WORKSHOP PROCEEDINGS
ACIDIC DEPOSITION, TRACE CONTAMINANTS
AM) THEIR INDIRECT HUMAN HEALTH EFFECTS:
RESEARCH NEEDS
held
June 19-22, 1984
at
Itpridge Conference Center
Paul Smith's, New York
Editor
Scott 0. Quinn
Bureau of Water Research
New York State Department of ESwironmental Conservation
Assistant Editor
Naomi Blocmtield, M.D.
Albany Medical College
Sponsors
U.S. Environmental Protection Agency
New York State Department of Environmental Conservation
Albany Medical Cbllege
August 1985
Project Officer
Robert T. Lackey
U.S. Environmental Protection Agency
Environmental Research laboratory - Corvallis
Corvallis, Oregon
ENVIRONMENTAL RESEARCH LABORATORY - CORVALLIS
CORVALLIS, OREGON
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DISCLAIMER
This workshop proceedings has been funded as part of the National Acid
Precipitation Assessment Program in part by the U.S. Environmental Protection
Agency under contract P.O. #4B1201NAEX to the N.Y.S. Department of Environmental
Conservation. This proceedings has been subject to the EPA's peer and
administrative review, and it has been approved for publication as an EPA
document. Although it has been approved for publication as an EPA document,
because it" is a workshop proceedings, it does not necessarily reflect the view
of the Agency and no official endorsement should be inferred.
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PREFACE
A workshop to examine the research needs for acidic deposition, trace
contaminants and their indirect human health effects was held at the Topridge
Conference Center in the Adirondack Mountains of New York State during June
19-22, 1984. The meeting was co-sponsored by the United States Environmental
Protection Agency, the New York State Department of Environmental Conservation,
and the Albany Medical College.
The purpose of the workshop was to: (1) review the current knowledge base;
and (2) propose a research agenda concerning the relationship between acidic
deposition and altered biogeochemical cycling of trace contaminants, as related
specifically to increased human exposure (post-deposition).
^rkshcp-afctsndees^were selected to-iepresent- a,broad. spectrum-of
scientists with an established record of expertise in acidic deposition
research. Most participants have been actively involved in national and
international efforts to develop research programs with respect to acidic
deposition.
The conference was organized into four working groups, each focusing on one
aspect of the acidic deposition phenomenon. Working groups in atmospheric
processes, biogeochemistry, bioaccumulatlon and human health effects developed
state-of-the-art sunmaries concerning acidic deposition effects. Each working
group summarized known and unknown aspects of the acidic deposition phenomenon
and identified research needs.
The four working groups collaborated to select specific trace contaminants
for study. They concluded that acidic deposition may influence human exposure
to trace contaminants via two main pathways: (1) the accumulation of
contaminants in the food chain leading to man; and (2) the contamination of
drinking water. After much discussion it was decided that those contaminants
requiring special attention are mercury, lead, aluminum, and cadmium. Thip
decision was based on several factors, including human toxicity and potential
for human exposure.
The research agenda which follows addresses gaps in the data concerning the
impacts of acidic deposition on the biogeochemical cycling of important trace
contaminants. The research needs identified, therefore, are those required to
elucidate the human health effects indirectly exacerbated by acidic deposition.
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TABLE OF CONTENTS
Page
PREFACE iii
WORKSHOP PARTICIPANTS ... ix
Section
I. CURRENT STATE OF KNOWLEDGE
1. Atmospheric Processes ... 1
1.1 General Introduction. 1
1.1.1 Brtissians 1
1.1.2 Transport... 3
1.1.3 Transformation 3
1.1.4 Deposition 3
1.1.5 Measurements.. 5
1.2 Mercury 5
1.2.1 Dnissions 5
1.2.2 Transport 5
1.2.3 Deposition 5
1.3 Lead 10
1.3.1 Emissions. 10
1.3.2 Transport.. 10
1.3.3 Deposition... 10
1.5 Aluminum 10
1.5.1 Emissions 10
1.5.2 Transport 10
1.5.3 Deposition 10
1.4 Cadmium . 10
1.4.1 Dnissions 10
1.4.2 Transport 10
1.4.3 Deposition... 11
iv
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2. Biogeochemistry 11
2.1 General Introduction 11
2.2 Mercury 13
2.2.1 Importance of Direct Atmospheric
Deposition to a Lake vs. Watershed
Weathering 13
2.2.2 Mercury Mobilization frcm Watershed
and Sediments 13
2.2.3 Influence of Acidification on Mercury
Speciatian 13
2.2.4 Effect of Acidification on Microbial
Transformations of Mercury 14
2.2.5 Effects of Acid Water and Increased
Atmospheric Loading on Mercury Levels
in Water Distribution Systems 15
2.3 Lead 15
2.3.1 Importance of Direct Atmospheric
Deposition to a Lake vs. Watershed
Weathering 15
2.3.2 Lead Mobilization from Watershed
and Sediments 15
2.3.3 Influence of Acidification on Lead
Speciatian .. 16
2.3.4 Effect of Acidification on Microbial
Transformations of Lead 16
2.3.5 Effect of Acid Water and Increased
Atncspheric Loading on Lead Levels in
Water Distribution Systems 16
2.4 Aluminum 16
2.4.1 Importance of Direct Atmospheric
Deposition to a Lake vs. Watershed
Weathering 16
2.4.2 Aluminum Mobilization frcm Watershed
and Sediments 16
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2.4.3 Influence of Acidification on Aluminum
Speciation 17
2.4.4 Effect of Acidification on Microbial
Transformations of Aluminum 17
2.4.5 Effect of Acid Water and Increased
Atmospheric Loading on Aluminum Levels
in Water Distribution Systems 17
2.5 Cadmium 17
2.5.1 Importance of Direct Atmospheric
Deposition to a Lake vs. Watershed
Weathering... 17
2.5.2 Cadmium Mobilization frcm Watershed
and Sediments 17
2.5.3 Influence of Acidification on Cadmium
Speciation 18
2.5.4 Effect of Acidification on Microbial
Transformations of Cadmium.... 18
2.5.5 Effect of Acid Water and Increased
Atmospheric Loading on Cadmium Levels
in Water Distribution Systems 18
Bioaocunulation 18
3.1 General Introduction. 18
3.2 Mercury -.. 19
3.2.1 Accumulation of Mercury by Fish 19
3.2.2 Effects of Acidification on Mercury
Bioaccumulation by Fish . 19
3.2.3 Effects of Productivity on Mercury
Uptake by Fish 23
3.2.4 Effects of Chemical Neutralization of
Lakes on Mercury Contamination of Fish..... 24
3.3 Other Metals... .24
Indirect Health Effects. 24
4.1 General Introduction. 24
4.2 Mercury 21
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4-2.1 Environmental Exposure and Uptake
by Humans . 27
4.2.1.1 Concentrations in Water. 27
4.2.1.2 Concentrations in Aquatic Biota.. 28
4.2.1.3 Human Uptake and Risk
Assessment 28
4.2.2 Effects on Human Health... .. 31
4.2.2.1 Metabolism in Humans 31
4.2.2.2 Effects and Dose-Response
in Humans 33
4.3 Lead. 35
4.3.1 Environmental Exposure and Uptake
by Humans 35
4.3.1.1 Concentrations in Waters 35
4.3.1.2 Concentrations in Aquatic.Biota.. 36
4.3.1.3 Human Uptake and Risk
Assessnent 36
4.3.2 Effects on Human Health 40
4.3.2.1 Metabolism in HLonans 40
4.3.2.2 Effects and Dose-Response
in Humans... 42
4.4 Aluminum 47
4.4.1 Environmental Exposure and Ujptake by
Humans 47
4.4.1.1 Concentrations in Water 47
4.4.1.2 Concentrations in Aquatic Biota.. 48
4.4.1.3 Human Uptake and Risk
Assessment 48
4.4.2 Effects on Human Health.. 49
4.4.2.1 Metabolism in Humans 49
4.4.2.2 Effects and Dose-Response in
Humans 49
4.5 Cadmium 50
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4.5.1 Environmental Exposure and Uptake by
Humans..... 50
4.5.1.1 Concentrations in Water. 50
4.5.1.2 Concentrations in Aquatic Biota.. 50
4.5.1.3 Human Uptake and Risk
Assessment 50
4.5.2 Effects on Human Health 51
4.5.2.1 Metabolism in Humans. 51
4.5.2.2 Effects and Dose-Response in
Humans .. 51
4.6 Other Trace Contaminants 51
4.6.1 Asbestos . 51
4.6.2 Nitrates. 52
II. SUMMARY OF RESEARCH RECOMMENDATIONS
Atmospheric Processes 53
Biogeochemistry 53
Bioaccumulatian 55
Indirect Health Effects. 55
III. LITERATURE CITED
References 58
Appendices
A. AGENDA FOR WORKSHOP 74
B. ACKfJOWLEDGEWENTS ... 76
viii
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WORKSHOP PARTICIPANTS
AJM3SPHERIC PROCESSES
Chairman
Perry J. Samson, Ph.D.
University of Michigan
Department of Atmospheric and
Ocean Science
2213 Space Research Building
Ann Arbor, Michigan 48109
Kenneth T. Knapp, Ph.D.
Supervisory Research Chemist
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Cliff I. Davidson, Ph.D.
Assistant Professor of Civil
Engineering & Public Policy
Carnegie-Mellon University
Pittsburgh, Pennsylvania 15213
BIOGEOCHEMtSTRY
Chairman
Peter G.C. Campbell, Ph.D.
Universite of Quebec
Institut national de la
recherche scientifique
C.P. 7500
Ste.-Fqy, Quebec, Canada G1V 4C7
Stephan A. Norton, Ph.D.
Department of Geological Sciences
University of Maine at Qrono
110 Boardman Hall
Qrono, Maine 04469
BIOftCCtMILATION
Chairman
James G. Wiener, Ph.D.
U.S. Fish & Wildlife Service
Colunbia Hat. Fish. Res. Lab.
Field Research Station
P.O. Box 936
La Crosse, Wisconsin 54601
Liquat Husain, Ph.D.
Director, Chemical Sciences Lab.
Center for Labs and Research
N.Y.S Department of Health
Foam D-539
Albany, New York 12201
Thomas M. Church, Ph.D.
College of Marine Sciences
University of Delaware
Newark, Delaware 19716
Philip J. Galvin, Ph.D.
Division of Air Resources
N.Y.S. Department of
Environmental Conservation
50 Wolf Road
Albany, New York 12233-0001
Hans Hultberg, Ph.D.
Research Group Leader
Swedish Environmental Research
Institute
P.O. Bok 5207
S-402 24 Gothenburg 5, Sweden
Michael R. Winfrey, Ph.D.
Department of Biology
University of Wisconsin
La Crosse, Wisconsin 54601
Pamela M. Stokes, Ph.D.
Institute for Environ. Studies
University of Toronto
Haul train Building
Toronto, Ontario, Canada M5S 1A6
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BIQACCUMUIATION (cqn't.)
Ronald J. Sloan, Ph.D.
Division of Fish & Wildlife
N.Y.S. Dept. Environmental
Conservation
50 Vfolf Road
Albany, New York 12233-0001
HUMAN HEALTH EFFECTS
Chairman
lianas W. Clarksori, Ph.D.
Division of Toxicology
University of Rochester
School of Medicine
P.O. Box RBB
Rochester, New York 14642
Gunar Nordberg, M.D.
Professor
Dept. of Environmental Medicine
Umea University
S-901 87 Unea, Sweden
Magnus Piscator, M.D.
Department of Environmental Hygiene
Karolinska Institute
P.O. Box 60400
S-104 01 Stockholm, Sweden
David A. Bennett, Ph.D.
Director
Acid Deposition Assessment Staff
U.S. Env. Protection Agency
401 M Street, SW (RD-676)
Washington, D.C. 20460
Harvey Olem, Ph.D.
Tennessee Valley Authority
248 401 Building
Chattanooga, Tennessee 37401
Shadad Joshi, M.D.
Albany Medical Center
New Scotland Avenue
Albany , New York 12209
Robert G. Miller/ Ph.D.
Health Effects Research Lab.
U.S. Env. Protection Agency
26 West St. Claire Street
Cincinnati, Ohio 45268
Robert A. Goyer, M.D.
-National Institute of
Environmental Health Sciences
P.O. Box 12233
Research Triangle Park, NC 27709
Paul D. Moskcwitz, Ph.D.
Health Effects Assessment
Brookhaven National Laboratory
Upton, New York 11973
Michael McDonald, Ph.D.
Center for Limnology
680 N. Park Street
University of Wisconsin
Madison, Wisconsin 53706
WORKSHOP ADMINISTRATION
Scott 0. Quinn Nacmi Blocmfield, M.D.
Bureau of Water Research Albany Medical College
N.Y.S. Department of New Scotland Avenue
Environmental Conservation Albany, New York 12209
Roan 317
50 Wblf Road
Albany, New York 12233-0001
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WORKSHOP ADMINISTRATION (con't.)
Joseph M. Eilers
U.S. Environmental Protection Agency
200 SW 35th Street
Corvallis, Oregon 97333
xi
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SECTION I: CURRENT STATE OF KNOWLEDGE
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1. ATMOSPHERIC PROCESSES (P.J. Samson, T.M. Church, C.I. Davidson, P.J. Galvin,
L. Husain, K.T. Knapp)
1.1 GENERAL INTRODUCTICN
Hie atmosphere plays a major role in the movement of trace elements in
the environment. Trace elements are known to be transported in the atmosphere
fron their source to their eventual deposition. This direct transfer ot
material to the soil nay be an important input ot seme elements including
sulfur, selenium, lead and cadmium which have large anthropogenic emissions.
Ftor other trace elements such as aluminum, the importance of direct atmospheric
input is thought to be secondary to the in-soil mobilization of the element by
atmospheric acids (see Section 2.). This section briefly discusses the role ot
the atmosphere in transporting and depositing of both trace elements and
atmospheric acids, and prioritizes the research agenda needed to gain a firmer
understanding of the atmospheric input of trace elements into the environment.
Brphasis is given to the elements mercury, lead, aluminum,and cadmium because
these have been identified as potentially toxic metals in the environment.
Sulfur is also discussed inasmuch as the acidity associated with sulfur oxides
is believed to be responsible for mobilizing trace metals in the soil.
The atmospheric input ot both trace elements and acidic materials into
the environment is the result of four major processes: emission, transport,
transformation and deposition. Each of the processes can be investigated
individually to identify areas where additional knowledge is required.
1.1.1 Bnissions — Data about the emissions of trace elements into the
atmosphere are limited in many aspects. Almost all studies of trace element
emissions have not addressed the form, size distribution, or speciation of the
emitted material. Only emissions ot sulfur dioxide, the precursor for
particulate sulfate, have been quantified with any confidence.
Lantzy and Mackenzie (1979) and Galloway et al. (1980) have estimated
the global magnitude of natural and anthropogenic emissions of many trace
elements. Table 1-1 lists their estimates of the global emission rates for
mercury, lead and cadmium. On this scale it is obvious that the atmospheric
source of these elements is largely anthropogenic, though natural emissions of
mercury in the vapor phase are relatively large. The more recent estimates for
global emission ot mercury by Lindquist et al. (1984) appears to be about an
order of magnitude lower than those estimated by Lantzy and MacKenzie.
Fitzgerald et al. (1983) have estimated global emissions at 8 x 10 kg y using
new data on biological emissions frcm the oceans. Thus the estimation of global
mercury emissions is still debated.
Data tor specific parts of the globe and for specific industries is
relatively unknown. Pacyna et al. (1984) have estimated the emissions rate of
several trace elements from European sources by source type. Table 1-2 lists
their estimates for lead and cadmium for 1979. Data tor sized particles frcm
coal-fired and oil-fired power plants, pulp and paper plants and sludge
incineration have been reported (Bennett and Knapp, 1978; Henry, 1979; Henry and
Knapp, 1980) but these data did not include several ot the important trace
elements. Few data exist on the elemental content ot non-point source emissions
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Table 1-1. Estimates of natural and anthropogenic global annugl emission
rates for mercury, lead and cadmium in units of 10 kg y as
derived frcm Galloway et al., (1980). Numbers in parentheses
represent estimates of vapor emissions.
Natural Anthropogenic
Mercury 0.4 (25) 11
Lead 5.9 2,000
Cadmium 0.3 6
Table 1-2. European emissions of trace elements^in 1979 as estimated by
Pacyna et al., (1984) in units of 10 kg y .
Cd
Pb
Coal Combustion
0:15
1.67
Oil Combustion
0.11
1.12
Wood Combustion
0.025
0.56
Gasoline Combustion
0.031
74.3
Mining
0.001
1.09
Copper-Nickel Smelters
0.60
9.25
Zinc-Cadmium Smelters
1.55
7.88
Lead Shelters
0.008
10.45
Manufacturing
0.058
14.66
Refuse Incinerators
0.84
0.80
Phosphate Fertilizers
0.027
0.01
Cement Production
0.015
0.75
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which could be a major contribution of certain trace elements in the
environment.
1.1.2 Transport — Atmospheric transport of pollutants from emission to
eventual deposition is dependent upon the speed of the transporting winds and
the rate of removal from the atmosphere. The removal of pollutants occurs by
dry or wet deposition. Hie pollutant can be deposited at the surface through
sedimentation or impaction and uptake (dry deposition) or can become
incorporated into hydrcraeteors (cloud droplets, raindrops, or snowflakes) and be
deposited to the surface (wet deposition). In general, the distance the
pollutant travels before it deposits depends, in part, on the height of
injection of the material into the atmosphere. This is because, in general, the
wind speed increases with height and hence the pollutant is carried further, and
also because the elevated releases require a longer time for the plume to be
mixed to the surface and dry deposited.
It is useful to ccnpare the transport of trace elements to that of the
more-studied sulfur oxides to identify the areas where our understanding of
trace element transport is still poor. This section also serves to illustrate
the transport scales of sulfur oxides responsible for the production of acidic
precipitation.
Sulfur oxides are emitted largely as gaseous sulfur dioxide (SOJ
though it is estimated that 1 to 5% of the sulfur oxide is emitted directly as
sulfate. The percentage of this primary sulfate emission is thought to be
highest for oil-fueled sources. While the bulk of the sulfur is transported
away frcm the source area, a fraction of the sulfur will be oxidized into
sulfate (SO.) and a fraction will be removed as 90. through dry deposition to
within 50 km of the source.
The distance the pollutant travels before depositing is dictated in
part by the physical and chemical characteristics of the particle being
transported. Particulate sulfate, for example, is created through a
gas-to-particle conversion process which yields very small aerosol particles.
These particles will collide and grow to a size of between 0.1 to 1.0 urn. In
this size range, the particles have very small deposition velocities and can
travel very long distances without significant losses due to dry deposition.
1.1.3 Transformation — Transformation involves a chemical change in the
emitted species. Such changes can affect the efficiency with which the material
is deposited by precipitation and the residence time of the material. Chemical
transformation is known to be important in the fornetion of atmospheric acids.
Most trace elements enter the atmosphere directly in a particulate
form or very rapidly condense into particles as the hot gaseous effluent cools.
An exception to this occurs with mercury and possibly cadmium (Galloway et al.,
1980). Hie reactivity of these trace elements is thought to be small compared
to. sulfur oxides (Lindquist et al., 1984).
1.1.4 Deposition — Trace elements are deposited at the earth's surface by
"wet" or "dry" processes. Wet deposition is usually confined to precipitation
(rain or snow) although dew and fog condensation are sometimes included. Dry
deposition includes processes such as sedimentation, interception, inertial
3
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impaction, and diffusion; thermophoratic and electrostatic forces may also be
important in sane instances. In all but very dry areas, wet deposition accounts
for at least half of the total deposition for those trace elements predominantly
conposed of submicron particles. For trace elements largely dominated by
natural crustal sources (e.g. Fe, Mn, and possibly Cu) dry deposition is as
great or even greater than wet deposition on an annual basis as evidenced from
secular increases in trace elements to aquatic systems during differential
transport by wind conditions. For other trace elements which have predominant
inventories in the gaseous phase (e.g. Hg, sane metalloids), dry deposition via
gas exchange with aquatic surfaces appears important. However, most estimates
of total deposition of trace elements rely on our very limited knowledge of
deposition velocities, and the ambient atmosphere concentrations of those
elements. Total deposition estimates based on soil profiles are also fraught
with geochemical uncertainties of diagenetic post-depositional mobilities of
many-trace elsrcnts (see Section 2.1).
Measurement of trace elements in precipitation at even remote ambient
concentrations is within the current state of analytical technique (eg. Church
et al., 1985? Davidson et al., 1984a). However, extrapolation of such trace
element concentrations to wet deposition assumes minor contamination during
collection and the use of adequate procedures to quantify the lack of background
concentration in blank samples. Procedures for collection of trace elements in
precipitation have been shewn to require strict protocols of trace element-free
and clean collector equipment, containers, and reagents. These protocols.need
to be established for individual areas and needs, but are new within the state
of the art for collection, processing, and analysis of trace elements in natural
waters at ultra trace concentrations. Nevertheless, extrapolation of trace
element concentrations in precipitation to actual wet deposition still has not
clearly resolved particulate versus dissolved trace element burdens or
speciation.
Measurement of trace element dry deposition is far more difficult than
measurement of wet deposition. Examples of dry flux assessment techniques,
include exposure of surrogate surfaces, vegetation washing, and micrcmeteoro-
logical methods. Exposure of surrogate surfaces is convenient and inexpensive,
but the technique has been criticized due to the inability to simulate natural
surfaces (Hicks et al., 1980). Recent data has suggested the possible utility
of the method for assessing fluxes of large particles (Davidson et al., 1984b).
Vegetation washing may provide direct deposition data on the surfaces
of interest. However, large variability among samples, and interferences caused
by leaching of internal material, limit the utility of the method for routine
monitoring (Lindberg et al., 1982).
Finally, examples of micrcmeteorological techniques include
measurement of concentration gradients (Garland, 1978), eddy correlation (Wesley
et al., 1977), and variance techniques (Wesley and Hicks, 1978). These methods
may provide estimates of vapor and submicron particle fluxes, but the strict
fetch and analytical precision requirements, as well as the expense of operating
such systems, limit their utility (Hicks et al., 1980). It is apparent that the
limitations of currently- available techniques prevent the acquisition of
reliable trace element dry deposition data on a routine basis.
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1.1.5 Measurements — Measured atmospheric concentrations of trace metals
have been sumarized by Galloway et'al. (1980). A similar compilation of
concentrations in precipitation has been presented by Galloway et al. (1982).
They report a range of concentrations in the atmosphere and precipitation as
illustrated in Figures 1-1 and 1-2, respectively, for mercury; Figures 1-3 and
1-4 for lead and Figures 1-5 and 1-6 for cadmium. As expected, the highest mean
concentrations measured are for urban areas. The "remote" locations cited have
concentrations frcm one to several orders of magnitude lower than "urban"
samples.
Lindberg (1982) has investigated the variations in lead concentrations
in rain. He shows that lead is higher in the sunnier months in response to
increased air stagnation and lcwer precipitation volume (hence less dilution)
during that season. Rusain and Samson (1979) illustrated that in a remote
location during the sunnier the concentrations of sane elements are governed by
the wind flew. At White face Mountain, New York extensive measurements have
shown that lead and several other elements concentrations are enhanced
approximately four-fold when windflow is frcm the Midwestern United States
(Husain, 1984). Atmospheric transport of trace metals frcm mid-latitude source
regions to the Arctic has been documented by several investigators (Rahn, 1981;
Barrie et al., 1981; Davidson et al., 1984a). Thus it is obvious that those
trace metals which are found as sutmicron particles are being transported long
distances frcm their source.
The acidity of precipitation may also be contributing to the
mobilization of metals in the soil (see Section 2.1). Current measurements of
hydrogen ion concentration generally show a maximum deposition rate in the upper
Ohio River basin. Figure 1-7 shows this distribution for the United States and
Canada in units of pH.
1.2 MERCURY
1.2.1 Emissions — Much of the emissions data for trace elements is .
derived frcm analyses of collected particles (fly ash). Thus, little data exist
on the emissions of mercury into the atmosphere because a sizable fraction of
the emitted material is in the vapor phase (Lindquisf'et al., 1984). Likewise,
although anthropogenic emissions of mercury are significant, there are
uncertainties about the natural emission rates.
1.2.2 Transport — Hie atmospheric transport of mercury has also not been
adequately defined because of the lack of vapor-phase data. Husain and Samson
(1979) found that particulate mercury behaved like no other trace element,
having its highest concentrations at times most other elements were
significantly below average.
1.2.3 Deposition — Mercury in the environment is partially due to direct
atmospheric input. This fraction is hard to quantify because of the lack of
reliable data. The increased groundwater input of mercury to aquatic systems is
augmented by increased acidification of historically deposited mercury. This
mercury has accumulated in the soil profiles over recent industrial times and
may be released into the aquatic systems as the pH of the precipitation
decreases (see Sections 2.2.1 and 2.2.2).
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ATMOSfHKWIO MtWCUWY CONCtNTWATION»
Figure 1-1. Minimum, median and maximum concentrations of airborne
mercury found in the literature by Galloway et al. (1980)
for remote, rural and urban locations. Median concentrations
are for total Hg while max and min include particulate Hg
only.
PMKCIPtTATION Mucunv CONCENTRATIONS
I *
i
Figure 1-2. Minimum, median and maximum concentrations of mercury in
precipitation found in the literature by Galloway et ad.
(1982) for remote, rural and urban locations.
6
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V I
IIUfUL URBAN
smc ocseRiFnON
MIM ~ MCOIAN
Figure 1-3. Minimum, median and naximum concentrations of airborne lead
found in the literature by Galloway et al. (1980) for renote,
rural and urban locations.
PREOIf>lTATION LEAD OONOiKTIUTlONi
S -
IT* DUeKIFDON
MIOMN
Figure 1-4.
Minimum, median and neximun concentrations of lead ill
precipitation found in the literature by Galloway et al.
(1982) for raiote, rural and urban locations.
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ATMOSPHERIC CADMIUM CONCENTRATIONS
NUKAL
• IT* DESCRIPTION
~ MEDIAN
URBAN
Figure 1 -5. Minimum, median and maximum concentrations of airborne cadmium
found in the literature by Galloway et al. (1980) for remote,
rural and urban locations.
PRECIPITATION CADMIUM CONCENTRATIONS
RURAL
srr* DESCRIPTION
~ MEDIAN .~
URBAN
Figure 1-6, Minimum, median and maxinon concentrations of cadmium in
precipitation found in the literature by Gallcway et al.
(1982) for remote/ rural and urban locations.
8
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,/y^
*630
//\~\
\
'-*.5^ —
J
• s»t
• Sj60
• 5.73
Figure 1-7. Annual mean value of pH in precipitation weighted by the
amount of precipitation in the United States and Canada for
1980 (from Calvert et al., 1983).
9
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1.3 LEAD
1.3.1 Bnissions — Compared to mercury the emissions of lead are probably
better understood. The emissions of lead are largely the result of vehicular
emissions, with smaller amounts frcm stationary sources such as coal ccmbustion,
snelting, and similar industries (Patterson, 1980).
1.3.2 Transport — The transport of lead in the atmosphere is thought to
be similar to other fine particles. The particles generally coagulate to a
diameter of 0.1 - 1.0 ym and thus are capable of being transported long
distances.
1.3.3 Deposition — The wet deposition of lead in rural areas of high
precipitation is probably dominate over dry deposition. For example, Peirson et
al. (1973) have shown that dry deposition of lead accounts for less than 10
percent of the total deposition at a rural site in England. Hcwever, dry
deposition may play a significant role in areas receiving less precipitation.
Roughly equal wet and dry annual fluxes of lead for Walker Branch Watershed,
Tennessee have been reported by Lindberg et al. (1982).
1.5 ALUMINUM
1.5.1 Bnissions — Aluminum, unlike the other trace metals discussed
above, is largely soil derived. There is evidence that aluminum is being
mobilized in the soil by the increased acidity of the precipitation.
1.5.2 Transport — Aluminum is generally composed of larger particles than
the other three trace metals discussed in this report. Thus, its ability to be
transported long distances is not as great.
1.5.3 Deposition — Both wet and dry deposition velocities of aluminum are
significantly greater than those of lead and cadmium, due to the larger particle
sizes associated with aluminum. Davidson et al. (1985b) have reported that dry
deposition velocities for aluminum are more than an order of magnitude greater
than those of lead or cadmium; Peirson et al. (1973) have reported a factor of
three difference.
These latter authors have also reported total annual dry deposition
fluxes for aluminum which are approximately one-fifth of the total deposition
fluxes at a rural site in England. These values may have been influenced 'by
re suspension of local soil, hcwever, and must be regarded as preliminary.
1.4 CADMIUM
1.4.1 Bnissions — Important emission sources of cadmium include smelters
and refuse incinerators. Incinerators may becane increasingly important as
cannunities deal with their growing need for solid waste disposal. Details of
the emissions of cadmium have been discussed by Nriagu (1980).
1.4.2 Transport — Cadmium in the atmosphere is associated with somewhat
larger particles than lead. A recent literature review by Davidson and Osborn
(1985a) has shown an average mass median aerodynamic diameter (MMAD) of 1.5 vim,
with 25 percent of the mass associated with particles smaller than 0.5 pan and
10
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that lead has an average MMAD of 0.53 yrn with 51 percent of the mass less than
0.5 van. Thus, one would expect cadmium to be significantly influenced by
long-range transport, although perhaps to a lesser extent than lead.
1.4.3 Deposition — Less is known about the deposition rates of cadmium
then is known about the deposition rates of lead. Lindberg et al. (1982)
reported that the annual dry deposition fluxes which are approximately one-fifth
of the annual wet deposition in Walker Branch Watershed, Tennessee. Davidson et
al. (1985b) report that cadmium dry deposition velocities onto surrogate
surfaces are cxxiparable to those of lead in Olympic National Park, Washington.
More information is needed before cadmium deposition rates are kncwn with any
degree of certainty.
2. BIOGBOCHEMISTHY (P.G.C. Campbell, H. Hultberg, S.A. Norton, M.F. Winfrey)
2.1 GENERAL INTRODUCTION
Workshop participants considered the relationship between acidic
deposition and altered biogeochemical cycling of trace contaminants (metals), as
related specifically to increased human exposure. For any organism, increased
exposure to a given metal inplies changes in the concentration and/or the
speciation of the specific contaminant. Changes in concentration within various
ecosystem components (atmosphere, water, biota) has been reported for sane
metals, related either to increased atmospheric deposition of the metal or to
its increased mobilization in response to acidification. Even if total metal
concentrations do not change, changes in metal speciation can be anticipated as
a result of acidification, due to shifts in the hydrolysis, ccmplexation, and
adsorption equilibria.
FDr the metals (Hg, Cd, Pb and Al) judged to be of potential concern
in the context of acidic deposition/health effects, the Biogeochemistry Working
Group considered the following specific-questions:
(1) importance of direct atmospheric deposition to a lake vs. inputs
frcm watershed weathering;
(2) metal mobilization frcm the watershed and from sediments;
(3) influence of acidification on metal speciation (and hence
bioavailability);
(4) effect of acidification on the microbial transformation of the
metal;
(5) effect of acid water and increased atmospheric loading on water
quality in water distribution systems.
The types of evidence considered in responding to these various questions are
indicated in Table 2-1. As can be appreciated on examination of this table, our
responses to questions (1), (2), (4) and (5) were straightforward. However, our
approach to question (3), dealing with the influence of acidification on metal
bioavailability, clearly merits seme additional explanation.
In principle acidification could directly influence metal
bioavailability by (i) shifting the equilibrium distribution of the dissolved
metal complexes, both inorganic and organic, towards more available forms (e.g.,
the aquo ion, M(H20) z ), or (ii) influencing metal-surface interactions at the
11
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Table 2-1. Evidence considered in addressing the questions posed by the
Biogeochemistry Wbrking Group.
Question Type of Evidence
(1) • whole-lake burdens of anthropogenic metals (g/m2 or g/m2/yr) in lakes
receiving similar atmospheric inputs, but differing in
[watershed/lake] area ratios
(2) • metal concentration profiles in undisturbed soils
• laboratory or field experiments in which soil columns have been
subjected to simulated or real precipitation events, and the chemical
ccnposition of the percolate followed over time
• regional surveys of soils, vegetation and/or surface waters in areas
of relative geologic homogeneity, along a gradient in precipitation
chemistry (increasing acidity)
• metal concentration profiles in lake sediment cores
• experimental acidification of streams, lakes
(3) • thermodynamic calculations and experimental determinations of the +
equilibrium distribution of various metal forms, as a function of [H ]
in the pH range 1 + 4 (assumption that bioavailability increases as
the contribution of free aquo ion, M^O) , increases)
(4) • laboratory studies of the kinetics of microbial transformations of
metals, as a function of [H ]
• regional surveys of metal transformation rates in situ, along a
gradient in sediment / water chemistry (increasing acidity)
(5) • regional surveys of water quality in public water supply systems and
in private (dug or drilled) wells
• evaluation of water quality changes occurring in water distribution
systems with and without flews
12
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cell/medium interface. Depending upon the pK values of the functional groups
involved in metal binding and/or transport, t$e hydrogen ion may ccmpete with
the metal for binding sites at the cell surface and thus play a protective role
analogous to the well-documented protection afforded by the hardness cations. A
decrease in pH can thus be expected to affect both metal speciation in solution
and biological sensitivity at the level of the cell surface. These two
responses to acidification are inherently antagonistic.
Acidification may also indirectly influence metal bioavailability. For
exanple, acidic deposition enhances sulfate reduction and denitrification in
anaerobic habitats (Kelly et al., 1982). Of these, only sulfate reduction would
have a significant effect on metal availability. Hydrogen sulfide, the product
of bacterial sulfate reduction, forms extremely insoluble precipitates with many
metals. This would effectively enhance migration of metals from the lower part
of tl'iO- water column to the sadimETit-snd reduce bioavailability .of thejnetal,
provided conditions remained anoxic. If aerobic conditions developed, the metal
vrould be released by bacterial oxidation of the sulfide. With mercury,
acidification may also influence metal bioavailability indirectly by affecting
the microbial transformations of the metal.
Clearly there is no simple answer to the question "Does acidification
affect metal bioavailability?". Vfe have limited our response by considering
only the biogeochemical aspects, i.e., the effect of acidification on metal
speciation in solution. Hie Bioaccumulation Working Group has considered many
of the other dimensions of the problem.
2.2 MERCURY
2.2.1 Importance of Direct Atnpspheric Deposition to a Lake vs. Watershed
Weathering — The atmospheric deposition rate of mercury to lakes"
generally exceeds watershed sources except under geologically unusual conditions
(Hg-enriched bedrock). Indigenous sources of mercury in lakes sensitive to
atmospheric deposition are generally small.
2.2.2 Mercury Mobilization fran Watershed and Sediments — Data for
mercury are inconclusive. Aimer et al. (1976) revised data for lake waters in
Sweden showing an increase in the total concentration of mercury (unfiltered
samples) as the lake pH decreased. In a subsequent publication, Dickson (1980)
noted that this trend applied to lakes with different pH values, subjected to
similar atmospheric loadings of mercury. However, this similarity of
atmospheric loadings, a key point if the data are to be interpreted
unambiguously in terms of acidification + mobilization frcm the watershed and/or
sediments, could not be verified as the relevant atmospheric deposition data
were not presented. Recently Lindqvist et al. (1984) concluded that reliable
data on the concentration of mercury in lake water in Sweden are virtually
nan-existent, casting doubt on the reported trend towards increasing mercury
concentrations with decreasing pH.
2.2.3 Influence of Acidification on Mercury Speciation — Chemical
equilibrium models can be used to calculate the distribution of mercury among a
variety of representative inorganic and organic ligands at various pH values.
Such simulations indicate that mercury speciation in inorganic media is very
pH-sensitive over the pH range 7 + 4. At circumneutral pH values Hg (OH)_ 3™*
HgClClr are the dominant forms, whereas as the pH decreases HgCl2° and HgCl
13
-------
increase in importance; even at pH 4, however, less than 1% of the dissolved
mercury is predicted to occur as the divalent aquo ion, Hg(H-O) a. The
biological significance of these speciation changes in unclear Since the
relative bioavailability of these different mercury forms is unknown.
The only experimental verification of these predictions performed
lander field conditions is that carried out in limnocorrals in the Experimental
Lakes Area, Ontario, Canada (Jackson et al., 1980; Schindler et al., 1980).
Lowering the pH within the enclosures frcm 6.7 ¦+• 5.7 5.1 decreased the ratio
of dissolved to particulate mercury but did not affect the rate of mercury loss
frcm the water column.
2.2.4 Effect of Acidification on Microbial Transformations of Mercury —
Mercury undergoes several microbially mediated transformations and.has been
studied mcrs-thsr, any othsr-hsavy: metals; Acidification affeetsthase
transformations, although the nature of the pH effects is largely unclear.
Three transformations of mercury (methylation, demethylation, and reduction) are
biologically mediated and will be discussed separately.
Methylation. Mercury methylation is predominantly a microbial process
that exhibits greatest activity in surficial anoxic sediments.
Although acidification has generally been thought to enhance
methylation (Fagerstran and Jernelov, 1972), recent evidence suggests
that net methylation is inhibited with decreasing pH (J.W.M. Rudd,
personal caimunication; Staff an and Winfrey, 1984).
This is, in contrast to a growing amount of evidence showing an inverse
relation between lake pH and mercury burdens in fish (Jernelov et al.,
1975; Tsai et al., 1975; Brouzes et al., 1977; Jernelov, 1980;
Tcmlinson et al., 1980; Akielaszek and Haines, 1981; Stokes et al.,
1983; Wiener, 1983; Lindqvist et al., 1984). Thus, the explanation
for the increased mercury in fish must lie in factors other than
enhanced bicmethylation. Possible mechanisms include altered
partitioning of methylmercury between sediments and water (Miller and
Akagi, 1979), increased gill permeability (Pagenkopf, 1983), decreased
trophic state of the lake (Hakanson, 1980), and decreased growth rate
of fish (Jernelov, 1980).
A second contradiction appears when the effects of organic nutrients
on methylation are examined. Increases in labile organic natter
stimulate methylation (Bishop and Kirsh, 1972; Shin and Krenkel, 1976;
Akagi et al., 1979; Furutani and Rudd, 1980; Wright and Hamilton,
1982), yet there is an inverse relation between the trophic state of a
late and mercury burdens in fish (D'ltri et al., 1971; Akielaszek and
Haines, 1981; Lindqvist et al., 1984).
Done thy lation. Methylmercury, once formed, may also be demethylated
by microorganisms (Billen et al., 1975; Jernelov et al., 1975; Kudo
et al., 1978; Jernelov and Martin, 1980). Little is known, however,
concerning this process or the effect of acidification on it.
Mercury Ion Reduction. Mercury ion is readily reduced to volatile
elemental mercury by mercury-resistant bacteria. This can result in
14
-------
release of mercury from an aquatic ecosystem. Steffan and Winfrey
(1984) showed that volatilization was not inhibited by a decrease in
pH in surface lake sediments. Little is known, however, about the
process of mercury reduction in nature or the effects of acidification
on it.
2.2.5 Effect of Acid Water and Increased Atmospheric Loading on tfercury
Levels In Water Distribution Systems — There are no sources of
mercury in water distribution systans; consequently, concentrations of this
metal are determined before entry into the distribution system, i.e., by the raw
water quality. Human exposure to mercury via potable water supplies will be
negligible because mercury concentrations are extremely low, even in lew pH
waters.
2.3 1EAD
2.3.1 Importance of Direct Atmospheric Deposition to a Lake vs. Watershed
Weathering — Dillon and EVans (1982) demonstrated that total lake
lead burdens (g/m2/y and g/m2) varied only by a factor of 1.5 in lakes in the
sane geographical region but with Various [watershed/lake] area ratios. This
suggests little transport of lead frcra the drainage basin to the lake. Lead
concentrations in precipitation are much greater than those in lakes and inlet
streams to lakes (Norton et al., 1981b; Kahl and Norton, 1983). Lead
concentrations in recent sediments greatly exceed those in older "pre-cultural"
sediments (many authors). Therefore, lead inputs to the lake water column are
primarily from direct atmospheric deposition to the lake surface; inputs via
tributaries or inflow will be negligible, because lead is retained in
terrestrial soils by various mechanisms.
2.3.2 Lead Mobilization from Watershed and Sediments — Data for lead are
conclusive: lead is clearly the least"mobile of the four metals considered.
In laboratory and field experiments involving soil columns subjepted
to simulated or real precipitation events, lead was immobile, even when the
columns were subjected to leaching with dilute acids (Fuller et al., 1976;
Tyler, 1978). In a prolonged field study, Tyler (1981) quantified the amount of
lead leached fran the A-horizon of a podzolic spruce forest in southern Sweden;
although scne downward movement of lead occured, the metal was retained in the B
horizon and not lost to the groundwater. For well-drained organic soils in
North America, Hanson et al. (1982) reported that essentially all of the
incoming lead in precipitation was retained in the organic soil layer,
regardless of the pH of the incident precipitation (see also Snith and Siccama,
1981). Andren et al. (1975) developed a lead budget for the Walker Branch
watershed near Oak Ridge, Tennessee; only 2 to 3% of the lead input was exported
fran the watershed by the stream
The paleolimnological data available for lake sediments in Norway
(Reuther et al., 1981), the northeastern U.S. (Hanson et al., 1982), Ontario
(Dillon and Evans, 1982) and Quebec (Ouellet and Jones, 1983) all suggest that
lead is strongly bound to lake sediments. The critical pH for release of lead
fran the sediment is less than the lowest observed water column pH values.
Experimental confirmation of this inmcbility was recently provided by Davis
et al. (1982), who subjected the upper strata fran two sediment cores to
15
-------
acidification in the laboratory; significant lead release (> 5% of total
concentration) occurred only at pH < 3.0. Kahl and Norton (1983) sampled
sediments frcm lakes of pH 6-6.5 and subjected the undisturbed cores and the
overlying water to acidification in the laboratory; significant release of lead
was noted only at pH 4^0, with lead concentrations increasing frcm initial
values of about 1 pg Z to equilibrium values of 5-10 pg I .
2.3.3 Influence of Acidification on Lead Speciation — Computer
simulations indicate that lead speciation should vary" moderately in inorganic
media over the pH range 7 + 4. At circumneutral pH values, the PbCO° and PbOH+
complexes+are predicted to exist in solution (> 50%), but belcw pH 6 the
Pb(H_0)a aquo ion dominates (> 90%). Ccnparable calculations for media
containing model organic ligands, similar to those used by Sposito (1981) to
represent the acid-base and metal-carplexing behavior of fulvic jacid, do not
reveal additional^ pE dependence of lead ccrplexation in the-presence of these
ligands. However, experimental measurements of lead speciation in synthetic
solutions containing humic acid do indicate an enhanced pH sensitivity in the
presence of natural organic matter over the critical region 7 + 4 (Guy and
Chakrabarti, 1976; O'Shea and Mancy, 1978); as would be anticipated, the degree
of ccmplexation decreased at the lower pH values.
2.3.4 Effect of Acidification on Microbial Transformations of Lead —
Several investigations have demonstrated bianethylation of lead (Wong et al.,
1975; Huber and Schmidt, 1976; Thompson and Crerar, 1980), although recent
studies have questioned the validity of these reports (Reisinger et al., 1981;
Jarvie et al., 1983). Chemical methylation, however, is known to occur (Ahmad
et al., 1980; Jarvie and Whitmore, 1981). Lead concentrations in fish were
greater in naturally acidic than circumneutral clear-water lakes in northern
Wisconsin (Wiener, 1983), although no data are available on the effect on pH on
lead methylation. Hie reported lead burdens in fish or waters subject to
cultural acidification do not appear to impose any human health hazard.
2.3.5 Effect of Acid Water and Increased Atmospheric Loading on Lead
Levels in Water Distribution Systems — Total lead concentrations in
surface and groundwaters are typically less than 2 pg (many surveys). In
contrast, elevated lead concentrations (up to 200 pg ) have been found in
domestic water, allowed to equilibrate overnight in the household distribution
system (Hultberg and Johansson, 1981; Norton et al., 1984). Thus the distri-
bution system is dominant in determining lead levels. However, secondary
effects of acidified water (e.g. sulfur compounds) may contribute to this
mobilization (see Sections 4.3.1.1 and 4.3.1.3).
2.4 ALUMINUM
2.4.1 Importance of Direct Atmospheric Deposition to a Lake vs. Watershed
Weathering — Hie atmospheric input of aluminum to lakes is small,
relative to groundwater and surface water inputs derived frcm watershed
weathering.
2.4.2 Aluminum Mobilization frcm Watershed and Sediments — The increased
geochemical mobility of aluminum in response to environmental acidification has
been shown in numerous studies, including laboratory experiments involving the
leaching of soil columns (Fuller et al., 1976), field studies of soil percolates
16
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(Tyler, 1981), regional surveys of soil, groundwater and surface waters (Aimer
et al., 1978; Cronan and Schofield, 1979; Dickson, 1980; Johnson et al., 1981),
the experimental acidification of lakes (Schindler and Turner, 1982) and streams
(Hall et al., 1980), and the experimental acidification of sediments (Kahl and
Norton, 1983).
2.4.3 Influence of Acidification on Aluminum Speciation — Chemical
equilibrium calculations indicate that the distribution of dissolved aluminum
among its various inorganic complexes (hydroxo- and fluoro-carplexes) should be
highly sensitive to pH changes. At neutral pH values, the predominant species
is Al (OH) whereas at pH < 4 the trivalent aquo ion, Al(H-0)n , daji^nates.
In the critical pH range-frem 7 to 4 both mononuclear carplexes (AlOH
Al(OH>2 , Al(OH) _ , AlT , AlF„ ) and polynuclear ocrplexes (Al (OH) )
have been suggested as dissolved -species. Introducing representative organic
iigands into the model system increases the degree of aluminum caiplexation at
pH ^ 7, but has relatively little effect at pH 4. The pH sensitivity of
aluminum speciation, already appreciable in the inorganic system, is thus
predicted to be even more marked in the presence o|_organic l^ands. The
increased contribution of species such as Al(H_0)n and AlOH at the lewer pH
values would be expected to increase bioavailability (Driscoll et al., 1980;
Heliwell et. al., 1983), provided there is not an antagonistic interaction with
the hydrogen ion at the cell surface.
2.4.4 Effect of Acidification on Microbial Transformations of Aluminum —
No biologically mediated transformations of aluminum have been reported.
2.4.5 Effect of Acid Water and increased Atmospheric loading on Aluminum
Levels In Water Distribution Systems — If no sources of aluminum
are present in water distribution systems, concentrations of this metal will be
determined by the initial water quality before into the distribution system.
However, food preparation with aluminum containers /utensils may introduce large
amounts of aluminum to food fluids; the resultant aluminum concentrations may be
orders of magnitude greater than those in acidified surface and ground waters
(see Sections 4.4.1.1 and 4.4.1.3).
2.5 CADMIUM
2.5.1 Importance of Direct Atmospheric Deposition to a Lake vs. Watershed
Weathering — Whole-Take taurdens of anthropogenic "c^dSTum
calculated on an areal basis for softwater lakes in southern Ontario (Evans
et al., 1983), ranged over a factor of 2 and were unrelated to the ratio of
watershed area to lake area, or to the water replenishment time of the lakes.
The authors concluded that atmospheric deposition to the lakes' surfaces was an
important input of cadmium to the lakes. Watershed contributions will also be
important, however, particularly in areas receiving acidic deposition (see
Section 2.5.2).
2.5.2 Cadmium Mobilization fron Vfetershed and Sediments — In laboratory
experiments involving the leaching of soil columns with synthetic rain water
(Fuller et al., 1976; Tyler, 1978), cadmium showed moderate mobility (similar to
Zn; less than Al, greater than Pb; i.e., Al > Cd ^ Zn > Pb). Field studies con-
firmed the sensitivity of cadmium mobility to minor changes in the pH of the soil
17
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percolate (Tyler, 1981); cadmium concentrations in soil leachates were
positively correlated with hydrogen ion concentrations.
Total cadmium concentrations in lake waters in Sweden (unfiltered
samples) increased in lew pH lakes (Aimer et al., 1978; Dickson, 1980),
suggesting that cadmium mobility increases with acidification. However, as
noted earlier for mercury (Section 2.2.2), atmospheric loadings of cadmium to
these lakes were not reported. It is thus difficult to distinguish between
"acidification + mobilization" and "acidification + concomitant increased
atmospheric deposition" as possible explanations for the increase in cadmium
concentrations at lewer pH. Cadmium concentration profiles in sediment cores
frcm several Norwegian lakes show a decrease in concentration in the most recent
sediment strata (Reuther et al., 1981); these data have been interpreted as
mobilization of cadmium (and zinc) frcm the surficial sediments in lakes where
thepH of the overlying water- column is presently belcv/4.9. Alternatively, it
nay be argued that there is less efficient sediment retention of the metal in
low pH lakes (Norton et al., 1981; Evans et al., 1983). In either case, cadmium
is clearly mobile in acidified environments.
2.5.3 Influence of Acidification on Cadmium Speciation — Computer
simulations, both for inorganic media and for" synt±e"t"ic solutions containing
representative organic ligands, indicate that cadmium speciation should be pH -
insensitive in the pH range 7-4; the free aquo ion, Cd(H_0) , is predicted to
dominate throughout this pH range. Experimental measurements in natural waters
(Engel et al., 1981), and in synthetic solutions with or without humic acid
(Gardiner, 1974; Qiy and Chakrabarti, 1976; O'Shea and Mancy, 1978), confirm
this relative insensitivity of cadmium speciation to pH changes.
2.5.4 Effect of Acidification on Microbial Transformations of Cadmium —
No biologically mediated transformations of cadmium have been reported.
2.5.5 Effect of Acid Water and Increased Atmospheric Loading on Cadmium
Levels in Water Distribution Systems — Cadmium concentrations, in
surface and ground waters are typically < 1 pg I , well below the reocrrmended
health limits (10 pg l~ ). Contact with water_distribution systems will on
occasionally elevate these values to 1-10 pg H~ (see Section 4.5.1.1). It is
not knewn whether this is due to natural corrosion or to secondary chemical
effects related to water acidification (e.g., sulfur compounds as contrasted
with CO^-charged water).
3. BIQftCCUMULATIQtv" (J.G. Wiener, P.M. Stokes, and R. Sloan)
3.1 GENERAL INTRODUCTICN
A toxin accumulates in an organism if the absorption rate exceeds the
excretion rate. The bioaccumulation of trace contaminants secondary to acid
deposition may result frcm ingestion of contaminated food, direct uptake frcm
water, or both.
The Bioaccumulation Working Group identified mercury as the
contaminant of primary concern with regard to human consumption of fish frcm
acid-sensitive waters. For mercury, consumption of fish is considered the
biological pathway of greatest importance to man. Dietary intake of
18
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methylmercury by man results almost entirely frcn consumption of fish and fish
products (Clarkson, 1983). Accumulation of methylmercury in edible fish flesh
during exposure of fish to mercury in water or food can be substantial (McKim et
al., 1976; Scherer et al., 1975; Huckabee et al., 1978, 1979; Lodenius et al.#
1983; Wfen et al., 1983; Phillips et al., 1980). A substantial amount of the
methylmercury in fresftwaters may be contained within the bicmass of the system
(Clarkson and Baker, 1983).
Concentrations of certain other metals, such as cadmium and lead, in
fish are inversely correlated with pH and related chemical characteristics of
lakes; however, these metals tend to accumulate in tissues and organs not
consumed by man (see Section 3.3). Aluminum can be acutely toxic to fish (via
the gills) in acidic surface waters. Because of this acute toxicity, aluminum
accumulation in fish is not significant from a human health perspective.
3.2 MERCURY
3.2.1 Accumulation of Mercury by Fish — Most of the mercury present in
edible fish flesh (generally > 80%) exists as methylmercury (Huckabee et al.,
1979). Fish cannot me thy late mercury in vivo (Huckabee et al., 1978), although
methylation of mercury in the gastrointestinal tract has been documented (Rudd
et al., 1980). Fish can accumulate methylmercury fron water and food (Phillips
and Buhler, 1978; Phillips and Gregory, 1979; Olson et al., 1973; Scherer et
al., 1975; Rodgers and Qadri, 1982; Turner and Swick, 1983). The direct uptake
of methylmercury fran water occurs almost entirely across the gills (Olson et
al., 1973). Total mercury concentrations in fish are often correlated with
mercury levels in organisms of lower trophic levels (Hultberg and Hasselrot,
1981; Stokes et al., 1983; MacCriimon et al., 1983), providing circumstantial
evidence of the importance of dietary uptake in natural waters. Within a given
fish ccnmunity, piscivorous fish species generally contain the highest mercury
concentrations in edible flesh (MacCrimmon et al., 1983; Lodenius et al., 1983;
Wren et al., 1983; Phillips et al., 1980). A number of North American fislji
species, in which mercury concentrations in muscle have exceeded 1.0 yg g wet
weight, are listed in Table 3-1.
Elimination of methylmercury by fish is very slew relative to the rate
of uptake. Consequently, «ercury Goncentrations in-edible flesh of a fish
species within a given water body generally increase With increasing age or body
size (Huckabee et al., 1979; Phillips and Buhler, 1978).
3.2.2 Effects of Acidification an Mercury Bioaccumulation by Fish —
Accumulation of mercury by fish may increase as a result of acidification.
Within a geographic area, mercury concentrations in edible fish flesh (of a
given species and age) are often inversely related to lake pH or alkalinity
(Table 3-2). Furthermore, total mercury concentrations in edible flesh of
piscivorous fishes inhabiting lew pH waters (pH £ 6.0) frequently exceed 0.5 or
1.0 pg g wet weight. This pattern has been observed in remote lakes having no
local anthropogenic sources of mercury (e.g., Bloamfield et al., 1980).
The relative importance of uptake from food and water in acidic lakes
is not understood. The direct uptake of methylmercury across the gills of fish
is apparently enhanced in waters having low pH and low calcium content (Drummond
et al., 1974; Rodgers and Beamish, 1983). An increased dietary uptake of
19
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table 3-1.
North American fishes
(wet weight)
for which mercury concentrations exceeding 1.0 gg/g
in edible muscle tissue have been reported.
Species
Location
Reference
Largemouth bass, Micropterme Balmoides
Smallmouth bass, Miaropteme dolomieui
Lake trout, Sdlvelinue namayeueh
Northern pike, Fbox lucine
Walleye, Stisoeted-ion vitveum vitveum
South Carolina reservoirs
Tadenac Lake, Ontario
New York State lakes and rivers
Tadenac Lake and other
Precambrian Shield lakes
Northern Maine lakes
Lake Hacksjon, Sweden
Tongue River reservoir, Montana
Swedish lakes
Tadenac Lake, Ontario
Minnesota lakes
Tongue River reservoir, Montana
Minnesota lakes
Southern Ontario lakes
Michigan lakes
Softwater lakes in northern
Wisconsin
Abernathy and Cumbie, 1977
Wren et al., 1983
Bloomfield et al., 1980
MacCriirtnon et al., 1983
Akielaszek and Haines, 1981
Gothberg, 1983
Eftillips et al., 1980
Lindqvist et al., 1984
Wren et al., 1983
Glazer and Bohlander, 1978
Phillips et al., 1980
Glazer and Bohlander, 1978
Scheider et al., 1979
Kelly et al., 1975
Wiener, 1983
-------
TABLE 3-2.
Reported relationships botwoon total mercury concentrations In fish and physlcochemical characteristics of surface waters in North America.
pH range
of waters studied Matrix and Surumry of findings Inference
Location (N)» species analyzed
lakes near Huakoka,
Ontario
S.1-7.5
(14)
Miole, yearling (1+ age)
yellow perch
Mercury concentration in yellow perch was significantly (P < 0.05)
correlated with epillmwtic pH (t* ¦ 0.63), aliimlnun
concentration in eplliimetlc water (p « 0.65), and watershed
area/lake volune ratio (>• = 0.92).
Suns et al., 1980
Lakes in south-central
Ontario
Not reported
300 poq/tTF. Mean mercury
concentrations of large (>48 cm) walleyes in low alkalinity
lakes equaled or exceeded 1.0 Mg/g>
Schelaer et al.,
1979
Lakes in north-central
Wisconsin
5.6-7.0
(5)
Axial muscle tissue of
walleye
Mercury concentrations Here significantly (P < 0.01) greater in
4-, 5-, and 7-year-old fish frcre tw naturally)acidic lakes
5.0. ttrne of the 200 brook trout analyzed
contained mercury concentrations exceeding 1.0, vo/g-
Sloan and Schofield,
1983
Lakes and beys in south-
central Ontario
5.6-8.4
(16)
Axial muscle tissue of
puilfclnseed
Mercury concentration in pnupklnsood was significantly (P < 0.05)
correlated with jjK
-------
methylmercury could also occur in acidic waters, due to increased mercury
accumulation by organisms in lower trophic levels and subsequent passage through
the food chain (Stokes et al., 1983).
Elevated mercury levels in fish inhabiting naturally acidic or
culturally acidified waters have been observed in Scandinavia, Ontario,
Wisconsin, and the Adirondack region of New York (Hakanson, 1980; Hultberg and
Hasselrot, 1981; Blocmfield et al., 1980; Scheider et al., 1979; Wiener, 1983;
Lindqvist et al., 1984; Jemelov, 1980; Sloan and Schofield, 1983). Multiple
correlation analysis of Swedish data by Jernelov (1980) indicated that lake pH
explained about 50% of the variation in mercury levels in 1-kilogram northern
pike. Hultberg and Hasselrot (1981) also evaluated factors affecting mercury
contamination of fish in 152 Swedish lakes. The mercury concentration in muscle
of northern pike was inversely correlated with lake pH, and it was estimated
that a one uriitdecrease in pH wculcj increase iiercury concentration in the
muscle tissue of pike by 0.14 vg g~ .
Several mechanisms have been proposed (Jernelov, 1980; Wood, 1980;
Haines, 1981) as potential explanations of the relation between the acidity of
surface waters and the bioaccumulation of mercury.
(1) Tcmlinson et al. (198j0) hypothesized that acidic precipitation
may scavenge mercury frcm the atmosphere more effectively than
nonacidic precipitation. This hypothesis does not explain the
large differences in the mercury content of fish from acidic and
nearby circumneutral lakes (e.g., Wiener, 1983).
(2) Jernelov (1980) proposed that formation of mononrethy Imercury by
microorganisms is pH-dependent, with the maximum rate occurring
at pH 6.0. Methylation is higher at pH 5.0-7.0 than at higher pH
values. ' In addition, it has been hypothesized that lew pH levels
(pH <_ 6.5) enhance the formation of monarethyImercury, rather
than dimethy Imercury, which is volatile and released frcm the
aquatic ;system to the atmosphere- (Fagerstrcm and Jernelov, 1972).
Thus, at these lower pH levels more moncmethyImercury — the form
most readily assimilated by fish — would presumably be present,
and bioaccumulation would be enhanced. Hcwever, this proposed
mechanism is not supported by recent investigations. In situ
studies by J. W. M. Rudd (pers. conn.) and laboratory studies by
Steffan and Winfrey (1984) have shown that decreases in pH (belew
6.5) slew the net rate of moncmethy la tion of 203„ added to
surficial sediments. Thus, the net rate of moncmetnfImercury
formation is probably not greater at low pH. Furthermore, Miller
and Akagi (1979) showed that methylmercury production was not
pH-dependent in three types of sediment across the pH range 5.0 -
7.5.
(3) Inorganic mercury is more soluble in acidic media (under aerobic
conditions) and more potentially available for methylation
reactions. Retention of mercury in the water column is thus
enhanced at low pH (Jackson et al., 1980), increasing the
exposure of fish to mercury. For example, Miller and Akagi
(1979) performed laboratory experiments on three sediment types
22
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and showed that the. equilibrium ,partition coefficient [MeHg] ../
fMeHg] decreased between pH 7 and 6, and between pH 6 ana 5
for earn 8i the sediment types. They hypothesized that the
effect of pH on methylmercury partitioning between water and
sediment was responsible for the high bioavailability of mercury
in acidic waters. Hcwever, their experiments were conducted with
very high mercury concentrations.
(4) Tcmlinson et al. (1980) proposed two potential mechanisms leading
to increased mercury contamination cf biota in acid waters.
First, reductions in bicmass in acidic lakes would increase
mercury availability per unit weight to the remaining bicmass,
resulting in higher mercury burdens in biota. For example, a
decrease in the'growth rate of fish in an acidified lake would
presumably be acccsnpanied by increased mercury concentration in
tissues —the accumulated mercury being contained in a smaller
bicmass. Second, as a lake acidifies, the forage base for
piscivorous fishes is altered such that they feed on larger fish
with higher mercury content. Bnpirical support for these
hypotheses is lacking.
There are presently no definitive experimental data showing a cause
and effect relationship between acidification and increased mercury accumulation
by aquatic biota. Experimental work on biogeochemical factors enhancing mercury
accumulation by fish in dilute, acidic waters is therefore needed. This need is
especially urgent in the United States, where few data are available.
3.2.3 Effects of Productivity on Mercury Uptake by Fish — Fish in
oligotrophic lakes often have higher mercury concentrations than fish in more
productive systems (D'ltri et al., 1971; Hakanson, 1974). TVo potential
explanations for this have been offered, one relating to mercury methylation and
the other relating to mercury ccrplexation. Akielaszek and Haines (1981)
hypothesized that methylation of mercury is enhanced in oligotrophic lakes.
Hakanson (1980) cited several studies showing a relation between productivity
and suspended organic matter, which provides sites for ccrplexation of mercury,
thus reducing the fraction of mercury in the water column available to fish.
In contrast, organic enrichment of mercury-polluted systems appears to
increase mercury uptake by fish. For exartple, addition of organic matter such
as wood chips, moss, or leaves to enclosures in mercury-polluted Clay Lake in
the Wcibigoon River system in Ontario increased the mercury levels in fish (Rudd
et al., 1983). Even in the absence of gross mercury pollution, the.addition of
organic matter to limnocorrals in the Southern Indian Lake reservoir in northern
Manitoba increased both methylation rates and mercury concentrations in fish
(Bodaly et al., 1984). These results were interpreted as the stimulation of
microbial or chemical methylation of mercury, where the quantity of mercury
present is presumably not the limiting factor. Father, the organic matter
available for methylation may be limiting. This interpretation is supported by
the observations of Hi Hultberg (pers. ccmm.) on fish and zooplankton and by P.
M. Stokes (unpublished data) on filamentous algae. Field surveys by both
Hultberg and Stokes revealed a positive correlation between dissolved organic
carbon in oligotrophic lakes and mercury concentration in biota.
23
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3.2.4 Effects of Chemical Neutralization of Lakes on Mercury
Contamination of Fish — The mercury content of fish has been
statistically related to pHf waterborne calcium, alkalinity, and other
physioochemical variables (Table 3-2). The effect of liming (the addition of
CaCO_, Ca(OH) or dolariite limestone) on mercury levels in fish should therefore
be evaluated. Decreases in the mercury content of fish after liming have been
reported (Suns et al., 1980). However, mercury levels in fish did not decrease
after liming in many Swedish lakes (H. Hultberg, pers. conn.). Without a better
understanding of the mechanisms by which aqueous chemistry affects the
availability of mercury to fish, predictions will remain speculative.
3.3 OTHER METALS
Certain other metals exhibit elevated concentrations in acid waters
and cooid-potentially accumulate in fish. For example, body burdens of cadmium
and lead are inversely correlated with pH and related chemical characteristics
of lakes in sane areas (Kelso et al., 1982; Wiener, 1983; U.S./Canada, 1983;
Kelso and Gunn, 1984). Increasing accumulation of inorganic lead with
decreasing pH has also been shewn in laboratory studies (Merlini and Pozzi,
1977; Hodson et al., 1978b). However, concentrations of cadmium and lead in
edible fish flesh (axial nuscle tissue) do not increase greatly with increasing
exposure concentration (Holccmbe et al., 1976; Jaakkola et al., 1972;
Bollingberg and Johansen, 1979; Sangalang and Freeman, 1979; Westernhagen et
al., 1980). Father, lead and cadmium tend to accumulate in tissues and organs
that are not generally consumed by man, such as the kidney, liver, and bone
(Holoaribe et al., 1976; Patterson and Settle, 1977; Hodson et al., 1978a; Kumada
et al., 1980; Fjerdingstad and Nilssen, 1983).
The extent to which aluminum is accumulated by fish has been little
studied. It is, hewever, known that the mode of aluminum toxicity to fish is
external (on the gills), not internal (Muniz and Leivestad, 1980). Aluminum can
be toxic to fish at waterborne concentrations as low as 0.1-0.2 mg £
(Schofield and Trojnar, 1980; Muniz and Leivestad, 1980; Baker and Schofield,
1982), which is within the range of concentrations measured in many acidic
surface waters. Buergel and Soltero (1983) reported aluminum concentration^ in
muscle tissue of rainbow trout (Salmo gairdneri) between 0.003 - 0.18 yg g dry
weight. Therefore, based on limited results, aluminum apparently"has little
tendency to accumulate in edible fish tissue.
4. INDIRECT HEALTH tar^CTS (T.W. Clarkson, D.A. Bennett, R.A. Gqyer, S. Joshi,
M. McDonald, R.G. Miller, P.D. Moskcwitz,
G. Nordberg, H. Olem, M. Piscator)
4.1 GENERAL INTRODUCTION
The most conclusive evidence of cause and effect relationship between
an environmental factor and human disease is obtained from observations made on
human populations. However, the indirect health effects fron the acidic
Footnote: Parts of this section regarding mercury, lead and drinking water
systems were extracted frcm Chapter E-6 of the Critical Assessment
Document (EPA-600/8-83-016B). Acknowledgement: T.W..Clarkson, J.P.
Baker, and w.E. Sharpe.
24
-------
deposition phenomenon do not provide opportunity for these types of
observations. Numerous factors, contribute to this restriction, the nature of
the contaminants involved and the types of illness that might be expected to be
observed.
Of particular interest are contaminants such as lead, cadmium, mercury
and other trace elements. Virtually everyone in the general population has
already accumulated sane level of these substances in body tissues; further,
there is nearly continuous exposure to these contaminants, although in scane
instances very small, from multiple sources, and are both natural and
anthropogenic. All trace contaminants have known toxic effects on man. But
with few exceptions (heme effects of lead) the expected health effects include
camion ailments experienced by the general population such as behavioral and
neurologic diseases, renal disease, reproductive disorders, neoplasia and other
common disorders. -Ebr each of these toxicants, -expected dose-response
relationships are established and surveillance of human populations permits
protection from expected health effects by elimination of likely sources.
The potential contribution of acidic deposition to human disease lies
in the role it may have in increasing human exposure to trace substances,
thereby making control of human exposure more difficult (contamination of water
supplies), or even so restrictive of life style as to impose other undesirable
effects (contamination of food supplies). If such extremes are permitted it may
be exceedingly costly both monetarily and healthwise to reverse such effects.
Therefore, investigation is warranted of the contributions of the acidic
deposition to human exposure to these contaminants.
The metals and other contaminants selected for discussion were chosen
because they may present a potential human health hazard. The toxicology of
these substances - mercury, lead, cadmium, aluminum, asbestos and nitrates. - was
discussed by the working group in sane detail. Mercury, in the form of
methylmercury, is toxic to the central nervous system in both the human adult
and fetus. Lead also can produce serious effects in the nervous system, as well
as in the hemopoietic system and kidneys. Cadmium has an uncoimon capacity to
accumulate in the human body virtually over the entire human lifespan with the
kidney as the major site of damage for cadmium. Aluminum continues to be
inplicated in degenerative diseases of the central nervous system. Asbestos,
when inhaled, is a potent carcinogen. Nitrates are known to cause
methemoglobinemia in infants. The Health Effects Working Group discussions
include the effects of these agents at different stages of the human life cycle.
The potential for prenatal damage was noted for methylmercury and for lead.
Cadmium may affect the function of the placenta.
These health effects will be described in further detail in this
report. Hie Human Health Effects Working Group recognized the potential for
direct health effects, however, believes that the main interest at this time is
in the potential for acidic deposition to mobilize these substances and thereby
increase human exposure. If future research identifies large increases in human
exposure due to acidic deposition, then the subject of health effects including
dose-effect and dose-response relationships will come sharply into focus.
Hie postdepositional pathways of human exposure are principally intake
frcm food and drinking water. Hemodialysis is an exception and is of interest
25
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with regard to exposure to aluminum. The mechanisms of mobilization of toxic
substances into food and drinking water are discussed by the Biogeochemistry and
Bioaccumulation working groups and will only be discussed here, where
appropriate.
Food. Mobilization of toxic substances into human food might occur
via the aquatic food chain. The accumulation of methylmercury in
edible tissues of fish and the relationship to acid deposition is
discussed by other working groups (see Sections 2. and 3.).
Information is sparse on mobilization into terrestrial food chains.
Human exposure might be affected by the state of combination of toxic
substances in food. For exairple, cadmium in certain edible animal
tissues nay be bound to metallothionein and in this form be
distributed within the human body in a different way frcm cadmium not
bound to metallothionein.
Drinking Water. An understanding of the modes of hydrologic
interactions between acid deposition and various types of water supply
systems is essential to assessing the potential indirect health
effects to users of drinking water obtained frcm each type of system.
In addition, the physical facilities used to store, treat, and
distribute water are of primary importanoe, as are the chemical
methods used to treat water prior to use. Principal water sources in
continental North America are usually either surface or groundwater,
with other sources such as direct use of precipitation of much lesser
importance. Health risk is directly related to the source of drinking
water.
Health risk in drinking water supplies is also closely related to the
management of the drinking water supply. Risks are generally greater
the smaller the water supply, with small privately owned water systems
serving a single dwelling at greatest risk. These systems typically
do not routinely monitor water quality nor do they provide even
rudimentary water treatment. Data on the impacts of atmospheric
deposition on drinking water quality are extremely scarce; hcwever, by
using available information on the impacts to surface water aquatic
ecosystems, we my assess impacts.
Precipitation was directly used for water supply in North America frcm
very early times. Its use is still carmen where there are no other
water supply alternatives. In the equatorial regions of the world,
island carrnunities still rely heavily on rainwater cisterns to supply
their freshwater needs, and this method of water supply is being
seriously considered as appropriate technology to supply drinking
water for the developing countries of the world.
Since, in most systems, precipitation is used directly with no
treatment, the quality of precipitation and the amount of dry
deposition on the catchment between precipitation events are of
paramount importance to the quality of drinking water at the user's
tap. The major impacts are twofold: (1) direct deposition of
atmospheric pollutants such as lead and copper may occur, and (2) the
26
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acid components of atmospheric deposition may cause increased
corrosion of metallic plumbing system canponents, as well as the
release of asbestos fibers fran sane types of roofing material into
cistern water supplies.
Surface water supplies may be impacted by acidic deposition in two
ways: (1) the quality of the source water may be impaired, and (2)
increases in the corrosivity of the water could increase corrosion of
metal plumbing systems. Corrosivity is probably the most significant
potential concern related to atmospheric deposition on surface water
supplies. The corrosivity of the dilute water often used for surface
water supplies in the northeastern United States is mostly controlled
by hydrogen ion concentration. As the hydrogen ion concentration
increases so does the corrosivity of the water.
Corrosivity in surface water supplies has been widely reported, and
its impacts are well documented. Where lead water distribution pipes
are in use, clinical lead poisoning of children has been reported as a
consequence of corrosive drinking water conveyance. A notable exanple
of such a problem is Boston, Massachusetts. Less well known is the
case of Mahanpy City, PA (Kuntz, 1983). A case of copper toxicity
frcm a corroded water fountain has also been reported by Semple et
al., (1960). Where pipes are of other metals such as copper, iron, or
galvanized steel the corrosion products of copper, lead, iron, zinc,
and cadmium can affect water quality. Where asbestos/cement piping is
used in central water distribution systems, asbestos fibers may appear
in the water. Because corrosion can lead to elevated concentrations
of toxic contaminants in drinking water, the U.S. Environmental
Protection Agency (1979a) has reccrmended that all drinking water
supplies be noncorrosive and that a minimum pH of 6.5 be maintained.
Taylor et al., (1984) in a study of acidic deposition and drinking
water quality in the eastern United States, found that generally when
the pH of raw water was lower than 6.5 the drinking water standards
for trace contaminants (U.S. EPA, 1979b) were not met.
Acidification of groundwater as a consequence of atmospheric
deposition has been reported in Sweden by Hultberg and Wenblad (1980),
such changes have not as yet been well documented in North America.
The need for documentation is greatest with respect to groundwater
impacts frcm atmospheric deposition. Although additional work is
indicated, preliminary information seems to indicate that adverse
impacts to drinking water quality are possible in water supplies using
shallow groundwater in areas edaphically and geologically sensitive to
atmospheric deposition.
4.2 MERCURY
4.2.1 Environmental Exposure and Uptake by Humans
4.2.1.1 Concentrations in Water — Bodies of fresh water having no kncwn
source of mercury generally yield values less than 200 ng I . Most values fall
in the range of 10 to 40 ng l~ and drinking water usually has values less than
30 ng 1 (WHO, 1976).
27
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Few reports exist on the speciation ot mercury in water, probably
because of analytical difficulties. A recent review by McLean et al., (1980)
found that methylmercury accounted tor a small fraction of the total
approximately in the order of 1 percent. However, a more recent report by Kudo
et al., (1982) found that methylmercury accounted for about 30 percent of total
inerc^ry in samples taken fran Canadian and Japanese rivers. Mercuric mercury
(Hg ) accounted tor about 50 percent.
Two important conclusions may be drawn fran these data. First, that
precipitation is an important source of mercury to fresh water, and second, that
mercury in drinking water offers no health threat (see Section 2.2.1).
Concentrations on the order ot a few hundred nanograms per liter would result in
a negligible intake of mercury on the assumed intake of two liters per day (U.S.
EPA, 1980a). This intake, less than.2 jjg day- , is well belcw the advised
maximum safe intake of 30 ?.ig Hg day (WHO, .1972)thus, additional mobilization
of mercury into water by acidic deposition should not pose a health threat in
terms of contaminated drinking water.
4.2.1.2 Concentrations in Aquatic Biota — The concentration of me thy1-
mercury in fish tissue is of special interest in terns of human exposure.
Bioaccumulation of methylmercury in fish is the principal if not the sole
source of human exposure, barring episodes of accidental discharge or misuse of
man-made methylmercury canpounds. Thus, factors that affect concentrations of
methylmercury in edible fish tissue are of considerable importance in assessing
potential human health risks fran this form of mercury (see Sections 2.2 and
3.2).
4.2.1.3 Human Uptake and Risk Assessment — Dietary intake accounts tor
the greatest fraction of total mercury intake~by man (Table 4-1). Methyl
mercury intake is exclusively fran the diet and almost entirely tron tish and
fish products. The evidence canes fran dietary studies shewing close
correlation ot blood levels with fish consumption (Swedish Expert Group, 1971)
and tron large-scale analyses of food items in several countries, indicating
that significant concentrations of methylmercury are found only in fish and
fish products (U.S. EPA, 1980a).
Based on data fran the National Marine Fisheries, Oordle et al.,
(1979) have reported a ranking of species of fish according to annual
consumption in the United States (Table 4-2). The table clearly demonstrates
that oceanic fish, especially tuna, account tor the major amount consumed.
However, when consumption is expressed according to the consumer use, a
different picture emerges. Cn this basis, freshwate^ fishes daninate the
rankings, with northern pike consumed at 17.4 g day , followed by trout ,
(freshwater) at 12.3 g day , bass (freshwater) and cattish at 12.1 g day .
The Ijdghest per user consumptions oj seafood are crabs and lobster at 10.6 g
day , with tuna dewn to 6.1 g day .
The highest average mercury concentrations are also found in
freshwater tish - northern pike at 0.61 pg Hg g wet weight and trout at 0.42
pg Hg g~ wet weight. Thus a northern pike consumer would have a daily average
intake of methylmercury ot 10.4 pg exclusively tron northern pike, and a trout
consumer would have had an average intake ot 5.2 pg Hg. These average values
are well below the recattnended maximum safe intake of 30 pg Hg day .
28
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TABLE 4-1. ESTIMATES OF AVERAGE AND MAXIMUM DAILY INTAKES OF MERCURY
BY TOE "70 kg MAN" IN THE' UNITED' STATES POPULATION
(ADAPTED FROM U.S. EPA 1980a)
Media Mercury intake pg day ^ 70 kg ^ Predominate
(average) form
Air 0.3 Hg°
Water 0.1 Hg2+
Pood 3.0
CH3Hg+
29
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TABLE 4-2. ESTIMATED FISH AND SHELLFISH CONSUMPTION IN THE UNITED STATES
STATES RANKED ACCORDING TO ANNUAL CONSUMPTION FOR THE PERIOD
SEPTEMBER 1973 TO AUGUST 1974 (ADAPTED FRCM U.S. EPA 1980a
AND CQRDLE ET AL. 1979).
Rank 106
Amount,
lb yr~
Percent of
total by
weight
Number of
actual users
(millions)
Mean Amount
per user,
(g day )
Average cone
of mercury
\iq Hg g
Total
2957
100.0
197.0
18.7
w
Tuna
0.14
(mainly canned)
1
634
21.4
130.0
6.1
0.27
Unclassified
0.35
(mainly breaded,
including fish '
sticks)
2
542
18.4
68.0
10.0
c
Shrimp ,
3
301
10.2
45.0
8.3
0.05
Ocean Perch
4
149
5.0
19.0
9.7
0.13
Flounder
5
144
4.9
31.0
8.6
0.10
Clams
6
113
3.8
18.0
7.6
0.05
Crabs/lobsters
7
110
3.7
13.0
10.6
0.07-0.14
Salmon
8
101
3.4
19.0
6.7
0.08
Oysters/scallops
9
88
3.0
14.0
7.8
0.03
Trout
9
88
3.0
9.0
12.3
0.42
Cod ,¦
11
78
2.7
12.0
8.1
0.14
Bass c
12
73
2.5
7.6
12.0
c
Catfish,
12
73
2.5
7.5
12.1
0.15
Haddock,
12
73
2.5
11.0
8.6
0.11
Pollock
15
60
2.0
11.0
6.8
0.14
Herring/smelt
Sardines
16
54
1.8
10.0
6.7
0.02
Pike ,
17
35
1.2
2.5
17.4
0.61
Halibut
18
32
1.1
5.0
8.0
0.19-0.53
Snapper
18
32
1.1
4.3
9.3
0.45-369
Whiting
20
25
0.9
3.2
9.7
c
All other
Classified
152
5.1
c
^.S. Chamber of Commerce (1978).
^Average values for skipjack, yellow fin, and white tuna, respectively.
°Data not available.
^Mainly imports.
eKing crab - all others, respectively.
^Fresh Water.
30
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The National Marine Fisheries Service developed an extensive data bank
on fish consumption by individuals according to fish species (U.S. Department of
Ccnmerce, 1978). These data were based on a Diary Panel Survey of approximately
25,000 individuals chosen to be representative of the U.S. population. These
data, along with additional information on mercury concentration of edible
tissues of various fish species, allowed a calculation of the number of
individuals who would be expected to exceed the maximum safe daily intake of 30
pg. It was calculated that 47 of the 25,000 individuals would exceed this limit
by a small margin frcm consumption of fish and that 23 of these were consumers
mainly of freshwater fish. According to calculations by Nordberg and Strangert
(Figure 4-1) the risk of parathesia (see Section 4.2.2.2) at this level of
intake will be small — on the order of 0.3 percent.
4.2.2 Effects on Human Health
4.2.2.1 Metabolism in Humans — The U.S. EPA (1980a) has reviewed
information on uptake, distribution, and excretion of methylmercury in man.
Methyl mercury is almost ccnpletely absorbed frcm the diet ( i.e., between 90 to
100 percent of the amount ingested is absorbed). After absorption in the
gastrointestinal tract, methylmercury passes into the bloodstream and is
distributed to all organs in the body. Approximately 5 percent of the absorbed
dose goes to the blood compartment and 10 percent to the brain — the target
organ for toxic effects.
After the initial distribution is completed, usually a matter of a few
days in man, the brain to blood concentration ratio is roughly constant, having
values between 5:1 to 10:1. Methyl mercury is accumulated in growing hair. At
the time of formation of the head hair, the ratio of the concentration of
mercury in hair to the simultaneous concentration in blood is roughly constant
and has an average value of about 250:1. Once incorporated into the hair, the
mercury concentration remains constant. Because human head hair grows about 1
centimeter per month, analyzing centimeter segments of hair can recapitulate
average monthly concentrations of methylmercury in blood. Measurements of
mercury in sanples of blood or hair are now routinely used to assess the body
burden of methylmercury in humans and as an indicator of brain concentrations.
Methyl mercury is excreted fran the body mainly in feces. Before
excretion in the feces, methylmercury is converted into inorganic mercury. The
site of this conversion is not known, but microflora in the lower gut are known
to possess this capability. The rate of elimination fran the body is directly
proportional to the body burden. It is well described by a single exponential
function characterized by a half-time of about 70 days. An important conclusion
from this kinetic information is that it will take about one year for humans to
attain a state of balance,i.e., to attain maximum steady, body burden of methyl-
mercury for any given daily intake in the diet. After cessation of given
exposure, it will take one year for the body burden to fall to pre-exposure
levels. Thus, dietary intake of methylmercury fran fish should be evaluated
over a matter of months. Intake on any one single day does not normally make an
important contribution to the overall body burden.
Considerable individual differences exist in biological half-times in
nan although the average value is 70 days with a range of 30 to 180 days. The
distribution is birrodal with 90 percent of the values distributed about an
31
-------
100
LU
o
a.
V)
LU
DC
a
UJ
»-
o
UJ
IXl
background
0.4 1.0
DAILY INTAKE.(mg)
Figure 4 -1. The calculated relationship betvveen frequency of
paresthesia in acquits and long-term average daily
intake of methyl mercury. The calculations were
performed by Nordberg and Strangert (1978). The
broken line is the estimated background frequency of
paresthesia in the population. Data are taken from
publications on the Iraqi outbreak of methyl mercury
poisoning (Bakir et al. 1973; Shahristani & Shihab
1974).
32
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average value of about 65 days and 10 percent distributed about an average value
of 120 days. The reasons for this wide range of biological half-times are not
kncwn, except that lactating women have a short half-time averaging about 40
days.
Methylmercury readily crosses the placental barrier and enters the
fetus. It distributes to all tissues in the fetus, including the fetal brain,
which is the principal target for prenatal toxicity of methylmercury. Levels
of methylmercury in cord blood are usually higher than the maternal blood
concentrations.
Methylmercury is secreted in milk. Thus body ,burdens of methyl-
mercury acquired by the infant before birth may be maintained by breast feeding
if the nursing mother continues to be exposed to methylmercury.
Hie rate of elimination of methylmercury frcm the human fetus and
suckling infant is not kncwn. Experiments on animals indicate that elimination
in suckling animals is much slower than in adults. The adult rate of excretion
appears to ccsrmence at the end of the suckling period.
In brief, methylmercury accumulates in the human body over a period
of about one year. Blood and hair analyses may be used as indicators of human
absorption of mercury. In assessing hazard to human health, chronic exposure
over weeks or months is important.
4.2.2.2 Effects and Dose-Response in Humans — Ms thy 1 mercury damages
primarily the human central nervous system. When ingested in sufficient
amounts, methylmercury destroys neuronal cells in certain areas of the brain,
the cerebellum and the visual cortex, resulting in permanent loss of function.
Symptans of damage include loss of sensation, constriction of the visual fields,
and impairment of hearing. Coordination functions of the brain are also
damaged, leading to ataxia and dysarthria. Severest damage causes mental
incapacitation, ccma, and death. The mildest and earliest effect in adults is
usually a ccnplaint of paresthesia, an unusual sensation in the extremities and
around the mouth. In the Japanese population poisoned by methylmercury fron
contaminated fish, paresthesia was usually permanent. In the Iraqi population,
paresthesia was frequently reported tobe transient. This population had
consumed homemade bread fron wheat contaminated with a methylmercury fungicide.
The effects on the fetal brain differ qualitatively frcm those seen in
adults. Methyl mercury interferes with the normal growing processes of the
brain. It inhibits migration of neuronal cells to their final destination, thus
affecting the brain's architecture. This damage manifests itself as diminished
head size (microcephaly) and gross neurological manifestations such as cerebral
palsy. The mildest effects are delayed achievement of developmental milestones
in children and the presence of abnormal reflexes and mild seizures.
Brain concentrations associated with the onset of human methylmercury
poisoning are in the range of 1 to 5 ug Hg g~ wet tissue. Blood concentrations
for the onset ..of the mildest effects have been established to be between 200 and
500 ng Hg ml" whole blood. Corresponding hair concentrations would be 50 to
125 ug Hg g~ hair (Table 4-3). The chronic daily intake of methylmercury that
would lead to a maximum blood level of 200 ng ml has been established to be
33
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TABLE 4-3. THE CX3NCTMTRATI0NS OF TOTAL MERCURY IN INDICATOR MEDIA AND
METHYL MERCURY ASSOCIATED WITH THE EARLIEST EFFECTS IN THE
MDST SENSITIVE GROUP IN THE ADULT POPULATION3
(ADAPTED FROM WHO 1976)
Concentrations in indicator media
Blood. Hair. Equivalent long-term daily intake
(ng ml ) (yg g ) (pg kg body weight)
-200 to 500 50 to 125 3 to 7
The risk of the earliest effects can be expected to be between 3 to
8 percent, i.e., between 3 to 8 percent of a population having blood
levels in.the range 200 to 500 mg ml" , or hair levels between 50 to
125 pg g~ would be expected to be affected (for further details,
see text).
34
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300 yg Hg. However the blood level in the mother associated with the earliest
damage to the fetus has not yet been determined.
A major epidemiological study was carried out in Northwestern Quebec
on Cree Indians exposed to methylmercury in freshwater fish (Methyl Mercury
Study Group,, 1980). The authors claim to find an association in men over age 30
and wcmen over age 40 of a set of neurological abnormalities and the estimated
exposure to methylmercury. However, it should be pointed out that this
association has been seen by only four of seven observers who reviewed video
taped recordings of the neurological screening tests. The severity of these
neurological abnormalities was assessed by neurologists as mild or questionable.
It was not possible to estimate any threshold body burden or hair levels because
this population had been exposed possibly for most of their lives; therefore
peak values in previous years are unknown. However, observations on this
papulation over_several years indicate thai- nasdnam blood concentrations are
be lew 600 ng ml and most below 200 ng ml (Vftieatley, 1979). A WHO expert group
(1980), on examining the reports fran these studies, raised the possibility that
this might be the first example of an endemic disease due to exposure to methyl-
mercury in freshwater fish. However, another epidemiological and clinical study
of the same population of Cree Indians failed to find any effects associated
with methylmercury (Kaufman, personal ccmnunication to EPA), which would
indicate that the background frequency of mental retardation would be increased
by less than 50 percent.
Estimates of increased rates risks due to acid precipitation would
depend upon a number of assumptions, including whether increases in freshwater
acidity would elevate levels of methylmercury in freshwater fish and by how
much, the effect of acidity on the supply of freshwater fish, as well as actions
taken by local state and Federal agencies to limit fishing and sales of fish if
methylmercury levels increase. Nevertheless, information on nethylmercury is
new reaching the point where rough estimates can be made of health risks in this
country for consumption of methylmercury from freshwater fish, and information
may be forthcoming on the impact of acidity on methylmercury levels in fish.
At least the direction of future research is new more clear — to obtain more
quantitative information on human dose-response relationship and to further test
hypotheses on cause-effect relationship between pH and methylmercury levels in
freshwater fish.
4.3 LEAD
4.3.1 Environmental Exposure and Uptake by Humans
4.3.1.1 Concentrations in Water — Ihe.U.S. national interim primary
drinking water standard for lead is 50 pg £ . The United States Environmental
Protection Agency (U.S. EPA, 1979a) summarized data in two surveys on lead in
drinking water. The median lead concentration in municipal drinking water
supplies is about 10 pg i~ . In certain areas, such as Metropolitan Boston,
drinking water may contain lead in excess of the 50 yg I standard. This is
believed to be due to very soft water (low pH) and the presence of lead piping
in the domestic water distribution system (The Nutrition Foundation Expert
Advisory Ccnmittee, 1982). Lead piping is no longer used for new potable water
systems in the United States (U.S. EPA, 1979a).
35
-------
A recent national survey cf Canadian drinking water supplies involving
71 municipalities representing 55 percent of the population, indicate^ a medial
level of lead < 1 iig 1 and values ranged frcm < 1 pg Jl to 7 pg £ .
In areas where the hone water supply is stored in lead lined tanks and
where it is conveyed to the household taps by lead pipes, the lead concentration
may reach several hundred micrograms per liter and even exceed 1000 pg £
(Beattie et al., 1972). The concentration of lead in water conveyed through
lead pipes is affected by several factors. The longer the water is held in the
pipes, the higher the lead concentrations (Wong and Berrang, 1976). The
so-called "first flush" sample generally has lead concentrations about three
times higher" than free-running tap water (Nutrition Foundation Expert -Advisory
Ccmrtittee, 1982). The lower the pH of the water and the Icwer the concentration
of dissolved salts, the greater the solubility of lead in water.
Leaching of lead frcm plastic pipes nas also been reported (Heusgem
and DeGraeve, 1973). The source of lead was probably lead stearate, which is
vised as a stabilizer in the manufacture of polyvinyl plastics.
4.3.1.2 Concentrations in Aquatic_Biota — Lead has been reported to bind
to a wide range of organic fractions in river water (Ramamoorthy and Kushner,
1975). As water pH decreases the fraction of heavy metals bound to organic
ccniponents also decreases and the concentration of free inorganic metal species
increases. This should increase lead levels in aquatic biota, possibly
affecting human dietary intake (see Sections 2.3 and 3.3).
4.3.1.3 Human Uptake and Risk Assessment — Mahaffey (1977) estimated that
the daily intake of drinking water ranged frcm 300 ml for children to as much as
2000 ml for adults. An expert group.of the National Acaderty of Sciences (NAS,
1980) stated a value of 1630 ml day for water intake of adults (not including
amounts used to prepare foods and beverages) and a range of 100 ml to 3000 ml
for children.
A study in Canada, by Armstrong and McCullough quoted by the Nutrition
Foundation's Expert Advisory Carmittee (1982) indicated that the total daily
intake including water used as a food ingredient was 760 ml averaged for 0 to 6
years, and 1140 ml for the 6- to 18-year-old group. The highest average daily
intake was 1570 ml for the 55 older age group. However, up to 3000 ml total
water per day was consumed by seme children in the 0- to 6-year-old age group
and up to 4300 ml total water was consumed by certain individuals in each of the
remaining age groups.
Using the NAS reported range of 100 to 3000 ml daily water intake for
children and a U.S. median level of 10 pg £ of lead, the range of intake for
children would be 1 to 30 pg and for adults for 16 pg, assuming a water intake
of 1600 ml day (Table 4-4). If average lead concentrations reached the
interim drinking water standard of 50 yg SL , these intake values would be five
times greater.
Lead levels in blood of children in the United States vary widely
(Mahaffey et al., 1982a). A criterion of 30 yg Pb 100 ml whole blood has been
used to indicate elevated blood lead (Center for Disease Control, 1978). If
this concentration of blood lead is accompanied by an erythrocyte protoporphyrin
36
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TABLE 4-4. DAILY INTAKE OF LEAD FROM DRINKING WATER
Age Group Daily Water Intake9 Daily Lead Intake*5
ml ng Pb
Children 100 - 3000 1 -3
Adults 1630 16
a National Academy of Science (1980).
k Assumes U.S. median concentration of lead in drinking water
to be 10 yg Pb 8. .
37
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concentration of 50 to 250 yg 100 ml~ of whole blood, the child is thought to
have undue lead absorption. Gcrmunity based lead poisoning prevention programs
report that approximately 75 percent of children with blood lead levels of
greater than or equal to 30 yg 100 ml also have erythrocyte protoporphyrin
values of greater than or equal to 50 yg 100 ml" (Mahaffey et al., 1982a).
A survey of blood lead levels in children in the years 1976 to 1980 in
the United States indicated that substantial ..numbers of children have blood
leads greater than or equal to 30 yg 100 ml" (Table 4-5). The prevalence of
elevated blood lead values is highest in children of low inccme families
(approximately 11 percent of children in families having an annual inccme less
than 6,000 dollars) and in children living in large cities (7.2 percent of
children living in cities of population more than one million). Hcwever,
elevated blood lead is widely distributed in the general population, including
diildreii in fsnu.j.xes cozniTig iforc "than- $15>000 annual income (1.2 percent; and
in children living in rural areas (2.1 percent).
Reduced pH increases' the corrosivity of water and can mobilize metals
such as lead, resulting in increased concentrations in drinking water. Lead
piping in heme plumbing is rare and no longer used in this country except in
certain parts of New England. However, lead can be mobilized fran other types
of piping where it is used as a solder (copper piping) or in stabilizers
(certain types of plastic pipes). Hemes using roof-catchment cisterns for
collecting drinking water seem especially vulnerable to corrosive rain water.
In a study of 40 roof-catchment cistern systems in western Pennsyl-
vania, Young and Sharpe (1984) report that lead in atmospheric deposition
accumulates in the sediments that collect at the bottoms of cisterns and that
this particulate lead could appear in the drinking water of cistern users when
conditions allowing the suspension of this material in cistern water are
present. They did not report on the frequency of such conditions, but they did
point out that in the systems they studied there were no safeguards to prevent
the ingestion of lead contaminated cistern sediments. Hcwever, cistern systems
with gross particulate filters for incoming catchment runoff had much lewer lead
concentrations in sediments.
Young and Sharpe measured the concentrations of lead in tapwater that
had stood in the cistern plumbing system overnight. In nine of the 40 systems
studied (22 percent) average lead concentrations exceeded the drinking water
standard of 50 yg Pb I (U.S. EPA, 1979b).
Taylor et al., (1984) found that lead concentrations in 7 of 129
household water samples frcm both surface and ground water sources exceeded the
federal drinking water standard of 50 yg Pb I (U.S. EPA, 1979b). Violation of
the drinking water standard occurred primarily in saitples that had remained
overnight in household plumbing.
Fran the point of view of human health risks, any increases of lead
concentrations in drinking water should be viewed as an additional burden of
lead. This is especially important with children, of whan substantial numbers_^
already have elevated blood levels. At the median concentration of 10 yg Pb £
drinking water already makes an appreciable contribution to blood lead levels
(approximately 30 percent of the total frcm all other sources of lead). Thus
38
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TABLE 4-5.
BLOOD LEAD LEVELS IN CHILDREN 6 MONTHS THROUGH 5 YEARS
BY ANNUAL FAMILY INCCME AND DEGREE OF URBANIZATION OF
PLACE OF RESIDENCE IN THE UNITED STATES FRCM 1976 TO 1980 .
Blood lead
levels < 30
Estimated No. of . yg 100 ml~
population persons Blooa lea£ % persons
Demographic variable (thousands) examined pg 100 ml examined
£
Annual Family Inocme
< $6000
2465
448
20
+
0.6
10.9
+
1.4
$1000 - 14,999
7534
1083
16
+
0.5
4.2
+
0.7
> 15,000
6428
774
14
+
0.4
1.2
+
0.4
Degree of Urbanization
urban > 10^ persons
4344
544
18
+
0.5
7.2
+
0.7
urban < 10^ persons
6891
944
16
+
0.7
3.5
+
0.6
Rural
5627
884
14
+
0.6
2.1
+
0.9
aAdapted fron Mahaffey et al. (1982a).
^Mean + S.E.M.
°All values shown for this variable reflect the exclusion (fron analysis and
tests of significance) of children in households that declined to report
their income.
39
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the drinking water standard of 50 pg Pb Z~ will not provide sufficient
protection to children already having elevated blood lead frcm other sources of
exposure.
Quantitative data are lacking on the contribution of acidic deposition to
lead in drinking water. Poof-catchment cistern systems believed to be widely
used in rural areas of Ohio and Western Pennsylvania are likely to be affected
by acidic deposition. Thus, it is of great importance to ascertain the extent
to which cistern systems are used in areas of the USA subject to acidic
deposition and to check the extent to which changed corrosivity of this water
affects lead levels in tap water.
4.3.2 Effects on Human Health
4.3.2.1 Metabolism in Hcpqns—¦ The uptake^ distribution, and excretion of
lead have recently been reviewed in detail"(Urs. EPA, 196ub). Approximately8
percent of dietary lead is absorbed in the gastrointestinal tract in adults.
Children absorb about 50 percent of ingested lead. Lead may be absorbed with
greater efficiency from water and other beverages than frcm food.
Lead is distributed to all tissues in the body and to all ocrnpartments
within cells. Most of the lead in blood is associated with the red blood cells.
The skeleton is the nain site of lead storage, with about 95 percent of the
total lead in the body in the skeleton of adults. Lead readily crosses the
placenta. It also crosses the blood-brain barrier but more readily in children
than in adults.
Lead is excreted in urine and feces, with the human urinary route
probably being more important. Hie half-time of lead retention in soft tissues
is about six weeks following exposure of a few months. The half-time may be
longer following years of occupational exposures to lead. Lead is accumulated
in the skeleton throughout most of the human life-span, and the half-time in
skeletal tissue is very long.
Lead concentration in whole blood is the most commonly used indicator
for assessing the burden of lead in soft tissues. The relative contributions of
airborne lead, lead in food, and other sources of lead are usually assessed in
terms of their contributions to the blood-level concentration.
A positive correlation exists between the concentration of lead in
domestic water supply and the concentration of lead in blood. Ifte U.S. EPA,
based on a study by Moore et al., (1977), has estimated blood concentrations
associated with various levels of lead in free-running tap water (Table 4-6).
If the U.S. EPA estimate is valid, the impact of lead concentrations
in running tapwater is greatest in the lower range of lead in water^- According
to Table 4-6, the median lead level inU.S. drinking water (10 yg t ) would
contribute approximately 7.4 ug 100 ml" . Assuming the median blood level in
the absence of the water contribution to be 11 yg 100 ml , the lead present in
tapwater at the current interim primary drinking water standard would contribute
about 10 pg 100 ml" to blood lead concentration, i.e., about equal to the lead
contribution frcm all other sources. However, blood levels in the United States
40
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TABLE 4-6. THE ESTIMATED RELATIONSHIP BETWEEN LEAD (XNCENTRATIONS
IN RUNNING TAP WATER AND HUMAN BLOOD LEAD LEVELS
(MOORE ET AL. 1977 IN U.S. EPA 1980b).
Lead in running
tap water
(pg * )
Total lead
in blood (PbB)
(yg 100 mi. )
Lead in Blood
due to water
(yg 100 mi. )
0
11
0
1
14.4
3.4
5
16.7
5.8
10
18.4
7.4
25
21.0
10.0
50
23.6
12.6
100
26.8
15.8
41
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are affected by a number of factors such as age, sex, and urban versus non-urban
locations. Urinary excretion of lead may be used on a group basis to indicate
the soft tissue burden. Lead in hair represents external contamination of the
hair sample.
4.3.2.2 Effects and Dose-Response in Humans — Lead damages a variety of
human organs and tissues. Damage to the" human hemopoietic system is usually the
first observable effect of lead (Figure 4-2). The inhibition of enzymes
involved in synthesizing hemoglobin results in the accumulation of precursor
substances: 6-aminolevulinic acid (6-ALA) in plasma and urine, and free
erythrocyte protoporphyrin (FEP) on the red blood cells. Measuring FEP has
become a routine method for checking the earliest effects of lead.
During recent years, measurement of FEP has cans into wide use as the
.npst .practical screening„tool both, in epidemiologic studies and in-monitoring
populations at high risk for lead toxicity. Figure 4-3 shews the curvilinear
relationship between FEP and lead concentration in blood. The curvilinear shape
is typical of the relationship between blood lead and other intermediate
metabolites of porphyrin synthesis, such as 6-aminolevulinic acid and urinary
ooprcporphyrin. Thus it is difficult to assign a specific blood level of lead
at which FEP or other metabolites attain abnormal values. At first, levels of
FEP increase slowly with blood lead, but as blood lead rises about 40 to 50 pg
100 ml" the rate of increment of FEP rises rapidly. Roels et al., (1978)
defined abnormal blood FEP levels as those in excess of the upper 95 percent
confidence limit of the controls and published a dose-response relationship
relating blood lead levels to the frequency of individuals having FEP values
equal to or in excess of their defined abnormal value (Figure 4-4). Children
and adult females tend to show a greater response than adult males. This
analysis indicates that most of a population having blood leads in the range of
30 to 40 yg 100 ml will have abnormally high FEP values.
Excessive doses of lead cause anemia and damage to both the peripheral
and central human nervous systems (Table 4-7). The central nervous system in
children appears to be more sensitive to damage from lead than the mature "
central nervous system. A growing body of knowledge suggests that lewer blood
levels of lead than those previously recognized are associated with altered
neuropsychological function and intelligence deficiencies. For exanple, reduced
general intelligence quotients, reduced auditory and speech processing, and
attention deficiences have been reported in children with higher dentine lead
than those with lower dentine lead (Needleman et al., 1979). In a study of 166
children whose blood lead levels ranged fran 7 to 33 yg 100 ml , YUle et al.,
(1981) reported decreases in attainment scores on tests of reading, spelling,
and intelligence as blood lead levels increased. Sane, but not all, of this
variability was removed after influence of the social class of the subject's
family was considered.
Picmelli (1980) has reported that heme synthesis is impaired in
children with blood lead levels less than 30 pg 100 ml consistent with
findings of Roels et al., (1978) reported in Figure 4-4. Several other
metabolic changes associated with low level lead exposure of children have been
identified. Plasma levels of the vitamin D metabolite, 1,25 dihydroxy vitamin
D, which is active in stimulating the gastrointestinal absorption of calcium and
phosphorus, decreased as the blood level increased (Rosen et al., 1980). Plasma
42
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ENZYMIC STEPS
INHIBITED
BY LEAD
NORMAL PATHWAYS
METABOLITES AND
ABNORMAL PRODUCTS ACCUMULATED
IN HUMAN LEAD POISONING
PROPHYRIN FORMATION
IRON UTILIZATION
•KREBS CYCLE
SUCC1NYL CoA ~ GLYCINE
| ALAS
t
COPROGENASE
Fe TRANSFERRIN
(SERUH) INTO
RETICULOCYTES
d-AMINOLEVULINIC ACID (ALA)
I ALAD
PORPHOBILINOGEN
lURO I SYN
IURO II COSYN
UROPORPHURINOGEN III
| UROGENASE
COPROPORPHYRINOGEN III
Senn Fe
my be Increased
ALA 1n urine (ALAU)
and serum Increased
urtne
COPRO in rbc urine (CPU)
PROTOPORPHYRIN 9
| HEMESYNTHETASE
I Fe4*
Pb
HEME
Pb
Pb
HEMOGLOBIN
2n Protoporphyrin
(ZnP).in RBC
Ferritin, Fe nlcelles
In rbc
Damaged Mitochondria and
isnature rbc fragronts
basophilic stippled cells)
Globln
Ficrure 4-2. The initial and final steps associated with disturbances in the
biosynthesis of hemoglobin due to lead are mediated by intramito-
chondrial enzymes arid the intermediate steps by cytoplasmic
enzymes. The enzymes most sensitive to lead (Steps 2 and 7) are
the SH-dependent enzymes, 6-amino-levulinate dehvdrase (ALAD)
and heme synthetase. Accumulation of the substrates of these
enzymes (ALA and FTP) is characteristic of human lead poisoning
as is increased urinary coproporhyrin excretion. Although zinc
protoporphyrin (ZnP) accumulates in erythrocytes in lead
ooisoning (and iron deficiency), it is usually measured as
"free" erythrocyte protoporphyrin (FEP). Lead reduces the
bioavailability of iron for heme formation. A compensatory
increase in the activity of the first enzyme in the pathway,
6-amino-levulinic acid synthetase (ALAS), occurs in response to
reduced heme formation. Other ocrnpensatory responses include
erythroid hyperplasia, reticulocytosis and microcytosis. J ton-
randan shortening of erythrocyte life span has been danonstrategl
in lead vorkers.' Amicrocytic, hypochromic anemia results
including seme morphological features noted above. Adapted
from Chisolm (1978).
43
-------
FEP - 0.043 x (blood lead)2* 0.45(blood lead) - 2
r - 0.79
n ¦ 1056
1200
1080
960
840
I
E
§
720
600
CT>
a
480-
#•
c.
240
120
BLOOD LEAD (yg 100 ml'1)
Figure 4-3. Free erythrocyte protoporphyrin (FEP) vs
blood level. Shoshone County, Idaho,
Aucrust 1974. Adaoted from Landriqan
et'al. (1976).
44
-------
100
g 80
3
O.
o
o.
t/i
Ui
3
CO
<
u.
o
Ul
o
60
40
20
0L
LEGEND
A - CHILDREN
B - ADULT FEMALES
C - ADULT MALES
B
10 20 30 40
LEAD IN BLOOD (jig 100 ml'l)
—I
50
Figure 4-4. The relationship between the percentaoe of
abnormally high free erythrocyte protoporphyrin
and average blood lead levels in A, children;
B, adult females; and C, adult males. Abnormal
values were defined as FEP values in excess of
82, 83, and 63 yg 100 ml"1 for children, votien,
and men, respectively. Adapted from Poels et al.
(1978).
45
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TABLE 4-7. NO DETECTED EFFECT LEVELS IN RELATION TO FfoB
(ADAPTED FROM WHO 1977)
No-detected effect
level (pg 100 ml )
Effect
Population
< 10
Erythrocyte ALAD inhibition
Adults, children
20-25
FEP
Children
20-30
FEP
Adults, female
25-35
PEP
Adults, male
30-40
Erythrocyte ATPase inhibition
General
40
ALA excretion in urine
Adults, children
40
CP excretion in urine
Adults
40
Anemia3
Children
40-50
Peripheral neuropathy
Adults
50
Anemia3
Adults
50-60
Minimal brain dysfunction
Children
60-70
Minimal brain dysfunction
Adults
60-70
Encephalopathy
Children
> 80
Encephalopathy
Adults
aTbe term anemia here is used to denote earliest statistically demonstrable
decrease in blood hemoglobin. In adult workers a decrease in blood hemoglobin
within the normal range has been reported during the first 100 days of employ-
ment. Other studies of workers indicate that frank anemia is not statistically
demonstrable until EbB > 100 pg, as cited elsewhere in the WHO report. An
increased frequency of early anemia has been reported at PbB > 40 pg of groups
of children in when concurrent iron deficiency anemia was ruled out but is
highly likely.
46
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levels of the vitamin D metabolite exhibited a strong negative^correlation with
blood lead concentrations in the range of' 12 to1120 yg 100 ml , with no
difference in slope of the regression line fran blood lead levels above or below
30 pg 100 ml (Mahaffey et al., 1982b).
Lead produces both acute and chronic effects on kidney function
(Nutritional Foundation's Expert Advisory Committee, 1982). The acute effects
manitested as dysfunction o± the proximal tubular cell such as amino aciduria,
glycoseria, and hyperphosphoturia, usually do not occur until blood levels
exceed 70 pg 100 ml . Chronic lead nephropathy is not usually recognized in
humans until it has reached an irreversible stage. Hie disease is characterized
by the slew development of contracted kidneys with pronounced arterioscelerotic
changes, fibrosis, glomerular atrophy and hyaline degeneration of blood levels.
These changes portend progressive disease sometimes resulting in acute renal
failure. Tfce duration of excessive exposure to lead is believed to play an
ijnportant role in the development of the disease. Although information on blood
levels is inadequate, it is unlikely that blood levels in the general child and
adult populations, even in the upper 2 to 5 percentile of the "normal" U.S.
range are sufficient to produce chronic renal effects.
Studies in the 19th and early 20th centuries indicated that
occupational exposures to lead (presumably higher than current exposures) caused
increased frequency of abortions and stillbirths (Oliver 1911). indeed, since
the publication of Oliver's findings, women were largely excluded from
occupational exposures to lead until very recently.
Lancranjan et al., (1975) have reported reduction in sperm counts and
abnormal sperm morphology in occupationally exposed men. The functional
significance on fertility is not known.
Prenatal exposure to lead may be associated with mental retardation in
children (Moore, 1980). The human data are consistent with experimental
findings on animals with modestly elevated blood levels, approximately 40 pg Ft)
100 ml , during prenatal and early postnatal life may be associated with subtle
and long lasting adverse consequences to the offspring.
Lead has been shown to be a carcinogen in animal tests , but
epidemiological studies have failed to reveal an association between lead
exposure and human cancer. Measurement of precursor metabolite of heme
isynthesis such as FEP or 6-ALA provide the earliest warning of the effects of
lead. Eftect on heme synthesis will protect against the more serious clinical
Effects of lead, such as anemia and encephalopathy.
4.4 ALUMINUM
4.4.1 Environmental Exposure and Uptake by Humans
4.4.1.1 Concentrations in Water — Aluminum is a significant element of
the earth's crust and is abundant in "clay soils. While the geochemical behavior
of aluminum is not precisely known, many investigators (e.g. Hutchinson et al.,
1978; Hultberg and Johansson, 1981; Driscoll et al., 1980) have observed an
increase in aluminum concentrations in water with decreasing solution pH. This
47
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phenomenon is consistent with theoretical and experimental solubility of
aluminum minerals (see Section 2.4).
There are no synoptic data on aluminum concentration in ground and
surface waters in acid sensitive regions across the United States. Initial
surveys on aluminum in raw and finished drinking waters have been prepared by
Fuhs et al., (1980), Taylor et al., (1984) and Miller et al., (1984). The Fuhs
study was limited to several samples collected in upstate New York^ The highest
aluminum concentrations found in rural drinking water was 1.4 mg I . Taylor et
al. sampled 119 surface and 14 groundwater supplies in the Northeast, both raw
and finished water types. This study found that the median level of aluminum in
water was < 0.1 mg Al l~ with a range of < 0.1 - 0.7 mg Al £. frczn 261
sanples. Miller et al. collected raw and finished water samples fran 186 water
utilities across the United States. The survey indicated that aluminum was more
likely to be found in~sarfai.de.' waters trian. in groundwaters and that there vias a
40 - 50 percent chance that alum coagulation would increase the aluminum
concentrations in finished waters above the original concentration in the raw
waters.
4.4.1.2 Concentrations in Aquatic Biota — The toxic properties of
aluminum are gplf-Hnvi-Hng with regard tn hirers ma ila-Hnn; when the aluminum
levels in water reach tcocic levels, the ensuing mortality of fish stops further
accumulation in aquatic food chains (see Sections 2.4 and 3.3).
4.4.1.3 Human Uptake and Risk Assessment — Acute or chronic disease in
man has not been related to normal dietary intake of aluminum frcm food or
drinking water. However, a potential risk may exist under the special
circumstances of patients with canprcmised kidney function who undergo regular
therapeutic dialysis. The average daily intake of aluminum via food and water
in the U.S. is approximately 5 mg, but considerable variation may occur (Greger
and Baier, 1983a). Thus drinking water containing in excess of 1 mg Al I
could make an important contribution to an individual's total intake of
aluminum. The widespread use of aluminum-oontaixiing antacids leads to a
substantial amount of the population -ingesting several hundred milligrams,
sometimes gram amounts, per day for long periods. Driscoll et al., (1980) have
reported levels of aluminum in natural bodies of freshwater in the Adirondack
Region of New York State to attain values as high as 800 pg_Al I under the
influence of acidic deposition. A concentration of 50 vg I of aluminum in
dialysis water is claimed to be dangerous (Registration Caimittee, European
Dialysis and Transplant Association, 1980). The American National Standards
Institute (1982) has established a limit of 10 pg I for aluminum content of
dialysates.
Of the various species of aluminum known to exist in bodies of natural
water, data are available only on aluminum hydroxide, which is absorbed across
the human gastrointestinal tract. In areas of the country where drinking water
is fluoridated or where elevated fluoride concentrations occur naturally, it is
likely that aluminum fluoride complexes will be present in tap water in
substantial amounts. Unfortunately, we knew nothing of the gastrointestinal
absorption of aluminum fluoride or about its potential toxicity in humans.
No systematic attempt has yet been made to develop quantitative
estimates of populations, exposure profiles and risks. Rural populations
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deriving their drinking water frcm private shallow wells are likely to be at
increased risk.
4.4.2 Effects on Human Health
4.4.2.1 Metabolian in Humans — Data on absorption, distribution, and
excretion of aluminum ccnpounds-ih man have been reviewed recently (Norseth,
1979). Aluminum is absorbed in the gastrointestinal tract. The fraction of
dietary intake absorbed into the blood stream is believed to be small, but
precise figures are not available. When aluminum was given as the hydroxide
salt to uremic patients, approximately 15 percent of the dose was absorbed, with
considerable differences among individuals (Clarkson et al., 1972). No
information is available on the absorption of other forms of aluminum or on
absorption in people with normal kidney function. Aluminum is distributed to
all tissues irr die: body and has been reported in fetal tissues. S&en aluminum
in food was given to rats, increased levels were reported in blood, brain,
liver, and testes (Ondreicka et al., 1966).
Little information is available on the relative importance of urine
and fecal pathways of aluminum excretion. Renal clearance of aluminum may be as
high as 10 percent of the glomerular filtration rate (creatinine clearance) as
indicated in patients with canprcmised renal function. These data would suggest
a high urinary rate of excretion in normal subjects and a correspondingly short
biological half-time (on the order of days or hours). Animal experiments
indicate that biliary excretion of aluminum contributes to fecal excretion of
the metal.
Aluminum is found in both cow and human milk. Nornal levels of
aluminum in human blood and other biological fluids exhibited a very wide range
of values, apparently affected by differences in the laboratories making the
analyses. Chance contamination by the ubiquitous metal causes important
problems in determining reliable values for the low levels in human plasma.
4.4.2.2 Effects and Dose-Response in Humans — Aluminum has long been
thought to be innocuous, consequently.. in the U.S. there are no established food
or drinking water standards for this pollutant. Because of recent studies
associating aluminum in water used for dialysis with senile dementia and
dialysis encephalopathy, and because aluminum continues to be implicated in
central nervous system-degenerative diseases such as Alzheimer1 s disease, the
U.S. EPA has proposed establishing National Primary Drinking Water Standards for
this pollutant (U.S. EPA, 1983).
Aluminum interacts with phosphate and fluoride in the gut. At intakes
around 100 mg Al day phosphate absorption may be decreased by a few percent,
but the regulation of phosphate by the-kidneys easily ccnpensates for that.
Daily intakes of several grams of aluminum for long periods may cause phosphate
depletion and has caused osteomalacia in wonen.
Systematic effects of aluminum absorbed frcm the gut have not been
found in persons with good renal functions. In persons with chronic renal
failure the decrease in renal elimination, possibly also increased absorption,
has caused accumulation in bone and brain (Alfrey et al., 1976). In uremic
patients given gram amounts of aluminum hydroxide to prevent absorption of
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phosphorus, this accumulation has caused osteomalacia and encephalopathy,
especially in children (Alfrey, 1984). Haemodialysis with an aluminum
concentration above 50 pg in the dialysate has caused several cases of
lethal encephalopathy. The dialysis encephalopathy is new prevented b^not
allowing the water used for the dialysate to contain more than 10 pg l .
Aluminum has also been suspected of causing Alzheimer's disease
(primary degenerative dementia), a disease of the elderly which is a ccnmon
cause of death in the U.S. (Crapper et al.f 1973). The most recent data
indicate that patients with this disease do not have, .higher concentrations of
aluminum in the brain than age-matched controls (Shore and Wyatt, 1983).
High intraneuronal concentrations of aluminum have been reported in
persons who suffered degenerative brain disease (amyotrophic lateral sclerosis
and f^kinsonisrc-dementia) in Guam (Perl et al., 1982)High concentrations of
aluminum and low concentrations of calcium and magnesium were stated to occur in
drinking water of that area. However, the concentrations of aluminum were not
reported. Because of the lack of quantitative information concerning exposure
the relationship between high aluminum exposure and the occurrence of
degenerative brain disease cannot be regardedas established.
4.5 CADMIUM
4.5.1 Environmental Exposure and Uptake by Humans
4.5.1.1 Concentrations in Water — Limited research to date has not
documented significant quantities of cadmium in tap Water for water supplies
(Young and Sharpe, 1984; Taylor et al.f 1984; Meranger et al., 1983).- Intake
via drinking water is generally less than 2 pg day" "(Sharrett et al., 1982)'.
Influx of cadmium through the combination of direct input of acidic deposition
and as corrosion products resulting frcm the contact of acidic water with
household plumbing is most likely in roof-catchment cistern systems.
Young and Sharpe (1984) also report accumulations of cadmium in "
cistern sediments, although less frequently than lead. The cadmium
concentrations in atmospheric deposition shown by the: Young and Sharpe study are
generally very low, indicating that seme other source such as corrosion of
galvanized gutters and downspouts might be present.
4.5.1.2 Concentrations in Aquatic Biota — The solubility of cadmium and
the concentration of free inorganic metal species in solution both increases as
water pH decreases. This should increase cadmium levels in aquatic' biota,
possibly increasing human dietary intake (see Sections: 2.5 and 3.3).
4.5.1.3 Human Uptake and Risk Assessment — The average dietary intake of
cadmium in the U.S. is around 20 pg, with large variations (Kcv/al et al., 1979;
Kjellstrcm, 1979; Spencer et al., 1979; Ryan et al., 1982^. The cadmium levels
in ambient ai r are generally less than 10 mg cubic meter , but tobacco contains
1-2 pg g and smoking a pack of cigarettes a day may ;cause an absorption of the
same magnitude as that frcm food since cadmium in the -inhaled smoke may be
absorbed up to 50 percent (Piscator, 1982).
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4.5.2 Effects on Human Health
4.5.2.1 Metabolism in Humans — Normally, only a few percent of the
cadmium in food is absorbed from the gastrointestinal tract, but iron deficiency
(Flanagan et al., 1978) may cause considerably higher absorption, up to 20
percent. Also cadmium deficiency may increase the absorption. Absorbed cadmium
is mainly deposited in the liver where it is bound to a low molecular weight
protein, metallothionein, which also can bind zinc, copper and mercury.
Frcrn the liver cadmium will be slowly released with a long biological
halftime, 5-10 years, and transported to the kidney, where it also will be
stored with similar halftime. Therefore the biological half-time in the whole
body may be around 20 years. If cadmium has been ingested bound to
metallothionein, the uptake in the kidney will be faster.
In the kidney the highest cadmium concentrations are in the renal
cortex. The average concentration in Americans is around 25 mg kg wet weight
(Piscator, 1982; Kjellstrcm, 1979) at age 50. In nonsmokers the concentration
in the cortex is around 15 mg Cd kg" wet weight and around 30 mg Cd kg" wet
weight in smokers (Kjellstrcm, 1979). The distribution of renal cortex
concentrations is log normal and in nonatokers the upper limit is probably
around 70 rag Od kg wet weight. A 3-fold increase in dietary intake would
cause a corresponding increase in renal levels of cadmium.
4.5.2.2 Effects and Dose-Response in Humans — The critical organ is the
kidney and the critical concentration in renal cortex is around 200 mg Od kg"
wet weight (Roels et al., 1983). At that level the reabsorption of proteins
fran the glcmeral filtrate may be slightly reduced, resulting in an increased
urinary excretion of especially lew molecular weight proteins, e.g.,
8-microglobulin. This effect probably occurs when the renal capacity to
synthesize metallothionein and keep cadmium in an inert form is exceeded. At
higher renal concentrations also other functions may be disturbed, e.g.
reabsorpticn of calcium and phosphorus, which may cause general disturbances in
bone metabolism. In carbination with nutritional deficiencies the renal damage
caused by cadmium was the cause of the Itai-Itai disease, an osteomalacia, in
Japan.
Cadmium has also been suspected as causing cancer of the prostate, but
the evidence is very weak, and there is no reason to believe that ingested
cadmium has an etiological role (Piscator, 1981).
4.6 OTHER TRACE CONTAMINANTS
4.6.1 Asbestos —Asbestos fibers may enter drinking water fran natural
erosion of asbestos-containing rock, mine processing wastes, corrosion of
asbestos-cement pipe, or disintegration of asbestos tile roof used to collect
and direct rain water into drinking water cisterns. While inhaled asbestos is
recognized to present carcinogenic risks to human health, the effect of asbestos
ingested fran drinking water is unclear. The role of asbestos in the etiology
of gastro- intestinal cancer has been a subject of scientific controversy. At
present there is no National Primary Drinking Water Quality Standard for
asbestos.
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U.S. exposure to asbestos in drinking water frcm distribution pipes
has, however, been estimated by Milette et al., (1981a, 1981b). Available data
indicate that 16.5 percent of U.S. water utilities distribute highly corrosive
water, and 52 percent distribute moderately corrosive water. In this same
study, Millette estimated that approximately 11 percent of the U.S. population
is exposed to asbestos fiber concentrations in excess of 10 million fibers £
Kanarek et al. (1980) published the first study concluding that a
statistically significant association (P < 0.01, t-test for multiple regression
and correlation coefficients) exists between asbestos in drinking water and
cancer. The water canes from aquifers or is stored in reservoirs which have
contact with naturally occurring asbestos rock. In contrast, several other
studies have not found significant associations between asbestos fiber
concentrations and cancer.
No data directly associate acidic deposition with asbestos-related
health effects. There are several explanations for this lack of association:
diagnosis- of asbestos-related cancer is difficult; acidic deposition health
effects and asbestos contamination of drinking water are each recently-
discovered concerns and no experimental attempt has yet been made to associate
them; a time lag may be involved between exposure of material to aggressive
water and corrosion onset or detection; or, indeed, no such association exists.
4.6.2 Nitrates —- The incremental increase in nitrates in drinking water
supplies due to deposition of nitrogen compounds is a concern due to infant
methemoglobinemia frcm ingestion of high levels of nitrate. However, the
relative importance of deposition-related nitrates to increases in nitrates frcm
agricultural fertilizer use and sewage disposal (septic system failure,
treatment plant discharges) has not been well documented. The additional input
of nitrates at the Hubbard Brook watershed due to deposition since the 1960's
has been estimated at about 1 mg £ as N (National Research Council, 1983).
The current U.S. maximum contaminant level for drinking water nitrates has been
set at 10 mg i. as N.
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SECTION II: SUMMARY OF RESEARCH RECOMMENDATIONS
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ATMOSPHERIC PROCESSES (reccnmendations apply to all trace contaminants
studied: mercury, lead, cadmium and aluminum).
(1) Interpret existing trace contaminant data relative to human exposure
and toxicity;
(2) Evaluate the applicability of existing emissions data for the
prediction of trace contaminant input into the environment;
(3) Develop techniques for measurement of dry and wet deposition (includes
both vapor phase and solid phase for mercury) ;
(4) Establish standardized methods to include quality control procedures
and archive all measurements of atmospheric concentration, dry
deposition and -wet deposition for^traee contaminant species;
(5) Establish or coordinate a national network of trace contaminant
measurements to determine their spatial variability;
(6) Evaluate the applicability of trace contaminants as unique tracers for
establishing source-receptor relationships.
BIOGEOCHEMISTRY
(1) Studies of temporal and regional spatial variability in atmospheric
deposition of trace metals, with canton methodologies, are not
available. We recommend such a study for Cd, Hg and Fb.. Hie
frequency of collection, the type of collectors, sample preservation,
metal speciation and bioavailability should be addressed. A proposal
of this type of survey work was prepared by National Atmospheric
Deposition Program personnel and has been under consideration by the
U.S. Environmental Protection Agency and U.S. Geological Survey for
two years. We strongly urge consideration of this proposal or one
similar to it.
(2) Taylor et al. (1984) have studied water quality in the northeastern
U.S. in public water supply systems. Their study investigated water
quality fran raw sources to the dcirestic; thus it evaluated effects of
acidic deposition, water treatment, and distribution systems on the
water quality. A large fraction of the population in predominantly
rural states (e.g. approximately 1/3 in Maine) uses water directly
frcm surface sources, shallow dug wells or individual drilled wells -
all without chemical treatment. Hultberg and Johansson (1981) studied
the water quality of private dug and drilled wells in southwestern
Sweden and found significant water acidification and associated high
concentrations of heavy metals and Al in the raw water. Effects due
to local distribution systems were also observed. A study of this
type should be conducted on a regional or state-by-state basis.
(3) Given the general relationship between mercury in fish and lakewater
pH, our lack of knowledge of the extent of mercury contamination of
fish stocks in remote lakes in the U.S., and our present inability to
predict mercury burdens in fish in such lakes, we reccrrmend that a
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national survey of mercury content in canton native fish species; be
undertaken, with particular reference to differences in:
- watershed geology, soil type and soil depth;
lakewater retention time;
acidity;
lake trophic status;
color (humic vs. clearwater lakes);
- mercury content in sediment and watershed soil.
Hie lakes should be selected as a subset of those initially sampled in
the U.S. National Lake Survey, Phase I; the proposed mercury study
could be incorporated into Phase II of the National Lake Survey
(biological parameters).
(4) Recent experimental results (Hultberg, personal ccntnunication)
indicate that biotic transformations of mercury may have an important
influence on mercury bioaccumulation (e.g. chemical methylation;
volatilization). Experiments should be undertaken to determine
whether or not chemical methylation occurs in the natural environment
(soils, sediments, surface waters) and to evaluate its relative
importance as a function of such factors as pH, fulvic acid
concentration, iron and manganese levels.
(5) Biogeochanical cycling of mercury is poorly understood and thus
predictions about fluxes, body burdens in fish, altered pathways (due
to acidification), and related questions cannot be adequately
addressed. To better understand the biogeochanical cycling of mercury
we reccrmend that the mercury flux through the ecosystem be studied at
calibrated watershed sites for a small number of lakes which differ
in:
productivity;
hydrology;
atmospheric loading of mercury;
acidity.
Depending on available funding the study could look at a variety of
processes, including:
relative importance of wet and dry deposition;
- chemical form of mercury in atmospheric deposition;
mercury fluxes between terrestrial, aquatic, biological,
sedimentary, and atmospheric oanpartments;
transformations of mercury (reduction, methylation) Within
ocrrpartments;
mechanisms of mercury uptake by fish;
food chain interactions with mercury forms;
methylnercury partitioning (sediment/water; soils/water;
periphyton/water).
degree of mercury adsorption/precipitation in sediments.
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Intensively studied sites should include the Little Pock Lake
Acidification Project in Wisconsin (U.S. EPA-funded) and an additional
site (soon to be named) in the northeastern U.S. Additionally,
several presently acidified lakes that are to be limed should be
studied. Additional funding should be provided for these studies at
the lake manipulation sites.
BIQACCUMUIATION
(1) In Scandinavia, Canada, and the United States, mercury concentrations
exceeding 1.0 pg g~ wet weight in piscivorous fishes have been
recorded, particularly in low pH lakes. However, there are no
comprehensive surveys of mercury contamination of fish across
different geographic regions containing susceptible softwater lakes.
Development ar.d extensive statistical analysis of a .centralized-data
base on mercury concentrations in fish in acid-sensitive lakes that
lack direct or local anthropogenic sources of mercury is needed to
define the geographic extent of this problem. In addition, Phase II
of the (U.S.) National Lake Survey should include sampling and
analysis of fish for total mercury concentration in axial muscle
tissue. A subset of the fish samples should be analyzed for
methylmercury.
(2) Mechanisms enhancing availability of me thy lmercury in poorly buffered
lakes should be elucidated. At a minimum, whole-lake experimental
acidification studies in the Upper Midwest and the northeastern United
States should include collection and analysis of fishes for total
mercury concentration at annual intervals.
(3) lb evaluate the effect of chemical neutralization on mercury
contamination of fishery resources, Task Group E (Objective E-5) of
the National Acid Precipitation Assessment Program should consider
inclusion of mercury analysis of fishes and other biota in recently
initiated and future mitigation studies.
INDIRECT HEALTH EFFECTS
(1) Freshwater fish with the highest user consumption figures and the
highest average methy lmercury levels are among the most likely species
to be affected by acidic deposition (see Section 4.2.1.3). More
research is needed to delineate the cause-effect relationship between
acidic deposition, pH, and ire thy lmercury levels in freshwater fish.
(2) Human patterns of freshwater fish consumption need documentation so
that populations at risk for me thy lmercury toxicity may be identified.
With these data, an average methy lmercury intake in freshwater fish
consumers could be calculated.
(3) Prenatal life is a more sensitive stage of the life cycle for methyl-
mercury (see Section 4.2.2.2). More information is needed on fish
consumption patterns of wcmen of child-bearing age in order to
55
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quantitatively assess the potential impact on human health of elevated
methylmercury levels in freshwater fish.
(4) Populations are at increased risk of being exposed to higher
concentrations of corrosive toxicants, such as lead and possibly
cadmium, where acidic surface waters are used for drinking without
corrosive control (see Section 4.3.1.3). Quantification of the extent
of lead and cadmium mobilization from water distribution systems and
domestic plumbing into drinking water is warranted. Hie populations
at risk and water distribution networks involved need to be
identified.
(5) People receiving drinking water from roof catchment cisterns are at
potential risk of increased intake Of lead and possibly cadmium in
areas of acidic deposition (see Section 4.3.1.3). The population at
risk needs to be identified and quantification of the extent of lead
and cadmium mobilization from roof catchments systems and domestic
plumbing is warranted.
(6) Individual water supplies emanating from shallow groundwater in areas
edaphically and geologically sensitive to acidic deposition, may
contain elevated concentrations of aluminum. Quantification of the
extent of aluminum contamination in these water supplies and
identification of populations at risk is warranted.
(7) The human toxicity of lead has been extensively studied (see Section
4.3.2.2). However, further study is necessary concerning lead effects
on certain physiological processes.
- Determine the functional significance of lead exposure
on infertility in the human population.
- Determine the effect of prenatal exposure to lead on
the development of the fetus.
(8) Aluminum toxicity to humans is of growing concern, particularly due to
its association with certain types of dementia: dialysis dementia,
senile dementia, Alzheimer's disease, and Parkinson's disease (see
Section 4.4.2.2). The following studies are warranted.
- Pharmacokinetic studies are needed in appropriate animal
models to establish the relationship between oral intake
of aluminum and the levels of aluminum in brain tissue.
These data, when obtained, should be extrapolated to man.
- Biochemical studies are needed to elucidate the mechanism
whereby aluminum can produce chronic degenerative
disease of the human nervous system.
- Studies to further clarify the relationship between
dialysis and aluminum toxicity are required.
- Development of reliable analytical methods for
56
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measuring low levels of aluminum in human plaana
are necessary.
- Studies to determine the significance of osteomalacia
in wcmen with large daily intaks of aluminum with
phosphate depletion are required.
Acidification of certain water supplies with cement-asbestos pipe
distribution systems may release asbestos fibers into the water (see
Section 4.6.1). The following studies are necessary to assess the
significance of this problem.
- Epidemiological and toxicological studies need to be
conducted to ascertain the relationship between
asbestos fiber ingestion and gastro-intestinal cancer.
- Studies to quantify the extent to which acidified
water is capable of mobilizing asbestos fibers frcm
cement-asbestos piping is required.
- The populations at risk and water distribution
networks involved with asbestos contamination
need to be identified.
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APPENDICES
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APPENDIX A: AGENDA FOR WORKSHOP
Research Needs Regarding Trace Contaminants and Acid Deposition
as Related to Increased Human Exposure
Top Ridge Conference Center
Paul Shiiths, New York
June 19-22, 1984
AGENDA
Tuesday, June 19
5:00 pro
5:30 - 7:30 pan
7:30 - 9:00 pm
Arrival at Top Ridge Conference Center.
Transportation will be provided frcro the
Albany, NY airport.
Reception and Dinner at Top Ridge Lodge.
Opening remarks and workshop organization.
Wednesday, June 20
7:30 am
8:00 am - Noon
Noon - 1:00 pm
1:00 - 5:30 p11
6:30 - 7:30 pm
Breakfast at Top Ridge Lodge.
Plenary session. Introduction of participants and
informal presentations.
Lunch at Top Ridge Lodge.
Review and development of state-of-the-art
sunmaries concerning trace contaminants and acid
deposition. Discussions will take place in
working groups focusing on atmospheric processes
and ecosystem effects leading to increased human
exposure.
Dinner at Top Ridge Lodge.
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Thursday, June 21
7:30 am
8:00 am - Noon
Noon - 1:00
1:00 - 3:00 pm
3:00 - 6:30 jm
7:30 - 8:30 pm
Breakfast at Top Ridge Lodge
Further discussion in working groups of research
needs and the development of working group
reports.
Lunch at Top Ridge Lodge
Break
General session, including working group reports,
general discussion and development of workshop
reccranendations.
Dinner at Top Ridge Lodge.
Friday, June 22
7:30 am
8:00 am
Noon
Breakfast at Top Ridge Lodge.
Adjournment.
Appraxinate arrival time at Albany, NY airport.
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APPENDIX B: ACKNOWLEDGEMENTS
The editors thank the workshop participants for their enthusiasm, diligence
and thoroughness in the development and review of this report. Continual
interest and encouragement have been provided by Joseph M. Eilers, U.S. EPA.
His patience and constructive input throughout the review process are greatly
appreciated. Finally, special recognition is given to supporting staff, Sharon
Hotaling and Denise Polsinelli, who worked long, tedious hours in preparation of
this report.
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