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 the marine ciliate Cristigera spp. at a concentration of
100 µg/litre, but growth was affected after exposure to 200 or
500 µg/litre. There was a synergistic interaction between mercury and
lead on this ciliate. Gray & Ventilla (1973) reported reductions in
growth rate of between 8% and 12% after exposing Cristigera to
mercuric chloride at 25 or 50 µg/litre. Persoone & Uyttersprot (1975)
found no effects of mercuric chloride, at concentrations up to
100 µg/litre, on the survival or reproduction of the marine ciliate
Euplotes vannus. However, all cells died after exposure to
1000 µg/litre.

5.1.2 Mixed cultures and communities

Singleton & Guthrie (1977) investigated the effects of inorganic
mercury, added as mercuric chloride at 40 µg/litre, on populations of
bacteria from fresh and brackish water. Water was taken from the two
sources and kept for 1 week in the laboratory before the metal salt
was added. Results were assessed by measuring total colony-forming
units (viable bacteria), percentage of chromagenic organisms, and
numbers of different colony types (species diversity). Control systems
maintained constant numbers of viable bacteria throughout the 14-day
test period. When mercury was added, the numbers of viable bacteria
from test samples increased and remained elevated throughout the test.
The effect was greater in brackish than in fresh water. Diversity
declined at the same time as total numbers increased. Some genera of
bacteria disappeared from the community, notably Flavobacterium and
Brevibacterium. Other organisms which disappeared or were greatly
reduced included Sarcina sp., Enterobacter sp., Achromobacter sp.,
and Escherichia sp. After mercury treatment, the percentage of
chromagenic species decreased in the population. Controls maintained
chromagens at a steady 20-25% of total bacteria. Chromagen percentage
declined most markedly after 9-10 days of mercury exposure.

Kuiper (1981) exposed a mixed community of marine plankton to
mercuric chloride (at 0.5, 5.0, or 50 µg mercury/litre) in 1400-litre
plastic bags suspended from a raft in a Netherlands harbour. The
addition of 50 µg mercury/litre resulted in complete inhibition of
phytoplankton activity. There was a decrease in phytoplankton biomass
because of settling of cells to the bottom of the bag. Phytoplankton
growth resumed after about 20 days when mercury concentrations were
still at 18 µg/litre. There was evidence for two possible mechanisms
for this: either mercury-resistant species were growing or mercury was
being adsorbed to inanimate particles or removed by chelation.
Addition of 5.0 µg/litre reduced phytoplankton growth rate. Biomass
decreased initially but began to increase again when the mercury
concentration decreased to about 1.5 µg/litre. Mercury at 5.0 µg/litre
delayed the phytoplankton peak by 9 days but relative carbon
assimilation by only 1 day. One possible explanation is that mercury
affected cell division more than carbon assimilation. Both 5.0 and
50 µg/litre altered the species composition of the growth peak; higher
mercury concentrations favoured the selection of larger species. The
first stage in the uptake and toxicity of mercury in phytoplankton is
adsorption on cell surfaces (e.g., cell walls); the smaller surface to
volume ratio of larger cells may explain why larger cells are more
resistant to higher mercury concentrations. Another possible
explanation involves predation; reduced numbers of predatory
zooplankton might favour larger phytoplankton cells which might be
preferred by predators. There was some evidence to support both
hypotheses. Zooplankton were also affected by mercury. There was
immediate death of most copepods after the addition of mercuric

chloride at 50 µg mercury/litre. Development of the copepods Temorus
longicornis and Pseudocalanus elongatus was delayed by
5.0 µg/litre. The results suggest that the major effect on these
zooplankton is a retardation of development rather than an increase in
mortality. Laboratory experiments simulating conditions in the bags
suggested that zooplankton grazing on phytoplankton was an important
factor in the productivity of the bags during the second, but not the
first, half of the experimental period. On day 10 of the experiment,
viable bacteria numbers were higher in bags with 5.0 and 50 µg
mercury/litre than in controls. This was probably due to the high
mortality of phytoplankton increasing the food source for bacteria.
Conversion rate of organic matter into ammonia was reduced. The author
concluded that the toxicity of mercury to plankton depends on mercury
concentration, total surface area for adsorption of mercury (and,
therefore, on the ratio between living and non-living particles
present and on absolute cell size), and on the metal species present
(Kuiper, 1981).

Hongve et al. (1980) added mercuric salt, alone or in combination
with humus or sediment, to cultures of a natural phytoplankton
community in lake water, and monitored photosynthetic carbon fixation
using a radiolabelled tracer. Mercury reduced carbon fixation by 50%
at the lowest dose tested (5 × 10^-9 mol/litre) and to less than 10%
of control levels at the highest dose tested (2 × 10^-7 mol/litre).
Addition of either humus or sediment to the cultures reduced mercury
toxicity presumably by binding the metal to surfaces.

Zelles et al. (1986) conducted a complex and comprehensive
experiment to compare different methods for assessing the overall
ecotoxicological effects of chemicals on soil microorganisms. Three
soil types were used in an 18-week experiment which investigated ATP,
heat production, respiration (as measured by carbon dioxide output),
and iron reduction in the soils under dry and moist conditions. Two
different dose levels of mercuric chloride were added to the soils
(2 and 20 mg/kg). Averaging the results obtained in the different
tests, adverse effects on microorganisms were least in peat soil and
greatest in sandy soil. Some stimulation of microbial activity
occurred in peat soil with both low and high concentrations of
mercuric chloride. At both 2 and 20 mg/kg mercury there was inhibition
of microbial activity in sandy soil. Effects were generally inhibitory
in clay soil at both concentrations of mercuric chloride. The authors
pointed out that it is not possible to assess the ecotoxicological
effects of mercury on soil by using a single method to assess soil
function.

5.2 Toxicity of Organic Mercury

Appraisal

Methylmercury is more toxic to microorganisms than are inorganic
mercury salts. This is probably because greater surface adsorption
enhances the availability and subsequent uptake of methylmercury.
This may explain why the toxicity of organomercury is inversely
correlated with cell density. As the surface area of the total cells
in the culture increases, so less mercury is available for uptake per
cell. In organomercury compounds, it is the mercury-containing
moiety, as opposed to the dissociable anion, which determines the
toxicity. A common toxic effect in phytoplankton is the inhibition of
growth, which may in turn often be due to reduced photosynthesis.

Methylmercury in water at 1 µg/litre has adverse effects on
microorganisms.

Ukeles (1962) tested the effect of Lignasan (ethylmercuric
phosphate 6.25%) on a variety of algae in pure culture. The cultures
were exposed to Lignasan at 0.6, 6.0, and 60 µg/litre for 10 days. The
highest dose of 60 µg/litre prevented all growth of cultures, and at
the end of the exposure, all cells were killed by the treatment. Three
out of the five algae tested were also killed by Lignasan at
6.0 µg/litre: Protococcus sp., Chlorella sp., and Monochrysis
lutheri. Growth of the other two species was reduced; Dunaliella
euchlora showed 31% of the growth of controls and Phaeodactylum
tricornutum 17% of control growth. At 0.6 µg/litre, Lignasan reduced
growth of four of the five cultures to between 55% and 86% of control
levels, Monochrysis alone being unaffected.

Nuzzi (1972) exposed Phaeodactylum tricornutum, Chlorella sp.
and Chlamydomonas sp. to phenylmercuric acetate (PMA) at
concentrations of 0.06-15.0 µg mercury/litre. P. tricornutum was
also tested against phenylacetate equivalent to the phenylacetate
content of the PMA, but this had no effect. All three organisms were
adversely affected by the mercury in PMA, growth being inhibited even
at the lowest dose tested. Chlamydomonas was totally inhibited by
3 µg mercury/litre. Chlorella sp. showed a steep decline in growth
as exposure increased from 0.06 to 3 µg/litre, where growth was about
25% of the control value. Phaeodactylum growth declined rapidly as
dose increased to 9 µg/litre, where growth was minimal.

Holderness et al. (1975) cultured the green alga Coelastrum
microporum with methylmercuric chloride (MMC) at 0.8, 3, 6, 12.6,
and 250 µg/litre. There was no significant effect on cell
concentration, as determined by transmittance, at 0.8 µg/litre, but
higher concentrations were inhibitory. There was a steady reduction in
cell concentration between 0 and 3 µg MMC/litre and a marked decline
between 3 and 6 µg/litre, with cell concentration changing from

125 µlitre/litre, at 3 µg MMC/litre to 31 µlitre/litre at
6 µg MMC/litre. It was noted, in three series of experiments, that MMC
caused increased storage of starch in the cells. A slight increase in
photosynthesis was found after exposure to 0.6 µg MMC/litre.

Delcourt & Mestre (1978) exposed cultures of Chlamydomonas
variabilis to concentrations of phenylmercuric acetate (PMA) between
10^-9 and 7.5 × 10^-8 mol/litre. Growth curves of the control
cultures were linear, with no evident lag phase, irrespective of the
cell concentration (which varied between 2000 and 100 000 cells/ml) in
the initial inoculum. The effect of mercury as PMA was initially
tested with cell concentration at 20 000 cells/ml. Under these
conditions, cultures exposed to PMA at 10^-9 or 2.5 × 10^-9 mol/litre
grew exactly the same as controls. However, at PMA concentrations of
5 × 10^-9 mol/litre or more, there was a dose-related lag phase. When
exponential growth did start, the curves were parallel to those of the
control. Final cell numbers were not affected, only the time taken to
reach maximum growth. Changing the initial cell concentration in the
cultures changed the toxic threshold of the PMA, PMA toxicity being
higher at lower algal cell concentrations. The authors considered that
there are a limited number of binding sites for mercury on the cell
surface and that this was the reason for the effect of cell
concentration on toxic threshold. Whilst the toxic threshold was
higher than likely exposure levels in natural waters at high algal
cell concentrations, the authors pointed out that the threshold would
be exceeded at low, spring algal concentrations.

Harriss et al. (1970) exposed a pure culture of the marine diatom
Nitzschia delicatissima and a natural phytoplankton community from a
freshwater lake to four organomercurial compounds at concentrations
between 0 and 50 µg/litre. The four compounds (PMA, methylmercury
dicyandiamide [Panogen], N-methylmercuric- 1,2,3,6-tetra hydro-3,6-
methano-3,4,5,6,7,8-hexachlorophthalimide [MEMMI], and
diphenylmercury) showed broadly similar effects on photosynthesis at
the same concentrations expressed in terms of mercury content. The
diphenylmercury was slightly less toxic than the other compounds. The
diatom was exposed to the mercurials for 24 h, and the phytoplankton
community was exposed for 24, 72, or 120 h, before estimating the
photosynthetic uptake of labelled hydrogen carbonate over 5 h. At
concentrations of 1 µg/litre, all four mercurials inhibited
photosynthesis of the natural phytoplankton. Photosynthetic uptake of
labelled carbon was between 35% and 55% of control levels for the four
compounds. At 50 µg/litre, all uptake of carbon stopped and cell
counts indicated cessation of growth in the case of all compounds
except diphenylmercury. Photosynthetic carbon uptake was about 40% of
control levels after exposure for 120 h to 50 µg diphenylmercury/
litre. The authors stated that the toxicity of diphenylmercury to the
natural phytoplankton was similar to that of mercuric chloride, but no
details of studies with inorganic mercury were given. Nitzschia was
similarly inhibited by all mercurials tested, except diphenylmercury,
at 1 µg/litre. The diatom showed virtually no carbon, at assimilation

in the presence of PMA, methylmercury dicyandiamide, or MEMMI
10 µg/litre. At 1 µg/litre, the carbon assimilation was 95% of the
control value with diphenylmercury, 60% with PMA, 23% with
methylmercury dicyandiamide, and < 10% with MEMMI. The authors noted
that the toxicity of mercurials to the natural phytoplankton community
decreased with increasing cell numbers, but no details were given.

6. TOXICITY TO AQUATIC ORGANISMS

Mercury is toxic to aquatic organisms, organic forms of the metal
being generally more toxic than inorganic forms. Effects are more
likely to be observed in soft freshwater, since the toxicity of the
metal is reduced in the presence of high salt concentrations. The
concentration of mercury that produces effects varies considerably
from one species to another.

6.1 Toxicity to Aquatic Plants

Appraisal

As in the case of microorganisms, mercury, at a wide range of
concentrations, has effects on various aspects of performance,
including development and survival. These are partly the result of
adverse effects on photosynthesis.

The presence of sediment or humic material reduces the
availability of mercury to aquatic plants because of adsorption. In
studies involving a dual medium, such as soil-water, actual exposures
are more difficult to determine than in studies with a single medium,
such as water alone.

Organic forms of mercury, such as methyl- or butylmercury
chloride are more toxic to aquatic plants than inorganic forms.

Boney (1971) exposed 2-day-old sporelings of the red alga
Plumaria elegans to mercuric chloride in solution, and found that
50% growth inhibition occurred after 6 h, approximately 12 h, and
approximately 24 h at concentrations of 1.0, 0.5, and 0.25 mg/litre,
respectively. Organic forms of mercury (methyl, butyl, and
propylmercuric chlorides) were also investigated, and found to be much
more toxic than inorganic mercury. Methylmercury gave 50% inhibition
after 17.5 and 25 min of exposure to 0.08 and 0.04 mg/litre,
respectively. Propylmercury, at 0.5 mg/litre, produced 50% growth
inhibition after 2.5 min of exposure and 70% inhibition after 5 min.
Butylmercury produced more marked inhibition than propylmercury (no
detailed results given). Hopkin & Kain (1978) found that the survival
of germinating gametophytes of the macroalga Laminaria hyperborea,
in culture, was reduced by 0.01 mg mercury/litre. The lowest effective
toxic level of mercury for the sporophyte culture was 0.05 mg/litre.

Stanley (1974) determined EC₅₀s, in the presence of a mercuric
salt, for various growth parameters of Eurasian watermilfoil
(Myriophyllum spicatum) grown in soil with water above. EC₅₀s
(in mg/litre) were 3.4 for root weight, 4.4 for shoot weight, 12 for
root length, 1.2 for shoot length. The author added mercury to the
water, to the soil, or to the water in a system containing ferric
silicate instead of soil. Comparison of the tissue concentrations of
mercury when the metal salt was added in these different ways

indicated a very strong tendency for mercury to be adsorbed onto soil.
There was no indication that the presence of soil affected mercury
uptake in any way other than by simple adsorption, i.e., no soil
component interacted with the mercury ions.

De et al. (1985) exposed the floating plant water cabbage
(Pistia stratiotes) for 2 days to mercuric chloride at
concentrations between 0.05 and 20.0 mg/litre. The highest dose of
mercury promoted plant senescence by decreasing chlorophyll content,
protein, RNA, dry weight, and catalase and protease activities, and by
increasing free amino acid content. Lesser, mostly non-significant,
effects on these parameters were recorded at lower doses. In studies
by Brown & Rattigan (1979), the aquatic macrophyte Elodea canadensis
(Canadian pond weed) and the free-floating duckweed Lemna minor were
exposed for 28 days and 14 days, respectively, to a range of
concentrations of mercuric chloride. Damage to the plants was assessed
visually on a coded scale ranging from 0 (no damage) to 10 (plant
killed). Water concentrations of 7.4 and 1.0 mg/litre produced 50%
damage to the two plants, respectively. In a separate study, Elodea
was exposed to mercury for 24 h in the dark and then oxygen evolution
in the light was measured. Levels of 0.8 and 1.69 mg mercury/litre
reduced photosynthetic oxygen evolution by 50% and 90%, respectively.
Czuba & Mortimer (1980, 1982) exposed plants of Elodea densa,
growing in flowing water, to concentrations of methylmercuric chloride
at 7.5 × 10^-10, 7.5 × 10^-9, or 7.5 × 10^-8mol/litre, for 25 days.
Toxicity was assessed by gross morphological examination and from the
examination of histological sections embedded in paraffin wax. There
was a difference in toxic effect between tissues. Apical cells were
most sensitive to the mercury and developed aberrant nuclear and
mitotic characteristics at lower concentrations than did roots. Root
meristems showed total inhibition of mitotic activity at the middle
concentration but no effect at the lowest concentration used. Mitotic
activity in bud meristems was absent in controls, but increased in the
presence of methylmercury; divisions were abnormal. Higher
concentrations of methylmercury chloride, up to 2.5 × 10^-6mol/litre,
stimulated the development of additional buds. The development of root
and bud initials was inhibited by methylmercury at 7.5 × 10^-8 and
2.5 × 10^-6mol/litre, respectively.

6.2 Toxicity to Aquatic Invertebrates

Appraisal

Factors which affect the toxicity of mercury to aquatic
invertebrates include the concentration and species of mercury, the
developmental stage of the organisms, and the temperature, salinity,
water hardness, and flow rate. Methylmercury is more toxic than aryl
or inorganic mercury. The larval stage is apparently the most
sensitive stage of the organism's life cycle. Mercury toxicity
increases with temperature and decreases with water hardness.
Toxicity, appears to be higher in flow-through systems than in static

systems. This effect is probably due mostly to the actual
concentration of mercury available to the organism, which is lower in
static systems. The fact that lower salinity seems to increase
toxicity may be due more to the stress that is placed on the
organism.

Levels of 1 to 10 µg/litre normally causes acute toxicity for
the most sensitive developmental stage of many different species of
aquatic invertebrates.

The acute toxicity of mercury to aquatic invertebrates is
summarized in Tables 3 and 4.

6.2.1 Acute and short-term toxicity to invertebrates

Wisely & Blick (1967) determined the concentration of mercury in
water required to kill 50% of larvae for some species of bryozoans
(Watersipora cucullata and Bugula neritina), tubeworms (Spirorbis
lamellosa and Galeolaria caespitosa), bivalve molluscs (Mytilus
edulis and Crassostrea commercialis), and the brine shrimp
(Artemia salina). The 2-h LC₅₀s for the larvae of these species
were 5 × 10^-7, 1 × 10^-6, 7 × 10^-7, 6 × 10^-6, 6.5 × 10^-5,
9 × 10^-4, and 9 × 10^-3 mol mercuric chloride/litre, respectively.

Howell (1984) exposed two species of marine nematodes, one
euryhaline^a (Enoplus brevis) and one stenohaline^b (Enoplus
communis) to mercuric chloride. E. brevis was collected from two
sites, one nonpolluted and one polluted with heavy metals. The
stenohaline species was more sensitive to mercuric chloride than the
related euryhaline species. At a concentration of 0.01 mg mercuric
chloride/litre, E. communis showed an LT₅₀ of approximately 65 h,
whereas 50% E. brevis collected from the nonpolluted site survived
for approximately 415 h at the same concentration. E. brevis from
the polluted area was even less sensitive, with an LT₅₀ of more than
600 h, suggesting the selection of resistant strains.

When Best et al. (1981) exposed the planarian Dugesia
dorotocephala to concentrations of methylmercury chloride of between
0 and 2 mg/litre, 100% deaths were reported at 0.5, 1, and 2 mg/litre
within 5 days, 1 day, and 5 h, respectively. No deaths occurred at
0.2 mg/litre over a 10-day-period, but other, non-lethal toxic

^a tolerant of a wide range of salinity
^b tolerant of only a narrow range of salinity

Table 3. Toxicity of inorganic mercury (as mercuric chloride) to marine invertebrates

Organism Lifestage Stat/ Temperature pH Salinity Dissolved Parameter Water Reference
flow^a (°C) (%) oxygen concentration
(mg/litre) (µg/litre)

Starfish adult stat 20 7.8 20 > 4 24-h LC₅₀ 1800 Eisler & Hennekey
(1977)
(Asterias adult stat 20 7.8 20 > 4 96-h LC₅₀ 60 Eisler & Hennekey
forbesi) (1977)
adult stat 20 7.8 20 > 4 168-h LC₅₀ 20 Eisler & Hennekey
(1977)

Hard clam embryo stat 25-27 7-8.5 25 48-h LC₅₀ 4.8 Calabrese & Nelson
(Mercenaria (3.8-5.6) (1974)
merceneria)

Softshell clam adult stat 20 7.8 20 > 4 24-h LC₅₀ 5200 Eisler & Hennekey
(1977)
(Mya arenaria) adult stat 20 7.8 20 > 4 96-h LC₅₀ 400 Eisler & Hennekey
(1977)
adult stat 20 7.8 20 > 4 168-h LC₅₀ 4 Eisler & Hennekey
(1977)

American oyster embryo stat 25-27 7-8.5 25 48-h LC₅₀ 5.6 Calabrese et al.
(Crassostrea (4.2-6.8) (1973)
virginica)

Pacific oyster embryo stat 19-21 7.9-8.3 33.7-33.8 6.5-8.0 48-h EC₅₀^c 5.7 Glickstein (1978)
(Crassostrea
gigas)

Oyster larvae stat 15 48-h LC₅₀ 1.0-3.3 Connor (1972)
(Ostrea edulis) adult stat 15 48-h LC₅₀ 4200 Portmann & Wilson
(1971)

Table 3 (cont'd)

Organism Lifestage Stat/ Temperature pH Salinity Dissolved Parameter Water Reference
flow^a (°C) (%) oxygen concentration
(mg/litre) (µg/litre)

Cockle adult stat 15 48-h LC₅₀ 9000 Portmann & Wilson
(Cardium edule) (1971)

Mud snail adult stat 20 7.8 20 > 4 24-h LC₅₀ 32 000 Eisler & Hennekey
(1977)
(Nassarius adult stat 20 7.8 20 > 4 96-h LC₅₀ 32 000 Eisler & Hennekey
obsoletus (1977)
adult stat 20 7.8 20 > 4 168-h LC₅₀ 700 Eisler & Hennekey
(1977)

American lobster stage I stat 18-22 29.5-31.5 7.6-8.6 96-h LC₅₀ 20 Johnson & Gentile
(Homarus larvae (1979)
americanus)

European lobster larvae stat 15 48-h LC₅₀ 33-100 Connor (1972)
(Homarus
gammarus)

Pink shrimp adult stat 15 48-h LC₅₀ 75 Portmann & Wilson
(Pandalus (1971)
montagui)

White shrimp post- stat 21-24 25 96-h LC₅₀ 17 Green et al.
(Penaeus larval (13-21) (1976)
setiferus)

Brown shrimp larvae stat 15 48-h LC₅₀ 10 Connor (1972)
(Crangon adult stat 15 48-h LC₅₀ 3300-10 000 Portmann & Wilson
crangon) (1971)
adult stat^b 15 96-h LC₅₀ 100-330 Portmann & Wilson
(1971)

Table 3 (cont'd)

Organism Lifestage Stat/ Temperature pH Salinity Dissolved Parameter Water Reference
flow^a (°C) (%) oxygen concentration
(mg/litre) (µg/litre)

Grass shrimp stage I 26.5-27 6.3-6.9 32.73-33.29 5.6 48-h LC₅₀ 10 Shealy & Sandifer
(Palaemonetes larvae (7.8-12.7) (1975)
vulgaris) unfed
stage I 27 6.4-6.7 33.99 5.8-7.6 48-h LC₅₀ 15.6 Shealy & Sandifer
larvae (12.7-19.3) (1975)
fed

Dungeness crab 1st stage 14-16 7.9-8.3 33.72-33.86 6.5-8.0 48-h LC₅₀ 21.1 Glickstein (1978)
(Cancer zoeae (19.7-22.5)
magister)

1st stage 14-16 7.9-8.3 33.72-33.86 6.5-8.0 96-h LC₅₀ 6.6 Glickstein (1978)
zoeae (5.6-4.6)

Shore crab larvae stat 15 48-h LC₅₀ 14 Connor (1972)
(Carcinus adult stat 15 48-h LC₅₀ 1200 Portmann & Wilson
maenus) (1971)

Hermit crab adult stat 20 7.8 20 > 4 24-h LC₅₀ 2200 Eisler & Hennekey
(1977)
(Pagurus adult stat 20 7.8 20 > 4 96-h LC₅₀ 50 Eisler & Hennekey
longicarpus) (1977)
adult stat 20 7.8 20 > 4 168-h LC₅₀ 50 Eisler & Hennekey
(1977)

Crab adult stat 26.5-29.5 7-7.2 24-h LC₅₀ 930 Krishnaja et al.
(1987)
(Scylla serrata) adult stat 26.5-29.5 7-7.2 48-h LC₅₀ 800 Krishnaja et al.
(740-860) (1987)

Table 3 (cont'd)

Organism Lifestage Stat/ Temperature pH Salinity Dissolved Parameter Water Reference
flow^a (°C) (%) oxygen concentration
(mg/litre) (µg/litre)

adult stat 26.5-29.5 7-7.2 72-h LC₅₀ 680 Krishnaja et al.
(1987)
adult stat 26.5-29.5 7-7.2 96-h LC₅₀ 680 Krishnaja et al.
(600-760) (1987)

Polychaete juv stat 7.8 96-h LC₅₀ 100^d Reish et al. (1976)
(Neanthes adult stat 7.8 96-h LC₅₀ 22^d Reish et al. (1976)
arenaceodentata) juv stat 7.8 28-day LC₅₀ 90^d Reish et al. (1976)
adult stat 7.8 28-day LC₅₀ 17^d Reish et al. (1976)

Polychaete larva stat 7.8 96-h LC₅₀ 14^d Reish et al. (1976)
(Capitella adult stat 7.8 96-h LC₅₀ >100^d Reish et al. (1976)
capitata) adult stat 7.8 28-day LC₅₀ 100^d Reish et al. (1976)

Sandworm adult stat 20 7.8 20 > 4 24-h LC₅₀ 3100 Eisler & Hennekey
(1977)
(Nereis virens) adult stat 20 7.8 20 > 4 96-h LC₅₀ 70 Eisler & Hennekey
(1977)
adult stat 20 7.8 20 > 4 168-h LC₅₀ 60 Eisler & Hennekey
(1977)

^a stat = static conditions (water unchanged for duration of test).
^b static conditions but test water changed every 24 h.
^c abnormal development.
^d with food.

Table 4. Toxicity of inorganic mercury to freshwater invertebrates^c

Organism/ Stat/ Temperature Alkalinity^f Hardness^f pH Dissolved Parameter Water Reference
weight (g) flow^a (°C) oxygen concentration
(mg/litre) (µg/litre)

Mussel stat 28-32 9.5-9.9 32-38 7-7.3 5.38-6.2 24-h LC₅₀ 7390 Ramamurthi
(Lamellidens stat 28-32 9.5-9.9 32-38 7-7.3 5.38-6.2 48-h LC₅₀ 5910 et al.
maginalis) stat 28-32 9.5-9.9 32-38 7-7.3 5.38-6.2 72-h LC₅₀ 3690 (1982)
32-35

Snail (egg) stat 17 50 7.6 6.2 24-h LC₅₀ 6300 Rehwoldt
(Amnicola sp.) stat 17 50 7.6 6.2 96-h LC₅₀ 2100 et al.
(adult) stat 17 50 7.6 6.2 24-h LC₅₀ 1100 (1973)
stat 17 50 7.6 6.2 96-h LC₅₀ 80 Rehwoldt et
al. (1973)

Snail stat 28-32 9.5-9.9 32-38 7-7.3 5.38-6.2 24-h LC₅₀ 1108 Ramamurthi
(Pila globosa) stat 28-32 9.5-9.9 32-38 7-7.3 5.38-6.2 48-h LC₅₀ 369 et al.
20-25 stat 28-32 9.5-9.9 32-38 7-7.3 5.38-6.2 72-h LC₅₀ 296 (1982)

Pulmonate snail 25.5-29.5 240-278 290-335 7.4-8.1 6.0-8.1 24-h LC₅₀ 330 Mathur et al.
(297-360) (1981)
(Lymnaea luteola) 25.5-29.5 240-278 290-335 7.4-8.1 6.0-8.1 48-h LC₅₀ 188 Mathur et al.
0.46-0.72 (163-210) (1981)
25.5-29.5 240-278 290-335 7.4-8.1 6.0-8.1 96-h LC₅₀ 135 Mathur et al.
(112-186) (1981)

Crab stat 28-32 9.5-9.9 32-38 7-7.3 5.38-6.2 24-h LC₅₀ 739 Ramamurthi
(Oziotelphusa stat 28-32 9.5-9.9 32-38 7-7.3 5.38-6.2 48-h LC₅₀ 591 et al.
senex senex) stat 28-32 9.5-9.9 32-38 7-7.3 5.38-6.2 72-h LC₅₀ 443 (1982)
35-88

Table 4 (cont'd)

Organism/ Stat/ Temperature Alkalinity^f Hardness^f pH Dissolved Parameter Water Reference
weight (g) flow^a (°C) oxygen concentration
(mg/litre) (µg/litre)

Crayfish flow^b 15-17 7.0 96-h LC₅₀ 20 Boutet &
(Austropotamobius flow^b 15-17 7.0 30-day LC₅₀ 2 Chaisemartin
pallipes pallipes) flow^b 15-17 7.0 30-day LC₅₀ < 2^d (1973)

Crayfish flow^b 15-17 7.0 96-h LC₅₀ 50 Boutet &
(Orconectes flow^b 15-17 7.0 30-day LC₅₀ 2 Chaisemartin
limosus) flow^b 15-17 7.0 30-day LC₅₀ < 2^d (1973)

Scud stat 17 50 7.6 6.2 24-h LC₅₀ 90 Rehwoldt
(Gammarus sp.) stat 17 50 7.6 6.2 96-h LC₅₀ 10 et al. (1973)

Copepod stat 10 0.58 meq/ 7.2 48-h LC₅₀ 2200 Baudouin &
(Cyclops abyssorum) litre (1500-3300) Scoppa (1974)

Water flea stat 10 0.58 meq/ 7.2 48-h LC₅₀ 5.5 Baudouin &
(Daphnia hyalina) litre (3.1-9.8) Scoppa (1974)

Water flea stat 48-h LC₅₀ 1.8-4.3 Canton &
Adema (1978)
(Daphnia magna) stat 17-19 41-50 44-53 7.4-8.2 48-h LC₅₀ 5 Biesinger &
stat 17-19 41-50 44-53 7.4-8.2 21-day LC₅₀ 13^e Christensen
(9-19) (1972)

stat 11.5-14.5 390-415 235-260 7.4-7.8 5.2-6.5 24-h LC₅₀ 4890 Khangarot &
(4190-5890) Ray (1987)
stat 11.5-14.5 390-415 235-260 7.4-7.8 5.2-6.5 48-h LC₅₀ 3610 Khangarot &
(2830-4400) Ray (1987)

Table 4 (cont'd)

Organism/ Stat/ Temperature Alkalinity^f Hardness^f pH Dissolved Parameter Water Reference
weight (g) flow^a (°C) oxygen concentration
(mg/litre) (µg/litre)

Water flea stat 48-h LC₅₀ 3.0 Canton &
(Daphnia pulex) Adema (1978)

Water flea stat 48-h LC₅₀ 3.2 Canton &
(Daphnia cucullata) Adema (1978)

Copepod stat 10 0.58 meq/ 7.2 48-h LC₅₀ 850 Baudouin &
(Eudiaptomus litre (710-1020) Scoppa (1974)
padanus)

Bristle worm stat 17 50 7.6 6.2 24-h LC₅₀ 1900 Rehwoldt
(Nais sp.) stat 17 50 7.6 6.2 96-h LC₅₀ 1000 et al. (1973)

Stone fly stat 16-20 40 44 7.25 9.2 96-h LC₅₀ 2000 Warnick &
(Acroneuria Bell (1969)
lycorias)

May fly stat 16-20 40 44 7.25 9.2 96-h LC₅₀ 2000 Warnick &
(Ephemerella Bell (1969)
subvaria)

Caddis fly stat 16-20 40 44 7.25 9.2 96-h LC₅₀ 2000 Warnick &
(Hydropsyche Bell (1969)
betteni)

Caddis fly stat 17 50 7.6 6.2 24-h LC₅₀ 5600 Rehwoldt
(unidentified sp.) stat 17 50 7.6 6.2 96-h LC₅₀ 1200 et al. (1973)

Table 4 (cont'd)

Organism/ Stat/ Temperature Alkalinity^f Hardness^f pH Dissolved Parameter Water Reference
weight (g) flow^a (°C) oxygen concentration
(mg/litre) (µg/litre)

Damsel fly stat 17 50 7.6 6.2 24-h LC₅₀ 3200 Rehwoldt
(unidentified sp.) stat 17 50 7.6 6.2 96-h LC₅₀ 1200 et al. (1973)

Midge stat 17 50 7.6 6.2 24-h LC₅₀ 60 Rehwoldt
(Chironomus sp.) stat 17 50 7.6 6.2 96-h LC₅₀ 20 et al. (1973)

Midge flow^b 20 50 6.8 24-h LC₅₀ 1074 Rossaro
(Chironomus (760-1520) et al. (1986)
riparius)
4th instar flow^b 20 50 6.8 48-h LC₅₀ 316 Rossaro
larvae (230-440) et al. (1986)

flow^b 20 50 6.8 96-h LC₅₀ 100 Rossaro
(50-180) et al. (1986)

stat 20 50 6.8 24-h LC₅₀ 1028 Rossaro
(880-1200) et al. (1986)

stat 20 50 6.8 48-h LC₅₀ 750 Rossaro
(660-850) et al. (1986)

stat 20 50 6.8 96-h LC₅₀ 547 Rossaro
(480-630) et al. (1986)

^a stat = static conditions (water unchanged for duration of test).
^b intermittent flow-through conditions.
^c mercuric chloride was used except in the studies of Rehwoldt et al. (1973) and Khangarot & Ray (1987) (salt used unspecified).
^d with food.
^e extrapolated value from three concentrations less than the LC₅₀, daphnids were fed at the beginning of each week.
^f alkalinity & hardness expressed as mg CaCO₃/litre unless otherwise stated.

responses, including varying degrees of head resorption, were observed
within 1 day. This was followed by some head regeneration within 10
days. After some animals were decapitated, regeneration was retarded
at 0.1 and 0.2 mg methylmercury chloride/litre. Although no deaths,
malformations, visible lesions, or gross behavioural abnormalities
were seen at 20 µg/litre or less, significant changes in fissioning
were noted, even at the lowest mercury concentration tested
(0.03 µg/litre). Fissioning was almost completely suppressed after
3 days in 0.1 µg/litre.

When Dorn (1974) exposed the bivalve mollusc Congeria
leucophaeata for 48 h to mercuric chloride at concentrations of
0, 0.001, 0.01, 0.1, and 1.0 mg/litre, there was a significant
increase, compared with controls, in respiration rate at all dose
levels. The effect was dose related over the entire range. Stromgren
(1982) exposed the mussel Mytilus edulis to mercuric chloride and
found after 5 days a significant reduction in growth rate at 0.3 µg
mercury/litre. At concentrations > 1.6 µg/litre, growth almost ceased
within 3 to 4 days of exposure, while at 25 µg/litre acute lethal
effects were observed within 24 h. Breittmayer et al. (1981)
investigated the effects of metal concentration, size of organism, and
seasonal differences on the toxicity of mercury to Mytilus edulis.
The most important factor for mercury toxicity was season, though all
factors interacted. MacInnes (1981) studied the effect of mercury on
embryos of the American oyster Crassostrea virginica. The test was
initiated 2 h after fertilization and continued for 48 h, the embryos
then being checked for abnormal development (they did not undergo
embryogenesis). The percentage of abnormal development for the test
concentrations of 5 and 10 µg/litre were 6 and 15.7% for the chloride
salt, and 2.9 and 9.8% for the nitrate. Dillon (1977) found that the
96-h LC₅₀ for the estuarine marsh clam Rangia cuneata exposed to
mercuric chloride was reduced from 0.122 to 0.058 mg/litre with an
increase in salinity from 2 to 15%. The pre-exposure of clams to
8.56 µg mercury/litre, followed by a period in clean water,
significantly enhanced the survival of Rangia experimentally exposed
to 0.87 mg mercury/litre. Results showed an LT₅₀ of 135 h for
unexposed clams compared to an LT₅₀ of 210 h for pre-exposed clams.

Biesinger & Christensen (1972) found that in waterfleas (Daphnia
magna) reproductive impairment was a more sensitive measure of the
toxicity of mercuric chloride than survival. EC₁₆ and EC₁₀₀ values
were 3.4 and 6.7 µg mercury/litre, respectively, for a 3-week
exposure. Biesinger et al. (1982) exposed Daphnia magna to mercury
(as mercuric chloride, methylmercuric chloride, or phenylmercuric
acetate) in a chronic experiment over 3 weeks. The lowest
concentrations of the three compounds to affect survival were 1.92,
0.2-0.98, and 2.25 µg/litre, respectively. Lowest concentrations
affecting reproduction were 0.72, 0.04, and 1.90 µg/litre,
respectively. All figures are in terms of mercury concentration in
water.

Pyefinch & Mott (1948) studied the effect of mercuric chloride on
the barnacles Balanus balanoides and Balanus crenatus. The
toxicity of mercury to cyprids of B. balanoides was reduced by
dilution of the sea water to reduce salinity. Older (11-12 day) larvae
were less resistant than 1-day-old larvae. A mercury concentration of
0.01 mg/litre reduced the number of cyprids settling. Exposure of
B. balanoides and B. crenatus after metamorphosis yielded median
lethal concentrations, over 6 h, of 0.36 and 1.35 mg/litre,
respectively.

Barnes & Stanbury (1948) found the median lethal concentration of
mercuric chloride to the harpacticoid copepod Nitocra spinipes to be
0.6 mg mercury/litre. When the mercuric salt was added with copper
sulfate, the chemicals acted synergistically. Lalande & Pinel-Alloul
(1986) collected Tropocyclops prasinus mexicanus from three
different Quebec lakes, two of low water hardness
(10 mg CaCO₃/litre) and one of high (120 mg CaCO₃/litre). The lake
with the high water hardness was polluted with human effluent. Animals
from the two unpolluted lakes showed mean 48-h EC₅₀s
(immobilization) of 0.015 and 0.045 mg/litre, whereas those from the
polluted lake with a high water hardness showed an EC₅₀ of
0.199 mg/litre.

When Sheally & Sandifer (1975) exposed newly-hatched grass shrimp
(Palaemonetes vulgaris) larvae to mercury, a concentration of
56 µg/litre was lethal to all larvae within 24 h. No deaths occurred
within 48 h when the shrimps were exposed to concentrations of
3.2 µg/litre or less. At 5.6 µg/litre, there were no deaths in fed
larvae but some deaths occurred among unfed animals. The authors found
that feeding slightly increased the resistance of P. vulgaris larvae
to mercury. In surviving larvae some delayed effects of mercury were
noted. Concentrations of 10 to 18 µg/litre caused a significant
reduction in survival to the post-larval stage, a delayed moult, an
extended development time, an increase in the number of larval
instars, and an increase in the occurrence of deformities.

Portmann (1968) found that a reduction in temperature from 22°C
to 5°C increased 5-fold the tolerance of brown shrimps to mercury
(added as mercuric chloride) within 48 h. With cockles the effect was
even more pronounced, increasing the 48-h LC₅₀ by a factor of 130.
It was also found that starving the animals reduced their tolerance to
mercury. The 48-h LC₅₀ for brown shrimps was halved (from 1.3 to
0.65 mg/litre) and reduced by a third in cockles (from 15.5 to
9.6 mg/litre). Larger shrimps were more resistant to mercury; the
LC₅₀ for the largest shrimps was 1.26 mg/litre, whereas that for the
smallest was 0.58 mg/litre.

Brown & Ahsanullah (1971) studied the effects of mercuric
chloride on the mortality of the adult brine shrimp (Artemis salina)
and the worm (Ophryotrocha labronica). After exposure to

1 mg mercury/litre, the LT₅₀s were 25 h for Artemia and 0.5 h for
Ophryotrocha. Green et al. (1976) found that a 60-day exposure of
post-larval white shrimp (Penaeus setiferus) to mercuric chloride
(at either 0.5 or 1.0 µg mercury/litre) did not significantly affect
respiratory rate, growth, or moulting rate.

In studies by Chinnayya (1971), mercuric chloride in freshwater
(at 1 × 10^-7mol/litre) reduced oxygen consumption of the shrimp
Caridina rajadhari from a control level of 0.485 ml/h per g wet
weight of shrimps to 0.377 ml/h per g. This concentration of mercury
caused no mortality over 10 days. The lowest concentration causing
mortality in this species was 2.5 × 10^-7mol/litre.

Barthalamus (1977) found that concentrations of 2 and 5 mg
mercuric chloride/litre killed 100% of grass shrimps Palaemonetes
pugio, within 24 h, and 1 and 0.5 mg/litre over a period of 96 h. He
calculated the 120 h LC₅₀ to be 0.2 mg/litre, and found that
0.05 mg/litre significantly impaired the conditioned avoidance
response.

Knapik (1969) studied the toxic effect of mercuric nitrate on
four species of crustaceans, using concentrations of 10, 100, 200, and
500 mg mercuric nitrate/litre. The most sensitive species was
Neomysis vulgaris (only 10% survived for 2 h at 10 mg/litre),
followed by Palaemonetes varians and Gammarus locusta.
Rhithropanopeus harrisi tridentatus was unaffected by a 3-h exposure
to 100 mg/litre and 23% of animals survived 1 h at 500 mg/litre.

When Doyle et al. (1976) exposed crayfish Orconectes limosus to
mercuric chloride, they observed 100% survival at 0.25 mg/litre over a
period of 96 h. Survivors of a 96-h exposure to 1 mg/litre
(the LC₆₀) showed a sluggish response to mechanical stimulation.
Only occasional ventilative movements were observed in survivors of
higher concentrations. All crayfish were dead within 96 h at
5 mg/litre.

Khayrallah (1985) studied the effect of both mercuric and
methylmercuric chloride on the amphipod Bathyporeia pilosa. The
toxicity of both inorganic and organic mercury was directly related to
both concentration (0.04-0.75 mg mercury/litre) and temperature
(1, 10, and 20°C) and inversely related to salinity (10, 20, and 30%)
and age (adult and juvenile).

Meadows & Erdem (1982) calculated LT₅₀s for Corophium
volutator in 1 µg mercuric chloride/litre of about 30 days and in
1000 mg/litre of about 3 h. Krishnaja et al. (1987) studied the acute
toxicity of phenylmercuric acetate to the intertidal crab Scylla
serrata and calculated the 24-h, 48-h, 72-h, and 96-h LC₅₀s to be
700, 580, 540, and 540 µg/litre, respectively. DeCoursey & Vernberg
(1972) exposed larval stages (zoea I, III, and V) of the fiddler crab
Uca pugilator to mercuric chloride at concentrations of 0.018, 1.8,

or 180 µg mercury/litre. No stage V, and only a few of stages I and
III, survived 180 µg/litre for longer than 24 h. Vernberg et al.
(1974) found that the adult fiddler crab Uca pugilator could survive
prolonged periods of time in sea water (at 25°C and a salinity of 30%)
and at a mercuric chloride concentration of 0.18 mg mercury/litre.
However, under temperature and salinity stress, survival periods were
reduced. At 5°C and 5%, LT₅₀s were 20 and 7 days, for females and
males, respectively, and these were further reduced to 8 and 6 days,
respectively, by the addition of 0.18 mg mercury/litre. When the
temperature was increased to 35°C, crabs survived to 28 days at low
salinity, but the addition of mercury at 0.18 mg/litre again reduced
survival, with LT₅₀s of 17 days for males and 26 days for females.
Exposure of larvae revealed that 0.18 mg/litre was fatal to stage I
zoeae, the LT₅₀ being < 24-h. At 1.8 and 0.0018 mg/litre the 50%
survival times were 8 days (stage II) and 11 days (stage III),
respectively, compared to a control value of 18 days (stage IV).

McKenney & Costlow (1981) found that the survival of the
megalopae stage of the blue crab Callinectes sapidus was highest at
a salinity of 30% and significantly reduced at 10%. Mercury at
10 µg/litre significantly increased the number of deaths of megalopae
developing at 10% but not those at salinities of 20-40%. At all
salinities, fewer megalopae completed metamorphosis at 20 µg
mercury/litre. Developmental times of the megalopae in the presence of
20 µg mercury/litre were increased to 8 to 10 days when the salinity
was reduced to 10%, and increased further, to nearly 13 days.
Following metamorphosis, the crabs were found to be more resistant.
There were no significant effects of salinity or mercury on survival
or developmental duration at the first two adult crab stages.

Depledge (1984a) found that exposure of the shore crab Carcinus
maenus to 0.05 mg mercuric sulfate/litre disrupted various
endogenous rhythms. Locomotor activity increased and the mean heart
rate rose from 32.1 beats/min to 44.7, although there was no change in
the heart stroke volume (as indicated by a lack of change in the trace
height of cardiograph readings). Exposure of crabs to 1 mg/litre
suppressed cardiac activity and oxygen consumption. Alternating
periods of bradycardia and tachycardia were observed together with
marked changes in the heart stroke volume. There was an increase in
the median perfusion index (volume of blood per unit volume of
dissolved oxygen). All of the animals died within 24 to 48 h, this
being associated with a loss of the ability to osmoregulate (Depledge,
1984b).

Weis (1980) exposed the fiddler crab Uca pugilator to a mixture
of methylmercuric chloride (0.5 mg mercury/litre) and as zinc chloride
(3 mg zinc/litre) and found the effect of the combination of metals on
the retardation of limb regeneration to be additive. The effect was
also additive at a reduced salinity (7-8%).

6.2.2 Behavioural effects

Appraisal

Mercury appears to increase the probability of prey organisms
being eaten by predators (at least in a single study). Prior exposure
of prey organisms leads to the selection of a resistant strain and
the effect of mercury, at the same concentration, disappears. The
development of tolerance in invertebrates in the field must be taken
into account when evaluating laboratory studies on test animals that
have not experienced exposure to mercury before.

Kraus & Kraus (1986) tested predator avoidance in adult grass
shrimps (Palaemonetes pugio) collected from two sites, one polluted
with mercury (sediment mercury levels "as high as 10.3 mg/kg") and the
other relatively pollution-free (sediment levels of 0.05 mg/kg). The
shrimps were maintained in water containing either mercuric chloride
or methylmercuric chloride (both at 0.01 mg/litre), for 96 h prior to
testing. Killifish, collected only from the nonpolluted area, were
then added to the tanks and the time between first and second captures
of shrimp were noted. This was significantly reduced by both inorganic
and organic mercury in shrimp from the nonpolluted area. Control
shrimp from the polluted area showed a reduced capture time compared
to shrimp from the nonpolluted area, which was not reduced further by
the mercury treatment. The overall survival of shrimps from the
nonpolluted area, over 60 or 120 min of exposure to the predator, was
not significantly affected by mercury treatment. In the shrimps from
the polluted area, only the survival of shrimps in organic mercury,
over the 60-min test period, showed a significant overall effect of
the predator.

6.3 Toxicity to Fish

Appraisal

Inorganic mercury is toxic to fish at low concentrations. The
96-h LC₅₀ s vary between 33 and 400 µg/litre for freshwater fish and
are higher for sea water fish. 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 have been reported after exposure of fish
to sublethal concentrations of mercury. Reproduction is also
adversely affected by mercury.

6.3.1 Acute and short term toxicity to fish

The acute toxicity of mercury to fish is summarized in Tables 5
and 6. Schweiger (1957) investigated the effects of mercury ions on
fish and their food organisms and suggests a concentration of 0.03 mg
mercury/litre as the toxic threshold for the various species tested.

Rodgers et al. (1951) investigated the toxicity of pyridyl
mercuric acetate to three different species of trout. No deaths
occurred in either brown trout or brook trout exposed to the compound
at 10 mg/litre for 1 h. Rainbow trout were more susceptible with 99%
mortality at 13°C and 33% mortality at 8.5°C. Deaths also occurred in
rainbow trout exposed to 5 mg/litre (3% at 8.5°C; 36% at 13°C) but
little mortality was noted at 2.5 mg/litre (0% at 8.5°C; 2% at 13°C).
MacLeod & Pessah (1973) exposed rainbow trout (Salmo gairdneri) to
mercuric chloride concentrations between 0 and 2 mg mercury/litre and
calculated 96-h LC₅₀s of 0.4, 0.28, and 0.22 mg/litre at
temperatures of 5, 10, and 20°C, respectively. At 10°C, the 24-h
LC₅₀ for mercuric chloride was approximately 30 times higher
(in terms of mercury concentration) than for phenylmercuric acetate.
Turnbull et al. (1954), using bluegill sunfish, calculated that the
24-h and 48-h LC₅₀s for pyridyl mercuric acetate were 12.5 and
11.3 mg/litre, respectively. Rehwoldt et al. (1972) measured the acute
toxicity of inorganic mercury to six species of fish (Table 5), and
found that it was less when tests were conducted at 15°C than at 28°C.
Amend et al. (1969) exposed Salmo gairdneri to 125 µg ethylmercury
phosphate/litre for 1 h, and found that increasing the temperature
from 13°C to 15°C tended to increase the acute toxicity of the mercury
solution. An increase in the water hardness from 23 to 120 mg
CaCO₃/litre also decreased the toxicity. But the dissolved oxygen
content of the water had the most pronounced effect. At saturation, no
deaths occurred, even at the highest water hardness, but at a
dissolved oxygen level of < 6 mg/litre substantial losses occurred
(72-76%) and even at the lowest temperature 37% of the trout died.

Jones (1940) found that the mean survival time for the minnow
Phoxinus phoxinus in mercuric chloride rose from 15 min for
10^-3mol/litre to 230 min at 5 × 10^-6mol/litre. The addition of
enough sodium chloride to convert the whole of the mercuric chloride
into a double-chloride sodium mercuric chloride, and even the addition
of ten times this amount, did not affect the toxicity of the solution.
The addition of a considerable excess of sodium chloride caused a
marked prolongation of the survival time, the maximum effect being
attained when the solution was approximately isotonic.

6.3.2 Reproductive effects and effects on early life stages

Appraisal

The data reveal an obvious difference between static and flow
test concentrations, with LC₅₀ values being up to 150 times lower
under flow conditions. The increased LC₅₀ in the static tests may
be explained by a combination of adsorption of the compound to
surfaces of the test vessels and to the gelatinous egg surface during
embryo development. As a result, the larvae are exposed to much lower
mercury concentrations at hatching time than are present at the
beginning of the experiment. By contrast, the concentration is
maintained throughout in a flow-through system.

Table 5. Toxicity of inorganic mercury to fish^d

Organism/ Stat/ Temperature Alkalinity^b Hardness^b pH Dissolved Parameter Water Reference
weight (g) flow^a (°C) oxygen concentration
(mg/litre) (µg/litre)

Tilapia
(Tilapia
mossambica)
2.2-3.5 stat 28-30 > 4.8 48-h LC₅₀ 1000 Menezes &
(792-1261) Qasim (1983)

10-13 stat 28-32 7.7-11.7 32-38 7-7.3 5.4-6.2 24-h LC₅₀ 1256 Ramamurthi
stat 28-32 7.7-11.7 32-38 7-7.3 5.4-6.2 48-h LC₅₀ 1108 et al.
stat 28-32 7.7-11.7 32-38 7-7.3 5.4-6.2 72-h LC₅₀ 739 (1982)

Catfish stat 28 96-h LC₅₀ 350 Das et al.
(Heteropneustes (1980)
fossilis) 25

Catfish stat 28-32 7.7-11.7 32-38 7-7.3 5.4-6.2 24-h LC₅₀ 1700 Subbaiah
(Sarotherodon stat 28-32 7.7-11.7 32-38 7-7.3 5.4-6.2 48-h LC₅₀ 1500 et al.
mossambicus) stat 28-32 7.7-11.7 32-38 7-7.3 5.4-6.2 72-h LC₅₀ 1000 (1983)
25 stat 28 96-h LC₅₀ 75 Das et al.
(1980)

Catfish 24-27.5 165-190 245-285 7.1-7.7 5.5-8.2 24-h LC₅₀ 860 Khangarot
(Channa marulius) (801-916) (1981)
3-4.5

24-27.5 165-190 245-285 7.1-7.7 5.5-8.2 96-h LC₅₀ 314 Khangarot
(271-371) (1981)

24-27.5 165-190 245-285 7.1-7.7 5.5-8.2 240-h LC₅₀ 131 Khangarot
(103-158) (1981)

Table 5 (cont'd)

Organism/ Stat/ Temperature Alkalinity^b Hardness^b pH Dissolved Parameter Water Reference
weight (g) flow^a (°C) oxygen concentration
(mg/litre) (µg/litre)

Rainbow trout
(Salmo gairdneri)
0.6-3.0 stat 9.3-10.7 70 101 8.55 > 8.0 24-h LC₅₀ 903 Wobeser
(783-1023) (1975a)

9.1-15.5 stat 5 90 7.5-7.8 48-h LC₅₀ 650 MacLeod &
stat 5 90 7.5-7.8 96-h LC₅₀ 400 Pessah
13.2-21.3 stat 10 90 7.5-7.8 48-h LC₅₀ 450 (1973)
stat 10 90 7.5-7.8 96-h LC₅₀ 280 MacLeod &
18.5-27.8 stat 20 90 7.5-7.8 48-h LC₅₀ 300 Pessah
stat 20 90 7.5-7.8 96-h LC₅₀ 220 (1973)
length: 51-76mm flow 82-132 6.4-8.3 4.8-9.0 96-h LC₅₀ 33^d Hale (1977)

Banded killifish stat 28 55 8.0 6.9 24-h LC₅₀ 270 Rehwoldt
(Fundulus stat 28 55 8.0 6.9 48-h LC₅₀ 160 et al.
diaphanus) stat 28 55 8.0 6.9 96-h LC₅₀ 110 (1972)

Striped bass stat 28 55 8.0 6.9 24-h LC₅₀ 220 Rehwoldt
(Roccus stat 28 55 8.0 6.9 48-h LC₅₀ 140 et al.
saxatilis) stat 28 55 8.0 6.9 96-h LC₅₀ 90 (1972)

Pumpkinseed stat 28 55 8.0 6.9 24-h LC₅₀ 410 Rehwoldt
(Lepomis stat 28 55 8.0 6.9 48-h LC₅₀ 390 et al.
gibbosus) stat 28 55 8.0 6.9 96-h LC₅₀ 300 (1972)

White perch stat 28 55 8.0 6.9 24-h LC₅₀ 420 Rehwoldt
(Roccus americanus) stat 28 55 8.0 6.9 48-h LC₅₀ 340 et al.
stat 28 55 8.0 6.9 96-h LC₅₀ 220 (1972)

Table 5 (cont'd)

Organism/ Stat/ Temperature Alkalinity^b Hardness^b pH Dissolved Parameter Water Reference
weight (g) flow^a (°C) oxygen concentration
(mg/litre) (µg/litre)

Carp stat 28 55 8.0 6.9 24-h LC₅₀ 330 Rehwoldt
(Cyprinus carpio) stat 28 55 8.0 6.9 48-h LC₅₀ 210 et al.
stat 28 55 8.0 6.9 96-h LC₅₀ 180 (1972)

American eel stat 28 55 8.0 6.9 24-h LC₅₀ 250 Rehwoldt
(Anguilla stat 28 55 8.0 6.9 48-h LC₅₀ 190 et al.
rostrata) stat 28 55 8.0 6.9 96-h LC₅₀ 140 (1972)

Mummichog 20 20^c 8.0 96-h LC₅₀ 2000 Klaunig
(Fundulus et al. (1975)
heteroclitus)
3.3-3.5
2-6 stat 20 20^c 7.8 < 4 24-h LC₅₀ 23 000 Eisler &
stat 20 20^c 7.8 < 4 96-h LC₅₀ 800 Henneky
stat 20 20^c 7.8 < 4 168-h LC₅₀ 800 (1977)

Flounder (adult) stat 15 48-h LC₅₀ 3300 Portmann &
(Platichthys Wilson (1971)
flesus)

^a stat = static conditions (water unchanged for duration of test); flow = flow-through conditions (mercury concentration in water
continuously maintained).
^b alkalinity & hardness expressed as mg CaCO₃/litre.
^c These figures are values for salinity (expressed in %), not alkalinity.
^d Mercuric chloride was used, except in the studies of Hale (1977), where the salt used was mercurous nitrate.

Table 6. Toxicity of organic mercury to fish^c

Organism/ Stat/ Temperature Alkalinity^b Hardness^b pH Dissolved Parameter Water Reference
weight (g) flow^a (°C) oxygen concentration
(mg/litre) (µg/litre)

Blue gourami stat 26-28 7.4 10 24-h LC₅₀ 123 Roales &
(Trichogaster (115.62-130.38) Perlmutter
trichopterus) stat 26-28 7.4 10 48-h LC₅₀ 94.2 (1974)
1.5-2 (85.5-102.9) Roales &
stat 26-28 7.4 10 96-h LC₅₀ 89.5 Perlmutter
(85.38-93.62) (1974)

Rainbow trout (fry) stat 9.3-10.7 70 101 8.55 > 8 24-h LC₅₀ 84 (81-87) Wobeser
(Salmo gairdneri) stat 9.3-10.7 70 101 8.55 > 8 48-h LC₅₀ 45 (36-54) (1975a)
stat 9.3-10.7 70 101 8.55 > 8 96-h LC₅₀ 24 (22-26) Wobeser
(fingerling) 0.6-3 stat 9.3-10.7 70 101 8.55 > 8 24-h LC₅₀ 125 (120-130) (1975a)
stat 9.3-10.7 70 101 8.55 > 8 48-h LC₅₀ 66 (63-69) Wobeser
stat 9.3-10.7 70 101 8.55 > 8 96-h LC₅₀ 42 (25-59) (1975a)
(juvenile) 22.9 flow 10 90 7.5-7.8 24-h LC₅₀ 25^c MacLeod &
Pessah
(1973)
stat 18 250 24-h LC₅₀ 5^c Alabaster
stat 18 250 48-h LC₅₀ 4^c (1969)

Table 6 (cont'd)

Organism/ Stat/ Temperature Alkalinity^b Hardness^b pH Dissolved Parameter Water Reference
weight (g) flow^a (°C) oxygen concentration
(mg/litre) (µg/litre)

Brook trout (juv.) flow 11-13 41-44 45-46 6.9-7.6 7.7 96-h LC₅₀ 75 McKim et al.
(Salvelinus (1976)
fontinalis)
Lamprey (larvae) flow 12 150 146 8-8.5 24-h LC₅₀ > 166 Mallatt
(Petromyzon marinus) flow 12 150 146 8-8.5 48-h LC₅₀ 88 et al.
0.3-3 flow 12 150 146 8-8.5 96-h LC₅₀ 48 (1986)

^a stat = static conditions (water unchanged for duration of test); flow = flow-through conditions (mercury concentration in water
continuously maintained).
^b alkalinity & hardness expressed as mg CaCO₃/litre.
^c methylmercuric chloride was used, except in the studies of MacLeod & Pessah (1973) and Alabster (1969), where phenyl mercuric
acetate was used.

Selenium may increase the toxicity of mercury to fish eggs at
higher concentrations of mercury. At low water concentrations
selenium effects are additive.

Table 7 summarizes the acute toxicity of mercury to embryolarval
stages of fish.

When Ram & Sathyanesan (1983) exposed adults of the freshwater
teleost Channa punctatus for 6 months to 0.01 mg mercuric
chloride/litre, the mercury prevented oocytes development in the ovary
and spermatogenesis in the testis. The number and activity of
gonadotrophs in the pituitary were also reduced, giving the appearance
of "resting phase" at a time when full reproductive development was
expected. McIntyre (1973) exposed sperm from Salmo gairdneri to
concentrations of methylmercuric chloride between 1 µg/litre and
10 mg/litre for 30 min. The sperm-containing solution was then added
to eggs, and the percentage of fertilization was determined 17 days
later. After exposure to 0.5 mg mercury/litre, there was an increase
in the percentage of unfertilized eggs from 9.1% in controls to 12.5%
in treated samples. This effect was enhanced with increasing mercury
concentration, reaching 100% nonfertile eggs at 5 mg/litre or greater
concentrations of mercury.

Kihlstrom & Hulth (1972) transferred eggs laid by mature
zebrafishes (Brachydanio rerio) into solutions containing 10, 20, or
50 µg phenylmercuric acetate (PMA)/kg. The frequency of hatching was
significantly higher in the 10 µg/kg group than in the controls and
the same as the controls in the 20 µg/kg group. None of the eggs
transferred to the solution containing 50 µg/kg hatched. Most eggs
hatched 3 days after fertilization, the frequency of eggs hatching up
to and including the third day being significantly higher in water
containing 10 or 20 µg PMA/kg when compared to the controls.

Weis & Weis (1977) exposed early embryos of the killifish
Fundulus heteroclitus to mercuric chloride at concentrations of
0.01, 0.03, 0.1, or 1.0 mg mercury/litre. Mercury was added to the
water at the start of the experiment and the solution was not
replaced. The authors cite Jackim et al. (1970) to indicate that the
loss of mercury from the solution would have amounted to about 26%
over the course of the 96-h test period. Embryos treated at stage 12
of development (the early blastula stage) showed a reduction in axis
formation in solutions of 0.01 and 0.03 mg mercury/litre and a severe
reduction at 0.1 mg/litre. There were no forebrain defects at 0.01 mg,
but 20% of embryos showed defects after exposure to 0.03 and
0.1 mg/litre. All the embryos exposed to 1 mg mercury/litre died
before gastrulation. Embryos treated at stage 14, the late blastula
stage of development, with concentrations of mercury of
0.01-0.1 mg/litre), showed no reduction in axis formation. Negligible
defects were noted at 0.01 and 0.03, but 20% of embryos were affected
at 0.1 mg mercury/litre.

Table 7. Toxicity of inorganic mercury to the embryo-larval
stages of fish

Organism Stat/ LC₅₀ 95% confidence
flow (µg/litre) limits

Rainbow trout stat^a 4.7 4.2-5.3
(Salmo gairdneri) flow^b < 0.1

Channel catfish stat^a 30.0 26.9-33.2
(Ictalurus punctatus) flow^b 0.3 0.2-0.4

Bluegill sunfish stat^a 88.7 73.5-106.3
(Lepomis macrochirus)

Goldfish stat^a 121.9 112.3-132.1
(Carassius auratus) flow^b 0.7 0.6-0.8

Redear sunfish stat^a 137.2 115.0-162.8
(Lepomis microlophus)

Largemouth bass stat^a 140.0 128.7-151.9
(Micropterus salmoides) flow^b 5.3 5.0-5.6

^a static conditions but water renewed every 12 h.
^b flow-through conditions (mercury concentration in water
continuously maintained).

Exposure was initiated 30 min to 2 h after spawning and continued
through to 4 days post-hatching. Hatching times were 24 days for
rainbow trout, 6 days for channel catfish, and 3 to 4 days for
the other fish. Therefore, total exposure was as follows: rainbow
trout 28 days, channel catfish 10 days, and the other fish 7 to 8
days. (Birge et al. 1979).

When Sharp & Neff (1980) exposed embryos (4-8 cell stage) of
Fundulus heteroclitus to mercuric chloride at concentrations of
0-100 µg mercury/litre for 1 to 32 days, survival was reduced at all
concentrations above 40 µg/litre. The hatching success of embryos
exposed for 32 days was significantly reduced at concentrations above
10 µg/litre. Reducing the duration of exposure from 5 days to 1 day
significantly increased the total hatchability of the eleutheroembryos
emerging after exposure for 32 days. Increases in the incidence of
spinal curvature were also noted at concentrations exceeding
20 µg/litre, which were significantly reduced if the exposure was

reduced to 5 days or less. The 24-h LC₅₀ for the embryos was
89.6 µg/litre, the 24-h EC₅₀ for spinal curvature was
61.45 µg/litre, and the 24-h EC₅₀ for hatching success was
71.6 µg/litre.

McKim et al. (1976) exposed three generations of brook trout
(Salvelinus fontinalis) to methylmercuric chloride concentrations of
0.03 to 2.93 µg/litre, over a 144-week period. At the highest dose,
deformities were observed during the first 39 weeks and 88% of the
first generation adults died. At 0.93 µg/litre, the second generation
fish showed deformities and all but one female died during a 108-week
exposure. No significant effects on survival, growth, or reproduction
were observed in second generation trout at concentrations lower than
0.93 µg/litre, and no toxic symptoms were found in the third
generation below 0.29 µg/litre. The authors established that the
maximum acceptable toxicant concentration (MATC) for brook trout
exposed to methylmercuric chloride (hardness = 45 mg CaCO₃/litre;
pH = 7.5) was between 0.29 and 0.93 µg/litre.

Weis & Weis (1984) measured the tolerance of eggs of the
killifish Fundulus heteroclitus to methylmercury in four successive
years of sampling in the same pond. There was considerable variation
in susceptibility between eggs from different females at the beginning
of the sampling period, some females producing resistant and some
susceptible eggs. After a period of heavy rainfall in the third year,
when heavy metals and pesticides were washed into the ponds, the
proportion of resistant eggs in the population increased. The authors
noted an initial correlation between production of resistant eggs and
numbers of fin rays in females. The same correlation indicated an
increase in these females in the population after exposure to metals.
Selection, rather than physiological adoption, had taken place.

Birge et al. (1979) investigated the effects of combinations of
mercury and selenium on the hatchability of eggs of the rainbow trout,
catfish, goldfish, and bass. Mercury and selenium were added to the
test medium in a 1:1 ratio over a wide range of concentrations (from
1 to 2500 µg/litre). From separate tests with mercury and selenium
alone, the author calculated additive values for the two materials and
compared the results with those observed with the mixture. For each
species, results were dependant on actual concentration. At lower
concentrations the interaction between mercury and selenium was
additive or antagonistic, whereas at higher concentrations interaction
was synergistic, with the mixture leading to much greater inhibition
of hatching than predicted. Calculated additive LC₅₀s for mercury
and selenium were 0.09 mg/litre for trout, 0.1 mg/litre for catfish,
0.67 mg/litre for goldfish, and 0.35 mg/litre for bass. Actual LC₅₀s
for the mixture of 1:1 mercury:selenium were 0.01 mg/litre for trout,
0.01 mg/litre for catfish, 0.16 mg/litre for goldfish, and
0.35 mg/litre for bass, in all cases substantially greater toxicities
than predicted. For the two most sensitive species, trout and catfish,
synergism became evident at water concentrations of 5 µg/litre and

increased in parallel with increasing concentration. At water
concentrations of 75 µg/litre, the predicted hatchability of eggs
(assuming mercury and selenium effects to be additive) was 44% for
trout and 57% for catfish. Actual observed hatchability at this
concentration was 0% for trout and 2% for catfish.

6.3.3 Behavioural effects

Weir & Hine (1970) pretrained goldfish (Carassius auratus) to
avoid electric shock with a light stimulus and then exposed them to
solutions of mercuric chloride. The lowest concentration of mercuric
chloride found to significantly impair the behavioural response was
3 µg/litre. The lowest concentration causing deaths, under the same
conditions, was 360 µg/litre. Hartman (1978) fed rainbow trout
(Salmo gairdneri) for a year on a diet containing ethylmercury
( p-toluene sulfonanilide) "Ceresan" at 0.5-25 mg/kg diet each day,
or 2.5 or 10 mg/kg delivered every fifth day of feeding. Fish
receiving 10 mg/kg every 5 days or 5 mg/kg or more per day were
unable, with few exceptions, to learn to avoid a shock preceded by a
signal of light. However, there was no evidence of the impairment of
general behaviour.

When Sharma (1984) exposed Channa punctatus to mercuric
chloride (at concentrations of 0.034, 0.068, 0.102, or 0.136 mg/litre
for 1, 7, 15, 30, or 45 days), hyperactive avoidance reaction was seen
after exposure to the two highest doses within 24 h. Similar reactions
occurred with the lower doses after 5 days (0.034 mg/litre) and 2 days
(0.068 mg/litre). Acute distress symptoms were noted at the lowest two
exposure levels during the last 5 days of the experiment. Feeding was
normal up to 20 days of exposure at 0.034 mg mercury/litre, 12 days at
0.068 mg/litre, 6 days at 0.102 mg/litre, and only 3 days at
0.136 mg/litre. There were deaths in all treated groups within
45 days, ranging from 16% at 0.034 mg/litre to 100% at 0.102 mg
mercury/litre or more. Growth was inhibited by all treatments in
proportion to the mercury dose. Blood glucose level showed an early
elevation followed by a significant reduction, the timing of the
effect varying according to dose. There was also a progressively
significant depletion in liver and muscle glycogen which was similarly
dose dependant.

6.3.4 Physiological and biochemical effects

Panigrahi & Misra (1978) found that concentrations of mercuric
nitrate of 5 mg/litre or more killed all test fish Anabas scandens
within 24 h. At 3 mg/litre, the fish survived but showed pathological
and biochemical disorders. The major clinical disorders (lack of
movement and reduced food consumption) showed themselves within 5 days
of exposure. After 3 weeks, 29% of the fish were blind and their
respiratory rate was greatly reduced; 71% were blind within 4 weeks.
When the fish were transferred to clean water, partial recovery to

normal respiratory rate occurred. Considerable reductions in blood
haemoglobin content, erythrocyte count, body weight, and body protein
content were recorded.

Lindahl & Hell (1970) exposed the roach Leuciscus rutilus to
phenylmercuric hydroxide at 1 mg/litre for 40 min, then killed the
fish and measured the respiration rate of isolated gill filaments.
Filament respiration was reduced by about 30%, the cause being damage
to secondary lamellae, and the oxygen content of blood was reduced by
82%. An in vitro study with erythrocytes showed that half the cells
haemolyzed after exposure for 55 min to 0.5 × 10^-4mol phenylmercuric
hydroxide/litre.

Hara et al. (1976) studied the effect of mercuric chloride on the
olfactory response of the rainbow trout Salmo gairdneri. Mercury
depressed the response, the lowest concentration to cause an
appreciable effect within 2 h being 100 µg/litre. The depression
increased with increases in mercury concentration and exposure time.

Hilmy et al. (1982) exposed the cyprinodont Aphanius dispar to
acute concentrations of mercury of 1-12 mg/litre for 96 h or chronic
concentrations of 1 mg/litre for up to 30 days. The acute treatment
caused significant increases in plasma sodium, calcium, and potassium
levels, which reached maxima of 3, 5, and 12 mg/litre, respectively.
At the chronic exposure, the levels of sodium, calcium, and potassium
initially rose, then fell to near normal levels by the end of the
30-day experiment.

Das et al. (1980) studied the acute and subacute toxicity of
mercuric chloride to the air-breathing fish Heteropneustes fossilis
and the non-air-breathing fish Sarotherodon mossambica. The air-
breathing fish was more resistant to mercury, giving a 96-h LC₅₀
value of 350 µg mercury/litre compared to 75 µg/litre for
Sarotherodon. The effect on several enzymes of mercury at
50 µg/litre was also studied. Gill lysosomal acid phosphatase and
liver microsomal glucose-6-phosphatase were significantly stimulated
in both species, whereas liver acid phosphatase and intestinal
alkaline phosphatase were significantly stimulated in Heteropneustes
and significantly inhibited in Sarotherodon. In both species serum
glucose levels were significantly increased and liver glycogen levels
decreased, while muscle glycogen levels were unaffected.

Gill & Pant (1985) exposed Barbus conchonius to concentrations
of mercuric chloride of 36, 60, or 181 µg/litre, the highest dose
corresponding to the 96-h LC₅₀ for the species. Acute exposure to
181 µg/litre for 24 or 48 h led to deformities in the erythrocytes:
vacuolation, nuclear deterioration, microcytosis, and collapsed
cytoplasmic membranes. There was also significant thrombocytosis and
neutropenia. Chronic exposure to 36 or 60 µg mercury/litre led to

poikilocytosis, hypochromia, fragmentation and nuclear displacement of
erythrocytes, thrombocytosis, lymphocytosis, neutropenia, and mild
basophilia.

When Ramalingam & Ramalingam (1982) exposed the catfish
Sarotherodon mossambicus to a concentration of mercuric chloride of
0.09 mg mercury/litre, they found no effect on the liver or muscle
total protein content over 24 h. There were, however, significant
decreases after both 7 and 15 days.

Verma et al. (1984) dosed the lungfish Notopterus notopterus
with mercuric chloride concentrations of 0.017-0.088 mg/litre for up
to 60 days. Concentrations of 0.022 or more caused significant
increases in serum glutamic oxaloacetic transaminase and serum
glutamic pyruvic transaminase activities within 15 days. The lowest
dose took at least 30 days to significantly increase the activity of
the same enzymes.

O'Connor & Fromm (1975) exposed rainbow trout Salmo gairdneri
to methylmercuric chloride, at 10 µg mercury/litre, in a flow-through
system. The fish were killed and assayed at 4, 8, and 12 weeks. There
was no significant difference in plasma electrolyte concentrations
(Na⁺, K⁺, Cl^-, Mg²⁺, and Ca²⁺) or between the in vitro
oxygen consumption of excised gill filaments from control and mercury-
treated fish determined in 10% or 100% phosphate-buffered saline.

In studies by Sastry et al. (1982), the freshwater murrel Channa
punctatus was exposed to mercury either directly once into the
intestinal sac (0.001-10 mmol/litre) or in the water at 3 µg/litre for
15 or 30 days. A significant decrease in the rate of intestinal
absorption of glucose, fructose, glycine, and tryptophan occurred at
the higher concentrations of 0.1, 1.0, and 10 mmol/litre. At 0.01 and
0.001 mmol/litre there was a reduction in absorption but this was not
significant except at 0.01 mmol/litre in the case of tryptophan. There
was a significant decrease of the absorption rate of all four
nutrients in the mercury solution, but only after a 30-day exposure.

Dawson et al. (1977) exposed juvenile striped bass Morone
saxatilis to 1.0, 5.0, or 10 µg mercuric chloride/litre for between
30 and 120 days. The fish were then allowed to recover for a further
30 days in clean running sea water. Fish exposed to the lowest dose
did not differ significantly from controls with regard to respiration
rate. Exposure to 5 µg/litre for 30 days significantly lowered the
respiration rate but the effect had disappeared after 60-days
exposure. Fish exposed to 10 µg mercury/litre showed a decreased
respiratory rate after 30 days, which was reversed until a significant
increase in rate was observed after 120 days of exposure. Mercury
exposure did not significantly affect liver activities of aspartate
aminotransferase, glucose-6-phosphatase, malic dehydrogenase, or
magnesium activation of aspartate aminotransferase.

Christensen (1975) examined a range of biochemical parameters in
brook trout (Salvelinus fontinalis) embryos and alevins exposed to
methylmercuric chloride at concentrations from 0.01 to 1.03 µg
mercury/litre. The fish were exposed as eggs for 16-17 days and then
for a further 21 days as alevins. There was a significant decrease in
glutamic oxaloacetic transaminase activity in embryos after exposure
to 1.03 µg/litre, and a significant increase in its activity in
alevins at 0.93 µg/litre. The alevin effect was accompanied by a
significant decrease in weight. Christensen et al. (1977) exposed
brook trout to methylmercuric chloride concentrations of 0.01 or
0.03 µg/litre and 2.93 µg/litre, for either 2 or 8 weeks. After
8 weeks, they found no significant effects on body weight, body
length, blood plasma glucose, chloride or sodium, or plasma lactic
dehydrogenase, and glutamic oxaloacetic transaminase activities. There
were, however, significant increases in haemoglobin and blood plasma
sodium and chloride after 2 weeks, but no effect on the other
parameters measured.

Varanasi et al. (1975) noted structural alterations in the
epidermal mucus of rainbow trout exposed to 1 mg of mercuric
chloride/litre. Mercury accumulated in the mucus and altered the
physical characteristics of the layer, which is important for
locomotion and protection of the fish. Lock & Overbeeke (1981) studied
the effects of methylmercuric chloride and mercuric chloride on mucus
production in rainbow trout. Of three measurements made, density of
mucus cells, mucus in the tissue, and release of mucus into water,
only the latter was affected by mercury. The effect was less with
organic than with inorganic mercury, where mucus production was
increased significantly. Exposure to 10 µg inorganic mercury/litre for
4 h increased mucus production, and greater exposure concentrations
and times enhanced the effect. Opercular movements increased with
increased mucus production, suggesting mucus-induced hypoxia. Lock et
al. (1981) attributed the osmoregulatory effect of mercury on fish as
an effect on the permeability of the gill to water, rather than as an
effect on active ionic transport.

Roales & Perlmutter (1977) found that methylmercury (9 µg/litre),
or methylmercury and copper combined, resulted in a decrease in the
immune response of blue gourami (Trichogaster trichopterus) to both
infectious pancreatic necrosis (IPN) virus and Proteus vulgaris. The
two toxicants jointly produced no greater or lesser effect than when
each was added alone.

6.4 Toxicity to Amphibia

Mercury has a toxicity for amphibian tadpoles similar to that
for fish. There is considerable species variability in susceptibility
to the metal. Sublethal effects and developmental effects have been
reported. There is no information on effects on adult amphibians.

Acute toxicity of mercury to amphibian tadpoles is summarized in
Table 8.

Birge et al. (1979) conducted embryo-larval bioassays on 14
species of amphibia. Exposure to inorganic mercury was maintained from
fertilization to 4 days after hatching, using static renewal
procedures (Table 9). Gastrophryne and five species of Hyla were the
most sensitive, with LC₅₀ values ranging from 1.3 to 2.8 µg/litre,
compared to an LC₅₀ value of 4.7 µg/litre for rainbow trout (the
exposure period was shorter than for the trout; 6.6 to 7.4 days
compared to 28 days.

Chang et al. (1974) dosed leopard frog (Rana pipiens) tadpoles
with methylmercuric chloride, either via the water at concentrations
of 0-1.0 mg mercury/litre or via injections of 0.025 mg mercury/day
for 10 days. There was 100% mortality after 48 h at a water
concentration of 50 µg/litre or more. At 1-10 µg/litre there was total
arrest of development and differentiation after 48 h, which continued
for 3 to 4 months. Mercury-injected tadpoles showed extensive
deposition of blood pigment in their livers. The authors suggest that
this was due to haemolysis of red blood cells caused by mercury,
followed by severe peripheral oedema and haemopoietic reactions in the
kidneys of the tadpoles. Dial (1976) exposed Rana pipiens embryos
(at the cleavage, blastula, gastrula, and neural-plate stages of
development) to concentrations of methylmercuric chloride of
0.5-200 µg/litre. Concentrations of 40 µg/litre or more were lethal to
embryos treated at the cleavage stage. Embryos at the blastula,
gastrula, and neural-plate stages were treated for 5 days at
concentrations of 5-30 µg/litre. Tadpoles treated with 5 µg/litre
showed only minor effects, whereas 10, 15, or 20 µg/litre caused
various effects, including exogastrulae, poor tail development, and
poor general development. Death rates increased with exposure time and
concentration. At 30 µg/litre many defects were observed after 24 h
and all tadpoles had died within 3 days.

6.5 Toxicity to Aquatic Mammals

There appears to be only a single experimental study on the
effects of methylmercury on aquatic mammals. Ronald et al. (1977) fed
harp seals on herring dosed with methylmercuric chloride. Two animals
were used as controls, two were fed 0.25 mg/kg body weight per day and
two fed 25.0 mg/kg body weight per day. Various blood parameters were
monitored and found to be unaffected by the lower dose. The two
animals on the higher dose died after 20 and 26 days of dosing. Prior
to death these animals exhibited toxic hepatitis, uremia, and renal
failure.

Table 8. Toxicity of mercuric chloride to amphibians

Organism Lifestage Stat/ Temperature Alkalinity^e Hardness^e pH Parameter Water Reference
flow^a (°C) concentration
(µg/litre)

Frog^c tadpole stat 13-16 24-40 13-80 6.2-6.7 24-h LC₅₀ 762 (677-837) Khangarot
(Rana hexadactyla) tadpole stat^b 13-16 24-40 13-80 6.2-6.7 48-h LC₅₀ 121 (93-151) et al. (1985)
tadpole stat^b 13-16 24-40 13-80 6.2-6.7 72-h LC₅₀ 68 (57-85) Khangarot
tadpole stat^b 13-16 24-40 13-80 6.2-6.7 96-h LC₅₀ 51 (33-53) et al. (1985)

Clawed toad 3- to stat 19-21 48-h LC₅₀ 100 de Zwart &
(Xenopus laevis) 4-week Slooff (1987)
larva

Toad^d tadpole stat 29-34 120-160 165-215 7.1-7.6 12-h LC₅₀ 69.8 Khangarot &
(Bufo melanostictus) tadpole stat 29-34 120-160 165-215 7.1-7.6 24-h LC₅₀ 52.8 Ray (1987)
(43.6-61.5)
tadpole stat 29-34 120-160 165-215 7.1-7.6 48-h LC₅₀ 45.6 Khangarot &
(40.9-56.7)
tadpole stat 29-34 120-160 165-215 7.1-7.6 96-h LC₅₀ 43.6 Ray (1987)
(36.8-58.5)

^a stat = static conditions (water unchanged for duration of test).
^b static conditions but test water renewed every 24 h.
^c tadpole length 15-25 mm, weight 350-800 mg (wet weight).
^d tadpole length 18-22 mm, weight 90-120 mg (wet weight).
^e alkalinity & hardness expressed as mg CaCO₃/litre.

Table 9. Toxicity of inorganic mercury to the embryo-larval stage
of amphibians

Organism LC₅₀ 95% confidence
(µg/litre) limits

Narrow-mouthed toad 1.3 0.9-1.9
(Gastrophryne carolinensis)
Southern grey tree frog 2.4 1.5-3.4
(Hyla chrysoscelis)
Squirrel tree frog 2.4 1.5-3.8
(Hyla squirrella)
Barking tree frog 2.5 1.7-3.4
(Hyla gratiosa)
Grey tree frog 2.6 1.2-4.2
(Hyla versicolor)
Spring peeper 2.8 1.9-3.9
(Hyla crucifer)
Leopard frog 7.3 4.8-10.0
(Rana pipiens)
Cricket frog 10.4 8.5-12.6
(Acris crepitans blanchardi)
Red-spotted toad 36.8 18.3-51.1
(Bufo punctatus)
Green toad 40.0 25.6-52.2
(Bufo debilis debilis)
River frog 59.9 53.8-65.9
(Rana heckscheri)
Fowlers toad 65.9 44.0-84.0
(Bufo fowleri)
Pig frog 67.2 54.3-79.5
(Rana grylis)
Marbled salamander 107.5 72.5-153.5
(Ambystoma opacum)

Exposure was under static conditions (but water renewed every
12 h), and was initiated 30 min to 2 h after spawning and
continued to 4 days post-hatching. Hatching times varied from 2.6
to 3.4 days, therefore total exposure was between 6.6 and 7.4
days. (Birge et al., 1979).

7. TOXICITY TO TERRESTRIAL ORGANISMS

7.1 Toxicity to Terrestrial Plants

Appraisal

The main problem with studies on the effects of mercury on
terrestrial plants is their relevance to the natural situation.
Mercury normally binds to soil particles, which may reduce its
availability to plants. In most studies, mercury has been
administered as a solution in hydroponic culture. Most of the
experiments have been on crop plants; wild plants might behave
differently.

Oberlander & Roth (1978) measured the uptake and translocation of
potassium and phosphate, into the roots and shoots of 7-day-old barley
plants, from doubly labelled (⁴²K, ³²P) nutrient solutions
containing mercuric chloride. Uptake and translocation was monitored
over 5 h during exposure to mercury at 10^-4mol/litre. Potassium and
phosphate uptake was significantly reduced to 21% and 31%,
respectively, of the control level. Potassium and phosphate
translocation was also significantly reduced to 6% and 8%,
respectively, of the control level.

Barker (1972) exposed explants of cauliflower inflorescence stem,
lettuce stem, secondary phloem of carrot root, and tubers of potato
for 20 days to mercuric chloride at concentrations between 0.005 and
50 mg mercury/litre of medium. There was a significant reduction in
growth (measured as mean fresh weight) after exposure to 0.5 mg/litre
or more, although carrot and potato showed significant increases in
growth at low levels (0.005 mg/litre) of mercury.

Mhatre & Chaphekar (1984) exposed young plants of three species
(a cereal Pennisetum typhoideum, a forage crop Medicago sativa, and
a vegetable Abelmoschus esculentus) to solutions containing mercuric
chloride at 1-1000 µg mercury/litre for 24 h. They then estimated the
percentages of leaf area injured and number of leaves injured.
Abelmoschus was found to be the least sensitive of the plants,
showing no damage at 10 µg/litre, whereas the other two species showed
injury at this concentration. All species showed increasing
percentages of leaf area injury and number of leaves injured with
increasing mercury exposure. At the highest dose, 1000 µg/litre, all
leaves were injured in Pennisetum and Abelmoschus and 50% of the
leaves of Medicago.

7.2 Toxicity to Terrestrial Animals

7.2.1 Toxicity to terrestrial invertebrates

Appraisal

The experimental information available on the effects of mercury
on terrestrial invertebrates is insufficient to make any proper
appraisal.

Marigomez et al. (1986) fed the terrestrial slug Arion ater for
27 days on a diet containing mercuric chloride at 0, 10, 25, 50, 100,
300, or 1000 mg/kg. The number of slugs dying was low in all
treatments (a maximum of three deaths out of 24 animals per treatment)
and unrelated to the dose. The results indicated that exposure of
slugs to mercury at levels likely to be found in the environment will
not kill them. A significant reduction in food consumption was noted
at mercury exposures > 10 mg/kg diet, the effect being dose-related.
A significant dose-related reduction in growth rate also occurred.
Only at the highest dose (1000 mg/kg diet) did mercury severely
disrupt growth.

Abbasi & Soni (1983) kept the earthworm Octochaetus pattoni in
cement tanks at a density of 120 animals/m³, the average density of
the species in the wild, and mixed mercuric chloride, into the soil
and animal dung mixture in the tanks to dose levels of 0, 0.5, 1.0,
2.0, or 5.0 mg mercury/kg. The experiment ran for 60 days and
estimates of mortality were used to give LC₅₀ values. There was less
than 50% mortality within 5 days. The LC₅₀ was 2.39 mg/kg at 10 days
and had fallen to 0.79 mg/kg over a 60-day exposure period. As the
mortality of adult earthworms progressed throughout the experimental
period, so the earthworms still alive reproduced more than the
controls. The reason for this effect is unclear; that the animals were
stressed by the metal is evidenced by the continuing deaths. Beyer et
al. (1985) exposed the earthworm Eisenia foetida to soil containing
methylmercuric chloride at 0, 1, 5, 25, or 125 mg/kg. All worms dosed
at 25 or 125 mg/kg died within 12 weeks. Survival at 12 weeks was 97%,
92%, and 79%, respectively, for doses of 0, 1, and 5 mg/kg.
Regeneration of amputated segments was normal after treatment with
methylmercuric chloride at 1 mg/kg soil, but reduced or eliminated by
5 mg/kg.

7.2.2 Effects of mercury on birds

Appraisal

Interpretation of the results of laboratory experiments on birds
should take into account that practically all studies have been
carried out using gallinaceous birds, which are unrepresentative of
bird species as a whole.

Birds fed inorganic mercury show a reduction in food intake and
consequently in growth. Many other sublethal effects have been
reported. Organomercury compounds are more toxic to birds and cause
reproductive impairment.

Acute toxicity to birds is summarized in Table 10. The majority
of tests have been carried out using organic mercury compounds, which
are generally much more toxic than inorganic salts. The 5-day dietary
toxicity of mercuric chloride was in excess of 3000 mg/kg diet for
those species tested. The organic mercury fungicidal preparations were
the most toxic, with 5-day dietary LC₅₀s as low as 50 mg/kg diet.

7.2.2.1 Inorganic and metallic mercury

When Beliles et al. (1967) exposed male Carneaux pigeons to
mercury vapour (0.1 mg/m³) for 6 h per day over 20 weeks, no
behavioural, histological, or gross signs of mercury toxicity were
noted. Armstrong et al. (1963) trained pigeons to respond to coloured
lights to obtain food. The birds were then exposed to mercury vapour
(17 mg/m³) for 2 h daily (5 days/week) for 30 weeks. Marked changes
in behaviour were observed, as measured by a decrease in the averaged
response rate. A return to normal response was found when exposure to
mercury ceased.

Ridgway & Karnofsky (1952) injected chicken eggs, after 4 and 8
days of development, with mercuric chloride solutions into the yolk
sac and, after 8 days of development, into the chorio-allantoic
membrane, and estimated LD₅₀s. These were 0.3, at day 4, and 3.1, at
day 8, expressed as molar equivalents of mercury, for the yolk sac
route, and 0.21 Meq, at day 8, for the chorio-allantoic route. The
result on day 4 is equivalent to a dose of 0.08 mg mercuric
chloride/egg. Birge & Roberts (1976) injected chicken eggs (into the
yolk sac), immediately prior to incubation, with mercuric chloride and
obtained an EC₅₀ for hatchability of 1.0 mg/litre yolk.

Grissom & Thaxton (1985) exposed 4-week-old male chickens to
mercuric chloride (0 or 500 mg mercury/litre) in their drinking water
for up to 15 days. Rates of growth, together with feed and water
consumption, decreased significantly within 3 days of the beginning of
mercury treatment and remained depressed throughout the study.
Mortality was greater in the mercury-treated group. Red blood cell
numbers, haematocrit, mean corpuscular volume, and haemoglobin level
increased within 3 days of the start of treatment. Mean corpuscular
haemoglobin concentration (as pg/cell) was unchanged, but mean
corpuscular haemoglobin (as % of cell) decreased.

Table 10. Toxicity of mercury to birds

Species Age Compound^a Parameter^b Concentration Reference
(mg/kg)

Japanese quail 14 days methyl mercuric chloride acute LD₅₀^c 18 (14-24) Hill & Soares (1984)
(Coturnix coturnix 14 days mercuric chloride acute LD₅₀^c 42 (33-54) Hill & Soares (1984)
japonica) 2 months ceresan M acute LD₅₀^c 668 (530-842) Hudson et al. (1984)
4 months ceresan L acute LD₅₀^c 1498 (1190-1888) Hudson et al. (1984)
14 days methyl mercuric chloride 5-day LC₅₀ 47 (36-60) Hill & Soares (1984)
14 days mercuric chloride 5-day LC₅₀ 5086 (3743-6912) Hill & Soares (1984)
14 days methoxyethylmercury chloride 5-day LC₅₀ approx. 1750 Hill et al. (1975)
14 days phenyl mercuric acetate 5-day LC₅₀ 614 (496-761) Hill & Camardese (1986)
14 days morsodren 5-day LC₅₀ 45 (40-52) Hill & Camardese (1986)
14 days ceresan M 5-day LC₅₀ 147 (120-180) Hill & Camardese (1986)

Pheasant 12 months ceresan M acute LD₅₀^c 360 Hudson et al. (1984)
(Phasianus 3-4 months ceresan L acute LD₅₀^c 1190 Hudson et al. (1984)
colchicus) 3-4 months phenyl mercuric acetate acute LD₅₀^c 169 (101-283) Hudson et al. (1984)
10 days mercuric chloride 5-day LC₅₀ 3790 (2768-5541) Hill et al. (1975)
10 days methoxyethylmercury chloride 5-day LC₅₀ 1102 (957-1263) Hill et al. (1975)
10 days phenyl mercuric acetate 5-day LC₅₀ approx. 2350 Hill et al. (1975)
10 days morsodren 5-day LC₅₀ 64 (55-73) Hill et al. (1975)
10 days ceresan M 5-day LC₅₀ 146 (127-167) Hill et al. (1975)

Mallard duck 6-8 days ceresan M acute LD₅₀^c > 2262 Hudson et al. (1984)
(Anas platyrhynchos) 3 months ceresan M acute LD₅₀^c > 2262 Hudson et al. (1984)
3 months ceresan L acute LD₅₀^c > 2000 Hudson et al. (1954)
3-4 months phenyl mercuric acetate acute LD₅₀^c 878 (169-4558) Hudson et al. (1984)
10 days mercuric chloride 5-day LC₅₀ > 5000 Hill et al. (1975)
10 days methoxyethylmercury chloride 5-day LC₅₀ approx. 280 Hill et al. (1975)
10 days phenyl mercuric acetate 5-day LC₅₀ approx. 1175 Hill et al. (1975)
5 days morsodren 5-day LC₅₀ 51 (43-60) Hill et al. (1975)

Table 10 (cont'd)

Species Age Compound^a Parameter^b Concentration Reference
(mg/kg)

10 days morsodren 5-day LC₅₀ 60 (47-76) Hill et al. (1975)
5 days ceresan M 5-day LC₅₀ approx. 54 Hill et al. (1975)
10 days ceresan M 5-day LC₅₀ approx. 50 Hill et al. (1975)

Bobwhite quail 2-3 months ceresan L acute LD₅₀^c 1060 (841-1330) Hudson et al. (1984)
(Colinus 14 days ceresan M 5-day LC₅₀ approx. 70 Hill et al. (1975)
virginianus)

Prairie chicken ceresan M acute LD₅₀^c 360 (233-566) Hudson et al. (1984)
(Tympanuchus cupido)

Chukar partridge 4 months ceresan M acute LD₅₀^c 841 Hudson et al. (1984)
(Alectoris chukar)

Grey partridge 9-20 months ceresan M acute LD₅₀^c 550 (385-786) Hudson et al. (1984)
(Perdix perdix)

Rock dove ceresan M acute LD₅₀^c 714 (437-1164) Hudson et al. (1984)
(Columba livia)

Fulvous whistling 3-6 months ceresan L acute LD₅₀^c 1680 Hudson et al. (1984)
duck
(Dendrocygna bicolor)

^a morsodren = cyano methylmercury guanidine (1.51% mercury);
ceresan M = N(ethylmercury)-p-toluenesulfonalide (3.2% mercury);
ceresan L = methylmercury 2,3-di-hydroxyl propyl mercaptide + methylmercury acetate (2.25% mercury).
^b concentrations expressed as mg/kg food, unless stated otherwise.
^c concentrations expressed as mg compound per kg body weight in a single oral dosage (i.e., birds were fed with a dosed diet for
5 days followed by a 'clean' diet for 3 days).

Grissom & Thaxton (1984) investigated the interaction of mercury
treatment (as mercuric chloride in the drinking water) and water
deprivation in chickens. Birds (3-weeks-old) were treated at a rate of
500 mg/litre water over 15 days. One group had water ad libitum,
while a second group were given limited water by intubation. Water
consumption increased as the birds grew during the experiment.
Monitored water intake was 25, 55, 70, 50, and 80 ml/kg body weight at
0-3, 3-6, 6-9, 9-12, and 12-15 days into the experiment for the
mercury-treated birds. Birds on water by intubation were given 20, 35,
60, 70, and 70 ml/kg water for the same periods of the experiment.
Water limitation resulted in a significant inhibition of the growth
rate of untreated birds within the first 3 days of the experiment and
this inhibition continued throughout the experiment. Mercury did not
cause a significant inhibition of growth until between 12 and 15 days
after the beginning of treatment. The only significant interaction
between the effects of mercury and water deprivation occurred at 15
days. Food consumption was significantly reduced in water-deprived
birds. Mercury caused a significant reduction in food intake during
the 9-12 and 12-15 day periods. Dehydration increased mortality of the
groups to 10% compared with 3.75% for controls on water ad libitum.
Mercury results in birds refusing to take water or food contaminated
with the metal. Therefore, the effects of mercury can be direct or
indirect. Direct mercury effects appear to need more than 2 weeks of
exposure to develop. Examination of the birds during a 14-day recovery
period on clean water showed incomplete restoration of normal patterns
of food and water consumption over this time.

Brake et al. (1977) treated juvenile chickens with mercuric
chloride in the drinking water (300 mg/litre) or by injection
(5 consecutive days at 3 or 12 mg/kg body weight). Growth was retarded
by the chronic treatment in drinking water and by the higher of the
two injection rates. Relative heart weights (the ratio of heart weight
to body weight) were increased by mercury in drinking water, decreased
by the higher injected dose, and unchanged by the lower injected dose.
Similar results were reported for relative aorta weights.
Electrocardiograms showed a consistent decrease in the amplitude of
R-S and T waves, with the greatest effect in the injected birds (both
doses). Histological examination of the hearts of treated birds showed
myocardial histopathological changes described as a myocarditis with
polymorphonuclear and lymphocytic infiltration and fatty degeneration.
The authors concluded that mercury causes cardiovascular disturbance
in chickens even when administered at doses which do not inhibit
growth.

Hill & Shafner (1975) fed Japanese quail from hatching to one
year of age on a diet containing mercuric chloride (0, 2, 4, 8, 16, or
32 mg mercury/kg). Food consumption, growth rate, weight maintenance,
hatchability, and eggshell thickness were unaffected. As dietary
mercuric chloride increased, so initial oviposition occurred at a
younger age. The average rate of egg production was also positively

related to the concentration of mercuric chloride. The rate of egg
fertilization, however, was generally depressed for all mercury
treatments above 4 mg/kg.

Kosba et al. (1982) dosed 8-month-old hens with mercuric chloride
in drinking water at 0, 150 or 250 mg mercury/litre. Dosing at
250 mg/litre caused a slight, but insignificant, decrease in body
weight and egg numbers. Birds given the maximum dose consumed less
food than controls, but birds on 150 mg/litre consumed more food than
controls. All treated birds laid significantly smaller eggs than
controls. Fertility and hatchability were adversely affected by
mercury, and chicks hatched from eggs laid by treated birds were
lighter.

Hill & Spares (1984) studied the sublethal effects of feeding
9-week-old Japanese quail with mercuric chloride in the diet. They
calculated EC₅₀s (a reduction to 50% of the activity of controls)
for the activities of aspartate aminotransferase,
alpha-hydroxybutyrate dehydrogenase, lactate dehydrogenase, and
ornithine carbamoyl-transferase, in blood plasma, of 8.6, 11.2, 3.0,
and 62.8 mg/kg diet, respectively.

Dieter (1974) fed male Japanese quail for 12 weeks on a diet
containing mercuric chloride at concentrations of 2, 4, and 8 mg/kg.
The dosed diets did not significantly effect the carcass or liver
weights or the blood haematocrit, and, although there was a
significant decrease in haemoglobin at the 4 mg/kg treatment, this was
not reflected in the other treatment groups. The treatments had no
significant effect on the activity of the plasma enzymes creatine
kinase, asparate aminotransferase, or fructose-diphosphate aldolase,
but cholinesterase and lactate dehydrogenase activities were altered.
The maximum decrease in cholinesterase activity amounted to 25% below
that in controls, and showed almost a linear relationship with the
logarithm of the dose. Irrespective of the mercuric chloride dose,
lactate dehydrogenase activity increased 3-fold above control values.

In studies by Scott (1977), Japanese quail were fed diets
containing mercuric sulfate (0, 100, or 200 mg mercury/kg). With the
highest dose, there was a significant reduction in the hatchability of
fertile eggs and the strength of the eggshells. There were no
significant effects on daily food intake, egg production, average egg
weight, or percentage of fertile eggs.

Nicholson & Osborn (1984) found kidney lesions in juvenile
starlings (Sturnus vulgaris) fed on a commercial diet contaminated
by mercury. Analysis of the food showed mercury levels at 1.1 mg/kg.
No signs of overt toxicity were seen in the birds. Damage to the
kidney was mainly confined to the proximal tubules, and was similar to
that found in mercury-contaminated sea birds in the field.

Bridger & Thaxton (1983) demonstrated the effects of mercuric
chloride on the humoral immune response of chickens. Three treatments
were employed: chronic treatment with mercuric chloride at
300 mg/litre of drinking water; acute low dose with five consecutive
daily injections of 3 mg mercury/kg body weight; and acute high dose
with five daily injections of 12 mg/kg. The drinking-water treatment
was inhibitory to growth, while the acute treatments were not.
Chronically treated birds also showed suppressed primary and secondary
responses to a challenge with sheep red blood cells. Immunoglobulin M
levels were reduced to a greater extent than immunoglobulin G in
chronically treated birds. The primary response to Brucellus abortus
was also suppressed in chronically treated birds, but the secondary
response was enhanced, with a greater titre of circulating antibodies.
Bridger & Thaxton (1982) exposed chicks to either mercuric chloride in
drinking water (300 mg/litre) or five consecutive daily injections of
mercuric chloride into pectoral muscle (3 or 12 mg mercury/kg body
weight). The authors found that these treatments did not significantly
affect cell-mediated immune responses, in contrast to the effects on
humoral immune responses.

7.2.2.2 Effect of organic mercury on birds

When Birge & Roberts (1976) injected chicken eggs, immediately
prior to incubation, with methylmercuric chloride, into the yolk sac
the EC₅₀ for hatchability was 0.1-0.5 mg/litre yolk.

Haegele et al. (1974) dosed female mallard ducks with 200 mg/kg
diet of Ceresan M (3.1% ethylmercury) and measured eggshell thickness
on days 76 and 85 of treatment. No significant effects were found.
When mercury was added to the diet along with DDE at 40 mg/kg, mercury
did not increase the effect of the organochlorine on shell thickness.

Mullins et al. (1977) dosed captive hen pheasants with
phenylmercuric acetate (PMA) either in capsules (20 mg/kg body weight)
or added to the diet (at the normal fungicidal treatment rate of
14.18 g/bushel of seed wheat). Birds given mercury by capsule showed
significant decreases in egg hatchability, eggshell thickness, and
chick weight and survival, but no effect on egg production, egg
volume, fertility, or chick behaviour. The mercury-dosed diet had no
effect on any of these reproductive parameters.

Hill & Soares (1984) studied the sublethal effects of feeding
9-week-old Japanese quail with methylmercuric chloride in the diet,
and calculated EC₅₀s for the activity of asparate aminotransferase,
alpha-hydroxybutyrate dehydrogenase, lactate dehydrogenase, and
ornithine carbamoyltransferase, in blood plasma, of 4.8, 6.1, 1.2, and
3.5 mg/kg diet, respectively.

In studies by Scott (1977), Japanese quail were fed diets
containing methylmercuric chloride (0, 10, or 20 mg mercury/kg). The
daily food intake, egg production, average egg weight, percentage of

fertile eggs, and the hatchability of fertile eggs were all
significantly reduced at 10 mg/kg. There were greater effects on all
these parameters with 20 mg/kg, but the difference was not significant
relative to the lower dose in terms of percentage fertility or
hatchability of fertile eggs. The strength of the eggshell was
significantly reduced by the 10 mg/kg dose after 3 weeks of dosing.
Insufficient eggs were laid by the group dosed at the higher rate to
monitor this factor.

Tejning (1967) studied the effects on domestic fowl of
methylmercuric-dicyandiamide (MMD)-treated grain (0-18.4 mg mercury/kg
diet). Food consumption was unaffected in birds treated with 0 or
4.4 mg mercury/kg, but fell gradually over 50 days, in birds treated
with 9.2 or 18.4 mg/kg. Food consumption returned to normal after
about 60-65 days, but then fell below control levels again later. Egg
production (eggs/hen per day) was unaffected by 4.4 mg mercury/kg or
by 8.8 or 9.2 mg/kg for the first 40 days of exposure. After treatment
at 17.6 or 18.4 mg/kg diet, egg production gradually fell over the
period of exposure. There was no effect on body weight of any of the
treated birds. Some birds on the highest doses showed ataxia with
difficulty in walking. In a study comparing three treatment levels of
MMD (0, 9.2, and 18.4 mg mercury/kg diet, various reproductive
parameters were monitored. There was an increase, relative to
controls, in the number of soft-shelled eggs of 17.1% at the highest
dose and 1.4% in the 9.2 mg/kg group. Percentages of deaths of embryos
in shell during the first 5 days of incubation were also increased
(values were 10.5% in controls, 43.7% in the birds dosed with
9.2 mg/kg, and 62.1% in the 18.4 mg/kg group). Mortality later in the
incubation period was similar in all groups. Overall hatchability was
reduced from 60% in the controls to 16% in the 9.2 mg/kg group and 10%
in the 18.4 mg/kg group.

Fimreite (1970) exposed leghorn cockerels to a diet dosed with
Panogen 15 (2.5% MMD) at concentrations of 6, 12, and 18 mg MMD/kg for
3 weeks, from 2 weeks of age. The total intake of mercury, based on
monitoring food consumption, was calculated to be 1.7, 3.4, and
5.1 mg/chick, respectively, for the three dosing levels. All treated
birds showed significant reductions in weight, but only at the highest
dose was there a significant increase in deaths. Fimreite (1971) fed
penned pheasant (Phasianus colchicus) breeder ration and treated
grain containing MMD at 2.25, 4.5, or 9 mg mercury/kg, for 2, 4, or 12
weeks. There was no weight reduction amongst adults, and food
consumption was only adversely affected by the highest dose. Some hens
fed the highest dose showed extensive demyelination of the spinal
cord. All treated birds showed reduced hatchability and egg
production, with a large number of shell-less eggs. There was a
significant reduction in the weight of eggs laid by mercury-treated
birds. The highest dose group laid eggs of an abnormal colour.

Spann et al. (1986) fed 12-day-old bobwhite quail on diets
containing methylmercuric chloride at 0, 5.4, or 20 mg/kg (equivalent
to 0, 4.3, or 16 mg mercury/kg). Birds dosed at the lower rate showed
low mortality, not significantly different from controls, whereas
birds dosed at the higher rate showed high mortality after 6 weeks (at
between 55% and 80% for three different vehicles: no solvent; corn
oil; and propylene glycol). When acetone was used as carrier, deaths
were significantly reduced (to about 30%), deaths in the control group
being < 10%.

When Mykkanen & Ganther (1974) fed 1-day-old Japanese quail a
diet containing 0-30 mg mercury/kg (as methylmercury hydroxide) for up
to 32 days, no effect on erythrocyte glutathione reductase activity
was found.

Fimreite & Karstad (1971) dosed chicks with MMD and then fed them
to red-tailed hawks for up to 12 weeks. Mercury levels in the liver of
the chicks were between 3.9 and 10 mg/kg. Three of the six birds,
given chicks with mercury in the liver at 10 mg/kg and died, one bird
out of six, given chicks with mercury in the liver at 7.2 mg/kg, died.
All the poisoned birds showed neurological symptoms, weakness in
extremities, and impaired coordination of muscular movement, and,
although the hawks did not lose their appetite, they had difficulty
feeding. There was no effect on food consumption, even poisoned birds
maintaining appetites until in an advanced stage of poisoning. Only
birds with overt signs of poisoning showed substantial body weight
loss. Borg et al. (1970) fed chickens with 8 mg MMD/kg diet for 5 to 6
weeks, and muscle and liver from the contaminated chickens were fed to
goshawks (Accipiter gentilis gentilis). Three goshawks receiving
muscle and liver averaging 13 mg mercury/kg died within 30, 38, and 47
days. One goshawk receiving muscle only (10 mg mercury/kg) died within
39 days. The major clinical symptoms, appearing after about two weeks,
were inappetance, muscular weakness, ataxia, and loss of body weight.
Autopsy revealed that the dominating effect was muscular atrophy,
which was presumably the main cause of weight loss. Pronounced
histological changes included demyelination and nerve cell
degeneration of the cerebellum and medulla oblongata and demyelination
of peripheral nerves. No lesions were found in the cerebrum.

When Heinz (1974) fed mallard ducks a dry mash diet containing
MMD (0.5 or 3.0 mg mercury/kg) for 21 weeks, the lower of the two dose
levels had no effects on reproduction but the higher reduced egg
laying and increased embryonic and duckling mortality. Eggs laid by
controls tended to be heavier than eggs laid by treated birds, but
there was no effect on eggshell thickness. Heinz (1976a) fed mallard
for 2 consecutive years on the same doses of MMD as above. There was
no significant effect on egg production or hatching success or on
approach behaviour of ducklings. Ducklings from females fed 3.0 mg/kg
were less likely to survive to 1 week than those from other groups.
Ducklings from parents fed the highest dose were hyper-responsive in
avoidance behaviour. Heinz (1976b) fed ducklings (from 9 days of age)

whose parents had been fed MMD at 0.5 mg/kg diet, on the same dosed
diet. Dosed second generation females laid a greater proportion of
their eggs on open ground outside the nest boxes. They also produced
fewer ducklings surviving to 1 week. In ducklings from second
generation females, there were no significant differences in behaviour
patterns such as approach response to maternal calls, avoidance
response to frightening stimuli, and open-field behaviour. There was a
reduction in growth of third generation ducklings. Heinz (1979) dosed
three generations of mallard with 0.5 mg MMD/kg diet. As in the second
generation, females laid a greater number of eggs outside the nest
box. They also laid fewer eggs and produced fewer ducklings. There was
some eggshell thinning in the third generation and a reduced response
of ducklings to maternal calls.

Prince (1981) tested mallard ducks through four generations in an
attempt to establish if resistance to the reproductive effects of
methylmercury was developed. The parental generation was exposed to
two doses of 8 mg methylmercuric chloride within a 2-week period. The
parents were split into two groups on the basis of the survival of
ducklings after exposure to mercury. Three further generations of each
line were produced. The percentage survival of ducklings exposed, via
the parent, to mercury tended to increase in the "resistant" strain in
successive generations. This suggested an ability of the birds to
adapt to mercury exposure over time.

Ganther et al. (1972) fed Japanese quail with diets containing up
to 20 mg methylmercury/kg. Some groups of quail were given tuna fish
as 17% of the total diet, while other groups had corn-soya instead of
the tuna. Mortality in the group fed corn-soya with 20 mg
methylmercury/kg diet was 61% over 6 weeks, the majority (52%) of
deaths occurring between 4 and 6 weeks of dosing. The same amount of
methylmercury added to the tuna diet led to only 14% mortality over
the 6-week-period of dosing. The authors ascribe the protective effect
of the fish diet to the high selenium level in the tuna. Selenium,
which in these diets amounted to 0.3-0.6 mg/kg, becomes toxic to birds
only at dietary concentrations more than 10 times higher than this.

7.2.3 Effects of mercury on non-laboratory mammals

Appraisal

Few studies have been published on truly wild, non-laboratory
mammals. Work is of most value when done not on mammal species that
have been changed by generations in captivity but on those that are
still found in the wild, or are genetically close to wild forms. The
only work in this last category is on mink and prairie vole. The
available evidence indicates that toxic effects, including
reproductive changes, can be produced. Methylmercury has been found
to be more toxic than inorganic mercury.

Aulerich et al. (1974) dosed the diet of mink with either 5 mg
methylmercury/kg (as contained in Ceresan L, which contains 2.25%
mercury) or 10 mg mercuric chloride/kg. No adverse effects attributed
to these diets were observed for 3 weeks. After 25 days, the mink
dosed with organic mercury showed signs of lack of coordination, loss
of balance, anorexia, and loss of weight. Within 4 days, ataxia,
paralysis, tremors, and, finally, death were observed. Attempts to
arrest these symptoms in the least affected mink by reverting to a
control diet, with either EDTA or methionine injections, had no effect
and the mink still died. Mink dosed with inorganic mercury showed no
clinical signs. The mercuric chloride treatment did not affect
reproductive performance and no teratological effect was noted. There
was a significant reduction in the weight of the kits from treated
parents, at birth, but this had recovered by 4 weeks of age.

Wren et al. (1987a) fed adult mink a daily diet containing 1 mg
methylmercury/kg for 3 months. Later, because of mortality, the dosed
diet was administered every other day for a further 3 months. The
initial, daily-dosed diet resulted in the death of 8 out of 12 females
and 1 out of 4 males. There were no observed effects of the treatment
on the thyroid, pituitary, or adrenal glands or on serum
triiodothyronine (T3) or thyroxine (T4) levels during the experimental
period. Mortality was thought to be caused by a combination of mercury
poisoning and cold stress (the animals were kept outside during the
winter). Under laboratory conditions, 1 mg/kg would not be considered
fatal to mink (Wobeser et al., 1976). Under the same experimental
conditions, Wren et al. (1987b) found that the fertility of adult male
mink, percentage of females whelped, and number of kits born per
female were not affected by the mercury treatment.

Hartke et al. (1976) calculated an acute LD₅₀ of 10 mg/kg body
weight for phenylmercuric acetate (PMA) in female prairie voles
(Microtus ochrogaster), after intraperitoneal injection. Female
voles were also injected on days 8, 9, and 10 of gestation with
0.06-5.0 mg PMA/kg body weight. Some normal foetuses and some
resorption sites (where implantation had occurred but the foetal
material had been reabsorbed) were found in voles injected with
0.5 mg/kg or less on days 8 and 9 of gestation. Animals treated with
> 1.0 mg/kg had no live foetuses, but all had resorption sites in
the uterus. Similar results were found for voles treated on day 10.
No resorption sites were found in voles treated with < 0.25 mg/kg.
To study the effects of dose and the stage of gestation when dosing
occurred, the authors further injected voles with 0.5 mg/kg on days
7, 11, and 12 of gestation. Normal embryos and some resorption and
abortion sites were found after dosing on days 7 and 11. Dosing on
day 12 of gestation produced no resorption or abortion and the
numbers of live foetuses accounted for all corpora lutea in the
ovary.

8. EFFECTS OF MERCURY IN THE FIELD

Appraisal

Fatalities and severe poisonings in birds have been reported in
association with outbreaks of human poisoning. In addition, the
agricultural use of organomercury fungicides has caused poisoning in
birds. A statistical association has been reported between the
mercury content of birds' eggs and reproductive failure. These eggs
also contained organochlorine residues, but these residues did not
correlate with the observed reproductive effects.

Methylmercury levels in fish in Japan have caused a major problem
for human health. During these incidents, there were also reports of
direct effects of mercury on wildlife in the area. Fish carrying
methylmercury were found dead or showed symptoms of mercury poisoning.
Fish-eating birds and scavenging birds were also killed (Harada,
1978). Birds found dead in the area showed the characteristic
pathological changes in the central nervous system of Minamata
disease, but no measurement of mercury content was made (Takeuchi et
al., 1957).

The use of organic mercury compounds as a fungicidal seed
dressing has led to deaths in the field of birds, mostly grain-eating
species. Some raptors, feeding on the poisoned birds, were also
casualties (Borg et al., 1969). Koeman et al. (1969) reported large
numbers of birds of prey killed by indirect poisoning with
organomercury fungicides in the Netherlands.

Mercury contamination has been implicated in the breeding failure
of some raptor species both in Europe and North America, where
residues have equalled those found to cause reproductive impairment in
laboratory species. These birds also contained organochlorine
insecticide residues and the separation of effects is difficult
(Newton, 1979). More recent work suggests more strongly that mercury
affects the breeding of birds of prey in the field. Merlins sampled in
Scotland contained organochlorines along with mercury in their eggs.
Statistical analysis of the data showed a clear inverse relationship
between mercury content of eggs and brood size; the higher the mercury
content, the less likelihood of successful breeding. Productivity fell
markedly when mercury residues in eggs exceeded 3 mg/kg. Productivity,
that is the number of young successfully reared, showed no
statistically significant relationship with residues of other
chemicals present in the eggs. Levels of mercury were highest in birds
sampled in Orkney and Shetland, but the relationship between mercury
residue and productivity remained when these, particularly high,
residue levels were excluded from the analysis (Newton & Haas, 1988).
The merlins were feeding on wading birds in estuaries and this was
presumed to be the source of the mercury. A similar, but not quite
significant, relationship was found in peregrine falcons breeding near
the coast.

Jefferies et al. (1973) sampled small mammals from fields sown
with mercury-treated grain. They express the view that residues were
sufficiently high to have caused deaths in small mammals feeding on
the grain. Some mammals were found dead and deemed to have been killed
by mercury poisoning.

9. EVALUATION

In evaluating the environmental hazard of mercury it is necessary
to extrapolate from laboratory experiments to ecosystems. This must be
done with extreme caution for the following reasons.

(i) Speciation of mercury and its adsorption to environmental
components such as soil, sediment, organic matter, and biota
limit its availability to organisms in the environment.

(ii) Environmental variables such as temperature, pH and chemical
composition of water, soil type, and geology have been shown
in limited studies on a narrow range of species to affect
both uptake and effect of mercury. There is insufficient
information to fully assess the probable affects of, for
example, tropical conditions and acid precipitation.

(iii) There are few data measuring mercury availability to
organisms. Most data represent nominal or total metal
concentration, rather than that component which could be
taken up by organisms. True exposure is, therefore,
difficult to assess.

(iv) There are limited data on the behaviour of mixtures of
metals from controlled experimental work; organisms in the
environment are exposed to mixtures.

(v) Experimental work seldom, if ever, is conducted on species
or communities that are either representative or key
components of natural communities and ecosystems. Studies do
not consider all of the interactions between populations and
all of the environmental factors affecting these
populations.

It is probable that subtle disturbances to the community occur at
much lower concentrations than those suggested in laboratory studies
on acute effect, perhaps as much as one order of magnitude lower.

9.1 The Marine Environment

Marine aquatic organisms at all levels accumulate mercury into
tissues. This mercury is retained for long periods if it is in an
organic form. A number of factors affect the susceptibility of aquatic
organisms to mercury. These include the life-cycle stage (the larval
stage being particularly sensitive), the development of tolerance,
water temperature, and salinity. Some incidents of severe pollution
have resulted in the death of fish at that time. Few follow-up studies
have been reported so that it is impossible to assess the long-term
hazards. Toxic effects have been produced experimentally only at
concentrations much higher than those found in the non-polluted marine
environment. Furthermore, most of the studies have been on acute

lethality and have used inorganic mercury compounds in the main.
Birds, particularly coastal species or those eating prey that feed in
estuaries, have been affected by mercury contamination. It has
adversely affected breeding and may have influenced population
stability.

9.2 The Freshwater Environment

Mercury compounds are acutely toxic to freshwater microorganisms.
Using photosynthesis and/or growth as parameters, the NOTEL
(No-observed-toxic-effect-level) for inorganic mercury lies between 1
and 50 µg/litre, depending on the organism, density of cells in
culture, and experimental conditions. Diversity of species in mixed
culture may be affected by 40 µg mercuric chloride/litre. For
organomercury compounds, the NOTEL is 10-100 times lower.

Aquatic plants sustain damage after exposure to inorganic mercury
at concentrations of 800 to 1200 µg/litre. Organomercury produces
toxic effects at concentrations 10-100 times lower.

Many aquatic invertebrates are sensitive to mercury toxicity,
particularly as larvae. Organic mercury compounds are toxic at
concentrations 10 to 100 times less than inorganic mercury. For the
most sensitive species, Daphnia magna, the NOTEL for reproductive
impairment is 3 µg/litre for inorganic mercury and < 0.04 µg/litre
for methylmercury.

Freshwater fish show lethal responses to mercury in acute nominal
concentrations from approximately 30 µg/litre. Larvae under the same
static conditions are 10 times more sensitive. In flow-through tests,
fish are up to 100 times more sensitive. In both static and flow-
through tests, organomercury compounds are approximately 10 times more
toxic than inorganic compounds. The NOTEL for the most sensitive
parameters may be well below 0.01 µg/litre.

Aquatic developmental stages of amphibia show sensitivity to
mercuric compounds similar to that of fish.

9.3 The Terrestrial Environment

Based on the current state of knowledge, it is not possible to
determine the true exposure or concentration of mercury available to
terrestrial organisms. It can, however, be stated that exposure via
soil, soil water, and food is most important; exposure via open water
and air is less important.

Mercury has been shown, in laboratory studies, to be toxic to
terrestrial organisms over a broad range of concentrations. However,
most of these studies are at high exposure levels (birds) or
environmentally unrealistic exposure routes (hydroponic culture of
plants).

It can be stated that acute effects would not be seen in
terrestrial plants growing in natural soils, nor in terrestrial birds
or mammals, other than by exposure to mercurials used as fungicidal
seed-dressings. Other effects seen in birds derive from mercury in the
marine environment.

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    See Also:
       Toxicological Abbreviations