Mercury - environmental aspects (EHC 86, 1989)

INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY

ENVIRONMENTAL HEALTH CRITERIA 86

Mercury - Environmental Aspects

This report contains the collective views of an international group of
experts and does not necessarily represent the decisions or the stated
policy of the United Nations Environment Programme, the International
Labour Organisation, or the World Health Organization.

First draft prepared at the National Institute of Health Sciences,
Tokyo, Japan, and the Institute of Terrestrial Ecology, Monk's Wood,
United Kingdom

Published under the joint sponsorship of the United Nations
Environment Programme, the International Labour Organisation, and the
World Health Organization

World Health Organization
Geneva, 1989

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CONTENTS

ENVIRONMENTAL HEALTH CRITERIA FOR MERCURY - ENVIRONMENTAL ASPECTS

1. SUMMARY AND CONCLUSIONS

1.1. Physical and chemical properties
1.2. Sources in the environment
1.3. Uptake, elimination, and accumulation in organisms
1.4. Toxicity to microorganisms
1.5. Toxicity to aquatic organisms
1.6. Toxicity to terrestrial organisms
1.7. Effects of mercury in the field

2. PHYSICAL AND CHEMICAL PROPERTIES

3. SOURCES OF MERCURY IN THE ENVIRONMENT

3.1. Natural and anthropogenic sources and cycling
3.2. Speciation
3.3. Levels in the environment
3.4. Methylation of mercury

4. UPTAKE, LOSS, AND ACCUMULATION IN ORGANISMS

4.1. Speciation of mercury
4.2. Uptake and loss in aquatic organisms
4.2.1. Microorganisms, plants, and invertebrates
4.2.2. Fish
4.2.2.1 Effects of environmental variables on
uptake by fish
4.2.3. Studies on more than one type of organism
4.3. Uptake and loss in terrestrial organisms
4.4. Accumulation in the field
4.4.1. General exposure
4.4.2. Mercury manufacturing and general industrial areas
4.4.3. Mining activity
4.4.4. Chloralkali plants
4.4.5. Mercurial fungicides

5. TOXICITY TO MICROORGANISMS

5.1. Toxicity of inorganic mercury
5.1.1. Single species cultures
5.1.2. Mixed cultures and communities
5.2. Toxicity of organic mercury

6. TOXICITY TO AQUATIC ORGANISMS

6.1. Toxicity to aquatic plants
6.2. Toxicity to aquatic invertebrates

6.2.1. Acute and short-term toxicity to
invertebrates
6.2.2. Behavioural effects
6.3. Toxicity to fish
6.3.1. Acute and short-term toxicity to fish
6.3.2. Reproductive effects and effects on
early life stages
6.3.3. Behavioural effects
6.3.4. Physiological and biochemical effects
6.4. Toxicity to amphibia
6.5. Toxicity to aquatic mammals

7. TOXICITY TO TERRESTRIAL ORGANISMS

7.1. Toxicity to terrestrial plants
7.2. Toxicity to terrestrial animals
7.2.1. Toxicity to terrestrial invertebrates
7.2.2. Effects of mercury on birds
7.2.2.1 Inorganic and metallic mercury
7.2.2.2 Effect of organic mercury on birds
7.2.3. Effects of mercury on non-laboratory mammals

8. EFFECTS OF MERCURY IN THE FIELD

9. EVALUATION

9.1. The marine environment
9.2. The freshwater environment
9.3. The terrestrial environment

REFERENCES

WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR MERCURY -
ENVIRONMENTAL ASPECTS

Participants

Dr L.A. Albert, Director, Environmental Pollution Programme, National
Institute for Research on Biotic Resources, Xalapa, Mexico

Professor T.W. Clarkson, Division of Toxicology, The University of
Rochester, School of Medicine and Dentistry, Rochester, USA
(Chairman)

Dr R. Elias, Environmental Criteria and Assessment Office, US
Environmental Protection Agency, Research Triangle Park, North
Carolina, USA

Dr J.H.M. Temmink, Department of Toxicology, Agricultural University,
Biotechnion, Wageningen, Netherlands

Dr G. Roderer, Fraunhofer Institute for Environmental Chemistry and
Ecotoxicology, Schmallenberg-Grafschaft, Federal Republic of
Germany

Dr R. Koch, Division of Toxicology, Research Institute for Hygiene and
Microbiology, Bad Elster, German Democratic Republic

Professor Y. Kodama, Department of Environmental Health, University of
Occupational and Environmental Health, Kitakyushu, Japan

Professor P.N. Viswanathan, Ecotoxicology Section, Industrial
Toxicology Research Centre, Lucknow, India

Observers

Mr D.J.A. Davies, Department of the Environment, London, United
Kingdom

Dr I. Newton, The Institute of Terrestrial Ecology, Monks Wood
Experimental Station, Huntingdon, United Kingdom

Secretariat

Dr S. Dobson, The Institute of Terrestrial Ecology, Monks Wood
Experimental Station, Huntingdon, United Kingdom (Rapporteur)

Dr M. Gilbert, International Programme on Chemical Safety, World
Health Organization, Geneva, Switzerland (Secretary)

Mr P.D. Howe, The Institute of Terrestrial Ecology, Monks Wood
Experimental Station, Huntingdon, United Kingdom

NOTE TO READERS OF THE CRITERIA DOCUMENTS

Every effort has been made to present information in the criteria
documents as accurately as possible without unduly delaying their
publication. In the interest of all users of the environmental health
criteria documents, readers are kindly requested to communicate any
errors that may have occurred to the Manager of the International
Programme on Chemical Safety, World Health Organization, Geneva,
Switzerland, in order that they may be included in corrigenda, which
will appear in subsequent volumes.

* * *

A detailed data profile and a legal file can be obtained from the
International Register of Potentially Toxic Chemicals, Palais des
Nations, 1211 Geneva 10, Switzerland (Telephone no. 988400 - 985850).

ENVIRONMENTAL HEALTH CRITERIA FOR MERCURY - ENVIRONMENTAL ASPECTS

A WHO Task Group on Environmental Health Criteria for Mercury -
Environmental Aspects met at the Institute of Terrestrial Ecology,
Monks Wood, UK, from 7 to 11 December 1987. Dr B.N.K. Davis welcomed
the participants on behalf of the host Institution, and Dr M. Gilbert
opened the meeting on behalf of the three co-sponsoring organizations
of the IPCS (ILO/UNEP/WHO). The Task Group reviewed and revised the
draft criteria document and made an evaluation of the risks for the
environment from exposure to mercury.

The first draft of this document was prepared by Dr S. Dobson and
Mr P.D. Howe, Institute of Terrestrial Ecology. Dr M. Gilbert and
Dr P.G. Jenkins, both members of the IPCS Central Unit, were
responsible for the overall scientific content and editing,
respectively.

* * *

Partial financial support for the publication of this criteria
document was kindly provided by the United States Department of Health
and Human Services, through a contract from the National Institute of
Environmental Health Sciences, Research Triangle Park, North Carolina,
USA - a WHO Collaborating Centre for Environmental Health Effects.

INTRODUCTION

There is a fundamental difference in approach between the
toxicologist and the ecotoxicologist concerning the appraisal of the
potential threat posed by chemicals. The toxicologist, because his
concern is with human health and welfare, is preoccupied with any
adverse effects on individuals, whether or not they have ultimate
effects on performance or survival. The ecotoxicologist, in contrast,
is concerned primarily with the maintenance of population levels of
organisms in the environment. In toxicity tests, he is interested in
effects on the performance of individuals - in their reproduction and
survival - only insofar as these might ultimately affect the
population size. To him, minor biochemical and physiological effects
of toxicants are irrelevant if they do not, in turn, affect
reproduction, growth, or survival.

It is the aim of this document to take the ecotoxicologist's
point of view and consider effects on populations of organisms in the
environment. No attempt has been made to link the conclusions reached
in this document with possible effects on human health. This will only
be feasible when Environmental Health Criteria 1: Mercury (WHO, 1976),
which considered the effects of mercury on human health, has been
updated. Due attention has been given to the persistence in the
environment and the bioaccumulation and transport of mercury in
aquatic food chains. These will have implications for human
consumption of the metal.

This document, although based on a thorough survey of the
literature, is not intended to be exhaustive in the material included.
In order to keep the document concise, only those data which were
considered to be essential in the evaluation of the risk posed by
mercury to the environment have been included. Concentration figures
for mercury in the environment, or in particular species of organism,
have not been included unless they illustrate specific toxicological
points. "Snap shot" concentration data, where a causal relationship
between the presence of the metal and an observed effect is not
clearly demonstrated, have been excluded.

The term bioaccumulation indicates that organisms take-up
chemicals to a greater concentration than that found in their
environment or their food. 'Bioconcentration factor' is a quantitative
way of expressing bioaccumulation: the ratio of the concentration of
the chemical in the organism to the concentration of the chemical in
the environment or food. Biomagnification refers, in this document, to
the progressive accumulation of chemicals along a food chain.

1. SUMMARY AND CONCLUSIONS

1.1 Physical and chemical properties

Mercury is a metal which is liquid at normal temperatures and
pressures. It forms salts in two ionic states mercury(I) and
mercury(II). Mercury(II), or mercuric, salts are very much more common
than mercury(I) salts, and hence it is mercuric salts which will be
mainly considered here. Mercury also forms organometallic compounds,
some of which have found industrial and agricultural use.
"Organometallic" is used here to indicate a covalently-bonded
compound, and does not include mercury bound to proteins nor salts
formed with organic acids. These organometallic compounds are stable,
though some are readily broken down by living organisms, while others
are not readily biodegraded. Elemental mercury gives rise to a vapour
which dissolves only slightly in water.

1.2 Sources in the Environment

Natural mercury arises from the degassing of the earth's crust
through volcanic gases and, probably, by evaporation from the oceans.
Local levels in water derived from mercury ores may also be high
(up to 80 µg/litre). Atmospheric pollution from industrial production
is probably low, but pollution of water by mine tailings is
significant. The burning of fossil fuels is a source of mercury. The
chloralkali industry and, previously, the wood pulping industry also
released significant amounts of mercury. Although the use of mercury
is reducing, high concentrations of the metal are still present in
sediments associated with the industrial applications of mercury. Some
mercury compounds have been used in agriculture, principally as
fungicides.

1.3 Uptake, Elimination, and Accumulation in Organisms

Mercuric salts, and, to a much greater extent, organic mercury,
are readily taken up by organisms in water. Aquatic invertebrates, and
most particularly aquatic insects, accumulate mercury to high
concentrations. Fish also take up the metal and retain it in tissues,
principally as methylmercury, although most of the environmental
mercury to which they are exposed is inorganic. The source of the
methylation is uncertain, but there is strong indication that
bacterial action leads to methylation in aquatic systems.
Environmental levels of methylmercury depend upon the balance between
bacterial methylation and demethylation. The indications are that
methylmercury in fish arises from this bacterial methylation of
inorganic mercury, either in the environment or in bacteria associated
with fish gills, surface, or gut. There is little indication that fish
themselves either methylate or demethylate mercury. Elimination of
methylmercury is slow from fish (with half times in the order of
months or years) and from other aquatic organisms. Loss of inorganic
mercury is more rapid and so most of the mercury in fish is retained

in the form of methylmercury. Terrestrial organisms are also
contaminated by mercury, with birds being the best studied. Sea birds
and those feeding in estuaries are most contaminated. The form of
retained mercury in birds is more variable and depends on species,
organ, and geographical site.

1.4 Toxicity to Microorganisms

The metal is toxic to microorganisms. Inorganic mercury has been
reported to have effects at concentrations of the metal in the culture
medium of 5 µg/litre, and organomercury compounds at concentrations at
least 10 times lower than this. Organomercury compounds have been used
as fungicides. One factor affecting the toxicity of the organometal is
the rate of uptake of the metal by cells. Mercury is bound to the cell
walls or cell membranes of microorganisms, apparently to a limited
number of binding sites. This means that effects are related to cell
density as well as to the concentration of mercury in the substrate.
These effects are often irreversible, and mercury at low
concentrations represents a major hazard to microorganisms.

1.5 Toxicity to Aquatic Organisms

The organic forms of mercury are generally more toxic to aquatic
organisms than the inorganic forms. Aquatic plants are affected by
mercury in the water at concentrations approaching 1 mg/litre for
inorganic mercury but at much lower concentrations of organic mercury.
Aquatic invertebrates vary greatly in their susceptibility to mercury.
Generally, larval stages are more sensitive than adults. The 96-h
LC₅₀s vary between 33 and 400 µg/litre for freshwater fish and are
higher for sea-water fish. However, organic mercury compounds are more
toxic. Toxicity is affected by temperature, salinity, dissolved
oxygen, and water hardness. A wide variety of physiological and
biochemical abnormalities has been reported after fish have been
exposed to sublethal concentrations of mercury, although the
environmental significance of these effects is difficult to assess.
Reproduction is also affected adversely by mercury.

1.6 Toxicity to Terrestrial Organisms

Plants are generally insensitive to the toxic effects of mercury
compounds. Birds fed inorganic mercury show a reduction in food intake
and consequent poor growth. Other, more subtle, effects on enzyme
systems, cardiovascular function, blood parameters, the immune
response, kidney function and structure, and behaviour have been
reported. Organomercury compounds are more toxic for birds than are
inorganic.

1.7 Effects of Mercury in the Field

Pollution of the sea with organomercury led to the death of fish
and fish-eating birds in Japan. Except for this incident at Minamata,
few follow-up studies of the effects of localised release have been
conducted. The use of organomercury fungicides as seed dressings in
Europe led to the deaths of large numbers of granivorous birds,
together with birds of prey feeding on the corpses. Residues of
mercury in birds' eggs have been associated with deaths of embryos in
shell. The presence of organochlorine residues in the same birds and
their eggs makes an accurate assessment of the effects of mercury
difficult. It is, however, thought to be a contributing factor in the
population decline of some species of raptors.

2. PHYSICAL AND CHEMICAL PROPERTIES

The physical and chemical properties of mercury have been
detailed in Environmental Health Criteria 1: Mercury (WHO, 1976). The
relevant chapter is summarized here.

Mercury can exist in a wide variety of physical and chemical
states. The different chemical and physical forms of this element all
have their intrinsic toxic properties and different applications in
industry and agriculture, and require a separate assessment of risk.

Mercury, along with cadmium and zinc, falls into Group IIb of the
Periodic Table. In addition to its elemental state, mercury exists in
the mercury (I) and mercury (II) states in which the mercury atom has
lost one and two electrons, respectively. The chemical compounds of
mercury (II) are much more numerous than those of mercury (I).

In addition to simple salts, such as chloride, nitrate and
sulfate, mercury (II) forms an important class of organometallic
compounds. These are characterized by the attachment of mercury to
either one or two carbon atoms to form compounds of the type RHgX and
RHgR' where are R and R' represent the organic moiety. The most
numerous are those of the type RHgX. X may be one of a variety of
anions. The carbon-mercury bond is chemically stable. It is not split
in water nor by weak acids or bases. The stability is not due to the
high strength of the carbon-mercury bond but to the very low affinity
of mercury for oxygen. The organic moiety, R, takes a variety of
forms, some of the most common being the alkyl, the phenyl, and the
methoxyethyl radicals. If the anion X is nitrate or sulfate, the
compound tends to be "salt-like" having appreciable solubility in
water; however, the chlorides are covalent, non-polar compounds that
are more soluble in organic solvents than in water. From the
toxicological standpoint, the most important of these organometallic
compounds is the subclass of short-chain alkyl mercurials in which
mercury is attached to the carbon atom of a methyl, ethyl, or propyl
group.

3. SOURCES OF MERCURY IN THE ENVIRONMENT

The sources of mercury have been detailed in Environmental Health
Criteria 1: Mercury (WHO, 1976). Relevant data are summarized here.

3.1 Natural and Anthropogenic Sources and Cycling

The major source of mercury is the natural degassing of the
earth's crust and amounts to between 25 000 and 125 000 tonnes per
year. Anthropogenic sources are probably less than natural sources.
World production of mercury by mining and smelting was estimated at
10 000 tonnes per year in 1973 and has been increasing at an annual
rate of about 2%. The chloralkali, electrical equipment, and paint
industries are the largest consumers of mercury, accounting for about
55% of the total consumption. Mercury has a wide variety of other uses
in industry, agriculture, military applications, medicine, and
dentistry.

Several of man's activities, not directly related to mercury,
account for substantial releases into the environment. These include
the burning of fossil fuel, the production of steel, cement, and
phosphate, and the smelting of metals from their sulfide ores.

Alkylmercury fungicides used as seed dressings are important
original sources of mercury in terrestrial food chains, although the
use of these materials has decreased considerably.

Two cycles are believed to be involved in the environmental
transport and distribution of mercury. One is global in scope and
involves the atmospheric circulation of elemental mercury vapour from
sources on land to the oceans. However, the mercury content of the
oceans is so large, at least 70 million tonnes, that the yearly
increases in concentration due to deposition from the global cycle are
not detectable.

The other cycle is local in scope and depends upon the
methylation of inorganic mercury mainly from anthropogenic sources.
Many steps in this cycle are still poorly understood, but it is
believed to involve the atmospheric circulation of dimethylmercury
formed by bacterial action.

3.2 Speciation

The following speciation among mercury compounds has been
proposed by Lindquist et al. (1984), where V stands for volatile, R
for water-soluble or particle-borne reactive species, and NR for non-
reactive species (Hg° is elemental mercury):

V: Hg°, (CH₃)₂Hg

R: Hg²⁺, HgX₂, HgX₃^-, and HgX₄^2-,

with X = OH^-, Cl^- and Br^-.

HgO on aerosol particles. Hg²⁺ complexes with organic
acids.

NR: CH₃Hg⁺, CH₃HgCl, CH₃HgOH and other organomercuric
compounds, Hg(CN)₂. HgS and Hg²⁺ bound to sulfur in
fragments of humic matter.

The main volatile form in air is elemental mercury but dimethylmercury
may also occur (Slemr et al., 1951).

Uncharged complexes, such as HgCl₂, CH₃HgOH etc., occur in
the gaseous phase, but are also relatively stable in fresh water (snow
and rain as well as standing or flowing water). HgCl₄^2- is the
dominant form in sea water.

3.3 Levels in the Environment

The following data have been extracted from Lindquist et al.
(1984) and are included here to indicate background levels of mercury
in the environment. Considerable local variations can occur and local
levels close to anthropogenic sources of mercury would be much higher.

Reliable data on mercury concentrations in the air are scarce.
Recent information suggests a background level at about 2 ng/m³ in
the lower troposphere of the northern hemisphere and about 1 ng/m³
in the southern hemisphere, at least over oceanic areas. In European
areas remote from industrial sources, such as the rural parts of
southern Sweden and Italy, concentrations most often lie in the range
from 2 to 3 ng/m³ in summer and from 3 to 4 ng/m³ in winter
(Brosset 1983, Ferrara et al., 1982). In urban air the concentrations
could be higher.

Deposition with precipitation is a major factor in removing
mercury from the atmosphere. The lowest concentrations of mercury in
rain water, around 1 ng/litre, have been reported from a coastal site
in Japan and from the islands of Samoa. Most other values reported lie
in the range between 5 and 100 ng/litre.

Recent measurements of mercury in aquatic systems have given
the following concentration ranges, which may be considered
representative for dissolved mercury:

Open ocean 0.5-3 ng/litre

Coastal sea water 2-15 ng/litre

Rivers and lakes 1-3 ng/litre

Local variations from these values are considerable, especially
in coastal sea water and in lakes and rivers where mercury associated
with suspended material may also contribute to the total load.

The mercury content in minerals forming ordinary rock and soils
is usually very low. The normal level in igneous rocks and minerals
seems to be less than 50 µg/kg, and in many cases is less than
10 µg/kg. Due to the strong binding of mercury to soil particles,
including organic matter, only small amounts of the metal are present
in soil solution; reported averages range between 20 and 625 µg/kg
soil.

Background levels in sediments are approximately the same as
levels in unpolluted surface soils. Average concentrations in ocean
sediments probably lie in the range between 20 and 100 µg/kg.

3.4 Methylation of Mercury

The methylation of inorganic mercury in the sediment of lakes,
rivers and other waterways, as well as in the oceans, is a key step in
the transport of mercury in aquatic food chains.

It was first demonstrated by Jensen & Jernelov (1967) that
microorganisms in lake sediments could methylate mercury. They later
showed that the degree of methylation correlated well with the overall
microbial activity in the sediment (Jensen & Jernelov, 1969). Detailed
mechanisms of methylation in microorganisms have been proposed by Wood
(1971) and Landner (1971). Some soil organisms capable of methylating
mercury have also been isolated (Kitamura et al., 1969; Yamada &
Tonamura, 1972).

The following general conclusions have been drawn by Bisogni &
Lawrence (1973) concerning methylation by microorganisms:

(a) mono-methylmercury is the predominant product of biological
methylation near neutral pH,

(b) the rate of methylation is greater under oxidising
conditions than under anaerobic conditions,

(d) temperature affects methylation as a result of its effect on
overall microbial activity,

(e) higher microbial growth rate increases mercury methylation,

(f) methylation rates are inhibited by the addition of sulfide
to anaerobic systems.

The formation of new or enlarged artificial lakes considerably
increases the production of methylmercury, although this increase was
found to be short-lived in new lakes in Finland (Simola & Lodenius,
1982; Alfthan et al., 1983). A similar problem of increased mercury in
new lakes, which was taken up by fish and fish-eating mammals,
occurred in the scheme to divert the Churchill River in Manitoba,
Canada (Canada-Manitoba, 1987). Methylation rates in one lake, which
had been flooded 20 years previously, had returned to normal.
Methylation rates in the new lake, which had flooded arboreal forest,
were high and were expected to remain high for decades. The source of
mercury in all of these artificial lakes appeared to be natural rather
than anthropogenic in origin. Anaerobic conditions after the flooding
of large amounts of organic material and the subsequent increase in
microbial activity are thought to be the causes of the increased
availability of mercury through methylation.

4. UPTAKE, LOSS, AND ACCUMULATION IN ORGANISMS

Background levels of naturally-occuring mercury in the
environment are generally low, except in the immediate vicinity of
mining sites and chloralkali plants for the industrial extraction of
mercury. The majority of mercury in the environment is natural rather
than the result of human activities. Inorganic mercury can be
methylated in the environment and the resultant methylmercury is
taken up into organisms readily; more readily than inorganic mercury.
Although environmental levels are low, the high capacity of organisms
to accumulate mercury means that the metal is found widely in both
aquatic and terrestrial animals and plants. Methylmercury is released
more slowly by aquatic organisms than inorganic mercury. Aquatic
invertebrates, and particularly aquatic insects, accumulate mercury
to a greater extent than fish.

Speciation of mercury is of great importance in determining the
uptake of the metal from water and soil. Much of the mercury in
natural waters and in soil is strongly bound to sediment or organic
material and is unavailable to organisms.

Mercury has been found in many terrestrial organisms, birds
being the subjects of most of the monitoring.

In many experimental studies, the concentrations of mercury
quoted are nominal rather than measured. Few attempts have been made
to estimate available mercury in experimental studies.

Because of the very extensive literature on the uptake of metals
into organisms, this section contains illustrative examples and is
not exhaustive.

Bioconcentration factors for mercury, determined in laboratory
experiments, are summarized in Tables 1 and 2.

Bioconcentration factors are simple ratios between the
concentration of mercury in an organism and the concentration in the
medium to which the organism was exposed. This means that results
should be treated with caution. A relatively low body burden resulting
from exposure to very low levels of mercury in the medium can give a
high bioconcentration factor. Conversely, exposure to very high
mercury levels in the medium can lead to a low bioconcentration
factor. Exposure to mercury under static test conditions will lead to
the removal of mercury during the course of the test, whereas flow-
through conditions maintain a constant level of exposure. Since
mercury is strongly bound to sediment in the field, it is unclear
which of these two exposure regimes is the most realistic. It is
probable that static exposure underestimates and flow-through exposure
overestimates mercury uptake. Most studies have failed to distinguish

Table 1. Accumulation of mercury into aquatic organisms

Organism Lifestage^c Stat/ Organ^b Temperature Compound^e Duration Exposure Bioconcentration Reference
flow^a (°C) (days) (µg/litre) factor^g

Alga mercuric chloride 2 164 8537^h Parrish & Carr
(Croomonas (1976)
salina)

Filamentous algae stat phenyl mercuric acetate 35 10 1200^f Hannerz (1968)
(Oedogonium sp.) flow 16.5 methoxyethylmercuric OH 18 0.58 2610^f Hannerz (1968)
flow 15.2 mercuric chloride 54 0.05 871^f Hannerz (1968)

Duckweed flow methylmercuric OH 32 3 2950^f Hannerz (1968)
(Lemna minor) flow 16.5 methoxyethylmercuric OH 24 0.58 480^f Hannerz (1968)
flow 15.2 mercuric chloride 41 0.05 70^f Hannerz (1968)

Water hyacinth mature stat roots 23-27 mercuric chloride 16 1000 580 Muramoto & Oki
(Eichhornia (1983)
crassipes)

Reed stat emergent phenyl mercuric acetate 35 10 0^f Hannerz (1968)
(Phragmites stat submerged phenyl mercuric acetate 35 10 850^f Hannerz (1968)
communis) flow emergent methylmercuric OH 32 3 25^f Hannerz (1968)
flow submerged methylmercuric OH 32 3 530^f Hannerz (1968)
flow emergent 16.5 methoxyethylmercuric OH 24 0.58 74^f Hannerz (1968)
flow submerged 16.5 methoxyethylmercuric OH 24 0.58 139^f Hannerz (1968)
flow emergent 15.2 mercuric chloride 41 0.05 56^f Hannerz (1968)
flow submerged 15.2 mercuric chloride 14 0.05 149^f Hannerz (1968)

Bulrush stat emergent phenyl mercuric acetate 35 10 90^f Hannerz (1968)
(Scirpus stat submerged phenyl mercuric acetate 35 10 790^f Hannerz (1968)
lucustris) flow emergent methylmercuric OH 32 3 8^f Hannerz (1968)
flow submerged methylmercuric OH 32 3 1250^f Hannerz (1968)

Table 1 (cont'd)

Organism Lifestage^c Stat/ Organ^b Temperature Compound^e Duration Exposure Bioconcentration Reference
flow^a (°C) (days) (µg/litre) factor^g

flow emergent 16.5 methoxyethylmercuric OH 24 0.58 39^f Hannerz (1968)
flow submerged 16.5 methoxyethylmercuric OH 24 0.58 190^f Hannerz (1968)
flow emergent 15.2 mercuric chloride 21 0.05 77^f Hannerz (1968)
flow submerged 15.2 mercuric chloride 41 0.05 70^f Hannerz (1968)

Yellow iris stat emergent phenyl mercuric acetate 35 10 20^f Hannerz (1968)
(Iris stat submerged phenyl mercuric acetate 35 10 40^f Hannerz (1968)
pseudacorus) flow emergent methylmercuric OH 32 3 18^f Hannerz (1968)
flow submerged methylmercuric OH 32 3 34^f Hannerz (1968)
flow emergent 16.5 methoxyethylmercuric OH 18 0.58 31^f Hannerz (1968)
flow submerged 16.5 methoxyethylmercuric OH 18 0.58 90^f Hannerz (1968)
flow emergent 15.2 mercuric chloride 49 0.05 18^f Hannerz (1968)
flow submerged 15.2 mercuric chloride 49 0.05 23^f Hannerz (1968)

Bloodworm stat phenyl mercuric acetate 35 10 12 700^f Hannerz (1968)
(Chironomidae) flow methylmercuric OH 32 3 3070^f Hannerz (1968)
flow 16.5 methoxyethylmercuric OH 89 0.58 988^f Hannerz (1968)

Annelid stat phenyl mercuric acetate 35 10 2030^f Hannerz (1968)
(Haemopis flow methylmercuric OH 32 3 450^f Hannerz (1968)
sanguisuga) flow 16.5 methoxyethylmercuric OH 89 0.58 1148^f Hannerz (1968)

Annelid flow methylmercuric OH 32 3 110^f Hannerz (1968)
(Glossosiphonia flow 16.5 methoxyethylmercuric OH 89 0.58 640^f Hannerz (1968)
complanata) flow 15.2 mercuric chloride 65 0.05 670^f Hannerz (1968)

Worm flow methylmercuric OH 32 3 1780^f Hannerz (1968)
(Oligochaeta) flow 16.5 methoxyethylmercuric OH 65 0.58 690^f Hannerz (1968)
flow 15.2 mercuric chloride 65 0.05 517^f Hannerz (1968)

Table 1 (cont'd)

Organism Lifestage^c Stat/ Organ^b Temperature Compound^e Duration Exposure Bioconcentration Reference
flow^a (°C) (days) (µg/litre) factor^g

Freshwater leech flow 15.2 mercuric chloride 65 0.05 534^f Hannerz (1968)
(Herpobdella
octoculata)

Mussel WB mercuric chloride 4 50 664 Tsuruga (1963)
(Mytilus edulis) flow WB 4 0.06 236^f Hannerz (1968)

Short-necked clam WB mercuric chloride 8 50 190 Tsuruga (1963)
(Venerupis
philippinarum)

Pond snail stat phenyl mercuric acetate 35 10 1280^f Hannerz (1968)
(Planorbis sp.) flow methylmercuric OH 32 3 3570^f Hannerz (1968)
flow 16.5 methoxyethylmercuric OH 31 0.58 1970^f Hannerz (1968)
flow 15.2 mercuric chloride 49 0.05 795^f Hannerz (1968)

Giant pond snail stat phenyl mercuric acetate 35 10 1800^f Hannerz (1968)
(Lymnaea flow methylmercuric OH 32 3 3480^f Hannerz (1968)
stagnalis) flow 16.5 methoxyethylmercuric OH 31 0.58 1178^f Hannerz (1968)
flow 15.2 mercuric chloride 14 0.05 297^f Hannerz (1968)

Snail flow 16.5 methoxyethylmercuric OH 24 0.58 4266^f Hannerz (1968)
(Physa flow 15.2 mercuric chloride 14 0.05 637^f Hannerz (1968)
fontinalis)

Water flea stat phenyl mercuric acetate 35 10 3570^f Hannerz (1968)
(Daphnia sp.)

Cladoceran flow 16.5 methoxyethylmercuric OH 89 0.58 286^f Hannerz (1968)
(Eurycerus)

Table 1 (cont'd)

Organism Lifestage^c Stat/ Organ^b Temperature Compound^e Duration Exposure Bioconcentration Reference
flow^a (°C) (days) (µg/litre) factor^g

Copepod stat WB 21-26 mercuric chloride 1 0.1 7600 Hirota et al.
(Acartia clausi) (1983)
stat WB 21-26 methyl mercuric chloride 1 0.1 249 000 Hirota et al.
(1983)

Grass shrimp WB mercuric chloride 3 1.5 333 Ray & Tripp
(Palaemonetes (1976)
pugio)

Mayfly naiad stat phenyl mercuric acetate 35 10 900^f Hannerz (1968)
(Ephemeridae) naiad flow methylmercuric OH 32 3 3290^f Hannerz (1968)
naiad flow 16.5 methoxyethylmercuric OH 24 0.58 680^f Hannerz (1968)
larva flow 15.2 mercuric chloride 65 0.05 138^f Hannerz (1968)

Lesser water stat phenyl mercuric acetate 35 10 4200^f Hannerz (1968)
boatman flow methylmercuric OH 32 3 8470^f Hannerz (1968)
(Corixa sp.) flow 16.5 methoxyethylmercuric OH 89 0.58 740^f Hannerz (1968)
flow 15.2 mercuric chloride 65 0.05 414^f Hannerz (1968)

Water boatman flow methylmercuric OH 32 3 2460^f Hannerz (1968)
(Notonecta flow 16.5 methoxyethylmercuric OH 89 0.58 674^f Hannerz (1968)
glaaca) flow 15.2 mercuric chloride 65 0.05 483^f Hannerz (1968)

Midge larva flow WB mercuric chloride 30 5.5 19 600 Rossaro et al.
(Chironomus (1986)
riparius) pupa flow WB mercuric chloride 30 5.5 15 600 Rossaro et al.
(1986)
adult flow WB mercuric chloride 30 5.5 7500 Rossaro et al.
(1986)

Table 1 (cont'd)

Organism Lifestage^c Stat/ Organ^b Temperature Compound^e Duration Exposure Bioconcentration Reference
flow^a (°C) (days) (µg/litre) factor^g

Caddisfly larva flow 16.5 methoxyethylmercuric OH 89 0.58 710^f Hannerz (1968)
(Trichoptera sp.) larva flow 15.2 mercuric chloride 49 0.05 513^f Hannerz (1968)

Dragonfly nymph flow 16.5 methoxyethylmercuric OH 89 0.58 1296^f Hannerz (1968)
(Odonata sp.)

Damselfly nymph flow 16.5 methoxyethylmercuric OH 89 0.58 1186^f Hannerz (1968)
(Odonata sp.) nymph flow 15.2 mercuric chloride 65 0.05 655^f Hannerz (1968)

Alderfly larva flow 16.5 methoxyethylmercuric OH 89 0.58 1270^f Hannerz (1968)
(Sialis lutaria)

Cranefly larva flow 16.5 methoxyethylmercuric OH 18 0.58 625^f Hannerz (1968)
(Tipula sp.) larva flow 15.2 mercuric chloride 41 0.05 840^f Hannerz (1968)

Great diving imago flow 16.5 methoxyethylmercuric OH 89 0.58 800^f Hannerz (1968)
beetle
(Dytiscus
marginalis)

larva flow 16.5 methoxyethylmercuric OH 89 0.58 3134^f Hannerz (1968)
larva flow 15.2 mercuric chloride 65 0.05 603^f Hannerz (1968)
imago flow 15.2 mercuric chloride 65 0.05 862^f Hannerz (1968)

Pond skater flow 16.5 methoxyethylmercuric OH 89 0.58 754^f Hannerz (1968)
(Gerris najas) flow 15.2 mercuric chloride 65 0.05 431^f Hannerz (1968)

Aquatic saw bug flow 16.5 methoxyethylmercuric OH 89 0.58 954^f Hannerz (1968)
(Asellus
aquaticus)

Table 1 (cont'd)

Organism Lifestage^c Stat/ Organ^b Temperature Compound^e Duration Exposure Bioconcentration Reference
flow^a (°C) (days) (µg/litre) factor^g

Water spiders flow 16.5 methoxyethylmercuric OH 89 0.58 624^f Hannerz (1968)
(Hydracnidae)

Pike stat liver 17.2 methoxyethylmercuric OH 10 0.4 7673^f Hannerz (1968)
(Esox lucius) stat kidney 17.2 methoxyethylmercuric OH 10 0.4 7230^f Hannerz (1968)
stat liver methylmercuric OH 10 0.3 2002^f Hannerz (1968)
stat kidney methylmercuric OH 10 0.3 2198^f Hannerz (1968)

Rainbow trout juv flow WB 5 methyl mercuric chloride 84 0.263 4525 Reinert et al.
(1974)
(Salmo gairdneri) juv flow WB 10 methyl mercuric chloride 84 0.258 6628 Reinert et al.
(1974)
juv flow WB 15 methyl mercuric chloride 84 0.244 8033 Reinert et al.
(1974)
WB^d 5 mercuric chloride 4 50 5 MacLeod &
Pessah (1973)
WB^d 10 mercuric chloride 4 50 12 MacLeod &
Pessah (1973)
WB^d 20 mercuric chloride 4 50 26 MacLeod &
Pessah (1973)

Bluegill sunfish stat WB 9 methyl mercuric chloride 28.6 0.5 222^f Cember et al.
(1978)

Table 1 (cont'd)

Organism Lifestage^c Stat/ Organ^b Temperature Compound^e Duration Exposure Bioconcentration Reference
flow^a (°C) (days) (µg/litre) factor^g

(Lepomis stat WB 21 methyl mercuric chloride 28.6 0.5 1138^f Cember et al.
macrochirus) (1978)
stat WB 33 methyl mercuric chloride 28.6 0.5 2454^f Cember et al.
(1978)

^a stat = static conditions (water unchanged for duration of experiment); flow = flow-through conditions (mercury concentration in water
continuously maintained).
^b WB = whole body.
^c juv = juvenile.
^d muscle, skin & bone.
^e OH = hydroxide.
^f radiometrically calculated.
^g bioconcentration factor = concentration in organism/concentration in medium.
^h dry weight.

Table 2. Accumulation of mercury into terrestrial organisms

Organism Route Organ Compound Duration Exposure^b Bioconcentration Reference
(days) (mg/kg) factor^c

Broccoli soil leaves mercuric chloride 60 20 0.002^d John (1972)
(Brassica oleracea) soil roots mercuric chloride 60 20 0.09^d John (1972)

Pea soil roots mercuric chloride 95 20 0.07^d John (1972)
(Pisum sativum)

Cauliflower soil leaves mercuric chloride 70 20 0.003^d John (1972)
(Brassica oleracea) soil roots mercuric chloride 70 20 0.12^d John (1972)

Spinach soil leaves mercuric chloride 55 20 0.03^d John (1972)
(Spinacia oleracea) soil roots mercuric chloride 55 20 0.05^d John (1972)

Chicken diet muscle methyl mercury dicyandiamide 35-42 8 1.25 Borg et al. (1970)
diet liver methyl mercury dicyandiamide 35-42 8 5 Borg et al. (1970)

Mallard diet liver methyl mercury dicyandiamide 14 8 2.1 Stickel et al. (1977)
(Anas platyrhynchos) diet kidney methyl mercury dicyandiamide 14 8 2.2 Stickel et al. (1977)

Redwinged blackbird diet liver methyl mercury dicyandiamide 11 40 2.3 Finley et al. (1979)
(Agelaius diet kidney methyl mercury dicyandiamide 11 40 2.1 Finley et al. (1979)
phoeniceus

Cowbird diet liver methyl mercury dicyandiamide 11 40 1.7 Finley et al. (1979)
(Molothrus ater) diet kidney methyl mercury dicyandiamide 11 40 1.5 Finley et al. (1979)

Grackle diet liver methyl mercury dicyandiamide 11 40 1.3 Finley et al. (1979)
(Quiscalus quiscula) diet kidney methyl mercury dicyandiamide 11 40 1.1 Finley et al. (1979)

Table 2 (cont'd)

Organism Route Organ Compound Duration Exposure^b Bioconcentration Reference
(days) (mg/kg) factor^c

Mink diet liver ceresan L^a 5 32 11.1 Aulerich et al.(1974)
(Mustela vision) diet kidney ceresan L^a 5 32 7.4 Aulerich et al.(1974)
(adult) diet liver mercuric chloride 10 135 0.3 Aulerich et al.(1974)
diet kidney mercuric chloride 10 135 3.2 Aulerich et al.(1974)

^a ceresan L = methylmercury 2,3-di-hydroxy propyl mercaptide + methylmercury acetate.
^b exposure as mg/kg of mercury soil or diet according to route.
^c concentration factors calculated on a wet weight basis unless otherwise stated; bioconcentration factor = concentration in
organism/concentration in medium.
^d dry weight.

between mercury taken into the tissues of the organism and mercury
adsorbed on external surfaces. This should also be taken into account
when interpreting results.

Taking these factors into account, it is still clear that
organisms take up both inorganic and organic forms of mercury from the
medium. This uptake can result in high concentration factors. Under
identical conditions, organic mercury is taken up by organisms to a
greater degree than inorganic mercury, although the latter may often
be strongly adsorbed to the outer surfaces.

4.1 Speciation of Mercury

Appraisal

Different species of mercury differ greatly in their
physicochemical properties: in their solubility, rates of
accumulation by organisms, and behaviour in ecosystems. It is in its
methyl form that mercury is most hazardous. Although not all sites of
methylation in the environment are fully known, several have been
identified in the aquatic environment.

Mercury accumulated in the tissues of fish is usually in the form
of methylmercury, while the source is usually inorganic mercury
(Huckabee et al., 1979). Several hypotheses of how and where
methylation occurs have been proposed. The main hypotheses are:

(a) biological methylation, bacterial in origin, which produces
methylmercury in the environment (methylmercury is taken up
by fish more readily than inorganic mercury),

(b) methylation by microorganisms associated with branchial
mucus of the fish or in the fish gut, and

It is generally agreed that methylation by fish, other than by
bacteria associated with the fish, either does not occur or accounts
for only an insignificant amount of the methylmercury produced. There
is good evidence for methylation by bacteria in aquatic systems.

Jernelov (1968) suggested that fish could not methylate mercury
themselves and this is generally accepted (Huckabee et al., 1979),
though not universally. Jernelov & Lann (1971) showed that 60% of the
mercury content of predator fish (northern pike) arose from prey fish.
This mercury was already methylated in the prey. The concentration of
mercury in predator species was similar to that in their prey. They
also measured the mercury content of organisms that were the food of
the prey fish. Mercury levels in benthic fauna were very low and
contributed less than 25% of the mercury in bottom-feeding fish. Most
of the mercury accumulated by non-predator species was, therefore,

accumulated directly from water. This conclusion was also reached by
Fagerstrom & Asell (1973). The question of where the methylation,
which gives rise to methylmercury residues in fish, occurs is still
unresolved. It is also generally accepted that fish do not demethylate
mercury either.

4.2 Uptake and Loss in Aquatic Organisms

Appraisal

The data presented on uptake by aquatic invertebrates are
difficult to interpret because most studies do not differentiate
between external adsorption and actual uptake into the organism. This
is especially important for methylmercury compounds for which uptake
seems to be correlated with surface adsorption capacity, as expressed
by the relative size of the organism.

The extrapolation of data on uptake to other organisms appears
risky because of a lack of knowledge regarding the mechanisms of
uptake. This is even true for phenomena that are apparently fairly
universal, e.g., the facilitating influence of chelators upon
uptake.

Most data on uptake by fish support the notion that uptake
correlates positively with available concentration, with exposure
time, and with temperature, although hardly any investigation
differentiates between nominal and available concentrations. The
importance of this distinction seems to be illustrated by the
positive influence of lowered pH upon uptake.

None of the studies address the problem of distinguishing
between adsorption to gills and slime on the one hand and real uptake
into the body on the other. Studies of mercury distribution between
organs are valuable for the potential effects of the total body
burden, but they give no reliable insight into the time-dependent
process of accumulation.

Data consistently show a higher uptake of methylmercury than of
inorganic mercury. However, other organic mercury compounds exhibit a
lower uptake, since they are adsorbed to a lesser extent.

4.2.1 Microorganisms, plants, and invertebrates

When Glooschenko (1969) exposed the marine diatom Chaetoceros
costatum to labelled mercury, he found no difference between uptake
in the light or the dark in non-dividing cells. Dead cells took up
twice as much mercury as living cells, presumably by surface
adsorption. As dividing cells in the light accumulated the labelled
mercury for longer than non-dividing cells, the author suggested the
possibility of some active uptake.

Hannerz (1968) demonstrated that there was no appreciable
assimilation of mercury into the tissues of aquatic plants. Although
concentrations were 10-20 times higher in submerged parts compared to
emergent parts, this was attributed to surface adsorption differences.
De et al. (1985) grew the plant Pistia stratiotes in nutrient
solution to which mercuric chloride had been added at concentrations
ranging from 0.05 to 20 mg/litre. They found that uptake gradually
increased with an increase in the mercury concentration. Maximum
accumulation occurred within one day. Maximum removal (approximately
90%) was recorded at 6 mg/litre or less, only 20% being lost from
plants receiving the highest concentration. Mercury accumulation into
the roots was about 4 times higher than into the shoots at lower
concentrations and about twice as high at 20 mg/litre.

Zuberik & O'Connor (1977) studied the accumulation of mercury in
aquatic organisms from the Hudson River, USA. The organisms were
maintained in filtered river water that contained mercury
concentrations of < 0.1 µg/litre (less than levels normally found in
the Hudson River). Planktonic organisms were exposed to various forms
of labelled mercury, and the concentration factors after 24 h ranged
from 10² to 10⁶. Mercury uptake was greater in microplankton and
algae than in macroplankton and fish larvae. An amphipod (Gammarus
sp.) was exposed for one day to each of four types of mercury, two
organic (phenylmercuric acetate and methylmercury chloride) and two
inorganic (mercurous nitrate and mercuric chloride). No differences in
uptake were found, but when the amphipod was exposed for a week the
organic forms were accumulated to 3 times the concentration of the
inorganic forms.

Riisgard et al. (1985) transferred mussels (Mytilus edulis)
from clean water to an area chronically polluted with mercury. The
mussels accumulated mercury readily during 3 months of exposure. They
were then transferred to clean water in the laboratory and the
elimination of the mercury was measured. The biological half-life was
293 days, but was only 53 days in the case of mussels contaminated by
a temporary massive mercury contamination. In both cases, 75% of the
mercury in the mussels was inorganic, but both inorganic and organic
species were immobilized in the mussels from the chronically polluted
area. In another study, only 6% of the total mercury in Macoma
balthica, a sediment-feeding bivalve, was methylated, a much lower
percentage than in Mytilus from the same area.

Hirota et al. (1983) exposed the copepod Acartia clausi to
inorganic (mercuric chloride) and organic (methylmercury chloride)
mercury at concentrations of 0.05-0.5 µg/litre for 24 h. The
bioconcentration factor for inorganic mercury was nearly constant
(approximately 7500), regardless of the mercury concentration in the
water or the density of the copepods. In contrast, the concentration
factor of methylmercury fluctuated, showing an inverse relationship
with density but no relationship with the mercury concentration in the
water.

DeFreitas et al. (1981) found a net assimilation of 70%-80% for
methylmercury and 38% for inorganic mercury when fed in the diet to
the shrimp Hyalella azteca. From water, inorganic mercury was
assimilated 2 to 3 times more slowly than methylmercury. Khayrallah
(1985) found that the accumulation of ethylmercuric chloride was
almost twice as rapid as that of mercuric chloride in the amphipod
Bathyporeia pilosa, although death occurred at similar levels of
mercury.

Ray & Tripp (1976) exposed the grass shrimp (Palaemonetes pugio)
to radioactively labelled methylmercury chloride and mercuric chloride
for 24 and 72 h. After 24 h, the methylated form was mostly
concentrated in the ventral nerve cord and to a lesser extent in the
gills. The reverse was true for mercuric chloride. The concentrations
of mercury accumulated in the other tissues (exoskeleton, foregut, and
remainder) were similar for both compounds, and were in decreasing
order of the above list. After 72 h the tissue distribution had
changed, and there was no consistent order of the relative tissue
concentrations. There was an increase in the mercury levels of the
exoskeleton, foregut, and remainder tissues, while that in the gills
remained about the same and that in the ventral nerve cord decreased.

Vernberg & O'Hara (1972) measured the uptake of labelled mercury
into the gills and hepatopancreas of fiddler crabs (Uca pugilator)
maintained in a solution containing 0.18 mg mercury/litre (as mercuric
chloride) for 72 h. Uptake was determined under various temperature
(5°C to 33°C) and salinity (5 and 30 g/litre) regimes. The total
mercury taken up by the gills and hepatopancreas pooled together was
unaffected by the different regimes. However, the ratio of uptake into
the two tissues was affected. At higher temperatures, the crabs seem
able to transport mercury from gill tissue to the hepatopancreas more
effectively than at low temperatures.

When Rossaro et al. (1986) exposed various life stages of the
midge Chironomus riparius to mercuric chloride for a period of
30 days, the levels were still increasing at the end of the
experiment. Both larvae and pupae accumulated mercury to about the
same levels, some accumulation being due to passive adsorption. In a
small experiment to illustrate this, larvae kept in a solution of
5 µg/litre for only 1 min accumulated 9.32 mg mercury/kg. The adults
accumulated only 40% of the levels found in the larval stage. The
authors suggested that this is because the adults have some means for
eliminating the mercury.

Getsova & Volkova (1964) reported concentration factors for the
accumulation of radioactively labelled mercury in four insect species.
A midge, Glyphotaelius punctatolineatus, accumulated 5240 times the
water concentration within 16 days, while a dragonfly, Leucorrhinia
rubicunda, accumulated 8310 times the concentration over 16 days.
Another dragonfly, Aeschna grandis, accumulated 4000 times the
waterborne mercury in 8 days, while a waste-water inhabiting fly,

Eristalis tenax, accumulated only 640 times the water concentration
after 4 days and the concentration factor had fallen to just 266 after
8 days. The authors stated that the concentration factors that they
found were in agreement with other Russian work on mercury
accumulation.

4.2.2 Fish

When Birge et al. (1979) exposed rainbow trout eggs to an
inorganic mercury concentration of 0.1 µg/litre in a flow-through
system, the eggs accumulated 42.4, 68.2, and 96.8 µg mercury/kg after
1, 4, and 7.5 days, respectively. Control eggs contained
18.6 µg mercury/kg. The bioconcentration factor over 7.5 days was 782,
taking into account the degree of contamination of controls. This
represented a daily uptake rate of about 20 µg/kg. There was no
evidence to suggest that the mercury penetrated the outer covering of
the eggs and there was a high probability that most of the "uptake"
was surface adsorption.

Backstrom (1969) found that the uptake by fish of various mercury
compounds was similar to that observed with birds (where methylmercury
is rapidly absorbed compared with phenylmercury, methoxyethylmercury,
and inorganic mercury), but the difference in uptake between
methylmercury and the other mercury compounds was less pronounced.
Mercury uptake into the spleen and the thyroids was greater than for
birds. Phenylmercury was also retained in the wall of the gall
bladder. In general the uptake of mercury into fish was far more
localized than in birds. The levels of methylmercury steadily
increased in the muscles and in the brain, whereas the other compounds
accumulated primarily in the kidneys, spleen, and liver. More mercury
accumulated in red flesh than white. There was also a high uptake of
mercury into the gills and pseudobranch.

Kramer & Neidhart (1975) demonstrated that methylmercury was
taken up from water by guppies (Lebistes reticulatus) 17 times
faster than inorganic mercury. Organic mercury was also eliminated
more slowly than inorganic. The authors suggested that some
methylation of mercury occurred in the fish.

Ribeyre & Boudou (1984) examined the uptake of mercury over time
into specific organs of the rainbow trout. The uptake was sigmoid with
a linear phase and a plateau. The majority (55% for inorganic and 60%
for methylmercury) of the metal was found in muscle and gills, while
blood contained 3%-12%, liver 2%-5%, and kidneys 2%-7%. Brain,
posterior intestine, and spleen together accounted for only 2% of
total mercury. Those organs which would eventually contain most
mercury accumulated their mercury exponentially. After the exposure,
some organs lost their mercury while others (the ones with most
mercury) continued to increase their mercury content. The organs which
lost mercury in clean water had accumulated the metal with a flatter
sigmoid curve.

Schindler & Alberts (1977) found that the mosquitofish (Gambusia
affinis) readily accumulated metallic mercury during short-term
continuous exposure. Within 24 h, 20 mg/kg wet weight had been taken
up from a solution containing 0.1 mg total mercury/litre. The uptake
curves for metallic mercury and mercuric chloride were very similar.
The authors suggested that uptake in the short-term is largely the
result of physical adsorption. This rate of uptake closely agrees with
that found by McKone et al. (1971) in goldfish (Carassius auratus)
where 22 mg mercuric chloride/kg was accumulated from a solution
containing 0.25 mg/litre over a period of 24 h.

When Schindler & Alberts (1977) periodically exposed (2 h/day for
10 days) mosquitofish to metallic mercury and mercuric chloride
(in separate experiments) at 100 µg/litre, the uptake of metallic
mercury was 5 times greater than that of the chloride. The authors
suggested that the metallic mercury remained unchanged and that its
high lipid solubility enabled it to penetrate the gill membrane,
whereas the salt bound more tightly to the mucoproteins of the gills
and penetration was restricted. The rate of elimination in mercury-
free water was about the same for both, with the half-time calculated
to be about 45 days.

McKim et al. (1976) exposed 3 generations of brook trout
(Salvelinus fontinalis) to methylmercury at concentrations measured
at < 0.01-2.93 µg/litre. The uptake was rapid and 2-week
concentration factors ranged from 1000 to 12 000, depending on the
tissue. There was a tendency for the uptake to reach a steady state
(that is the tissue content reached a constant level) over 20-28
weeks. There was no significant elimination over this period.

In studies by Pentreath (1976), the thornback ray (Raja clavata)
readily absorbed both inorganic mercuric chloride and organic
methylmercuric chloride from sea water. Methylmercury, in contrast to
inorganic mercury, was readily absorbed from food and slowly
eliminated. The half-lives of elimination of mercury taken up from
food were 61.6 days for inorganic and 323 days for organic components.

Thellen et al. (1981) found that methylmercuric chloride rapidly
accumulated in the organs and muscular tissue of rainbow trout exposed
to 1 mg/kg diet. However, mercuric chloride, at the same
concentration, did not accumulate. During exposure to a continuous
sublethal concentration of 0.25 µg mercury/litre, both organic and
inorganic mercury accumulated, primarily in the internal organs and to
a lesser extent the muscle tissue. Mercuric chloride was detected in
the muscle at half of the concentration of organic mercury. Wobeser
(1975b) fed rainbow trout fingerlings a diet containing methylmercuric
chloride (at 4, 8, 16, or 24 mg mercury/kg) over a 15-week period. The
total accumulation of mercury in muscle tissue was directly related to
the concentration of mercury in the food, as was the rate of
accumulation. Mercury was accumulated in muscle to a higher
concentration than there had been in the diet.

When Amend (1970) exposed juvenile sockeye salmon (1 h per day
for 12 to 15 days) to 1 mg/litre of lignasan (6.25% ethylmercury
phosphate), the fish contained highest levels in the kidneys and
liver. One week after the cessation of treatment, these levels were
36.5 and 20.4 mg/kg for the kidney and liver, respectively. Three
years later, the fish having migrated, levels were still higher than
normal but had returned to normal after 4 years. Similar studies using
coho and chinook salmon yielded similar results. When Kendall (1975)
injected channel catfish intraperitoneally with methylmercury chloride
at 15 mg/litre, the mean concentration of mercury in the kidneys was
51.03 µg/g after 24 h and fell to 14.24 mg/kg after 96 h.

4.2.2.1 Effects of environmental variables on uptake by fish

Appraisal

Environmental variables such as temperature and pH increase the
uptake of mercury, particularly methylmercury, by fish. This is of
potentially considerable importance in the field.

Reinert et al. (1974) found that yearling rainbow trout (Salmo
gairdneri) exposed to methylmercury chloride for 12 weeks accumulate
more mercury at 15°C than at 5°C (Table 1). When Cember et al. (1978)
exposed bluegill sunfish (Lepomis macrochirus) to methylmercury
chloride at concentrations ranging from 0.2 to 50 µg/litre for up to
688 h, mercury accumulation was not affected by the different mercury
concentrations. It did, however, increase when the temperature was
increased from 9°C to 33°C (Table 1). MacLeod & Pessah (1973) found an
increase in mercury accumulation, in response to an increase in
temperature (from 5 to 20°C), in rainbow trout exposed to
concentrations of between 50 and 200 µg/litre for 4 days. The authors
also interpolated (from 7-day data) a 4-day bioconcentration factor
for phenylmercuric acetate of 100, when the fish were exposed to
5 µg/litre mercury at 10°C. Tsai et al. (1975) studied the effect of
pH on the accumulation of inorganic mercury (mercuric chloride) at a
concentration in water of 1500 µg mercury/litre for 15 min. The
accumulation increased as pH decreased. At pHs of 5, 6.5, and 7.5,
fathead minnow accumulated whole body residues of 2.7, 1.8, and
0.4 mg mercury/kg, calculated on a wet weight basis, respectively. A
similar result was found for the emerald shiner (Nicropterus
atherinoides).

Rodgers & Beamish (1981) found that the uptake of methylmercury
by rainbow trout was increased when the hardness of the water was
decreased from 385 mg/litre to 30 mg/litre. The addition of inorganic
mercuric chloride increased the uptake of methylmercury in both hard
and soft water. Kudo & Mortimer (1979) exposed guppies to mercury in a
double chambered system, with an exchange of water. Only in one
chamber did the fish have access to sediment. After being exposed for

20 days to a sediment mercury concentration of 1.023 mg/kg, the fish
without direct access to the sediment showed a concentration factor of
57 and those with access a factor of 570.

4.2.3 Studies on more than one type of organism

Cultures of the alga Croomonas salina, grown for 48 h in the
presence of mercuric chloride (164 µg mercury/litre), retained about
half of the mercury (1400 mg/kg dry weight) (Parrish & Carr 1976).
When the alga was fed to the copepod Acartia tonsa for 5 days,
neither the copepods nor their eggs or faeces retained mercury in
detectable amounts.

Boudou et al. (1979) exposed mosquitofish (Gambusia affinis) to
methylmercury directly from the water and via food organisms and water
in a simple model ecosystem. More mercury was taken up at higher
temperatures. The authors calculated mercury uptake from water as a
percentage of the "global" uptake from both water and food. This
percentage varied with temperature, being 83% at 10°C, 40% at 18°C,
and 11% at 26°C.

In studies by Boudou & Ribeyre (1984), alevins of rainbow trout
(Salmo gairdneri) were exposed to a constant water concentration of
methylmercuric or mercuric chloride at 1 µg/litre for 83 days. Mercury
uptake was faster with organic than inorganic and both were initially
linear. A plateau was eventually achieved in both cases. Uptake was
negatively related to fish weight, although the authors pointed out
that in the field there is usually a positive relationship.

Fang (1973) maintained the pond weed Elodea canadensis, snail
Helisoma campanulata, coontail plant Ceratophyllum demersum, and
guppy Lebistes reticulatus in solutions containing labelled
phenylmercuric acetate (PMA) at concentrations between 5 × 10^-8 and
5 × 10^-7mol/litre. All of the organisms readily accumulated PMA and
the uptake was related to the length of exposure and the
concentration. The absorbed PMA was largely converted to inorganic
mercury. Although the uptake curves were very similar, pond weed and
coontail both accumulated much more PMA than guppy or snail. The half-
life of Hg²⁰³ residues ranged from 43 to 58 days. When Fang (1974)
exposed Lebistes reticulatus and Ceratophyllum demersum to
labelled ethylmercuric chloride (EMC), the uptake was positively
related to the time of exposure over 200 h and the concentration up to
5 × 10^-7mol/litre. Highest concentrations were accumulated in the
internal organs. The half-life of EMC was 20-23 days. Both organisms
converted EMC to inorganic mercury, 34% being converted by the
coontail and 29% by the guppy over a 7-day period. When the same
organisms were exposed to methylmercury chloride, little or no
breakdown to inorganic mercury occurred.

4.3 Uptake and Loss in Terrestrial Organisms

Appraisal

The accumulation of mercury in plants increases with increasing
soil mercury concentration. Soil type has a considerable influence on
this process, a high organic matter content decreasing the uptake.
Generally, the highest concentrations of mercury are found at the
roots, but translocation to other organs (e.g., leaves) occurs. In
contrast to higher plants, mosses take up mercury via the
atmosphere.

In exposed birds, the highest mercury levels are generally found
in liver and kidneys. Methylmercury is more readily absorbed than
inorganic mercury and it exhibits a longer biological half-time.
Depending on speciation, mercury occurs in different compartments of
birds' eggs; methylmercury tends to concentrate in the white and
inorganic mercury in the yolk.

Huckabee & Janzen (1975) found that the mat-forming moss
Dicramum scoparium did not take up radioactively labelled mercury
from substrate. The authors concluded that the uptake of mercury into
this point was mostly from the atmosphere. This is commonly true for
mosses, which have been used extensively as monitor organisms for
atmospheric pollutants in the field. Weaver et al. (1984) maintained
bermuda grass (Cynodon dactylon) in three types of soil (clay, silt
loam, and fine sand) treated with mercuric chloride (1-50 mg/kg).
Mercury was accumulated into the roots from silt loam, clay, and sand
in increasing order. The accumulation increased with increasing
mercury concentration. At 50 mg/kg the concentration of mercury in
(and on the surface of) the roots was 800 mg/kg, when the grass was
grown in sand.

John (1972) grew eight types of food crop in soil treated with
mercuric chloride at 4 or 20 mg mercury/kg, and uptake was measured
after 35 to 130 days, depending on the plant species. Higher
concentrations of mercury were found in the roots compared to the
above-ground samples. At the highest treatment level the mercury
content of the roots, calculated on a dry weight basis, ranged from
0.387 mg/kg for lettuce to 2.447 mg/kg for cauliflower. Of the edible
plant parts, spinach leaves and radish tubers contained the highest
concentrations (0.695 and 0.663 mg/kg mercury, respectively).

Siegel & Siegel (1985) found that the seed-pods of several
leguminous species exposed to soil mercury concentrations of
10-69 µg/kg lost 75-85% of their tissue water during maturation but
showed no loss of mercury content. However, the seeds not only lost
most of their water but also at least 75% of their mercury. The
authors suggested that the elimination was by "bio-volatilisation",
i.e., loss of elemental mercury as vapour rather than by
translocation.

Nuorteva et al. (1980) reared blowfly (Lucilia illustris) on
trout flesh contaminated with mercury (0.66 mg/kg). Levels rose from
0.14 to 1.18 mg/kg during the larval feeding period, whereas pupae and
freshly emerged adults contained 0.99 and 1.01 mg/kg, respectively.
When adults were then fed honey, mercury levels were reduced to a
third within 2 days. The authors found that it was easier for the
flies to eliminate inorganic mercury than methylmercury. Nuorteva &
Nuorteva (1982), after rearing blowfly larvae on mercury-contaminated
fish flesh and obtaining mercury levels of 2, 6.3, and 13.3 mg/kg in
different groups, fed the flies to staphylinid beetles (Creophilus
maxillosus) for a 1-week period. This gave residues of 6.9, 17.4,
and 33.4 mg/kg, respectively, in the beetles.

Kiwimae et al. (1969) fed white leghorn hens for 140 days on a
diet containing 400 or 1600 µg of mercury per day as either mercury
nitrate, phenylmercury hydroxide, or methoxyethylmercury hydroxide.
The total mercury accumulated in the egg-whites of eggs laid was 0.31,
0.53, and 0.46 mg/kg, respectively, for the lower dose and 0.44, 0.85,
and 0.88 mg/kg for the higher dose. At the higher dose, the mercury
residue in the egg yolks was 2.12, 4.53, and 2.89 mg/kg, for the three
mercury compounds, respectively.

Backstrom (1969) administered labelled mercury compounds, either
parenterally or perorally, to Japanese quail and studied the tissue
uptake and elimination. The route of administration did not affect the
final uptake or subsequent elimination. Methylmercury was readily
absorbed and was stable, while the other compounds, phenylmercury,
methoxyethylmercury, and inorganic mercury, were less well absorbed,
and the phenylmercury was rapidly decomposed to inorganic mercury.
Methylmercury was characterized by an even tissue distribution and a
slow excretion, which was enhanced in egg-laying hens. The author
attributed this to an increased concentration of methylmercury in the
egg-white. Little of the other compounds were taken up into the brain,
but methylmercury slowly reached a high concentration. The other
mercury compounds were accumulated in the yolks of the eggs laid,
and also in the liver and kidneys of the adult birds, and were rapidly
excreted. The plumage and other keratinised structures strongly
concentrated mercury, irrespective of the compound. These structures
seem to be an important excretion route, especially for methylmercury.

Nicholson & Osborn (1984) fed juvenile starlings (Sturnus
vulgaris) on a mercury-contaminated synthetic diet
(1.1 mg mercury/kg) and analysed the birds after 8 weeks. The highest
mercury levels were found in the kidneys and the liver (36.3 and
6.55 mg/kg dry weight, respectively).

In studies by Finley & Stendell (1978), black ducks (Anas
rubripes) were fed a diet containing 3 mg mercury/kg (as
methylmercury dicyandiamide) for periods of 28 weeks over two
consecutive breeding seasons, during which time any ducklings that
hatched were also fed the dosed diet. Mercury levels were highest in

the feathers of the adult birds (61 mg/kg wet weight), followed by the
liver and kidneys (22 and 14 mg/kg, respectively). Similarly the
highest levels were also found in the feathers, liver, and kidneys of
first-year ducklings. Eggs and embryos analysed during the first year
revealed mercury levels of 6.14 and 9.62 mg/kg, respectively. Mercury
residues in eggs, embryos, and ducklings were, on average, about 30%
lower during the second year. Stickel et al. (1977) dosed mallard
(Anas platyrhynchos) with 8 mg mercury/kg for 2 weeks, and found
that the highest levels of mercury were accumulated in the liver
(16.5 mg/kg wet weight) and the kidney (17.6 mg/kg wet weight). One
week later the liver and kidney had retained 64 and 66%, of the
mercury, respectively. No significant additional loss was noted during
the next 8 weeks.

Adams & Prince (1976) showed that ring-necked pheasants
(Phasianus colchicus) accumulated more mercury in the tissues after
consuming methylmercury dicyandiamide than after consuming the
corresponding mass of phenylmercuric acetate. This reflects the
greater toxicity of alkyl mercury compounds than aryl ones.

When Borg et al. (1970) fed goshawks (Accipiter gentilis) liver
and muscle from chickens dosed with methylmercury (average dietary
mercury content 13 mg/kg), the hawks died within 6-7 weeks. The
highest residues of mercury were found in the liver at 113 mg/kg wet
weight (102 mg methylmercury/kg), and the kidneys at 129 mg/kg
(98 mg methylmercury/kg). Substantially higher levels of mercury were
found in the skeletal muscle and brain of treated birds than in those
of controls. The reproductive organs also showed an ability to
accumulate mercury.

4.4 Accumulation in the Field

Appraisal

Observations on given species of marine and freshwater fish
indicate that all tissue concentrations of mercury increase with
increasing age (as inferred from length) of the fish. In certain
species males have been found to have higher levels than females.

In aquatic systems, fish-eating birds tend to have higher
mercury levels than non-fishing birds. In terrestrial systems, seed-
eating birds, small mammals, and their predators can have high levels
in areas where methylmercury fungicides are used.

Bird feathers are useful for biological monitoring for
methylmercury exposure. Analysis of feathers, especially using
neutron activation, can allow recapitulation of past exposure. In
general liver and kidney have higher levels than other bird tissues.

Sea mammals are reported to have a wide range of total mercury
concentrations in liver (0.4 to over 300 mg/kg), only a small
fraction (2-17%) being in the methylated form. Selenium and mercury
have been found in seal livers in a consistent 1:1 atomic ratio. A
number of studies have indicated that selenium plays a protecting
role.

Point sources of mercury pollution often lead to elevated
mercury levels in organisms living in the affected area. There are
some circumstances where toxic effects have been produced. These
effects should be taken into account in various countries during the
process of industrialization.

4.4.1 General exposure

Gilmartin & Revelante (1975) analysed Northern Adriatic anchovy
(Engraulis encrasicholus) and sardine (Sardina pilchardus) for
mercury content. Seasonal distribution of mercury in various tissues
of both anchovy and sardine ranged between 5 and 610 ng/g wet weight,
the highest concentrations of mercury being in the liver and kidney.
Perttila et al. (1982) found that mercury levels in the Baltic herring
(Clupea harengus) increased significantly with age. Bache et al.
(1971) observed that concentrations of both total mercury and
methylmercury increased with the age of lake trout (Salvelinus
namaycush), the proportion of methylmercury to total mercury
increasing with age. However, Westoo (1973) did not find that the
proportion of methylmercury to total mercury in salmon (Salmo salar)
and sea trout (Salmo ocla) was dependant on age.

Forrester et al. (1972) found a correlation between length and
mercury concentration in Squalus acanthias (the spurdog, an
elasmobranch fish). Olsson (1976) analysed northern pike (Esox
lucius) in 1968 and 1972 and found a correlation between mercury
levels and length of fish, and that males contained significantly more
mercury than females. It was considered that, during a general
decrease of mercury levels within pike population, the age of the fish
is not a suitable parameter for estimating mercury levels. This is
because uptake and retention of mercury is dependant on body size but
loss of accumulated mercury is less dependent on fish size. May &
Mckinney (1981) sampled freshwater fish, in 1976 and 1977, from
selected sites throughout the United States, and found mercury levels
of 0.01-0.84 mg/kg wet weight.

Berg et al. (1966) analysed feathers from Swedish birds collected
over a period of 100 years, and found roughly constant levels of
mercury during the period 1840 to 1940. However, a well documented
increase of 10-20 times appeared in the 1940s and 1950s, which the
authors concluded was due to the use of alkylmercury seed dressings.
Martin (1972) and Martin & Nickerson (1973) sampled starlings
throughout the United Slates in 1970 and 1971 and found that most of
the birds had mercury levels of < 0.5 mg/kg (76% of the birds

analysed in 1971 contained levels of < 0.05 mg/kg). Lindsay & Dimmick
(1983) found mercury in the liver, breast muscle, and body fat of wood
duck taken from the area of the Holston River, Tennessee, USA. The
highest levels were in juveniles (0.42, 0.15, and 0.1 mg/kg, for the
three tissues, respectively. Local sediment contained
0.76 mg mercury/kg, black fly larvae and aquatic plants < 0.1 mg/kg.

Osborn & Nicholson (1984) sampled puffin from the islands of
St. Kilda and May, off the British coast, and found liver and kidney
mercury levels of approximately 1.25 mg/kg dry weight (in both
tissues) for the Isle of May, and 3.75 and 5 mg/kg dry weight,
respectively, for St. Kilda. Braune (1987) analysed tissues of nine
species of sea birds sampled in New Brunswick, Canada, for total
mercury content, and found highest levels in the liver (0.046 to
0.606 mg/kg) and kidney (0.242 to 5.345 mg/kg). Birds which fed on
benthic invertebrates or fish had the highest levels, while those
feeding mainly on pelagic invertebrates had the lowest.

Fimreite et al. (1982) sampled eggs from a Norwegian gannet
colony for mercury in 1972, 1978, and 1979, and obtained values of
0.58, 0.8, and 0.36 mg/kg, respectively. Ohlendorf (1986) analysed
eggs from three Hawaiian seabird species in 1980, and found mercury in
all eggs, ranging from 0.122 to 0.359 mg/kg wet weight. Koeman et al.
(1975) analysed oiled seabirds (guillemot and razorbill) from the
Dutch coast for mercury residues and reported levels ranging from 1.8
to 2.4 mg/kg wet weight. Hoffman & Curnow (1979) analysed the levels
of mercury in the tissues of herons, egrets, and their food collected
from two sites near Lake Erie, USA. One population fed on Lake Erie
(food items, 0.02-0.81 mg/kg wet weight; bird livers, 3.0-16.5 mg/kg
wet weight). The other population fed predominantly on bordering
marshland (food items, up to 0.24 mg/kg; bird livers,
1.03-8.22 mg/kg).

Honda et al. (1986) sampled striped dolphin (Stenella
coeruleoalba) and found that the accumulation of total mercury in
bone correlated significantly with age. Levels rose to 1.44 and
1.55 mg/kg for adult male and female, respectively, and similar trends
were seen for methylmercury, levels reaching 0.27 mg/kg in adults.
Falconer et al. (1983) found that in common porpoise (Phocoena
phocoena) highest mercury levels were in the liver, where mean
levels for females were 6.03 mg/kg and for males 3.42 mg/kg.
Heppleston & French (1973) analysed tissues of common and grey seals,
from the British coast for mercury and found highest levels in the
livers (4.9-113 mg/kg). Koeman et al. (1975) determined mercury levels
of 0.37-326 mg/kg in the livers of marine mammals (seals, dolphins,
and porpoises) and also reported an almost perfect correlation between
mercury and selenium content of these mammals (1:1 ratio between
mercury and selenium concentrations). The authors suggested that
selenium uptake may protect marine mammals from the toxic effects of

mercury. Gaskin et al. (1974) found liver total mercury levels ranging
from 13 to 157 mg/kg in short-finned pilot whales and long-snouted
dolphins from the Lesser Antilles. Between 2% and 17% of the total
mercury was methylated.

4.4.2 Mercury manufacturing and general industrial areas

Yeaple (1972) analysed bryophytes from various localities of
eastern USA for mercury content and found that highest levels
(1.45 mg/kg) were in plants from a large city. Levels in cities and
industrial areas were higher than those in rural areas (e.g.,
< 0.05 mg/kg in a high, isolated mountain area). Kraus et al. (1986)
collected leaves of the salt marsh cordgrass (Spartina alterniflora)
from two sites in the USA, one site near a heavily industrialized area
and the other in a non-industrialized area. The mean soil
concentrations of mercury for the two sites were 18.17 and 0.22 mg/kg,
respectively, while the residues in the leaves were 0.16 and
0.02 mg/kg, respectively. Salts collected from the surface of plants
in the contaminated area contained 0.11 mg mercury/kg; laboratory
studies have shown the plant capable of mercury excretion.

Nuorteva et al. (1980) analysed trout (Salmo trutta) from the
Idrijca River, Yugoslavia, about 3 km downstream from a mercury
distillation plant. The fish had a mercury content of 0.66 mg/kg in
the flesh, and highest levels were found in the spleen and kidney
(17.5 and 24 mg/kg, respectively). Three samples of ephemerids, taken
6 km from the plant, contained 0.27, 0.36, and 0.56 mg/kg wet weight,
and a sample containing 4.28 mg/kg was found 1 km from the plant.
These were lower levels than those reported previously, presumably
because of six months inactivity at the plant. The same authors
analysed blow flies from various polluted and non-polluted localities.
From an unpolluted area mercury levels were < 0.1 mg/kg, near a
Finnish pulp factory, 0.2 mg/kg, and near a caustic soda factory,
0.3 mg/kg. Higher levels (0.8 mg/kg) were found close to a mercury
mine and distillation plant in Yugoslavia, whereas levels were near
normal 1 km upstream or downstream from the mine.

Doi et al. (1984) analysed feathers from birds collected over a
period of 25 years from the mercury-polluted shores of the Shiranui
Sea, Japan. Relatively high levels were found until the late 1970s
even though the draining of water containing methylmercury from a
local factory was stopped in 1968. Mean mercury levels were: fish-
eating birds, 7.1 mg/kg; omnivorous waterfowl, 5.5 mg/kg; predatory
birds, 3.6 mg/kg; omnivorous terrestrial birds, 1.5 mg/kg; and
herbivorous waterfowl, 0.9 mg/kg.

Fimreite et al. (1971) analysed 156 fish and 48 bird livers from
the Great Lakes area of Canada in 1968 and 1969. Elevated mercury
levels were found in all fish samples, highest levels occurring in
lake trout, pumpkinseed sunfish, and walleye (10.5, 7.09, and

5.01 mg/kg, respectively). Levels were generally highest in fish
collected downstream from suspected sources. The highest mercury level
in a fish-eating bird was found in a red-necked grebe, where the liver
level was 17.4 mg/kg. Three grebes sampled showed a range of
0.45-17.4 mg/kg. Lower concentrations were found in cormorants,
herons, murrelets, terns, kingfisher, and other fish-eating birds, but
mean mercury liver burden was greater in these birds than in non fish-
eating species.

4.4.3 Mining activity

Huckabee et al. (1983) monitored levels of mercury in vegetation
in the vicinity of the mercury mine at Almaden in Spain. Mean
concentrations of total mercury in vegetation ranged from > 100 mg/kg
within 0.5 km of the mine to 0.20 mg/kg 20 km from the mine. There was
still a significantly higher mercury content in vegetation 25 km
upwind from the mine (about 10 times the background level). Mosses
were found to contain the greatest concentration of mercury
(7.58 mg/kg), and woody plants accumulated less of the metal
(0.72 mg/kg) than herbaceous plants (2.25 mg/kg). The figures given
are for samples collected in spring. There was a correlation between
distance from the mine and plant mercury content for woody plants and
mosses but not for herbaceous plants. No methylmercury, at
quantifiable levels, was found in any of the plants analysed, although
traces were seen in several samples indicating a methylmercury content
of less than 10 pg per sample.

When Phillips & Buhler (1980) analysed rainbow trout (Salmo
gairdneri), stocked in a reservoir contaminated by a disused mine,
for mercury, they found that lateral muscle tissue levels increased
linearly during the first five months that the fish were in the
reservoir. Trout sampled 7, 19, or 31 months after being introduced
showed levels that did not differ significantly (mean level = 1.25 mg
mercury/kg). Matsunaga (1975) analysed crucian carp, dace, and zacco
temmincku from two rivers receiving discharge from mercury mines in
Japan. Total mercury levels in the fish were approximately
0.2-4.5 mg/kg and reflected the levels of mercury in the water
(4-50 ng/kg).

Hesse et al. (1975) determined total mercury concentrations in
the muscle, liver, and kidney of 22 species of birds collected from a
western South Dakota watershed contaminated by mining activity.
Elevated mercury levels were found in fish-eating birds, especially
double-crested cormorants. Levels in non fish-eating birds were lower
but still significantly higher than background. In general, greater
accumulations occurred in the livers of fish-eating birds (0.89 to
30.9 mg/kg) and in the kidneys of non-fish-eating birds (0.27 to
0.60 mg/kg).

4.4.4 Chloralkali plants

Gardner et al. (1978) analysed sediment, plants, and animals from
a salt marsh contaminated by a chloralkali plant in Brunswick,
Georgia, USA. Chloralkali plants produce metallic mercury from salts.
Sediment levels ranged from 0.27 to 1.7 mg/kg dry weight for the top
5 cm and they varied according to distance from plant and depth of
sample. The roots of Spartina alterniflora, the marsh grass,
contained the highest levels (0.07-1.47 mg/kg dry weight) within the
plant. Of the animals analysed from the contaminated marsh and nearby
river, the invertebrates contained 0.3-9.4 mg/kg dry weight, the fish
0.3-1.9 mg/kg dry weight, the birds 2.4-37.0 mg/kg dry weight (liver)
and the mammals 3.8-15 mg/kg dry weight (liver). Methylmercury levels
were low (< 0.002 mg/kg) in sediment and plants but accounted for
most of the mercury found in the tissues of higher organisms.

Hildebrand et al. (1980) sampled fish and invertebrates from the
Holston River, USA, above and below an inactive chloralkali plant.
Rock bass and hog sucker contained total mercury levels at less than
1 mg/kg above the plant, and 1-3 mg/kg immediately below it. Benthic
invertebrates gave a similar pattern, lower levels being found above
the plant and the higher levels below it. Total mercury concentrations
in the individual taxonomic groups of the invertebrates ranged from a
maximum of 3.75 mg/kg (Hydropsychidae, 3.7 km below the plant) to a
minimum of 0.016 mg/kg (Psepheridae, 5.5 km above the plant). Total
mercury concentrations in fish and invertebrates decreased with
distance down stream of the plant. Mercury in the methyl form
comprised 91.7% of total mercury in the fish and 50% in the
invertebrates.

Wallin (1976) reported that samples of the carpet-forming moss
Hypnum cupressiforme from sites around six Swedish chloralkali
plants all contained similar mercury levels. Levels were highest
(1-15 mg/kg) close to the plants and decreased with increasing
distance from each plant. Background levels for the region
(90-150 µg/kg) were reached at distances of 9-15 km from the plants.
The author calculated that only a small part of the annual fallout
(< 10%) was deposited locally. Shaw & Panigrahi (1986) analysed soil
and five species of dwarf plants, from an area adjacent to a
chloralkali factory, for mercury content. Soil from around the roots
of the plants was analysed, and the mercury content was found to be
very variable (2.13-893 mg/kg dry weight). Uptake into the roots,
stem, leaf, and fruit of all plants in the area was significant.
Leaves contained the highest levels of mercury, ranging between 2.32
and 38.8 mg/kg dry weight. Greater accumulation of mercury was found
in the stem than roots of Croton sp. and Jatropha sp.; similar
amounts in both stem and roots of Argemone sp., and more mercury in
the roots than the stem of Ipomoea sp. and Calotropis sp. No
correlation was found between the soil mercury level and plant uptake.
Bull et al. (1977) measured mercury in soil, grass, earthworms, and

small mammals near a chloralkali factory. At a distance of < 0.5 km
from the factory, mean mercury levels in surface soil (3.81 mg/kg dry
weight), grass (4.01 mg/kg dry weight), earthworms (1.29 mg/kg wet
weight) and moss bags (63 ng/dm² per day) were significantly higher
than levels found 10 to 30 km from the works. Levels of mercury at
this distance were comparable with those found at sites not associated
with mercury sources. Mercury levels in all tissues analysed, except
muscle of bank voles (Clethrionomys glareolus) and woodmice
(Apodemus sylvaticus) were significantly higher in the study area
than control areas. The authors also found elevated levels of
methylmercury in small mammals and earthworms in the study area,
suggesting methylation of the inorganic mercury fall-out.

4.4.5 Mercurial fungicides

Fimreite et al. (1970) found that seed-eating birds, and their
avian predators, had higher liver mercury levels in areas where
treated grain (mercurial fungicide) had been sown compared with areas
using untreated grain. Jefferies & French (1976) analysed specimens of
the long-tailed field mouse (Apodemus sylvaticus) taken from a wheat
field that had been drilled two months previously with wheat dressed
with dieldrin and mercury. Whole body mercury concentrations were much
higher (0.83 ± 0.44 mg/kg wet weight) than those found immediately
after drilling (0.39 ± 0.04 mg/kg wet weight).

5. TOXICITY TO MICROORGANISMS

Mercury in an inorganic form is toxic to microorganisms. It is
much more toxic in an organic form, owing to increased availability of
the metal to cells. The following are illustrative examples, rather
than an exhaustive cover, of research into the effects of mercury on
microorganisms.

Wood (1984) discussed six protective mechanisms available to
microorganisms (and certain higher organisms) that increase their
resistance to metal ions in general, and specifically to mercury.
These mechanisms are biochemical in nature and, generally, render the
mercury ion ineffective in disturbing the normal biochemical processes
of the cell. The mechanisms are: (a) efflux pumps that remove the ion
from the cell, a process which requires energy, (b) enzymatic
reduction to the less toxic elemental form; (c) chelation by
intracellular polymers (not firmly established for mercury);
(d) binding of mercury to cell surfaces; (e) precipitation of
insoluble inorganic complexes, usually sulfides and oxides, at the
cell surface; and (f) biomethylation with subsequent transport through
the cell membrane by simple diffusion. It is this last mechanism,
biomethylation, which renders the mercury more toxic to higher life-
forms.

5.1 Toxicity of Inorganic Mercury

Appraisal

Inorganic mercury is toxic to microorganisms over a wide range
of concentrations. Its effects on development and survival are
modified by environmental factors such as temperature, light
intensity, pH, and chemical composition of the medium, and by cell-
related factors such as genetic variation. Through selective effects
on particular species, it can change the composition of a plankton
community. The mechanism of action is not fully understood.

5.1.1 Single species cultures

Kamp-Nielson (1971) demonstrated a time-dependent effect of
mercuric chloride, added at 300 µg/litre, on the photosynthesis of
Chlorella pyrenoidosa. There was little effect in the first hour of
incubation, a pronounced drop in photosynthetic rate in the second
hour, and a period of little further effect between 2 and 5 h. An
overall rate reduction of about 50% occurred after 5 h with a cell
density of 6.5 × 10⁷ cells/litre. There was a greater effect on
photosynthesis at lower cell densities. It was also found that
photosynthesis had to occur for the effect to develop, since exposure
to mercuric chloride for 2 h in the light had the same effect as
exposure to the same concentration of mercury for 2 h in the dark
followed by 2 h in the light. Similar results were found after 1-h
exposures in light and darkness followed by light. There was an effect

of light intensity; in short-term experiments mercury had a
deleterious effect on photosynthesis only at high light intensities.
Mercury also affected photosynthesis at low light intensity, but only
after 20-h exposures. Mercury affected photosynthesis adversely at
concentrations between about 50 and 300 µg/litre, but had no greater
effect at concentrations up to 1000 µg/litre (the highest tested). The
effect was dependant on cell density, pH, light intensity, and
duration of exposure. Potassium and sodium in the growth medium had no
effect on mercury toxicity to Chlorella. Increasing the
concentration of mercuric chloride in the medium increased the
"leakage" of potassium from the cells of Chlorella. This was maximal
at a mercury concentration of about 300 µg/litre and was considered to
be the main toxic effect of mercury. The effect on potassium leakage
occurred equally in darkness and light and was, therefore, independent
of the photosynthetic effect. Mercury increased the length of the
lag-phase during the growth of Chlorella pyrenoidosa cultures. A
greater effect was seen at 660 than at 330 µg/litre, the only two doses
tested. This effect was also demonstrated by Osokina et al. (1984) in
the green alga Scenedesmus quadricauda. The effect was highly
dependant on the cell density of the original inoculum.

Rai et al. (1981) exposed Chlorella vulgaris to mercuric
chloride concentrations between 100 and 1000 µg/litre for 3 weeks, and
monitored growth and survival. LC₅₀ for survival was at 400 µg/litre
of mercuric chloride. The growth rate was 92% of the control value at
100 µg/litre and 31% at 800 µg/litre, and there was no growth at
1000 µg mercuric chloride/litre. The chlorophyll content of the cells
was reduced throughout the dose range. There was a greater toxic
effect of mercuric chloride at low pH, with the greatest amelioration
of toxicity at pH 9. There was also a protective effect of calcium and
phosphate in the medium and, to a lesser extent, of magnesium. Both
calcium and phosphate increased the yield of algae, in the presence of
sublethal concentrations of mercury, when added at concentrations up
to 20 mg/litre. At higher concentrations of both calcium and
phosphate, the protection was less marked. Den Dooren de Jong (1965)
determined the no-observed-effect-level (NOEL) for mercuric chloride
on Chlorella vulgaris to be 50 µg/litre. Hannon & Patouillet (1972)
emphasized the irreversibility of the effects of mercuric chloride on
Chlorella pyrenoidosa. If mercury was present in sufficient
concentration to affect growth of the alga, then no recovery was found
following transfer in clean medium. Similar effects were reported for
three species of marine unicellular algae. Mercury toxicity was
dependant on cell numbers in the initial inoculum (Kuiper, 1981). In
studies with unialgal cultures of Chlamydomonas sp., there was a
relationship between cell concentration and mercury toxicity. The
author attributed this to a surface area effect, the metal is being
adsorbed onto cell walls to cause its effect on the unicellular algae.

Huisman et al. (1980) investigated the effect of temperature on
the toxicity of mercuric salts to the green alga Scenedesmus acutus.
Mercury concentration in the cultures was kept constant by a

mercury(II) buffer system, and the growth and photosynthesis of the
alga were monitored. Toxicity increased with increasing temperature
over the range 15-30°C. There was no effect observed in this study on
the lag phase, no later increase in growth, and no effect of initial
cell numbers. This was attributed to the buffer system which prevented
changes in free mercury concentrations over time. The authors also
examined the binding of mercury to algal cells. Metal bound to the
cell wall consists of two fractions: one which can be washed off with
cysteine solution and one which cannot. The amount of mercury which
can be washed off the cell wall increase with increasing temperature.
The mercury bound to cell walls, but washable with cysteine, appears
to be the toxic fraction. The total mercury content of algal cells
does not correlate with effect. A total mercury content not lethal at
15°C causes complete inhibition of growth and photosynthesis at 30°C.
Recovery occurs under circumstances where the cells retain the non-
washable mercury, indicating that the washable fraction is the toxic
component. The authors suggested that the reversibility of the action
of cysteine-washable mercury indicates that the metal is bound to
carboxyl or phosphate groups and not to sulfhydryl groups. These
mercury ions can be readily exchanged for other metal ions, leading to
a decreased inhibition by mercury. Therefore, in media with a high
concentration of dissolved salts, mercury appears to be less toxic.
The authors postulated another mechanism by which mercury might be
toxic to algal cells. Interference with potassium-sodium-dependent
ATPase in the cell membrane influences the active transport of
nutrients. This would give rise to disturbances of nitrogen metabolism
and also of photosynthesis. The delayed action of mercury on cultures
could be ascribed to their being initially rich in nitrogen, and,
therefore, less susceptible to nitrogen starvation.

Nuzzi (1972) exposed Phaeodactylum tricornutum, Chlorella sp.,
and Chlamydomonas sp., isolated from the lower Hudson River, New
York, USA, to mercuric chloride. The growth of all three organisms was
severely inhibited by mercury at 7.5 µg/litre (to between 50% and 75%
of control growth). The growth of Chlamydomonas sp. was completely
inhibited by 15 µg mercuric chloride/litre and the other two species
by 22 µg/litre.

Gray & Ventilla (1971) found no effect of mercuric chloride on
growth of