INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY
ENVIRONMENTAL HEALTH CRITERIA 1
This report contains the collective views of an international
group of experts and does not necessarily represent the
decisions or the stated policy of either the World Health
Organization or the United Nations Environment Programme
Published under the joint sponsorship of
the United Nations Environment Programme
and the World Health Organization
World Health Organization Geneva, 1976
ISBN 92 4 154061 3
(c) World Health Organization 1976
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BACKGROUND AND PURPOSE OF THE WHO ENVIRONMENTAL HEALTH
ENVIRONMENTAL HEALTH CRITERIA FOR MERCURY
1. SUMMARY AND RECOMMENDATIONS FOR FURTHER RESEARCH
1.1. Some definitions
1.2.1. Analytical methods
1.2.2. Sources of environmental pollution
1.2.3. Environmental distribution and transport
1.2.4. Environmental exposure levels
1.2.5. Metabolism of mercury
1.2.6. Experimental studies on the effects of mercury
1.2.7. Epidemiological and clinical studies
1.2.8. Evaluation of health risks to man and guidelines
for health protection
1.3. Recommendations for further research
1.3.1. Environmental sources and pathways of mercury
1.3.2. Metabolic models in man
1.3.3. Epidemiological studies
1.3.4. Interaction of mercury with other environmental
1.3.5. Biochemical and physiological mechanisms of
2. PROPERTIES AND ANALYTICAL METHODS
2.1. Chemical and physical properties
2.2. Purity of compounds
2.3. Sampling and analysis
2.3.1. Sample collection
2.3.2. Analytical methods
2.3.3. Analysis of alkyl mercury compounds in the presence
of inorganic mercury
3. SOURCES OF ENVIRONMENTAL POLLUTION
3.1. Natural occurrence
3.2. Industrial production
3.3. Uses of mercury
3.4. Contamination by fossil fuels, waste disposal, and
4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
4.1. Distribution between media -- the global mercury cycle
4.2. Environmental transformation -- the local mercury cycle
4.3. Interaction with physical or chemical factors
5. ENVIRONMENTAL LEVELS AND EXPOSURES
5.1. Levels in air, water, and food
5.2. Occupational exposures
5.3. Estimate of effective human exposure
6. METABOLISM OF MERCURY
6.1.1. Uptake by inhalation
6.1.2. Uptake by ingestion
6.1.3. Absorption through skin
6.2. Distribution in organisms
6.3. Elimination in urine and faeces
6.4. Transplacental transfer and secretion in milk
6.5. Metabolic transformation and rate of elimination
6.6. Accumulation of mercury and biological half-time (metabolic
6.7. Individual variations -- strain and species comparisons
7. EXPERIMENTAL STUDIES ON THE EFFECTS OF MERCURY
7.1. Experimental animal studies
7.1.1. Acute studies
7.1.2. Subacute and chronic studies
22.214.171.124 Reversible damage
126.96.36.199 Irreversible damage
188.8.131.52 Interactions with physical and chemical
7.1.3. Biochemical and physiological mechanisms of
8. EFFECTS OF MERCURY ON MAN -- EPIDEMIOLOGICAL AND CLINICAL STUDIES
8.1. Epidemiological studies
8.1.1. Occupational exposure to mercury vapour,
alkylmercury vapour and other exposures
8.1.2. General population
8.1.3. Children and infants with in utero exposure
8.2. Clinical studies of effects of mercury binding compounds
8.3. Pathological findings and progression of disease
8.3.1. Psychiatric and neurological disturbances
8.3.2. Eye and visual effects
8.3.3. Kidney damage
8.3.4. Skin and mucous membrane changes
9. EVALUATION OF HEALTH RISKS TO MAN FROM EXPOSURE TO MERCURY
AND ITS COMPOUNDS
9.1. General considerations
9.1.1. Elemental mercury vapour
9.1.2. Methylmercury compounds
9.1.3. Ethylmercury compounds and other short-chain
9.1.4. Inorganic mercury, aryl- and alkoxyalkylmercurials
9.2. Summary and guidelines
BACKGROUND AND PURPOSE OF THE WHO ENVIRONMENTAL HEALTH
ORIGIN AND OBJECTIVES OF THE PROGRAMME
During the last two decades, evaluation of the health hazards from
chemical and other environmental agents has received considerable
attention in several WHO programmes. High priority was given to
drinking water quality (1), food additives (2), and pesticide residues
(3), to occupational exposure (4), air quality in urban areas (5),
and, more recently, to the carcinogenic risk of chemicals to man (6).
In most instances, man's total exposure to a given agent, from
different media or conditions (air, water, food, work, home), was not
considered. The inadequacy of this approach is obvious for pollutants
that may reach man by several pathways, as is the case with lead,
cadmium, and some other metals, and certain persistent organic
compounds. In response to a number of World Health Assembly
resolutions (WHA23.60, WHA24.47, WHA25.58, WHA26.68) and taking into
consideration the relevant recommendations of the United Nations
Conference on the Human Environment (7) held at Stockholm in 1972, and
of the Governing Council of the United Nations Environment Programme
(UNEP) (8), an integrated and expanded programme on the assessment of
health effects of environmental conditions was initiated in 1973 under
the title of: WHO Environmental Health Criteria Programme, with the
(i) to assess existing information on the relationship between
exposure to environmental pollutants (or other physical and
chemical factors) and man's health, and to provide guidelines
for setting exposure limits consistent with health protection,
i.e., to compile environmental health criteria documents;
(ii) to identify new or potential pollutants by preparing
preliminary reviews on the health effects of agents likely to
be increasingly used in industry, agriculture, in the home or
(iii) to identify gaps in knowledge concerning the health effects of
recognized or potential pollutants or other environmental
factors, to stimulate and promote research in areas where
information is inadequate, and
(iv) to promote the harmonization of toxicological and
epidemiological methods in order to obtain research results
that are internationally comparable.
a Prepared by the WHO Secretariat. References are listed on page 14.
The general framework of the Environmental Health Criteria
Programme was formulated by a WHO meeting held in November 1972 (9),
and further elaborated by a WHO Scientific Group that met in April
DEFINITIONS, TERMINOLOGY, AND UNITS
In the framework of the WHO Environmental Health Criteria
Programme, it is understood that the term "criteria" designates the
relationship between exposure to a pollutant or other factor and the
risk or magnitude of undesirable effects under specified circumstances
defined by environmental and target variables (9). This corresponds to
the definition proposed by the Preparatory Committee for the United
Nations Conference on the Human Environment (11). Other Preparatory
Committee definitions of immediate interest to the criteria programme
-- " exposure: the amount of a particular physical or chemical agent
that reaches the target";
-- " target (or receptor): the organism, population, or resource to
be protected from specific risks";
-- " risk: the expected frequency of undesirable effects arising
from a given exposure to a pollutant".
The WHO Scientific Group on Environmental Health Criteria (10)
accepted these definitions for the purposes of its discussions, but
felt that they were not altogether satisfactory, and recommended that
WHO, in collaboration with other international organizations, should
reconsider them, along with other necessary definitions, at an
appropriate international meeting. In accordance with this
recommendation, the WHO Secretariat is preparing a list of basic terms
to be used in the Environmental Health Criteria Programme that will be
submitted to the national institutions and other international
organizations for discussion.
The Scientific Group (10) found the definition of "exposure"
particularly inadequate and considered that it should be expanded to
include the concepts of concentration and length of exposure in
addition to the amount of the agent.
The WHO Secretariat considers it useful to attach specific
meanings to the terms "effect", "response" and "dose" as was done by
the Subcommittee on the Toxicology of Metals of the Permanent
Commission and International Association on Occupational Health at the
Tokyo meeting (12). These terms will be used in the following sense
unless indicated differently in specific criteria documents:
-- " effect: a biological change caused by (or associated with)a an
-- " response: the proportion of a population that demonstrates a
-- " dose: the amount or concentration of a given chemical at the
site of the effect".
The concept of "response" as defined above is generally accepted
but the terminology used to describe this concept varies widely. Many
toxicologists use the terms "effect" and "response" interchangeably to
denote a specific biological change associated with exposure, whereas
different terms are used to indicate the proportion of a population
affected (e.g., incidence, cumulative response frequency, response
There is no general agreement as to the use of the term "dose" for
chemical agents. Its common usage is to express the amount of
substance administered, for instance, to an experimental animal (e.g.,
oral dose, injected dose, etc.). In most cases, the amount or
concentration of a given agent at the site where its presence induces
a given effect cannot be determined by direct measurement and has to
be estimated from experimental, occupational, or general environmental
exposure, or from measurements in biological indicator media such as
blood, urine, faeces, sweat, or hair (12). To avoid misunderstanding,
it is, therefore, necessary in each case to make as clear as possible
the way in which the "dose" is measured or estimated, including the
Because of the existing differences in the use of terms, no
attempt has been made at this stage to impose a uniform terminology in
all criteria documents. Until an internationally agreed terminology
becomes available, the task groups on specific criteria documents are
given freedom to choose their terminology, provided the terms are
defined and used consistently throughout the document under
a Added by the WHO Secretariat.
An attempt has been made to express all numerical values in a
uniform fashion, for instance, the concentrations are always expressed
as mass concentrations in units acceptable to the SI system (e.g.
mg/litre or mg/kg) (13). Some departures from this are made where the
introduction of new units would cause confusion, e.g., lead in blood
is expressed in µg/100 ml and not in µg/litre.
Considering the large number of environmental agents and factors
that may adversely influence human health, a practical programme for
the preparation of criteria documents must be based on clearly defined
priorities. The list of priorities has been established by a WHO
Scientific Group (10), and is based on the following considerations:
-- " Severity and frequency of observed or suspected adverse effects
on human health. Of importance are irreversible or chronic
effects, such as genetic, neurotoxic, carcinogenic, and
embryotoxic effects including teratogenicity. Continuous or
repeated exposures generally merit a higher priority than isolated
or accidental exposures.
-- Ubiquity and abundance of the agent in man's environment. Of
special concern are inadvertently produced agents, the levels of
which may be expected to increase rapidly, and agents that add to
a natural hazard.
-- Persistence in the environment. Pollutants that resist
environmental degradation and accumulate, in man, in the
environment, or in food chains, deserve attention.
-- Environmental transformations or metabolic alterations. Since
these alterations may lead to the production of chemicals that
have greater toxic potential, it may be more important to
ascertain the distribution of the derivatives than that of the
-- Population exposed. Attention should be paid to exposures
involving a large portion of the general population, or
occupational groups, and to selective exposures of highly
vulnerable groups represented by pregnant women, the newborn,
children, the infirm or the aged."
The full list contains some 70 chemicals and physical hazards, and
it will be periodically reviewed. In preparing this list, it was
realized that each country must assess environmental health problems
in the light of its own national situation and establish its own
priorities, which may not have been covered by this list.
SCOPE AND CONTENT OF ENVIRONMENTAL HEALTH CRITERIA DOCUMENTS
As stated on page 5, the purpose of the criteria documents is to
compile, review, and evaluate available information on the biological
effects of pollutants and other environmental factors that may
influence man's health, and to provide a scientific basis for
decisions aimed at protecting man from the adverse consequences of
exposure to such environmental factors, both in the occupational and
general environment. Although attainment of this objective entails
consideration of a wide range of data, no attempt is made to include
in the documents an exhaustive review of all published information on
the environmental and health aspects of specific agents. In the
process of collecting the required information, the available
literature has been carefully evaluated and selected as to its
validity and its relevance to the assessment of human exposure, to the
understanding of the mechanism of biological effects, and to the
establishment of dose-effect and dose-response relationships.
Environmental considerations are limited to information that can help
in understanding the pathways leading from the natural and man-made
sources of pollutants to man. Non-human targets (e.g., plants,
animals) are not considered unless the effects of their contamination
are judged to be of direct relevance to human health. For similar
reasons much of the published information on the effects of chemicals
on experimental animals has been omitted.
The criteria documents consist of three parts:
(i) A summary, which highlights the major issues, followed by
recommendations for research to fill existing gaps in
(ii) The bulk of the report, which contains the findings on which
the evaluation of the health risks is based. This part has a
similar structure in all the criteria documents on chemical
agents and contains the following chapters: chemical and
physical properties and analytical methods; sources of
environmental pollution; environmental transport, distribution
and transformation; metabolism; experimental studies of
effects; and epidemiological and clinical studies of the
effects. The subdivision of these chapters differs from
document to document.
(iii) Evaluation of health risks to man from exposure to the
specific agent. This part of the criteria document states the
considered opinion of the task group, which examined the
findings contained in the second part (see (ii) above), and
typically contains the following sections: relative
contributions to the total dose from air, food, water, and
other exposures; dose-effect relationships; dose-response
relationships and, whenever possible, guidelines on exposure
or dose limits.
Chemical and physical data
The chemical and physical data included in the criteria documents
are limited to the properties that are considered relevant to the
assessment of exposure and to the understanding of the effects. Where
applicable, the impurities that may occur in commercial products are
examined. Analytical techniques are discussed only to the extent
needed to understand and evaluate data on levels in the environment
and biological samples. The methods described should not be considered
as recommended procedures. Where feasible, information is included on
the applicability of a given method for the analysis of different
types of sample, on detection limits, precision, and accuracy. The
detection limit represents the smallest total amount the method is
able to determine. In most cases, the amount of sample is limited so
that it is useful in practice to express the smallest concentration
that can be determined by that method. Precision of a method is
defined in terms of the standard deviation or the coefficient of
variation of a number of analyses made on the sample. Accuracy denotes
systematic deviation of the measured values from the true value. It is
impossible to ascertain the accuracy with absolute certainty; the
evidence for the accuracy of a method is often circumstantial and is
based either on inter-laboratory data-quality control studies or on
the agreement of results obtained with procedures using different
approaches. The results of one "accurate" procedure should agree with
those of another "accurate" procedure for a given set of samples.
Production, use, and environmental levels
Data on the production, use, and levels in the environment of
pollutants are reported only to illustrate the magnitude and extent of
the problem and are not meant to represent an exhaustive and critical
review. It is hoped that, in the future, better data will be available
and that closer collaboration will be established with other
governmental and non-governmental organizations qualified to supply
Although every effort is made to review the whole literature, it
is possible that some publications have been overlooked. Some studies
have purposely been omitted because the information contained therein
was not considered valid or relevant to the scope of the criteria
documents, or because they only confirmed findings already described.
In general, the information is summarized as given by the author;
however, certain shortcomings of reporting or of experimental design
are also pointed out. The data on carcinogenicity have been examined
and evaluated in consultation with the International Agency for
Research on Cancer.
Whenever possible, the dose-effect and the dose-response
relationships reported in the criteria documents are based on
epidemiological and other human studies, and animal data are used, in
general, as supporting evidence.
ARRANGEMENTS FOR THE PREPARATION OF CRITERIA DOCUMENTS
In order to obtain balanced and unbiased information, the
collection and evaluation of information is done in close
collaboration with national scientific and health institutions. About
20 Member States of WHO have designated national focal points for
collaboration in the WHO Environmental Health Criteria Programme.
Without this collaboration no progress could have been made in its
In addition, a number of WHO collaborating centres on
environmental health effects have been designated to extend and
complement the expertise available in the WHO Secretariat.
Two procedures have been used in preparing the criteria documents.
One is based on the consolidation of national contributions and the
other on a draft criteria document prepared by consultants or the
collaborating centres in association with the Secretariat.
Procedure based on national contributions
Criteria documents are prepared in four stages: (1) the
preparation of national contributions by focal points in the Member
States reviewing all relevant research results obtained in these
countries; (2) consolidation of the national contributions into a
draft document, which is done on a contractual basis with individual
experts or WHO collaborating centres; (3) the draft criteria documents
are circulated to the national focal points for comments and
additions, based on which a second draft is prepared, and (4) the
second draft document is reviewed and the information assessed at a
meeting of internationally recognized experts (the task group
National contributions to the criteria documents consist of a
review of data on health effects of environmental agents, as revealed
by experimental, clinical, and epidemiological studies, and of other
relevant information on research carried out in each country and
published in scientific journals or official publications. In order to
facilitate the integration of national contributions into draft
criteria documents, detailed outlines are prepared for each
environmental agent considered, and the national focal points are
requested to follow these outlines as closely as possible and to
attach all publications referred to in the review in the form of
reprints or microfiches.
Procedure for drafts prepared by the Secretariat
With the exception of steps 1 and 2 (which are replaced by the
preparation of a draft criteria document by individual experts or WHO
collaborating centres), the procedure is the same as described above.
This procedure is applied in cases where much preparatory work has
been done in Member States and where criteria-like documents (WHO or
national) already exist.
Task group meetings
The task group meetings that are convened to complete the criteria
documents have the following terms of reference:
(i) to verify, as far as possible, that all available data have
been collected and examined;
(ii) to select those data relevant to the criteria documents;
(iii) to determine whether the data, as summarized in the draft
criteria document, will enable the reader to make his own
judgement concerning the adequacy of an experimental,
epidemiological, or clinical study;
(iv) to judge the health significance of the information contained
in the draft criteria document, and
(v) to make an evaluation of the dose-effect, dose-response
relationships and of the health risks from exposure to the
environmental agents under examination.
Members of task groups serve in a personal capacity, as experts
and not as representatives of their governments or of any organization
with which they are affiliated. In addition to the first and second
draft criteria documents, the members of the task group are requested
to refer to the original publications whenever they deem that
necessary, and to review national and other comments on the first
draft criteria document to make sure that no significant information
is omitted and that the final document properly reflects the work done
in different countries.
Collaboration with the United Nations Environment Programme (UNEP) and
other international organizations
The WHO Environmental Health Criteria Programme has received
substantial financial assistance from UNEP which is acknowledged with
appreciation. In addition, the programme has been planned from the
outset in consultation with the UNEP Secretariat. The UNEP Secretariat
receives all the drafts of criteria documents and their comments are
carefully considered in the preparation of the final documents. UNEP
is regularly invited to be represented at the task group meetings.
The United Nations, their subsidiary bodies and specialized
agencies, and the IAEA are as a rule invited to provide comments on
the draft criteria documents and to participate in the task group
meetings. The same applies to selected nongovernmental organizations
in official relationship with WHO.
Note to readers of the criteria documents
While every effort has been made to present information in the
criteria documents as accurately as possible without unduly delaying
their publication, mistakes might have occurred and are likely to
occur in the future. In the interest of all users of the environmental
health criteria documents, readers are kindly requested to communicate
any errors found to the Division of Environmental Health, World Health
Organization, Geneva, Switzerland, in order that they may be included
in corrigenda which will appear in subsequent volumes.
In addition, experts in any particular field dealt with in the
criteria documents are kindly requested to make available to the WHO
Secretariat any important published information that may have
inadvertently been omitted and which may change the evaluation of
health risks from exposure to the environmental agent under
examination, so that the information may be considered in the event of
updating and re-evaluation of the conclusions contained in the
1. International Standards for Drinking Water, third edition,
Geneva, World Health Organization, 1971.
2. WHO Technical Report Series, Nos: 129 (1957), 228 (1962), 281
(1964), 309 (1965), 339 (1966), 373 (1967), 383 (1968), 430
(1969), 445 (1970), 462 (1971), 488 (1972), 505 (1972), 539
3. WHO Technical Report Series, Nos: 370 (1967), 391 (1968), 417
(1969), 458 (1970), 474 (1971), 502 (1972), 525 (1973), 545
(1974), 574 (1975), 592 (1976).
4. WHO Technical Report Series, No.: 415 (1969).
5. WHO Technical Report Series, No.: 506 (1972).
6. INTERNATIONAL AGENCY FOR RESEARCH ON CANCER. IARC Monographs on
the Evaluation of Carcinogenic Risk of Chemicals to Man,
Vol. 1-11 (1972-76).
7. UNITED NATIONS GENERAL ASSEMBLY. Report of the United Nations
Conference on the Human Environment held at Stockholm, 5-16
June 1972 A/CONF.48/14, 3 July 1972.
8. UNITED NATIONS ENVIRONMENT PROGRAMME. Report of the Governing
Council of the United Nations Environment Programme (First
session) UNEP/GC/10, 3 July 1973.
9. The WHO Environmental Health Criteria Programme (unpublished
WHO document EP/73.1).
10. Environmental Health Criteria. Report of a WHO Scientific Group
(unpublished WHO document EP/73.2).
11. UNITED NATIONS GENERAL ASSEMBLY. Report of the Preparatory
Committee for the United Nations Conference on the Human
Environment on its Third Session. United Nations document
A/CONF.48/PC/13, 30 September 1971.
12. NORDBERG, G. F., ed. Effects and dose-response relationships of
toxic metals, Proceedings from an international meeting
organized by the Sub-committee on the Toxicology of Metals
of the Permanent Commission and International Associations
on Occupational Health, Tokyo, 18-23 November 1974.
Amsterdam, Oxford, New York, Elsevier Scientific Publishing
13. LOWE, D. A. A guide to international recommendations on names and
symbols for quantities and on units of measurement. Geneva,
World Health Organization, 1975, 314pp. (Progress in
Standardization No. 2.)
WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR MERCURY
Geneva 4-10 February 1975
Professor T. Beritic, Institute for Medical Research and
Occupational Medicine, Zagreb, Yugoslavia
Dr H. Blumenthal, Division of Toxicology, Bureau of Foods, Food
and Drug Administration, Department of Health, Education and
Welfare, Washington, DC, USA (Rapporteur)
Dr J. Bouquiaux, Department of the Environment, Institute of
Hygiene and Epidemiology, Brussels, Belgium
Dr G. J. van Esch, Laboratory for Toxicology, National Institute
of Public Health, Bilthoven, Netherlands
Professor L. Friberg, Department of Environmental Hygiene,
Karolinska Institute, Stockholm, Sweden (Chairman)
Professor G. L. Gatti, Istituto Superio di Sanità, Rome, Italy
Dr L. Magos, Toxicology Research Unit, Medical Research Council
Laboratories, Carshalton, Surrey, England
Dr J. Parizek, Institute of Physiology, Czechoslovak Academy of
Sciences, Prague, Czechoslovakia
Dr J. K. Piotrowski, Department of Biochemistry, Institute of
Environmental Research, Medical Academy in Lodz, Lodz, Poland
Dr E. Samuel, Health Protection Branch, Department of National
Health and Welfare, Ottawa, Ontario, Canada
Dr S. Skerfving, Department of Internal Medicine, University
Hospital, Lund, Sweden
Dr T. Tsubaki, Brain Research Insitiute, Niigata University,
Professor H. Valentin, Institute for Occupational and Social
Medicine, Erlangen, Federal Republic of Germany
Representatives from other organizations
Dr A. Berlin, Health Protection Directorate, Commission of the
European Communities, Luxembourg
Dr D. Djordjevic, Occupational Health and Safety Branch, ILO,
Dr W. J. Hunter, Commission of the European Communities,
G. D. Kapsiotis, Senior Officer, Food Policy and Nutrition
Division, FAO, Rome, Italy
Dr E. Mastromatteo, Chief, Occupational Health and Safety Branch,
ILO, Geneva, Switzerland
Dr T. Clarkson, University Center in Environmental Health
Sciences, The University of Rochester, School of Medicine and
Dentistry, Rochester NY, USA (Temporary Adviser)
Dr F. C. Lu, Chief, Food Additives, WHO, Geneva, Switzerland
Dr B. Marschall, Medical Officer, Occupational Health, WHO,
ENVIRONMENTAL HEALTH CRITERIA FOR MERCURY
A WHO Task Group on Environmental Health Criteria for Mercury met
in Geneva from 4-10 February 1975. Dr B. H. Dietrich, Director,
Division of Environmental Health, opened the meeting on behalf of the
Director-General. The Task Group reviewed and amended the second draft
criteria document and made an evaluation of health risks from exposure
to mercury and its compounds. The revised draft was sent for comments
to all members of the Task Group.
A group of WHO temporary advisers (Dr T. Clarkson, Dr L. Friberg,
Dr A. Jernelöv,a Dr L. Magos, and Dr G. Nordbergb) assisted the
Secretariat in the final scientific editing of the document. They met
in Geneva on 13 and 14 November 1975.
The first and second draft criteria documents were prepared by
Dr T. Clarkson, Environmental Health Sciences Centre, the University
of Rochester School of Medicine and Dentistry, Rochester, New York,
USA. The comments on which the second draft was based were received
from the national focal points for the WHO Environmental Health
Criteria Programme in Bulgaria, Czechoslovakia, the Federal Republic
of Germany, Italy, Japan, the Netherlands, New Zealand, Poland,
Sweden, the USA, and the USSR; and from the United Nations
Industrial Development Organization (UNIDO), Vienna, and the United
Nations Scientific, Educational and Cultural Organization (UNESCO),
Paris. Comments from the International Labour Organisation, Geneva,
the United Nations Food and Agriculture Organization, Rome, and the
Commission of the European Communities Health Protection Directorate,
Luxembourg, were submitted at the task group meeting.
Comments were also received, at the request of the Secretariat,
from Dr L. Amin-Zaki, Iraq, Dr G. J. van Esch, Netherlands, Dr K.
Kojima, Japan, and Dr S. I. Shibko, USA.
The collaboration of these national institutions, international
organizations, WHO collaborating centres and individual experts is
gratefully acknowledged. Without their assistance the document could
not have been completed. The Secretariat wishes to thank in particular
Dr T. Clarkson for his help in all phases of the preparation of the
a Institute for Water and Air Pollution Research, Stockholm, Sweden.
b Department of Environmental Hygiene, Karolinska Institute,
This document is based primarily on original publications listed
in the reference section. However, several recent publications broadly
reviewing health aspects of mercury and its compounds have also been
used. These include reviews by the Swedish Expert Group (1971).,
Hartung & Dinman (1972), IAEA (1972), and Wallace et al. (1971).
Reviews devoted primarily to the biological effects of mercury have
been published by Clarkson (1972a, 1972b) and Miller & Clarkson
(1973). Furthermore, several recent symposia have provided extensive
reviews of the environmental aspects of mercury (Bouquiaux, 1974;
D'Itri, 1972; Krenkel, 1975). A systematic review of various
environmental health aspects of mercury, including a broad review of
the accessible literature up to 1971, has been presented by Friberg &
1. SUMMARY AND RECOMMENDATIONS FOR FURTHER RESEARCH
1.1 Some definitions
In order to clarify the meaning of certain terms used in the
document, some definitions are given below. However, it should be
noted that these definitions have not been formally adopted by WHO.
The terms critical effects, critical organ, and critical organ
concentration have recently been defined by the Sub-Committee on
Toxicology of Metals of the Permanent Commission and International
Association of Occupational Health (Nordberg, 1976). The term
"critical" as defined by the Committee differs from its usual meaning
in clinical medicine, where it refers to a situation in which the
patient's condition may deteriorate suddenly and dramatically. It also
differs in meaning from that used in the field of radiation
protection, where the "critical" organ is defined as the organ of the
body whose damage by radiation results in the greatest injury to the
individual. In this document, the term "critical" does not refer to a
life-threatening situation, but to a key decision point for taking
preventive action. For example, at some point in the dose-effect
relationship, a critical effect can be identified. The appearance of
an effect in an individual signals the point at which measures should
be taken to reduce or prevent further exposure.
1.2.1 Analytical methods
The method of choice for determining total mercury in
environmental and biological samples is flameless atomic absorption.
The technique is rapid and sensitive and the procedure is technically
simple. Neutron activation is now principally used as a reference
method against which the accuracy of atomic absorption procedures may
be checked. Gas-liquid chromatography combined with an
electron-capture detector is the most widely used method for
identifying methylmercury in the presence of other compounds of
The methods of sampling require careful consideration of the type
of exposure to be monitored and the material to be analysed. Errors
arising in collection, storage, and transportation of samples may be
as important as instrument errors in contributing to the total error
in the measurement of mercury in the sample. These include
contamination of the sample, and the loss of mercury by adsorption on
the walls of the container, and by volatilization. In estimating human
exposure, special care should be taken to see that the sample is truly
representative, e.g. the mercury vapour concentration in the breathing
zone and the concentration of methylmercury in the daily diet.
1.2.2 Sources of environmental pollution
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 by 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
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. It was
extimated that the total anthropogenic release of mercury would amount
to 20 000 tonnes per year in 1975.
1.2.3 Environmental distribution and transport
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 seventy million tonnes, that the yearly
increases in concentration due to deposition from the global cycle are
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
The methylation of inorganic mercury in the sediment of lakes,
rivers, and other waterways and in the oceans is a key step in the
transport of mercury in aquatic food chains leading eventually to
human consumption. Methylmercury accumulates in aquatic organisms
according to the trophic level, the highest concentrations being found
in the large carnivorous fish.
Alkylmercury fungicides used as seed dressings are important
original sources of mercury in terrestrial food chains. Mercury is
passed first to seed eating rodents and birds and subsequently to
Accumulation of methylmercury in aquatic and terrestrial food
chains represents a potential hazard to man by consumption of certain
species of oceanic fish, of fish or shellfish from contaminated
waters, and of game birds in areas where methylmercury fungicides are
1.2.4 Environmental exposure levels
The concentration of mercury in the atmosphere is usually below
50 ng/m3 and averages approximately 20 ng/m3. A concentration of
50 ng/m3 would lead to a daily intake of about 1 µg. "Hot spots" near
mines, smelting works, and refineries require further investigation
but could lead to daily intakes as high as 30 µg. Daily intakes would
be higher for occupational exposures to mercury vapour. An average
mercury concentration in air of 0.05 mg/m3 would lead to an average
daily intake via inhalation of about 480 µg. The highest occupational
exposures usually occur in mining operations but over 50 specific
occupations or trades involve frequent exposure to mercury vapour.
Mercury in drinking water would contribute less than 0.4 µg to the
total daily intake. Bodies of fresh water for which there is no
independent evidence of contamination contain mercury at less than
200 ng/litre. Oceanic mercury is usually less than 300 ng/litre.
Food is the main source of mercury in nonoccupationally exposed
populations, and fish and fish products account for most of the
methyl-mercury in food. Mercury in food other than fish is usually
present at concentrations below 60 µg/kg. Mercury is present in
freshwater fish from uncontaminated waters at concentrations of
between 100 and 200 µg/kg wet weight. In contaminated areas of
freshwater, mercury levels between 500 and 700 µ/kg wet weight are
often described and in some cases, concentrations are even higher.
Most species of oceanic fish have mercury levels of about 150 µg/kg.
However, the large carnivorous species (e.g. swordfish and tuna)
usually fall in the range of 200-1500 µg/kg. With few exceptions
methylmercury accounts for virtually all the mercury in both
freshwater and marine fish.
Intake of mercury from food is difficult to estimate with
precision. Daily intake from food other than fish is estimated as 5 µg
but the chemical form of mercury is not known. Most of the
methylmercury in diet probably comes from fish and fish products. The
median daily intake of methylmercury in Sweden has been estimated as
5 µg. In most countries the daily intake is less than 20 µg but in
subgroups in certain countries where there is an unusually high fish
intake (dieters) the daily intake may rise to 75 µg and may even be as
high as 200-300 µg (in coastal villages dependent on large oceanic
fish as the main source of protein). In areas of high local pollution,
daily intakes could be well in excess of 300 µg and these levels have
led to two recorded outbreaks of methylmercury poisoning.
1.2.5 Metabolism of mercury
Approximately 80% of inhaled mercury vapour is retained.
Information on pulmonary retention of other forms of mercury in man is
lacking. Absorption of inorganic mercury compounds from foods is about
7% of the ingested dose. In contrast, gastrointestinal absorption of
methylmercury is practically complete. Little information is available
on skin absorption although it is suspected that most forms of mercury
can penetrate the skin to some extent. In the case of methylmercury,
poisoning has resulted from skin application.
Animal data indicate that the kidneys accumulate the highest
tissue concentrations no matter what form of mercury is administered.
The distribution of mercury between red cells and plasma depends upon
the form of mercury. The red cell to plasma ratio is highest for
methylmercury (approximately 10) and lowest for inorganic mercury
(approximately 1) in man.
The hair is a useful indicator medium for people exposed to
methylmercury. The concentration of mercury in hair is proportional to
the concentration in the blood at the time of formation of the hair.
The relationship between hair and blood concentrations is not known
for other forms of mercury.
Most forms of mercury are predominantly eliminated with urine and
faeces. In workers exposed over a long period to mercury vapour,
urinary excretion slightly exceeds faecal elimination. On a group
basis, mercury excretion in urine is proportional to the time-weighted
average air concentration. Large individual fluctuations are common in
daily mercury excretion in urine in people under the same exposure
Faecal elimination accounted for approximately 90% of total
mercury elimination in volunteers given a single dose of
methylmercury. Urinary concentrations of total mercury do not
correlate with blood levels after exposure to methylmercury.
Animal data indicate that elemental mercury vapour rapidly crosses
the placenta. The transplacental transfer of methylmercury compounds
is well documented in man. The mercury concentrations in plasma in the
mother and the newborn infant are similar but the concentration in the
fetal red blood cells is approximately 30% higher than in those of the
Details on transmission into breast milk are available only for
methylmercury. The concentration of mercury in breast milk is
approximately 5% of the simultaneous mercury level in blood in the
mother, and infants can accumulate dangerously high blood
concentrations by suckling if their mothers are heavily exposed.
Tracer studies in volunteers and in exposed populations have
established the main features of the metabolic model for methylmercury
in man. Clearance half-times from the whole body and from blood are
about 70 days. Daily intakes of methylmercury will lead to a
steady-state balance in about one year, when the body burden will be
approximately one hundred times the daily intake. In steady-state, the
numerical value of the concentration of mercury in whole blood in
µg/litre is virtually equal to the numerical value of the daily intake
in µg/day/70 kg body weight. Considerable individual variation around
these average values has been noted, which must be taken into account
in the estimation of risk in exposed populations.
The metabolic models for other forms of mercury are less well
1.2.6 Experimental studies on the effects of mercury
Reversible and irreversible toxic effects may be caused by mercury
and its compounds, depending upon the dose and duration of exposure.
Reversible behavioural changes may be produced in animals by exposure
to mercury vapour.
Methylmercury compounds produce irreversible neurological damage
in animals. Many of the neurological signs seen in man have been
reproduced in animals. Methylmercury is equally toxic to animals
whether it is given in the pure chemical state or in fish where it has
accumulated naturally. A latent period lasting weeks or months is
observed between cessation of exposure and onset of poisoning.
Morphological changes have been seen in the brain before onset of
signs. This phenomenon has been referred to as "silent damage". Animal
data support epidemiological evidence from Japan, that the fetus is
more sensitive than the adult.
Little is known about the physical and chemical factors affecting
the toxicity of mercury. Selenium is believed to be protective against
inorganic and methylmercury compounds.
1.2.7 Epidemiological and clinical studies
The classic symptoms of poisoning by mercury vapour are erethism
(irritability, excitability, loss of memory, insomnia), intentional
tremor, and gingivitis. Most effects of mercury vapour are reversible
on cessation of exposure, although complete recovery from the
psychological effects is difficult to determine. Recovery may be
accelerated by treatment with penicillamine and unithiol
Studies of occupational exposure to mercury vapour reveal that the
classic symptoms of mercurialism do not occur below a time-weighted
average mercury concentration in air of 0.1 mg/m3. Symptoms such as
loss of appetite and psychological disturbance have been reported to
occur at mercury levels below 0.1 mg/m3.
The most common signs and symptoms of methylmercury poisoning are
paraesthesia, constriction of the visual fields, impairment of
hearing, and ataxia. The effects are usually irreversible but some
improvement in motor coordination may occur. Complexing and chelating
agents may be useful in prevention if given early enough after
exposure but BAL is contraindicated in cases of methylmercury
poisoning as it leads to increased brain levels of mercury.
Epidemiological investigations have been made on populations in
whom the intensity and duration of exposure to methylmercury through
diet differs, for example, a population in Iraq having-high daily
mercury intakes (as high as 200 µg/kg/day) for a brief period (about 2
months), populations in Japan having lower daily intakes with exposure
for several months or years, and several fish-eating populations
having daily intakes of mercury usually below 5 µg/kg but with
exposure lasting for the lifetime of the individual. The results of
these studies indicate that the effects of methylmercury in adults
become detectable in the most sensitive individuals at blood levels of
mercury of 20-50 µg/100 ml, hair levels from 50-120 mg/kg, and body
burdens between about 0.5 and 0.8 mg/kg body weight.
Observations on the Minamata outbreak in Japan indicate that the
fetus is more sensitive to methylmercury than the adult but the
difference in degree of sensitivity has not yet been established.
1.2.8 Evaluation of health risks to man from exposure to mercury
and its compounds
Adverse health effects have not yet been identified in workers
occupationally exposed to a time-weighted average air concentration of
mercuryof 0.05mg/m3. This air concentration is equivalent to an
average mercury concentration in blood of 3.5 µ/100 ml and an average
mercury concentration in urine of 150 µg/litre on a group basis. The
corresponding ambient air concentration of mercury for exposure of the
general population would be 0.015 mg/m3.
It is estimated that the first effects associated with long-term
daily intake of methylmercury should occur at intake levels between 3
and 7 µg/kg/day. The probability of an effect (paraesthesia) at this
intake level is about 5% or less in the general population. These
figures apply only to adults. Prenatal life may be the most sensitive
stage of the life cycle to methylmercury. Furthermore experiments on
animals indicate a potential for genetic damage by methylmercury.
1.3 Recommendations for Further Research
1.3.1 Environmental sources and pathways of mercury intake
More information is needed on the physical and chemical forms of
mercury in air, food, and water. With the exception of fish tissue,
little is known of the proportion of total mercury in the diet that is
in the form of methylmercury.
The concentration of mercury in the air in "hot spots" near points
of industrial release is not yet adequately documented. The few
reports reviewed in this criteria document indicate that people living
near points of emission may receive substantial exposure to airborne
mercury. Levels of mercury in the oceans are still inadequately
documented. The pathways of methylation of mercury in the ocean and
its uptake by fish of different trophic levels are poorly understood.
Studies are needed to estimate quantitatively the dietary intake
of methylmercury in populations dependent on fish for their main
source of protein. Average dietary intakes for the populations of
several industrialized countries have been reported. However, of much
greater importance are the identification of those subgroups of the
population having unusually high dietary intakes of methylmercury and
the careful quantitative estimation of average daily intake in these
1.3.2 Metabolic models in man
The kinetic parameters of uptake, distribution, and excretion of
methylmercury in man are documented in much more detail than for other
forms of mercury. However, questions still remain on the linearity of
this metabolic model at high toxic doses of methylmercury.
Specifically, the applicability of the metabolic model derived from
human tracer-dose studies should be verified at higher dose levels.
Information on this point would greatly facilitate the interpretation
of results of epidemiological studies on heavily exposed populations.
Recent findings of large individual variations in clearance
half-times of methylmercury from blood are of considerable importance
in the estimation of risks from long-term dietary intake. Further
studies are needed to establish the statistical parameters of the
distribution of individual clearance half-times, and on the biological
mechanisms underlying these differences.
A more complete metabolic model for inhaled mercury vapour in man
is urgently needed. Despite the continuous occupational exposure of
thousands of workers annually and the long history of man's exposure
to this form of mercury, we still do not have sufficient information
to relate mercury concentrations in air to accumulated body burdens
and to identify the most appropriate indicator media for levels of
mercury vapour in the target organ (the brain). Animal experiments
have indicated the ability of the inhaled vapour to cross the
placenta; no information is available on human subjects concerning
this important question.
1.3.3 Epidemiological studies
Several types of epidemiological study are needed. Long-term
studies on adults should concentrate on those areas of the
dose-response relationship where the effects of methylmercury become
just detectable. There are still uncertainties concerning the
concentrations of total mercury in indicator media and the equivalent
long-term daily intake of mercury as methylmercury associated with the
earliest effects in the most sensitive group in the adult population.
So far, dose-response relationships in human populations have been
based on outbreaks of poisoning in which daily exposure was high and
limited to months or a few years at the most. To extrapolate these
relationships to the general population, more information is needed on
the potential influence of long-term exposure.
In addition to continuing studies on mature adults, groups of the
population specially sensitive to methylmercury should be identified.
Special studies should be made on the relationship between the dose
received by the expectant mother and the effect on her infant
including the development and growth of the child.
Further epidemiological studies are needed on groups
occupationally exposed to mercury vapour. Whenever possible,
collaborative studies should be carried out in which cohorts should be
followed in time and different groups related to each other.
1.3.4 Interaction of mercury with other environmental factors
The extrapolation to the general population of epidemiological
data from outbreaks of methylmercury poisoning that have occurred in
certain parts of the world is fraught with uncertainties, unless the
possible interaction of local environmental factors can be taken into
account. For example, the conditions under which selenium exerts
antagonistic and synergistic effects and its mode of action should be
studied. Alcohol influences the metabolism of mercury and may affect
the toxicity of inhaled vapour in man. Genetic factors should also be
considered. Acatalasaemic individuals may metabolize inhaled mercury
vapour differently from normal individuals.
Mercury, along with other heavy metals, has the potential to alter
the activity of drug metabolizing enzymes. Studies should be made on
these potential effects with special emphasis on those individuals
carrying high body burdens of mercury.
1.3.5 Biochemical and physiological mechanisms of toxicity
Long-term investigations of the mode of toxic action of mercury
and its compounds are needed to give an insight into the causes of
individual differences in sensitivity to mercury and into differences
in metabolism such as clearance half-times. Methylmercury is known to
produce "silent damage" in that morphological changes can be seen in
the brains of experimental animals before functional disturbances are
detectable. Biochemical disturbances such as inhibition of protein
synthesis precede overt signs of damage. There is a great need to
develop sensitive biochemical and physiological tests, especially in
the case of methylmercury poisoning.
A deeper understanding of the toxic action of mercury should lead
to the development of more effective means of treatment. Present
methods depend mainly on prevention, using complexing and chelating
agents to remove the metal from the body before serious damage has
2. PROPERTIES AND ANALYTICAL METHODS
2.1 Chemical and Physical Properties
Mercury can exist in a wide variety of physical and chemical
states. This property presents special problems to those interested in
assessing the possible risk to public health. The different chemical
and physical forms of this element all have their intrinsic toxic
properties and different applications in industry, agriculture, and
medicine, and require a separate assessment of risk.
The chemistry of mercury and its compounds has been outlined in
several standard chemistry texts (Rochow et al., 1957; Gould, 1962;
Cotton & Wilkinson, 1972). Mercury, along with cadmium and zinc, falls
into Group IIb of the Periodic Table. In addition to its elemental
state, mercury exists in the + 1 (mercury(I)) and +2 (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 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 (only 15-20 cal/mol and actually weaker
than zinc and cadmium bonds) 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 alkylmercurials in which mercury is attached to the carbon
atom of a methyl, ethyl, or propyl group.
An expert committee, considering occupational hazards of mercury
compounds, distinguished two major classes of mercury compounds --
"organic" and "inorganic" (MAC Committee, 1969). Inorganic mercury
compounds included the metallic form, the salts of mercury(I) and
mercury(II) ions, and those complexes in which mercury(II) was
reversibly bound to such tissue ligands as thiol groups and protein.
Compounds in which mercury was directly linked to a carbon atom by a
covalent bond were classified as organic mercury compounds. This
distinction is of limited value because the toxic properties of
elemental mercury vapour differ from those of the inorganic salts and,
furthermore, the short-chain alkylmercurials differ dramatically from
other mercurials that fall within the definition of organic mercury.
From the standpoint of risk to human health, the most important forms
of mercury are elemental mercury vapour and the short-chain
Mercury in its metallic form is a liquid at room temperature. Its
vapour pressure is sufficiently high to yield hazardous concentrations
of vapour at temperatures normally encountered both indoors and
outdoors under most climatic conditions. For example, at 24°C, a
saturated atmosphere of mercury vapour would contain approximately
18 mg/m3 -- a level of mercury 360 times greater than the average
permissible concentration of 0.05 mg/m3 recommended for occupational
exposure by the National Institutes of Safety and Health, USA (NIOSH,
1973). Apart from the noble gases, mercury is the only element having
a vapour which is monatomic at room temperature. However, little is
known about the chemical and physical states of mercury found in the
ambient air and in the air where occupational exposure occurs.
Elemental mercury vapour is generally regarded as insoluble.
Nevertheless, small amounts dissolved in water and other solvents are
important from the toxicological point of view. At room temperatures,
in air-free water, its solubility is approximately 20 µg/litre. In the
presence of oxygen, metallic mercury is rapidly oxidized to the ionic
form -- mercury(II) -- and may attain concentrations in water as high
as 40 µg/litre.
Calomel or mercury(I) chloride (Hg2Cl2) is the best known
mercury(I) salt. Widely used in the first half of this century in
teething powders and in anthelmintic preparations, the low toxicity of
this compound is due principally to its very low solubility in water.
Mercury(I) forms few complexes with biological molecules. However, in
the presence of protein and other molecules containing SH groups, it
gives one atom of metallic mercury and one mercury(II) ion. In
general, an equilibrium is established between Hg0, Hg2++ and Hg++
in aqueous solution. The distribution of mercury between the three
oxidation states is determined by the redox (oxidation-reduction)
potential of the solution and the concentration of halide, thiol, and
other groups that form complexes with Hg++. The dissociation of
mercury(I) chloride by thiol groups should be understood in this
context. Extra halide and thiol compounds, added to solution, form
complexes with mercury(II) ions and the mercury(I) chloride splits to
restore the equilibrium between Hg0, Hg2++ and Hg++. The split
results in the formation of one atom of mercury for every mercury(I)
chloride molecule dissociated.
The mercury(II) ion, Hg++, is able to form many stable complexes
with biologically important molecules. Mercury(II) chloride (corrosive
sublimate), a highly reactive compound, readily denatures proteins and
was extensively used in the past century as a disinfectant. It is
soluble in water and, in solution, forms four different complexes with
chloride, HgCI+, HgCl2, HgCl3- and HgCl4=. It has been
suggested that the negatively charged chlorine complexes are present
in sea water (see section 5).
Phenylmercury compounds have a low volatility. However, the halide
salts of methyl-, ethyl-, and methoxyethylmercury can give rise, at
20°C, to saturated mercury vapour concentrations of the order of 90,
8, and 26 mg/m3, respectively (Swensson & Ulfvarsson, 1968). In the
case of methylmercury this saturated vapour concentration is several
orders of magnitude greater than the maximum allowable concentration
in the working atmosphere. This hazardous property of the halide salts
of the short-chain alkylmercurials is not always fully appreciated in
industrial and agricultural use and even in research laboratories
(Klein & Hermen, 1971). In contrast, methylmercury dicyandiamide,
previously widely used as a fungicide, has a much lower vapour
pressure, being 340 times less volatile than the chloride salt.
Although the carbon-mercury bond is chemically stable, in the
living animal, the bond is subject to cleavage (for review, see
Clarkson, 1972a). The nature of the R radical is all important. If R
is a phenyl or methoxyalkyl group, rapid breakdown occurs in animal
tissues so that most of the organic compound has disappeared within a
few days. Enzymes that break the carbon mercury bond have been
discovered and isolated (Tonomura et al., 1968a, 1968b, 1968c). The
short-chain alkylmercurials undergo the slowest breakdown in vivo
with methylmercury being the most stable. Differences in the stability
of the carbon-mercury bond play an important role in determining the
toxicity and mode of action in man. The rapid breakdown of phenyl- and
methoxymercury results in toxic effects similar to those of inorganic
mercury salts. The relative stability of the alkylmercurials is one
important factor in their unique position with regard to toxicity and
risks to human health.
The organic and inorganic cations of mercury, in common with other
heavy metal cations, will react reversibly with a variety of organic
ligandsa found in biologically important molecules. The chemical
affinity of mercury(II) and of its monovalent alkylmercury cations for
a variety of biologically occurring ligands is so great that free
mercury would be present in vivo at concentrations so low as to be
undetectable by present methods.
2.2 Purity of Compounds
Impurities in mercury and its compounds are not important in
assessing the hazards to man. Those compounds of mercury used in
industry and agriculture have impurities of less than 10%. Bakir et
al. (1973) reported that a methylmercury fungicide responsible for an
epidemic of poisoning in Iraq contained 10% or less of ethylmercury as
an impurity. Inorganic mercury usually amounts to no more than 1% of
the total mercury in organomercurial preparations and rarely exceeds
Impurities are of importance in the preparation of standard
solutions for analytical procedures and in experimental research in
animals where impurities in radioactive mercury may give misleading
results. Preparations of methylmercury labelled with the isotope 203Hg
are subject to radiolytic breakdown to inorganic compounds depending
on the pH. This instability must be taken into account in the
interpretation of some original reports in which the purity of the
radioisotope was not checked properly.
2.3 Sampling and Analysis
Before reviewing various aspects of sample collection and analysis
it may be worth taking an overview of the various sources of error in
the determination of mercury content. Not only are there errors in the
instrumental determination of mercury and in the laboratory
procedures, but significant and often major errors occur during the
collection, transportation, and storage of the samples. The accuracy
of the determination of mercury in environmental samples should be
assessed from this broad point of view. The error will be the sum of
a Ligands are chemical groups within a molecule that are capable of
donating electrons to a metal cation to form a chemical bond.
Examples of biologically important ligands are the carboxyl, and
especially with regard to heavy metals, the sulfhydryl (SH)
the errors in collection, storage, transportation and, in the
instrumental determination. It is of the greatest importance to
determine the greatest source of error in each particular case. This,
in itself, may lead to considerable improvement in the overall
accuracy of the determination. For example, the introduction of a new
and more sensitive instrumental technique may allow the collection of
smaller samples and thus facilitate storage and transport. On the
other hand, there is little value in proceeding further with
improvements in instrumental measurements if major errors remain at
the collection, storage, or transport stages.
2.3.1 Sample collection
Methods of sample collection for the determination of mercury in
air have recently been reviewed (NIOSH, 1973). A recommended method
for the determination of total mercury in air is presented.
Essentially the method consists of using two bubblers in series,
containing sulfuric acid and potassium permanganate. The mercury in
these traps is subsequently determined by atomic absorption
procedures. Problems of the determination of mercury in air are
critically evaluated. Included in these problems is the fact that
numerous chemical and physical forms of mercury may exist in air and
that these are subject to interconversion. The volatility of mercury
and its compounds is a special problem in the determination of mercury
bound to particles. The separation of particulates from air, such as
by filtration, may result in the loss of mercury by volatilization
from the particulate. Published methods of sample collection consist
of removal of mercury from the air by passing it through scrubbing
devices, or direct collection of the air sample, for example in a
plastic bag or syringe. The scrubbing device may take the form of
bubblers, filters, absorbants, or amalgam collectors. Unfortunately
many of the published procedures do not report collection efficiency.
Attention is drawn to the importance of the use of standard dust
chambers to check the efficiency of absorption.
The procedure recommended by NIOSH (1973) has a collection
efficiency for total mercury of more than 90%, when mercury is in the
form of elemental vapour or inorganic salts. Organomercurials in air
are collected with an efficiency of more than 80%, except in the case
of the short-chain alkylmercurials. Bramen (1974) has described a
procedure for separating and measuring different physical and chemical
forms of mercury in air. Previous reports distinguishing between
mercury vapour and particle-bound mercury have not reported the
efficiency of collection.
An early method (Polesajev, 1936) for the determination of mercury
in air involved absorption in iodine and subsequent determination of
the coloured complex in the sediment. This method is still widely used
in the Soviet Union and some countries of eastern Europe.
Commercially available portable monitoring devices are used to
determine mercury directly in air. The air is pumped through an
optical cell that measures the absorption of light emitted from a
mercury vapour lamp. These units, although convenient, measure only
elemental mercury vapour and are subject to a wide variety of
interferences and interfering substances many of which are likely to
be present in the working environment. These units should be
calibrated each time before use. The commercial units also suffer from
the deficiency that they sample only small volumes of air that may not
give a representative picture of the working environment. Research
should be directed towards the development of personal monitoring
devices. These devices should be small and portable so that they can
be carried by workmen throughout the working day and thereby give a
cumulative picture of the exposure of each individual. In most cases
it would be necessary only to devise systems for collecting total
The method of Wolf et al. (1974) allows the direct detection of
mercury using reactive tubes (Draeger tubes) providing a simple
screening method for determining mercury in working places at sporadic
The collection of samples for the determination of mercury in
water must take into account the following factors; (a) the low
concentration of mercury in water, normally of the order of
10 ng/litre; (b) the tendency of mercury to adsorb on to the surface
of the collection vessel at these low concentrations; (c) the
possibility, if not likelihood, of volatilization of mercury from the
sample (Toribara et al., 1970) and (d) the type of collection vessel.
Greenwood & Clarkson (1970) have reported on the rates of loss of
mercury from containers made from ten different materials and
suggested that Pyrex, polycarbonate, and Teflon are the best materials
for storing and handling mercury. Further studies of possible losses
of organomercurials through the walls of some plastic containers
should, however, be studied. Losses due to volatilization may be
reituced by the addition of oxidizing substances such as potassium
permanganate (Toribara et al., 1970). Lamm & Ruzika (1972) have
recommended that radioactive-tracer mercury be added to the sample to
check the losses discussed above. They note that this procedure has
rarely been adopted to date.
For the collection and storage of food samples, acceptable
procedures are usually followed. The most important food items for
determination of mercury are those containing fish and fish products.
Mercury levels in other foodstuffs usually do not amount to a
significant fraction of daily exposure unless the food has accidently
been contaminated, such as by the use of pesticides. In the collection
and storage of food samples prior to analysis, care should be taken to
avoid bacterial growth leading either to the breakdown of organic
mercury compounds or to the volatilization of mercury (Magos et al.,
Samples of blood, hair, and urine have been used to monitor the
exposure of human beings to mercury. The methods of collecting and
storing these samples are of great importance. With respect to blood
samples, care should be exercised to avoid any clot formation. If this
does occur, the sample should be homogenized thoroughly before
analysis. It is useful, in certain situations, to determine mercury in
the red cells and plasma and it is thus important to avoid any
haemolysis of the blood sample. The nature of the anticoagulants used
does not affect the mercury determinations, of either the total
mercury in whole blood or the distribution of mercury between plasma
and red blood cells. "Vacutainers"a are convenient for blood
collection and allow storage of the blood samples in Pyrex tubing
under aseptic conditions. Blood samples that have been contaminated by
microorganisms and stored in the refrigerator at 4°C for a month or
more may give misleading results due to the breakdown of methylmercury
and other organic mercury compounds (Clarkson, personal communication,
1974). The storage of blood samples in the frozen state or
freeze-dried is suitable providing that mercury is determined only for
whole blood. Significant losses of mercury do not occur during
freeze-drying procedures (Albanus et al., 1972).
Measurement of mercury in urine samples has been used as a measure
of exposure to mercury under industrial conditions. The popularity of
this approach in early studies was mainly due to the case of digestion
of the urine sample. However, there are serious problems in the
collection and storage of urine samples that may seriously influence
the results. The following factors have been recognized; (a) the
time of day of urine collection (Piotrowski et al., 1975),
(b) bacterial contamination, which might give rise to significant
losses of mercury by volatilization (Magos et al., 1964), (c) the
nature of the container (Greenwood & Clarkson, 1970),
(d) contamination from mercury in workers' clothing and from the
collection of urine samples under working conditions. It should be
noted that urine samples do not give a reliable indication of exposure
to methylmercury (Bakir et al., 1973).
Hair samples are becoming the samples of choice in determining
exposure to methylmercury through diet. Depending upon the length of
the hair sample, it is possible to recapitulate exposure to
methylmercury for several yearsb. The concentration of mercury in
hair when formed is directly proportional to the concentration of
a Trade name of heparinized test-tube manufactured by Becton &
Dickinson, USA, and used for collection of blood samples.
b The average rate of growth of hair is approximately 1 cm
per month (Giovanoli et al., 1974; Shahristani & Shihab,
mercury in the blood, the concentration in hair being about 250 times
the concentration in blood. The ratios are well established for
exposure to methylmercury but only limited information is available
for inorganic mercury. Attention has been drawn to the errors
introduced during the collection and transportation of hair samples
(Giovanoli & Berg, 1974). Usually 50-100 strands of hair are needed
for analysis. Differential rates of growth for each strand and lateral
displacement of the samples during cutting and transportation of the
hair will affect the longitudinal profiles of mercury in the hair
sample. Giovanoli & Berg (1974) have described a computerized
procedure for the correction of these artifacts.
2.3.2 Analytical methods
Methods of analysis are usually classified according to the type
of instrument used in the final measurement. This convenient
classification will be used here. However this approach tends to
belittle the role of the skill and experience of the analyst. In fact
a poor method in the hands of a highly skilled analyst is more likely
to yield accurate results than a good method in the hands of a poor
analyst. In recent years it has become a practice to test methods by a
"round robin" distribution of a standard sample. Comparison of results
from the participating laboratories is more likely to give information
on the competence of the analysts in the laboratory than it is to give
a critical evaluation of the method itself.
Measurement of the very low levels of mercury found in the
non-contaminated environment makes special demands both on the skills
of the analyst and the resources of the method employed. No matter how
frequently used, a method for the determination of mercury in nanogram
quantities cannot be regarded as a routine procedure. Continued
vigilance over the results is an absolute requirement. Furthermore,
where conditions allow, it is highly desirable that the results with
one method and from one laboratory be checked against those with a
different method from another laboratory. One useful combination of
different procedures is the analysis of total and inorganic mercury by
selective atomic absorption and the selective analysis of organic
mercury compounds (usually methylmercury and other short-chain
mercurials) by gas chromatography (Giovanoli et al., 1974).
The literature is full of papers concerning methods of determining
mercury. Several recent reviews have appeared (D'Itri, 1972; NIOSH,
1973; Burrows, 1975, Swedish Expert Group, 1971; Wallace et al., 1971;
CEC Working Group of Experts, 1974). The most frequently used methods
for measurements of total mercury are colorimetric (dithizone),
flameless atomic absorption, and neutron activation. The flameless
atomic absorption method has become the "work-horse" for measurement
of environmental samples. Difficulties might arise in the measurement
of mercury owing to the fact that it is strongly bound to the organic
materials in most samples. Many procedures require the destruction of
organic materials by wet oxidation or by high temperatures. Loss of
mercury by volatilization may occur. If the wet oxidation is too mild
the result will be inadequate recovery. A high reagent blank may be
introduced by the chemicals used for oxidation. In certain procedures
involving atomic absorption or neutron activation the digestion of the
sample or heating of the sample is not necessary. These procedures
have the advantage of having a low blank but problems of variable
recovery or interference may arise.
The determination of mercury by colorimetric measurement of a
mercury dithizonate complex has been the basis of most of the methods
in the 1950s and in the 1960s. Other related methods using dithizone
for measuring mercury in environmental samples have been described by
Kudsk (1964) and Smart et al. (1969). The above procedures all make
use of wet oxidation of the sample followed by extraction of mercury
in an organic solvent as a dithizonate complex and finally the
colorimetric determination of the complex itselfa. Selectivity for
mercury is obtained by adjusting the conditions of extraction. Copper
is the metal most likely to interfere with mercury measurement by
The dithizone procedure has an absolute sensitivity of about
0.5 µg of mercury. A sample size of 10 g is suitable for most
digestion procedures so that mercury can be determined at the
0.05 mg/kg level in most foodstuffs and tissues.
Kudsk (1964) has described a dithizone procedure for measuring
mercury in air that will measure as little as 0.05 µg of mercury. With
the usual sample size of 0.1 m3, the detection limit would be
0.5 µg/m3. This is more than adequate sensitivity for monitoring air
in the working environment with the MAC levels in force. The quoted
recovery rates from foodstuffs and tissues are in the range of 85-99%
and the reproducibility can yield a coefficient of variation of as low
as 2%. On account of its long history of use, the dithizone procedure
has been used to measure mercury in virtually all types of
environmental samples including air, water, food, tissues, and soils.
It suffers from the disadvantage that it is time consuming and its
sensitivity is not high when compared with atomic absorption
a The organic material may also be destroyed by combustion in an
oxygen flask (Gutenmann & Lisk, 1960; White & Lisk, 1970; and
Fujita et al., 1968). This allows all biological materials to be
treated alike but has the disadvantage of requiring dried
The latest developments in atomic absorption procedures have
recently been reviewed by Burrows (1975). The most commonly used
method in the USA is that of Hatch & Ott (1968) as modified by Uthe et
al. (1970). The procedure involves oxidative digestion ("wet ashing"),
followed by reduction, aeration, and measurement of mercury vapour
absorption at 253.7 nm. The detection limit is approximately 1-5 ng of
mercury. The wide popularity of cold vapour atomic absorption has
resulted in a large number of publications dealing with various
applications of this procedure to the measurement of mercury in
sediments, soils, and biological samples (including foodstuffs). Of
the 16 publications reviewed by Burrows (1975), 13 reported recoveries
of 90% or more. The relative standard deviation was 10% or less in
half of the published procedures, and was less than 20% in more than
90% of these procedures.
The measurement of very low levels of mercury in water samples
requires some preconcentration. This may be achieved by dithizone
extraction (Chau & Saiton, 1970; Thomson & McComas, 1973), by
electrodeposition (Doherty & Dorsett, 1971) and by an amalgamation on
silver wire (Hinkle & Learned, 1969; Fishman, 1970), in each case
permitting detection limits of 1 ng/litre-10 ng/litre. Winter &
Clements (1972) have described a procedure that will measure mercury
in water in the range of 200 ng/litre and does not require
Magos (1971) has described a reduction technique that selectively
determines total and inorganic mercury in biological samples without
digestion of the material. This technique has been modified by Magos &
Clarkson (1972) to permit determination of mercury in blood samples at
the low levels found in unexposed populations (0.1-1.0 µg/100 ml). The
technique has a sensitivity of approximately 0.5 ng of mercury.
Recently it has been successfully applied to the measurement of total
and inorganic mercury in hair samples (Giovanoli et al., 1974). The
relative standard deviation was 2% and the recovery rates were quoted
as being close to 100%. The technique has the advantage of high speed
-- each determination taking less than 2 minutes -- high sensitivity,
and the apparatus involved is light, portable, and suitable for field
applications. Its widest application to date has been in the
measurement of mercury in biological samples in the large Iraq
outbreak (Bakir et al., 1973). Since the procedure does not require
digestion of the biological sample, internal standards are used in
each determination. The rates in this procedure must be checked for
each new biological matrix.
The atomic absorption techniques referred to above are subject to
interference. The most common interfering substances are benzene and
other aromatic hydrocarbons that absorb strongly in the 253.7 nm
region. Interference from a variety of organic solvents has been
reported by Kopp et al. (1972).
The combustion-amalgamation method has undergone a series of
developments to avoid difficulties due to interfering substances.
Reference may be made to the work of Lidmus & Ulfvarson (1968), Okuno
et al. (1972), and Willford (1973) who developed techniques for
oxidation of the biological sample, and the trapping of mercury vapour
on silver or gold followed by its release into an atomic absorption
measuring device. All these methods have sensitivities down to the
1 µg/litre level and avoid the risk of interference from other
substances. However, as pointed out by Burrows (1975), care must be
taken in the design and operation of the combustion tube to avoid
losses of volatile mercury derivatives.
In summary, a wide variety of applications of atomic absorption
procedures have now been published. The technique is rapid and
sensitive and the procedure is technically simple. Procedures are
available for avoiding difficulties due to interfering substances.
Most procedures have a detection limit in the range of 0.5-5 ng of
mercury and a relative standard deviation of about 10% or less.
Recovery rates are usually of the order of 95-100% depending on the
technique used in the preparation of the biological sample and the
rate of release of mercury from it.
Procedures for neutron activation analysis of total mercury have
recently been reviewed by Wallace et al. (1971), Swedish Expert Group
(1971), Westermark & Ljunggren (1972), and Burrows (1975). The method
is based on the principle that when natural mercury (a mixture of
stable isotopes) is exposed to a high flux of thermal (slow) neutrons,
it is converted to a mixture of radioactive isotopes, principally
197Hg and 203Hg, which have decay half-lives of 65 hours and 47 days,
respectively. The Sjostrand (1964) technique has been used most in the
measurement of environmental samples. After the sample has been
irradiated with neutrons, a precise weight of carrier mercury is added
and the sample subjected to digestion and organic destruction. On
completion of digestion, mercury is isolated by electrodeposition on a
gold foil and the radioactivity is determined with a gamma counter.
The use of carrier mercury corrects for any losses of mercury during
the digestion, extraction, and isolation procedures. The limit of
detection is 0.1-0.3 ng of mercury. The sample size is 0.3 g giving a
concentration limit of 0.3-1 µg/kg in most biological samples. The
relative standard deviation in samples of kale, fish, minerals, oil,
blood, and water is less than 10%. Samuel (unpublished data)
decomposed biological material irradiated with neutrons using fuming
sulfuric acid and hydrogen peroxide and after the addition of hydrogen
bromide, distilled the mercury as bromide together with other trace
elements. This method, which is suitable for series analysis, is
characterized by high recovery (96%) and good reproducibility. Trace
mercury in biological and environmental materials can also be rapidly
and satisfactorily determined through isolation as mercury(II) oxide
or mercury(II) sulfide after digestion and clean-up procedures
following neutron activation (Pillay et al., 1971; Samuel, unpublished
In general, the analyst is faced with three major options in the
use of neutron activation procedures; (a) destruction or
non-destruction of the sample, (destruction and isolation of the
mercury is usually required in samples containing less than 1 µg of
mercury); (b) the choice of isotope 197Hg (if the longer-lived
isotope, 203Hg, is used the sample may be allowed to stand to avoid
interference from short-lived elements activated along with the
mercury -- however, 203Hg requires a more intense neutron flux or a
longer irradiation time to achieve the same activity as the 197Hg);
(c) the choice of detector (the sodium iodide (thallium) detector
does not have as high a resolution as the germanium (lithium)
detector, although its sensitivity is significantly higher).
Interference may come from the following elements, produced at the
same time as the radioactive mercury isotopes, 24Na, 82Br, 32P, and
75Se. Interference from these isotopes may be avoided, as in the
Sjostrand (1964) procedure, by chemical isolation of the radioactive
isotope. However, 75Se may not be completely removed by the isolation
procedures and might interfere if the sodium iodide (thallium)
detector is used. The better resolution of the germanium (lithium)
detector allows correction for 75Se interference through use of other
lines in the 75Se spectrum. For samples containing more than 1 µg of
mercury, the required selectivity can be achieved without destruction
of the sample, i.e., by instrumental analysis only. One procedure is
to measure the 203Hg isotope, after allowing the sample to stand for
approximately one month to eliminate interference due to sodium,
phosphorous, and bromine. Another procedure is to make use of the
discriminating germanium (lithium) detector when the gamma irradiation
from the radioactive isotope may be determined to the exclusion of
most of the interfering radioactivity.
A recent non-destructive procedure for measuring mercury in coal
makes use of a low-energy photon detector to estimate levels at the
100 µg/kg level with a precision of 10% (Weaver, 1973).
Burrows (1975) has recently reviewed 11 publications describing
the application of neutron activation to a variety of environmental
samples. Non-destructive (instrumental) determination was used in only
two of these publications. In 9 of these publications the 197Hg
isotope was determined. Mercury levels were reported in lake water
(4 µg/litre, relative standard deviation 23%), in glacial ice
(0.2 µg/kg, relative standard deviation 90%), in coal (100 µg/kg,
relative standard deviation 10%), in whole blood (0.7 µg/100 ml,a
relative standard deviation 10%), in fish (1-3 mg/kg, relative
standard deviation less than 10%). Many environmental samples were
measured by neutron activation, especially in Sweden, before the
introduction of the atomic absorption technique (Westermark &
Compared with other methods reviewed here, the neutron activation
procedure has the following advantages; (1) high sensitivity
(approximately 0.5 µg/kg); (2) no reagent blank; (3) independence from
the chemical form of the element; and (4) non-destructive instrumental
methods applicable to samples containing 1 µg of mercury or more. It
has the disadvantages that it cannot be adapted to field use and, that
if there are large numbers of samples, special radiation facilities
and data processing are required. It is generally agreed that the
neutron activation procedure finds its most important use as a
reference method against which other procedures can be checked.
A variety of other instrumental techniques, such as X-ray
fluorescence, mass spectrometry, and atomic fluorescence, for the
measurement of total mercury have been reviewed by Lamm & Ruzicka
(1972) and by Burrows (1975). In general, some of these methods may
have a potentially higher sensitivity or selectivity for mercury. The
fact is that, at the time of writing, these procedures have not yet
found useful application in the measurement of mercury in
To summarize the present methods for the determination of total
mercury in environmental samples, it would appear that the method of
choice is that of flameless atomic absorption. No single procedure is
appropriate, however, in all circumstances. The methods of sample
handling depend upon the particular biological matrix to be analysed.
Neutron activation is principally of use as a reference method against
which atomic absorption methods may be checked.
a In this document the concentration of mercury in blood is
expressed in µg/100 ml although in some original papers the values
are given in µg/100 g. For practical purposes the difference of
about 5% can be neglected.
2.3.3 Analysis of alkylmercury compounds in the presence of
Techniques for the identification and measurement of alkylmercury
compounds in the presence of other compounds of mercury have been
reviewed recently (Swedish Export Group, 1971; Tatton, 1972; Sumino,
1975; Westöö, 1973). In general, three methods are available for the
identification of alkylmercury compounds. These include (a) paper
chromatography (Kanazawa & Sato, 1959; Sera et al., 1962), (b) thin
layer chromatography (Johnson & Vickery, 1970; Westöö, 1966, 1967;
Tatton & Wagstaffe, 1969), (c) gas-liquid chromatography (Westöö,
1966, 1967; Sumino, 1968; Tatton & Wagstaff, 1969). The paper
chromatographic techniques have given way to thin-layer chromatography
(TLC) for qualitative identification of the organomercurial compounds.
Most quantitative work is now carried out using TLC techniques, and
also gas-liquid chromatography (Westöö, 1966, 1967; Sumino, 1968;
Tatton & Wagstaffe, 1969; Solomon & Uthe, 1971). However, the method
of Magos & Clarkson (1972) that selectively determines organic mercury
by cold vapour atomic absorption is frequently applicable to the
determination of methylmercury at levels occurring in fish and blood.
Methylmercury is the only organic form of mercury present in fish.
Blood samples from people exposed to methylmercury contain only
inorganic mercury and methylmercury compounds. Thus the determination
of organic mercury by this procedure is an accurate measure of
methylmercury in these situations.
The basic procedures for samples of food, soil, and biological
materials are first, homogenization of the sample, acidification by a
hydrogen halide acid followed by extraction with an organic solvent,
usually benzene, a clean-up step involving the conversion of the
organomercurial compound to a water soluble compound usually the
hydroxide or sulfate or a cysteine complex, and re-extraction with
benzene. The benzene layer is now ready for analysis by thin-layer
chromatography for qualitative purposes or by gas-liquid
chromatography if quantitative measurements are required. A recent
variant by Rivers et al. (1972) converts the organic into inorganic
mercury and then makes use of cold vapour atomic absorption for final
The gas-liquid chromatographic system is the one most commonly
used. Problems may be encountered both in the pre-treatment of the
sample and in the gas chromatographic determination itself. All these
techniques involve non-destructive extraction of mercury from the
sample. Thus recovery rates have to be checked for every different
type of sample matrix. The efficiency of extraction of mercury is
determined by both the nature of the sample matrix and the extraction
procedures themselves. Von Burg et al. (1974) introduced the idea of
adding a tracer amount of radioactively labelled methylmercury to the
homogenate and counting the final benzene extract to check variations
in the efficiency of extraction. This procedure is well worth
consideration for routine use as it is most difficult to check
extraction recovery rates.
Acidification of the homogenate is usually achieved by the
addition of a hydrogen halide acid (usually HCl). At this point
mercury(II) chloride may be added to either the homogenate or the
benzene to tie up excess sulfur compounds and prevent recombination of
methylmercury with sulfur. Westöö (1968) has shown that this approach
may give high recovery rates but cannot be used with liver as there is
a danger of methylation of the inorganic mercury. Clean-up of the
first benzene extract is usually achieved by using solutions of
cysteine. However, this complexing agent is subject to oxidation,
particularly by substances in muds. A more suitable system in the
presence of oxidizing agents is the ammonium hydroxide-sodium sulfate
solution described by Westöö. No problems are usually encountered in
the reextraction ofmethylmercury from cysteine to benzene using
3 mol/litre hydrochloric acid. However, in the extraction procedures,
volumetric errors may arise especially when the concentration of
hydrochloric acid is low (1 mol/litre) and when small amounts of
methyl-mercury are extracted from large volumes (Westöö, 1973).
In gas chromatography, the main object is to produce sharp peaks
and attain high sensitivity. Tatton (1972) has noted that most
commercial preparations ofalkylmercury salts are not pure enough to
use as standards. Sumino (1973) prepares pure methylmercury from the
combination of inorganic mercury with tetramethyl lead salts. The peak
is identified by electron-capture detectors using tritium or nickel as
the source of beta particles. These detectors are subject to
overloading and not more than 100 ng of mercury should be determined
at one time (Tatton, 1972). Absolute confirmation of the identity of
the peak should be made by mass fragmentation methods (Sumino, 1975).
The detection limit in the Westöö procedure is approximately
1-5 µg per kilogram of sample using a 10 g sample. The precision is 3%
at the 0.05 mg/kg level for fish samples. Recovery rates are generally
above 90% but do vary with the sample matrix. Solomon & Uthe (1971)
developed a semimicro-method for the rapid determination of
methylmercury in fish tissues. Samples of about 2 g were used. A
precision of 2% was reported with recovery rates of about 99%. Samples
such as blood, liver, and kidney are much more difficult to extract
than fish tissues.
Thin-layer chromatography usually requires, for optimum spot size,
2 µg of mercury for each type of compound.
3. SOURCES OF ENVIRONMENTAL POLLUTION
The sources of mercury leading to environmental pollution have
been the subject of several recent reviews (Wallace et al., 1971;
D'Itri et al., 1972; Joint FAO/WHO Expert Committee on Food Additives,
1972; Heindryckx et al., 1974; Korringa & Hagel, 1974). Estimates of
both natural and anthropogenic sources of mercury are subject to
considerable error. In the first place the levels of mercury in
environmental samples such as ice from Greenland are extremely low and
close to the limit of sensitivity of the analytical methods. These low
values are then converted by large multiplication factors (annual
total global rainfall, 5.2 x 105 km3) so as to obtain values for the
global sources and turnover of mercury. Enormous fluctuations may be
seen in samples such as coal and oil, which are believed to be an
important anthropogenic source of mercury. Values quoted by D'Itri
(1972) indicate ranges of concentrations of mercury in crude oil
varying by a factor of 1000 and ranges in coal even greater than this.
Estimates of industrial production and consumption of mercury are
subject to the vagaries of the economic market and in recent years to
government regulation because of concern over mercury pollution.
Nevertheless, despite all the assumptions and approximations in these
procedures, the general picture that emerges from a variety of
independent calculations is that the natural sources of mercury are at
least as great as, and may substantially outweigh, the anthropogenic
sources. However, man-made sources may be of considerable importance
in terms of local contamination of the environment. For example,
Korringa & Hagel (1974) have calculated that the man-made release of
mercury in the Netherlands is 100 times greater than the release of
mercury by natural degassing processes.
3.1 Natural Occurrence
A recent review by the Joint FAO/WHO Expert Committee on Food
Additives (1972) quotes the major source of mercury as the natural
degassing of the earth's crust and quotes figures in the range of
25 000-150 000 tonnes of mercury per year. These figures originate
from a paper by Weiss et al. (1971) on concentrations of mercury in
Greenland ice that was deposited prior to 1900. The most recent
calculations on natural sources of mercury have been published by
Korringa & Hagel (1974). These authors also made use of the figures of
Weiss et al. (1971) to calculate the annual amount of mercury reaching
the earth's surface due to precipitation of rainfall and arrived at a
figure of approximately 30 000 tonnes. It was admitted that the
sources of this atmospheric mercury are not yet clearly established
but that volcanic gases and evaporation from the oceans are probably
significant sources. It was also calculated by these authors that the
run-off of mercury from rivers having a "natural mercury" content of
less than 200 ng/litre would account for approximately 5000 tonnes of
mercury per year. Measurements of the concentrations of mercury in air
attached to aerosols (Heindryckx et al., 1974) indicate that soil
dispersion to the atmosphere is not an important source of mercury.
Significant local contamination may result from natural sources of
mercury. For example, Wershaw (1970) has shown that water sources
located near mercury ore deposits may contain up to 80 µg/litre as
compared with the levels of 0.1 µg/litre in non-contaminated sources.
3.2 Industrial Production
According to a recent review by Korringa & Hagel (1974), world
production averaged about 4000 tonnes per year over the period
1900-1940. Production in 1968 was 8000 tonnes per year and, in 1973,
attained 10 000 tonnes per year. Although considerable yearly
fluctuations were noted, the average rate of increase since 1950 has
been about 2% per year. Recent concern over environmental problems
related to the use of mercury seems to have stabilized production
rates and to have led to a dramatic fall in the price of mercury. For
example, according to figures quoted by Korringa & Hagel (1974), the
1966 price was $452 per flask (a flask is 34.5 kg), the 1969 price had
risen to $510.00 but by 1972 it had fallen dramatically to $202 per
It is difficult to estimate the amount of mercury released into
the environment as a result of the mining and smelting of this metal.
High levels of mercury in lake and stream waters have been attributed
to the dumping of materials and tailings (for review, see Wallace et
al., 1971). It has been estimated that stack losses during smelting
operations should not exceed 2-3%. Thus, based on a production figure
for mercury of 10 000 tonnes in 1973, one might expect to find losses
to the atmosphere of the order of 300 tonnes per year.
3.3 Uses of Mercury
Wallace et al. (1971) have attempted to give a picture of the use
of mercury in the USA. They note that 26% of the mercury mined is not
reusable. They point out, however, that at least from the theoretical
point of view most of the remaining mercury (i.e. 74% of the mercury
mined) is reusable. To what extent these theoretical possibilities are
attained is debatable at the present moment.
Rauhut & Wild (1973) reported on the consumption and fate of
mercury in the Federal Republic of Germany in 1971. Flewelling (1975)
noted that the chloralkali industry, one of the largest users of
mercury, has been able to cut losses in water effluent by at least 99%
in the last two or three years; consequently losses from chloralkali
plants now occur predominantly by emission into the atmosphere. Losses
by volatilization into the atmosphere have been reduced (approximately
50%) by the introduction of cooling systems for effluent gases.
Korringa & Hagel (1974) take a more pessimistic point of view and
conclude that there is every reason to assume that by about 1975 all
the 10 000 to 11 000 tonnes of mercury produced per year due to mining
operations will finally find its way into the environment,
predominantly via the atmosphere.
Average consumption patterns for industrialized countries have
been summarized by Korringa & Hagel (1974) as follows: chloralkali
plants, 25%; electrical equipment, 20%; paints, 15%; measurements and
control systems, such as thermometers and blood pressure meters, 10%;
agriculture, 5%; dental, 3%; laboratory, 2%; and other uses including
military uses as detonators, 20%. This pattern of consumption in
industrialized countries is similar to that published by D'Itri (1972)
for the consumption in the USA in 1968. Included in "other uses" are
mercury compounds in catalysts, preservatives in paper pulp
industries, pharmaceutical and cosmetic preparations, and in
amalgamation processes. The use of mercury in the paper pulp
industries is dramatically declining and it was banned in Sweden in
1966 (Swedish Expert Group, 1971). Hasanen (1974) has reported that no
mercury compounds have been used in the paper pulp industry in Sweden
and Finland since 1968.
3.4 Contamination by Fossil Fuels, Waste Disposal, and
Industrial activities not directly related to mercury can give
rise to substantial releases of this metal into the environment. The
most significant source is probably the burning of fossil fuels.
Heindryckx et al. (1974) calculated the following approximate figures
based on reports published in 1971 and 1972 (Joensuu, 1971; Cardozo,
1972): the combustion of coal and lignite, 3000 tonnes per year; the
refining and combustion of petroleum and natural gas, 400 tonnes per
year; the production of steel, cement, and phosphate, 500 tonnes per
year. Korringa & Hagel (1974) made similar calculations from published
material (Joensuu, 1971; Filby et al., 1970; Cardozo, 1972; Weiss et
al., 1971). They estimated for the year 1970, an annual release of
3000 tonnes of mercury from coal burning, 1250 tonnes from mineral
oil, and 250 tonnes from the consumption of natural gas. They expected
that, by 1975, a total of 5000 tonnes of mercury would be emitted from
burning fossil fuels.
Smelting of metals from their sulfate ores should contribute some
2000 tonnes annually and the making of cement and phosphate and other
processes involving heating should have contributed another 5000
tonnes per year by 1975.
D'Itri (1972) points out that the disposal of sewage might be an
important source of environmental mercury. Calculations from data in
the literature indicate that somewhere between 200 and 400 kg of
mercury per million population may be released from sewage disposal
units. This would amount to approximately 40-80 tonnes per year for
the entire poptilation of the USA. He further points out that sewage
sludge can retain high amounts of mercury according to published
studies from Sweden (6-20 mg/kg). This sludge is sometimes used as a
fertilizer resulting in widespread dispersal of mercury or is
sometimes heated in multiple hearth furnaces when most of the mercury
would probably be released into the atmosphere. If the United States
production is taken as being roughly 30% of world consumption, one
might extrapolate the sewage release figure for the United States to
indicate that something of the order of 1000 tonnes of mercury may be
released frow sewage systems on a global scale.
The anthropogenic release of mercury has been well summarized in a
recent article by Korringa & Hagel (1974) and will be briefly stated
here. The total global release of mercury is taken as the sum of the
global production (following their pessimistic view that all will be
released into the environment) plus the release from fossil fuels and
natural gas and release from non-mercury related industries.
It was calculated that by 1975 the total anthropogenic release of
mercury on a global scale would be about 20 000 tonnes per year. These
figures should be compared with a minimum estimated release of 25 000
to 30 000 tonnes per year from natural sources. The latter figure may,
in fact, be as high as 150 000 tonnes per year, given the
uncertainties in calculations on the natural global release of
4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
Jenson & Jernelov (1972) have suggested different types of cycle
for the distribution of mercury. One cycle is global in scope and
depends upon the atmospheric circulation of elemental mercury vapour.
The other cycle is local and is based on an assumed circulation of
volatile dimethylmercury compounds. In the global cycle most of the
mercury is derived from natural sources whereas the local cycle is
predominantly concerned with man-made release.
4.1 Distribution between Media -- the Global Mercury Cycle
Recent calculations on the global circulation of mercury have been
reported by Korringa & Hagel (1974). Their calculations are based
principally on data giving mercury levels in ice samples collected in
Greenland and in the Antarctic as reported by Weiss et al. (1971). The
circulation of mercury from natural sources was calculated using a
figure of 0.06 µg of mercury per kilogram of Greenland ice samples
collected prior to the year 1900. Using a reported figure for the
global precipitation of water as 5.2 x 105 km3 per year, they
estimated that minimum transport from the atmosphere to the earth
should have been about 30 000 tonnes annually, prior to 1900. The
contribution by dust particles was regarded as insignificant, an
assumption now supported by the findings of Heindryckx et al. (1974).
Based on a published figure of 4.1 x 105 km3 for annual
precipitation over the oceans, these authors estimated the annual
delivery of mercury to the oceans as 25 000 tonnes.
Korringa & Hagel (1974) also calculated the contribution of the
man-made release of mercury to the atmospheric transport cycle. They
assumed that 16 000 tonnes of mercury is now released per year to the
atmosphere from man-made sources and that the mercury is returned to
the continental land surfaces and would soon re-evaporate to the
atmosphere. The 16 000 tonnes per year would eventually find its way
into the oceans and thus the annual delivery to the oceans from both
natural and man-made sources would be 25 000 plus 16 000 tonnes which
on a proportional basis should increase the background level from the
0.06 µg/kg observed prior to the 1900s in Greenland ice to a predicted
level of 0.1 µg/kg. However, they point out that since most of the
man-made release is probably in the northern hemisphere, the present
level in Greenland ice should be somewhat higher than 0.1 µg/kg. They
note that this estimate agrees well with the observations of Weiss et
al. (1971) who found present levels in Greenland ice to range from
0.09 to 0.23 µg/kg with an average of 0.125 µg/kg. Thus, from these
rough estimates, it would appear that present day "background" levels
in rainwater, and presumably in the atmosphere, have a substantial
component related to man-made release (approximately one-third).
Observations on "background" mercury levels in the atmosphere tend
to confirm the quantitative features of this global picture
(Heindryckx et al., 1974). These authors assume that 50 000 tonnes are
released each year from the continental land masses, that the mercury
mixes up to a height of 1 km and that, in effect, the 50 000 tonnes
are located over the continental land masses that account for 30% of
the earth's surface.a The assumption of the location of this mercury
over the land masses is not in contradiction with the calculations of
Korringa & Hagel (1974). It assumes only that the atmosphere above the
land masses is in steady state, and receives 50 000 tonnes of mercury
a year as evaporation and loses 50 000 tonnes per year to the
atmosphere over the oceans. Their figure of 50 000 tonnes per year
comes from the publication of Bertini & Goldberg (1971) and agrees
well with the figure of 41 000 tonnes per year as indicated above.
With these assumptions, Heindryckx et al. (1974) concluded that the
background continental levels of mercury vapour plus aerosols should
be 10 ng/m3. The assumed mixing height of 1 km is probably the
maximum level and they suggest that the actual level of mercury in air
would lie between 1 and 10 ng/m3. These figures are in good agreement
with the published air levels as indicated in section 5.1.
Korringa & Hagel (1974) estimate the amount of mercury transported
by rivers to the oceans to be 5000 tonnes per year based on quoted
figures of 37 000 km3 of water flow via the rivers and a natural
mercury content of less than 0.2 µg/litre in river water. They note
that this figure does not change substantially if one takes into
account the fact that most of the mercury in river water is adsorbed
to suspended matter with a mercury content of 200-500 µg/kg and that
some 1010-1011 tonnes of sediment are carried each year to the
oceans. In fact river transport of mercury to the oceans may be less
than 5000 tonnes per year. Heindryckx et al. (1974) noted that the
concentrations of mercury in the North Sea and in the coastal areas
around the North Sea were far less than would be predicted if all the
mercury in the rivers entering this area were, in fact, delivered into
the oceans. Presumably a considerable amount of mercury observed in
river water is retained in sediments in the rivers and estuaries and
does not reach the ocean by normal flow of the river. Thus it would
appear that the major pathway of global transport of mercury is
metallic mercury transported in the atmosphere.
a Recent studies in Sweden cast some doubt on the validity of this
An important conclusion from these calculations on the global
cycle of mercury is that the concentration of mercury in the oceans
should not change substantially in the foreseeable future, and that
the mercury concentration in the oceans has not changed significantly
since the beginning of the industrial era. The amount of mercury in
the oceans has been calculated as 70 million tonnes using a figure for
total ocean volume of 1.37 x 109 km3 and taking the average mercury
content of ocean water as 50 ng/litre. Thus contrary to what has been
observed for the mercury content of the atmosphere, it will be a long
time before the mercury content in sea water is significantly
increased. Since water is thought to remain in the surface layers of
the ocean for 10-50 years, these authors concluded that the mercury
resulting from man-made activities should be well distributed in the
water of all the oceans and therefore should not lead to high local
This conclusion is consistent with the findings reported in
section 5.1 that mercury levels in swordfish and tuna fish caught at
the beginning of the century fall within the same range as mercury
levels reported in recent catches.
The origin of mercury released by natural processes is not well
established. Volcanic emissions are a possible source in view of the
high concentrations of mercury vapour reported in the vicinity of
volcanoes (for review, see Jonasson & Boyle, 1971). The general
"degassing" of the earth's surface is probably a major source (Weiss
et al., 1971). Levels of metallic mercury vapour in the atmosphere
over soils rich in mercury (the humus layers of topsoil) have been
reported in the range of 20-200 ng/m3 as compared to background
levels of 5 ng/m3 according to a report by Barber et al. (quoted by
Vostal, 1972). Korringa & Hagel (1974) have raised the possibility
that evaporation from the oceans may make a contribution to the
mercury present in the atmosphere in view of the substantial
quantities of water vapour that evaporate (4.48 x 105 km3). However,
it seems unlikely that mercury would evaporate at the same rate as
water in view of the fact that it is believed to be in a complex form
in the oceans (see section 5.1). Furthermore, the observations of
Williston (1968) (referred to in section 5.1) indicate that the
mercury content of the atmosphere over the oceans is considerably
lower than that over land (industrialized and rural areas).
The mechanisms of volatilization of mercury from the land masses
are not well understood. Presumably release of mercury from volcanoes
is due to the high temperatures associated with volcanic activity.
Vostal (1972) has suggested two major mechanisms, firstly the
reduction of mercury in soils by a chemical process depending on the
local redox potential, and secondly reduction by the activity of
microorganisms. The quantitative importance of these two processes is
not known. Mercury-volatilizing microorganisms are known to exist and
have been identified (Magos et al., 1964; Furukawa et al, 1969;
Tonomura & Kanzaki, 1969).
4.2 Environmental Transformationthe Local Mercury Cycle
Mercury is present naturally in the environment and released from
manmade sources in a variety of chemical and physical states. The
principal mercury ore is cinnabar, which is mercury sulfide. Andersson
(1967) has shown that mercury in soils is complexed to the organic
(humus) content. Metallic mercury may be discharged into the
environment from natural sources as discussed above and also from
man-made sources such as chloralkali plants. A variety of
organomercurial compounds are also discharged into the environment as
a result of human activities (see section 3). Both the inorganic forms
of mercury (such as metallic mercury vapour and cinnabar) and the
organic forms of mercury are subject to conversion in the environment.
Jensen & Jernelov (1972) have summarized the major pathways of
transformation. The inorganic forms of mercury (Hg0 and HgS) undergo
transformations in the environment mainly by oxidation-reduction
reactions. Mercury vapour is oxidized to ionic divalent mercury
(Hg++) in water in the presence of oxygen. Concentrations as high as
40 g/litre have been attained when water saturated with oxygen was
exposed to mercury vapour (Wallace et al., 1971). As pointed out by
Jensen & Jernelov (1972) the oxidation of metallic mercury to
inorganic divalent mercury is greatly favoured when organic substances
are present in the aquatic environment.
Ionic mercury, once present in water, is capable of forming a wide
variety of complexes and chelates with organic materials. Of
considerable importance is its reaction with the sulfide (S--) ion to
form highly insoluble mercury(II) sulfide. This reaction is likely to
occur in anaerobic aquatic environments owing to the presence of
hydrogen sulfide gas. This sulfide complex of mercury is highly stable
and will not normally become involved in transformation under
anaerobic conditions. However, in the presence of oxygen, the
insoluble mercury(II) sulfide can become oxidized to the soluble
sulfite and sulfate salts of mercury which allow the metal to ionize
and enter subsequent chemical reactions.
In addition to the oxidation of metallic vapour, inorganic mercury
(Hg++) can be formed by the breakdown of a variety of organic
mercury compounds. The alkoxyalkylmercury compounds are very unstable
in acid conditions and it has been reported (see Jensen & Jernelov,
1972) that, in humid soil (pH = 5), methoxyethylmercury has a
half-life of only 3 days. Aryl- and alkylmercury compounds can all be
degraded in the environment by chemical and physical processes and by
biologically mediated processes.
Divalent inorganic mercury (Hg++) can undergo two important
reactions in the environment. The first is the reduction to metallic
mercury vapour, a reaction that will occur in nature under appropriate
reducing conditions. As mentioned above, certain bacteria,
particularly of the genus Pseudomonas, can convert divalent mercury
into metallic mercury (Magos et al., 1964; Furukawa et aI., 1969). The
formation of inorganic divalent mercury in nature and its reduction to
metallic mercury vapour are probably key processes in the global cycle
of mercury. The reduction to metallic mercury vapour must be the key
step in the release of mercury because of degassing of the earth's
surface. The oxidation of metallic mercury vapour to divalent ionic
mercury must be the critical step in the uptake of mercury vapour in
rainwater and in the oceans. Unfortunately, other than these crude
generalizations, little is known of the details of the kinetics of
these processes in nature.
The second important reaction that ionic divalent mercury (Hg++)
undergoes in nature is its conversion to methylmercury and
dimethylmercury compounds and the interconversions between these
compounds. These reactions play a critical role in the so called
"local cycle" of mercury and are worth further discussion. Some
countries, particularly those in Scandinavia, that used methylmercury
fungicides extensively, experienced a general rise in the mercury
content of their agricultural products. High levels were also noted in
some species of birds. The increase corresponded with the onset of the
use of methylmercury fungicides. However, it was discovered that
mercury levels in fish were also high and that these fish were
obtained in areas where methylmercury compounds were not used (Jensen
& Jernelov, 1969). It was subsequently discovered that methylmercury
was the predominant form of mercury in fish regardless of the nature
of the mercury pollutant. This was the first evidence that
transformations of mercury compounds must occur in the environment and
that, indeed, they must be of great significance. It has now been
demonstrated that biological methylation of mercury occurs in the
organic sediments of aquaria and in sediments from freshwater and
coastal waters of Sweden (Jensen & Jernelov, 1967, 1969; Jernelov,
Two biochemical pathways of methylation of mercury have been
identified, one anaerobic the other aerobic. The anaerobic pathway
involves the methylation of inorganic mercury by methylcobalamine
compounds produced by methanogenic bacteria in a mildly reducing
environment (Wood et al., 1968). The process is non-enzymic and is
strictly anaerobic. The aerobic pathway has been described by Landnet
(1971) in studies of Neurospora crassa. His findings indicate that
methylmercury bound to homocysteine becomes methylated by those
processes in the cell normally responsible for the formation of
methionine. In other words, the methylmercury-homocysteine complex is
methylated by "mistake".
Despite the fact that an anaerobic pathway for methylmercury
production is well known, it seems unlikely that significant amounts
of methylmercury are formed in the aquatic environment under anaerobic
conditions. The chief reason for this, as pointed out by Jensen &
Jernelov (1972), is that, in natural water when oxygen is exhausted,
hydrogen sulfide is formed and divalent mercury becomes bound up as
mercury(II) sulfide. In this sulfide form, mercury is not available
for methylation under anaerobic conditions (Jernelov, 1968; Rissanen,
quoted by Jensen & Jernelov, 1972), and methylation is slow even under
aerobic conditions (Fagerstrom & Jernelov, 1971).
In an aquatic environment under aerobic conditions, it must be
borne in mind that the upper sedimentary layers and sedimentary
particles suspended in the water may be both aerobic and anaerobic,
the exterior being well oxygenated and the interior deficient in
oxygen. Thus both pathways, aerobic and anaerobic, are possible routes
of methylation in water that is oxygenated.
The ability to methylate mercury is not confined to a limited
number of species of microorganism. Thus, conditions that promote
bacterial growth in general, will lead to enhanced methylation of
mercury. The highest rates of methylation in the aquatic environment
are, therefore, seen in the uppermost part of the organic sediments
and on suspended organic material in water (Jernelov, 1973).
The formation of dimethylmercury from monomethylmercury compounds
has been shown to occur in decomposing fish (Jensen & Jernelov, 1968),
and from (originally) inorganic mercury in sediments. The anaerobic
pathway using methylcobalamines is one means by which dimethylmercury
can be synthesized. The reaction is greatly favoured by high pH
whereas the formation of monomethylmercury is favoured by a low pH
The ability to methylate mercury at a high rate correlates with
the resistance of the microorganism to concentrations of inorganic
mercury (for review, see Jernelov, 1973).
The observations, reviewed above, of the interconversion of the
various mercury compounds in nature have led to a hypothesis for a
local cycle (Jensen & Jernelov, 1972). Inorganic divalent mercury is
formed either by the oxidation of metallic mercury vapour by
physico-chemical processes or by the cleavage of the carbon-mercury
bond in organomercurial compounds either chemically or enzymatically.
The divalent ionic mercury becomes attached to sediments either
suspended in the water or in the sedimentary layers. The upper
sedimentary layers are biologically active but it is postulated that,
with the passage of time, large quantities of inorganic mercury will
penetrate down to the inorganic mineral layers of the sediments where
the mercury should remain inactive. In the surface layers of the
sediment, part of the inorganic mercury becomes methylated.
Methylation significantly increases the ability of mercury to cross
biological membranes. This is why aquatic organisms contain mainly
If conditions of pH are appropriate, dimethylmercury will be
formed. Dimethylmercury is water insoluble, possesses a very high
volatility, and is postulated to diffuse from the aquatic environment
into the atmosphere. Once in the atmosphere, it is subject to removal
by rainfall. If the rainwater is acidic, the dimethylmercury is
converted to monomethylmercury compounds and is thereby returned to
the aquatic environment completing the cycle. In the presence of
mercury(II), dimethylmercury is converted to two methylmercury
molecules (Jensen & Jernelov, 1969).
Key parts of this local cycle remain conjectural. It is known that
dimethylmercury compounds can be formed and that the conditions for
their formation can exist in an aquatic environment. Unfortunately
analytical data are sparse but Bramen & Johnson (1974) have identified
both mono-and dimethyl compounds in the atmosphere both outdoors and
indoors in the USA. Evidence is still lacking for methylmercury
compounds in rainwater. The analytical difficulties are considerable.
Nevertheless the present weight of evidence supports the existence of
a local cycle for the transport of mercury involving dimethylmercury
as the key intermediary for the atmospheric turnover in this cycle.
The observations available today on this cycle refer to local bodies
of water such as lakes and rivers and the cycle itself would represent
the best available explanation for the presence of methylmercury
compounds in freshwater fish.
The origin of methylmercury compounds in oceanic fish has not been
well described. Inorganic mercury is available in unlimited quantities
in the oceans, as has been indicated in the calculations reported in
section 4.1. The site of methylation of this mercury is not known.
Sediment suspended in oceanic water would seem to be a prime suspect.
Methylation of mercury is also known to occur in the slime covering
fish but it does not occur in the fish tissues themselves (Jensen &
Jernelov, 1972). It would seem an important research priority to
describe the methylation pathways in ocean waters. Only then will it
be possible to state whether the rate of formation of methylmercury in
ocean waters and uptake in oceanic fish is related to the total
deposit of mercury in the oceans (70 million tonnes) or whether it is
related to a very small sub-fraction of the mercury in the oceans that
may respond to man's activities more dramatically than the total ocean
4.3 Interaction with Physical or Chemical Factors
The interaction of mercury with physical or chemical factors has
been referred to frequently in the previous section, so that only a
brief summary will be given here. In terms of the global distribution
of mercury, such physicochemical factors as temperature, pH, redox
potential, and chemical affinities for the organic materials in soil
will interact to determine the degree of volatility of mercury under
specific local conditions and the rate of release of mercury from the
earth's crust as elemental mercury vapour. The interplay between these
factors is so complex that studies of mercury volatilization from soil
and from the earth's crust, in general, do not lend themselves easily
to experimental work. Once in the atmosphere, metallic mercury is
liable to both physical and chemical interactions. Physically it may
be adsorbed on to particulate materials in air but evidence reviewed
in section 5.1 indicates that the aerosol fraction of mercury is 5% or
less of the total mercury in air. Metallic mercury vapour should
distribute more or less evenly between air and water providing it
remains in the unoxidized metallic state (Hughes, 1957). However, the
reported levels in rainwater (see section 5.1) are higher than the
background level by a factor of at least 2 or 3. This is no doubt a
consequence of the oxidation of metallic mercury to ionic mercury in
the water in the presence of oxygen. Once deposited in the ocean from
rainwater, any remaining metallic mercury should be liable to
oxidation to ionic mercury whereupon it will undergo rapid chemical
combination with various chemical compounds in ocean water. Sillen
(1963) has estimated that the mercury may be present as negative
chloride complexes (section 5.1). However, it seems probable that,
because of its affinity for sulfhydryl groups, mercury will also bind
strongly to living organisms in ocean waters.
Another aspect that should be considered is the relationship
between mercury and selenium. Recent data indicate that selenium
compounds known to detoxify mercury, increase mercury retention in
some organisms changing the tissue distribution (Parizek et al.,
1971). High mercury concentrations were accompanied by high selenium
concentrations in tissues of several animal species (Ganther et al.,
1972; Koeman et al., 1972, 1973) and also in man (Kosta et al., 1975;
Byrne & Kosta, 1974). This relationship is further discussed in
section 7 of this document.
In the local cycle of mercury, the same physico-chemical factors
will be operative. Oxygen tension in the aquatic environment will
determine the degree of formation of insoluble mercury(II) sulfide
that will limit the rate of methylation. The pH of the aquatic
environment and also of the rainwater will determine the distribution
of the methylated forms of mercury between dimethyl and monomethyl
The short-chain alkylmercurials, especially methylmercury
compounds, have a strong tendency to bioaccumulation since they
possess a group of properties that makes them unique among the mercury
compounds. Methylmercury is very efficiently absorbed through
biological membranes. In mammals, absorption of methylmercury from
food is virtually complete. Methylmercury is degraded much more slowly
into inorganic mercury than are the other classes of organomercurial
compounds. It is excreted from living organisms much more slowly than
other mercury compounds. It possesses a very high chemical affinity
for the sulfhydryl group. Since this group occurs mainly in proteins
in living organisms, methylmercury, once it has entered the organism,
is soon convened to a non-diffusible protein-bound form. However, even
though most of the methylmercury is bound to protein, a small fraction
remains in a diffusible form. Methylmercury rapidly equilibrates
between diffusible and non-diffusible binding sites and thus retains
its mobility within animal tissues.
In view of its ability to accumulate in living organisms, one
would, in general, expect to see higher concentrations of
methylmercury at higher trophic levels in natural food chains.
Qualitatively, this generalization appears to be true but quantitative
predictions are not possible because of the complex interplay of a
host of factors that influence the accumulation and movement of
mercury in food chains. For example, remarkably large species
differences exist in biological half-times which vary from
approximately 7 days in the mouse, to 70 days in the monkey and man,
500 days in seals, and over 1000 days in some species of fish (for
review, see Clarkson, 1972a).
The origin of methylmercury in terrestrial food chains is
predominantly the use of mercury fungicides in the treatment of seed
grain (D'Itri, 1972). The seeds are consumed by grain-eating birds or
rodents and the rodents themselves become victims of the large
carnivorous birds. The dramatic increase in the concentration of
mercury in feathers of carnivorous birds in Sweden was associated with
the introduction of methylmercury fungicides in 1940 (for review, see
Swedish Expert Group, 1971). High concentrations of mercury in
pheasants and other game birds are also a result of this terrestrial
food chain and have led to restrictions on hunting in certain areas of
North America. The replacement of methylmercury by the
alkoxyalkylmercury compounds in Sweden led to a diminished level in
this terrestrial food chain. Generally speaking, alkoxyalkyl- and
phenylmercury compounds are either less well absorbed or more easily
degraded to inorganic mercury and more rapidly excreted.
The accumulation of methylmercury compounds in aquatic food chains
has been the subject of a recent review (Fagerstrom & Larsson,
unpublished report). This chain or group of chains is considerably
more complex than the terrestrial ones. Nevertheless, several
tentative generalizations seem plausible at this time. Once
methylmercury is formed in the upper sedimentary layers or in
suspended sediments in water, it readily leaves the sedimentary
particle (Gavis & Ferguson, 1973). The reason for this is not fully
established but Fagerstrom & Larsson suggest that it may be due to the
pathway of synthesis of methylmercury compounds. For example, if
methylmercury is formed by the pathway proposed by Landner (1971), it
will be in the form of a diffusible complex with homo-cysteine. In
contrast, inorganic mercury in the sediment is probably bound to large
macromolecules. Once methylmercury has diffused from the sedimentary
particle into the water, it must be rapidly accumulated by living
organisms. This accumulation is so efficient that methylmercury has
never been detected in filtered water. Fagerstrom & Larsson, in
reviewing recent experimental work on methylmercury accumulation,
noted that this form of mercury accumulates in all species, whether
plant or animal, that possess membranes for gas exchange with their
The accumulation of methylmercury in food chains in freshwater
systems has been proposed as a three-step process by Fagerstrom &
Larsson. The first step is an accumulation by bottom fauna that are in
closest proximity to the active sedimentary layers where the
methylmercury is formed. Accumulation in the bottom fauna, including
plankton, would be followed by accumulation in species such as the
roach and finally in the large carnivorous fish such as the northern
pike. The authors point out that the relative importance of uptake of
methylmercury directly from water through the gill membranes, as
opposed to intake from food, should depend upon the trophic level of
the fish. The higher the trophic level the more important the intake
from food. However, for the overall food chain, uptake through the
gills is the key process. If for some reason there is a dramatic
change in the environmental layers of methylmercury, the authors
predict that it would take from 10-15 years for the levels in the top
predators to readjust to the new environment.
These generalizations on freshwater species should be expected to
apply to oceanic fish. The remarkably high levels of methylmercury
seen in swordfish and tuna fish are due to a variety of factors. First
these species are large carnivorous fish at the end of a food chain.
They live for a relatively long time compared with other species of
fish and it is well established that methylmercury levels show a
positive correlation with age (and or weight) of the fish. They are
highly active fish having insatiable appetites. Because of their
activity, large quantities of oceanic water pass through the gill
membranes each day. Thus it is possible that tuna fish, swordfish and
related species have a high intake of methylmercury both from their
food supply and from the surrounding water.
Accumulation of mercury in the terrestrial and aquatic food chains
(Fagerstrom & Larsson) results in risks for man mainly through the
consumption of: game birds in areas where methylmercury fungicides are
in use; fish from contaminated waters, especially predator species,
tuna fish, swordfish and other large oceanic fish even if caught
considerably off shore; other seafoods including muscles and crayfish;
fish-eating birds and mammals; and eggs of fish-eating birds.
Space does not permit a full discussion of the important questions
concerning the chain of mercury transport from soil to plant to
domestic animals and ultimately to man. Important parameters in this
transport include absorption and availability in the soil, intake and
distribution in the plant, toxic effects on the plant, and intake by
domestic animals and by man. The maximum amounts tolerated in the soil
may be key factors in determining the possible enrichment in food
chains and the ultimate hazards to man (Koronowski, 1973; Kloka,
5. ENVIRONMENTAL LEVELS AND EXPOSURES
The levels of mercury in the environment have been reviewed either
partially or completely by: Swedish Expert Group (1971), Joint FAO/WHO
Expert Committee on Food Additives (1972), Holden (1972), D'Itri
(1972), Petersen et al. (1973), Bouquiaux (1974), and CEC (1974). The
principal findings may be summarized as follows. The concentration of
mercury vapour in the atmosphere is so low that it does not contribute
significantly to human intake of mercury. A few "hot spots" may exist
but these require further investigation. Concentrations of mercury in
water, particularly drinking water, are also sufficiently low as not
to contribute significantly to human exposure. The industrial release
of methylmercury compounds into a sheltered ocean bay (Minamata Bay)
and into a river (the Agano River) in Japan have led to extremely high
concentrations of methylmercury in fish (up to 20 000 µg/kg wet
weight) and resulted in human poisonings and fatalities. The
industrial release of a variety of chemical and physical forms of
mercury into inland waters has led to local pollution, to mercury
levels in fish occasionally over 10 000 µg/kg but usually less than
5000 µg/kg, and to the restriction of fishing for sport and commercial
fishing in these areas. The mercury level in most freshwater and
oceanic fish is below 200 µg/kg. However, in large carnivorous fish
such as tuna, swordfish, halibut, and shark, levels are usually above
200 µg/kg and can be as high as 5000 µg/kg wet weight. The general
population face no significant hazards from the consumption of
methylmercury in the diet. However, certain sub-populations, either
those eating locally contaminated fish or those with an unusually high
consumption of large carnivorous oceanic fish eventually develop blood
levels of mercury in the range of the lowest levels associated with
signs and symptoms of poisoning in the Japanese outbreak. It is
estimated that the average daily intake of the general population is
less than 20 µg of mercury per day in the diet. An appreciable amount
of this would be methylmemury. However, individuals in certain
sub-populations having unusually high exposure may ingest daily
amounts of mercury of up to 200 µg, mainly as methylmercury compounds.
5.1 Levels in Air, Water, and Food
The average concentration of mercury in the general atmosphere was
reported by Stock & Cucuel (1934) to be 20 ng/m3. These results were
confirmed by Eriksson (1967) in Sweden. Sergeev (1967) noted
concentrations of 10 ng/m3 in the USSR. Fujimura (1964) reported
concentrations of 0-14 ng/m3 in non-industrialized regions of Japan.
The lowest reported levels are those reported by McCarthy (1968) in
Denver, USA, of 2-5 ng/m3. Williston (1968) reported mercury levels
in the vicinity of San Francisco, USA, of 0.5-50 ng/m3, the level
depending greatly on the direction of the wind. Williston's method
would have detected only mercury vapour.
Levels of particle-bound mercury have also been reported.
Goldwater (1964) noted that airborne dust in New York City contained
from 1 to 41 ng/m3 and that outdoors the concentration was from 0 to
14 ng/m3. Brar et al. (1969) noted that particle-bound mercury in air
above Chicago ranged from 3 to 39 ng/m3. Heindryckx et al. (1974), in
the most recent study, found that aerosol mercury levels corresponding
to remote background levels in Norway and Switzerland were as low as
0.02 ng/m3. In a heavily industrialized area of Belgium, near Liege,
the aerosol mercury levels noted were as high as 7.9 ng/m3. Other
sampling stations in Belgium reported values roughly an order of
magnitude below this. Unfortunately it is not known to what extent
particle-bound mercury contributes to total mercury levels in the
atmosphere. An indirect reference to Jervis by Heindryckx et al.
(1974) indicates that aerosol mercury accounts for only 5% of total
mercury in the atmosphere. All the particle-bound mercury reported by
Heindryckx et al. (1974) had a particle size of less than 0.4 µm.
"Hot spots" of mercury concentration have been reported in
atmospheres close to industrial emissions or above areas where mercury
fungicides have been used extensively. Fujimura (1964) reported air
levels up to 10 000 ng/m3 near rice fields where mercury fungicides
had been used and values of up to 18 000 ng/m3 near a busy super
highway in Japan. McCarthy et al. (1970) noted air values of up to 600
and 1500 ng/m3 near mercury mines and refineries. Fernandez et al.
(1966) reported maximum values of 800 000 ng/m3 in a village close to
a large mercury mine in Spain. The remarkably high mercury vapour
levels reported by these authors indicate the need for further studies
into localized high concentrations of mercury in the atmosphere.
Limited data are available for concentrations of mercury in
rainwater and snow. First reported values were 50-500 ng/litre (Stock
& Cucuel, 1934). Eriksson (1967) found values from 0 to 200 ng/litre.
Brune (1969) noted values of approximately 300 ng/litre in rainwater
in Sweden. Values for mercury in snow have been reported by Johnels et
al. (1967) as 70 ng/kg and by Byrne & Kosta (quoted by Holden, 1972)
as 1000-3000 ng/kg in centrifuged melted snow. It is probable that
mercury levels in snow depend greatly on the collection conditions and
upon how long the snow has laid on the ground. For example, Strabya
noted values of 80 ng/kg in fresh snow but 400-500 ng/kg in snow that
a STRABY, A. (1968) Analysis of snow and water. In: Westermark, T.
& Ljunggeren, K., ed. Development of analytical methods for
mercury and studies of its dissemination from industrial sources.
Stockholm, Swedish Technical Research Council, mimeographed
may have partly melted or evaporated over the winter. Analysis of ice
deposited in Greenland prior to the 1900s (Weiss et al., 1971)
indicates values of 60 ng/kg.
Bodies of freshwater for which there is not independent evidence
for mercury contamination, contain levels of mercury of less than
200 ng/litre. Stock & Cucuel (1934) reported 10-50 ng/litre in
well-water and 100 ng/litre in the River Rhine. Dall'Aglio (1968) in
measurements of 300 samples from natural water in Italy found values
in the range of 10-50 ng/litre. Voege (1971) reported levels up to
40 ng/litre for uncontaminated Canadian waters. Durum et al. (1971)
have reported data on the concentration of mercury in surface waters
of the USA. In areas where mercury mineralization was present, values
of up to 200 ng/litre were seen. The results of the CEC International
Symposium, reviewed by Bouquiaux (1974), indicate that the purest
surface water (drinking quality) contains less than 30 ng/litre based
on over 700 samples collected from drinking reservoirs in the Federal
Republic of Germany. Rivers believed to have low contamination, such
as the Danube, and bodies of water such as the Boden See, have values
close to 150 ng/litre based on the analysis of 152 samples. The rivers
in the lowland countries of Western Europe that flow into the North
Sea have mercury values in the range of 400-700 ng/litre no doubt
reflecting the high industrialization of this area (Schramel et al.,
1973). Reports by Hasselrota, Fonds (1971), and Smith et al. (1971a),
indicate that mercury is predominantly particle-bound in contaminated
water-ways. In the Federal Republic of Germany the mercury
concentration measured was around 400 ng/litre in inland waters,
between 100 ng and 1800 ng/litre in rivers, and 600 ng/litre in a
sample of potable water. (Reichert, 1973; Schramel et al., 1973.)
Data for mercury concentrations in ocean waters are not as
extensive as those reported for freshwater. Findings of Stock & Cucuel
(1934) giving a mean value of 30 ng/litre were confirmed by Sillen
(1963). Sillen, on the basis of physico-chemical arguments, suggested
that most of the mercury in seawater would be present as negatively
charged halide complexes. Hosohara (1961) noted the following levels
in the Pacific Ocean: at the surface, 80-150 ng/litre; at a depth of
500 metres, 60-240 ng/litre and at a depth of 3000 metres,
150-270 ng/litre. Levels reported at the CEC International Symposium
(reviewed by Bouquiaux, 1974) were 20 ng/litre in 14 samples from the
a HASSELROT, T. (1971) Mercury in fish, water, and bottomless
sediments. Investigations at the research laboratories of the
National Swedish Environment Protection Board (mimeographed
English Channel but were as high as 150 ng/litre in samples taken from
the Belgian shoreline and the Waddenzee in the Netherlands. Other
references such as Burton & Leatherland (1971) and Leatherland et al.
(1971) also support the general rule that oceanic levels are below
300 ng/litre. Higher concentrations have been produced as a result of
local contamination such as in Minamata Bay where Hosohara et al.
(1961) have reported values up to 600 ng/litre.
In view of questions, discussed earlier, on the total mercury
content of the ocean, the stability of mercury levels in the ocean
over the past 50 years, and on the high mercury levels in species of
oceanic fish, the paucity of data on oceanic levels of mercury is
remarkable. This would seem to be one area for future studies of
environmental levels of mercury. These efforts should include attempts
to analyse the different physical (particulate, or soluble) and
chemical (inorganic, or methyl) forms of mercury.
Food (except fish)
Smart (1968) has reviewed data concerning mercury concentrations
in foods and the most recent data from Europe have been summarized by
Bouquiaux (1974). Mercury levels in milk products (81 samples from the
Federal Republic of Germany and the United Kingdom) ranged from 0 to
40 µg/kg with a median value of 6 µg/kg. Levels in eggs (440 samples,
taken from Denmark, the Federal Republic of Germany and the United
Kingdom, ranged from 0 to 100 µg/kg with most of the values between 10
and 20 µg/kg. Levels in meat, meat products, and prepared meat
products (318 samples from the United Kingdom) ranged from 0 to
50 µg/kg with most values lying between 10 and 20 µg/kg. Various kinds
of cereal and flour (2133 samples, taken from the Federal Republic of
Germany and the United Kingdom) ranged from 0 to 20 µg/kg with most
values being close to 3 µg/kg. Mercury levels in cereal products from
the same countries (52 samples) ranged up to 50 µg/kg with most values
close to 20 µg/kg. Vegetables and fruits (288 samples) from Belgium,
the Federal Republic of Germany, and the United Kingdom had mercury
levels up to 50 µg/kg with most values close to 7 µg/kg. The analysis
of nearly 1400 foods, excluding fish, in Canada during 1970 showed
mercury residues to be less than 60 µg/ kg in bread, flour, grains,
and eggs and less than 40 µg/kg in meats and vegetables (Somers,
A Swedish Expert Group (1971) has reviewed Swedish experience on
the effects of widespread use of methylmercury fungicides on food
levels of mercury. As a result of a ban on the use of methylmercury
fungicides, food levels fell by a factor of three. For example, the
mercury levels in Swedish hen eggs (whole) averaged 29 µg/kg prior to
April 1966. Between October 1967 and September 1969, following the ban
on methylmercury fungicides instituted in 1966, the level in Swedish
hen eggs fell to 9 µg/kg.
The chemical form of mercury in foodstuffs other than fish has not
been well identified. The reason is that the levels are, in general,
so low as to preclude gas chromatographic identification. However,
Westöö (quoted by a Swedish Expert Group, 1971) has noted that
methylmercury accounts for over half the total mercury in samples of
pork chop and liver, filet of beef, and egg white. Inorganic mercury
can account for more than half the total mercury in pig kidney, pig
brain, ox liver, and egg yolk.
The earliest reported mercury levels for freshwater fish are those
of Stock & Cucuel (1934) and Raeder & Snekvik (1949) and range from 30
to 180 µg/kg wet weight. Upper limits for mercury levels have been
quoted as, 200 µg/kg wet weight (Lofroth, 1969) for Sweden, 150 µg/kg
(Sprague & Carson, 1970) for Canada, and 100 µg/kg (Ui, 1967) for
Japan. These are probably to be regarded as normal levels, i.e. for
fish in uncontaminated water. The WHO Regional Office for Europe
(1973) has summarized references indicating that fish caught in
contaminated freshwater areas may have values of 200-5000 µg/kg and,
where the water is heavily polluted, values may be as high as
20 000 µg/kg.
The CEC International Symposium (Bouquiaux, 1974) quote levels in
freshwater fish caught in Western Europe as ranging from 0 to
1000 µg/kg with most values being between 200 and 400 µg/kg wet
weight. Canned fish, excluding tuna taken from several Western
European countries (597 samples), had values up to 500 µg/kg with an
average close to 50 µg/kg wet weight. Canned tuna from the same areas
(1798 samples) had values ranging up to 4000 µg/kg with most values
falling into the range of 200-500 µg/kg. Salmon appears to have
remarkably low levels of mercury. Measurements of some 260 samples of
Atlantic Ocean, Canadian, and Baltic Sea salmon had mercury levels
ranging up to 150 µg/kg with most values being close to 50 µg/kg. On
the other hand, pike caught in contaminated rivers in Denmark had
average mercury values of 5000 µg/kg, results which are in agreement
with experiences summarized by a Swedish Expert Group (1971) in
contaminated freshwater areas in Sweden and Finland. The concentration
of mercury in marine fish showed marked variations. Not all the
factors responsible for these variations are understood but it is
generally realized that the species of fish, the geographical
location, and the age and/or weight of the fish are important. The
highest values of mercury are usually seen in those fish at the end of
a long food-chain such as the large carnivorous species.
The concentration of mercury in marine fish has been the subject
of intense study in recent years. The first measurements reported by
Stock & Cucuel (1934) and Raeder & Snekvik (1941) are in agreement
indicating levels from 44 to 150 µg/kg wet weight. The most recent
reports (Peterson et al., 1973; Bouquiaux, 1974) indicate that mercury
levels in most species of oceanic fish fall in the range of
0-500 µg/kg wet weight with most values close to 150 µg/kg wet weight
(more than 1600 samples). The most important exceptions to this rule
are swordfish, tuna fish, and halibut, whose values usually range from
200 to 1500 µg/kg (reviewed by the Joint FAO/WHO Expert Committee on
Food Additives, 1972). Skipjack, white tuna, and yellow fin tuna (911
samples) ranged from 0 to 1000 µg/kg with most values ranging from 200
to 300 µg/kg. These samples were caught in the Atlantic, Pacific, and
Indian Oceans. Bluefin tuna from the Bay of Biscay (285 samples)
ranged from 200 to 800 µg/kg with most values close to 500 µg/kg. The
same species caught in the Mediterranean Sea (136 samples) ranged from
500 to 2500 µg/kg with most values close to 1100 µg/kg. Big-eye tuna
(20 samples from various origins) had mercury values ranging from 400
to 1000 µg/kg. Over 5200 samples of tuna, variety not specified but
originating from Italy, had levels in the range of 0-1750 µg/kg with
most values ranging from 300 to 500 µg/kg wet weight.
Swordfish caught in the western Atlantic (210 samples) had mercury
values ranging from 50 to 4900 µg/kg with a mean value of 1150 µg/kg.
40 samples of swordfish, originating near Italy, had values ranging
from 650 to 1750 µg/kg with most values close to 1100 µg/kg wet
The geographical location appears to be important. This is
illustrated by mercury analysis of cod (Dalgaard-Mikkelsen, 1969).
Samples recovered from the strait between Denmark and Sweden, which is
heavily contaminated, had values up to 1290 µg/kg; cod caught in the
area of Greenland had values of 12-36 µg/kg whereas North Sea cod had
values in the range of 150-195 µg/kg wet weight. Peterson et al.
(1973) quote evidence that halibut caught in the southern areas of the
Northern Pacific had higher mercury levels than those caught in the
North. Beckett & Freeman (quoted by Peterson et al., 1973) in a study
of 210 swordfish from six areas extending from the Caribbean Sea to
the Grand Banks noted significant variations from one area to another
in average mercury levels.
Metabolic differences may also affect mercury levels. For example,
Barber et al. (1972) noted differences in mercury content in different
species of benthopelagic fish despite the fact that they had identical
feeding habits and ecological requirements and were exposed to mercury
in the same area for the same length of time.
The age (or weight) of the fish appears to be an important
determinant of mercury levels. A positive correlation between mercury
concentrations and the weight of the fish has been demonstrated by
Beckett & Freeman (quoted by Peterson et al., 1973) for swordfish,
halibut, benthopelagic morid (Barber et al., 1972), spiny dogfish
(Forrester et al., 1972), blue marlin (Rivers et al., 1972), and tuna
(quoted by Peterson et al., 1973). In the last study, mercury levels
were measured in 88 yellowfin tuna whose sizes ranged up to 100 kg.
Tuna having weights below 25 kg had mercury levels not exceeding
250 µg/kg; tuna having body weights below 50 kg had mercury levels not
exceeding 500 µg/kg. Tuna with body weights above 60 kg had values
ranging up to 1000 µg/kg. However, large variations in mercury content
were noted in tuna with body weights in the range of 60-100 kg. A
relationship between mercury content and body weight has previously
been noted for freshwater fish (Johnels, 1967; Kleinart, 1972; Bache
et al., 1971).
Mercury content may also differ with the sex of the fish. For
example, Forrester et al. (1972), in studies of spiny dogfish on the
coast of British Columbia, noted that males had a higher mercury
content than females for a given body weight. These authors suggested
that this difference may be due to the fact that the males grow more
slowly than the females.
Mercury in fish appears to be predominantly in the form of
methylmercury. Swedish measurements of freshwater fish, summarized by
a Swedish Expert Group (1971), indicated that virtually all of the
mercury is present in the form of methylmercury compounds. Smith et
al. (1971b) confirmed these findings for fish on the North American
continent and for swordfish and tuna fish. Exceptions to this rule are
Pacific marlin caught off the coast of Hawaii where methylmercury
accounts for only a small fraction of the total mercury (Rivers et
al., 1972) and also lake trout where methylmercury seems to account
for only 21-35% of total mercury (Bache et al., 1971).
Interpretation of the results of observations on museum specimens
of tuna fish and swordfish caught at the turn of the century (Miller
et al., 1972) indicates that mercury levels in these species of fish
have not changed significantly throughout the twentieth century.
Specimens from preserved fish of this age are necessarily limited. In
seven samples of tuna reported by Miller et al. (1972), the mercury
concentrations ranged from 180 to 640 µg/kg. These compare with
present values in tuna ranging roughly from 200 µg/kg to over
1000 µg/kg wet weight. Given this variation, it is true to say that
there is no statistically significant difference between samples
caught in 1900 and those caught in 1970. However, because of the wide
range of values, the data at present available do not preclude the
possibility that some change may have taken place and that the change
might be quite substantial.
5.2 Occupational Exposures (See also section 8.1.1)
Occupational exposure to elemental mercury vapour is still the
principal hazard to human health when mercury is considered. More than
50 specific occupations or trades involving frequent exposure to
mercury have been described by Gafafer (1966). Diseases caused by
mercury or its toxic compounds are classical occupational diseases and
in most countries are notifiable and qualify for compensation.
Reporting of occupational poisoning by mercury has been inadequate, as
is the case with all other occupational diseases, particularly in
developing countries where there is evidence that large numbers of
workers are exposed to high concentrations of mercury leading to
poisoning. Occurrence of occupational mercury poisoning in a wide
variety of industries in different parts of the world has been
reported. In accordance with the information available, most people
exposed to elemental mercury vapour appear to be employed in the
mining industry, or in chloralkali plants (McGill et al., 1964; Ladd
et al., 1966; West & Lim, 1968; Smith et al., 1970) and in the
manufacturing of instruments where mercury finds application. These
publications, all appearing within the last ten years, indicate that
mercury levels in air may attain values as high as 5 mg/m3. The
highest mercury concentrations in air are reported in papers on
exposure in mining operations. The concentration of mercury in urine
may attain levels as high as 2175 µg/litre.
In mining for metals other than mercury (e.g. copper), mercury ore
may be present in the mine and give rise to occupational exposure.
Donovan (1974) has reported levels of mercury in urine samples (91
samples, number of workers not stated) ranging from 30 to 700 µg/litre
in a non-mercury related mining operation. In the two years (1972-73),
seven urine samples were found with mercury levels in excess of
250 µg/litre and some of the miners were admitted to hospital.
Ladd et al. (1964) have reported on occupational exposure to
phenylmercury compounds. Air mercury concentrations ranged up to
0.1 mg/m3 and urinary mercury levels ranged from 1 to 788 µg/litre. A
total of 67 workers were involved in these studies. Phenylmercury
compounds continue to be used as fungicides in the paint industry (for
review, see Goldwater, 1973) so that occupational exposure to
phenylmercury compounds is still significant.
The Swedish Expert Group (1971) have summarized reports on
occupational exposures to methyl- and ethylmercury compounds. All
these reports were published within the period 1940-60 except for the
reports on laboratory personnel published by Edwards in 1865 and 1866.
Restrictions on the agricultural application of ethyl- and
methylmercury compounds by various industrialized countries probably
accounts for the lack of recent reports on occupational exposure.
5.3 Estimate of Effective Human Exposure
The daily intake of elemental mercury vapour by the general
population may be calculated from the published data on ambient air
levels discussed above and on the assumption that 80% of the inhaled
mercury vapour is retained and that the daily ventilation in the
average person is 20 m3 of air. The ambient air level, except in
polluted areas, appears to be of the order of 20 ng/m3 and appears
not to exceed 50 ng/m3 (see section 5.1). Assuming an ambient air
level of 50 ng/m3, the average daily intake of metallic mercury
vapour would amount to 1 µg/day due to inhalation. The average daily
intake of those sub-groups of the general population living in
specially polluted areas is difficult to estimate with any accuracy.
If we use the figures of McCarthy (1970), it is possible to find
mercury levels as high as 0.0015 mg/m3 close to points of emission.
Individuals living continuously in these areas would have intakes of
30 µg/day. Daily intake from occupational exposure is almost
impossible to estimate because of the wide variation in exposure
conditions in industry (see section 5.2). Assuming that, generally,
the time-weighted average threshold limit value of 0.05 mg/m3 (ACGOH,
1976) is being followed, average occupational exposure would lead to
an average daily intake of 300 µg of mercury or less, assuming a
ventilation of 10 m3/day at work and 225 working days per year. The
published reports are insufficient to estimate occupational daily
intake from other forms of mercury. The proposed guideline of
0.1 mg/m3 for phenylmercury (MAC Committee, 1969) should lead to an
intake in workers exposed to phenylmercury compounds of 500 µg/day or
The intake of mercury from drinking water by the general
population is more difficult to estimate but it is probably very low
in comparison with intake from diet. The major problem is that the
chemical form of mercury in water has not always been identified and
the efficiency of absorption from the gastrointestinal tract depends
greatly on the form of mercury. Methylmercury compounds are absorbed
almost completely whereas absorption of inorganic mercury may be 15%
or less. In making the following calculations the worst case will be
assumed, namely that all mercury in drinking water is methylmercury.
it will also be assumed that the daily intake of water in adults is
2 litres/day (Joint FAO/WHO Expert Committee on Food Additives, 1972).
Published reports indicate that pure well-water and drinking water
from reservoirs have mercury levels not exceeding 50 ng/litre (see
section 5.1). Thus the daily intake of mercury from drinking water
would not normally exceed 0.1 µg/day. However, drinking water in
certain areas may derive either from natural waters such as those
reported in Italy that have levels as high as 300 ng/litre because of
exposure to mineralized mercury deposits, or from rivers in heavily
industrialized areas reported to have values up to 700 ng/litre (see
section 5.1). Taking the highest reported figure and assuming that
mercury is not removed during purification of the water, the highest
daily intake would be close to 1.4 µg/day. The advised upper limit for
mercury in drinking water is 1 µg/litre (World Health Organization,
1971) which would allow intakes of up to 2 µg/day from this source.
The intake of mercury from food is the most difficult of all to
estimate because of the different levels of mercury in different
classes of foodstuffs and different dietary habits of individuals in
the general population. The one important generalization that emerges
is that the intake of mercury as methylmercury is related to fish
intake. Thus normal levels for intake of mercury cannot be stated in
general without some reference to the fish intake of the population in
Over the past forty years, various estimates have been made on the
intake of mercury by the general population assuming that fish intake
is close to the average values for that population. These reports have
been reviewed by a Joint FAO/WHO Expert Committee on Food Additives
(1972) and indicate that the range of daily intake of mercury in the
general population is from 1 to 20 µg/day. The most complete reviews
of dietary intake published to date are those of a Swedish Expert
Group (1971) and Jonsson et al. (1972). The reports refer specifically
to the Swedish population. It was noted that the intake of mercury in
the diet from sources other than fish in Sweden is about 5 µg/day and
that the methylmercury content is not known precisely. The median
supply of methylmercury from fish is stated to be 5 µg/day or less. As
fish consumption exceeds the median value for Sweden, the daily intake
of methylmercury will increase in proportion. It was noted that the
average daily intake of fish flesh was 30 g, that 10% of the adult men
might consume between 80 and 100 g and that a few individuals may
consume as much as 500 g/day.
Epidemiological studies summarized by a Swedish Expert Group
(1971) indicate that in fishermen and their families, daily intakes of
methylmercury can rise to values of 200 µg/day and that one individual
had an unusually high intake of 800 µg/day. Another example of a
Swedish fish eater with very heavy methylmercury exposure has now been
published (Skerfving, 1974b).
Dietary intake of mercury in other countries is not as well
documented as that in Sweden. Recent studies reported in a CEC
Symposium (Bouquiaux, 1974) indicate that average dietary intake in
the United Kingdom, based on total diet samples, is less than
20 µg/day. Observations on fish eating groups, such as fishermen based
in American Samoa, indicate that blood mercury levels of up to
20 µg/100 ml can be obtained through fish intake (Clarkson et al.,
1975). Such blood levels would be equivalent to a daily intake of
between 200 and 300 µg/day of methylmercury in fish. McDuffie (1973)
has reported on intakes of mercury in dieters in the United States who
consume substantial amounts of tuna and swordfish. He estimated that
in the 40 dieters, who had the highest daily intake of fish, 25%
consumed 9-16 µg/day, that the second quartile consumed 17-26 µg/day,
the third quartile consumed 27-38 µg/day and that the highest quartile
consumed 40-75 µg/day. On the basis of radio-chemical measurements,
Diehl & Schellenz (1974) estimate the total intake of mercury with
food in the Federal Republic of Germany to be between 57 and 192 µg
per person per week.
Some industrial countries appear to have an average daily intake
of less than 20 µg/day but sub-groups in these countries with
unusually high fish intakes (dieters, fishermen's families) may have
intakes rising to 75 µg/day (dieters) and even to 800 µg/day (an
extremely heavy fish-eater in Sweden).
In countries depending greatly on fish as the major source of
dietary protein, there is a great need for dietary studies including
the measurement of mercury in the diet of these populations. Initial
studies from a South American country indicate that coastal villages
have populations that are comparable to the Swedish fishermens'
families in terms of daily intake of methylmercury (Turner et al.,
6. METABOLISM OF MERCURY
6.1.1 Uptake by Inhalation
Inhalation is the most important route of uptake for elemental
mercury vapour. From what is known of the general principles governing
pulmonary retention of vapours, the high diffusibility and appreciable
lipid solubility of metallic mercury vapour should ensure a high rate
of absorption in the alveolar regions of the lung (Task Group on Metal
Accumulation, 1973). Calculations made by Nordberg & Skerfving (1972)
indicate that mercury vapour should be distributed between air and
body tissues in the proportion of 20 to 1 in fayour of tissue
deposition. Experiments on animals confirm that the major site of
absorption is alveolar tissue where virtually complete absorption of
the vapour takes place (Magos, 1967; Berlin et al., 1969; Hayes &
Rothstein, 1962). If mercury vapour is completely absorbed across the
alveolar membranes, one would expect that, owing to the physiological
dead space, 80% of the inhaled vapour would be retained. This has been
confirmed by observations in man where retention of the inhaled vapour
was in the range of 75-85%, at mercury concentrations between 50 and
350 µg/m3. (Teisinger & Fierova-Bergerova, 1965; Kudsk, 1965a). The
retention of mercury vapour in man can be reduced by moderate amounts
of alcohol (Kudsk, 1965b). Magos et al. (1973) have shown that the
action of alcohol is due to the inhibition or, oxidation of the vapour
in the red blood cells and other tissues. More recently Magos et al.
(1974) have shown that the herbicide, aminotriazole, has a similar
action to that of alcohol.
No specific data are available on the monoalkylmercury compounds.
However, it is generally believed that absorption is high, of the
order of 80% of the inhaled amount (Task Group on Metal Accumulation,
1973). Ostlund (1969a, 1969b) reported a high retention of inhaled
dimethylmercury in mice. The inorganic and organic compounds of
mercury may also exist in the atmosphere in particulate form (see
section 5). No detailed studies have been reported on pulmonary
retention and clearance of mercury aerosols. In general, one would
expect that aerosols of mercury should follow the general physical
laws governing deposition in the respiratory system.
Particulates with a high probability of deposition in the upper
respiratory tract should be cleared quickly. For particulates
deposited in the lower respiratory tree, longer retention will be
expected, the length of which will depend on solubility, among other
factors (Task Group on Lung Dynamics, 1966). Approximately 45% of a
mercury(II) oxide aerosol having a mean diameter of 0.16 µm was
cleared in less than 24 hours and the remainder cleared with a
half-time of 33 days according to experiments on dogs by Morrow et al.
(1964). Information on pulmonary retention of aerosols of the
organomercurials is lacking. Pulmonary absorption of monoalkylmercury
must be significant to judge from the incidents of poisoning resulting
from occupational exposures to dusts and vapours of the alkylmercury
fungicides. It should be noted that the gastrointestinal route may
include those particulates of mercury compounds that have been cleared
from the lung in the bronchociliary tract.
6.1.2 Uptake by ingestion
The general principles underlying the gastrointestinal absorption
of mercury and its compounds are not clearly understood. Probably the
formation of soluble salts and complexes is a prerequisite for
absorption of metals ingested from food.
Liquid metallic mercury has long been considered to be poorly
absorbed from the gastrointestinal tract. Based on the data of
Bornmann et al. (1970), in animals given gram quantities by mouth,
Friberg & Nordberg (1973) have calculated that less than 0.01% of an
administered dose of metallic mercury was absorbed. Persons who had
accidently ingested several grams of metallic mercury showed increased
blood levels of mercury (Suzuki & Tanaka, 1971).
The efficiency of absorption from food depends greatly upon the
type of mercury compound (Clarkson, 1972a). Studies on mice revealed
that the absorption of inorganic salts of mercury from food was 15% or
less in contrast with 80% or more in the case of phenyl- or
methylmercury compounds. Observations on volunteers given tracer doses
of inorganic mercury revealed that the efficiency of absorption was
the same with both free and protein-bound mercury. The absorption from
food in these volunteers was an average of about 7%, (Rahola et al.,
Aberg et al. (1969) and Miettinen (1973) have reported on the
absorption of radioactive methylmercury compounds in volunteers given
oral doses. The absorption of the administered dose was 95%
irrespective of whether the methylmercury was administered as a salt
dissolved in water or in a protein-bound form. Information on the
absorption in humans of other organic compounds of mercury including
the other short-chain alkylmercurials is not available. As episodes of
accidental poisoning due to ingestion of food contaminated with
ethylmercury compounds have occurred, absorption must be significant.
The Task Group on Metal Accumulation (1973) considered the
possibility that the gastrointestinal absorption of one metal may be
influenced by the presence of another. Studies on animals and animal
tissues (Sahagain et al., 1966, 1967) suggest the possibility that
some interaction may occur between zinc, manganese, cadmium, and
6.1.3 Absorption through skin
Debate has persisted throughout most of the present century about
the importance of skin as a route for entry of metallic mercury into
the body. Early studies on man (Juliusberg, 1901) and animals
(Schamberg et al., 1918), where inhalation of mercury vapour was
prevented, indicated that appreciable skin absorption of metallic
mercury took place. It would appear that metallic mercury can cross
the skin barrier but to what extent is not known.
Studies on experimental animals reveal that inorganic salts of
mercury, principally mercury(II) chloride, may be absorbed in
significant amounts through skin. For example, Friberg et al. (1961)
and Skog & Wahlberg (1964) indicate that 5% of mercury in a 2% water
solution of mercury(II) chloride was absorbed through intact skin of
guinea-pig over a 5-hour period. Such a penetration rate, if
applicable to man, could result in absorption of substantial amounts
of mercury under conditions of high exposure.
Friberg et al. (1961) and Wahlberg (1965) have demonstrated in
guinea-pigs that methylmercury dicyandiamide was absorbed from a water
solution through intact skin, the rate was more or less the same as
that for mercury(II) chloride reported above. No information is
available on animals with respect to ethyl- or other alkylmercury
No quantitative data are available for skin absorption of the
short-chain alkylmercurials in man. People have been poisoned by
administration of methylmercury compounds locally to the skin such as
methyl-mercury thioacetamide (Tsuda et al., 1963; Ukita et al, 1963;
Okinaka et al., 1964; Suzuki & Yoshino, 1969; Suzuki et al., 1970).
The methylmercury compound was absorbed in sufficient amounts to cause
severe poisoning although the possibility of some inhalation exposure
cannot be excluded.
6.2 Distribution in the Organism
Details on the organ distribution of mercury have been recently
reviewed (Clarkson, 1972a; Nordberg & Skerfving, 1972). New
publications since that time have not substantially changed the
general picture. Methylmercury and its homologous short-chain
alkylmercurials, which are much more uniformly distributed throughout
the body than are the other organomercurials, and inhaled elemental
mercury vapour are distinguished from other types of mercury compound
in their ability to cross the blood-brain barrier and placenta
Organ distribution is not only affected by the type of mercury
compound ingested or inhaled but also changes with time after
exposure. For example, the phenylmercurials are subject to rapid
conversion in the body to inorganic mercury so that the distribution
of mercury following administration of these compounds and related
organomercurials approaches that of inorganic mercury with increasing
time after exposure (for details, see Clarkson, 1972b).
The distribution between cells and plasma (the red cell/plasma
ratio) depends upon the form of mercury to which the subject is
exposed. Studies on fish-eating populations reported by Birke et al.
(1972) and on a heavily exposed population in Iraq (Bakir et al. 1973)
indicate that the cell to plasma ratio for methylmercury is
approximately 10, as was found in human volunteers given tracer doses
of radioactive methylmercury (Aberg et al., 1969; Miettinen, 1973).
The red cell to plasma ratio in human volunteers given radioactive
inorganic mercury salts was 0.4 (Miettinen, 1973).
The distribution of mercury between hair and blood tends to follow
a constant ratio in people exposed to methylmercury (Table 1). In
various populations having a broad range of dietary methylmercury
intake from fish, the concentration of total mercury in hair is
proportional to the concentration in whole blood. The ratio of hair to
blood concentration is about 250 as determined by linear regression
analysis. The data in Table 1 are from populations of individuals,
most of whom probably have a steady concentration of methylmercury in
hair and blood. In the Iraq epidemic, hair and blood concentrations
underwent rapid changes. Two cases have been reported in Iraq in which
blood and hair concentrations were measured when both were declining
following cessation of heavy exposure. (Amin-Zaki et al., in press.)
The ratios of hair to blood concentrations were constant and the value
of the ratio was close to 250. However it should be noted that, when
hair and blood concentrations are changing, it is important to choose
the segment of hair for analysis that corresponds to the blood sample.
Depending on the length of hair segment used for analysis and the rate
of growth of hair, there is a delay of about 2-4 weeks between the
time of sampling the blood, and the emergence of the appropriate
segment of hair above the scalp (Amin-Zaki et al., in press).
In people occupationally exposed to metallic mercury vapour, the
red cell to plasma ratio may be as high as 2 (Lundgren et al., 1967;
Suzuki et al., 1970; Einarsson et al., 1974). Work on experimental
animals has shown that the ratio was higher in animals given
radioactive vapour compared with those given salts of inorganic
Table 1. Relationship between concentrations of mercury in samples of blood and hair
in people having long-term exposure to methylmercury from fish
No. of Whole blood Hair (y)
subjects (x) (mg/kg) (mg/kg) Linear regression References
12 0.004-0.65 1-180 y = 280x - 1.3 Birke et al. (1972)
51 0.004-0.11 1-30 y = 230x + 0.6 Swedish Expert Group (1971)
50 0.005-0.27 1-56 y = 140x + 1.5 Swedish Expert Group (1971)
45 0.002-0.80 20-325 y = 260x + 0 Tsubaki (1971)
60 0.044-5.5 1-142 y = 230x - 3.6 Skerfving (1974b)
Studies on a variety of experimental animals indicate that the
kidney is the chief depository of mercury after the administration of
inorganic salts and exposure to elemental mercury vapour. Over 50% of
the body burden of mercury can be found in the kidneys of rats exposed
to mercuric salts and metallic mercury vapour a few days after
receiving the dose. This percentage may rise to 90% or more as the
length of time after exposure increases (Rothstein & Hayes, 1960;
Hayes & Rothstein, 1962; Trojanovska, 1966). However, it should be
noted that in experimental animals, the brain levels of mercury
following exposure to elemental mercury vapour were ten times higher
than brain levels after equal doses of inorganic salts (Berlin et al.,
1966; Magos, 1967; Nordberg & Serenius, 1969). A more uniform
distribution of methylmercury throughout the body also results in much
higher brain levels for a given body burden of mercury as compared
with inorganic salts.
Little information is available on the distribution of mercury in
human organs following exposure to elemental mercury vapour. Takahata
et al. (1970) and Watanabe (1971) have reported mercury levels in the
brain several times higher than those in the liver and other organs
(except the kidney) of miners with long-term exposure to high
concentrations of mercury vapour. These concentration ratios were
maintained even several years after cessation of exposure. High
mercury concentrations in the thyroid and pituitary glands in persons
connected with mercury mining have been reported (Kosta et al., 1975).
It should be noted that organ distribution of mercury after inhalation
of elemental mercury vapour can be dramatically affected by moderate
intakes of alcohol and small doses of the herbicide aminotriazole, as
shown in animals (Magos et al., 1973, 1974). These agents reduce
levels in the lung and increase levels in the liver several-fold.
The percentage of the body burden of methylmercury found in the
brain is much higher in primates than in other animal species (Swedish
Expert Group, 1971). Observations on human volunteers given tracer
doses of radioactive methylmercury (Aberg et al., 1969) indicate that
10% of the radioactivity in the whole body is located in the posterior
part of the head. Probably not all of this represents methylmercury in
the brain but would include methylmercury attached to the hair.
Studies by Miettinen's group (quoted by a Swedish Expert Group, 1971),
on volunteers given tracer doses of radioactive methylmercury,
indicate that an initial rapid distribution throughout the body is
followed by a further slow redistribution of methylmercury to the
6.3 Elimination in Urine and Faeces
Urine and faeces are the principal routes of elimination of
mercury from the body. The contribution of each pathway to total
elimination depends upon the type of mercury compound and the time
that elapses after exposure. Experiments in animals indicate that
elimination of inorganic mercury by the gastrointestinal tract depends
on the size of the dose and the time after exposure. The faecal route
is dominant soon after exposure. The urinary route is favoured when
high doses are given (Prickett et al., 1950; Friberg, 1956; Rothstein
& Hayes, 1960; Ulfvarson, 1962; Cember, 1962; Phillips & Cember, 1969;
Nordberg & Skerfving, 1972).
Data obtained on rats subjected to a single exposure of labelled
203Hg vapour indicated that about 4 times more mercury was eliminated
in the faeces than in the urine (Hayes & Rothstein, 1962). In
prolonged exposure of rats, the proportion changed in favour of
urinary excretion (Gage, 1961). In workers exposed to mercury vapour,
the output of mercury in urine slightly exceeded that in the faeces
(Tejning & Ohman, 1966), High individual variation and great
fluctuation from day to day were the principal features of urinary
excretion in workers under similar exposure conditions (Goldwater et
al., 1963; Jacobs et al., 1964). There is evidence that, on a group
basis, urinary excretion is roughly proportional to exposure (air
concentration) to elemental vapour (MAC Committee, 1969). Occupational
exposure of at least 6 months, 5 days per week at average air
concentrations of mercury of 0.05 mg/m3, should lead to average
urinary concentrations of mercury of about 150 µg/litre.
Piotrowski et al. (1975) have reported changes in urinary rates of
excretion in workmen following exposure to elemental mercury vapour.
They noted that urinary excretion could be described by a two-term
exponential equation with rate constants equivalent to half-times of 2
and 70 days. The short half-time compartment accounted for about
20-30% of the excretion rate under conditions of steady-state
excretion. Piotrowski et al. (1975) suggested that there is variation
in urinary mercury excretion in individuals and that this can be
greatly reduced by collecting the urine samples at the same time in
Mercury exhalation found in animals after exposure to the
elemental vapour (Clarkson & Rothstein, 1964) has also been confirmed
in man (Hursh et al., 1975). This pathway of excretion accounted for
about 7% of the total excretion of mercury in volunteers following
inhalation of a tracer dose. Recent observations indicate that the
concentration of mercury in sweat may be sufficiently high to be taken
into account in the overall mercury balance in workers exposed to
elemental mercury vapour (Lovejoy et al., 1974).
The faecal route is most important in the elimination of mercury
after acute or chronic dosing with methylmercury. Studies on human
volunteers (Aberg et al., 1969; Miettinen, 1973) indicate that
approximately 90% of the elimination takes place via the faeces. This
proportion does not change with time after exposure. Concentrations of
total mercury in urine showed no correlation with blood mercury in
people heavily exposed to methylmercury (Bakir et al., 1973).
6.4 Transplacental Transfer and Secretion in Milk
The transplacental movement of mercury in women exposed to
elemental mercury vapour has not been studied thoroughly.
Experiments on animals reveal that after brief (approximately 20
minutes) exposure to radioactive elemental mercury vapour, the
radioactive mercury easily penetrates the placental barrier (Clarkson
et al., 1972). These authors report that, after equal exposure of
pregnant rats, the fetal uptake was 10-40 times higher after exposure
to elemental mercury vapour than to inorganic salts. In contrast, the
placental content of mercury after exposure to elemental mercury
vapour was only about 40% of that after exposure to inorganic salts of
The alkylmercuric compounds have been known for some time to
penetrate the placenta readily as indicated from studies on
experimental animals (for review, see Swedish Expert Group, 1971). In
a recent study, Childs (1973) noted that the level of methylmercury in
the fetus may be twice that in the maternal tissues when low levels of
methylmercury are fed to rats in a tuna fish matrix. At higher dose
levels, the ratio between fetal and maternal tissues becomes close to
unity. Transplacental movement of methylmercury in women has been
sufficient to cause several cases of prenatal poisonings in various
countries (Engleson & Herner, 1952; Harada, 1968; Bakulina, 1968;
Snyder, 1971; Bakir et al., 1973; Amin-Zaki et al., 1974a). Tejning
(1970) has reported methylmercury levels in fetal blood cells to be
30% higher than in maternal cells in studies on women having normal
pregnancies and a low to moderate fish intake. The relatively higher
concentrations in fetal blood have been confirmed in a study by Suzuki
et al. (1971). It was noted that the plasma levels in both types of
blood were similar and that differences arose only in terms of
concentrations in the red blood cells.
Information on the transplacental movement of other compounds of
mercury in women is lacking. Animal experiments indicate that those
compounds rapidly converted to inorganic mercury in the body, such as
phenylmercury compounds, behave in this respect like inorganic mercury
(for review, see Clarkson, 1972b).
Mercury has been reported in breast milk in women exposed to
methylmercury from fish (Harada, 1968; Skerfving, 1974a) and from
bread contaminated with methylmercury fungicides in the 1971-72
outbreak in Iraq (Bakir et al., 1973). In Iraq, it was noted that
levels of total mercury in milk correlated closely with levels in
whole blood and averaged 5% of simultaneous concentrations in.
maternal blood. The total mercury in milk consisted of two fractions
identified as inorganic mercury (40%) and methylmercury (60%).
Skerfving's (1973) observations on 15 lactating mothers exposed to
methylmercury in fish are in general agreement with the findings in
Iraq except that methylmercury accounted for only 20% of the total
mercury in milk.
Despite the relatively low concentration in milk as compared with
maternal blood, the suckling infants accumulated high concentrations
of mercury in their blood if their mothers were heavily exposed
(Amin-Zaki et al., 1974b). Some Iraqi infants, exposed only through
maternal milk, had blood levels in excess of 100 µg/100 ml. In
prenatally exposed infants, intake of methylmercury by suckling is one
factor responsible for the slower decline in blood levels as compared
with the mother (Amin-Zaki et al., 1974a).
6.5 Metabolic Transformation and Rate of Elimination
The most dramatic example of metabolic transformation is the
conversion of metallic mercury to divalent ionic mercury in the body.
This oxidation reaction has been shown to take place in vitro in the
red cells (Clarkson et al., 1961). More recent studies indicate that
it probably takes place in most other tissues (for details, see Kudsk,
1973). The process is enzyme mediated and the catalase complex is the
most likely site of biochemical oxidation (Kudsk, 1973; Magos et al.,
Studies on the biotransformation of elemental mercury make it
possible to develop a picture of the role of the oxidation process in
the accumulation of mercury vapour in the body and its transport to
the site of action (for details, see Clarkson, 1972a). Elemental
mercury vapour, after inhalation, is absorbed into the blood stream.
Despite the rapid oxidation that has been shown to take place in the
red blood cells, some elemental mercury remains dissolved in the blood
long enough for it to be carried to the blood-brain barrier and to the
placenta. Its lipid solubility and high diffusibility allow rapid
transit across these barriers. Tissue oxidation of the mercury vapour
in brain and fetal tissues converts it to the ionic form which is much
less likely to cross the blood-brain and placental barriers. Thus
oxidation in these tissues serves as a trap to hold the mercury and
leads to accumulation in brain and fetal tissues.
Most studies on the metabolic transformation of organomercury
compounds have concentrated on measurements of the rate of cleavage of
the carbon-mercury bond. There is no evidence in the literature
supporting the possibility of the synthesis of organomercury compounds
in human or mammalian tissues.
The absolute rates of cleavage of the carbon-mercury bond in man
or experimental animals is not known. The relative rates of cleavage
of different mercury compounds have been estimated by measurements of
the amounts of inorganic mercury deposited in tissues following single
doses of organomercury compounds. In general, these studies reveal
that the phenyl-(aryl) and the methoxyethyl compounds are converted
rapidly to inorganic mercury in the body (for reviews, see Gage, 1974;
Clarkson, 1972a). The short-chain alkylmercurials are converted more
slowly to inorganic mercury with the methylmercury compounds being
converted the most slowly of all. The phenyl- and
methoxyethylmercurials are probably converted to inorganic mercury
more or less completely within a few days whereas methylmercury can be
detected in human tissue months after exposure has stopped (Bakir et
al., 1973). Suzuki et al. (1973) have reported the only case in which
the metabolic conversion of ethylmercury has been studied in man.
Proportional values of inorganic mercury to total mercury ranging from
12 to 69% were detected in red cells, plasma, brain, spleen, liver,
and kidney in a patient exposed for about 3 months to
The role of biotransformation in determining the toxicity of
organomercurials is not well understood (for discussion, see Clarkson,
1972b). The rapid conversion of phenylmercury to inorganic mercury
probably accounts for the fact that, in chronic studies on animals,
the effects of this organomercury compound on kidneys were similar to
those of inorganic mercury (Fitzhugh et al., 1950).
The conversion of organic to inorganic mercury may increase or
decrease the total rate of excretion of mercury from the body. If the
intact molecule of an organomercurial is more rapidly excreted than
inorganic mercury, biotransformation will decrease the overall
excretion rate. This has been demonstrated in the case of the
diuretic, chlormerodrin, where the intact molecule is almost
completely excreted within 24 hours, but inorganic mercury remains in
the animal for much longer periods (Clarkson et al., 1965). The
phenyl- and methoxyethylmercury compounds are excreted at a rate
similar to that of inorganic mercury according to studies on
experimental animals. In the case of methylmercury, biotransformation
may play an important part in determining the rate of excretion of
total mercury from the body (Swensson & Ulvarson, 1968, 1969).
Inorganic mercury accounts for approximately 50% of the total mercury
in faeces, the principal pathway of excretion following single or
chronic doses of methylmercury compounds. Methylmercury undergoes
extensive enterohepatic recirculation in rats but inorganic mercury
does not (Norseth & Clarkson, 1971). Thus a small rate of metabolic
transformation in the liver leading to biliary excretion of inorganic
mercury could make an important contribution to the faecal elimination
6.6 Accumulation of Mercury and Biological Half-time
The body accumulates a metal when uptake exceeds elimination. At a
certain stage a steady state may be reached when uptake and
elimination are equal. A common way to express the accumulation is in
terms of biological half-time. The biological half-time for mercury
would be the time taken for the amount of mercury in the body to fall
by one-half. The concept of biological half-time is meaningful,
however, only if the elimination can be approximated to a single
exponential first-order function. This will be true if the
distribution and turnover of a metal in different tissues of the body
are faster than the elimination from the body as a whole. If
elimination from one organ is slow compared with that from other
organs then the calculation of a biological half-time for the whole
body may be completely misleading from the toxicological point of view
(Task Group on Metal Accumulation, 1973; Nordberg, 1976).
Studies on experimental animals and volunteers indicate that, for
methylmercury compounds, the elimination can be approximated to a
single exponential first-order function (Miettinen, 1973; Aberg et
al., 1969; for reviews, see Clarkson, 1972b; Swedish Expert Group,
1971; Task Group on Metal Accumulation, 1973). Observations on
experimental animals indicate that the elimination of mercury after
exposure to mercury vapour, inorganic mercury salts, and the phenyl
and methoxyethyl compounds does not follow such a pattern and thus the
accumulation and elimination of mercury ("the metabolic model") is
much more complex. The pattern of elimination of these mercury
compounds, when administered to animals, is dose- and time-dependent
(Rothstein & Hayes, 1960; Ulfvarson, 1962; Piotrowski et al., 1969).
In cases where the elimination of a metal such as methylmercury
follows a single exponential first order function, the concentration
in an organ at any time can be expressed by the following equation;
C = Co. e-b.t (1)
C = concentration in the organ at time t
Co = concentration in the organ at time o
b = elimination constant
t = time
The relation between the elimination constant and the biological
half-time is the following:
T = 1n 2/ b
T = biological half-time
1n 2 (natural logarithm of 2) = 0.693
If data on exposure and absorption of the metal are known, then it
will be possible to predict the body burden of the metal at constant
exposure over different time periods. If a constant fraction of the
intake is taken up by a certain organ, the accumulated amount in that
organ can also be calculated. The following expression gives the
accumulated amount of metal in the total body (or organ):
A = ( a/b)(1 - exp ( -b.t)) (2)
A = accumulated amount
a = amount taken up by the body (or organ) daily
At steady state the following applies:
A = a/b (3)
In other words, the steady state amount in the body (or organ)
A is proportional to the average daily intake and inversely
proportional to the elimination rate. The latter point will be taken
up later (section 9) in discussing hazards to man, as large individual
variations in elimination rates imply large individual variation in
steady state body burden, even in people having the same average daily
Equations (1), (2), and (3) are illustrated graphically in Fig. 1.
During the period of steady daily intake (assumed to be 10 µg/70 kg
body weight), the amount in the body rises rapidly at first, reaching
half its maximum (steady state) value in a time equivalent to one
elimination half-time (assumed to be 69 days for methylmercury in
man). After an exposure period equivalent to 5 elimination half-times
(approximately one year for methylmercury), the body burden is within
3% of its final steady state value. The steady state value is one
hundred times the average daily intake assuming an elimination
half-time of 69 days. On cessation of exposure, the body burden will
immediately begin to fall following an exponential curve that is an
inverse image of the accumulation curve. Thus the body burden will
have returned to within 3% of pre-exposure values in 5 half-times.
In this example, it is assumed that the hair to blood ratio is
constant and equal to 250 and that 1% of the body burden is found in
1 litre of blood in a 70-kg man.
That this model provides a reasonable approximation to the
accumulation of methylmercury in man over a wide range of daily
intakes is indicated by the data in Tables 2 and 3. Data on
elimination rates for the whole body reported by Aberg et al. (1969)
on 5, and by Miettinen (1973) on 15 volunteers were in good agreement
indicating average values close to the value of 69 days used in Fig.
1. An average value of 50 ± 7 days for clearance half-time from blood
was reported by Miettinen (1973) in volunteers receiving a single
tracer dose. Blood clearance values are difficult to measure
accurately with tracer doses owing to the low counting rates in the
blood samples. Skerfving (1974b) reported clearance from whole blood
ranging from 58 to 87 days in 4 people having high intake (up to
5 µg/kg body weight) of methylmercury from fish, one individual had a
clearance half-time of 164 days. Bakir et al. (1973) reported that
patients having very high blood levels (over 100 µg/100 ml) in Iraq,
had clearance half-times in the same range (45-105 days, mean 65
days). "Clearance" from hair is estimated by analysis of consecutive
short (0.2-1 cm) segments of hair samples and plotting the mercury
concentrations against the distance from the scalp on semilogarithmic
paper (for details, see Birke et al., 1972). A straight line is
usually obtained, the slope of which is equivalent to a biological
half-time if the growth rate of the hair is known. "Clearance"
half-times from hair are assumed to reflect clearance half-times for
blood. Data from a fish eating population (daily intake up to 5 µg/kg)
and on a highly exposed population in Iraq (daily intake up to
50 µg/kg) are compatible with this assumption given the wide range of
individual variations (Table 2).
The relationship between steady state body burden (A) and
average daily intake is given by equation (3); using data derived from
tracer observations on volunteers (Aberg et al., 1969; Miettinen,
1973), one would predict that the steady state blood level ( y ng/ml)
is numerically equal to the average daily intake ( x µg/day/70 kg
body weight) as indicated in Table 3. This calculation assumes a
69-day elimination half-time from the whole body, that 1% of the body
burden is found in 1 litre of blood in a 70-kg "standard man".
Observed steady state relationships between blood concentration
(y) and daily intake (x) are given in Table 3 for several
populations. These populations consist of fishermen and their families
who had had a high dietary intake of fish for many years. The range of
intake between different individuals is high -- up to 800 µg/day. The
relationship between blood concentrations and average daily intake was
found to be linear for each population studied. Linear regression
analysis reveals that the observed relationship between y and x is
less than that predicted by tracer studies. The coefficient of
x lies between 0.5 and 0.8 in the fish eating populations as
compared with the predicted value of unity from tracer studies.
Table 2. Mercury intake and clearance
Clearance half-times (days)
No. of Hg intake
subjects (µg/kg/day) Body Blood Hair References
5 tracer 70 -- -- Aberg et al. (1969)
15 tracer 76 50 -- Miettinen (1973)
5 up to 5 -- -- (33-120) Birke et al. (1972)
5 up to 5 -- seea -- Skerfving (1974)
16 up to 50 -- 65 -- Bakir et al. (1973)
48 up to 50 -- -- 72b Shahristani & Shihab
a One person had a biological half-time of 164 days. The other four were in the
range of 58-87 days.
b The data were distributed bimodally. One group accounting for 89% of the samples
had a mean value of 65 days and the other group had a mean value of 119 days.
Table 3. Relationship of steady state blood concentrations to daily intake of
Ave. Hg intake Steady blood
No. of Time of (µg/day/ concentration
subjects exposure 70 kg B.W.) (ng/ml) References
6 + 26b years 0-800 y = 0.7x + 1 Birke et al. (1967)
139 + 26b years 0-400 y = 0.3x + 5 Tejning (1967, 1969a,
6 + 14b years 0-800 y = 0.8x + 1 Birke et al. (1972)
725c years 0-800 y = 0.5x + 4 Estimated from Kojima &
Araki (unpublished data)
22 years 0-800 y = 0.5x + 10 Skerfving (1974b)
30d 1-2 months 0-2340 r = 0.8x Estimated from Shahristani
& Shihab (1974) and
Shahristani et al. (1976)
15 single y = 1.0x Estimated from Miettinen
tracer dose (1973)
a For details of these calculations, see text.
b None or low fish consumers.
c Estimated from data on hair concentrations and daily intake. The hair to blood
concentration was assumed to be 250 and the average body weight of the population
under study to be 60 kg.
d Estimated from data on hair concentrations and daily intake. The hair to blood
concentration ratio was assumed to be 250.
Given the difficulties in the accurate measurement of dietary
intake and the uncertainty in tracer studies based on counting blood
samples, it is likely that differences between the observed and the
tracer values are not real. This conclusion is supported by the fact
that the Iraqi populations (Table 3, Shahristani et al., 1974), having
an extremely high dietary intake yielded a factor of 0.8 suggesting
that the relationship between y and x is not substantially changed
at high doses.
In summary, a considerable body of evidence exists to support the
linearity of the metabolic model for methylmercury in man. No
definitive evidence is yet available that refutes this conclusion.
However, we cannot exclude the possibility that the mean values of the
parameters of the metabolic model could change by about a factor of
two over a wide dose range. We have, however, taken the predicted
value from the tracer data, since this approach would offer a greater
margin of safety in estimates of hazards to human health.
Biological half-times for other mercury compounds are not well
established and this is particularly true for those organs that are of
toxicological importance. It seems, however, that the biological
half-time for the greatest part of retained salts of inorganic mercury
has an average value of about 40 days (Miettinen, 1973). In five
female volunteers the average half-time was 37 days and in a similar
number of males the average half-time was 48 days.
The biological half-time of both methyl and inorganic salts of
mercury does not appear to be affected whether the compound is
administered in an ionic form or bound to protein (Miettinen, 1973).
Limited information is available on biological half-times of
mercury in the body following exposure to elemental vapour. Five
volunteers inhaled radioactive mercury vapour for 10-15 minutes and
were subjected to whole body counting for up to 43 days after exposure
(Hursh et al., 1975). Elimination from the body followed a single
exponential process having a biological half-time of 58 days with a
range of individual values of 35-90 days.
As noted above, whole body elimination half-times may not be a
reliable guide to accumulation in specific organs. For example, the
fact that Takahata et al. (1970) and Watanabe (1971) found high
mercury concentrations in the brain in an individual 10 years after
cessation of exposure indicates that accumulation in the brain does
not follow the same kinetics as seen by whole-body counting.
Observations on experimental animals also indicate that the half-time
in brain is longer than in other organs (Task Group on Metal
6.7 Individual Variations -- Strain and Species Comparisons
As a general rule, the processes of absorption by inhalation or
ingestion do not appear to be subject to large species and strain
differences (for detailed review, see Clarkson, 1972b). Elemental
mercury vapour and methylmercury compounds are well absorbed across
the pulmonary epithelium and the gastrointestinal tract, respectively,
in a variety of animal species. Distribution in the body tissues is
subject to species differences. A very pronounced example is the case
of the red cell to plasma ratios of methylmercury where the ratio can
be as high as 300 in the rat and as low as 10 in primates. Differences
in red cell to plasma ratio may account for species differences in
blood to brain ratios (Vostal, 1972). The blood to brain ratio has
been reported to be approximately 10-20 for the rat, approximately
unity for the cat, 0.5 for the dog and pig, and 0.1 for primates (from
data reviewed by a Swedish Expert Group, 1971). Careful and accurate
quantitative comparison of species differences and distribution is not
possible because of the different experimental conditions in these
The oxidation of elemental mercury vapour to ionic mercury and the
cleavage of the carbon-mercury bond in a variety of mercurials has
been described for several different species of animal. However the
observations in this field are not adequate to allow quantitative
comparison of metabolic rates of breakdown in different species.
The rate of elimination of mercury from the body is subject to
wide species variation (for detailed review, see Clarkson, 1972a). In
general, animals of small body weight tend to excrete mercury more
rapidly than larger animals and the cold blooded species, particularly
fish, appear to retain mercury for an extremely long time.
Species differences in the elimination of methylmercury have been
reported. The mouse and the rat have half-times between 8 and 16 days
as compared with 70 days in primates. The seal is reported to have a
half-time of 500 days and fish and crustaceans appear to have
half-times ranging from 400 to over 1000 days. These species
differences indicate that we cannot extrapolate parameters describing
the metabolic fate of mercury in animals to that in man. Furthermore,
since these parameters determine the amount of mercury accumulated in
the body, it would appear that quantitative information on toxicities
cannot be directly extrapolated from animals to man.
The half-time of clearance of mercury from blood and hair varies
considerably between individuals exposed to methylmercury (Birke et
al., 1972; Skerfving, 1974). In the outbreak in Iraq the half-time of
clearance from blood ranged from 40-105 days in 16 subjects (Bakir et
al., 1973) and from hair the range was from 35-189 days (Shahristani &
Shihab, 1974). The Iraqi data on blood samples (Bakir et al., 1973)
were obtained about 1-2 months after the termination of exposure.
Average biological half-times for groups of at least 15
individuals seem to be remarkably constant. The average half-time in
blood was 65 days (Bakir et al., 1973), in hair 72 days (Shahristani &
Shihab, 1974), and in the whole body in 15 subjects given a tracer
dose, 76 days (Miettinen, 1973).
7. EXPERIMENTAL STUDIES ON THE EFFECTS OF MERCURY
7.1 Experimental Animal Studies
7.1.1 Acute studies
Little information is available on the acute toxicity of elemental
mercury vapour to animals. Ashe et al. (1953) reported evidence of
damage to brain, kidney, heart, and lungs in rabbits exposed to
mercury vapour at a mercury concentration of 29 mg/m3 of air. This
concentration under the circumstances of their experiments would
represent an atmosphere saturated with mercury vapour. The first
effects were seen within 1 hour of exposure and subsequent severe
changes resulted after longer exposures.
Information on the LD50 (the dose of mercury that kills half the
test population) has been reviewed by a Swedish Expert Group (1971).
The results reported for different mercury compounds are not easily
comparable since different animal species and different routes of
administration have been used for the test. Nevertheless, despite all
these differences, the LD50 lies between approximately 10 and 40
mg/kg body weight for all compounds tested to date including inorganic
mercury, arylmercury, alkoxyalkyl- and alkylmercury compounds. The
remarkable similarity in LD50s of these various types of mercury
compound is probably due to the fact that when given in acute massive
doses, mercury in whatever chemical form will denature proteins,
inactivate enzymes, and cause severe disruption of any tissue with
which it comes into contact in sufficient concentration. The symptoms
of acute toxicity usually consist of shock, cardiovascular collapse,
acute renal failure and severe gastrointestinal damage. A variety of
complexing and chelating agents, all of which contain sulfhydryl
groups can modify the LD50s of mercury and its compounds (for review,
see Clarkson, 1972a). These agents are most effective when given
either prior to the mercury dose or in the few hours following a
single dose of mercury. The importance of time of administration is to
be expected since the effects of these agents are to reduce the
reactivity of mercury in the body and to do so before irreversible
damage has been inflicted on the tissue (see also, section 9.2).
Irreversible damage is used in this document to define cellular or
organ damage that is not repaired even after the cessation of
exposure; some improvement in function or in the condition of the
poisoned tissue may occur but recovery is never complete. In the case
of reversible damage, regeneration of cells and restoration of
function takes place after the cessation of exposure.
7.1.2 Subacute and chronic studies
184.108.40.206 Reversible damage
This section deals with toxic effects known to be reversible, at
least up to a certain dose and/or duration of exposure. It should be
noted, however, that at higher doses or longer duration of exposure,
the damage can surpass the stage of reversibility.
Studies by Trahtenberg (1969) (reviewed by Friberg & Nordberg,
1973) indicate that exposure of rats to concentrations of elemental
mercury vapour in the range of 0.1-0.3 mg/m3 for over 100 days
increased uptake of radioactive iodine by the thyroid. Friberg &
Nordberg (1973) also refer to unpublished observations by Aveckaja
which, under different conditions including prolonged exposure,
indicate the opposite effect. Kournossov reported in 1962 (reviewed by
Friberg & Nordberg, 1973) effects on the behaviour of rats at mercury
concentrations in air as low as 0.005 mg/m3. Studies by Armstrong et
al. (1963) on pigeons showed irreversible behavioural changes only at
vapour levels well in excess of recommended maximum allowable
concentrations. It is clear that many more studies need to be carried
out on behavioural and other subtle changes resulting from exposure to
mercury vapour at these low concentrations.
Fitzhugh et al. (1950) studied the toxicity of mercury(II)
chloride and phenylmercury acetate when added to the diet of rats for
periods of up to 2 years. Morphological changes were induced in kidney
tissue at approximately the same mercury levels for both compounds.
The similarity in the effects of the two compounds on the kidney is
probably due to the fact that phenylmercury compounds are rapidly
converted to inorganic mercury in animal tissues.
The main effect of alkylmercurials is the irreversible action on
the central nervous system. However, Lucier et al. (1971, 1972, 1973)
have reported that subacute doses (with no neurological signs) caused
a marked decrease in rat liver mixed-function oxidase activity. This
effect was shown to be due to an increased degradation rate of
cytochrome P-450 in vivo. Methylmercury also depressed the activity
of enzymes dependent upon cytochrome P-450. Ultrastructural changes
involving the endoplasmic reticulum in liver have been reported by
Chang & Yamaguchi (1974) and these effects were reversible.
In rats, administration of methylmercury can produce kidney damage
manifested by tubular degeneration in the distal convoluted tubules
after daily doses of 10 mg/kg for 7 days (Klein et al., 1972). With
lower doses of methylmercury, morphological and functional damage is
produced in kidney tissue in the absence of any signs of neurological
dysfunction (Fowler, 1972a, 1972b; Magos & Butler, 1972; Klein et al.,
1973). Ultrastructural studies by Fowler (1972a) showed that female
rats given 2 mg/kg methylmercury in their diet were more sensitive to
methylmercury than males. The primary lesion was characterized by
extrusion of cyto-plasmic masses from proximal tubular cells. It has
been suggested that the nephrotoxic effect is due to inorganic mercury
split from methyl-mercury in vivo (Klein et al, 1973).
220.127.116.11 Irreversible damage
With the exception of massive doses of inorganic compounds or
prolonged exposure to extremely high concentrations of elemental
mercury vapour, as in the experiments of Ashe et al. referred to
below, the effects of inorganic mercury on tissues are generally
Ashe et al. (1953) reported microscopically detectable changes in
the organs of dogs, rabbits, and rats exposed to concentrations of
elemental mercury vapour ranging from 0.1 to 30 mg/m3 for different
periods of time. Severe damage was noted in kidneys and brains at
mercury levels in air of 0.9 mg/m3 after an exposure period of about
12 weeks. After longer periods of exposure to 0.1 mg/m3, no
microscopically detectable effects could be seen.
The short-chain alkylmercurials are primarily neurotoxic in man.
After a single dose there is a latency period of days or weeks before
signs of poisoning occur. Many of the signs of methylmercury poisoning
observed in man can be reproduced in animals under appropriate
conditions. For example, Berlin et al. (1973) noted that a sudden
visual disturbance occurred in monkeys given subacute doses of
methylmercury. Prolonged exposure to methylmercury resulted in a
gradual constriction of the visual field and impaired motor
coordination and possibly sensory disturbances. Neurological signs of
damage have also been produced in the mouse, rat, ferret, cat, and dog
by feeding them methylmercury compounds (Chang et al., 1974; for
review, see Swedish Expert Group, 1971). The toxicity of methylmercury
to animals does not appear to be affected whether it is given to them
as a pure chemical, e.g. methylmercury chloride, or whether it has
accumulated naturally as in fish such as the Northern Pike (Swedish
Expert Group, 1971; Albanus et al., 1972).
Grant (1973), in his studies of primates experimentally poisoned
with methylmercury, reported findings confirming those of Hunter &
Russell (1954). Neuronal damage and destruction was observed in the
visual cortex, and the granular layer of the cerebellum. The dose-rate
of methylmercury, the period of dosing, and the animal species all
influence the pattern of pathological damage. For example, in monkeys
after short periods of exposure to high doses, there is an abrupt
visual change over two days leading to blindness. This is accompanied
by damage to the neurons in the visual cortex. Longer exposure to
lower daily doses of methylmercury leads to more generalized damage to
the cortex and is accompanied by gradual onset of visual changes and
other signs of central nervous involvement such as ataxia.
Recent studies on rats reviewed by Somjen et al. (1973a) have
confirmed the findings of Hunter et al. (1940) that the earliest
neurological effects in these animals is damage to the peripheral
sensory nerves. Later the disease affects other parts of the central
and peripheral nervous systems. There is now evidence that the primary
site of the disease is the cell bodies in the dorsal root ganglia with
secondary deterioration in their fibres (Chang & Hartman, 1972a,
1972b). Consistent with this interpretation, Somjen et al. (1973b)
found that the spinal dorsal root ganglia contained the highest
concentrations of mercury. Electrophysiological investigations
confirmed the findings drawn from morphological evidence that the cell
bodies in the spinal ganglia are the primary sites of action (Somjen
et al., 1973a).
Morphological, electrophysiological and biochemical changes have
been demonstrated in animals prior to the onset of overt signs of
poisoning. These phenomena, especially with respect to morphological
changes, have been referred to as "silent damage". For example,
Nordberg et al. (1971) and Grant (1973) noted that morphological
damage was present in certain of the test monkeys before signs of
visual impairment could be detected. Somjen et al. (1973a) reported
electrophysiological manifestations of methylmercury intoxication in
rats preceding clinical signs. Yoshino et al. (1966) noted a decreased
uptake of amino acids in brain slices taken from animals given high
doses of methylmercury at a time before signs of poisoning had
appeared. Cavanagh & Chen (1971) reported that incorporation of amino
acids into protein was impeded in spinal root ganglia of rats, treated
with methylmercury before signs of poisoning were present. Chang &
Hartman (1972c) noted damage to the blood-brain barrier as early as 12
hours after a dose of 1 mg/kg of methylmercury chloride to rats. If
these observations on experimental animals may be extrapolated to man,
the possibility must be considered that significant damage to the
central and peripheral nervous systems may take place prior to the
onset of clinical signs and symptoms.
Methylmercury compounds have been demonstrated to disturb mitosis
in the plant cell, in human leucocytes treated in vivo, and in human
cells in tissue culture. The short-chain alkylmercurials cause
chromosome breakage in plant cells and point mutations in Drosophila
(for detailed reviews, see Swedish Expert Group, 1971; Ramel, 1972).
Clegg (1971) has given a detailed review of the embryotoxicity of
the short-chain alkylmercurials. In general, animal experiments
confirm the idea derived from epidemiological observations in the
Minamata epidemic that much more damage was inflicted on the fetus
than on the mother. Spyker et al. (1972) have recently reported on
performance deficits in mice treated prenatally with methylmercury.
The alkylmercury compounds may also damage the gametes prior to
fertilization (Khera, 1973). Virtually no information is available on
the morphological and biochemical factors related to prenatal damage
in experimental animals. The results, however, do point to incipient
hazards to human fetuses exposed before birth.
In discussing the biological effects of methylmercury compounds,
species differences should be considered. Although man, monkeys, and
pigs may become blind at high exposures, similar visual disturbances
in cats were not detected (Albanus et al., 1972; Charbonneau et al.,
1974). Pronounced morphological changes are seen in the peripheral
nervous system of rats (Somjen et al., 1973a), but again, were not
detected in cats (Albanus et al., 1972; Charbonneau, 1974). However,
one cannot exclude the possibility that qualitative differences,
reported in studies of different species including man, may reflect
differences in degree and intensity of exposure in man and in
experimental conditions in animal studies.
18.104.22.168 Interactions with physical and chemical factors
Parizek & Ostadalova (1967) reported that selenite salts could
protect experimental animals against the toxic effects of inorganic
mercury. Selenium also depressed the passage of inorganic mercury into
fetuses and its secretion into milk (for review, see Parizek et al.,
1969, 1971, 1974). Parallelism between tissue concentrations of
mercury and selenium has been reported in human subjects exposed to
elemental mercury vapour (Kosta et al., 1975). Several recent
publications have claimed that selenite added to the diet protects
experimental animals against methylmercury compounds (Ganther et al.,
1972; El Bergerami et al., 1973; Potter & Metrone, 1973). Ganther &
Sunde (1974) have reviewed evidence indicatting that the content of
selenium in tuna fish is sufficiently high to provide substantial
protection against methylmercury. However, the protective factor in
tuna fish, whether selenium or some other substance, has yet to be
Toxicological interactions have been reported between
methylmercury and the chlorinated hydrocarbon pesticide, dieldrin.
Rats dosed with both dieldrin and methylmercury, showed less
morphological damage of the pars recta tubule than animals given only
methylmercury. However, there was degeneration in proximal tubular
cells (Fowler, 1972b). It has been demonstrated recently that
phenobarbital administration increases the biliary excretion of
methylmercury compounds (Magos & Clarkson, 1973).
Estrogenic hormones (Lehotzky, 1972) and spironolactone (Selye,
1970) protect the kidney from methoxyethylmercury salts and
mercury(II) chloride, respectively. The mechanism of these actions is
The question of the interaction of the physical and chemical
factors on the toxicity of methylmercury is important and should be a
major priority in future research studies. Extrapolation of
epidemiological and toxicological data from populations in Japan and
in Iraq suffering from methylmercury poisoning is fraught with
difficulties when the possible interactions of local factors are not
taken into account.
7.1.3 Biochemical and physiological mechanisms of toxicity
A physiological basis for the action of mercury and other heavy
metals had already been propounded prior to 1967 (for reviews, see
Hughes, 1957; Passow et al., 1961; Peters, 1963; Webb, 1966). Two
general concepts on the mechanisms of action of mercury and other
heavy metals are discussed in these reviews. The first dates back to
the 1940s and is attributed to Peters, 1963. The toxic sequelae of
heavy metal action on tissues result from a primary "biochemical
lesion" whereby a critical enzyme or metabolic process is inhibited.
Unfortunately, despite a considerable amount of research work (see
Webb, 1966), it has not been possible to locate the biochemical lesion
associated with the toxic actions of mercury.
An alternative general concept mainly proposed by Passow et al.
(1961) is that the cell membrane is the first site of attack by heavy
metals. Topographically, this would seem reasonable. Furthermore, the
membrane is known to contain sulfhydryl groups that are essential to
the normal permeability and transport properties of the cell membrane.
These same sulfhydryl groups are known to have a very high affinity
for mercury and its compounds. Passow et al. (1961) summarized a great
many experimental studies that support this general idea. However, it
must be admitted that most experimental work testing this idea is
based on in vitro studies on isolated cells and tissues so that the
role of membrane damage in the pathogenesis of heavy metal poisoning
remains to be established. The effect of mercury in intercellular
membranes is also of some interest.
The affinity of mercury for thiol groups in proteins and other
biological molecules is far in excess of its affinity for other
biologically occurring ligands (Clarkson, 1972b). As pointed out by
Rothstein (1973), the affinity of mercury(II) cations for the
sulfhydryl groups of proteins creates a "severe logistics problem" for
those interested in elucidating the mechanisms of action of
mercurials. "Although mercurials are highly specific for sulfhydryl
groups, they are highly unspecific in terms of proteins. Almost all
proteins contain sulfhydryl groups that are metal-reactive.
Furthermore, because most sulfhydryl groups are important in most
protein functions, mercurials can disturb almost all functions in
which proteins are involved. Thus almost every protein in the body is
a potential target". In other words, the mercurials are potent but
non-specific enzyme poisons. Mercury will inflict cellular damage
wherever it accumulates in sufficient concentrations. This reasoning
has given rise to the idea that the selective toxicity of mercury is
related to its selective distribution. In general, there seems to be
some truth in this. Inorganic mercury compounds are avidly accumulated
by the kidney which is the target organ for this compound.
Studies by Somjen et al. (1973b) on the microdistribution of
mercury in the nervous system also lend credence to the importance of
distribution of methylmercury in that the spinal root ganglia, the
site of peripheral nerve damage, are also the area of highest
accumulation of methylmercury in rats.
However, it seems that distribution factors alone cannot give a
complete explanation for the toxicity of methylmercury. The kidney is
always the site of the highest accumulation of mercury irrespective of
the form of the mercury compound involved. For example, kidney levels
of methylmercury are much higher than brain levels and yet kidney
damage, except in the rat, is much less than that seen in the central
In recent years, interest has arisen in the biocomplexes of
mercury in the body. The toxicity of any mercury compound will be
determined by its chemical activity close to its site of action. For
example the chloride salts of mercury compounds when added in vitro
to tissue preparations are highly toxic, whereas when mercury is added
in the presence of sulfhydryl compounds the toxicity is very much less
(for review of this subject, see Clarkson & Vostal, 1973). Thus the
chemical state of combination of mercury in plasma and other body
fluids may be of primary importance in determining the particular site
of action of the mercurial.
Mercury accumulated in the kidney is contained there partly in
form of a metallothioneine-like complex (Jakubowski et al., 1970;
Wisniewska, et al., 1970). In the rat, binding by this protein is
especially effective in repeated exposure to mercury(II) chloride,
owing to the induction of higher levels of the metallothioneine-like
protein by mercury (Piotrowski et al., 1974a, 1974b). This form of
binding probably also occurs in the case of exposure to elemental
mercury vapour since this exposure results in enhancement of the
metallothioneine level in the kidneys of rats (Sapota et al., 1974).
Binding of inorganic mercury in other organs may also involve storage
of a similar form, as found in the liver (Wisniewska et al., 1972) and
brain of the rat (Sapota et al., 1974).
The binding of mercury by metallothioneine-like protein of the
kidneys is enhanced by the presence of cadmium (Shaikh et al., 1973)
and therefore may play an important role in man whose kidneys
accumulate, in normal conditions, considerable amounts of cadmium
(Piscator & Lind, 1972). The above applies also to organic mercurials,
which are rapidly converted into inorganic mercury, as in the case of
phenylmercury acetate (Piotrowski & Bolanowska, 1970).
The primary biochemical lesions associated with mercury poisoning
have not yet been established. Virtually nothing is known of the
biochemical disturbances associated with exposure to metallic mercury
vapour. Studies referred to in section 22.214.171.124 by Cavanagh & Chen
(1971) and Yoshino et al. (1966) suggest that protein synthesis may
undergo an early biochemical change preceding clinical signs and
symptoms of methylmercury poisoning. This may give an explanation for
the latent period associated with this form of mercury poisoning.
8. EFFECTS OF MERCURY ON MAN -- EPIDEMIOLOGICAL AND
8.1 Epidemiological Studies
8.1.1 Occupational exposure to mercury vapour, alkylmercury
vapour and other compounds
Occupational exposures to elemental mercury vapour have been the
subject of recent reviews by Friberg & Nordberg (1972, 1973) and NIOSH
(1973). Many studies dating back to the 1930s have related the
frequency of signs and symptoms of mercury poisoning to exposure.
These studies, involving observations of more than one thousand
individuals, indicate that the classical signs and symptoms of
elemental mercury vapour poisoning (objective tremors, mental
disturbances, and gingivitis) may be expected to appear after chronic
exposure of workers to air concentrations of mercury above 0.1 mg/m3
(Neal et al., 1937; Smith & Moskowitz, 1948; Smith et al., 1949;
Friberg, 1951; Bidstrup et al., 1951; Vouk et al., 1950; Kesic &
Heusler, 1951; Baldi et al., 1953; Seifert & Neudert, 1954; McGill et
al., 1964; Ladd et al., 1966; Copplestone & McArthur, 1967; Smith et
al., 1970). The industries involved included the chloralkali industry,
the manufacture of thermometers and graduated scientific glassware,
the repair of DC electrical meters, the mining and milling of mercury,
the manufacture of artificial jewellery, the felt hat industry and
others (NIOSH, 1973). Most of the publications referred to above do
not report time-weighted average exposures and few give information as
to the physical and chemical forms of mercury in the atmosphere.
Different methods of measurement of mercury in air were employed some
of which measured only mercury vapour, while others attempted to
include particulate forms of mercury. Most of the studies, if not all,
assumed that exposure occurred only during the working day. However,
evidence has now come to light that, in certain industries, metallic
mercury may be entrapped in the clothing and contaminate the home,
particularly in those industries actually handling liquid metallic
mercury (West & Lim, 1968; Danzinger & Possick, 1973).
Effects of elemental mercury vapour, other than those designated
as classical mercurialism, have been reported (Smith et al., 1970;
Trahtenberg, reviewed by Friberg & Nordberg, 1973). The study of Smith
and co-workers involved observations on 567 workers exposed to mercury
in chloralkali plants. The air concentrations of mercury (measured by
a mercury vapour meter) ranged from less than 0.01 to 0.27 mg/m3 and
time-weighted averages were calculated for each worker. A significant
increase in the frequency of objective tremors was noted at mercury
levels in air above 0.1 mg/m3 in agreement with previous reports on
occupational exposure. However, a significant increase was observed at
mercury concentrations in air of 0.06-0.1 mg/m3 in such non-specific
signs and symptoms as loss of appetite, weight loss, and shyness.
Studies related to assessment of the occurrence of a so-called
"asthenic-vegetative syndrome" or "micromercurialism" have been
reported by Trahtenberg (1969). This syndrome may occur in persons
with or without mercury exposure. For a diagnosis of mercury-induced
asthenic vegetative syndrome Trahtenberg (1969) (reviewed by Friberg &
Nordberg, 1972) required that not only neurasthenic symptoms should be
present but as supporting evidence three or more of the following
clinical findings; tremor, enlargement of the thyroid, increased
uptake of radio-iodine in the thyroid, labile pulse, tachycardia,
dermographism, gingivitis, haemotological changes, and excretion of
mercury in the urine which was above normal or increased 8-fold after
medication with unithiol.
Trahtenberg, in her monograph (1969) considered that this syndrome
would be more frequently found in persons exposed to mercury
concentrations between 0.004-12 mg/m3 but, upon detailed scrutiny of
her data (see review by Friberg & Nordberg, 1972), there does not seem
to be any difference between exposed and control groups that can be
related to mercury exposure.
Studies on the prevalence of a similar syndrome defined as a
combination of insomnia, sweating, and emotional lability have been
reported by Trahtenberg, et al., quoted by Trahtenberg (1969), in
workers exposed to mercury levels of 0.006-0.01 mg/m3 and
temperatures of 40-42°C in the summer and 28-38°C in the winter. They
found 28-50% prevalence of this syndrome in this exposed group and
only 13% of the same syndrome in a control group exposed to only
38-42°C temperatures. Details about the selection and evaluation of
these workers are not known.
The studies of Bidstrup et al. (1951) and Turrian et al. (1956)
also indicate that psychological disturbances may be seen at air
concentrations of mercury below 0.10 mg/m3. Thus it is impossible, at
this time, to establish a lower exposure limit at which no effects
occur. There is a continuous need for research studies on the effects
of exposure of people to mercury vapour concentrations below
Short-chain alkyl compounds have been the subject of recent
reviews by a Swedish Expert Group (1971) and Kurland (1973), but no
new information has appeared in the literature on occupational
exposures to methylmercury or other short-chain alkylmercury
compounds. Following the description of the first two cases of
occupational exposure to diethylmercury compounds in 1865 by Edwards
(1866), occupational exposures have been infrequent and usually
limited to a few individuals. For example, exposure has occurred in
laboratory personnel, workers, and farmers involved in either the
production of alkylmercury fungicides or their application to cereal
seeds, in people in seed testing institutes, and in workers in pulp
mills and saw mills. Exposure is presumed to be mainly by inhalation
of the vapour or dust but it is possible that, in some cases,
absorption of the liquid preparation of the fungicide may have
occurred through the skin.
Occupational exposures to alkylmercury compounds, although not
important numerically or epidemiologically, have been the occasion for
the description of the signs and symptoms of poisoning. For example,
four workmen exposed to methylmercury fungicide were the basis of the
now classic reports of Hunter et al. (1940) and Hunter & Russell
(1954) which gave the detailed pathology of methylmercury poisoning in
Occupational exposures to the phenyl- and methoxyethylmercury
compounds have been the subject of a recent review (Goldwater, 1973).
The hazards from industrial exposure to these compounds appear to be
very low. Ladd et al. (1964), in a study of 67 workers occupationally
exposed to phenylmercury compounds, found no evidence of adverse
health effects. Mercury levels in air in this study were mainly below
0.1 mg/m3. Comparison of readings with a mercury vapour detector and
estimates of total mercury in the atmosphere revealed that elemental
mercury vapour was the principal form of mercury. Goldwater (1964,
1973) makes reference to seven workers who had spent approximately six
weeks preparing and packaging a batch of material containing
methoxyethylmercury chloride. Four weeks after they had completed this
task their whole-blood mercury levels were in the range of
34-109 µg/100 ml with an average of 65 µg/100 ml. At no time did any
toxic sign or symptom appear. These workers may also have had a
limited exposure to phenylmercury compounds.
At this point it is worth while to quote from the Task Group on
Metal Accumulation (1973) indicating the need for carefully controlled
studies of occupationally exposed groups. "There is a need in
industrially exposed populations for standardized, wherever possible
collaborative, epidemiological studies, where cohorts can be followed
in time and where groups can be related to each other. With some
occupational exposures to the less common metals, only small groups
may be available for study in any one country, so that international
collaboration in epidemiological studies would again be of value".
8.1.2 General population
Epidemics of poisoning in the general population due to exposure
to phenyl- and methoxyethylmercury compounds have not been reported.
Two outbreaks of poisoning due to elemental mercury vapour occurred in
the 19th century, one due to a fire in the mercury mines in Idria, the
other being caused by spillage of metallic mercury in a British
warship in the early 1800s (Bidstrup, 1964). Fernandez et al. (1966)
have reported that, in the village of Almaden, the site of the large
mercury mines in Spain, air levels exceeded 0.1 mg/m3. However, there
are no reports as yet about the health status of the population in the
Methyl- and ethylmercury compounds have been the cause of several
major epidemics of poisoning in the general population due either to
the consumption of contaminated fish or to eating bread prepared from
cereals treated with alkylmercury fungicide. The two major epidemics
of methylmercury poisoning in Japan in Minamata Bay (Katsuna, 1968)
and in Niigata (Niigata Report, 1967) were caused by the industrial
release of methyl- and other mercury compounds into Minamata Bay and
into the Agano River followed by accumulation of the mercury by edible
fish. The median level of total mercury in fish caught in Minamata Bay
at the time of the epidemic has been estimated as 11 mg/kg fresh
weight and in the Agano river in Niigata as less than 10 mg/kg fresh
weight (Swedish Expert Group, 1971).
A recent report by Tsubaki (1971) indicates that follow-up
observations on exposed people in Niigata revealed a much larger
number having mild signs and symptoms than the original 46 that had
been reported. These milder cases may only have had paraesthesia. By
1971 a total of 269 cases of methylmercury poisoning had been reported
in Minamata and Niigata, of which 55 proved fatal. By 1974, more than
700 cases of methylmercury poisoning had been identified in Minamata
and more than 500 cases had been identified in Niigata (personal
communication, Tsubaki, 1975).a The two Japanese epidemics have been
the subject of intensive studies on the effects of methylmercury on
man and have resulted in important conclusions concerning
dose-response relationships (Swedish Expert Group, 1971).
a It has not been possible for this group to review data on the new
cases reported since the publication of the Niigata and Minamata
Epidemics resulting in the largest number of cases of poisoning
and of fatalities have been caused by the ingestion of contaminated
bread prepared from wheat and other cereals treated with alkyl-
(methyl- or ethyl-) mercury fungicides. The largest recorded epidemic
took place in the winter of 1971-72 in Iraq resulting in the admission
of over 6000 patients to hospital and over 500 deaths in hospital
(Bakir et al., 1973). Previous epidemics have occurred in Iraq (Jalili
& Abbasi, 1961), in Pakistan (Haq, 1963), in Guatemala (Ordonez et
al., 1966), and on a limited scale in other countries (Snyder, 1971).
Reports on these epidemics have resulted in interesting clinical
findings but quantitative studies relating exposure to effects have
been reported only on the recent epidemic in Iraq (Bakir et al., 1973;
Kazantzis et al., 1976; Mufti et al., 1976; Shahristani et al., 1976).
In the Iraqi outbreak, the mean methylmercury content of the wheat
was 7.9 mg/kg with most samples falling between 3.7 and 14.9 mg/kg.
The mean methylmercury content of wheat flour samples was 9.1 mg/kg
with a range of 4.8-14.6 mg/kg in 19 samples (Bakir et al., 1973). The
average weight of the home-made loaves was about 200 g with a moisture
content of about 30% of the fresh weight (Damluji, 1962). The range of
daily intake of bread varied widely. In an epidemiological survey of a
heavily affected village, Mufti et al. (1976) reported that the
average total ingested dose of a group of 426 people was about 150 mg
of mercury but some people may have consumed as much as 600 mg. The
average daily intake of contaminated loaves was 3.2 loaves although
individual variation was large, some people eating up to 10 loaves per
day. The daily intake of methylmercury would vary greatly. The average
daily intake of mercury in this village would be 80 µg/kg assuming a
body weight 50 kg for the population, with extremes of daily intake
attaining 250 µg/kg. In the most severely affected group, reported by
Bakir et al. (1974), the highest daily intake of mercury was about
130 µg/kg body weight. The average period of consumption for groups of
patients reported by Bakir et al. (1973) ranged from 43-68 days. Mufti
et al. (1976) reported mean consumption periods in villages to be
about 32 days but some people continued for up to 3 months. Birke et
al. (1972) and Skerfving (1974b) have reported on families in Sweden
consuming fish containing mercury levels of 0.3-7 mg/kg. Daily intake
ranged up to approximately 5 µg/kg body weight. In two cases, intake
was as high as 10-20 µg/kg. The highest recorded blood level of
mercury was 1.2 µg/g of red cells or approximately 60 µg/100 ml of
whole blood. A total of 188 people were referred to in these studies.
No signs or symptoms of poisoning attributable to methylmercury were
Clarkson et al. (1975) and Marsh et al. (1974) have offered
preliminary information on 163 fishermen based in American Samoa who
ingested unusually high amounts of fish containing methylmercury. Data
on daily intake of mercury were not reported but the mercury levels in
blood in this population ranged as high as 28 µg/100 ml. No signs or
symptoms of poisoning could be ascribed to methylmercury.
Turner et al. (1974) have reported on neurological examinations of
186 persons living in two fishing villages in northern Peru.
Concentrations of total mercury, methyl- and inorganic mercury were
measured in blood samples from 141 of these villagers. The
concentration range of methylmercury in blood was from 1.1 to
27.5 µg/100 ml with a mean of 8.9 µg/ 100 ml. The mean intake of fish
was approximately 10 kg per family per week and the average family
size was six.
Fifty-one persons from a "control" village were also examined. The
mean intake of fish was 1.0 kg per family per week and the mean family
size was 6.4. Methylmercury concentrations in blood averaged
0.99 µg/100 ml with a range of 0.33-2.5 µg/100 ml. No correlation was
observed between blood levels of methylmercury and the frequency of
signs and symptoms usually associated with methylmercury poisoning
(paraesthesia, ataxia of gait and limbs, impaired vision, and
Paccagnella et al. (1974) have reported blood concentrations of
mercury in a community in the Mediterranean island of S. Peitro
(Cagliari). The average dietary weekly intake of fish was 300 g. Tuna
fish, with an average total content of mercury of 1.23 mg/kg and a
methylmercury content of 0.92 mg/kg was an important dietary item.
Other fish in their diet had an average total mercury concentration of
0.33 mg/kg. The concentration of total mercury in blood was measured
in 115 people aged from 10 years upwards. The mean mercury
concentration in blood in the males was approximately 8 µg/100 ml and
the maximum recorded concentration was 23 µg/100 ml. The females had
average mercury levels in blood of 6 µg/100 ml with a maximum level of
24 µg/100 ml. Three individuals in the sample population of 115 had
blood levels higher than 20 µg/100 ml.
Epidemiological studies show that these exposed populations may be
classified into three categories distinguished by intensity and
duration of exposure to the short-chain alkylmercurials. Populations
consuming contaminated grain had a high daily intake of mercury
(reaching over 200 µg/kg for brief periods averaging 1-2 months). The
outbreaks in Japan fall into the second category where daily mercury
intakes ranged from 5 to 100 µg/kg with a median of 30 µg/kg/day with
exposure times lasting from several months to years, in Niigata,
although doubts on the accuracy of those figures have been expressed
by a Swedish Expert Group (1971). The third category includes
populations having unusually high fish intakes for years if not, for
most of their lives, such as in parts of Sweden (Skerfving, 1974) the
Samoan fishermen (Marsh et al., 1974), and in Peru (Turner et al.,
1974). A small proportion of those in this category may attain mercury
levels up to 5 µg/kg/day or even higher. Quantitative studies relating
frequency of signs and symptoms to various indices of exposure have
been reported for all three categories so that it is now possible to
compare the effects of differences in intensity and duration.
Before discussing these studies, it would be well to point out
some of the attendant difficulties. Methylmercury and the other
short-chain alkylmercurials produce effects unique among the mercury
compounds. However, some, if not all, of these effects can be caused
by agents other than mercury or by certain disease states. Thus, in
examining the population for neurological changes, it must be borne in
mind, that there could be, at least in theory, many causes of the
observed neurological effects other than methylmercury itself.
Dose-response relationships derived from these epidemiological studies
imply a cause-effect relationship. In fact, the only proof we have
that methylmercury caused certain effects in these populations is that
(1) these effects coincide in time with exposure to methylmercury, (2)
the frequency of these effects in a given population increases with
increasing exposure to methylmercury, and (3) the major signs have
been reproduced in some animal models. One of the key problems in
these studies is to distinguish between the background frequency of a
sign or symptom and the increase in that frequency due to increased
exposure to methylmercury.
The studies in the Iraqi population (typical of category one) were
made 2-3 months after cessation of exposure and in most cases,
sometime after the onset of signs and symptoms in the patients. Thus
the investigators were faced with the problem of recapitulation of
exposure and of determining the dose received by the individual.
Bakir et al. (1973) classified the population into cohorts
according to blood levels of mercury. The first blood samples were
collected at various times after cessation of exposure. They were
corrected to an average point in time corresponding to 65 days after
cessation of exposure using a clearance half-time of 65 days. Back
extrapolation to times earlier than 65 days was not attempted because
the pattern of mercury clearance from blood was not established for
this period. In fact it was noted that in 11 cases, blood mercury
concentrations did not exhibit statistically significant decline
during the first 20 days of March 1972. These individuals had stopped
consumption of bread 45 days before collection of the first blood
Thus, instead of attempting to back extrapolate the blood samples
to the time of onset of symptoms (for example in the Niigata studies
to be discussed below), a different approach was made to recapitulate
exposure. It was noted that 58 individuals (approximately half the
population studied by Bakir et al., 1973) gave sufficient information
on their consumption of contaminated bread to allow estimation of the
ingested dose of mercury. When the blood levels in these individuals,
corrected to the time point of 65 days after exposure, were plotted
against the ingested dose as reported by the patients, the
relationship between blood levels and estimated dose was linear for
both adults and children, the results for the children giving a
steeper slope consistent with a smaller volume of distribution of
methylmercury. The correlation coefficient for people over 18 years of
age was 0.85 and for people of 10-15 years it was 0.89 (Fig. 2). This
empirical correlation between blood mercury and ingested dose was used
to estimate ingested dose from observed blood levels (corrected to 65
days after exposure).
Bakir et al. (1973) proceeded to estimate the average amount of
methylmercury for each group of the people by use of the exponential
equation (equation 1) discussed in section 6. They assumed complete
absorption of methylmercury from the diet and that the average
elimination half-time was 70 days. This estimated body burden, plotted
on a logarithmic scale, was related to the frequency of signs and
symptoms in each group of the population (see Fig. 3). The signs and
symptoms were paraesthesia, ataxia, visual changes, dysarthria,
hearing defects, and death. It was noted that there was a background
frequency of signs or symptoms that were not related to the mercury
level, and that the frequency increased in relation to the mercury
levels, at high doses of mercury. The thresholda body burden for this
mercury-related increase in frequency was estimated for each sign of
symptom. Paraesthesia had the lowest threshold body burden. From the
dose-response curve, the onset of symptoms was estimated to occur at
approximately 25 mg of mercury or 0.5 mg/kg body weight.
Bakir et al. (1973) noted that the empirical relationship between
blood level and ingested dose did not conform with the relationship
expected from Miettinen's tracer studies on volunteers. Bakir et al.
(1973) pointed out that this difference may be due "either to
differences in conditions of exposure between the Iraqi patients and
the volunteers given labelled methylmercury, or to underestimations of
dose in Iraq, or to both causes". They noted that, in 14 patients, the
average exposure period reported by the patients was 48 days as
compared with 66 days calculated from hair analysis of the same
patients. Thus, an alternative dose-response relationship was plotted
by increasing the estimates of ingested dose by a factor of about 1.6,
which made their empirically observed blood-ingested dose relationship
identical to that calculated from Miettinen. This yielded a threshold
body burden for paraesthesia of 40 mg of mercury or 0.8 mg/kg body
An alternative calculation may be made directly from the blood
level, to try to estimate the minimum concentration of mercury in
blood at which paraesthesia became detectable. If the mean values
(estimated as geometric means) of the blood mercury for each cohort,
as reported by Bakir et al., are plotted on a logarithmic scale
against the frequency of paraesthesia, the relationship has basically
the same pattern as observed when body burden was used (Fig. 4). The
horizontal portion of the line relates to the background frequency at
a mean blood mercury level of between 1.0 and 22 µg/100 ml. The next
points lie significantly above the background frequency level. A
least-squares linear regression line through these points intercepts
the background level at a mercury level of approximately 24 µg/100 ml.
However, the estimated threshold value is for mercury concentrations
in the blood 65 days after cessation of exposure. If correction is
made using a 65 day clearance half-time, the actual threshold level
could be twice as high, i.e. 48 µg/100 ml. Evidence noted above
suggested that blood mercury may have been cleared at longer
half-times than 65 days. Thus the actual threshold mercury value for
blood probably lies between 24 and 48 µg/100 ml.
a The phrase "threshold body burden" is meant to indicate the value
of the body burden at which effects due to methylmercury become
detectable above the background frequency. It is not intended to
mean that methylmercury does not produce effects in some
individuals below this level.
Mufti et al. (1976) and Kazantzis et al. (1976) have reported the
results of a survey of 956 persons in a heavily affected village in
Iraq, on an additional 207 persons living nearby, and on 1014 persons
in a control village that did not receive the treated grain. Mufti et
al. (1976) reported on 427 persons who had eaten contaminated bread.
They were divided into groups according to the total consumption of
contaminated bread (loaves per day x period of consumption) and the
frequency of parasthesia was reported for each group. The total
quantity of mercury consumed can be estimated from the mercury content
of each load.
Figure 5 is a plot of the frequency of paraesthesia in this
population against the log of the mean total ingested dose for each
group plotted on a logarithmic scale. A horizontal line is drawn
through the point corresponding to 4% paraesthesia assuming this to be
the background frequency in this population. The points at higher
ingested doses lie significantly above this line. A linear regression
line drawn through these points intercepts the background frequency at
a total infested dose of 37 mg of mercury. Assuming 50 kg to be the
average body weight for this population (Mufti et al., 1976), this
threshold dose would be 0.7 mg/kg. However, during the period of
consumption (average 32 days), some excretion of methylmercury took
place, so that the maximum body burden must have been less than the
total ingested. The difference would be small in view of the short
period of consumption.
Thus the study of this large population, carried out approximately
6 months after the study by Bakir et al. (1973), would also be
compatible with a body burden of mercury in the range of 0.5-0.8 mg/kg
Shaharistani et al. (1976) have reported on 184 persons in rural
Iraq, 143 of whom consumed the contaminated bread.a The signs and
symptoms were classified as mild, moderate, and severe. People
classified as mild cases complained of numbness of the extremities and
had slight tremors and mild ataxia. Moderate cases had difficulty of
hearing, tunnel vision, and partial paralysis. The severe cases
generally suffered from a combination of the following; complete
paralysis, loss of vision, loss of hearing, loss of speech, and coma.
The dose was expressed as the peak hair concentration of mercury
determined by neutron activation of consecutive 1-cm segments of the
hair sample, and graphical determination of the peak concentration as
described by Giovanoli & Berg (1974).
a The clinical observations were made by a local physician in the
rural district and by a resident physician of the hospital where
the hair samples were collected (Shahristani, personal
Shahristani et al. (1976) did not formulate the usual
dose-response relationships. Instead the population was classified
according to the signs and symptoms of poisoning into four groups (no
symptoms, mild, moderate, and severe symptoms). Figure 6 is redrawn
from a similar figure presented by Shahristani et al. to indicate the
range of peak mercury concentrations in hair for each group. The group
having no signs and symptoms attributable to mercury poisoning had
hair values in the range 1-300 mg/kg, the mild group in the range of
120-600 mg/kg, the moderate in the range 200-600 mg/kg and the severe
in the range 400-1600 mg/kg. Unfortunately, insufficient information
is given in the paper to allow formulation of the usual dose-response
relationship in which the population is classified according to dose,
and so these results cannot be compared quantitatively with the
results of Bakir et al. (1973) and Mufti et al. (1976). However, they
do indicate that mild cases have been reported with peak hair levels
as low as 120 mg/kg.
Shahristani et al. gave sufficient information on the 30 cases in
the group they studied to allow estimation of their daily intake of
methylmercury from contaminated bread. It was possible to compare the
mercury concentrations in hair, as they increased during the ingestion
period, with the daily intake of mercury. It was found that the
concentration of mercury in hair was related to daily intake by an
exponential equation similar to that described in section 6 relating
body burden to daily intake. Thus they were able to calculate the
ratio of mercury concentration in hair (mg/kg) to average
concentration of mercury in the body (mg/kg). The average ratio was
found to be 137 with a range of 82-268 in the 30 individuals. Using
this average ratio, the hair level of 120 mg/kg at which mild symptoms
were first observed would be equivalent to a body burden of mercury of
0.8 mg/kg body weight.
Observations on the Niigata outbreak of methylmercury poisoning
included figures on concentrations of mercury in samples of blood and
hair as well as detailed clinical reports. A Swedish Expert Group
(1971) estimated the blood levels in patients at the time of onset of
symptoms. This involved a graphical procedure in which the
concentration of mercury in samples of blood, collected from the
patient at various times after admission to hospital, were plotted
against the time of collection on semilogarithmic paper (Fig. 7). The
decline in blood levels of mercury corresponded to a clearance
half-time of 70 days, although there was considerable scatter of the
data. Back extrapolation of time of onset of symptoms revealed that
the lowest group (about 3 patients) had blood mercury levels in the
range of 20-40 µg/100 ml.
Data on hair concentrations in the Niigata outbreak were analysed
and discussed by the Swedish Expert Group (1971). One hair sample,
collected close to the time of onset of symptoms, contained mercury at
52 mg/kg. Unfortunately, no corresponding blood sample was available.
Accurate comparison of hair to blood concentrations in the Niigata
samples was not possible because the hair specimens were not
representative of the current blood levels. Tsubaki (1971) has
presented data relating fish consumption to hair values in the Niigata
outbreak which also identified cases of poisoning (also published by
the Swedish Expert Group, 1971). The hair concentrations cannot be
related to the signs and symptoms because they were not extrapolated
back to the time of onset of symptoms.
Several fish-eating populations (category 3) have been examined
for signs and symptoms of poisoning, and determinations made of
concentration of mercury in samples of blood and hair. A Swedish
population of fish eaters had blood levels of mercury up to
56 µg/100 ml. Skerfving (1972) reported a dose-response curve
calculated from data for this population, which had no cases of
mercury poisoning, and for the Niigata cases reviewed above. The
frequency of signs and symptoms was related to concentrations of
mercury in the hair. The apparent threshold effect corresponded to
50-90 mg/kg in the hair and thus to a blood mercury level of
20-36 µg/100 ml.
The studies by Turner et al. (1974) and Marsh et al. (1974) on
populations of fish eaters in Peru and Samoa are consistent with the
relationship published by Skerfving (1972). Thus Turner et al. noted
that 8 people had mercury blood levels between 20 and 30 µg/100 ml and
Marsh et al. also found 2 people having blood levels in the same
Paccagnella et al. (1974) could not find any connexion between the
prevalence of neurological defects and concentrations of mercury in
samples of blood and hair in the high fish consumers on S. Pietro
Island. No neurological deficits were reported in the three
individuals having blood mercury levels between 20 and 24 µg/100 ml.
However the dose-response data reviewed above indicate that the risk
of poisoning at these levels is small.
Table 4 records the results obtained in studies on the populations
discussed above. The studies on the Niigata population and the report
of Shahristani et al. on the Iraqi population identify cases of
methylmercury poisoning. The quoted levels in hair and blood are those
seen in the most sensitive individuals in that population. The results
(Table 4) indicate that such "sensitive" individuals may exhibit
symptoms of methylmercury poisoning at blood mercury levels in the
range of 20-40 µg/100 ml and at hair levels of 50-60 mg/kg.
Unfortunately these studies do not indicate the percentage of the
general population that is sensitive. Studies on high fish consumers
in Sweden, Peru, Samoa, and Italy, suggest a low probability of
symptoms at mercury levels of 20-40 µg/100 ml blood. At least 15
persons having blood levels in this range did not exhibit symptoms of
Table 4. Summary of concentrations of mercury in samples of blood and hair and the body
burden of mercury associated with effects (usually paraesthesia) in the most
sensitive group in the populationa
Total No. Blood Hair the body
Population studied (µg/100 ml) (mg/kg) (mg/kg) References
Niigata 17 20-40 52 -- Swedish Expert Group
Iraq 184 -- 120 0.8 Shahristani et al.
Iraq 125 24-48 -- 0.5-0.8 Bakir et al. (1973)
Iraq 427 -- -- 0.7 Mufti et al. (1976)
a The numbers quoted in this table should not be considered independently of the
The studies by Bakir et al. (1973) and Mufti et al. (1976) on the
Iraqi population followed a different approach. Individual cases of
methylmercury poisoning were not reported. Instead, the population was
divided into groups according to observed levels of mercury in the
blood and according to estimated dose, and the frequency of specific
signs and symptoms was reported for each group. Using this approach,
it is necessary to distinguish between the background frequency of
signs and symptoms and the increase of frequency due to methylmercury.
The figures quoted in Table 4 represent a graphical estimate (see
Fig. 3, 4, and 5) of the blood level, a body burden or dose of
methylmercury where the frequency of paraesthesia emerges above the
background level. Variation in observed background frequencies ranged
from 4% (Fig. 5) to 9.5% (Fig. 3), thus setting a practical limit to
the accuracy of such graphical estimates. Thus the numbers quoted in
Table 4 are compatible with frequencies of paraesthesia due to
methylmercury of 5% or less.
The data in Table 4 apply only to neurological signs and symptoms
in adults. They do not apply to infants exposed either prenatally or
in the early postnatal period.
Skerfving et al. (1974) have reported on the cytogenetic effects
of methylmercury in 23 people exposed through intakes of various
amounts of fish containing methylmercury (0.5-7 mg/kg) and in 16
people with a low or moderate intake of mainly oceanic fish. The
mercury levels in blood in the "exposed" subjects ranged from 1.4 to
11.6 µg/100 ml and in the "non-exposed" from 0.3 to 1.8 µg/100 ml. A
statistical relationship was found between frequency of chromosome
breaks and blood mercury levels. In a study carried out in Iraq
(Firman, unpublished report), no statistically significant difference
was noted in an exposed group compared with a control population with
regard to chromosome damage.
8.1.3 Children and infants with in utero exposure
Of the cases of poisoning reported from Minamata (Harada, 1968),
23 were due to prenatal exposure to methylmercury. These infants had
severe cerebral involvement (palsy and retardation) whereas their
mothers had mild or no manifestations of poisoning. However, there is
a possibility that slight symptoms in the mothers might have been
overlooked (Harada, 1971). It should also be noted that the infants
were examined some years after birth and that no mercury levels are
available either for the mothers or for the affected infants. A case
of prenatal methylmercury poisoning has been reported for a family
that consumed meat from pigs that ate grain treated with methylmercury
fungicide (Pierce et al., 1972). The mother was exposed in early
pregnancy and had a hair mercury level of 186 mg/kg. It was stated
that the mother had no symptoms other than a slight slur of speech
which occurred during two weeks in early pregnancy. The child, which
was delivered normally after an uneventful pregnancy, exhibited
tremulous movement of the extremities in the first few days of life
and subsequently developed myclonic convulsions (Snyder, 1971). At one
year of age the infant exhibited normal physical growth but could not
sit up and was blind.
Cases of prenatal poisoning have been referred to in a preliminary
report on the Iraqi epidemic where it was noted that blood levels in
the infants at birth and in the first few months after birth could be
considerably higher than those of the mother (Bakir et al., 1973).
Quantitative data on the prenatal exposure of the infants is not yet
available. These observations, however, have led to the belief that
prenatal life in man is more sensitive than adult life but the
difference in the degree of sensitivity has not yet been
Amin-Zaki et al. (1974a) have reported clinical examinations of 15
infant-mother pairs in which the infant was exposed prenatally to
methyl-mercury. Clinical manifestations were evident in 6 of the
mothers and in at least 6 of the infants. Five of the infants were
severely affected having gross impairment of motor and mental
development. However, in only 1 infant-mother pair was the infant
affected and the mother free of signs and symptoms.
These observations from the 1971-72 Iraq outbreak must be regarded
as preliminary. It may take time for some of the consequences of
prenatal exposure to manifest themselves. Harada (1971) scrutinized
the chromosomes of 7 victims with congenital Minamata disease and 1
infant victim with severe noncongenital Minamata disease, and reported
that the chromosome patterns were within normal range.
8.2 Clinical Studies of Effects of Mercury-binding Compounds
Attempts to treat mercury poisoning have generally involved the
use of antidotes that reduce the amount of mercury in the target
tissue, either by forming an inactive complex with mercury, or by
enhancing its removal from the tissue. Such antidotes are of course
used in conjunction with general supportive therapy. Ideally the
antidote should have a sufficiently high affinity for mercury so that
nontoxic doses are able to remove mercury from tissue binding sites.
The mercury chelate so formed should be less toxic than mercury and
preferably should be rapidly excreted. The agent should be
metabolically stable so that dosing should not be too frequent and
preferably the agent should be given by mouth. These antidotes are
most effective when given early after exposure to mercury. Clearly the
removal of mercury is without much advantage if irreversible damage
has already occurred.
The first effective antidote, 2,3-dimercaptopropanol (British
Antilewisite-BAL), is a sulfur-containing compound (a dithiol
molecule, possessing a remarkably high affinity for divalent ionic
mercury) and was developed on the basis that mercury and other heavy
metals combined with sulfur groups in the body (for detailed review,
see Levine, 1970). This compound is life saving in cases of acute
mercury(II) chloride poisoning, alleviates symptoms from overdoses of
mercury diuretics, and dramatically relieves certain symptoms of
acrodynia. For alkylmercury poisoning BAL is contraindicated since it
increases brain mercury levels (Berlin & Rylander, 1964; Magos, 1968).
It also does not alleviate neurological disorders caused by mercury
vapour exposure (Hay et al., 1963; Glomme & Gustavson, 1959). Unithiol
(2,3, dimercaptopropansulfonate) is a water soluble derivative of BAL
that is apparently more effective in mobilizing mercury (Trojanowska &
Azendzikowski, in press; Dutkiewicz & Oginski, 1967). Furthermore
unithiol does not produce redistribution to the brain, as has been
observed after BAL treatment. Unithiol is effective in the treatment
of occupational mercurialism (Fesenko, 1969), but there are no reports
on its effects on alkylmercury poisoning.
The penicillamines (D-penicillamine and N-acetyl-DL-
penicillamine) are effective in increasing the excretion of mercury
after exposure to mercury vapour and in relieving the symptoms of
chronic mercury vapour poisoning (Smith & Miller, 1961; Parameshvera,
1967). Fatal brain damage can be prevented in the offspring of rats
treated with methylmercury, by D-penicillamine according to the
experiments of Matsumoto et al. (1967). Recent literature reports
(Bakir et al., 1973; Suzuki & Yoshino, 1969) indicate that the
penicillamines are capable of mobilizing mercury from tissues and
increasing the excretion of mercury in cases of methylmercury
poisoning in man. Thus it appears that the penicillamines offer
advantages over BAL in that they are orally effective, less toxic, and
effective in treating mercury vapour poisoning and probably
alkylmercury poisoning, when administered immediately following
A slight increase in the urinary excretion of mercury has been
noted in methylmercury poisoned patients with "Minamata disease"
(Katsuna, 1968), after administration of EDTA.
Thioacetamide increases urinary excretion of mercury in animals
dosed with mercury(II) chloride but probably one of the major causes
of this effect is kidney damage caused by the combination of the toxic
effects of thioacetamide and mercury leading to an increased
exfoliation of renal tubular cells (Trojanowska et al., 1971).
Some interesting new ideas in the realm of antidotes to mercury
are worth noting. Aaseth (1973) described the application of large
molecular weight mercaptodextran in the successful treatment of
mercury(II) chloride poisoning in animals. This agent does not enter
the intracellular spaces and achieves removal of mercury from the body
without redistribution. However, its effectiveness is limited by the
time of administration. For example, if given more than 2 hours after
exposure to mercury, this compound is totally ineffective whereas BAL
is still useful.
A second approach is to give a nonabsorbable mercury-binding
compound in the diet, in order to trap the mercury that is secreted in
the bile, to prevent its reabsorption, and to greatly increase the
faecal elimination (Takahashi & Hirayama, 1971; Clarkson et al.,
1973a). A polystyrene resin containing fixed sulfhydryl groups has
been shown to increase mercury excretion in experimental animals given
methylmercury, and to reduce blood mercury levels in the victims of
methylmercury poisoning in Iraq (Clarkson et al., 1973a; Bakir et al.,
1973). There is, however, variation in response among patients. More
recently it has been demonstrated that phenobarbital can increase the
biliary excretion of methylmercury compounds (Magos et al., 1974).
A new technique for the removal of methylmercury directly from
blood has been proposed for use in methylmercury poisoning (Kostyniak
et al, 1975). The simultaneous combination of extracorporeal regional
complexation of methylmercury with haemodialysis has been effective in
producing a rapid removal of mercury in both experimental animals and
man (Kostyniak et al., 1974; Al-Abbasi et al., 1974, unpublished
8.3 Pathological Findings and Progression of Disease
The signs and symptoms of acute toxicity, severe gastrointestinal
damage, shock, cardiovascular collapse, and acute renal failure, after
large doses of divalent mercury and pulmonary irritation after
inhalation of massive doses of the vapour, reflect the fact that all
mercury compounds are chemically reactive, can denature proteins,
inactivate enzymes, and disrupt cell membranes, leading to cellular
death and the destruction of any tissue with which they come into
contact in sufficient concentration. The pathological effects
following long-term exposure to lower doses of mercury compounds are
more subtle and depend upon the type of mercury compound to which the
subjects are exposed. The remainder of this section will be concerned
with long-term exposure, as this type of exposure is most important in
helping to assess the hazards to man of the presence of mercury in
food and air.
Pathological findings on human subjects exposed to mercury and its
compounds have been reviewed by a Swedish Expert Group (1971) and by
Friberg & Vostal (1972).
Pathological findings demonstrate that methyl- and ethylmercury
compounds are primarily neurotoxic and produce similar types of lesion
in man. The main pathological features consist of the destruction of
neurological cells in the cortex particularly in the visual areas of
the occipital cortex and various degrees of damage to the granular
layer in the cerebellum. Damage to the peripheral nerves may occur as
indicated by clinical signs but no definitive pathological
observations are available for man. Takeuchi (1970) has reported on
changes in the diameter of the peripheral nerves in patients suffering
from heavy exposure to methylmercury in the Minamata Bay epidemic.
However, Von Burg & Rustam (quoted by Bakir et al., 1973) could not
find any changes in conduction velocities in the patients in Iraq who
received very high exposure to methylmercury.
Brain concentrations of mercury associated with the onset of
pathological changes following exposure to the vapour or to doses of
the alkylmercury compounds are not fully known. Takahata et al. (1970)
demonstrated mean mercury levels of 11 mg/kg wet weight in the brains
of two workers, poisoned by exposure in a mercury mine, who had had no
known exposure to mercury for 10 years prior to death. Studies on
occupationally poisoned individuals (Swedish Expert Group, 1971), and
the patients who died in the Minamata epidemic indicate that the onset
point of signs and symptoms corresponds to an average brain level of
approximately 5 mg/kg wet weight. These findings are in general
agreement with threshold methylmercury concentrations in studies on
experimentally poisoned animals.
8.3.1 Psychiatric and neurological disturbances
Trahtenberg (as reviewed by Friberg & Nordberg, 1972) reported
minor psychiatric disturbances such as insomnia, shyness, nervousness,
and dizziness in workers exposed to elemental mercury vapour
concentrations of the order of 0.1 mg/m3. The classical literature
(for review, see Friberg & Nordberg, 1973) contains detailed accounts
of the consequences of long exposure to higher concentrations of
elemental mercury vapour where the full syndrome of erethism is seen.
Individual variation in exposed people is the rule but the most
commonly reported syndrome includes loss of memory, insomnia, lack of
self-control, irritability and excitability, anxiety, loss of
self-confidence, drowsiness, and depression. In the most severe cases
delirium with hallucinations, suicidal melancholia, or even
manic-depressive psychoses have been described.
The presence of tremor is one of the most characteristic features
of mercurialism and usually follows the minor psychological
disturbances referred to above. With continuing exposure to elemental
mercury vapour, the tremor develops gradually in the form of fine
trembling of the muscles interrupted by coarse shaking movements every
few minutes. It may be seen in the fingers, but also on the closed
eyelids, lips, and on the protruding tongue. The frequency is of the
order of 5-8 cycles per second. It is intentional and stops during
sleep. On cessation of exposure, the tremor gradually disappears.
Dramatic alterations in the steadiness of the handwriting may be seen
in persons suffering from mercurial tremor.
The most common signs and symptoms in cases of poisoning due to
methyl- or ethylmercury compounds are paraesthesia, loss of sensation
in the extremities and around the mouth, ataxia, constriction of the
visual fields, and impairment of hearing. In the Japanese experience
the effects of alkylmercury poisoning are usually irreversible but the
coordination may improve after rehabilitation (Kitagawa, 1968). On the
other hand, in Iraq, improvement in motor disturbances was often
spontaneous. In addition, paraesthesia was often reversible in Iraq
but a persistent symptom in Japan (Damluji, 1974; Tsubaki, 1971).
8.3.2 Eye and visual effects
Occupational exposure to elemental mercury vapour causes the
appearance of a greyish-brown or yellow haze on the anterior surface
of the lens of the eye (Atkinson, 1943). It appears usually after
long-term exposure and the depth of colour depends upon the length of
time and the air concentration of mercury to which the worker has been
exposed. The presence of this coloured reflex may or may not be
associated with signs and symptoms of poisoning.
The narrowing of the visual fields is a classic sign of poisoning
due to short-chain alkylmercurial compounds (Hunter et al., 1940). In
cases of severe exposure, the constriction may proceed to complete
8.3.3 Kidney damage
Kazantzis et al. (1962) reported four cases of proteinuria in two
groups of workmen exposed to elemental mercury vapour. Exposure
conditions were not specified, but all four cases were excreting
mercury in urine in excess of 1000 µg/litre at the time of the first
examination. The proteinuria disappeared after the workers were
removed from exposure. Joselow & Goldwater (1967) noted that the mean
urinary protein concentration (90 mg/litre) in a group of workers
exposed to elemental mercury vapour was significantly higher than the
mean protein concentrations (53 mg/litre) in a nonexposed group. The
urinary protein correlated with urinary mercury levels.
Renal involvement after exposure to methylmercury compounds is
very rare (Bakir et al., 1973). Cases of renal damage have been
reported in an outbreak of ethylmercury poisoning (Jalili & Al-Abbasi,
8.3.4 Skin and mucous membrane changes
Dermatitis has been reported after occupational exposure to
phenylmercurials (for review, see Goldwater, 1973). However dermatitis
may result from exposure to inorganic mercury (Hunter, 1969).
Sensitivity to metallic mercury in tooth fillings has resulted in
facial and intra-oral rashes. Skin reactions due to sensitivity to
phenylmercury have also been reported (for review, see Clarkson,
Dermatitis has been reported after skin contact with the
alkylmercurials (for review, see Swedish Expert Group, 1971). Oral
ingestion of methyl- and ethylmercury compounds may also result in
this condition, as observed in the Iraqi epidemics (Jalili &
Al-Abbasi, 1961; Damluji et al., 1976).
9. EVALUATION OF HEALTH RISKS TO MAN FROM EXPOSURE TO MERCURY
AND ITS COMPOUNDS
9.1 General Considerations
In order to control risks to human health from the presence of
mercury in the environment, it is necessary to attempt to define the
degree of risk in any given environmental situation. The health risk
evaluation depends in part on a knowledge of dose-effect,
dose-response relationships. It also depends on a knowledge of the
variation in exposure (or intake) in any given situation. A general
review of sampling and analytical techniques (section 2), and of
environmental sources and exposure levels (sections 3 and 5) has been
presented in this document. However, specific assessment of exposure
or intake must be made by the public health authorities responsible
for any given population or group. This section of the report is
concerned, therefore, primarily with a summary of those dose-effect,
dose-response relationships that are relevant to human populations in
both occupational and general environmental exposure to mercury and
Species differences in the metabolism and toxicity of mercury and
its compounds are so great that this health risk evaluation is based
primarily on data for man. Animal data have been included or
considered only when data for man are lacking, and have generally been
used in a qualitative way. Thus animal data may in certain cases
indicate the potential for genetic effects, or that a certain stage of
the life cycle, for example, the fetus appears to be the most
sensitive stage. General patterns of deposition of mercury in the
body, as seen from animal experiments, have been considered to be
broadly applicable to man in a qualitative sense.
Quantitative estimates of the dose-response relationship in man
are fraught with many difficulties. Observations of
occupationally-exposed persons should offer the best possibilities for
well controlled studies. However, the exposure range and population
sizes are limited and difficulties are usually encountered in
obtaining accurate estimates of time-weighted average exposure to
airborne mercury. The most difficult situation for controlled study
arises in the case of methylmercury where our knowledge derives
primarily from observations on the unexpected outbreaks of poisoning
from contaminated food in Japan and Iraq. Generally the studies
commenced some time after the start of the outbreak or after the end
of the exposure. Attempts had to be made either to recapitulate
exposure by the back-extrapolation of blood levels, or to estimate the
ingested dose from the patient's ability to remember. Doubts have also
been expressed on the reliability of analytical methods in the earlier
Estimates will be made of minimum effect exposures or
concentrations in indicator media since these numbers may be of value
to authorities in setting safety standards. These minimum effect
figures (e.g. Tables 5 and 6) should be viewed in terms of the overall
dose-response relationship, namely, that as the dose is decreased, so
also is the probability of poisoning, and that the minimum effect
level is that dose (expressed as exposure level, daily intake,
concentration in indicator media) that is associated with the first
detectable effect in the population under study. The effect will be
present at some specified frequency in the population. The value of
the minimum effect dose is the dose derived from the observed
dose-response relationship. (It will be subject to more statistical
uncertainty than, for example, the dose giving 50% frequency of the
effect.) Its value will depend on the size of the population under
study; the larger the population, the more likely it will be that
effects will become evident in the more sensitive individuals.
9.1.1 Elemental mercury vapour
The central nervous system is the critical organ for the toxic
effects of inhaled elemental mercury. Effects on the kidney such as
proteinuria have been reported but only at doses higher than those
associated with the onset of signs and symptoms from the central
nervous system. No information is available with regard to mercury
levels in the central nervous system at the time of onset of signs and
symptoms or at death in man. Furthermore, we do not have indicator
media such as urine or blood the mercury levels of which would reflect
those in the brain. Observations on animals exposed to elemental
mercury vapour reveal large regional differences in distribution
within the brain so that average brain concentrations, even if they
were available for man, might not be of much value.
Our evaluation of the health risks from exposure to elemental
mercury vapour in man has therefore to be based on an empirical
relationship between air levels (exposure) and the frequency of signs
and symptoms in exposed populations. Concentrations of mercury in
urine and blood are related, on a group basis, to average air
A complete metabolic model relating air levels to absorption,
accumulation, and excretion of mercury in man following exposure to
elemental mercury vapour is not available. However, some useful
generalizations are beginning to emerge from recent observations. Most
of the inhaled vapour (approximately 80%) is retained in the lung.
Once absorbed into the blood stream, it is rapidly oxidized to ionic
mercury. The limited information on biological half-times in man
suggests that a workman exposed to a constant average concentration of
mercury vapour in his working environment would not reach a state of
balance (steady state) until after one year of exposure. Consequently
one would expect the concentrations of mercury in blood or urine to
exhibit a consistent relationship to air levels after the worker had
been exposed for at least one year. Unfortunately, most of the
publications in the literature do not indicate the period of
employment of the worker. However, general experience in occupational
health studies (for review, see MAC Committee, 1969) reveals that
exposure of workers to an average air concentration of mercury of
0.05 mg/m3 are associated, on a group basis, with blood levels of
approximately 3.5 µg/100 ml, and with urinary concentrations of
150 µg/ litre. Linear relationship has been observed between urinary
and blood concentrations of mercury in people occupationally exposed
to elemental mercury vapour. Thus, measurements of mercury
concentrations in the working atmosphere and in samples of blood and
urine from the workmen may be used as an index of exposure. These
values in turn may be compared with the frequency of clinical signs
and symptoms in workers experiencing different degrees of exposure.
Such studies, reported in detail in section 8 and briefly reviewed
below, form the basis for establishing "threshold limit values" or
maximal allowable air concentrations of mercury in occupational
Early studies dating back to the 1930s indicate that cases of
poisoning occurred at atmospheric mercury levels above 0.1 mg/m3 (for
details, see section 8.1.1). Recent data also demonstrate that there
was an increase in complaints of appetite-loss and insomnia in a group
of workers exposed to time-weighted average air concentrations of
mercury between 0.06 and 0.1 mg/m3 as compared with two lower
exposure groups (0.01 and 0.05 mg/m3). These findings are in
agreement with data published in the 1940s and 1950s indicating that
mercury intoxication occurred in workers exposed to mercury in air
concentration less than 0.2 mg/m3, but no data were given on the
lower exposure limits.
At this time it is difficult, if not impossible, to establish a
lower limit at which no effects occur. Studies reviewed in section
8.1.1 indicate that mental disturbances may be seen at extremely low
mercury concentrations in air. However, the problem in the
interpretation of these reports is that, as the air concentration of
mercury decreases, it becomes more difficult to correlate effects with
exposure with any degree of confidence. For example, it would appear
that effects of mercury levels below 0.05 mg/m3 have not been
Concentrations of mercury in blood and urine equivalent to average
mercury concentrations in air of 0.05 and 0.1 mg/m3 are given in
Table 5. In general these relationships are not seen in indivduals but
only when averaged over a substantial number of workers.
An occupational limit for mercury in air concentration of
0.05 mg/m3 would be equivalent to an ambient air level for the
general population of approximately 0.015 mg/m3. This calculation is
based on the assumptions of a daily ventilation of 10 m3 during
working hours and 20 m3 for a 24-hour day and that there are 225
working days in the year. The data of Smith et al. (1970) would
indicate that the probability of seeing adverse effects at air levels
of, 0.05 mg/m3 for occupational exposure, and 0.015 mg/m3 for
continuous environmental exposure, is low for such symptoms as loss of
appetite, weight loss, and shyness. However, this calculation does not
take into account the sensitive groups in the general population. It
should be noted that concentrations found in ambient air are far below
these levels (see section 5).
Table 5. The time-weighted average air concentrations associated
with the earliest effects in the most sensitive adults
following long-term exposure to elemental mercury vapour.
The table also lists the equivalent blood and urine
Air Blood Urine Earliest effects
(mg/m3) (µg/100 ml) (µg/litre)
0.05 3.5 150 non-specific symptoms
0.1-0.2 7-14 300-600 tremor
a Blood and urine values may be used on a group basis owing to
gross individual variations. Furthermore, these average values
reflect exposure only after exposure for a year or more. After
shorter periods of exposure, air concentrations would be
associated with lower concentrations in blood and urine.
9.1.2 Methylmercury compounds
The estimate of risks to human health from methylmercury compounds
is important for several reasons. First, many thousands of people have
been poisoned following accidental consumption of food contaminated
with methylmercury fungicides or the consumption of fish contaminated
by industrial release of methylmercury. Second, methylmercury probably
accounts for a significant part of mercury in the human diet and is
especially important in fish and fish products. Third, the
risk-benefit calculations with regard to methylmercury in fish are of
critical importance in those countries and areas of the world where
fish is an important dietary source of protein, or where the fish
industry is of economic importance.
This evaluation of risk from exposure to methylmercury compounds
is based primarily upon data in man. The data consist of observations
on the frequency of signs and symptoms in populations exposed to a
wide range of mercury intake, observations on the concentrations of
mercury in hair and blood samples, and on estimates of dietary intake
In estimates of risks to human health, the custom has been
followed of attempting to determine the lowest concentration in
indicator media or the lowest daily intake associated with the onset
of toxic signs and symptoms in man, along with information on maximal
intakes which produce no effects. This procedure has been followed in
estimating the data presented in Table 6. The levels in blood and
hair, and the amount of the body burden of mercury associated with the
onset of signs and symptoms are taken from Table 4.
The reports on the Minamata outbreak to the effect that infants
were born having cerebral palsy due to methylmercury whereas their
mothers lacked or had only slight symptoms led to the belief that the
fetus was the stage of the life cycle that was most sensitive to
methylmercury. Studies on animals, exposed during the gestation period
to methylmercury led to qualitative confirmation of this conclusion.
Table 6 lists our conclusions on concentrations of total mercury
in indicator media associated with the earliest effects of
methylmercury in the most sensitive group in the adult population. As
discussed below, the prevalence of the earliest effects would be
expected to be approximately 5%. The equivalent long-term daily intake
quoted in Table 6 was calculated on the most conservative relationship
(quoted in Table 3) i.e. that the steady state blood concentration
(ng/ml) is numerically equal to the average daily intake (µg/day/70 kg
However, Nordberg & Strangert (1976) have proposed an alternative
approach to estimating risks of poisoning on long-term exposure. The
relationship between long-term daily intake and steady state blood
levels (and hair levels and body burden) depends, inter alia, on the
biological half-time of methylmercury in man. Biological half-time
times are subject to individual variation as discussed in section 6.
Thus an individual having a long biological half-time (a slow excretor
of methylmercury) would accumulate higher steady state levels than one
having a short biological half-time. Thus the statistical distribution
of biological half-times should be taken into account in estimates of
risk of poisoning.
Variations also occur in threshold concentrations for the
appearance of signs and symptoms in individuals in the population.
These may be estimated from empirical relationships relating frequency
of signs and symptoms to concentrations of methylmercury in indicator
media (blood and hair) or body burdens (e.g. Fig. 3, 4, and 5). On the
assumptions that the distribution of biological half-times was normal,
that the distribution of individual threshold values for paraesthesia
was log-normal, and that these distributions were independent of each
other, Nordberg & Strangert (1974) estimated the overall probability
of an individual developing symptoms of paraesthesia. Their
calculations were based on data of Shahristani & Shihab (1974) and
gave the distribution of biological half-times estimated from analysis
of hair samples after the Iraqi outbreak, and the statistical
distribution of individual threshold values estimated from the data of
Bakir et al. (1973) on the relationship of frequency of paraesthesia
to body burdens. Their results indicated that with a long-term daily
intake of 4 µg/kg would yield the risk of paraesthesia of about 8%.
Subsequent calculations based on the data by Mufti et al. (1976) would
indicate a risk of between 3% and 4% for the same daily intake.a
The figures estimated by Nordberg & Strangert may somewhat
overestimate the risk of poisoning. The errors to be expected in both
field and laboratory observations would tend to decrease the slope in
dose-response relationships. This would lead to an overestimate of the
variance of threshold values in the general population. Nevertheless,
the estimate of risk by Nordberg & Strangert is in reasonable
agreement with the more empirical estimates discussed in section 8 and
indicates that, with a long-term daily intake as listed in Table 6,
the prevalence of the earliest effects could be expected to be 5% or
Occupational hazards have arisen mainly from airborne
concentrations of alkylmercurials. Skin contact has also been noted,
but the quantitative importance of skin contact and percutaneous
absorption cannot be estimated. Hazards from occupational exposures
can be estimated only by reference to data already discussed above
with respect to dietary intake of methylmercury compounds. The data
summarized in Table 6 indicate that the first effect (paraesthesia)
associated with long-term intake of methylmercury arises at an intake
level of approximately 5 µg/kg body weight per day. Assuming a daily
ventilation at work of 10 m3 of air, 80% retention of the inhaled
mercurial, 225 working days to the year, the average time-weighted air
concentrations that would give rise to this intake would be
0.07 mg/m3. Consideration of occupational risks should also take into
account the possibility of a high but brief exposure to methylmercury
a Nordberg & Strangert, personal communication.
Table 6. The concentrations of total mercury in indicator media
and the equivalent long-term daily intake of mercury as
methylmercury associated with the earliest effects in
the most sensitive group in the adult populationa,b
Concentrations in indicator media
Blood Hair Equivalent long-term daily intakec
(µg/100 ml) (µg/g) (µg/kg body weight)
20-50 50-125 3-7
a The prevalence of the earliest effects could be expected to be
b The WHO Task Group specifically urged that this table should
not be considered independently of the text in section 8.
c A Japanese group has recently concluded that a daily intake of
mercury of 5 µg/kg is the "minimal toxic dose", following a
ten-year follow-up study of the Minamata outbreak (Research
Committee on Minamata Disease, 1975).
The estimates in Table 6 apply only to adults. As discussed
previously, prenatal life may be the stage of the life-cycle most
sensitive to methyl-mercury. It would be prudent therefore to follow
the advice of the MAC Committee (MAC, 1969), not to expose females of
child bearing age occupationally to methylmercury compounds.
Studies by Skerfving et al. (section 8) have indicated that
chromosome breaks may be associated with exposure to methylmercury.
There are, however, other studies in which no such relationship was
found. Furthermore the health significance of chromosome breakage is
not known. However, as reviewed elsewhere in this criteria document,
experiments on animals and other forms of life do indicate the
potential for genetic damage by methylmercury.
9.1.3 Ethylmercury compounds and other short-chain
Insufficient information is available to allow risk calculations
based on data from human exposure to ethyl- or higher short-chain
alkylmercurial compounds. Suzuki et al. (1973) have reported
observations on five patients poisoned with ethylmercury compounds.
The picture of distribution between plasma and red cells and
observations on autopsy tissue indicate that the metabolism and
patterns of distribution of ethylmercury are generally similar to
those of methylmercury compounds. However, in one individual the
clearance-time from blood was only 10 days. Evidence reviewed in
section 6.6 indicates that ethylmercury compounds may be more rapidly
converted to inorganic mercury in the body than methylmercury
compounds. Thus, these limited observations suggest that ethylmercury
compounds are probably less hazardous than the methylmercury compounds
in so much as they remain in the body for a shorter time because of
transformation to inorganic mercury and more rapid excretion. Thus,
standards set for methylmercury compounds will probably be sufficient
to control the hazards from other short-chain alkylmercurial
9.1.4 Inorganic mercury, aryl- and alkoxyalkylmercurials
The risk to human health from long-term ingestion of inorganic,
aryl-, and alkoxyalkyl- compounds of mercury in the diet is difficult
to estimate because there are not any recorded cases of human
poisoning under these circumstances. Data from animals cannot be used
for exact quantitative extrapolation because of species differences in
the metabolism and toxicity of these compounds. However, certain
qualitative conclusions based on animals can probably be extrapolated
to man. For example, the arylmercurials are rapidly converted to
inorganic mercury in mammals. The penetration of mercury across the
blood-brain and placental barriers is less after doses of inorganic
and aryl compounds than after equivalent doses of elemental mercury
vapour and short-chain alkylmercurials. Animal studies indicate that
the kidney is the critical target organ for exposure to inorganic,
alkyl-, and alkoxy-alkylmercurials. Kidney involvement appears to be
minimal in workers exposed to concentrations of mercury vapour
(0.05-0.1 mg/m3) that elicit the first signs and symptoms of damage
to the central nervous system. Thus guidelines for health protection,
set for long-term exposure to elemental mercury vapour, should offer
an even greater safety margin for equivalent exposures to inorganic,
alkyl, and alkoxyalkyl compounds. In fact, recognizing the lower toxic
potential of these forms of mercury, the MAC Committee (1969) advised
a maximum allowable concentration for occupational exposure of
0.1 mg/m3 -- twice as high as that for elemental mercury vapour.
In the discussion of guidelines for exposure to elemental mercury
vapour, it was concluded that long-term exposure of the general
population to 0.015 mg/m3 was equivalent, in terms of average daily
mercury intake, to the occupational limit of 0.05 mg/m3. Assuming an
average pulmonary retention of 80%, the average daily amount entering
the blood stream is 280 µg based on a daily ventilation of 20 m3.
The equivalent daily intake in diet of phenylmercury compounds
would also be about 240 µg, as animal studies indicate virtually
complete absorption in the gastrointestinal tract. The alkoxyalkyl and
aryl compounds are probably absorbed equally well from food. Tracer
studies on volunteers indicate that approximately 10% of inorganic
mercury compounds are absorbed from the diet, so that dietary intakes
approximately ten times greater than those of phenylmercury compounds
would offer no greater risk of poisoning.
A daily intake of mercury of 240 µg is in the same range as the
daily intake listed for methylmercury compounds in Table 6. The
biological half-time for inorganic mercury, based on tracer studies in
man, appears to be less than that for tracer doses ofmethylmercury.
Animal data suggest that aryl and alkoxyaryl compounds have biological
half-times lower than that for methylmercury and similar to that for
inorganic mercury. Thus the long-term ingestion of inorganic, aryl,
and alkoxyaryl compounds should offer no greater hazards and probably
substantially less than the hazards from ingestion of methylmercury
9.2 Summary and Guidelines
In the case of elemental mercury and alkylmercury the Task Group
was able to construct tables (Tables 5 and 6) that related exposure to
symptoms as well as to concentrations in indicator media in the human
body. In the case of inorganic mercury, arylmercurials, and
alkoxyalkyl mercurials, this could not be done because of the
inconsistency in the animal data available as well as a paucity of
data in man.
The tables for elemental mercury and alkylmercury are given above.
In constructing these tables, the Task Group evaluated the results of
studies summarized in this document, and drawing on their experience
and judgement, identified the concentration and amounts of mercury
associated with certain observed effects. There is insufficient
information available to permit precise quantification of this risk.
Usually, the proportion that may be expected to be affected is small.
The expected health effects for elemental mercury at a level in
air of 0.05 mg/m3 have been quoted only for occupational exposure
assuming 8 hours per day and 225 working days a year. The equivalent
environmental mercury levels in air for continuous exposure would be
approximately 0.015 mg/m3 to give the same degree of risk. The urine
and blood values, of course, would be the same as those quoted in
Table 5. Even though the figures do not only take into account
specific sensitive groups, it is highly unlikely that the
concentration in the general environment approaches levels of
The ranges of minimum effect values quoted in Table 6 for
methylmercury reflect the uncertainty in estimations.
Although it was not possible to identify even approximate minimum
effect values for inorganic, aryl-, and alkoxyalkylmercurials, the
Task Group concluded that the limited experience for occupational
exposure suggested that these forms of mercury were probably less
hazardous than either elemental mercury vapour or methylmercury
compounds. Thus the figures for occupational exposure to elemental
mercury vapour given in Table 5 would serve as conservative figures
for occupational exposure to these forms of mercury and those in Table
6 would offer conservative figures for dietary intake.
A Joint FAO/WHO Expert Committee on Food Additives (1972)
established a provisional tolerable weekly intake of 0.3 mg of total
mercury per person of which no more than 0.2 mg should be present as
methylmercury (expressed as mercury); these amounts are equivalent to
5 µg and 3.3 µg, respectively, per kg of body weight. Where the total
mercury intake in the diet is found to exceed 0.3 mg per week, the
level of methylmercury compounds should also be investigated. If the
excessive intake is attributable entirely to inorganic mercury, the
above provisional limit for total mercury no longer applies and will
need to be reassessed in the light of all prevailing circumstances.
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