fachetti
Section Head - Isotope Analysis
Chemistry Division
Joint Research Center
121020 Ispra
Varese, Italy
Dr. Virgil Ferm
Department of Anatomy and Cytology
Dartmouth Medical School
Hanover, NH 03755
Mr. Jerry Cole
International Lead-Zinc Research
Organization
292 Madison Avenue
New York, NY 10017
Dr. Max Costa
Department of Pharmacology
University of Texas Medical
School
Houston, TX 77025
Dr. Anita Curran
Commissioner of Health
Westchester County
White Plains, NY 10607
Dr. Warren Galke
Department of Biostatisties
and Epidemiology
School of Allied Health
East Carolina University
Greenville, NC 27834
Mr. Eric Goldstein
Natural Resources Defense
Council, Inc.
122 E. 42nd Street
New York, NY 10168
Dr. Harvey Gonick
1033 Gayley Avenue
Suite 116
Los Angeles, CA 90024
Dr. Robert Goyer
Deputy Director
NIEHS
P.O. Box 12233
Research Triangle Park, NC 27709
Dr. Stanley Gross
Hazard Evaluation Division
Toxicology Branch
U.S. Environmental Protection
Agency
Washington, DC 20460
Dr. Paul Hammond
University of Cincinnati
Kettering Laboratory
Cincinnati, OH 45267
xxv
-------
Dr. Alf Fischbein
Environmental Sciences Laboratory
Mt. Sinai School of Medicine
New York, NY 10029
Dr. Jack Fowle
Reproductive Effects Assessment Group
U.S. Environmental Protection Agency
RD-689
Washington, DC 20460
Dr. Bruce Fowler
Laboratory of Pharmacology
NIEHS
P.O. Box 12233
Research Triangle Park, NC 27709
Dr. Kristal Kostial
Institute for Medical Research
and Occupational Health
Yu-4100 Zagreb
Yugoslavia
Dr. Lawrence Kupper
Department of Biostatisties
UNC School of Public Health
Chapel Hill, NC 27514
Dr. Phillip Landrigan
Division of Surveillance,
Hazard Evaluation and Field Studies
Taft Laboratories - NIOSH
Cincinnati, OH 45226
Dr. David Lawrence
Microbiology and Immunology Dept.
Albany Medical College of Union
University
Albany, NY 12208
Dr. Jane Lin-Fu
Office of Maternal and Child Health
Department of Health and Human Services
Rockville, MD 20857
Dr. Don Lynam
Air Conservation
Ethyl Corporation
451 Florida Boulevard
Baton Rouge, LA 70801
Dr. Ronald D. Hood
Department of Biology
The University of Alabama
University, AL 35486
Dr. V. Houk
Centers for Disease Control
1600 Clifton Road, NE
Atlanta, GA 30333
Dr. Loren D. Koller
School of Veterinary Medicine
University of Idaho
Moscow, ID 83843
Dr. Chuck Nauman
Exposure Assessment Group
U.S. Environmental Protection
Agency
Washington, DC 20460
Dr. Herbert L. Needleman
Children's Hospital of Pittsburgh
Pittsburgh, PA 15213
Dr. H. Mitchell Perry
V.A. Medical Center
St. Louis, MO 63131
Dr. Jack Pierrard
E.I. duPont de Nemours and
Compancy, Inc.
Petroleum Laboratory
Wilmington, DE 19898
Dr. Sergio Piomelli
Columbia University Medical School
Division of Pediatric Hematology
and Oncology
New York, NY 10032
Dr. Magnus Piscator
Department of Environmental Hygiene
The Karolinska Institute 104 01
Stockholm
Sweden
xxvi
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Dr. Kathryn Mahaffey
Division of Nutrition
Food and Drug Administration
1090 Tusculum Avenue
Cincinnati, OH 45226
Dr. Ed McCabe
Department of Pediatrics
University of Wisconsin
Madison, WI 53706
Dr. Paul Mushak
Department of Pathology
UNC School of Medicine
Chapel Hill, NC 27514
Dr. John Rosen
Division of Pediatric Metabolism
Albert Einstein College of Medicine
Montefiore Hospital and Medical Center
111 East 210 Street
Bronx, NY 10467
Dr. Stephen R. Schroeder
Division for Disorders
of Development and Learning
Biological Sciences Research Center
University of North Carolina
Chapel Hill, NC 27514
Dr. Anna-Maria Seppalainen
Institutes of Occupational Health
Tyoterveyslaitos
Haartmaninkatu 1
00290 Helsinki 29
Finland
Dr. Ellen Silbergeld
Environmental Defense Fund
1525 18th Street, NW
Washington, DC 20036
Dr. Robert Putnam
International Lead-Zinc
Research Organization
292 Madison Avenue
New York, NY 10017
Dr. Michael Rabinovntz
Children's Hospital Medical
Center
300 Longwood Avenue
Boston, MA 02115
Or. Harry Roels
Unite de Toxicologie
Industrielle et Medicale
Universite de Louvain
Brussels, Belgium
Dr. Ron Snee
E.I. duPont Nemours and
Company, Inc.
Engineering Department L3167
Wilmington, DE 19898
Mr. Gary Ter Haar
Toxicology and Industrial
Hygiene
Ethyl Corporation
451 Florida Boulevard
Baton Rouge, LA 70801
Mr. Ian von Lindern
Department of Chemical
Engineering
University of Idaho
Moscow, ID 83843
Dr. Richard P. Wedeen
V.A. Medical Center
Tremont Avenue
East Orange, NJ 07019
Chapter 8: Effects of Lead on Ecosystems
Principal Author
Dr. Robert Eli as
Environmental Criteria and Assessment Office
MD-52
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
xxv ii
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Contributing Author
Dr. J.H.B Garner
Environmental Criteria and Assessment Office
MD-52
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
The following persons reviewed this chapter at EPA's request:
Dr. Clarence A. Hall
Air Conservation Division
Ethyl Corporation
1600 West 8-Mile Road
Ferndale, MI 48220
Dr. Derek Hodgson
Department of Chemistry
University of North Carolina
Chapel Hill, NC 27514
Dr. David E. Koeppe
Department of Plant and Soil Science
P.O. Box 4169
Texas Technical University
Lubbock, TX 79409
Dr. Samuel Lestz
Department of Mechanical Engineering
Pennsylvania State University
University Park, PA 16802
Dr. Ben Y. H. Liu
Department of Mechanical Engineering
University of Minnesota
Minneapolis, MN 55455
Dr. Michael Oppenheimer
Environmental Defense Fund
444 Park Avenue, S.
New York, NY 10016
Dr. William R. Pierson
Research Staff
Ford Motor Company
P.O. Box 2053
Dearborn, MI 48121
Or. Keturah Reinbold
Illinois Natural History Survey
Urbana, IL 61801
Dr. Gary Rolfe
Department of Forestry
University of Illinois
Urbana, IL 61801
Dr. Glen Sanderson
Illinois Natural History Survey
University of Illinois
Urbana, IL 61801
Dr. William H. Schlesinger
Department of Botany
Duke University
Durham, NC 27706
Dr. Rodney K. Skogerboe
Department of Chemistry
Colorado State University
Fort Collins, CO 80521
Dr. William H. Smith
Greeley Memorial Laboratory
and Environmental Studies
Yale University, School of
Forestry
New Haven, CT 06511
Dr. Gary Ter Haar
Toxicology and Industrial Hygiene
Ethyl Corporation
451 Florida Boulevard
Baton Rouge, LA 70801
Dr. James Wedding
Engineering Research Center
Colorado State University
Fort Collins, CO 80523
xxvm
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2. INTRODUCTION
According to Section 108 of the Clean Air Act of 1970, as amended in June 1974, a cri-
teria document for a specific pollutant or class of pollutants shall
. . . accurately reflect the latest scientific knowledge useful in indicating
the kind and extent of all identifiable effects on public health or welfare which
may be expected from the presence of such pollutant in the ambient air, in varying
quantities.
Air quality criteria are of necessity based on presently available scientific data, which
in turn reflect the sophistication of the technology used in obtaining those data as well as
the magnitude of the experimental efforts expended. Thus air quality criteria for atmospheric
pollutants are a scientific expression of current knowledge and uncertainties. Specifically,
air quality criteria are expressions of the scientific knowledge of the relationships between
various concentrations—averaged over a suitable time period—of pollutants in the same atmos-
phere and their adverse effects upon public health and the environment. Criteria are issued
to help make decisions about the need for control of a pollutant and about the development of
air quality standards governing the pollutant. Air quality criteria are descriptive; that is,
they describe the effects that have been observed to occur as a result of external exposure at
specific levels of a pollutant. In contrast, air quality standards are prescriptive; that is,
they prescribe what a political jurisdiction has determined to be the maximum permissible
exposure for a given time in a specified geographic area.
In the case of criteria for pollutants that appear in the atmosphere only in the gas
phase (and thus remain airborne), the sources, levels, and effects of exposure must be con-
sidered only as they affect the human population through inhalation of or external contact
with that pollutant. Lead, however, is found in the atmosphere primarily as inorganic parti-
culate, with only a small fraction normally occurring as vapor-phase organic lead. Conse-
quently, inhalation and contact are but two of the routes by which human populations may be
exposed to lead. Some particulate lead may remain suspended in the air and enter the human
body only by inhalation, but other lead-containing particles will be deposited on vegetation,
surface waters, dust, soil, pavements, interior and exterior surfaces of housing—in fact, on
any surface in contact with the air. Thus criteria for lead must be developed that will take
into account all principal routes of exposure of the human population.
This criteria document is a revision of the previous Air Quality Criteria Document for
Lead (EPA-600/8-77-017) published in December, 1977. This revision is mandated by the Clean
Air Act (Sections 108 and 109), as amended U.S.C. §§7408 and 7409. The criteria document sets
forth what is known about the effects of lead contamination in the environment on human health
2-1
-------
and welfare. This requires that the relationship between levels of exposure to lead, via all
routes and averaged over a suitable time period, and the biological responses to those levels
be carefully assessed. Assessment of exposure must take into consideration the temporal and
spatial distribution of lead and its various forms in the environment.
This document focuses primarily on lead as found in its various forms in the ambient
atmosphere; in order to assess its effects on human health, however, the distribution and bio-
logical availability of lead in other environmental media have been considered. The rationale
for structuring the document was based primarily on the two major questions of exposure and
response. The first portion of the document is devoted to lead in the environment--its physi-
cal and chemical properties; the monitoring of lead in various media; sources, emissions, and
concentrations of lead; and the transport and transformation of lead within environmental
media. The later chapters are devoted to discussion of biological responses and effects on
ecosystems and human health.
In order to facilitate printing and distribution of the present materials, this Draft
Final version of the revised EPA Air Quality Criteria Document for Lead is being released in
the form of four volumes. The first volume (Volume I) contains the executive summary and con-
clusions chapter (Chapter 1) for the entire document. Volume II (the present volume) contains
Chapters 2-8, which include: the introduction for the document (Chapter 2); discussions of
the above listed topics concerning lead in the environment (Chapters 3-7); and evaluation of
lead effects on ecosystems (Chapter 8). The remaining two volumes contain Chapters 9-13,
which deal with the extensive available literature relevant to assessment of health effects
associated with lead exposure. In addition to the above materials, there is appended to
Chapter 1 an addendum specifically addressing: the complex relationship between blood lead
level and blood pressure; and the effects of fetal and pediatric exposures on growth and
neurobehavioral development.
An effort has been made to limit the document to a highly critical assessment of the
scientific data base through December, 1985. The references cited do not constitute an
exhaustive bibliography of all available lead-related literature but they are thought to be
sufficient to reflect the current state of knowledge on those issues most relevant to the
review of the air quality standard for lead.
The status of control technology for lead is not discussed in this document. For infor-
mation on t,he subject, the reader is referred to appropriate control technology documentation
published by the Office of Air Quality Planning and Standards (OAQPS), EPA. The subject of
adequate margin of safety stipulated in Section 108 of the Clean Air Act also is not explicit-
ly addressed here; this topic will be considered in depth by EPA's Office of Air Quality Plan-
ning and Standards in documentation prepared as a part of the process of revising the National
Ambient Air Quality Standard for Lead.
2-2
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3. CHEMICAL AND PHYSICAL PROPERTIES
3.1 INTRODUCTION
Lead is a gray-white metal of silvery luster that, because of its easy isolation and low
melting point (327.5°C), was among the first of the metals to be placed in the service of
civilization. The Phoenicians traveled as far as Spain and England to mine lead as early as
2000 B.C. The Egyptians also used lead extensively; the British Museum contains a lead figure
found in an Egyptian temple which possibly dates from 3000 B.C. The most abundant ore is
galena, in which lead is present as the sulfide (PbS); metallic lead is readily smelted from
galena. The metal is soft, malleable, and ductile, a poor electrical conductor, and highly
impervious to corrosion. This unique combination of physical properties has led to its use in
piping and roofing, and in containers for corrosive liquids. By the time of the Roman Empire,
it was already in wide use in aqueducts and public water systems, as well as in cooking and
storage utensils. Solder, type metal, and various antifriction materials are manufactured
from alloys of lead. Metallic lead and lead dioxide are used in storage batteries, and
metallic lead is used in cable covering, plumbing and ammunition. Because of its high nuclear
cross section, the lead atom can absorb a broad range of radiation, making this element an
effective shield around X-ray equipment and nuclear reactors.
This chapter does not attempt to describe all of the properties of lead for each
environmental medium. Additional discussions of the chemical properties of lead, as they
pertain to specific media such as air and soil, may be found in chapters 6 and 8.
3.2 ELEMENTAL LEAD
In comparison with the most abundant metals in the earth's crust (aluminum and iron),
lead is a rare metal; even copper and zinc are more abundant by factors of five and eight,
respectively. Lead is, however, more abundant than the other toxic heavy metals; its
abundance in the earth's crust has been estimated (Moeller, 1952) to be as high as 160 ug/g,
although some other authors (Heslop and Jones, 1976) suggest a lower value of 20 ug/g. Either
of these estimates suggests that the abundance of lead is more than 100 times that of cadmium
or mercury, two other significant systemic metallic poisons. More important, since lead
occurs in highly concentrated ores from which it is readily separated, the availability of
lead is far greater than its natural abundance would suggest. The environmental significance
of lead is the result both of its utility and of its availability to mankind. Lead ranks
fifth among metals in tonnage consumed, after iron, copper, aluminum and zinc; it is,
therefore, produced in far larger quantities than any other toxic heavy metal (Dyrssen, 1972).
The properties of elemental lead are summarized in Table 3-1.
3-1
-------
TABLE 3-1. PROPERTIES OF ELEMENTAL LEAD
Property
Description
Atomic weight
Atomic number
Oxidation states
Density
Melting point
Boiling point
Covalent radius (tetradehral)
Ionic radii
Resistivity
207.19
82
+2, +4
11.35 g/cm3 at 20 °C
327.5 °C
1740 °C
1.44 A
1.21 A (+2), 0.78 A (+4)
_6
21.9 x 10 ohm/cm
Natural lead is a mixture of four stable isotopes: 204Pb (~1.5 percent), 206Pb (23.6
percent), 207Pb (22.6 percent), and 208Pb (52.3 percent). There is no radioactive progenitor
for 204Pb, but 206Pb, 207Pb, and 208Pb are produced by the radioactive decay of 238U, 235U,
and 232Th, respectively. There are four radioactive isotopes of lead that occur as members of
these decay series. Of these, only 210Pb is long lived, with a half-life of 22 years. The
others are 211Pb (half-life 36.1 min), 212Pb (10.64 hr), and 214Pb (26.8 min). The stable
isotopic compositions of naturally occurring lead ores are not identical, but show variations
reflecting geological evolution (Russell and Farquhar, 1960). Thus, the observed isotopic
ratios depend upon the U/Pb and Th/Pb ratios of the source from which the ore is derived and
the age of the ore deposit. The 206Pb/204Pb isotopic ratio, for example, varies from
approximately 16.5 to 21 depending on the source (Doe, 1970). The isotopic ratios in average
crustal rock reflect the continuing decay of uranium and thorium. The differences between
crustal rock and ore bodies, and between major ore bodies in various parts of the world, often
permit the identification of the source of lead in the environment.
3.3 GENERAL CHEMISTRY OF LEAD
Lead is the heaviest element in Group IVB of the periodic table; this is the group that
also contains carbon, silicon, germanium, and tin. Unlike the chemistry of carbon, however,
the inorganic chemistry of lead is dominated by the divalent (+2) oxidation state rather than
3-2
-------
the tetravalent (+4) oxidation state. This important chemical feature is a direct result of
the fact that the strengths of single bonds between the Group IV atoms and other atoms
generally decrease as the atomic number of the Group IV atom increases (Cotton and Wilkinson,
1980). Thus, the average energy of a C-H bond is 100 kcal/mole, and it is this factor that
stabilizes CH4 relative to CH2; for lead, the Pb-H energy is only approximately 50 kcal/mole
(Shaw and Allred, 1970), and this is presumably too small to compensate for the Pb(II) •*
Pb(IV) promotional energy. It is this same feature that explains the marked difference in the
tendencies to catenation shown by these elements. Though C-C bonds are present in literally
millions of compounds, lead catenation occurs only in organolead compounds. Lead does,
however, form compounds like Na4Pb9 which contain distinct polyatomic lead clusters (Britton,
1964), and Pb-Pb bonds are found in the cationic cluster [Pb60(OH)6>4 (01 in and Soderquist,
1972).
A listing of the solubilities and physical properties of the more common compounds of
lead is given in Appendix 3A (Table 3A-1) (Weast, 1982). As can be discerned from those data,
most inorganic lead salts are sparingly soluble (e.g., PbF2, PbCl2) or virtually insoluble
(PbS04, PbCr04) in water; the notable exceptions are lead nitrate, Pb(N03)2, and lead acetate,
Pb(OCOCH3)2. Inorganic lead (II) salts are, for the most part, relatively high-melting-point
solids with correspondingly low vapor pressures at room temperatures. The vapor pressures of
the most commonly encountered lead salts are also tabulated in Appendix 3A (Table 3A-2)
(Stull, 1947). The transformation of lead salts in the atmosphere is discussed in Chapter 6.
3.4 ORGANOMETALLIC CHEMISTRY OF LEAD
The properties of organolead compounds (i.e., compounds containing bonds between lead and
carbon) are entirely different from those of the inorganic compounds of lead; although a few
organolead(II) compounds, such as dicyclopentadienyllead, Pb(C5Hs)2, are known, the organic
chemistry of lead is dominated by the tetravalent (+4) oxidation state. An important property
of most organolead compounds is that they undergo photolysis when exposed to light (Rufman and
Rotenberg, 1980).
Because of their use as antiknock agents in gasoline and other fuels, the most important
organolead compounds have been the tetraalkyl compounds tetraethyllead (TEL) and
tetramethyllead (TML). As would be expected for such nonpolar compounds, TEL and TML are
insoluble in water but soluble in hydrocarbon solvents (e.g., gasoline). These two compounds
are manufactured by the reaction of the alkyl chloride with lead-sodium alloy (Shapiro and
Frey, 1968):
4NaPb + 4C2H5C1 -» (C2Hs)4Pb + 3Pb + 4NaCl (3-1)
3-3
-------
The methyl compound, TML, is also manufactured by a Grignard process involving the
electrolysis of lead pellets in methylmagnesium chloride (Shapiro and Frey, 1968):
2CH3MgCl + 2CH3C1 + Pb -»• (CH3)4Pb + 2MgCl2 (3-2)
A common type of commercial antiknock mixture contains a chemically redistributed mixture
of alky Head compounds. In the presence of Lewis acid catalysts, a mixture of TEL and TML
undergoes a redistribution reaction to produce an equilibrium mixture of the five possible
tetraalkyl1ead compounds. For example, an equimolar mixture of TEL and TML produces a product
with a composition as shown below:
Component Mol percent
(CH3)4Pb 4.6
(CH3)3Pb(C2H5) 24.8
(CH3)2Pb(C2H5)2 41.2
(CH3)Pb(C2H5)3 24.8
(C2H5)4Pb 4.6
These lead compounds are removed from internal combustion engines by a process called
lead scavenging, in which they react in the combustion chamber with halogenated hydrocarbon
additives (notably ethylene dibromide and ethylene dichloride) to form lead halides, usually
bromochlorolead(II). Mobile source emissions are discussed in detail in Section 5.3.3.2.
Several hundred other organolead compounds have been synthesized, and the properties of
many of them are reported by Shapiro and Frey (1968). The continuing importance of organolead
chemistry is demonstrated by a variety of recent publications investigating the syntheses
(Hager and Huber, 1980; Wharf et al., 1980) and structures (Barkigia et al., 1980) of
organolead complexes, and by recent patents for lead catalysts (Nishikido et al. , 1980).
3.5 FORMATION OF CHELATES AND OTHER COMPLEXES
The bonding in organometallic derivatives of lead is principally covalent rather than
ionic because of the small difference in the electronegativities of lead (1.8) and carbon
(2.6). As is the case in virtually all metal complexes, however, the bonding is of the
donor-acceptor type, in which both electrons in the bonding orbital originate from the carbon
atom.
The donor atoms in a metal complex could be almost any basic atom or molecule; the only
requirement is that a donor, usually called a ligand, must have a pair of electrons available
3-4
-------
for bond formation. In general, the metal atom occupies a central position in the complex, as
exemplified by the lead atom in tetramethyllead (Figure 3-la) which is tetrahedrally
surrounded by four methyl groups. In these simple organolead compounds, the lead is usually
present as Pb(IV), and the complexes are relatively inert. These simple ligands, which bind
to metal at only a single site, are called monodentate ligands. Some ligands, however, can
bind to the metal atom by more than one donor atom, so as to form a heterocyclic ring
structure. Rings of this general type are called chelate rings, and the donor molecules which
form them are called polydentate ligands or chelating agents. In the chemistry of lead,
chelation normally involves Pb(II), leading to kinetically quite labile (although
thermodynamically stable) octahedral complexes. A wide variety of biologically significant
chelates with ligands, such as amino acids, peptides, nucleotides and similar macromolecules,
are known. The simplest structure of this type occurs with the amino acid glycine, as
represented in Figure 3-lb for a 1:2 (metal:ligand) complex. The importance of chelating
agents in the present context is their widespread use in the treatment of lead and other metal
poisoning.
Metals are often classified according to some combination of their electronegativity,
ionic radius and formal charge (Ahrland, 1966, 1968, 1973; Basolo and Pearson, 1967; Nieboer
and Richardson, 1980; Pearson, 1963, 1968). These parameters are used to construct empirical
classification schemes of relative hardness or softness. In these schemes, "hard" metals form
strong bonds with "hard" anions and likewise "soft" metals with "soft" anions. Some metals
are borderline, having both soft and hard character. Pb(II), although borderline,
demonstrates primarily soft character (Figure 3-2) (Nieboer and Richardson, 1980). The terms
Class A and Class B may also be used to refer to hard metals and soft metals, respectively.
Since Pb(II) is a relatively soft (or class B) metal ion, it forms strong bonds to soft donor
atoms like the sulfur atoms in the cysteine residues of proteins and enzymes; it also coordin-
ates strongly with the imidazole groups of histidine residues and with the carboxyl groups of
glutamic and aspartic acid residues. In living systems, therefore, lead atoms bind to these
peptide residues in proteins, thereby preventing the proteins from carrying out their
functions by changing the tertiary structure of the protein or by blocking the substrate's
approach to the active site of the protein. As has been demonstrated in several studies
(Jones and Vaughn, 1978; Williams and Turner, 1981; Williams et al. , 1982), there is an
inverse correlation between the LDSO values of metal complexes and the chemical softness
parameter (ap) (Pearson and Mawby, 1967). Thus, for both mice and Drosophlla. soft metal ions
like lead(II) have been found to be more toxic than hard metal ions (Williams et al., 1982).
This classification of metal ions according to their toxicity has been discussed in detail by
Nieboer and Richardson (1980). Lead(II) has a higher softness parameter than either
cadmium(II) or mercury(II), so lead(II) compounds would not be expected to be as toxic as
their cadmium or mercury analogues.
3-5
-------
X
>b
/\
H3C CH3
Pb
(a)
NH2
H2O
(b)
Figure 3-1. Metal complexes of lead.
JE
X
iu
O
Z
UJ
I
DC
O
eo
0
9.U
4.5
4.0
3.5
3.0
2.5
2.0
1.5
1.0
0.5
0
• I I I I I I I I " I " I
> Au
r
• Ag W p(j _ ^
_ • ' Pb(IV)
^TC Hg>
• Ti'
_0Cu CLASS B _
— Sn> • mr .
Cd"« A»'""
F.'««C°' '* * •
— ^ •Ni1' • f) Fe' SnIIVI —
Ti" •*• Zn"
~" M"'* V' Ga>* BORDERLINE ~~
Qd»- Lu> —
Mg- „• «se'- •
Ct Ba' • • y,. A,,.
%Na Sr' •
• Be'
— Li
CLASS A
I I I I I I I I -,l ..I
6 8 10 12 14 16 20
CLASS A OR IONIC INDEX. Z'/r
Figure 3-2. Softness parameters of metals.
Source: Nieboer and Richardson (1980).
3-6
23
-------
o o
0-C-CH2 CHz-C-0- CH3 O
\-CH2-CH2-/ HS-C-CH-C^
-0-C-CH2 CHz-C-0- CH3 NH2 OH
II jl
o o
EDTA PENICILLAMINE
Figure 3-3. Structure of chelating agents.
The role of the chelating agents is to compete with the peptides for the metal by forming
stable chelate complexes that can be transported from the protein and eventually be excreted
by the body. For simple thermodynamic reasons (see Appendix 3A), chelate complexes are much
more stable than monodentate metal complexes, and it is this enhanced stability that is the
basis for their ability to compete favorably with proteins and other ligands for the metal
ions. The chelating agents most commonly used for the treatment of lead poisoning are
ethylenediaminetetraacetate ions (EDTA), D-penicillamine (Figure 3-3) and their derivatives.
EDTA is known to act as a hexadentate ligand toward metals (Lis, 1978; McCandlish et al.,
1978). X-ray diffraction studies have demonstrated that D-penicillamine is a tridentate
ligand binding through its sulfur, nitrogen and oxygen atoms to cobalt (de Meester and
Hodgson, 1977a; Helis et al., 1977), chromium (de Meester and Hodgson, 1977b), cadmium
(Freeman et al., 1976), and lead itself (Freeman et al., 1974), but both penicillamine and
other cysteine derivatives may act as bidentate ligands (Carty and Taylor, 1977; de Meester
and Hodgson, 1977c). Moreover, penicillamine binds to mercury only through its sulfur atoms
(Wong et al., 1973; Carty and Taylor, 1976).
It should be noted that both the stoichiometry and structures of metal chelates depend
upon pH, and that structures different from those manifest in solution may occur in crystals.
It will suffice to state, however, that several ligands can be found that are capable of
sufficiently strong chelation with lead present in the body under physiological conditions to
permit their use in the effective treatment of lead poisoning.
3-7
-------
3.6 REFERENCES
Ahrland, S. (1966) Factors contributing to (b)-behaviour in acceptors. Struct. Bonding
(Berlin) 1: 207-220.
Ahrland, S. (1968) Thermodynamics of complex formation between hard and soft acceptors and
donors. Struct. Bonding (Berlin) 5: 118-149.
Ahrland, S. (1973) Thermodynamics of the stepwise formation of metal-ion complexes in aqueous
solution. Struct. Bonding (Berlin) 15: 167-188.
Barkigia, K. M.; Fajer, J.; Adler, A. D.; Williams, G. J. B. (1980) Crystal and molecular
structure of (5,10,15,20-tetra-n-propylporphinato)lead(II): a "roof" porphyrin. Inorg.
Chem. 19: 2057-2061.
Basolo, F.; Pearson, R. G. (1967) Mechanisms of inorganic reactions: a study of metal com-
plexes in solution. New York, NY: John Wiley & Sons, Inc.; pp. 23-25, 113-119.
Britton, D. (1964) The structure of the Pb9" ion. Inorg. Chem. 3: 305.
Carty, A. J.; Taylor, N. J. (1976) Binding of inorganic mercury at biological sites: crystal
structures of Hg2+ complexes with sulphur amino-acids. J. Chem. Soc. Chem. Commun. (6):
214-216.
Carty, A. J.; Taylor, N. J. (1977) Binding of heavy metals at biologically important sites:
synthesis and molecular structure of aquo(bromo)-DL-penicillaminatocadmium(II) dihydrate.
Inorg. Chem. 16: 177-181. ~
Cotton, F. A.; Wilkinson, G. (1980) Advanced inorganic chemistry: a comprehensive text. 4th
ed. New York, NY: John Wiley & Sons, Inc. pp. 374-406.
de Meester, P.; Hodgson, D. J. (1977a) Model for the binding of D-penicillamine to metal ions
in living systems: synthesis and structure of L-histidinyl-D-penicillaminatocobalt(III)
monohydrate, [Co(L-his)(D-pen)] H20. J. Am. Chem. Soc. 99: 101-104.
de Meester, P.; Hodgson, D. J. (1977b) Synthesis and structural characterization of
L-histidinato-D-penicillaminatochromium (III) monohydrate. J. Chem. Soc. Dalton Trans.
(17): 1604-1607.
de Meester, P.; Hodgson, D. J. (1977c) Absence of metal interaction with sulfur in two metal
complexes of a cysteine derivative: the structural characterization of Bis(S-methyl-L-
cysteinato)cadmium(II) and Bis(S-methyl-L-cysteinato)zinc(II). J. Am. Chem. Soc. 99:
6884-6889.
Doe, B. R. (1970) Lead isotopes. New York, NY: Springer-Verlag. (Engelhardt, W.; Hahn, T.;
Roy, R.; Winchester, J. W.; Wyllie, P. J., eds. Minerals, rocks and inorganic materials:
monograph series of theoretical and experimental studies: v. 3).
Dyrssen, D. (1972) The changing chemistry of the oceans. Ambio 1: 21-25.
Freeman, H. C.; Stevens, G. N.; Taylor, I. F., Jr. (1974) Metal binding in chelation therapy:
the crystal structure of D-pemcillaminatolead(II). J. Chem. Soc. Chem. Commun. (10):
366-367.
3-8
-------
Freeman, H. C.; Huq, F.; Stevens, G. N. (1976) Metal binding by D-penicillamine: crystal
structure of D-penicillaniinatocadim'um(II) hydrate. J. Chem. Soc. Chem. Commun. (3):
90-91.
Hager, C.-D.; Huber, F. (1980) Organobleiverbindungen von MercaptocarbonsSuren [Organolead
compounds of mercaptocarboxylic acids]. Z. Naturforsch. 35b: 542-547.
Helis, H. M.; de Meester, P.; Hodgson, D. J. (1977) Binding of penicillainine to toxic metal
ions: synthesis and structure of potassium(D-penicillaminate) (L-Penicillaminato)cobalt-
ate(III) dihydrate, K[Co(D-pen)(L-pen)]-2H20. J. Am. Chem. Soc. 99: 3309-3312.
Heslop, R. B.; Jones, K. (1976) Inorganic chemistry: a guide to advanced study. New York, NY:
Elsevier Science Publishing Co.; pp. 402-403.
Jones, M. M.; Vaughn, W. K. (1978) HSAB theory and acute metal ion toxicity and detoxification
processes. J. Inorg. Nucl. Chem. 40: 2081-2088.
Lis, T. (1978) Potassium ethylenediaminetetraacetatomanganate(III) dihydrate. Acta Crystal-
logr. Sect. B 34: 1342-1344.
McCandlish, E. F. K.; Michael, T. K.; Neal, J. A.; Lingafelter, E. C.; Rose, N. J. (1978) Com-
parison of the structures and aqueous solutions of [o-phenylenediaminetetraacetato(4-)]
iSS? and tethylenedl'am1netetraacetat°C4-)] cobalt(II) ions. Inorg. Chem. 17:
X Jo<3" JL394.
Moeller, T. (1952) Inorganic chemistry: an advanced textbook. New York, NY: John Wiley & Sons,
x nc •
Nieboer, E.; Richardson, D. H. S. (1980) The replacement of the nondescript term heavy metals
by a biologically and chemically significant classification of metal ions. Environ.
Pollut. Ser. B. 1: 3-26.
Nishikido J.; Tamura, N.; Fukuoka, Y. (1980) (Asahi Chemical Industry Co. Ltd.) Ger. patent
no. 2,936,652.
Olin, A.; SBderquist, R. (1972) The crystal structure of B-[Pb60(OH)6](C104)4 H20. Acta Chem.
Scand. 26: 3505-3514.
Pearson, R. G. (1963) Hard and soft acids and bases. J. Am. Chem. Soc. 85: 3533-3539.
Pearson, R. G. (1968) Hard and soft acids and bases, HSAB, part 1: fundamental principles. J.
Chem. Educ. 45: 581-587.
Pearson, R. G.; Mawby, R. J. (1967) The nature of metal-halogen bonds. In: Gutmann, V., ed.
Halogen chemistry: volume 3. New York, NY: Academic Press, Inc.; pp. 55-84.
Rufman, N. M.; Rotenberg, Z. A. (1980) Special kinetic features of the photodecomposition of
organolead compounds at lead electrode surfaces. Sov. Electrochem. Engl. trans!. 16:
309-314.
Russell, R. D.; Farquhar, R. M. (1960) Introduction. In: Lead isotopes in geology. New York,
NY: Interscience; pp. 1-12.
3-9
-------
Shapiro, H.; Frey, F. W. (1968) The organic compounds of lead. New York, NY: John Wiley &
Sons. (Seyferth, D., ed. The chemistry of organometallic compounds: a series of mono-
graphs).
Shaw, C. F., III; Allred, A. L. (1970) Nonbonded interactions in organometallic compounds of
Group IV B. Organometallic Chem. Rev. A 5: 95-142.
Wharf, I.; Onyszchuk, M.; Miller, J. M.; Jones, T. R. B. (1980) Synthesis and spectroscopic
studies of phenyllead halide and thiocyanate adducts with hexamethylphosphoramide. J.
Organomet. Chem. 190: 417-433.
Williams, M. W.; Turner, J. E. (1981) Comments on softness parameters and metal ion toxicity.
J. Inorg. Nucl. Chem. 43: 1689-1691.
Williams, M. W.; Hoeschele, J. D.; Turner, J. E.; Jacobson, K. B.; Christie, N. T.; Paton, C.
L.; Smith, L. H.; Witschi, H. R.; Lee, E. H. (1982) Chemical softness and acute metal
toxicity in mice and Drosophila. Toxicol. Appl. Pharmacol. 63: 461-469.
Wong, Y. S.; Chieh, P. C.; Carty, A. J. (1973) Binding of methylmercury by amino-acids: X-ray
structures of DL-penicillaminatomethylmercury(II). J. Chem. Soc. Chem. Commun. (19):
741-742.
3-10
-------
3A.1 DATA TABLES
APPENDIX 3A
PHYSICAL/CHEMICAL DATA FOR LEAD COMPOUNDS
TABLE 3A-1. PHYSICAL PROPERTIES OF INORGANIC LEAD COMPOUNDS
Solubility, g/100 ml
Compound
Lead
Acetate
Azide
Bromate
Bromide
Carbonate
Carbonate,
basic
Chloride
Chlorobromide1
Chromate
Chromate ,
basic
Cyanide
Fluoride
Fluorochloride
Formate
Hydride
Hydroxide
lodate
Iodide
Nitrate
Formula
Pb
Pb(C2H302)2
Pb(N3)2
Pb(Br03)2-H20
PbBr2
PbC03
2PbC03-Pb(OH)2
PbCl2
PbClBr
PbCr04
PbCr04-PbO
Pb(CN)2
PbF2
PbFCl
Pb(CH02)2
PbH2
Pb(OH)2
Pb(I03)2
PbI2
Pb(N03)2
M.W.
207. 19
325.28
291.23
481.02
367.01
267.20
775.60
278. 10
322.56
323.18
546.37
259.23
245.19
261.64
297.23
209. 21
241. 20
557.00
461. 00
331.20
S.G.
11.35
3.25
-
5.53
6.66
6.6
6.14
5.85
-
6.12
6.63
-
B.24
7.05
4.63
-
-
6.155
6.16
4.53
M.P.
(°C)
327.5
280
expl
dl80
373
d315
d400
501
430
844
-
-
855
601
d!90
*d
d!45
d300
402
d470
Cold
water
i
44.3
0.023
1.38
0.8441
0.00011
i
0.99
0.6619
6x10" 6
i
si s
0.064
0.037
1.6
-
0.0155
0.0012
0.063
37.65
Hot
water
i
221s0
0.0970
si s
4.71100
d
i
3.34100
1.0343
i
i
s
-
0.1081
20
-
si s
0.003
0.41
127
Other
solvents
sa
s glyc
-
-
sa
sa,alk
s HN03
i al
-
sa.alk
sa,alk
s KCN
s HN03
-
i al
-
sa.alk
s HN03
s,alk
s.alk
3A-1
-------
TABLE 3A-1. (continued)
Solubility, g/100 ml
Compound
Nitrate, basic
Oxalate
Oxide
Dioxide
Oxide (red)
Phosphate
Sulfate
Sulfide
Sulfite
Thiocyanate
Formula
Pb(OH)N03
PbC204
PbO
Pb02
Pb304
Pb3(P04)2
PbS04
PbS
PbS03
Pb(SCN)2
M.W.
286.20
295.21
223.19
239.19
685.57
811.51
303.25
239.25
287.25
323.35
S.G.
5.93
5.28
9.53
9.375
9.1
7
6.2
7.5
-
3.82
M.P.
(°C)
d!80
d300
888
d290
d500
1014
1170
1114
d
d!90
Cold
water
19.4
0.00016
0.0017
i
i
1.4X10"5
0.00425
8.6xlO"5
i
0.05
Hot
water
s
-
-
i
i
1
0.0056
i
0.2
Other
solvents
sa
sa
s,alk
sa
sa
s.alk
sa
sa
s.alk
Belting point and solubility data from Corrin and Natusch (1977)
Abbreviations: a - acid; al - alcohol; alk - alkali; d - decomposes;
expl - explodes; glyc - glycol; i - insoluble; s - soluble;
si s - slightly soluble; M.W. - molecular weight;
S.G. - specific gravity; and M.P. - melting point.
Source: Weast, 1982.
3A-2
-------
TABLE 3A-2. TEMPERATURE AT WHICH SELECTED LEAD COMPOUNDS REACH
DESIGNATED VAPOR PRESSURES
Name
Lead
Lead
Lead
Lead
Lead
Lead
Lead
bromide
chloride
fluoride
iodide
oxide
sulfide
Formula
Pb
PbBr2
PbCl2
PbF2
PbI2
PbO
PbS
M.P.
(°C)
327.4
373
501
855
402
890
1114
Vapor Pressure (mm Hg)
1 mm
973°C
513
547
solid
479
943
852
(solid)
10 mm
1162°C
610
648
904
571
1085
975
(solid)
40 mm
1309°C
686
725
1003
644
1189
1048
(solid)
100 mm
142 1°C
745
784
1080
701
1265
1108
(solid)
400 mm
1630°C
856
893
1219
807
1402
1221
760 mm
1744°C
914
954
1293
872
1472
1281
Source: Stull, 1947.
3A.2. THE CHELATE EFFECT
The stability constants of chelated complexes are normally several orders of magnitude
higher than those of comparable monodentate complexes; this effect is called the chelate
effect, and is very readily explained in terms of kinetic considerations. A comparison of the
binding of a single bidentate ligand with that of two molecules of a chemically similar mono-
dentate ligand shows that, for the monodentate case, the process can be represented by the
equations:
M + B;
2t M-B
(3A-1)
M-B + B;
MB,
(3A-2)
The related expressions for the bidentate case are:
M + B-B;
* M-B-B
(3A-3)
3A-3
-------
M-B-B ^ - -M
^ D
k4
The overall equilibrium constants, therefore, are:
KI = kakc. K2 = klk3
ka
3 (3A-4)
^^ k2k4
For a given metal, M, and two ligands, B and B-B, which are chemically similar, it is
established that kt and k have similar values to each other, as do k2 and k^ and k4 and k.;
each of these pairs of terms represents chemically similar processes. The origin of the
chelate effect lies in the very large value of k3 relative to that of kc> This comes about
because k3 represents a unimolecular process, whereas kc is a bimolecular rate constant.
Consequently, K2 » Kx.
This concept can, of course, be extended to polydentate ligands; in general, the more
extensive the chelation, the more stable the metal complex. Hence, one would anticipate,
correctly, that polydentate chelating agents such as penicillamine or EDTA can form extremely
stable complexes with metal ions.
3A.3 REFERENCES
Corrin, M. L.; Natusch, 0. F. S. (1977) Physical and chemical characteristics of environmental
lead. Washington, D.C.: National Science Foundation; report no. NSF/RA 770214; pp.
7-31. Available from NTIS, Springfield, VA; PB-278278.
Stull, D. R. (1947) Vapor pressure of pure substances: organic compounds. Ind. Eng. Chem.
39: 517-540.
Weast, R. C., ed. (1982) Handbook of chemistry and physics. 63rd edition. Cleveland, OH:
The Chemical Rubber Co.
3 A-4
-------
4. SAMPLING AND ANALYTICAL METHODS FOR ENVIRONMENTAL LEAD
4.1 INTRODUCTION
Lead, like all criteria pollutants, has a designated reference method for monitoring and
analysis as required in State Implementation Plans for determining compliance with the lead
National Ambient Air Quality Standard. The reference method [C.F.R. (1982) 40:§50] uses a
high volume sampler (hi-vol) for sample collection and atomic absorption spectrometry for
analysis. Inductively coupled plasma emission spectroscopy and X-ray fluorescence are also
reference methods for analysis. These and several other analytical procedures are discussed
in this chapter. The reference method for sample collection may be revised to require col-
lection of a specific size fraction of atmospheric particles.
Airborne lead originates principally from manmade sources (about 75 - 90 percent comes
from automobile exhaust; see Section 5.3.3.1) and is transported through the atmosphere to
vegetation, soil, water, and animals. Knowledge of environmental concentrations of lead and
the extent of its movement among various media is essential to control lead pollution and
assess its effects on human populations.
The collection and analysis of environmental samples for lead require a rigorous quality
assurance program [C.F.R. (1982) 40:§58]. It is essential that the investigator recognize all
sources of contamination and use every precaution to eliminate them. Potential lead contamin-
ation occurs on the surfaces of collection containers and devices, on the hands and clothing
of the investigator, in the chemical reagents, in the laboratory atmosphere, and on the lab-
ware and tools used to prepare the sample for analysis. General procedures for controlling
this contamination of samples in trace metal analysis are described by Zief and Mitchell
(1976); specific details are given in Patterson and Settle (1976). In the following discus-
sion of methods for sampling and analysis, it is assumed that all procedures are carried out
with precise attention to contamination control.
In the following sections, the specific operation, procedure and instrumentation involved
in monitoring and analyzing environmental lead are discussed. Site selection criteria are
treated only briefly, due to the lack of verifying data. Much remains to be done in estab-
lishing valid criteria for sampler location. The various types of samples and substrates used
to collect airborne lead are described. Methods for collecting dry deposition, wet deposi-
tion, and aqueous, soil, and vegetation samples are also reviewed along with current sampling
methods specific to mobile and stationary sources. Finally, advantages and limitations of
techniques for sample preparation and analysis are discussed.
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4.2 SAMPLING
The purpose of sampling is to determine the nature and concentration of lead in the envi-
ronment. Sampling strategy is dictated by research needs. This strategy encompasses site
selection, choice of instrument used to obtain representative samples, and choice of method
used to preserve sample integrity. In the United States, sampling stations for air pollutants
have been operated since the early 1950's. These early stations were a part of the National
Air Surveillance Network (NASN), which has now become the National Filter Analysis Network
(NFAN). Two other types of networks have been established to meet specific data requirements.
State and Local Air Monitoring Stations (SLAMS) provide data from specific areas where pollu-
tant concentrations and population densities are the greatest and where monitoring of compli-
ance to standards is critical. The National Air Monitoring Station (NAMS) network is designed
to serve national monitoring needs, including assessment of national ambient trends. SLAMS
and NAMS stations are maintained by state and local agencies and the air samples are analyzed
in their laboratories. Stations in the NFAN network are maintained by state and local agen-
cies, but the samples are analyzed by laboratories in the U.S. Environmental Protection
Agency, where quality control procedures are rigorously maintained.
Data from all three networks are combined into one data base, the National Aerometric
Data Bank (NADB). These data may be individual chemical analyses of a 24-hour sampling period
arithmetically averaged over a calendar period, or chemical composites of several filter
samples used to determine a quarterly composite. Data are occasionally not available because
they do not conform to strict statistical requirements. A summary of the data from the NADB
appears in Section 7.2.1.
4.2.1 Regulatory Siting Criteria for Ambient Aerosol Samplers
In September of 1981, EPA promulgated regulations establishing ambient air monitoring and
data reporting requirements for lead [C.F.R. (1982) 40:§58] comparable to those already estab-
lished in May of 1979 for the other criteria pollutants. Whereas sampling for lead is accomp-
lished when sampling for total suspended particulates (TSP), the designs of lead and TSP moni-
toring stations must be complementary to insure compliance with the NAMS criteria for each
pollutant, as presented in Table 4-1, Table 4-2, and Figure 4-1.
In general, the criteria with respect to monitoring stations designate that there must be
at least two SLAMS sites for lead in any area which has a population greater than 500,000 and/
or any area where lead concentration currently exceeds the ambient lead standard (1.5 ug/m3)
or has exceeded it since January 1, 1974. In such areas, the SLAMS sites designated as part
of the NAMS network must include a microscale or middlescale site located near a major roadway
[£30,000 average daily traffic (ADT)], as well as a neighborhood scale site located in a
highly populated residential sector with high traffic density (£30,000 ADT).
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TABLE 4-1. DESIGN OF NATIONAL AIR MONITORING STATIONS
Category A
TSP
Pb
j
Category B
Pb
Conditions
High traffic and
population density
Major roadway
Major roadway
High traffic and
population density
Minimum number of
Spatial scale stations required
Neighborhood see Table 4-2
Microscale One
Middlescale One
Neighborhood One
Traffic
density
^30000
230000
^10000
20000
540000
S10000
20000
£40000
Required Siting of Station
Meters from Meters above
edge of roadway ground level
see Figure
5-15
15-50
15-75
15-100
> 50
> 75
>100
4-1
2-7
2-15
2-15
2-15
2-15
2-15
2-15
Source: C.F.R. (1982) 40:§58 App E.
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TABLE 4-2. TSP NAMS CRITERIA
Population Category
High -- >500,000
Medium — 100-500,000
Low — 50-100,000
Approximate
High1
6-8
4-6
2-4
number of stations
Concentration
Medium2
4-6
2-4
1-2
per area
Low3
0-2
0-2
0
:When TSP concentration exceeds by 20% Primary Ambient Air Standard of 75 ug/m3 annual
geometric mean.
2TSP concentration > Secondary Ambient Air Standard of 60 ug/m3 annual geometric mean.
3TSP concentration < Secondary Ambient Air Standard.
Source: C.F.R. (1982) 40:§58 App D.
With respect to the siting of monitors for lead and other criteria pollutants, there are
standards for elevation of the monitors above ground level, setback from roadways, and setback
from obstacles. A summary of the specific siting requirements for lead is presented in Table
4-1 and summarized below:
• Samples must be placed between 2 and 15 meters from the ground and greater than 20
meters from trees.
• Spacing of samplers from roads should vary with traffic volume; a range of 5 to
100 meters from the roadway is suggested.
• Distance from samplers to obstacles must be at least twice the height the obstacle
protrudes above the sampler.
• There must be a 270° arc of unrestricted air flow around the monitor to include
the prevailing wind direction that provides the maximum pollutant concentration to
the monitor.
• No furnaces or incineration flues should be in close proximity to the monitor.
4-4
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-pi
Ul
ZONE C (UNACCEPTABLE)
ZONE B (NOT RECOMMENDED
10 20 25 30
DISTANCE FROM EDGE OF NEAREST TRAFFIC LANE, meters
Figure 4-1. Acceptable zone for siting TSP monitors where the average daily traffic exceeds 3000
vehicles/day.
Zone A: Recommended for neighborhood, urban, regional and most middle spatial scales. All NAMS are in this zone.
Zone B: If SLAMS are placed in Zone B they have middle scale of representativeness.
Source: C.F.R. (1982) 40: § 58
-------
To clarify the relationship between monitoring objectives and the actual siting of a mon-
itor, the concept of a spatial scale of representativeness was developed. The spatial scales
are described in terms of the physical dimensions of the air space surrounding the monitor
throughout which pollutant concentrations are fairly similar. Table 4-3 describes the scales
of representativeness while Table 4-4 relates monitoring objectives to the appropriate spatial
scale [C.F.R. (1982) 40:§58].
The time scale may also be an important factor. A study by Lynam (1972) illustrates the
effect of setback distance on short-term (15-minute) measurements of lead concentrations
directly downwind from the source. They found sharp reductions in lead concentration with in-
creasing distance from the roadway. A similar study by PEDCo Environmental, Inc. (1981) did
not show the same pronounced reduction when the data were averaged over monthly or quarterly
time periods. The apparent reason for this effect is that windspeed and direction are not
consistent. Therefore, siting criteria must include sampling times sufficiently long to
include average windspeed and direction, or a sufficient number of samples must be collected
over short sampling periods to provide an average value consistent with a 24-hour exposure.
4.2.2 Ambient Sampling for Particulate and Gaseous Lead
Airborne lead is primarily inorganic particulate matter (PM) but may occur in the form of
organic gases. Devices used for collecting samples of ambient atmospheric lead include the
standard hi-vol and a variety of other collectors employing filters, impactors, impingers, or
scrubbers, either separately or in combination. Some samplers measure total particulate
matter gravimetrically; thus the lead data are usually expressed in ug/g PM or ug/m3 air.
Other samplers do not measure PM gravimetrically; therefore, the lead data can only be
expressed as ug/m3. Some samplers measure lead deposition expressed in |jg/cm2. Some instru-
ments separate particles by size. As a general rule, particles smaller than 2.5 urn are
defined as fine, and those larger than 2.5 (jm are defined as coarse.
In a typical sampler, the ambient air is drawn down into the inlet and deposited on the
collection surface after one or more stages of particle size separation. Inlet effectiveness,
internal wall losses, and retention efficiency of the collection surface may bias the
collected sample by selectively excluding particles of certain sizes.
4.2.2.1 High Volume Sampler (hi-vol). The present SLAMS and NAMS employ the standard hi-vol
sampler (Robson and Foster, 1962; Silverman and Viles, 1948; U.S. Environmental Protection
Agency, 1971) as part of their sampling networks. As a Federal Reference Method Sampler, the
hi-vol operates with a specific flow rate range of 1.13 - 1.70 mVmin, drawing air through a
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TABLE 4-3. DESCRIPTION OF SPATIAL SCALES OF REPRESENTATIVENESS
Microscale
Middle scale
Neighborhood scale
Urban scale
Regional scale
National and global
scales
Personal
Defines ambient concentrations in air volumes associated
with areas ranging from several to 100 m2 in size.
Defines concentrations in areas from 100 to 500 m2
(area up to several city blocks).
Defines concentrations in an extended area of uniform
land use, within a city, from 0.5 to 4.0 km2 in
size.
Defines citywide concentrations, areas from 4-50
km2 in size. Usually requires more than one
site.
Defines concentrations in a rural area with homogeneous
geography. Range of tens to hundreds of km2.
Defines concentrations characterizing the U.S. and the
globe as a whole.
Defines air proximate to human respiration, usually
sampled with a portable pump.
Source: C.F.R. (1982) 40:§58 App. D; personal scale added in this report.
TABLE 4-4. RELATIONSHIP BETWEEN MONITORING OBJECTIVES AND
APPROPRIATE SPATIAL SCALES
Monitoring objective
Appropriate spatial scale for siting air monitors
Highest concentration
Population
Source impact
General (background)
Micro, Middle, Neighborhood (sometimes Urban).
Neighborhood, Urban
Micro, Middle, Neighborhood
Neighborhood, Regional
Source: C.F.R. (1982) 40:§58 App. D.
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200 x 250 mm glass fiber filter. At these flow rates, 1600 - 2500 m3 of air per day are
sampled. Many hi-vol systems are presently equipped with mass flow sensors to control the
total flow rate through the filter.
The present hi-vol approach has been shown during performance characterization tests to
have a number of deficiencies. Wind tunnel testing by Wedding et al. (1977) has shown that
the collection characteristics of hi-vol samples are strongly affected by particle size, wind
speed and direction, and inlet size. However, since most lead particles have been shown to
have a mass median aerodynamic diameter (MMAO) in the range of 0.25 - 1.4 urn (Lee and
Goranson, 1972), the hi-vol sampler should present reasonably good estimates of ambient lead
concentrations. For particles larger than 5 urn, the hi-vol system is unlikely to collect
representative samples (McFarland et al., 1979; Wedding et al., 1977).
4.2.2.2 Dichotomous Sampler. The dichotomous sampler collects two particle size fractions,
typically 0 - 2.5 urn and 2.5 urn to the upper cutoff of the inlet employed (normally 10 urn).
The impetus for the dichotomy of collection, which approximately separates the fine and coarse
particles, was provided by Whitby et al. (1972) to assist in the identification of particle
sources. A 2.5 (jm cutpoint for the separator was also recommended by Miller et al. (1979) be-
cause it satisfied the requirements of health researchers interested in respirable particles,
provided adequate separation between two naturally occurring peaks in the size distribution,
and was mechanically practical. Because the fine and coarse fractions collected in most loca-
tions tend to be acidic and basic, respectively, this separation also minimizes potential par-
ticle interaction after collection.
The particle separation principle used by this sampler was described by Hounam and
Sherwood (1965) and Conner (1966). The version now in use by EPA was developed by Loo et al.
(1979). The separation principle involves acceleration of the particles through a nozzle.
Ninety percent of the flowstream is diverted to a small particle collector, while the larger
particles continue by inertia toward the large particle collection surface. The inertial vir-
tual impactor design causes 10 percent of the fine particles to be collected with the coarse
particle fraction. Therefore, the mass of fine and coarse particles must be adjusted to allow
for their cross contamination. This mass correction procedure has been described by Dzubay et
al. (1982).
Teflon membrane filters with pore sizes as large as 2.0 urn can be used in the dichoto-
mous sampler (Dzubay et al, 1982; Stevens et al., 1980) and have been shown to have essen-
tially 100 percent collection efficiency for particles with an aerodynamic diameter as small
as 0.03 urn (Liu et al., 1976; see Section 4.2.5). Because the sampler operates at a flowrate
of 1 mVhr (167 1/min) and collects sub-milligram quantities of particles, a microbalance with
a 1 M9 resolution is recommended for filter weighing (Shaw, 1980). Removal of the fine par-
ticles via this fractionation technique may result in some of the collected coarse particles
4-8
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falling off the filter if care is not taken during filter handling and shipping. However
Dzubay and Barbour (1983) have developed a filter coating procedure which eliminates particle
loss during transport. A study by Wedding et al. (1980) has shown that the Sierra® inlet to
the dichotomous sampler was sensitive to windspeed. The 50 percent cutpoint (D50) was found
to vary from 10 to 22 pm over the windspeed range of 0 to 15 km/hr.
Automated versions of the sampler allow timely and unattended changes of the sampler
filters. Depending on atmospheric concentrations, short-term samples of as little as 4 hours
can provide diurnal pattern information. The mass collected during such short sample periods
however, is extremely small and highly variable results may be expected.
4-2.2.3 Impactor Samplers. Impactors provide a means of dividing an ambient particle sample
into subfractions of specific particle size for possible use in determining size distribution.
A jet of air is directed toward a collection surface, which is often coated with an adhesive
or grease to reduce particle bounce. Large, high-inertia particles are unable to turn with
the airstream; consequently, they hit the collection surface. Smaller particles follow the
airstream and are directed toward the next impactor stage or to the filter. Use of multiple
stages, each with a different particle size cutpoint, provides collection of particles in
several size ranges.
For determining particle mass, removable impaction surfaces may be weighed before and
after exposure. The particles collected may be removed and analyzed for individual elements.
The selection and preparation of these impaction surfaces have significant effects on the
impactor performance. Improperly coated or overloaded surfaces can cause particle bounce to
lower stages resulting in substantial cutpoint shifts (Dzubay et al. , 1976). Additionally,
coatings may cause contamination of the sample. Marple and Willeke (1976) showed the effect
of various impactor substrates on the sharpness of the stage cutpoint. Glass fiber substrates
can also cause particle bounce or particle interception (Dzubay et al., 1976) and are subject
to the formation of artifacts, due to reactive gases interacting with the glass fiber, similar
to those on hi-vol sampler filters (Stevens et al., 1978).
Cascade impactors typically have 2 to 10 stages, and flowrates for commercial low-volume
versions range from about 0.01 to 0.10 mVmin. Lee and Goranson (1972) modified a commer-
cially available 0.03 rtrVmin low-volume impactor and operated it at 0.14 m3/min to obtain
larger mass collections on each stage. Cascade impactors have also been designed to mount on
a hi-vol sampler and operate at flowrates as high as 0.6 - 1.1 m3/min.
Particle size cutpoints for each stage depend primarily on sampler geometry and flowrate.
The smallest particle size cutpoint routinely used is approximately 0.3 pm, although special
low-pressure impactors such as that described by Hering et al. (1978) are available with cut-
points as small as 0.05 pm. However, due to the low pressure, volatile organics and nitrates
4-9
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are lost during sampling. A membrane filter is typically used after the last stage to collect
the remaining small particles.
4.2.2.4 Dry Deposition Sampling. Dry deposition may be measured directly with surrogate or
natural surfaces, or indirectly using micrometeorological techniques. The earliest surrogate
surfaces were dustfall buckets placed upright and exposed for several days. The Health and
Safety Laboratory (HASL) wet-dry collector is a modification which permits one of a pair of
buckets to remain covered except during rainfall. These buckets do not collect a representa-
tive sample of particles in the small size range where lead is found because the rim perturbs
the natural turbulent flow of the main airstream (Hicks et al., 1980). They are widely used
for other pollutants, especially those found primarily on large particles, in the National
Atmospheric Deposition Program.
Other surrogate surface devices with smaller rims or no rims have been developed recently
(Elias et al., 1976; Lindberg et al., 1979; Peirson et al., 1973). Peirson et al. (1973)
used horizontal sheets of filter paper exposed for several days with protection from rainfall.
Elias et al. (1976) used Teflon® disks held rigid with a 1 cm Teflon® ring. Lindberg et al.
(1979) used petri dishes suspended in a forest canopy. In all of these studies, the calcu-
lated deposition velocity (see Section 6.3.1) was within the range expected for small aerosol
particles.
A few studies have measured direct deposition on vegetation surfaces using chemical wash-
ing techniques to remove surface particles. These determinations are generally 4-10 times
lower than comparable surrogate surface measurements (Elias et al., 1976; Lindberg et al.,
1979), but the reason for this difference could be that natural surfaces represent net accumu-
lation rather than total deposition. Lead removed by rain, dripping dew, or other processes
such as foliar uptake would result in an apparently lower deposition rate. In the Lindberg
et al. (1979) study, leaves were collected during rainless periods and could not have been
influenced by rain washoff. Removal by dew or intercepted fog dripping from the leaves could
not be ruled out, but the explanation given by a subsequent report (Lindberg and Harriss,
1981) was that some dry deposition was absorbed by the foliage, that is, foliar uptake was
occurring.
There are several micrometeorological techniques that have been used to measure particle
deposition. They overcome a deficiency of surrogate surfaces, the lack of correlation between
the natural and artificial surfaces, but micrometeorological techniques require expensive
equipment and skilled operators. They measure instantaneous or short-term deposition only,
and this deposition is inferred to be to a plane-projected surface area only, not necessarily
to vegetation surfaces.
Of the five micrometeorological techniques commonly used to measure particle deposition,
only two have been used to measure lead particle deposition. Everett et al. (1979) used the
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profile gradient technique by which lead concentrations are measured at two or more levels
within 10 m above the surface. Parallel meteorological data are used to calculate the net
flux downward. Droppo (1980) used eddy correlation, which measures fluctuations in the ver-
tical wind component with adjacent measurements of lead concentrations. The calculated dif-
ferences of each can be used to determine the turbulent flux. These two micrometeorological
techniques and the three not yet used for lead, modified Bowen, variance, and eddy accumula-
tion, are described in detail in Hicks et al. (1980).
4.2.2.5 Gas Collection. When sampling ambient lead with systems employing filters, it is
likely that vapor-phase organolead compounds will pass through the filter media. The use of
bubblers downstream of the filter containing a suitable reagent or absorber for collection of
these compounds has been shown to be effective (Purdue et al., 1973). Organolead may be col-
lected on iodine crystals, adsorbed on activated charcoal, or absorbed in an iodine mono-
chloride solution (Skogerboe et al., 1977b).
In one experiment, Purdue et al. (1973) operated two bubblers in series containing iodine
monochloride solution. One hundred percent of the lead was recovered in the first bubbler.
It should be noted, however, that the analytical detection sensitivity was poor. In general,
use of bubblers limits the sample volume (and consequently the sample collection period) due
to losses by evaporation and/or bubble carryover. Birch et al. (1980) addressed this problem
by increasing the volume of iodine monochloride solution and modifying the inlet impinger to
reduce foaming. These authors reported a 97-99 percent collection efficiency of 100 to 500 ng
Pb in the first of two bubblers in series. Under ambient sample conditions, this procedure
can be used to collect a 24-48 hour sample, provided precautions are taken to retard the
decomposition of iodine monochloride by avoiding exposure to light. The sensitivity was
0.25 ng Pb/m3 for a 48-hour sample.
These procedures do not identify the specific organolead compound collected. Blaszkewicz
and Neidhart (1983) have described a technique for the quantitative identification of four
organolead compounds: tetramethyl lead, tetraethyl lead, trimethyl lead, and triethyl lead.
This technique requires a minimum sample size of 70 ng Pb, which is probably higher than
ambient under most collection conditions for a 24- to 48-hour sampling period.
4.2.3 Source Sampling
Sources of atmospheric lead include automobiles, smelters, coal-burning facilities, waste
oil combustion, battery manufacturing plants, chemical processing plants, facilities for scrap
processing, and welding and soldering operations (see Section 5.3.3). A potentially important
secondary source is fugitive dust from mining operations and from soils contaminated with
automotive emissions (Olson and Skogerboe, 1975). Chapter 5 contains a complete discussion of
4-11
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sources of lead emissions. The following sections discuss the sampling near potential sta-
tionary and mobile sources. Neither indoor nor personal monitoring for lead is performed
routinely for ambient situations, although Roy (1977) and Tosteson et al. (1982) discuss the
techniques used for personal sampling under special circumstances (see Section 7.2.1.3.3).
4.2.3.1 Stationary Sources. Sampling of stationary sources for lead requires the use of a
sequence of samplers at the source of the effluent stream. Since lead in stack emissions may
be present in a variety of physical and chemical forms, source sampling trains must be de-
signed to trap and retain both gaseous and particulate lead. A sampling probe is inserted
directly in the stack or exhaust stream. In the tentative ASTM method for sampling for atmos-
pheric lead, air is pulled through a 0.45 urn membrane filter and an activated carbon adsorp-
tion tube (American Society for Testing and Materials, 1975a).
4.2.3.2 Mobile Sources. Three principal procedures have been used to obtain samples of auto
exhaust aerosols for subsequent analysis for lead compounds: a horizontal dilution tunnel,
plastic sample collection bags, and a low residence time proportional sampler. In each proce-
dure, samples are air-diluted to simulate roadside exposure conditions. In the most commonly
used procedure, a large horizontal air dilution tube segregates fine combustion-derived parti-
cles from larger lead particles ablated from combustion chamber and exhaust deposits. In one
example of this procedure (Habibi, 1970), hot exhaust is ducted into a 56-cm diameter, 12-m
long, air dilution tunnel and mixed with filtered ambient air in a 10-cm diameter mixing
baffle in a concurrent flow arrangement. Total exhaust and dilution airflow rate is 28 - 36
nrVmin, which produces a residence time of approximately 5 sec in the tunnel. At the down-
stream end of the tunnel, samples of the aerosol are obtained by means of isokinetic probes
using filters or cascade impactors.
In recent years, various configurations of the horizontal air dilution tunnel have been
developed. Several dilution tunnels have been made of polyvinyl chloride with a diameter of
46 cm, but these are subject to wall losses due to charge effects (Gentel et al., 1973; Moran
et al., 1972; Trayser et al., 1975). Such tunnels of varying lengths have been limited by
exhaust temperatures to total flows above approximately 11 mVmin. Similar tunnels have a
centrifugal fan located upstream, rather than a positive displacement pump located downstream
(Trayser et al., 1975). This geometry produces a slight positive pressure in the tunnel and
expedites transfer of the aerosol to holding chambers for studies of aerosol growth. However,
turbulence from the fan may affect the sampling efficiency. Since the total exhaust plus
dilution airflow is not held constant in this system, potential errors can be reduced by main-
taining a very high dilution air/exhaust flow ratio (Trayser et al., 1975).
There have also been a number of studies using total filtration of the exhaust stream to
arrive at material balances for lead with rather low back-pressure metal filters in an air
4-12
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distribution tunnel (Habibi, 1973; Hirschler etal., 1957; Hirschler and Gilbert, 1964;
Sampson and Springer, 1973). The cylindrical filtration unit used in these studies is better
than 99 percent efficient in retaining lead particles (Habibi, 1973). Supporting data for
lead balances generally confirm this conclusion (Kunz et al., 1975).
In the bag technique, auto emissions produced during simulated driving cycles are air-
diluted and collected in a large plastic bag. The aerosol sample is passed through a filtra-
tion or impaction sampler prior to lead analysis (Ter Haar et al., 1972). This technique may
result in errors of aerosol size analysis because of condensation of low vapor pressure
organic substances onto the lead particles.
To minimize condensation problems, a third technique, a low residence time proportional
sampling system, has been used. It is based on proportional sampling of raw exhaust, again
diluted with ambient air followed by filtration or impaction (Ganley and Springer, 1974;
Sampson and Springer, 1973). Since the sample flow must be a constant proportion of the total
exhaust flow, this technique may be limited by the response time of the equipment to operating
cycle phases that cause relatively small transients in the exhaust flow rate.
4.2.4 Sampling for Lead in Water, Soil. Plants, and Food
Other primary environmental media that may contain airborne lead include precipitation,
surface water, soil, vegetation, and foodstuffs. The sampling plans and the sampling metho-
dologies used in dealing with these media depend on the purpose of the experiments, the types
of measurements to be carried out, and the analytical technique to be used. General
approaches are given below in lieu of specific procedures associated with the numerous possi-
ble special situations.
4.2.4.1 Precipitation. Methods developed and used at the Oak Ridge National Laboratory for
precipitation collection and analysis for lead are described in Lindberg et al. (1979),
Lindberg (1982), and Lindberg and Turner (1983). The investigation should be aware that dry
deposition occurs continuously, that lead at the start of a rain event is higher in concentra-
tion than at the end, and that rain striking the canopy of a forest may rinse dry deposition
particles from the leaf surfaces. Rain collection systems should be designed to collect pre-
cipitation on an event basis and to collect sequential samples during the event. They should
be tightly sealed from the atmosphere before and after sampling to prevent contamination from
dry deposition, falling leaves, and flying insects. Samples for total lead analysis should be
acidified to pH less than 2 with nitric acid and refrigerated immediately after sampling.
Samples to be separated for particulate and dissolved lead analysis should be filtered prior
to acidification. All collection and storage surfaces should be thoroughly cleaned and free
of contamination.
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Two automated rain-collecting systems have been in use for some time. The Sangamo Pre-
cipitation Collector, Type A, collects rain in a single bucket exposed at the beginning of the
rain event (Samant and Vaidya, 1982). These authors reported no leaching of lead from the
bucket into a solution of 0.3N HN03. A second sampler, described by Coscio et al. (1982),
also remains covered between rain events; it can collect a sequence of eight samples during
the period of rain and may be fitted with a refrigeration unit for sample cooling. No reports
of lead analyses were given. Because neither system is widely used for lead sampling, their
monitoring effectiveness has not been thoroughly evaluated.
4.2.4.2 Surface Water. Atmospheric lead may be dissolved in water as hydrated ions, chemical
complexes, and soluble compounds, or it may be associated with suspended matter. Because the
physicochemical form often influences environmental effects, there is a need to differentiate
among the various chemical forms of lead. Complete differentiation among all such forms is a
complex task that has not yet been fully accomplished. The most commonly used approach is to
distinguish between dissolved and suspended forms of lead. All lead passing through a 0.45 pm
membrane filter is operationally defined as dissolved, while that retained on the filter is
defined as suspended (Kopp and McKee, 1983). Figura and McDuffie (1979, 1980) broadened this
scheme to encompass four categories of metal lability that are presumably more representative
of uptake by biological systems. These categories are: very labile, moderately labile,
slowly labile, and inert. Distinctions between categories are made experimentally by column
ion exchange, batch ion exchange, and anodic stripping voltammetry. The key point is the
kinetics of the experimental process. If the metal complex can be made to dissociate within
milliseconds (anodic stripping voltammetry), then it is considered very labile. Assuming that
biological systems take up metals in the free ion state rather than as metal complexes, this
scheme can provide important information on the bioavailability of lead in natural waters.
Cox et al. (1984) provide evidence that Donnan dialysis, which uses an ion exchange membrane
rather than a resin column, may provide a better estimate of lability for lead in natural
waters than Chelex-100.
When sampling water bodies, flow dynamics should be considered in the context of the pur-
pose for which the sample is collected. Water at the convergence point of two flowing
streams, for example, may not be well mixed for several hundred meters. Similarly, the heavy
metal concentrations above and below the thermocline of a lake may be very different. Thus,
several samples should be selected in order to define the degree of horizontal or vertical
variation. The final sampling plan should be based on the results of pilot studies. In cases
where the average concentration is of primary concern, samples can be collected at several
points and then mixed to obtain a composite.
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Containers used for sample collection and storage should be fabricated from essentially
lead-free plastic or glass, e.g., conventional polyethylene, Teflon , or quartz. These con-
tainers must be leached with hot acid for several days to ensure minimum lead contamination
(Patterson and Settle, 1976). If only the total lead is to be determined, the sample may be
collected without filtration in the field. Nitric acid should be added immediately to reduce
the pH to less than 2; the acid will normally dissolve the suspended lead. Otherwise, it is
recommended that the sample be filtered upon collection to separate the suspended and dissol-
ved lead and the latter preserved by acid addition as above (U.S. Environmental Protection
Agency, 1978). It is also recommended that water samples be stored at 4°C until analysis to
avoid further leaching from the container wall (Fishman and Erdmann, 1973; Kopp and Kroner,
1967; Lovering, 1976; National Academy of Sciences, 1972; U.S. Environmental Protection
Agency, 1978).
4.2.4.3 Soils. The distance from emission sources and depth gradients associated with lead
in soil must be considered in designing the sampling plan. Vegetation, litter, and large
objects such as stones should not be included in the sample, depth samples should be collected
at 2 cm intervals to preserve vertical integrity, and the samples should be air dried and
stored in sealed containers until analyzed. Brown and Black (1983) have addressed the problem
of quality assurance and quality control in the collection and analysis of soil samples. A
twelve-step procedural protocol and a three-step data validation process were recommended to
obtain the most accurate results, and some suggestions were made for handling data bias,
precision and uncertainty. Eastwood and Jackson (1984) reported the results of an inter-
laboratory study that showed greater variations between laboratories than within a laboratory,
especially when different analytical procedures are followed.
The chemical similarities between lead complexes in natural waters and in the water
associated with soil are not clearly established in the literature. In the more concentrated
medium of soil moisture, the lability of lead may change, favoring higher percentages of inert
or slowly labile lead (see Section 6.5.1). Although there are many procedures for the analy-
sis of bulk soil samples and for extruding metals from soils in a manner that simulates plant
uptake, there are few reports on the collection and analysis of soil moisture at the site of
root uptake. The techniques developed by Hinkley and Patterson (1973) for sampling the film
of moisture surrounding soil particles have been used by Elias et al. (1976, 1978, 1982) and
Elias and Patterson (1980) for the analysis of lead in small volumes of moisture extracted
from soil particles in the root zone.
4.2.4.4 Vegetation. Because most soil lead is in forms unavailable to plants, and because
lead is not easily transported by plants, roots typically contain very little lead and shoots
even less (Zimdahl, 1976; Zimdahl and Koeppe, 1977). Before analysis, a decision must be made
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as to whether or not the plant material should be washed to remove surface contamination from
dry deposition and soil particles. If the plants are sampled for total lead content (e.g., if
they represent animal food sources), they cannot be washed. If the effect of lead on internal
plant processes is being studied, the plant samples should be washed. In either case, the
decision must be made at the time of sampling, as washing cannot be effective after the plant
materials have dried. Fresh plant samples cannot be stored for any length of time in a tight-
ly closed container before washing because molds and enzymatic action may affect the distribu-
tion of lead on and in the plant tissues. Freshly picked leaves stored in sealed polyethylene
bags at room temperature generally begin to decompose in a few days. Storage time may be
increased to approximately 2 weeks by refrigeration. Samples that are to be stored for ex-
tended periods of time should be oven dried to arrest enzymatic reactions and render the plant
tissue amenable to grinding. Storage in sealed containers is required after grinding. For
analysis of surface lead, fresh, intact plant parts are agitated in dilute nitric acid or EDTA
solutions for a few seconds.
4.2.4.5 Foodstuffs. Analyses for lead in food have been included in the Food and Drug Admin-
istration's Total Diet Study since 1972. Initially, this survey involved sampling of foods
representing the average diet of a 15 to 20 year-old male, i.e., the individual who on a
statistical basis eats the greatest quantity of food (Kolbye et al., 1974). Various food
items from the several food classes were purchased in retail stores in various cities across
the nation. The foods were cooked or otherwise prepared as they would be in the kitchen, then
composited into 12 food classes and analyzed chemically. Other FDA sampling programs are
required for different investigative purposes, e.g., enforcement of regulations. For those
foods where lead may be deposited on the edible portion, typical kitchen washing procedures
are used. This survey procedure has been replaced by one involving separate analyses of 234
individual foods and covering 8 age-sex groups (Pennington, 1983). It is this revised sam-
pling and analytical format that is the basis for food exposure estimates in Section 7.3.1.2.
4.2.5 Filter Selection and Sample Preparation
In sampling for airborne lead, air is drawn through filter materials such as glass fiber,
cellulose acetate, or porous plastic (Skogerboe et al., 1977b, Stern, 1976). These materials
often include contaminant lead that can interfere with the subsequent analysis (Gandrud and
Lazrus, 1972; Kometani et al. 1972; Luke et al., 1972; Seeley and Skogerboe, 1974). If a
large mass of particulate matter is collected, then the effects of these trace contaminants
may be negligible (Witz and MacPhee, 1976). Procedures for cleaning filters to reduce the
lead blank rely on washing with acids or complexing agents (Gandrud and Lazrus, 1972). The
type of filter and the analytical method to be used often determines the washing technique.
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In some methods, e.g., X-ray fluorescence, analysis can be performed directly on the filter if
the filter material is suitable (Dzubay and Stevens, 1975). Skogerboe (1974) provided a
general review of filter materials.
The main advantages of glass fiber filters are low pressure drop and high particle col-
lection efficiency at high flow rates. The main disadvantage is variable lead blank, which
makes their use inadvisable in many cases (Kometani et al., 1972; Luke at al., 1972). This
has placed a high priority on the standardization of a suitable filter for hi-vol samples
(Witz and MacPhee, 1976). Other investigations have indicated, however, that glass fiber
filters are now available that do not present a lead interference problem (Scott et al.,
A
1976b). Teflon filters have been used since 1975 by Dzubay et al. (1982) and Stevens et al.
(1978), who have shown these filters to have very low lead blanks (<2 ng/cm2). The collection
efficiencies of filters, and also of impactors, have been shown to be dominant factors in the
quality of the derived data (Skogerboe et al., 1977a).
Sample preparation usually involves conversion to a solution through wet ashing of solids
with acids or through dry ashing in a furnace followed by acid treatment. Either approach
works effectively if used properly (Kometani et al., 1972; Skogerboe et al., 1977b). In one
j&
investigation of porous plastic Nuclepore filters, some lead blanks were too high to allow
measurements of ambient air lead concentrations (Skogerboe et al., 1977b).
4.3 ANALYSIS
The choice of analytical method depends on the nature of the data required, the type of
sample being analyzed, the skill of the analyst, and the equipment available. For general
determination of elemental lead, atomic absorption spectroscopy is widely used and recommended
[C.F.R. (1982) 40:§50]. Optical emission spectrometry (Scott et al., 1976b) and X-ray fluore-
scence (Stevens et al., 1978) are rapid and inexpensive methods for multi-elemental analyses.
X-ray fluorescence can measure lead concentrations reliably to 1 ng/m3 using samples col-
lected with commercial dichotomous samplers. Other analytical methods have specific advan-
tages appropriate for special studies. Only those analytical techniques receiving widespread
current use in lead analysis are described below. More complete reviews are available in the
literature (American Public Health Association, 1971; Lovering, 1976; Skogerboe et al., 1977b;
National Academy of Sciences, 1980).
With respect to measuring lead without sampling or laboratory contamination, several in-
vestigators have shown that the magnitude of the problem is quite large (Patterson and Settle,
1976; Patterson et al., 1976; Pierce et al., 1976; Patterson, 1983; Skogerboe, 1982). It ap-
pears that the problem may be caused by failure to control the blank or by failure to stan-
dardize instrument operation (Patterson, 1983; Skogerboe, 1982). The laboratory atmosphere,
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collecting containers, and the labware used may be primary contributors to the lead blank pro-
blem (Murphy, 1976; Patterson, 1983; Skogerboe, 1982). Failure to recognize these and other
sources such as reagents and hand contact is very likely to result in the generation of arti-
ficially high analytical results. Samples with less than 100 ug Pb should be analyzed in a
clean laboratory especially designed for the elimination of lead contamination. Moody (1982)
has described the construction and application of such a laboratory at the National Bureau of
Standards.
For many analytical techniques, a preconcentration step is recommended. Leyden and
Wegschneider (1981) have described several procedures and the associated problems with
controlling the analytical blank. There are two steps to preconcentration. The first is the
removal of organic matter by dry ashing or wet digestion. The second is the separation of
lead from interfering metallic elements by coprecipitation, co-crystallization, solvent
extraction of chelate, electro-deposition or passing through a chelating ion exchange resin
column. New separation techniques are continuously being evaluated, many of which have
application to specific analytical problems. Torsi and Palmisano (1984) have described
electrochemical deposition directly on a glassy carbon crucible during atomic absorption
spectrometry. Yang and Yen (1982) have described a polyacrylamide-hydrous-zirconia (PHZ)
composite ion exchanger suitable for high phosphate solutions. Corsini et al. (1982)
evaluated a macroreticular acrylic ester resin capable of removing free and inorganically
bound metal ions directly from aqueous solution without prior chelation.
Occasionally, it is advantageous to automate the sample preparation and preconcentration
process. Tyson (1985) has reviewed the use of flow injection analysis techniques specific for
atomic absorption spectrometry. Another promising technique involves a flow-injection system
in conjunction with an ion-exchange column and flame atomic absorption (Fang et al., 1984a;
1984b). For aqueous samples, preconcentration factors of 50 to 100-fold were achieved while
maintaining a sample frequency of 60 samples per hour.
The application of these and other new techniques can be expected to shed further light
on the chemistry and biological availability of lead in natural systems.
4.3.1 Atomic Absorption Spectroscopy (AAS)
Atomic absorption spectroscopy (AAS) is a widely accepted method for the measurement of
lead in environmental sampling (Skogerboe et al., 1977b). A variety of lead studies using AAS
have been reported (Kometani et al., 1972; Zoller et al., 1974; Huntzicker et al., 1975; Scott
et al., 1976b; Lester et al., 1977; Hirao et al., 1979; Compton and Thomas, 1980; Bertenshaw
and Gelsthorpe, 1981).
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The lead atoms in the sample must be vaporized either in a precisely controlled flame or
in a furnace. Furnace systems in AAS offer high sensitivity as well as the ability to analyze
small samples (Lester et al., 1977; Rouseff and Ting, 1980; Stein et al., 1980; Bertenshaw and
Gelsthorpe, 1981). These enhanced capabilities are offset in part by greater difficulty in
analytical calibration and by loss of analytical precision.
Pachuta and Love (1980) collected particles on cellulose acetate filters. Disks (0.5
cm2) were punched from these filters and analyzed by insertion of the nichrome cups containing
the disks into a flame. Another application involves the use of graphite cups as particle
filters with the subsequent analysis of the cups directly in the furnace system (Seeley and
Skogerboe, 1974; Torsi et al., 1981). These two procedures offer the ability to determine
particulate lead directly with minimal sample handling.
In an analysis using AAS and hi-vol samplers, atmospheric concentrations of lead were
found to be 0.076 ng/m3 at the South Pole (Maenhaut et al., 1979). Lead analyses of 995 par-
ticulate samples from the NASN were accomplished by AAS with an indicated precision of 11
percent (Scott et al., 1976a; see also Section 7.2.1.1). More specialized AAS methods have
been described for the determination of tetraalkyl lead compounds in water and fish tissue
(Chau et al., 1979) and in air (Birnie and Noden, 1980; Rohbock et al., 1980).
Atomic absorption requires as much care as other techniques to obtain highly precise
data. Background absorption, chemical interference, background light loss, and other factors
can cause errors. A major problem with AAS is that untrained operators use it in many labor-
atories without adequate quality control.
Techniques for AAS are still evolving. An alternative to the graphite furnace, evaluated
by Jin and Taga (1982), uses a heated quartz tube through which the metal ion in gaseous
hydride form flows continuously. Sensitivities were 1-3 ng/g for lead. The technique is
similar to the hydride generators used for mercury, arsenic, and selenium. Other nonflame
atomization systems, electrodeless discharge lamps, and other equipment refinements and tech-
nique developments have been reported (Horlick, 1982). A promising technique for the analysis
of samples with high salt content has been developed by 01 sen et al. (1983) using flow injec-
tion analysis. In an automated system, these authors reported a detection limit of 10 ng/g in
seawater while analyzing 30 to 60 samples per hour. This sensitivity is not as low as AAS
with a graphite furnace, so the technique would not improve the analysis of air samples with-
out further refinement.
4.3.2 Emission Spectroscopy
Optical emission spectroscopy is based on the measurement of the light emitted by ele-
ments when they are excited in an appropriate energy medium. The technique has been used to
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determine the lead content of soils, rocks, and minerals at the 5-10 ug/g level with a rela-
tive standard deviation of 5 - 10 percent (Jolly, 1963); this method has also been applied to
the analysis of a large number of air samples (Scott et al., 1976b; Sugimae and Skogerboe,
1978). The primary advantage of this method is that it allows simultaneous measurement of a
large number of elements in a small sample (Ward and Fishman, 1976).
In a study of environmental contamination by automotive lead, sampling times were short-
ened by using a sampling technique in which lead-free porous graphite was used both as the
filter medium and as the electrode in the spectrometer (Copeland et al., 1973; Seeley and
Skogerboe, 1974). Lead concentrations of 1 - 10 ug/m3 were detected after a half-hour flow at
800 to 1200 ml/min through the filter.
Scott et al. (1976a) analyzed composited particulate samples obtained with hi-vol sam-
plers for 24 elements, including lead, using a direct reading emission spectrometer. Over
1000 samples collected by the NASN in 1970 were analyzed. Careful consideration of accuracy
and precision led to the conclusion that optical emission spectroscopy is a rapid and practi-
cal technique for particle analysis.
More recent activities have focused attention on the inductively coupled plasma (ICP)
system as a valuable means of excitation and analysis (Garbarino and Taylor, 1979; Winge et
al., 1977). The ICP system offers a higher degree of sensitivity with less analytical inter-
ference than is typical of many of the other emission spectroscopic systems. Optical emission
methods are inefficient when used for analysis of a single element, since the equipment is
expensive and a high level of operator training is required. This problem is largely offset
when analysis for several elements is required, as is often the case for atmospheric aerosols.
However, the ICP procedure does not provide the sensitivity required for determining the
levels of lead in foods (Jones and Boyer, 1978; Jones et al., 1982).
4.3.3 X-Ray Fluorescence (XRF)
X-ray emissions that characterize the elemental content of a sample also occur when atoms
are irradiated at sufficient energy to excite an inner-shell electron (Hammerle and Pierson,
1975; Jaklevic et al., 1973; Skogerboe et al., 1977b; Stevens et al., 1978). This fluores-
cence allows simultaneous identification of a range of elements including lead.
X-ray fluorescence may require a high-energy irradiation source. But with the X-ray
tubes coupled with fluorescers (Jaklevic et al., 1973; Dzubay and Stevens, 1975; Paciga and
Jervis, 1976) very little energy is transmitted to the sample; thus sample degradation is kept
to a minimum (Shaw et al., 1980). Electron beams (McKinley et al., 1966) and radioactive iso-
tope sources (Kneip and Laurer 1972) have been used extensively as energy sources for XRF
analysis (Birks et al., 1971; Birks, 1972). To reduce background interference, secondary
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fluorescers have been employed (Birks et al. , 1971; Dzubay and Stevens, 1975). The fluor-
escent X-ray emission from the sample may be analyzed with a crystal monochromator and
detected with scintillation or proportional counters, or with low-temperature semiconductor
detectors that discriminate the energy of the fluorescence. The latter technique requires a
very low level of excitation (Dzubay and Stevens, 1975; Toussaint and Boniforti, 1979).
X-ray emission induced by charged-particle excitation (proton-induced X-ray emission or
PIXE) offers an attractive alternative to the more common techniques (Barfoot et al., 1979;
Hardy et al., 1976; Johansson et al., 1970). The potential of heavy-particle bombardment for
excitation was demonstrated by Johansson et al. (1970), who reported an interference-free
signal in the picogram (10 12 g) range. The excellent capability of accelerator beams for
X-ray emission analysis is partially due to the relatively low background radiation associated
with the excitation. The high particle fluxes obtainable from accelerators also contribute to
the sensitivity of the PIXE method. Literature reviews (Folkmann et al., 1974; Gilfrich et
al., 1973; Herman et al., 1973; Walter et al., 1974) on approaches to X-ray elemental analysis
agree that protons of a few MeV energy provide a preferred combination for high sensitivity
analysis under conditions less subject to matrix interference effects. As a result of this
premise, a system designed for routine analysis has been described (Johansson et al., 1975)
and papers involving the use of PIXE for aerosol analysis have appeared (Hardy et al., 1976;
Johansson et al., 1975). The use of radionuclides to excite X-ray fluorescence and to deter-
mine lead in airborne particles has also been described (Havranek and Bumbalova, 1981;
Havranek et al., 1980).
X-radiation is the basis of the electron microprobe method of analysis. When an intense
electron beam is incident on a sample, it produces several forms of radiation, including
X-rays, whose wavelengths depend on the elements present in the material and whose intensities
depend on the relative quantities of these elements. An electron beam that gives a spot size
as small as 0.2 urn is possible. The microprobe is often incorporated in a scanning electron
microscope that allows precise location of the beam and comparison of the sample morphology
with its elemental composition. Under ideal conditions, the analysis is quantitative, with an
accuracy of a few percent. The mass of the analyzed element may range from 10 14 to 10 16 g
(McKinley et al., 1966).
Electron microprobe analysis is not a widely applicable monitoring method. It requires
expensive equipment, complex sample preparation procedures, and a highly trained operator.
The method is unique, however, in providing compositional information on individual lead par-
ticles, thus permitting the study of dynamic chemical changes and perhaps allowing improved
source identification.
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Advantages of X-ray fluorescence methods include the ability to detect a variety of ele-
ments, the ability to analyze with little or no sample preparation, low detection limits (2 ng
Pb/m3) and the availability of automated analytical equipment. Disadvantages are that the
X-ray analysis requires liquid nitrogen (e.g., for energy-dispersive models) and highly
trained analysts. The detection limit for lead is approximately 9 ng/cm2 of filter area
(Jaklevic and Walter, 1977), which is well below the quantity obtained in normal sampling
periods with the dichotomous sampler (Dzubay and Stevens, 1975).
4.3.4 Isotope Dilution Mass Spectrometry (IDMS)
Isotope dilution mass spectrometry (IDMS) is an absolute measurement technique. It
serves as the standard to which other analytical techniques are compared. No other techniques
serve more reliably as a comparative reference. Its use for analyses at subnanogram concen-
trations of lead and in a variety of sample types has been reported (Chow et al., 1969, 1974;
Facchetti and Geiss, 1982; Hirao and Patterson, 1974; Murozumi et al., 1969; Patterson et al.,
1976; Rabinowitz et al., 1973).
The isotopic composition of lead peculiar to various ore bodies and crustal sources may
also be used as a means of tracing the origin of anthropogenic lead. Other examples of IDMS
application are found in several reports cited above, and in Rabinowitz and Wetherill (1972),
Stacey and Kramers (1975), and Machlan et al. (1976).
4.3.5 Colorimetric Analysis
Colorimetric or spectrophotometric analysis for lead using dithizone (diphenylthiocarba-
zone) as the reagent has been used for many years (Jolly, 1963; Williams, 1984; Sandell,
1944). It was the primary method recommended by a National Academy of Sciences (1972) report
on lead, and the basis for the tentative method of testing for lead in the atmosphere by the
American Society for Testing and Materials (1975b). Prior to the development of the IDMS
method, Colorimetric analysis served as the reference by which other methods were tested.
The procedures for the Colorimetric analysis require a skilled analyst. The ASTM conduc-
ted a collaborative test of the method (Foster et al., 1975) and concluded that the procedure
gave satisfactory precision in the determination of particulate lead in the atmosphere. In
addition, the required apparatus is simple and relatively inexpensive, the absorption is
linearly related to the lead concentration, large samples can be used, the method is easily
sensitive to a few micrograms of lead, and interferences can be removed (Skogerboe et al.,
1977b). Realization of these advantages depends on meticulous attention to the procedures and
reagents.
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4.3.6 Electrochemical Methods: Anodic Stripping Voltammetry (ASV). Differential Pulse
Polarography (DPP)
Analytical methods based on electrochemical phenomena are found in a variety of forms
(Sawyer and Roberts, 1974; Willard et al., 1974). They are characterized by a high degree of
sensitivity, selectivity, and accuracy derived from the relationship between current, charge,
potential, and time for electrolytic reactions in solutions. The electrochemistry of lead is
based primarily on Pb(II), which behaves reversibly in ionic solutions having a reduction po-
tential near -0.4 volt versus the standard calomel electrode (Skogerboe et al., 1977b). Two
electrochemical methods generally offer sufficient analytical sensitivity for most lead mea-
surement problems. Differential pulse polarography (DPP) relies on the measurement of the
faradaic current for lead as the voltage is scanned while compensating for the nonfaradaic
(background) current produced (McDonnell, 1981). Anodic stripping voltammetry (ASV) is a two
step process in which the lead is preconcentrated onto a mercury electrode by an extended but
selected period of reduction. After the reduction step, the potential is scanned either
linearly or by differential pulse to oxidize the lead and allow measurement of the oxidation
(stripping) current. The preconcentration step allows development of enhanced analytical
signals; when used in combination with the differential pulse method, lead concentrations at
the subnanogram level can be measured (Florence, 1980).
The ASV method has been widely applied to the analysis of atmospheric lead (Harrison et
al., 1971; Khandekar et al., 1981; MacLeod and Lee, 1973). Landy (1980) has shown the applic-
ability to the determination of Cd, Cu, Pb, and Zn in Antarctic snow, while others have
analyzed rain water (Nguyen et al., 1979; Nurnberg, 1984a; 1984b) and snow samples (Nguyen et
al., 1979). Green et al. (1981) have used the method to determine Cd, Cu, and Pb in sea
water. The ASV determination of Cd, Cu, Pb, and Zn in foods has been described (Jones et al.,
1977; Capar et al., 1982; Mannino, 1982, 1983; Satzger et al., 1982), and the general accuracy
of the method summarized by Holak (1980). An ASV method for lead and cadmium in foods has
been collaboratively studied and has been adopted as an official method by the Association of
Official Analytical Chemists (Capar et al., 1982; Williams, 1984). Current practice with
commercially available equipment allows lead analysis at subnanogram concentrations with
precision at the 5 to 10 percent level on a routine basis (Skogerboe et al., 1977b). New
developments center around the use of microcomputers in controlling the stripping voltage
(Kryger, 1981) and conformational modifications of the electrode (Brihaye and Duyckaerts,
1982, 1983). Wang et al. (1983) applied flow-injection techniques to anodic stripping voltam-
metry to achieve a rate of ten samples per hour.
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4.3.7 Methods for Compound Analysis
The majority of analytical methods are restricted to measurement of total lead and cannot
directly identify the various compounds of lead. The electron microprobe and other X-ray
fluorescence methods provide approximate data on compounds on the basis of the ratios of
elements present (Ter Haar and Bayard, 1971). Gas chromatography (GC) using the electron cap-
ture detector has been demonstrated to be useful for organolead compounds (Shapiro and Frey,
1968). The use of atomic absorption as the GC detector for organolead compounds has been
described by DeJonghe et al. (1981) and Hewitt and Harrison (1985), while a plasma emission
detector has been used by Estes et al. (1981). In addition, Messman and Rains (1981) have
used liquid chromatography with an atomic absorption detector to measure organolead compounds.
Mass spectrometry may also be used with GC (Mykytiuk et al., 1980).
Powder X-ray diffraction techniques have been applied to the identification of lead com-
pounds in soils by Olson and Skogerboe (1975) and by Linton et al. (1980). X-ray diffraction
techniques were used (Harrison and Perry, 1977; Foster and Lott, 1980; Jaklevic et al., 1981)
to identify lead compounds collected on air filters.
4.4 CONCLUSIONS
To monitor lead particles in air, collection with the hi-vol and dichotomous samplers and
analysis by atomic absorption spectrometry and X-ray fluorescence methods have emerged as the
most widely used methods. Sampling with the hi-vol has inherent biases in sampling large par-
ticles and does not provide for fractionation of the particles according to size, nor does it
allow determination of the gaseous (organic) concentrations. Sampling with a dichotomous
sampler provides size information but does not permit measurement of gaseous lead. The size
distribution of lead aerosol particles is important in considering inhalable particulate
matter. X-ray fluorescence and optical emission spectroscopy are applicable to multi-element
analysis. Other analytical techniques find application for specific purposes.
There is no routine monitoring program in the United States for ambient concentrations of
gaseous lead. Such measurements would require the addition of a chemical scrubber to the
particulate sampling device, a procedure that is used only under special circumstances. Dis-
cussion of the concentrations of gaseous lead are found in Section 6.3.2.
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4.5 REFERENCES
American Public Health Association. (1971) Standard methods for the examination of water and
wastewater; 13th ed. New York, NY: American Public Health Association.
American Society for Testing and Materials. (1975a) Standard method for collection and anal-
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standards; part 26. gaseous fuels; coal and coke; atmospheric analysis. Philadelphia, PA:
American Society for Testing and Materials; pp. 517-521.
American Society for Testing and Materials. (1975b) Tentative method of test for lead in the
atmosphere by colorimetric dithizone procedure; D 3112-72T. In: 1975 annual book of ASTM
standards; part 26. gaseous fuels; coal and coke; atmospheric analysis. Philadelphia, PA:
American Society for Testing and Materials; pp. 633-641.
Barfoot, K. M.; Mitchell, I. V.; Eschbach, H. L.; Mason, P. I.; Gilboy, W. B. (1979) The anal-
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Bertenshaw, M. P.; Gelsthorpe, D. (1981) Determination of lead in drinking water by atomic-
absorption spectrophotometry with electrothermal atomisation. Analyst (London) 106:
^tJ~<3X*
Birch, J.; Harrison, R. M.; Laxen, D. P. H. (1980) A specific method for 24-48 hour analysis
of tetraalkyl lead in air. Sci. Total Environ. 14: 31-42.
Birks, L. S. (1972) X-ray absorption and emission. Anal. Chem. 44: 557R-562R.
Birks, L. S.; Gilfrich, J. V.; Nagel, D. J. (1971) Large-scale monitoring of automobile
exhaust particulates: methods and costs. Washington, DC: Naval Research Laboratory; NRL
memorandum report 2350. Available from: NTIS, Springfield, VA; AD-738801.
Birnie, S. E.; Noden, F. G. (1980) Determination of tetramethyl- and tetraethyllead vapours in
air following collection on a glass-fibre-iodised carbon filter disc. Analyst (London)
105: 110-118.
Blaszkewicz, M.: Neidhart, B. (1983) A sensitive method for simultaneous determination of
airborne organolead compounds; part 1: chromatographic separation and chemical reaction
detection. Int. J. Environ. Anal. Chem. 14: 11-21.
Brihaye, C.; Duyckaerts, G. (1982) Determination of traces of metals by anodic stripping volt-
ammetry at a rotating glassy carbon ring-disc electrode: part 1. method and instrumenta-
tion with evaluation of some parameters. Anal. Chim. Acta 143: 111-120.
Brihaye, C.; Duyckaerts, G. (1983) Determination of traces of metals by anodic stripping volt-
ammetry at a rotating glassy carbon ring-disc electrode: part 2. Comparison between
linear anodic stripping voltammetry with ring collection and various other stripping
techniques. Anal. Chim. Acta 146: 37-43.
Brown, K. W.; Black, S. C. (1983) Quality assurance and quality control data validation pro-
cedures used for the Love Canal and Dallas lead soil monitoring programs. Environ. Monit.
Asses. 3: 113-122.
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Capar, S. G.; Gajan, R. J.; Madzsar, E.; Albert, R. H.; Sanders, M.; Zyren, J. (1982) Determi-
nation of lead and cadmium in foods by anodic stripping volammetry: II. Collaborative
study. J. Assoc. Off. Anal. Chem. 65: 978-986.
Chau, Y. K.; Wong, P. T. S.; Bengert, G. A.; Kramar, 0. (1979) Determination of tetraalkyllead
compounds in water, sediment, and fish samples. Anal. Chem. 51: 186-188.
Chow, T. J.; Earl, J. L.; Bennet, C. F. (1969) Lead aerosols in marine atmosphere. Environ.
Sci. Technol. 3: 737-740.
Chow, T. J.; Patterson, C. C.; Settle, D. (1974) Occurrence of lead in tuna [letter]. Nature
(London) 251: 159-161.
Code of Federal Regulations. (1982) Ambient air quality surveilance. C. F. R. 40: § 58.
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4-38
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5. SOURCES AND EMISSIONS
5.1 HISTORICAL PERSPECTIVE
The history of global lead emissions has been assembled from chronological records of
deposition in polar snow strata, marine and freshwater sediments, coral skeleton bands, and
the annual rings of trees. These records are important for two reasons. They aid in
establishing natural background levels of lead in air, soils, plants, animals, and humans.
They also place current trends in atmospheric lead concentrations in the perspective of
historical changes. Most chronological records document the sudden increase in atmospheric
lead at the time of the industrial revolution, and a later burst starting in the 1920's when
lead-alkyls were first added to gasoline.
Tree ring analyses are not likely to show the detailed year-by-year chronological record
of atmospheric lead increases. In situations where ring-porous trees (species that retain the
nutrient solution only in the most recent annual rings) grow in heavily polluted areas where
soil lead has increased 100-fold, significant increases in the lead content of tree rings over
the last several decades have been documented. Rolfe (1974) found 4-fold increases in both
rural and urban tree rings using pooled samples from the period of 1910-20 compared to samples
from the period from 1963-73. Symeonides (1979) found a 2-fold increase during a comparable
interval at a high lead site but no increase at a low lead site. Baes and Ragsdale (1981)
found significant post-1930 increases in oak (Quercus) and hickory (Carya) with high lead
exposure, but only in hickory with low lead exposure. Dodge and Gilbert (1984) reported a
chronological increase in lead deposited in the annual bands of coral skeletons near St.
Croix, U.S. Virgin Islands. The 2-fold increase from 1950 to 1980 in the coral at the
relatively unpolluted site appeared to reflect regional or global deposition.
Pond sediment analyses (Shirahata et al., 1980) have shown a 20-fold increase in lead
deposition during the last 150 years in the western United States (Figure 5-1), documenting
not only the increasing use of lead since the beginning of the industrial revolution in that
region, but also the relative fraction of natural vs. anthropogenic lead inputs. Other
studies have shown a similar magnitude of increasing deposition in freshwater sediments
(Christensen and Chien, 1981; Galloway and Likens, 1979; Edgington and Robbins, 1976; Dominik
et al., 1984; Wong et al., 1984), and marine sediments (Ng and Patterson, 1982). The pond and
marine sediments of Shirahata et al. (1980) and Ng and Patterson (1982) also document the
shift in isotopic composition caused by the recent opening of the New Lead Belt in Missouri
(see Section 5.3.3.2) where the ore body has an isotopic composition substantially different
from other ore bodies of the world.
5-1
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Ul
in
i
IM
c
e
u
e
cc
1.0
0.9
0.8
0.7
0.6
0.5
0.4
0.3
0.2
0.1
1750
1775
1800
1825
1850
1875
1900
1925
1950
1975
YEAR
Figure 5-1. Chronological record of the relative increase of lead in snow strata, pond and lake sedi-
ments, marine sediments, and tree rings. The data are expressed as a ratio of the latest year of the
record and should not be interpreted to extend back in time to natural or uncontaminated levels of
lead concentration.
Source: Adapted from Murozumi et al. (1969) . Ng and Patterson (1982) (A), and Rolfe (1974) (•).
-------
Perhaps the best and certainly the most controversial chronological record is that of the
polar ice strata of Murozumi et al. (1969), which extends nearly three thousand years back in
time (Figure 5-1). In a comprehensive review of chronological studies of global pollution in
polar snow and ice, Wolff and Peel (1985) concluded that, although a few samples in the
Greenland study of Murozumi et al. (1969) may have been contaminated, the results are valid
and have been confirmed by later studies (Ng and Patterson, 1981). Intermediate studies that
reported much higher concentrations were probably erroneous.
In Antarctica, lead concentrations in snow and ice are about one tenth of the values from
polar regions in the northern hemisphere. This phenomenon has been attributed to the
restricted interchange in the atmospheric circulation patterns between the northern and
southern hemispheres, and to the fact that 90 percent of the global industrial activity occurs
in the northern hemisphere (Wolff and Peel, 1985). Recent studies by Wolff and Peel (1985)
confirmed the values of 5 pg Pb/g snow found by Boutron and Patterson (1983), repudiating many
previous studies that reported higher values.
It is likely that prehistoric concentrations of lead in snow and ice of Greenland and
Antarctica were a maximum of 1.4 and 1.2 pg/g (Ng and Patterson, 1983; Boutron and Patterson,
1983), while present concentrations are 200 pg/p in Greenland (Murozumi et al., 1969) and 5-6
pg/g in Antarctica (Boutron and Patterson, 1983). Data for Antarctica agree with atmospheric
measurements of Maenhaut et al. (1979), who found air concentrations of 0.000076 ug/m3
suggested by Patterson (1980) and Servant (1982) as the natural lead concentration in the
atmosphere.
In summary, it is likely that atmospheric lead emissions have increased 2000-fold since
the pre-Roman era, that even at this early time the atmosphere may have been contaminated by a
factor of three over natural levels (Murozumi et al. 1969), and that global atmospheric
concentrations have increased dramatically since the 1920's.
The history of global emissions may also be determined from total production of lead, if
the amounts of lead released to the atmosphere during the smelting process, released during
industrial consumption, and emitted from non-lead sources are known. The historical picture
of lead production has been pieced together from many sources by Settle and Patterson (1980)
(Figure 5-2). They used records of accumulated silver stocks to estimate the lead production
needed to support coin production. Until the industrial revolution, lead production was
determined largely by the ability or desire to mine lead for its silver content. Since that
time, lead has been used as an industrial product in its own right, and efforts to improve
smelter efficiency, including control of stack emissions and fugitive dusts, have made lead
production more economical. This improved efficiency is not reflected in the chronological
record because of atmospheric emissions of lead from many other anthropogenic sources,
especially gasoline combustion (see Section 5.3.3). From this knowledge of the chronological
5-3
-------
To*
IU
o
o
c
i
Z 10*
g 10*
10"
i r i i
•SPANISH PRODUCTION
OF SILVER
IN NEW WORLD
EXHAUSTION
OF ROMAN
LEAD MINES
INDUSTRIAL
REVOLUTION
SILVER
PRODUCTION
IN GERMANY
DISCOVERY OF
CUPELLATION
INTRODUCTION
OF COINAGE
RISE AND FALL
OF ATHENS
ROMAN REPUBLIC
AND EMPIRE
H---'I i i i i i
5500 6000 4500 4000 3500 3000 2500 2000 1500 1000 500 0
YEARS BEFORE PRESENT
Figure 5-2. The global lead production has changed historically in response to
major economic and political events. Increases in lead production (note log
scale) correspond approximately to historical increases in lead emissions shown
in Figure 5-1.
Source: Adapted from Settle and Patterson (1980).
record, it is possible to sort out contemporary anthropogenic emissions from natural sources
of atmospheric lead.
5.2 NATURAL SOURCES
Lead enters the biosphere from lead-bearing minerals in the lithosphere through both
natural and man-made processes. Measurements of soil materials taken at 20-cm depths in the
continental United States (Lovering, 1976; Shacklette et al. 1971) show a median lead
concentration of 15 - 16 pg Pb/g soil. Ninety-five percent of these measurements show 30 pg/g
of lead or less, with a maximum sample concentration of 700 pg/g.
5-4
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In natural processes, lead is first incorporated in soil in the active root zone, from
which it may be absorbed by plants, leached into surface waters, or eroded into windborne
dusts (National Academy of Sciences, 1980; Chamberlain, 1970; Patterson, 1965; Chow and
Patterson, 1962).
Natural emissions of lead from volcanoes have been estimated by Nriagu (1979) to be 6400
metric tons (t)/year based on enrichment over crustal abundance. That is, 10 X 109 kg/year of
volcanic dust are produced, with an average lead concentration of 640 ug/g, or 40 times the
crustal abundance of 16 M9/9- The enrichment factor is based on Lepel et al. (1978), who
measured lead in the plume of the Augustine volcano in Alaska. Settle and Patterson (1980)
have calculated emissions of only 1 t/year, based on a measured Pb/S ratio of 2 X 10
(Buat-Menard and Arnold, 1978), and estimated sulfur emissions of 6 X 106 t/year. The
estimate of Settle and Patterson (1980) is more direct, and perhaps more reliable, because it
depends on estimates of sulfur emissions rather than total volcanic dust.
Calculations of natural contributions using geochemical information indicate that natural
sources contribute a relatively small amount of lead to the atmosphere. For example, if the
typical 25 - 40 ug/m3 of rural airborne particulate matter consisted solely of wind-entrained
soils containing 15 (jg/g (and rarely more than 30 ug of lead/g), as cited above, then the
natural contribution to airborne lead would range from 0.0004 to 0.0012 ug/m3. It has been
estimated from geochemical evidence that the natural particulate lead level is less than
0.0005 ug/m3 (National Academy of Sciences, 1980; United Kingdom Department of the
Environment, 1974). In fact, levels as low as 0.000076 ug/m3 have been measured at the South
Pole in Anarctica (Maenhaut et al. , 1979). In contrast, lead concentrations in urban
suspended particulate matter may be as high as 6 ug/m3 (Akland, 1976; U.S. Environmental
Protection Agency, 1979, 1978). Most of this urban particulate lead stems from manmade
sources.
5.3 MANMADE SOURCES
5.3.1 Production
Lead occupies an important position in the U.S. economy, ranking fifth among all metals
in tonnage used. Approximately 85 percent of the primary lead produced in this country is
from native mines; it is often associated with minor amounts of zinc, cadmium, copper,
bismuth, gold, silver, and other minerals (U.S. Bureau of Mines, 1975). Missouri lead ore
deposits account for approximately 80 to 90 percent of the domestic production. Approximately
40 to 50 percent of annual lead production is recovered and eventually recycled.
5-5
-------
5.3.2 Utilization
The 1971-1982 uses of lead are listed by major product category in Table 5-1 (U.S. Bureau
of Mines, 1972-1984). Total utilization averaged approximately 1.29 x 106 t/yr over the
12-year period, with storage batteries and gasoline additives accounting for ~70 percent of
total use. The gasoline antiknock compounds listed in Table 5-1 include additives for both
domestic and import markets. The additive fraction of total lead utilization has decreased
from greater than 18 percent in 1971-1973 to less than 9.5 percent in 1981. Certain products,
especially batteries, cables, plumbing, weights, and ballast, contain lead that is
economically recoverable as secondary lead. This reserve of lead in use is estimated at 3.8
million metric tons. Of the one million metric tons of lead used in commercial products
annually, 0.5 to 0.8 million tons are recovered. Lead used in pigments, gasoline additives,
ammunition, foil, solder, and steel products is widely dispersed and therefore is largely
unrecoverable.
5.3.3 Emissions
Lead or its compounds may enter the environment at any point during mining, smelting,
processing, use, recycling, or disposal. Estimates of the dispersal of lead emissions into
the environment by principal sources indicate that the atmosphere is the major initial
recipient. Estimated lead emissions to the atmosphere are shown in Table 5-2. Mobile and
stationary sources of lead emissions, although found throughout the nation, tend to be
concentrated in areas of high population density, with the exception of smelters. Figure 5-3
shows the approximate locations of major lead mines, primary and secondary smelters and
refineries, and alkyl lead plants (International Lead Zinc Research Organization, 1982).
5.3.3.1 Mobile Sources. The majority of lead compounds found in the atmosphere result from
leaded gasoline combustion. Several reports indicate that transportation sources, which
include light-duty, heavy-duty, and off-highway vehicles, contribute over 80 percent of the
total atmospheric lead (Nationwide [lead] emissions report, 1980, 1979; U.S. Environmental
Protection Agency, 1977). Other mobile sources, including aviation use of leaded gasoline and
diesel and jet fuel combustion, contribute insignificant lead emissions to the atmosphere.
The detailed emissions inventory in Table 5-2 shows that 89 percent of the lead emissions in
the United States are from gasoline combustion. Cass and McRae (1983) assembled emissions
inventory data on the Los Angeles Basin and determined that 83 percent of the fine particle
emissions originated from highway vehicles. Lead is added to gasoline as an antiknock
additive to enhance engine performance in the form of two tetralkyl lead compounds, tetraethyl
and tetramethyl lead (see Section 3.4). Lead is emitted from vehicles primarily in the form
of inorganic particles, although a very small fraction (<10 percent) of lead emissions are
released as volatile organic compounds, i.e., lead alkyls (see Section 6.3.2)
5-6
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TABLE 5-1. U.S. UTILIZATION OF LEAD BY PRODUCT CATEGORY (1971-1980)
(metric tons/yr)
Product category
Storage batteries
Gasoline antiknock
additives
Pigments and ceramics
Ammunition
Solder
Cable coverings
Caulking lead
Pipe and sheet lead
Type metal
Brass and bronze
Bearing metals
Other
TOTAL
1971
616,581
239,666
73,701
79,423
63,502
47.998
27,204
41,523
18,876
18,180
14,771
56,958
1,298,383
1972
661,740
252,545
80,917
76,822
64,659
41,659
20,392
37,592
18,089
17,963
14,435
63,124
1,349,846
1973
697,888
248.890
98,651
73,091
65,095
39,006
18,192
40,529
19,883
20,621
14,201
61,019
1,397,876
1974
772,656
227,847
105,405
78,991
60,116
39,387
17,903
34,238
18,608
20,172
13,250
62,106
1,450,679
1975
634,368
189,369
71,718
68,098
52,011
20,044
12,966
35,456
14,703
12,157
11,051
54,524
1,176,465
1976
746,085
217,508
95,792
66,659
57,448
14,452
11,317
34,680
13,614
14,207
11,851
68,181
1,351,794
1977
858,099
211,296
90,704
62,043
58,320
13,705
8,725
30.861
11,395
15,148
10,873
64,328
1,435,497
1978
879,274
178,473
91,642
55,776
68,390
13,851
9,909
23,105
10,795
16,502
9,510
75,517
1,432,744
1979
814,332
186,945
90,790
53,236
54,278
16,393
8,017
27,618
10,019
18,748
9,630
68,329
1,358,335
1980
645,357
127,903
78,430
48,662
41,366
13,408
5,684
28,393
8,997
13,981
7.808
50,314
1,070,303
1981
770,152
111,367
80,165
49,514
29,705
12,072
5,522
28,184
7,838
13,306
6,922
52,354
1,167,101
1982
704,323
119,234
60,866
44,237
28,500
15,181
4,056
23,838
2,766
11,352
6,133
54,922
1,075,408
1983
806,999
89,118
68,694
43,697
28,490
10,505
3,572
27,261
2,540
10,980
5,844
50,887
1,148,487
1984
865, 547
78,933
76,808
47,828
24,441
12,270
3,966
28,323
2,162
6,954
4,677
55,124
1,207,033
aIncludes additives for both domestic and export markets.
Source: U.S. Bureau of Mines (1972-1984).
-------
TABLE 5-2. ESTIMATED ANTHROPOGENIC LEAD EMISSIONS TO THE ATMOSPHERE FOR THE
UNITED STATES, 1984
Annual (1984) Percentage of
emissions, total U.S.
Source Category (t/yr) emissions
Gasoline combustion 34,881 89.4%
Waste oil combustion
Solid waste disposal
Coal combustion
Oil combustion
Gray iron production
Iron and steel production
Secondary lead smelting
Primary copper smelting
Ore crushing and grinding
Primary lead smelting
Zn smelting
Other metallurgical
Lead alkyl manufacture
Lead acid battery manufacture
Portland cement production
Miscellaneous
Total
781
352
265
115
54
427
278
29
116
1150
116
11
224
112
70
35
39,016a
2.0
0.9
0.7
0.3
0.1
1.1
0.7
0.1
0.3
2.8
0.3
0.1
0.6
0.3
0.2
0.1
100%
Inventory does not include emissions from exhausting workroom air, burning of lead-painted
surfaces, welding of lead-painted steel structures, or weathering of painted surfaces.
Source: Updated from Battye (1983).
5-8
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• MINES (11)
A SMELTERS AND REFINERIES (5)
O SECONDARY SMELTERS AND REFINERIES (39)
• LEAD ALKYL PLANTS (4)
Figure 5-3. Locations of major lead operations in the United States.
Source: International Lead Zinc Research Organization (1985).
-------
Commercial lead antiknock additives of all types contain halogens designated as
scavengers that serve to reduce the accumulation of decomposition products of the lead alkyls
in certain critical areas of the engine combustion chamber. The most commonly used additive
package contains enough ethylene dibromide to tie up all of the lead as PbBr2, and enough
ethylene dichloride to tie up 1.5 times the amount of lead as PbCl2.
The factors which affect both the rate of particulate lead emissions and the
physicochemical properties of the emissions are: lead content of the fuel, other additives,
vehicle fuel economy, the driving speed or conditions, and type of vehicle, as well as design
parameters, maintenance, and ages of the engine, exhaust, and emission control systems. The
major types of vehicles are light-duty (predominantly cars) and heavy-duty (trucks and buses).
The important properties of the particulate emissions include the total amount emitted, the
size distribution of the particles, and the chemical composition of these particles as a
function of particle size. The most commonly used index of particle size is the mass median
aerodynamic diameter (MMAD), which is defined as the point in the size distribution of
particles such that half the mass lies on either side of the MMAD value (National Air
Pollution Control Administration, 1969). Table 5-3 summarizes a recent study estimating the
particulate emission rates and particle composition for light-duty vehicles operated on a
leaded fuel of 1.8 g Pb/gallon (Hare and Black, 1981). Table 5-4 estimates particulate
emission rates for heavy-duty vehicles (trucks) operated on a leaded fuel of 1.8 g Pb/gallon
(Hare and Black, 1981). The lead content of 1.8 g Pb/gallon was chosen to approximate the
lead concentration of leaded gasoline during 1979 (Table 5-5). Another recent study utilizing
similar composite emission factors provides estimates of motor vehicle lead emissions for
large areas (Provenzano, 1978).
The fate of emitted lead particles depends upon their particle size (see Section 6.3.1).
Particles initially formed by condensation of lead compounds in the combustion gases are quite
small (well under 0.1 urn in diameter, see Section 6.3.1) (Pierson and Brachaczek, 1983).
Particles in this size category are subject to growth by coagulation and, when airborne, can
remain suspended in the atmosphere for 7-30 days and travel thousands of miles from their
original source (Chamberlain et al., 1979). Larger particles are formed as the result of
agglomeration of smaller condensation particles and have limited atmospheric lifetimes
(Harrison and Laxen, 1981). The largest vehicle-emitted particles, which are greater than 100
urn in diameter, may be formed by materials flaking off from the surfaces of the exhaust
system. As indicated in Table 5-3, the estimated mass median equivalent diameter of leaded
particles from light-duty vehicles is <0.25 urn, suggesting that such particles with relatively
long atmospheric lifetimes have the potential for long-distance transport. Similar values for
MMAD in automobile exhausts were found in Britain (0.27 pm) (Chamberlain et al. 1979) and
5-10
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TABLE 5-3. LIGHT-DUTY VEHICULAR PARTICULATE EMISSIONS*
Rate or property
Exhaust participate emissions,
Particle mass median equivalent
g/mi (g/km)
diameter, urn
Data by
Pre-1970
0.29 (0.47)
<0.25
vehicle category
1970 & later
without catalyst
0.13 (0.21)
<0.25
Percent of particulate mass as:
Lead (Pb)
Bromine (Br)
Chlorine (Cl)
Trace metals
Carbon (C), total
Sulfate (S042-)
Soluble organics
22 or greater
11 or greater
4 or greater
1
33 or greater
1.3
~30 or less
36 or greater
18 or greater
6 or greater
1 or greater
33 or less
1.3 or greater
-10
*Rate estimates are based on 1.8 g Pb/gal (0.42 g/1) fuel.
Source: Hare and Black (1981).
TABLE 5-4. HEAVY-DUTY VEHICULAR PARTICULATE EMISSIONS*
[g/mi (g/km)]
Heavy-duty category
Particulate emissions by model year
Pre-1970 1970 and later
Medium-duty trucks .
(6,000 to 10,000 Ib)1
Heavy-duty trucks .
(over 10,000 lb)T
0.50 (0.80)
0.76 (1.2)
0.40 (0.64)
0.60 (0.96)
*Rate estimates are based on 1.8 g Pb/gal (0.42 g/1) fuel.
'Gross vehicular weight.
Source: Hare and Black (1981).
5-11
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TABLE 5-5. RECENT AND PROJECTED CONSUMPTION OF GASOLINE LEAD
Gasoline volume
Calendar
year
1975a
1976
1977
1978
1979
1980
1981.
1982°
1983
1984
1985d
1986
1987
1988
1989
1990
10s gal
Total
102.3
107.0
113.2
115.8
111.2
110.8
102.6
98.7
102.4
105.7
100.6
100.3
100.0
99.3
99.0
99.0
Leaded
92.5
87.0
79.7
75.0
68.1
57.5
51.0
52.5
47.5
43.8
32.2
28.8
25.6
22.4
19.2
16.4
Average lead content
g/gai
Pooled
1.62
1.60
1.49
1.32
1.16
0.71
0.59
0.61
0.51
0.44
0.26
0.03
0.03
0.02
0.02
0.02
Leaded
1.81
1.97
2.12
2.04
1.90
1.37
1.19
1.14
1.10
1.05
0.80
0.10
0.10
0.10
0.10
0.10
Total lead
103t
167.4
171.4
168.9
153.0
129.4
78.8
60.7
59.9
52.3
46.0
25.8
2.9
2.6
2.2
1.9
1.6
Air- lead
ug/m3
1.23
1.22
1.20
1.13,.
0.74C
0.66C
0.51C
0.53C
0.40C
0.36C
Data for the years 1975-1981 are taken from U.S. Environmental Protection Agency
(1983).
°Data for 1982-1984 are taken from U.S. Environmental Protection Agency (1985).
:Data from U.S. EPA (1986), discussed in Chapter 7, are the maximum quarterly
average lead levels from a composite of 147 sampling sites. Earlier reports
for the period 1975-78 were based on a different, although comparable group
of sites.
dData for 1985-1990 are estimates taken from F.R. (1985 March 7).
Italy (0.33 urn) (Facchetti and Geiss, 1982). Particles this small deposit by Brownian
diffusion and are generally independent of gravitation (see Section 6.4.1.1).
The size distribution of lead particles is essentially bimodal at the time of exhaust
(Pierson and Brachaczek, 1976, 1983) and depends on a number of factors, including the
particular driving pattern in which the vehicle is used and its past driving history (Ganley
and Springer, 1974; Habibi, 1973, 1970; Ter Haar et al. , 1972; Hirschler and Gilbert, 1964;
Hirschler et al., 1957). As an overall average, it has been estimated that during the
lifetime of the vehicle, approximately 35 percent of the lead contained in'the gasoline burned
by the vehicle is emitted as small particles (<0.25 urn MMAD), and approximately 40 percent is
emitted as larger particles (>10 urn MMAD) (Ter Haar et al., 1972). The remainder of the lead
5-12
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consumed in gasoline combustion is deposited in the engine and exhaust system. Engine deposits
are, in part, gradually transferred to the lubricating oil and removed from the vehicle when
the oil is changed. A flow chart depicting lead-only emissions per gallon of fuel charged
into the engine is shown in Figure 5-4. It is estimated that 10 percent of the lead consumed
during combustion is released into the environment via disposal of used lubricating oil
(Piver, 1977). In addition, some of the lead deposited in the exhaust system gradually flakes
off, is emitted in the exhaust as extremely large particles, and rapidly falls into the
streets and roads where it is incorporated into the dust and washed into sewers or onto
adjacent soil.
Although the majority (>90 percent on a mass basis) of vehicular lead compounds are
emitted as inorganic particles (e.g., PbBrCl), some organolead vapors (e.g., lead alkyls) are
also emitted. The largest volume of organolead vapors arises from the manufacture, transport,
and handling of leaded gasoline. Such vapors are photoreactive and their presence in local
atmospheres is transitory; i.e., the estimated atmospheric half-lives of lead alkyls, under
typical summertime conditions, are less than half a day (Nielsen, 1984). Organolead vapors
are most likely to occur in occupational settings (e.g., gasoline transport and handling
operations, gas stations, parking garages) and have been found to contribute less than 10
percent of the total lead present in the atmosphere (Gibson and Farmer, 1981; National Academy
of Sciences, 1972).
The use of lead additives in gasoline, which increased in volume for many years, is now
decreasing as automobiles designed to use unleaded fuel constitute the major portion of the
fleet (Table 5-1). The decline in the use of leaded fuel is the result of two regulations
promulgated by the U.S. Environmental Protection Agency (F.R., 1973, December 6). The first
required the availability of unleaded fuel for use in automobiles designed to meet federal
emission standards with lead-sensitive emission control devices (e.g., catalytic converters);
the second required a reduction or phase-down of the lead content in leaded gasoline. The
phase-down schedule of lead in gasoline was modified in 1982 (F.R., 1982, October 29),
replacing the 0.5 g/gal standard for the average lead content of all gasoline with a standard
of 1.10 g Pb/gal for leaded gasoline alone, and again in 1985 (F.R., 1985, March 7), calling
for a reduction to 0.5 g Pb/gal leaded gas by July 1985 and 0.1 g Pb/gal leaded gas by
January 1986.
The trend in lead content for U.S. gasolines is shown in Figure 5-5 and Table 5-5. Of
the total gasoline pool, which includes both leaded and unleaded fuels, the average lead
content has decreased 73 percent, from an average of 1.62 g/gal in 1975 to 0.44 g/gal in 1984
(Table 5-5, Figure 5-5). Accompanying the phase-down of lead in leaded fuel has been the
5-13
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LEADED FUEL
(Pb
AUTO
ENGINE
1000 mg 1100%)-
TOTAL MASS OF LEAD
CHARGED INTO THE
ENGINE
TAILPIPE DEPOSITION ^ 18% /
160 mg RETAINED ON
INTERIOR SURFACES OF
ENGINE AND EXHAUST
SYSTEM
V
360 mfl Pb EMITTED
TO ATMOSPHERE AS
LEAD AEROSOL WITH
MASS MEDIAN DIAMETER
OF <0.26 urn. POTENTIAL,
FOR LONG RANGE
TRANSPORT/POLLUTION.
400 mg Pb EMITTED TO
ROADWAY AS PARTICLES
WITH MASS MEDIAN
DIAMETERS >10 ym
LOCALIZED POLLUTION
100 mg Pb RETAINED BY
LUBRICATING OIL
EXHAUST PRODUCTS
•\>7B%(7BOmg TOTAL
Pb EMITTED)
Figure 5-4. Estimated lead-only emissions distribution per gallon of combusted fuel.
5-14
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2.40
2.00 -
1
cri
1.60 -
8
O
HI
3
E
2
1.00 -
0.50 -
0.00
SALES WEIGHTED TOTAL
GASOLINE POOL
(LEADED AND UNLEADED
"AVERAGE")
1976 1976 1977 1978 1979 1980 1981 1982 1983 1984
CALENDAR YEAR
Figure 5-5. Trend in lead content of U.S. gasolines, 1975-1984.
Source. U.S. EPA (1985).
5-15
-------
increased consumption of unleaded fuel, from 10 percent of the total gasoline pool in 1975 to
59 percent in 1984 (Table 5-5 and Figure 5-6). Since 1975, when the catalytic converter was
introduced by automobile manufacturers for automotive exhaust emissions control, virtually all
new passenger cars have been certified on unleaded gasoline (with the exception of a few
diesels and a very few leaded-gasoline vehicles).
Data describing the lead consumed in gasoline and average ambient lead levels (composite
of maximum quarterly values) vs. calendar year are listed in Table 5-5 and plotted in Figure
5-7. The 1975 through 1979 composite quarterly lead averages are based on 105 lead-monitoring
sites, primarily urban. The 1980 through 1984 composite average is based on 147 sites with
valid annual data. Between 1975 and 1984, the lead consumed in gasoline decreased 73 percent
(from 167,400 to 46,000 metric tons) while the corresponding composite maximum quarterly
average of ambient air lead decreased 71 percent (from 1.23 to 0.36 ug/m3). This indicates
that control of lead in gasoline over the past several years has effected a direct decrease in
peak ambient lead concentrations, at least for this group of monitoring sites.
5.3.3.2 Stationary Sources. As shown in Table 5-2 (based on 1984 emission estimates),
primary lead smelting, coal combustion, and combustion of waste oil are the principal
contributors of lead emissions from stationary sources. Coal-fired electric power stations
typically burn 5,000 to 10,000 tons of coal per day. Pulverized coal is mixed with hot air
and passed into a burning chamber or boiler, where the mixture is ignited. Some of the
unburned residue falls to the bottom of the boiler, where it is removed as 'bottom ash1. The
residue that passes through the boiler is called 'fly ash1, much of which is removed by
electrostatic precipitators and other pollution control devices. In a well-designed system,
99.8 percent of the original inorganic mass of the coal is retained by the system. At 10 g
Pb/ton of coal, very little lead would be emitted. However, the remaining 0.2 percent of the
coal mass that is emitted from the stack is highly enriched in lead, compared to the original
coal. Although data on stack emissions of lead are limited, the concentration of lead in fly
ash may provide a reasonable indication of stack lead emissions. Klusek et al. (1983)
reported an enrichment of 6.1 between coal and fly ash. On this basis, a typical power plant
consuming 10,000 tons of coal per day would emit 1.2 kg Pb/day (10,000 t/day x 0.002 x 6.1 x
10 g Pb/t coal). Turner and Lowry (1983) reported enrichment factors of 17 to 75 for
conventional coal-fired power plants in Pennsylvania and New Hampshire.
The manufacture of consumer products such as lead glass, storage batteries, and lead
additives for gasoline also contributes significantly to stationary source lead emissions.
Since 1970, the quantity of lead emitted from the metallurgical industry has decreased
somewhat because of the application of control equipment and the closing of several plants,
particularly in the zinc and pyrometallurgical industries.
5-16
-------
TOTAL GASOLINE SALES
1975 1976
1977 1978 1979 1980
CALENDAR YEAR
Figure 5-6. Trend in U.S. gasoline sales. 1975 1984.
Source: U.S. EPA (1985)
1981 1982 1983 1984
5-17
-------
180
160
140
(0
I
§ 120
100
0 80
UJ
(A
8 60
O
UJ
40
20
i i i i r
CONSUMED IN GASOLINE
AMBIENT AIR
LEAD CONCENTRATION
••*•*
i i
j i i i
1.2
n
1.1 -|
3"
1.0 S
0.9 uj
IU
O
0.8 c
UJ
0.7
0.6
K
ID
o
0.5 I
s
0.4 5
HI
0.3 O
02
o
u
0 1
0
1975 1976 1977 1978 1979 1980 1981 1982 1983 1984
CALENDAR YEAR
Figure 5-7. Lead consumed in gasoline and ambient lead concentrations, 1975-1984.
Source: U.S. Environmental Protection Agency (1985, 1986).
5-18
-------
In the United States, a new source for lead emissions emerged in the mid-1960s with the
opening of the "Viburnum Trend" or "New Lead Belt" in southeastern Missouri. The presence of
eight mines and three accompanying lead smelters in this area makes it the largest lead-
producing district in the world and has moved the United States into first place among the
world's lead-producing nations.
Although some contamination of soil and water occurs as a result of such mechanisms as
leaching from mine and smelter wastes, quantitative estimates of the extent of this
contamination are not available. Spillage of ore concentrates from open trucks and railroad
cars, however, is known to contribute significantly to contamination along transportation
routes. For example, along two routes used by ore trucks in southeastern Missouri, lead
levels in leaf litter ranged from 2000 - 5000 ug/g at the roadway, declining to a fairly
constant 100 - 200 yg/g beyond about 400 ft from the roadway (Wixson et al., 1977).
Another possible source of land or water contamination is the disposal of particulate
lead collected by air pollution control systems. The potential impact on soil and water
systems from the disposal of dusts collected by these control systems has not been quantified.
5.4 SUMMARY
There is no doubt that atmospheric lead has been a component of the human environment
since the earliest written record of civilization. Atmospheric emissions are recorded in
glacial ice strata and pond and lake sediments. The history of these global emissions seems
closely tied to production of lead by industrially oriented civilizations.
Although there are conflicting reports of the amount of lead emitted from natural
sources, even the more liberal estimate (25 X 103 t/year, Nriagu, 1979) is dwarfed by the
global emissions from anthropogenic sources (450 X 103 t/year).
Production of lead in the United States has remained steady at about 1.2 X 106 t/year for
the past decade. The gasoline additive share of this market has dropped from 18 to 6.5
percent during the period 1971 - 1984. The contribution of gasoline lead to total atmospheric
emissions has remained high, at 89 percent, as emissions from stationary sources have
decreased at the same pace as from mobile sources. The decrease in stationary source
emissions is due primarily to control of stack emissions, whereas the decrease in mobile
source emissions is a result of switchover to unleaded gasolines. The decreasing use of lead
in gasoline is projected to continue through 1990.
5-19
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Shirahata, H.; Elias, R. W.; Patterson, C. C.; Koide, M. (1980) Chronological variations in
concentrations and isotopic compositions of anthropogenic atmospheric lead in sediments
of a remote subalpine pond. Geochim. Cosmochim. Acta 44: 149-162.
Symeonides, C. (1979) Tree-ring analysis for tracing the history of pollution: application to
a study in northern Sweden. J. Environ. Qual. 8: 482-486.
Ter Haar, G. L.; Lenane, D. L.; Hu, J. N.; Brandt, M. (1972) Composition, size and control of
automotive exhaust particulates. J. Air Pollut. Control Assoc. 22: 39-46.
Turner, R. R.; Lowry, P. D. (1983) Comparison of coal gasification and combustion residues. J.
Environ. Eng. 109: 305-320.
U. S. Bureau of Mines. (1972-1984) Lead. In: Minerals yearbook; volume I. metals and minerals.
Washington, DC: U. S. Department of the Interior.
U. S. Environmental Protection Agency. (1977) Control techniques for lead air emissions:
volumes I and II. Durham, NC: Office of Air Quality Planning and Standards; EPA report
nos. EPA-450/2-77-012A and EPA-450/2-77-012B. Available from: NTIS, Springfield, VA;
PB80-197544 and PB80-197551.
U. S. Environmental Protection Agency. (1978) Air quality data for metals 1975, from the
National Air Surveillance Networks. Research Triangle Park, NC: Office of Research and
Development; EPA report no. EPA-600/4-78-059. Available from: NTIS, Springfield, VA;
PB-293106.
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U. S. Environmental Protection Agency. (1979) Air quality data for metals 1976, from the
National Air Surveillance Networks. Research Triangle Park, NC: Office of Research and
Development; EPA report no. EPA-600/4-79-054. Available from NTIS, Springfield, VA;
PB80-147432.
U. S. Environmental Protection Agency. (1986) National air quality and emission trends report,
1984. Research Triangle Park, NC: Office of Air Quality Planning and Standards; EPA
report no. EPA 450/4-86-001.
U. S. Environmental Protection Agency. (1985) Summary of lead additive reports for refineries.
Washington, DC: Office of Mobile Source: draft report.
United Kingdom Department of the Environment, Central Unit on Environmental Pollution. (1974)
Lead in the environment and its significance to man. London, United Kingdom: Her
Majesty's Stationery Office; pollution paper no. 2.
Wixson, B. G.; Bolter, E.; Gale, N. L.; Hemphill, D. D.; Jennett, J. C. ; Koirtyohann, S. R.;
Pierce, J. 0.; Lowsley, I. H., Jr.; Tranter, W. H. (1977) The Missouri lead study: an
interdisciplinary investigation of environmental pollution by lead and other heavy metals
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(London) 313: 535-540.
Wong, H. K. T.; Nriagu, J. 0.; Coker, R. D. (1984) Atmospheric input of heavy metals
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Geol. 44: 187-201.
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6. TRANSPORT AND TRANSFORMATION
6.1 INTRODUCTION
This chapter describes the transition from the emission of lead particles into the atmos-
phere to their ultimate deposition on environmental surfaces, i.e., vegetation, soil, house-
hold dust, or water. Lead emissions at the tailpipe are typically around 24,000 ug/m3 (38 x
104 ug Pb/Kg gas x 0.0838 Kg gas/m3 air x 0.75 tailpipe efficiency), while in city air,
ambient lead values are usually between 0.1 and 10 ug/m3 (Dzubay et al., 1979; Reiter et al.,
1977; also see Section 7.2.1.1.1). These reduced concentrations are the result of dilution of
effluent gas with clean air and the removal of particles by wet or dry deposition. Charac-
teristically, lead concentrations are highest in confined areas close to sources and are pro-
gressively reduced by dilution or deposition in air masses more removed from sources.
At any particular location and time, the concentration of lead found in the atmosphere
depends on the proximity to the source, the amount of lead emitted from sources, and the
degree of mixing provided by the motion of the atmosphere. It is possible to describe quanti-
tatively the physics of atmospheric mixing in a variety of ways and, with some limiting
assumptions, to develop simulation models that predict atmospheric lead concentrations. These
models are not sensitive to short-term variations in air motion over a period of weeks or
months because these variations are suppressed by integration over long periods of time.
In highly confined areas such as parking garages or tunnels, atmospheric lead concentra-
tions can be 10-1000 times greater than values measured near roadways or in urban areas. In
turn, atmospheric lead concentrations are usually about 2h times greater in the central city
than in residential suburbs. Rural areas have even lower concentrations.
Because lead emissions in the United States have declined dramatically in the past few
years, the older lead concentration data on which recent dispersion studies are based may seem
irrelevant to existing conditions. Such studies do in fact illustrate principles of at-
mospheric dispersion and are valid when applied to existing concentrations of lead with appro-
priate corrections (see Section 7.2.1.1).
Transformations that may occur during dispersion are physical changes in particle size
distribution, chemical changes from the organic to the inorganic phase, and chemical changes
in the inorganic phase of lead particles. Particle size distribution stabilizes within a few
hundred kilometers of the sources, although atmospheric concentration continues to decrease
with distance. Concentrations of organolead compounds are relatively small (1-6 percent of
total lead) except in special situations where gasoline is handled or where engines are
started cold within confined areas. Ambient organolead concentrations decrease more rapidly
6-1
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than inorganic lead, suggesting conversion from the organic to the inorganic phase during
transport. Inorganic lead appears to convert from lead halides and oxides to lead sulfates.
Lead is removed from the atmosphere by wet or dry deposition. The mechanisms of dry
deposition have been incorporated into models that estimate the flux of atmospheric lead to
the earth's surface. Of particular interest is deposition on vegetation surfaces, since this
lead may be incorporated into food chains. Between wet and dry deposition, it is possible to
calculate an atmospheric lead budget that balances the emission inputs discussed in Section
5.3.3. with deposition outputs.
6.2 TRANSPORT OF LEAD IN AIR BY DISPERSION
6.2.1 Fluid Mechanics of Dispersion
Particles in air streams are subject to the same principles of fluid mechanics as parti-
cles in flowing water (Friedlander, 1977). On this basis, the authors of several texts have
described the mathematical arguments for the mixing of polluted air with clean air (Benarie,
1980; Dobbins, 1979; Pasquill, 1974). If the airflow is steady and free of turbulence, the
rate of mixing is constant along a concentration gradient and is a function of particle size
(Dobbins, 1979). If the steady flow of air is interrupted by obstacles near the ground, tur-
bulent eddies or vortices may be formed. Diffusivity is no longer constant with particle size
and concentration but may be influenced by windspeed, atmospheric stability, and the nature of
the obstacle. By making generalizations of windspeed, stability, and surface roughness, it is
possible to construct models using a variable transport factor called eddy diffusivity (K), in
which K varies in each direction, including vertically. There is a family of K-theory models
that describe the dispersion of particulate pollutants.
The simplest K-theory model, which assumes that the surface is uniform and the wind is
steady (Pasquill, 1974), produces a Gaussian plume, where the concentration of the pollutant
decreases according to a normal or Gaussian distribution in both the vertical and horizontal
directions. Although these models are the basis for most of the air quality simulations per-
formed to date (Benarie, 1980), the assumptions of steady windspeed and smooth surface limit
their use.
Some theoretical approaches, circumventing the constraints of the Gaussian models, have
been adapted for studying long range transport (LRT) (more than 100 km) of pollutants.
Johnson (1981) discusses 35 LRT models developed during the 1970s to describe the dispersion
of atmospheric sulfur compounds. One family of models is based on the conservative volume
element approach, where volumes of air are seen as discrete parcels having conservative meteo-
rological properties, such as water vapor mixing ratio, potential temperature, and absolute
vorticity (Benarie, 1980). The effect of pollutants on these parcels is expressed as a mixing
6-2
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ratio. These parcels of air may be considered to move along a trajectory that follows the ad-
vective wind direction. These models are particularly suitable for dealing with surface
roughness, but they tend to introduce artifact diffusion or pseudodiffusion, which must be
suppressed by calculation (Egan and Mahoney, 1972; Liu and Seinfeld, 1975; Long and Pepper,
1976).
An approach useful for estimating dispersion from a roadway derives from the similarity
approach of Prandtl and Tietjens (1934). A mixing length parameter is related to the distance
traveled by turbulent eddies during which violent exchange of material occurs. This mixing
length is mathematically related to the square root of the shear stress between the atmosphere
and the surface. Richardson and Proctor (1926) formulated these concepts in a law of at-
mospheric diffusion which was further extended to boundary layer concepts by Obukhov (1941).
At the boundary layer, the turbulent eddy grows and its energy decreases with distance away
from the source.
Although physical descriptions of turbulent diffusion exist for idealized circumstances
such as isolated roadways and flat terrain, the complex flow and turbulence patterns of cities
have defied theoretical description. The permeability of street patterns and turbulent eddy
development in street canyons are two major problem areas that make modeling urban atmospheres
difficult. Kotake and Sano (1981) have developed a simulation model for describing air flow
and pollutant dispersion in various combinations of streets and buildings on two scales. A
small scale, 2-20 m, is used to define the boundary conditions for 2-4 buildings and asso-
ciated roadways. These subprograms are combined on a large scale of 50-500 meters. Simula-
tions for oxides of nitrogen show nonlinear turbulent diffusion, as would be expected. The
primary utility of this program is to establish the limits of uncertainty, the first step
toward making firm predictions. It is likely that the development of more complete models of
dispersion in complex terrains will become a reality in the near future.
None of the models described above have been tested for lead. The reason for this is
simple. All of the models require sampling periods of 2 hours or less in order for the sample
to conform to a well-defined set of meteorological conditions. In most cases, such a sample
would be below the detection limits for lead. The common pollutant used to test models is
SOg, which can be measured over very short, nearly instantaneous, time periods. The question
of whether gaseous S02 can be used as a surrogate for particulate lead in these models remains
to be answered.
6-3
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6.2.2 Influence of Dispersion on Ambient Lead Concentrations
Dispersion within confined situations, such as parking garages, residential garages and
tunnels, and away from expressways and other roadways not influenced by complex terrain fea-
tures depends on emission rates and the volume of clean air available for mixing. These fac-
tors are relatively easy to estimate and some effort has been made to describe ambient lead
concentrations that can result under selected conditions. On an urban scale, the routes of
transport are not clearly defined, but can be inferred from an isopleth diagram, i.e., a plot
connecting points of identical ambient concentrations. These plots always show that lead con-
centrations are maximum where traffic density is highest.
Dispersion beyond cities to regional and remote locations is complicated by the facts
that there are no monitoring network data from which to construct isopleth diagrams, that re-
moval by deposition plays a more important role with time and distance, and that emissions
from many different sources converge. Some techniques of source reconciliation are described,
but these become less precise with increasing distance from major sources of lead. Dispersion
from point sources such as smelters and refineries results in a concentration distribution
pattern similar to urban dispersion, although the available data are notably less abundant.
6.2.2.1 Confined and Roadway Situations. Ingalls and Garbe (1982) used a variety of box and
Gaussian plume models to calculate typical levels of automotive air pollutants that might be
present in microscale (within 100 m of the source) situations with limited ventilation, such
as garages, tunnels, and street canyons. Table 6-1 shows a comparison of six exposure situa-
tions, recomputed for a flat-average lead emission factor of 6.3 mg/km for roadway situations
and 1.0 mg/min for garage situations. The roadway emission factor chosen corresponds roughly
to values chosen by Dzubay et al. (1979) and Pierson and Brachaczek (1976) scaled to 1979
lead-use statistics. The parking garage factor was estimated from roadway factors by correc-
tion for fuel consumption (Ingalls and Garbe, 1982).
Confined situations, with low air volumes and little ventilation, allow automotive pollu-
tant concentrations to reach one to three orders of magnitude higher than are found in open
air. Thus, parking garages and tunnels are likely to have considerably higher ambient lead
concentrations than are found in expressways with high traffic density or in city streets.
Purdue et al. (1973) found total lead levels of 1.4-2.3 ug/m3 in five of six U.S. cities in
1972. In similar samples from an underground parking garage, total lead was 11-12 ug/m3.
Vaitkus et al. (1974) developed a model for the transport of automotive lead that predicted an
exponential decrease in air lead concentrations with distance, up to 100 m downwind from the
roadway. Dzubay et al. (1979) found lead concentrations of 4-20 (jg/m3 in air over Los
Angeles freeways in 1976; at nearby sites off the freeways, concentrations of 0.3-4.7 pg/m3
were measured.
6-4
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TABLE 6-1. SUMMARY OF MICROSCALE CONCENTRATIONS
Air lead
concentration
Situation (ug/m3)
Residential garage (1 mg Pb/min)
Typical (30 second idle time) 80
Severe (5 min idle time) 670
Parking garage (1 mg Pb/min)
Typical 40
Severe 560
Roadway tunnel (6.3 mg Pb/km)
Typical 11
Severe 29
Street canyon (sidewalk receptor) (6.3 mg Pb/km)
Typical a) 800 vehicles/hr 0.4
b) 1,600 vehicles/hr 0.9
Severe a) 800 vehicles/hr 1.4
b) 1,600 vehicles/hr 2.8
On expressway (wind: 315 deg. rel., 1 m/sec) (6.3 mg Pb/km)
Typical 2.4
Severe 10
Data are recalculated from Ingalls and Garbe (1982) using 1979 lead emission factors. They
show that air lead concentrations in a garage or tunnel can be two or three orders of magni-
tude higher than on streets or expressways. Typical conditions refer to neutral atmospheric
stability and average daily traffic volumes. Severe conditions refer to maximum hourly
traffic volume with atmospheric inversion. Emission rates are given in parentheses.
Tiao and Hillmer (1978) and Ledolter and Tiao (1979) have analyzed 3 years (1974-1977) of
ambient air lead data from one site on the San Diego Freeway in Los Angeles, California.
Particulate lead concentrations were measured at five locations: in the median strip and at
distances of 8 and 30-35 m from the road edge on both sides of the road. Average lead con-
centrations at the 35 meter point were two- to fourfold lower than at the 8 m location (Tiao
and Hillmer, 1978). An empirical model involving traffic count and traffic speed, which are
related to road emissions, required only windspeed as a predictor of dispersion conditions.
Witz et al. (1982) found that meteorological parameters in addition to windspeed, such as
inversion frequency, inversion duration, and temperature, correlate well with ambient levels
of lead. At a different site near the San Diego freeway in Los Angeles, monthly ambient
particulate lead concentrations and meteorological variables were measured about 100 meters
6-5
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from the roadway through 1980. Multiple linear regression analysis showed that temperature at
6 AM, windspeed, wind direction, and a surface-based inversion factor were important variables
in accurately predicting monthly average lead concentrations. In this data set, lead values
for December were about fivefold higher than those measured in the May to September summer
season, suggesting that seasonal variations in wind direction and the occurrence of surface-
based inversions favor high winter lead values. Unusually high early morning temperatures and
windspeed during the winter increased dispersion and reduced lead concentration.
In a study of a newly constructed freeway near Melbourne, Australia, Clift et al. (1983)
found that lead concentrations in the top centimeter of soil one meter from the road edge in-
creased in lead concentration from 60 ug/g in March 1974 to 1250 ug/g in November 1980. The
traffic density in November 1980 was 37,000 vehicles per day and the typical pattern of lead
concentrations decreasing exponentially with distance from the road edge had developed. At
4-6 meters from the road edge, lead concentrations decreased to constant values, although
these values were significantly higher than the pre-1974 concentrations.
In Philadelphia, a recent study of dispersion away from a major highway showed the zone
of influence may extend farther downwind than previously expected (Burton and Suggs, 1984).
The Philadelphia Roadway study was designed to measure the vertical (15 m) and horizontal
(175 m) dispersion of large and small particles (Figure 6-1). Horizontally, air concentra-
tions decreased exponentially at 2 m height for fine, coarse, and total Pb according to the
following equations:
Coarse Pb C = 0.187 - (0.029 x InD)
Fine Pb C = 0.715 - (0.106 x InO)
Total Pb C = 0.903 - (0.135 x InD)
where C is the concentration of lead in air (ug/m3) at the downwind distance, D (m), measured
from the edge of the road. The numerical coefficients are specific for this site, and were
found to vary with windspeed and traffic density.
Vertical profiles showed decreasing lead concentrations with increasing height for
coarse, fine, and total particulate lead at 5 and 25 m downwind, although the effect was less
pronounced at 25 m. Bullin et al. (1985) found similar results in Houston, somewhat tempered
by greater mixing due to the presence of tall buildings.
6.2.2.2 Dispersion of Lead on an Urban Scale. In cities, air pollutants, including lead,
that are emitted from automobiles tend to be highest in concentration in high traffic areas.
Most U.S. cities have a well-defined central business district (CBD) where lead concentrations
are highest. To illustrate the dispersion of lead experienced in cities, two cases are
presented below.
6-6
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I I I I I I
220mr \ 0
UPWIND CENTER OF
(BACKGROUND) ROADWAY
20
40
60
80 100 120 140 160 180
HORIZONTAL DOWNWIND DISTANCE (2 m HEIGHT), m
20
15
i
UJ
1
10
uj 5
I III
5 m DOWNWIND
T
I
I I
i i r r
25 m DOWNWIND
I I
UPWIND Pb
CONCENTRATION
COARSE = 0.022
FINE = 0.074
TOTAL = 0.096
I I I I I I I
0.1 0.2 0.3 0.4 0.5 0.6 0.7 0 0.1 0.2 0.3 0.4 0.5 0.6 0.7
Pb MASS CONCENTRATION, yg/m»
Figure 6-1. Vertical and horizontal distribution of lead downwind
from a roadway in Philadelphia, PA.
Source: Burton and Suggs (1984).
6-7
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For the South Coast Basin of Southern California, the area of high traffic density is
more widespread than is characteristic of many cities. Ambient concentrations of lead tend to
be more uniform. For example, Figures 6-2 and 6-3 show the average daily traffic by grid
square and the contour plots of annual average lead concentration, respectively, for 1969
(Kawecki, 1978). In addition, Figure 6-3 shows annual average lead measured at nine sites in
the basin for that year. It is clear that the central portion had atmospheric particulate
lead concentrations of about 3 ug/m3; the outer areas were in the range of 1-2 ug/m3.
Reiter et al. (1977) have shown similar results for the town of Fort Collins, Colorado,
for a 5.5-hr period in May of 1973. In that study, modeling results showed maximum lead con-
centrations in the center of town around 0.25 ug/m3, which decreased to 0.1 ug/m3 in the
outermost region. Presumably, still lower values would be found at more remote locations.
Apparently, then, lead in the air decreases 2- to 3-fold from maximum values in center
city areas to well populated suburbs, with a further 2-fold decrease in the outlying areas.
These modeling estimates are generally confirmed by measurement in the cases cited above and
in the data presented in Section 7.2.1.
6.2.2.3 Dispersion from Smelter and Refinery Locations. The 11 mines and 5 primary smelters
and refineries shown in Figure 5-3 are not located in urban areas. Most of the 39 secondary
smelters and refineries are likewise non-urban. Consequently, dispersion from these point
sources should be considered separately, but in a manner similar to the treatment of urban
regions. In addition to lead concentrations in air, concentrations in soil and on vegetation
surfaces are often used to determine the extent of dispersion of plumes from smelters and
refineries. In a study of smelters in Missouri, Dorn et al. (1976) found that 66 percent of
the mass of lead was on particles smaller than 4.7 urn on a farm near a smelter (800 m from the
smelter stack), whereas 73 percent were smaller than 4.7 urn on the control farm. These
authors also noted seasonal differences in particle size distributions, with larger differ-
ences between the test and control farms during the winter than the spring or summer.
6.2.2.4 Dispersion to Regional and Remote Locations. Beyond the immediate vicinity of urban
areas and smelter sites, lead in air declines rapidly to concentrations of 0.1 to 0.5 ug/m3.
Two mechanisms responsible for this change are dilution with clean air and removal by deposi-
tion (Section 6.4). In the absence of monitoring networks that might identify the sources of
lead in remote areas, two techniques of source identification have been used. Vector gradient
analysis was attempted by Everett et al. (1979) and source reconciliation has been reported by
Sievering et al. (1980) and Cass and McRae (1983). A third technique, isotopic composition,
has been used to identify anthropogenic lead in air, sediments, soils, plants, and animals in
urban, rural, and remote locations (Chow et al., 1975). Whereas this technique can often
identify the source of lead, it has not yet been used to determine the mechanism of transport.
6-8
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Figure 6-2. Spatial distribution of surface street and freeway traffic in the Los
Angeles Basin I103 vehicle miles traveled/day) for 1979.
Source: Kawecki (1978).
6-9
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KEY TO CONTOUR CONCENTRATIONS
Figure 6-3. Annual average suspended lead concentrations for 1969 in the Los Angeles Basin,
calculated from the model of Cass (1975). The white zones between the patterned areas are
transitional zones between the indicated concentrations.
Source: Kawecki (1978).
6-10
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In vector gradient analysis, the sampler is oriented to the direction of the incoming
wind vector, and samples are taken only during the time the wind is within a 30° arc of that
vector. Other meteorological data are taken continuously. As the wind vector changes, a dif-
ferent sampler is turned on. A 360° plot of concentration vs. wind direction gives the direc-
tion from which the pollutant arrives at that location. Only one report of the use of this
technique for lead occurs in the literature (Everett et al., 1979), and analysis of this
experiment was complicated by the fact that in more than half the samples, the lead concentra-
tions were below the detection limit. The study was conducted at Argonne National Laboratory
and the results reflected the influence of automobile traffic east and northeast of this loca-
tion.
Source reconciliation is based on the concept that each type of natural or anthropogenic
emission has a unique combination of elemental concentrations. Measurements of ambient air,
properly weighted during multivariate regression analysis, should reflect the relative amount
of pollutant derived from each of several sources (Stolzenburg et al., 1982). Sievering et
al. (1980) used the method of Stolzenberg et al. (1982) to analyze the transport of urban air
from Chicago over Lake Michigan. They found that 95 percent of the lead in Lake Michigan air
could be attributed to various anthropogenic sources, namely auto emissions, coal fly ash,
cement manufacture, iron and steel manufacture, agricultural soil dust, construction soil
dust, and incineration emissions. This information alone does not describe transport pro-
cesses, but the study was repeated for several locations to show the changing influence of
each source.
Cass and McRae (1983) used source reconciliation in the Los Angeles Basin to interpret
1976 NFAN data (see Sections 4.2.1 and 7.2.1.1) based on emission profiles from several
sources. They developed a chemical element balance model, a chemical tracer model, and a
multivariate statistical model. The chemical element balance model showed that 20 to 22 per-
cent of the total suspended particle mass could be attributed to highway sources. The chemi-
cal tracer model permitted the lead concentration alone to represent the highway profile,
since lead comprised about 12 percent of the mass of the highway generated aerosol. The
multivariate statistical model used only air quality data without source emission profiles to
estimate stoichiometric coefficients of the model equation. The study showed that single
element concentrations can be used to predict the mass of total suspended particles.
Pacyna et al. (1985) used a receptor-oriented Lagrangian model to predict air concentra-
tions in Spitsbergen, Norway, based on estimated emissions from the U.S.S.R. Compared to mea-
sured concentrations, the model was accurate for some metals, but overestimated the air
concentration of lead by an average factor of 1.8. The consistent pattern in the ratio of
estimated to measured air concentration led the authors to suggest that a more accurate esti-
mate of lead emissions might correct the discrepancy.
6-11
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A type of source reconciliation, chemical mass balance, has been used for many years by
geochemists in determining the anthropogenic influence on the global distribution of elements.
Two studies that have applied this technique to the transport of lead to remote areas are
Murozumi et al. (1969) and Shirahata et al. (1980). In these studies, the influence of
natural or crustal lead was determined by mass balance, and the relative influence of anthro-
pogenic lead was established. In the Shirahata et al. (1980) study, the influence of anthro-
pogenic lead was confirmed quantitatively by analysis of isotopic compositions in the manner
of Chow et al. (1975).
Harrison and Williams (1982) determined air concentrations, particle size distributions,
and total deposition flux at one urban and two rural sites in England. The urban site, which
had no apparent industrial, commercial, or municipal emission sources, had an air-lead concen-
tration of 3.8 ug/m3, whereas the two rural sites were about 0.15 ug/m3. The average particle
size became smaller toward the rural sites, as the mass median aerodynamic diameter (MMAD)
shifted downward from 0.5 urn to 0.1 pm. The total deposition flux will be discussed in Sec-
tion 6.4.2.
Knowledge of lead concentrations in the oceans and glaciers provides some insight into
the degrees of atmospheric mixing and long range transport. Tatsumoto and Patterson (1963),
Chow and Patterson (1966), and Schaule and Patterson (1980) measured dissolved lead concentra-
tions in sea water in the Mediterranean, in the Central North Atlantic (near Bermuda), and in
the northeast Pacific, respectively. The profile obtained by Schaule and Patterson (1980) is
shown in Figure 6-4. Surface concentrations in the Pacific (14 ng/kg) were found to be higher
than those of the Mediterranean or the Atlantic, decreasing abruptly with depth to a relative-
ly constant level of 1-2 ng/kg. The vertical gradient was found to be much less in the
Atlantic (Figure 6-5). Tatsumoto and Patterson (1963) had earlier estimated an average sur-
face lead concentration of 200 ng/kg in the northern hemispheric oceans. Chow and Patterson
(1966) revised this estimate downward to 70 ng/kg. Below the mixing layer, there appears to
be no difference between lead concentrations in the Atlantic and Pacific. These studies are
significant in that they show that seawater concentrations during prehistoric times (below the
mixing layer) were constant and much lower than modern seawater concentrations at the surface.
From these data, it is possible to calculate present and prehistoric atmospheric inputs to the
oceans (Schaule and Patterson, 1980), and by inference, the prehistoric concentrations of lead
in air. They estimated the present inputs are 60-68 ng/cm2-yr, which is 10-20 times the pre-
historic rate.
Wiersma and Davidson (1985) have reviewed published data on trace metal concentrations
(including lead) in the atmosphere at remote northern and southern hemispheric sites. The
natural sources for such atmospheric trace metals include the oceans and the weathering of the
6-12
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1000
S 2000
I
x
0.
D 3000
4000
6000
TvlI
-//I
• DISSOLVED Pb
D PARTICIPATE Pb
I I I I I I I I
024 6 S 10 12 14 16 0
CONCENTRATION, ng Pfa/kg
Figure 6-4. Profile of lead concentrations in the
central northeast Pacific. Values below 1000 m
are an order of magnitude lower than reported by
Tatsumoto and Patterson (1963) and Chow and
Patterson 11966).
Source: Schaule and Patterson (1980).
1000
2000
3000
4000
O ATLANTIC (BERMUDA)
D MEDITERRANEAN
40°39 N OS1>48 E
A PACIFIC
29°13'N. 117°37'W
01 02 03
LEAD IN SEA WATER, wg kg
04
Figure 6-5. Lead concentration profiles in oceans show exten-
sive contamination above the mixing layer l~ 1000 m)
Source: Chow and Patterson 11966).
6-13
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earth's crust, while the anthropogenic source is particulate air pollution. Enrichment fac-
tors for concentrations relative to standard values for the oceans and the crust were calcula-
ted (Table 6-2); the crustal enrichment factors for the northern and southern hemispheres sug-
gest that 90 percent of the particulate pollutants in the global troposphere are injected in
the northern hemisphere. Since the residence times for particles in the troposphere are much
less than the interhemi spheric mixing time (Poet et al., 1972), it is unlikely that signifi-
cant amounts of particulate pollutants can migrate from the northern to the southern hemi-
sphere via the troposphere; however, this does not rule out stratospheric transfer.
TABLE 6-2. ENRICHMENT OF ATMOSPHERIC AEROSOLS OVER CRUSTAL ABUNDANCE (EF t ,)
IN REMOTE AREAS OF THE NORTHERN AND SOUTHERN HEMISPHERES cruswl
Element
Al
Si
Fe
Mn
Ca
Co
V
Cr
Cu
Zn
Sb
Pb
Cd
Se
Remote air concentration
range , ng/m3
0.3-1200
21-3900
0.25-660
0.0067-190
1.9-1600
0.0017-1.0
0.001-1.5
0.01-7.0
0.06-110
0.035-110
0.002-0.9
0.027-97
0.02-2.2
0.006-1.4
Global
EFcrustal
1.0
0.84
1.3
1.5
1.8
1.9
3.3
3.6
25
50
211
320
1100
3500
Remote
continental
EFcrustal
1.0
0.7 .
1.5(N)D
l.O(S)
2.0(N)
l.O(S)
1.5
1.5(N)
0.9(S)
1.5
6.0(N)
l.O(S)
20
80
500
2000(N)
80(S)
2000
1000
Remote
marine
EFcrustal
1.0
0.7
2.5(N)
l.O(S)
3.0(N)
l.O(S)
8.0
4.0(N)
0.9(S)
15(N)
1.5(5)
20(N)
l.O(S)
150
400
2000
2000(N)
150(S)
5000
6000
aSee text for explanation of the relationship between air concentration and
(N) = northern hemisphere; (S) = southern hemisphere.
Source: data from Wiersma and Davidson (1985).
6-14
-------
I
o
iu
1 ..........
AGE OF SAMPLES
Figure 6-6. Lead concentration profile in snow strata
of Northern Greenland.
Source: Murozumi et al. (1969).
Murozumi et al. (1969) have shown that long range transport of lead particles emitted
from automobiles has significantly polluted the polar glaciers. They collected samples of
snow and ice from Greenland and the Antarctic. As shown in Figure 6-6, they found that the
concentration of lead in Greenland varied inversely with the geological age of the sample.
The authors attribute the gradient increase after 1750 to the Industrial Revolution and the
accelerated increase after 1940 to the increased use of lead alkyls in gasoline. The most
recent levels found in the Antarctic snows (not shown on Figure 6-6) were less than those
found in Greenland by a factor of 10 or more. Before 1940, the concentrations in the Antarc-
tic were below the detectable level (<0.001 ug/kg) and have risen to 0.2 ug/kg in recent snow.
Evidence from remote areas of the world suggests that lead and other fine particle
components are transported substantial distances, up to thousands of kilometers, by general
weather systems. The degree of surface contamination of remote areas with lead depends both
6-15
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on weather influences and on the degree of air contamination. However, even in remote areas,
man's primitive activities can play an important role in atmospheric lead levels. Davidson et
al. (1982) have shown that there are significant levels of fine particle lead, up to 0.5
(jg/m3, in remote villages in Nepal. The apparent source is combustion of dried yak dung,
which contains small amounts of naturally occurring lead derived from plant life in those
remote valleys.
6.3 TRANSFORMATION OF LEAD IN AIR
6.3.1 Particle Size Distribution
Whitby et al. (1975) placed atmospheric particles into three different size regimes: the
nuclei mode (<0.1 urn), the accumulation mode (0.1-2 urn) and the large particle mode (>2
|jm). At the source, lead particles are generally in the nuclei and large particle modes.
Large particles are removed by deposition close to the source and particles in the nuclei mode
diffuse to surfaces or agglomerate while airborne to form larger particles of the accumulation
mode. Thus it is in the accumulation mode that particles are dispersed great distances.
Pierson and Brachaczek (1983) reported particle size distributions for ambient air that
were skewed farther to the right (more large particles) than in a roadway tunnel, where
vehicle exhaust must be dominant (Figure 6-7). The large particles may have been deposited in
the roadway itself and small particles may have agglomerated during transport away from the
roadway (see Section 5.3.3.1). Since 40 to 1,000 urn particles are found in gutter debris,
deposition of large particles appears confirmed (Pierson and Brachaczek, 1976, 1983).
Particle size distributions reported by Huntzicker et al. (1975) show bimodal distribu-
tions for on-roadway samples, with peak mass values at about 0.1 and 10 urn. For off-roadway
Pasadena samples, there is no evidence of bimodality and only a broad maximum in lead mass
between 0.1 and 1 urn.
In cities or in rural areas, there is a remarkable consistency in lead particle size
range. For example, Robinson and Ludwig (1964) report cascade impactor MMAD values for lead
ranging from 0.23 to 0.3 urn in six U.S. cities and three rural areas. Stevens et al. (1978)
have reported dichotomous sampler data for six U.S. cities, as shown in Table 6-3, and Stevens
et al. (1980, 1982) have reported similar results for remote locations. Virtually every other
study reported in the literature for Europe, South America, and Asia has come to the conclu-
sion that ambient urban and rural air contains predominantly fine particles (Cholak et al.,
1968; De Jonghe and Adams, 1980; Durando and Aragon, 1982; Lee et al., 1968; Htun and
Ramachandran, 1977). The size distributions of lead-bearing particles in ambient air from
several global locations are discussed further in Section 7.2.1.3.1 and shown in Figure 7-5.
6-16
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a
•a
0.6
0.5
0.4
0.3
0.2
0.1
0
0.7
0.6
0.5
0.4
0.3
0.2
0.1
1 I I \ I I
AMBIENT
AEROSOL Pb
1
1 1
VEHICLE
AEROSOL Pb
0.01 0.02 0.05 0.1 0.2 0.5 1 2 5 10 20
AERODYNAMIC DIAMETER (dp), pm
50 100
Figure 6-7. Typical airborne mass size distribution patterns for
ambient and vehicle aerosol lead. AC represents the airborne lead
concentrations in each size range. Cps the total airborne lead
concentration in all size ranges, and dp is the aerodynamic particle
diameter.
Source: Data from Pierson and Brachaczek (1983).
6-17
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TABLE 6-3. DISTRIBUTION OF LEAD IN TWO SIZE FRACTIONS AT SEVERAL SITES IN THE UNITED STATES
Location
New York, NY
Philadelphia, PA
South Charleston, WV
St. Louis, MO
Portland, OR
Glendora, CA
Average
Date
2/1977
2-3/1977
4-8/1976
12/1975
12/1977
3/1977
Fine3
1.1
0.95
0.62
0.83
0.87
0.61
Coarse3
0.18
0.17
0.13
0.24
0.17
0.09
F/C ratio
6.0
5.6
4.6
3.4
5.0
6.7
5.2
aData are in ug/m3.
Source: Stevens et al. (1978).
The data in Table 6-3 indicate that there is about five times more lead associated with
small particles than large particles in urban atmospheres. It appears that lead particle size
distributions are stabilized close to roadways and remain constant with transport into remote
environments (Gillette and Winchester, 1972).
6.3.2 Organic (Vapor Phase) Lead in Air
Small amounts of lead additives may escape to the atmosphere by evaporation from fuel
systems or storage facilities. Tetraethyllead (TEL) and tetramethyllead (TML) photochemically
decompose when they reach the atmosphere (Huntzicker et al., 1975; National Air Pollution
Control Administration, 1965). The lifetime of TML is longer than that of TEL. Laveskog
(1971) found that transient peak concentrations of organolead up to 5,000 MS/m3 in exhaust gas
may be reached in a cold-started, fully choked, and poorly tuned vehicle. If a vehicle with
such emissions were to pass a sampling station on a street where the lead level might typical-
ly be 0.02-0.04 ug/m3, a peak of about 0.5 ug/m3 could be measured as the car passed by. The
data reported by Laveskog were obtained with a procedure that collected very small (100 ml),
short-time (10 min) air samples. Harrison et al. (1975) found levels as high as 0.59 ug/m3
(9.7 percent of total lead) at a busy gasoline service station in England. Grandjean and
Nielsen (1979), using GC-MS techniques, found elevated levels (0.1 ug/m3) of TML in city
streets in Denmark and Norway. These authors attributed these results to the volatility of
TML compared with TEL.
6-18
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A number of studies have used gas absorbers behind filters to trap vapor-phase lead com-
pounds (see Section 4.2.2.5). Because it is not clear that all the lead captured in the back-
up traps is, in fact, in the vapor phase in the atmosphere, "organic" or "vapor phase" lead is
an operational definition in these studies. Purdue et al. (1973) measured both particulate
and organic lead in atmospheric samples. They found that the vapor phase lead was about 5
percent of the total lead in most samples (see Section 5.3.3.1). The results are consistent
with the studies by Huntzicker et al. (1975) who reported an organic component of 6 percent of
the total airborne lead in Pasadena for a 3-day period in June, 1974, and by Skogerboe (1976),
who measured fractions in the range of 4 to 12 percent at a site in Fort Collins, Colorado. It
is noteworthy, however, that in an underground garage, total lead concentrations were approxi-
mately five times greater than those in ambient urban atmospheres, and the organic lead
increased to approximately 17 percent.
Harrison et al. (1979) report typical organolead percentages in ambient urban air of 1-6
percent. Rohbock et al. (1980) reported higher fractions, up to 20 percent, but the data and
interpretations have been questioned by Harrison and Laxen (1981). Rohbock et al. (1980) and
De Jonghe and Adams (1980) report one to two orders of magnitude decrease in organolead con-
centrations from the central urban areas to residential areas. A review by Nielsen (1984)
documents the concentrations of organolead in partially enclosed areas such as gas stations,
parking garages, car repair shops, and tunnels, and in open urban and rural areas in the U.S.
and Europe. Mean concentrations varied from 0.15 to 3.5 ug organolead/m3 in enclosed areas
and 0.014 to 0.47 in open urban areas.
6.3.3 Chemical Transformations of Inorganic Lead in Air
Lead is emitted into the air from automobiles as lead halides and as double salts with
ammonium halides (e.g., PbBrCl • 2NH4C1). From mines and smelters, PbS04, PbO'PbS04, and PbS
appear to be the dominant species. In the atmosphere, lead is present mainly as the sulfate
with minor amounts of halides. It is not completely clear just how the chemical composition
changes in transport.
Biggins and Harrison (1978, 1979) have studied the chemical composition of lead particles
in exhaust and in city air in England by X-ray diffractometry. These authors reported that
the dominant exhaust forms were PbBrCl, PbBrCl'2NH4C1, and ot-2PbBrCl'NH4Cl, in agreement with
the earlier studies of Hirschler and Gilbert (1964) and Ter Haar and Bayard (1971).
At sampling sites in Lancaster, England, Biggins and Harrison (1978, 1979) found
PbS04-(NH4)2S04, and PbS04'(NH4)2BrCl together with minor amounts of the lead halides and
double salts found in auto exhaust. These authors suggested that emitted lead halides react
with acidic gases or aerosol components (S02 or H2S04) on filters to form substantial levels
6-19
-------
2-
of sulfate salts. It is not clear whether reactions with S04 occur in the atmosphere or on
the sample filter.
The ratio of Br to Pb is often cited as an indication of automotive emissions. From the
mixtures commonly used in gasoline additives, the mass Br/Pb ratio should be 0.4-0.5 (Pierson
and Brachaczek, 1976, 1983; Dzubay et al. , 1979; Dietzmann et al. , 1981; Lang et al., 1981).
However, several authors have reported loss of halide, preferentially bromine, from lead salts
in atmospheric transport (Dzubay and Stevens, 1973; Pierrard, 1969; Ter Haar and Bayard,
1971). Both photochemical decomposition (Lee et al., 1971; Ter Haar and Bayard, 1971) and
acidic gas displacement (Robbins and Snitz, 1972) have been postulated as mechanisms. Chang
et al. (1977) have reported only very slow decomposition of lead bromochloride in natural sun-
light; currently the acid displacement of halide seems to be the most likely mechanism.
O'Connor et al. (1977) have compared roadside and suburban-rural aerosol samples from western
Australia and reported no loss in bromine; low levels of S02 and sulfate aerosol could account
for that result. Harrison and Sturges (1983) warn of several other factors that can alter the
Br/Pb ratio. Bromine may pass through the filter as hydrogen bromide gas, lead may be
retained in the exhaust system, or bromine may be added to the atmosphere from other sources,
such as marine aerosols. They concluded that Br/Pb ratios are only crude estimates of automo-
bile emissions, and that this ratio would decrease with distance from the highway from 0.39 to
0.35 at less proximate sites to 0.25 in suburban residential areas. For an aged aerosol, the
Br/Pb mass ratio is usually about 0.22.
Habibi et al. (1970) studied the composition of auto exhaust particles as a function of
particle size. Their main conclusions follow:
1. Chemical composition of emitted exhaust particles is related to particle size.
a. Very large particles, greater than 200 urn, have a composition similar to
lead-containing material deposited in the exhaust system, confirming that
they have been emitted from the exhaust system. These particles contain
approximately 60 to 65 percent lead salts, 30 to 35 percent ferric oxide
(Fe203), and 2 to 3 percent soot and carbonaceous material. The major
lead salt is lead bromochloride (PbBrCl), with (15 to 17 percent) lead
oxide (PbO) occurring as the 2PbO-PbBrCl double salt. Lead sulfate and
lead phosphate account for 5 to 6 percent of these deposits. (These
compositions resulted from the combustion of low-sulfur and low-phosphorus
fuel.)
b. PbBrCl is the major lead salt in particles of 2 to 10 urn equivalent diame-
ter, with 2PbBrCl'NH4C1 present as a minor constituent.
c. Submicrometer-sized lead salts are primarily 2PbBrCl-NH4Cl.
6-20
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2. Lead-halogen molar ratios in particles of less than 10 urn MMAD indicate that
much more halogen is associated with these solids than the amount expected from
the presence of 2PbBrCl-NH4C1, as identified by X-ray diffraction. This is
particularly true for particles in the 0.5 to 2 pm size range.
3. There is considerably more soot and carbonaceous material associated with
fine-mode particles than with coarse mode particles re-entrained after having
been deposited following emission from the exhaust system. This carbonaceous
material accounts for 15 to 20 percent of the fine particles.
4. Particulate matter emitted under typical driving conditions is rich in carbona-
ceous material. There is substantially less material emitted under continuous
hot operation.
5. Only small quantities of 2PbBrCl-NH4Cl were found in samples collected at the
tailpipe from the hot exhaust gas. Its formation therefore takes place primar-
ily during cooling and mixing of exhaust with ambient air.
Foster and Lott (1980) used X-ray diffractometry to study the composition of lead com-
pounds associated with ore handling, sintering, and blast furnace operations around a lead
smelter in Missouri. Lead sulfide was the main constituent of those samples associated with
ore handling and fugitive dust from open mounds of ore concentrate. The major constituents
from sintering and blast furnace operations appeared to be PbS04 and PbO«PbS04, respectively.
6.4 REMOVAL OF LEAD FROM THE ATMOSPHERE
Before atmospheric lead can have any effect on organisms or ecosystems, it must be trans-
ferred from the air to a surface by wet or dry deposition.
6.4.1 Dry Deposition
6.4.1.1 Mechanisms of Dry Deposition. The theory and mechanics of particle deposition from
the atmosphere to smooth surfaces are fairly well understood (Friedlander, 1977). Transfer by
dry deposition requires that the particle move from the main airstream through the boundary
layer to a surface. The boundary layer is defined as the region of minimal air flow imme-
diately adjacent to that surface. The thickness of the boundary layer depends mostly on the
windspeed and roughness of the surface. Schack et al. (1985) have extended particle deposi-
tion theory to include completely rough surfaces, such as terrestrial surfaces.
Airborne particles do not follow a smooth, straight path in the airstream. On the con-
trary, the path of a particle may be affected by micro-turbulent air currents, gravitation, or
inertia. There are several mechanisms that may alter the particle path enough to cause trans-
fer to a surface. These mechanisms are a function of particle size, windspeed, and surface
characteristics.
6-21
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Particles larger than a few micrometers in diameter are influenced primarily by sedimen-
tation, where the particle accelerates downward until aerodynamic drag is exactly balanced by
gravitational force. The particle continues at this velocity until it reaches a surface.
Sedimentation is not influenced by horizontal windspeed or surface characteristics. Particles
moving in an airstream may be removed by impaction whenever they are unable to follow the air-
stream around roughness elements of the surface, such as leaves, branches, or tree trunks. In
this case, the particle moves parallel to the airstream and strikes a surface perpendicular to
the airstream. A related mechanism, turbulent inertial deposition, occurs when a particle en-
counters turbulence within the airstream causing the particle to move perpendicular to the
airstream. It may then strike a surface parallel to the airstream. In two mechanisms, wind
eddy diffusion and interception, the particle remains in the airstream until it is transferred
to a surface. With wind eddy diffusion, the particle is transported downward by turbulent
eddies. Interception occurs when the particle in the airstream passes within one particle
radius of a surface. This mechanism is more a function of particle size than windspeed. The
final mechanism, Brownian diffusion, is important for very small particles at very low wind-
speeds. Brownian diffusion is motion, caused by random collision with molecules, in the di*
rection of a decreasing concentration gradient.
Transfer from the main airstream to the boundary layer is usually by sedimentation or
wind eddy diffusion. From the boundary layer to the surface, transfer may be by any of the
six mechanisms, although those that are independent of windspeed (sedimentation, interception,
Brownian diffusion) are more likely. Determining deposition onto rough surfaces requires
information of the height, shape and density of protrusions from the surface into the boundary
layer (Schack et al. , 1985). If dry deposition is seen as a two-step process, diffusion
through the boundary layer and interception by the surface, then for rough surfaces with rapid
eddy diffusion, interception by the protrusion surfaces becomes the rate-limiting step. Con-
sequently, surfaces such as water, grass, or bare rocks can be evaluated using a general
correlation with a reference surface (Schack et al., 1985), and a more complete understanding
of dry deposition to natural surfaces may be possible with the application of these experimen-
tal results.
6.4.1.2 Dry deposition models. A particle influenced only by sedimentation may be considered
to be moving downward at a specific velocity usually expressed in cm/sec. Similarly, parti-
cles transported to a surface by any mechanism are said to have an effective deposition velo-
city (V.), which is an expression of the rate of particle mass transfer measured by accumula-
tion on a surface as a function of time and air concentration. This relationship is expressed
in the equation:
6-22
-------
Vd = J/C (6-1)
where J is the flux or accumulation expressed in ng/cm2*s and C is the air concentration in
ng/cm3. The units of Vd become cm/sec.
Several recent models of dry deposition have evolved from the theoretical discussion of
Fuchs (1964) and the wind tunnel experiments of Chamberlain (1966). From those early works,
it was obvious that the transfer of particles from the atmosphere to the Earth's surface in-
volved more than rain or snow. The models of Slinn (1982) and Davidson et al. (1982) are par-
ticularly useful for lead deposition and were strongly influenced by the theoretical discus-
sions of fluid dynamics by Friedlander (1977). Slinn1s model considers a multitude of vegeta-
tion parameters to find several approximate solutions for particles in the size range of 0.1-
1.0 urn. In the absence of appropriate field studies, Slinn (1982) estimates deposition velo-
cities of 0.01-0.1 cm/sec.
The model of Davidson et al. (1982) is based on detailed vegetation measurements and wind
data to predict a V. of 0.05-1.0 cm/sec. Deposition velocities are specific for each vegeta-
tion type. This approach has the advantage of using vegetation parameters of the type made
for vegetation analysis in ecological studies (density, leaf area index (LAI), height, dia-
meter) and thus may be applicable to a broad range of vegetation types for which data are al-
ready available in the ecological literature.
Both models show a decrease in deposition velocity with decreasing particle size down to
about 0.1-0.2 urn, followed by an increase in V. with decreasing diameter from 0.1 to 0.001
cm/sec. On a log plot of Vd versus diameter, this curve is v-shaped (Sehmel, 1980), and the
plots of several vegetation types show large changes (10X) in minimum V., although the minima
commonly occur at about the same particle diameter (Figure 6-8). Although shown on the dia-
gram, particles larger than 0.1 urn diameter are not likely to have a density as great as 11.5
g/km3.
In summary, it is not correct to assume that air concentration and particle size alone
determine the flux of lead from the atmosphere to terrestrial surfaces. The type of vegetation
canopy and the influence of the canopy on windspeed are important predictors of dry deposi-
tion. Both of these models predict deposition velocities more than one order of magnitude
lower than reported in several earlier studies (e.g., Sehmel and Hodgson, 1976).
6.4.1.3 Calculation of Dry Deposition. The data required for calculating the flux of lead
from the atmosphere by dry deposition are leaf area index (LAI), windspeed, deposition velo-
city, and air concentration by particle size. The LAI should be total surface rather than up-
facing surface, as used in photosynthetic productivity measurements. LAI's should also be
expressed for the entire community rather than by individual plant, in order to incorporate
6-23
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UPPER LIMIT:
NO RESISTANCE BELOW AND
ATMOSPHERIC DIFFUSION FROM
1 cm TO 1 m
N
BELOW AND
'USION
INDICATED HEIGHT
ATMOSPHERIC DIFFUSION ABOVE
ATMOSPHERE
ROUGHNESS
HEIGHT, cm
PARTICLE DENSITY
ROUGHNESS HEIGHT
FRICTION VELOCITY
10
10" 1
PARTICLE DIAMETER, urn
Figure 6-8. Predicted deposition velocities at 1 m for M* = -
and particle densities of 1, 4, and 11.5 cm '.
Source: Sehmel (1980).
cm s
6-24
-------
variations in density. Some models use a more generalized surface roughness parameter, in
which case the deposition velocity may also be different.
The value selected for v"d depends on the type of vegetation, usually described as either
short (grasses or shrubs) or tall (forests). For particles with an MMAD of about 0.5 urn,
Hicks (1979) gives values for tall vegetation deposition velocity from 0.1-0.4 cm/sec.
Lannefors et al. (1983) estimated values of 0.2-0.5 cm/sec in the particle size range of
0.06-2.0 urn in a coniferous forest. For lead, with an MMAD of 0.55 urn, they measured a depo-
sition velocity of 0.41 cm/sec. In a series of articles (Wiman and Agren, 1985; Wiman and
Lannefors, 1985; and Wiman et al., 1985), this research group has described the modeling para-
meters required to define deposition in coniferous forests. They found a significant deple-
tion of aerosols from the forest edge to the interior, and distinct edge effects, for larger
particles, but nearly negligible depletion and edge effects for submicron particles. This
suggests that because lead is borne primarily on particles less than 1 urn, lead deposition
within a forest may be comparable to open grasslands and other vegetation types.
6.4.1.4 Field Measurements of Dry Deposition on Surrogate and Natural Surfaces. Several in-
vestigators have used surrogate surface devices similar to those described in Section 4.2.2.4.
These data are summarized in Table 6-4. The few studies available on deposition to vegetation
surfaces show deposition rates comparable to those of surrogate surfaces and deposition velo-
cities in the range predicted by the models discussed above. A study to compare vegetation
washing and several types of surrogate surfaces was reported by Dolske and Gatz (1984). Al-
though the study emphasizes sulfate particles, the devices and techniques are similar to those
used for lead. One important observation was that surrogate surface devices may be more re-
presentative of actual deposition if the device has a very shallow rim or no rim at all.
Therefore, the data of Table 6-4 do not include measurements made with deposition buckets. In
Section 6.4.3, these data are used to show that global emissions are in approximate balance
with global deposition. It is reasonable to expect that future refinements of field measure-
ments and model calculations will permit more accurate estimates of dry deposition in specific
regions or under specific environmental conditions.
6.4.2 Wet Deposition
Wet deposition includes removal by rainout and washout. Rainout occurs when particulate
matter is present in the supersaturated environment of a growing cloud. The small particles
(0.1 to 0.2 urn) act as nuclei for the formation of small droplets, which grow into raindrops
(Junge, 1963). Droplets also collect particles under 0.1 urn by Brownian motion and by the
water-vapor gradient. These processes are referred to as rainout. Washout, on the other
hand, occurs when falling raindrops collect particles by diffusion and impaction on the way to
6-25
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TABLE 6-4. SUMMARY OF SURROGATE AND VEGETATION SURFACE DEPOSITION OF LEAD
Deposition, Air cone, Deposition velocity,
Depositions! surface ng Pb/cm2-day ng/m3 cm/sec Reference
Tree leaves (Paris)
Tree leaves (Tennessee)
Plastic disk (remote
California)
Plastic plates
0.38
0.29-1.2
0.02-0.08
0.29-1.5
—
—
13-31
110
0.086
—
0.05-0.4
0.05-0.06
1
2
3
4
(Tennessee)
Tree leaves (Tennessee)
Snow (Greenland)
Grass (Pennsylvania)
Coniferous forest (Sweden)
—
0.004
—
0.74
110
0.1-0.2
590
21
0.005
0.1
0.2-1.1
0.41
4
5
6
7
1. Servant, 1975.
2. Lindberg et al., 1982.
3. Elias and Davidson, 1980.
4. Lindberg and Harriss, 1981.
5. Davidson et al., 1981.
6. Davidson et al., 1982.
7. Lannefors et al., 1983.
the ground. The limited data on the lead content of precipitation indicate a high variabi-
lity.
Wet deposition in rural and remote areas can often indicate regional or global processes
that remove lead from the atmosphere. Talbot and Andren (1983) measured lead in air and rain
at a semiremote site in Wisconsin. They found that wet deposition appeared to represent 80
percent of the total deposition of lead, and the total atmospheric flux of lead was
8 mg/iriVyr. There was a sharp increase in lead deposition during the summer months.
Deposition to a snowpack can be informative if the measurement correctly samples lead
that was deposited with the snow during a period of no snowmelt (Barrie and Vet, 1984). These
authors reported deposition of 1.8 mg/m2/yr to a snowpack of the East Canadian shield. Al-
though they meticulously avoided collecting melted snow, it was not clear how they accounted
for dry deposition deposited between periods of snowfall.
6-26
-------
A study of cloud droplet capture by vegetation (Lovett, 1984) suggests a possible mechan-
ism of deposition not included in wet or dry deposition. Although data on lead are not avail-
able, the mass transfer of water by this mechanism (0.01 cmVhr) suggests that, at a concen-
tration of 1 ug Pb/kg, the flux of lead could be 0.01 ng/cm2 for each hour of cloud droplet
exposure.
Lazrus et al. (1970) sampled precipitation at 32 U.S. stations and found a correlation
between gasoline used and lead concentrations in rainfall in each area. Similarly, there is
probably an inverse correlation between lead concentration in rainfall and distance from large
stationary point sources. The authors pointed out that at least twice as much lead is found
in precipitation as in water supplies, implying the existence of a process by which lead is
removed from water in the soil after precipitation reaches the ground. Russian studies
(Konovalov et al., 1966) point to the insolubility of lead compounds in surface waters and
suggest removal by natural sedimentation and filtration.
Lindberg et al. (1979) evaluated the deposition of Pb by wet and dry processes in a study
at Walker Branch Watershed in eastern Tennessee during the period 1976-1977. The mean annual
precipitation in the area is approximately 140 cm. Results for a typical year are reported in
Table 6-5. Wet deposition was estimated to contribute approximately 50 percent of the total
atmospheric input during this period, but on a seasonal basis ranged from 20 percent to 60
percent of total deposition. Further details on these studies have been published (Lindberg
et al., 1982; Lindberg, 1982).
6.4.3 Global Budget of Atmospheric Lead
The geochemical mass balance of lead in the atmosphere may be determined on a global
basis from quantitative estimates of inputs and outputs. Inputs are from natural and anthro-
pogenic emissions described in Section 5.2 and 5.3. They amount to 450,000-475,000 metric
tons annually (Nriagu, 1979). This simple procedure is an informative exercise that shows
whether the observed emission rates and deposition rates can, by making a minimum number of
reasonable assumptions, be brought into arithmetic balance. Each assumption can be tested in-
dependently, within the constraints of the overall model. For example, Table 6-6 assumes an
average concentration of 0.4 (jg Pb/kg precipitation. The total mass of rain and snowfall is
5.2 x 1017 kg/yr, so the amount of lead removed by wet deposition is approximately 208,000
t/yr. The average concentration of lead in precipitation cannot be greater than 0.8 ug/kg
(although values higher than this are commonly found in the scientific literature), since this
would exceed the estimates of global emissions. Furthermore, a value this high would preclude
dry deposition. For dry deposition, a crude estimate may be derived by dividing the surface
of the earth into three major vegetation types based on surface roughness or LAI. Oceans,
polar regions, and deserts have a very low surface roughness and can be assigned a deposition
6-27
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TABLE 6-5. ANNUAL AND SEASONAL DEPOSITION OF Pb AT WALKER BRANCH WATERSHED, mg/m2
Period
Atmospheric deposition of Pb
Wet
Dry
MEAN Daily deposition
Winter
Spri ng
Summer
Fall
Total year
Mean daily deposition
1.9 x 01-2
2.4
0.3
2.7
1.6
7.0 2
1.9 x 10"
2.2 x 10-2
1.8
1.5
3.1
1.6
8.0 2
2.2 x 10"
Calculated for a typical year from data collected during 1976-1977.
Winter = November-February, Spring = March and April, Summer = May-August, Fall = September
and October.
Source: Lindberg et al., 1979.
TABLE 6-6. ESTIMATED GLOBAL DEPOSITION OF ATMOSPHERIC LEAD
Wet
To oceans
To continents
Dry
To oceans, ice caps,
deserts
Grassland, agricultural
areas, and tundra
Forests
Mass of water,
10 17 kg/yr
4.1
1.1
Area,
1012 m2
405
46
59
Lead concentration,
10-6 g/kg
0.4
0.4
Total wet:
Deposition rate,
10-3 g/m2-yr
0.22
0.71
1.5
Total dry:
Total wet:
Global:
Lead deposition,
106 kg/yr
164
44
208
Deposition,
106 kg/yr
89
33
80
202
208
410
Source: This report.
6-28
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velocity of 0.01 cm/sec, which gives a flux of 0.2 ug/m2«yr, assuming 75 ng Pb/m3
air concentration. Grasslands, tundra, and other areas of low-lying vegetation have a some-
what higher deposition velocity; forests would have the highest. Values of 0.3 and 0.65 can
be assigned to these two vegetation types, based on the data of Davidson et al. (1982).
Whittaker (1975) lists the global surface area of each of the three types as 405, 46, and 59 x
1012 m2, respectively. In the absence of data on the global distribution of air concentra-
tions of lead, an average of 0.075 ug/m3 is assumed. Multiplying air concentration by depo-
sition velocity gives the deposition flux for each surface roughness type shown on Table 6-6.
The combined wet and dry deposition is 410,000 metric tons, which compares favorably with the
estimated 450,000-475,000 metric tons of emissions.
The data used above are not held to be absolutely firm. Certainly, more refined esti-
mates of air concentrations and deposition velocities can be made in the future. On the other
hand, the calculations above show some published calculations to be unreasonable. In particu-
lar, if the values for lead in rain (36 (jg/kg) reported by Lazrus et al. (1970) were applied
to this global model, more than 50 times the total global emissions would be required for mass
balance. Likewise, deposition fluxes of 0.95 ug/cm2-yr reported by Jaworowski et al. (1981)
would account for 10 times global emissions. Chemical mass balance budgets are an effective
aid to evaluating reports of environmental lead data.
6.5 TRANSFORMATION AND TRANSPORT IN OTHER ENVIRONMENTAL MEDIA
6.5.1 Soil
The accumulation of lead in soils is primarily a function of the rate of deposition,
since most lead is retained by the soil and very little passes through into surface or ground
water. The wet and dry deposition rates discussed in Section 6.4 would apply provided the
surface roughness and location (urban, rural, remote) are considered. A value of 8 mg/m2 yr
such as that measured by Talbot and Andren (1983) in a semiremote location in Wisconsin con-
verts to 0.8 ug/cm2 yr. It is difficult to generalize on the depth of penetration of lead in
undisturbed soils, but if it is assumed that most of the lead is retained in the upper 5 cm
(Reaves and Berrow, 1984; Garcia-Miragaya, 1984), then the accumulation rate of 0.8 ug/cm2 yr
becomes 0.16 ug/cm3 yr, or 0.16 ug/g if a density of 1 is assumed for soil. Ewing and Pearson
(1974) reported an accumulation of 13 ug/g soil from the 1920's to the late 1960's, or an
annual rate of about 0.26 ug/g in a rural setting. It should be noted that the atmospheric
concentration of lead increased 20-fold during this period (Shirahata et al., 1980). Page and
Ganje (1970) found an accumulation of 0.83 ug/g during the same time for a site near high
traffic density.
6-29
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These accumulation rates are discussed further in Section 7.2.2.1. Understanding the
distinction between atmospheric and natural lead in soil can provide some insight into the
mechanisms regulating transport in soil. Of particular importance are solubility and the sta-
bility of lead complexes with humic substances.
Soils have both a liquid and solid phase, and trace metals are normally distributed be-
tween these two phases. In the liquid phase, metals may exist as free ions or as soluble com-
plexes with organic or inorganic ligands. Organic ligands are typically humic substances such
as fulvic or humic acid; inorganic ligands may be iron or manganese hydrous oxides. Since
lead rarely occurs as a free ion in the liquid phase (Camerlynck and Kiekens, 1982), its mobi-
lity in the soil solution depends on the availability of organic or inorganic ligands. The
liquid phase of soil often exists as a thin film of moisture in intimate contact with the
solid phase. The availability of metals to plants depends on the equilibrium between the
liquid and solid phase.
In the solid phase, metals may be incorporated into crystalline minerals of parent rock
material, into secondary clay minerals, or precipitated as insoluble organic or inorganic com-
plexes. They may also be adsorbed onto the surfaces of any of these solid forms. Of these
categories, the most mobile form is in the film of moisture surrounding soil particles, where
lead can move freely into plant roots or soil microorganisms with dissolved nutrients. The
least mobile is parent rock material, where lead may be bound within crystalline structures
over geologic periods of time. Intermediate are the lead complexes and precipitates. Trans-
formation from one form to another depends on the chemical environment of the soil. For exam-
ple, at pH 6-8, insoluble organic-Pb complexes are favored if sufficient organic matter is
available; otherwise hydrous oxide complexes may form or the lead may precipitate with the
carbonate or phosphate ion. In the pH range of 4-6, the organic-Pb complexes become soluble.
Soils outside the pH range of 4-8 are rare. The interconversion between soluble and insoluble
organic complexes affects the equilibrium of lead between the liquid and solid phase of soil.
Dong et al. (1985) found that only 0.2 percent of the total lead in soil can be released
to solution by physical shaking. Even if 99.99 percent of the total lead in soil is immobil-
ized, 0.01 percent of the total lead in soil can have a significant effect on plants and mi-
croorganisms if the soils are heavily contaminated with lead (see Section 8.3.1).
The water soluble and exchangeable (as determined by chemical extraction) forms of metals
are the forms generally considered potentially available for plant uptake. It is important
not to confuse the term "extractable" with "plant uptake." Lead that can be extracted from
soil by chemical treatment may not be taken up by plants, even though the same chemical treat-
ment is known to release other metals to plants. Because little is known of this relation-
ship, lead that is extractable by chemical means normal for other metals is considered only
6-30
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potentially available for plant uptake. Camerlynck and Kiekens (1982) demonstrated that in
normal soils, only a small fraction of the total lead is in exchangeable form (about 1 ug/g)
and none exists as free lead ions. Of the exchangeable lead, 30 percent existed as stable
complexes, 70 percent as labile complexes. The organic content of these soils was low (3.2
percent clay, 8.5 percent silt, 88.3 percent sand). In heavily contaminated soils near a mid-
western industrial site, Miller and McFee (1983) found that 77 percent of the lead was in
either the exchangeable or organic form, although still none could be found in aqueous solu-
tion. Soils had a total lead content from 64 to 360 ug/g and an organic content of 7-16 per-
cent.
There is evidence that atmospheric lead enters the soil system as PbS04 or is rapidly
converted to PbS04 at the soil surface (Olson and Skogerboe, 1975). Lead sulfate is more
soluble than PbC03 or Pb3(P04)2 and thus could remain mobile if not transformed. Lead could
be immobilized by precipitation as less soluble compounds [PbC03, Pb3(P04)2], by ion exchange
with hydrous oxides or clays, or by chelation with humic and fulvic acids. Santillan-Medrano
and Jurinak (1975) discussed the possibility that the mobility of lead is regulated by the
formation of Pb(OH)2, Pb3(P04)2, Pb5(P04)3OH, and PbC03. This model, however, did not con-
sider the possible influence of organic matter on lead immobilization. Zimdahl and Skogerboe
(1977), on the other hand, found lead varied linearly with cation exchange capacity (CEC) of
soil at a given pH, and linearly with pH at a given CEC (Figure 6-9). The relationship
between CEC and organic carbon is discussed below.
If surface adsorption on clays plays a major role in lead immobilization, then the capa-
city to immobilize should vary directly with the surface-to-volume ratio of clay. In two
separate experiments using the nitrogen BET method for determining surface area and size frac-
tionation techniques to obtain samples with different surface-to-volume ratios, Zimdahl and
Skogerboe (1977) demonstrated that this was not the case. They also showed that precipitation
as lead phosphate or lead sulfate is not significant, although carbonate precipitation can be
important in soils that are carbonaceous in nature or to which lime (CaC03) has been added.
Of the two remaining processes, lead immobilization is more strongly correlated with or-
ganic chelation than with iron and manganese oxide formation (Zimdahl and Skogerboe, 1977).
It is possible, however, that chelation with fulvic and humic acids is catalyzed by the pre-
sence of iron and manganese oxides (Saar and Weber, 1982). This would explain the positive
correlation for both mechanisms observed by Zimdahl and Skogerboe (1977). The study of Miller
and McFee (1983) discussed above indicates that atmospheric lead added to soil is distributed
to organic matter (43 percent) and ferro-manganese hydrous oxides (39 percent), with 8 percent
found in the exchangeable fraction (determined by chemical extraction) and 10 percent as in-
soluble precipitates.
6-31
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If organic chelation is the correct model of lead immobilization in soil, then several
features of this model merit further discussion. First, the total capacity of soil to immobi-
lize lead can be predicted from the linear relationship developed by Zimdahl and Skogerboe
(1977) (Figure 6-9) based on the equation:
N = 2.8 x 10"6 (A) + 1.1 x 10"5 (B) - 4.9 x 10"5 (6-2)
where N is the saturation capacity of the soil expressed in moles/g soil, A is the CEC of the
soil in meq/100 g soil, and B is the pH. Because the CEC of soil is more difficult to deter-
mine than total organic carbon, it is useful to define the relationship between CEC and or-
ganic content. Pratt (1957) and Klemmedson and Jenny (1966) found a linear correlation be-
tween CEC and organic carbon for soils of similar sand, silt, and clay content. The data of
Zimdahl and Skogerboe (1977) also show this relationship when grouped by soil type. They show
that sandy clay loam with an organic content of 1.5 percent might be expected to have a CEC of
12 meq/100 g. From the equation, the saturation capacity for lead in soil of pH 5.5 would be
45 umoles/g soil or 9,300 M9/9- The same soil at pH 4.0 would have a total capacity of 5,900
When the soil contains sufficient organic matter, the adsorption of lead and other metals
is a function of pH. In reality, it is not the total organic matter but the number of reac-
tive sites that determines metal absorption (Harter, 1983), and there appears to be a weak re-
lationship between cation exchange capacity and metal retention. In competition with other
metals, lead would normally be more strongly favored for retention, in accordance with the
Irving-Williams series (Irving and Williams, 1953). Gamble et al. (1983) have shown that the
Irving-Williams series becomes somewhat distorted when the binding sites are chemically dis-
similar.
The soil humus model also facilitates the calculation of lead in soil moisture using
values available in the literature for conditional stability constants with fulvic acid (FA).
The term conditional is used to specify that the stability constants are specific for the con-
ditions of the reaction. Conditional stability constants for humic acid (HA) and FA are com-
parable. The values reported for log K are linear in the pH range of 3 to 6 (Buffle and
Greter, 1979; Buffle et al., 1976; Greter et al., 1979), so that interpolations in the criti-
cal range of pH 4 to 5.5 are possible. Thus, at pH 4.5, the ratio of complexed lead to ionic
lead is expected to be 3.8 x 103. For soils of 100 ug/g, the ionic lead in soil moisture
solution would be 0.03 M9/9- The significance of this ratio is discussed in Section 8.3.1.1.
6-32
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pH = 8
pH = 6
pH = 4
75
100
125
CEC, meq/100 g
Figure 6-9. Variation of lead saturation capacity with cation exchange
capacity (CEC) in soil at selected pH values.
Source: Data from Zimdahl and Skogerboe (1977).
6-33
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It is also important to consider the stability constant of the PtrFA complex relative to
other metals. Schnitzer and Hansen (1970) showed that at pH 3, Fe3 is the most stable in the
sequence Fea+ > A13+ > Cu2+ > Ni2+ > Co2+ > Pb2+ > Ca2+ > Zn2+ > Mn2* > Mg2*. At pH 5, this
sequence becomes Ni2+ = Co2+ > Pb2+ > Cu2+ > Zn2+ = Mn2+ > Ca2+ > Mg2+. This means that at
normal soil pH levels of 4.5-8, lead is bound to FA and HA in preference to many other metals
that are known plant nutrients (Zn, Mn, Ca, and Mg). Furthermore, if lead displaces iron in
this scheme, an important function of FA may be inhibited at near saturation capacity (above
6000 pg Pb/g.) Fulvic acid is believed to play a role in the weathering of parent rock mate-
rial by the removal of iron from the crystalline structure of the minerals, causing the rock
to weather more rapidly. In the absence of this process, the weathering of parent rock mate-
rial and the subsequent release of nutrients to soil would proceed more slowly. Bizri et al.
(1984) found stability constants for humic substances were log Kj = 4.2 and log K2 = 3.7. For
humic materials in aquatic systems, Alberts and Giesy (1983) reported conditional stability
constants of log Kx = 5.09 and log K2 = 2.00.
6.5.2 Water
6.5.2.1 Inorganic. The chemistry of lead in an aqueous solution is highly complex because
the element can be found in a multiplicity of forms. Hem and Durum (1973) have reviewed the
chemistry of lead in water in detail; the aspects of aqueous lead chemistry that are germane
to this document are discussed in Section 3.3.
Lead in ore deposits does not pass easily to ground or surface water. Any lead dissolved
from primary lead sulfide ore tends to combine with carbonate or sulfate ions to (1) form in-
soluble lead carbonate or lead sulfate, or (2) be absorbed by ferric hydroxide (Lovering,
1976). An outstanding characteristic of lead is its tendency to form compounds of low solu-
bility with the major anions of natural water. Hydroxide, carbonate, sulfide, and more rarely
sulfate may act as solubility controls in precipitating lead from water. The amount of lead
that can remain in solution is a function of the pH of the water and the dissolved salt con-
tent. Equilibrium calculations show that at pH > 5.4, the total solubility of lead is about
30 ug/1 in hard water and about 500 ug/1 in soft water (Davies and Everhart, 1973). Lead sul-
fate is present in soft water and limits the lead concentration in solution. Above pH 5.4,
PbC03 and Pb2(OH)2C03 limit the concentration. The carbonate concentration is in turn depend-
ent on the partial pressure of C02 as well as the pH. Calculations by Hem and Durum (1973)
show that many river waters in the United States have lead concentrations near the solubility
limits imposed by their pH levels and contents of dissolved C02. Because of the influence of
temperature on the solubility of C02, observed lead concentrations may vary significantly from
theoretically calculated ones.
6-34
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Concentrations as high as 330 |jg/l could be stable in water with pH near 6.5 and an alka-
linity of about 25 mg bicarbonate ion/1 of water. Water having these properties is common in
runoff areas of New York State and New England; hence, the potential for high lead concentra-
tions exists there. In other areas, the average pH and alkalinity are so high that maximum
concentrations of lead of about 1 ug/1 could be retained in solutions at equilibrium
(Levering, 1976).
A significant fraction of the lead carried by river water may be in an undissolved state.
This insoluble lead can consist of colloidal particles or larger undissolved particles of lead
carbonate, -oxide, -hydroxide, or other lead compounds incorporated in other components of
particulate lead from runoff; it may occur either as sorbed ions or surface coatings on sedi-
ment mineral particles or be carried as a part of suspended living or nonliving organic matter
(Lovering, 1976). A laboratory study by Hem (1976) of sorption of lead by cation exchange in-
dicated that a major part of the lead in stream water may be adsorbed on suspended sediment.
Figure 6-10 illustrates the distribution of lead outputs between filtrate and solids in water
from both urban and rural streams, as reported by Getz et al. (1977). The majority of lead
output is associated with suspended solids in both urban and rural streams, with very little
dissolved in the filtrate. The ratio of lead in suspended solids to lead in filtrate varies
from 4:1 in rural streams to 27:1 in urban streams.
Soluble lead is operationally defined as that fraction which is separated from the in-
soluble fraction by filtration. However, most filtration techniques do not remove all colloi-
dal particles. Upon acidification of the filtered sample, which is usually done to preserve
it before analysis, the colloidal material that passed through the filter is dissolved and is
reported as dissolved lead. Because the lead in rainfall can be mainly particulate, it is
necessary to obtain more information on the amounts of insoluble lead (Lovering, 1976) before
a valid estimate can be obtained of the effectiveness of runoff in transporting lead away from
areas where it has been deposited by wet and dry deposition.
6.5.2.2 Organic. The bulk of organic compounds in surface waters originates from natural
sources (Neubecker and Allen, 1983). The humic and fulvic acids that are primary complexing
agents in soils are also found in surface waters at concentrations from 1-5 mg/1, occasionally
exceeding 10 mg/1 (Steelink, 1977), and have approximately the same chemical characteristics
(Reuter and Perdue, 1977). The most common anthropogenic organic compounds are nitrilotri-
acetonitrile (NTA) and ethylenediaminetetraacetic acid (EDTA) (Neubecker and Allen, 1983).
There are many other organic compounds such as oils, plasticizers, and polymers discharged
from manufacturing processes that may complex with lead.
6-35
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0)
u
<
GC
o
u
o
100
75
50
25
SUSPENDED SOLIDS
FILTRATE
URBAN
RURAL
Figure 6-10. Lead distribution between filtrate and suspended
solids in stream water from urban and rural compartments.
Source: Getz et al. (1977).
6-36
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The presence of fulvic acid in water has been shown to increase the rate of solution of
lead sulfide 10 to 60 times over that of a water solution at the same pH that did not contain
fulvic acid (Bondarenko, 1968; Lovering, 1976). At pH values near 7, soluble Pb-FA complexes
are present in solution. At initial pH values between 7.4 and about 9, the Pb-FA complexes
are partially decomposed, and lead hydroxide and carbonate are precipitated. At initial pH
values of about 10, the Pb-FA complexes again increase. This increase is attributed to dis-
sociation of phenolic groups at high pH values, which increases the complexing capacity of the
FA. But it also may be due to the formation of soluble lead-hydroxyl complexes.
Beijer and Jernelov (1984) review the evidence for the microbial methylation of lead in
aquatic systems. The transformation of inorganic lead, especially in sediment, to tetra-
methyllead (TML) has been observed and biomethylation has been postulated (Schmidt and Huber,
1976; Wong et al., 1975). Reisinger et al. (1981) have reported extensive studies of the
methylation of lead in the presence of numerous bacterial species known to alkylate mercury
and other heavy metals. In these experiments no biological methylation of lead was found
under any condition. Chemical alkylation from methylcobalamine was found to occur in the
presence of sulfide or of aluminum ion; chemical methylation was independent of the presence
of bacteria.
Jarvie et al. (1975, 1981) have recently shown that tetraalkyllead compounds are unstable
in water. Small amounts of Ca2 and Fe2+ ions and sunlight have been shown to cause decompo-
sition of TEL over time periods of 5-50 days. The only product detected was triethyllead,
which appears to be considerably more stable than the TEL. Tetramethyllead is decomposed much
more rapidly than TEL in water, to form the trimethyl lead ion. Initial concentrations of
10 molar were reduced by one order of magnitude either in the dark or light in one day.
Tetramethyllead was virtually undetectable after 21 days. Apparently, chemical methylation of
lead to the trialkyHead cation does occur in some water systems, but evolution of TML appears
insignificant.
Lead occurs in riverine and estuarial waters and alluvial deposits. Laxen and Harrison
(1977) and Harrison and Laxen (1981) found large concentrations of lead (~1 mg/1) in rainwater
runoff from a roadway; but only 5-10 percent of this is soluble in water. Concentrations of
lead in ground water appear to decrease logarithmically with distance from a roadway. Rain-
water runoff has been found to be an important transport mechanism in the removal of lead from
a roadway surface in a number of studies (Bryan, 1974; Harrison and Laxon, 1981; Hedley and
Lockley, 1975; Laxen and Harrison, 1977). Apparently, only a light rainfall, 2-3 mm, is suf-
ficient to remove 90 percent of the lead from the road surface to surrounding soil and to
waterways (Laxen and Harrison, 1977).
6-37
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The Applied Geochemistry Research Group (1978) has reported elevated lead concentrations
(40 (jg/9 and above) in about 30 percent of stream bed sediment samples from England and Wales
in a study of 50,000 such samples. Abdullah and Royle (1973) have reported lead levels in
coastal areas of the Irish sea of 400 |jg/g and higher.
Evidence for the sedimentation of lead in freshwater streams may be found in several re-
ports. Laxen and Harrison (1983) found that lead in the effluent of a lead-acid battery plant
near Manchester, England, changed drastically in particle size. In the plant effluent, 53
percent of the lead was on particles smaller than 0.015 urn and 43 percent on particles greater
than 1 (jm. Just downstream of the plant, 91 percent of the lead was on particles greater than
1 urn and only 1 percent on particles smaller than 0.015 urn. Under these conditions, lead
formed or attached to large particles at a rate exceeding that of Cd, Cu, Fe or Mn.
The lead concentrations in off-shore sediments often show a marked increase corresponding
to anthropogenic activity in the region (Section 5.1). Rippey et al. (1982) found such in-
creases recorded in the sediments of Lough Neagh, Northern Ireland, beginning during the
1600's and increasing during the late 1800's. Corresponding increases were also observed for
Cr, Cu, Zn, Hg, P, and Ni. For lead, the authors found an average anthropogenic flux of 72
mg/m2-yr, of which 27 mg/m2-yr could be attributed to direct atmospheric deposition. Prior to
1650, the total flux was 12 mg/m2-yr, so there has been a 6-fold increase since that time.
Ng and Patterson (1982) found prehistoric fluxes of 1-7 mg Pb/m2-yr in three offshore
basins in southern California, which have now increased 3 to 9-fold to 11-21 mg/m2«yr. Much
of this lead is deposited directly from sewage outfalls, although at least 25 percent probably
comes from the atmosphere.
6.5.3 Vegetation Surfaces
The deposition of lead on the leaf surfaces of plants where the particles are often re-
tained for a long time must also be considered (Dedolph et al., 1970; Page et al., 1971;
Schuck and Locke, 1970). Many studies have shown that plants near roadways exhibit consid-
erably higher levels of lead than those further away. In most instances, the higher concen-
trations were due to lead particle deposition on plant surfaces (Schuck and Locke, 1970).
Studies have shown that particles deposited on plant surfaces are difficult to remove by
typical kitchen washing techniques. (Arvik and Zimdahl, 1974; Page et al., 1971; Lagerwerff
et al., 1973). Leaves with pubescent surfaces seem able to retain particles via an electro-
static mechanism. Other types of leaves are covered with a cuticular wax physically suitable
for retaining particles. Rainfall does not remove all of the particles on the leaf surface.
It appears that there is a buildup with time of surface deposition on leafy vegetation.
Animals consuming the leafy portions of such plants can certainly be exposed to higher than
normal levels of lead.
6-38
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The uptake of soluble lead by aquatic plants can be an important mechanism for depleting
lead concentrations in downstream waterways. Gale and Wixson (1979) have studied the influ-
ence of algae, cattails, and other aquatic plants on lead and zinc levels in wastewater in the
New Lead Belt of Missouri. These authors report that mineral particles become trapped by
roots, stems, and filaments of aquatic plants. Numerous anionic sites on and within cell
walls participate in cation exchange, replacing metals such as lead with Na , K , and H ions.
Mineralization of lead in these Missouri waters may also be promoted by water alkalinity.
However, construction of stream meanders and settling ponds have greatly reduced downstream
water concentrations of lead, mainly because of absorption in aquatic plants (Gale and Wixson,
1979).
6.6 SUMMARY
From the source of emission to the site of deposition, lead particles are dispersed by
the flow of the airstream, transformed by physical and chemical processes, and removed from
the atmosphere by wet or dry deposition. Under the simplest of conditions (smooth, flat ter-
rain), the dispersion of lead particles has been modeled and can be predicted (Benarie, 1980).
Dispersion modeling in complex terrains is still under development and these models have not
been evaluated (Kotake and Sano, 1981).
Air lead concentrations decrease logarithmically away from roadways (Edwards, 1976) and
smelters (Roberts et al., 1974). Within urban regions, air concentrations decrease from the
central business district to the outlying residential areas by a factor of 2-3. From urban to
rural areas, air concentrations decrease from 1-2 ug/m3 down to 0.1-0.5 ug/m3 (Chapter 7).
This decrease is caused by dilution with clean air and removal by deposition. During
dispersion to remote areas, concentrations decrease to 0.01 |jg/m3 in the United States (Elias
and Davidson, 1980), to 0.001 ug/m3 in the Atlantic Ocean (Duce et al., 1975), and to 0.000076
ug/m3 in Antarctica (Maenhaut et al., 1979).
Physical transformations of lead particles cause a shift in the particle size distribu-
tion. The bimodal distribution of large and small particles normally found near the roadway
changes with time and distance to a single mode of intermediate sized particles (Huntzicker et
al., 1975). This is probably because large particles deposit near roadways and small parti-
cles agglomerate to medium sized particles with an MMAD of about 0.2-0.3 urn.
Particles transform chemically from lead halides to lead sulfates and oxides. Organolead
compounds constitute 1-6 percent of the total airborne lead in ambient urban air (Harrison
et al., 1979).
6-39
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On a regional or global basis, wet deposition accounts for about half of the removal of
lead particles from the atmosphere. The other half of the atmospheric lead is removed by dry
deposition. Mechanisms may be gravitational for large particles or a combination of gravita-
tional and wind-related mechanisms for small particles (Elias and Davidson, 1980). Models of
dry deposition predict deposition velocities as a function of particle size, windspeed, and
surface roughness. Because of their large surface area/ground area ratio, grasslands, crop-
lands, and forested areas receive the bulk of dry deposited particles over continental areas.
Lead enters soil as a moderately insoluble lead sulfate and is immobilized by complexa-
tion with humic and fulvic acids. This immobilization is a function of pH and the concentra-
tion of humic substances. At low pH (~4) and low organic content (<5 percent), immobilization
of lead in soil may be limited to a few hundred ug/g (Zimdahl and Skogerboe, 1977), but at 20
percent organic content and pH 6, 10,000 pg Pb/g soil may be found.
In natural waters, lead may precipitate as lead sulfate or carbonate, or it may form a
complex with ferric hydroxide (Lovering, 1976). The solubility of lead in water is a function
of pH and hardness (a combination of Ca and Mg content). Below pH 5.4, concentrations of dis-
solved lead may vary from 30 ug/1 in hard water to 500 ug/1 in soft water at saturation
(Lovering, 1976).
6-40
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6.7 REFERENCES
Abdullah, M. I.; Royle, L. G. (1973) The occurrence of lead in natural waters. In: Barth, 0.;
Berlin, A.; Engel, R.; Recht, P.; Smeets, J., eds. Enviromental health aspects of lead:
proceedings, international symposium; October 1972; Amsterdam, The Netherlands. Luxem-
bourg: Commission of the European Communities; pp. 113-124.
Alberts, J. J.; Giesy, J. P. (1983) Conditional stability constants of trace metals and
naturally occurring humic materials: application in equilibrium models and verification
with field data. In: Christman, R. F.; Gjessing, E. T. Aquatic and terrestrial humic
materials: symposium; November 1981; Chapel Hill, NC. Woburn, MA: Butterworth; pp.
333-348.
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6-51
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7. ENVIRONMENTAL CONCENTRATIONS AND POTENTIAL PATHWAYS TO HUMAN EXPOSURE
7.1 INTRODUCTION
In general, typical levels of human lead exposure may be attributed to four components of
the human environment: food, inhaled air, dusts of various types, and drinking water. This
chapter presents information on the ranges and temporal trends of concentrations in ambient
air, soil, and natural waters, and discusses the pathways from each source to food, inhaled
air, dust, and drinking water. The ultimate goal is to quantify the contribution of anthropo-
genic lead to each source and the contribution of each source to the total lead consumed by
humans. These sources and pathways of human lead exposure are diagrammed in Figure 7-1.
Chapters 5 and 6 discuss the emission, transport, and deposition of lead in ambient air.
Some information is also presented in Chapter 6 on the accumulation of lead in soil and on
plant surfaces. Because this accumulation is at the beginning of the human food chain, it is
critical to understand the relationship between this lead and lead in the human diet. It is
also important where possible to project temporal trends.
In this chapter, a baseline level of potential human exposure is determined equivalent to
that for a normal adult eating a typical diet and living in a non-urban community. This base-
line exposure is deemed to be unavoidable by any reasonable means. Beyond this level, addi-
tive exposure factors can be determined for other environments (e.g., urban, occupational,
smelter communities), for certain habits and activities (e.g., pica, smoking, drinking, and
hobbies), and for variations due to age, sex, or socioeconomic status.
7.2 ENVIRONMENTAL CONCENTRATIONS
Quantifying human exposure to lead requires an understanding of ambient lead levels in
environmental media. Of particular importance are lead concentrations in ambient air, soil,
and surface or ground water. The following sections discuss environmental lead concentrations
in each of these media in the context of anthropogenic vs. natural origin, and the contribu-
tion of each to potential human exposure.
7.2.1 Ambient Air
Ambient airborne lead concentrations may influence human exposure through direct inhala-
tion of lead-containing particles and through ingestion of lead that has been deposited from
the air onto surfaces. Although a plethora of data on airborne lead is now available, our
7-1
-------
INDUSTRIAL
EMISSIONS
SURFACE AND
GROUND WATER
/
DRINKING
WATER
Figure 7-1. Principal pathways of lead from the environment to human consump-
tion. Heavy arrows are those pathways discussed in greatest detail in this chapter.
7-2
-------
understanding of the pathways to human exposure is far from complete because most ambient mea-
surements were not taken in conjunction with studies of the concentrations of lead in man or
in components of his food chain. However, that is the context in which these studies must now
be interpreted to shed the most light possible on the concentrations likely to be encountered
in various environmental settings.
The most complete set of data on ambient air concentrations may be extracted from the
National Filter Analysis Network (NFAN) and its predecessors (see Section 4.2.1). These data,
which are primarily for urban regions, have been supplemented with published data from rural
and remote regions of the United States. Because some stations in the network have been in
place for about 15 years, information on temporal trends is available but sporadic. Ambient
air concentrations in the United States are comparable to other industrialized nations. In
remote regions of the world, air concentrations are two or three orders of magnitude lower,
lending credence to estimates of the concentration of natural lead in the atmosphere. In the
context of the NFAN data base, the conditions are considered that modify ambient air, as mea-
sured by the monitoring networks, to air as inhaled by humans. Specifically, these conditions
are changes in particle size distributions, changes with vertical distance above ground, and
differences between indoor and outdoor concentrations.
7.2.1.1 Total Airborne Lead Concentrations. A thorough understanding of human exposure to
airborne lead requires detailed knowledge of spatial and temporal variations in ambient con-
centrations. The wide range of concentrations is apparent from Table 7-1, which summarizes
data obtained from numerous independent measurements, and Tables 7-2 and 7-3, which show air
concentrations in specific locations throughout the United States. Concentrations vary from
0.000076 M9/m3 in remote areas to over 13 pg/m3 near sources such as smelters. Many of the
remote areas are far from human habitation and therefore do not reflect human exposure. How-
ever, a few of the regions characterized by low lead concentrations are populated by indivi-
duals with primitive lifestyles; these data provide baseline airborne lead data to which
modern American lead exposures can be compared. Examples include some of the data from South
America and the data from Nepal. A more extensive review of atmospheric lead in remote areas
has been compiled by Wiersma and Davidson (1984).
Urban, rural, and remote airborne lead concentrations in Table 7-1 suggest that human ex-
posure to lead has increased as the use of lead in inhabited areas has increased. This is
consistent with published results of retrospective human exposure studies. For example,
Ericson et al. (1979) have analyzed the teeth and bones of Peruvians buried 1600 years ago.
Based on their data, they estimate that the skeletons of present-day American and British
adults contain roughly 500 times the amount of lead that would occur naturally in the absence
of widespread anthropogenic lead emissions. Grandjean et al. (1979) and Shapiro et al. (1980)
7-3
-------
TABLE 7-1. ATMOSPHERIC LEAD IN URBAN, RURAL, AND REMOTE AREAS OF THE WORLD
Location
Urban
New York
Boston
St. Louis
Houston
Chicago
Los Angeles
Ottowa
Toronto
Montreal
Brussels
Turin
Riyadh, Saudi Arabia
Rural
New York Bight
United Kingdom
Italy
Belgium
Illinois
Remote
White Mtn. , CA
High Sierra, CA
Olympic Nat. Park, WA
Great Smoky Mtns. Nat.
Park, TN
Glacier Nat. Park, MT
South Pole
Thule, Greenland
Thule, Greenland
Prins Christian-
sund, Greenland
Dye 3, Greenland
Eniwetok, Pacific Ocean
Kumjung, Nepal
Bermuda
Abastumani Mtns. USSR
Sampling period
1978-79
1978-79
1973
1978-79
1979
1978-79
1975
1975
1975
1978
1974-79
1983
1974
1972
1976-80
1978
1973-74
1969-70
1976-77
1980
1979
1981
1974
1965
1978-79
1978-79
1979
1979
1979
1973-75
1979
Lead cone. ,
(pg/m3)
1.1
0.8
1.1
0.9
0.8
1.4
1.3
1.3
2.0
0.5
4.5
5.5
0.13
0.13
0.33
0.37
0.23
0.008
0.021
0.0022
0.015
0.0046
0.000076
0.0005
0.008
0.018
0.00015
0.00017
0.00086
0.0041
0.019
Reference
NEDS, 1982
NEDS, 1982
NEDS, 1982
NEDS, 1982
NEDS, 1982
NEDS, 1982
NAPS, 1971-1976
NAPS, 1971-1976
NAPS, 1971-1976
Roels et al. , 1980
Facchetti and Geiss, 1982
El-Shobokshy, 1984
Duce et al., 1975
Cawse, 1974
Facchetti and Geiss, 1982
Roels et al. , 1980
Hudson et al. , 1975
Chow et al., 1972
Eli as and Davidson, 1980
Davidson et al. , 1982
Davidson et al. , 1985
Davidson et al. , 1985
Maenhaut et al . , 1979
Murozumi et al . , 1969
He i dam, 1983
Heidam, 1983
Davidson et al . , 1981c
Settle and Patterson, 1982
Davidson et al. , 1981b
Duce et al. , 1976
Dzubay et al. , 1984
7-4
-------
report lead levels in teeth and bones of contemporary populations to be elevated 100-fold over
levels in ancient Nubians buried before 750 A.D. On the other hand, Barry and Connolly (1981)
report excessive lead concentrations in buried medieval English skeletons; one cannot discount
the possibility that the lead was absorbed into the skeletons from the surrounding soil.
The remote area concentrations reported in Table 7-1 do not necessarily reflect natural,
preindustrial lead. Murozumi et al. (1969) measured a 200-fold increase over the past 3000
years in the lead content of Greenland snow, confirmed by Ng and Patterson (1981). In the
opinion of these authors, this lead originates in populated mid-latitude regions, and is
transported over thousands of kilometers through the atmosphere to the Arctic. All of the
concentrations in Table 7-1, including values for remote areas, have been influenced by
anthropogenic lead emissions.
Studies referenced in Table 7-1 are limited in that the procedures for determining the
quality of the data are generally not reported. In contrast, the two principal airborne lead
data bases described in Section 4.2.1 include measurements subjected to documented quality as-
surance procedures. The U.S. Environmental Protection Agency's National Filter Analysis Net-
work (NFAN) provides comprehensive nationwide data on long-term trends. The second data base,
EPA's National Aerometric Data Bank, contains information contributed by state and local
agencies, which monitor compliance with the current ambient airborne standard for lead (1.5
ug/m3 averaged over a calendar quarter) promulgated in 1978.
7.2.1.1.1 Distribution of air lead in the United States. Figure 7-2 categorizes the urban
sites with valid annual averages (4 valid quarters) into several annual average concentration
ranges (Akland, 1976; Shearer et al. 1972; U.S. Environmental Protection Agency, 1978, 1979;
Quarterly averages of lead from NFAN, 1982). Nearly all of the sites reported annual averages
below 1.0 ug/m3. Although the decreasing number of monitoring stations in service in recent
years could account for some of the shift in averages toward lower concentrations, trends at
individual urban stations, discussed below, confirm the apparent national trend of decreasing
lead concentration.
The data from these networks show both the maximum quarterly average to reflect compli-
ance of the station to the ambient airborne standard (1.5 ug/m3), and quarterly averages to
show trends at a particular location. Valid quarterly averages must include at least five
24-hour sampling periods evenly spaced throughout the quarter. The number of stations comply-
ing with the standard has increased, the quarterly averages have decreased, and the maximum
24-hour values appear to be smaller since 1977.
Long-term trends and seasonal variations in airborne lead levels at urban sites can be
seen in Figure 7-3. The 10th, 50th, and 90th percentile concentrations are graphed, using
quarterly composite and quarterly average data from an original group of 92 urban stations
(1965-1974) updated with data for 1975 through 1980. Note that maximum lead concentrations
7-5
-------
(A
O
I
u.
O
<0.5
- 0.5 0.9
•• 1.0-1.9
2.0-3.9 ug/m3
1966 67 68 69 70 71 72 73 74 75 76 77 78 79 80
(95) (146) (159) (180) (130) (162) (72) (57)
YEAR
Figure 7-2. Percent of urban stations reporting indicated concentration interval.
7-6
-------
typically occur in the winter, while minima occur in the summer. In contrast, automotive
emissions of lead would be expected to be greater in the summer for two reasons: (1) gasoline
usage is higher in the summer; and (2) lead content is raised in summer gasolines to replace
some of the more volatile high-octane components that cannot be used in summertime gasolines.
Apparently, the troposphere has a greater capacity to disperse submicron particles in the
summer than in the winter.
Figure 7-3 also clearly portrays the significant decrease in airborne lead levels over
the past decade. This trend is attributed to the decreasing lead content of regular and pre-
mium gasoline, and to the increasing usage of unleaded gasoline. The close parallel between
these two parameters is discussed in detail in Chapter 5. (See Figure 5-7 and Table 5-5.)
Table 7-2 shows lead concentrations in the atmospheres of several major metropolitan
areas of epidemiological interest. Some of the data presented do not meet the stringent re-
quirements for quarterly averages and occasionally there have been changes in site location or
sampling methodology. Nevertheless, the data are the best available for reporting the history
of lead contamination in these specific urban atmospheres. Further discussions of these data
appear in Chapter 11.
4.0
3.0
2.0
1.0
iii|iii|iri|iii JT i
ii(TiiriiiiMi
_ 10th PERCENTILE
i . I . . i I i . il . . . I . . i I II .1 iiilii i_li ,, I ili
66666768697071 7273 74 757677787980
YEAR
Figure 7-3. Seasonal patterns and trends in quarterly average urban lead concentrations.
7-7
-------
7.2.1.1.2 Global distributions of air lead. Other industrialized nations have maintained
networks for monitoring atmospheric lead. For example, Kretzschmar et al. (1980) reported
trends from 1972 to 1977 in a 15-station network in Belgium. Annual averages ranged from 0.16
(jg/m3 at rural sites to 1.2 ug/m3 near the center of Antwerp. All urban areas showed a maxi-
mum near the center of the city, with lead concentrations decreasing outward. The rural back-
ground levels appeared to range from 0.1 to 0.3 M9/1"3- Representative data from other nations
appear in Table 7-1.
7.2.1.1.3 Natural concentrations of lead in air. There are no direct measurements of prehis-
toric natural concentrations of lead in air. Air lead concentrations that existed in prehis-
toric times must be inferred from available data. Table 7-1 lists several values for remote
areas of the world, the lowest of which is 0.000076 ug/m3 at the South Pole (Maenhaut et al.,
1979). Two other reports show comparable values: 0.00017 ug/m3 at Eniwetok in the Pacific
Ocean (Settle and Patterson, 1982) and 0.00015 at Dye 3 in Greenland (Davidson et al., 1981a).
Since each of these studies reported some anthropogenic influence, it may be assumed that
natural lead concentrations are somewhat lower than these measured values.
Another approach to determining natural concentrations is to estimate global emissions
from natural sources. Nriagu (1979) estimated emissions at 24.5 x 106 kg/yr, whereas Settle
and Patterson (1980) estimated a lower value of 2 x 106 kg/yr. An average tropospheric
volume, to which surface-generated particles are generally confined, is about 2.55 x 1010m3.
Assuming a residence time of 10 days (see Section 6.3), natural lead emissions during this
time would be 6.7 x 1014 M9- The air lead concentrations would be 0.000263 M9/m3 using the
values of Nriagu (1979) or 0.0000214 ug/m3 using the data of Settle and Patterson (1980). It
seems likely that the concentration of natural lead in the atmosphere is between 0.00002 and
0.00007 ug/m3. A value of 0.00005 ug/m3 will be used for calculations regarding the contri-
bution of natural air lead to total human uptake in Section 7.3.1.
7.2.1.2 Compliance with the 1978 Air Quality Standard. Figure 7-4 shows percentile distri-
butions for the maximum average quarterly lead concentrations by year for a select group of 36
sites for which the data are available during the entire time period, 1975-1984. These data
show that not only did the composite average maximum average quarterly values decrease during
the .period 1975-1984, but the maximum average quarterly lead concentrations for all percen-
tiles showed a comparable pattern of decrease. From Figure 7-4, it may be concluded that most
stations reported average quarterly lead concentrations below the NAAQS standard of 1.5
pg/m3. Those that did not are shown on Table 7-3. Table 7-3 lists stations operated by
state and local agencies where one or more quarterly averages exceeded 1.0 ug/m3 or the cur-
rent standard of 1.5 ug/m3 from 1979 to 1984. A portion of each agency's compliance monitor-
ing network consists of monitors sited in areas expected to yield high concentrations associ-
ated with identifiable sources. In the case of lead, these locations are most likely to be
7-8
-------
3.5
3.0
O
e
UJ
g 2.5
8
Q
2.0
u
ee
UJ
cc
c
O
1.0
I 0.5
1 I I I
95th PERCENT!LE
90th PERCENTILE
75th PERCENTILE
COMPOSITE AVERAGE
MEDIAN
25th PERCENTILE
10th PERCENTILE
5th PERCENTILE
-NAAQS--
I
I
I
1975 1976 1977 1978 1979 1980 1981 1982 1983 1984
YEAR
Figure 7-4. Comparison of trends in maximum quarterly average lead concentrations
at 36 sites, 1975 • 1984.
Source: U.S. Environmental Protection Agency (1986).
7-9
-------
TABLE 7-2. AIR LEAD CONCENTRATIONS IN MAJOR METROPOLITAN AREAS
jjg/m3 quarterly averages
•vj
I—»
o
Boston
MA
Year
1970
1971
1972
1973
1974
1975
Quarter
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
1
0.8
1.2
1.2
0.7
1.0
0.6
2.5
0.6
0.9
1.0
1.2
0.61
l.O1
0.91
New York
NY
1
1.2
1.5
1.9
1.4
1.6
1.8
1.7
0.9
1.3
1.0
1.1
0.8
1.3
0.9
0.5
1.1
0.9
0.9
0.8
0.8
1.0
1.1
Phila. Wash.
PA DC
1 4 1
0.
0.
1.
1.
1.
1.
2.
1.
1.
1.
0.
1.
9
9
2
1
3
3
1
7
2
1
5
1
Detroit Chicago Houston
MI IL TX
1
1.2
1.4
1.4
1.3
1.0
1.8
1.6
2.2
0.9
0.9
0.8
0.7
1.2
1.2
Station Type
1231
1.
4
8
2.0
1.
2.
1.
1.
1.
2.
2.
1.
0.
2.
2.
1.
1.
1.
1.
2.
1.
2.
2.
1.
2.
2.
9
5
9
6
7
7
3
0
9
3
9
8
7
7
8
0 0.61
8 0.6
6 0.5
I1 0.7
7 0.7
1 0.6
4 1.1
Dal las/Ft. Worth
TX
1 2 4
3.8
2.3
2.8
3.7
3.4
1.8
2.5
2.7
3.4
1.8
2.2
2.8
1.9
1.3
1.9
1.3
1.4 0.21
2.8 0.4
3.3 0.6
2.9 0.3
2.3 0.3
3.0 0.4
2.9 0.5 0.3
Los Angeles
CA
1
5.7
3.5
5.1
3.9
6.0
2.9
3.3
6.3
3.1
2.0
2.6
4.7
2.7
2.0
2.7
1.9
2.0
1.4
3.2
1.2
1.9
3.2
2
3.2
2.2
3.3
1.9
1.6
1.5
2.1
1.6
2.5
1.6
1.7
1.9
2.6
1.7
1.2
1.7
2.2
-------
TABLE 7-2. (continued)
Boston
MA
Year
1976
1977
1978
1979
1980
1981
Quarter
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
1
0.61
0.7
0.8
l.O1
0.9
1.3
1.0
0.4
0.6
0.81
0.91
0.5
0.6
0.4
0.3
New York Phi la.
NY PA
1 1
1.3
1.6
1.4
1.3
1.2
1.1
1.4
1.3 1.6
l.O1 1.1
0.9 1.2
1.0
1.2
0.7
0.4
0.7
0.7
0.5
0.4
0.4
0.4
4
1.0
0.8
0.9
1.0
0.8
0.7
0.7
1.2
0.7
0.6
0.6
0.8
0.4
0.4
0.4
0.5
0.41
0.3
0.2
0.3
Wash.
DC
1
1.21
1.4
0.41
1.2
0.91
2.1
2.2
1.1
1.1
3.3
1.8
1.3
1.6
1.9
Detroit Chicago Houston
MI IL TX
1
1.1
0.9
1.0
0.3
0.3
0.3
0.41
0.3
0.3
0.3
0.31
Station Type
12314
0.81 0.5
0.71 0.5
1.1 0.7
0.31 0.2
0.8 0.3
1.3 0.7
1.0 0.5
0.8 0.4
0.8 0.5
1.7 0.9
0.9 0.4
0.7 0.9 0.8 0.8 0.4
0.5 0.6 0.8 0.51 0.61
0.71 0.5
0.4 0.3 0.3 0.61 0.3
0.7 0.4 0.6 0.31 0.31
1.0 0.5 0.5 0.2
0.5 0.4 0.4 0.4
0.2 0.3 0.2 0.7 0.5
0.4 0.3 0.3 0.2 0.2
0.3 0.3 0.2 0.5 0.3
0.4 0.21 0.3 0.8 l.O1
Dall as/Ft. Worth
TX
1
0.71
0.7
l.l1
2.3
1.2
1.1
1.61
1.71
1.1
1.3
1.7
1.21
0.61
l.l1
0.51
0.31
0.61
0.3
0.4
0.6
0.3
2
0.3
0.3
0.3
0.2
0.2
0.5
0.4
0.4
0.4
0.5
0.4
0.2
0.4
0.3
0.3
0.1
0.1
0.3
0.3
0.1
0.2
0.3
4
0.2
0.4
0.3
0.2
0.2
0.5
0.3
0.3
0.3
0.6
0.4
0.3
0.5
0.4
0.2
0.2
0.1
0.3
0.3
0.2
0.3
0.4
Los Angeles
CA
1
4.1
3.3
1.7
1.8
3.8
2.21
1.6
1.9
1.5
0.9
l.O1
0.61
0.7
l.l1
1.3
0.7
0.8
1.3
2
3.0
2.4
1.4
1.6
2.9
1.6
1.1
0.8
1.0
1.7
1.0
0.7
0.8
1.1
-------
TABLE 7-2. (continued)
I
J—»
ro
Boston
MA
New York Phila.
NY PA
Wash.
be
Detroit Chicago
MI IL
Houston
TX
Dal las/Ft. Worth
TX
Los Angeles
CA
Station Type
Year
1982
1983
1984
Station
Quarter
1
2
3
4
1
2
3
4
1
2
3
4
type: 1.
2.
3.
4.
1
1.0
0.5
0.4
0.6
0.5
0.4
0.4
0.5
0.5
center
center
center
1 1
0.5
0.5
0.6
0.3
0.3
0.3
0.5
0.3
0.4
0.3
0.4
city commercial
city residential
city industrial
4
0.3
0.3
0.3
0.4
0.3
0.2
0.2
0.3
0.3
0.2
0.2
0.3
i
0.5
0.3
0.3
0.51
0.2
0.2
0.2
0.3
0.2
0.2
1 1
0.4
0.2
0.3
0.4
0.4
0.3
0.4
0.4
0.1 0.3
0.1 0.2
0.1 0.2
0.3 0.3
2
0.3
0.4
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.3
0.2
0.3
3
0.3
0.3
0.2
0.3
0.4
0.3
0.3
0.2
0.3
0.3
0.3
0.3
1 4
0.2
0.3
0.4 0.3
0.2
0.2
0.3
0.4
0.3 0.1
0.3 0.2
0.3 0.1
1
0.5
0.6
1.0
0.6
0.5
0.6
0.1
0.4
0.4
2
0.2
0.3
0.6
0.2
0.5
0.6
0.3
0.4
4
0.2
0.2
0.3
0.2
0.2
0.5
0.1
0.2
1
0.8
0.5
0.8
1.1
1.0
0.4
0.6
0.8
0.7
0.1
2
0.7
0.6
0.6
0.3
0.4
0.7
0.3
0.4
suburban residential
less than required number of 24-hour sampling periods to meet composite criteria.
-------
TABLE 7-3. STATIONS WITH AIR LEAD CONCENTRATIONS GREATER THAN 1.0
1979 Max 1980 Max 1981 Max 1982
No. of quarters qtrly No. of quarters qtrly No. of quarters qtrly Mo. of quarters
Station* >1.0 >1.5 ave >1.0 >1.5 ave >1.0 >1.5 ave >1.0 >1.5
BtnNinghaa, AL
Leeds, AL
w II
Troy, AL
Fairbanks, AK
Fairbanks, AK
Glendale, A?
Phoenix, AZ
ii ii
ii n
ii ii
ii ii
Scottsdale, AZ
Tucson, AZ
Nogales, AZ
Los Angeles, CA
Anahein, CA
Lennox , CA
Los Angeles, CA
Los Angeles , CA
Lynwood, CA
Pico Rivera, CA
Adams Co, CD
Arapahoe Co, CO
Arvada, CO
Brighton, CO
Colorado Springs,
CO
Denver, CO
"
"
11
"
"
(028)
(004)
(005)
(003)
(010)
(016)
(001)
(002A)
(002G)
(004)
(013)
(017)
(003)
(DOS)
(004)
(001)
(001)
(001)
(103A)
(1031)
(001)
(001)
(001)
(001)
(001)
(001)
(004)
(003)
(002)
(003)
(009)
(010)
(012)
2
1
1
2
2
2
2
1
1
1
2
1
1
1
1
2
4
3
1
2
2
2
0
1
0
0
0
i
0
1
0
1
0
0
1
3
1
1
1
1
1 0
2.78 2 1 1.89 4 3 3.34 4 4
1 0
1 0
1.06
1.54 2 0 1.29 1 0 1.17
2.59 2 0 1.49 2 0 1.39
1.48 1 0 1.04
1.55 1 0 1.06
1 0
1.41 1 0 1.13 1 0 1.08
MB
1 0 1. 10
1.51 2 0 J.43
1.11
2 1
1 1
1 0
1 0
1 0
.77
.10
.60
.17
.37
.70
.47 2 1 1.53
2.13 1 0 1.03
1.57 2 0 1.23
1.67
1.67 1 0 1.10
Max 1983 Max 1964 Max
qtrly No. of quarters ctrly No. of quarters qtrly
ave >1.0 >1.5 ave >1.0 >1 5 ave
1.32 1 0 1.04
2 2 3.04 1 1 5.33
3 2 4.17 2 1 2.96
3.67 3 2 5.44 4 3 7.08
1.01
1.09
1.24 1 0 1.08 1 0 1.29
1.68 2 0 1.10
1.65
1.05
1 25 1 0 1.03 1 0 1.27
1.15 1 0 1.02
-------
TABLE 7-3. (continued)
1979 Max 1980 Max 1981 Max 1982 Max 1983 Max 1984 Max
do. of quarters qtrly No. of quarters qtrly No. of quarters qtrly No. of quarters qtrly No. of quarters qtrly No. of quarters qtrly
Englewood, CO
Garfield. CO
Grand Junction, CO
LongMnt , CO
Pueblo, CO
H a
Routt Co. CO
Nev Haven, CT
Watertaury. CT
Wilmington, OE
Washington, DC
Dade Co, FL
Dade Co, FL
MiMi, FL
Perriiw, FL
Hillstaorough, FL
Jacksonville, FL
Taw*, FL
Ta»pa, FL
Boise, 10
Kellogg, 10
H H
Slwshone Co, ID
(001)
(001)
(010)
(001)
(001)
(003)
(003)
(123)
(123)
(002)
(005)
(007)
(008)
(Oil)
(015)
(017)
(020)
(024)
(016)
(002)
(082)
(084)
(043)
(060)
(003)
(004)
(006)
(015)
(016)
(017)
(020)
(021)
(027)
f 1 - U
1
1
2
2
1
1
1
3
2
2
1
4
1
2
2
1
1
3
1
2
3
4
4
2
1
4
2
4
4
••A. J
I
0
1
0
0
0
0
0
0
0
1
0
0
0
0
0
0
0
4
0
1
4
dve
.80
.20
.53
.07
.03
.03
.33
.57
.41
1.21
1.49
1.89
1.90
1.44
1.06
1.4S
1.16
1.46
1.01
1.31
1.60
9.02
8.25
1.21
2.27
4.57
4.11
13.54
10.81
1 0 1.27
1 fl 1 11
1 1 1.51 3 U * J3
2 0 1. 10
1 ° 1'09 3 j j ?2 2 0 1.15 2 0 1.26
1 0 1.07
1 0 1.10
1 0 1.01
2 6.88
4 4 8. 72 4 4 6.67
1 0 1.02
3 3.33 2 2 1.54
2 2. 15 1 0 1.49
4 4 13.67 4 4 11.82 21 1 75
3 7.18
I
I—»
-p»
-------
TABLE 7-3. (continued)
Station 9
Chicago, IL
Cicero, IL
Elgin, IL
Granite City, IL
II H
H n
II II
Jeffersonville, IN
East Chicago, IN
U II
II H
II II
HaMond, IN
n n
II M
Indianapolis, IN
Council Bluffs, IA
Des Koines, IA
Buechel, KY
Coving ton, KY
H H
Greenup Co, KY
Jefferson Co, Ky
Louisville, KY
11
11
11
11
"
Newport, KY
Okolona, KY
(022)
(030)
(005)
(036)
(037)
(001)
(004)
(007)
(009)
(010)
(Oil)
(001)
(001)
(003)
(004)
(006)
(004)
(006)
(Oil)
(030)
(017)
(051)
(001)
(001)
(008)
(003)
(029)
(004)
(009)
(019)
(020)
(021)
(028)
(002)
(001)
1979 Max 1980
No. of quarters qtrly No. of quarters
1
1
1
1
1
4
4
4
3
2
2
1
2
2
1
1
1
2
1
1
1
1
1
1
1
1
0
0
0
0
0
0
4
0
0
0
1
0
0
0
0
0
0
0
0
0
0
0
0
0
1
1.05
1.02
1.14
1.00
1.04
1.15
3.17
1.33
1.38
2.19
1.42
1.67
1.34
1.18
1.46
1.16
1.30
1.12
1.16
1.42
1.05
1.01
1.29
1.06
1.06
1.51
1 0
1 0
1 1
3 2
1 0
1 0
1 0
1 1
1 1
1 1
1 1
1 1
1 1
2 1
Max 1981 Max 1982 Max 1983 Max 1984 Max
qtrly No. of quarters qtrly No. of quarters qtrly No. of quarters qtrly No. of quarters qtrly
1.02
1.06
1.95
2.97 4 3 7.27 1 0 1.48
1.43 1 0 1.13
1.04
3 2 2.95
1 1 1.59 1 0 1.20
1.41
1.78
2.41
1.75
1.59
2.52
1.42
2.31
-------
TABLE 7-3. (continued)
1979 Max 1980 Max 1981 Max 1982
No. of quarters qtrly No. of quarters qtrly No of quarters qtrly No. of quarters
Station i >1.0 >1 ,5 ave >1.0 >1.5 a»e >1.0 >1.S ave >1.0 >1.5
Paducha, KV
ii ii
St. Matthews, KY
Sluvely, KY
Baton Rouge, LA
Portland, HE
Anne Arundel Co, MO
ii »
Baltimore, MB
ii ii
II H
11 H
II II
Cheverly, MO
Essex, HO
HyatUville, MO
Springfield, MA
Springfield, HA
Boston. HA
Boston, MA
Eagan, HN
Minneapolis, HH
" "
Richfield, MM
St. Louis Park, MN
St. Paul. MN
ii H
Iron Co. MO
ii u
H ii
•' M
Jefferson Co, K>
11 I,
ii ii
H M
(004)
(020)
(004)
(002)
(OOZ)
(009)
<001)
(003)
(001)
(006)
(008)
(009)
(01B)
(004)
(001)
(001)
(002)
(007)
(002)
(012)
(001)
(027)
(055)
(004)
(007)
(031)
(038
(016)
(020)
(021)
(OK)
(005)
(009)
(Oil)
/nni
1
1
1
1
1
2
1
2
2
1
1
1
2
4
2
2
1
1
1
4
2
1
1
0
0
0
1
1
0
0
0
0
0
0
0
0
1
0
0
1
0
1
0
0
1.41
1.22
1.20 1 1 1.83
1.56
1.57
1.02
1.27
1.45
1.06
1.09
1.24
1.08
1.12
1.51
1.15
1.18
1.68 1 0 1.04
1 0
2 0
1.01
2.44
3 2.41 3 1 1.52
1.95 2 0 1.18
2.87 4 3.04
1.04
1-36 3 1.82 2 2 311 1 1
Max 1983 Max 1984 Max
qtrly No. of quarters qtrly No. of quarters qtrly
a« >1.0 >1.S ave >1.0 >1.5 ave
1.26 1 0 1.00 1 0 1 29
1.08
1 0 1.01
7 97
1 1 2.39 3 2 2.21
2 1 1.73
3 2 2.85 1 0 1.26
2 1 1.54 1 0 1.10
4 3 4.J3 1 0 1.28
1 0 1.26 1 0 1.11
4 3 6.70 1 1 2.41
2 1 2.56 1 1 1.60
-------
TABLE 7-3. (continued)
station
LewisiClark Co, NT
LewisiClark Co, NT
M II
II II
tl II
LewisUlark Co, NT
LewisiClark Co, NT
ii ii
Omaha, NE
Omaha, NE
Las Vegas, NV
Clifton, NJ
Newark, NJ
New Brunswick, NJ
Perth Amtray, NJ
PaUrson, KJ
Elizabeth, NJ
Sale* Co, NJ
Albuquerque, NM
Dona Ana Co, NM
Orange Co, NY
Yonkers, NY
Cincinnati, OH
Portland, OR
Laureldale, PA
Reading, PA
E.Coneiaugh, PA
Throop. PA
Lancaster City, PA
New Castle, PA
Montgomery Co, PA
Pottstown, PA
Phi la , PA
it ii
(002)
(007)
(008)
(714)
(716)
(719)
(722)
(724)
(Oil)
(034)
(001)
(002)
(001)
(003)
(001)
(001)
(002)
(003)
(022)
(015)
(001)
(001)
(001)
(082)
(717)
(712)
(804)
(019)
(315)
(015)
(103)
(101)
(026)
(028)
1979 Max 1980
No. of quarters qtrly No. of quarters
4
1
1
1
1
1
1
1
1
4
1
3
3
1
1
1
1
3
4
' *. J
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
4.19 4
1 0
1.08
1.15
1.17
1.08
1.42
1.16
1.08
1.15
3.30 2
1.11
1.28
1.13
1.18
1.01
1.23
1.16
1.21
2.71 3 0
Max 1981 Max 1982 Max 1983
qtrly No. of quarters qtrly No. of quarters qtrly No. of quarters
2.75 2 2 3.19 2
1.19
2
2
1
1
1
1
3
1
4
1.86 4 3 2.18 3
1.26 1 0 1.30 3
r I. •*
2
2
2
0
0
0
0
0
0
2
0
1
ave
2.25
2.69
2.34
1.21
1.17
1.24
1.01
1.34
1.03
1.63
1.49
1.57
4
2
.2
1
4
2
3
1
2
4
4
3
4
2
0
1
4
2
2
0
1
0
0
2
Max 1984
qtrly No. of quarters
3.12 4 4
5. 26 4 4
1.31 3 0
1.99
3.39
1.84 4 3
2.96 3 4
1.23
1 1
1.81
1.37 3 1
1.25 1 0
3.66 4 4
Max
qtrly
ave
3.87
6.83
1.48
4.63
3.23
1.73
1.58
1.40
5.13
-------
TABLE 7-3. (continued)
St.it ion 4
(031)
(038)
Guaynatao Co, PR (001)
Ponce, PR (002)
San Juan Co. , PR (003)
E. Providence, RI (008)
Providence, RI (007)
(015)
Greenville, SC (001)
Memphis, TN (035)
Nashville/Davidson,
TN (006)
San Antonio, TX (034)
Dallas, TX (018)
El
(029)
(035)
(046)
(049)
(050)
(057)
(060)
Paso, TX (002 A)
(002F)
(002G)
(018)
(021)
(022)
(023)
(027)
(028)
(030)
(031)
(033)
(037)
1979
No. of quarters
>1 n vi E
2
1
2
1
4
2
4
1
2
1
2
1
1
4
2
1
2
2
2
1
1
1
•» A. y
0
0
0
0
0
0
0
0
1
0
0
0
0
0
1
1
0
0
1
1
Max 1980 Max 1981 Max 1982
qtrly No. of quarters qtrly No. of quarters qtrly No. of quarters
ave >1.0 >1.5 ave >1.0 >1.5 ave >1.0 >1.5
1.29
1.06
1.60 1 0 1.06 1 0 1.02
1.08
3.59 1 1
1.10
1.92 2 0 1.16 20
1.34
1.38
1 0
1.05
1.23
1.59
1.07
1.12
1.22
1.01
1.13
1.90 2.12
1.90 4 1 1.79
2.60
1.91
1.02
1.84
2.12
2.15 2 1.74 4 2 1.75
1 0 1.16
1.02
2.47
1.97
Max 1983 Max 1984 Max
qtrly No. of quarters qtrly No. of quarters qtrly
1.69 4 0 1.27 4 0 1.30
1.11
1.30 1 0 1.17 3 1 6.19
1 0 1.23
1 0 1.44 3 1 l.M
2 1 1.37
1 0 1.39
3 1 1.54 31 1.54
1 0 1.02
-------
HBLE 7-3. (continued)
Houston, TX
ii ii
it ii
41 II
Ft. Worth, IX
Seattle, WA
Taco«, WA
Charleston, WV
Station »
(001)
(002)
(037)
(M9)
(003)
(057)
(004)
(001)
1979 Max 1980 Max 1981 Max 1982 Max 1983 Max 1984 Max
No. of quarters qtrly No. of quarters qtrly Ho. of quarters qtrly No. of quarters qtrly Ho. of quarters qtrly No. of quarters qtrly
>1.0 >1.5 ave >1.0 >1.S ave >1.0 >1.5 ave >l.O >1.5 ave >1.0 >1.5 ave >1.0 >1.S ave
2 0 1.35
2 0
1 0
3 0
2 C
1 0
.39
• 26 3 1 1 €0
.13 1 1 1.96
.14 . „ , „,
.36 i u l.m
1 0 1.06
1 0 1.09
"Where data are not given, reported quarterly averages were less than 1.0
valid quarterly averages.
or there were insufficient reports for
-------
near stationary point sources such as smelters, or near routes of high traffic density. Both
situations are represented in Table 7-3; e.g., the Idaho data reflect predominantly stationary
source emissions, whereas the Washington, D.C. data reflect predominantly vehicular emissions.
Table 7-4 summarizes the maximum quarterly lead values for those stations reporting 4
valid quarters in 1979, 1980, and 1981, grouped according to principal exposure orientation or
influence—population, stationary source, or background. The sites located near stationary
sources clearly dominate the concentrations over 2.0 (jg/m3; however, new monitor siting guide-
lines, discussed in Section 7.2.1.3.2, will probably effect some increase in the upper end of
the distribution of values from population-oriented sites by adding monitoring sites closer to
traffic emissions.
The effect of the 1978 National Ambient Air Quality Standard for Lead has been to reduce
the air concentration of lead in major urban areas. Similar trends may also be seen in urban
areas of lesser population density. Continuous monitoring at non-urban stations has been in-
sufficient to show a trend at more than a few locations. There are two reports that reflect a
trend toward decreasing atmospheric lead concentrations. Eisenreich et al. (1986) report de-
creasing concentrations of lead in rain during the period 1979-83 from 29 to 4.3 ug/liter in
urban areas and 5.7 to 1.5 in rural areas. All sites were in Minnesota. Trefry et al. (1985)
reported a decrease in the lead concentration of Mississippi River sediment layers for the
post 1970 period. They estimated that the Mississippi River carried 40 percent less lead in
suspended sediments in 1982-83 than in 1974-1975.
7.2.1.3 Changes in Air Lead Prior to Human Uptake. There are many factors that can cause
differences between the concentration of lead measured at a monitoring station and the actual
inhalation of air by humans. The following sections show that air lead concentrations usually
decrease with vertical and horizontal distance from emission sources, and are generally lower
indoors than outdoors. A person working on the fifth floor of an office building would be ex-
posed to less lead than a person standing on a curb at street level. The following discus-
sions will describe how these differences can affect individual exposures in particular cir-
cumstances.
7.2.1.3.1 Airborne particle size distributions. The effects of airborne lead on human health
and welfare depend upon the sizes of the lead-containing particles. As discussed in Chapter
6, large particles are removed relatively quickly from the atmosphere by dry and wet deposi-
tion processes. Particles with diameter smaller than a few micrometers tend to remain air-
borne for long periods (see Section 6.3.1).
Figure 7-5 summarizes airborne lead particle size data from the literature (Davidson and
Osborn, 1984). Minimum and maximum aerodynamic particle diameters of 0.05 |jm and 25 urn, re-
spectively, have been assumed unless otherwise specified in the original reference. Note that
most of the airborne lead mass is associated with small particles. There is also a distinct
7-20
-------
TABLE 7-4. DISTRIBUTION OF AIR LEAD CONCENTRATIONS BY TYPE OF SITE FOR 1980-83
Category
Neighborhood scale
Middle scale
Stationary source
Microscale roadside
Other1
Total
iO.5
38
13
99
5
666
820
Concentration
>0.5 >
20
14
25
12
190
262
ranges (ug/m3)
1.5 ^2! 0
3 0
6 0
13 5
8 4
30 15
59 24
>2.0
1
0
17
1
4
24
Total no. of
site-years
62
33
159
30
905
1189
Percentage of sites in
concentration range 69% 22% 5% 2% 2% 100%
1Data are the number of site years during 1980-83 falling within the designated quarterly
average concentration range. To be included, a site year must have four valid quarters
of data.
Source: SAROAD system.
peak of large particles in the upper end of many of the distributions. Two separate cate-
gories of sources are responsible for these distributions: the small particles result from
nucleation of vapor phase lead emissions (predominantly automotive), while the larger parti-
cles may originate directly from soil dust, coal particles, and other coarse materials, or
indirectly by the attachment of anthropogenically emitted submicron particles with high lead
content to larger particles, such as soil particles. Large particle peaks may also indicate
fly ash with a surface coating of condensed lead (Linton et al., 1980).
Information associated with each in the distributions in Figure 7-5 may be found in Table
7A-1 of Appendix 7A. The first six distributions were obtained by an EPA cascade impactor
network established in several cities during the calendar year 1970 (Lee et al., 1972). These
distributions represent the most extensive size distribution data base available. However,
the impactors were operated at excessive air flow rates that most likely resulted in particle
bounceoff, biasing the data toward smaller particles (Dzubay et al., 1976). Many of the later
distributions, although obtained by independent investigators with poorly defined quality con-
trol, were collected using techniques that minimize particle bounceoff and hence may be more
reliable. It is important to note that a few of the distributions were obtained without back-
up filters that capture the smallest particles. These distributions are likely to be inaccu-
rate, since an appreciable fraction of the airborne lead mass was probably not sampled. The
distributions of Figure 7-5 have been used with published lung deposition data to estimate the
7-21
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» QKCAT SMOKIES
NAT'L PARK. TN
0.1 1 10 0.01 0.1 1 10 0.01 0.1 1 10 0.01 0.1 1 10 0.01 0.1 1 10
Figure 7-5. Airborne mass size distributions for lead taken from the literature. AC represents
the airborne lead concentration in each size range. Cj is the total airborne lead concentra-
tion in all size ranges, and dp is the aerodynamic particle diameter.
Source: Davidson and Osborn (1984).
7-22
-------
fraction of inhaled airborne lead deposited in the human respiratory system (see Section
10.2.1).
7.2.1.3.2 Vertical gradients and siting guidelines. New guidelines for placing ambient air
lead monitors went into effect in July, 1981 (C.F.R. (1984) 40: §58, see Section 4.2.1).
"Microscale" sites, placed between 5 and 15 meters from thoroughfares and 2 to 7 m above the
ground, are prescribed, but until now few monitors have been located close to heavily traveled
roadways. Many of these microscale sites might be expected to show higher lead concentrations
than that measured at nearby middlescale urban sites, due to vertical gradients in lead con-
centrations near the source. One study (PEDCo, 1981) gives limited insight into the relation-
ship between a microscale location and locations further from a roadway. The data in the
lower half of Table 7-5 summarize total suspended particulates and particulate lead concen-
trations in samples collected in Cincinnati, Ohio, on 21 consecutive days in April and May,
1980, adjacent to a 58,500 average daily traffic (ADT) expressway connector. Simple interpo-
lation indicates that a microscale monitor as close as 5 meters from the roadway and 2 m above
the ground would record concentrations some 20 percent higher than those at a "middle scale"
site 21.4 m from the roadway. On the other hand, these data also indicate that although lead
concentrations very close to the roadway (2.8 m setback) are quite dependent on the height of
the sampler, the averages at the three selected heights converge rapidly with increasing dis-
tance from the roadway. In fact, the average lead concentration (1.07 ug/m3) f°r tne one mon"
itor (6.3 m height, 7.1 m setback) that satisfies the microscale site definition does not
prove to be significantly different from the averages for its two companions at other heights
but the same 7.1 m setback, or from the averages for any of the three monitors at the 21.4 m
setback. It also appears that distance from the source, whether vertical or horizontal, can
be the primary determining factor for changes in air lead concentrations. At 7.1 m setback
distance, the samplers at heights of 1.1 and 6.3 m would be about 7 and 11 m, respectively,
from the road surface. The values at these vertical distances are only slightly lower than
the corresponding values for comparable horizontal distances.
Other urban locations around the country with their own characteristic wind flow patterns
and complex settings, such as multiple roadways, may produce situations where the microscale
site does not record the highest concentrations. Collectively, however, the addition of these
microscale sites to the nation's networks can be expected to shift the distribution of report-
ed quarterly averages toward higher values. This shift will result from the change in com-
position of the networks and is a separate phenomenon from downward trend at long-established
sites described above, reflecting the decrease in lead additives used in gasoline.
Two other studies show that lead concentrations decrease with vertical distance from the
source (PEDCo Environmental, 1977; Sinn, 1980). PEDCo Environmental (1977) measured lead
concentrations at heights of 1.5 and 6.1 m at sites in Kansas City, MO and Cincinnati, OH (top
7-23
-------
TABLE 7-5. VERTICAL DISTRIBUTION OF LEAD CONCENTRATIONS
Location
Kansas City
east of road
west of road
Cincinnati
east of road
west of road
Cincinnati0
Cincinnati0
Cincinnati0
Setback
distance
(m)
,4
3.0d
3.0d
3.0d
3.0d
2.8
7.1
21.4
Height
(m)
6.1
1.5
6.1
1.5
6.1
1.5
6.1
1.5
10.5
6.3
1.1
10.5
6.3
1.1
10.5
6.3
1.1
Effective'
distance
from
source
(m)
6.4
3.2
6.4
3.2
6.4
3.2
6.4
3.2
10.4
6.4
2.9
12.3
9.2
7.1
23.6
22.2
21.4
l
Air lead
cone.
(ug/m3)
1.7
2.0
1.5
1.7
0.9
1.4
0.6
0.8
0.81
0.96
1.33
0.93
1.07
1.16
0.90
0.97
1.01
Ratio to
source
0.85
Se
0.88
S
0.64
S
0.75
S
0.61
0.72
S
0.69
0.80
0.87
0.68
0.73
0.77
Effective distance was calculated assuming the source was the edge of the roadway at a
height of 0.1 m.
Source: PEDCo Environmental (1977).
cSource: PEDCo (1981).
Assumed setback distance of 3.0 m.
Station closest to source used to calculate ratio.
half of Table 7-5). The sampling sites in Kansas City were described as unsheltered, unbiased
by local pollution influences, and not immediately surrounded by large buildings. The
Cincinnati study was conducted in a primarily residential area with one commercial street.
Samplers were operated for 24-hour periods; however, a few 12-hour samples were collected from
8 AM to 8 PM. Data were obtained in Kansas City on 35 days and in Cincinnati on 33 days. The
measured concentrations were greater at 1.5 m than at 6.1 m, and the difference between the
east side and west side of the street was approximately the same as the difference between 1.5
m and 6.1 m in height.
7-24
-------
Sinn (1980) investigated airborne lead concentrations at heights of 3 and 20 m above a
road in Frankfurt, Germany. Measurements conducted in December 1975, December 1976, and Janu-
ary 1978 gave monthly mean values of 3.18, 1.04, and 0.66 ug/m3, respectively, at 3 m height.
The corresponding values at 20 m height were 0.59, 0.38, and 0.31 ug/m3, showing a substantial
reduction at this height. The decrease in concentration over the 2-year period was attributed
to a decrease in the permissible lead content of gasoline from 0.4 to 0.15 g/liter beginning
in January, 1976.
Two reports show no relationship between air concentration and vertical distance
(Barltrop and Strehlow, 1976; Ter Haar, 1979). From August, 1975 to July, 1976, Barltrop and
Strehlow (1976) conducted an air sampling program in London at a proposed nursery site under
an elevated motorway. The height of the motorway was 9.3 m. Air samplers were operated at
five to seven sites during the period from Monday to Friday, 8 AM to 6 PM, for one year. The
maximum individual value observed was 18 ug/m3. The 12-month mean ranged from 1.35 to 1.51
ug/m3, with standard deviations of 0.91 and 0.66, respectively. The authors reported that the
airborne concentrations were independent of height from ground level up to 7 m.
Ter Haar (1979) measured airborne lead at several heights above the ground, using sam-
plers positioned 6 m from a heavily traveled road in Detroit. A total of nine 8-hour daytime
samples were collected. The overall average airborne lead concentrations at heights of 0.3,
0.9, 1.5, and 3.0 m were 4.2, 4.8, 4.7, and 4.6 MS/1"3, respectively, indicating a uniform con-
centration over this range of heights at the measurement site. It should be noted that at any
one height, the concentration varied by as much as a factor of 10 from one day to the next;
the importance of simultaneous sampling when attempting to measure gradients is clearly demon-
strated.
Data that show variations with vertical distance reflect the strong influence of the geo-
metry of the boundary layer, wind, and atmospheric stability conditions on the vertical gradi-
ent of lead resulting from automobile emissions. The variability of concentration with height
is further complicated by the higher emission elevation of smokestacks. Concentrations mea-
sured from sampling stations on the roofs of buildings several stories high may not reflect
actual human exposure conditions, but neither would a single sampling station located at
ground level among several buildings. The height variation in concentration resulting from
vertical diffusion of automobile emissions is likely to be small compared to temporal and
spatial variations resulting from surface geometry, wind, and atmospheric conditions. Our
understanding of the complex factors affecting the vertical distribution of airborne lead is
extremely limited, but the data of Table 7-5 indicate that air lead concentrations are pri-
marily a function of distance from the source, whether vertical or horizontal.
7-25
-------
7.2.1.3.3 Indoor/outdoor relationships and personal monitoring. Because people spend much of
their time indoors, ambient air sampled outdoors may not accurately represent actual inhala-
tion exposure to airborne lead. Table 7-6 summarizes the results of several indoor/outdoor
airborne lead studies. In nearly all cases, the indoor concentration is substantially lower
than the corresponding value outdoors; the only indoor/outdoor ratio exceeding unity is for a
high-rise apartment building, where air taken in near street level is rapidly distributed
through the building air circulation system. Some of the studies in Table 7-6 show smaller
indoor/outdoor ratios during the winter, when windows and doors are tightly closed. Overall,
the data suggest indoor/outdoor ratios of 0.6-0.8 are typical for airborne lead in houses
without air conditioning. Ratios in air conditioned houses are expected to be in the range of
0.3-0.5 (Yocom, 1982). The available data imply that virtually all airborne lead found in-
doors is associated with material transported from the outside. Because of the complexity of
factors affecting infiltration of air into buildings, however, it is difficult to predict
accurately indoor lead concentrations based on outdoor levels. Rabinowitz et al. (1984) found
a correlation between indoor air lead in Boston homes and the amount of lead sold in gasoline
in Massachusetts.
Even detailed knowledge of indoor and outdoor airborne lead concentrations at fixed loca-
tions may still be insufficient to assess human exposure to airborne lead. The study of
Tosteson et al. (1982) in Table 7-6 included measurement of airborne lead concentrations using
personal exposure monitors carried by individuals going about their day-to-day activities. In
contrast to the lead concentrations of 0.092 and 0.12 ug/m3 at fixed locations, the average
personal exposure was 0.16 ug/m3. The authors suggest this indicates an inadequacy of using
fixed monitors at either indoor or outdoor locations to assess exposure.
Rohbock (1981) reported that, whereas a parked car may exhibit properties similar to
buildings in reducing internal air concentrations, a moving car quickly reaches the same air
lead concentration inside as outside, suggesting a rapid exchange of air in a moving vehicle.
7.2.2 Lead in Soil
Much of the lead in the atmosphere is transferred to terrestrial surfaces where it 1s
eventually passed to the upper layer of the soil surface. The mechanisms that determine the
transfer rate of lead to soil are described in Section 6.4.1 and the transformation of lead 1n
soil in Section 6.5.1. The uptake of lead by plants and its subsequent effect on animals may
be found in Sections 8.3 and 8.4, respectively. The purpose of this section is to discuss the
distribution of lead in U.S. soils and the impact of this lead on potential human exposures.
7-26
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TABLE 7-6. COMPARISON OF INDOOR AND OUTDOOR AIRBORNE LEAD CONCENTRATIONS
Airborne lead concentration
(ug/m3)
Type of building Indoor . Outdoor
Library
City hall
Office building 1
Office building 2
House 1
House 2
Apartment building 1
Second floor
Roof
Apartment building 2
Third floor
Eleventh floor
Eighteenth floor
Roof
1.12
1.31
0.73
0.55
1.37
0.94
1.46
1.50
—
1.68
1.86
—
2.44
1.87
1.44
1.09
2.48
1.34
2.67
1.38
1.21
—
—
1.42
Indoor/outdoor
ratio Location Ref
0.46
0.70
0.51
0.51
0.55
0.70
0.55
1.09
--
—
—
--
Hartford, CT 1
n
n
"
n
n
New York, NY 2
n
n
n
n
n
New air conditioned
apartment
Older non-air condi-
tioned apartment
Air conditioned public
building
Non-air conditioned
storeroom in public
building
Houses
University buildings
Public schools
Store
Commercial office
Houses
Houses with gas stoves
Houses with electric
stoves
Office buildings
0.12-0.40 0.13-0.50
0.14-0.51 0.17-0.64
0.15-0.79 0.33-1.18
0.45-0.98 0.38-1.05
0.092
0.12
0.82
0.87
0.63
0.81
0.53
0.28
0.28
0.31
0.27
0.74
0.65
0.68
0.42
New York, NY
n
n
M
Pittsburgh, PA
n
n
n
n
Topeka, KS
Boston, MA
n
n
3
4
5
6
7-27
-------
TABLE 7-6. (continued)
Type of building
House 1
Before energy conser-
Airborne lead concentration
(ug/m3)
Indoor Outdoor
Sources:
1. Yocom et al., 1971.
2. General Electric Company, 1972.
3. Halpern, 1978.
4. Cohen and Cohen, 1980.
5. Tosteson et al., 1982.
6. Geomet, Inc., 1981.
7. Berk et al., 1981.
Indoor/outdoor
ratio Location
Ref
vation retrofit
After energy conser-
vation retrofit
House 2
Before energy conser-
vation retrofit
After energy conser-
vation retrofit
0.039
0.037
0.035
0.038
0.070
0.084
0.045
0.112
0.56
0.44
0.78
0.34
Medford, OR 7
ii
n
n
7.2.2.1. Typical Concentrations of Lead in Soil
7.2.2.1.1 Lead in urban, smelter, and rural soils. Shacklette et al. (1971) sampled soils at
a depth of 20 cm to determine the elemental composition of soil materials derived from the
earth's crust, not the atmosphere. The range of values probably represent natural levels of
lead in soil, although there may have been some contamination with anthropogenic lead during
collection and handling. Lead concentrations in soil ranged from less than 10 to greater than
70 ug/g. The arithmetic mean of 20 pg/g and geometric mean of 16 ug/g reflect the fact that
most of the 863 samples were below 30 ug/g at this depth. McKeague and Wolynetz (1980) found
the same arithmetic mean (20 pg/g) for 53 uncultivated Canadian soils. The range was 5 to 50
H9/g and there was little variation with depth between the A, B and C horizons in the soil
profile.
Studies discussed in Section 6.5.1 have determined that atmospheric lead is retained in
the upper 2-5 cm of undisturbed soil, especially soils with at least 5 percent organic matter
and a pH of 5 or above. There has been no general survey of this upper 2-5 cm of the soil
surface in the United States, but several studies of lead in soil near roadsides and smelters
7-28
-------
and a few studies of lead in soil near old houses with lead-based paint can provide the back-
gound information for determining potential human exposures to lead from soil.
Because lead is immobilized by the organic component of soil (Section 6.5.1), the concen-
tration of anthropogenic lead in the upper 2-5 cm is determined by the flux of atmospheric
lead to the soil surface. Near roadsides, this flux is largely by dry deposition and the rate
depends on particle size and concentration. These factors vary with air concentration and
average windspeed (see Section 6.4.1). In general, deposition drops off abruptly with in-
creasing distance from the roadway. This effect is demonstrated in studies that show that
surface soil lead decreases exponentially up to 25 m from the edge of the road. The original
work of Quarles et al. (1974) showed decreases in soil lead from 550 to 40 ug/g within 25 m
alongside a highway with 12,500 vehicles/day in Virginia. Pierson and Brachaczek (1976) found
that lead concentrations in topsoil adjacent to a major artery decreased exponentially from 0
to 12 m away from the highway (Figure 7-6). These findings were confirmed by Wheeler and
Rolfe (1979), who observed an exponential decrease linearly correlated with traffic volume.
Agrawal et al. (1981) found similar correlations between traffic density and roadside proxim-
ity in Baroda City, as did Garcia-Miragaya et al. (1981) in Venezuela and Wong and Tarn (1978)
in Hong Kong. Little and Wiffen (1978) found additional relationships between particle size
and roadside proximity and decreases with depth in the soil profile. The general conclusion
from these studies is that roadside soils may contain atmospheric lead from 30-2000 ug/g in
excess of natural levels within 25 m of the roadbed, all of which is in the upper layer of the
soil profile. It is assumed that particles deposited directly on the roadway are washed to
the edge of the pavement, but do not migrate beyond the shoulder.
Near primary and secondary smelters, lead in soil decreases exponentially within a
5-10 km zone around the smelter complex. Soil lead contamination varies with the smelter
emission rate, stack height, length of time the smelter has been in operation, prevailing
windspeed and direction, regional climatic conditions, and local topography (Roberts, 1975).
Little and Martin (1972) observed decreases from 125 to 10 ug/g in a 6 km zone around a
smelting complex in Great Britain; all of the excess lead was in the upper 6 cm of the soil
profile. Roberts (1975) reported soil lead between 15,000 and 20,000 ug/g near a smelter in
Toronto. Kerin (1975) found 5,000-9,000 ug/g adjacent to a Yugoslavian smelter; the contami-
nation zone was 7 km in radius. Ragaini et al. (1977) observed 7900 ug/g near a smelter in
Kellogg, Idaho; they also observed a 100-fold decrease at a depth of 20 cm in the soil pro-
file. Palmer and Kucera (1980) observed soil lead in excess of 60,000 ug/g near two smelters
in Missouri, decreasing to 10 ug/g at 10 km.
Urban soils may be contaminated from a variety of atmospheric and non-atmospheric
sources. The major sources of soil lead seem to be paint chips from older houses and deposi-
tion from nearby highways. Lead in soil adjacent to a house decreases with distance from the
7-29
-------
I
CO
O
o
£
1 I I I I I I I I \ I T
10*
I I I I I I I I I
4 6 8 10
METERS FROM EDGE OF ROAD
I I I
12
14
Figure 7-6. Change in soil lead concentrations with increasing distance from a
roadway.
Source: Pierson and Brachaczek (1976).
-------
house; this may be due to paint chips or to dust of atmospheric origin washing from the roof-
top (Wheeler and Rolfe, 1979).
Andresen et al. (1980) reported lead in the litter layer of 51 forest soils in the north-
eastern United States. They found values from 20-700 ug/g, which can be compared only qual-
itatively to the soil lead concentration cited above. This study clearly shows that the major
pathway of lead to the soil is by the decomposition of plant material containing high concen-
trations of atmospheric lead on or within their tissues. Because this organic matter is a
part of the decomposer food chain, and because the organic matter is in dynamic equilibrium
with soil moisture, it is reasonable to assume that lead associated with organic matter is
biologically more mobile than lead tightly bound within the crystalline structure of inorganic
rock fragments.
Finally, a definitive study that describes the source of soil lead was reported by Gulson
et al. (1981) for soils in the vicinity of Adelaide, South Australia. In an urban to rural
transect, stable lead isotopes were measured in the top 10 cm of soils over a 50 km distance.
By their isotopic compositions, three sources of lead were identified: natural, non-automo-
tive industrial lead from Australia, and tetraethyl lead manufactured in the United States.
The results indicated that most of the soil surface lead originated from leaded gasoline.
Similar studies have not been conducted in the United States.
7.2.2.1.2 Natural and anthropogenic sources of soil lead. Although no study has clearly
identified the relative concentrations of natural and anthropogenic lead in soil, a few clari-
fying statements can be made with some certainty. Lead may be found in inorganic primary
minerals, on humic substances, complexed with Fe-Mn oxide films, on secondary minerals or in
soil moisture. All of the lead in primary minerals is natural and is bound tightly within the
crystalline structure of the minerals. Most of this lead can be released only by harsh treat-
ment with acids. The lead on the surface of these minerals is leached slowly into the soil
moisture. Atmospheric lead forms complexes with humic substances or on oxide films that are
in equilibrium with soil moisture, although the equilibrium strongly favors the complexing
agents. Consequently, the ratio of anthropogenic to natural lead in soil moisture depends
mostly on the amounts of each type of lead in the complexing agents and very little on the
concentration of natural lead in the inorganic minerals.
Except near roadsides and smelters, only a few micrograms of atmospheric lead have been
added to each square centimeter of soil surface. Several studies indicate that this lead is
available to plants (Section 8.3.1.1). Even with small amounts of atmospheric lead, as much
as 75 percent of the lead in soil moisture is of atmospheric origin (Elias et al., 1982). A
conservative estimate of 50 percent is used in the discussions in Section 7.3.1.2. A break-
down of the types of lead in soil may be found in Table 7-7.
7-31
-------
TABLE 7-7. SUMMARY OF SOIL LEAD CONCENTRATIONS
Atmospheric
Matrix
Total soil
Primary minerals
Humic substances*
Soil moisture
Natural
lead
8-25
8-25
20
0.0005
lead
Rural
3-5
-
60
0.0005
Urban
50-150
-
2000
0.0150
Total
lead
Rural
10-30
8-25
80
0.001
Urban
150-300
8-25
2000
0.0155
*Assumes 5% organic matter, pH 5.0; may also include lead in Fe-Mn oxide films.
7.2.2.2 Pathways of Soil Lead to Human Consumption
7.2.2.2.1 Crops. On the surfaces of vegetation, most lead may be of atmospheric origin. In
the internal tissues, lead may be a combination of atmospheric and soil origin. As with
soils, lead on vegetation surfaces decreases exponentially with distance away from roadsides
and smelters (Cannon and Bowles, 1962; Nasralla and Ali, 1985; see also Chapter 8). For many
years, plant surfaces have been used as indicators of lead pollution (Garty and Fuchs, 1982;
Pilegaard, 1978; Ratcliffe, 1975; Ruhling and Tyler, 1969; Tanaka and Ichikuni, 1982). These
studies all show that lead on the surface of leaves and bark is proportional to traffic den-
sity and distance from the highway, or more specifically, to air lead concentrations and par-
ticle size distributions. Other factors such as surface roughness, wind direction and speed
are discussed in Chapter 6. The data also show that lead in internal plant tissues is direct-
ly, although not always linearly, related to lead in soil. Nicklow et al. (1983) found a
linear relationship between extracted soil lead and several food crops.
In a study to determine the background concentrations of lead and other metals in agri-
cultural crops, the Food and Drug Administration (Wolnik et al., 1983, 1985), in cooperation
with the U.S. Department of Agriculture and the U.S. Environmental Protection Agency, analyzed
over 1500 samples of the most common crops taken from a cross section of geographic locations.
Collection sites were remote from mobile or stationary sources of lead. Soil lead concentra-
tions were within the normal range (8-25 ug/g) of U.S. soils. Extreme care was taken to avoid
contamination during collection, transportation, and analysis. The concentrations of lead in
crops found by Wolnik et al. (1983, 1985) are shown as "Total" concentrations in Table 7-8.
The breakdown by source of lead is discussed below. The total concentration data should
probably be seen as representing the lowest concentrations of lead in food available to
Americans. From harvest to packaging, the lead concentration in food increases by a factor of
2-12 (see Section 7.3.1.2). A small portion of this increase may occur because: (1) some
7-32
-------
TABLE 7-8. BACKGROUND LEAD IN BASIC FOOD CROPS AND MEATS
(ug/g fresh weight)
Crop
Wheat
Potatoes
Field corn
Sweet corn
Soybeans
Peanuts
Onions
Rice
Carrots
Tomatoes
Spinach
Lettuce
Beef (muscle)
Pork (muscle)
Natural
Pb
0.0015
0.0045
0.0015
0.0015
0.021
0.005
0.0023
0.0015
0.0045
0.001
0.0015
0.0015
0.0002
0.0002
Indirect
atmospheric
0.0015
0.0045
0.0015
0.0015
0.021
0.005
0.0023
0.0015
0.0045
0.001
0.0015
0.0015
0.002
0.002
Direct
atmospheric
0.034
--
0.019
--
—
—
—
0.004
—
--
0.042
0.010
0.02
0.06
Totalt
0.037
0.009
0.022
0.003
0.042
0.010
0.0046
0.007
0.009
0.002
0.045
0.013
0.02*
0.06*
'Except as indicated, data are from Wolnik et al. (1983, 1985).
*Data from Penumarthy et al. (1980).
crops are grown closer to highways and stationary sources of lead than those sampled by Wolnik
et al. (1983, 1985); (2) some harvest techniques used by farmers might add more lead to the
crop than did Wolnik et al.; and (3) some crops are grown on soils significantly higher in
lead than those of the Wolnik et al. study because of a history of fertilizer additions or
sludge applications.
Because the study reported by Wolnik et al. was a systematic effort that covers a broad
spectrum of agricultural practices in the United States and was conducted with appropriate
attention to quality assurance, it serves in this report as the sole basis for background crop
data. There are many other reputable studies that describe the impact of lead on crops under
specific circumstances or with a variety of control measures. Generally, these studies report
that the lead concentrations are highest in leafy crops, lowest in fruits, with root crops
somewhat intermediate (e.g., Nicklow et al., 1983). It is important to recognize that root
crops such as radishes and potatoes are specialized structures for the storage of photosynthe-
tic products, and are functionally different from the roots that absorb water and nutrients.
These latter roots usually have lead concentrations higher than shoots or leaves and form a
reasonably effective barrier to soil lead. Reports of lead in food crops from other countries
have found patterns similar to those in the United States (Nasralla and Ali, 1985; Wong and
Koh, 1982).
7-33
-------
Studies that specifically apply to roadside or stationary source conditions can be evalu-
ated in the context of these recent background findings by Wolnik et al. (1983, 1985).
Studies of the lead associated with crops near highways have shown that both lead taken up
from soil and aerosol lead delivered by deposition are found associated with the edible por-
tions of common vegetable crops. However, there is enormous variability in the total amount
of lead associated with such crops and in the relative amounts of lead in the plants versus on
the plants. The variability depends upon several factors, the most prominent of which are the
plant species, the traffic density, the meteorological conditions, and the local soil condi-
tions (Welch and Dick, 1975; Rabinowitz, 1974; Dedolph et al. , 1970; Motto etal., 1970;
Schuck and Locke, 1970; Ter Haar, 1970). These factors, coupled with the fact that many
studies have not differentiated between lead on plants versus lead in plants, make it diffi-
cult to generalize on the relationship between lead in crops and lead in soil or air. Data of
Schuck and Locke (1970) suggest that in some cases (e.g., tomatoes and oranges) much of the
surface lead is readily removed by washing. But as noted in Section 6.4.3, this is not uni-
versally true; in some cases, much more vigorous washing procedures would be necessary to re-
move all or most of the surface lead.
Ter Haar (1970) found that inedible portions of several plants (bean leaves, corn husks,
soybean husks, and chaff from oats, wheat, and rice) had two to three times the lead concen-
tration when grown near a busy highway compared with similar plants grown 160 m from the high-
way or in a greenhouse supplied with filtered air. The edible portions of these and other
plants showed little or no difference in lead content between those grown in ambient air and
those grown in the filtered air. However, the lead concentrations found by Ter Haar (1970)
for edible portions of crops grown in filtered air in the greenhouse were generally one to two
orders of magnitude higher than those of the same types of crops taken from actual agricul-
tural situations by Wolnik et al. (1983, 1985). Dedolph et al. (1970) found that while rye-
grass and radish leaves grown near a busy highway contained deposited airborne lead, the
edible portion of the radish was unaffected by variations in either soil lead or air lead.
The accumulation of lead by edible portions of crops was measured by Ter Haar (1970), who
showed that edible plant parts not exposed to air (potatoes, corn, carrots, etc.) do not accu-
mulate atmospheric lead, while leafy vegetables do. These results were confirmed by McLean
and Shields (1977), who found that most of the lead associated with food crops is on leaves
and husks. The general conclusion from these studies is that lead associated with food crops
varies according to exposure to the atmosphere and in proportion to the effort taken to
separate husks, chaff, and hulls from edible parts during processing for human or animal con-
sumption.
To estimate the distribution of natural and atmospheric lead in food crops (Table 7-8),
it is necessary to recognize that some crops of the Wolnik et al. study have no lead from
7-34
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direct atmospheric deposition, but rather that all lead found in these crops comes through
soil moisture. The lowest concentrations of lead are found in those crops where the edible
portion grows above ground and it does not accumulate lead from atmospheric deposition (sweet
corn and tomatoes). Belowground crops are protected from atmospheric deposition but have
slightly higher concentrations of lead, partly because lead accumulates in the roots of plants
(potatoes, onions, carrots). Leafy aboveground plants (lettuce, spinach, wheat) have even
higher lead concentrations presumably because of increased exposure to atmospheric lead. The
assumption that can be made here is that, in the absence of atmospheric deposition, exposed
aboveground plant parts would have lead concentrations similar to sweet corn and tomatoes.
The data on these ten crops suggest that root vegetables have lead concentrations between
0.0046 and 0.009 ug/g. This is all lead of soil origin, of which presumably half is natural
and half anthropogenic (called indirect atmospheric lead here). Aboveground parts not exposed
to significant amounts of atmospheric deposition (sweet corn and tomatoes) have less lead in-
ternally, also equally divided between natural and indirect atmospheric lead. If it is
assumed that this same concentration is the internal concentration for aboveground parts for
other plants, it is apparent that five crops (wheat, field corn, rice, spinach, and lettuce)
have direct atmospheric deposition in proportion to surface area and estimated duration of
exposure. The deposition rate of only 0.04 ng/cm2«day, which is much smaller than would
normally be expected in rural environments (see Section 6.4.1) could account for these amounts
of direct atmospheric lead. In this scheme, soybeans are anomalously high. Soybeans grow
inside a sheath and should have an internal lead concentration similar to sweet corn.
These discussions lead to the conclusion that root parts and protected aboveground parts
of edible crops contain natural lead and indirect atmospheric lead, both derived from the
soil. For exposed aboveground parts, any lead in excess of the average found on unexposed
aboveground parts is considered to be the result of direct atmospheric deposition.
Near smelters, Merry et al. (1981) found a pattern different from roadside studies cited
above. They observed that wheat crops contained lead in proportion to the amount of soil
lead, not vegetation surface contamination. A similar effect was reported by Harris (1981).
7.2.2.2.2 Livestock. Lead in forage was found to exceed 950 ug/g within 25 m of roadsides
with 15,000 or more vehicles per day (Graham and Kalman, 1974). At lesser traffic densities,
200 ug/9 were found. Other reports have observed 20-660 ug/g with the same relationship to
traffic density and distance from the road (see review by Graham and Kalman, 1974). A more
recent study by Crump and Barlow (1982) showed that the accumulation of lead in forage is di-
rectly related to the deposition rate, which varied seasonally according to traffic density.
The deposition rate was measured using the moss bag technique, in which bags of moss are
exposed and analyzed as relative indicators of deposition flux. Rain was not effective in
removing lead from the surface of the moss. The ratio of atmospheric lead to total lead in
7-35
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meat products is partly a function of the same ratio in forage. The fact that most lead in
cattle is stored in bones and not eaten by man does not alter the ratio of atmospheric to
total lead in meat.
Factors that might add non-atmospheric lead would be soil ingestion by cattle, processed
food given to cattle in feedlots, and lead added during processing. Thornton and Abrahams
(1983) estimated that 1 to 18 percent of the dry matter ingested by cattle is soil, based the
titanium content of feces. Soil ingestion increases when overgrazing is permitted because of
dry weather, seasonal changes or other farm management practices. Most of the ingested soil
would be from the upper 1-5 cm. In a normal pasture this soil layer would contain a signifi-
cant fraction of atmospheric lead.
7.2.3 Lead in Surface and Ground Water
Lead occurs in untreated water in either dissolved or particulate form. Dissolved lead is
operationally defined as that which passes through a 0.45 urn membrane filter. Because atmos-
pheric lead in rain or snow is retained by soil, there is little correlation between lead in
precipitation and lead in streams which drain terrestrial watersheds. Rather, the important
factors seem to be the chemistry of the stream (pH and hardness) and the volume of the stream
flow. For groundwater, chemistry is also important, as is the geochemical composition of the
water-bearing bedrock.
Of the year-round housing units in the United States, 84 percent receive their drinking
water from a municipal or private supply of chemically treated surface or ground water. The
second largest source is privately owned wells (U.S. Bureau of the Census, 1982). In some
communities, the purchase of untreated bottled drinking water is a common practice. The ini-
tial concentration of lead in this water depends largely on the source of the untreated water.
7.2.3.1. Typical Concentrations of Lead in Untreated Water
7.2.3.1.1 Surface water. Durum et al. (1971) reported lead concentrations in the range of
1-55 |jg/l in 749 surface water samples in the United States. Very few samples were above
50 ug/1, and the average was 3.9 ug/1. Chow (1978) reviewed other reports with mean values
between 3 and 4 pg/l- The National Academy of Sciences (1980) reported a mean of 4 ug/1, with
a range from below detection to 890 pg/1. Concentrations of 100 pg/1 were found near sites of
sewage treatment, urban runoff, and industrial waste disposal.
Because 1 pg lead/1 was at or below the detection limit of most investigators during the
1970's, it is likely that the mean of 3-4 pg/1 was unduly influenced by a large number of
erroneously high values at the lower range of detection. On the other hand, Patterson (1980)
reports values of 0.006-0.05 |jg/l for samples taken from remote streams. Extreme care was
taken to avoid contamination and analytical techniques sensitive to less than 0.001 pg/1 were
used.
7-36
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Streams and lakes are influenced by their water chemistry and the lead content of their
sediments. At neutral pH, lead moves from the dissolved to the particulate form and the part-
icles eventually pass to sediments. At low pH, the reverse pathway generally takes place.
Hardness, which is a combination of the Ca and Mg concentration, also can influence lead con-
centrations. At higher concentrations of Ca and Mg, the solubility of lead decreases. Fur-
ther discussion of the chemistry of lead in water may be found in Sections 6.5.2.1 and 8.2.2.
7.2.3.1.2 Ground water. Municipal and private wells account for a large percentage of the
drinking water supply. This water typically has a neutral pH and somewhat higher hardness
than surface water. Lead concentrations are not influenced by acid rain, surface runoff, or
atmospheric deposition. Rather, the primary determinant of lead concentration is the geochem-
ical makeup of the bedrock that is the source of the water supply. Ground water typically
ranges from 1 to 100 |jg lead/1 (National Academy of Sciences, 1980). Again, the lower part of
the range may be erroneously high due to difficulties of analysis. It is also possible that
the careless application of fertilizers or sewage sludge to agricultural lands can cause con-
tamination of ground water supplies.
7.2.3.1.3 Natural vs. anthropogenic lead in water. Although Chow (1978) reports that the na-
tural lead concentration of surface water is 0.5 ug/1, this value may be excessively high. In
a discussion of mass balance considerations (National Academy of Sciences, 1980), natural lead
was suggested to range from 0.005 to 10 ug/1. Patterson (1980) used further arguments to
establish an upper limit of 0,02 ug/1 for natural lead in surface water. This upper limit
will be used in further discussions of natural lead in drinking water.
Because ground water is free of atmospheric lead, lead in ground water should probably be
considered natural in origin as it occurs at the well head, unless there is evidence of sur-
face contamination.
7.2.3.2 Human Consumption of Lead in Water. Whether from surface or ground water supplies,
municipal waters undergo extensive chemical treatment prior to release to the distribution
system. There is no direct effort to remove lead from the water supply. However, some treat-
ments, such as flocculation and sedimentation, may inadvertently remove lead along with other
undesirable substances. On the other hand, chemical treatment to soften water increases the
solubility of lead and enhances the possibility that lead will be added to water as it passes
through the distribution system.
7.2.3.2.1 Contributions to drinking water. For samples taken at the household tap, lead con-
centrations are usually higher in the initial volume (first daily flush) than after the tap
has been running for some time. Water standing in the pipes for several hours is intermediate
between these two concentrations (Sharrett et al., 1982; Worth et al., 1981). Common plumbing
materials are galvanized and copper pipe; lead solder is usually used to seal the joints of
7-37
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copper pipes. Lead pipes are seldom in service in the United States, except in the New
England states (Worth et al., 1981), and as a flexible fitting between the main line and the
house service pipe.
Average lead content of running water at the household tap is generally lower (8 ug/1)
than in some untreated water sources (25-30 pg/1) (Sharrett et al., 1982). Water treatment
removes lead associated with the suspended solids in raw surface waters. If first flush or
standing water is sampled, the lead content may be considerably higher. Sharrett et al.
(1982) showed that in both copper and galvanized pipes, lead concentrations were increased by
a factor of two when the sample was taken without first flushing the line (see Section
7.3.1.3).
The age of the plumbing is an important factor. New copper pipes with lead solder
exposed on the inner surface of the joints produce the highest amount of lead in standing
water. After about six years, this lead is either leached away or covered with calcium
deposits, and copper pipes subsequently have less lead in standing water than galvanized
pipes. Because lead pipes are rarely used in the United States, exposure from this source
will be treated as a special case in Section 7.3.2.1.4. The pH of the water is also impor-
tant; the acid water of some eastern and northwestern United States localities can increase
the leaching rate of lead from lead pipes or lead solder joints and prevent the buildup of a
protective coating of calcium carbonate plaque.
Table 7-9 summarizes the contribution of atmospheric lead to drinking water. In this
determination, the maximum reported value for natural lead (0.02 ug/1) was used, all addi-
tional lead in untreated water is considered to be of atmospheric origin, and it is assumed
that treatment removes 85 percent of the original lead, and that any lead added during distri-
bution is non-atmospheric anthropogenic lead.
7.2.3.2.2 Contributions to food. The use of treated water in the preparation of food can be
a significant source of lead in the human diet. There are many uncertainties in determining
this contribution, however. Water used in food processing may be from a municipal supply or a
private well. This water may be used to merely wash the food, as with fruits and vegetables,
or as an actual ingredient. Water lead may remain on food that is partially or entirely de-
hydrated during processing (e.g., pasta). Water used for packing or canning may be used with
the meal or drained prior to preparation. It is apparent from discussions in Section 7.3.1.3
that, considering both drinking water and food preparation, a significant amount of lead can
be consumed by humans from treated water. Only a small fraction of this lead is of atmos-
pheric origin, however.
7-38
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TABLE 7-9. SUMMARY OF LEAD CONCENTRATIONS IN DRINKING WATER SUPPLIES
(M9/D
Source
Untreated
Lakes
Rivers
Streams
Groundwater
Treated
Surface
Ground
Natural
lead
0.02
0.02
0.02
3
0.003
0.45
Indirect
atmospheric
lead
15
15
2.5
0
2.5
0
Direct
atmospheric
lead
10
15
2.5
0
1.5
0
Non- atmospheric
anthropogenic
lead
0
0
0
0
4
7.5
Total
lead
25
30
5
3
8
8
Source: Text
7.2.4 Summary of Environmental Concentrations of Lead
Lead concentrations in environmental media that are in the pathway to human consumption
are summarized in Table 7-10. These values are estimates derived from the preceding discus-
sions. A single value has been used, rather than a range, in order to facilitate further
estimates of actual human consumption. This use of a single value is not meant to imply a
high degree of certainty in its determination or homogeneity within the human population. The
units for water are converted from ug/1 as in Table 7-9 to ug/g to facilitate the discussions
of dietary consumption of water and beverages.
TABLE 7-10. SUMMARY OF ENVIRONMENTAL CONCENTRATIONS OF LEAD
Medium
Urban air (ug/m3)
Rural air (ug/m3)
Total soil (ug/g)
Food crops (M9/g)
Surface water (ug/g)*
Ground water (ug/g)*
Natural
lead
0.00005
0.00005
8-25
0. 0025
0. 00002
0.003
Atmospheric
lead
0.3-1.1
0.15-0.3
3-5
0.00-0.042
0.005-0.030
0.00
Total
lead
0.3-1.1
0.15-0.3
10-30
0.002-0.045
0.005-0.030
0.001-0.1
*Note change in units from Table 7-9.
7-39
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Because concentrations of natural lead are generally three to four orders of magnitude
lower than anthropogenic lead in ambient rural or urban air, all atmospheric contributions of
lead are considered to be of anthropogenic origin. Natural soil lead typically ranges from 10
to 30 ug/g, but much of this is tightly bound within the crystalline matrix of soil minerals
at normal soil pHs of 4-8. Lead in the organic fraction of soil is part natural and part
atmospheric. The fraction derived from fertilizer is considered to be minimal. In undis-
turbed rural and remote soils, the ratio of natural to atmospheric lead is about 1:1, perhaps
as high as 1:3. This ratio persists in soil moisture and in internal plant tissues. Thus,
some of the internal lead in crops is of anthropogenic origin, and some is natural. Informa-
tion on the effect of fertilizer on this ratio is not available. Lead in untreated surface
water is 99 percent anthropogenic. Except near municipal waste outfalls, this anthropogenic
lead is mostly atmospheric. It is possible that 75 percent of this lead is removed during
treatment. Lead in untreated ground water is presumed to be natural in the absence of evi-
dence of groundwater contamination.
In tracking air lead through pathways to human exposure, it is necessary to distinguish
between lead of atmospheric origin that has passed through the soil (indirect atmospheric
lead), and atmospheric lead that has deposited directly on crops or water. Because indirect
atmospheric lead will remain in the soil for many decades, this source is insensitive to pro-
jected changes in atmospheric lead concentrations. Regulation of ambient air lead concentra-
tions will not affect indirect atmospheric lead concentrations over the next several decades.
The method used in this document for calculating the relative contribution of atmospheric
lead to total potential human exposure relies partially on the relationship between air con-
centration and deposition flux described on Section 6.4. Estimates of contributions from
other sources are usually based on the observed value for total lead concentration from which
the estimated contribution of atmospheric lead is subtracted. The forms of lead subject to
the greatest human exposure are atmospheric lead, lead in food cans, and lead in paint pig-
ments. There is little evidence for the substantial contribution of other forms of anthropo-
genic lead to the total lead consumption by the general U.S. population.
7.3 POTENTIAL PATHWAYS TO HUMAN EXPOSURE
The preceding section discussed ambient concentrations of lead in the environment, focus-
ing on levels in the air, soil, food crops, and water. In this section, environmental lead
concentrations are examined from the perspective of pathways to human exposure (Figure 7-1).
Exposure is a measure of the amount of pollutant available at the interface between the human
7-40
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and the human environment. The estimation of exposure requires a knowledge of pollutant con-
centrations of each environmental component, the amounts of each environmental component con-
sumed, and the time budgets or other specific activities normal for humans (Moschandreas,
1981). For this analysis, a current baseline exposure scenario is described for an individual
with a minimum amount of daily lead consumption. It is assumed that this person lives and
works in a nonurban environment, eats a normal diet of food taken from a typical grocery
shelf, and has no habits or activities that tend to increase lead exposure. Without drastic
changes in lifestyle, lead exposure at the baseline level is considered unavoidable without
further reductions of lead in the atmosphere or in canned foods. Most of the baseline lead is
of anthropogenic origin, although a portion is natural, as discussed in Section 7.3.1.5.
7.3.1 Baseline Human Exposure
To arrive at a minimum or baseline exposure for humans, it is necessary to begin with the
environmental components (air, soil, food crops, and water) that are the major sources of lead
consumed by humans (Table 7-10). These components are measured frequently, even monitored
routinely in the case of air, so that many data are available on their concentrations. But
there are several factors that modify these components prior to actual human exposure. We do
not breathe air as monitored at an atmospheric sampling station. We may be closer to or far-
ther from the source of lead than is the monitor. We may be inside a building, with or with-
out filtered air; the water we drink does not come directly from a stream or river. It has
passed through a chemical treatment plant and a distribution system. A similar type of pro-
cessing has modified the lead levels present in our food.
It is inappropriate to assess human exposure to lead from a single source or through a
single medium without a simultaneous assessment from other sources (Laxon et al., 1985). Our
ability to monitor the environment depends on the available technology. But our knowledge of
human exposure depends on the correct understanding of the transfer of a pollutant from the
environmental component to the human body. In the past, exposure to air pollutants have been
interpreted strictly in the context of inhalation, with little consideration given to other
routes of exposure. This document attempts to assess the total human exposure to lead from
all sources and through all pathways.
Besides the atmospheric lead in environmental components, there are two other anthropo-
genic sources that contribute to this baseline of human exposure: paint pigments and lead
solder (Figure 7-7). Solder contributes directly to the human diet through canned food and
copper water distribution systems. Chips of paint pigments are discussed later under special
environments. But paint and solder are also a source of lead-bearing dusts. The most common
dusts in the baseline human environment are street dusts and household dusts. They originate
as emissions from mobile or stationary sources, as the oxidation products of surface exposure,
7-41
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I
-p.
t\>
CRUSTAL
WEATHERING
INDUSTRIAL
EMISSIONS
SURFACE AND
GROUND WATER
Figure 7-7. Paint pigments and solder are two additional sources of potential lead exposure which
are not of atmospheric origin. Solder is common even in baseline exposures and may represent 30
to 45 percent of the baseline human consumption. Paint pigments are encountered in older
houses and in soils adjacent to older houses.
-------
or as products of frictional grinding processes. Dusts are different from soil in that soil
derives from crustal rock and typically has a lead concentration of 10-30 ng/g, whereas dusts
come from both natural and anthropogenic sources and vary from 1,000 to 10,000 M9/9-
The discussion of the baseline human exposure traces the sequence from ambient air to in-
haled air, from soil to prepared food, from natural water to drinking water, and from paint,
solder and aerosol particles to dusts. At the end of this section, Table 7-18 summarizes the
four sources by natural and anthropogenic contributions, with the atmospheric contribution to
the anthropogenic fraction identified. Reference to this table will guide the discussion of
human exposure in a logical sequence that ultimately presents an estimate of the exposure of
the human population to atmospheric lead. To construct this table, it was necessary to make
decisions based on sound scientific judgment, extracted from the best available data. This
method provides a working approach to identifying sources of lead that can be easily modified
as more accurate data become available.
7.3.1.1 Lead in Inhaled Air. A principal determinant of atmospheric lead concentration is
distance from the source. At more than 100 m from a major highway or more than 2 km from a
stationary source, lead concentrations generally drop to constant levels (see Section 6.3),
and the particle size distribution shifts from a bimodal distribution to a unimodal one with
an MMAD of about 0.2 urn. Because the concentration of atmospheric lead at nonurban stations
is generally 0.05-0.15 pg/m3, a value of 0.1 |jg/m3 may reasonably be assumed. A correction
can be made for the indoor/outdoor ratio assuming the average individual spends 20-22 hours/
day in an unfiltered inside atmosphere and the average indoor/outdoor ratio for a nonurban
location is 0.5 (Table 7-6). The adjusted air concentration becomes 0.05 ug/m3 for baseline
purposes.
The concentration of natural lead in the atmosphere, discussed in Section 7.2.1.1.3, is
probably about 0.00005 ug/m3. This is an insignificant amount compared to the anthropogenic
contribution of 0.2 (jg/m3. A summary of lead in inhaled air appears in Table 7-11.
TABLE 7-11. SUMMARY OF INHALED AIR LEAD EXPOSURE
Population
Children (2 year-old)
Adult, working inside
Adult, working outside
Adjusted
air Pb
cone.*
(|jg/m3)
0.05
0.05
0.10
Amount
inhaled
(mVday)
10
20
20
Total
lead
exposure
(ug/day)
0.5
1.0
2.0
Natural
Pb
(ug/day)
0.001
0.002
0.004
Direct
atmospheric
Pb
(ug/day)
0.5
1.0
2.0
*Values adjusted for indoor/outdoor ratio of lead concentrations and for daily time spent
outdoors.
Source: Text
7-43
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7.3.1.2 Lead in Food. The route by which many people receive the largest portion of their
daily lead intake is through foods. Several studies have reported average dietary lead in-
takes in the range 100 to 500 ug/day for adults, with individual diets covering a much greater
range (Schroeder and Tipton, 1968; Mahaffey, 1978; Nutrition Foundation, Inc. 1982). Gross
(1981) analyzed results of the extensive lead mass balance experiments described by Kehoe
(1961), which were conducted from 1937 to 1972. According to these data, total dietary lead
intake decreased from approximately 300 jjg/day in 1937 to 100 ug/day in 1970, although there
is considerable variability in the data. Only a fraction of this lead is absorbed, as dis-
cussed in Chapter 10.
The amount of lead typically found in plants and animals is discussed in Section 7.2.2.2.
The sources of this lead are air, soil, and untreated waters (Figure 7-1). Food crops and
livestock contain lead in varying proportions from the atmosphere and natural sources. From
the farm to the dinner table, lead is added to food as it is harvested, transported, pro-
cessed, packaged, and prepared. The sources of this lead are dusts of atmospheric and indus-
trial origin, metals used in grinding, crushing, and sieving, solder used in packaging, and
water used in cooking.
The American diet is extremely complex and variable among individuals. Pennington (1983)
has described the basic diets, suppressing individual variation but identifying 234 typical
food categories, for Americans grouped into eight age/sex groups. These basic diets are the
foundation for the Food and Drug Administration's revised Total Diet Study, often called the
"market basket" study, beginning in April, 1982. The diets used for this document include
food, beverages, and drinking water for 2-year-old children, teen-age males and females, adult
males and females (25-30 years of age), and adult males and females (60 - 65 years of age).
The 201 typical food categories that constitute the basic diets are an aggregation of 3500
categories of food actually consumed by participants in the two surveys that formed the basis
of the Pennington study. Lead concentration data are given for each of these 201 food cate-
gories in Table 7D-1 of Appendix 7D and are from a preliminary report of the 1982 and 1983
Total Diet Study provided by the U.S. Food and Drug Administration for the purpose of this
document.
In 1982, the Nutrition Foundation published an exhaustive study of lead in foods, using
some data from the National Food Processors Assocation and some data from Canadian studies by
Kirkpatrick et al. (1980) and Kirkpatrick and Coffin (1974, 1977). A summary of the available
data for 1973-1980 was prepared in an internal report to the FDA prepared by Beloian and
McDowell (1981). Portions of these reports were used to interpret the contributions of lead
to food during processing.
The following section evaluates the amounts of lead added during each step of the process
from the field to the dinner table. In the best case, reliable data exist for the specific
7-44
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situation in question and conclusions are drawn. In some cases, comparable data can be used
with a few reasonable assumptions to formulate acceptable estimates of lead contributions.
For a portion of the diet, there are no acceptable data and the contributions of lead must,
for the time, be listed as of undetermined origin.
7.3.1.2.1 Lead added during handling and transportation to processor. Between the field and
the food processor, lead is added to the food crops. It is assumed that this lead is all of
direct atmospheric origin. Direct atmospheric lead can be lead deposited directly on food
materials by dry deposition, or it can be lead on dust that has collected on other surfaces,
then transferred to foods. For the purposes of this discussion, it is not necessary to distin-
guish between these two forms, as both are a function of air lead concentration.
There are no clear data on how much lead is added during transportation, but some obser-
vations are worth noting. First, some fresh vegetables (e.g., potatoes, lettuce, carrots,
onions) undergo no further processing other than trimming, washing and packaging. If washed,
water without soap is used; no additives or preservatives are used. An estimate of the amount
of atmospheric lead added during handling and transportation of all food crops can be made
from the observed increases in lead on those fresh vegetables where handling and transporta-
tion would be the only source of added lead. Because atmospheric lead deposition is a func-
tion of time, air concentration, and exposed surface area, there is an upper limit to the
maximum amount of direct atmospheric lead that can be added, except by the accumulation of
atmospheric dusts,
7.3.1.2.2 Lead added during preparation for packaging. For some of the canned food items,
data are available on lead concentrations just prior to the filling of cans. In the case
where the food product has not undergone extensive modification (e.g., cooking, added ingre-
dients), the added lead was most likely derived from the atmosphere or from the machinery used
to handle the product. As with transportation, the addition of atmospheric lead is limited to
reasonable amounts that can be added during exposure to air, and reasonable amounts of atmos-
pheric dust accumulation on food processing surfaces. One process that may increase the expo-
sure of the food to air is the use of air in separating food items, as in wheat grains from
chaff.
Where modification of the food product has occurred, the most common ingredients added
are sugar, salt, and water. It is reasonable that water has a lead concentration similar to
drinking water reported in Section 7.3.1.3 (0.008 ug/g) and that sugar (Boyer and Johnson,
1982) and salt have lead concentrations of 0.01 pg/g. Grinding, crushing, chopping, and cook-
ing may add lead from the metallic parts of machinery and from industrial greases. A summary
of the data (Table 7-12) indicates that about 30 percent of the total lead in canned goods is
the result of prepacking processes.
7-45
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TABLE 7-12. ADDITION OF LEAD TO FOOD PRODUCTS*
(|jg/g fresh weight)
Food
In the
field
(A)
After
preparation
for packaging
(B)
After
packaging
(C)
After
kitchen
preparation
(D)
Total lead
added
after harvest
(E)
Soft Packaged
Wheat 0.037
Field corn 0.022
Potatoes 0.009
Lettuce 0.013
Rice 0.007
Carrots 0.009
Beef 0.01
Pork 0.06
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
0.065
0.14
0.018
0.07
0.10
0.05
0.07
0.10
0.025
0.02
0.015
0.084
0.017
0.035
0.06
0.003
0.011
0.002
0.077
0.008
0.025
Metal cans
Sweet corn
Tomatoes
Spinach
Peas
Applesauce
Apricots
Mixed fruit
Plums
Green beans
0.003
0.002
0.045
N/A
N/A
N/A
N/A
N/A
N/A
0.04
0.06
0.43
0.08
0.08
0.07
0.08
0.09
0.16
0.27
0.29
0.68
0.19
0.24
0.17
0.24
0.16
0.32
0.28
—
0.86
0.22
0.17
--
0.20
—
0.16
0.28
—
0.82
0.14
0.09
0.10
0.12
0.07
~ ™
'"This table summarizes the stepwise addition of lead to food products at several stages
between the field and the dinner table. Data in column A are from Wolnik et al. (1983,
1985), columns B and C from National Food Processors Association (1982), and column D from
U.S. FDA (1985). Column E is calculated as column D - column A. Where data are not
available in column A, the values in column B were used. For the most part, column C
values closely approximate column D values, even though they are from separate studies,
suggesting most of the lead in food production is added prior to kitchen preparation.
N/A: data not available.
Occasionally, the processing or preparation of food may separate lead into a single pro-
duct or byproduct. Hayashi et al. (1982) found that lead in milk is isolated during process-
ing from butter and ends up in the buttermilk. Thus lead in butter is typically lower than
and buttermilk higher than normal sweet milk. Smart et al. (1981) have found that foods
cooked in water adsorb the lead in that water. Consequently, when pasta or similar items are
cooked then drained, the lead content of the prepared food is the sum of the dry food and the
7-46
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water. Conversely, when only the water is retained, as with tea bags, the final beverage may
have less lead than the original water.
7.3.1.2.3 Lead added during packaging. From the time a product is packaged in bottles, cans
or plastic containers, until it is opened in the kitchen, it may be assumed it receives
atmospheric lead. Most of the lead that is added during this stage comes from the solder used
to seal some types of cans. Estimates by the U.S. FDA, prepared in cooperation with the
National Food Processors Association, suggest that lead in solder contributes more than 66
percent of the lead in canned foods where a lead solder side seam is used. This lead was
thought to represent a contribution of 20 percent to the total lead consumption in foods
(F.R., 1979 August 31).
The full extent of the contribution of the canning process to overall lead levels in
albacore tuna was reported in a benchmark study by Settle and Patterson (1980). Using rigor-
ous clean laboratory procedures, these investigators analysed lead in fresh tuna, as well as
in tuna packaged in soldered and unsoldered cans. The data, presented in Table 7-13, show
that lead concentrations in canned tuna are elevated above levels in fresh tuna by a factor of
4,000, and by a factor of 40,000 above natural levels of lead in tuna. Nearly all of the in-
crease results from leaching of the lead from the soldered seam of the can; tuna from an un-
soldered can is elevated by a factor of only 20 compared with tuna fresh from the sea. Note
that when fresh tuna is dried and pulverized, as in the National Bureau of Standards reference
material, lead levels are seen to increase by a factor of 400 over fresh sea tuna. Table 7-13
also shows the results of analyses conducted by the National Marine Fisheries Service.
TABLE 7-13. PREHISTORIC AND MODERN CONCENTRATIONS IN HUMAN FOOD FROM A MARINE FOOD CHAIN
(ng/g fresh weight)
Source
Surface seawater
Albacore muscle, fresh
Albacore muscle from die-punched unsoldered can
Albacore muscle, lead-soldered can
Anchovy from albacore stomach
Anchovy from lead-soldered can
Estimated
prehistoric
0.0005
0.03
--
--
2.1
--
Modern
0.005
0.3
7.0
1400
21
4200
Source: Settle and Patterson (1980).
7-47
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7.3.1.2.4 Lead added during kitchen usage and storage. Although there have been several
studies of the lead concentrations in food after typical meal preparation, most of the data
are not amenable to this analysis because there are no data on lead concentrations before meal
preparation. As a part of its compliance program, the U.S. FDA has conducted the Total Diet
Study of lead and other trace contaminants in kitchen-prepared food each year since 1973.
Because the kitchen-prepared items were composited by category, there is no direct link be-
tween a specific food crop and the dinner table. Since April, 1982, this survey has analyzed
each food item individually (Pennington, 1983).
Other studies that reflect contributions of lead added during kitchen preparation have
been conducted. Capar (1978) showed that lead in acidic foods that are stored refrigerated in
open cans can increase by a factor of 2-8 in five days if the cans have a lead-soldered side
seam not protected by an interior lacquer coating. Comparable products in cans with the
lacquer coating or in glass jars showed little or no increase.
7.3.1.2.5 Recent changes in lead in food. As a part of its program to reduce the total lead
intake by children (0-5 years of age) to less than 100 (jg/day by 1988, the U.S. FDA estimated
lead intakes for individual children in a large-scale food consumption survey (Beloian and
McDowell, 1981). To convert the survey of total food intakes into lead intake, 23 separate
government and industry studies, covering 1973-78, were statistically analyzed. In spite of
the variability that can occur among individuals grouped by age, the authors estimated a base-
line (1973-78) daily lead intake .of 15 ug/day for infants aged 0-5 months, 59 ug/day for
children 6-23 months, and 82 pg/day for children 2-5 years. Between 1973 and 1978, intensive
efforts were made by the food industry to remove sources of lead from infant food items. By
1980, there had been a 47 percent reduction in the lead concentration for food consumed by the
age group 0-5 months and a 7 percent reduction for the 6-23 month age group (Table 7-14).
Most of this reduction was accomplished by the discontinuation of soldered cans used for in-
fant formula.
The 47 percent reduction in dietary lead achieved for infants prior to 1980 came about
largely because there are relatively few manufacturers of foods for infants and it was compar-
atively simple for this industry to mount a coordinated program in cooperation with the FDA.
There has not yet been a similar decrease in adult foods (Table 7-14) because only a few manu-
facturers have switched to lead-free cans. As the switchover increases, lead in canned food
should decrease to a level as low as 30 percent of the pre-1978 values, and there should be a
corresponding decrease of lead in the total adult diet, perhaps' as much as 20 percent. The
use of lead-soldered cans in the canning industry has decreased from 90 percent in 1979 to 63
percent in 1982. Within the next few years, the two leading can manufacturers expect to pro-
duce no more lead-soldered cans for the food industry. A two-year time lag is expected before
7-48
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TABLE 7-14. RECENT TRENDS OF MEAN LEAD CONCENTRATIONS
IN CANNED ADULT AND INFANT FOOD ITEMS
(ug/g)
Canned food*
Green beans
Beans w/pork
Peas
Tomatoes
Beets
Tomato juice
Applesauce
Citrus juice
Infant food
Formula concentrate
Juices
Pureed foods
Evaporated milk
Early 70' s
0.32
0.64
0.43
0.71
0.38
0.34
0.32
0.14
0.10
0.30
0.15
0.52
1976-77
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
0.055
0.045
0.05
0.10
1980-81
0.32
0.26
0.19
0.29
0.24
0.08
0.04
0.11
0.01
0.015
0.02
0.07
1982
0.16
0.17
0.22
0.21
0.12
0.067
0.17
0.04
N/A
N/A
N/A
N/A
*Boyer and Johnson (1982); 1982 data from U.S. Food and Drug Administration 1985 (see
Appendix 7D).
tData from early 70's and 1976-79 from Jelinek (1982); 1980-81 data from Schaffner (1981).
N/A = data not available.
the last of these cans disappears from the grocery shelf. Some of the 23 smaller manufac-
turers of cans have announced similar plans over a longer period of time. It is likely that
any expected decrease in the contribution of air lead to foods will be complemented by a de-
crease in lead from soldered cans.
7.3.1.2.6 Summary of lead in food. There are two major sources of lead in food and bev-
erages: atmospheric lead and lead from cans with lead soldered seams. The data of Wolnik et
al. (1983, 1985) provide some insight into the amount of atmospheric lead on food crops (Table
7-8). The FDA analyses of foods by category (Table 7D-1 in appendix 7D) clearly show the in-
fluence of solder on canned foods compared to fresh foods of the same type. The total food
consumption data of Pennington (1983) for 201 adult food categories were multiplied by the
mean lead concentrations from Table 7D-1 to determine the total daily exposure of seven age/
sex categories of Americans to lead in food and beverages.
For each food category, a separate source coefficient was assigned for direct atmos-
pheric, solder and metallic, indirect atmospheric, and natural lead. Any fraction of lead
that could not be otherwise assigned was considered lead of undetermined origin. In this
7-49
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manner, the lead content of 201 food categories was determined for five sources. To simplify
the presentation of this data, the 201 food categories have been combined into nine groups
based on the scheme of Table 7D-2 in appendix 7D-2 in appendix 7D. The nine categories were
specifically selected to emphasize the most probable source of lead. Therefore canned foods
were placed in one category to isolate metallic lead, and crop foods in another to isolate
atmospheric lead.
The total consumption for the seven age/sex categories and nine food categories is shown
in Table 7-15, adapted from Pennington (1983). The amount of lead that is consumed with the
food and beverages in Table 7-15 is shown on Table 7-16. This calculation is based solely on
the average lead concentrations of each food item of Table 7D-1, the data provided by FDA. To
determine the source of this lead, the individual source coefficients for each food item were
multiplied by the average lead concentration in Table 7D-1 and by the amount consumed
(Pennington, 1983) to get the amount of lead consumed from each source for each age/sex cate-
gory. An average was taken of each age/sex category then the 201 food items were condensed
into the nine food categories and presented on Table 7-17.
TABLE 7-15. TOTAL CONSUMPTION, BY AGE AND SEX, OF FOOD AND BEVERAGES
(g/day)
Major
food category
Dairy products
Meat products
Food crops
Canned food
Canned juices
Frozen juices
Soda
Canned beer
Watert
Totals
*7.5 g baby food
Child*
2 yrs
390
133
282
72
54
65
65
0
441
1502
and infant
Female
14-16
405
182
386
77
28
53
232
0
596
1959
formula were
Male
yrs
645
269
528
104
30
75
274
17
743
2685
not included
Female
25-30
245
194
390
73
28
66
228
51
903
2178
in this
Male
yrs
351
319
518
103
27
73
315
318
1061
3086
evaluation.
Female
60-65
208
172
437
99
17
72
78
18
1166
2267
Male
yrs
279
252
532
119
12
61
85
116
1244
2700
tlncludes coffee, tea, and powdered drinks.
Source: Data are summarized from Pennington (1983) according to Table 7D-2.
7-50
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TABLE 7-16. TOTAL CONSUMPTION, BY AGE AND SEX, OF LEAD IN MILK AND FOOD AND BEVERAGES
(ug/day)
Major
food category
Dairy products
Meat products
Food crops
Canned food
Canned juices
Frozen juices
Soda
Canned beer
Watert
Totals
Child*
2 yrs
2.8
3.4
5.5
7.3
2.7
0.5
0.7
0.0
2.1
25.0
Female
14-16
3.5
4.8
8.1
8.1
1.4
0.5
2.3
0.0
2.5
31.2
Male
yrs
5.4
7.4
11.7
11.8
1.5
0.7
3.0
0.1
3.2
44.8
Female
25-30
2.5
5.0
7.9
8.8
1.4
0.6
2.1
0.7
3.0
32.0
Male
yrs
3.4
7.4
11.3
12.0
1.4
0.7
2.9
2.5
3.6
45.2
Female
60-65
2.3
4.0
7.8
11.6
0.9
0.7
0.9
0.3
3.9
32.4
Male
yrs
3.1
5.4
9.6
14.4
0.6
0.5
0.9
1.0
4.2
39.7
tlncludes coffee, tea, and powdered drinks.
It is apparent that about 43 percent of lead in food and beverages milk and food can be
attributed to direct atmospheric deposition, compared to 42 percent from solder or other metal
sources. Of the remaining 5 percent for which the source is as yet undetermined, it is likely
that further research will show this lead to be part atmospheric in origin and part from
solder and other industrial metals.
This dietary lead consumption is used to calculate the total baseline human exposure in
Section 7.3.1.5 and is the largest baseline source of lead. Possible additions to dietary
lead consumption are discussed in Section 7.3.2.1.3 with respect to urban gardens.
Because the U.S. FDA is actively pursuing programs to decrease lead in adult foods, it is
probable that there will be a decrease in total dietary lead consumption over the next decade
independent of projected decreases in atmospheric lead concentration. With both sources of
lead minimized, the lowest reasonable estimated dietary lead consumption would be 10-15 ug/day
for adults and children. This estimate is based on the assumption that about 90 percent of
the direct atmospheric lead, solder lead, and lead of undetermined origin would be removed
from the diet, leaving 8 ug/day from these sources and 3 ug/day of natural and indirect atmos-
pheric lead.
7-51
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TABLE 7-17. SUMMARY BY SOURCE OF LEAD CONSUMED FROM FOOD AND BEVERAGES
((jg/day)
Major
food category
Dairy
Meat
Food crops
Canned foods
Canned juices
Frozen Juices
Soda
Canned Beer
Water
Total
Percent
Total
lead*
3.3
5.3
8.8
10.6
1.4
0.6
1.8
0.7
3.2
35.7
Natural
lead
0.030
0.040
0.880
0.120
0.001
0.001
0.005
0.001
0.010
1.088
3.0%
Atmospheric lead
indirect
0.030
0.040
0.880
0.120
0.062
0.110
0.280
0.140
0.850
2.512
7.0%
direct
2.74
4.11
6.60
0.92
0.04
0.07
0.21
0.05
0.54
15.28
42.8%
Lead from
solder and
other metals
0.00
0.41
0.00
9.40
1.30
0.42
1.30
0.51
1.80
15.14
42.4%
Lead of
undetermined
origin
0.50
0.70
0.44
0.04
0.00
0.00
0.00
0.00
0.00
1.68
4.7%
*Based on average
atmospheric lead
lead consumption by 7 age/sex groups. There may be some direct
and solder lead in the category of undetermined origin.
7.3.1.3 Lead in Drinking Water. The U.S. Public Health Service standards specify that lead
levels in drinking water should not exceed 50 yg/1. The presence of detectable amounts of
lead in untreated public water supplies was shown by Durum et al. (1971) to be widespread, but
only a few samples contained amounts above the 50 ug/1 standard.
The major source of lead contamination in drinking water is the water distribution
system. Water that is corrosive can leach considerable amounts of lead from lead plumbing and
lead compounds used to join pipes. Moore (1977) demonstrated the effect of water standing in
pipes overnight. Lead concentrations dropped significantly with flushing at 10 1/min for five
minutes (Figure 7-8). Lead pipe currently is in use in some parts of New England for water
service lines and interior plumbing, particularly in older urban areas. The contributions of
lead plumbing to potential human exposure are considered additive rather than baseline and are
discussed in Section 7.3.2.1.4.
7-52
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10
TIME OF FLUSHING, minutes
Figure 7-8. Change in drinking water lead concentration in a house
with lead plumbing for the first use of water in the morning.
Flushing rate was 10 liters/minute.
Source: Moore (1977).
7-53
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There have been several studies in North America and Europe of the sources of lead in
drinking water. A recent study in Seattle, WA by Sharrett et al. (1982) showed that the age
of the house and the type of plumbing determined the lead concentration in tap water. Stand-
ing water in copper pipes from houses newer than five years averaged 31 ug/1; those less than
18 months average about 70 MS/I- Houses older than five years and houses with galvanized pipe
averaged less than 6 ug/1. The source of the water supply, the length of the pipe, and the
use of plastic pipes in the service line had little or no effect on the lead concentrations.
It appears certain that the source of lead in new homes with copper pipes is the solder used
to join these pipes, and that this lead is either leached away with age or isolated by accumu-
lated deposits within the pipes. A study of copper pipes in cottages using local lakes for a
water source revealed a similar pattern of lead with increased standing time in the pipes
(Meranger et al., 1983). The lead concentration in the first liter drawn continued to in-
crease with standing time, even up to ten days.
The Sharrett et al. (1982) study of the Seattle population also provided data on water
and beverage consumption which extended the scope of the Pennington (1983) study of all Ameri-
cans. While the total amount of liquids consumed was slightly higher in Seattle (2200 g/day
vs. 1800 g/day for all Americans), the breakdown between water consumed inside and outside the
home can prove useful. Men, women, and children consume 53, 87, and 87 percent, respectively,
of their water and beverages within the home.
Bailey and Russell (1981) have developed a model for population exposure to lead in home
drinking water. The model incorporates data for lead concentration as a function of stagna-
tion time in the pipes, as well as probability distributions for times of water use throughout
the day. Population surveys conducted as part of the United Kingdom Regional Heart Survey
provided these water-use distributions.
Other studies have been conducted in Canada and Belgium. Lead levels in water boiled in
electric kettles were measured in 574 households in Ottawa (Wigle and Charlebois, 1978). Con-
centrations greater than 50 ug/1 were observed in 42.5 percent of the households, and exces-
sive lead levels were associated with kettles more than five years old.
7.3.1.4 Lead in Dusts. By technical definition, dusts are solid particles produced by the
disintegration of materials (Friedlander, 1977) and appear to have no size limitations. Al-
though dusts are of complex origin, they may be placed conveniently into a few categories re-
lating to human exposure. Generally, the most convenient categories are household dusts, soil
dust, street dusts, and occupational dusts. In each case, the lead in dust arises from a com-
plex mixture of fine particles of soil, flaked paint, and airborne particles of industrial or
automotive origin. It is a characteristic of dust particles that they accumulate on exposed
surfaces and are trapped in the fibers of clothing and carpets. Ingestion of dust particles,
7-54
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rather than inhalation, appears to be the greater problem in the baseline environment, espe-
cially ingestion during meals and playtime activity by small children.
Two other features of dust are important. First, they must be described in both concen-
tration and amount. The concentration of lead in street dust may be the same in a rural and
urban environment, but the amount of dust may differ by a wide margin. Secondly, each cate-
gory represents a different combination of sources. Household dusts contain some atmospheric
lead, some paint lead, and some soil lead. Street dusts contain atmospheric, soil, and occa-
sionally paint lead. This apparent paradox does not prevent the evaluation of exposures to
dust, but it does confound efforts to identify the amounts of atmospheric lead contributed to
dusts. For the baseline human exposure, it is assumed that workers are not exposed to occupa-
tional dusts, nor do they live in houses with interior leaded paints. Street dust, soil dust,
and some household dust are the primary dust sources for baseline potential human lead expo-
sure.
In considering the impact of street dust on the human environment, the obvious question
arises as to whether lead in street dust varies with traffic density. In a transect through
Minneapolis/St. Paul, Mielke etal. (1984) found soil lead concentrations 10 to 1000-fold
higher near major interstate highways. Nriagu (1978) reviewed several studies of lead in
street dust. Warren et al. (1971) reported 20,000 ug Pb/g street dust in a heavily trafficked
area. In the review by Nriagu (1978), street dust lead concentrations ranged from 300 to
18,000 ug/g in several cities in the United States. More recent studies have attempted to
characterize lead in street dust in greater detail. Franz and Hadley (1981) separated street
dust by particle size and found that smaller particles contain greater concentrations of lead.
One-third of the mass was less than 150 urn and contained 37 percent of the total lead. The
average concentration in the Albuquerque street dust was 5000 |jg/g, 20 percent of which was
attributed to curb paint. Dong et al. (1984) separated street dusts by mechanical sieve and
found, with one exception, 50 percent or more of the lead on clay-sized particles, the small-
est fraction both in particle size and in total mass (5-6 percent). There was, however, con-
siderable variation in the absolute concentrations of lead in the samples from the same loca-
tion taken four days apart.
There are several reports of street dusts outside the United States that show similar
relationships. Fergusson and Ryan (1984) found concentrations in small urban cities in
Canada, New Zealand and Jamaica ranged from 700 to 2000 |jg/g, while in New York and London the
range was from 2000 to 4000 ug/g. Sequential extractions showed much of the lead (44 percent)
was on the Fe-Mn oxide fraction, but that 36 percent was on the exchangeable and carbonate
fractions that are more readily available. Gibson and Farmer (1984) also found 41 percent of
the street dust lead in Glascow, Scotland to be on the exchangeable and carbonate fractions.
7-55
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Duggan (1984) attempted to relate London street dust lead to airborne lead concentrations
and found that airborne variations with time were greater than for dust, but spatial varia-
tions were greater for dust. The results suggested that dust may be an adequate measure of
long-term (three month) ambient concentrations, but that several samples over a wide area must
be taken. In a related study, Thornton et al. (1985) recommended the adoption of guidelines
for urban dust lead concentrations to the Greater London Council. The recommendation was that
lead concentrations of 500 ug/g in the fraction smaller than 0.5 mm justified further investi-
gation, whereas concentrations above 5000 (jg/g justified control measures. Duggan et al.
(1985) reported that the amount of lead on children's hands was proportional to the concentra-
tion of lead in playground dust. This relationship was nearly linear up to 4000 ug Pb/g dust.
In Hong Kong, lead in street dust ranged from 960 to 7400 |jg/g with no direct relation-
ship to traffic volume (Ho, 1979). In other reports from Hong Kong, Lau and Wong (1982) found
values from 130 ug/g at 20 vehicles/day to 3900 ug/g at 37,000 vehicles/day. Fourteen sites
in this study showed close correlation with traffic density.
In the United Kingdom, lead in urban and rural street dusts was determined to be 970 and
85 ug/g, respectively, by Day et al. (1975). A later report by this group (Day et al., 1979)
discusses the persistency of lead dusts in rainwashed areas of the United Kingdom and New
Zealand and the potential health hazard due to ingestion by children. They concluded that,
whereas the acidity of rain was insufficient to dissolve and transport lead particles, the
potential health hazard lies with the ingestion of these particles during the normal play
activities of children residing near these areas. A child playing at a playground near a
roadside might consume 20-200 |jg lead while eating a single piece of candy with unwashed
hands. It appears that in nonurban environments, lead in street dust ranges from 80 to 130
|jg/g, whereas urban street dusts range from 1,000 to 20,000 ug/g. For the purpose of esti-
mating potential human exposure, an average lead value of 90 ug/g in street dust is assumed
for baseline exposure on Table 7-18, and 1500 ug/g in the discussions of urban environments in
Section 7.3.2.1.
Dust is also a normal component of the home environment. It accumulates on all exposed
surfaces, especially furniture, rugs and windowsills. For reasons of hygiene and respiratory
health, many homemakers take great care to remove this dust from the household. Because there
are at least two circumstances where these measures are inadequate, it is important to con-
sider the possible concentration of lead in these dusts in order to determine potential expo-
sure to young children. First, some households do not practice regular dust removal, and
secondly, in some households of workers exposed occupationally to lead dusts, the worker may
carry dust home in amounts too small for efficient removal but containing lead concentrations
much higher than normal baseline values.
7-56
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TABLE 7-18. CURRENT BASELINE ESTIMATES OF POTENTIAL HUMAN EXPOSURE TO DUSTS
Child
Household dusts
Street dust
Occupational dust
Total
Percent
Adult
Household dusts
Street dust
Occupational dust
Total
Percent
Dust
lead
cone.
300
90
150
300
90
150
Dust
ingested
(9/day)
0.05
0.04
0.01
0.10
0.01
0.0
0.01
0.02
Dust
lead
consumed
(ug/day)
15
4.5
1.5
21.0
100%
3
0
1.5
4.5
100%
Natural
(ug/day)
0.5
0.0
0.1
0.6
2.8%
0.1
0.0
0.1
0.2
4.5%
Source of
Atmos .
(ug/day)
14.5
4.5
0.0
19.0
90.5%
2.9
0.0
0.0
2.9
64.4%
lead
Undetermined
(ug/day)
0.0
0.0
1.4
1.4
6.7%
0.0
0.0
1.4
1.4
31.1%
In Omaha, Nebraska, Angle and Mclntire (1979) found that lead in household dust ranged
from 18 to 5600 pg/g. Clark et al. (1985) found household dusts in Cincinnati ranged from 70
to 16000 ug/g, but that much of the variations could be attributed to housing quality. Public
housing averaged 350 ug/g, rehabilitated 600 (jg/9> and averages in private housing ranged from
1400 to 3000 based on external estimates of condition from satisfactory to deteriorating to
dilapidated. In Lancaster, England, a region of low industrial lead emissions, Harrison
(1979) found that household dust ranged from 510 to 970 ug/g, with a mean of 720 ug/g. They
observed that dust contained soil particles (10-200 urn in diameter), carpet and clothing
fibers, animal and human hairs, food particles, and an occasional chip of paint. The previous
Lead Criteria Document (U.S. Environmental Protection Agency, 1977) summarized earlier reports
of lead in household dust showing residential suburban areas ranging from 280 to 1,500 ug/g,
urban residential from 600 to 2,000 ug/g, and urban industrial from 900 to 16,000 ug/g.
Brunekreef (1983) summarized studies of simultaneous measurements of air lead, soil lead, and
household dust lead. With some exceptions, the household dust lead concentrations ranged from
400 to 700 ug/g per 1 ug/m3 of lead in air. The relationship between household dust and soil
dust was much broader. Because of the diverse nature of the studies, care should be taken in
extrapolating these observations to more general circumstances. In El Paso, Texas, lead in
household dust ranged from 2,800 to 100,000 ug/g within 2 km of a smelter (Landrigan et al.
7-57
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1975). Davies et al. (1985) found a correlation between soil dust and household dust in an
old lead mining area of North Wales, Great Britain, where a tenfold increase in soil lead was
associated with a twofold increase in household dust lead.
It appears that most of the values for lead in dust in nonurban household environments
fall in the range of 50-500 |jg/g. A mean value of 300 pg/g is assumed. The only natural lead
in dust would be some fraction of that derived from soil lead. A value of 10 (jg/g seems
reasonable, since some of the soil lead is of atmospheric origin. Since very little paint
lead is included in the baseline estimate, most of the remaining dust lead would be from the
atmosphere. Table 7-18 summarizes these estimates of human exposure to dusts for children and
adults. It assumes that children ingest about five times as much dust as adults, most of the
excess being street dusts from sidewalks and playgrounds. Exposure of children to occupa-
tional lead would be through contaminated clothing brought home by parents. Most of this lead
is of undetermined origin because no data exist on whether the source is dust similar to
household dust or unusual dust from the grinding and milling activities of factories.
7.3.1.5 Summary of Baseline Human Exposure to Lead. The values derived or assumed in the
proceeding sections are summarized in Table 7-19. These values represent only consumption,
not absorption, of lead by the human body. The key question of what are the risks to human
health from these baseline exposures is addressed in Chapter 13. The approach used here to
evaluate potential human exposure is similar to that used by the National Academy of Sciences
(1980) and the Nutrition Foundation (1982) in their assessments of the impact of lead in the
human environment.
7.3.2 Additive Exposure Factors
There are many conditions, even in nonurban environments, where an individual may in-
crease his lead exposure by choice, habit, or unavoidable circumstance. The following sec-
tions describe these conditions as separate exposures to be added as appropriate to the base-
line of human exposure described above. Most of these additive exposures clearly derive from
air or dust, while a few derive from water or food.
7.3.2.1 Living and Working Environments With Increased Lead Exposure. Ambient air lead con-
centrations are typically higher in an urban than a rural environment. This factor alone can
contribute significantly to the potential lead exposure of Americans, through increases in
inhaled air and consumed dust. Produce from urban gardens may also increase the daily con-
sumption of lead. Some environmental exposures may not be related only to urban living, such
as houses with interior lead paint or lead plumbing, residences near smelters or refineries,
or family gardens grown on high-lead soils. Occupational exposures may also occur in an urban
or rural setting. These exposures, whether primarily in the occupational environment or
7-58
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TABLE 7-19. SUMMARY OF BASELINE HUMAN EXPOSURES TO LEAD
(pg/day)
I
CJ1
IO
Soil
Source
Child-2 yr old
Inhaled air
Food, Water &
beverages
Dust
Total
Percent
Adult female
Inhaled air
Food, Water &
beverages
Dust
Total
Percent
Adult male
Inhaled air
Food, Water &
beverages
Dust
Total
Percent
Total
lead
consumed
0.5
25.1
21.0
46.6
100%
1.0
32.0
4.5
37.5
100%
1.0
45.2
4.5
50.7
100%
Natural
lead
consumed
0.001
0.71
0.6
1.3
2.8%
0.002
0.91
0.2
1.2
3.1%
0.002
1.42
0.2
1.6
3.1*
Indirect
atmospheric
lead*
-
1.7
-
1.7
3.5%
-
2.4
-
2.5
6.6%
-
3.5
-
3.5
6.8%
Direct
atmospheric
lead*
0.5
10.3
19.0
29.8
64.0%
1.0
12.6
2.9
17.4
46.5%
1.0
19.3
2.9
23.2
45.8%
Lead from
solder or
other metals
-
11.2
-
11.2
24.0%
-
8.2
-
13.5
36.136
-
18.9
-
18.9
37.2%
Lead of
undetermi ned
origin
-
1.2
1.4
2.6
5.6%
-
1.5
1.4
2.9
7.8%
-
2.2
1.4
3.6
7.0%
*Indirect atmospheric lead has been previously incorporated into soil, and will probably remain in the
soil for decades or longer. Direct atmospheric lead has been deposited on the surfaces of vegetation
and living areas or incorporated during food processing prior to human consumption.
Source: This report.
-------
secondarily in the home of the worker, would be additive with other exposures in an urban
location or with special cases of lead-based paint or plumbing.
7.3.2.1.1 Urban atmospheres. Urban atmospheres have more airborne lead than do nonurban
atmospheres, therefore there are increased amounts of lead in urban household and street dust.
Typical urban atmospheres contain 0.5-1.0 ug Pb/m3. Other variables are the amount of indoor
filtered air breathed by urban residents, the amount of time spent indoors, and the amount of
time spent on freeways. Reported means of urban dusts range from 500 to 3000 ug Pb/g. it is
not known whether there is more or less dust in urban households and playgrounds than in rural
environments. Whereas people may breathe the same amount of air, or eat and drink the same
amount of food and water, it is not certain that urban residents consume the same amount of
dust as nonurban. Nevertheless, in the absence of more reliable data, it has been assumed
that urban and nonurban residents consume the same amount of dusts.
The indoor/outdoor ratio of atmospheric lead for urban environments is about 0.8 (Table
7-6). Assuming 2 hours of exposure/day outdoors at a lead concentration of 0.75 \ig/m3, 20
hours indoors at 0.6 ug/m3, and 2 hours in a high traffic density area at 5 ug/m3, a weighted
mean air exposure of 1.0 ug/m3 appears to be typical of urban residents.
7.3.2.1.2 Houses with interior lead paint. In 1974, the Consumer Product Safety Commission
collected household paint samples and analyzed them for lead content (National Academy of
Sciences; National Research Council, 1976). Analysis of 489 samples showed that 8 percent of
the oil-based paints and 1 percent of the water-based paints contained greater than 0.5 per-
cent lead (5000 ug Pb/g paint, based on dried solids), which was the statutory limit at the
time of the study. The current statutory limit for Federal construction is 0.06 percent. The
greatest amounts of leaded paint are typically found in the kitchens, bathrooms, and bedrooms
(Tyler, 1970; Laurer et al., 1973; Gilbert et al., 1979).
Some investigators have shown that flaking paint can cause elevated lead concentrations
in nearby soil. For example, Hardy et al. (1971) measured soil lead levels of 2000 ug/g next
to a barn in rural Massachusetts. A steady decrease in lead level with increasing distance
from the barn was shown, reaching 60 ug/g at fifty feet from the barn. Ter Haar and Aronow
(1974) reported elevated soil lead levels in Detroit near eighteen old wood frame houses
painted with lead-based paint. The average soil lead level within two feet of a house was
just over 2000 ug/g; the average concentration at ten feet was slightly more than 400 ug/g.
The same authors reported smaller soil lead elevations in the vicinity of eighteen brick
veneer houses in Detroit. Soil lead levels near painted barns located in rural areas were
similar to urban soil lead concentrations near painted houses, suggesting the importance of
leaded paint at both urban and rural locations. The baseline lead concentration for household
dust of 300 ug/g was increased to 2000 ug/g for houses with interior lead-based paints. The
7-60
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additional 1700 pg/g would add 85 pg Pb/day to the potential exposure of a child (Table 7-20).
This increase would occur in an urban or nonurban environment and would be in addition to the
urban residential increase if the lead-based painted house were in an urban environment.
7.3.2.1.3 Family gardens. Several studies have shown potentially higher lead exposure
through the consumption of home-grown produce from family gardens grown on high lead soils or
near sources of atmospheric lead. Mielke et al. (1983) surveyed the lead content of urban
garden soils in Baltimore, finding concentrations ranging from 1 to 10,900, with a median of
100 pg/g. The soil sample was a mixture through 20-30 cm of the soil profile. The values
greater than 100 pg/g were concentrated near the center of the city. Kneip (1978) found ele-
vated levels of lead in leafy vegetables, root crops, and garden fruits associated qualita-
tively with traffic density and soil lead. Spittler and Feder (1978) reported a linear corre-
lation between soil lead (100-1650 pg/g) and lead in or on leafy or root vegetables. Freer et
al. (1980) found a threefold increase in lead concentrations of leafy vegetables (from 6 to 16
pg/g) in the soil lead range from 150 to 2200 pg/g. Chaney et al. (1984) have reviewed the
recent studies on lead in urban gardens. In none of these studies were the lowest soil lead
concentrations in the normal range of 10-25 pg/g, nor were any lead concentrations reported
for vegetables as low as those of Wolnik et al. (1983, 1985) (see Table 7-8).
In family gardens, lead may reach the edible portions of vegetables by deposition of at-
mospheric lead directly on aboveground plant parts or on soil, or by the flaking of lead-
containing paint chips from houses. Traffic density and distance from the road are not good
predictors of soil or vegetable lead concentrations (Freer et al., 1980). Air concentrations
and particle size distributions are the important determinants of deposition on soil or vege-
tation surfaces. Even at relatively high air concentrations (1.5 pg/m3) and deposition velo-
city (0.5 cm/sec) (see Section 6.4.1), it is unlikely that surface deposition alone can
account for more than 2-5 pg/g lead on the surface of lettuce during a 21-day growing period.
It appears that a significant fraction of the lead in both leafy and root vegetables derives
from the soil.
Using the same air concentration and deposition velocity values, a maximum of 1000 pg
lead has been added to each cm2 of the surface of the soil over the past 40 years. With cul-
tivation to a depth of 15 cm, it is not likely that atmospheric lead alone can account for
more than a few hundred pg/g of soil in urban gardens. Urban soils with lead concentrations
of 500 pg/g or more must certainly have another source of lead. In the absence of a nearby
(<5 km) stationary industrial source, paint chips seem the most likely explanation. Even if
the house no longer stands at the site, the lead from paint chips may still be present in the
soil.
Studies of family gardens do not agree on the concentrations of lead in produce. At the
higher soil concentrations, Kneip (1978) reported 0.2-1 pg/g in vegetables, Spittler and Feder
7-61
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TABLE 7-20. SUMMARY OF POTENTIAL ADDITIVE EXPOSURES TO LEAD
(yg/day)
Exposure
Total
lead
consumed
Atmospheric
lead
consumed
Other
lead
sources
Baseline exposure:
Child
Inhaled air
Food, water & beverages
Dust
Baseline exposure:
Adult male
Inhaled air
Food, water & beverages
Dust
Total baseline
0.5
25.1
21.0
1.0
54.7
4.5
60.2
0.5
10.3
19.0
1.0
20.3
2.9
24.2
14.8
2.0
Total baseline
Additional exposure due to:
Urban atmospheres1
Family gardens2
Interior lead paint3
Residence near smelter4
Secondary occupational5
46.6
91
48
110
880
150
29.8
91
12
880
16.8
36
110
34.4
1.6
36.0
Additional exposure due to:
Urban atmospheres1
Family gardens2
Interior lead paint3
Residence near smelter4
Occupational6
Secondary occupational5
Smoking7
Wine consumption8
28
120
17
100
1100
44
30
100
28
30
100
1100
27
17
Includes lead from household (1000 ug/g) and street dust (1500 ug/g) and inhaled air
(0.75 ug/m3).
2Assumes soil lead concentration of 2000 ug/g; all fresh leafy and root vegetables, and sweet
corn of Table 7-12 replaced by produce from garden. Also assumes 25% of soil lead is of
atmospheric origin.
3Assumes household dust rises from 300 to 2000 ug/g. Dust consumption remains the same
as baseline.
4Assumes household and street dust increase to 10,000 ug/g.
5Assumes household dust increases to 2400 ug/g.
6Assumes 8-hr shift at 10 ug Pb/m3 or 90% efficiency of respirators at 100 ug Pb/m3, and
occupational dusts at 100,000 ug/m3.
70ne and a half packs per day.
8Assumes unusually high consumption of one liter per day.
7-62
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(1978) reported 0.8-4.5 ug/g, and Freer et al. (1980) found 0.1-0.8 |jg/g (all values converted
to fresh weight). Since the Spittler and Feder (1978) and Freer et al. (1980) studies dealt
with soils of about 2000 ug/g, these data can be used to calculate a worst case exposure of
lead from family gardens. Assuming 0.8 ug/g for the leafy and root vegetables [compared to
0.01-0.05 ug/g of the Wolnik et al. (1983, 1985) study] family gardens could add 100 ug/day if
the 137 g of leafy and root vegetables, sweet corn and potatoes consumed by adult males
(Table 7D-1) were replaced by family garden products. Comparable values for children and
adult females would be 40 and 80 ug/day, respectively. No conclusive data are available for
vine vegetables, but the ranges of 0.8 to 0.1 ug/g for tomatoes suggest that the contamination
by lead from soil is much less for vine vegetables than for leafy or root vegetables. Chaney
et al. (1984) recommended that special precautions (extra washing and peeling) be taken with
produce from urban gardens with soil lead from 500 to 1000 ug/g. They also recommended that
leafy and root vegetables not be grown in gardens over 3000 ug/g.
7.3.2.1.4 Houses with lead plumbing. The Glasgow Duplicate Diet Study (United Kingdom De-
partment of the Environment, 1982) reports that children approximately 13 weeks old living in
houses with lead plumbing consume 6-480 ug Pb/day. Concentrations of lead in water ranged
from less than 50 to over 500 ug/1 for the 131 homes studied. Those children and mothers
living in the homes containing high water-lead concentrations generally had greater total lead
consumption and higher blood lead levels, according to the study. Breast-fed infants were
exposed to much less lead than bottle-fed infants. Because the project was designed to inves-
tigate child and maternal blood lead levels over a wide range of water lead concentrations,
the individuals studied do not represent a typical cross-section of the population. However,
results of the study suggest that infants living in homes with lead plumbing may have exposure
to considerable amounts of lead. This conclusion was also demonstrated by Sherlock et al.
(1982) in a duplicate diet study in Ayr, Scotland.
7.3.2.1.5 Residences near smelters and refineries. Air lead concentrations within 2 km of
lead smelters and refineries average 5-15 ug/m3. Assuming the same indoor/outdoor ratio of
atmospheric lead for nonurban residents (0.5), residents near smelters would be exposed to in-
haled air lead concentrations of about 6 ug/m3, compared to 0.05 [ig/m3 for the background
levels. Household dust concentrations at El Paso, TX range from 3000 to 100,000 ug/g in 1982
(Landrigan et al., 1975). Morse et al. (1979) found that, with this installation of engi-
neering improvements and pollution control, the dust lead was reduced to 1500-2000 by 1977. A
value of 10,000 ug/g is assumed for household dust near a smelter. Between inhaled air and
dust, a child in this circumstance would be exposed to 900 ug Pb/day above background levels.
Exposures for adults would be much less, since they consume only 20 percent of the dusts
children consume.
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7.3.2.1.6 Occupational exposures. The highest and most prolonged exposures to lead are found
among workers in the lead smelting, refining, and manufacturing industries (World Health
Organization, 1977). In all work areas, the major route of lead exposure is by inhalation and
ingestion of lead-bearing dusts and fumes. Airborne dusts settle out of the air onto food,
water, the workers' clothing, and other objects, and may be transferred subsequently to the
mouth. Therefore, good housekeeping and good ventilation have a major impact on exposure. It
has been found that concentrations might be quite high in one factory and low in another
solely because of differences in ventilation, or differences in custodial practices and worker
education. The estimate of additional exposure in Table 7-20 is for an 8-hour shift at 100 \ig
Pb/m3. Occupational exposure under these conditions is primarily determined by occupational
dust consumed. Even tiny amounts (e.g., 10 mg) of dust containing 100,000 ug Pb/g dust can
account for 1,000 |jg/day exposure.
7.3.2.1.6.1 Lead mining, smelting, and refining. Roy (1977) studied exposures during
mining and grinding of lead sulfide at a mill in the Missouri lead belt. Primary smelting
operations were 4 km from the mill, hence the influence of the smelter was believed to be
negligible. The total airborne lead levels were much greater than the concentrations of
respirable lead, indicating a predominance of coarse material.
The greatest potential for high-level exposure exists in the process of lead smelting and
refining (World Health Organization, 1977). The most hazardous operations are those in which
molten lead and lead alloys are brought to high temperatures, resulting in the vaporization of
lead. This is because condensed lead vapor or fume has, to a substantial degree, a small
(respirable) particle size range. Although the total air lead concentration may be greater in
the vicinity of ore-proportioning bins than it is in the vicinity of a blast furnace in a
smelter, the amount of particle mass in the respirable size range may be much greater near the
furnace.
A measure of the potential lead exposure in smelters was obtained in a study of three
typical installations in Utah (World Health Organization, 1977). Air lead concentrations near
all major operations, as determined using personal monitors worn by workers, were found to
vary from about 100 to more than 4000 ug/m3. Obviously, the hazard to these workers would be
extremely serious if it were not for the fact that the use of respirators is mandatory in
these particular smelters. Maximum airborne lead concentrations of about 300 ug/m3 were mea-
sured in a primary lead-zinc smelter in the United Kingdom (King et al., 1979). These authors
found poor correlations between airborne lead and blood lead in the smelter workers, and con-
cluded that a program designed to protect these workers should focus on monitoring of biologi-
cal parameters rather than environmental concentrations.
Spivey et al. (1979) studied a secondary smelter in southern California that recovers
lead mainly from automotive storage batteries. Airborne lead concentrations of 10-4800 ug/m3
7-64
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were measured. The project also involved measurement of biological parameters as well as a
survey of symptoms commonly associated with lead exposure; a poor correlation was found
between indices of lead absorption and symptom reporting. The authors suggested that such
factors as educational level, knowledge of possible symptoms, and biological susceptibility
may be important factors in influencing symptom reporting. In a second article covering this
same study, Brown et al. (1980) reported that smokers working at a smelter had greater blood
lead concentrations than nonsmokers. Furthermore, smokers who brought their cigarettes into
the workplace had greater blood lead concentrations than those who left their cigarettes else-
where. It was concluded that direct environmental contamination of the cigarettes by lead-
containing dust may be a major exposure pathway for these individuals (See Section 7.3.2.3.1).
Secondary lead smelters in Memphis, Tennessee and Salt Lake City, Utah were studied by
Baker et al. (1979). The Memphis plant extracted lead principally from automotive batteries,
producing 11,500 metric tons of lead in the eleven months preceding the measurements. The
Salt Lake City plant used scrap to recover 258 metric tons of lead in the six months preceding
the measurements. Airborne concentrations of lead in the Tennessee study exceeded 200 \ig/m3
in some instances, with personal air sampler data ranging from 120 ug/m3 for a battery wrecker
to 350 ug/m3 for two yard workers. At the Utah plant, airborne lead levels in the office,
lunchroom, and furnace room (furnace not operating) were 60, 90, and 100 ug/m3, respectively.
When charging the furnace, this value increased to 2650 ug/m3. Personal samplers yielded con-
centrations of 17 ug/m3 for an office worker, 700 ug/m3 for two welders, and 2660 ug/m3 for
two furnace workers. Some workers in both plants showed clinical manifestations of lead poi-
soning; a significant correlation was found between blood lead concentrations and symptom
reporting.
High levels of atmospheric lead are also reported in foundries in which molten lead is
alloyed with other metals. Berg and Zenz (1967) found in one such operation that average con-
centrations of lead in various work areas were 280 to 600 ug/m3. These levels were subse-
quently reduced to 30 to 40 ug/m3 with the installation of forced ventilation systems to ex-
haust the work area atmosphere to the outside.
7.3.2.1.6.2 Welding and cutting of metals containing lead. When metals that contain
lead or are protected with a lead-containing coating are heated in the process of welding or
cutting, copious quantities of lead in the respirable size range may be emitted. Under condi-
tions of poor ventilation, electric arc welding of zinc silicate-coated steel (containing 4.5
mg Pb/cm2 of coating) produced breathing-zone concentrations of lead reaching 15,000 ug/m3,
far in excess of 450 ug/m3, which is the current occupational short-term exposure limit (STEL)
in the United States (Pegues, 1960). Under good ventilation conditions, a concentration of
140 ug/m3 was measured (Tabershaw et al., 1943).
7-65
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In a study of salvage workers using oxyacetylene cutting torches on lead-painted struc-
tural steel under conditions of good ventilation, breathing-zone concentrations of lead aver-
aged 1200 ug/m3 and ranged as high as 2400 pg/m3 (Rieke, 1969). Lead poisoning in workers
dismantling a painted bridge has been reported by Graben et al. (1978). Fischbein et al.
(1978) discuss the exposure of workers dismantling an elevated subway line in New York City,
where the lead content of the paint was as great as 40 percent. The authors report that one
m3 of air can contain 0.05 g lead at the source of emission. Similarly, Grandjean and Kon
(1981) report elevated lead exposures of welders and other employees in a Baltimore, Maryland
shipyard.
7.3.2.1.6.3 Storage battery industry. At all stages in battery manufacture except for
final assembly and finishing, workers are exposed to high air lead concentrations, particular-
ly lead oxide dust. For example, Boscolo et al. (1978) report air lead concentrations of
16-100 (jg/m3 in a battery factory in Italy, while values up to 1315 ug/m3 have been measured
by Richter et al. (1979) in an Israeli battery factory. Excessive concentrations, as great as
5400 Hg/m3, have been reported by the World Health Organization (1977).
7.3.2.1.6.4 Printing industry. The use of lead in typesetting machines has declined in
recent years (see Table 5-1). Air concentrations of 10 to 30 (jg/m3 have been reported where
this technique is used (Parikh et al., 1979). Lead is also a component of inks and dyes used
in the printing industry, and consequently can present a hazard to workers handling these
products.
7.3.2.1.6.5 Alkyl lead manufacture. Workers involved in the manufacture of alkyl lead
compounds are exposed to both inorganic and alkyl lead. Some exposure also occurs at the
petroleum refineries where the two compounds are blended into gasoline, but no data are avail-
able on these blenders.
The major potential hazard in the manufacture of tetraethyl lead and tetramethyl lead is
from skin absorption, which is minimized by the use of protective clothing. Linch et al.
(1970) found a correlation between an index of organic plus inorganic lead concentrations in a
plant and the rate of lead excretion in the urine of workers. Significant concentrations of
organic lead in the urine were found in workers involved with both tetramethyl lead and tetra-
ethyl lead; lead levels in the tetramethyl lead workers were slightly higher because the reac-
tion between the organic reagent and lead alloy takes place at a somewhat higher temperature
and pressure than that employed in tetraethyl lead production.
Cope et al. (1979) used personal air samplers to assess exposures of five alkyl lead
workers exposed primarily to tetraethyl lead. Blood and urine levels were measured over a
six-week period. Alkyl lead levels in air ranged from 1.3 to 1249 pg/m3, while inorganic lead
varied from 1.3 to 62.6 (jg/m3. There was no significant correlation between airborne lead
(either alkyl or inorganic) and blood or urine levels. The authors concluded that biological
7-66
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monitoring, rather than airborne lead monitoring, is a more reliable indicator of potential
exposure problems.
7.3.2.1.6.6 Other occupations. In the rubber products industry and the plastics
industry there are potentially high exposures to lead. The potential hazard of the use of
lead stearate as a stabilizer in the manufacture of polyvinyl chloride was noted in the 1971
Annual Report of the British Chief Inspector of Factories (United Kingdom Department of
Employment, Chief Inspector of Factories 1972). The inspector stated that the number of
reported cases of lead poisoning in the plastics industry was second only to that in the lead
smelting industry. Scarlato et al. (1969) reported other individual cases of exposure. The
source of this problem is the dust that is generated when the lead stearate is milled and
mixed with the polyvinyl chloride and the plasticizer. An encapsulated stabilizer which
greatly reduces the occupational hazard is reported by Fischbein et al. (1982).
Sakurai et al. (1974), in a study of bioindicators of lead exposure, found ambient air
concentrations averaging 58 |jg/m3 in the lead-covering department of a rubber hose manufactu-
ring plant. Unfortunately, no ambient air measurements were taken for other departments or
the control group.
The manufacture of cans with leaded seams may expose workers to elevated ambient lead
levels. Bishop (1980) reports airborne lead concentrations of 25-800 ug/m3 in several can
manufacturing plants in the United Kingdom. Between 23 and 54 percent of the airborne lead
was associated with respirable particles, based on cyclone sampler data.
Firing ranges may be characterized by high airborne lead concentrations, hence instruc-
tors who spend considerable amounts of time in such areas may be exposed to lead. For exam-
ple, Smith (1976) reports airborne lead concentrations of 30-160 \ig/m3 at a firing range in
the United Kingdom. Anderson et al. (1977) discuss lead poisoning in a 17 year old male
employee of a New York City firing range, where airborne lead concentrations as great as 1000
ug/m3 were measured during sweeping operations. Another report from the same research group
presents time-weighted average exposures of instructors of 45-900 ug/m3 in three New York City
firing ranges (Fischbein et al., 1979).
Removal of leaded paint from walls and other surfaces in old houses may pose a health
hazard. Feldman (1978) reports an airborne lead concentration of 510 ug/m3, after 22 minutes
of sanding an outdoor post coated with paint containing 2.5 mg Pb/cm2. After only five min-
utes of sanding an indoor window sill containing 0.8-0.9 mg Pb/cm2, the air contained 550
ug/m3. Homeowners who attempt to remove leaded paint themselves may be at risk of excessive
lead exposure. Garage mechanics may be exposed to excessive lead concentrations. Clausen and
Rastogi (1977) report airborne lead levels of 0.2-35.5 ug/m3 in ten garages in Denmark; the
greatest concentration was measured in a paint workshop. Used motor oils were found to
contain 1500-3500 ug Pb/g, while one brand of unused gear oil contained 9280 ug Pb/g. The
7-67
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authors state that absorption through damaged skin could be an important exposure pathway.
Other occupations involving risk of lead exposure include stained glass manufacturing and re-
pair, arts and crafts, and soldering and splicing.
7.3.2.1.7 Secondary occupational exposure. Winegar et al. (1977) examined environmental con-
centrations as well as biological indicators and symptom reporting in workers in a secondary
lead smelter near St. Paul, Minnesota. The smelter recovers approximately 9000 metric tons of
lead per year from automotive batteries. The lead concentrations in cuff dust from trousers
worn by two workers were 60,000 and 600,000 M9/9- The amount of lead contained in pieces of
cloth 1 cm2 cut from the bottoms of trousers worn by the workers ranged from 110 to 3000 ug,
with a median of 410 ug. In all cases, the trousers were worn under coveralls. Dust samples
from 25 households of smelter workers ranged from 120 to 26,000 |jg/g, with a median of 2400
ug/g. No significant correlations were found between dust lead concentrations and biological
indicators, or between symptom reporting and biological indicators. However, there was an in-
creased frequency of certain objective physical signs, possibly due to lead toxicity, with in-
creased blood lead level. The authors also concluded that the high dust lead levels in the
workers' homes are most likely due to lead originating in the smelter.
7.3.2.2 Additive Exposure Due to Age, Sex, or Socio-Economic Status
7.3.2.2.1 Quality and quantity of food. The quantity of food consumed per body weight varies
greatly with age and somewhat with sex. A 14 kg, 2-year-old child eats and drinks 1.5 kg food
and water per day. This is 110 g/kg, or 3 times the consumption of an 80 kg adult male, who
eats 39 g/kg. Teenage girls consume less than boys and elderly women eat more than men, on a
body weight basis.
It is likely that poor people eat more canned foods and less frozen and pre-prepared
foods. Rural populations probably eat more home-grown foods and meats packed locally.
7.3.2.2.2 Mouthing behavior of children. Children place their mouths on dust-collecting sur-
faces and lick non-food items with their tongues. This fingersucking and mouthing activity
are natural forms of behavior for young children that expose them to some of the highest con-
centrations of lead in their environment. A single gram of dust may contain ten times more
lead than the total diet of the child.
7.3.2.3 Special Habits or Activities. Rabinowitz and Needleman (1984) found a positive
correlation between cord blood lead and such maternal exposure factors as use of tobacco, hard
alcohol, coffee, and amount of lead in dust. Factors unrelated to cord blood lead levels were
amount of dust, tap water lead, air lead, and lead paint. One or more of the above exposure
factors may be correlated with other factors, such as race, marital status, schooling, or
maternal age. Of these, race and marital status demonstrated a relationship to blood lead.
Whereas this study did not attempt to quantify actual exposure, it does identify several acti-
vities that are likely to increase human exposure to lead.
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7.3.2.3.1 Smoking. Lead is also present in tobacco. The World Health Organization (1977)
estimates a lead content of 2.5-12.2 |jg per cigarette; roughly two to six percent of this lead
may be inhaled by the smoker. The National Academy of Sciences (1980) has used these data to
conclude that a typical urban resident who smokes 30 cigarettes per day may inhale roughly
equal amounts of lead from smoking and from breathing urban air.
7.3.2.3.2 Alcoholic beverages. Reports of lead in European wines (Olsen et al., 1981;
Boudene et al., 1975; Zurlo and Griffini, 1973) show concentrations averaging 100-200 ug/1 and
ranging as high as 300 ug/1. Measurements of lead in domestic wines were in the range of
100-300 |jg/l for California wines with and without lead foil caps. The U.S. Food and Drug
Administration (1983) found 30 ug/1 in the 1982 Market Basket Survey. The average daily con-
sumption of table wine by a 25- to 30-year-old adult in the U.S. is about 12 g. Even with a
lead content of 0.1 ug/g, which is ten times higher than drinking water, wine does not appear
to represent a significant potential exposure to lead. At one liter per day, however, lead
consumption in wine would be greater than the total baseline consumption.
McDonald (1981) points out that older wines with lead foil caps may represent a hazard,
especially if they have been damaged or corroded. Wai et al. (1979) found that the lead con-
tent of wine rose from 200 to 1200 ug/1 when the wine was allowed to pass over the thin ring
of residue left by the corroded lead foil cap. Newer wines (1971 and later) use other means
of sealing. If a lead foil is used, the foil is tin-plated and coated with an acid-resistant
substance. Lead concentrations in beer are generally lower than those in wine; Thalacker
(1980) reports a maximum concentration of 80 ug/1 in several brands of German beer. The U.S.
Food and Drug Administration (1983) found 13 ug/1 in beer consumed by Americans (Table 70-1).
7.3.2.3.3 Pica. Pica is the compulsive, habitual consumption of non-food items, such as
paint chips and soil. This habit can present a significant lead exposure to the afflicted
person, especially to children, who are more apt to have pica. There are very little data on
the amounts of paint or soil eaten by children with varying degrees of pica. Exposure can
only be expressed on a unit basis. Billick and Gray (1978) report lead concentrations of 1000
to 5000 ug/cm2 surface area in lead-based paint pigments. To a child with pica, a single chip
of paint can represent greater exposure than any other source of lead. A gram of urban soil
may have 150 to 2000 ug lead.
7.3.2.3.4 Glazed earthenware vessels. Another potential source of dietary lead poisoning is
the use of inadequately glazed earthenware vessels for food storage and cooking. An example
of this danger involved the severe poisoning of a family in Idaho that resulted from drinking
orange juice that had been stored in an earthenware pitcher (Block, 1969). Similar cases,
sometimes including fatalities, have involved other relatively acidic beverages such as fruit
juices and soft drinks, and have been documented by other workers (Klein et al., 1970; Harris
and Elsen, 1967). Because of these incidents, the U.S. Food and Drug Administration (1980)
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has established a maximum permissible concentration of 2.5-7 pg Pb/ml in solution after leach-
ing with 4 percent acetic acid in the kitchenware for 24 hours, depending on the shape and
volume of the vessel.
Inadequately glazed pottery manufactured in other countries continues to pose a signifi-
cant health hazard. For example, Spielholtz and Kaplan (1980) report 24-hour acetic acid-
leached lead concentrations as great as 4400 ug/g in Mexican pottery. The leached lead
decreased with exposure time, and after several days appears to asymptotically approach a
value which may be as high as 600 ug/g. These investigators have also measured excessive lead
concentrations leached into acidic foods cooked for two hours in the same pottery. Similarly,
Acra et al. (1981) report that 85 percent of 275 earthenware vessels produced in primitive
Lebanese potteries had lead concentrations above the 7 pg/g limit set by the U.S. FDA. How-
ever, only 9 percent of 75 vessels produced in a modern Beirut pottery exceeded the limit.
Cubbon et al. (1981) have examined properly glazed ceramic plates in the United Kingdom, and
have found a decrease in leached lead with exposure time down to very low levels. The authors
state that earthenware satisfying the 7 ug/g limit can contribute about 3 ug/day to the
dietary intake of the average consumer.
7.3.2.3.5 Hobbies. There are a few hobbies where the use of metallic lead or solder may pre-
sent a hazard to the user. Examples are electronics projects, stained glass window construc-
tion, and firing range ammunition recovery. There are no reports in which the exposure to
lead has been quantified during these activities.
7.3.3 Summary of Additive Exposure Factors
Beyond the baseline level of human exposure, additional amounts of lead consumption are
largely a matter of individual choice or circumstance. Many of these additional exposures
arise from the ingestion of atmospheric lead in dust. In one or more ways probably 90 percent
of the American population are exposed to lead at greater than baseline levels. A summary of
the most common additive exposure factors appears in Table 7-20. In some cases, the additive
exposure can be fully quantified and the amount of lead consumed can be added to the baseline
consumption. These may be continuous (urban residence), or seasonal (family gardening) expo-
sures. Some factors can be quantified only on a unit basis because of wide ranges in exposure
duration or concentration. For example, factors affecting occupational exposure are air lead
concentrations (10-4000 ug/m3), use and efficiency of respirators, length of time of exposure
dust control techniques, and worker training in occupational hygiene.
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7.4 SUMMARY
Ambient airborne lead concentrations have shown no marked trend from 1965 to 1977. Over
the past five years, however, distinct decreases have occurred. The mean air concentration
has dropped from 1.3 ng/m3 in 1977 to 0.40 ug/m3 in 1984. This decrease reflects the lower
lead emissions from mobile sources in recent years. Airborne size distribution data indicate
that most of the airborne lead mass is found in submicron particles.
Atmospheric lead is deposited on vegetation and soil surfaces, entering the human food
chain through contamination of grains and leafy vegetables, of pasture lands, and of soil
moisture taken up by all crops. Lead contamination of drinking water supplies appears to
originate mostly from within the distribution system.
Most people receive the largest portion of their lead intake through foods. Unprocessed
foods such as fresh fruits and vegetables receive lead by atmospheric deposition as well as
uptake from soil; crops grown near heavily traveled roads generally have greater lead levels
than those grown at greater distances from traffic. For many crops the edible internal por-
tions of the plant (e.g., kernels of corn and wheat) have considerably less lead than the
outer, more exposed parts, such as stems, leaves, and husks. Atmospheric lead accounts for
about 45 percent of the total adult lead exposure, and 65 percent of the exposure for
children. Processed foods have greater lead concentrations than unprocessed foods, due to
lead inadvertently added during processing. Foods packaged in soldered cans have much greater
lead levels than foods packaged in other types of containers. About 35 percent of the base-
line adult exposure to lead results from the use of solder lead in packaging food and distri-
buting drinking water.
Significant amounts of lead in drinking water can result from contamination at the water
source and from the use of lead solder in the water distribution system. Atmospheric deposi-
tion has been shown to increase lead in rivers, reservoirs, and other sources of drinking
water; in some areas, however, lead pipes pose a more serious problem. Soft, acidic water in
homes with lead plumbing may have excessive lead concentrations. Besides direct consumption
of the water, exposure may occur when vegetables and other foods are cooked in water contain-
ing lead.
All of the categories of potential lead exposure discussed above may influence or be in-
fluenced by dust and soil. For example, lead in street dust is derived primarily from vehic-
ular emissions, while in house dust may originate from nearby stationary or mobile sources.
Food and water may include lead adsorbed from soil as well as deposited atmospheric material.
Flaking lead-based paint has been shown to increase soil lead levels. Natural concentrations
of lead in soil average approximately 15 ug/g; this natural lead, in addition to anthropogenic
lead emissions, influences human exposure.
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Americans living in rural areas away from sources of atmospheric lead consume 35-55 \ig
Pb/day from all sources. Circumstances that can increase this exposure are urban residence
(25-100 ug/day), family garden on high-lead soil (40-100 pg/day), houses with interior lead-
based paint (20-85 (jg/day), and residence near a smelter (400-900 (jg/day). Occupational
settings, smoking, and wine consumption also can increase consumption of lead according to the
degree of exposure.
A number of manmade materials are known to contain lead, the most important being paint
and plastics. Lead-based interior paints, although no longer used, are a major problem in
older homes. Small children who ingest paint flakes can receive excessive lead exposure.
Incineration of plastics may emit large amounts of lead into the atmosphere. Because of the
increasing use of plastics, this source is likely to become more important. Other manmade
materials containing lead include colored dyes, cosmetic products, candle wicks, and products
made of pewter and silver.
The greatest occupational exposures are found in the lead smelting and refining indus-
tries. Excessive airborne lead concentrations and dust lead levels are occasionally found in
primary and secondary smelters; smaller exposures are associated with mining and processing of
the lead ores. Welding and cutting of metal surfaces coated with lead-based paint may also
result in excessive exposure. Other occupations with potentially high exposures to lead in-
clude the manufacture of lead storage batteries, printing equipment, alkyl lead, rubber pro-
ducts, plastics, and cans; individuals removing lead paint from walls and those who work in
indoor firing ranges may also be exposed to lead.
Environmental contamination by lead should be measured in terms of the total amount of
lead emitted to the biosphere. American industry contributes several hundred thousand tons of
lead to the environment each year: 55,000 tons from petroleum additives, 50,000 tons from am-
munition, 45,000 tons in glass and ceramic products, 16,000 tons in paint pigments, 8,000 tons
in food can solder, and untold thousands of tons of captured wastes during smelting, refining,
and coal combustion. These are uses of lead that are generally not recoverable, thus they
represent a permanent contamination of the human or natural environment. Although much of
this lead is confined to municipal and industrial waste dumps, a large amount is emitted to
the atmosphere, waterways, and soil, to become a part of the biosphere.
Potential human exposure can be expressed as the concentrations of lead in these environ-
mental components (air, dust, food, and water) that interface with man. It appears that, with
the exception of extraordinary cases of exposure, about 80 to 100 ug of lead are consumed
daily by each American, including additional exposure above baseline.
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Commission of the European Communities; pp. 93-98.
7-86
-------
APPENDIX 7A
SUPPLEMENTAL AIR MONITORING INFORMATION
7A.1 AIRBORNE LEAD SIZE DISTRIBUTION
In Section 7.2.1.3.1, several studies of the particle size distributions for atmospheric
lead were discussed. The distributions at forty locations were given in Figure 7-5. Supple-
mentary information from each of these studies is given in Table 7A-1.
7A-1
-------
TABLE 7A-1.
INFORMATION ASSOCIATED WITH THE AIRBORNE LEAD SI2E
DISTRIBUTIONS OF FIGURE 7-5
Graph
no. Reference
1 Lee et al. (1972)
2 Lee et al. (1972)
3 Lee et al. (1972)
4 Lee et al. (1972)
Dates of sampling
Jan. - Dec. 1970
Average of 4 quarterly
composited samples,
representing a total of
21 sampling periods of
24 hours each
Mar. - Dec. 1970
Same averaging as
Graph 1, total of 18
sampling periods
Jan. - Dec. 1970
Same averaging as
Graph 1, total of
21 sampling periods
Mar. - Dec. 1970
Same averaging as
Location of sampling
Chicago, Illinois
Cincinnati, Ohio
Denver, Colorado
Philadelphia,
Pennsylvania
Type of sampler
Modified Anderson
impactor with backup
filter
Modified Andersen
impactor with backup
filter
Modified Andersen
inpactor with backup
filter
Modified Andersen
impactor with backup
T Approx.
pg/m3 MMD u«
3.2 0.68
1.8 0.48
1.8 0.50
1.6 0.47
Lee et al. (1972)
Lee et al. (1972)
Graph 1, total of 20
sampline periods
Jan. - Dec. 1970
Same averaging as
Graph 1, total of 22
sampling periods
Jan. - Dec. 1970
Same averaging as
Graph .1, total of 23
sampling periods
St. Louis, Missouri
Washington, D.C.
filter
Modified Andersen 1.8
impactor with backup
filter
Modified Andersen 1.3
impactor with backup
filter
0.69
0.42
-------
TABLE 7A-1. (continued)
Graph
no Reference
Dates of sampling
Location of sampling
Type of sampler
ug/m3
Approx.
HMD \OK
7 Lee et al. (1968)
8 Lee et al. (1968)
9 Peden (1977)
10 Peden (1977)
>
j
11 Peden (1977)
12 Peden (1977)
13 Peden (1977)
14 Peden (1977)
September 1966
Average of 14 runs,
24 hours each
February 1967
Average of 3 runs
4 days each
Summer 1975
Average of 4 runs,
average 8 days each
Summer 1972
Average of 3 runs,
average 10 days each
Summer 1973
Average of 2 runs
average 5 days each
Summer 1973
Average of 2 runs,
average 6 days each
Summer 1972
Average of 9 runs,
average 9 days each
Summer 1975
Average of 4 runs,
average 8 days each
Cincinnati , Ohio
Fairfax, Ohio
suburb of Cincinnati
Alton, Illinois,
industrial area near
St. Louis
Centreville, Illinois,
downwind of a zinc
smelter
Collinsville, Illinois
industrial area near
St. Louis
KMOX radio transmitter,
Illinois, industrial
area near St. Louis
Pere Marquette State
Park, lllionis, upwind
of St. Louis
Wood River, Illinois,
industrial area near
St. Louis
Andersen impactor with 2.8
backup filter, 1.2m
above the ground
Andersen Impactor with 0.69
backup filter, 1.2n
above the ground
Andersen impactor 0.24
no backup filter
Andersen impactor 0.62
with backup filter
Andersen impactor 0.67
with backup filter
Andersen impactor 0.60
with backup filter
Andersen impactor 0.15
with backup filter
Andersen impactor, 0.27
no backup filter
0.29
0.42
2.1
0.41
0.24
0.31
0.51
1.8
-------
TABLE 7A-1 (continued)
Graph
no Reference
15 Cholak et al.
(1968)
16 McDonald and
Duncan (1979)
17 Dorn et al. (1976)
2 18 Dorn et al. (1976)
-Pa
19 Daines et al .
(1970)
20 Martens et al.
(1973)
21 Lundgren (1970)
22 Huntzicker et al.
(1975)
Dates of sampling
April 1968
average of several runs,
3 days each
June 1975
One run of 15 days
Winter, spring,
sumer 1972
Average of 3 runs,
27 days each
Winter, spring,
suwner 1972
Average of 3 runs,
14 days each
1968
Average of continuous
1-week runs over an
8-Bonth period
July 1971
One run of 4 days
November 1968
Average of 10 runs,
16 hours each
Hay 1973
One run of 8 hours
Location of sampling
3 sites: 10,400 and
3300M fro* Interstate
75, Cincinnati, Ohio
Glasgow, Scotland
Southeast Missouri,
BOOM fro» a lead
sutelter
Southeast Missouri,
75 k» fro* the lead
sue Her of Graph 17
3 sites: 9, 76, and
530* fro* U.S. Route 1,
New Brunswick,
New Jersey
9 sites throughout
San Francisco area
Riverside, California
Shoulder of Pasadena
Freeway near downtown
Type of sampler
Andersen inpactor
with backup filter
Case 11 a inpactor
with backup filter,
30n above the ground
Andersen inpactor,
no backup filter,
1.7* above the ground
Andersen inpactor,
no backup filter,
1.7n above the ground
Cascade i*pactor with
backup filter
Andersen inpactor
with backup filter
Lundgren i*pactor
Andersen iapactor
with backup filter,
C
T
M9/"3
7.8*
1.7
1.1
0.53
1.0
0.11
4.5
2.2
1.5
0.84
0.59
14.0
Approx.
HMD \m
1
0.32
0.51
3.8
2.4
0.35
0.49
0.50
0.32
Los Angeles, California
2m above the ground
-------
TABLE 7A-1 (continued)
Graph
no
23
24
25
i
'26
27
28
Reference
Huntzicker et al.
(1975)
Davidson (1977)
Davidson et al.
(1980)
Davidson et al.
(1981a)
Davidson et al.
(1981b)
Goold and
Davidson (1982)
Dates of sampling
Februray 1974
One run of 6 days
Hay and July 1975
Average of 2 runs,
61 hours each
October 1979
One run of 120 hours
July-Sep. 1979
Average of 2 runs,
90 hours each
December 1979
One run of 52 hours
June 1980
One run of 72 hours
Location of sampling
Pasadena, California
Pasadena, California
Clingman's Dome
Great Smokies National
Park, elev. 2024m
Pittsburgh, Pennsylvania
Nepal Himalayas
elev. 3962m
Export, Pennsylvania
rural site 40 km
C
T
Type of sampler vg/m3
Andersen inpactor 3.5
with backup filter,
on roof of 4 story
building
Modified Andersen 1.2
impactor with backup
filter on roof of 4
story building
2 Modified Andersen 0.014
impactors with backup
filters, 1.2m above
the ground
Modified Andersen 0.60
impactor with backup
filter, 4m above the
ground
Modified Andersen 0.0014
impactor with backup
filter, 1.2i» above
the ground
2 Modified Andersen 0.111
impactors with backup
Approx.
HMD (im
0.72
0.97
1.0
0.56
0.54
1.2
29 Goold and
Davidson (1982)
July 1980
One run of 34 hours
east of Pittsburgh
Packwood, Washington
rural site in Gifford
Pinchot National Forest
filters, 1.2m above
the ground
Modified Andersen
impactor with backup
filter, l.5m above
the ground
0.016
0.40
-------
TABLE 7A-1 (continued)
Graph
no
30
31
32
33
34
35
36
37
Reference
Goold and
Davidson (1982)
Duce et al.
(1976)
Duce et al.
(1976)
Harrison et al.
(1971)
Gillette and
Winchester (1972)
Gillette and
Winchester (1972)
Gillette and
Winchester (1972)
Johansson et al.
(1976)
Dates of sampling
July-Aug. 1980
One run of 92 hours
May - June 1975
One run of 112 hours
July 1975
One run of 79 hours
April 1968
Average of 21 runs,
2 hours each
Oct. 1968
Average of 15 runs,
24 hours each
May - Sept. 1968
Average of 10 runs,
8 hours each
Oct. 1968
Average of 3 runs,
24 hours each
June - July 1973
Average of 15 runs,
Location of sampling
Hurricane Ridge
Olympic National
Park elev. 1600m
Southeast coast of
Bermuda
Southeast coast of
Bermuda
Ann Arbor, Michigan
Ann Arbor, Michigan
Chicago, Illinois
Lincoln, Nebraska
2 sites in Tallahassee,
Florida
C
T Approx.
Type of sampler MS/"3 MMD u«
Modified Andersen 0.0024 0.87
impactor with backup
filter, 1.5m above
the ground
Sierra high-volume 0.0085 0.57
impactor with backup
filter, 20m above the
ground
Sierra high-volume 0.0041 0.43
impactor with backup
filter, 20m above the
ground
Modified Andersen 1.8 0.16
impactor with backup
filter, 20m above the
ground
Andersen impactor with 0.82 0.28
backup filter
Andersen impactor with 1.9 0.39
backup filter
Andersen impactor with 0.14 0.42
backup filter
Delron Battelle-type 0.24 0.62
impactor, no backup
average 50 hr each
filter, on building roofs
-------
TABLE 7A-1 (continued)
Graph
no Reference
38 Cawse et al.
(1974)
39 Pattenden et al .
^ (1974)
40 Bernstein and
Rahn (1979)
Dates of sampling
July - Dec. 1973
May - Aug. 1973
Average of 4 runs,
1 month each
Aug. 1976
Average of 4 runs,
1 week each
C
T
Location of sampling Type of sampler ug/m3
Chilton, England Andersen impactor with 0.16
backup filter, 1.5m above
the ground
Trebanos, England Andersen impactor with 0.23
backup filter, 1.5m above
the ground
New York City Cyclone sampling 1.2
system with backup
filter, on roof on
15 story building
Approx.
MHO MM
0.57
0.74
0.64
•Airborne concentrations for filters run at the same sites as the impactor, but during different time periods. Impactor concentrations not available.
-------
APPENDIX 7B
SUPPLEMENTAL SOIL AND DUST INFORMATION
Lead in soil, and dust of soil origin, is discussed in Section 7.2.2. The data show
average soil concentrations are 8-25 ng/g, and dust from this soil rarely exceeds 80-100 |jg/g.
Street dust, household dust, and occupational dusts often exceed this level by one to two
orders of magnitude. Tables 7B-1 and 7B-2 summarize several studies of street dust. Table
7B-3 shows data on household and residential soil dust. These data support the estimates of
mean lead concentrations in dust discussed in Section 7.3.1.4. Table 7B-4 gives airborne lead
concentrations for an occupational setting, which are only qualitatively related to dust lead
concentrations.
7B-1
-------
TABLE 7B-1. LEAD DUST ON AND NEAR HEAVILY TRAVELED ROADWAYS
Sampling site
Washington, DC:
Busy intersection
Many sites
Chicago:
Near expressway
Philadelphia:
Near expressway
Brooklyn:
Near expressway
New York City:
Near expressway
Detroit:
Street dust
Philadelphia:
Gutter (low pressure)
Gutter (high pressure)
Miscellaneous U.S. Cities:
Highways and tunnels
Netherlands:
Heavily traveled roads
TABLE 7B-2. LEAD
No. of
Site samples
Car parks 4
16
Garage forecourts 2
7
Town centre streets 13
Main roads 19
Residential areas 7
Rural roads 4
Concentration
M9 Pb/g
13,000
4000-8000
6600
3000-8000
900-4900
2000
970-1200
210-2600
280-8200
10,000-20,000
5000
CONCENTRATIONS IN STREET
Range of
concentrations
39,700 - 51,900
950 - 15,000
44,100 - 48,900
1,370 - 4,480
840 - 4,530
740 - 4,880
620 - 1,240
410 - 870
Reference
Fritsch and Prival (1972)
Kennedy (1973)
Lombardo (1973)
Pinkerton et al. (1973)
Ter Haar and Aronow (1974)
Shapiro et al. (1973)
Shapiro et al. (1973)
Buckley et al. (1973)
Rameau (1973)
DUST IN LANCASTER, ENGLAND
Standard
Mean deviation
46,300 5,900
4,560 3,700
46,500
2,310 1,150
2,130 960
1,890 1,030
850 230
570 210
Source: Harrison (1979).
7B-2
-------
TABLE 7B-3. LEAD DUST IN RESIDENTIAL AREAS
Sampling site
Concentration
(M9 Pb/g)
Reference
Philadelphia:
Classroom
Playground
Window frames
Boston and New York:
House dust
Brattleboro, VT:
In home
New York City:
Middle Class
Residential
Philadelphia:
Urban industrial
Residential
Suburban
Derbyshire, England:
Low soil lead area
High soil lead area
2000
3000
1750
1000-2000
500-900
610-740
930-16,000
290-1000
280-1500
130-3000
1050-28,000
Shapiro et al. (1973)
Needleman and Scanlon (1973)
Darrow and Schroeder (1974)
Pinkerton et al. (1973)
Needleman et al. (1974)
Needleman et al. (1974)
Needleman et al. (1974)
Barltrop et al. (1975)
Barltrop et al. (1975)
•TABLE 7B-4. AIRBORNE LEAD CONCENTRATIONS BASED ON PERSONAL SAMPLERS, WORN BY
EMPLOYEES AT A LEAD MINING AND GRINDING OPERATION IN THE MISSOURI LEAD BELT
(|jg/m3)
Occupation
Mill operator
Flotation operator
Filter operator
Crusher operator
Sample finisher
Crusher utility
Shift boss
Equipment operator
N*
6
4
4
4
2
1
5
1
High
300
750
2450
590
10,000
—
560
™ ™
Low
50
100
380
20
7070
--
110
"
Mean
180
320
. 1330
190
8530
70
290
430
*N denotes number of air samples.
Source: Roy (1977).
7B-3
-------
APPENDIX 7C
STUDIES OF SPECIFIC POINT SOURCES
OF LEAD
This collection of studies is intended to extend and detail the general picture of lead
concentrations in proximity to identified major point sources as portrayed in Chapter 7.
Because emissions and control technology vary between point sources, each point source is
unique in the degree of environmental contamination. The list is by no means all-inclusive,
but is intended to be representative and to supplement the data cited in Chapter 7. In many
of the studies, blood samples of workers and their families were taken. These studies are
also discussed in Chapter 11.
7C.1 SMELTERS AND MINES
7C.1.1 Two Smelter Study
The homes of workers of two unidentified secondary lead smelters in different geograph-
ical areas of the United States were studied by Rice et al. (1978). Paper towels were used to
collect dust from surfaces in each house, following the method of Vostal et al. (1974). A
total of 33 homes of smelter workers and 19 control homes located in the same or similar
neighborhoods were investigated. The geometric mean lead levels on the towels were 79.3 ug
(smelter workers) versus 28.8 ug (controls) in the first area, while in the second area mean
values were 112 ug versus 9.7 ug. Also in the second area, settled dust above doorways was
collected by brushing the dust into glassine envelopes for subsequent analysis. The geometric
mean lead content of this dust in 15 workers' homes was 3300 ug/g, compared with 1200 ug/g
in eight control homes. Curbside dust collected near each home in the second area had a
geometric mean lead content of 1500 ug/g, with no significant difference between worker and
control homes. No significant difference was reported in the paint lead content between
worker and control homes. The authors concluded that lead in dust carried home by these
workers contributed to the lead content of dust in their homes, despite showering and changing
clothes at the plant, and despite work clothes being laundered by the company. Storage of
employee street clothes in dusty lockers, walking across lead-contaminated areas on the way
home, and particulate settling on workers' cars in the parking lot may have been important
factors. Based on measurement of zinc protoporphyrin levels in the blood of children in these
homes, the authors also concluded that the greater lead levels in housedust contributed to in-
creased child absorption of lead.
7C-1
-------
7C.1.2 British Columbia. Canada
Neri et al. (1978) and Schmitt et al. (1979) examined environmental lead levels in the
vicinity of a lead-zinc smelter at Trail, British Columbia. Total emissions from the smelter
averaged about 135 kg Pb/day. Measurements were conducted in Trail (population 12,000), in
Nelson, a control city 41 km north of Trail (population 10,000), and in Vancouver. The annual
mean airborne lead concentrations in Trail and in Nelson were 2.0 and 0.5 ug/m3, respectively.
Mean lead levels in surface soil were 1320 pg/g in Trail (153 samples), 192 ug/g in Nelson (55
samples), and 1545 ug/g in Vancouver (37 samples).
Blood lead measurements show a positive correlation with soil lead levels for children
aged 1-3 years and for first graders, but no significant correlation for ninth graders. The
authors concluded that small children are most likely to ingest soil dust, and hence deposited
smelter-emitted lead may pose a potential hazard for the youngest age group.
7C.1.3 Netherlands
Environmental lead concentrations were measured in 1978 near a secondary lead smelter in
Arnhem, Netherlands (Diemel et al., 1981). Air and dust were sampled in over 100 houses at
distances of 450-1000 meters from the smelter, with outdoor samples of air, dust, and soil
collected for comparison. Results are presented in Table 7C-1. Note that the mean indoor
concentration of total suspended particulates (TSP) is greater than the mean outdoor concen-
tration, yet the mean indoor lead level is smaller than the corresponding outdoor level. The
authors reasoned that indoor sources such as tobacco smoke, consumer products, and decay of
furnishings are likely to be important in affecting indoor TSP; however, much of the indoor
lead was probably carried in from the outside by the occupants, e.g., as dust adhering to
shoes. The importance of resuspension of indoor particles by activity around the house was
also discussed.
7C.1.4 Belgium
Reels et al. (1978, 1980) measured lead levels in the air, in dust, and on childrens1
hands at varying distances from a lead smelter in Belgium (annual production 100,000 metric
tons). Blood data from children living near the smelter were also obtained. Air samples were
collected nearly continuously beginning in September 1973. Table 7C-2 lists the airborne con-
centrations recorded during five distinct population surveys between 1974 and 1978, while
Figure 7C-1 presents air, dust, and hand data for Survey #3 in 1976. Statistical tests showed
that blood lead levels were better correlated with lead on childrens' hands than with air
lead. The authors suggested that ingestion of contaminated dust by hand-to-mouth activities
7C-2
-------
TABLE 7C-1. LEAD CONCENTRATIONS IN INDOOR AND OUTDOOR AIR, INDOOR AND OUTDOOR
DUST, AND OUTDOOR SOIL NEAR THE ARNHEM, NETHERLANDS SECONDARY LEAD SMELTER
(INDOOR CONCENTRATIONS)
Parameter
Arithmetic
mean
Range
Suspended particulate matter
dust concentration ((jg/m3)
lead concentration (pg/m3)
dust lead content (pg/kg)
Dustfall
dust deposition (mg/m2*day)
lead deposition (pg/m2-day)
dust lead content (mg/kg)
Floor dust
amount of dust (mg/m2)
amount of lead (pg/m2)
140
0.27
2670
15.0
9.30
1140
356
166
20-570
0.13-0.74
400-8200
1.4-63.9
1.36-42.4
457-8100
41-2320
18-886
101
101
106
105
105
105
107
101
Dust lead content (mg/kg)
in "fine" floor dust
in "coarse" floor dust
1050
370
463-4740
117-5250
107
101
*N number of houses.
(OUTDOOR CONCENTRATIONS)
Parameter
Arithmetic mean
Range
Suspended particles
dust concentration (pg/m3)
lead concentraton (pg/m3)
(high-volume samplers, 24-hr samples, 2 month's
average)
Lead in dustfall
(pg/m2-day)
(deposit gauges, weekly samples, 2 months'
average)
Lead in soil
(mg/kg 0-5 cm)
Lead in streetdust
(mg/kg <0.3 mm)
64.5
0.42
508
322
860
53.7-73.3
0.28-0.52
208-2210
21-1130
77-2670
Source: Diemel et al. (1981).
7C-3
-------
Pb IN AIR
c
Pb IN DUST
(
Pb ON HAND
(
n
18 cr
20 c
I I I
1 2 3 Mg m>
1 1 1
) 760 1600 2260 Mg g
1 1 1
) 150 300 450 Hg hind
AT LESS THAN 1km FROM LEAD SMELTER
ill! itT
'i 'i Uy
i il iilillH!;i!!lll liiiiilii!! II"!!!!!!!1!! UEM1ESE3
..-..• : : : ! '•: ''•'. : i-x-x-x • : • . .... :j • -:. : : ! x-x-.- Wj : :•;•••::• : ' ! • --; ••' .: "- . ' x-:-:: -:-.;: . :• :. J
^mm. v\ — i
AT 25 km FROM LEAD SMELTER
26 cr
16 9
17 cr
9 9
.
^4<
iJp
i«
-:":' ' ::;r: AIR
URBAN - BRUSSELS (CONTROL) w DUST
if
I^<-»<^§x HAND cr
y%Mm> HAND?
RURAL - HERENT (CONTROL)
21 o*
23 9
CHILDREN 1976
3RD SURVEY
Figure 7C-1. Concentrations of lead in air. in dust, and on children's hands, measured during
the third population survey of Table E. Values obtained less than 1 km from the smelter, at 2.5
km from the smelter, and in two control areas are shown. The number of children (n) is shown
by sex.
Source: Roels et al. (1980).
7C-4
-------
TABLE 7C-2. AIRBORNE CONCENTRATIONS OF LEAD DURING FIVE
POPULATION SURVEYS NEAR A LEAD SMELTER IN BELGIUM*
Study populations
1 Survey
(1974)
2 Survey
(1975)
3 Survey
(1976)
4 Survey
(1977)
5 Survey
(1978)
<1 km
2.5 km
Rural
<1 km
2.5 km
Rural
<1 km
2.5 km
Urban
Rural
<1 km
2.5 km
<1 km
2.5 km
Urban
Rural
Pb-Air
4.06
1.00
0.29
2.94
0.74
3.67
0.80
0.45
0.30
3.42
0.49
2.68
0.54
0.56
0.37
*Additional airborne data in rural and urban areas obtained as controls are also shown.
Source: Roels et al. (1980).
such as nail-biting and thumb-sucking, as well as eating with the hands, may be an important
exposure pathway. It was concluded that intake from contaminated hands contributes at least
two to four times as much lead as inhalation of airborne material.
7C.1.5 Meza River Valley, Yugoslavia
In 1967, work was initiated in the community of Zerjav, situated in the Slovenian Alps on
the Meza River, to investigate contamination by lead of the air, water, snow, soil, vegeta-
tion, and animal life, as well as the human population. The smelter in this community pro-
duces about 20,000 metric tons of lead annually; until 1969 the stack emitted lead oxides
without control by filters or other devices. Five sampling sites with high-volume samplers
operating on a 24-hr basis were established in the four principal settlements within the Meza
River Valley (Figure 7C-2): (1) Zerjav, in the center, the site of the smelter, housing 1503
inhabitants, (20 Rudarjevo, about 2 km to the south of Zerjav with a population of 100;
(3) Crna, some 5 km to the southwest, population 2198, where there are two sites (Crna-SE and
Crna-W); and (4) Mezica, a village about 10 km to the northwest of the smelter with 2515
7C-5
-------
inhabitants. The data in Table 7C-3 are sufficient to depict general environmental contami-
nation of striking proportions.
7C.1.6 Kosova Province, Yugoslavia
Popovac et al. (1982) discuss lead exposure in an industrialized region near the town of
Kosova Mitrovica, Yugoslavia, containing a lead smelter and refinery, and a battery factory.
In 1979, 5800 kg of lead were emitted daily from the lead smelter alone. Ambient air concen-
trations in the town were in the range 21.2-29.2 ug/m3 in 1980, with levels occasionally
reaching 70 pg/m3. The authors report elevated blood lead levels in most of the children
tested; some extremely high values were found, suggesting the presence of congenital lead
poisoning.
7C.1.7 Czechoslovakia
Wagner et al. (1981) measured total suspended particulate and airborne lead concentra-
tions in the vicinity of a waste lead processing plant in Czechoslovakia. Data are shown in
Table 7C-4. Blood lead levels in 90 children living near the plant were significantly greater
than in 61 control children.
7C.1.8 Australia
Heyworth et al. (1981) examined child response to lead in the vicinity of a lead sulfide
mine in Northhampton, Western Australia. Two samples of mine tailings measured in 1969
contained 12,000 ug/g and 28,000 ug/g lead; several additional samples analyzed in 1978 con-
tained 22,000-157,000 ug/g lead. Surface soil from the town boundary contained 300 ug/g,
while a playground and a recreational area had soil containing 11,000 ug/g and 12,000 ug/g
lead, respectively.
Blood lead levels measured in Northhampton children, near the mine, were slightly greater
than levels measured in children living a short distance away. The Northhampton blood lead
levels were also slightly greater than those reported for children in Victoria, Australia
(DeSilva and Donnan, 1980). Heyworth et al. (1981) concluded that the mine tailings could
have increased the lead exposure of children living in the area.
7C.2 BATTERY FACTORIES
7C.2.1 Southern Vermont
Watson et al. (1978) investigated homes of employees of a lead storage battery plant in
southern Vermont in August and September, 1976. Lead levels in household dust, drinking
7C-6
-------
RIVERS
SETTLEMENTS
Figure 7C-2. Schematic plan of lead mine and smelter from Meza Valley,
Yugoslavia, study.
Source: Fugas (1977).
7C-7
-------
TABLE 7C-3. ATMOSPHERIC LEAD CONCENTRATIONS (24-hour) IN THE
MEZA VALLEY, YUGOSLAVIA, NOVEMBER 1971 TO AUGUST 1972
(ug/m3)
Site
Mezica
Zerjav
Rudarjevo
Crna SE
Crna W
Minimum
0.1
0.3
0.5
0.1
0.1
Maximum
236.0
216.5
328.0
258.5
222.0
Average
24.2
29.5
38.4
33.7
28.4
Source: Fugas (1977).
TABLE 7C-4. CONCENTRATIONS OF TOTAL AIRBORNE DUST AND OF AIRBORNE LEAD IN THE
VICINITY OF A WASTE LEAD PROCESSING PLANT IN CZECHOSLOVAKIA,
AND IN A CONTROL AREA INFLUENCED PREDOMINANTLY BY AUTOMOBILE EMISSIONS
Exposed
Control
n
x (ug/m3)
S
range
95% c.i.
n
x (ug/m3)
S
range
95% c.i.
TSP
300
113.6
83.99
19.7-553.4
123.1-104.1
56.0
92.0
40.5
10-210
102.7-81.3
Lead
303
1.33
1.9
0.12-10.9
1.54-1.11
87
0.16
0.07
0.03-0.36
0.17-0.14
n = number of samples; x = mean of 24-hour samples;
s = standard deviation; 95% confidence interval.
Source: Wagner et al. (1981).
7C-8
-------
water, and paint were determined for 22 workers' homes and 22 control homes. The mean lead
concentration in dust in the workers' homes was 2,200 ug/g, compared with 720 ng/g in the
control homes. Blood lead levels in the workers' children were greater than levels in the
control children, and were significantly correlated with dust lead concentrations. No sig-
nificant correlations were found between drinking water lead and blood lead, or between paint
lead and blood lead. It is noteworthy that although 90 percent of the employees showered and
changed clothes at the plant, 87 percent brought their work clothes home for laundering. The
authors concluded that dust carried home by the workers contributed to increased lead absorp-
tion in their children.
7C.2.2 North Carolina
Several cases of elevated environmental lead levels near point sources in North Carolina
have been reported by Dolcourt et al. (1978, 1981). In the first instance, dust lead was
measured in the homes of mothers employed in a battery factory in Raleigh; blood lead levels
in the mothers and their chldren were also measured. Carpet dust was found to contain 1,700-
48,000 ug/g lead in six homes where the children had elevated blood lead levels (>40
ug/dl). The authors concluded that lead carried home on the mothers' clothing resulted in
increased exposure to their children (Dolcourt et al., 1978). In this particular plant, no
uniforms or garment covers were provided by the factory; work clothing was worn home.
In a second case, discarded automobile battery casings from a small-scale lead recovery
operation in rural North Carolina were brought home by a worker and used in the family's
wood-burning stove (Dolcourt et al., 1981). Two samples of indoor dust yielded 13,000 and
41,000 |jg/g lead. A three-year-old girl living in the house developed encephalopathy
resulting in permanent brain damage.
In a third case, also in rural North Carolina, a worker employed in an automobile battery
reclamation plant was found to be operating an illicit battery recycling operation in his
home. Reclaimed lead was melted on the kitchen stove. Soil samples obtained near the house
measured as high as 49 percent lead by weight; the driveway was covered with fragments of
battery casings. Although no family member had evidence of lead poisoning, there were
unexplained deaths among chickens who fed where the lead waste products were discarded
(Dolcourt et al., 1981).
7C.2.3 Oklahoma
Morton et al. (1982) studied lead exposure in children of employees at a battery manu-
facturing plant in Oklahoma. A total of 34 lead-exposed children and 34 control children were
examined during February and March, 1978; 18 children in the lead-exposed group had elevated
blood lead levels (>30 ug/dl), while none of the controls were in this category.
7C-9
-------
It was found that many of the battery factory employees also used lead at home, such as
casting lead into fishing sinkers and using leaded ammunition. A significant difference in
blood lead levels between the two groups of children was found even when families using lead
at home were deleted from the data set. Using the results of personal interviews with the
homemaker in each household, the authors concluded that dust carried home by the employees
resulted in increased exposure of their children. Merely changing clothes at the plant was
deemed insufficient to avoid transporting appreciable amounts of lead home: showering and
shampooing, in addition to changing clothes, was necessary.
7C.2.4 Oakland. California
Environmental lead contamination at the former site of a wet-cell battery manufacturing
plant in Oakland, California was reported by Wesolowski et al. (1979). The plant was opera-
tional from 1924 to 1974, and was demolished in 1976. Soil lead levels at the site measured
shortly after demolition are shown in Table 7C-5. The increase in median concentrations with
depth suggested that the battery plant, rather than emissions from automobiles, was respons-
ible for the elevated soil lead levels. The levels decreased rapidly below 30 cm depth. The
contaminated soil was removed to a sanitary landfill and replaced with clean soil; a park has
subsequently been constructed at the site.
TABLE 7C-5. LEAD CONCENTRATIONS IN SOIL AT THE FORMER SITE OF A WET-CELL
BATTERY MANUFACTURING PLANT IN OAKLAND, CALIFORNIA
Depth
Surface
15 cm
30 cm
N*
24
23
24
Range
57-96,000
13-4200
13-4500
Mean
4300
370
1100
Median
200
200
360
*N = number of samples.
Source: Wesolowski et al. (1979).
7C.2.5 Manchester, England
Elwood et al. (1977) measured lead concentrations in air, dust, soil, vegetation, and tap
water, as well as in the blood of children and adults, in the vicinity of a large battery
factory near Manchester. It was found that lead levels in dust, soil, and vegetation de-
creased with increasing distance from the factory. Airborne lead concentrations did not show
7C-10
-------
a consistent effect with downwind distance, although higher concentrations were found downwind
compared with upwind of the factory. Blood lead levels were greatest in the households of
battery factory employees: other factors such as distance from the factory, car ownership,
age of house, and presence of lead water pipes were outweighed by the presence of a leadworker
in the household. These results strongly suggest that lead dust carried home by the factory
employees is a dominant exposure pathway for their families. The authors also discussed the
work of Burrows (1976), who demonstrated experimentally that the most important means of lead
transport from the factory into the home is via the workers' shoes.
7C-11
-------
APPENDIX 7D
SUPPLEMENTAL DIETARY INFORMATION FROM THE
U.S. FDA TOTAL DIET STUDY
The U.S. Food and Drug Administration published a new Total Diet Food List (Pennington,
1983) based on over 100,000 daily diets from 50,000 participants. Thirty five hundred
categories of foods were condensed to 201 adult food categories for 8 age/sex groups.
Summaries of these data were used in Section 7.3.1.2 to arrive at lead exposures through food,
water, and beverages. For brevity and continuity with the crop data of Section 7.2.2.2.1, it
was necessary to condense the 201 categories of the Pennington study to 25 categories in this
report.
The preliminary lead concentrations for all 201 items of the food list were provided by
U.S. Food and Drug Administration (1985). These data represent four Market Basket Surveys,
each from a different geographic location. Means of these values have been calculated by
EPA, using one-half the detection limit for values reported below detection limit. These data
appear in Table 7D-1.
In condensing the 201 categories of Table 7D-1 to the 9 categories of Table 7-17,
combinations and fractional combinations of categories were made according to the scheme
of Table 7D-2. In this way, specific categories of food more closely identified with farm
products were summarized. The assumptions made concerning the ingredients in the final
product, (mainly water, flour, eggs, and milk) had little influence on the outcome of the
summarization.
7D-1
-------
TABLE 7D-1. FOOD LIST AND PRELIMINARY LEAD CONCENTRATIONS
—I
o
Category
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
; 16
> 17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
Food
Whole milk
Low fat milk
Chocolate milk
Skim milk
Buttermilk
Yogurt, plain
Milkshake
Evaporated milk
Yogurt, sweetened
Cheese, American
Cottage cheese
Cheese, Cheddar
Beef, ground
Beef, chuck roast
Beef, round steak
Beef, sirloin
Pork, ham
Pork chop
Pork sausage
Pork, bacon
Pork roast
Lamb chop
Veal cutlet
Chicken, fried
Chicken, roasted
Turkey, roasted
Beef liver
Frankfurters
Bologna
Salami
Cod/haddock filet
Tuna, canned
Shrimp
Fish sticks, frozen
Eggs, scrambled
Eggs, fried
Eggs, soft boiled
Lead concentration*
(pg/g)
0.02
0.06
0.08
0.04
0.03
0.05
0.04
0.09
0.03
0.05
0.04
0.11
0.02
0.18
0.03
T T
0.04
0.05
0.07 0.18
0.11
0.03
0.03
0.03
0.05
0.22
0.03
0.12
0.07
0.27 0.08
0.10
0.03
T
T
T
T
T
0.03
0.03 0.05
0.10 0.03
T
T
T
0.04
0.02
0.03
0.08 0.06
0.04
0.37
0.05
0.02
0.02 0.02
0.09 0.02
0.05 0.05 0.03
0.07 0.07 0.06
T
0.03 0.05
0.02 0.02
0.08
0.02
0.04
0.04 0.04
0.03
0.02 0.04
0.02
0.03 0.03
0.06 0.14 0.09
0.05
0.03 0.06
0.08 0.03
0.24 0.06 0.07
0.04 0.04
0.02 0.02
0.07
0.06 0.03
Mean
0.003
0.007
0.010
0.008
0.016
0.006
0.040
0.083
0.009
0.016
0.014
0.021
0.016
0.019
0.007
0.002
0.006
0.006
0.021
0.035
0.006
0.006
0.009
0.011
0.013
0.002
0.083
0.002
0.015
0.013
0.024
0.159
0.030
0.010
0.002
0.014
0.013
-------
TABLE 7D-1. (continued)
Category
38
39
40
41
42
43
44
45
46
47
48
49
50
51
^ 52
-------
TABLE 7D-1. (continued)
Category
76
77
78
79
80
81
82
83
84
85
86
87
88
89
>4 9°
3 91
*• 92
93
94
95
96
97
98
99
100
101
102
103
104
105
106
107
108
109
110
111
112
113
Food
Granola
Oat ring cereal
Apple, raw
Orange, raw
Banana, raw
Watermelon, raw
Peach, canned
Peach, raw
Applesauce, canned
Pear, raw
Strawberries, raw
Fruit cocktail, canned
Grapes, raw
Cantaloupe, raw
Pear, canned
Plums, raw
Grapefruit, raw
Pineapple, canned
Cherries, raw
Raisins, dried
Prunes, dried
Avocado, raw
Orange juice, frozen
Apple juice, canned
Grapefruit juice, frozen
Grape juice, canned
Pineapple juice, canned
Prune juice, bottled
Orange juice, canned
Lemonade, frozen
Spinach, canned
Spinach, frozen
Collards, frozen
Lettuce, raw
Cabbage, raw
Coleslaw
Sauerkraut, canned
Broccol i , frozen
Lead concentration*
(M9/9)
0.03
0.03
0.04
0.18
0.02
0.21
0.02
0.03
0.23
0.03
0.24
T
0.03
0.10
0.04
0.05
0.03
0.02
0.06
0.03
0.06
0.08
0.02
0.05
0.04
0.80
0.05
0.05
0.03
0.13
0.77
0.04
0.02
0.04
0.03
0.23
0.04
0.19
0.03
0.24
0.02
0.08
0.22
0.08
0.03
0.07
0.09
0.04
0.11
0.02
0.03
0.07
1.65
0.10
0.27
0.39
0.03
0.02
0.04
0.02
0.02
0.28
0.10
0.13
0.17
0.05
0.04
0.04
0.02
0.04
0.05
0.02
0.02
0.12
0.06
0.04
0.12
0.03
T
T
0.29
0.05
0.04
T
0.18
T
0.10
0.04
0.05
0.03
0.02
0.02
0.02
0.03
1.34
0.06
0.03
0.02
0.02
0.64
T
O.OZ
0.03
T
T
T
0.44
0.05
T
0.02
0.41
T
0.17
0.04
0.04
0.04
0.02
T
0.09
T
0.03
0.03
0.03
0.14
0.02
0.38
0.14
0.08
0.04
0.04
0.02
0.84
0.03
0.03
0.02
0.12
T
0.12
T
0.10
T
0.03
0.22
0.28
T
T
0.04
0.04
0.03
0.04
T
0.11
0.11
0.06
T
0.49
0.04
0.02
T
T
0.46
T
0.02
0.12
0.02
0.05
0.02
0.21
0.09
0.01
0.03
0.07
0.02
0.02
T
0.06
T
0.25
0.02
0.04
0.93
0.03
0.03
T
0.03
0.12
0.03
0.05
0.02
0.04
0.15
T
0.08
T
0.05
0.04
0.02
T
T
0.03
0.04
T
0.09
0.16
0.06
0.06
T
T
0.04
0.04
T
1
Mean
j
0.021
0.021
0.015
0.027
0.006
0.009
0.223
0.022
0.094
0.017
0.019
0.221
0.007
0.015
0.169
0.005
0.008
0.093
0.006
0.033
0.038
0.023
0.007
0.048
0.011
0.053
0.040
0.015
0.053
0.019
0.649
0.066
0.074
0.011
0.014
0.026
0.524
0.016
-------
TABLE 7D-1. (continued)
Category
114
115
116
117
118
119
120
121
122
123
124
125
126
127
^ 128
-------
TABLE 7D-1. (continued)
Category
149
150
151
152
153
154
155
156
157
158
159
160
161
162
163
164
165
166
167
168
169
170
171
172
173
174
175
176
177
178
179
180
181
182
183
184
185
Food
Spaghetti in tomato sauce,
canned
Chicken noodle casserole
Lasagne
Potpie, frozen
Pork chow mein
Frozen dinner
Chicken noodle soup, canned
Tomato soup, canned
Vegetable beef soup, canned
Beef bouillon, canned
Gravy mix
White sauce
Pickles
Margari ne
Salad dressing
Butter
Vegetable oil
Mayonnaise
Cream
Cream substitute
Sugar
Syrup
Jelly
Honey
Catsup
Ice cream
Pudding, instant
Ice cream sandwich
Ice milk
Chocolate cake
Yellow cake
Coffee cake
Doughnuts
Danish pastry
Cookies, choc, chip
Cookies, sandwich type
Apple pie, frozen
Lead concentration*
(Mg/g)
0.06
0.11
0.04
0.32
0.02
0.07
0.04
0.02
0.05
0.10
0.06
0.03
0.06
0.10
0.07
0.06
0.12
0.03
0.05
0.07
0.13
0.16
0.04
0.02
0.06
0.04
0.03
0.04
0.02
0.04
0.06
0.03
0.03
0.02
0.02
0.04
0.02
0.02
0.09
0.06
0.06
0.14
0.04
0.05
0.05
0.06
0.02
0.02
0.04
0.03
0.03
0.03
0.03
0.
0.
0.
T
0.
0.
0.
0.
0.
03
04
06
04
02
03
02
05
0.03
0.
0.
04
02
0.06
0.04
0.02
T
0.07
0.03
T
T
0.06
T
T
0.02
0.03
0.05
0.03
0.04
0.05
0.03
0.03
0.03
0.03
0.03
0.04
0.04
0.02
0.08
0.05
0.04
0.07
0.04
0.02
0.02
0.06
0.03
0.03
0.04
0.04
0.07
0.03
T
0.10
T
0.04
0.10
0.03
0.07
0.04
0.02
0.06
0.02
T
T
0.08
0.02
0.03
0.02
0.03
0.03
0.02
0.10
0.05
0.03
0.04
0.08
0.06
T
T
0.02
0.02
0.29
T
0.02
0.02
0.03
0.04
0.12
T
0.04
0.07
T
0.05
0.04
0.05
0.18
T
0.03
0.08
0.15
0.04
0.15
0.09
0.03
0.02
0.03
Mean
0.016
0.017
0.070
0.012
0.076
0.010
0.044
0.030
0.073
0.021
0.005
0.014
0.044
0.017
0.013
0.019
0.002
0.028
0.010
0.024
0.017
0.009
0.008
0.031
0.010
0.044
0.008
0.058
0.023
0.035
0.025
0.040
0.010
0.028
0.035
0.048
0.017
-------
TABLE 7D-1. (continued)
Category
186
187
188
189
190
191
192
193
! 194
i
195
196
197
198
199
200
201
Food
Pumpkin pie
Candy, milk chocolate
Candy, caramels
Chocolate powder
Gelatin dessert
Soda pop, cola, canned
Soda pop lemon-lime, canned
Soft drink powder
Soda pop, cola, low cal. ,
canned
Coffee, instant
Coffee, instant, decaf.
Tea
Beer, canned
Wine
Whiskey
Water
Lead concentration*
(Mg/g)
0.05
0.09
0.06
0.02
0.13
0.05
0.02
0.03
0.02
T
0.02
0.04
0.04
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.03
0.03
0.09
0.04
0.08
T
0.02
0.03
0.05 0.03 0.05 0.03 0.06
0.11 0.08 0.05 0.05 0.07
0.05 0.02 0.06 0.03 0.03
0.07 0.07 0.07 0.06
T
0.02
T
0.03 0.04 0.03 0.05 0.09
T
0.01
Mean
0.040
0.073
0.034
0.055
0.006
0.004
0.023
0.007
0.011
0.002
0.004
0.002
0.007
0.041
0.005
0.004
Individual values for four Market Basket Surveys. "T" means only a trace detected, missing value means below
detection limit.
+Means determined by EPA using 0.002 (k of detection limit) for values below detection limit and 0.01 for detection
of trace value.
-------
TABLE 7D-2. SCHEME FOR THE CONDENSATION OF 201 CATEGORIES OF
FOOD FROM TABLE 7D-1 INTO 9 CATEGORIES
Whole
Dairy
Meat
1-12,
13-37
164,
167
Category
, 174,
176, 177
0.
0.
0.
1
3
1
Partial
(68-70, 152)
(144, 146),
(143, 155),
»
0
0
Category
0.2 (151
.5 (156)
.2 (144,
, 178-187)
146, 151,
Food Crops
Canned Foods
Canned Juices
Frozen Juices
Soda
Canned Beer
Water
38, 40-44, 46-54, 57-67,
71-81, 83, 85, 86, 88, 89, 91,
92, 94-97, 107-111, 113-117,
121, 123-128, 132-141, 159-163
165, 166, 168-173, 175, 188-190
39, 45, 55, 56, 82, 84, 87, 90,
93, 106, 112, 118-120, 122,
129-131
99, 101, 102, 104
98, 100, 103, 105
191, 192, 194
198
193, 195-197, 199-201
178-187), 0.3 (68-70, 145, 153, 154,
158), 0.4 (152) 0.5 (150), 0.6 (142,
147, 148, 149)
0.2 (148), 0.3 (142, 144, 146,
149-151), 0.4 (147), 0.5 (143,
152), 0.6 (68-70, 178-187),
0.7 (153, 154)
0.1 (142, 145, 149), 0.2 (144, 148,
150, 151), 0.5 (155-157)
0.1 (151), 0.2 (146), 0.4 (143,
155), 0.5 (157), 0.6 (145),
0.7 (158)
7D-8
-------
7E REFERENCES
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8. EFFECTS OF LEAD ON ECOSYSTEMS
8.1 INTRODUCTION
8.1.1 Scope of Chapter 8
This chapter describes the potential effects of atmospheric lead inputs on several types
of ecosystems. An effect is any condition attributable to lead that causes an abnormal
physiological response in individual organisms or that perturbs the normal processes of an
ecosystem. A distinction is made among natural, cultivated, and urban ecosystems, and extend-
ed discussions are included on the mobility and bioavailability of lead in ecosystems.
There are many reports on the effects of lead on individual populations of plants and
animals and a few studies on the effects of lead in simulated ecosystems or microcosms.
However, the most realistic studies are those that examine the effects of lead on entire
ecosystems, as they incorporate all of the ecological interactions among the various popu-
lations and all of the chemical and biochemical processes relating to lead (National Academy
of Sciences, 1981). Unfortunately, these studies have also had to cope with the inherent
variability of natural systems and the confounding frustrations of large-scale projects.
Consequently, there are only a handful of ecosystem studies on which to base this report.
Effects at the ecosystem level are usually seen as a form of stress. In nearly every
case of stress caused by pollutants, the initial effect is to cause cytological or biochemical
changes in specific cells of individual organisms. Mclaughlin (1985) has summarized some of
the effects on forest ecosystems that have been caused by air pollutants. Examples of cyto-
logical or biochemical changes are reduction in enzyme activity, a change in membrane perme-
ability or osmotic potential, or a loss of organelle integrity. These cellular changes cause
some disruption of physiological function, such as photosynthesis, respiration, transpiration,
root uptake, the opening and closing of stomata, or a disruption of resource allocation, such
as growth, reproduction, or defense mechanisms. In turn, the growth of the individual may be
directly or indirectly affected, either in amount, timing, or quality. These effects on the
individual can cause a change in the productivity of the entire community. Some of the expec-
ted effects on the community as a whole or populations within the community might be reduced
growth, increased mortality, unbalanced competition, delayed succession, or reduced rate of
reproduction.
Because of the complexity of processes that can affect an ecosystem, it is difficult to
predict the mechanism by which a specific air pollutant might influence an ecosystem. General
categories of effects are those that predispose an ecosystem to stress, those that incite
stress, and those that contribute directly to stress (from Manion, 1981, as modified by
Mclaughlin, 1985). Examples of predisposition are chronic weakening caused by changes in
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climate, soil moisture, soil nutrients, or competition. Inciting factors are triggering
episodes, such as insect defoliation, frost, drought, mechanical injury or increased salinity.
Those factors that directly contribute to effects generally do so by accelerating the process-
es already taking place, such as infections of bark beetles, canker fungi, viruses, root decay
fungi, or increased competition. As a general rule, air pollutants are either predisposing or
inciting types of agents and are noticed only when a change is triggered by the effect of the
pollutant. The effects of air pollutants may go unnoticed for decades, causing only a chronic
weakening that cannot be detected by normal methods of evaluating ecosystem stability.
The principle sources of lead entering an ecosystem include the following: the atmo-
sphere (largely from automotive emissions), paint chips, spent ammunition, the application of
fertilizers and pesticides, and the careless disposal of lead-acid batteries or other indus-
trial products. Atmospheric lead is deposited on the surfaces of vegetation as well as on
ground and water surfaces. In terrestrial ecosystems, this lead is transferred to the upper
layers of the soil surface, where it may be retained for a period of several years. The move-
ment of lead within ecosystems is influenced by the chemical and physical properties of lead
and by the biogeochemical processes within the ecosystem. Lead is persistent, but in the
appropriate chemical environment, may undergo transformations that affect its solubility
(e.g., formation of lead sulfate in soils), its bioavailability (e.g., chelation with humic
substances), or its toxicity (e.g., chemical methylation).
Because the effects of lead on ecosystems begin with some initial effect on specific
cells of individuals within the ecosystem, there are a number of mechanisms or strategies
whereby individuals or populations may have developed a resistance to lead toxicity. Wood
(1984) has described six potential strategies for resistance to toxic metals: 1) the cell may
pump the metal out through the cell membrane, a process that requires energy; 2) the metal may
be enzymatically oxidized or reduced to a less toxic form; 3) the cell may synthesize a poly-
mer to trap and remove the metal; 4) the metal may be bound to the cell surface; 5) the metal
may be precipitated as an insoluble metal complex; 6) the metal may be biomethylated and
transported through the cell membrane by diffusion a process that requires less energy than
actively pumping. The evidence for the biomethylation of lead is circumstantial at best and
clearly not conclusive (Craig and Wood, 1981, Reisinger et al., 1981, Chau, 1986).
The previous Air Quality Criteria for Lead (U.S. Environmental Protection Agency, 1977)
recognized the problems of atmospheric lead exposure incurred by all organisms including man.
Emphasis in the chapter on ecosystem effects was given to reports of toxic effects on specific
groups of organisms, e.g. domestic animals, wildlife, aquatic organisms, and vascular and non-
vascular plants. Forage containing lead at 80 ug/g dry weight was reported to be lethal to
horses, whereas 300 |jg/9 dry weight caused lethal clinical symptoms in cattle. This report
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will attempt to place the data in the context of sublethal effects of lead exposure, to extend
the conclusions to a greater variety of domestic animals, and to describe the types and ranges
of exposures in ecosystems likely to present a problem for domestic animals.
Research on lead in wildlife has traditionally fallen into the following somewhat arti-
ficial categories: waterfowl; birds and small mammals; fish; and invertebrates. In all these
categories, no correlation could be made in the 1977 report between toxic effects and environ-
mental concentrations. Some recent toxicity studies have been completed on fish and inverte-
brates and the data are reported below, but there is still little information on the levels of
lead that can cause toxic effects in small mammals or birds.
Information on the relationship between soil lead and plants can be expanded somewhat
beyond the 1977 report, primarily due to a better understanding of the role of humic sub-
stances in binding lead. Although the situation is extremely complex, it is reasonable to
state that most plants cannot survive in soil containing 10,000 ug/g dry weight if the pH is
below 4.5 and the organic content is below 5 percent. The specifics of this statement are
discussed more extensively in Section 8.3.1.2.
Before 1977, natural levels of lead in environmental media other than soil were not well
known. Reports of sublethal effects of lead were sparse and there were few studies of total
ecosystem effects. Although several ecosystem studies have been completed since 1977 and many
problems have been overcome, it is still difficult to translate observed effects under speci-
fic conditions directly to predicted effects in ecosystems. Some of the known effects, which
are documented in detail in the appropriate sections, are summarized here.
8.1.1.1 Plants. The basic effect of lead on plants is to stunt growth. This may be through
a reduction of photosynthetic rate, inhibition of respiration, cell elongation, or root deve-
lopment, or premature senescence. Lead tolerance in ecotypes suggests some effects on popula-
tion genetics. All of these effects have been observed in isolated cells or in hydroponically-
grown plants in solutions comparable to 1 to 2 ug/g soil moisture. These concentrations are
well above those normally found in any ecosystem except near smelters or roadsides. Terres-
trial plants take up lead from the soil moisture and most of this lead is retained by the
roots. There is some evidence for foliar uptake of lead and little evidence that lead can be
translocated freely to the upper portions of the plant. Soil applications of calcium and
phosphorus may reduce the uptake of lead by roots.
8.1.1.2 Animals. Lead affects the central nervous system of animals and their ability to
synthesize red blood cells. Blood concentrations above 0.4 ppm (40 ug/dl) can cause observ-
able clinical symptoms in domestic animals. Calcium and phosphorus can reduce the intestinal
absorption of lead. The physiological effects of lead exposures in laboratory animals are
discussed in extensive detail in Chapters 10 and 12 of this document.
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8.1.1.3 Microorganisms. There is evidence that lead at environmental concentrations occa-
sionally found near roadsides and smelters [10,000 - 40,000 |jg/g dw (dry weight)] can elimi-
nate populations of bacteria and fungi on leaf surfaces and in soil. Many of those micro-
organisms play key roles in the decomposition food chain. It is likely that the affected
microbial populations are replaced by others of the same or different species, perhaps less
efficient at decomposing organic matter. There is also evidence that microorganisms can
mobilize lead by making it more soluble and more readily taken up by plants. This process
occurs when bacteria exude organic acids that lower the pH in the immediate vicinity of the
plant root.
8.1.1.4 Ecosystems. There are three known conditions under which lead may perturb ecosystem
processes. At soil concentrations of 1,000 ug/g or higher, delayed decomposition may result
from the elimination of a single population of decomposer microorganisms. Secondly, at con-
centrations of 500 - 1,000 ug/g, populations of plants, microorganisms, and invertebrates may
shift toward lead-tolerant populations of the same or different species. Finally, the normal
biogeochemical process that purifies and repurifies the calcium pool in grazing and decomposer
food chains may be circumvented by the addition of lead to vegetation and animal surfaces.
This third effect can be measured at all ambient atmospheric concentrations of lead.
Some additional effects may occur due to the uneven distribution of lead in ecosystems.
It is known that lead accumulates in soil, especially soil with high organic content.
Although no firm documentation exists, it is reasonable to assume the following from the known
chemistry of lead in soil: 1) other metals may be displaced from the binding sites on the
organic matter; 2) the chemical breakdown of inorganic soil fragments may be retarded by the
interference of lead with the action of fulvic acid on iron-bearing crystals; and 3) lead in
soil may be in equilibrium with moisture films surrounding soil particles and thus be avail-
able for uptake by plants.
To aid the reader in understanding the effects of lead on ecosystems, sections have been
included that discuss such important matters as how ecosystems are organized, what processes
regulate metal cycles, what criteria are valid in interpreting ecosystem effects, and how soil
systems function to regulate the controlled release of nutrients to plants. The informed
reader may wish to turn directly to Section 8.3, where the discussion of the effects of lead
on organisms begins.
8.1.2 Ecosystem Functions
8.1.2.1 Types of Ecosystems. Based on ambient concentrations of atmospheric lead and the dis-
tribution of lead in the soil profile, it is useful to distinguish among three types of eco-
systems: natural, cultivated, and urban. Natural ecosystems include aquatic and terrestrial
8-4
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ecosystems that are otherwise unperturbed by man, and those managed ecosystems, such as com-
mercial forests, grazing areas, and abandoned fields, where the soil profile has remained un-
disturbed for several decades. Cultivated ecosystems include those where the soil profile is
frequently disturbed and those where chemical fertilizers, weed killers, and pest-control
agents may be added. In urban ecosystems, a significant part of the exposed surface includes
rooftops, roadways, and parking lots from which runoff, if not channeled into municipal waste
processing plants, is spread over relatively small areas of soil surface. The ambient air
concentration of lead in urban ecosystems is 5-10 times higher than in natural or cultivated
ecosystems (See Chapter 7). Urban ecosystems may also be exposed to lead from other than
atmospheric sources, such as paint, discarded batteries, and used motor oil. The effects of
atmospheric lead depend on the type of ecosystems examined.
8.1.2.2 Energy Flow and Biogeochemical Cycles. To function properly, ecosystems require an
adequate supply of energy, which continually flows through the system, and an adequate supply
of nutrients, which for the most part, cycle within the ecosystem. There is evidence that
lead can interfere with both of these processes. Energy usually enters the ecosystem in the
form of sunlight and leaves as heat of respiration. Stored chemical energy may be transported
into or out of an ecosystem (e.g., leaf detritus in a stream) or be retained by the ecosystem
for long periods of time (e.g., tree trunks). Energy flow through an ecosystem may give
structure to the ecosystem by establishing food webs that efficiently regulate the transfer
of energy. Segments of these food webs are called food chains. Energy that flows along a
grazing food chain is diverted at each step to the detrital food chain.
Unlike energy, nutrient and non-nutrient elements are recycled by the ecosystem and
transferred from reservoir to reservoir in a pattern usually referred to as a biogeochemical
cycle (Brewer, 1979, p. 139). The reservoirs correspond approximately to the food webs of
energy flow. Although elements may enter (e.g., weathering of soil) or leave the ecosystem
(e.g., stream runoff), the greater fraction of the available mass of the element is usually
cycled within the ecosystem.
Two important characteristics of a reservoir are the amount of the element that may be
stored in the reservoir and the rate at which the element enters or leaves the reservoir.
Some reservoirs may contain a disproportionately large amount of a given element. For exam-
ple, most of the carbon in a forest is bound in the trunks and roots of trees, whereas most of
the calcium may be found in the soil (Smith, 1980, p. 316). Some large storage reservoirs,
such as soil, are not actively involved in the rapid exchange of the nutrient element, but
serve as a reserve source of the element through the slow exchange with a more active reser-
voir, such as soil moisture. When inputs exceed outputs, the size of the reservoir increases.
Increases of a single element may reflect instability of the ecosystem. If several elements
increase simultaneously, this expansion may reflect stable growth of the community.
8-5
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Reservoirs are connected by pathways that represent real ecosystem processes. Figure
8-1 depicts the biogeochemical reservoirs and pathways of a typical terrestrial ecosystem.
Most elements, especially those with no gaseous phase, do not undergo changes in oxidation
state and are equally available for exchange between any two reservoirs, provided a pathway
exists between the two reservoirs. The chemical environment of the reservoir may, however,
regulate the availability of an element by controlling solubility or binding strengths. This
condition is especially true for soils.
Ecosystems have boundaries. These boundaries may be as distinct as the border of a pond
or as arbitrary as an imaginary circle drawn on a map. Many trace metal studies are conducted
in watersheds where some of the boundaries are determined by topography. For atmospheric
inputs to terrestrial ecosystems, the boundary is usually defined as the surface of vegeta-
tion, exposed rock, or soil. The water surface suffices for aquatic ecosystems.
Non-nutrient elements differ little from nutrient elements in their biogeochemical cy-
cles. Quite often, the cycling patterns are similar to those of a major nutrient. In the
case of lead, the reservoirs and pathways are very similar to those of calcium.
There are three important questions concerning the effect of lead on ecosystems: Does
atmospheric lead interfere with the normal mechanisms of nutrient cycles? How does atmo-
spheric lead influence the normal lead cycle in an ecosystem? Can atmospheric lead interfere
with the normal flow of energy through an ecosystem?
8.1.2.3 Biogeochemistry of Lead. Naturally occurring lead from the earth's crust is commonly
found in soils and the atmosphere. Lead may enter an ecosystem by weathering of parent rock
or by deposition of atmospheric particles. This lead becomes a part of the nutrient medium of
plants and the diet of animals. All ecosystems receive lead from the atmosphere. More than
99 percent of the current atmospheric lead deposition is now due to human activities (National
Academy of Sciences, 1980). In addition, lead shot from ammunition may be found in many
waterways and popular hunting regions, leaded paint chips often occur in older urban regions,
and lead in fertilizer may contaminate the soil in agricultural regions.
In prehistoric times, the contribution of lead from weathering of soil was probably about
4 g Pb/ha-yr and from atmospheric deposition about 0.02 g Pb/ha-yr, based on estimates of
natural and anthropogenic emissions in Chapter 5 and deposition rates discussed in Chapter 6.
Weathering rates are presumed to have remained the same, but atmospheric inputs are believed
to have increased to 180 g/ha*yr in natural and some cultivated ecosystems, and 3,000 g/ha«yr
in urban ecosystems and along roadways (see Chapter 6). In every terrestrial ecosystem of the
Northern Hemisphere, atmospheric lead deposition now exceeds weathering by a factor of at
least 10, sometimes by as much as 1,000.
8-6
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GRAZERS
INORGANIC
NUTRIENTS
Figure 8-1. This figure depicts cycling processes within the major components of a terrestrial
ecosystem, i.e. primary producers, grazers and decomposers. Nutrient and non-nutrient
elements are stored in reservoirs within these components. Processes that take place within
reservoirs regulate the flow of elements between reservoirs along established pathways. The
rate of flow is in part a function of the concentration in the preceding reservoir. Lead
accumulates in decomposer reservoirs (DrD4) which have a high binding capacity for this
metal. When the flow of nutrients is reduced at I. II, or III. the rate of flow of inorganic
nutrients to primary producers is reduced.
Source: Adapted from Swift et at. (1979).
8-7
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Many of the effects of lead on plants, microorganisms, and ecosystems arise from the fact
that lead from atmospheric and weathering inputs is retained by soil. Geochemical studies
show that less than 3 percent of the inputs to a watershed leave by stream runoff (Siccama and
Smith, 1978; Shirahata et al., 1980). In prehistoric times, stream output nearly equalled
weathering inputs and the lead content of soil probably remained stable, accumulating at an
annual rate of less than 0.1 percent of the original natural lead (reviewed by Nriagu, 1978).
Due to human activity, lead in natural soils now accumulates on the surface at an annual rate
of 5 - 10 percent of the natural lead. One effect of cultivation is that atmospheric lead is
mixed to a greater depth than the 0 - 3 cm of natural soils.
Most of the effects on grazing vertebrates stem from the deposition of atmospheric parti-
cles on vegetation surfaces. Atmospheric deposition may occur by either of two mechanisms.
Wet deposition (precipitation scavenging through rainout or washout) generally transfers lead
directly to the soil. Dry deposition transfers particles to all exposed surfaces. Large
particles (>4 urn) are transferred by gravitational mechanisms; small particles (<0.5 urn) are
also deposited by wind-related mechanisms.
About half of the foliar dry deposition remains on leaf surfaces following normal rain-
fall (Elias et al., 1976; Peterson, 1978), but heavy rainfall may transfer the lead to other
portions of the plant (Elias and Croxdale, 1980). Koeppe (1981) has reviewed the literature
and concluded that less than 1 percent of the surface lead can pass directly into the internal
leaf tissues of higher plants. The cuticular layer of the leaves may be an effective barrier
to aerosol particles and even to metals in solution on the leaf surface (Arvik and Zimdahl,
1974), and passage through the stomata does not appear to account for a significant fraction
of the lead inside leaves (Carlson et al., 1976; 1977).
When particles attach to vegetation surfaces, transfer to soil is delayed from a few
months to several years. Due to this delay, large amounts of lead are diverted to grazing
food chains, bypassing the soil moisture and plant root reservoirs (Elias et al., 1982).
8.1.3 Criteria for Evaluating Ecosystem Effects
As it is the purpose of this chapter to describe the levels of atmospheric lead that may
produce adverse effects in plants, animals, and ecosystems, it is necessary to establish the
criteria for evaluating these effects. The first step is to determine the connection between
air concentration and ecosystem exposure. If the air concentration is known, ecosystem inputs
from the atmosphere can be predicted over time and under normal conditions. These inputs and
those from the weathering of soil determine the concentration of lead in the nutrient media of
plants, animals, and microorganisms. It follows that the concentration of lead in the nutri-
ent medium determines the concentration of lead in the organism and this in turn determines
the effects of lead on the organism.
8-8
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The fundamental nutrient medium of a terrestrial ecosystem is the soil moisture film
that surrounds organic and inorganic, soil particles. This film of water is in equilibrium
with other soil components and provides dissolved inorganic nutrients to plants. It is chemi-
cally different than ground water or rain water and there is little reliable information on
the relationship between lead in soil and lead in soil moisture. Thus, it appears impossible
to quantify all the steps by which atmospheric lead is transferred to plants. Until more
information is available on lead in soil moisture, another approach may be more productive.
This involves determining the degree of contamination of organisms by comparing the present
known concentrations with calculated prehistoric concentrations.
Prehistoric concentrations of lead have been calculated for only a few types of organ-
isms. However, the results are so low that any normal variation, even of an order of magni-
tude, would not seriously influence the calculation of the degree of contamination. The link
between lead in the prehistoric atmosphere and in prehistoric organisms may allow us to pre-
dict concentrations of lead in organisms based on present or future concentrations of
atmospheric lead.
It is reasonable to infer a relationship between degree of contamination and physio-
logical effect. It seems appropriate to assume that natural levels of lead that were safe
for organisms in prehistoric times would also be safe today. It is also reasonable that some
additional atmospheric lead can be tolerated by all populations of organisms with no ill
effects, that some populations are more tolerant than others, and that some individuals within
populations are more tolerant of lead effects than others.
For nutrient elements, the concept of tolerance is not new. The Law of Tolerance
(illustrated in Figure 8-2) states that any nutrient may be present at concentrations either
too low or too high for a given population and that the ecological success of a population is
greatest at some optimum concentration of the nutrient (Smith, 1980, p. 35). In a similar
manner, the principle applies to non-nutrient elements. Although there is no minimum concen-
tration below which the population cannot survive, there is a concentration above which the
success of the population will decline (point of initial response) and a concentration at
which the entire population will die (point of absolute toxicity). In this respect, both
nutrients and non-nutrients behave in a similar manner at concentrations above some optimum.
Certain variables make the points of initial response and absolute toxicity somewhat
imprecise. The point of initial response depends on the type of response investigated. This
response may be at the molecular, tissue, or organismic level, with the molecular response
occurring at the lowest concentration. Similarly, at the point of absolute toxicity, death
may occur instantly at high concentrations or over a prolonged period of time at somewhat
8-9
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MAXIMUM
NON-NUTRIENT
INITIAL
RESPONSE
O
_i
U
S
8
/
/
/ NUTRIENT
/
LOW
HIGH
CONCENTRATION OF ELEMENT
Figure 8-2. The ecological success of a population depends in part on the availability
of all nutrients at some optimum concentration. The dashed line of this diagram
depicts the rise and decline of ecological success (the ability of a population to grow,
survive and reproduce) over a wide concentration range of a single element. The
curve need not be symmetrically bell-shaped, but may be skewed to the right or left.
Although the range in concentration that permits maximum success may be much
wider than shown here, the important point is that at some high concentration, the
nutrient element becomes toxic. The tolerance of populations for high concentrations
of non-nutrients (solid line) is similar to that of nutrients, although there is not yet
any scientific basis for describing the exact shape of this portion of the curve.
Source: Adapted from Smith (1980).
8-10
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lower concentrations. Nevertheless, the gradient between these two points remains an appro-
priate basis on which to evaluate known environmental effects, and any information that
correctly positions this part of the tolerance curve will be of great value.
The normal parameters of a tolerance curve, i.e., concentration and ecological success,
can be replaced by degree of contamination and percent physiological dysfunction, respectively
(Figure 8-3). Use of this method of expressing degree of contamination should not imply that
natural levels are the only safe levels. It is likely that some degree of contamination can
be tolerated with no physiological effect.
I
*
100
ARBITRARY ZONE OF ASSUMED
SAFE CONCENTRATION
I 1
NATURAL /N
CONCENTRATION / v
INITIAL V X
RESPONSE v,
OBSERVED x.
DYSFUNCTION
K*-
-OEQREE OF CONTAMINATION -
N^
ABSOLUTE
x TOXICTT*
10
100
1.000
10.000
OBSERVED CONCJNATURAl CONC
Figure 8-3. Thfo figure attempt* to reconstruct the right portion of • tolerance curve, ttmnar to
Figure 8-2 but plotted on a temitog wale, lor e population wing a limited amount of information.
If the natural concentration It known for a population and H H It arbitrarily ataumed that lOx
natural concentration is also tafe. then the zone of attorned tafe concentration deflnet the
region.
Data reported by the National Academy of Sciences (1980) are used to determine the typi-
cal natural lead concentrations shown in various compartments of ecosystems in Table 8-1.
These data are from a variety of sources and are simplified to the most probable value within
the range reported by NAS. The actual prehistoric air concentration was probably near the low
end of the range (0.02-1.0 ng/m3), as present atmospheric concentrations of 0.3 ng/m3 in the
Southern Hemisphere and 0.07 ng/m3 at the South Pole (Chapter 5), would seem to preclude natu-
ral lead values higher than this.
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TABLE 8-1. ESTIMATED NATURAL LEVELS OF LEAD IN ECOSYSTEMS
Component Range Best estimate
Air 0.01-1.0 ng/m3 0.07
Soil
Inorganic 5-25 M9/9 12-°
Organic 1 ng/g 1.0
Soil moisture 0.0002 M9/9 0.0002
Plant leaves 0.01-0.1 ug/g dw 0.05
Herbivore bones 0.04-0.12 ug/g dw °-12
Carnivore bones 0.01-0.03 ng/g dw 0.03
Source: Ranges are from the National Academy of Sciences (1980); best estimates are
discussed in the text. Units for best estimates are the same as for ranges.
In prehistoric times, the rate of entry of lead into the nutrient pool available to
plants was predominantly determined by the rate of weathering of inorganic minerals in frag-
ments of parent rock material. Geochemical estimates of denudation and adsorption rates
(Chapter 6) suggest a median value of 12 ug/g as the average natural lead content of total
soil, with the concentration in the organic fraction at approximately 1 ug/g.
Studies have shown the lead content of leafy vegetation to be 90 percent anthropogenic,
even in remote areas (Crump and Barlow, 1980; Elias et al., 1976, 1978). The natural lead
content of nuts and fruits may be somewhat higher than leafy vegetation, based on internal
lead concentrations of modern samples (Elias et al., 1982). The natural lead concentrations
of herbivore and carnivore bones were reported by Elias et al. (Elias and Patterson, 1980;
Elias et al., 1982). These estimates are based on predicted Pb/Ca ratios calculated from the
observed biopurification of calcium reservoirs with respect to Sr, Ba, and Pb, on the system-
atic evaluation of anthropogenic lead inputs to the food chain (Section 8.5.3), and on
measurements of prehistoric mammalian bones.
8.2 LEAD IN SOILS AND SEDIMENTS
8.2.1 Distribution of Lead in Soils
Because lead in soil is the source of most effects on plants, microorganisms, and eco-
systems, it is important to understand the processes that control the accumulation of lead in
soil. The major components of soil are the following: 1) fragments of inorganic parent rock
8-12
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material; 2) secondary inorganic minerals; 3) organic constituents, primarily humic sub-
stances, which are residues of decomposition or products of decomposer organisms; 4) Fe-Mn
oxide films, which coat the surfaces of all soil particles and appear to have a high binding
capacity for metals; 5) soil microorganisms, most commonly bacteria and fungi, although
protozoa and soil algae may also be found; and 6) soil moisture, the thin film of water sur-
rounding soil particles that is the nutrient medium of plants. Some watershed studies con-
sider that fragments of inorganic parent rock material lie outside the forest ecosystem,
because transfer from this compartment is so slow that much of the material remains inert for
centuries.
The concentration of natural lead ranges from 5 to 30 ug/g in the top 5 cm of most soils
not adjacent to ore bodies, where natural lead may reach 800 ug/g. Aside from surface deposi-
tion of atmospheric particles, plants in North America average about 0.5-1 ug/g dw
(Peterson, 1978) and animals roughly 2 ug/g (Forbes and Sanderson, 1978). Thus, soils contain
the greater part of total ecosystem lead. In soils, lead in parent rock fragments is tightly
bound within the crystalline structures of the inorganic soil minerals. It is released to the
ecosystem only by surface contact with soil moisture films.
The evidence for atmospheric inputs of lead to soil rests mainly with the accumulation of
lead in the soil profile. There are several reports that lead accumulates in the upper
layers, usually about 2-5 cm, of the soil, just below the litter layer. This is the soil
layer that is usually highest in organic content. Many soils develop by podzolization, char-
acterized by distinct soil horizons caused by the separation and segretation of organic and
inorganic compounds, including metal salts and metal-organic complexes. Siccama et al.,
(1980), and Friedland et al. (1984a, 1984b), found that lead in the forest floors (the litter
layer above the mineral soil) of New England have increased during the 1960-1980 at about the
same rate that atmospheric concentrations of lead increased. Friedland et al. (1984a, 1984b)
found that copper, zinc, and nickel also increased over the same time period, as did the total
organic content. They concluded that lead and perhaps other metals may have inhibited
decomposition.
Soils adjacent to smelters may be contaminated at a distance of several kilometers away
from the source and to a depth of ten or more centimeters. Hogan and Wotton (1984) found
elevated concentrations of lead at a distance of 38 kilometers from a Cu-Zn smelter on the
surface of the soil, and up to six kilometers at a depth of fifteen centimeters. McNeilly
et al. (1984) reported an exponential decrease in lead concentrations of surface soil from 0
to 75 meters for mine spoils. Effects of the spoils were detectable even at a depth of 20
centimeters.
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Hutchinson (1980) has reviewed the effects of acid precipitation on the ability of soils
to retain cations. Excess calcium and other metals are leached from the A horizon of soils by
rain with a pH more acidic than 4.5. Most soils in the eastern United States are normally
acidic (pH 3.5 to 5.2) and the leaching process is a part of the complex equilibrium main-
tained in the soil system. By increasing the leaching rate, acid rain can reduce the availa-
bility of nutrient metals to organisms dependent on the top layer of soil. Tyler (1978)
reports the effect of acid rain on the leaching rate (reported as residence time) for lead and
other metals. Simulated rain of pH 4.2 to 2.8 showed the leaching rate for lead increases with
decreasing pH, but not nearly as much as that of other metals, especially Cu, Mn, and Zn.
This would be as expected from the high stability constant of lead relative to other metals in
humic acids (see Section 6.5.1). It appears from this limited information that acidification
of soil may increase the rate of removal of lead from the soil, but not before several major
nutrients are removed first. The effect of acid rain on the retention of lead by soil mois-
ture is not known.
8.2.2 Origin and Availability of Lead in Aquatic Sediments
Atmospheric lead may enter aquatic ecosystems by wet or dry deposition (Dolske and
Sievering, 1979) or by the erosional transport of soil particles (Baier and Healy, 1977). In
waters not polluted by industrial, agricultural, or municipal effluents, the lead concentra-
tion is usually less than 1 ug/1. Of this amount, approximately 0.02 M9/1 is natural lead and
the rest is anthropogenic lead, probably of atmospheric origin (Patterson, 1980). Surface
waters mixed with urban effluents may frequently reach lead concentrations of 50 ug/1, and
occasionally higher (Bradford, 1977).
In aqueous solution, virtually all lead is divalent, as tetravalent lead can exist only
under extremely oxidizing conditions (reviewed by Rickard and Nriagu, 1978; Chapter 3). At pH
higher than 5, divalent lead can form a number of hydroxyl complexes, most commonly PbOH+,
Pb(OH)2, and Pb(OH)3 . At pH lower than 5, lead exists in solution as hydrated Pb. In still
water, lead is removed from the water column by the settling of lead-containing particulate
matter, by the formation of insoluble complexes, or by the adsorption of lead onto suspended
organic particles. The rate of sedimentation is determined by temperature, pH, oxidation-
reduction potential, ionic competition, the chemical form of lead in water, and certain bio-
logical activities (Jenne and Luoma, 1977). McNurney et al. (1977) found 14 ug Pb/g in stream
sediments draining cultivated areas and 400 ug/g in sediments associated with urban eco-
systems. Small sediment grain size and high organic content contributed to increased reten-
tion in sediments.
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8.3 EFFECTS OF LEAD ON PLANTS
8.3.1 Effects on Vascular Plants and Algae
Some physiological and biochemical effects of lead on vascular plants have been detected
under laboratory conditions at concentrations higher than normally found in the environment.
The commonly reported effects are the inhibition of photosynthesis, respiration, or cell elon-
gation, all of which reduce the growth of the plant (Koeppe, 1981). Lead may also induce
premature senescence, which may affect the long-term survival of the plant or the ecological
success of the plant population. To provide a meaningful evaluation of these effects, it is
necessary to examine the correlation between laboratory conditions and typical conditions in
nature with respect to form, concentration, and availability of lead. First, the reader must
understand what is known of the movement of lead from soil to the root to the stem and finally
to the leaf or flower. Most notably, there are specific barriers to lead at the soil:soil
moisture interface and at the root:shoot interface that retard the movement of lead and reduce
the impact of lead on photosynthetic and meristematic (growth and reproduction) tissue.
8.3.1.1 Uptake by Plants. Most of the lead in or on a plant occurs on the surfaces of leaves
and the trunk or stem. The surface concentration of lead in trees, shrubs, and grasses
exceeds the internal concentration by a factor of at least five (Elias et al., 1978). Foliar
uptake was believed to account for less than 1 percent of the uptake by roots (Arvik and
Zimdahl, 1974; reviewed by Koeppe, 1981; Zimdahl, 1976). Krause and Kaiser (1977) were able
to show foliar uptake and translocation of lead mixed with cadmium, copper, and manganese
oxides when applied in extremely large amounts (122 mg/m2) directly to leaves. This would be
comparable to 100,000 days accumulation at a remote site (0.12 ng/cm2'd) (Elias et al., 1978).
However, recent isotopic evidence by Facchetti and Geiss (1982) and Patterson (1982) and mass
balance interpretations from watershed data (Lindberg and Harriss, 1981) suggest that lead can
be absorbed across the leaf surface into internal plant tissues. Nevertheless, the major
effect of surface lead at ambient concentrations seems to be on subsequent components of the
grazing food chain (Section 8.4.1) and on the decomposer food chain following litterfall
(Elias et al., 1982). (See also Section 8.4.2.)
In the soil, the availability of metals to plants is generally controlled by the concen-
tration and form of the metal, which are in turn influenced by such soil forming processes as
gleying, leaching, podzolization, and the accumulation of organic matter at the surface.
Other factors such as pH and the presence of other cations may also be important. The amount
of lead that enters plants by this route is determined by the availability of lead in soil,
with apparent variations according to plant species. Soil cation exchange capacity, a major
factor, is determined by the relative size of the clay and organic fractions, soil pH, and the
amount of Fe-Mn oxide films present (Nriagu, 1978). Of these, organic humus and high soil pH
8-15
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are the dominant factors in immobilizing lead (see Section 6.5.1). Under natural conditions,
most of the total lead in soil would be tightly bound within the crystalline structure of
inorganic soil fragments, unavailable to soil moisture. Available lead, bound on clays,
organic colloids, and Fe-Mn films, would be controlled by the slow release of bound lead from
inorganic rock sources. Since before 3000 B.C., atmospheric lead inputs through litter decom-
position have increased the pool of available lead bound on organic matter within the soil
reservoir (see Section 5.1).
Because lead is strongly immobilized by humic substances, only a small fraction (perhaps
0.01 percent in soils with 20 percent organic matter, pH 5.5) is released to soil moisture
(see Section 6.5.1). In soil moisture, lead may pass along the pathway of water and nutrient
uptake on either a cellular route through the cell membranes of root hairs (symplastic route)
or an extracellular route between epidermal cells into the intercellular spaces of the root
cortex (apoplastic route) (Foy et al., 1978). Lead probably passes into the symplast by mem-
brane transport mechanisms similar to the uptake of calcium or other bivalent cations.
In soils with lead concentrations within the range of natural lead (15-30 ug/g), only
trace amounts of lead are absorbed by plants. The amount absorbed increases when the concen-
tration of lead in soil increases or when the binding capacity of soil for lead decreases.
Uptake by root systems does not necessarily mean the lead reaches the stems, leaves or fruits.
Rather, the process should be seen as a soil-plant continuum that strongly favors retention of
lead by the soil and the root system.
When viewed from the perspective of the uptake of nutrients such as calcium, there are at
least three mechanisms whereby lead can be taken up by roots: transpirational mass flow,
diffusion, and active transport (Jenny, 1980). Probably the most significant is transpira-
tional mass flow. In the process of absorbing and transporting water from the soil to the
leaves, the plant absorbs relatively large amounts of ions in solution. Since plants take up
about 100 times their weight in water each growing season, this process could account for
twice the normal amounts of lead found in vegetation, assuming equilibrium between the soil
and the soil solution. For example, a lettuce plant transpired 100 liters of water during the
season, which contained 5000 ug of lead (Rabinowitz, 1972). But the plant itself contained
only 2500 ug, most of which was in the roots.
Diffusion can occur along a concentration gradient whenever the transpiration stream is
idle, e.g., during the night or during periods of high humidity. Because the concentration of
lead in the soil solution is usually higher than in the plant, and because lead bound on the
cellulose matrix of the cell wall would not effect the concentration gradient, the flow of
lead would probably be toward the root. Although the third mechanism, active transport, is an
important process for nutrient elements, there is no evidence that such a process occurs for
8-16
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lead or any other non-nutrient. This process requires energy, and it is unreasonable that a
plant would expend energy to take up non-nutrient elements.
The soil-root continuum is a complex structure that consists of the soil particles, the
soil solution, the mutigel or other remnants of root exudates, the epidermal cells with elon-
gated root hairs, and the root cortial cells. The walls of the epidermal cells are a loose
matrix of cellulose and hemicellulose fibers. Much of this continuum is of biological origin
and contains compounds active in ion exchange, such as hemicelluloses and pectic substances
that are heavily endowed with -COOH groups, and proteins that also have charged groups. As a
cation moves from the soil particle to the root cortex, whether by mass flow or diffusion, it
is continually proximate to root structures with a high binding capacity. Lead is more tight-
ly bound at these sites than other cations, even calcium. Consequently, relatively little
lead passes through the roots into the shoot. It appears that most of the soil lead is
retained within the root. However, some plants may allow more lead to translocate than
others. Rabinowitz (1972) found that for lettuce and wild oats growing in soils with in-
creasing lead concentrations, the lettuce translocated very little soil lead but the wild oats
translocated proportionately greater amounts. The author was able to distinguish isotopically
between soil and atmospheric lead, and found also that more than half the lead in plants,
after water washing, was of atmospheric origin when the plants were grown at 30 meters from a
freeway.
At 500 ug Pb/g nutrient solution, lead has been shown to accumulate in the cell walls of
germinating Raphinus sativus roots (Lane and Martin, 1982). This concentration is much higher
than that found by Wong and Bradshaw (1982) to cause inhibition of germinating root elongation
(less than 2.5 ug/g), absence of root growth (5 |jg/g), or 55 percent inhibition of seed ger-
mination (20-40 ug/g) in the rye grass, Coliurn perenne. Lane and Martin (1982) also observed
lead in cytoplasmic organelles that because of their osmiophillic properties, appeared to have
a storage function. It was suggested that the organelles eventually emptied their contents
into the tonoplast.
The accumulation of lead in cell walls and cytoplasmic bodies has also been observed in
blue-green algae by Jensen et al. (1982), who used X-ray energy dispersive analysis in con-
junction with scanning electron microscopy to observe high concentrations of lead and other
metals in these single-celled procaryotic organisms. They found the lead concentrated in the
third of the four-layered cell wall and in polyphosphate bodies (not organelles, since they
are not membrane-bound) which appeared to be a storage site for essential metals. The nutri-
ent solution contained 100 pg Pb/g. The same group (Rachlin et al., 1982) reported morpholo-
gical changes in the same blue-green alga (Plectonema boryanum). There was a significant
increase in cell size caused by the lead, which indicated that the cell was able to detoxify
its cytoplasm by excreting lead with innocuous cell wall material.
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It appears that two defensive mechanisms may exist in the roots of plants for removing
lead from the stream of nutrients flowing to the above-ground portions of plants: lead may be
deposited with cell wall material exterior to the individual root cells, or may be sequestered
in organelles within the root cells. Any lead not captured by these mechanisms would likely
move with nutrient metals cell-to-cell through the symplast and into the vascular system.
Uptake of lead by plants may be enhanced by symbiotic associations between plant roots
and mycorrhizal fungi. The three primary factors that control the uptake of nutrients by
plants are the surface area of the roots, the ability of the root to absorb particular ions,
and the transfer of ions through the soil. The symbiotic relationship between mycorrhizal
fungi and the roots of higher plants can increase the uptake of nutrients by enhancing all
three of these factors (Voigt, 1969). The typical ectomycorrhiza consists of a mantle or
sheath of mycelia that completely surrounds the root. The physical extension of the sheath
may increase the volume of the root two to three times (Voigt, 1969). Mycorrhizal roots often
show greater affinities for nutrients than do uninfected roots of the same species grown in
the same conditions. In many soil systems, where the bulk of the nutrients are bound up in
parent rock material, efficient uptake of these nutrients by plants depends on the ability of
organisms in the rhizosphere (plant roots, soil fungi, and bacteria) to increase the rates of
weathering. Mycorrhizal fungi are known to produce and secrete into their environment many
different acidic compounds (e.g., malic and oxalic acids). In addition, mycorrhizal roots
have been shown to release more carbon dioxide into the rhizosphere than do non-mycorrhizal
roots as a result of their increased rates of respiration. Carbon dioxide readily combines
with soil moisture to produce carbonic acid. All of these acids are capable of increasing the
weathering rates of soil particles such as clays, and altering the binding capacity of organic
material, thereby increasing the amount of nutrients and other cations in the soil solution.
Mycorrhizae are known to enhance the uptake of zinc by pine roots (Bowen et al., 1974), and it
is likely that lead uptake is similarly increased, by inference to the ability of mycorrhizae
to enhance the uptake of calcium by pine roots (Melin and Nilsson, 1955; Melin et al., 1958).
The translocation of lead to aboveground portions of the plant is not clearly understood.
Lead may follow the same pathway and be subject to the same controls as a nutrient metal such
as calcium. This assumption implies that the plant root has no means of discriminating
against lead during the uptake process, and it is not known that any such discrimination
mechanism exists. There may be several mechanisms, however, that excrete lead back out of the
root or that prevent its translocation to other plant parts. The primary mechanisms may be
storage in cell organelles or adsorption on cell walls. The apoplast contains an important
supply of plant nutrients, including water. Lead in the apoplast remains external to the
cells and cannot pass to vascular tissue without at least passing through the cell membranes
8-18
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of the endodermis. Because this extracellular region is bounded on all sides by cell walls,
the surface of which is composed of layers of cellulose strands, the surface area of the
apoplast is comparable to a sponge. It is likely that much of the lead in roots is adsorbed
to the apoplast surface. Dictyosomes (cytoplasmic organelles that contain cell wall
material) may carry lead from inside the cell through the membrane to become a part of the
external cell wall (Malone et al., 1974), possibly replacing calcium in calcium pectate. Lead
may also be stored and excreted as lead phosphate in dictyosome vesicles (Malone et al.,
1974). Nevertheless, some lead does pass into the vascular tissue, along with water and
dissolved nutrients, and is carried to physiologically active tissue of the plant.
Evidence that lead in contaminated soils can enter the vascular system of plants and be
transported to above-ground parts may be found in the analysis of tree rings. Rolfe (1974)
found fourfold increases in both rural and urban trees, comparing 10-year increments of annual
rings for the period 1910-20 to annual rings of the period 1963-73. Symeonides (1979) found a
twofold increase from 1907-17 to 1967-77 in trees at a high-lead site, with no increase in
trees from a low-lead site. Baes and Ragsdale (1981), using only ring porous species, found
significant post-1930 increases in Quercus and Carya with high lead exposure, but only in
Carya with low-lead exposure. These chronological records confirm that lead can be translo-
cated from roots to the upper portions of the plant and that the amounts translocated are in
proportion to the concentrations of lead in soil.
8.3.1.2 Physiological Effects on Plants. Because most of the physiologically active tissue of
plants is involved in growth, maintenance, and photosynthesis, it is expected that lead might
interfere with one or more of these processes. Indeed, such interferences have been observed
under optimal growth conditions in laboratory experiments at lead concentrations greater than
those normally found in the field, except near smelters or mines (Koeppe, 1981). It is likely
that because these are the physiological processes studied more vigorously than others, more
is known of these effects. Studies of lead effects on other plant processes, especially main-
tenance, flowering, and hormone development, have not been conducted and no conclusion can be
reached concerning these processes.
Inhibition of photosynthesis by lead may be by direct interference with the light reac-
tion or the indirect interference with carbohydrate synthesis. With 21 M9 pb/g reaction solu-
tion, Miles et al. (1972) demonstrated substantial inhibition of photosystem II near the site
of water splitting, a biochemical process believed to require manganese. Homer et al. (1979)
found a second effect on photosystem II at slightly higher concentrations of lead. This
effect was similar to that of DCMU [3-(3,4-dichlorophenyl)-l,l-dimethylurea], a reagent com-
monly used to uncouple the photosynthetic electron transport system. Bazzaz and Govindjee
(1974) suggested that the mechanism of lead inhibition was a change in the conformation of the
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thylakoid membranes, separating and isolating pigment systems I and II. Wong and Govindjee
(1976) found that lead also interferes with P700 photooxidation and re-reduction, a part of
the photosystem I light reaction. Homer et al. (1981) found a lead tolerant population of the
grass Phalaris arundinacea had lowered the ratio of chlorophyll a/chlorophyll b, believed to
be a compensation for photosystem II inhibition. There was no change in the total amount of
chlorophyll, but the mechanism of inhibition was considered different from that of Miles et
al. (1972). Hampp and Lendzian (1974) found that lead chloride inhibits the synthesis of
chlorophyll b more than that of chlorophyll a at concentrations up to 100 mg Pb/g. Devi
Prasad and Devi Prasad (1982) found 10 percent inhibition of pigment production in three spe-
cies of green algae at 1 |jg/g, increasing to 50 percent inhibition at 3 ug/g. Bazzaz et al.
(1974, 1975) observed reduced net photosynthesis which may have been caused indirectly by
inhibition of carbohydrate synthesis. Without carbohydrates, stomatal guard cells remain
flaccid, transpiration ceases, carbon dioxide fixation decreases, and further carbohydrate
synthesis is inhibited.
In the quantification of growth inhibition, one can measure either the concentration of
lead in the nutrient medium or in the tissue that is growth inhibited. Lead concentrations in
the nutrient medium relate directly to the degree of environmental contamination, but the more
precise measurement is in the tissue, since there would be a more direct correlation between
the lead concentration and the physiological processes inhibited. Burton et al. (1983) deter-
mined that when tissue concentrations in the shoots of Sitka-spruce seedlings exceeded about
20 pg Pb/g dw, growth inhibition became significant, and lethal at about 40 ug/g. This narrow
range between the onset of inhibition and lethality was attributed to the sequestering of lead
in the roots and shoots up to 19 ug/g, above which any additional lead would be more available
and extremely toxic. The stunting of plant growth may be by the inhibition of the growth
hormone IAA (indole-3-ylacetic acid). Lane et al. (1978) found a 25 percent reduction in
elongation at 10 ug/g lead as lead nitrate in the nutrient medium of wheat coleoptiles. This
effect could be reversed with the addition of calcium at 18 ug/g. Lead may also interfere
with plant growth by reducing respiration or inhibiting cell division. Miller and Koeppe
(1971) and Miller et al. (1975) showed succinate oxidation inhibition in isolated mitochondria
as well as stimulation of exogenous NADH oxidation with related mitochondrial swelling.
Hassett et al. (1976), Koeppe (1977), and Malone et al. (1978) described significant inhibi-
tion of lateral root initiation in corn. Inhibition increased with the simultaneous addition
of cadmium.
Sung and Yang (1979) found that lead at 1 ug/g can complex with and inactivate ATPase to
reduce the production and utilization of ATP in kidney bean (Phaseolus vulgaris) and buckwheat
leaves (Fagopyrum esculentum). The lead was added hydroponically at concentrations up to
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1,000 ug/g. Kidney bean ATPase showed a continued response from 1 to 1,000 ug/g, but buck-
wheat leaves showed little further reduction after 10 ug/g. Neither extracted ATP nor chemi-
cally added ATP could be used by the treated plants. Lee et al. (1976) found a 50 percent
increase in the activity of several enzymes related to the onset of senescence in soybean
leaves when lead was added hydroponically at 20 ug/g. These enzymes were acid phosphatase,
peroxidase, and alpha-amylase. A build-up of ammonia was observed along with a reduction in
nitrate, calcium, and phosphorus. Glutamine synthetase activity was also reduced by 65 per-
cent. Continued increases in effects were observed up to 100 ug/g, including a build-up of
soluble protein. Paivoke (1979) also observed a 60 percent increase in acid phosphatase acti-
vity during the first 6 days of pea seedling germination (Pisum sativum) at 2 ug/g, under low
nutrient conditions. The accumulation of soluble protein was observed and the effect could be
reversed with the addition of nutrients, including calcium.
Scarponi and Perucci (1984) reported that lead can interfere with the synthesis of ALA-
dehydratase in corn, but does not appear to affect the activity of this enzyme. This enzyme
catalyzes the conversion of 6-aminolevulinic acid to porphobilinogen, an intermediate in
chlorophyll synthesis. The concentration of lead was above 10,000 ug/1 in the nutrient solu-
tion.
The interaction of lead with calcium has been shown by several authors, most recently by
Garland and Wilkins (1981), who demonstrated that calcium could partially overcome the effects
of lead on growth in barley seedlings (Hordeum vulgare). Seedlings that were growth-inhibited
at 2 ug Pb/g sol. with no added calcium, grew at about half the control rate with 17 ug Ca/g
sol. This relation persisted up to 25 ug Pb/g sol. and 500 ug Ca/g sol.
Chaney and Strickland (1984) measured the effects of lead on the germination on red pine
pollen. Following exposure in an aqueous nutrient medium, two parameters were measured:
pollen germination and germ tube elongation. Pollen germination was inhibited by greater than
10 percent only at relatively high concentrations of lead, about 1,000 ug/1, but the most
significant effect was shown for germ tube elongation, which showed 10 percent inhibition at
about 150 ug/1.
These studies of the physiological effects of lead on plants all show some effect at
concentrations from 2 to 10 ug/g in the nutrient medium of hydroponically-grown agricultural
plants. It is probable that no effects would have been observed at these concentrations had
the lead solutions been added to normal soil, where the lead would have been bound by humic
substances. There is no firm relationship between soil lead and soil moisture lead, because
each soil type has a unique capacity to retain lead and to release that lead to the soil
moisture film surrounding the soil particle. Once in soil moisture, lead seems to pass freely
to the plant root according to the capacity of the plant root to absorb water and dissolved
substances (Koeppe, 1981).
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Chapter 6 discusses the many parameters controlling the release of lead from soil to soil
moisture, but so few data are available on observed lead concentrations in soil moisture that
no model can be formed. It seems reasonable that there may be a direct correlation between
lead in hydroponic media and lead in soil moisture. Hydroponic media typically have an excess
of essential nutrients, including calcium and phosphorus, so that movement of lead from hydro-
ponic media to plant root would be equal to or slower than movement from soil moisture to
plant root. Hughes (1981) adopted the general conclusion that extractable soil lead is typi-
cally 10 percent of total soil lead. However, this lead was extracted chemically under lab-
oratory conditions more rigorous than the natural equilibrium between soil and soil moisture.
Ten percent should therefore be considered the upper limit, where the ability of soil to
retain lead is at a minimum. A lower limit of 0.01 percent is based on the only known report
of lead in both soil and soil moisture (16 ug/g soil, 1.4 ng/g soil moisture; Elias et al.,
1982). This single value shows neither trends with different soil concentrations nor the soil
component (organic or inorganic) that provides the lead to the soil moisture. But the number
(0.01 percent) is a conservative estimate of the ability of soil to retain lead, since the
conditions (pH, organic content) were optimum for retaining lead. A further complication is
that atmospheric lead is retained at the surface (0-2 cm) of the soil profile (Martin and
Coughtrey, 1981), whereas most reports of lead in soil pertain to samples from 0 to 10 cm as
the "upper" layer of soil. Any plant that absorbs solely from the top few centimeters of soil
obviously is exposed to more lead than one with roots penetrating to a depth of 25 cm or more.
Agricultural practices that cultivate soil to a depth of 25 cm blend in the upper layers with
lower to create a soil with average lead content somewhat above background.
These observations lead to the general conclusion that even under the best of conditions
where soil has the highest capacity to retain lead, most plants would experience reduced
growth rate (inhibition of photosynthesis, respiration, or cell elongation) in soils of 10,000
ug Pb/g or greater. Khan and Frankland (1983) observed stunted growth in radish plants at
1000 ug pb/9 soil when the lead was added as chloride, with complete growth inhibition at
5000 pg/g. The effects were less severe when lead oxide was added to the soil. Concentra-
tions approaching these values typically occur around smelters (Martin and Coughtrey, 1981)
and near major highways (Wheeler and Rolfe, 1979). These conclusions pertain to soil with the
ideal composition and pH to retain the maximum amount of lead. Acid soils or soils lacking
organic matter would inhibit plants at much lower lead concentrations.
The rate at which atmospheric lead accumulates in soil varies from 1.1 mg/m2-yr average
global deposition (Table 6-6) to 3,000 mg/m2-yr near a smelter (Patterson et al., 1975).
Assuming an average density of 1.5 g/cm3, undisturbed soil to a depth of 2 cm (20,000 cm3/m2)
would incur an increase in lead concentration at a rate of 0.04 to 100 ug/g soil-yr. This
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means remote or rural area soils may never reach the 10,000 ug/g threshold but that undis-
turbed soils closer to major sources may be within range in the next 50 years.
8.3.1.3 Lead Tolerance in Vascular Plants. Some plant species have developed populations
tolerant to high-lead soils (Antonovics et al., 1971). In addition to Homer et al. (1981)
cited above, Jowett (1964) found populations of Agrostis tenuis in pure stands on acidic spoil
banks near an abandoned mine. The exclusion of other species was attributed to root inhibi-
tion. Populations of A. tenuis from low-lead soils had no tolerance for the high-lead soils.
Several other studies suggest that similar responses may occur in populations growing in
lead-rich soils (reviewed in Peterson, 1978). A few have suggested that crops may be culti-
vated for their resistance to high-lead soils (Gerakis et al., 1980; John, 1977).
Using populations taken from mine waste and uncontaminated control areas, some authors
have quantified the degree of tolerance of Agrostis tenuis (Karataglis, 1982) and Festuca
rubra (Wong, 1982) under controlled laboratory conditions. Root elongation was used as the
index of tolerance. At 36 ug Pb/g nutrient solution, all populations of A. tenuis were com-
pletely inhibited. At 12 ug Pb/g, the control populations from low-lead soils were completely
inhibited, but the populations from mine soils achieved 30 percent of their normal growth
(growth at no lead in nutrient solution). At 6 ug/g, the control populations achieved 10 per-
cent of their normal growth; tolerant populations achieved 42 percent. There were no measure-
ments below 6 M9/9- Wong (1982) measured the index of tolerance at one concentration only,
2.5 ug Pb/g nutrient solution, and found that non-adapted populations of Festuca rubra that
had grown on soils with 47 ug/g total lead content were completely inhibited, populations from
soils with 350 - 650 ug/g achieved 3-7 percent of normal growth, and populations from 5,000
ug/g soil achieved nearly 40 percent of normal growth. Tolerance indices should be used with
caution because they depend on two measurements that may be genetically independent.
Humphreys and Nicholls (1984) suggested that different genes regulated root elongation in a
control solution and in the heavy-metal solution.
These studies support the conclusion that inhibition of plant growth begins at a lead
concentration of less than 1 ug/g soil moisture and becomes completely inhibitory at a level
between 3 and 10 ug/g. Plant populations that are genetically adapted to high-lead soils may
achieve 50 percent of their normal root growth at lead concentrations above 3 ug/g. These
experiments did not show the effect of reduced root growth on total productivity, but they did
show that exposure to high-lead soils is a requirement for genetic adaptation and that, at
least in the case of F. rubra. plant lead concentrations increase with increasing concentra-
tions in the soil.
There are a few plants known to be hyperaccumulators of metals (Reeves and Brooks, 1983).
These plants appear to show no adverse effects even when their tissue concentrations reach
8-23
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1000 ug/g dry weight. About 100 species of plants are known to hyperaccumulate nickel, fif-
teen each for copper and cobalt. Reeves and Brooks (1983) describe two species that hyper-
accumulate lead and mention three others reported in the literature. The fact that many of
these species belong to the genus Alyssum suggests a genetic mechanism of metal tolerance.
8.3.1.4 Effects of Lead on Forage Crops. In the 1977 Air Quality Criteria Document for Lead
(U.S. Environmental Protection Agency, 1977), there was a general awareness that most of the
lead in plants was surface lead from the atmosphere. Most studies since then have addressed
the problem of distinguishing between surface and internal plant lead. The general conclusion
is that, even in farmlands remote from major highways or industrial sources, 90 - 99 percent
of the total plant lead is of anthropogenic origin (National Academy of Sciences, 1980).
Obviously, the critical agricultural problem concerns forage crops and leafy vegetables. In
Great Britain, Crump and Barlow (1982) determined that, within 50 m of the highway, surface
deposition is the major source of lead in forage vegetation. Beyond this range, seasonal
effects can obscure the relative contribution of atmospheric lead. The atmospheric deposition
rate appears to be much greater in the winter than in the summer. Two factors may explain
this difference. First, deposition rate is a function of air concentration, particle size
distribution, windspeed, and surface roughness. Of these, only particle size distribution is
likely to be independent of seasonal effects. Lower windspeeds or air concentration during
the summer could account for lower deposition rates. Second, it may be that the deposition
rate only appears to change during the summer. With an increase in biomass and a greater
turnover in biomass, the effective surface area increases and the rate of deposition, which is
a function of surface area, decreases. During the winter, lead may not build up on the sur-
face of leaves as it does in summer, even though the flux per unit of ground area may be the
same.
8.3.1.5 Effects on Algae. Sicko-Goad and Lazinsky (1981) have presented cytological evidence
that lead can be incorporated into polyphosphate bodies in some algal species (Diatoma tenue
var. elongatum, Scenedesmus sp.), presumably as a tolerance mechanism. They also report the
immobilization of lead in cell vacuoles. At high concentrations (207 ug/g), Roderer (1984a)
found deformations of cell organelles, especially nuclei and mitochondria, and increased
autolytic activity in the chrysophyte Poterioochromonas malhamensis. a unicellular alga. In
the same study, organolead compounds, TriEL and TEL were found to cause an increase in number
and size of nuclei, contractile vacuoles, chloroplasts and dictyosomes, as well as a marked
accumulation of lipid droplets and lysosomes. The concentrations for these effects were 10 uM
TriEL and 100 uM TEL. Similar results were reported in a review of the toxic effects of
organolead compounds by Roderer (1984b).
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8.3.1.6 Summary of Plant Effects. When soil conditions allow lead concentrations in soil
moisture to exceed 2-10 ug/g, most plants experience reduced growth due to the inhibition of
one or more physiological processes. Excess calcium or phosphorus may reverse the effect.
Plants that absorb nutrients from deeper soil layers may receive less lead. Acid rain is not
likely to release more lead until after major nutrients have been depleted from the soil. A
few species of plants have the genetic capability to adapt to high lead soils.
8.3.2 Effects on Bacteria and Fungi
Wood and Wang (1983) discuss possible mechanisms for microbial resistance to metals,
noting that some metals (e.g., Al, Pb, Sn, Be) occur at crustal abundances greater than other
metals known to be required nutrients. Abundance alone is not a sufficient condition for the
evolution of a nutritive requirement. A second condition is solubility in anaerobic condi-
tions. Except at low pH, aluminum, lead, and tin are insoluble in an anaerobic solution and
would not have been available to primitive microorganisms during the early stages of their
evolution.
8.3.2.1 Effects on Decomposers. Tyler (1972) explained three ways in which lead might inter-
fere with the normal decomposition processes in a terrestrial ecosystem. Lead may be toxic to
specific groups of decomposers, it may deactivate enzymes excreted by decomposers to break
down organic matter, or it may bind with the organic matter to render it resistant to the
action of decomposers. Because lead in litter may selectively inhibit decomposition by soil
bacteria at 2,000 - 5,000 ug/g (Smith, 1981, p. 160), forest floor nutrient cycling processes
may be seriously disturbed near lead smelters (Bisessar, 1982; Watson et al., 1976). This is
especially important because approximately 70 percent of plant biomass enters the decomposer
food chain (Swift et al., 1979, p. 6). If decomposition of the biomass is inhibited, then
much of the energy and nutrients remain unavailable to subsequent components of the food
chain. There is also the possibility that the ability of soil to retain lead would be re-
duced, as humic substances are byproducts of bacterial decomposition.
Babich et al. (1983) introduced the concept of ecological dose as it applies to the
effects of metals on ecological processes in soil. The inhibition of microbe-mediated pro-
cesses can be used to quantify the effects of environmental pollutants on natural ecosystems.
The ecological dose 50 percent (EcD50) is the concentration of a toxicant that inhibits a
microbe-mediated ecological process by 50 percent. Since microbes are an integral part of the
biogeochemical cycling of elements and the flow of energy through an ecosystem, they are an
important indicator of the productivity of the ecosystem. This concept is superior to the
lethal dose (LD) concept because it is based on an assemblage of heterogeneous populations
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that are important to the ecosystem and that might be comparable to similar population assem-
blages of other ecosystems. The LD concept relies on the elimination of a single population
that may be insignificant to the ecosystem or not comparable to other ecosystems.
Using published data, Babich et al. (1983) determined that the EcD50 for nitrification
inhibition was 100 ug/g as soluble lead extracted from soil, based on the data of Chang and
Broadbent (1982). The data of Doelman and Haanstra (1979) suggested an EcD50 for inhibition
of respiration ranging from 0 to 7,500 pg/g total lead in soil, depending on the soil type.
Peat soils showed no inhibition, sandy soils showed the most.
During decomposition, plant tissues are reduced to resistant particulate matter, as solu-
ble organic and inorganic compounds are removed by the chemical action of soil moisture and
the biochemical action of microorganisms (Odum and Drifmeyer, 1978). Each group of micro-
organisms specializes in the breakdown of a particular type of organic molecule. Residual
waste products of one group become the food for the next group. Swift et al. (1979, p. 101)
explained this relationship as a cascade effect with the following generalized pattern (Figure
8-4). Organisms capable of penetrating hard or chemically resistant plant tissue are the
primary decomposers. These saprotrophs, some of which are fungi and bacteria that reside on
leaf surfaces at the initial stages of senescence, produce a wide range of extracellular
enzymes. Others may reside in the intestinal tract of millipedes, beetle larvae, and termites
capable of mashing plant tissue into small fragments. The feces and remains of this group and
the residual plant tissue are consumed by secondary decomposers, i.e., the coprophilic fungi,
bacteria, and invertebrates (including protozoa) specialized for consuming bacteria. These
are followed by tertiary decomposers. Microorganisms usually excrete enzymes that carry out
this digestive process external to their cells. They are often protected by a thick cell
coat, usually a polysaccharide. Because they are interdependent, the absence of one group in
this sequence seriously affects the success of subsequent groups, as well as the rate at which
plant tissue decomposes. Each group may be affected in a different way and at different lead
concentrations. Lead concentrations toxic to decomposer microbes may be as low as 1 - 5 ug/g
or as high as 5,000 ug/g (Ooelman, 1978).
Crist et al. (1985) found no inhibition due to lead during the early stages of deciduous
leaf decomposition. Green leaves were ground to a compost and innoculated with microbes from
the same location. Loss of biomass was about 30 percent after 18 weeks for the controls and
all lead concentrations (0 to 1,000 ug/g) of lead added as lead sulfate. The sulfate salt was
considered the most common form of lead available to the decomposing leaves in the natural
system. No intermediate biomass measurements were made, however. In another study, Doelman
and Haanstra (1984) observed an initial inhibition of decomposition, measured by soil respira-
tion, during the first eight weeks, followed by nearly complete recovery by about 70 weeks.
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This effect was greatest for sandy and sandy loam soils, somewhat mediated in clay and sandy
peat soils and virtually nonexistent in silty loam soils. No effects were observed below
1,000 ug/g. In this case, lead was added as the chloride salt.
Some studies have measured the effects of lead on specific decomposition enzymes or sub-
strates. Haanstra and Doelman (1984) reported 50 percent inhibition (doubling of decomposi-
tion time) of glutamic acid decomposition in sandy soils at 3,500 ug/g. There was a small but
distinct effect in clay soils and no effect in a calcareous silty loam soil. Frankenburger
and Tabutabai (1985) measured a 5 percent inhibition of free soil amidase at about 1,000 ug/g
soil. Bacterial amidase was inhibited 30 percent at about 800 mg Pb/1 substrate/enzyme
system.
Under conditions of mild contamination, the loss of one sensitive bacterial population
may result in its replacement by a more lead-tolerant strain. Inman and Parker (1978) found
that litter transplanted from a low-lead to a high-lead site decayed more slowly than high-
lead litter, suggesting the presence of a lead-sensitive microorganism at the low-lead site.
When high-lead litter was transplanted to the low-lead site, decomposition proceeded at a rate
faster than the low-lead litter at the low-lead site. In fact, the rate was faster than the
high-lead litter at the high-lead site, suggesting even the lead-tolerant strains were some-
what inhibited. The long-term effect is a change in the species composition of the ecosystem,
which will be considered in greater detail in Section 8.5.3.
Delayed decomposition has been reported near smelters (Jackson and Watson, 1977), mine
waste dumps (Williams et al., 1977), and roadsides (Inman and Parker, 1978). This delay is
generally in the breakdown of litter from the first stage (00 to the second (02), with intact
plant leaves and twigs accumulating at the soil surface. The substrate concentrations at
which lead inhibits decomposition appear to be very low. Williams et al. (1977) found inhibi-
tion in 50 percent of the bacterial and fungal strains at 50 ug Pb/ml nutrient solution. The
community response time for introducing lead-tolerant populations seems very fast, however.
Doelman and Haanstra (1979a,b) found lead-tolerant strains had replaced non-tolerant bacteria
within three years of lead exposure. These new bacteria were predominately thick-coated
gram-negative strains and their effectiveness in replacing lead-sensitive strains was not
evaluated in terms of soil decomposition rates.
Tyler (1982) has also shown that many species of wood-decaying fungi do not accumulate
Pb, Ca, Sr, or Mn as strongly as they do other metals, even the normally toxic metal, cadmium.
Accumulation was expressed as the ratio of the metal concentration in the fungus to its sub-
strate. A ratio of greater than one implies accumulation, less than one, exclusion. Of 11
species, manganese was excluded by ten, strontium by nine, lead by eight, and calcium by
seven. Potassium, at the other end of the spectrum, was not excluded by any species. The
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RAW
DETRITUS
0,
GROUP I
GROUP 11
-68-
GROUP III
INORGANIC
NUTRIENTS
Figure 8-4. Within the decomposer food chain, detritus is progressively broken down in
a sequence of steps regulated by specific groups of decomposers. Because of the cascade
effect of this process, the elimination of any decomposer interrupts the supply of organic
nutrients to subsequent groups and reduces the recycling of inorganic nutrients to plants.
Undecomposed litter would accumulate at the stages preceding the affected decomposer.
Source: Adapted from Swift et al. (1979).
species which appeared to accumulate calcium and lead were described as having harder, less
ephemeral tissues.
This relationship among calcium, strontium, and lead is consistent with the phenomenon of
biopurification described in Section 8.5.2. From the data of Tyler (1982) it appears that
some of the species of fungi receive lead from a source other than the nutrient medium, per-
haps by direct atmospheric deposition.
8.3.2.2 Effects on Nitrifying Bacteria. The conversion of ammonia to nitrate in soil is a
two-step process mediated by two genera of bacteria, Nitrosomonas and Nitrobacter. Nitrate is
required by all plants, although some maintain a symbiotic relationship with nitrogen-fixing
bacteria as an alternate source of nitrogen. Those that do not would be affected by a loss
of free-living nitrifying bacteria, and it is known that many trace metals inhibit this nitri-
fying process (Liang and Tabatabai, 1977,1978). Lead is the least of these, inhibiting nitri-
fication 14 percent at concentrations of 1,000 ug/g soil. Many metals, even the nutrient
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metals, manganese and iron, show greater inhibition at comparable molar concentrations.
Nevertheless, soils with environmental concentrations above 1,000 ug Pb/g are frequently
found. Even a 14 percent inhibition of nitrification can reduce the potential success of a
plant population, as nitrate is usually the limiting nutrient in terrestrial ecosystems. In
cultivated ecosystems, nitrification inhibition is not a problem if nitrate fertilizer is
added to soil, but could reduce the effectiveness of ammonia fertilizer if the crops rely on
nitrifying bacteria for conversion to nitrates. Rother et al. (1983) found that lead concen-
trations as high as 30,000 ug/g soil did not affect symbiotic nitrogen fixation in white
clover (Trifolium repens).
8.3.2.3 Methylation by Aquatic Microorganisms. While methyl lead is not a primary form of
environmental lead, methylation greatly increases the toxicity of lead to aquatic organisms
(Wong and Chau, 1979; Thayer and Brinckman, 1982). There is some uncertainty about whether
the mechanism of methylation is biotic or abiotic. Some reports (Wong and Chau, 1979;
Thompson and Crerar, 1980) conclude that lead in sediments can be methylated by bacteria.
Reisinger et al. (1981) report that biomethylation of lead under aerobic or anaerobic condi-
tions does not occur and such reports are probably due to sulfide-induced chemical conversion
of organic lead salts. These authors generally agree that tetramethyl lead can be formed
under environmental conditions when another tetravalent organolead compound is available, but
methylation of divalent lead salts such as Pb(N03)2 does not appear to be significant. Jarvie
et al. (1983) also report that they were unable to produce any definite evidence for bio-
methylation of lead.
8.3.2.4 Summary of Effects on Microorganisms. It appears that microorganisms are more sen-
sitive than plants to soil lead pollution and that changes in the composition of bacterial
populations may be an early indication of lead effects. Delayed decomposition may occur at
750 ug Pb/g soil and nitrification inhibition at 1,000 ug/g. Many of the environmental vari-
ables that can raise or lower these estimates are not yet known. In certain chemical en-
vironments, the highly toxic tetramethyllead can be formed, but this process does not appear
to be mediated by aquatic microorganisms.
8.4 EFFECTS OF LEAD ON DOMESTIC AND WILD ANIMALS
8.4.1 Vertebrates
8.4.1.1 Terrestrial Vertebrates. Forbes and Sanderson (1978) have reviewed reports of lead
toxicity in domestic and wild animals. Lethal toxicity can usually be traced to consumption
of lead battery casings, lead-based paints, oil wastes, putty, linoleum, pesticides, lead
shot, or forage near smelters. Except for lead shot ingestion, these problems can be solved
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by proper management of domestic animals. However, the 3,000 tons of lead shot falling annually
along waterways and other hunting grounds continues to be a problem.
A single pellet of lead shot weighs about 110 mg, and 70 percent of this may be eroded in
ringed turtle dove gizzards over a period of 14 days (Kendall et al., 1982). Their data
showed an immediate elevation of blood lead and reduction of aminolevulinic acid dehydrogenase
(ALA-D) activity within one day of swallowing two pellets. Feierabend (1983) reviewed 97
reports on the effects of lead shot on waterfowl. Of the estimated 80 to 125 million water-
fowl in North America, 1.5 to 38 million die each year from lead poisoning. Many more are
greatly impaired by chronic sublethal exposure. Reichell et al. (1984) reported that 17 of
293 bald eagles sampled had lead concentrations in their liver high enough to suspect lead
poisoning. The 293 specimens were found dead or nearly dead during 1978 to 1983. The most
common causes of death were trauma from being hit by a motor vehicle (20 percent) and shooting
(19 percent). Bjorn et al., (1982) also reported the uptake of lead shot by grazing cattle
near a trapshooting site.
Bull et al. (1983) and Osborn et al. (1983) reported extensive bird mortality that could
be attributed to alkyl lead pollution of the Mersey Estuary in the United Kingdom. Bull
et al. (1983) found 3-18 ug/g alkyl lead in dead birds, 1-14 pg/g in sick birds, and 0.3-
1.2 ug/g in apparently healthy birds. Osborn et al. (1983), in laboratory studies, found that
2000 jjg/day alkyl lead in the diet caused heavy mortality and 200 ug/day caused tremors,
impaired balance, and feeding irregularities, although no mortality was observed. Tissue con-
centrations of alkyl lead at the lower dose were in the range of 0.2 to 5.4 ug/g. The authors
concluded that many of the apparently healthy wild birds were experiencing symptoms likely to
impair their chances for survival.
Awareness of the routes of uptake is important in interpreting the exposure and accumula-
tion in vertebrates. Inhalation rarely accounts for more than 10 - 15 percent of the daily
intake of lead (National Academy of Sciences, 1980). Much of the inhaled lead is trapped on
the walls of the bronchial tubes and passes to the stomach embedded in swallowed mucus.
Because lead concentrations in lakes or running stream water are quite low, intake from drink-
ing water may also be insignificant unless the animal drinks from a stagnant or otherwise con-
taminated source.
Food is the largest contributor of lead to animals. The type of food an herbivore eats
determines the rate of lead ingestion. More than 90 percent of the total lead in leaves and
bark may be due to surface deposition, but relatively little surface deposition may be found
on some fruits, berries, and seeds that have short exposure times. Roots intrinsically have
no surface deposition. Similarly, ingestion of lead by a carnivore depends mostly on deposi-
tion on herbivore fur and somewhat less on lead in herbivore tissue. Harrison and Dyer (1984)
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estimated that mule deer grazing in the Rocky Mountain National Park would exceed acceptable
lead exposure by grazing on roadside vegetation for just 1 to 2 percent of the time. This
estimate was based on the assumption that the upper limit of exposure should be 3,000 ug
Pb/day. Mule deer grazing on non-roadside forage would consume about 1,500 ug/day.
The type of food eaten is a major determinant of lead body burdens in small mammals.
Goldsmith and Scanlon (1977) and Scanlon (1979) measured higher lead concentrations in insect-
ivorous species than in herbivorous species, confirming the earlier work of Quarles et al.
(1974), which showed body burdens of granivores < herbivores < insectivores, and Jeffries and
French (1972) that granivores < herbivores. Animals in these studies were analyzed whole
minus the digestive tract. It is likely that observed diet-related differences were somewhat
diluted by including fur in the analysis, because fur lead might be similar for small mammals
from the same habitats with different feeding habits.
Since 1977, there has been a trend away from whole body analyses toward analyses of iso-
lated tissues, especially bones and blood. Bone concentrations of lead are better than blood
as indicators of long-term exposure. Because natural levels of blood lead are not well known
for animals and blood is not a good indicator of chronic exposure, blood lead is poorly suited
for estimating total body burdens. One experiment with sheep shows the rapid response of
blood to changes in lead ingestion and the relative contribution of food and air to the total
blood level. Ward et al. (1978) analyzed the blood in sheep grazing near a highway (0.9 ug/g
ml) and in an uncontaminated area (0.2 ug/ml). When sheep from the uncontaminated area were
allowed to graze near the roadway, their blood levels rose rapidly (within 1 day) to about 3.0
ug/ml, then decreased to 2.0 Mg/m1 during the next 2 days, remaining constant for the remain-
der of the 14-day period. Sheep from the contaminated area were moved to the uncontaminated
area, where upon their blood dropped to 0.5 ug/ml in 10 days and decreased to 0.3 ug/ml during
the next 180 days. Sheep in the uncontaminated area that were fed forage from the roadside
experienced an increase in blood lead from 0.2 to 1.1 ug/ml in 9 days. Conversely, sheep from
the uncontaminated area moved to the roadside but fed forage only from the uncontaminated site
experienced an increase from 0.2 to 0.5 ug/ml in 4 days. These data show that both air and
food contribute to lead in blood and that blood lead concentrations are a function of both the
recent history of lead exposure and the long-term storage of lead in bone tissue.
Beyer et al. (1985) reported a decrease in red blood cell ALA-D activity for 14 small
mammals and 15 songbirds in a habitat near a smelter. There were no changes in packed cell
volume or hemoglobin concentrations and little evidence of gross or microscopic lesions that
could be attributed to metal poisoning. Intranuclear inclusion bodies were found in a kidney
of one shrew. The soil concentrations at the surface were 1,200-2,700 ug/g. Foliage concen-
trations were 21 ug/g and the fruits and berries averaged 4 ug/g, a typical pattern for the
8-31
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distribution of lead of atmospheric origin. The authors attributed the relatively minor
effects of lead on the animals to the fact that the mice, shrews, and songbirds were eating
primarily fruits and berries, not leaves. They also considered the possibility that popula-
tions of some species had been previously reduced or eliminated by the emissions from the
smelter.
Chmiel and Harrison (1981) showed that, for small mammals, the highest concentrations of
lead occurred in the bones (Table 8-2), with kidney and liver concentrations somewhat less.
They also showed greater bone concentrations in insectivores than herbivores, both at the
control and contaminated sites. Clark (1979) found lead concentrations in shrews, voles, and
brown bats from roadside habitats near Washington, D.C., to be higher than any previously
reported. His estimates of dosages (7.4 mg Pb/kg-day) exceed those that normally cause mor-
tality or reproductive impairment in domestic mammals (1.5-9 mg Pb/g-day) (Hammond and
Aronson, 1964; James et al., 1966; Kelliher et al., 1973). Traffic density was the same as
reported by Chmiel and Harrison (1981), nearly twice that of Goldsmith and Scanlon (1977)
(Table 8-2). The body lead burden of shrews exceeded mice, which exceeded voles. Beresford
et al. (1981) found higher lead in box turtles within 500 m of a lead smelter than in those
from control sites. Bone lead exceeded kidney and liver lead as in small mammals.
Kusseberth et al. (1984) reported that lead in the bones of small mammals indigenous to a
habitat near a battery reclamation plant decreased exponentially with distance from the
battery plant. The observed pattern was similar to that reported for lead in roadside soils
and vegetation reported in Section 7.2.2.1.1. They also reported findings of intranuclear
inclusions in renal tubular epithelial tissue in one vole and four deer mice.
There are few studies reporting lead in vertebrate tissues from remote sites. Elias et
al. (1976, 1982) reported tissue concentrations in voles, shrews, chipmunks, tree squirrels,
and pine martens from the remote High Sierra. Bone concentrations were generally only 2 per-
cent of those reported from roadside studies and 10 percent of the controls of roadside
studies (Table 8-2), indicating the roadside controls were themselves contaminated to a large
degree. Furthermore, biogeochemical calculations suggest that even animals in remote areas
have bone lead concentrations 50 to 500 times natural background levels. The natural concen-
tration of lead in the bones of herbivores is about 0.04 ng/g dry weight (Table 8-1). This
value may vary regionally with geochemical anomalies in crustal rock, but provides a reason-
able indicator of contamination. Natural levels of lead in carnivore bone tissue should be
somewhat lower, with omnivores generally in between (Elias and Patterson, 1980; Elias et al.,
1982).
Table 8-2 shows the results of several studies of small animal bone tissue. To convert
reported values to a common basis, assumptions were made of the average water content, calcium
8-32
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TABLE 8-2. ESTIMATES OF THE DEGREE OF CONTAMINATION OF HERBIVORES,
OMNIVORES, AND CARNIVORES
Data are based on published concentrations of lead in bone tissue (corrected to dry weight as
indicated). Degree of contamination is calculated as observed/natural Pb. Natural lead con-
centrations are from Table 8-1. Concentrations are in pg Pb/g dw.
Organism
Herbivores
Vole-roadside
Vole-roadside
-control
Vole-orchard
-control
Vole- remote
Deer mouse- roadside
-control
Deer mouse- near battery
plant
-control
Deer mouse- roadside
-control
Deer mouse- roadside
-control
Mouse- roadside
-control
Mouse- roadside
-control
Averaae herbivore
roadside (7)
control (7)
remote (2)
Omni vores/f rugi vores
Woodmouse- roads i de
-control
Compos i te- roads i de
-control
Chipmunk-remote
Tree squirrel -remote
Feral pigeon-urban
-rural
Feral pigeon-urban
-suburan
-rural
Bone
Pb cone.
38
17
5
73
9
2
25
5.7
80
2
29
7.2
52
5
19
9.3
109
18
41
8.5
2
67
25
22
3
2
1.3
670
5.7
250
33
12
Ref.
1
2
2
5
5
11
2
2
13
13
3
3
4
4
2
2
2
2
1
1
7
7
1
11
6
6
12
12
12
Estimated degree of
contamination
bone
320
140
42
610
75
17
210
48
650
18
240
60
430
42
160
78
910
150
340
71
17
840
310
280
37
25
16
8400
71
3100
410
150
(continued)
8-33
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TABLE 8-2. (continued)
Estimated degree of
Bone
Organism Pb cone.
Star ling- roadside
-control
Rob in- roadside
-control
Sparrow-roadside
-control
Blackbird- roadside
-control
Grackle-roadside
-control
Rats-roadside
-control
Average omnivore
roadside (7)
urban (1)
control (7)
remote (2)
Carnivores
Box turtle-smelter
-control
Egret-rural
Gull-rural
Mink- rural
Shrew- roadside
-control
Shrew- roadside
-control
Shrew- remote
Pine marten-remote
Average carnivore
roadside (3)
smelter (1)
rural (2)
control (4)
remote (2)
aDry weight calculated from published
1. Chmiel and Harrison, 1981
2. Getz et al. , 1977b
3. Welch and Dick, 1975
4. Mierau and Favara, 1975
5. Elfving et al., 1978
6. Mutton and Goodman, 1980
7. Getz et al . , 1977a
210
13
130
41
130
17
90
7
63
22a
310a
15a
102
670
18
1.7
91" a
V
12a
lla
1.5
67
12
193
41
4.6
1.4
190
91
11
18
3
fresh weights
contamination
Ref.
7
7
7
7
7
7
7
7
7
7
9
9
8
8
10
10
14
2
2
1
1
1
11
assuming 35 percent
8. Beresford et
9. Mouw et al.,
10. Hulse et al. ,
11. Elias et al. ,
12. Johnson et al
13. Kisseberth et
14. Ogle et al. ,
bone
2600
160
1600
510
1600
200
1100
88
790
280
10000
500
1260
8400
230
21
3000
190
400
370
50
2200
400
6400
1400
150
47
6200
3000
385
620
99
water.
al . , 1981
1975
1980
1982
. , 1982b
al. , 1984
1985
8-34
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concentration, and average crustal concentration. Because ranges of natural concentrations of
lead in bones, plants, soils, and air are known with reasonable certainty (Table 8-1), it is
possible to estimate the degree of contamination of vertebrates from a wide range of habitats.
It is important to recognize that these are merely estimates that do not allow for possible
errors in analysis or anomalies in regional crustal abundances of lead.
8.4.1.2 Effects on Aquatic Vertebrates. Lowe et al. (1985) reported the results of a nation-
wide survey of metal concentrations in freshwater fish during the period 1979 to 1981. At 112
monitoring stations they found an average (geometric mean) of 0.19 ug Pb/g wet weight for the
period 1978-79 and 0.17 pg/g for 1980-81. Several laboratories have reported experiments that
measure the effects of lead on freshwater fish. Two requirements limit the evaluation of
literature reports of lead effects on aquatic organisms. First, any laboratory study should
incorporate the entire life cycle of the organism studied. It is clear that certain stages of
a life cycle are more vulnerable than others (Hodson, 1979, Hodson et al., 1979). For fish,
the egg or fry is usually most sensitive. Secondly, the same index must be used to compare
results. Christensen et al. (1977) proposed three indices useful for identifying the effects
of lead on organisms. A molecular index reports the maximum concentration of lead causing no
significant biochemical change; residue index is the maximum concentration showing no continu-
ing increase of deposition in tissue; and a bioassay i ndex is the maximum concentration
causing no mortality, growth change, or physical deformity. These indices are comparable to
those of physiological dysfunction (molecular, tissue, and organismic) discussed in Section
8.1.3.
From the standpoint of environmental protection, the most useful index is the molecular
index. This index is comparable to the point of initial response discussed previously and is
equivalent to the "safe concentration" originally described by the U.S. Environmental Protec-
tion Agency (Battelle, 1971) as being the concentration that permits normal reproduction,
growth, and all other life-processes of all organisms. It is unfortunate that very few of the
toxicity studies in the aquatic literature report safe concentrations as defined above.
Nearly all report levels at which some or all of the organisms die.
Hematological and neurological responses are the most commonly reported effects of ex-
tended lead exposures in aquatic vertebrates. Hematological effects include the disabling and
destruction of mature red blood cells and the inhibition of the enzyme ALA-D required for
hemoglobin synthesis. At low exposures, fish compensate by forming additional red blood
cells. These red blood cells often do not reach maturity. At higher exposures, the fish
become anemic. Symptoms of neurological responses are difficult to detect at low exposure,
but higher exposure can induce neuromuscular distortion, anorexia, and muscle tremors. Spinal
curvature eventually occurs with time or increased concentration (Hodson 1979; Hodson et al.,
8-35
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1977). Weis and Weis (1982) found spinal curvature in developing eggs of killifish when the
embryos had been exposed to 10 (jg Pb/ml during the first 7 days after fertilization. All
batches showed some measure of curvature, but those that were most resistant to lead were
least resistant to the effects of methylmercury. Sippel et al. (1983) reported that black fin
and spinal curvature in rainbow trout were the most reliable clinical tests for lead toxicity
at low levels. These effects appear at about 120 ug/1 before effects on red blood cells,
liver function, or histopathological indications in the liver, spleen, kidneys, gills, brain,
spinal cord, or gastrointestinal tract.
The biochemical changes used by Christensen et al. (1977) to determine the molecular
index for brook trout were 1) increases in plasma sodium and chloride and 2) decreases in
glutamic oxalacetic transaminase activity and hemoglobin. They observed effects at 0.5 ug/1,
which is 20-fold less than the lower range (10 |jg/l) suggested by Wong et al. (1978) to cause
significant detrimental effects. Hodson et al. (1978a) found tissue accumulation and blood
parameter changes in rainbow trout at 13 ug/1. This was the lowest experimental level, and
only slightly above the controls, which averaged 4 H9/1- They concluded, however, that
because spinal curvature does not occur until exposures reach 120 pg/1, rainbow trout are ade-
quately protected at 25 ug/1.
Aside from the biochemical responses discussed by Christensen et al. (1977), the lowest
reported exposure concentration that causes hematological or neurological effects is 8 ug/1
(Hodson, 1979). Christensen's group dealt with subcellular responses, whereas Hodson's group
dealt primarily with responses at the cellular or higher level. Hodson et al. (1978a) also
reported that lead in food is not available for assimilation by fish, that most of their lead
comes from water, and that decreasing the pH of water (as in acid rain) increases the uptake
of lead by fish (Hodson et al., 1978b). Patrick and Loutit (1978), however, reported that
tissue lead in fish reflects the lead in food if the fish are exposed to the food for more
than a few days. Hodson et al. (1980) also reported that, although the symptoms are similar
(spinal deformation), lead toxicity and ascorbic acid deficiency are not metabolically
related.
8.4.2 Invertebrates
Insects have lead concentrations that correspond to those found in their habitat and
diet. Herbivorous invertebrates have lower concentrations than do predatory types (Wade et
al., 1980). Among the herbivorous groups, sucking insects have lower lead concentrations than
chewing insects, especially in regions near roadsides, where more lead is found on the sur-
faces of vegetation. Williamson and Evans (1972) found gradients away from roadsides are not
the same as with vertebrates, in that invertebrate lead decreases more slowly than vertebrate
8-36
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lead relative to decreases in soil lead. They also found great differences between major
groups of invertebrates. Wood lice in the same habitat, eating the same food, had eight times
more lead than millipedes.
There are a few isolated reports on the effects of lead on the physiology of insects.
Hopkin and Martin (1984) fed hepatopancreas tissue from the woodlouse Qm'scus asellus L. to
centipedes (Lithobius variegatus) and found that the lead was not assimilated by the centipede
but passed directly through the midgut within four days of consumption. The centipedes were
fed 1 to 16 ug Pb. Bengtsson et al. (1983) observed delayed growth in populations of
Onychiurus armatus (Tullb.), a soil insect that feeds on detritus and microorganisms. The
insects were fed a diet of fungi that had been grown on media from 0 to 150 ug Pb/g and had
accumulated mycelial concentrations of 8 to 3100 ug/g in direct proportion to media concentra-
tions. The Fl and F2 generations experienced a marked decline in growth rate, measured as
length versus age, but eventually acheived the same maximum length as the controls. Lead was
stored for the first two weeks during the life cycle, then excreted. The reduction in growth
rate, or delay in achieving maximum length, was seen to be significant to the reproductive
process because the length at first egg-laying appeared to be relatively constant at 0.90 to
0.97 mm. This evidence suggests that reduced growth rate might be accompanied by delayed
sexual maturity.
The distribution of lead among terrestrial gastropod tissues was reported by Ireland
(1979). He found little difference among the foot, skin, mantle, digestive gland, gonad, and
intestine. There are no reports of lead toxicity in soil invertebrates. In a feeding experi-
ment, however, Coughtrey et al. (1980) found decreased tolerance for lead by microorganisms
from the guts of insects at 800 ug Pb/g food. Many roadside soils fall in this range.
In Cepaea hortensis, a terrestrial snail, Williamson (1979) found most of the lead in the
digestive gland and gonadal tissue. He also determined that these snails can lose 93 percent
of their whole body lead burden in 20 days when fed a low-lead diet in the laboratory. Since
no analyses of the shell were reported, elimination of lead from this tissue cannot be evalu-
ated. A continuation of the study (Williamson, 1980) showed that body weight, age, and day-
length influenced the lead concentrations in soft tissues.
Beeby and Eaves (1983) addressed the question of whether uptake of lead in the garden
snail, Helix aspersa, is related to the nutrient requirement for calcium during shell forma-
tion and reproductive activity. They found concentrations of both metals were strongly corre-
lated with changes in dry weight and little evidence for correlation of lead with calcium
independent of weight gain or loss. Lead in the diet remained constant.
Gish and Christensen (1973) found lead in whole earthworms to be correlated with soil
lead, with little rejection of lead by earthworms. Consequently, animals feeding on earth-
worms from high-lead soils might receive toxic amounts of lead in their diets, although there
8-37
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was no evidence of toxic effects on the earthworms (Ireland, 1977). Ash and Lee (1980)
cleared the digestive tracts of earthworms and still found direct correlation of lead in
earthworms with soil lead; in this case, soil lead was inferred from fecal analyses. These
authors found differences among species of earthworms. Ireland and Richards (1977) also found
species differences in earthworms, as well as some localization of lead in subcellular organ-
el les of chloragogue and intestinal tissue. In view of the fact that chloragocytes are be-
lieved to be involved with waste storage and glycogen synthesis, the authors concluded that
this tissue is used to sequester lead in the manner of vertebrate livers. Species differences
in whole body lead concentrations could not be attributed to selective feeding or differential
absorption, unless the differential absorption occurs only at elevated lead concentrations.
The authors suggested that the two species have different maximum tolerances for body lead but
gave no indication of physiological dysfunction when the maximum tolerance was reached. In
soils with a total lead concentration of 1,800 ug/g dry weight (Ireland, 1975), Lumbricus
rubellus had a whole body concentration of 3,600 ug/g, while Dendrobaene rubida accumulated
7,600 ug/g in the same location (Ireland and Richards, 1977). Because this difference was not
observed at the control site (15 ug/g soil), it can be assumed that at some soil concentration
between 15 and 1,800 ug/g, different species of earthworms begin to accumulate different
amounts of lead. The authors concluded that D. rubida can simply tolerate higher tissue lead
concentrations, implying that soil concentrations of 1,800 ug/g are toxic to L. rube11 us.
This concentration would be considerably lower than soil lead concentrations that cause
effects in plants, and similar to that which can affect soil microorganisms. Ma et al. (1983)
found that the uptake of lead by populations of earthworms near a zinc smelter complex was
related to soil pH and organic content. In the observed range of 3.5 to 6.1, low soil pH
increased the accumulation of lead by L rubellus. Likewise, for the range of 2.2 to 8.6
percent organic matter, earthworms accumulated more lead when exposed to soil at the lower end
of the range. Kruse and Barrett (1985) measured greater lead concentrations in cleared earth-
worms from sludge-treated soils. The sludge amended soil was 1.5 times the lead content of
the control soil, and the corresponding earthworms were about 3.5 times higher.
Aquatic insects appear to be resistant to high levels of lead in water. To be conclu-
sive, toxicity studies must observe invertebrates through an entire life cycle, although this
is infrequently done. Anderson et al. (1980) found LC50's for eggs and larvae of Tanytarsus
dissimilis, a chironomid, to be 260 ug/1. This value is 13 - 250 times lower than previously
reported by Warnick and Bell (1969), Rehwoldt et al. (1973), and Nehring (1976). However,
Spehar et al. (1978) found that mature amphipods (Gamitiarus pseudolimnaeus) responded nega-
tively to lead at 32 ug/1. Fraser et al. (1978) found that adult populations of a freshwater
isopod (Asellus aquaticus) have apparently developed a genetic tolerance for lead in river
sediments.
8-38
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Newman and Mclntosh (1982) investigated freshwater gastropods, both grazing and burrow-
ing. Lead concentrations in the grazers {Physa integra, Pseudosuccinea columella, and Helisoma
trivolvis) were more closely correlated with water concentrations than with lead in the food.
Lead in the burrowing species, Campeloma decisum, was not correlated with any environmental
factor. These authors (Newman and Mclntosh, 1983) also reported that both Physa integra and
Campeloma decisum are able to eliminate lead from their soft tissue when transferred to a
low-lead medium, but that tissue lead stabilized at a level higher than found in populations
living permanently in the low-lead environment. This would seem to indicate the presence of a
persistent reservoir of lead in the soft tissues of these gastropods. Tessier et al. (1984)
measured metal accumulation in the tissues of the freshwater bivalve Elliptio complanata and
concluded that concentrations of lead in the bivalve were directly related to concentrations
of lead in that fraction of the sediment that can be most easily extracted. The highest
concentrations of lead were in the gills, mantle, and hepatopancreatic tissue. They concluded
further that lead may enter the organism through the gills more so than through the digestive
tract, and that the presence of amorphous iron oxyhydroxides reduces the concentration of
metals in the bivalve tissues by selectively competing for the binding sites.
Everard and Denny (1984) observed that freshwater snails ( Lymnaea peregra) accumulate
lead in their digestive glands, feet, and shells when fed a diet enriched with lead. These
snails are efficient grazers of Aufwuchs, the epiflora and epifauna that coat all submerged
surfaces of the euphotic zone. Granular bodies, thought to be precipitated lead phosphate,
were observed in the gut epithelium, gut lumen, digestive gland, and the foot of those snails
fed a lead-rich diet, but not in the controls. Snails transferred from a lead-contaminated
environment to a lead-free environment could be cleared of lead in their soft tissues in about
four weeks, but the concentration of lead in the shells did not decrease. Borgmann et al.
(1978) found increased mortality in a freshwater snail, Lymnaea palutris, associated with
stream water with a lead content as low as 19 ug/1. Full life cycles were studied to estimate
population productivity. Although individual growth rates were not affected, increased
mortality, especially at the egg hatching stage, effectively reduced total biomass production
at the population level. Production was 50 percent at 36 ug/1 and 0 percent at 48 |jg Pb/1.
The relationship between LC50 and initial physiological response is not immediately
obvious. It is certain that some individuals of a population experience physiological dys-
function at concentrations well below that where half of them die. For example, Biesinger and
Christensen (1972) observed minimum reproductive impairment in Daphnia at 6 percent of the
(450 ug/1) for this species.
8-39
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8.4.3 Summary of Effects on Animals
While it is impossible to establish a safe limit of daily lead consumption, it is reason-
able to generalize that a regular diet of 2-8 mg Pb/kg-day body weight over an extended
period of time (Botts, 1977) will cause death in most animals. Animals of the grazing food
chain are affected most directly by the accumulation of aerosol particles on vegetation sur-
faces and somewhat indirectly by the uptake of lead through plant roots. Many of these
animals consume more than 1 mg Pb/kg-day in habitats near smelters and roadsides, but no toxic
effects have been documented. Animals of the decomposer food chain are affected indirectly by
lead in soil which can eliminate populations of microorganisms preceding animals in the food
chain or occupying the digestive tract of animals and aiding in the breakdown of organic
matter. Invertebrates may also accumultate lead at levels toxic to their predators.
Aquatic animals are affected by lead at water concentrations lower than previously con-
sidered safe (50 pg Pb/1) for wildlife. These concentrations occur commonly, but the contri-
bution of atmospheric lead to specific sites of high aquatic lead is not clear.
8.5 EFFECTS OF LEAD ON ECOSYSTEMS
There is wide variation in the mass transfer of lead from the atmosphere to terrestrial
ecosystems. Even within the somewhat artificial classification of undisturbed, cultivated,
and urban ecosystems, reported fluxes in undisturbed ecosystems vary by nearly 20-fold. Smith
and Siccama (1981) report 270 g/ha-yr in the Hubbard Brook forest of New Hampshire; Lindberg
and Harriss (1981) found 150 g/ha-yr in the Walker Branch watershed of Tennessee; and Elias
et al. (1976) found 15 g/ha-yr in a remote subalpine ecosystem of California. Jackson and
Watson (1977) found 1,000,000 g/ha-yr near a smelter in southeastern Missouri. Getz et al.
(1977c) estimated 240 g/ha-yr by wet precipitation alone in a rural ecosystem largely culti-
vated and 770 g/ha-yr in an urban ecosystem.
One factor causing great variation is remoteness from source, which translates to lower
air concentrations, smaller particles, and greater dependence on wind as a mechanism of depo-
sition (Elias and Davidson, 1980). Another factor is type of vegetation cover. Deciduous
leaves may, by the nature of their surface and orientation in the wind stream, be more suit-
able deposition surfaces than conifer needles. Davidson et al. (1982) discussed the influence
of leaf surface on deposition rates to grasses.
The history of lead contamination in roadside ecosystems has been reviewed by Smith
(1976). Recent studies have shown three areas of concern where the effects of lead on eco-
systems may be extremely sensitive (Martin and Coughtrey, 1981; Smith, 1981). First, decom-
position is delayed by lead, as some decomposer microorganisms and invertebrates are inhibited
8-40
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by soil lead. Secondly, the natural processes of calcium biopurification are circumvented by
the accumulation of lead on the surfaces of vegetation and in the soil reservoir. Thirdly,
some ecosystems experience subtle shifts toward lead tolerant plant populations. These pro-
blems all arise because lead in ecosystems is deposited on vegetation surfaces, accumulates in
the soil reservoir, and is not removed with the surface and ground water passing out of the
ecosystem. Other potential effects are discussed that may occur because of the long-term
build-up of lead in soil.
8.5.1 Delayed Decomposition
The flow of energy through an ecosystem is regulated largely by the ability of organisms
to trap energy in the form of sunlight and to convert this energy from one chemical form to
another (photosynthesis). Through photosynthesis, plants convert light to stored chemical
energy. Starch is only a minor product of this energy conversion. The most abundant sub-
stance produced by net primary production is cellulose, a structural carbohydrate of plants.
Terrestrial ecosystems, especially forests, accumulate a tremendous amount of cellulose as
woody tissue of trees. Few animals can digest cellulose and most of these require symbiotic
associations with specialized bacteria. It is no surprise then, that most of this cellulose
must eventually pass through the decomposer food chain. Litter fall is the major route for
this pathway. Because 80 percent or more of net primary production passes through the decom-
posing food chain (Swift et al., 1979), the energy of this litter is vital to the rest of the
plant community and the inorganic nutrients are vital to plants.
The amount of lead that causes litter to be resistant to decomposition is not known.
Although laboratory studies show that 50 ug Pb/ml nutrient medium definitely inhibits soil
bacterial populations, field studies indicate little or no effect at 600 |jg/g litter (Doelman
and Haanstra, 1979b). One explanation is that the lead in the laboratory nutrient medium was
readily available, while the lead in the litter was chemically bound to soil organic matter.
Indeed, Doelman and Haanstra (1979a) demonstrated the effects of soil lead content on delayed
decomposition: sandy soils lacking organic complexing compounds showed a 30 percent inhibition
of decomposition at 750 (jg/g, including the complete loss of major bacterial species, whereas
the effect was reduced in clay soils and non-existent in peat soils. Organic matter maintains
the cation exchange capacity of soils. A reduction in decomposition rate was observed by
Doelman and Haanstra (1979a) even at the lowest experimental concentration of lead, leading to
the conclusion that some effect might have occurred at even lower concentrations.
When decomposition is delayed, nutrients may be limiting to plants. In tropical regions
or areas with sandy soils, rapid turnover of nutrients is essential for the success of the
forest community. Even in a mixed deciduous forest, a significant portion of the nutrients,
8-41
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especially nitrogen and sulfur, may be found in the litter reservoir (Likens et al. 1977).
Annual litter inputs of calcium and nitrogen to the soil account for about 60 percent of root
uptake. With delayed decomposition, plants must rely on precipitation and soil weathering for
the bulk of their nutrients. Furthermore, the organic content of soil may decrease, reducing
the cation exchange capacity of soil.
8.5.2 Circumvention of Calcium Biopurification
Biopurification is a process that regulates the relative concentrations of nutrient to
non-nutrient elements in biological, components of a food chain. In the absence of absolute
knowledge of natural lead concentrations, biopurification can be a convenient method for esti-
mating the degree of contamination. Following the suggestion by Comar (1965) that carnivorous
animals show reduced Sr/Ca ratios compared to herbivorous animals which, in turn show less
than plants, Eli as et al. (1976, 1982) developed a theory of biopurification, which hypothe-
sizes that calcium reservoirs are progressively purified of Sr, Ba, and Pb in successive
stages of a food chain. In other words, if the Sr/Ca and Ba/Ca ratios are known, the natural
Pb/Ca ratio can be predicted and the observed Pb/Ca to natural Pb/Ca ratio is an expression of
the degree of contamination. Elias et al. (1976, 1982) and Elias and Patterson (1980)
observed continuous biopurification of calcium in grazing and detrital food chains by the pro-
gressive exclusion of Sr, Ba, and Pb (Figure 8-5). It is now believed that members of grazing
and decomposer food chains are contaminated by factors of 30 - 500, i.e., that 97 - 99.9 per-
cent of the lead in organisms is of anthropogenic origin. Burnett and Patterson (1980) have
shown a similar pattern for a marine food chain.
The mechanism of biopurification relies heavily on the selective transport of calcium
across membranes, the selective retention of non-nutrients at physiologically inactive binding
sites, and the reduced solubility of non-nutrient elements in the nutrient medium of plants
and animals. For example, lead is bound more vigorously to soil organic complexes and is less
soluble in soil moisture (Section 6.5.1). Lead is also adsorbed to cell walls in the root
apoplast, is excluded by the cortical cell membrane, and is isolated as a precipitate in sub-
cellular vesicles of cortical cells (Koeppe, 1981). Further selectivity at the endodermis
results in a nutrient solution of calcium in the vascular tissue that is greatly purified of
lead. Similar mechanisms occur in the stems and leaves of plants, in the digestive and circu-
latory systems of herbivores and carnivores, and in the nutrient processing mechanisms of
insects.
Atmospheric lead circumvents the natural biopurification of calcium. Deposition on plant
surfaces, which accounts for 90 percent of the total plant lead, increases the ratio of Pb/Ca
in the diet of herbivores. Deposition on herbivore fur increases the Pb/Ca ratio in the diet
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10-
10'
10"
" 10-
10§
10-'
10- —
ROCKS SOIL PLANT HERBI
MOISTURE LEAVES VORES
CARNI
VORES
Figure 8-5. The atomic ratios Sr/Ca, Ba/Ca and Pb/Ca (O)
normally decrease by several orders of magnitude from the
crustal rock to ultimate carnivores in grazer and decomposer
food chains. Anthropogenic lead in soil moisture and on the
surfaces of vegetation and animal fur interrupt this process
to cause elevated Pb Ca ratios (•) at each stage of the
sequence. The degree of contamination is the ratio of Total
Pb/Ca vs. Natural Pb/Ca at any stage. Ba/Ca ;md Sr/Ca ratios
are approximate guidelines to the expected natural Pb/Ca
ratio.
Source: Adapted from Elias et al. (1982).
8-43
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of carnivores. Atmospheric lead consumed by inhalation or grooming, possibly 15 percent of
the total intake of lead, represents sources of lead that were non-existent in prehistoric
times and therefore were not present in the food chain.
8.5.3 Population Shifts Toward Lead-Tolerant Populations
It has been observed that plant communities near smelter sites are composed mostly of
lead tolerant plant populations (Antonovics et al., 1971). In some cases, these populations
appear to have adapted to high-lead soils, since populations of the same species from low-lead
soils often do not thrive on high-lead soils (Jowett, 1964). Similar effects have been ob-
served for soils enriched to 28,000 |jg/g dry weight with ore lead (Hdiland and Oftedal, 1980)
and near roadsides at soil concentrations of 1,300 ug/g dry weight (Atkins et al. , 1982). In
these situations, it is clear that soil lead concentration has become the dominant factor in
determining the success of plant populations and the stability of the ecological community.
Soil moisture, soil pH, light intensity, photoperiod, and temperature are all secondary fac-
tors (Antonovics et al., 1971). Strategies for efficient use of light and water, and for
protection from temperature extremes, are obliterated by the succession of lead-tolerant plant
populations. Smith and Bradshaw (1972) concluded that lead-tolerant plant populations of
Festuca rubra and Agrostis tenuis can be used to stabilize toxic mine wastes with lead concen-
trations as high as 80,000 ug/g.
8.5.4 Biogeochemical Distribution of Lead in Ecosystems
Inputs of natural lead to ecosystems, approximately 90 percent from rock weathering and
10 percent from atmospheric sources, account for slightly more than the hydrologic lead out-
puts in most watersheds (Patterson, 1980). The difference is small and accumulation in the
ecosystem is significant only over a period of several thousand years. In modern ecosystems,
with atmospheric inputs exceeding weathering by factors of 10 - 1000, greater accumulation
occurs in soils. This reservoir must be treated as lacking a steady state condition
(Heinrichs and Mayer, 1977, 1980; Siccama and Smith, 1978). Odum and Drifmeyer (1978)
describe the role of detrital particles in retaining a wide variety of pollutant substances,
and this role may be extended to include non-nutrient substances.
It appears that plant communities have a built-in mechanism for purifying their own
nutrient medium. As a plant community matures through successional stages, the soil profile
develops a stratified arrangement that retains a layer of organic material near the surface.
This organic layer becomes a natural site for the accumulation of lead and other non-nutrient
metals that might otherwise interfere with the uptake and utilization of nutrient metals.
But the rate accumulation of lead in this reservoir may eventually exceed the capacity of the
8-44
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reservoir. Johnson et al. (1982a) have established a baseline of 80 stations in forests of
the northeast United States. In the litter component of the forest floor, they measured an
average lead concentration of 150 ug/g. Near a smelter, they measured 700 ng/9 and near a
highway, 440 ug/g. They presented some evidence from buried litter that predevelopment con-
centrations were 24 ng/g. On an area basis, the present concentrations range from 0.7 to
1.8 g Pb/m2. Inputs of 270 g/ha-yr measured in the Hubbard Brook forest would account for
1.0 g Pb/m2 in forty years if all of the lead were retained. The 80 stations will be moni-
tored regularly to show temporal changes. Evidence for recent changes in litter lead concen-
trations is documented in the linear relationship between forest floor lead concentration and
age of forest floor, up to 100 years.
Lead in the detrital reservoir is determined by the continued input of atmospheric lead
from the litter layer, the passage of detritus through the decomposer food chain, and the rate
of leaching into soil moisture. There is strong evidence that soil has a finite capacity to
retain lead (Zimdahl and Skogerboe, 1977). Harrison et al. (1981) observed that most of the
lead in roadside soils above 200 ug/g is found on Fe-Mn oxide films or as soluble lead car-
bonate. Elias et al. (1982) have shown that soil moisture lead is derived from the Teachable/
organic fraction of soil, not the inorganic fraction. Lead is removed from the detrital
reservoir by the digestion of organic particles in the detrital food chain and by the release
of lead to soil moisture. Both mechanisms result in a redistribution of lead among all of the
reservoirs of the ecosystem at a very slow rate. A closer look at the mechanisms whereby lead
is bound to humic and fulvic acids leads to the following conclusions: 1) because lead has a
higher binding strength than other metals, lead can displace other metals on the organic
molecule (Schnitzer, 1978); 2) if calcium is displaced, it would be leached to a lower soil
horizon (B), where it may accumulate as it normally does during the development of the soil
profile; and 3) if other nutrient metals, such as iron or manganese, are displaced, they may
become unavailable to roots as they pass out of the soil system.
Fulvic acid plays an important role in the development of the soil profile. This organic
acid has the ability to remove iron from the lattice structures of inorganic minerals, result-
ing in the decomposition of these minerals as a part of the weathering process. This break-
down releases nutrients for uptake by plant roots. If all binding sites on fulvic acid are
occupied by lead, the role of fulvic acid in providing nutrients to plants will be circum-
vented. While it is reasonably certain that such a process is possible, there is no informa-
tion about the soil lead concentrations that would cause such an effect.
Ecosystem inputs of lead by the atmospheric route have established new pathways and
widened old ones. Insignificant amounts of lead are removed by surface runoff or ground water
seepage. It is likely that the ultimate fate of atmospheric lead will be a gradual elevation
8-45
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in lead concentration of all reservoirs in the system, with most of the lead accumulating in
the detrital reservoir.
8.6 SUMMARY
Because there is no protection from industrial lead once it enters the atmosphere, it is
important to fully understand the effects of industrial lead emissions. Of the 450,000 tons
emitted annually on a global basis, 115,000 tons of lead fall on terrestrial ecosystems.
Evenly distributed, this would amount to 0.1 g/ha-yr, which is much lower than the range of
15-1,000,000 g/ha-yr reported in ecosystem studies in the United States. Lead has permeated
these ecosystems and accumulated in the soil reservoir where it will remain for decades.
Within 20 meters of every major highway, up to 10,000 ug Pb have been added to each gram of
surface soil since 1930 (Getz et al., 1977c). Near smelters, mines, and in urban areas, as
much as 130,000 ug/g have been observed in the upper 2.5 cm of soil (Jennett et al., 1977).
At increasing distances up to 5 kilometers away from sources, the gradient of lead added since
1930 drops to less than 10 ug/g (Page and Ganje, 1970), and 1-5 |jg/g have been added in
regions more distant than 5 kilometers (Nriagu, 1978). In undisturbed ecosystems, atmospheric
lead is retained by soil organic matter in the upper layer of soil surface. In cultivated
soils, this lead is mixed with soil to a depth of 25 cm.
Because of the special nature of the soil reservoir, it must not be regarded as an infi-
nite sink for lead. On the contrary, atmospheric lead that is already bound to soil will
continue to pass into the grazing and detrital food chains until equilibrium is reached,
whereupon the lead in all reservoirs will be elevated proportionately higher than natural
background levels. This conclusion applies also to cultivated soils, where lead bound within
the upper 25 cm is still within the root zone.
Few plants can survive at soil concentrations in excess of 20,000 ug/g, even under opti-
mum conditions. Some key populations of soil microorganisms and invertebrates die off at 1000
ug/g. Herbivores, in addition to a normal diet from plant tissues, receive lead from the sur-
faces of vegetation in amounts that may be 10 times greater than from internal plant tissue.
A diet of 2-8 mg/daykg body weight seems to initiate physiological dysfunction in many
vertebrates.
Whereas previous reports have focused on possible toxic effects of lead on plants,
animals, and humans, it is essential to consider the degree of contamination as one measure of
safe concentration. Observed toxic effects occur at environmental concentrations well above
levels that cause no physiological dysfunction. Small animals in undisturbed ecosystems are
contaminated by factors of 20-600 over natural background levels, and in roadside and urban
8-46
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ecosystems by 300-6200. Extrapolations based on sublethal effects may become reliable when
these measurements can be made with controls free of contamination. The greatest impact may
be on carnivorous animals, which generally have the lowest concentrations of natural lead, and
may thus have the greatest percent increase when the final equilibrium is reached.
Perhaps the most subtle effect of lead is on ecosystems. The normal flow of energy
through the decomposer food chain may be interrupted, the composition of communities may shift
toward more lead-tolerant populations, and new biogeochemical pathways may be opened, as lead
flows into and throughout the ecosystem. The ability of an ecosystem to compensate for atmos-
pheric lead inputs, especially in the presence of other pollutants such as acid precipitation,
depends not so much on factors of ecosystem recovery, but on undiscovered factors of ecosystem
stability. Recovery implies that inputs of the perturbing pollutant have ceased and that the
pollutant is being removed from the ecosystem. In the case of lead, the pollutant is not
being eliminated from the system nor are the inputs ceasing. Terrestrial ecosystems will
never return to their original, pristine levels of lead concentrations.
8-47
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