Polychlorinated biphenyls and terphenyls (EHC 140, 1992, 2nd edition)

INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY

ENVIRONMENTAL HEALTH CRITERIA 140

POLYCHLORINATED BIPHENYLS AND TERPHENYLS
(SECOND EDITION)

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

First draft prepared by Dr S. Dobson, Institute of Terrestrial
Ecology, United Kingdom, and Dr G.J. van Esch, Bilthoven, The
Netherlands

World Health Organization
Geneva, 1993

The International Programme on Chemical Safety (IPCS) is a joint
venture of the United Nations Environment Programme, the International
Labour Organization, and the World Health Organization. The main
objective of the IPCS is to carry out and disseminate evaluations of
the effects of chemicals on human health and the quality of the
environment. Supporting activities include the development of
epidemiological, experimental laboratory, and risk-assessment methods
that could produce internationally comparable results, and the
development of manpower in the field of toxicology. Other activities
carried out by the IPCS include the development of know-how for coping
with chemical accidents, coordination of laboratory testing and
epidemiological studies, and promotion of research on the mechanisms
of the biological action of chemicals.

WHO Library Cataloguing in Publication Data

Polychlorinated Biphenyls and Terphenyls. -- 2nd ed.

(Environmental health criteria; 140)

1.Environmental exposure 2.Environmental pollutants 3.Polychlorinated
biphenyls -- adverse effects 4.Polychlorinated biphenyls -- toxicity
5.Polychloroterphenyl compounds -- adverse effects
6.Polychloroterphenyl compounds -- toxicity I.Series

ISBN 92 4 157140 3 (NLM Classification: QV 633)
ISSN 0250-863X

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CONTENTS

INTRODUCTION

1. SUMMARY AND EVALUATION, CONCLUSIONS, RECOMMENDATIONS
1.1 Summary and evaluation
1.1.1 Introduction
1.1.2 Identity, physical, and chemical properties
1.1.3 Analytical methods
1.1.4 Production and uses
1.1.5 Environmental transport, distribution, and transformation
1.1.6 Environmental levels and human exposure
1.1.7 Kinetics and metabolism
1.1.8 Effects on organisms in the environment
1.1.8.1 Laboratory studies
1.1.8.2 Field studies
1.1.9 Effects on experimental animals and in vitro systems
1.1.9.1 Single exposure
1.1.9.2 Short-term exposure
1.1.10 Reproduction, embryotoxicity, and teratogenicity
1.1.11 Mutagenicity
1.1.12 Carcinogenicity
1.1.13 Special studies
1.1.14 Factors modifying toxicity, mode of action
1.1.15 Effects on humans
1.2 Conclusions
1.2.1 Distribution
1.2.2 Effects on experimental animals
1.2.3 Effects on humans
1.2.4 Effects on the environment
1.3 Recommendations

2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, ANALYTICAL METHODS
2.1 Identity
2.1.1 Chemical formula and structure
2.1.2 Relative molecular mass
2.1.3 Common name
2.1.4 Chemical composition
2.1.5 Technical product
2.1.6 Purity and impurities
2.2 Physical and chemical properties
2.2.1 Log n-octanol/water partition coefficient
2.2.2 Conversion factors

2.3 Analytical methods
2.3.1 Sampling strategy and sampling methods
2.3.1.1 Extraction procedures
2.3.1.2 Sample clean-up
2.3.2 Separation and identification
2.3.2.1 Chromatographic separation
2.3.2.2 Gas-liquid chromatography
2.3.3 Quantification
2.3.4 Accuracy of PCB determinations
2.3.5 Confirmation
2.3.6 Detection limits
2.4 Codex questionnaire on analytical methods
2.4.1 Interpretation and comparability of data
2.5 Activities of the WHO Regional Office for Europe
2.6 Appraisal

3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE
3.1 Natural occurrence
3.2 Man-made sources
3.2.1 Production levels and processes, uses
3.2.1.1 World production figures
3.2.1.2 Manufacturing processes
3.2.2 Uses
3.2.2.1 Completely closed systems
3.2.2.2 Nominally closed systems
3.2.2.3 Open-ended applications
3.2.2.4 Contamination of other compounds
3.2.3 Loss into the environment
3.2.3.1 Routes of environmental pollution
3.2.3.2 Release of PCBs into the atmosphere
3.2.3.3 Leakage and disposal of PCBs in industry
3.2.4 Thermal decomposition of PCBs

4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
4.1 Transport and distribution between media
4.1.1 Transport in air
4.1.1.1 Dry deposition
4.1.1.2 Precipitation deposition
4.1.2 Transport in soil
4.1.3 Transport in water
4.1.4 Transport between media
4.2 Biotransformation
4.2.1 Biodegradation
4.2.1.1 Bacteria
4.2.2 Biodegradation; individual congeners
4.2.2.1 Bacteria
4.2.2.2 Fungi

4.2.3 Photodegradation
4.2.4 Bioaccumulation, distribution in organisms, and elimination
4.2.4.1 Microorganisms
4.2.4.2 Plants
4.2.4.3 Aquatic invertebrates
4.2.4.4 Fish
4.2.4.5 Birds
4.2.4.6 Mammals
4.2.5 Appraisal

5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE
5.1 Levels in the environment
5.1.1 Air
5.1.1.1 Rain and snow
5.1.1.2 Natural gas
5.1.2 Water
5.1.3 Soil
5.1.4 Aquatic and terrestrial organisms
5.1.4.1 Effect of dredging-contaminated sediment on organisms
5.1.4.2 Relationship to lipid content of organisms
5.1.4.3 Residues in different trophic levels and effects of diets
5.1.4.4 Effects of age, sex, and reproductive status on uptake and elimination
5.1.4.5 Time trends in residues
5.1.4.6 Seasonal patterns in residues
5.1.5 Appraisal
5.2 Levels in animal feed
5.3 Levels in human food
5.3.1 General
5.3.2 Drinking-water
5.3.3 Dairy products
5.3.4 Fish and shellfish
5.3.5 Influence of food processing
5.3.6 Food contamination by packaging materials
5.3.7 Appraisal
5.4 General population exposure
5.4.1 Air
5.4.2 Drinking-water
5.4.3 Intake by infants through mother's milk
5.4.4 Infant and toddler total diet
5.4.5 Total intake by adults via food
5.4.6 Total diet/market-basket studies
5.4.7 Total intake of major congeners by adults via food
5.4.8 Time trends in different matrices

5.5 Concentrations in the body tissues of the general population
5.5.1 Adipose tissue
5.5.1.1 PCBs in the fetus
5.5.1.2 Congeners in adipose tissue
5.5.2 Blood of the general population
5.5.3 Human milk
5.5.3.1 Major PCB congeners in human milk
5.5.3.2 Factors that influence the intake of PCBs with milk
5.5.4 Other tissues
5.6 Accidental exposures (Yusho and Yu-Cheng)
5.7 Occupational exposure
5.7.1 Accidental exposure
5.7.2 Occupational exposure during manufacture and use
5.7.2.1 Adipose tissue
5.7.2.2 Blood

6. KINETICS AND METABOLISM
6.1 Absorption
6.1.1 Inhalation
6.1.2 Dermal
6.1.3 Oral
6.2 Distribution
6.2.1 Inhalation (rat)
6.2.2 Oral (rat)
6.2.3 Oral (monkey)
6.2.4 Oral (humans)
6.2.5 Individual congeners of PCBs
6.2.6 Appraisal
6.3 Placental transport
6.3.1 Laboratory animals
6.3.2 Wildlife
6.3.3 Humans
6.4 Excretion and elimination
6.4.1 Following oral dosing
6.4.2 Following parenteral dosing
6.4.3 Humans
6.4.4 Elimination via milk (animals)
6.4.4.1 Elimination via breast milk
6.5 Metabolic transformation
6.5.1 PCBs
6.5.2 Dichlorobiphenyls
6.5.3 Tetrachlorobiphenyls
6.5.4 Hexachlorobiphenyls and higher chlorinated compounds
6.5.5 Retention and turnover
6.5.6 Appraisal

7. EFFECTS ON ORGANISMS IN THE ENVIRONMENT
7.1 Toxicity for microorganisms
7.1.1 Freshwater microorganisms
7.1.2 Marine and estuarine microorganisms
7.1.3 Soil microorganisms
7.1.4 Plankton communities
7.1.5 Interactions with other chemicals
7.1.6 Tolerance
7.2 Toxicity for aquatic organisms
7.2.1 Aquatic plants
7.2.2 Aquatic invertebrates
7.2.2.1 Short- and long-term toxicity
7.2.2.2 Response to temperature and salinity
7.2.2.3 Reproduction
7.2.2.4 Moulting
7.2.2.5 Behaviour
7.2.2.6 Population structure
7.2.2.7 Interactions with other chemicals
7.2.3 Fish
7.2.3.1 Short- and long-term toxicity
7.2.3.2 Carcinogenicity
7.2.3.3 Effects on developmental stages and reproduction
7.2.3.4 Physiological and biochemical effects
7.2.3.5 Behavioural effects
7.2.3.6 Interactions with other chemicals
7.2.4 Amphibians
7.2.5 Aquatic mammals
7.3 Toxicity for terrestrial organisms
7.3.1 Plants
7.3.2 Terrestrial invertebrates
7.3.3 Birds
7.3.3.1 Short-term toxicity
7.3.3.2 Egg production
7.3.3.3 Hatchability and embryotoxicity
7.3.3.4 Eggshell thinning
7.3.3.5 Effects on the male
7.3.3.6 The effects of stress
7.3.3.7 Physiological, biochemical, and behavioural effects
7.3.3.8 Interactive effects with other chemicals
7.3.4 Terrestrial mammals
7.3.4.1 Short-term toxicity
7.3.4.2 Reproductive effects
7.3.4.3 Physiological effects

7.4 Effects on organisms in the field
7.4.1 Plants
7.4.2 Fish
7.4.3 Birds
7.4.4 Mammals
7.4.4.1 Appraisal

8. EFFECTS ON EXPERIMENTAL ANIMALS AND IN VITRO TEST SYSTEMS
8.1 Single exposures
8.1.1 Oral
8.1.2 Inhalation
8.1.3 Dermal
8.1.4 Other routes
8.2 Short-term exposures
8.2.1 Oral
8.2.1.1 Aroclors
8.2.1.2 Individual congeners
8.2.2 Intraperitoneal: reconstituted PCB mixtures
8.2.3 Dermal exposure
8.2.4 Appraisal
8.3 Skin and eye irritation, sensitization
8.4 Reproduction, embryotoxicity, and teratogenicity
8.4.1 Reproduction and embryotoxicity
8.4.1.1 Oral
8.4.2 Teratogenicity
8.4.2.1 Aroclors (oral)
8.4.2.2 Aroclors (subcutaneous)
8.4.2.3 Individual congeners (oral)
8.4.3 Appraisal
8.4.4 Mutagenicity and related end-points
8.4.4.1 DNA damage
8.4.4.2 Mutagenicity tests
8.4.4.3 Cell transformation
8.4.4.4 Cell to cell communication
8.4.4.5 Interaction
8.4.4.6 Cell division parameters
8.5 Carcinogenicity
8.5.1 Long-term toxicity/carcinogenicity
8.5.2 Tumour promotion/anticarcinogenic effects
8.5.3 Initiation, promotion, and other special studies on individual congeners
8.5.4 Skin carcinogenicity
8.5.5 Appraisal
8.6 Special studies: target-organ effects
8.6.1 Liver
8.6.1.1 PCB mixtures
8.6.1.2 Individual congeners

8.6.2 Enzyme induction
8.6.2.1 Effects on liver enzymes of PCBs
8.6.2.2 Effects on liver enzymes of "biologically filtered" PCB mixtures
8.6.2.3 Effects of individual congeners on liver enzymes
8.6.2.4 Appraisal
8.6.3 Effects on vitamins and mineral metabolism
8.6.3.1 Effects of PCB mixtures
8.6.3.2 Effects of individual congeners
8.6.4 Effects on the gastrointestinal tract
8.6.5 Effects on lipid metabolism
8.6.5.1 Effects of PCB mixtures
8.6.5.2 Effects of individual congeners
8.6.6 Effects on porphyrin metabolism
8.6.6.1 Effects of PCB mixtures
8.6.6.2 Effects of individual congeners
8.6.7 Effects on the endocrine system
8.6.7.1 Effects of PCB mixtures
8.6.7.2 Effects of individual congeners
8.6.8 Immunotoxicity
8.6.8.1 Effects of PCB mixtures
8.6.8.2 Effects of individual congeners
8.6.8.3 Appraisal
8.6.9 Neurotoxic effects
8.6.10 Skin effects
8.6.11 Effects on the lung
8.6.12 Miscellaneous
8.7 Factors modifying toxicity; mode of action
8.7.1 Factors modifying toxicity
8.7.2 Mechanisms of toxicity
8.7.3 Toxicity of impurities in commercial PCBs

9. EFFECTS ON HUMANS
9.1 General population exposure
9.1.1 Acute effects - poisoning incidents
9.1.2 Effects of short- and long-term exposure
9.1.2.1 Yusho and Yu-Cheng accidents
9.1.2.2 Effects of PCBs on babies and infants
9.1.3 Appraisal
9.2 Occupational exposure
9.2.1 Acute toxicity - poisoning incidents
9.2.1.1 Acute dermal effects
9.2.2 Effects of short- and long-term exposure
9.2.3 Appraisal

9.2.4 Special studies (target organ effects)
9.2.4.1 Effects on the liver
9.2.4.2 Immunotoxicity
9.2.4.3 Effects on the respiratory system
9.2.4.4 Neurotoxicity
9.2.4.5 Blood pressure
9.2.5 Mortality studies
9.2.6 Appraisal

10. PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES

POLYCHLORINATED TERPHENYLS

1. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, ANALYTICAL METHODS

1.1 Identity
1.2 Physical and chemical properties
1.3 Analytical methods

2. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE

3. ENVIRONMENTAL TRANSPORT, DISTRIBUTION AND TRANSFORMATION

4. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE
4.1 Residues in the environment
4.2 Residues in food
4.3 Concentrations in adipose tissue
4.4 Concentrations in blood

5. KINETICS AND METABOLISM
5.1 Absorption
5.2 Distribution
5.3 Biotransformation

6. EFFECTS ON ORGANISMS IN THE ENVIRONMENT
6.1 Marine and estuarine organisms
6.2 Terrestrial invertebrates
6.3 Birds

7. EFFECTS ON EXPERIMENTAL ANIMALS AND IN VITRO TEST SYSTEMS
7.1 Single oral exposures
7.2 Short-term oral exposures
7.2.1 Rat
7.2.2 Monkey
7.3 Teratogenicity
7.4 Carcinogenicity
7.5 Miscellaneous effects

REFERENCES

ANNEX 1

RESUME

RESUMEN

WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR POLYCHLORINATED
BIPHENYLS (PCBs) AND POLYCHLORINATED TERPHENYLS (PCTs)

Members

Dr L.A. Albert, Consultores Ambientales Asociados, Xalapa, Veracruz,
Mexico

Professor U.G. Ahlborg, Institute of Environmental Medicine,
Karolinska Institute, Stockholm, Sweden

Dr V. Benes, Department of Toxicology and Reference Laboratory,
Institute of Hygiene and Epidemiology, Prague, Czechoslovakia
(Vice-Chairman)

Dr S. Dobson, Institute of Terrestrial Ecology, Monks Wood
Experimental Station, Abbots Ripton, Huntingdon, United Kingdom
(Chairman)

Dr Yuzo Hayashi, Division of Pathology, National Institute of Hygienic
Sciences, Tokyo, Japan

Dr T. Lakhanisky, Division of Toxicology, Institute of Hygiene and
Epidemiology, Brussels, Belgium

Dr J. McKinney, US Environmental Protection Agency, Research Triangle
Park, North Carolina, USA

Dr Pang Ying Fa, Chinese Academy of Preventive Medicine, Beijing,
China

Dr T. Vermeire, National Institute of Public Health and Environmental
Protection, Bilthoven, Netherlands (Co-Rapporteur)

Dr E. Yrjänheikki, Regional Institute of Occupational Health, Oulu,
Finland

Observers

Dr M. Martens (Representative from ECETOC), Monsanto Services
International, Brussels, Belgium

Mrs H. B. Sundmark (Representative from ECETOC), Norsk Hydro a.s.
Porsgrunn, Research Centre, Porsgrunn, Norway

Secretariat:

Dr G.J. van Esch, Bilthoven, Netherlands (Co-Rapporteur and
Secretary)

Dr M. Kogevinas, Unit of Analytical Epidemiology, International Agency
for Research on Cancer (IARC), Lyon, France

NOTE TO READERS OF THE CRITERIA MONOGRAPHS

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

* * *

A detailed data profile and a legal file can be obtained from the
International Register of Potentially Toxic Chemicals, Palais des
Nations, 1211 Geneva 10, Switzerland (Telephone no. 7988400/7985850).

ENVIRONMENTAL HEALTH CRITERIA FOR PCBs AND PCTs

A WHO Task Group on Environmental Health Criteria for PCBs and PCTs
met in Brussels from 28 May to 1 June 1990. The meeting was convened
in the Institute of Hygiene and Epidemiology in Brussels and sponsored
by the Belgian Ministry of Health. Mrs A.-M. Sacré-Bestin of the
Ministry opened the meeting and welcomed the participants on behalf of
the host country. Dr G.J. van Esch welcomed the participants on behalf
of the Heads of the three IPCS cooperating organizations
(UNEP/ILO/WHO). The Group reviewed and revised the draft Environmental
Health Criteria monograph and the companion Health and Safety Guide
and made an evaluation of the risks for human health and the
environment from exposure to PCBs and PCTs.

The first draft of the EHC monograph was prepared by Dr S. Dobson
(environmental aspects) and Dr G.J. van Esch (other sections) and was
based on contributions from several authors and countries. It was
prepared in close cooperation with the WHO Regional Office for Europe,
in Copenhagen.

The second draft was prepared by Dr G.J. van Esch, incorporating
comments received following the circulation of the first draft to the
IPCS contact points for Environmental Health Criteria monographs.
Dr K. Jager, Central Unit, IPCS, was responsible for the scientific
content of the final monograph and Mrs M.O. Head, Oxford, for the
editing.

The efforts of all who helped in the preparation and finalization of
the documents are gratefully acknowledged.

INTRODUCTION

The commercial production of the polychlorinated biphenyls (PCBs)
began in 1930, and, during the 1930s, cases of poisoning were reported
among men engaged in their manufacture. The nature of this
occupational disease was characterized by a skin affection with
acneiform eruptions; occasionally the liver was involved, in some
cases with fatal consequences. Subsequent safety precautions appear
largely to have prevented further outbreaks of this disease in
connection with the manufacture of PCBs, but, since 1953, cases have
been reported in Japanese factories manufacturing condensers.

The distribution of PCBs in the environment was not recognized until
Jensen started an investigation in 1964 to ascertain the origins of
unknown peaks, observed during the gas-liquid chromatographic
separation of organochlorine pesticides from wildlife samples. In
1966, he and his colleagues succeeded in attributing these to the
presence of PCBs. Since then, investigations in many parts of the
world have revealed the widespread distribution of PCBs in
environmental samples.

The serious outbreaks of poisoning in humans and in domestic animals
from the ingestion of food, accidentally contaminated with PCBs, have
stimulated investigations into the toxic effects of PCBs on animals
and on nutritional food chains. This has resulted in the limitation of
the commercial exploitation of PCBs and polychlorinated terphenyls
(PCTs), and in regulations to limit the residues in human and animal
food.

In recent years, many industrial nations have taken steps to control
the flow of PCBs into the environment. PCBs and PCB-containing
formulations are restricted (an exception is sometimes made for mono-
and dichloro-PCBs) for most uses. Now they are almost entirely
restricted to use in closed systems, such as isolating oils in
transformers, capacitors, and other electrical systems, and as a heat
transfer medium and hydraulic liquid. The most influential forces
leading to these restrictions have probably been the 1973 and 1987
decision-recommendations from the Organisation for Economic
Co-operation and Development (OECD).

The environmental impact of the PCBs and PCTs has been discussed at a
number of regional and international meetings and has been the subject
of several reviews, including: ATSDR (1989), DFG (1988), IARC (1978),
IRPTC (1988), Kimbrough (1987), Lorenz & Neumeier (1983a,b), NIOSH
(1987), NTIS (1972), OECD (1982), Slorach & Vaz (1983), WHO (1985a,b,
1986a,b) & WHO/EUR (1987).

In 1976, the World Health Organization published Environmental Health
Criteria 2: Polychlorinated biphenyls (PCBs) and terphenyls (PCTs)
(WHO, 1976), discussing and evaluating the data then available on
exposure levels and the effects of PCBs and PCTs on human beings, and,
to a lesser extent, on the environment.

Since then, a wealth of new information has become available.

The IPCS decided to update the above-mentioned EHC and also to produce
a Health and Safety Guide (HSG) and to do this in close coordination
with the WHO Regional Office for Europe, which prepared "PCBs, PCDDs
and PCDFs, prevention and control of accidental and environmental
exposures" as No. 23 of their Environmental Health Series (WHO/EURO,
1987). This publication includes a set of guidelines to assist Member
States in the development of strategies to reduce the probability of
accidents involving the environmental release of PCBs, PCDDS, and
PCDFs and also the severity of their hazardous effects, should such
accidents occur. In particular, it is intended to guide occupational
safety and health personnel and other staff, in workplaces and
environments where PCBs and/or PCB-containing equipment are in use, to
develop adequate safety measures, contingency planning, effective and
relevant accident response, and appropriate rehabilitation.

Within the scope of the present EHC on PCBs and PCTs, the PCDDs and
PCDFs have been mentioned where relevant. Full discussion of these
compounds and evaluation, however, can be found in the IPCS EHC 88:
Polychlorinated dibenzo- para-dioxins and dibenzofurans (WHO, 1989).

1. SUMMARY AND EVALUATION, CONCLUSIONS, RECOMMENDATIONS

1.1 Summary and evaluation

1.1.1 Introduction

Polychlorinated biphenyls (PCBs) were discovered before the turn of
the century and their usefulness for industry, because of their
physical properties, was recognized early. The PCBs have been used
commercially, since 1930, as dielectric and heat-exchange fluids and
in a variety of other applications. They have become widely
distributed in the environment throughout the world, and are
persistent and accumulate in food webs. Human exposure to PCBs has
resulted largely from the consumption of contaminated food, but also
from inhalation and skin absorption in work environments. PCBs
accumulate in the fatty tissues of humans and other animals and have
caused toxic effects in both, particularly if repeated exposure
occurs. The skin and liver are the major sites of pathology, but the
gastrointestinal tract, the immune system, and the nervous system are
also targets. Polychlorinated dibenzofurans (PCDFs), which are
contaminants in commercial PCB mixtures, contribute significantly to
their toxicity. The results of studies on rodents suggest that some
PCB congeners may be carcinogenic and that they can promote the
carcinogenicity of other chemicals.

It is clear from available data on polychlorinated biphenyls (PCBs)
and polychlorinated terphenyls (PCTs) that, in an ideal situation, it
would be preferable not to have these compounds in food at any level.
However, it is equally clear that the reduction of PCBs or PCTs
exposure from food sources to "zero" or to a level approaching zero,
would mean the elimination (prohibition of the consumption) of large
amounts of important food items, such as fish, but more importantly
breast milk. National and international scientific committees have to
decide where the proper balance lies between providing an adequate
degree of public health protection and avoiding excessive losses of
food.

No levels of PCBs or PCTs exposure that can provide an absolute
assurance of safety can be identified on the basis of the available
data.

1.1.2 Identity, physical, and chemical properties

PCBs are mixtures of aromatic chemicals, manufactured by the
chlorination of biphenyl in the presence of a suitable catalyst. The
chemical formula of PCBs can be presented as C₁₂H_10-nCl_n, where n is
a number of chlorine atoms within the range of 1-10.

Theoretically, 209 congeners are possible, but only about 130
congeners are likely to occur in commercial products. In addition,
PCBs may contain polychlorinated dibenzofurans (PCDFs) and chlorinated
quarterphenyls as impurities. These impurities are relatively stable
and resistant to chemical reactions, under normal conditions. All
congeners of PCBs are lipophilic and have a very low water solubility.
As a result, they easily enter the food chain and accumulate in fatty
tissues.

Commercial PCB mixtures contain PCDFs at levels ranging from a few
mg/kg up to 40 mg/kg. Polychlorinated dibenzo- p-dioxins (PCDDs), are
not found in commercial PCBs. However, when PCBs are mixed with other
chlorinated compounds, such as the chloro-benzenes used in
transformers, PCDDs can be found in the case of accidental fires and
during incineration.

Commercial PCB mixtures are light yellow or dark yellow in colour.
They do not crystallize, even at low temperatures, but turn into solid
resins. PCBs are, in practice, fire resistant, with rather high flash
points. They form vapours heavier than air, but they do not form any
explosive mixtures with air. They have very low electrical
conductivity, rather high thermal conductivity, and extremely high
resistance to thermal break-down. PCBs are chemically very stable
under normal conditions; however, when heated, other toxic compounds,
such as PCDFs, can be produced.

1.1.3 Analytical methods

In 1966, the discovery of PCBs in environmental samples raised
interest in the analysis of these compounds and their toxicity for
human beings and their environment.

Because of differences in the analytical methodology used, existing
data are not directly comparable; nevertheless, they can be used for
the establishment of control and preventive measures and for the
preliminary assessment of health and environmental risks associated
with these chemicals.

PCBs have been determined using gas chromatography (GC) techniques
with electron capture detection, often using packed columns, though
more sophisticated methods, such as capillary column and GC coupled
with mass-spectrometry (GC-MS), have been used in recent studies to
identify the individual congeners, to improve the comparability of the
analytical data from different sources, and to establish a basis for
toxicity assessment.

An extensive quality assurance programme is required for these
analyses and intercalibration studies have been implemented and
recommended. The quality and utility of the analytical data depend
critically on the validity of the sample and the adequacy of the
sampling. Furthermore, it is essential to have a planned and well
documented sampling programme; a detailed sampling procedure is
described in WHO/EURO (1987).

1.1.4 Production and uses

The commercial production of the PCBs began in 1930. They have been
widely used in electrical equipment, and smaller volumes of PCBs are
used as fire-resistant liquid in nominally closed systems.

By the end of 1980, the total world production of PCBs was in excess
of 1 million tonnes and, since then, production has continued in some
countries. Despite increasing withdrawal of the use, and restrictions
on the production, of PCBs, very large amounts of these compounds
continue to be present in the environment, either in use or as waste.

In recent years, many industrialized countries have taken steps to
control and restrict the flow of PCBs into the environment. The most
influential force leading to these restrictions has probably been a
1973 recommendation from the Organisation for Economic Co-operation
and Development (OECD) (WHO, 1976; IARC, 1978; OECD, 1982). Since
then, the 24 OECD member countries have restricted the manufacture,
sales, importation, exportation, and use of PCBs, as well as
establishing a labelling system for these compounds.

Current sources of PCB release include volatilization from landfills
containing transformer, capacitor, and other PCB-wastes, sewage
sludge, spills, and dredge spoils, and improper (or illegal) disposal
to open areas. Pollution may occur during the incineration of
industrial and municipal waste. Most municipal incinerators are not
effective in destroying PCBs. Explosions or overheating of
transformers and capacitors may release significant amounts of PCBs
into the local environment.

PCBs can be converted to PCDFs under pyrolytic conditions. The highest
yield of PCDFs under laboratory conditions was obtained at a
temperature between 550 and 700°C. Thus, the uncontrolled burning of
PCBs can be an important source of hazardous PCDFs. It is therefore
recommended that destruction of PCB-contaminated waste should be
carefully controlled, especially with regard to the burning
temperature (above 1000°C), residence time, and turbulence.

1.1.5 Environmental transport, distribution, and transformation

In the atmosphere, PCBs exist primarily in the vapour phase; the
tendency to adsorb on particulates increases with the degree of
chlorination. The virtually universal distribution of PCBs suggests
transport in air.

At present, the major source of PCB exposure in the general
environment appears to be the redistribution of PCBs, previously
introduced into the environment. This redistribution involves
volatilization from soil and water into the atmosphere with subsequent
transport in air and removal from the atmosphere via wet/dry
deposition (of PCBs bound to particulates) and then re-volatilization.
Concentrations of PCBs in precipitation range from 0.001 to
0.25 µg/litre. Since the volatilization and degradation rates of PCBs
vary between congeners, this redistribution leads to an alteration in
the composition of PCB mixtures in the environment.

In water, PCBs are adsorbed on sediments and other organic matter;
experimental and monitoring data have shown that PCB concentrations in
sediment and suspended matter are higher than those in associated
water columns. Strong adsorption on sediment, especially in the case
of the higher chlorinated PCBs, decreases the rate of volatilization.
On the basis of their water solubilities and n-octanol-water
partition coefficients, the lower chlorinated PCB congeners will sorb
less strongly than the higher chlorinated isomers. Although adsorption
can immobilize PCBs for relatively long periods in the aquatic
environment, desorption into the water column has been shown to occur
by both abiotic and biotic routes. The substantial quantities of PCBs
in aquatic sediments can therefore act as both an environmental sink
and a reservoir of PCBs for organisms. Most of the environmental load
of PCBs has been estimated to be in aquatic sediment.

The low solubility and the strong adsorption of PCBs on soil particles
limits leaching in soil; lower chlorinated PCBs will tend to leach
more than the highly chlorinated PCBs.

Degradation of PCBs in the environment is dependent on the degree of
chlorination of the biphenyl. In general, persistence of PCB congeners
increases as the degree of chlorination increases. In the atmosphere,
the vapour phase reaction of PCBs with hydroxyl radicals (which are
photochemically formed by sunlight) may be the dominant transformation
process. Estimated half-lives for this reaction in the atmosphere
range from about 10 days for a monochlorobiphenyl to 1.5 years for a
heptachlorobiphenyl.

In the aquatic environment, hydrolysis and oxidation do not
significantly degrade PCBs. Photolysis appears to be the only viable
abiotic degradation process in water; however, available experimental
data are not sufficient to determine its rate or importance in the
environment.

Microorganisms degrade mono-, di-, and trichlorinated biphenyls
relatively rapidly and tetrachlorobiphenyls slowly, whilst higher
chlorinated biphenyls are resistant to biodegradation. Chlorine
substitution positions on the biphenyl ring appear to be important in
determining the biodegradation rate. PCBs containing chlorine atoms in
the para positions are preferentially biodegraded. Higher
chlorinated congeners are biotransformed anaerobically, by a reductive
dechlorination, to lower chlorinated PCBs, which may then be
biodegradable by aerobic processes.

Several factors determine the degree of bioaccumulation in adipose
tissues: the duration and level of exposure, the chemical structure of
the compound, and the position and pattern of substitution. In
general, the higher chlorinated congeners are accumulated more
readily.

Experimentally determined bioconcentration factors of various PCBs in
aquatic species (fish, shrimp, oyster) range from 200 up to 70 000 or
more. In the open ocean, there is bioaccumulation of PCBs in higher
trophic levels with an increased proportion of higher chlorinated
biphenyls in higher ranking predators.

Transfer of PCBs from soil to vegetation takes place mainly by
adsorption on the external surfaces of terrestrial plants; little
translocation takes place.

1.1.6 Environmental levels and human exposure

Because of their high persistence, and their other physical and
chemical properties, PCBs are present in the environment all over the
world.

Globally, PCBs are found in air concentrations of 0.002 up to
15 ng/m³. In industrial areas, levels are higher (up to µg/m³). In
rain water and snow, PCBs are found in the range of nd (1 ng)-
250 ng/litre.

Under occupational conditions, the levels in the air may be much
higher. Under certain conditions, for instance, in the manufacturing
of transformers or capacitors, levels of up to 1000 µg/m³ have been
observed. In acute emergencies, concentrations of up to 16 mg/m³ have
been measured. In case of fires and/or explosions, soot may be
produced that contains high levels of PCBs. Levels of 8000 mg PCBs/kg
soot have been found. In the latter situation, PCDFs will also be
present. Polychlorinated dioxins (PCDDs) will be found in accidents
with transformers containing chlorinated benzenes, as well as PCBs.

In these emergency situations, ingestion, skin contamination, or
inhalation of soot particles may occur and result in serious exposure
of personnel. However, the exposure of the general population via air
will be very low.

Surface water may be contaminated by PCBs from atmospheric fallout,
from direct emissions from point sources, or from waste disposal.
Under certain conditions, levels of up to 100-500 ng/litre water have
been measured. In the oceans, levels of 0.05-0.6 ng/litre have been
found.

In non-contaminated areas, drinking-water contains less than 1 ng
PCBs/litre, but levels of up to 5 ng/litre have been reported. Soil
and sediments in different areas and depending on local conditions,
contain levels of PCBs ranging from <0.01 up to 2.0 mg/kg. In
polluted areas, the levels have been much higher, i.e., up to
500 mg/kg.

In past years, many thousands of samples of different foodstuffs have
been analysed in several countries for contaminants including PCBs.
Most samples have been taken from individual food items, especially
fish and other foods of animal origin, such as meat and milk. Human
food has become contaminated with PCBs by 3 main routes:

(a) uptake from the environment by fish, birds, livestock (via
food-chains), and crops;

(b) migration from packaging materials into food (mainly below
1 mg/kg, but, in some cases, up to 10 mg/kg);

The levels for the most important PCB-containing food items were:
animal fat, 20-240 µg/kg; cow's milk, 5-200 µg/kg; butter,
30-80 µg/kg; fish, 10-500 µg/kg, on a fat basis. Certain fish species
(eel) or fish products (fish liver and fish oils) contain much higher
levels, up to 10 mg/kg. Vegetables, cereals, fruits, and a number of
other products contained levels of <10 µg/kg. The major foods in
which contamination with PCBs needs consideration are fish, shellfish,
meat, milk, and other dairy products. Median levels in fish, reported
in various countries, are of the order of 100 µg/kg (on a fat basis).
When comparisons have been made, it appears that the levels of PCBs in
fish are slowly decreasing.

PCBs concentrate in human adipose tissue and breast milk. The
concentrations of PCBs in the different organs and tissues depend on
their lipid contents, with the exception of the brain. PCB residues in
the adipose tissue of the general population in industrialized
countries range from less than 1 up to 5 mg/kg, on a fat basis.

The average concentrations of total PCBs in human milk fat are in the
range of 0.5-1.5 mg/kg fat, depending on the donor's residence,
life-style, and the analytical methods used. Women who live in
heavily-industrialized, urban areas, or who consume a lot of fish,
especially from heavily-contaminated waters, may have higher PCB
concentrations in their breast milk.

The composition of most PCB extracts from environmental samples does
not resemble that of the commercial PCB mixtures. It has also been
shown, using high-resolution gas chromatography analysis, that the
congener composition and the relative concentrations of the individual
components in adipose tissues and breast milk differ markedly from
those in the commercial PCBs. The GC patterns of PCBs in human adipose
tissue and breast milk contain relatively high concentrations of
mainly the higher chlorinated PCBs, such as: 2,4,5,3',4'-pentachloro
biphenyl; 2,4,5,2',4',5'-hexachlorobiphenyl, and 2,3,4,2',4',5'-
hexachlorobiphenyl; 2,3,4,5,2',4',5'-hepta- and 2,3,4,5,2',3',4'-
heptachlorobiphenyl. A few other PCB congeners are present in
much lower quantities, such as the most toxic, coplanar PCBs:
3,4,3',4'-tetra-, 3,4,5,3',4'-penta-, and 3,4,5,3',4',5'-
hexachlorobiphenyl.

It has been calculated that the daily intake of PCBs by infants from
breast milk, is of the order of 4.2 µg/kg body weight (5.2 µg/100 Kcal
consumed) (WHO/EURO, 1988). The average total of ingested PCBs from
breast milk, during the first 6 months of life, is 4.5 mg compared
with the calculated intake of 357 mg of PCBs over the subsequent
life-time (0.2 µg/kg per day from the diet of a 70-kg person over a
70-year life-time). Therefore, the nursing period contributes about
1.3% of the life-time intake, which is not large, in the light of the
benefits of breast-feeding (WHO/EURO, 1988).

On the basis of the evaluated background data, for adults the average
dietary intake of PCBs amounts to a maximum of 100 µg per week, or
approximately 14 µg/person per day. For a 70-kg person, this is an
intake equivalent to a maximum of 0.2 µg/kg body weight per day
(WHO/EURO, 1988).

1.1.7 Kinetics and metabolism

Animal studies have been reported involving mainly oral, inhalation,
and dermal exposures to both PCB mixtures and individual congeners. In
general, PCBs appear to be rapidly absorbed, particularly by the
gastrointestinal tract after oral exposure. It is clear that
absorption does occur in humans, but information on the rates of human
absorption of PCBs is limited.

From the available studies, the data on the distribution of PCBs,
suggest a biphasic kinetic process with rapid clearance from the blood
and accumulation in the liver and the adipose tissue of various
organs. There is also evidence of placental transport, fetal
accumulation, and distribution to milk. In some human studies, the
skin contained a high concentration of PCBs, but the concentration in
the brain was lower than that expected on the basis of the lipid
content.

Mobilization of PCBs from fat appears to depend largely on the rates
of metabolism of the individual PCB congeners. Excretion depends on
the metabolism of PCBs to more polar compounds, such as phenols,
conjugates of thiol compounds, and other water-soluble derivatives.

Metabolic pathways include hydroxylation, and conjugation with thiols
and other water-soluble derivatives, some of which can involve
reactive intermediates, such as the arene oxides. Rates of metabolism
have been shown to depend on the PCB structure and reflect both the
degree and position of chlorine substituents. The polar metabolites of
the more highly chlorinated PCBs appear to be eliminated primarily in
the faeces, but excretion in the urine can also be significant. An
important elimination route, is via (breast) milk. Certain PCB
congeners can also be eliminated via hair.

The available kinetic studies indicate that there is a wide divergence
in biological half-life among the individual congeners and this can
reflect differences in structure-dependent metabolism, tissue
affinities, and other factors affecting mobilization from storage
sites. Persistence in tissues is not always correlated with high
toxicity, and differences in toxicity between PCB congeners may be
associated with specific metabolites and/or their intermediates.

1.1.8 Effects on organisms in the environment

PCBs are universal, environmental contaminants and are present in most
environmental compartments, abiotic and biotic, throughout the world.
Since many countries have controlled both use and release, new input
into the environment is on a reduced scale compared with the past.
However, the available evidence suggests that the cycling of PCBs is
causing a gradual redistribution of some congeners towards the marine
environment. There is a trend for the highest chlorinated congeners to
accumulate preferentially. While much of the PCB is adsorbed on to
particulates in sediment, it is still bioavailable to organisms and
will continue to be accumulated in higher trophic levels.

1.1.8.1 Laboratory studies

Effects of PCB mixtures on microorganisms are highly variable with
some species adversely affected by a level of 0.1 mg/litre and others
unaffected by 100 mg/litre; effects on different species do not vary
consistently with the degree of chlorination of the mixtures. Almost
all of the studies of the effects of PCBs on aquatic organisms have
been concerned with Aroclor mixtures. Results have been extremely
variable with no consistent relationship between percentage
chlorination or environmental conditions and toxicity, even with
closely-related organisms. Over 96 h, under static conditions, LC₅₀
values have ranged between 12 µg/litre and >10 mg/litre for various
aquatic invertebrate species and different Aroclor mixtures.

Flow-through conditions increased the toxicity of the PCBs. Generally,
the most toxic mixtures were Aroclors in the mid-range of
chlorination; low and high percentage chlorination mixtures were less
toxic. This was also true for sub-lethal effects, such as reproduction
effects in Daphnia. Crustaceans seem to be more susceptible to PCBs
during moult. In model populations, the community structure of
estuarine species changed on exposure to Aroclor 1254, with the
numbers of amphipods, bryozoans, crabs, and molluscs decreasing and
those of annelids, brachyopods, coelenterates, echinoderms, and
nemerines unaffected. Too few of the groups have been included in
acute tests to determine whether the results represent variation in
susceptibility to PCBs or differences in interaction between species.

There is a similar variation in the toxicity of PCB mixtures for fish,
with 96-h LC₅₀s varying between 0.008 and >100 mg/litre. Long-term
tests have shown that acute exposure, particularly in static
conditions, considerably underestimates the toxicity of the PCB.
Rainbow trout was particularly susceptible, with embryo-larval stages
showing a 22-day LC₅₀ of 0.32 µg/litre for Aroclor 1254 and a
no-observed-effect level (NOEL) over 22 days of 0.01 µg/litre for
Aroclors 1016, 1242, and 1254.

Freshwater fathead minnow showed NOELs of 5.4, 0.1, 1.8, and
1.3 µg/litre for Aroclors 1242, 1248, 1254, and 1260, respectively;
the estuarine sheephead minnow showed NOELs of 3.4 and 0.06 µg/litre
for Aroclors 1016 and 1254, respectively.

Experimental evidence has confirmed field observations demonstrating
reproductive impairment in seals fed on fish containing PCBs
accumulated in the wild. The effect occurs late in reproduction,
preventing implantation of the embryo in the uterine wall.

In short-term tests, the toxicity of Aroclor for birds increased with
increasing percentage chlorination; 5-day dietary LC₅₀s ranged from
604 to >6000 mg/kg diet. The main reproductive effects of PCBs on
birds were reduced hatchability of eggs and embryotoxicity. These
effects continued after dosing ended, as the hens reduced their PCB
load via the eggs. There is no evidence that Aroclors cause egg-shell
thinning, directly; effects on the food consumption and body weight of
hens have an indirect effect on shell thickness. Sub-lethal effects on
behaviour and hormone secretion have been reported.

The acute toxicity of Aroclors for mink decreases with increasing
percentage chlorination, acute oral LD₅₀s varying between >750 and
4000 mg/kg body weight; the ferret is less sensitive. Aroclor reduces
food consumption and, thus, the growth rate of young mink.
Reproduction of mink is reduced or eliminated by Aroclors, either
given directly or as natural contaminants in fish. Higher percentage
chlorinated Aroclors (notably 1254) have a greater effect. The
reproductive rate returns to normal after cessation of feeding with
Aroclor.

Bats are susceptible to Aroclor released from fat during migration.

Because the great majority of laboratory tests on aquatic and
terrestrial organisms were carried out using PCB mixtures, it is not
possible to identify which specific components of the mixtures were
responsible for effects. Similarly, because tests were conducted in
environmentally unrealistic conditions (e.g., beyond the solubility of
congeners and without sediment present in aquatic tests), it is
difficult to extrapolate from laboratory to field. However, it can
reasonably be assumed that any effects on populations of organisms,
likely to occur more generally in the environment in the future, will
already have been observed in local populations exposed to high PCB
levels in the past.

1.1.8.2 Field studies

Results suggesting effects of PCBs on fish populations in the field
are inconclusive. Interpretation of field data on birds is difficult,
since residues of many different organochlorines are also present.
Most authors have shown a correlation between effects (embryotoxicity)
and total organochlorine residues. Of the organochlorine compounds
present, PCB residues correlate best with the effects on embryos, but
the results cannot be regarded as proved field effects of the PCBs.

There is evidence (confirmed in laboratory studies) that PCBs reduce
the reproductive capacity of sea mammals. The effect is on the
implantation of the embryo, but there can also be physical changes in
the female reproductive tract.

Extrapolation from laboratory, acute and short-term tests to effects
at the population level in the field is not possible. Uncertainties
about which components of the PCB mixtures cause effects, the specific
congeners present in the environment, and the bioavailability of PCB
components to organisms, all combine to make estimates of likely
environmental exposures and effects difficult. The effects on sea
mammal populations can be regarded as proved, but the component(s) of
the PCB mixtures that are responsible are not yet known.

Given the trends towards increased contamination of the marine
environment, attention should be concentrated on the effects on marine
organisms. There is clear laboratory and field evidence of
reproductive effects on populations of sea mammals in heavily-polluted
areas. The residues and effects of PCBs on other populations of sea
mammals are likely to increase in the future. It is less clear whether
effects will be seen in other organisms, such as birds feeding on
marine prey.

Population and community effects on lower organisms, phytoplankton,
and zooplankton, would be expected to occur on the basis of laboratory
experiments. Both the extent and significance of such changes are
difficult to assess. From currently available information, effects on
fish populations would not be expected, though fish will act as a
route of exposure of fish-eating mammals and birds.

Previously reported effects on terrestrial species, fish-eating,
freshwater mammals and migratory bats, for example, should be less
evident as residues of PCBs are redistributed. Residues in terrestrial
biota currently show little decline overall, but information on
changes in congeners is scarce or absent. Declines in higher
chlorinated congeners would be expected to be slow.

1.1.9 Effects on experimental animals and in vitro systems

1.1.9.1 Single exposure

The acute toxicity of Aroclors, after a single oral exposure, is
generally low in rats. Young animals appear to be more sensitive
(LD₅₀: 1.3-2.5 g/kg body weight) than adults (LD₅₀: 4-11 g/kg body
weight). The lowest LD₅₀ reported for Aroclor 1254 in adult rats was
1.0 g/kg body weight. No differences between the sexes were observed.

The dermal LD₅₀ in rabbits ranged from >1.26 to <2 g/kg body weight
for Aroclor 1260 (in corn oil) and from 0.79 to <3.17 g/kg body
weight for some other undiluted PCB mixtures. With intravenous
application, an LD₅₀ of 0.4 g/kg body weight for Aroclor 1254 was
shown in rats; the LD₅₀ after intraperitoneal injection in the mouse
varied from 0.9 to 1.2 g/kg body weight.

1.1.9.2 Short-term exposure

The main targets in mammals, with short-term, oral exposure to PCB
mixtures or congeners, were the liver, the skin, the immune system,
and the reproductive system. The Rhesus monkey was the most sensitive
species tested, females being more sensitive than males. Adult female
Rhesus monkeys exposed to a diet containing Aroclor 1248 at a level of
2.5 mg/kg, or 0.09 mg/kg body weight per day, for 6 months, showed an
increased mortality rate, growth retardation, alopecia, acne, swelling
of the Meibomian glands, and possibly immunosuppression.

Microscopically, enlarged fatty liver with focal necrosis, and
epithelial hyperplasia, and keratinization of hair follicles were
found. At higher exposure levels, microscopic changes have also been
observed in other epithelial tissues, such as the sebaceous and
Meibomian glands, the gastric mucosa, gall bladder, bile duct, nail
beds, and the ameloblast. Serum levels of total lipid triglycerides
and cholesterol were decreased. Short-term exposure to commercial PCB
mixtures induced an increase in the concentrations of total lipids,
triglycerides, cholesterol, and/or phospholipids in the liver. Among
the PCB congeners, 3,4,3',4'-tetrachlorobiphenyl 3,4,5,3',4',5'-, and
2,4,6,2',4',6'-hexachlorobiphenyl were the most potent. Aroclor 1254,
at a dose level of 0.2 mg/kg body weight per day, also showed several
other effects, such as lymphoreticular lesions, fingernail detachment,
and gingival effects, but no acne and alopecia. A NOEL for the general
toxicity of Aroclor 1242 of 0.04 mg/kg body weight per day was
established in Rhesus monkeys. Relatively mild effects were shown in
suckling Rhesus monkeys, exposed to a much higher dose of Aroclor 1248
of 35 mg/kg body weight per day. Effects in the liver have been best
investigated in rats and include hypertrophy, fatty degeneration,
proliferation of the endoplasmic reticulum, porphyria, adenofibrosis,
bile-duct hyperplasia, cysts, and preneoplastic and neoplastic
changes. In studies on rats and mice, individual PCB congeners caused
effects in the liver, spleen, and thymus, the planar congeners being
most toxic. In monkeys, planar congeners, at doses of 1-3 mg/kg diet,
induced effects similar in character and severity to those produced by
Aroclor 1242, at a dose of 100 mg/kg diet, and Aroclor 1248, at a dose
of 25 mg/kg diet.

Following dermal exposure of rabbits and mice, PCB mixtures and some
congeners caused effects on the skin and liver, similar to those found
after oral exposure. In rabbits, thymic atrophy, a reduction of
germinal centres of the lymph nodes, and leukopenia were also
observed.

1.1.10 Reproduction, embryotoxicity, and teratogenicity

1.1.10.1 Reproduction and embryotoxicity

Comprehensive reproduction and teratogenicity studies have not been
conducted. In a 2-generation reproduction study on rats, a NOEL of
0.32 mg/kg body weight, based on reproductive parameters (Aroclor
1254) and a NOEL of 7.5 mg/kg body weight (Aroclor 1260) were
established. However, the lowest tested dose of 0.06 mg/kg body weight
resulted in increased relative liver weights in weanlings.

In Rhesus monkeys exposed to Aroclor 1016, a NOEL of 0.03 mg/kg body
weight was established, on the basis of reproductive parameters.
However, at this level, decreased birth weight was observed and the
lowest dose tested, of 0.01 mg/kg body weight, resulted in skin
hyperpigmentation.

For Aroclor 1248 (containing PCDFs), a NOEL of 0.09 mg/kg body weight
was established in Rhesus monkeys, 1 year after exposure ceased.

1.1.10.2 Teratogenicity

Available studies on rats and monkeys did not indicate any teratogenic
effects, when animals were dosed orally during organogenesis. A NOEL
of 50 mg/kg body weight for Aroclor 1254 was demonstrated in rats with
regard to pup weight, and a LOEL of 2.5 mg/kg body weight, on the
basis of fetotoxicity (lesion in thyroid follicular cells) could be
assumed.

In teratogenicity tests with individual congeners on mice, rats, and
Rhesus monkeys, no NOEL was demonstrated. In Rhesus monkeys a dose of
0.07 mg/kg body weight resulted in maternal toxic effects
(3,4,3',4'-tetrachlorobiphenyl).

1.1.11 Mutagenicity

PCB mixtures did not cause mutation or chromosomal damage in a variety
of test systems. Chromosome breakage was induced in human lymphocytes
in vitro by 3,4,3',4'-tetrachlorobiphenyl. High concentrations of
PCB mixtures may cause primary DNA damage, as evidenced by DNA single
strand breaks in alkaline elution assays.

1.1.12 Carcinogenicity

The interpretation of the available animal data involving commercial
PCB mixtures is often complicated by lack of information concerning
the presence, or contribution, of chlorinated dibenzofuran impurities
as well as variations in congener composition.

A number of long-term carcinogenicity studies have been carried out on
mice and rats. The PCB mixtures used were Kanechlors 300, 400, and
500, Aroclors 1254 and 1260, and Clophens A30 and A60. The Clophens
were reported to be free of PCDFs, but no data were provided on the
purity of the other PCB mixtures.

A significant increase in hepatocellular adenomas and/or carcinomas
was observed in mice fed a diet containing Kanechlor 500 and Aroclor
1254 at dose levels of approximately 15-25 mg/kg body weight. No
neoplasms could be detected in mice treated with Kanechlors 300 and
400.

In rats, an increase in hepatocellular adenomas and/or carcinomas was
noted in studies on Aroclors 1254 and 1260, and Clophen A30, with an
exposure period of more than one year. The increase in the incidence
of tumour-bearing animals in these studies was not considered to be
statistically significant, however, it was in the case of 2 other
studies. An increase in the incidence of hepatocellular (trabecular)
carcinomas and adenocarcinomas was demonstrated with Aroclor 1260 and
Clophen A60 administered at a dose level of approximately 5 mg/kg body
weight.

The liver tumours concerned were considered to be non-aggressive
(benign or of low malignancy, no metastasis) and not life shortening.
Adenofibrosis, a preneoplastic lesion and/or neoplastic nodules in the
liver were reported in some of the studies. In one test with Aroclor
1254, a dose-related increase in intestinal metaplasia and
adenocarcinomas of the glandular stomach was demonstrated in the rat.

There is a substantial body of evidence to support the enhancing
effects of PCBs on liver carcinogenesis in rodents pretreated with
hepatocarcinogens. There is weak evidence for the initiating activity
of PCB-mixtures in rodents. From the genotoxicity studies reported, it
can be concluded that PCB-mixtures can be regarded as non-genotoxic.
These results imply that the association of liver tumours with the
administration of PCBs in rodents is attributable to some epigenetic
mechanisms involving enforcement of cell proliferation in the liver
and other manifestations of liver toxicity, hence a threshold approach
can be followed in the evaluation of PCB toxicity. The possibility
that PCBs might enhance carcinogenesis in tissues other than the
liver, in animals pre-exposed to various tissue-specific carcinogens,
needs to be addressed. The anticarcinogenic activity of PCBs shown in
some studies, where PCBs were given to animals during, and prior to,
the administration of carcinogens, may be related to the microsomal,
enzyme-inducing properties of PCBs resulting in an increase in
detoxification.

Overall, there is reason to exercise caution in extrapolating the
available animal data on the carcinogenic potential of PCBs to humans.

1.1.13 Special studies

Lesions induced after exposure to PCB mixtures or individual congeners
concern the liver, skin, immune system, reproductive system, oedema
and disturbances of the gastrointestinal tract, and thyroid gland.

PCBs are able to induce various enzymes in the liver. This has been
demonstrated, in rats, mice, guinea-pigs, rabbits, dogs, and monkeys,
for Aroclors 1248, 1254, 1260, and Kanechlor 400 (induction of
cytochrome P450 and P448). The inducing ability increases with the
chlorine content in the molecule. It is also dependent on the congener
composition, congeners with chlorine in the para- and meta-
position inducing the P450 enzyme. For AHH induction, the position of
the chlorine seems to be more important than the degree of
chlorination. Congeners with both para- and at least two meta-
positions substituted by chlorine, are the most potent inducers of
AHH. Distinct inter-species variations have been demonstrated. The
lowest NOEL (0.025 mg/kg body weight) was found for Aroclor 1260 in
Osborn-Mendel rats.

Effects on the endocrine system are seen as alterations in hormonal
receptor binding and in steroid hormone balance. Direct and indirect
evidence for a weak estrogenic activity was observed for various
Aroclors. Decreased levels of gonadal hormones and increased relative
testes weight were found in rats exposed to 75 mg Aroclor 1242/kg diet
for 36 weeks. Decreased plasma corticosteroid levels without increased
adrenal weight, was found in female mice exposed to Aroclor 1254
(25 mg/kg diet) for 3 weeks. Increased adrenal weight was found in
another strain given a diet containing 200 mg/kg for 2 weeks.

PCB mixtures have shown an immunosuppressive effect in various animal
species, monkeys and rabbits being the most sensitive. The lowest NOEL
in monkeys was 0.1 mg/kg body weight, and, in rabbits, 0.18 mg/kg body
weight.

Depressed motor-activity was seen in mice administered a single oral
dose of 500 mg Aroclor 1254/kg body weight. This was probably in
relation to inhibition of the uptake and release of neurotransmitters.

PCB mixtures were found to decrease the levels of vitamins A and B₁
in the blood and liver of rats. Decreased levels of vitamins A, B₁,
B₂, and B₆ were seen in rats and mice exposed to PCB mixtures.

1.1.14 Factors modifying toxicity, mode of action

Commercial PCBs show a spectrum of toxic responses, partly resembling
that of PCDDs and PCDFs. In addition, the analogous structure-activity
relations of PCB congeners, with respect to most of their toxic
responses and to their potency in inducing P448-dependent AHH,
indicate that PCB congeners that are approximate stereoisomers of
2,3,7,8,-TCDD are the most active. These findings suggest a common
mechanism of action based on the affinity of these compounds for the
cytosolic AH-receptor protein. Toxic equivalence factors relating to
2,3,7,8-TCDD have been proposed for these coplanar PCB congeners. The
nature of the likely interactions between PCBs, PCDFs, and PCDDs has
not been adequately investigated. As PCBs can stimulate microsomal
enzyme activity, they can influence the action of other chemicals that
undergo microsomal metabolism. Other so-called, non-planar PCB
congeners may cause other more subtle toxicities. In addition, PCB
congeners, especially the lower chlorinated ones, may be metabolized
through arene oxide intermediates and methylsulfonyl metabolites.

1.1.15 Effects on humans

The toxicological evaluation of PCBs presents many problems. PCBs
usually occur as mixtures of many congeners, and many of the data on
the toxicity of the PCBs are based on the testing of these mixtures.
Some components of the mixtures are more easily degraded in the
environment than others. Thus, the general population may be exposed
to mixtures that are different from those to which workers, working
with PCBs, are exposed.

The general population is exposed to PCBs mainly through contaminated
food (aquatic organisms, meat and dairy products). The daily intake of
PCBs is of the order of some micrograms per person for most of the
industrialized countries. Such exposures have not been associated with
disease. The infant is exposed to PCBs through its mother's milk.
Daily intake of PCBs may be some micrograms/kg body weight.

There are great difficulties in assessing human health effects
separately for PCBs, PCDFs, or PCDDs, since, quite frequently, PCB
mixtures contain PCDFs. The presence of PCDDs has also been seen
occasionally, in accidents with certain mixtures. Commercial PCBs have
been shown to be contaminated with PCDFs and, therefore, in many
cases, it is not clear which effects are attributable to the PCBs
themselves and which to the much more toxic PCDFs. Thus, much of the
data that can be retrieved from large episodes of intoxication in
humans, e.g., the Yusho, Yu-Cheng, and other intoxications, probably
reflect effects of exposure to both PCDFs and PCBs.

The signs of intoxication in Yusho and Yu-Cheng patients were
hypersecretion of the Meibomian glands of the eyes, swelling of the
eyelids and pigmentation of the nails and mucous membranes,
occasionally associated with fatigue, nausea, and vomiting. This was
usually followed by hyperkeratosis and darkening of the skin with
follicular enlargement and acneiform eruptions. Furthermore, oedema of
the arms and legs, liver enlargement and liver disorders, central
nervous disturbances, respiratory problems e.g., bronchitis-like
disturbances, and changes in the immune status of the patients were
also observed. In children of Yusho- and Yu-Cheng patients, diminished
growth, dark pigmentation of the skin and mucous membranes, gingival
hyperplasia, xenophthalmic oedematous eyes, dentition at birth,
abnormal calcification of the skull, rocker bottom heel, and a high
incidence of low birth weight were observed. Whether or not a
correlation existed between the exposure and the occurrence of
malignant neoplasms in these patients could not be definitely
concluded, because the number of deaths was too small. However, a
statistically significant increase was observed in male patients, with
regard to mortality from all neoplasms, liver and lung cancer.

Under occupational conditions, skin rashes occurred a few hours after
acute exposure. Furthermore, itching, burning sensations, irritation
of the conjunctivae, pigmentation the fingers and nails, and chloracne
were found after exposure to high PCB concentrations. Chloracne is one
of the most prevalent findings among PCB-exposed workers. Besides
these dermal signs of intoxication, different authors have found liver
disturbances, immunosuppressive changes, transient irritation of the
mucous membranes of the respiratory tract, neurological and unspecific
psychological or psychosomatic effects, such as headache, dizziness,
depression, sleep and memory disturbances, nervousness, fatigue, and
impotence. The overall conclusion is that continuous occupational
exposure to high PCB and PCDF concentrations may result in effects on
the skin and liver.

Two large mortality studies were carried out on cohorts of workers.
When exposure to Aroclor 1254, 1242, and 1016 occurred, increased
mortality from cancer of the liver and gall bladder was observed in
one study and from neoplasms and cancer of the gastrointestinal tract
in the other. None of the available epidemiological studies provide
conclusive evidence of an association between PCB exposure and
increased cancer mortality, because of the small number of deaths in
exposed populations, the lack of dose-response relationships, and the
problem of contaminants in the PCB mixtures.

1.2 Conclusions

1.2.1 Distribution

Because of their physical and chemical properties, PCBs have become
dispersed globally, throughout the environment.

PCBs are almost universally present in organisms in the environment
and are readily bioaccumulated. Biomagnification in food chains has
also been demonstrated.

Higher chlorinated congeners accumulate preferentially.

1.2.2 Effects on experimental animals

The results of animal studies suggest that PCBs are immunosuppressive,
as assessed by alterations in gross measures of immune function
(spleen weight, thymus weight, and lymphocyte counts). NOELs in
monkeys have been estimated at 100 µg/kg for Aroclor 1248 and
<100 µg/kg body weight for Aroclor 1254. Immunosuppression appears to
be a congener-specific effect.

Reproductive toxicity is, in general, only seen at doses producing
systemic toxicity in the mother. Neonates feeding on contaminated
mother's milk (in monkeys and other animal species, used as models)
appear to be particularly sensitive to PCBs and show reduced growth
with other toxic symptoms. The NOEL for Aroclor 1016 on reproductive
effects is 30 µg/kg body weight for monkeys; no NOEL could be
established for the reproductive effects of Aroclor 1248.

PCBs are not genotoxic and there is inconclusive evidence for action
as tumour initiators. PCBs do act as tumour promoters. It can be
concluded that the toxicity of PCB mixtures can be evaluated on a
threshold basis.

1.2.3 Effects on humans

Exposure of the general population to PCBs will be principally through
food items. Babies will be exposed through the mother's milk.

Two large episodes of intoxication in humans have occurred in Japan
(Yusho) and Province of Taiwan (Yu-Cheng). The main symptoms in Yusho
and Yu-Cheng patients have frequently been attributed to contaminants
in the PCB mixtures, specifically, to PCDFs. The Task Group concluded
that symptoms may have been caused by the combined exposure to PCBs
and PCDFs. However, some of the symptoms, principally, the chronic
respiratory effects, may have been caused specifically by the
methylsulfone metabolites of certain PCB congeners.

1.2.4 Effects on the environment

While there have been reports of effects on local populations of
birds, the most important effect of PCBs on organisms in the
environment has been reproductive failure in sea mammals. This has
been observed principally in semi-enclosed seas and has led to
population declines, locally. The prediction that residues of PCBs in
the environment will gradually be redistributed towards the marine
environment indicates an increasing hazard for sea mammals in the
future.

1.3 Recommendations

* International agreement on analytical procedures to improve the
comparability of results of monitoring programmes is recommended.
Methodology for congener-specific analysis should continue to be
developed, though the value of analysis based on mixtures is
recognized.

* In order to ensure the reliability of analytical data,
inter-laboratory quality control studies are strongly recommended.
It is also recommended that an international network of technical
support and supervision is established, to allow developing
countries to participate in monitoring.

* Long-term studies using specific congeners, and studies on the
mechanism of action of constituents of PCBs mixtures, with special
regard to tumour promotion, are recommended to improve the
precision of the risk assessment of PCBs.

* Epidemiological studies to better assess the risk to neonates are
required, since new-born infants appear to be the most vulnerable
sector of the general population, because of high exposure through
milk.

* Sensitive and specific biomarkers for some of the more subtle
types of PCB toxicity (such as reproductive, immunological, and
neural toxicity) should be developed for use in future
epidemiological studies.

* Disposal of PCBs should be carried out by incineration in properly
designed and run facilities that can guarantee the constant high
temperatures (above 1000°C), residence time, and turbulence
necessary to ensure complete breakdown.

* Methods to remove PCBs already contained in landfills should be
investigated.

* Monitoring of PCBs in the environment and in wildlife should be
encouraged globally, to follow the expected redistribution of
residues already present.

* Marine mammals are susceptible to reproductive failure as a result
of PCB contamination. Studies on the population size and
reproductive success of cetaceans should be encouraged, together
with further research to establish which congeners are responsible
for the effects.

2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, ANALYTICAL METHODS

2.1 Identity

2.1.1 Chemical formula and structure

The chlorination of biphenyl can lead to the replacement of 1-10
hydrogen atoms by chlorine; the conventional numbering of substituent
positions is shown in the diagram:

The chemical formula can be presented as C₁₂H_10-nCl_n, where n, the
number of chlorine atoms in the molecule, can range from 1 to 10.

2.1.2 Relative molecular mass

The relative molecular mass depends on the degree of substitution.

Monochlorobiphenyl has a relative molecular mass of 188, while
completely chlorinated biphenyl (C₁₂Cl₁₀) has a relative molecular
mass of 494 (US EPA, 1980).

2.1.3 Common name

Common name: polychlorinated biphenyls (PCBs)
CAS Registry number: 1336-36-3
RTECS Registry number: TQ 1350000

2.1.4 Chemical composition

The PCBs are chlorinated hydrocarbons, manufactured commercially by
the progressive chlorination of biphenyl in the presence of a suitable
catalyst (e.g., iron chloride). Depending on the reaction conditions,
the degree of chlorination can vary between 21 and 68% (w/w). The
yield is always a mixture of different isomers and congeners. Thus, a
total of 209 theoretically different chemical components exist, but
only about 130 of these are likely to occur in commercial products or
mixtures of such compounds (Safe, 1990).

Seventy-eight out of the possible 209 PCB congeners can exist as
rotational isomers that are enantiomeric to each other. Nineteen PCBs,
of which 9 are components of commercial PCB formulations, have been
predicted to be stable at room temperature (Kaiser, 1974).

Puttmann et al. (1988) separated the atropisomers of
2,3,4,6,2',4'-hexachlorobiphenyl and demonstrated that they possess
different biological effects with regard to in vivo enzyme induction
(aminopyrine N-demethylase, aldrin epoxidase, cytochrome P-450
content, morphine UDP-glucuronosyl transferase) in Sprague-Dawley
rats.

Unlike the dioxins or dibenzofurans, the phenyl rings of a PCB are not
constrained through ring fusions and have relatively unconstrained
rotational freedom. Chlorines at the ortho (2,2', 6,6') positions
introduce constraints on rotational freedom that can hinder
coplanarity of the phenyl rings. X-ray crystallographic studies
(McKinney & Singh, 1981) indicate that the preferred conformation for
all PCBs, including those without ortho-substituents, is
noncoplanar. The proportion of molecules of a particular congener
assuming a coplanar configuration becomes increasingly small as the
degree of ortho-substitution and the energetic cost of conforming
increases. However, PCBs without ortho-substitution are often
referred to in the biological literature as the planar (or coplanar)
PCBs and all others as the nonplanar (or noncoplanar) PCBs. This
terminology, though somewhat misleading, is also used throughout this
publication for convenience and ease of referring back to the
published literature. It is widely recognized that certain biological
activities of the PCBs vary, at least quantitatively, with
stereochemical differences in the congeners.

Individual manufacturers have their own system of identification for
their products. In the Aroclor series, a 4-digit code is used;
biphenyls are generally indicated by 12 in the first 2 positions,
while the last 2 numbers indicate the percentage by weight of chlorine
in the mixture; thus, Aroclor 1260 is a polychlorinated-biphenyl
mixture containing 60% of chlorine. An exception to this
generalization is Aroclor 1016, which is a distillation product of

Aroclor 1242 containing only 1% of components with 5 or more chlorine
atoms (Burse et al., 1974). With other commercial products, the codes
may indicate the approximate mean number of chlorine atoms in the
components; thus Clophen A60, Phenochlor DP6, and Kanechlor 600 are
biphenyls with an average of about 6 chlorine atoms per molecule
(equivalent to 59% chlorine by weight).

Ballschmiter & Zell (1980) proposed a numbering system for the PCB
congeners, that was later adopted by the International Union of Pure
and Applied Chemists (IUPAC). The number, structure, and isomer group
are given for each congener in the paper of McFarland & Clarke (1989)
(see Appendix A). In the literature, the structure of a congener is
given in 2 ways; for example 2,2',5,5' or 2,5,2',5' (No 52).

Individual PCBs have been synthesized for use as reference samples in
the identification of gas-liquid chromatographic peaks, for
toxicological investigations, and for studying their metabolic fate in
living organisms, for which purpose they have been prepared labelled
with carbon-14 (Hutzinger et al., 1971; Jensen & Sundström, 1974a;
Sundström & Wachtmeister, 1975).

The proportions of PCBs with 1-9 chlorine substituents in the Aroclors
are shown in Table 1.

It is apparent, from gas chromatographic analyses of commercial
products, that PCB mixtures differ with respect to the individual
congeners present and their relative concentrations (Jensen &
Sundström, 1974a; Albro & Parker, 1979; Ballschmiter & Zell, 1980;
Albro et al., 1981; Mullin et al., 1984; Safe et al., 1985a;
Alford-Stevens, 1986).

There have been several investigations to identify individual PCBs in
commercial products. The components of the Aroclors were separated by
column and gas-liquid chromatography and many of the peaks
characterized by high-resolution mass spectrometry and nuclear
magnetic resonance, and also by comparison with synthesized PCBs
(Table 2) (see also DFG, 1988).

Jensen & Sundström (1974a) recognized that conventional gas-liquid
chromatography was not suitable for separating all the components, so
they devised a preliminary fractionation on a charcoal column, which
separated the component PCBs according to the number of chlorines in
the 2,6,2' or 6' positions in the molecule ( o-chlorines). They
compared the gas-liquid chromatographic peaks with those of 90
synthesized PCBs, and were able to characterize and quantify 60
components of Clophens A50 and A60.

Table 1. Approximate percentages (w/v) of Aroclors with different degrees of
chlorination^a

Number of Chlorine
chlorine weight Aroclor
atoms in (%)
molecule 1221 1232 1016 1242 1248 1254 1260

0 0 10 - -
1 18.8 50 26 2 3
2 31.8 35 29 19 13 2
3 41.3 4 24 57 28 18
4 48.6 1 15 22 30 40 11
5 54.4 22 36 49 12
6 59.0 4 4 34 38
7 62.8 6 41
8 66.0 8
9 68.8 1

^a From: WHO/EURO (1987).

2.1.5 Technical product

Major trade names

The PCBs manufactured commercially are known by a variety of trade
names including: Aroclor, Pyranol, Pyroclor (USA), Phenoclor, Pyralene
(France), Clophen, Elaol (Germany), Kanechlor, Santotherm (Japan),
Fenchlor, Apirolio (Italy), and Sovol (USSR). Table 3 contains the
most common trade names for commercial products, some of which are not
in use any more (Brinkman & De Kok, 1980; WHO/EURO, 1987).

2.1.6 Purity and impurities

Commercial PCBs are not sold according to a composition specification,
but according to their physical properties. The composition of
Aroclors and Clophens has been presented in recent papers; the
composition of 5 Aroclors is shown in Tables 1 and 2. In Table 1, the
approximate composition is expressed as the percentage of chlorine
weight, and, in Table 2, the composition of the chlorine substitution
pattern is expressed in mol % (Albro & Parker, 1979; Albro et al.,
1981; Jones, 1988). The composition of the chlorine substitution
pattern for 4 Clophens is described by Duinker & Hillebrand (1983) and

Jones (1988). It should be kept in mind that nothing can be said about
the variations in the different lots of these mixtures. Impurities
known to be present in commercial PCBs are chlorinated dibenzofurans
and chlorinated naphthalenes (Vos et al., 1970; Bowes et al., 1975;
Albro & Parker, 1979; Albro et al., 1981; Duinker & Hillebrand, 1983;
Rappe et al., 1985a). The concentrations of PCDFs in Aroclor, Clophen,
Phenoclor, and Kanechlor are summarized in Tables 4 and 5.

Different authors have examined the presence of PCDFs in PCB mixtures.
Bowes et al. (1975) found 0.8-2.0 mg/kg in samples of Aroclor 1248 and
1260, but none in Aroclor 1016, 8.4 mg/kg in Clophen A60, and
13.6 mg/kg in Phenoclor DP-6. Rappe et al. (1985a) and Bentley (1983)
found levels of PCDFs up to 40 mg/kg in a number of commercial PCBs.
Recently, Wakimoto et al. (1988) found a number of extremely toxic
PCDFs in several Japanese and American commercial PCB preparations.
These isomer-specific analyses revealed the 2,3,7,8-tetra-,
1,2,4,7,8-penta-, 1,2,3,7,8-penta-, 2,3,4,7,8-penta-, and
1,2,3,6,7,8-hexachlorodibenzofurans. The concentrations in unused
Kanechlor 300, 400, 500, and 600, were 7.5, 26, 7.2, and 5.4 mg/kg,
respectively, and those in Aroclors 1242, 1248, 1254, and 1260, were
0.6, 3.7, 4.2, and 7.5 mg/kg, respectively. Brown et al. (1988) found
that the electrical use of PCB dielectric fluids in transformers and
capacitors did not increase the PCDFs content significantly.

More data about the occurrence of PCDFs in the different commercial
PCB mixtures are summarized in WHO/EURO (1987).

There are no reports on the presence of PCDDs in commercial mixtures
(Bowes et al., 1975). Wakimoto et al. (1988) could not find PCDDs in
the above samples of Kanechlors and Aroclors with a detection limit of
<2 µg/kg.

2.2 Physical and chemical properties

Individual pure PCB congeners are colourless, often crystalline
compounds, but commercial PCBs are mixtures of these congeners with a
clear, light yellow or dark colour. They do not crystallize at low
temperatures, but turn into solid resins. Because of the chlorine
atoms in the molecule, their density is rather high. PCBs are, in
practice, fire resistant with rather high flash-points (170-380°C).
They form vapours heavier than air, but do not form any explosive
mixtures with air. They possess very low electrical conductivity and
an extremely high resistance to thermal breakdown, and it is on the
basis of these properties that they are used as cooling liquids in
electrical equipment (US EPA, 1980; WHO/EURO, 1987; DFG, 1988).

Table 2. PCB compositions of aroclors in mol %^a

IUPAC Chlorine Aroclor
No. substitution
pattern 1242 1016 1248 1254 1260

BP 0.01 0.50
1 2 0.68 0.80
2 3 0.04 0.10
3 4 0.22 1.00
4 2.2' 3.99 4.36 0.25
6 2.3' 1.24 1.37 0.69 0.07
7 2.4 1.04 1.16
8 2.4' 8.97 10.30 0.18
9 2.5 0.31 0.34 trace
10 2.6 0.13 0.20
12 3.4 0.09 0.11
13 3.4' 0.12 0.12
14 3.5 0.35 0.37
15 4.4' 0.99 1.07
16 2.3.2' 3.25 3.50 0.84
17 2.4.2' 2.92 3.14 0.19
18 2.5.2' 9.36 10.87 9.95 0.07
19 2.6.2' 0.97 1.08
20 2.3.3' 3.64 3.99
22 2.3.4' 2.64 2.80 1.24 trace trace
25 2.4.3' 1.68 1.79
26 2.5.3' 0.55 0.62 0.75
27 2.6.3' 0.54 0.58
28 2.4.4' 13.30 14.48 trace
31 2.5.4' 4.53 4.72 9.31 0.72
32 2.6.4' 2.15 2.31 1.46
33 3.4.2' 2.83 3.08
35 3.4.3' 0.66 0.38
37 3.4.4' 1.62 1.89 1.28 0.20 0.09
39 3.5.4' 1.03 1.08
40 2.3.2'.3' 0.15 0.18 1.12 0.26 0.04
41 2.3.4.2' 1.67 2.00
42 2.3.2'.4' 7.05 2.18 0.66
43 2.3.5.2' 0.44 0.47
44 2.3.2'.5' 1.06 1.14
45 2.3.6.2' 0.90 1.00 5.73 0.15
46 2.3.2'.6' 0.31 0.33
47 2.4.2'.4' 1.65 1.8 3.18 0.52 0.88
48 2.4.5.2' 1.33 1.41

Table 2. (cont'd).

IUPAC Chlorine Aroclor
No. substitution
pattern 1242 1016 1248 1254 1260

? 2.5.2'.4' - - 3.81 1.63 0.44
49 2.4.2'.5' 3.28 3.48
52 2.5.2'.5' 4.08 4.35 8.36 4.36 1.91
53 2.5.2'.6' 0.97 1.07 6.30 0.13
54 2.6.2'.6' 0.17 0.19
55 2.3.4.3' 0.11 0.43 0.12
56 2.3.3'.4' 0.60 trace 0.18 0.03
60 2.3.4.4' 0.21
66 2.4.3'.4' 0.81 0.14 4.95 2.24 0.22
70 2.5.3'.4' 1.11 6.38 4.75 0.85
71 2.6.3'.4' 0.65
72 2.5.3'.5' 0.33 2.10 1.01 0.28
74 2.4.5.4' 2.02 1.35 0.25 0.30 0.09
75 2.4.6.4' 2.18 2.40
76 3.4.5.2' trace trace 0.18 0.01
77 3.4.3'.4' 0.34 0.47 0.12 0.04
78 3.4.5.3' 0.52
79 3.4.3'.5' 0.24 trace 0.23 0.04
80 3.5.3'.5' trace trace trace
81 3.4.5.4' 0.28
83 2.3.5.2'.3' trace 0.32 0.09
84 2.3.6.2'.3' 0.38 0.01 0.71 1.72 0.69
85 2.3.4.2'.4' 0.40 0.55 2.15 0.31
? 2.3.4.3'.5' 0.02 0.55 0.14
87 2.3.4.2'.5' 0.09 1.05 3.81 1.10
91 2.3.6.2'.4' trace 1.78 5.00 3.22
92 2.3.5.2'.5' 0.12 0.20 0.63 0.21
95 2.3.6.2'.5' 0.53 0.18
97 2.4.5.2'.3' 0.78 2.59 0.63
98 2.4.6.2'.3' 0.13 0.04
99 2.4.5.2'.4' 0.55 2.52 6.10 0.82
101 2.4.5.2'.5' 0.27 1.50 6.98 5.04
102 2.4.5.2'.6' trace trace trace
105 2.3.4.3'.4' 0.25
106 2.3.4.5.3' 0.40 0.06
108 2.3.4.3'.5' 0.46 0.16
110 2.3.6.3'.4' 1.69 8.51 3.57
113 2.3.6.3'.5' 0.39 0.01 3.10 trace 0.01
114 2.3.4.5.4' 0.25 0.03
118 2.4.5.3'.4' 8.09 2.00
120 2.4.5.3'.5' 0.31 trace 0.15 3.01
121 2.4.6.3'.5' 0.92 4.32 3.51 0.57

Table 2. (cont'd).

IUPAC Chlorine Aroclor
No. substitution
pattern 1242 1016 1248 1254 1260

123 3.4.5.2'.4' 0.36
? 3.4.5.2'.3' trace 0.76 1.88
126 3.4.5.3'.4' 0.03 0.16 1.59
127 3.4.5.3'.5' 0.05
128 2.3.4.2'.3'.4' 1.31 0.47
131 2.3.4.6.2'.3' 0.14 0.01
132 2.3.4.2'.3'.6' trace 2.00 2.77
133 2.3.5.2'.3'.5' 1.13 0.03 0.06
134 2.3.5.6.2'.3' 0.11 0.38 1.01
135 2.3.5.2'.3'.6' 0.20 0.29
136 2.3.6.2'.3'.6' 0.20 0.34 1.12
138 2.3.4.2'.4'.5' 0.08 0.19 4.17 5.01
143 2.3.4.5.2'.6' 0.07
148 2.3.5.2'.4'.6' 0.12 0.07 0.06
149 2.4.5.2'.3'.6' 0.77 3.59 9.52
151 2.3.5.6.2'.5' trace 0.33 0.06
153 2.4.5.2'.4'.5' 0.02 0.13 3.32 8.22
154 2.4.5.4'.6' 0.14
156 2.3.4.5.3'.4' 0.41
157 2.3.4.3'.4'.5' 0.18 0.03
158 2.3.4.6.3'.4' 0.46 0.18
159 2.4.5.2'.3'.5' 0.75 1.48
163 2.3.5.6.3'.4' trace
167 2.4.5.3'.4'.5' 0.21 0.17
168 2.4.6.3'.4'.5' 0.56 4.23 0.59
170 2.3.4.5.2'.3'.4' 0.43 0.62
171 2.3.4.6.2'.3'.4' 0.30 4.31
174 2.3.4.5.2'.3'.6' trace 0.09
176 2.3.4.6.2'.3'.6' 0.09 trace 0.57
177 2.3.5.6.2'.3'.4' trace
179 2.3.5.6.2'.3'.6' 0.56 0.83
180 2.3.4.5.2'.4'.5' 0.76 7.20
181 2.3.4.5.6.2'.4' 0.28 2.72
182 2.3.4.5.2'.4'.6' trace 0.47
183 2.3.4.6.2'.4'.5' 1.16 2.58
185 2.3.4.5.6.2'.5' 1.11 5.65
186 2.3.4.5.6.2'.6' trace trace 0.37
187 2.3.5.6.2'.4'.5' 0.48 1.12
189 2.3.4.5.3'.4'.5' 0.13
190 2.3.4.5.6.3'.4' 0.02
192 2.3.4.5.6.3'.5' 0.20 0.97

Table 2. (cont'd).

IUPAC Chlorine Aroclor
No. substitution
pattern 1242 1016 1248 1254 1260

193 2.3.5.6.3'.4'.5' 2.30
194 2.3.4.5.2'.3'.4'.5' 2.21
195 2.3.4.5.6.2'.3'.4' trace
196 2.3.4.5.2'.3'.4'.6' 0.79
197 2.3.4.6.2'.3'.4'.6' 0.30
198 2.3.4.5.6.2'.3'.5' 1.00 0.15
199 2.3.4.5.6.2'.3'.6' 0.38
200 2.3.4.6.2'.3'.5'.6' trace 0.15
202 2.3.5.6.2'.3'.5'.6' trace 0.31
203 2.3.4.5.6.2'.4'.5' 0.08
204 2.3.4.5.6.2'.4'.6' trace 0.13
205 2.3.4.5.6.3'.4'.5' 0.01
206 2.3.4.5.6.2'.3'.4'.5' 0.51
207 2.3.4.5.6.2'3'.4'.6' 1.15
208 2.3.4.5.6.2'.3'.5'.6' 1.64
? 2.3.4.5.6.2'.3'.5'.6' 0.18

^a From: Albro & Parker (1979); Albro et el. (1981).

Table 3. The trade marks of PCB products and mixtures containing PCBs^a

Aceclor (t) Disconon (c) PCBs
Apirolio (t,c) Dk (t,c) Phenoclor (t,c)
Aroclor (t,c) Duconol (c) Polychlorinated biphenyl
Arubren Dykanol (t,c) Polychlorobiphenyl
Asbestol (t,c) EEC-18 Pydraul^c
Askarel Elemex (t,c) Pyralene (t,c)
Bakola 131 (t,c) Eucarel Pyranol (t,c)
Biclor (c) Fenchlor (t,c) Pyroclor (t)
Chlorextol (t) Hivar (c) Saf-T-Kuhl (t,c)
Chlorinated Biphenyl Hydol (t,c) Santotherm FR^b
Chlorinated Diphenyl Inclor Santovac 1 and 2
Chlorinol Inerteen (t,c) Siclonyl (c)
Chlorobiphenyl Kanechlor (t,c) Solvol (t,c)
Clophen (t,c) Kennechlor Sovol
Clorphen (t) Montar Therminol FR^b
Delor Nepolin
Diaclor (t,c) No-Flamol (t,c)
Dialor (c) PCB

^a From: WHO/EURO (1987).
^b Previous products (FR-series) used as pressure oil contained PCBs, but current
products are a different series and do not contain PCBs.
^c Previous products (A-series) e.g., PYDRAUL A-200 contained PCBs, but current
commercial products are B, C, or D-series and do not contain any chlorinated
compounds.

(t) Used in transformers.
(c) Used in capacitors.

Table 4. Concentrations of chlorinated dibenzofurans^a in Aroclor, Clophen, and
Phenoclor^b

PCB 4-Cl 5-Cl 6-Cl Total

Aroclor 1248 (1969) 0.5 (25) 1.2 (60) 0.3 (15) 2.0
Aroclor 1254 (1969) 0.1 (6) 0.2 (12) 1.4 (82) 1.7
Aroclor 1254 (1970) 0.2 (13) 0.4 (27) 0.9 (60) 1.5
Aroclor 1260 (1969) 0.1 (10) 0.4 (40) 0.5 (50) 1.0
Aroclor 1260 (lot AK3) 0.2 (25) 0.3 (38) 0.3 (38) 0.8
Aroclor 1016 (1972) ND ND ND
Clophen A-60 1.4 (17) 5.0 (59) 2.2 (26) 8.4
Phenoclor DP-6 0.7 (5) 10.0 (74) 2.9 (21) 13.6

^a Expressed as mg PCB/kg. Values in parentheses represent quantity as percentage
of total dibenzofurans.
^b From: Bowes et al. (1975).
ND = not detected (0.001 mg/kg).

Table 5. Concentrations of chlorinated dibenzofurans in Kanechlors^a

Kanechlor Chlorodibenzofurans Concentration
(mg/kg)

Di- Tri- Tetra- Penta- Hexa- Hepta- b c

300 + + 1 1.5
400 + + + + 18 17
500 + + + + 4 2.5
600 + + + + 5 3

^a From: Nagayama et al. (1975).
^b Calculated from peak heights.
^c Calculated by perchlorination method.

PCBs have a high degree of chemical stability under normal conditions.
They are very resistant to a range of different oxidants and other
chemicals. According to laboratory tests, they stay chemically
unchanged, even in the presence of oxygen or some active metals at
high temperatures (up to 170°C) and for protracted periods (WHO/EURO,
1987).

PCBs are practically insoluble in water, whereas they dissolve easily
in hydrocarbons, fats, and other organic compounds and they are
readily absorbed by fatty tissues (WHO/EURO, 1987).

Some physical and chemical data for a number of Aroclors are presented
in Table 6.

Foreman & Bidleman (1985) estimated the liquid phase vapour pressures,
at 25°C, of 134 PCB congeners found in 5 Aroclor fluids, using a
capillary gas chromatographic method in conjunction with published
retention indices of PCBs.

Burkhard et al. (1985) predicted Henry's Law Constants from the ratio
of the liquid (or subcooled liquid) vapour pressure and aqueous
solubility for PCB congeners. The predicted values were in fair
agreement with experimental values and the error for these constants
was estimated to be a factor of 5 in the temperature range of 0-40°C.
For the PCB congeners, Henry's Law Constants were independent of the
relative molecular mass and increased approximately an order of
magnitude with a 25°C increase in temperature.

Aqueous solubility is considered an essential parameter for predicting
the fate and transport of organic chemicals in the environment. As
already stated, some physical and chemical data are given for 6
Aroclor mixtures in Table 6 (Alford-Stevens, 1986). However, during
the last 5 years, much more information on aqueous solubility, melting
points, entropies of melting, Henry's law constants, and vapour
pressures has become available. This information concerns not only PCB
mixtures but also individual congeners.

Opperhuizen et al. (1988) studied the aqueous solubilities of 45
chlorinated biphenyls and the relationships between activity
coefficient and chemical structure parameters (total surface area
(TSA) and total molecular volume (TMV)) of hydrophobic chemicals, to
understand the nature of hydrophobicity. The aqueous solubilities of
PCBs showed a linear relationship between logarithms of aqueous
activity coefficients or TSA and TMV.

Table 6. Physical and chemical properties of a number of Aroclors^a

Substance Water Vapour Density Appearance Henry's Law Refractive index Boiling point
Aroclor solubility pressure (g/cm³) constant (distillation
(mg/litre) (torr) 25°C 25°C (atm-m³/mol range) (750
25°C at 25°C)^b torr, °C)

1016 0.42 4.0 × 10^-4 1.33 Clear, mobile oil 2.9 × 10^-4 1.6215-1.6235 325-356
(at 25°C)

1221 0.59^c 6.7 × 10^-3 1.15 Clear, mobile oil 3.5 × 10^-3 1.617-1.618 (at 20°C) 275-320

1232 0.45 4.1 × 10^-3 1.24 Clear, mobile oil unknown unknown 290-325

1242 0.24 4.1 × 10^-3 1.35 Clear, mobile oil 5.2 × 10^-4 1.627-1.629 (at 20°C) 325-366

1248 0.054 4.9 × 10^-4 1.41 Clear, mobile oil 2.8 × 10^-3 unknown 340-375

1254 0.021 7.7 × 10^-5 1.50 Light yellow 2.0 × 10^-3 1.6375-1.6415 365-390
viscous oil (at 25°C)

1260 0.0027 4.0 × 10^-5 1.58 Light yellow 4.6 × 10^-3 unknown 385-420
sticky resin

^a From: IARC (1978); WHO/EURO (1987); ATSDR (1989).
^b These Henry's Law Constants were estimated by dividing the vapour pressure by the water solubility. The first water solubility
given in this table was used for the calculation. The resulting estimated Henry's law constant is only an average for the
entire mixture; the individual chlorobiphenyl isomers may vary significantly from the average. Burkhard et al. (1985)
estimated the following Henry's Law Constants (atm-m3/mol) for various Aroclors at 25°C: 1221 (2.28 × 10^-4), 1242 (3.43 × 10^-4),
1248 (4.4 × 10^-4), 1254 (2.83 × 10^-4), 1260 (4.15 × 10^-4).
^c At 24°C.

Dickhut et al. (1986) studied the solubilities of 6 higher chlorinated
biphenyl congeners at different temperatures and found that the
solubility increased exponentially with temperature in the range of
0.4-80°C. From the temperature dependence of solubility, enthalpies of
solution were calculated. The same results were found by Doucette &
Andren (1988), who determined the aqueous solubilities of a few PCBs,
using a generator-column technique, at temperatures of 4.0, 25.0, and
40.0°C.

The dissolution of extremely hydrophobic chemicals that may be
associated with a relatively constant endothermic enthalpy of solution
and an endothermic enthalpy of fusion that is proportional to the
solute's melting point is discussed by Opperhuizen et al. (1987) and
Dickhut et al. (1987).

Dunnivant & Elzerman (1988) estimated the aqueous solubilities and
Henry's Law Constants (HLC) for 26 selected PCB congeners for the
evaluation of quantitative structure-property relationships (QSPRs).
Aqueous solubilities (as solids at 25°C, column generation technique),
determined for the 26 congeners, ranged from 1.08 × 10^-5 to
9.69 × 10^-10 mol/litre and generally decreased with relative molecular
mass. HLCs (25°C, gas purge technique), determined for 20 congeners,
ranged from 0.3 × 10^-4 to 8.97 × 10^-4 atm.m³/mol. Measured HLCs were
not correlated with relative molecular mass, but increased with the
degree of ortho-chlorine substitution within a relative molecular
mass class.

Vapour pressures calculated from the product of solubility (mol/m³)
and HLC (atm-m³/mol) data, generally decreased with relative
molecular mass and increased with increasing degree of
ortho-chlorine substitution (Dunnivant & Elzerman, 1988; Hawker,
1989). Westcott et al. (1981) used a semimicro gas saturation method
to determine the vapour pressures of 3 PCB isomers and 2 Aroclor
mixtures.

Experimental data were tabulated and the relationships between the
environmentally relevant physical chemical properties of the PCBs
critically reviewed by Shui & Mackay (1986). Aqueous solubility,
vapour pressure, Henry's law constant, and octanol-water partition
coefficient were discussed and recommended values given for 42 of the
209 congeners; procedures were suggested for estimating the properties
of the other congeners.

2.2.1 Log n-octanol/water partition coefficient

The environmental fate of PCBs is governed primarily by the
partitioning process. Partitioning processes that are of particular
interest with regard to environmental problems include: the octanol/
water partition coefficient and the aqueous solubility. The octanol/
water partition coefficient is a measure of the hydrophobicity of a
substance and, in this respect, it has been used to predict the extent
of bioconcentration of organic pollutants in organisms. Miller et al.
(1984) studied the octanol/water partition coefficients for 16 PCBs
and Hawker & Connell (1988) for 13 PCB congeners, using the generator
column method. These partition coefficients were used to confirm a
highly significant linear relationship between log K_ow and the
logarithm of the relative retention time on a nonselective gas
chromatographic stationary phase. The total surface areas (TSA) for
all the PCB congeners were determined by assuming planar molecules,
van der Waal's radii for component atoms, and appropriate values for
solvent radius, bond angles, and distances. The TSA was highly
significantly correlated with log K_ow and the relationship was used
to calculate log K_ow values for all the PCB congeners. In the report
of Hawker & Connell (1988), log K_ow values are summarized for all 209
PCB congeners. These log K_ow values range from 4.46 to 8.18.

2.2.2 Conversion factors^a

Aroclor
1016 1 mg/m³ = 0.095 ppm
1221 1 mg/m³ = 0.12 ppm
1232 1 mg/m³ = 0.105 ppm
1242 1 mg/m³ = 0.092 ppm
1248 1 mg/m³ = 0.008 ppm
1254 1 mg/m³ = 0.075 ppm
1260 1 mg/m³ = 0.065 ppm

2.3 Analytical methods

Reviews have been published on the methods used for the determination
of organochlorine compounds including PCBs in environmental samples
(Panel on Hazardous Trace Substances, 1972; Holden, 1973; US DHEW,
1978; Slorach & Vaz, 1983; Jensen, 1984, 1985; Erickson 1985;
Alford-Stevens, 1986; NIOSH, 1987; DFG, 1988; WHO/EURO, 1987, 1988).

^a These air conversion factors were calculated by using the average
molecular mass at 25°C.

No two laboratories used identical methods, though all the methods
have features in common. The techniques appear to be those previously
developed for the determination of organochlorine pesticides, with
appropriate modifications for the presence of PCBs, and the studies on
PCBs sometimes form part of a wider programme for monitoring
persistent organochlorine compounds in the environment. In the past,
the major difficulty in the determination of PCBs was to obtain a
single quantitative figure from a variable mixture of components. The
PCBs were chlorinated with antimony pentachloride to decachloro-
biphenyl, which was measured as a single peak (Greve & Wegman, 1983;
Tuinstra, 1983). At the moment, chemists and toxicologists are no
longer trying to derive a single quantitative figure, preferring
instead to quantify individual congeners. The legislation in certain
countries is now based on quantifying a few selected congeners,
instead of reporting "total PCBs". It is also felt that for
pinpointing areas with high levels of contamination, in order to rank
them into low, medium, or high priority areas for action, highly
accurate laboratory analyses are not necessary; instead, analytical
competence and the use of adequate controls and standards, resulting
in consistent, reasonably accurate results would be enough. Of course,
for complicated research, especially involving laboratories in
different countries, standardization of techniques through
collaborative and comparative studies would be necessary.

Jones (1988) and Safe et al. (1985a) studied the occurrence of
specific PCB congeners in commercial formulations or mixtures. The
congener composition of commercial formulations differs from
batch-to-batch, between manufacturing processes, and with the level of
chlorination. The presence of congeners in the environment will depend
on the eventual use of commercial formulations, the quantity of each
formulation manufactured, as well as on the isomer composition of the
source.

On the basis of a literature review of the occurrence of PCB congeners
in environmental and biological samples and human tissues, and
consideration of the relative toxicity and persistence of the
congeners, suggestions were made by Jones (1988), with regard to the
most relevant components to be quantified in human foodstuffs and
tissues, using a selective analytical approach.

The congeners reported (Safe et al., 1985a; Duinker et al., 1988;
McFarland & Clarke, 1989) as being the most abundant in human tissues
and which are most important, are compounds with IUPAC numbers 28, 52,
74, 77, 99, 101, 105, 118, 126, 128, 138, 153, 156, 169, 170, 179, and
180 (comprising >70% of total PCBs and being of greatest

toxicological significance). Because of their reported occurrence or
toxicity, congeners with IUPAC numbers 8, 37, 44, 49, 60, 66, 70, 82,
87, 114, 158, 166, 183, 187, and 189 might also be considered. Duinker
et al. (1988) were also of the opinion that toxicity should be
considered as a criterion for the selection of PCB congeners for
analysis in environmental samples. Most of these congeners can be
accurately determined with the application of the multidimensional,
high-resolution GC-ECD techniques.

PCB reference materials are necessary for the qualitative and
quantitative calibration of analytical apparatus and methods (e.g.,
determination of retention times, response factors, and reference
spectra in chromatographic and spectroscopic analyses) and for the
study of biological activity. Lindsey & Wagstaffe (1989) described the
production and certification of 10 high-purity PCBs with IUPAC numbers
8, 20, 28, 35, 52, 101, 118, 138, 153, and 180.

Mes et al. (1989a) described an analytical method to determine 34
isomers of PCB congeners in fatty foods. A sample was extracted with
an acetone:hexane mixture and the extracts washed and dried; this was
followed by a clean-up and determination by gas chromatography. GC/MS
was used for confirmation.

Environmental PCB residues are often expressed in terms of relative
Aroclor composition. Schwartz et al. (1987) assessed the similarity of
Aroclors with class models derived for fish and turtles, to ascertain
if the PCB residues in the samples could be described by an Aroclor or
Aroclor mixture. The PCB residues in fish and turtles were analysed
with Soft Independent Modelling of Class Analogy, a principal
components analysis (PCA) technique. Using PCA, it was inappropriate
to report these samples as an Aroclor or Aroclor mixture.

2.3.1 Sampling strategy and sampling methods

The quality and usefulness of analytical data, especially in the
microgram-nanogram range, or even lower, depend critically on the
validity of the sample and the adequacy of the sampling programme. The
purpose of sampling is to obtain specimens that represent the
situation being studied. Sampling plans may require that systematic
samples be obtained at specified times and places, or simple random
sampling may be necessary. Generally, the sample should be an unbiased
representative of the situation of interest (WHO/EURO, 1987). Slorach
(1984) described the problems encountered with the sampling and
determination of PCBs in breast milk (see also WHO/EURO, 1985, 1988).

All aspects of a sampling programme should be planned and documented
in detail, and the expected relationship of the sampling protocol to
the analytical result should be defined. A sampling programme should
include reasons for choosing sampling sites, the number and type of
samples, the timing of sample acquisition, and the sampling equipment
used. A detailed sampling procedure should include a description of
the sampling situation, the sampling methodology, labelling of
samples, field blank preparation, pretreatment procedures,
transportation, and storage (WHO/EURO, 1987).

The quality assurance programme should include means to demonstrate
that containers or storage procedures do not alter the qualitative or
quantitative composition of the sample. Special transportation and
storage procedures (refrigeration or exclusion of light) should be
described (WHO/EURO, 1987).

Because environmental samples are typically heterogeneous, a
sufficiently large number of samples (10 or more) must be analysed to
obtain meaningful composition data. The number of individual samples
that should be analysed will depend on the kind of information
required. If an average composition value is required, a number of
randomly selected individual samples may be obtained, combined, and
blended to provide a homogeneous composite sample, from which a
sufficient number of subsamples are analysed. If composition profiles,
time trends, or the variability of the sample population is of
interest, many samples need to be collected and analysed individually.

If field blanks are not available, efforts should be made to obtain
blank samples that best simulate a sample that does not contain the
analyte. In addition, measurements should be made to ascertain
whether, and to what extent, any reagent or solvent used may
contribute or interfere with the analytical results (laboratory and
solvent blanks). The recovery tests are frequently used and are
necessary to evaluate the analytical methodology. Uncontaminated
samples from control sites that have been spiked with the analyte of
interest provide the best information, because they simulate any
matrix effect. When feasible, isotopically labelled (¹³C, ³⁷Cl)
analytes spiked into the sample provide the greatest accuracy, since
they are subjected to the same matrix effects as the analytes. The
¹³C-labelled compounds can be used to:

(a) validate sampling (sampling surrogate);

(b) validate analytical waste (clean-up surrogate);

Only a small number of laboratories in the world have access to, and
experience in working with, these complicated analyses. In order to be
able to compare data generated in different laboratories, the same
quantitative standard compounds should be used. Interlaboratory
calibrations, or "round-robin" studies, have been performed in a few
cases (WHO/EURO, 1987).

2.3.1.1 Extraction procedures

Air

The sampling device used to collect and determine PCBs in air consists
of a glass fibre filter and a Florisil stick. The glass fibre filter,
held in a stainless steel holder, removes particles larger than
0.3 µm. The air passes from the filter to the Florisil stick, which is
made in 2 sections, to provide information on migration and trapping
efficiency for PCBs. Each section contains 0.4 g of Florisil preceded
and followed by a glass wool plug. The front and back sections are
separated by 2 plugs of glass wool. The front is spiked with 0.1 µg of
p,p'-DDE as a surrogate for recovery measurement and as an indication
of analyte migration. The detection limit for PCBs in air is reported
to be 0.3 ng/m³ (Anon., 1985; WHO/EURO, 1987; NIOSH, 1987).

Particulate fallout from air has been trapped on 200 µm nylon net
coated with silicone oil, and the PCBs then extracted with hexane
(Södergren, 1972). Separate determinations of particulate and vapour
phase PCBs in air have been made by passing a large volume of air
through a filter followed by an impinger containing hexane or toluene
(Rappe et al., 1985c), a polyurethane plug (Bidleman & Olney, 1974),
or ceramic saddles coated with OV 17 silicone (Harvey & Steinhauer,
1974) to absorb the vapour.

Surface sampling

Surface sampling of PCBs can be carried out using a wet-wipe procedure
with a cotton gauze pad that has been dampened with hexane before
collecting the sample. The sampled area is 0.25 m². The wet-wipe
sampling procedure collects both the contaminants from the surface and
the contaminants that can be extracted from pores in the material.
Materials such as waxes and plasticizers may interfere with the
chemical analysis (WHO/EURO, 1987).

Another sampling method has been described by Rappe et al. (1985c),
where a dry filter paper or Kleenex tissue is used first, for wiping,
followed by a wet wipe with water-dampened material.

Water

PCBs have been extracted from water by passing a sample through a
filter of undecane and Carbowax 400 monostearate supported on
Chromosorb W (Ahling & Jensen, 1970) or a porous plug of polyurethane
coated with a suitable gas-liquid chromatographic stationary phase, or
Amberlite XAD-2 resin (Harvey et al., 1973) followed by elution of the
PCBs with a solvent. Ahnoff & Josefsson (1974, 1975) have described
liquid-liquid extraction into cyclohexane.

Soil and sediment

In a study by Huckins et al. (1988), sediment samples were thawed at
room temperature and placed in a hexane-rinsed foil pan and air dried
for 5 days. The sediment was broken up, homogenized, and mixed with
anhydrous disodium sulfate until dry, for column extraction. The
samples were extracted with methylene chloride. PCB residues were
enriched by adsorption column chromatography on silica gel and
sulfuric acid silica gel. Prior to GC analysis, nitric acid-rinsed
copper wool was added to the sediment extract to remove elemental
sulfur. An aliquot of the PCB residues was diluted in a mixture
methylene chloride: cyclohexane (1:1) and the bulk of the o,o-Cl
substituted PCB components eliminated by eluting the column with
different solvents. The different PCB congeners were determined by
GC-ECD.

The feasibility of cleaning PCB-contaminated soils using a solvent
extraction method was studied by Reilly et al. (1986). Compared with
direct incineration of the sludge, the solvent extraction route has a
number of shortcomings; the detailed design of the extraction plant as
well as its operation will be quite challenging as an extremely
leak-tight operation is essential, considering the nature of the
material handled. Direct incineration will clean the solids much more
thoroughly than is feasible by solvent extraction under ambient
conditions. Furthermore, it is inevitable that some residual solvent
will remain in the solids after processing. The solvent extraction
process costs essentially the same as direct incineration.

Biological samples

Most analysts have used standard methods, developed for organochlorine
pesticides, in which the PCBs are extracted together with the fat; the
sample is ground with anhydrous sodium sulfate and extracted with
petroleum ether or hexane. Porter et al. (1970) studied the optimal
conditions for this procedure. A dehydrating solvent may be included
to facilitate the breakdown of cell structures; ethanol (Norén &
Westöö, 1968) and acetone (Jensen et al., 1973) have been used.

Reznicek (1987) described a method to extract and determine PCBs in
blood. The sensitivity of the method was 10 µg/litre.

2.3.1.2 Sample clean-up

Diverse extraction and clean-up procedures have been devised to
preferentially remove co-extractives that are present in different
matrices and interfere with routine quantitative gas chromatographic
and gas chromatographic-mass spectrometric analysis.

The analysis of lipid-containing matrices for residues of
organochlorine pesticides and PCBs is a common procedure. All the
methods require the separation of the residues from the lipids prior
to the determination of the PCBs by gas chromatography. The removal of
the lipids is usually carried out by low-resolution column
chromatography using an adsorbent, such as silica, alumina, or
Florisil as the stationary phase. Low-resolution gel permeation
chromatography has also been used. An electron-capture detector is the
most commonly used detector, but clean-up procedures may still leave
electron-capturing species in the extract, so the identities of the
eluting peaks must be confirmed. In order to overcome some of these
problems, perchlorination of the PCBs has been used, giving rise to
one GC peak (decachlorobiphenyl), which is well removed from most
interfering peaks, but this technique has been found to be
qualitatively and quantitatively unreliable and unsatisfactory.
Seymour et al. (1986b,c) attempted to simplify clean-up procedures by
using high performance liquid chromatography (HPLC) coupled with gas
chromatography-mass spectroscopy. This latter technique is less
expensive than it used to be and is the only technique that can
possibly identify each peak as a PCB before quantification is carried
out, thereby improving the quality of the result. It is also capable,
when used in the selective ion monitoring mode (SIM), of detecting
only PCBs, even in the presence of pesticides, so that sample clean-up
is further simplified.

Seymour et al. (1986a) described a clean-up procedure, with a
preparative, high-performance liquid chromatographic (HPLC) separation
method for selected pairs of chlorobiphenyl isomers, produced by
Cadogen coupling in the preparation of individual congeners, to be
used as standards in congener-specific determination using capillary
GC methods.

A routine method for the determination of PCBs in breast milk,
described by Seymour et al. (1987), is less labour-intensive and more
cost effective than the traditional methods. These advantages were
achieved by adsorption of the milk on a polar substrate prior to
Soxhlet extraction, using a polymeric HPLC column for the clean-up of
the extract, followed by highly selective capillary GC-MS analysis.

Methods for the removal of fat from the extract include solvent
partitioning between hexane and acetonitrile or dimethylformamide, or
treatment with strong sulfuric acid or ethanolic potassium hydroxide.
Gel permeation has also been used (Stalling et al., 1972), and Holden
& Marsden (1969) removed fat on dry, partially deactivated, alumina
columns. Certain pesticides, such as dieldrin, are destroyed by the
sulfuric acid treatment, so this method cannot be used if such
pesticides are to be determined together with PCBs (Jensen et al.,
1973).

Huckins et al. (1988) described the clean-up of fish samples. Tissue
samples were thawed, mixed, dried with sodium sulfate, and extracted
in glass columns with methylene chloride. The extract was evaporated
and the lipid content was determined gravimetrically. Gel permeation
chromatography was used for removal of lipid from fish sample
extracts. PCB residues were enriched by adsorption column
chromatography on silica gel and sulfuric acid silica gel, eluted with
a mixture of methylene chloride and cyclohexane, and determined by
GC-ECD.

PCBs can be separated from organochlorine pesticides by column
chromatography on Florisil (Mulhern et al., 1971), silica gel (Holden
& Marsden, 1969; Armour & Burke, 1970; Collins et al., 1972) or on
charcoal (Berg et al., 1972; Jensen & Sundström, 1974a). Several
laboratories have reported difficulties in repeating results obtained
by other investigators; the ease of separation appears to depend on
the characteristics of the absorbent, of the eluting solvent, and of
the sample extract, though there does not appear to be any difficulty
in separating all interfering substances, except DDE, a metabolite of
DDT. Thin-layer chromatography has been used for separation by Norén &
Westöö (1968), Bagley et al. (1970), and Reinke et al. (1973).

In many environmental samples, DDE is present in larger amounts than
the PCBs, and must be removed before their quantitative determination.
Oxidation procedures have been used to convert DDE to dichlorobenzo-
phenone; recommended oxidants are potassium dichromate and sulfuric
acid (Westöö & Norén, 1970b) and chromium (II)oxide and acetic acid
(Mulhern et al., 1971). Jensen & Sundström (1974a), who were
interested in determining DDT/PCB ratios in environmental samples,
preferred sodium dichromate in acetic acid with a trace of sulfuric
acid. They claimed that this does not destroy DDT and its metabolite
DDD, which may be present in extracts after clean-up with strong
sulfuric acid, and that using this mixture makes possible the
quantitative determination of the dichlorobenzophenone from the
oxidation of DDE.

Conversion of DDT to DDE can be achieved by treatment with ethanolic
potassium hydroxide, which also removes interference from elemental
sulfur (Ahling & Jensen, 1970). Sulfur may also be removed by
activated Raney nickel (Ahnoff & Josefsson, 1975) or by metallic
mercury.

Beck & Mathar (1985) used gel permeation chromatography to clean
extracts of food of animal origin.

2.3.2 Separation and identification

2.3.2.1 Chromatographic separation

Numerous gas chromatographic studies using packed or capillary columns
have confirmed the complexity of all commercial PCB formulations. The
accuracy in determining PCB levels is highly variable and matrix
dependent. Many factors including: the water solubility, volatility,
and biodegradability of individual PCBs, will alter the composition of
a commercial PCB preparation introduced as a pollutant into the
environment. Thus, the composition of PCB extracts from environmental
matrices will vary widely and often do not resemble any commercial
mixture. Quantitative analyses on these mixtures is usually determined
by pattern- or peak-matching methods, using artificially reconstituted
mixtures of different commercial formulations. High-resolution, glass
capillary gas chromatographic analysis can provide a solution.
Capillary gas chromatography columns, currently in use, are made of
fused silica, chemically bonded with various stationary phases, to
achieve a range of different selectivities towards complex samples. In
general, packed columns have been replaced by capillary columns,
because of their far superior efficiency. The identities of the
individual peaks must then be determined by using synthetic standards
and by retention index addition methods. This latter technique
predicts the relative retention times (RRT) of specific PCBs and has
been used to assign the structures of individual PCB congeners. The
method relies on the RRT values that have been determined for
synthetic PCB standards. On this basis, Safe et al. (1985a) reported
the first congener-specific analysis of a PCB preparation and PCBs in
human milk.

Some workers use GC with mass selective detection (MSD), which
quantifies the level of chlorination in a sample extract
(Alford-Stevens, 1986). Tanabe et al. (1987) and Kannan et al. (1987)
described a method to determine the 3 toxic, non- ortho-chlorine-
substituted, coplanar PCBs, 3,4,3',4'-tetra, 3,4,5,3',4'-penta-,
and 3,4,5,3',4',5'-hexachlorobiphenyl, which are biologically active
congeners. The method comprised alkali digestion, carbon
chromatography, and high-resolution gas-chromatography. Using
this method, it is possible to determine ppt levels of these toxic
residues in biological samples. Duinker et al. (1988) used
multidimensional gas chromatography with ECD to determine levels of
all congeners in some Clophen and Aroclor mixtures and found
considerable differences between their composition of congeners and
those in an extract of a seal blubber sample. Using this technique,
congeners were identified that had, hitherto, been undetected, using
other analytical techniques. It was possible to identify the toxic
congeners in the samples studied, even when the relative contribution
of each congener to the cluster was as low as 0.01%.

2.3.2.2 Gas-liquid chromatography

Most analysts use gas-liquid chromatography with an electron-capture
detector for the separation of PCBs from the extract after clean-up.
Stationary phases commonly used are silicones or their derivatives,
for example, DC 200, SF 96, OV 1, and QF 1, or Apiezon L. Jensen &
Sundström (1974a) stated that, with a mixture of SF 96 and QF 1, 14
peaks could be obtained from Clophen A50, but that Apiezon L gave much
better resolution. They obtained better peak separation by prior
fractionation on a charcoal column, which separated the PCBs according
to the number of o-chlorine substituents; they regarded such
refinements as unnecessary in PCB residue analysis, but they may be of
value in the study of the selective, environmental degradation of
PCBs. Column temperatures used ranged between 170°C and 230°C. Glass
capillary columns are superior to packed columns giving better
separation of closely-related congeners; they also give good
separation of PCBs from DDT and its metabolites (Zell et al., 1977;
Dunn et al., 1984; Beck & Mathar, 1985; Alford-Stevens, 1986; Tanabe
et al., 1987; Duinker et al., 1988).

A gas chromatography/electron impact mass spectrometry (GC/EIMS)
method was used by Erickson et al. (1988) for the determination of
by-product (non-Aroclor) PCBs. In this method, the recovery of 4
¹³C-labelled PCBs was measured to assure adequate recovery of the
native PCBs from diverse matrices. The complexity of the matrices and
the high probability of chlorinated organic interferents precluded the
use of GC/ECD. The best available technique for universal application
to commercial products, and associated waste, is GC/EIMS. During the
validation work, the anticipated difficulty of qualitative and
quantitative data interpretation was confirmed. In addition to the
inherent problems resulting from extrapolation from 11 standards to
209 analytes, interpretation of the complex peak clusters is tedious.

2.3.3 Quantification

An electron-capture detector (ECD) is the most commonly used
instrument for the quantification of PCBs. However, the response of
this detector varies according to the number and location of the
chlorine atoms in the PCB molecule, resulting in difficulties when the
sample under investigation contains PCBs that have degraded (Zitko et
al., 1971).

Various principles have been used to quantify PCB residues:

* comparison of a single peak in the residue with the corresponding
peak in a commercial reference PCB (Aroclor, Clophen);

* comparison of the total response for several peaks in the residue
with the total response of the corresponding peaks in a reference
standard;

* comparison of the response of all peaks in the sample with those
in the reference standard;

* perchlorination of PCBs to decachlorobiphenyl followed by
quantification of this single compound.

The results obtained using these various methods differ; consequently,
the precision in these analyses is not very good. Recently, Dunn et
al. (1984) described a method for the quantification of PCBs using gas
chromatography data, based on a pattern recognition technique and
partial least squares in latent variables. The data to which it was
applied were gas chromatograms of Aroclor 1242, 1248, 1252, and 1260.
This technique also allows the classification of unknown samples
(WHO/EURO, 1987).

Fait et al. (1989) investigated whether the results obtained for total
PCBs using FSCGC/ECD (see section 2.3), differed significantly from
those determined using packed column gas chromatography electron
capture (PCGC/ECD) techniques, within 3 exposure groups. The
concentrations of individual PCBs were determined in both the serum
and adipose tissue from 35 transformer repair workers and 17 previous
repair workers, exposed mainly to Aroclor 1260, in comparison with 56
non-exposed workers. Eighty-nine PCB peaks were identified. The total
serum PCBs determined by FSCGC/ECD greatly exceeded that from standard
PCGC/ECD. The median concentrations in serum were: 43.7, 30.0, and
16.1 µg/litre, and the median concentrations in adipose tissue were:
3180, 888, and 821 µg/kg, respectively. In all workers,
hexachlorinated and heptachlorinated congeners predominated followed
by octachlorinated and pentachlorinated species. The 7 major peaks in
serum and adipose tissue were 2,3,5,6,3',4',5'/ 2,3,4,5,2',4',5'/
2,3,4,5,2',3',4'-heptachloro-; 2,3,4,2',3',5'-hexachloro-;
2,4,6,3',4',5'/ 2,4,5,2',4',5'-hexachloro-; 2,3,4,5,2',3',5',6'/
2,3,4,5,6,2',3',5'-octachloro-; 2,4,5,3',4'/ 3,4,5,2',3'-pentachloro-
and 2,3,4,2',3',4'/ 2,3,5,6,2',4',5'/ 2,3,4,5,2',4',6'
multichlorobiphenyls.

The response of the electron capture detector is not equal for all PCB
components, being much affected by the degree of chlorination, as
already mentioned (Zitko et al., 1971). This does not lead to
difficulties when the sample under investigation has been directly
contaminated by a commercial PCB mixture, as this mixture can be used
as a standard. Difficulties are encountered when the PCBs in the
sample have undergone selective environmental degradation. Several
investigators have noted that the pattern of peaks from such samples
resembles fairly closely that of one or other of the higher
chlorinated PCB mixtures, such as Aroclor 1254, and they have compared
the total area of the peaks with that of the nearest commercial

product, in order to determine the amount of PCBs in the sample
(Armour & Burke, 1970; Tuinstra, 1983). Collins et al. (1972) observed
that, under their conditions, the area of peaks usually encountered in
extracts of tissue samples was very similar to that of an equivalent
amount of DDE, thus, DDE could be used for calibration. In order to
overcome the uncertainties of these procedures, Rote & Murphy (1971)
divided the peaks into groups according to the number of chlorine
atoms in the molecule, as determined from mass spectrographic data,
and calculated the PCB content of each group from the theoretical
response of the detector to chlorine content. Jensen et al. (1973)
selected a commercial PCB that included all the peaks from the
extract; they determined the PCB content of each peak by combined mass
spectrometry and coulometry, and determined the total PCBs in the
sample by comparing the height of each peak obtained with the extract
with those obtained with the reference sample. Simpler methods have
been used including that of Koeman et al. (1969), who compared the
height of a single peak, obtained with the extract, with that of a
peak with the same retention time obtained with a commercial PCB
mixture, and those of others who averaged out more than one peak for
this calculation (Reynolds, 1971; Reinke et al., 1973). Rote & Murphy
(1971) calculated that such procedures may give more than double the
values obtained by a more accurate method.

In the characterization of PCB components in PCB mixtures, the
retention properties of the components of the mixtures, as well as a
great number of synthesized components, were used to predict a
complete analysis of mixtures as Aroclors 1242, 1254, and 1260. Jensen
& Sundström (1974a) synthesized a large number of reference substances
and were able to identify almost 60 components in Clophen A50 and A60.

Attempting to account for unidentified peaks, authors have used the
chromatographic retention indices of available components to calculate
such data for missing ones. The identity of many peaks could not,
however, be determined unambiguously. Some of these uncertainties have
been resolved by the application of techniques other than the
comparison of retention times e.g., MS, NMR, and IR. The efficiency of
packed columns in GLC is not sufficient to allow their use for the
accurate analysis of complex mixtures, in most cases. Another approach
to the use of packed columns involves the use of columns with various
selectivities. In this way, complete analysis of all components in
Aroclors has been claimed with the use of up to 12 columns. The
strongly increased GLC separation offered by capillary columns has
been used to advantage in the analysis of technical formulations, in
some cases the eluate was analysed by MS. To identify individual
congeners, gas-liquid (using glass capillaries with different
coatings) chromatography (GLC) was used by Albro & Parker (1979) and
Albro et al. (1981). Hydrogen flame ionization detection (HFID) and
electron capture detection (ECD) and MS were used by Duinker &
Hillebrand (1983).

2.3.4 Accuracy of PCB determinations

A group of 8 analysts, engaged in an investigation of pollution in the
North Sea, undertook a collaborative study to determine the PCB
content of a sample of fish oil, using the methods currently employed
in their laboratories (International Council for the Exploration of
the Sea, 1974). The PCB values obtained ranged from 1.0 to 3.9 mg/kg
with a mean of 1.97 mg/kg and a standard deviation of 0.93 mg/kg.
Better agreement was obtained with the same fish oil fortified with
PCBs at a concentration of 10 mg/kg; the mean of the results for the
fortified sample was 10.0 mg/kg with a standard deviation of
1.1 mg/kg.

A probable source of error is incomplete initial extraction of PCBs
from a sample (Holden & Marsden, 1969). Another source of variation
between laboratories lies in the method used to quantify gas-liquid
chromatographic peaks; Van Hove Holdrinet (1975) considered this to be
the major source of error.

It is evident that caution should be exercised in accepting the
analytical results from a laboratory, particularly for samples with a
low PCB content, until the competence of the laboratory has been
established by an inter-laboratory collaborative study (Tuinstra,
1983).

Schulte & Malisch (1984) described a method to determine the real PCB
contents of environmental samples. A technical PCB mixture of known
composition was used for calibration. The PCB concentrations were
determined in samples of human milk and butter and the calculated
contents were 50% and 40% lower, respectively, than the values
obtained by the usual calculation based on evaluation of some higher
peaks of technical PCB mixtures.

2.3.5 Confirmation

Since Jensen first identified as PCBs hitherto unknown substances that
interfered in the glass-liquid chromatographic determination of
organochlorine pesticides using mass spectrographic data, other
investigators have confirmed the presence of PCBs in environmental
samples by combining gas-liquid chromatography with mass spectrometry
(Bagley et al., 1970) and with coulometry, to measure the chlorine
content. The conversion of PCBs to bicyclohexyl and decachlorobiphenyl
is further confirmation (Berg et al., 1972). The widespread
distribution of PCBs is now well established, and, as adequate methods
are available to remove interference from organochlorine pesticides,

there is no evidence of the presence of other interfering substances
in the types of sample that have so far been analysed, down to a limit
of detection of around 0.01 mg/kg. This does not necessarily apply to
other types of sample, particularly when very low levels are being
sought; Ahnoff & Josefsson (1973) reported a number of unknown
interfering substances, when measuring PCBs in water at levels below
1 ng/litre. One of these substances was subsequently identified as
elemental sulfur. They recommend confirmation by mass fragmentography
for such samples.

2.3.6 Detection limits

The limits of determination using low or high resolution mass
spectrometry are 0.01-1 pg per injection of each congener. The
detection levels in samples depend on the sample size and matrix.
Using an air sampling device described by Rappe et al. (1985b), a
detection level of 0.05 pg/m³ per congener could be determined in
ambient air (WHO/EURO, 1987).

In general, other substances are not considered to interfere at levels
of about 0.01 mg/kg. In river water and air, levels of 1 ng/litre and
0.3 ng/m³, respectively, are reported to be the detection limits of
PCBs (WHO/EURO, 1987). Tuinstra (1983) found a limit of detection for
individual chlorobiphenyls in environmental and biological samples, of
less than 1 µg/kg (see Table 7).

The results for sewage sludge, eel, grass, cow's milk, and human fat
are given in Table 7 (Tuinstra, 1983). Individual chlorobiphenyls were
also estimated in the monitoring programme for environmental and
biological samples in the Netherlands.

2.4 Codex questionnaire on analytical methods

2.4.1 Interpretation and comparability of data

Monitoring data are available from many sources in many countries.
They have been obtained using various methodologies, such as different
sampling techniques and different methods of analysis and
quantification. Limits of determination reported vary by a factor of
1000 or more.

Given this situation, data on levels of PCBs have to be interpreted
with the greatest care. Comparisons can only be made between data from
the same laboratory, using the same validated technique over a long
period. Comparisons between data from different laboratories have to
be limited to the very few cases, where very strict inter-laboratory
checks have been made on the basis of the same sampling and analytical
techniques. Indications about trends can only be obtained when taking
into account these basic considerations (Beck & Mathar, 1985; Tuinstra
et al., 1985b,c).

In June 1985, a questionnaire was distributed to all Codex Contact
Points with the aim of providing background information on PCBs for
the ad hoc working group on contaminants to compare such factors as
methods of analysis, quantification, monitoring, etc. Eighteen out of
22 countries responded to the questionnaire.

In some cases, the information given was incomplete, but it is
apparent that a variety of clean-up methods is employed. Where good
laboratory practices are followed and tests indicate close to 100%
recovery of standards from spiked samples, the main effect of
different clean-up procedures will be on the limit of detection.

For gas chromatography, 6 countries reported that they used capillary
columns as alternative or confirmatory systems. Among the respondents,
the Netherlands and the Federal Republic of Germany routinely used
capillary columns and specific PCB isomers as regulatory standards.
The types of packed column materials used varied considerably. With
respect to quantification, pattern comparison with standards of
various PCB formulations was the method most favoured, though some
countries specified the use of certain combinations of peaks. In
several cases, the methods being used were stated to have been
collaboratively tested, or checked by inter-laboratory ring tests.

During the sixties, packed column chromatography was the most widely
used method in the determination of PCBs. Results obtained with this
technique varied widely between laboratories, and were much influenced
by the method of quantification chosen and by the PCB mixture used as
a standard. Chemical conversion methods, especially perchlorination,
have also been used. These methods are quite sensitive, but do not
allow for peak pattern identification. Another drawback of
perchlorination is that conversion of less chlorinated biphenyls is
not quantitative.

Sensitivity is sufficient, if adequate clean-up methods are used.
Combined gas chromatography/mass spectrometry has a somewhat lower
sensitivity, needs more expensive equipment, and is not considered
suitable for routine work. The results obtained using these techniques
may vary widely and most of them can only be used as rough estimates.

When capillary columns are used with temperature programming, almost
all PCB isomers and congeners normally present in samples can be
identified. This method is now considered to be the best available
technique. However, it is important to decide which isomers should be
used as guiding substances.

Table 7. Typical values of individual chlorobiphenyls in Dutch environmental and
biological samples. Peak numbering according to IUPAC rules^a

PCB Structure Sewage Eel Grass Cow's Human
compound sludge milk fat
µg/kg µg/kg µg/kg µg/kg µg/kg
(dm)^b product (dm)^b,c fat^c fat

28^d 2,4,4' 60 35 -^c -^c 45
52^d 2,5,2'5' 22 110 0.4 2.1 10
44 2,3,2'5' 20 34 0.2 0.9 10
95 2,3,6,2'5' 58 130 0.7 1.6 30
101^d 2,4,5,2'5' 30 85 0.6 3.1 15
151 2,3,5,6,2'5' 9 24 0.2 0.6 10
149 2,3,6,2'4'5' 42 90 0.6 2.5 15
118 2,4,5,3'4' 20 110 0.3 -^c 80
153^d 2,4,5,2'4'5' 54 180 0.7 13 295
141 2,3,4,5,2'5' 10 40 0.2 0.6 <5
138^d 2,3,4,2'4'5' 45 200 0.7 11 235
128 2,3,4,2'3'4' 7 20 <0.1 1.2 15
180^d 2,3,4,5,2'4'5' 33 80 0.5 6.4 205
170 2,3,4,5,2'3'4' 10 30 0.2 1.8 90
201 2,3,4,5,2'3'5'6' <5 10 <0.1 <0.5 20

^a From: Tuinstra (1983).
^b dm = dry matter.
^c nd = not determined.
^d Monitoring compounds.

2.5 Activities of the WHO Regional Office for Europe

The WHO Regional Office for Europe (WHO/EURO) has an ongoing programme
related to PCBs, as well as to other chlorinated hydrocarbons,
including polychlorinated- para-dibenzodioxins (PCDDs) and
polychlorinated dibenzofurans (PCDFs). Within this programme,
practical guidelines to prevent and control accidental and
environmental exposures to these chemicals have been published in the
Environmental Health Series of WHO/EURO (1987). The other important
project within this programme dealt with the assessment of the health
risks to infants associated with contamination of mother's milk. This
assessment was completed by a WHO/EURO Expert Consultation held in
Abano Terme, Italy, in 1987, and the output of this consultation has

been published in the Environmental Health Series of WHO/EURO (1988).
In order to produce more data on exposure levels through human milk,
WHO/EURO has been coordinating analytical field studies in which
several countries have participated. The results of these studies have
been published in the Environmental Health Series of WHO/EURO (1989).
This document also includes the results of interlaboratory quality
control studies on levels of PCBs, PCDFs, and PCDDs in human milk. In
the first series of studies, 12 laboratories were involved. The second
round of the quality control studies has been completed, with the
participation of additional laboratories, and the results will be
published. Furthermore, the repetition of the analytical field studies
on the levels of PCBs, PCDFs, and PCDDs in human milk will be
implemented in 1991 and coordinated at WHO/EURO.

2.6 Appraisal

Since the congener composition and relative concentrations of the
individual components in PCB extracts from environmental and
biological samples differ markedly from those in commercial PCB
mixtures, the quantitative determination of the PCB contents of such
samples presents a special problem. Various approaches to the
quantitative determination of PCBs have been reported including:
attempts to determine the total PCB concentration through
perchlorination of the mixture; identification of selected
chromatographic peaks through gas chromatographic techniques with
packed columns using certain commercial products as standards; as well
as attempts to carry out congener-specific analysis, based on high
resolution chromatographic separation followed by identification and
quantification by mass spectrometry using synthetic standards. This
last method is considered the best at present, though it is not
feasible for all laboratories. Although the concentration values
obtained from the various methods might be similar, such comparison
will be limited and is of questionable value for most purposes. The
occurrence of specific PCB congeners in various samples and a
consideration of the relative toxicity and persistence of the
congeners have been suggested as a basis for a congener-specific
analytical approach. While this approach can be useful, particularly
in risk/hazard assessment exercises, it must be realized that it is
based on the present knowledge about the occurrence, persistence, and
toxicity of specific congeners. It does not take into consideration
potentially unrecognized toxicities associated with the same or
different congeners, which may be present in a sample, also it is not
feasible in some countries. Therefore, further research in this area
should continue to improve the basis for monitoring programmes and for
a congener-specific approach.

In the selection of areas with high levels of contamination, in order
to establish priorities for action, it is considered that analytical
competence and the use of adequate controls and standards is more
important than highly accurate laboratory analysis. Also, the quality
and usefulness of analytical data depend critically on the validity of
the samples and the adequacy of the sampling programme. A quality
assurance programme and collaborative studies should be part of any
long-term study on PCBs, since there are several possible sources of
error. In this situation, data on levels of PCBs have to be
interpreted with the greatest care and, in general, definitive
comparison can only be made between data from laboratories using the
same techniques and interpretation of results.

3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE

3.1 Natural occurrence

Polychlorinated biphenyls are aromatic chemicals that do not occur
naturally in the environment.

3.2 Man-made sources

3.2.1 Production levels and processes, uses

The first chlorinated biphenyl was synthesized in 1864, but it was not
until 1929/1930 that the PCBs were produced commercially for use:

(a) as dielectrics in transformers and large capacitors;

(b) in heat transfer and hydraulic systems;

(d) as plasticizers in paints, and as ink solvent/carriers in
carbonless copy paper, adhesives, sealants, flame retardants, and
plastics (Hutzinger et al., 1974; Pomerantz et al., 1978).

An extensive review of the uses of PCBs is given in DFG (1988).

3.2.1.1 World production figures

Over one million tonnes of PCBs have been produced commercially under
a number of trade names, such as Aroclor, Fenchlor, Clophen, and
Kanechlor.

Details of the production and uses of PCBs in the USA have been
released, and have been summarized by Nisbet & Sarofim (1972). Annual
production increased steadily from 1930 and reached a maximum in 1970
of 33 000 tonnes. Of this, 56% was used as a dielectric (36% in
capacitors and 20% in transformers). Various plasticizer outlets
accounted for 30%, hydraulic fluids and lubricants, 12%, and heat
transfer liquids, 1.5%. During this peak year, 65% of the production
was of the 42% chlorinated type, 25% was less chlorinated, and the
remainder more chlorinated. After 1970, production decreased sharply
owing to the voluntary limitation of sales by the Monsanto Company,
the major manufacturer in the USA.

Following the restriction of sales for dissipative uses, the
percentage of PCBs sold as dielectrics rose to 77% in 1971 and the
proportion of highly chlorinated products was considerably reduced;
Aroclor 1016 replaced Aroclor 1242. In Japan, 44 800 tonnes of PCBs
were used from 1962 to 1971; of this, 65.4% was used in the electrical
industry, 11.3% in heat exchangers, 17.9% in carbonless copying paper,
and 5.4% for other dissipative uses (Ishi, 1972).

During the period 1980-84, the production in EEC member states was as
follows: France, 16 200; Federal Republic of Germany, 24 200; Italy,
4500; and Spain, 3400 tonnes. After 1984, production was continued
only in France and Spain (Bletchly, 1985; WHO/EURO, 1987).

By the end of 1980, the total amount of PCBs produced was 1 054 800
tonnes (of which approximately half was used in transformers and
capacitors, see Table 8), divided between the following countries (in
tonnes): USA, 647 700; Federal Republic of Germany, 130 800; France,
101 600; United Kingdom, 66 800; Japan, 59 300; Spain, 25 100; and
Italy, 23 500 (Bletchly, 1983).

In addition, Czechoslovakia and the USSR have manufactured PCBs for
their domestic market under the trade names of Delor and Sovol,
respectively, but the data on production quantities are not available.

According to an OECD report, transformers and capacitors provided the
major outlets for PCBs in most OECD countries in 1971. In 1972,
several countries restricted sales; in Sweden the importation and use
of PCBs were restricted by law; in the United Kingdom, as in the USA,
sales were voluntarily restricted to the lower chlorinated PCBs for
use as dielectrics in enclosed systems, and, in the USA in 1979,
manufacture, use, handling, storage, and disposal were promulgated. As
late as 1985, a final rule concerning the restriction and conditions
on the use of PCB transformers was published (USEPA, 1985). In Japan,
the production and use of PCBs were banned in 1972.

The 24 OECD countries adopted a Decision in 1973, limiting the use of
PCBs to certain specific applications and asking for the control of
the manufacture, import, and export of bulk PCBs, for adequate waste
treatment and for a special labelling system for PCBs and
PCB-containing products. On 13 February 1987, the Council of the
Organization for Economic Co-operation and Development (OECD) adopted
a further Decision-Recommendation (C(87)2(final)) on "Further measures
for the protection of the environment by control of polychlorinated

Table 8. Estimated usage of PCBs in transformers and large capacitors in
a number of OECD countries in 1930-80 (in tonnes)^a

Country Usage in Usage in Total
transformers capacitors

France 50 700 8 800 59 500
Federal Republic 44 400 17 700 62 100
of Germany
Italy 10 400 1 500 11 900
Japan 37 200^b 37 200
Spain 20 100 3 400 23 500
United Kingdom 5 800 8 100 13 900
United States 125 800 130 400 256 200
of America

Total 294 400 169 900 464 300

^a From: WHO/EURO (1987).
^b Includes the usage in both transformers and capacitors

biphenyls". With this Decision-Recommendation, the OECD Member
countries committed themselves to ban virtually all new uses of PCBs,
accelerate the phasing out of PCBs from existing uses, control PCBs in
contaminated products, articles, or equipment, and ensure appropriate
disposal methods for PCB-containing waste. The uses of PCBs have been
virtually restricted to those in "closed systems". In 1976, an EEG
Directive made the limitations of the use compulsory for the EEG
Member States. Other Directives, such as those on waste treatment and
disposal, followed (van der Kolk, 1984a, Personal communication).

3.2.1.2 Manufacturing processes

Industrial manufacturing of PCBs is based on the chlorination of
biphenyl by anhydrous chlorine, under heated reaction conditions and
in the presence of suitable catalysts (e.g., iron-chloride). Depending
on the reaction conditions, a degree of chlorination varying between
21% and 68% (weight percentage, w/w) can be achieved.

The yield is always a mixture of different compounds and congeners.
Commercial mixtures generally have been purified by filtration and
fractional distillation, but, in spite of this, they have been found
to contain many impurities (WHO/EURO, 1987). In general, commercial
PCB products contain impurities, mainly polychlorinated dibenzofurans
(PCDFs).

Rappe et al. (1985d) cf. WHO/EURO (1987) analysed a series of
commercial PCBs, using a new clean-up technique based on reverse-phase
chromatography on a carbon column followed by a fluorosil column. In
all PCB products, PCDFs were found at levels varying from a few mg/kg
up to 40 mg/kg. The chlorination pattern of the PCDFs was found to
vary with the chlorination level of the PCBs. In most products,
2,3,7,8-substituted tetra-, penta-, and hexa-CDFs were the major
constituents.

3.2.2 Uses

PCBs have been widely used in electrical equipment, such as capacitors
and transformers. These have often been considered to be closed
systems, though small amounts of PCBs can frequently be found on the
outer metal surface of such equipment.

Smaller volumes of PCBs have often been used as fire-resistant liquid
in nominally closed systems, such as hydraulic and heat exchange
systems (WHO/EURO, 1988).

Broadhurst (1972) reviewed the many technical applications of PCBs
that appear in the literature and in patent specifications, and
indicate the possibility of a widespread, non-occupational, low-level
exposure to PCBs, other than that derived from the diet. PCBs are used
in the home in ballast capacitors for fluorescent lighting, and
exposure from pressure-sensitive copying paper has not been limited to
office workers. The valuable properties of PCBs as plasticizers has
led to their use in furnishings, interior decoration, and building
construction; examples are surface treatment for textiles, adhesive
for waterproof wall coatings, paints, and sealant putties. PCBs have
been used as plasticizers for plastic materials and in the formulation
of printing inks.

The value of PCBs for industrial applications depends on their
chemical inertness, resistance to heat, non-flammability, low vapour
pressure (particularly with the higher chlorinated compounds), and
high dielectric constant.

Data on the usage of technical PCB mixtures in Europe are scarce. In
the 1960s and early 1970s, PCBs were used in (WHO/EURO, 1987):

(a) completely closed systems;
(b) nominally closed systems;
(c) open-ended applications.

3.2.2.1 Completely closed systems

PCBs have been widely used in electrical equipment, such as capacitors
and transformers, which are considered to be completely closed
systems. Historically, capacitors are the single largest PCB-use
category. The PCB mixtures used for this purpose are, for example,
Pyralene 3010, Aroclor 1016, 1221, and, earlier, also Aroclor 1242 and
1254. The amounts used in a number of OECD countries are presented in
Table 8 (OECD, 1982; Bletchly, 1983; Callahan et al., 1983).

Since the late 1970s and the beginning of the 1980s, PCB-filled
capacitors have largely been superseded by capacitors with a non-PCB
dielectric fluid. The tendency for this substitution varies from
country to country, for example, it started in Sweden and Finland in
1982, and in Norway in 1985.

The technical PCB mixtures used in transformers are mostly highly
chlorinated like Aroclor 1254 and 1260. In general, the PCBs are used
in combination with tri- and tetrachlorobenzenes as mixtures called
Askarel.

The amounts of PCBs used in transformers differ in different
countries. In France, where most transformers are placed indoors, the
major dielectric fluid is PCBs or Askarels, which are both flame
retardants, while in Scandinavia, where most capacitors are placed
outdoors, mineral oils (with a lower melting point) are frequently
used.

During the 1980s, there has been a marked interest in replacing the
PCBs, mainly in indoor transformers, as a result of serious accidents,
for example, in Binghamton, San Francisco, Miami in the USA, and Reims
in France. Various products are used for this exchange, such as
mineral oils, silicone oils, perchloroethylene, and other chlorinated
products (WHO/EURO, 1987).

3.2.2.2 Nominally closed systems

Smaller volumes of PCBs have frequently been used as fire-resistant
liquid in nominally closed systems, such as hydraulic and heat
transfer exchange systems (for example, trade names Pydraul and
Therminol FR, containing Aroclor 1242, 1248, 1254, and 1260). PCBs are
used as a working fluid in vacuum pumps (Aroclor 1248, 1254), which
can also be considered as nominally closed systems (WHO/EURO, 1987).

3.2.2.3 Open-ended applications

With open-ended applications of PCB, both the emissions into the
environment and the levels of occupational exposure are more
pronounced. The major open-ended applications include use as a
plasticizer (in PVC, neoprene, and other artificial chlorinated
rubbers). Other open-ended uses, such as surface coatings, paints,
inks, adhesives, pesticide extenders, microencapsulation of dyes, and
carbonless copy paper contribute smaller volumes into the environment.
PCBs have also been used in immersion oils for microscopes, as
catalysts in the chemical industry, in casting waxes in the iron/steel
industry (decachlorobiphenyl), and in cutting and lubricating oils
(WHO/EURO, 1987).

3.2.2.4 Contamination of other compounds

In addition to the above uses of PCBs, numerous halogenated compounds
may contain PCBs in small amounts as a contaminant (US EPA, 1983).

3.2.3 Loss into the environment

PCBs are dispersed into the environment through atmospheric transport
and, on a more regional scale, following release into water. PCBs are
also mobilized in the soil or landfills, but the rates of dispersion
and subsequent transfer to biota and humans are difficult to estimate.

More highly chlorinated forms become most prevalent in compartments
further along the pathway chains. The analytical methods used to
quantify PCBs in the environment and biota vary greatly within, and
between, countries. Thus, comparisons can only be made in a very broad
sense and could, to some extent, be erroneous (WHO/EURO, 1988).

An overview of prevention and control measures of accidental and
environmental exposures is given in WHO/EURO (1987).

3.2.3.1 Routes of environmental pollution

Surveys of the sources of environmental pollution with PCBs were made
before production and use became limited, and the information
available may not now apply in North America and elsewhere. Only 20%
of the annual production in the USA can be regarded as a net increase
in current usage, and the remainder is balanced by a loss to the
environment. More than one-half of this entered dumps and landfills
and it has been calculated that 0.3 million tonnes of PCBs have
accumulated in such locations in North America, since 1930 (Nisbet &
Sarofim, 1972). Much of this was originally enclosed in containers,
such as capacitors, or was in plasticized resins and will not be
released until the containing medium decays. The diffusion of PCBs
from landfills is likely to be slow, on account of their low
volatility and low water solubility. Carnes et al. (1973) found little
leaching from the one site that they tested.

The concentration of PCBs in emissions from several municipal sanitary
landfills and refuse and sewage sludge incinerators were determined in
the Midwest of the USA. Sanitary landfills continuously emit the
gaseous products of anaerobic fermentation together with other
volatile materials into the atmosphere. A projection, based on the
amount of methane generated annually from landfills and a PCB to
methane ratio of 0.3 µg PCBs/m³ of methane found from the landfills
sampled, indicates that the annual PCB emissions from sanitary
landfills in the USA are of the order of 10-100 kg/year. The
concentrations of PCBs from the incinerator stacks ranged from
0.3-3 µg/m³ and the annual emissions per stack were 0.25 kg/year.
These estimates are very small in comparison with the 900 000 kg
PCBs/year estimated to cycle through the atmosphere over the USA,
annually (Murphy et al., 1985).

Scrap transformer fluid containing PCBs has been used in the USA in
amounts of about 10 tonnes/year in pesticide formulations (Panel on
Hazardous Trace Substances, 1972, cf. WHO/EURO, 1988), and this
unauthorized use has led to the local contamination of milk supplies.

Pressure sensitive duplicating paper (carbonless copying paper)
containing PCBs has found its way into waste paper supplies and has
been recycled into paper and board used as food packaging materials,
but not since 1970; paints for coating the bottom of ships contained
3-5% of PCBs, about 3% of the annual quantity imported into Sweden has
been used for this purpose, and this has been a source of plankton
contamination (Jensen et al., 1972a).

Schecter (1987) described the contamination of drinking-water by the
use of submersible water pumps which, in certain instances, contained
PCBs in the oil. When the pumps leak, PCBs may be released into the
drinking-water.

In addition, the US EPA, in 1980, estimated that over 1 000 000 wells
in the USA may have PCB capacitors in the well motors. Levels recorded
in drinking-water range from 0.26 to 57 µg/litre compared with
1 µg/litre considered safe in the guidelines for New York State. The
oil from these pumps contained 630 000-24 000 000 µg/kg of PCBs.

Stehr et al. (1985) studied the possibility of contamination with PCBs
of oils and oil-filled devices used by amateur radio operators. Two of
77 oil samples contained more than 50 mg/kg.

3.2.3.2 Release of PCBs into the atmosphere

There appears to be little atmospheric contamination during the
manufacture and processing of PCBs, but this can occur during their
subsequent use and disposal. Although PCBs have a low volatility,
there may be an appreciable loss to the atmosphere during the lifetime
of a PCB-plasticized resin, particularly of the lower chlorinated
products. Further pollution may occur during the incineration of
industrial and municipal waste. Most municipal incinerators are not
very effective in destroying PCBs; efficient incinerators can be
designed for this purpose (Oehme et al., 1987), though the higher
chlorinated PCBs are more resistant to pyrolysis. Secondary sources of
atmospheric pollution are volatilization from soil, and the drying of
sewage sludge. Furthermore, there is evidence that, even at ambient
temperatures, PCBs will enter the atmosphere by volatilization from
soils and water bodies, landfill sites etc. (section 4.1.1).

3.2.3.3 Leakage and disposal of PCBs in industry

Eschenroeder et al. (1986) analysed PCB risks using estimates of human
intake of PCBs originating from accidental spills from electrical
equipment. Equipment spills without controls resulted in a human
intake of PCBs of, at the most, 2 ng/day via the water exposure
pathway. This was negligible in comparison with the intakes calculated
on the basis of fish consumption. The inhalation exposure of
approximately 100 persons living in the vicinity of a spill in
Southern California was determined to equal the PCB intakes of a
fish-eating population.

3.2.4 Thermal decomposition of PCBs

It has been found by Buser et al. (1978a,b) that PCBs can be converted
to PCDFs under pyrolytic conditions. The pyrolysis of a commercial PCB
mixture in a sealed quartz ampoule, in the presence of air, yielded a
mixture including about 30 major and more than 30 minor PCDF
congeners.

Buser & Rappe (1979) studied the pyrolysis (at 600°C) of 15 individual
PCB isomers and demonstrated the presence of PCDFs via intramolecular
cyclizations, where m + n varies from 4 to 8 (Fig. 1). The
thermochemical generation of PCDFs from PCBs was found to follow 4
general reaction routes including loss of ortho-Cl; loss of HCl
involving a 2,3-chlorine shift at the benzene nucleus; loss of
ortho-HCl and loss of ortho-H (Buser, 1985; Hutzinger et al.,
1985).

The maximum yield of PCDFs was about 10%, calculated on the amount of
PCBs decomposed, and the optimal temperature was between 550 and
650/700°C (Bentley, 1983). Thus, the uncontrolled burning of PCBs can
be an important occupational and environmental source of toxic and
hazardous PCDFs and it is recommended that all destruction of
PCB-contaminated waste should be carefully controlled, especially with
regard to the burning temperature (above 1000°C), residence time, and
turbulence (Bentley, 1983; WHO/EURO, 1987).

In the temperature range 300-400°C, Morita et al. (1978) reported that
the yield of conversion seemed to be in the mg/kg range. However,
Nagayama et al. (1981) reported a dramatic increase in the levels of
PCDFs at these rather low temperatures, in the presence of stainless
steel or nickel.

No, or very low levels of, PCDDs have been reported from the pyrolysis
of PCBs. However, pyrolysis of a mixture of PCBs and chlorobenzenes
(product Askarel) can yield both PCDFs and PCDDs (Buser, 1979).

Rappe et al. (1985b) found that various types of industrial
incinerators, such as copper smelters and steel mills generate PCDFs
and PCDDs. Pyrolysis of chlorinated polymers like polyvinylchloride
(PVC) and Saran also generate these compounds and exhaust gases of
motor cars and their motor oil may contain PCDDs and PCDFs (WHO/EURO,
1987).

In a State Office Building in the centre of Binghamton, New York, a
fire, in conjunction with several explosions, occurred in the basement
mechanical room, in 1981. Approximately 750 litres of Askarel, a
dielectric fluid composed of 65% PCBs (Aroclor 1254) and 35%
polychlorinated benzenes, leaked from a transformer and caught fire.
Pyrolysis of the Askarel led to the formation of a fine oily soot
that spread throughout the building via 2 ventilation shafts. Samples
taken several days after the fire showed average concentrations of
PCBs in the air of the building of 1.5 µg/m³. The average result for
surfaces ranged from 4.6 to 162.2 µg/m². TCDFs and PCDDs were also
present. The soot samples were analysed for pyrolysis products. They
contained average levels of 3 mg TCDD/kg and 199 mg 2,3,7,8-TCDF/kg
(Fitzgerald et al., 1989). Achilles (1983) reported the following
levels in the deposited smut; 2160 mg PCDFs/kg and 20 mg PCDDs/kg
(including 0.6 mg 2,3,7,8-TCDD/kg).

In the soot from the Binghamton, Reims, and Stockholm accidents, high
levels of polychlorinated biphenylenes (PCBPs) were identified as well
as the PCDFs (Fig. 2) (Rappe et al., 1982, 1985).

Between 1981 and 1985, a number of accidents in electrical equipment
were reported from different countries; 28 accidents were mentioned in
WHO/EURO (1987) including actual capacitor explosions, capacitor
fires, and transformer accidents. In all eases, the accident site was
contaminated by PCDFs, average levels of total PCDFs being in the
range of 1-5 µg/m².

Hutzinger et al. (1985) also mentioned the presence of polychlorinated
pyrenes (PCPYs).

In the period 1977-85, particulates and flue gas from municipal
incinerators and hazardous waste incinerators in Canada, Denmark,
Netherlands, Sweden, and Switzerland were investigated. It was found
that emissions from incinerators contained many different PCDF and
PCDD isomers. The total levels ranged from ng/m³ to µg/m³. Fly-ash
contained levels of 0.1-0.6 mg/kg (Buser & Bosshardt, 1978; Rappe et
al., 1985c; WHO/EURO, 1987).

Rappe et al. (1985b) studied the emissions of the municipal solid
waste incinerator in Umea, Sweden. The levels of PCDDs and PCDFs
varied under different burning conditions. The amount of dioxins
formed seems to be dependent on the chlorine content in the waste, as
well as the construction of the incinerator. The critical parameters
seem to be temperature, residence time, turbulence, and excess air
(oxygen).

The 2,3,7,8-tetra-CDD was always found to be a very minor constituent,
whereas the 1,2,3,7,8-penta-CDD in all samples gave a medium-sized
peak. The 2,3,7,8-substituted PCDFs were always middle or major
components (WHO/EURO, 1987).

The fact that PCBs may be thermally converted to PCDFs has raised
concern that similar conversions might occur in electrical equipment,
such as capacitors and transformers, in which the dielectric fluids
used are subjected to modest temperature rises accompanied by
electrical stress. Brown et al. (1988) investigated the presence of
PCDFs in both used and unused capacitors and transformers and did not
find any evidence of an increase in PCDFs levels in the heavily used
capacitor or the transformer PCBs compared with levels in unused
samples.

For a number of years, concern has been expressed regarding the
release of PCBs and other dangerous compounds when fluorescent light
ballasts "burn out". The breakdown products may contain vapours and
condensed particles of PCBs and asphalt. In response to concern at a
school, the US EPA met with officials of Blaine Elementary School,
because of material leaking from some fluorescent light fixtures. It
was determined that the leaking material ("oil") contained PCBs
(Aroclor 1242 or 1260). Air samples collected following the burn out
of such lights, at different distances from the light fixture, gave
concentrations of 0.166 and 0.012 mg/m³, respectively, 1 and 6 m from
the light. Three days later, levels of 0.004-0.001 mg/m³ were still
found. In a second series of tests, both burn-out and non-burn-out
ballasts were heated to 150°C, 300°C, and 400°C, in a chamber. No PCBs
were detected at 150°C. At 300°C, concentrations ranged from 0.55 to
1.70 mg/m³ and, at 400°C, 2.54 to 28.2 mg/m³. Wipe samples were
taken in schoolrooms after burn-outs; average concentrations of
Aroclor of 0.34 and 1.22 µg/cm² were found. It is obvious that PCBs
and asphalt contamination, both surface and atmospheric, can occur
when fluorescent lamp ballasts burn out.

The most serious potential contamination results when thermal runaway
takes place. Thermal runaway volatilizes the asphalt potting compound
and may rupture the capacitor. When the potting compound and the PCBs
are exposed to high temperatures, some of both materials vapourizes.
As the vapours pass through the atmosphere they condense into freely
divided aerosols, less than 1 µm in diameter. Much of the visible
fumes results from volatilization of the asphalt (Anon., 1987).

4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION

4.1 Transport and distribution between media

A more detailed review of transport mechanisms can be found in Jury et
al. (1987).

4.1.1 Transport in air

The virtually universal distribution of PCBs throughout the world,
including the arctic and other remote areas, suggests that PCBs are
transported in air (Risebrough & de Lappe, 1972). The ability of PCBs
to co-distill, volatilize from landfills into the atmosphere
(adsorption to aerosols with particle size of less than 0.05-20 µm),
and resist degradation at low incinerating temperatures, makes
atmospheric transport the primary mode of global distribution within
the troposphere and stratosphere (Nisbet & Sarofim, 1972; Eisenreich
et al., 1981). PCBs have been measured in air samples at Eniwetok
Atoll in the North Pacific Ocean (Atlas & Giam, 1981), over the North
Atlantic (Giam et al., 1978), and in the Gulf of Mexico (Giam et al.,
1978, 1980). Murphy et al. (1985) estimated that approximately
18 000 kg of PCBs are present in the atmosphere over the USA, at any
given time. The authors also estimated that, if these PCBs had an
atmospheric residence time of one week, then about 900 000 kg/year of
PCBs cycle through the atmosphere of the USA.

Nisbet & Sarofim (1972) suggested that most of the airborne PCBs will
be adsorbed on any particles present. The half-life of particles in
the air will depend greatly on the size of the particles and the
extent of atmospheric precipitation. Most will be deposited within 2-3
days in their areas of origin (usually urban), the small amount
attached to fine particles will last in the atmosphere for longer
periods and can be transported to more remote regions.

Södergren (1972) collected airborne fallout in southern Sweden and
found regional differences in PCB levels, with mean monthly levels
ranging from 620 ng/m² per month to 10 510 ng/m² per month. The
lowest level was in a remote forest area. Industrialized areas had
high levels but so too did some agricultural regions. Higher levels
were generally found in the western part of the study region,
suggesting that some PCB fallout may have originated from further
afield and be dependent on the prevailing winds. Seasonal variations
in fallout correlated well with precipitation. Lower levels of PCB
precipitation were found in Iceland by Bengtson & Södergren (1974).
The highest level was found in Northern Iceland at 1050 ng/m² per
month and, like other sites sampled, showed a seasonal trend with
highest levels in the summer.

Harvey & Steinhauer (1974) measured PCBs in the atmosphere over the
western North Atlantic. They found that concentrations decreased
exponentially with distance from land and concluded that wind
transport is the major method of transport over the oceans. They also
suggested that PCBs are transported primarily in the vapour phase.

4.1.1.1 Dry deposition

Atmospheric input into the Great Lakes has been studied extensively,
because the lakes, as a whole, represent the largest surface area of
any freshwater body in the world, with the lake surface area
comprising from 27% (Ontario) to 64% (Superior) of the total basin
area, and ranging from 19 000 km² (Ontario) to 82 100 km²
(Superior). Eisenreich et al. (1981) estimated that more than 80% of
the annual mean total input of PCBs in Lake Michigan originated from
the atmosphere. They estimated that approximately 56% of the
9000 kg/year of PCB input in Lake Michigan was in the form of wet
deposition and that 30% of the 6600-8300 kg/year input in Lake
Superior was also in this form. However, Andren (1982) calculated a
precipitation input of 650 kg/year for Lake Michigan, again assuming
that all PCBs were on 0.5 µm airborne particles. Even assuming the
lowest estimate for the annual input of PCBs into the lake,
approximately 60% of the total input might be atmospheric deposition.

Andren (1982) also measured the input of PCB into an isolated lake
(Crystal Lake, Wisconsin), to calibrate a dry deposition model. The
model was then applied to Lake Michigan and the author concluded that,
assuming all particulate inputs of PCB are associated with 0.5 mm
particles, dry deposition inputs were significantly less than wet
inputs.

Manchester-Neesvig & Andren (1989) collected and analysed air samples
from a remote site in the Great Lakes watershed during 1984 and 1985.
Total PCB concentrations varied from 1.82 ng/m³ in the summer to
0.135 ng/m³ in the winter. They found that, on average, 92% of the
PCBs detected were in the vapour phase. When these data were compared
with data collected over the previous 7 years, no significant changes
in PCB concentrations were found. The authors concluded that, on the
basis of the short residence time and the relatively constant annual
average levels of PCBs, repeated cycling between earth and atmosphere
takes place.

Murphy (1984) reviewing data from the Great Lakes region on the
relative distribution of airborne PCBs between particulate matter and
vapour, concluded that they are transported predominantly in vapour.
He stated that there was reasonable evidence to suggest that the
atmosphere is the major source of the PCBs found in Lakes Michigan,
Superior, and Huron, Siskiwit Lake on Isle Royale, and probably in the
upper Great Lakes too.

Using liquid-coated collecting plates in near-shore areas of Lakes
Huron and Michigan, close to urban centres, more PCBs were found on
the upper plates suggesting that much of the dry deposit of PCBs was
associated with large particles (20 µm). This sampling technique also
indicated that, for the areas studied, dry deposition inputs were
higher than wet inputs (Murphy, 1984).

Duinker & Bouchertall (1989) analysed filtered air, particulates, and
rain, in the city of Kiel, Federal Republic of Germany for 14
different PCB congeners. They found that congeners with a low degree
of chlorination were dominant in filtered air, whereas, congeners with
a high degree of chlorination dominated in aerosols and rainfall. The
vapour phase represented up to 99% of the more volatile congeners
(i.e., those with a lower degree of chlorination). The particulates
were found to carry relatively more of the less volatile congeners.
Particle scavenging was the dominant source of PCBs in rain water
despite the small contribution of particulate PCBs to the overall
atmospheric concentration of PCBs (only 1 or 2%).

In a study by Södergren (1973), most of the PCB deposited on a south
Swedish lake was in the form of dry deposit, with 11% as particulate
matter in the precipitation and 2% from precipitation water. McClure
(1976) stated that, on the basis of flux measurements and model
calculations, most of the PCB fallout is in the form of dry deposition
and that most of the dry deposition of aerosol PCB introduced into the
troposphere falls within 100 km of its source.

4.1.1.2 Precipitation deposition

Precipitation scavenging of chlorinated hydrocarbons in the atmosphere
is complex. Scavenging of particles by cloud droplets and by rain
drops in, and below, clouds, and the scavenging of the vapour phase by
rain occurs (Murphy 1984). Thus chlorinated hydrocarbons are
concentrated in precipitation rather than in the atmosphere, resulting
in rainfall levels of many ng/litre. Swain (1978) and Strachan &
Huneault (1979) measured levels in rainfall ranging between 0 (not
detectable) and 230 ng/litre in the Great lakes area.

Murphy (1984) pointed out that variables, such as the amount of
particulate material and PCBs in the atmosphere, the type of rain, and
the rate of rainfall, will affect the precision of precipitation
estimates.

Levels of PCBs in the rainfall throughout Canada during 1984 were
monitored by Strachan (1988). Levels ranged from nd to 17 ng/litre, no
geographical trends were apparent.

4.1.2 Transport in soil

PCBs in soil, derive from particulate deposition (often concentrated
in urban areas), wet deposition, the use of sewage sludge as a
fertilizer, and leaching from landfill sites.

Significant amounts of PCBs are deposited on soil by particulate
deposition (see previous section). Fujiwara (1975) analysed soil
samples in Japan, and found that the main sources of PCB contamination
of agricultural soils are the industries using PCBs. Other sources
include treatment of soil with sewage sludge and accidental spills.
The 15% of soil samples in Indiana (USA) that contained more than
50 mg/kg had been treated with PCB-contaminated dried sludge (Bergh &
People, 1977).

Tucker et al. (1975a) found that, during a 4-month period following
the addition of Aroclor 1016 to soil, the PCBs were not readily
leached by percolating water and that only the lower chlorinated
isomers were leached. The ease of leaching from different soils was in
the order sandy loam > silty loam > silty clay loam.

The behaviour of ¹⁴C-labelled PCB in flooded soils was studied by
Ogiso et al. (1976). The amounts of PCB volatilized occurred in the
following order: water > subsoil > soil. The addition of compost
powder to soil reduced the amount that volatilized.

Haque et al. (1974) studied the adsorption of Aroclor 1254 on various
soil particle types in an aqueous solution of 56 µg PCB/litre.
Delmonte sand and silica gel did not adsorb any PCB. Woodburn soil
adsorbed the highest amount followed by illite, montmorillonite, and
kaolinite clays, in decreasing order. The high adsorptive capacity of
Woodburn soil was attributed to the presence of organic matter and
lipophilic or hydrophobic materials. Moza et al. (1976a) found that, 2
years after the application of ¹⁴C-labelled dichlorobiphenyl to a
loamy sand soil at 1 mg/kg, most of the detectable PCB was in the top
10 cm of the soil and only 0.2% had reached a depth of 40 cm. In
another study, Suzuki et al. (1977) found that Aroclors 1242 and 1254
did not move upwards through uncontaminated sand deposited over
contaminated soil. The leaching of water from soil may lead to a
downward movement of PCBs, depending on the soil type and clay content
(Pal et al., 1980).

A large spill of Askarel (containing 70% Aroclor 1254 and 30% tri-
and tetrachlorobenzenes) occurred at a transformer-manufacturing
facility in Canada, in 1976. Condie silt from near the site of the
spill was studied with respect to the sorption partition coefficients
and the transport retardation factors. The sorption partition
coefficient values for 2,5,2',5'-tetrachloro-, 2,4,5,2',5'-penta-
chloro-, and 2,4,5,2',4',5'-hexachlorobiphenyl were 5000, 9400, and
26 000, respectively. The mean transport retardation factors for these
3 congeners were 2.7 E + 04, 5.0 E + 04, and 1.4 E + 05, respectively.
This implies that dissolved PCBs will move only very slowly through
unfractured Condie silt (Anderson & Pankow, 1986).

4.1.3 Transport in water

PCBs enter water mainly from discharge points of industrial and urban
wastes into rivers, lakes, and coastal waters. In static water, PCBs
are more concentrated in the surface micro-layer than in subsurface
samples (Bidleman & Olney, 1974). This is probably due to deposition
from the air rather than redistribution in the water. On account of
their low water solubility and high specific activity, it is expected
that most of the PCBs discharged will be adsorbed by sediment at the
bottom of rivers or lakes and transport will be mainly via waterborne
particles (Nisbet & Sarofim, 1972). The bulk of the PCBs will sink to
the bottom sediments. The sinking rate of PCBs from the surface to
deeper layers in the open ocean is relatively slower in tropical
waters than in high-latitude waters (Tanabe, 1985).

Oloffs et al. (1973) added 0.1 mg Aroclor 1260/litre to water samples
in the presence of sediment. After 6 weeks, all of the PCBs had been
adsorbed by the sediment, none being given off to the atmosphere. The
degree of PCBs sorption is inversely related to the size of the
particles (Haque et al., 1974) and the solubility of PCBs in water
(Haque & Schmedding, 1975). Smaller particles have a relatively larger
surface area and so adsorb more PCBs (Steen et al., 1978). Nau-Ritter
et al. (1982) found the adsorption and retention of PCBs to be
directly related to the particle organic content. A significant
correlation was found by Larsen et al. (1985) between PCB levels and
total organic carbon in the deepwater sediments of the Gulf of Maine,
PCBs were concentrated on finer grain particles. Organic carbon and,
therefore, the PCB concentration were also correlated with depth.
Wildish et al. (1980) found that estuarine sediments, especially those
containing higher levels of organic matter, readily adsorbed Aroclor
1254. The PCBs were found to be tightly bound to the sediment with
virtually no desorption. Horzempa & Di Toro (1983) found that the
adsorption of hexachlorobiphenyl was correlated with both sediment
surface area and organic content. Adsorption was found to be
significantly greater at 40°C than at 1°C. Hexachlorobiphenyl is
strongly adsorbed on sediment and weakly desorbed. There is no simple
reversible reaction.

Fisher et al. (1983) found that the rate of release of PCBs from
contaminated sediment was a function of sediment PCB concentration,
chlorine substitution pattern, and degree of chlorination. In the
absence of disturbance, even very low deposition rates of new sediment
will quickly remove PCB-contaminated sediments from diffusional
communication with overlying water. Little change was found (Nimmo et
al., 1971a) in the PCB concentration in sediment at a point downstream
of a contamination source over a period of 9 months. The very small
amounts of PCBs leached from sediment into overlying water may be
taken up by organisms.

Hom et al. (1974) stated that the annual inputs of PCBs into the
southern California bight from waste water and from surface runoff in
1970-71 were estimated to be 10 and 0.25 tonnes, respectively.

Sewage treatment appears to remove PCBs from waste water,
concentrating them in the sludge. However, often, the sludge is then
discharged into open water (Ahling & Jensen, 1970). Holden (1970)
found an average of 3 mg PCBs/kg in wet sewage sludge dumped in the
Clyde estuary, in the United Kingdom, and calculated that this would
be equivalent to approximately one tonne per year. A similar annual
discharge of PCBs in the sludge on the Californian coast was
calculated by Schmidt et al., (1971).

Dredging of inland rivers and harbours may lead to a significant
transfer of PCBs from contaminated sediments, especially when dumped
at sea (Nisbet & Sarofim, 1972). Rice & White (1987) found that there
was an increase in water concentrations of PCBs immediately following
the dredging of sediment in the Shiawassee River, Michigan. The
availability of PCBs for clams and fish, as measured by an increase in
uptake, was found for up to 6 months following dredging.

4.1.4 Transport between media

In a model ecosystem, Södergren & Larsson (1982) found that the
presence of bottom-living organisms, such as Chironomus and
Tubifex, resulted not only in the uptake of PCBs from the sediment
but also in the release of PCBs into the water and to the surface
microlayer, compared with a system without organisms. PCBs were
transported to the air via jet drops from bursting bubbles in the
surface microlayer.

A similar pattern was found using large outdoor artificial ponds
(Larsson, 1985a). Following the addition of Clophen A50 to sediment,
the transport of PCBs from sediment to water followed a seasonal
cycle, with higher levels in the summer than in the winter. The
processes that transfer PCBs across the sediment/water interface
(bioturbation, desorption, and gas convection) are positively related

to temperature. Transfer from water to air was probably dominated by
volatilization with maximum concentrations of PCBs in air at the
highest water concentrations, lower chlorinated biphenyls achieving
the highest concentrations in air. The majority of the airborne phase
was presumed to be in the gaseous phase as it passed through particle
filters. In the same ponds, Larsson & Okla (1987) measured the rate at
which PCBs volatilized from water to air. PCB compounds volatilized at
a rate of 0.9 to 9.6 ng/m² per h, the rate increasing with the
temperature of the water and the concentration of PCBs. The transport
rate during the day exceeded the rate at night and was positively
correlated with the air temperature (Okla & Larsson, 1987).

Larsson (1985b) added Clophen A50 to the sediment in a model ecosystem
comprising sediment, water, benthic macroinvertebrates, and fish. PCBs
were detected in the water. The transport of PCBs from the water to
air included at least 2 routes, volatilization and jet drop transport.
Both routes were of the same magnitude (0.2-1.0 µg/week). However,
though the PCBs transported by volatilization consisted of lower
chlorinated isomers, those transported by jet drops were identical to
those in the sediment and water.

In an earlier study, Larsson (1984) measured the uptake of PCBs from
sediment by chironomid midge larvae and the concentrations of PCBs
from larva to adult. In the field, chironomid larvae contained
114 µg/kg fresh weight at a sediment concentration of 39 µg/kg wet
weight. Different sediments affected the amount of PCBs available to
the organisms. Adult chironomids sampled near a sewage plant contained
251 µg/kg fresh weight. The chironomid larval population was estimated
to be 9900 per m² and the authors calculated that these would move
20 µg PCB/m² per year into the terrestrial compartment of the
environment.

A model, based on the fugacity concept, was described and illustrated
by applying it to the time-varying fate of PCBs in Lake Ontario over
the period 1940-2000. Expressions are included for a great number of
variables, such as loadings and the partitioning of the contaminant
between the phases of air, aerosols, water, suspended and bottom
sediments, various trophic levels of aquatic organisms, and gull eggs.
Also included are expressions for transformation rates, and transport
rates for diffusion between water and sediment, and water and air wet
and dry atmospheric deposition, sediment deposition, burial, and
resuspension, and water and the inflow and outflow of suspended
matter. The results obtained by numerical integration and by assuming
reasonable loading and air concentrations were in accordance with
data. It was shown that PCBs cycle appreciably between the atmosphere
and water by wet and dry deposition and volatilization, and between
water and sediment by deposition, resuspension, and diffusion.
Biomonitors were shown to be particularly valuable indicators of
contamination levels in the ecosystem (MacKay, 1989).

4.2 Biotransformation

4.2.1 Biodegradation

Nissen (1973) did not find any alteration in Aroclor 1254 after a
9-week incubation period in soil. Iwata et al. (1973) added Aroclor
1254 to various soil types. They did not find any change after one
year in soils containing high amounts of organic matter (10.8-19.5%).

Biotransformation had occurred, causing the disappearance of the lower
chlorinated biphenyls, in soils with a low organic matter content
(0.1-3.3%), as diverse as loamy sand and clay. The authors concluded
that, after one year, the material remaining in loamy sand (0.1%
organic matter) consisted of mainly penta- and hexachlorobiphenyl
isomers.

4.2.1.1 Bacteria

The biodegradation of PCB isomers, which is possible with some aerobic
bacteria, depends on the degree of chlorination and the position of
chlorine substitution. Degradation decreases with increasing
chlorination. Dechlorination of PCBs occurs in anaerobic sediments.
Here bacterial activity is preferentially targeted towards PCB
congeners with higher levels of chlorination. Products of
dechlorination are, therefore, more readily degraded by aerobic
systems.

Early experiments were carried out to study the biodegradation of PCBs
using activated sludge inocula; some degradation was found (Baxter et
al., 1975). However, the presence of PCBs in sewage sludge shows that
they are not all readily transformed by microorganisms. Fries (1972)
analysed silage containing PCBs (Aroclor 1254) that had undergone
normal fermentation. The gas chromatogram of the standard was
identical to that of the silage sample. The authors suggested that, if
anaerobic degradation had taken place, it would have been unlikely to
have been uniform for all components. They stated, however, that this
test may not have been a good indication of possible anaerobic
degradation because DDT showed much less degradation, under the same
conditions, compared with other degradation test systems.

Lunt & Evans (1970) postulated a metabolic pathway, used by
microorganisms, for biphenyl oxidation, which was later confirmed by
the findings of Gibson et al. (1973) using a bacterium isolated from a
polluted stream. Lunt & Evans (1970) found that a Gram-negative
bacterium oxidized biphenyl to phenylpyruvic acid with the
intermediary formation of 2,3-dihydroxybiphenyl and
alpha-hydroxy-ß-phenylmuconic semialdehyde. Catelani et al. (1971)

found that the metabolism of biphenyl by Pseudomonas putida was
different, in that, though the intermediate products were the same,
benzoic acid was isolated, not phenylpyruvic acid. Ahmed & Focht
(1973a) isolated 2 species of Achromobacter from sewage effluent
using biphenyl and p-chlorobiphenyl as the sole carbon source. They
found that both sources were rapidly degraded, biphenyl being oxidized
to benzoic acid and both mono and dichlorinated biphenyls to
p-chlorobenzoic acid. In a second study, Ahmed & Focht (1973b)
investigated the biodegradation of other isomers of PCBs, with 2-5
chlorine atoms. The extent of oxidation seemed to be somewhat
dependent on the presence of unsubstituted biphenyl rings. Because of
the absence of chloride in all the supernatants, they concluded that
the bacterium was unable to dechlorinate the PCBs. The fact that
increasing chlorine substitution rendered the molecule more resistant
to microbial attack was used to support this argument. However, Kaiser
& Wong (1974), studying the degradation of Aroclor 1242 by a bacterial
culture, isolated from lake water, showed that the PCBs were degraded
into several metabolites (aliphatic and aromatic hydrocarbons), none
of which contained chlorine. Dechlorination had already taken place at
an early stage of metabolism.

Wong & Kaiser (1975) found that lake water bacteria could use both
Aroclor 1221 and 1242, but not 1254, as a sole carbon source for
growth, but that only 1% of the bacterial culture had this ability.
The authors then followed the degradation of Aroclor 1221. After one
month, the mixture had been totally degraded to several compounds of
low relative molecular mass. Unchlorinated biphenyls were degraded
faster than chlorinated forms.

Tucker et al. (1975b) observed the degradation rates of Aroclors 1221,
1016, 1242, and 1254, and MCS 1043 (a non-commercial mixture). They
found a clear relationship between the level of chlorination and the
relative degradability, when degradation rate was plotted against
percentage chlorine by weight. Volatilization rates fell within the
95% confidence limits of overall disappearance rates and so could be
ruled out. Analysis of the Aroclors, following exposure to the
activated sludge, revealed a redistribution of the dominant PCBs. For
example, the chromatograms for Aroclor 1221 and 1242 were very similar
showing that the lower chlorinated biphenyls were more rapidly
degraded. Furthermore, since Aroclor 1221 was found to be rapidly
degraded, a closer study was performed that showed that most of the
degradation occurred within 24 h.

The degradation of polychlorinated biphenyls by either Nocardia spp.
or Pseudomonas spp. was studied by Baxter et al. (1975). They found
that, under experimental conditions, many of the lower chlorinated
biphenyls (<3 chlorine atoms/molecule) were degraded very readily
and some biphenyls containing as many as 6 chlorine atoms could be
degraded, if the conditions were suitable. When PCB mixtures Aroclor
1016 and 1242 were used, a different pattern of degradation was
observed with an enhanced ability of the microorganisms to degrade.
For example, 4,4'-dichlorobiphenyl degraded to 50% in about 2 days,
when presented to Nocardia spp. as a component of Aroclor 1242, but
it was virtually unaffected after 12 days exposure as the pure isomer.
The authors suggested that mutual solubilization might play some part.

Sayler et al. (1977) found that an estuarine Pseudomonas sp. was
able to degrade both mixtures of PCBs (Aroclor 1254) and pure isomers
of hexachlorobiphenyl. Degradation was dependent on incubation time
and the purity and degree of chlorination of the biphenyl. Appreciable
degradation occurred at all substrate concentrations of the Aroclor
(10, 100, and 1000 µg/litre) within 22 days. Although, over this
22-day period, only 9% had been degraded at the lowest concentration
compared with 30-40% for the other concentrations, after 60 days, this
was reversed with 84% being degraded at 10 µg/litre, 70% at
100 µg/litre, and 63% at 1000 µg/litre. When compared with the pure
isomer, degradation of the Aroclor mixture proceeded at a slower rate.
Even though average chlorination was less, the authors speculated that
this could be owing to the substitution positions of the chlorines.
Chromatographic tracings showed that degradation of the lower
chlorinated components of the Aroclor occurs before degradation of the
more highly chlorinated biphenyls.

Furukawa et al. (1978a,b) examined 31 PCB isomers (mono to
pentachlorobiphenyl) for biodegradability by 2 bacterial species,
Alcaligenes and Acinetobacter. They found the following
relationship between chlorine substitution and biodegradability.

i. Degradation decreased as chlorine substitution increased.

ii. Isomers containing two chlorines at the ortho position of
either a single ring or on both rings showed very poor
degradability.

iii. Isomers, in which all the chlorines were on one ring, were
generally degraded faster.

iv. Molecules with non-chlorinated rings or rings with few chlorines
underwent preferential ring fission.

v. The 4'-chloro-substituted PCBs formed and accumulated a yellow
intermediate during degradation.

vi. Only with respect to 2,4,6-trichlorobiphenyl was there a
significant difference in ability to degrade between the 2
bacteria. This compound was mostly metabolized within 1 h by
Acinetobacter, but was degraded very slowly by Alcaligenes.

It was demonstrated by Carey & Harvey (1978) that mixed cultures of
marine bacteria were capable of metabolizing both pure isomers (tri-
and tetrachlorobiphenyl) and mixtures (Aroclor 1254). They isolated
and partially characterized an acid lactone metabolite. They did not
find any change in the chromatogram trace for the Aroclor but
suggested that this might be related to the insensitivity of the
method, since even if each of the isomers in the mixture had been
metabolized to the same extent as pure isomers, this would still not
have been detectable on the trace. The authors also found that no
metabolism occurred when a chlorobiphenyl isomer in an anaerobic
marine mud was incubated for 6 weeks. Degradation of Aroclor 1242 by
mixed microbial cultures, isolated from soil and river water samples,
was demonstrated by Clark et al. (1979). The predominant organisms in
the cultures were Alcaligenes odorans, Alcaligenes denitrificans,
and an unidentified bacterium. The lower chlorinated isomers were not
only degraded at a faster rate but were also more completely utilized
by the bacteria. In general, the rate of degradation was much faster
than in previous studies. Co-metabolism in the presence of sodium
acetate was studied; greatly enhanced degradation was found for the
more highly chlorinated isomers. Liu (1980) found that sodium
ligninsulfonate also greatly enhanced the biodegradation of commercial
PCB mixtures.

The same author found that a Pseudomonas sp. could oxidize Aroclors
1221, 1016, 1242, and 1254, at a rapid rate. A kinetic study using
resting cells revealed that Aroclor 1221 was degraded much faster
(980 µg/h per mg cell dry weight) than Aroclor 1254 (43 µg/h per mg
cell dry weight). The degradation of the higher chlorinated PCB
(Aroclor 1254) could be enhanced by the addition of Aroclor 1221. Liu
(1981) observed that the oxidation of Aroclor 1221 by the bacteria was
10 times faster than with sewage. Two possible explanations for this
difference were that the sewage contained toxic chemicals that
inhibited the bacteria, but this was found not to be the case, or, the
bacteria preferred Aroclor 1221 to the other substrates. This second
explanation is a possibility, for glucose, a substrate used readily by
most bacteria was poorly oxidized by this bacterium. Pseudomonas
oxidized Aroclor 1221 readily between 15 and 35°C, the rate increasing
with temperature. Reducing the temperature to 4 and 10°C drastically
retarded, but did not halt, degradation. Adjusting the concentrations
of phosphorus and nitrogen from 2 mg to 20 mg/litre (the lower
concentration being that found normally in sewage) did not alter the

rate of degradation by Pseudomonas spp. in raw sewage. But
increasing nitrogen and phosphorus gave more reproducible results,
suggesting that the compounds are on the border of limiting
degradation rates in raw sewage. The oxygen content was found not to
affect degradation at concentrations over 1 mg/litre (oxygen levels
are generally maintained at between 2 and 3 mg/litre in activated
sludge reactors, under the operational conditions of sewage-treatment
plants). Liu (1982) found that, under a limited substrate supply,
Pseudomonas spp. degraded all 7 of the major components of Aroclor
1221. However, with excessive amounts of nutrient, preferential
degradation of certain components was observed. The author stated that
one of the main factors influencing this selective biodegradability
was the position of chlorine substitution on the biphenyl.

4.2.2 Biodegradation; individual congeners

4.2.2.1 Bacteria

In a study by Parsons & Sijm (1988), the co-metabolism was
investigated of several different mono-, di- and tetrachlorobiphenyls
in chemostat continuous cultures of a Pseudomonas strain (JB1). They
found that chemostat conditions favoured degradation compared with
exposure of the Pseudomonos in batch culture, where little or no
degradation was recorded. Using benzoate as the carbon source, results
varied widely, with repeat incubations showing different degrees of
degradation of chlorobiphenyls and, sometimes, no breakdown at all. In
cultures that did degrade the materials, the monosubstituted
4-chlorobiphenyl was rapidly degraded. Of the disubstituted
dibiphenyls, 3,5-dichlorobiphenyl was more readily broken down than
2,5-dichlorobiphenyl. Changing the carbon source available to the
Pseudomonas sp. improved the reproducibility of the results. The
authors reviewed the literature relative to their own findings and
concluded that repeated culture on benzoate leads to the loss of the
ability of the Pseudomonas sp. to degrade biphenyl by meta
cleavage; ortho cleavage is retained. Coding for the meta cleavage
resides on plasmids, which can be lost, whereas coding for the ortho
cleavage is chromosomal. Growth of the Pseudomonas sp. on a
3-methylbenzoate substrate improved degradation of the biphenyls.
3-Methylbenzoate can only be degraded by a meta cleavage favouring
retention of the plasmid. Comparison of degradation of 4
tetrachlorobiphenyls showed the influence of the positions of the
chlorine substitutions. The relative degradability of the 5 compounds,
shown in Fig. 3, was: 2,3,2',3'-tetrachloro- >2,5,3',4'-tetrachloro-
> 2,5,2',5'-tetrachloro- approx. 2,6,2',6'-tetrachloro- approx.
3,4,3',4'-tetrachlorobiphenyl. The authors stated, from the
literature, that the first reaction in the degradation of
chlorobiphenyls is, in most cases, 2,3-dioxygenation, eventually
leading to the formation of chlorobenzoates. Chlorines in the ortho
and meta positions will, therefore, offer steric hindrance to this
reaction.

The low degradation rate of 3,4,3',4'-tetrachlorobiphenyl is not
explained by this mechanism, since it has 2 adjacent unoccupied 2,3
positions, but is more likely explained by its toxicity. Steric
influence on enzyme binding is offered as an explanation in this case.
Similarly, Furukawa et al. (1978a) did not find any degradation of
this compound in initial studies, though they did find degradation to
a dichlorobenzoic acid by Acinetobacter in a later study (Furukawa
et al., 1978b; Rogers, undated(a)).

Brown et al. (1987a,b) examined patterns of PCB congeners remaining in
sediments after spills of commercial mixtures of Aroclor. Sediment
from 5 different sites was examined. Shifts in gas chromatographic
peak distribution were indicative of dechlorination of congeners by
anaerobic bacteria in the sediment. Analysis of sediment from
different depths indicated less difference from the original traces in
superficial layers and the greatest shift in deeper layers of the
sediment cores. They concluded that dechlorination had taken place and
deduced several different processes involved by comparison between
sites. Six of these processes have been characterized in detail, each
presumed to be mediated by different populations of anaerobic
bacteria, with different selectivity for different congeners in the
PCB mixture. The point of most interest was that congeners with high
degrees of chlorination were selectively dechlorinated by these
anaerobic organisms. Whilst dechlorination still leaves the mass of
PCB intact, congeners with lower chlorination can be more readily

degraded by aerobic bacteria. This anaerobic dechlorination,
therefore, enables further degradation to take place elsewhere and
contributes significantly to the detoxification of the PCBs. While the
combined meta- para selective dechlorinating/oxidizing action of
sediment microbes for PCB residues is likely to be detoxifying, with
respect to dioxin-like effects, there are reservations about whether
this action would be detoxifying in respect of other, more subtle
toxic effects of PCBs and their degradation products, known (such as
the potential reproductive toxicity of the hydroxylated,
ortho-enriched PCBs from sediment microbe action) and unknown. This
is why it is important to study not only the disappearance of PCBs,
but also the exact nature and amounts of the degradation products
(McKinney et al., 1990). Two broad categories of transformation have
been observed: the first dechlorinates in the ortho, meta, and para
positions and the potential for the dechlorination of biphenyls is
related to the reduction potential of the compound, the second
dechlorinates only in the meta and para positions, and the
reactivities of the congeners relate to the molecular shape. The
second category suggested to the authors an active site on a
dechlorinating agent that would be roughly conical with a reducing or
hydrogenating site at the apex. In this schema, para-substituted
molecules could enter the site directly, enough rotation of the
molecule would be possible for the accommodation of meta, but not
ortho, substitution. Quensen et al. (1988) demonstrated this
dechlorinating capacity of anaerobic bacteria from Hudson River
sediments in the laboratory. Dechlorination occurred primarily from
the meta and para positions; ortho-substituted congeners
accumulated selectively. The fastest rate of dechlorination occurred
at the highest exposure used (700 mg Aroclor 1242/kg); 53% of the
total chlorine was removed over a 16-week incubation period. During
incubation, the proportion of mono- and dichlorobiphenyls increased
from 9 to 88%. The authors believed that a sequential anaerobic to
aerobic system could be devised for the biological degradation of
PCBs.

4.2.2.2 Fungi

Wallnofer et al. (1973) incubated a soil fungus Rhizopus japonicus
in a medium containing ³H-labelled 4-chlorobiphenyl or
4,4'-dichlorobiphenyl. After incubation for 1 week, the fungal
mycelium was filtered out. Scans of TLC plates indicated a
hydroxybiphenyl derivative present in the filtrate of both cultures.
To further identify the metabolite, larger amounts of unlabelled
4-chlorobiphenyl were added to a similar culture. The NMR and mass
spectra were identical to a synthetic sample of 4- chloro-4'-hydroxy-
biphenyl; mixed melting point determination showed no depression.
Further positive identification of the product was not possible,
because of limited material, but the experiment indicates the
probability of degradation of biphenyl to a hydroxy derivative by a
fungus.

4.2.3 Photodegradation

Several authors have reported that simple chlorinated biphenyls, as
well as complex commercial PCB mixtures, undergo photoreduction in
organic solvents (Safe & Hutzinger, 1971; Hustert & Korte, 1972; Ruzo
et al., 1972, 1974, 1975; Sawai & Sawai, 1973; Koshioka et al., 1987)
and aqueous systems (Crosby & Moilanen, 1973; Bunce, 1978) in the
laboratory. Herring et al. (1972) found that PCBs degraded faster in
hexane solution than in aqueous solution and slower in benzene
solution.

Bunce et al. (1978) posed the question of the environmental
significance of the photodegradation of PCBs and tried to estimate the
likely degree of photolysis under real environmental conditions,
rather than in solution in organic solvents at high concentrations.
The current best estimate suggests that significant amounts,
particularly of higher chlorinated PCB congeners, might be degraded in
water by the action of sunlight.

4.2.4 Bioaccumulation, distribution in organisms, and elimination

Polychlorinated biphenyls accumulate in almost all organisms, because
of their high lipid solubility and slow rate of metabolism and
elimination. They accumulate preferentially in fat-rich tissues.

Bioconcentration factors (BCFs) should be interpreted with caution,
since they are simple ratios. The exposure concentration, therefore,
makes a marked difference to the BCF obtained; very low exposure
concentrations are likely to lead to high BCFs, since all the PCBs are
absorbed, whilst high exposure concentration will tend to minimize the
BCFs.

Experimental data on the bioconcentration of PCB mixtures and pure
chlorinated biphenyls are presented in Table 9 for microorganisms,
Table 10 for aquatic organisms, and Table 11 for plants, birds, and
mammals.

4.2.4.1 Microorganisms

Uptake of both pure chlorinated biphenyl isomers and commercial PCB
mixtures by microorganisms is rapid, and high bioconcentration factors
are achieved. While there is a suggestion in studies on some species
that PCB congeners with higher levels of chlorination are taken up
preferentially, in the majority of studies, all PCBs appear to be
taken up equally. Uptake is true absorption; adsorption onto the
surface of the organisms represents little of the uptake. Since
resistant forms of microorganisms take up less PCBs than sensitive
forms and dead cells accumulate more PCBs than live ones, there is
some capacity to exclude the compounds.

Harding & Phillips (1978b) studied the uptake of ¹⁴C-labelled
2,4,5,2',5'-pentachlorobiphenyl, at concentrations of 0.31 or
9.86 µg/litre water, by 11 marine phytoplankton species including:
diatoms, green algae, chrysophytes, haptophytes, and dinoflagellates.
The cell density of each culture was maintained at 106-109 cells/
litre. Equilibrium between water and cell concentrations of biphenyl
was reached very rapidly after 0.5-2 h; small motile forms reached
equilibrium within 1 h and large centric diatoms after approximately
2 h. Exposure concentration and cell density, within the range given
above, had little effect on the time-course of uptake. Substantial
interspecies differences in adsorptive capacity were shown by
differences in the Freundlich adsorption constant (log K). A large
centric diatom, Coscinodiscus sp., had the highest log K. Nitzschia
longissima, a penate diatom that has been shown to be resistant to
PCBs (Harding & Phillips, 1978a), had the lowest log K value. The
flagellates, with the exception of Monochrysis lutheri, which has
been shown to be very sensitive to the effects of PCBs, had much lower
log K values than diatoms. Concentration factors, calculated from the
Freundlich adsorption isotherms, ranged between 12 300 and 2 410 000.

Biggs et al. (1980) exposed mixed species of estuarine phytoplankton
(numerically dominated by the diatom Skeletonema costatum) to
¹⁴C-labelled PCB (approximately 54% chlorine by weight) at
concentrations of 5.8 or 11.6 µg/litre. At a particle concentration of
25 mg/litre, 19-22% of the labelled-PCB was sorbed on the particles
after a 1-h exposure, with 70-72% in the water. At 4 times the above
particle concentration, 66-69% was sorbed on particles and only 22-23%
was retained in the water. Doubling the amount of ¹⁴C-PCB doubled the
mean amount of labelled-PCB in both the particles and the water. The
authors calculated an index of sorption (the ratio of ¹⁴C-PCB sorbed
on particles to that in an equal volume of water) at an average of
2 ± 1 × 10⁴. The authors suggested that the higher uptake (88%) of
PCBs found by Södergren (1971) was probably the result of an
unnaturally high cell concentration. Phytoplankton sampled in the
surface waters of Long Island Sound, USA, varied seasonally in
concentration from about 0.5 to 30 mg/litre.

Lederman & Rhee (1982) calculated bioconcentration factors for 3
species of Great Lakes planktonic algae (Table 9). In the case of
Fragilaria crotonensis, the uptake of hexachlorobiphenyl into the
frustule (the siliceous wall of the diatom) was investigated. The
bioconcentration factors for frustules were lower by an order of
magnitude than the factors for live and dead cells. It appears,
therefore, that adsorption on the cell surface contributes only a
little to the bioaccumulation of hexachlorobiphenyl.

Table 9. Bioaccumulation of PCBs: Microorganisms

Organism Biomass Temperature PCB type Duration Exposure Bioconcentration Reference
(cells/ml) (°C) (µg/litre) factor^a

Green alga 2 × 10⁶ 20-25 TeCB 1 h 10 3200 Urey et al. (1976)
Chlorella 2 × 10⁶ 20-25 HeCB 1 h 10 7000 Urey et al. (1976)
pyrenoidosa 2 × 10⁶ 20-25 OcCB 1 h 10 1600 Urey et al. (1976)
2 × 10⁶ 20-25 DeCB 1 h 10 5200 Urey et al. (1976)

Algae 3.2 × 10⁵ HeCB 19 h 1 117 000^b Lederman & Rhee
Fragilaria 1.6 × 10⁵ HeCB 19 h 1 313 000^b (1982)
crotonensis Lederman & Rhee
(1982)

Algae 3.4 × 10⁵ HeCB 6 h 1 619 000^b Lederman & Rhee
Ankistrodesmus 1.7 × 10⁵ HeCB 6 h 1 959 000^b (1982)
falcatus 8.5 × 10⁴ HeCB 6 h 1 1 207 000^b Lederman & Rhee
(1982)
Lederman & Rhee
(1982)

Algae 1.1 × 10⁶ HeCB 6 h 1 129 000^b Lederman & Rhee
Mycrocystis sp. 5.5 × 10⁵ HeCB 6 h 1 170 000^b (1982)
2.8 × 10⁵ HeCB 6 h 1 264 000^b Lederman & Rhee
(1982)
Lederman & Rhee
(1982)

Table 9. (cont'd).

Organism Biomass Temperature PCB type Duration Exposure Bioconcentration Reference
(cells/ml) (°C) (µg/litre) factor^a

Fungus 22-25 Aroclor 1254 24 h 0.007 mg/kg 1327^b,c Pinkney et al. (1985)
Fusarium 22-25 Aroclor 1254 48 h 0.007 mg/kg 1144^b,c Pinkney et al. (1985)
oxysporum

^a Concentration of PCBs in organism/concentration of PCBs in medium or food; bioconcentration factors calculated on a wet weight
basis unless otherwise stated.
^b Calculated on a dry weight basis.
^c Radioactive isotope used to calculate bioconcentration factor.

Table 10. Bioaccumulation of PCBs: Aquatic organisms

Organism Stat/ Organ^b Temperature PCB Type Duration Exposure Bioconcentration Reference
flow^a (°C) (µg/litre) factor^c

American oyster flow WB Aroclor 1016 96 h 0.6 6666 Hansen et al. (1974b)
Crassostrea WB Aroclor 1254 56 d 0.01 165 000 Parrish (1973)
virginica WB Aroclor 1254 392 d 0.01 89 000 Parrish (1973)

Polychaete stat WB Aroclor 1254 5 d 1.1 236 Courtney & Langston
Arenicola marina stat WB Aroclor 1254 5 d 1 mg/kg^d 0.24 (1978)

Polychaete stat WB Aroclor 1254 5 d 1.1 373 Courtney & Langston
Nereis stat WB Aroclor 1254 5 d 1 mg/kg^d 0.36 (1978)
diversicolor

Water flea flow WB 20-22 Aroclor 1254 96 h 1.1 47 000^e* Sanders & Chandler
Daphnia magna (1972)

Amphipod (M) stat^f WB Aroclor 1254 24 h 0.03 8700 Pinkney et al. (1985)
Gammarus stat^f WB Aroclor 1254 24 h 195.8 mg/kg 0.118 Pinkney et al. (1985)
tigrinus

Scud flow WB 20-22 Aroclor 1254 96 h 1.6 24 000^e* Sanders & Chandler
Gammarus flow WB 20-22 Aroclor 1254 21 d 1.6 27 000^e (1972)
pseudolimnaeus

Glass shrimp flow WB 20-22 Aroclor 1254 96 h 1.3 12 300^e* Sanders & Chandler
Palaemonetes flow WB 20-22 Aroclor 1254 21 d 1.3 16 600^e* (1972)
kadiekensis

Table 10. (cont'd).

Organism Stat/ Organ^b Temperature PCB Type Duration Exposure Bioconcentration Reference
flow^a (°C) (µg/litre) factor^c

Brown shrimp flow WB Aroclor 1016 96 h 0.9 4222 Hansen et al. (1974b)
Penaeus aztecus

Grass shrimp flow WB 17-28 Aroclor 1254 7 d 2.3 11 000 Nimmo et al. (1974)
Palaemonetes flow WB 17-28 Aroclor 1254 16 d 1.3 14 000 Nimmo et al. (1974)
pugio flow WB 17-28 Aroclor 1254 28 d 0.62 17 450 Nimmo et al. (1974)
flow WB 17-28 Aroclor 1254 35 d 0.62 26 580 Nimmo et al. (1974)
flow WB Aroclor 1016 96 h 0.4 2750 Hansen et al. (1974b)

Crayfish flow WB 20-22 Aroclor 1254 96 h 1.2 1700^e* Sanders & Chandler
Orconectes nais flow WB 20-22 Aroclor 1254 21 d 1.2 5100^e* (1972)

Stonefly flow WB 20-22 Aroclor 1254 96 h 2.8 2500^e* Sanders & Chandler
Pteronarcys flow WB 20-22 Aroclor 1254 21 d 2.8 2800^e* (1972)
dorsata

Dobsonfly flow WB 20-22 Aroclor 1254 96 h 1.1 4600^e* Sanders & Chandler
Corydalus flow WB 20-22 Aroclor 1254 21 d 1.1 6800^e* (1972)
cornutus

Phantom midge flow WB 20-22 Aroclor 1254 96 h 1.3 23 600^e* Sanders & Chandler
Chaoboruspuncti (1972)
pennis

Mosquito larvae flow WB 20-22 Aroclor 1254 96 h 1.5 18 000^e* Sanders & Chandler
Culex tarsalis (1972)

Table 10. (cont'd).

Organism Stat/ Organ^b Temperature PCB Type Duration Exposure Bioconcentration Reference
flow^a (°C) (µg/litre) factor^c

Mayfly flow WB 8 Clophen A50 6 d 0.526 2940 Södergren &
Ephemera danica Svensson (1973)

Pinfish flow WB Aroclor 1016 96 h 0.8 2750 Hansen et al. (1974b)
Lagodon flow WB Aroclor 1016 28 d 1 n 25 000 Hansen et al. (1974b)
rhomboides flow WB Aroclor 1016 56 d 1 n 17 000 Hansen et al. (1974b)

Sheepshead flow^g WB Aroclor 1016 33 d 1 n 26 000 Hansen et al. (1975)
minnow flow^g WB Aroclor 1016 28 d 1 n 54 000 Hansen et al. (1975)
Cyprinodon flow^g WB Aroclor 1016 28 d 1 n 22 000 Hansen et al. (1975)
variegatus

Spot flow WB Aroclor 1254 7 d 1 n 7200 Hansen et al. (1971)
Leiostomus flow WB Aroclor 1254 14 d 1 n 17 000 Hansen et al. (1971)
xanthurus flow WB Aroclor 1254 28 d 1 n 37 000 Hansen et al. (1971)
flow WB Aroclor 1254 56 d 1 n 27 000 Hansen et al. (1971)

Atlantic salmon flow WB 10-15 Aroclor 1254 33 d 10 mg/kg 0.39 Zitko (1974)
Salmo salar

Table 10. (cont'd).

Organism Stat/ Organ^b Temperature PCB Type Duration Exposure Bioconcentration Reference
flow^a (°C) (µg/litre) factor^c

Coho salmon flow WB 17 Aroclor 1254 112 d 0.048 mg/kg 9.79 Mayer et al. (1977)
Oncorhynchus flow WB 17 Aroclor 1254 112 d 4.8 mg/kg 0.79 Mayer et al. (1977)
kisutch WB TeCB 17 d 1 mg/kg 0.144 Gruger et al. (1976)
WB TeCB 35 d 1 mg/kg 0.139 Gruger et al. (1976)
WB PeCB 35 d 1 mg/kg 0.162 Gruger et al. (1976)
WB HeCB 35 d 1 mg/kg 0.151 Gruger et al. (1976)

Channel catfish flow WB 26 Aroclor 1232 150 d 2.4 mg/kg 1.875 Mayer et al. (1977)
Ictalurus flow WB 26 Aroclor 1232 193 d 2.4 mg/kg 1.3 Mayer et al. (1977)
punctatus flow WB 26 Aroclor 1248 193 d 2.4 mg/kg 0.79 Mayer et al. (1977)
flow WB 26 Aroclor 1254 193 d 2.4 mg/kg 2 Mayer et al. (1977)
flow WB 26 Aroclor 1260 193 d 2.4 mg/kg 1.46 Mayer et al. (1977)
flow WBh 24-26 Aroclor 1242 130 d 20 mg/kg 0.72 Hansen et al. (1976a)
flow WB Aroclor 1248 77 d 5.8 56 370* Mayer et al. (1977)
flow WB Aroclor 1254 77 d 2.4 61 190* Mayer et al. (1977)

Table 10. (cont'd).

Organism Stat/ Organ^b Temperature PCB Type Duration Exposure Bioconcentration Reference
flow^a (°C) (µg/litre) factor^c

Fathead (M) flow WB 25 Aroclor 1248 250 d 3 approx. 60 000 DeFoe et al. (1978)
minnow (M) flow WB 25 Aroclor 1260 250 d 2.1 approx. 160 000 DeFoe et al. (1978)
Pimephales (F) flow WB 25 Aroclor 1248 250 d 3 approx. 120 000 DeFoe et al. (1978)
promelas (F) flow WB 25 Aroclor 1260 250 d 2.1 approx. 270 000 DeFoe et al. (1978)

d = Days; M = Male; F = Female; DiCB = dichlorobiphenyl; TeCB = tetrachlorobiphenyl; PeCB = pentachlorobiphenyl;
HeCB = hexachlorobiphenyl; OcCB = octachlorobiphenyl; DeCB = decachlorobiphenyl.
^a Stat = static conditions (water unchanged for duration of experiment); flow = flow-through conditions (PCB concentration
in water continously maintained).
^b WB = whole body.
^c Bioconcentration factor = concentration of PCBs in organism/concentration of PCBs in medium or food; bioconcentration
factors calculated on a wet weight basis unless otherwise stated. * Radioactive isotope used to calculate
bioconcentration factor.
^d Sediment.
^e Calculated on a dry weight basis.
^f Static conditions, but test solution changed at intervals.
^g Intermittent flow-through conditions.
^h Not including stomach.

Södergren (1971) maintained the unicellular freshwater green alga
Chlorella pyrenoidosa in water (at a cell concentration of
approximately 900 mg/litre) with added nutrient medium containing
3.7 µg Clophen A50/litre, over a period of 7 days. By the end of the
experiment, 88% of the PCBs had been taken up by the alga. The
remaining PCBs were detected in the water, none being found in the air
samples taken. In another study, Urey et al. (1976) found that both
tetrachloro- and hexachlorobiphenyl isomers, at 10 µg/litre, were
concentrated by dead Chlorella pyrenoidosa cells by 6000 and 15 000
times, respectively, after a 1-h exposure. These concentration factors
are approximately twice those for living cells (Table 9). Similar
findings have been noted with other species of algae (Biggs et al.,
1980; Lederman & Rhee, 1982).

The ciliate Tetrahymena pyriformis was exposed to Aroclors 1248
(0.01, 0.1, and 1 mg/litre) and 1260 (0.001, 0.01, 0.1, and
1 mg/litre) for 7 days (Cooley et al., 1973). Uptake of the toxicant
increased linearly with increasing concentration. Concentration
factors ranged from 14.8 to 40.6 for Aroclor 1248 and from 21 to 79
for Aroclor 1260. Approximately 15-20% of Aroclor 1248 was absorbed at
each concentration compared with means of 37-53%, with increasing
concentration, for Aroclor 1260. If the data from Cooley et al. (1972)
on the uptake from Aroclor 1254 is included, it is clear that
T. pyriformis accumulates more PCBs with increasing degree of
chlorination.

Dive et al. (1976) studied the accumulation of 16 pure isomers of PCB
and one commercial product, Pyralene 3010, by the ciliate protozoan
Colpidium campylum, at concentrations of 0.1, 1, or 10 mg/litre for
43 h. The amount of PCBs taken up at 0.1 mg/litre was very similar for
each of the PCB isomers and the commercial product, ranging from 29.4
to 49%. The percentage uptake did not change greatly for the higher
exposures.

4.2.4.2 Plants

Uptake of PCBs into plants from soil is positively correlated with the
soil concentration of the PCBs. Roots accumulate more than stems and
foliage. Bioconcentration factors are low. More lower chlorinated
congeners of the PCBs are taken up, probably because of their greater
mobility in the soil. Much of the uptake is adsorption on the surfaces
of roots and there is little translocation. PCBs found in leaves have
volatilized from the soil. Uptake on root surfaces can be reduced or
eliminated by adding activated charcoal to the soil.

Lawrence & Tosine (1977) found that plants took up significant amounts
of PCBs (30-140% of the applied PCB concentration) from soil treated
with sewage sludge. In a waste PCB spill besides a North Carolina
highway, levels as high as 4700 mg/kg were recorded in the top 3 cm of
soil. Seven months later, the PCB concentrations were unchanged; the
authors believed that this was because the PCBs were bound to
activated carbon that had been used to treat the spill (Pal et al.,
1980).

Strek & Weber (1982) analysed statistically the data from several
literature sources (Iwata et al., 1974; Wallnofer & Koniger, 1974;
Wallnofer et al., 1975; Iwata & Gunther, 1976; Moza et al., 1976a,
1979a,b; Weber & Mrozek, 1979) on PCB uptake by plants, with the
following conclusions.

i. The PCB content of the plant is significantly dependent on the
soil PCB concentration.

ii. There is a significant difference between plant species,
carrots taking up more PCBs than other plants.

iii. There appears to be a limit of the PCB concentration in the
soil at which no detectable PCBs are taken up by the plants.

iv. Roots take up more PCBs than tops.

v. most of the PCBs in roots may, in fact, be adsorbed on the
surface and not actually taken up.

vi. There is a general trend of increasing PCB content with
decreasing chlorination, for pure PCB congeners.

vii. The amount of chlorination seems to have an effect on the
mobility of PCBs within plant parts. Since lower chlorinated
PCBs have been reported to be more mobile in soils than highly
chlorinated PCBs, they may be more readily transported and
available for plant uptake.

Larsson (1987) maintained the macroalga Cladophora glomerata in a
flowing-water, outdoor pool. Sediment contaminated with Clophen A50 at
2.7 mg/kg dry weight was added and PCB residues in the alga were
monitored. The algal concentration was 3.55 mg/kg dry weight within 3
months. Residues had fallen a year later to 0.2 mg/kg, reflecting the
water levels of PCBs. The authors concluded that a partitioning
process governed the uptake of PCBs by C. glomerata in this
experiment, because the alga accumulated the same PCBs and the same
proportion of PCBs that were present in the water.

Table 11. Bioaccumulation, of PCBs: Plants, birds, and mammals

Organism Organ PCB type Duration Exposure Bioconcentration Reference
(mg/kg) factor^a

Soil^b

Beet (Beta vulgaris) plant top Aroclor 1254 39 days 20 0.041^c Strek et al. (1981)

Sorghum (Sorghum bicolor) plant top Aroclor 1254 39 days 20 0.003^c Strek et al. (1981)

Peanut (Arachis hypogaea) plant top Aroclor 1254 78 days 20 0.024^c Strek et al. (1981)

Corn (Zea mays) plant top Aroclor 1254 13 days 20 0.001^c Strek et al. (1981)

Carrot root DiCBd 112 days 0.118 2^c Moza et al. (1976a)
leaves DiCBd 112 days 0.118 0.92^c Moza et al. (1976a)

Food

White pelican carcase Aroclor 1254 70 days 144 14.8 Greichus et al. (1975)
(Pelecanus erythrorhynchos)

Table 11. (cont'd).

Organism Organ PCB type Duration Exposure Bioconcentration Reference
(mg/kg) factor^a

Chicken fat Aroclor 1242 28 days 100 2.83 Harris & Rose (1972)
fat Aroclor 1254 28 days 100 5.15 Harris & Rose (1972)
fat Aroclor 1260 28 days 100 4.82 Harris & Rose (1972)

Big brown bat carcase Aroclor 1254 37 days 9.4 6.6 Clark & Prouty
(Eptesicus fuscus) (1977)

Mink fat Aroclor 1254 approx. 56 days 1.5 16.5 Hornshaw et al.
(Mustela vison) fat Aroclor 1254 approx. 126 days 1.5 28.5 (1983)
Hornshaw et al.
(1983)

^a Bioconcentration factor = concentration of PCBs in organism/concentration of PCBs in medium or food; bioconcentration
factors calculated on a wet weight basis, unless otherwise stated.
^b Calculated on a dry weight basis.
^c Radioactive isotope used to calculate bioconcentration factor.
^d DiCB = dichlorobiphenyl.

Red mangrove (Rhizophora mangle) seedlings were grown for 6 weeks in
soil treated with Aroclor 1242 at concentrations of between 0.038 and
6 mg/kg (Walsh et al., 1974). Low levels (detection limit was
0.1 mg/kg) of the PCBs were detected in the roots at exposure
concentrations of 3 or 6 mg/kg, during the exposure period, but no
residues were found in the stems. Residues were detected in both the
hypocotyls and leaves at application rates of 0.3 mg/kg or more. Leaf
residues did not change with time, but PCB concentrations in the
hypocotyls showed an increase. The highest mean residues of 1.5 mg/kg
were found in the hypocotyl in the highest exposure group.

Iwata et al. (1974) treated soil with Aroclor 1254, at a concentration
of 100 mg/kg, and sowed carrots in the plot 7 months later. The
carrots were harvested 3 or 4 months after seeding. The authors found
that the lower chlorinated biphenyls were more readily taken up from
the soil into the carrot root. Analysis of the carrot peel revealed
approximately 97% of the PCB residue, showing that there is little
translocation within the plant; 23 months after sowing carrots in soil
contaminated with 100 mg/kg Aroclor 1254, dissipation from soil
appeared to parallel the degree of chlorination (Iwata & Gunther,
1976). Analysis of the soil revealed that the lower chlorinated
biphenyls were slowly dissipated while the more highly chlorinated
biphenyls appeared to be unaffected. Small amounts of PCBs were found
in carrot foliage and the authors suggested that the PCB composition
indicated that they came from soil dust. Suzuki et al. (1977) also
found that lower chlorinated biphenyls were preferentially taken up by
plants, following exposure of soybean sprouts to soil contaminated
with Aroclor 1254 or 1242 at 100 mg/kg.

Moza et al. (1976a) found that carrot bioconcentrated
2,2'-dichlorobiphenyl (0.118 mg/kg soil) from soil by a factor of 2
(Table 11). No bioaccumulation was found in sugar beet, but the soil
residue was only 0.029 mg/kg. Carrots were grown in soil amended with
either ¹⁴C-labelled 2,5,4'-trichlorobiphenyl at 1.28 kg/ha or
2,4,2',4',6-pentachlorobiphenyl at 1.12 kg/ha for one season (Moza et
al., (1979a). Only 32.5% of the trichlorobiphenyl was recovered, the
rest being lost through volatilization. The carrots had taken up 3.1%
of the applied ¹⁴C, representing a concentration factor of 2.8. For
the pentachlorobiphenyl, 58.5% was recovered, 1.4% of which had been
taken up by the carrots. Sugar beet grown in the soil the following
year accumulated only 0.4% of the applied ¹⁴C.

In a study by Weber & Mrozek (1979), ¹⁴C-labelled Aroclor 1254 was
applied to Lakeland soil at a rate of 20 mg/kg. Activated carbon was
mixed with half the pots at a rate of 3.7 t/ha (3333 mg/kg). The pots
were seeded with either soybean or fescue. After harvesting at 16 days
for soybean and 50 days for fescue, the amounts of labelled-PCBs,
recovered from the plant tops, were 0.016% and 0.17% for the 2

species, respectively. The addition of activated carbon to the soil
reduced the uptake of ¹⁴C-PCBs, the recovery of labelled-PCBs being
0.001% and 0% for soybean and fescue, respectively. Strek et al.
(1981) also applied ¹⁴C-labelled Aroclor 1254, at the same rate, to
Lakeland soil; several species of crop plants were grown in the soil
and bioaccumulation factors, calculated (Table 11). Addition of
activated carbon (3.7 t/ha), equivalent to 3333 mg/kg to replicate
pots, reduced the uptake of the labelled-PCBs by 80-100%.

When approximately 1 mg ¹⁴C-labelled Aroclor 1254/kg was applied to
the centre leaflet of the first trifoliate leaf of 18-day-old soybean
plants, only 6.7% was recovered from the plant after 12 days, 76% of
which was still present in the treated leaf (Weber & Mrozek, 1979).

Mrozek & Leidy (1981) transferred the marsh plant Spartina
alterniflora from the field into soil containing 1 mg Aroclor
1254/kg (dry weight) and harvested the plants after a growth period of
90 days. The plants were found to take up selectively the lesser
chlorinated biphenyls. The authors stated that a further shifting of
the chromatographic pattern of PCBs towards the lesser chlorinated
components in aerial tissues suggested that some alteration of the PCB
mixture occurs in the plant. Mrozek et al. (1982) also found that
Spartina accumulates PCBs from both contaminated sand and mud-soil
systems. The total ¹⁴C-radioactivity accumulated in plants grown in
sand systems was higher than that in plants grown in mud. The level of
radioactivity accumulated in the green parts of the plants was similar
in both soil systems.

Moza et al. (1976b) applied 76 mg/kg of ¹⁴C-labelled 2,5,4'-tri-
chlorobiphenyl or 133 mg/kg of labelled 2,4,6,2',4'-pentachloro-
biphenyl to the leaves of the marsh plant Veronica beccabunga. Six
weeks later, the total recovery from plant, water, and soil was 3.7
and 18.3%, respectively, 86 and 95% of which was recovered from the
plant. In an earlier study, Moza et al. (1974) applied ¹⁴C-labelled
2,2'-dichlorobiphenyl in water or soil to 2 higher water plant species
( Ranunculus fluitans and Callitriche sp.) at concentrations of
13.7 and 14.5 mg/kg, respectively. Four weeks after application, the
results showed that the dichlorobiphenyl was metabolized more readily
after addition to water; the authors suggested the involvement of
aquatic bacteria. When applied in soil, 1.2% of the dichlorobiphenyl
was metabolized. This was contributed to the plant.

Moza et al. (1979b) grew 3-year-old spruce trees (Picea abies) in
soil containing ¹⁴C-labelled PCBs at approximately 4.2 mg/litre in
sewage sludge. When analysed 4 years later, only 0.8% (0.5% in
needles, 0.3% in stems) of the applied radioactivity was found in the
trees. Leaching of radioactive substances from the soil was less than
0.1% in the first 2 years and undetectable for the remainder of the
study.

In another study, Fries & Marrow (1981) grew soybean (Glycine max)
in pots, to determine residue contamination in plant tops from
¹⁴C-labelled 2,5,2'-trichlorobiphenyl, 2,5,2',5'-tetrachlorobiphenyl
or 2,4,5,2',5'-pentachlorobiphenyl, applied to the surface or
subsurface soil. Each compound was added to the soil at a rate of
2-3 mg/kg and the plants were harvested after a period of 52 days.
Detectable residues were only found in plants grown in surface-treated
soil, and concentrations in the plants increased with increasing
chlorination. Little of the labelled PCBs was lost from
subsurface-treated soil, but 20-30% of the surface-treated PCBs were
lost through volatilization. The authors concluded that the PCB
residues in the plant tops were, therefore, due to foliar
contamination from vapour rather than the uptake from the soil via the
roots. Miyazaki et al. (1975) came to the same conclusion when they
found no absorption or translocation of PCBs in sesame or rice seeds,
following the application of 4 types of Kanechlor (KC300, 400, 500,
and 600) at rates of between 0.1 and 100 mg/kg. But the rice straws
contained PCB levels of 0.02-0.08 mg/kg, which were the same as levels
found in plants from untreated soils.

Beets ( Beta vulgaris L.), turnips ( Brassica rapa L.), and beans
( Phaseolus vulgaris L.) were grown (Sawhney & Hankin, 1984) in soil
to which lake sediment contaminated with PCBs had been added. The
plants were exposed to Aroclor 1248 at a concentration of 80 µg/kg,
Aroclor 1254 at 1880 µg/kg, and Aroclor 1260 at 14 440 µg/kg. When
beets and turnips were grown in the soil for 6 months, the plants
showed greater uptake in the leaves than in the roots. For example,
beet roots contained 15, 16, and 35 µg/kg of Aroclors 1248, 1254, and
1260, respectively, while beet leaves contained 22, 94, and 52 µg/kg,
respectively. Total concentrations of the 3 Aroclors in beet roots and
leaves and in turnip roots and leaves were 66 and 168 µg/kg,
respectively, and 66 and 99 µg/kg, respectively. During a second
growing season, turnips and beans were grown for 6 months without any
additional PCB-contaminated sediment. Comparing PCB levels in turnips
between the 2 growing seasons showed a decrease in Aroclor 1248 uptake
relative to Aroclors 1254 and 1260. This was primarily because of a
large reduction in the amount of Aroclor 1248 in the soil after 1
year, due to degradation and volatilization. In beans, higher PCB
levels were found in the leaves and pods than in the stems and seeds.

Ten sludge application sites were selected within the Ontario area to
determine background heavy metal and PCB concentrations in the soils
and crops. Control sites (without sludge application) were adjacent to
the sludge application sites. Grab samples of liquid sludges applied
at each of the sites were taken for analysis. The soil samples were
taken at a depth of 15 cm. Twenty core samples were taken at 20-m
intervals and combined to form 1 sample. Eight of the application

sites were cropped with corn, one with oats, and one was left without
a crop. At the control sites, 7 were cropped with corn, 1 with oats,
and 2 left without a crop. PCB concentrations in the sludges ranged
from 0.13 to 1.61 mg/kg dry solids. PCB concentrations were in the
range of 0.007-0.025 mg/kg in the soil without sludges, and in the
range of 0.018-0.453 mg/kg air-dry weight in the soil with sludges.
The PCB levels in the crops were close to the control values (Webber
et al., 1983).

Bacci & Gaggi (1985) assessed the influence of translocation on the
concentrations of PCBs in the foliage of different plant species.
Beans, broad beans, tomatoes, and cucumbers were grown, either in soil
with a nominal added concentration of 500 mg/kg Fenclor 64 (similar to
Aroclor 1260), or in clean sand, for 28 days, enclosed in a glass box
with a constant turnover of air. The plants grown in clean sands were
exposed to PCBs by volatilization from other pots containing PCBs,
which were in the same growing box. The PCB peak pattern of both sand
and roots was similar to that of Fenclor 64, whereas the peak pattern
for foliage and air had moved towards lesser chlorinated congeners.
The concentrations of PCBs in the roots of tomatoes grown in
contaminated soil ranged from 105 to 168 mg/kg dry weight. But
translocation through the plants does not seem to be very likely since
there was no significant difference in foliage uptake of PCBs between
plants grown in contaminated soil and plants grown in clean soil.
Foliar uptake ranged from 13.8 to 42.6 mg PCB/kg (dry weight) for the
different species in PCB-fortified soil and from 11.8 to 47.1 mg/kg
for plants grown in clean soil.

4.2.4.3 Aquatic invertebrates

Bioconcentration factors are high for PCBs taken up by aquatic
invertebrates exposed to either pure chlorinated biphenyl isomers or
commercial mixtures in the water. Since PCBs are strongly bound to
sediments, this method of exposure is unrealistic. Addition of
sediment to test tanks decreases the uptake of PCBs, particularly by
organisms living in the upper water. However, there is clear evidence
that PCBs can also be readily absorbed into invertebrates from both
sediment and food. For organisms living on or in, sediment, uptake can
take place from the sediment, via food organisms that have absorbed
the PCBs, and from interstitial water or water immediately above the
sediment layer. A high content of organic matter in sediment decreases
the availability of PCBs for organisms. Uptake is rapid in most cases
and equilibrium is often reached in hours, though it may take weeks in
other examples. Uptake increases with increasing temperature. The
route of uptake is often via the gills, but varies among species. Loss
of PCBs is slow, but residues do decrease on cessation of exposure.
PCB uptake by aquatic invertebrates is transferred to predators and
can also be transferred to the terrestrial environment.

(a) Uptake from water

Vreeland (1974) exposed American oysters (Crassostrea virginica) to
various PCB isomers at concentrations of 5.5, 17, or 60 ng/litre
(which is within the range found in coastal waters) for 65 days.
Equilibrium was reached after approximately 1 month of exposure, with
concentration factors ranging from 1200 to 48 000 for PCB isomers with
2-6 chlorine atoms/molecule. The PCB concentration, after equilibrium
had been reached, was directly proportional to the amount of PCBs
added to the water. Lowe et al. (1972) found a linear pattern of
uptake in young American oysters exposed to Aroclor 1254 at 5 µg/litre
for 24 weeks, followed by a further 32 weeks in clean water. The
oysters already contained 17 mg/kg from a previous exposure and, by
the end of the 24-week exposure period, had accumulated 425 mg/kg (a
steady state was not established). By the end of the 32-week period in
clean water, no PCB residues could be detected. In another study on
uncontaminated young oysters, concentration factors of up to 101 000
were achieved after a 25-week exposure to 1 µg Aroclor 1254/litre.
After 12 weeks in clean water, whole-body residues were reduced to
0.2 mg/kg.

Courtney & Denton (1976) fed hard clams (Mercenaria mercenaria)
Aroclor 1254 adsorbed on the surface of alumina particles, at 1.25 and
12.5 µg/litre, for 21 days. The maximum concentration factor was 1800
for visceral mass, when the clams had been exposed to 1.25 µg/litre
for 18 days. The visceral mass accumulated a 1.4-5.3 times greater
concentration of PCBs per unit time than the muscular foot. Tissue
samples contained relatively more lower chlorinated isomers than the
Aroclor 1254 standard and, faeces and mud samples contained more
higher chlorinated isomers. Following exposure, clams from the lowest
dose group showed little change in PCB content after 3 months in clean
seawater. However, at the higher dose level, there was a significant
reduction in the PCB levels found in the foot after 1 month, but PCB
residues in the visceral mass remained unchanged for 6 months.

Pink shrimp (Penaeus duorarum) were exposed to Aroclor 1254 at a
concentration of 2.5 µg/litre, in flowing water, for 22 days (Nimmo et
al., 1971b). Accumulation was linear for the hepatopancreas and whole
body, but a plateau was reached after 2 days in muscle. Residues in
the hepatopancreas reached 510 mg/kg over the exposure period,
representing a concentration factor of 204 000; over the same period,
50% of the shrimps died. In a separate study, the shrimps were exposed
to 7.5 µg/litre for 16 days followed by an elimination period of 5
weeks in clean water. When calculated on the basis of the total tissue
burden of PCBs, an 80% reduction in the hepatopancreas was found,
concomitant with a doubling of the PCB levels in remaining tissues.
However, when data were presented as a concentration, a linear loss
from the hepatopancreas was seen, with the concentration in other
tissues remaining constant. The authors calculated a half-life for
loss of PCBs from the hepatopancreas of 17 days.

Nimmo et al. (1975) sampled shrimp from Pensacola estuary, USA, and
measured the relative concentration of PCBs in the various tissues.
The hepatopancreas contained the greatest amounts (50-75%) followed by
the ventral nerve. The authors studied the uptake of PCBs by pink
shrimp, experimentally, using various regimes with dosed food or dosed
water. They found the same tissue distribution in pink shrimp that had
been exposed to 0.2 µg Aroclor 1254/litre, in water, for 50 days. They
concluded that most of the PCBs were taken up directly from the water
in both the "wild" and laboratory situation. However, they did not
exclude the possibility of some PCBs being taken up from food, which
was found under some of the laboratory regimes.

To determine whether there was a concentration below which shrimps
would be unable to accumulate PCBs, grass shrimp (Palaemonetes pugio)
were exposed to flowing water concentrations of 0.04, 0.09, or
0.62 µg/litre. Whole-body residues of 0.2, 1.0, and 10 mg/kg,
respectively, were accumulated within 3-5 weeks. Even at the lowest
dose, shrimps accumulated more PCBs than the residues found in control
shrimp. Concentrations in the shrimp did not reach equilibrium during
the 5-week exposure, but the rate of accumulation decreased towards
the end of the exposure. When transferred to clean water, the shrimps
lost most of the PCBs within 4 weeks (Nimmo et al., 1975).

Gammarus oceanus were exposed by Wildish & Zitko (1971) to Aroclor
1254 concentrations of 2.5, 10, or 20 mg/litre for up to 6 h. Uptake
increased with increasing PCB concentration. Uptake decreased to half
of the initial rate after 4-6 h exposure at 20 mg/litre. There was
little or no uptake by dead animals. Although uptake was related to
branchial surface area, branchiae were not necessary sites of uptake,
since uptake could occur at an unchanged rate following branchial
removal. The authors did not find any change in the rate of uptake
during the intermoult stage.

Zhang et al. (1983) exposed Daphnia magna to ¹⁴C-labelled
2,2'-dichlorobiphenyl, 2,5,4'-trichlorobiphenyl, 2,4,6,2'-tetra-
chlorobiphenyl, or 2,4,6,2',4'-pentachlorobiphenyl at 50 µg/litre.
Equilibrium was reached after 20 h for all except the pentachloro-
biphenyl, which had not reached equilibrium within 24 h.
Bioaccumulation factors at equilibrium ranged from 3741, for the
dichlorobiphenyl, to 18 144, for the trichlorobiphenyl. Concentration
factors were not significantly related to the water solubility or
chlorine content of the biphenyl, but there was a tendency for the
bioaccumulation factor to increase with chlorine content and
decreasing water solubility. The authors studied the rate of
depuration and found it to increase with increasing water temperature
between 2 and 22°C. The rate of depuration was also faster for the
dichlorobiphenyl than for the pentachlorobiphenyl; after 48 h, the
amount of PCBs remaining in Daphnia was 22% and 77% (at 10-11°C) for
the 2 chlorobiphenyls, respectively.

(b) Uptake from sediment

Sediment was collected from the field and spiked with Phenochlor DP-5
to achieve a final PCB concentration of 0.65 mg/kg dry weight,
compared with 0.2 mg/kg in unspiked sediment (Elder et al., 1979).
Worms (Nereis diversicolor) were then added to aquaria containing
the sediment under flowing seawater. Equilibrium was reached within
40-60 days, by which time both groups had concentrated the PCBs by 3.5
times. Upon transfer from spiked to unspiked sediment, the worms took
2 months to attain body levels of PCBs comparable with those of the
unspiked group. A half-life of approximately 27 days was calculated
for incorporated PCBs.

Fowler et al. (1978) exposed Nereis diversicolor to spiked sediment
containing 9.3 or 80 mg Phenochlor DP-5/kg (dry weight), for 120 days,
compared with 0.11 mg PCB/kg in unspiked sediment. At the beginning of
the study, worms in the unspiked sediment had body residues of
0.59 mg/kg dry weight and reached a steady state at 1.2 mg/kg. Those
exposed to spiked sediment reached a steady state after a period of
approximately 2 months, with concentration factors ranging from 3 to
4. The worms maintained at the highest level of PCBs all died within a
90-day exposure period. When transferred to unspiked sediment for a
2-month period, the worms that had taken up PCBs from the unspiked
sediment lost PCBs exponentially. In a separate study, worms were
exposed to PCBs in water alone at a concentration of 0.57 µg/litre. A
steady state was reached much more quickly (2 weeks) than it was in
the presence of sediment, with a concentration factor of approximately
800. By comparing these results with field monitoring, the authors
calculated the relative importance of the 2 media. They stated that
approximately 99% of the PCBs in these studies was taken up from the
sediment. When the water overlying the spiked sediment was monitored,
28 ng PCBs/litre was measured (not leached, but a contaminant in the
experimental system) reducing the figure of uptake from sediment to
89%.

In a study by Courtney & Langston (1978), 1.1 mg Aroclor 1254/kg was
incorporated into intertidal sand. Specimens of 2 intertidal
polychaetes (Arenicola marina and Nereis diversicolor) containing
mean residues of 0.017 and 0.11 mg PCBs/kg (wet weight), respectively,
were collected. After 5 days in the spiked sediment, they contained
0.24 and 0.36 mg/kg, and, after a further 5 days, 0.39 and 0.49 mg/kg,
respectively. During a 3-week post-exposure period, there was no
significant loss of these PCB residues. The authors achieved
comparable PCB residues in these polychaetes after exposure to
1 µg/litre water or 1 mg/kg sediment.

McLeese et al. (1980) exposed the polychaete worm (Nereis virens)
and the shrimp (Crangon septemspinosa) to sediment containing
0.016-0.58 mg Aroclor 1254/kg (dry weight) for 32 days. Uptake was
found to be dependent on the exposure concentration and, in the case
of the worms, on the exposure period. The accumulation of PCBs was
inversely related to animal size; at 32 days, concentration factors
for worms ranged from 10.8 for 0.6-g worms to 3.8 for 4.7-g worms
following exposure to 0.17 mg PCB/kg. Factors of 3.5 and 1.9 were
found for shrimps weighing 0.1 and 2.9 g, respectively, after exposure
to 0.13 mg Aroclor/kg. Shrimps were found to accumulate, on average,
60% less PCBs than worms per unit weight. During the 26 days following
exposure, there was not any obvious loss of PCBs from the worms.

Rubinstein et al. (1983) collected sediments containing various levels
of pollutants (PCBs, 0.46-7.28 mg/kg dry weight; Cd; Hg) and organic
matter (5.5-22.3%). During a 100-day exposure period, only small
increases in PCB concentrations were detected in hard clam
(Mercenaria mercenaria) and grass shrimp (Palaemonetes pugio).
Higher concentrations of PCBs were accumulated by Nereis virens.
Uptake was found to be more dependent on the organic content of the
sediment than on the exposure concentration. Concentration factors
ranged from 1.59 in a low organic sediment to 0.15 in a high organic
sediment. The authors also calculated the maximum water exposure
concentration eluted from each of the sediments. On the basis of a
concentration factor of 800, calculated by Fowler et al. (1978) for
the uptake from water of Nereis sp., body residues of between 0.007
and 0.034 mg PCBs/kg (wet weight) would have been expected if
accumulation were dependent purely on direct partitioning from water.
However, whole-body residues of PCBs were found to be 0.4-0.63 mg/kg,
suggesting that pathways other than direct uptake from water (e.g.,
ingestion and sorption) contributed significantly to the accumulation
of PCBs by the polychaete.

Freshwater prawns (Macrobrachium rosenbergii) and clams (Corbicula
fluminea) were exposed to contaminated sediments for 48-50 days
(Tatem, 1986). Prawns were exposed to sediment containing
approximately 62 mg PCBs/kg (dry weight) and to the same sediment
diluted with sand to 50 and 10% of the original. Clams were exposed to
100, 50, or 10% of another sediment containing approximately 2 mg
PCBs/kg at 100%. The amount of PCBs accumulated was related to the
exposure concentration, with the highest concentration factors at the
lowest exposure (10%) level. Bioaccumulation factors for prawns ranged
from 0.1 to 0.9 for Aroclor 1242 and from 0.2 to 2.4 for Aroclor 1254,
relative to sediment concentrations. Exposed clams accumulated PCBs
(Aroclors 1242 and 1254) at concentration factors of 0.54-12.52,
relative to sediment. When tissues were analysed for Aroclor 1242 and
1254, maximum concentrations in prawns were attained at 7 and 40 days
for the 2 Aroclors, respectively. Exposure of prawns at 100 and 50%
dilution of sediment killed all the animals after 62 and 70 days,
respectively. Clams survived exposure.

Clark et al. (1986) investigated the accumulation of sediment-bound
PCBs by fiddler crabs (Uca pugilator) and (Uca minax). Mud and
mud/sand sediments were used; both were naturally contaminated with
PCBs and no further PCBs were added. Both species were exposed to a
mud sediment containing 1.04 mg PCBs/kg and to a mud/sand sediment
containing 0.37 mg/kg (dry weight). Concentration factors, after a
28-day exposure, were 0.19 and 0.79, for U. minax, and 0.2 and 0.59,
for U. pugilator, for the 2 sediments, respectively. In a second
study, using mud with 0.97 mg PCBs/kg and mud/sand with 0.55 mg
PCBs/kg, U. pugilator showed concentration factors of 0.58 and 0.71,
respectively, after 28 days. The authors did not find any detectable
PCBs in the overlying water, suggesting that the PCBs are tightly
bound to the sediment and leach out only very slowly. Following
transfer to uncontaminated sediment on day 42, no PCB residues were
detected in U. pugilator on day 56, or in U. minax on day 63.

Lynch & Johnson (1982) exposed the amphipod (Gammarus pseudolimnaeus)
to 2,4,5,2',4',5'-hexachlorobiphenyl added to sediment in flow-through
bioassays. Water overflowing from the tank containing the contaminated
sediment was directed into a second tank where further amphipods were
exposed without sediment. The hexachlorobiphenyl was labelled with
¹⁴C and added to the sediment at 1 mg/kg; the system was allowed to
equilibrate for 7-15 days prior to addition of amphipods, which were
sampled from the tanks after 24, 48, 96, and 192 h. In the initial
studies, the specific activity of the labelled hexachlorobiphenyl was
insufficient to detect the hexachlorobiphenyl concentrations in water.
However, it was clear that amphipods in the tank with the sediment
accumulated more hexachlorobiphenyl than animals exposed only to the
water overflow (8.8-10.5 times more PCBs). Removal of organic matter
from the sediment, by combustion, before addition of the PCB,
increased uptake of the hexachlorobiphenyl by increasing the
availability of the material to the Gammarus. In later studies,
specific activity was increased and water concentrations could be
measured. These were very low, ranging between 11 and 35 ng/litre in
the upper tank and 9 and 25 ng/litre in the lower tank. The lower end
of this range was found later in the exposure period suggesting that
less hexachlorobiphenyl was released over time. There was little
difference in concentration between water taken from the surface and
that sampled close to the sediment suggesting rapid mixing of the
overlying water. In this later series of studies, the authors
demonstrated that both the organic matter content of the sediment and
the presence of smaller particle sizes (silt and clay) reduced uptake
of hexachlorobiphenyl by the amphipods. Organic matter was the more
important factor. Adding maple leaves, to give about 70% organic
content in the sediment, reduced hexachlorobiphenyl uptake to between
10 and 20% of that in sediment without organic matter. Very high

bioconcentration factors were calculated relative to the very low
water concentrations of hexachlorobiphenyl (ranging between 27 000 and
1 000 000 in the upper tank and 2000 and 460 000 in the lower tank,
increasing with exposure period). These factors would be very low
relative to sediment concentrations of the PCB. However, it is clear
that the amphipod can accumulate hexachlorobiphenyl, leaching in very
small amounts from contaminated sediment.

Cores of lake sediment complete with overlying water were taken by
Larsson (1984) and transported back to the laboratory, still in the
sampling tube. PCBs were introduced at different dose levels by
injection through silicon septa in the walls of the tubes and spread
evenly 10 mm below the surface. The cores were allowed to stabilize in
the dark for 1 week at which time 80-100 chironomid larvae were
introduced. After 8 weeks, the systems were moved and kept at 20°C in
the light. After 2 days, the chironomid larvae began to pupate and
emerge. The study was terminated after 10 weeks. PCBs were measured in
sediment, larvae, adults, and exuviae (discarded skins after
emergence). Ranges of PCBs in sediment were between 0.5 and 14 mg/kg
giving rise to residues in larvae, exuviae, and adults directly
related to sediment concentrations. There was "biomagnification"
between larvae and adult. There was loss of body weight between the
final larval stage and the adult, but little loss of PCBs (only 17%
was retained in the exuviae). The author stated that the low variation
in uptake between animals is an indication of passive physicochemical
factors being involved in the handling of PCBs by chironomids. Active
uptake via ingestion would be expected to lead to more variation in
results. Meier & Rediske (1984) also monitored the uptake of PCBs from
contaminated sediment into chironomid larvae (Glyptotendipes
barbipes). Concentration factors for Aroclor 1242 from sediment
ranged between 20 and 130 for exposures of between 0.01 and 1.0 mg/kg,
considerably lower than concentration factors relative to water
(10 000 for these organisms) (Sanders & Chandler, 1972). Addition of
oil, commonly found in polluted areas where PCBs spills are likely,
reduced the uptake of PCBs from the sediment.

A detritus diet containing 17 µg Aroclor 1242/kg (wet weight) was fed
to male fiddler crabs (Uca pugnax) for 34 days (Marinucci & Bartha,
1982). The Spartina detritus was placed in the culture system at the
start of the study and, because of rapid depletion, was renewed after
19 days of exposure. Since PCBs leached continually from the food
source into the water, a second study was carried out to examine the
uptake of PCBs from water alone. Contaminated detritus was mixed
thoroughly with water and allowed to equilibrate for 24 h. Water
levels were found to be 14-15 µg/litre. Aroclor 1242 was accumulated
at a more rapid rate from PCB-laden detritus than from water alone.

The linear accumulation rate from litter was calculated to be 1 µg
PCBs/day per animal whereas, from water alone, the uptake was 0.1 µg
PCBs/day per animal. Aroclor 1242 was highly concentrated in the
hepatopancreatic tissue. It was found that the PCB residue in the
crabs was inversely related to their weight. Comparison of the
concentrations of PCBs in animals of the same weight shows that, at
the end of the 34 day exposure, those exposed to water alone had taken
up approximately half of the PCBs of those exposed to detritus. The
authors concluded that the crabs in the study accumulated a similar
amount of PCBs from both the food and the water.

Pinkney et al. (1985) exposed the amphipod Gammarus tigrinus to
Aroclor 1254 (¹⁴C-labelled) in fungus (Fusarium oxysporum) as a
food item. The fungus contained 195.8 mg Aroclor/kg dry weight.
Accumulation of PCBs was rapid, reaching a constant level in the
amphipods of 23 mg/kg after 9-24 h. Similar exposure of the amphipods,
but with exclusion from direct contact with the fungus by Teflon mesh
(to monitor uptake of PCBs leached into the water), resulted in
residues of between 0.16 and 3.3 mg/kg (from concentrations in the
water at 0.03 µg/litre), representing between 0.6 and 13.9% of uptake
from water and food combined. The PCB residues in the amphipods were
also monitored over 144 h on uncontaminated food to measure the
elimination rate. The water was changed every 24 h. Within this
period, 57% of the accumulated PCBs was eliminated.

(d) Comparison of different routes of uptake

In a study by Wyman & O'Connors (1980), the uptake by the marine
copepods Acartia tonsa and Acartia clausi of ¹⁴C-labelled Aroclor
1254 from water, inorganic sediment, and food, was monitored over a
period of 48 h. Acartia were exposed to water concentrations of
10 µg PCBs/litre. An asymptotic uptake curve was observed; equilibrium
was reached after 36 h, corresponding to whole-body residues of 248 mg
PCBs/kg (dry weight) for A. tonsa and 223 mg/kg for A. clausi.
During exposure, water concentrations fell rapidly to 5 or 6 µg/litre.
A similar pattern of uptake was found after exposure to sediment
contaminated with 20 mg PCBs/kg with maximum levels of PCBs in
A. tonsa of 22 mg/kg after 30 h. As in the water exposure, levels of
PCBs in sediment fell rapidly from 20 mg/kg to 14 mg/kg and then
slowly to 7 mg/kg at the end of the study. Water levels were initially
0.62 µg/litre and fell to 0.15 µg/litre. Uptake of PCBs by A. tonsa
from phytoplankton contaminated with 80 mg PCBs/kg (wet weight) was
very rapid and reached a maximum after 5 h at 61 mg/kg, but
subsequently declined after exhaustion of the food supply. PCB
concentrations in water were similar to those found when copepods were
exposed to contaminated sediment, copepods exposed to these water
concentrations alone accumulated significantly less PCBs than those
fed PCB-dosed phytoplankton.

McManus et al. (1983) exposed the marine copepod Acartia tonsa to
¹⁴C-Aroclor 1254 either in the food, as phytoplankton containing
approximately 1.3 mg PCBs, or in water at 1.5 µg/litre, for a period
of 30 h. For copepods exposed to contaminated phytoplankton, PCB
levels ranged from 117 to 163 mg/kg dry weight. For copepods exposed
to contaminated water alone, levels ranged from 82 to 104 mg/kg. When
transferred to clean water, the authors found that copepods lost PCBs
at a significantly faster rate if they were fed during depuration;
after 36 h, PCB concentrations in copepods fed during deputation were
10 mg/copepod whereas those starved contained 30 mg/copepod. No
significant difference in depuration rate was found between those
exposed via food and those exposed via water. In a second study,
elimination in males and females was compared. Although both sexes
contained similar residues at the start of depuration (117 mg/kg and
95 mg/kg, respectively), after 36 h, females contained significantly
lower levels of PCBs than males. During depuration, faecal pellets and
eggs were analysed; similar levels of PCBs were found in both male and
female faecal pellets during this period, but levels of PCBs more than
four times that in the females were found in eggs (407.5 mg/kg dry
weight after 4 h), indicating that egg production is an important
route for PCB elimination.

4.2.4.4 Fish

Fish of all life stages have been shown to take up PCBs readily from
water; bioconcentration factors are high. Time taken to reach
equilibrium is variable, but often long, in excess of 100 days. PCBs
with greater chlorination are more readily taken up and retained. PCB
body burden tends to increase with age and levels are higher in fish
with a greater lipid content. The accumulated PCBs are concentrated in
lipid-rich tissues. PCBs of lower chlorination are eliminated more
rapidly. Loss of PCBs is evident when exposure ends; an initial rapid
loss is followed by a slower rate of loss. Half-life estimates,
therefore, vary greatly, from a few weeks to several years.
Reproduction, with the production of a large mass of eggs or sperm,
allows loss of substantial amounts of the PCB residue. Depending on
the species, habitat, and behaviour, PCBs can be taken up from water,
sediment, or food to different degrees.

(a) Uptake from water

Califano et al. (1980) maintained larval striped bass (Morone
saxatilis) in Hudson river water (filtered and unfiltered)
contaminated with ¹⁴C-Aroclor 1254 at 1.36 µg/litre for a period of
48 h. Whole-body residues for filtered and unfiltered water were not
significantly different at 5 mg/kg and 5.9 mg/kg, respectively. Uptake
between 34 and 48 h was very slow, suggesting a steady state had
already been reached. Exposure of fish for a further 72 h in
unfiltered water, supported this theory. Elimination was slow, only
18% being lost in 48 h following a 24-h exposure.

The PCB uptake pattern in lake trout (Salvelinus namaycush) sac fry
was studied by Mac & Seelye (1981) by exposing them to a nominal
concentration of 50 ng Aroclor 1254/litre for 48 days. Patterns of
accumulation were similar, regardless of how the data were expressed
(wet weight, dry weight, or body burden). PCBs levels increased
slowly, reaching a peak after 32 days (just before completion of yolk
absorption), and then decreased by day 48.

Hansen et al. (1975) exposed different life-stages of sheepshead
minnow (Cyprinodon variegatus) to Aroclor 1016 (Table 10). After a
4-week exposure to nominal concentrations of 1, 3.2, or 10 µg/litre,
adult fish laid eggs containing on average 4.2, 17, and 66 mg/kg,
respectively. DeFoe et al. (1978) exposed fathead minnow (Pimephales
promelas) to Aroclor 1248 or 1260 at concentrations of
0.1-3 µg/litre, for 240 days (life cycle). Bioconcentration factors for
the uptake of PCBs were independent of the PCB concentration in the
water. Residues in the fish reached an apparent steady state within
about 100 days of exposure and growth. Females accumulated about twice
as much PCBs as males, because of their higher body lipid content. The
variability of residues in females reflected the variability of their
lipid content. Although mechanisms for uptake were similar for both
Aroclors, greater body burdens were always achieved with exposure to
Aroclor 1260. Bioconcentration factors ranged from 60 000 to 160 000
for males and from 120 000 to 270 000 for females. After transfer to
clean water, 18% of Aroclor 1248 was lost within 28 days and 15% of
Aroclor 1260 in 42 days. The authors stated that, because of
variations between fish, this 10-20% decline in total body burden of
PCBs was insufficient to indicate definite PCB elimination over this
period.

De Kock & Lord (1988) exposed an estuarine fish, the Cape stumpnose
(Rhabdosargus holubi) to a flowing water concentration of 1 µg
Aroclor 1260/litre for 90 days followed by a 90-day period in clean
water. Equilibrium was reached at 90 days with a concentration factor
of 24 000. The depuration rate was calculated to be 0.014 days,
producing a half-life of 50 days.

Goldfish (Carassius auratus) were exposed to Clophen A50 at levels
of 0.01, 0.05, 0.1, or 0.5 mg/litre for 18 days (Hattula & Karlog,
1973). Rapid uptake was observed with concentration factors of over
1000 at 18 days, but equilibrium was not achieved within this period.
Nearly all the fish exposed to 0.5 mg/litre died within 7 days. After
transfer to clean water, fish that had been exposed to 0.1 mg/litre
for 13 days and had attained body residues of 70 mg/kg lost half of
the PCBs within 3 weeks, but still retained levels of approximately
15 mg/kg, after 70 days.

Yoshida et al. (1973) exposed carp (Cyprinus carpio) to ¹⁴C-PCBs
(equivalent to Aroclor 1254) in water or in food. By measuring the
radioactivity, they found similar tissue patterns of uptake from both
water and diet. PCBs were localized in the gall bladder, adipose
tissue, and hepatopancreas and, in particular, the adipose tissue of
the skull.

Hansen et al. (1971) exposed spot (Leiostomus xanthurus) to Aroclor
1254 at 1 µg/litre, for 56 days. Maximum tissue levels of PCBs were
achieved between days 14 and 28. Highest levels were found in the
liver (210 mg/kg, after 28 days) followed by the gills, whole fish,
heart, brain, and muscle. Aroclor 1254 was slowly lost from tissues;
after 84 days in clean water, levels of PCBs had dropped by 73%.

In a study by Braun & Meyhofer (1977), rainbow trout (Salmo
gairdneri) fingerlings were exposed to water concentrations of 2 or
20 µg Clophen C/litre, for 8 weeks. Tissue PCB concentrations for
gills, muscle, and liver were found to be 0.62, 0.82, and 3.47 mg/kg,
respectively, for the lower dose and 12.3, 7.6, and 10.6 mg/kg, for
the higher dose. When fish were held in clean water for 10 weeks,
following exposure to 2 µg/litre for 8 weeks, residues decreased by
half in the liver and had disappeared completely from the gills, but
there was no change in the PCB levels in muscle.

Rainbow trout (Salmo gairdneri) were exposed by Guiney et al. (1977)
to ¹⁴C-labelled 2,5,2',5'-tetrachlorobiphenyl at 0.5 mg/litre for
36 h. The tissue distribution of ¹⁴C was measured at regular
intervals after transfer to clean water. Carcase, muscle, skin, lower
gastrointestinal tract, and fat contained most of the radioactivity
(88%). During the first 14 days after exposure, radioactivity
increased in adipose tissue, carcase, and eyes. Elimination from most
tissues appeared to be biphasic with a 30% loss within 2 weeks
followed by a loss of only 6% in the following 126 days. Losses from
the bile and blood were very rapid and nearly complete within 14 days.
Based on the initial rate of loss, the authors calculated a half-life
of 1.55 days, however, the second phase of eliminated PCBs suggested a
half-life at 2.66 years. In a similar study, Guiney et al. (1979)
calculated half-lives of 1.76 and 1.43 years for female and male
rainbow trout, respectively, based on fish sampled 2-34 weeks after
exposure. For both sexes the half-life of elimination was recalculated
to 0.52 and 0.54 years between weeks 38 and 52 after exposure (the
spawning season). The increased elimination appeared to be because of
loss via eggs and sperm. Vodicnik & Peterson (1985) found a similar
result after dosing yellow perch (Perca flavescens); an elimination
half-life of 22 weeks was calculated. This was later recalculated to
be <0.7 weeks during spawning, returning to 16.3 weeks after the
completion of spawning.

(b) Uptake from sediment

The uptake of Aroclor 1254 from suspended solids by juvenile Atlantic
salmon (Salmo salar) was studied by Zitko (1974). Aroclor 1254 was
mixed with suspended solids (simulated by SilicAR CC7) in hexane at
5 mg/ml. Fish were exposed to contaminated solids at 1 g/litre for up
to 144 days. Over this exposure period, the salmon accumulated 134 mg
Aroclor 1254/kg.

Stein et al. (1984) exposed English sole (Parophrys vetulus) to a
sediment concentration of 1 mg ¹⁴C-Aroclor 1254/kg (dry weight).
Seawater was allowed to flow over the sediment for 6 days before the
fish were added. A steady state of PCBs accumulated in the tissues of
the fish was achieved after 10 days of exposure. Highest residue
concentrations were found in the bile and the liver. Concentration
factors were 10 for the bile and 4 for the liver, with other tissues
individually concentrating PCBs by factors of 3 or less. Simultaneous
exposure of sole to PCBs and ³H-benzo[ a]pyrene (3 mg/kg, dry
weight) reduced the amount of PCBs accumulated. Stein et al. (1987)
collected urban sediment containing aromatic hydrocarbons and PCBs at
32 mg/kg and 2.2 mg/kg dry weight, respectively. English sole
accumulated hepatic concentrations of 1.4 mg PCBs/kg (wet weight) over
a period of 108 days exposure to the urban sediment. This was 8 times
the PCBs accumulated by sole exposed to the control sediment, which
did not contain any detectable PCBs. In another study, the same
authors added a ¹⁴C-labelled PCBs tracer to the urban sediment. The
concentration of PCB-derived radioactivity in the liver reached a
steady state after 14 days of exposure; the steady state concentration
in the carcase was found to be significantly lower.

Lieb et al. (1974) fed rainbow trout Salmo gairdneri on a diet
containing 15 mg Aroclor 1254/kg for 16 or 32 weeks. PCB levels in the
lipid fraction increased rapidly for the first 8 weeks, reaching
equilibrium at about 95 mg/kg. The absolute quantity of PCBs continued
to increase as the fish grew. The trout had retained 68% of the total
PCBs ingested at equilibrium. No elimination was found after transfer
to uncontaminated food at 16 weeks (for a period of 16 weeks), or
after starving the fish for 8 weeks following exposure for 32 weeks.
Reductions in PCB levels were found, but these were cancelled out by
concomitant reductions in lipid content.

Coho salmon (Oncorhynchus kisutch) parr were fed 10 mg chloro-
biphenyls/kg (containing equal parts by weight of 3,4,3',4'-tetra-
chlorobiphenyl, 2,4,5,2',4',5'-hexachlorobiphenyl, and
2,4,6,2',4',6'-hexachlorobiphenyl) for up to 165 days (Gruger et al.,
1975). Most of the PCBs were accumulated in the adipose tissue of the
salmon (51.1 mg/kg total chlorobiphenyls after 165 days). Tissue

levels of tetrachlorobiphenyl were about half those of either of the
two hexachlorobiphenyls throughout the exposure period. When fish were
starved for 48 days, the data indicate mobilization or transformation,
with, for example, chlorobiphenyls in the spleens lowered by half and
in adipose tissue increased 5-fold. Most tissues showed an increase in
PCB levels, especially blood levels. In contrast, when a second group
of salmon were fed on a clean diet, chlorobiphenyls were released from
adipose tissue and levels increased in some other tissues, such as the
lateral line dark muscle tissue. The ratio of the different
chlorobiphenyls remained unchanged during both of these post-exposure
treatments. Gruger et al. (1976) fed juvenile coho salmon diets
containing a mixture of 2,5,2',5'-tetrachlorobiphenyl, 2,4,5,2',5'-
pentachlorobiphenyl, and 2,4,5,2',4',5'-hexachlorobiphenyl
at 1, 2, and 12 mg/kg, for up to 72 days. A steady state appeared to
have been reached between 17 and 35 days at the lowest dose (a whole
body concentration of approximately 0.45 µg/kg (wet weight)); steady
state was not achieved at the other 2 dose levels. All 3
chlorobiphenyls were accumulated to similar levels. Comparing these
data with the study by Gruger et al. (1975), suggests that the
position of the chlorine substitution is an important factor.

Hansen et al. (1976a) fed channel catfish (Ictalurus punctatus) on a
diet contaminated with 20 mg Aroclor 1242/kg. The total burden of PCBs
(µg PCB/fish) increased exponentially with exposure time. When fish
were placed on a clean diet (from day 84 for 56 days) a slight net
decrease in body burden was observed, but levels remained constant
when fish were placed on a clean diet for 56 days after 140 days
exposure. On return to a PCB-contaminated diet, accumulation rates
returned to those previously observed. The authors noted that, during
PCB-free periods, there was a shift in residues from edible carcase to
offal.

Mayer et al. (1977) fed fingerling coho salmon with Aroclor 1254 at
concentrations ranging between 1.45 and 14 500 µg/kg body weight.
Equilibrium was reached after 112 days at concentrations of 1.45,
14.5, and 145 µg/kg, with whole body residues of 0.47, 0.5, and
3.8 mg/kg, respectively. A steady state was reached at the 2 highest
dose levels of 1450 and 14 500 µg/kg after 200 days, with
corresponding residues of 57 and 659 mg/kg. In another study, channel
catfish (Ictalurus punctatus) were exposed to Aroclors 1232, 1248,
1254, and 1260 in the diet at concentrations of 48 or 480 µg/kg body
weight, for 193 days. Equilibrium was only achieved at the lowest
exposure dose of Aroclor 1232, within 150 days, with a whole-body
burden of 4.5 mg/kg. Similar whole-body residues were achieved at the
lowest dose of the other Aroclors, but no steady state was reached. At
the higher dose, accumulation increased in the order Aroclor 1232 =
1248 < 1254 < 1260, with residues ranging from 13 to 32 mg/kg after
193 days.

When Zitko (1974) fed juvenile Atlantic salmon (Salmo salar) diets
containing 10 or 100 mg Aroclor 1254/kg, accumulation reached
equilibrium within 30 days at the lower dose, with a whole-body
residue of approximately 3.8 mg/kg. Equilibrium was not reached within
200 days at 100 mg PCBs/kg. A whole-body residue of 30 mg/kg was
recorded at 181 days.

Zinck & Addison (1974) administered a mixture of 2-, 3-, and
4-chlorobiphenyl to thorny skate (Raja radiata) and winter skate
(Raja ocellata) by intravenous injection. All three congeners were
cleared rapidly from blood plasma, 3-chlorobiphenyl consistently being
cleared more rapidly than the other two. Less than 6% of
3-chlorobiphenyl remained in the plasma after 15 min compared with 30%
for the other chlorobiphenyls. All three accumulated in the other
tissues of R. radiata, principally in the liver and muscle. Tissue
levels of 3-chlorobiphenyl were consistently less than the others
during the 53-h sampling period.

In a study by Guiney & Peterson (1980), both yellow perch (a non-fatty
fish) and rainbow trout (a fatty fish) were dosed with 0.8 µg of
¹⁴C-labelled 2,5,2',5'-tetrachlorobiphenyl, either orally or by
intraperitoneal injection. Whole-body elimination was found to be
similar for both species and routes. A 20-30% elimination was observed
after 3-4 days with virtually no more PCBs being eliminated during the
rest of the 32-day sampling period. Tissue distribution varied between
the 2 species; uptake in the perch was mainly concentrated in the
viscera and carcase, whereas, in the trout, skeletal muscle and
carcase were the major sites of uptake.

Niimi & Oliver (1983) calculated the biological half-life of 31
dichloro- to decachlorobiphenyl congeners, 105 days after a single
oral dose of 46-261 mg/kg was administered to rainbow trout (Salmo
gairdneri). Whole-body half-lives increased from 5 days to >1000
days as the number of chlorines on the biphenyl increased. From
structure-activity analysis of half-lives in whole fish, the authors
concluded that elimination was enhanced for congeners with a lower
chlorine content and no chlorine substitutions in the ortho
positions, and for those with 2 unsubstituted carbons adjacent on the
biphenyl.

4.2.4.5 Birds

PCBs are taken up from contaminated food or water and concentrated in
the fatty tissues of birds. PCBs of higher chlorination levels are
accumulated to a greater extent. Egg-laying females can lose
substantial amounts of PCBs from body tissues by transferring the PCBs
to the eggs. Redistribution of residues occurs on starvation (of

significance during the migration of birds in the wild). Expressed as
a whole-body concentration, PCB residues fall during starvation.
However, expressed as a concentration in fat, residues rise. Most
critically, PCB residues in the brain increase during starvation and
this may kill the birds without further intake of PCBs.

Brunström et al. (1982a) injected the yolk of developing hens' eggs,
on day 4 of incubation, with ¹⁴C-labelled 2,4,2',5'-tetra-
chlorobiphenyl at a concentration of 5 mg/kg. One hour after
injection, radioactivity was found in the sub-blastodermic fluid, the
highest concentrations being in amniotic membranes. None was present
in the yolk, albumen, or embryonic tissues. Uptake was uniform
throughout the embryo, after one day, and, as tissues developed,
became concentrated in certain of them, such as the liver, kidney, and
fluid brain vesicles, by day 7. ¹⁴C was found uniformly in the yolk
after 11-14 days and was highly concentrated in the first bile
produced on day 11. The labelled PCBs accumulated in fatty tissue as
it developed from day 14 onwards. In the hatched chick, large amounts
of radioactivity were found to be concentrated in the gall bladder,
intestine, cloaca, and the coiling of the gizzard. When either
3,4,3',4'-tetrachlorobiphenyl or 2,4,2',5'-tetrachlorobiphenyl was
injected into the air sac of hens' eggs on day 14 of incubation at
0.4 mg/kg, no difference in distribution pattern was observed 1-5 days
later (Brunström & Darnerud, 1983). The highest amounts of
radioactivity were found in the fatty tissue, liver, kidneys, and the
gall bladder, ¹⁴C was also found in the bone marrow, the adrenals,
and the gonads, but to a lesser extent. The yolk contained less
radioactivity than the yolk analysed in the previous study by
Brunström et al. (1982a), because the PCBs were administered via the
air sac.

White leghorn hens were exposed to 50 mg Aroclor 1254/litre in their
water for 6 weeks (Tumasonis et al., 1973). PCB residues in the yolks
of eggs laid increased during the exposure period to a peak, after 6
weeks, of approximately 205 mg/kg. When hens were given clean water,
the yolk levels of PCBs quickly dropped within 5 weeks to
approximately 100 mg/kg, and then more slowly until, after 20 weeks
without Aroclor 1254 in their water, the hens laid eggs containing
0.7 mg/kg.

During a 4-week exposure to Aroclor 1242, 1254, or 1260, in the feed
of one-day-old chicks, Harris & Rose (1972) found that PCBs
accumulated in the fat and that this accumulation increased with
increasing exposure concentrations of 100, 200, and 400 mg/kg. At the
2 highest dose levels, the hens accumulated more of Aroclor 1260 than
of the other 2 Aroclors (i.e., 482, 1427, and 2151 mg Aroclor 1260/kg
at the 3 exposure concentrations, respectively). At the highest dose,
there was high mortality during exposure to Aroclor 1242 and 1254 and
this might have affected the residues found.

Greichus et al. (1975) fed white pelicans (Pelecanus erythrorhynchos)
on a fish diet containing 100 mg Aroclor 1254/day, for 10 weeks. PCB
residues were measured in the carcase, liver, feathers, and brain;
mean residues found were 2130, 290, 120, and 110 mg/kg wet weight,
respectively.

In a study by Dahlgren et al. (1972), 11-week-old pheasant (Phasianus
colchicus) were dosed with one capsule per day containing 210 mg of
Aroclor 1254. Birds that died between days 1 and 5 contained, on
average, PCB residues of 520 mg/kg in the brain, 2500 mg/kg in the
liver, and 140 mg/kg in muscle. Birds that were sacrificed over the
same period had mean brain, liver, and muscle PCB levels of 370, 1900,
and 83 mg/kg, respectively. All birds dosed with only 10 mg of Aroclor
1254 per day died within 180 days and contained average brain and
liver residues of 360 and 1200 mg/kg, respectively.

Södergren & Ulfstrand (1972) fed robins (Erithacus rubecula)
mealworms containing 1 µg of Clophen A50/day for 15 days. Brain,
breast muscle, and carcase were analysed and contained mean PCB
residues of 0.35, 0.55, and 4.5 mg/kg fresh weight, respectively. A
second group of robins was starved following dosing and all died
within 48 h. PCB levels were higher in the brain and breast muscle at
1.1 and 1.3 mg/kg, respectively, but carcase PCB levels were lower on
a fresh weight basis at 2.6 mg/kg. When the carcase lost some of its
fat content during starvation, PCB levels in terms of fresh weight
decreased. Consequently, because of the low remaining fat content,
residue levels in terms of fat weight increased. Another group of
birds were fed both PCBs and DDT (10.5 µg/day) for 15 days and then
starved. PCB levels in all 3 tissues analysed were higher than those
in birds administered PCBs alone followed by starvation; residues
were: brain, 9.3 mg/kg fresh weight, breast muscle, 8.8 mg/kg, and
carcase, 4.5 mg/kg.

Cormorants (Phalacrocorax carbosinensis) were kept on a fish diet
contaminated with PCBs for one month, followed by gelatin capsules of
PCBs administered daily for the remainder of the exposure (Koeman et
al., 1973). After 14 weeks, the dose rate of Clophen A60 was increased
periodically during the exposure period from 200 to 500 mg/kg. The
birds died between days 55 and 124, and overall residues of PCBs
increased in the tissues, the longer the birds survived. Total-body
residues ranged from 850 to 2750 mg PCBs/kg (wet weight) at death.
Brain and liver residues ranged from 76 to 180 mg/kg and from 210 to
290 mg/kg, respectively. The fat of 2 birds was analysed for PCBs and
was found to contain 10 300 and 20 500 mg/kg.

Harris & Osborn (1981) dosed wild puffins (Fratercula arctica) by
implantation with 30-35 mg of Aroclor 1254. PCBs were quickly taken up
in fat, with concentrations rising to 10-14 times that in control
birds (highest fat residue 654 mg/kg wet weight), and remaining at
this level for up to 10 months. Levels slowly declined, but were still
twice those of controls after 34 months. PCB concentrations in the
liver and muscle tissue were highest shortly after dosing (48.4 and
25.2 mg/kg, respectively) and declined until, after 16 months, no PCBs
were detectable. Levels of PCBs in the kidneys and brain were variable
with no consistent trends.

Common grackles (Quiscalus quiscula), starlings (Sturnus vulgaris),
red-winged blackbirds (Agelaius phoeniceus), and brown-headed
cowbirds (Molothrus ater), were fed diets containing 1500 mg Aroclor
1254/kg over an 8-day period (Stickel et al., 1984). PCB residues in
the brains of birds that died were found to be higher than those in
birds that were sacrificed over a similar period. PCB residues ranged
from 349 to 763 mg/kg in birds that died and from 54 to 301 mg/kg in
birds sacrificed. Liver and whole-body residues tended to be higher in
birds that died, but they overlapped to a large extent. PCB residues
in whole bodies on a lipid basis showed the most clear-cut difference,
ranging from 22 600 to 98 600 mg/kg for birds that died and from 6690
to 22 500 mg/kg for those sacrificed. PCB residues in grackles
declined slowly, when the birds were placed on a clean diet. From a
whole-body level of 1300 mg/kg, residues declined to 169 mg/kg, 224
days later. The rate of decline was irregular, but a half-life was
estimated at 89 days over this period of loss.

4.2.4.6 Mammals

Olsson et al. (1979) fed mink (Mustela vison) on a diet containing
11 mg PCBs/kg for 66 days. Mink accumulated 310 mg PCBs/kg in
extractable fat over the exposure period. Control mink were found to
contain 14 mg PCBs/kg, and, when the control feed was analysed, it was
found to contain 0.05 mg PCBs/kg. The authors also found a significant
increase in cadmium uptake in the kidneys of PCB-treated animals
compared with controls. In another study on mink (Mustela vison),
Hornshaw et al. (1983) administered various PCB-contaminated fish
diets containing between 0.21 and 1.5 mg PCBs/kg. Adipose tissue
samples were taken after 6-8 weeks and after 18 weeks exposure (Table
11). The amount of PCBs accumulated was directly related to the amount
of PCBs in the diet; mean PCB residues ranging from 4 to 24.8 mg/kg
after 6-8 weeks and from 8.1 to 42.8 mg/kg after 18 weeks. When
expressed as individual congeners, it can be seen that the mink showed
the highest accumulation of the PCBs with the chromatographic peak
corresponding to 2,4,5,2',4',5'-hexachlorobiphenyl. To determine the
rate of PCB elimination, male mink that had been on a fish diet

containing 1.5 mg PCBs/kg for 10 weeks were transferred to a control
diet. Over this period, adipose tissue residues of 32 mg PCBs/kg had
accumulated. Over the 16-week elimination period, 60.3% of the total
PCB burden of the adipose tissue was eliminated. This consisted of a
loss of 87.2% of 2,5,2',5'-tetrachlorobiphenyl, 88.9% of
2,3,6,2',5'-pentachlorobiphenyl, and 55.4% of the hexachlorobiphenyl.
The half-life for total PCBs in mink adipose tissue was calculated to
be 98 days.

Wren et al. (1987a,b) fed mink on a commercial mink food supplemented
with 1 mg Aroclor 1254/kg for a period of 6 months. Male mink had
liver residues of 1.98 mg PCBs/kg after 118 days and 2.8 mg/kg after
183 days exposure. The liver of a female, analysed on day 161
contained a residue of 3.1 mg PCBs/kg. Liver PCB levels in 5-week-old
kits were similar to those in adult mink fed the experimental diet for
several months. Bleavins et al. (1981) measured the relative
importance of placental transfer and milk in the transfer of PCB
residues from mother mink to offspring. Newborn kits contained less
than 0.1% of a dose of PCBs injected into the mother mink. At 2 weeks
of age, the kits contained 1.2% of the dose given to the mother,
suggesting that lactation is a major route of exposing the young to
PCBs and a major route for the loss of PCBs from the mother. Placental
transfer of PCBs was greater in the ferret than in the mink (Bleavins
et al., 1984). The ratio of placental to mammary transfer was 1:15 for
offspring whose mothers were dosed during the first trimester of
pregnancy and 1:7 for mothers exposed during the last trimester.

Big brown bats (Eptesicus fuscus) were fed on mealworm diets
containing 9.4 mg Aroclor 1254/kg for up to 37 days (Clark & Prouty,
1977). In bats sacrificed on day 37, residues ranged from 29 to 121 mg
PCBs/kg (wet weight) for the carcase and from not detectable to
4.2 mg/kg in the brain. Bats that were starved following exposure
showed a significant correlation between increasing brain PCB
concentrations and carcase lipid concentrations. The authors stated
that PCBs increased in brain tissue as carcase fat was metabolized.
Clark (1978) exposed pregnant big brown bats to a mealworm diet
containing 6.36 mg Aroclor 1260/kg for approximately 18-28 days, until
the young were born. Mean carcase levels of PCBs were 20.34 mg/kg in
parent females and 4.38 mg/kg in litters. Levels of PCBs in both
adults and young continued to rise throughout the sampling period; the
longer the gestation time, the higher the PCB level in the sample.

4.2.5 Appraisal

Experimental work on mammals has been concentrated on terrestrial
species. Problems with PCB toxicity are important for marine mammals,
but these are less convenient for experimental study. Results in this
section, therefore, have to be related to field observations on marine
species.

Mink take up more chlorinated components of PCB mixtures and can
accumulate large residues of PCBs. On cessation of exposure, more
tetrachloro- and pentachlorobiphenyls were eliminated than
hexachlorobiphenyl. The half-life for total PCBs was calculated to be
98 days. PCB residues are transferred from mother to offspring. The
relative importance of transplacental transfer and transfer in milk
varies between species. Redistribution of residues takes place on
starvation, which is of significance for migratory species; brain
residues, which may be fatal with no further intake of PCBs, increase
as animals are starved.

5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

5.1 Levels in the environment

PCBs were detected in the environment in the late 1960s (Risebrough et
al., 1968; Jensen et al., 1969) and, within a short time, were
reported as contaminants in almost every component of the global
ecosystem including air, water, soil, fish, wildlife, human blood,
adipose tissues, and milk (Holdrinet et al., 1977; Wassermann et al.,
1979; Ballschmitter et al., 1981; Buckley, 1982; Safe, 1982b; Bush et
al., 1985; Kannan et al., 1988; Tanabe, 1988).

The lipophilic properties of PCBs are the basis of the bioaccumulation
and biomagnification that has been demonstrated and, thus, numerous
sources within the environment can lead to human exposure.

High-resolution, gas-chromatographic analysis has shown that the
congener composition and relative concentrations of the individual
components in many PCB extracts from environmental samples differ
markedly from those in the commercial PCBs (Jensen & Sundström, 1974a;
Wolff et al., 1982a; Safe et al., 1985a; Brown et al., 1987a,b).

A major problem with data concerning PCB levels in environmental
samples is that they normally are only available for "total PCBs" and
that there are much fewer data on actual "PCB patterns". Moreover,
when comparing results produced from different laboratories or from
the same laboratory at different times, an additional difficulty may
arise from differences in the sampling and analytical techniques used.

It is difficult, if not impossible, to compare data obtained with
different analytical methods, from different laboratories, and
countries. Nowadays, the older data seem less reliable, especially in
the light of the use of improved analytical methods and better
sampling techniques (WHO/EURO, 1987). A comprehensive review of world
PCB levels was published by Wassermann et al. (1979).

5.1.1 Air

PCB concentrations in air differ markedly from location to location,
with the lower levels found over the oceans or over non-industrialized
regions, such as the Canadian Northwest territories. In general,
levels over industrialized areas or over landfills are the highest.
Apparently, these levels influence PCB levels in rainwater and there
is a gradient of values in air from industrial to rural areas. Some
typical values can be found in Table 12.

MacLeod (1981) described a method for the analysis of PCBs using
low-volume, indoor air sampling to estimate the presence of PCBs in
indoor air in work-places and homes in the USA. Three facilities, an
industrial research facility, an academic facility, and a shopping
complex were sampled. The periods of sampling ranged from 2 days up to
6 months. The average concentrations (calculated as Aroclor 1242 plus
Aroclor 1254) ranged from 44 up to 240 ng/m³. Outdoor levels of up to
18 ng/mg³ were found. In the homes, air samples from 14 areas (of
which 9 were kitchens) were also analysed. The average concentrations
in the kitchens ranged from 150 up to 500 ng/m³ and, in the other
rooms, from 39 to 170 ng/m³. In a library, a level of 400 ng/m³ was
found.

The levels of PCB exposure that may occur in buildings in the USA were
determined by Oatman & Roy (1986). Air samples and surface wipe
samples were taken in 5 state-owned, office buildings and 2 elementary
schools. The average levels of airborne PCBs in buildings with PCB
transformers were nearly twice the levels in buildings without
transformers, i.e., 457 ± 223 and 229 ± 106 ng/m³, respectively. The
mean of the surface wipes taken in buildings without PCB transformers
was 0.17 and that in the buildings with transformers 0.23 µg/100 cm².
There was a wide variation between the different buildings and, as
shown above, the presence of transformers influenced the indoor PCB
concentrations.

5.1.1.1 Rain and snow

In the Netherlands, at Bilthoven, the PCB-concentrations in rainwater
ranged from 0.01 to 1.5 µg/litre (van Zorge, cf. WHO/EURO, 1985). In
the Federal Republic of Germany, concentrations of 0-4 ng/litre were
found (DFG, 1988).

Table 12. PCB levels in air in several countries

Country Location and/or type of sample PCB levels References
average and/or range

Canada Northwestern territories 0.002-0.07 ng/m³ Bidleman et al. (1978)

Germany Industrial area (Ruhr area) 3.3 ng/m³ DFG (1988)
Non-contaminated area 0.003 ng/m³

Japan Within industrial plants: - PCB vapours 13-540 µg/m³ Tatsukawa & Watanabe
- PCBs on airborne particulates 4-650 µg/m³ (1972)

North Pacific, South Pacific, Indian, 0.1-0.3 ng/m³ Tatsukawa & Tanabe
Antarctic and South Atlantic Oceans (1983)

North Atlantic Ocean 0.5 ng/m³ Tatsukawa & Tanabe
(1983)

Sweden Several locations 0.8^a-3.9 ng/m³ Ekstedt & Odén (1974)

USA Near the North-East Coast 5 ng/m³ Harvey & Steinhauer
Over the Atlantic Ocean, 2000 km away 0.05 ng/m³ (1974)
from the industrial complex

several locations 1-50 ng/m³ Panel on Hazardous
Substances (1972)
cf WHO/EURO (1988)

Yugoslavia Bela Krajina: - 300 m from an industrial plant 4-7 µg/m³ Jan et al. (1988b)
- air near a waste landfill 45 µg/m³
- over the River Kruga 2-5 µg/m³

^a Limit of determination.

5.1.1.2 Natural gas

PCBs were first identified in gas pipelines in January 1981, when a
PCB-containing oil condensate was found in the gas meters of some
residential customers in Long Island, New York. Voluntary monitoring
of condensate and natural gas by 33 transmission companies, showed the
presence of PCBs in 12 companies. PCBs were also found in gas
pipelines. Condensate is a mixture of heavier hydrocarbons and other
liquids, such as water, that condenses, because the gas is transmitted
under pressure. This condensate tends to collect in pools in the
pipes. In the period 1981-83, 1841 samples of condensate from gas
pipelines were analysed: 659 (35.8%) of the samples contained
< 25 mg/kg; 65.8% of the samples contained <1000 mg/kg, and 0.4%,
> 10 000 mg/kg. The maximum level that was found was 42 394 mg/kg
(Versar Inc., 1984).

In the period 1981-83, 138 samples of natural gas in transmission were
analysed. In 29 samples, PCBs were found with a minimum concentration
of <0.004 µg/m³ and a maximum concentration of 1050 µg/m³. Natural
gas in distribution lines was also analysed in the same period. Out of
528 samples, 224 did not contain any PCBs. The levels ranged from
<0.02 to 51 µg/m³.

Indoor concentrations (kitchens, etc.) were measured in 419 samples in
the period 1981-83. No PCBs could be detected in 49 samples, but, in
the others, levels ranged from <0.01 to 1.08 µg/m³ (Versar Inc.,
1984).

5.1.2 Water

Surface water may become contaminated with PCBs from atmospheric
fall-out or from direct emissions from point sources. Because of
adsorption on suspended particles, PCB concentrations in heavily
contaminated waters may be several times greater than their
solubility. Södergren (1973) reported a seasonal variation, which was
attributed to aerial fall-out.

It has been shown that polluted rivers, lakes, and estuaries have
higher PCB values than non-polluted waters (Table 13). On the basis of
scanty information on PCBs and reinforced by extensive analogue
information on DDT, it has been estimated that, for the Great Lakes of
North America, non-polluted freshwaters might contain less than
5 ng/litre, moderately polluted rivers and estuaries, 50 ng/litre, and
highly polluted rivers, 500 ng/litre. These values can be used to
evaluate those reported by several authors and presented in Table 13.

Table 13. PCB levels in water in several countries

Country Location and/or type of PCB levels References
sample average and/or range

Germany Several rivers 5-103 ng/litre Lorenz & Neumeier (1983)

Netherlands River Rhine (1976/1977) 100-500 ng/litre Wegman & Greve (1980)

Sweden Water entering a treatment 0.5 ng/litre Ahling & Jensen (1970)
plant

Tap water produced at the plant 0.33 ng/litre Ahling & Jensen (1970)

Several rivers 0.1-0.3 ng/litre Ahnoff & Josefsson (1974)

USA Polluted coastal area 100-450 ng/litre Panel on Trace Hazardous
Lake Michigan (1970)^a Substances (1972) (cf.
WHO/EURO, 1988)

Distribution system feeding the Fort Edwards reservoirs < 12-160 ng/litre Brinkman et al. (1980, 1981)
in New York (1978)

Hudson River at Fort Edward up to 530 ng/litre Brinkman et al. (1980, 1981)

^a Followed by a marked decrease in 1971.

5.1.3 Soil

Soil may become contaminated with PCBs from atmospheric fall-out or
from direct emissions from point sources. The presence and behaviour
of these compounds in the soil depend on substance (congener)-specific
characteristics and on a number of soil parameters. Sorption and
condensation processes in the soil also play a role in the removal of
PCBs. Some values of PCB levels in soil can be found in Table 14.

Klein (1983) found that PCBs accumulate in the sediments of rivers and
lakes in the Federal Republic of Germany and that these levels
indirectly reflect the contamination of water by PCBs. Some values for
PCBs in sediments can also be found in Table 14.

An important, though localized, source of PCB contamination of soil,
can be the use of sewage sludge as a fertilizer in agriculture. PCB
levels varying from 0.1 to 765 mg/kg (dry weight) have been reported
in sewage sludge from different countries, the usual range being 0.1
to 9.0 mg/kg (WHO/EURO, 1987). In the USA, 16 sewage sludge samples
from cities contained a mean Aroclor 1254 concentration of 5.2 mg/kg
dry weight (range 0.01-23.1 mg/kg). Other authors reported a range of
1.5-27.3 µg/litre in 36 raw sewage sludges. Some levels that have been
found for PCBs in sludges are presented in Table 14 (WHO/EURO, 1987).

Five sediment samples were collected from the Waukegan Harbour of Lake
Michigan, Illinois, in 1978. Residues of 3,4,3',4'-tetrachlorobiphenyl
ranged from 0.005 to 27.5 mg/kg and residues of 2,3,4,3',4'-penta-
chlorobiphenyl, from 0.102-131 mg/kg. The total PCB contents of the
sediment ranged from 10.6 to 13 360 mg/kg (Huckins et al., 1988).

5.1.4 Aquatic and terrestrial organisms

PCBs have been measured in a wide variety of biota from many different
locations throughout the world. Only a few illustrative examples are
given here, more comprehensive lists of PCB residues can be found in
reviews by Risebrough et al. (1968); Peakall (1975); and Eisler
(1986). Tanabe et al. (1987) reported that the highly toxic, coplanar
PCBs are as widely spread as general PCB pollution.

In the biota of a small upstate New York public water supply system,
which is near the polluted section of the Hudson River and a disposal
site of PCB-containing waste, PCBs were found in detectable
concentrations (Table 13). Five samples of algae showed Aroclor 1254
levels of <25 (nd)-120 µg/kg dry weight, macro-invertebrates showed
levels between <200 and 3800 µg/kg and vertebrates, between <25 and
1100 µg/kg dry weight (Brinkman et al., 1980, 1981).

Table 14. PCB levels in soils, sediments, and sewage sludge in several countries

Country Location and/or type of sample PCB levels References
average and/or range

Germany Soil without sewage sludge 0.02-0.08 mg/kg^a Markard (1988)
Soil with sewage sludge 0.05-3.0 mg/kg^a
Sewage sludge ndb-19 mg/kg
Sediments of contaminated waters 0.1-1.0 mg/kg^a Klein (1983)
Sediments of several rivers 0.16-0.59 mg/kg DFG (1988)
Agricultural soil 0.03 mg/kg DFG (1988)

Japan Agricultural soil < 1 mg/kg Fukada et al. (1973)
Soil near a factory making electrical components 510 mg/kg Fukada et al. (1973)

Netherlands Sediments from several surface waters < 0.01-1.2 mg/kg^a Greve & Wegman (1983)

United Soil from a waste disposal area with chemical treatment 4.5-44.8 µg/kg Eduljee et al. (1986);
Kingdom and incineration facilities Badsha et al. (1986);
(Scotland) Grass samples from the same area (foliage) 2.9-64.7 µg/kg Badsha & Eduljee (1986)
Soil of rural areas 8 µg/kg^a (1-23)
Grass of rural areas 9 µg/kg^a (7-16)
Soil of urban areas 52 µg/kg (11-141)
Soil of industrial locations 41 µg/kg (20-67)

United Surface soil 2.5 µg/kg Jones (1989)
Kingdom (0.2-12.2)
(Wales)

Table 14. (cont'd).

Country Location and/or type of sample PCB levels References
average and/or range

USA Sediments near a point of accidental release of PCBs 1.4-61 mg/kg Nimmo et al. (1971a)
(Florida)
Escambia Sediments 16 km downstream of this point 0.6 mg/kg
river
Escambia Soil samples from the bank, 6.5 km downstream from the 1.4-1.7 mg/kg
Bay point

^a Dry weight
^b Not detectable

Serious environmental contamination has been documented in enclosed
water bodies close to urban and industrialized areas, such as the
Great Lakes, the Baltic Sea, and Tokyo Bay. PCB levels in aquatic
organisms reflect these localized high concentrations.

Nimmo et al. (1971a) reported that PCB levels in shrimp from Escambia
Bay, Florida (contaminated by an industrial plant on the Escambia
River) contained between 0.6 and 120 mg Aroclor 1254/kg in 1969 and
fiddler crabs, collected in 1970, contained 0.45-1.5 mg/kg.

When fish, sampled throughout the USA, were analysed by Schmitt et al.
(1983, 1985), the highest levels of PCBs were found in the
North-eastern industrialized areas. Delfino (1979) reported
concentrations ranging from 26 to almost 1000 mg PCBs/kg in fish
collected from the Sheboygan River, Wisconsin, contaminated by a
die-casting plant.

Wiemeyer et al. (1975) analysed osprey eggs in 1968-69 and found
average levels of 2.6 mg/kg in Maryland compared with an average level
of 15 mg/kg in eggs from Connecticut. PCB residues in Connecticut eggs
had not changed significantly compared with those collected in 1964.

Buckley (1982) analysed aspen, sumac, and golden rod plants growing at
various distances (< 1200 m) and in different directions from a PCB
dump in New York State, USA. All the plants were growing beyond a
natural drainage ravine, which prevented contamination of soil and
water by PCBs. Downwind of the site, PCB levels in the plants were
found to be approximately 100 mg/kg dry weight (over 600 times
background levels in plants). Levels above background concentrations
were also found in directions from the site less obviously
contaminated by airborne dust.

Eggs of terrestrial birds collected in a rural environment in Canada
contained lower PCB levels than those sampled from urban areas (Frank
et al., 1975).

In the Great Lakes, the highest levels of PCBs were found in Lakes
Michigan and Ontario for fish (Delfino, 1979) and Lake Ontario for
birds (Weseloh et al., 1979); both lakes receive input from industrial
and urban sites. Glooschenko et al. (1976) found concentrations of up
to 8.1 mg/kg in microorganisms from the middle of Lake Huron.

Weseloh et al. (1983) found that the PCB levels in double-crested
cormorant eggs, collected from Lake Superior during 1972 (average of
23.8 mg/kg fresh weight), were higher than those in cormorant eggs
analysed in other Canadian colonies. Mineau et al. (1984) found that
the locations of herring gull colonies with the greatest mean levels
of PCBs, in each of the Great Lakes, corresponded with the locations
of major sources of the contaminant, as indicated by elevated residues
in sediment.

Muir et al. (1988) determined PCB levels in pooled Arctic cod muscle
(Boreogadus saida) and polar bear fat (Ursus maritimus), and in
the blubber and liver of ringed seals (Phoca hispida) from 3
locations in the East/Central Canadian Arctic. The mean arithmetic
concentrations of total-PCBs in the muscle of Arctic cod of 2
locations were 3 and 5 µg/kg wet weight. The mean concentrations shown
in the tabulation below were found in the blubber and liver of ringed
seals.

Year Number of Sex Arithmetic mean ± SD
samples (µg/kg wet weight)
(blubber)

1972 3 female 639 ± 249
1975/76 5 female 600 ± 99
1983 10 male 794 ± 879
16 female 308 ± 138
1984 19 male 568 ± 287
14 female 375 ± 172
(liver)
1984 19 male 6 ± 4
14 female 4 ± 3

The presence of PCBs in polar bears (Ursus maritimus) was studied by
Norström et al. (1988) in the Northwest territories of Canada. Liver
and adipose tissue specimens were obtained by Inuit hunters from 12
zones over the period 1982-84. A total of 121 samples was obtained.
The mean concentrations of total PCBs in pooled samples ranged from
3.24 to 8.25 mg/kg, on a lipid weight basis. The adipose tissue of
polar bear (10 pooled samples collected in 1982 and 10 samples, in
1984) contained 4.42 and 4.57 mg/kg wet weight, respectively. From
these results, biomagnification factors for the food-chain of the
Arctic cod/ringed seal/polar bears were calculated. For total PCBs,
these factors ranged from 3.7 to 8.8 for fish to seal; from 7.4 to
13.9 for seal to bear, and 49.2 for fish to bear. For individual PCB
homologues, for instance, for fish to bear, these factors ranged from
<0.5 (tetrachlorinated PCBs) to 263.4 for heptachlorinated PCBs.

Niimi & Oliver (1989b) monitored the presence of 92 monochloro- to
decachlorobiphenyl congeners in brown and lake trout, small and large
rainbow trout, and small and large coho salmon from Lake Ontario. Each
sample consisted of 8-12 fish. The highest concentrations were among
the penta- and hexachlorobiphenyl homologues, with 2,4,5,2',4',5'-
hexachlorobiphenyl the most common congener.

Total congener concentrations ranged from 1 to 10 mg/kg in whole fish
and from 0.3 to 4 mg/kg in muscle. The 10 most common PCB isomers were
84, 87/97, 101, 110, 118, 138, 149, 153, and 180, and represented 52%
of the total content. This value did not appear to be influenced by
species or by total concentration.

Huckins et al. (1988) collected fish (1-6 fish of 7 species) from the
Waukegan Harbour of Lake Michigan, Illinois in 1978. The fish samples
were analysed for the presence of 3,4,3',4'-tetrachloro- and
2,3,4,3',4'-pentachlorobiphenyl. Total PCB congener residues averaged
33.4 (2.4-56.6) mg/kg. The concentrations of 3,4,3',4'-tetra-
chlorobiphenyl averaged 45.3 µg/kg (2-89 µg/kg) in the whole body. The
concentrations for 2,3,4,3',4'-pentachlorobiphenyl averaged 229 µg/kg
(80-483 µg/kg).

Five times as much PCBs were found in herrings caught in
industrialized areas of Sweden (near Stockholm) compared with those
caught in the cleaner waters off the Swedish west coast. Levels in
plankton fell progressively with increasing distance from
industrialized areas (Jensen et al., 1972a).

Holden (1973) found levels of up to 235 mg/kg in the blubber of seals
sampled in the polluted coastal areas of the United Kingdom compared
with lower levels (2 mg/kg) from unpolluted areas. Higher levels, (up
to 88 mg/kg) were found in the blubber of toothed whales sampled in
the North Sea, but none was detectable in similar species sampled off
New Zealand and Surinam (Koeman et al., 1972).

Peakall (1975) mapped out the global distribution of PCB levels in
marine plankton. The values for the open North Atlantic (300-450 mg/kg
lipid) were found to be very similar to those collected from polluted
areas, such as the Baltic sea and the Firth of Clyde, in the United
Kingdom. Values in the South Atlantic (12-64 mg/kg) were considerably
lower. The highest values shown were for the Eastern coast of the USA
(up to 3050 mg/kg). There were no values for the Pacific Ocean.

When monitoring PCB levels in fish from the Mediterranean, Albaiges et
al. (1987) found that territorial species reflected local inputs of
the pollutant, but migratory species had baseline levels.

Risebrough & de Lappe (1972) reported PCB levels higher than 3 mg/kg
in fish from the industrialized areas of Tokyo Bay and New York Sound.

Tanabe et al. (1986a) analysed Antarctic minke whales and found that
they contained lower PCB levels than those caught in the Northern
hemisphere (Tanabe et al., 1983). McClurg (1984) also found low levels
of PCB in the Antarctic; Ross seals contained 0.09 mg/kg (in blubber).
Mean levels of 0.69 mg PCB/kg (wet weight), found by Smillie & Waid
(1987) in Australian fur seal blubber, were much lower than levels
found in seals from the temperate Northern hemisphere. Similarly,
Antarctic fish had very low PCB residues, ranging from 0.08 to
0.77 µg/kg wet weight (Subramanian et al., 1983).

PCB residues in biota are usually highest near industrial sources, but
this geographical distribution is becoming less pronounced. In fact,
O'Shea et al. (1980) and Tanabe et al. (1988) found PCB levels in
small oceanic cetaceans to be higher than those reported for
terrestrial mammals and birds. For example, Tanabe et al. (1988) found
the mean level of PCBs in the fatty tissue of the striped dolphin to
be 36 mg/kg wet weight.

Subramanian et al. (1986) analysed subcutaneous fat from Adelie
penguins from the Antarctic and found PCB levels of 0.05 mg/kg fat
weight. This is a factor of 100 lower than that in auks caught in the
northern North Pacific (Tanaka & Ogi, 1984) and a factor of 10 000
lower than residues found in the pectoral muscle (on a lipid weight
basis) of herring gulls in the Baltic (Lemmetyinen et al., 1982).

5.1.4.1 Effect of dredging-contaminated sediment on organisms

Dredging to remove contaminated sediments from the Shiawassee River,
Michigan, increased the availability of PCBs, and, thus, residue
levels, in freshwater clams (64.5-88 mg/kg dry weight) and in fish
(fathead minnow; 13.8-18.3 mg/kg), both during dredging and up to 6
months afterwards (Rice & White, 1987).

5.1.4.2 Relationship to lipid content of organisms

PCBs are accumulated in lipid-rich tissues and care must be taken when
interpreting results between species with different amounts of body
fat. Jensen et al. (1969) found that PCB levels in herring and cod,
from the same area of the Baltic Sea, were 0.27 and 0.033 mg/kg, on a
wet weight basis, respectively, even though the cod is at a higher
trophic level. The 2 species were found to have body fat contents of
4.4 and 0.32%, respectively, and when the PCB residues were
recalculated on a lipid weight basis, herring contained 6.8 mg/kg and
cod, 11 mg/kg.

PCBs are particularly accumulated in animals with large amounts of
fat, such as seals, dolphins, porpoises, and whales (Tanabe, 1988) and
in Arctic and Antarctic birds and mammals. Subramanian et al. (1986)
found PCBs in all Adelie penguins sampled in the Antarctic, an area
known to be relatively low in PCBs; the PCBs were mainly concentrated
in fat-rich tissues. Kawai et al. (1988) measured PCBs in striped
dolphins and found that the tissue level of PCBs depended entirely on
their lipid content and, especially, on the amount of triglycerides in
tissues.

Redistribution of PCBs, from fat to other tissues, occurs in animals
during periods of enforced starvation, such as seasonal food shortage,
hibernation, migration, incubation, and the feeding of offspring.
Subramanian et al. (1986) found that, as individuals Adelie penguins
starved during incubation, residues of PCBs increased with declining
fat reserves concomitant with tissue redistribution. Llorente et al.
(1987) found that migratory duck species had a smaller percentage of
the body burden of PCBs in adipose tissue than a resident species. A
similar redistribution during starvation has been shown in the
laboratory in European robins (Södergren & Ulfstrand, 1972) and big
brown bats (Clark & Prouty, 1977) (see sections 4.2.4.5 and 4.2.4.6).

5.1.4.3 Residues in different trophic levels and effects of diets

In a study by Shaw & Connell (1982), bioaccumulation was increasingly
evident in upper trophic level organisms, such as gulls and pelicans,
in an Australian estuary compared with organisms from lower trophic
levels. Veith et al. (1977) found typical PCB concentrations in Lake
Superior biota to be 0.1 mg/kg for large zooplankton, 0.3 mg/kg for
bottom fish, such as sculpins, and 1 mg/kg for pelagic fish.

When various insects were sampled for PCB residues (Morse et al.,
1987), levels in honey bees ranged from <0.1 to 1.5 mg/kg dry weight.
PCB residues in other species ranged from <0.1 to 2.6 mg/kg, with
predatory wasps containing the highest residues.

Prestt et al. (1970) analysed the livers from various bird species in
the United Kingdom. The highest PCB residues were found in freshwater,
fish-eating species (up to approximately 900 mg/kg). The authors did
not find any geographical pattern of distribution of PCBs in the
species studied.

Frank et al. (1975) collected birds' eggs from the Niagara peninsula
in 1971. Eggs from carnivorous species of birds at the top of the
aquatic food chain contained the highest levels of PCBs
(3.5- 74 mg/kg). Terrestrial carnivores contained lower, but still
relatively high, residues (0.2-1 mg/kg). Eggs from herbivorous and
insectivorous birds contained much lower residues of PCBs. Again, eggs

from terrestrial birds tended to contain lower levels (0.05-2 mg/kg)
than those feeding on aquatic prey (0.14-4 mg/kg). Focardi et al.
(1988) compared the PCB residues in the eggs of 8 species of water
bird. The residues were found to be higher in fish-eating birds than
in invertebrate feeders. The invertebrate feeders tended to contain
higher percentages of the lower chlorinated congeners. Bird species
that fed on other birds or fish had higher liver residues of PCBs than
those feeding on mammals (Cooke et al., 1982). Peregrine falcons,
herons, sparrowhawks, kingfishers, and great crested grebes had
relatively high residues of PCBs. By contrast, golden eagles were only
very lightly contaminated with PCBs.

Bowes & Jonkel (1975) found a similar pattern in Arctic and subarctic
food chains with PCB levels following the pattern: Arctic charfish
< seals < adult polar bears < polar bear cubs.

Mean PCB concentrations of 0.0018 mg/kg were found by Tanabe et al.
(1984) in zooplankton, 0.048 mg/kg in myctophid, 0.068 mg/kg in squid,
and 3.7 mg/kg in striped dolphin (all based on a whole-body, wet
weight basis) sampled from the western North Pacific. The authors
concluded that the bioaccumulation of chlorinated hydrocarbons was
dependent on physical and chemical factors, such as water solubility
and lipophilicity, in the lower trophic levels, whereas, in higher
trophic levels, accumulation was affected by biochemical factors, such
as the biodegradability of pollutants and the metabolizing capability
of the organism.

5.1.4.4 Effects of age, sex, and reproductive status on uptake and
elimination

Bache et al. (1972) found that the burden of PCBs increased with age
in lake trout from Cayuga lake, Ithaca, New York, sampled in 1970
(residues ranged from 0.6 to 30.4 mg PCBs/kg). An age- and
length-related increase in PCBs was found in striped bass from the
Hudson River and Long Island Sound; the author (Connell, 1987) stated
that this observed relationship was due to the slow rate of
bioaccumulation of the PCBs, particularly the higher chlorinated
congeners.

PCBs have been shown to accumulate with age in marine mammals, such as
pinnipeds (Addison et al., 1973; Frank et al., 1973; Helle et al.,
1983) and cetaceans (Gaskin et al., 1983; Aguilar & Borrell 1988;
Subramanian et al., 1988). Helle et al. (1983) found mean levels of
5.1 mg PCBs/kg (in extractable fat of blubber) in newly-born ringed
seal pups, 17.3 mg/kg in seals of 2-4 months of age, and 65.3 mg/kg in
sexually mature adults (4-12 years). However, lower levels of PCBs
have been found in females compared with males (Martineau et al.,
1987) and the age-related increase has often not been found in females

(Addison & Smith, 1974). In many studies, while levels of PCBs in
males have increased with age, those measured in females have fallen
(Born et al., 1981; Gaskin et al., 1983; Aguilar & Borrell, 1988).
Gaskin et al. (1983) found that PCB levels in the blubber of male
harbour porpoises increased from 48.4 mg/kg at birth to 161 mg/kg
after 8 years, whereas, in females, levels fell from 51 to 14.7 mg/kg.
A significant decrease in the PCB levels was found by Subramanian et
al. (1988) in female Dall's porpoises from 2 years of age onwards; 2
years is required for the animals to reach sexual maturity. Excretion
of PCBs during reproduction is known, from the laboratory, to be an
important means of females losing residues. This PCB loss has been
shown to be because of the transfer of PCBs to offspring via milk
during lactation (Addison & Brodie, 1977). Addison & Brodie (1977)
calculated that female grey seals excreted about 15% of their body
burden of PCBs via lactation. In striped dolphins, females transferred
between 72 and 98% of their body burden to the offspring (Fukushima &
Kawai, 1981; Tanabe et al., 1982). It was suggested by Tanabe (1988)
that such large transfer was because of the very high lipid content of
the milk. Relocation of the PCB burden during pregnancy is generally
thought not to be as important; in grey seals, the mother transfers
only about 1% of her body burden to her offspring (Donkin et al.,
1981) and in striped dolphins, only 4-9% (Fukushima & Kawai, 1981;
Tanabe et al., 1982). However, Duinker & Hillebrand (1979) suggested
that a much bigger percentage of female body burden (up to 15%) could
be transferred to the fetus across the placenta of Harbour porpoise.

Clark & Lamont (1976) calculated that female big brown bats
transferred between 17 and 32% of their body burden of PCBs to their
young, during gestation. The concentration of PCBs in adult females
plus their litters declined with increasing age of the female. PCB
levels were 0.83-3.6 mg Aroclor 1260/kg (wet weight) in adults and
0.22-3.3 mg/kg in litters.

When Passino & Kramer (1980) measured PCBs in deepwater ciscoes from
Lake Superior, male fish contained significantly higher levels of PCBs
(2.3 mg/kg wet weight) than females (1.2 mg/kg), eggs containing
0.51 mg/kg. Lemmetyinen et al. (1982) found annual rates of
elimination via egg production of 45% in the female Arctic tern and
24% in the herring gull. Adelie penguins eliminated only 4% of their
PCB body burden after laying their annual clutch of 2 eggs (Tanabe et
al., 1986b). Elimination was thought to be dependent on the relative
weights of the egg and mother.

5.1.4.5 Time trends in residues

Buckley (1983) analysed various species of terrestrial plants from New
York state. Total decreases of 42% in PCB residues were found between
1978 and 1980.

PCB levels in fish in the Hudson River, New York declined between 1977
and 1981. The PCB levels were much higher in the Upper Hudson River
(4217-1431 mg/kg of lipid), near to a major discharge of PCBs, than in
the Lower Hudson River (1604-319 mg/kg) (Sloan et al., 1983).

Frank et al. (1978) measured PCB levels in various fish species from
Lakes Huron and Superior during the period 1968-76. PCB residues
declined in lake trout and lake whitefish in Lake Superior between
1971 and 1975, but increased slightly over the same period in bloaters
and white sucker. In Lake Huron, PCB levels decreased between 1968 and
1971, and, in alewife, rainbow smelt, and walleye, between 1975 and
1976. In some of the study areas, residues increased in cisco, yellow
perch, coho salmon, and splake but, at most locations, and, for other
species analysed, no trends in PCB levels were found. St Amant et al.
(1984) analysed fish from Lake Michigan between 1971 and 1981. An
overall decrease in PCB levels was found for all species monitored
except the walleye. Levels decreased from a maximum of 22.4 mg/kg at
the beginning of the study to 3.8 mg/kg or less in 1981.

Fish from all over the USA were analysed in 1980-81 by Schmitt et al.
(1985) who found a significant downward trend (0.88-0.53 mg/kg PCB;
wet weight) when mean residues were compared with fish collected
between 1976 and 1977 (Schmitt et al., 1983). A similar downward
pattern in residues was found in the Baltic when Moilanen et al.
(1982) compared residues found in pike and herring caught between 1978
and 1982 with those in fish sampled between 1972 and 1978 (Paasivirta
& Linko, 1980). Haahti & Perttila (1988) found a continued decline in
PCB residues between 1979 and 1986, when residues in herring muscle
tissue decreased from 2.7-3.7 mg/kg to 0.3-1.1 mg/kg.

An overall fall in PCB levels was found by Newton & Bogan (1978) in
sparrowhawk eggs during the period 1971-74. Cooke et al. (1982)
analysed liver samples from grey herons, kestrels, and barn owls for
PCB residues during the period 1967-77. They found a significant
decline in PCB residues over the sampling period in all 3 species. The
mean residues in heron, kestrel, and barn owl for the period 1967-71
were 5.77, 1.57, and 0.44 mg/kg, respectively, and for 1977, 0.56,
0.6, and 0.15 mg/kg, respectively. However, Newton et al. (1986), when
analysing sparrowhawk eggs from 1971-80, found that, although levels
had fallen in the early 1970s, they had risen again in the late 1970s
(mean PCB residues in eggs ranged from 16 to 293 mg/kg in lipid). Data
on PCB residues in the livers of kestrel, sparrowhawk, heron,
kingfisher, and the great crested grebe, collected from the late 1960s
up to 1987, were analysed statistically by Newton & Haas (1989). For
the great crested grebe, a significant overall decline in PCB residues
was found when comparing data from 1987 with that from the 1960s. For
the other species, there was no significant difference. Spitzer et al.

(1978) reported that there was no significant change in PCB levels in
osprey eggs collected from the Connecticut-New York area during the
period 1969-76. Similarly, Wiemeyer et al. (1987) did not find any
change in the carcase levels of PCBs in ospreys from the Eastern
United States when comparing the 1971-73 and 1975-82 periods. They did
find that adults contained significantly higher concentrations of PCBs
than immature ospreys.

Blus et al. (1979) analysed brown pelican eggs from South Carolina and
Florida between 1969 and 1976. The highest levels of PCBs were found
in South Carolina (means ranged from 5.25 to 7.63 mg/kg wet weight),
but no significant trend was found during the study period. In
Florida, the authors did not find any significant change in eggs
collected from colonies in Florida Bay and on the Gulf Coast over the
study period (means ranged from 0.62 to 1.18 mg/kg), but the Atlantic
coastal colony showed a significant increase in PCB residues (from a
mean of 2.68 to 6.12 mg/kg) between 1969 and 1976.

In analysing herring gull eggs from the Great Lakes between 1974 and
1978, Weseloh et al. (1979) found a significant decline in PCB
residues from colonies on all the lakes. Lake Ontario, the most
contaminated, showed the biggest decline from 170 to 75 mg PCBs/kg at
one of the colonies, with other less contaminated Lakes, Huron,
Superior, and Erie, showing levels in the range of 50-86 mg/kg in 1974
and 32-46 mg/kg in 1978.

Moksnes & Norheim (1986) analysed herring gull eggs collected from the
Norwegian Coast between 1979 and 1981 and found that the PCB levels
were not significantly different from those in eggs collected in 1969;
mean PCB residues ranged from 1.2 to 6.7 mg/kg wet weight. They found
a small but significant increase in the most persistent congeners and
a significant decrease in DDE and the DDE/PCB ratio, but not in total
PCB levels.

An analysis of the eggs of double-crested cormorant (an inshore-
subsurface feeder), Leach's storm petrel (an offshore-surface
feeder) and Atlantic puffin (an offshore-subsurface feeder) was
carried by Pearce et al. (1989), every 4 years, between 1968 and 1984.
In the Bay of Fundy, Canada, PCB levels declined significantly during
this period in all 3 species. PCB levels in the cormorant were
consistently higher throughout than those in the other 2 species,
ranging from 4 to 29.5 mg/kg (wet weight). Petrel and puffin eggs
collected from the Atlantic Coast of Newfoundland showed lower levels
than those in eggs from both the Bay of Fundy and the St Lawrence
River estuary; as in the St Lawrence River, no significant trend in
PCB levels was observed. A significant decline in PCB residues was
found in gannet eggs collected during the same period from the gulf of
St Lawrence (Elliott et al., 1988).

The frequency of occurrence of measurable PCB residues has increased
in large-scale sampling exercises; PCBs in mallard wings increased
from 39% in 1976-77 (White, 1979) to 95% in 1979-80 (Cain, 1981). Cain
& Bunck (1983) found that, in 1976, 21% of European starlings
collected in the USA contained PCBs compared with 83% in 1979.

Addison et al. (1986) analysed the blubber of Arctic ringed seals
(Phoca hispida) from Holman Island, NWT, Canada, in 1981. They found
mean PCB levels of 0.58 mg/kg (wet weight) in the females and
1.28 mg/kg in the males. These concentrations were significantly lower
than those detected in the same species from this area in 1972. Over
this same period, pp'-DDE levels, although at lower levels, also
fell significantly, but it should be noted that total DDT levels in
blubber are much lower than PCB levels and have not changed
significantly.

5.1.4.6 Seasonal patterns in residues

Jensen et al. (1969) observed that there was considerable seasonal
variation in the fat content of herring caught in the Baltic Sea,
ranging from 1% in the spring to 10% in the autumn and that this
seasonal change in fat content led to seasonal changes in the tissue
levels of PCBs.

Cooke et al. (1982) found a seasonal pattern of PCB levels in European
kestrels. Residues in both fat and liver were low in the autumn, but
increased from about January, with a peak almost invariably occurring
during the second quarter of the year (April, May, or June). Seasonal
patterns were based on samples collected over a 10-year period.
Similar trends were found in sparrowhawks and barn owls, but fewer
samples were available.

5.1.5 Appraisal

PCB contamination is widespread and has been measured in a wide
variety of biota between the 1960s and the present day. They are
present throughout the world and, although initially concentrated in
areas of high industrial activity, are now found in organisms living
in remote areas, such as the oceans and the polar regions. In the
past, PCB levels were positively correlated with areas of heavy
industry and consequent discharge but, with the implementation of PCB
controls, in some countries, these geographical differences are
becoming less clear. Generally, levels of PCBs are declining in areas
previously high in PCBs. However, time-trend analysis for the general
environment shows little change in total PCBs since the late 1960s.
The ratio of congeners is, as would be expected, changing, with lower
chlorinated isomers disappearing and the more highly chlorinated ones
becoming more dominant in environmental samples.

PCBs are persistent and bioaccumulate in many organisms, because of
their high lipid solubility and low biodegradability, and usually
enter food-chains from water containing industrial discharge and by
precipitation.

Because of their hydrophobic nature, PCBs are associated with both
oildrop-like aggregates in the surface microlayer of water and with
sediment on the bottom.

They are accumulated by micro- and macroplankton organisms that live
in the surface microlayer and by bottom-living organisms.

5.2 Levels in animal feed

The effects of pollution are seen in the use of fish-meal in poultry
and fish farming. Kolbye (1972) sated that this may contain PCB levels
of 0.6-4.5 mg/kg.

Hansen et al. (1981) studied the transfer of PCBs in swine foraging on
sewage sludge amended soils in 1975-76. Sixteen Berkshire sows were
overwintered for 2 seasons on 4 experimental plots that had been
treated with 0, 126, 252, or 504 tonnes/hectare (on a dry solids
basis) of Chicago sewage sludge for the 8 preceding years. The
estimated PCB residues in the soils of the 4 plots (average of 3-4
samples) were 1.62, 1.88, 2.13, and 2.81 mg/kg dry weight (mean values
of 3-4 samples/plot). The mean concentrations in fat of 3-4 sows per
plot were 36 ± 9, 106 ± 64, 191 ± 97 and 389 ± 118 µg/kg fat basis. Of
the 12 individual congeners that were present in the fat, 3 accounted
for more than 50% of the congeners, e.g., 2,3,4,2',4',5'-,
2,4,5,2',4',5'-hexachlorobiphenyl and 2,3,4,5,2',4',5'-
heptachlorobiphenyl.

In vegetable animal feed (155 samples) originating from 5 areas of the
world, samples, collected in 1984/85, contained PCB levels of 0.0009
(Africa) up to 0.0093 mg/kg dry weight (Europe). In feed from North
and South America and the Far-East, the levels were between 0.0024 and
0.0066 mg/kg. Different types of feed originating from agriculture in
the Federal Republic in Germany, collected in 1985, contained PCB
levels of the order of 0.02 mg/kg dry weight. In feed (301 samples)
originating from animals (exclusive fish meals), collected in 1985,
0.021-0.036 mg/kg dry weight was found (DFG, 1988).

Levels of 10-100 µg/kg are given for groats, soybeans, and cotton
seed, and a mean value of 18 µg/kg is given for mixed feedstuffs. Fish
meal contained levels of 110-330 µg/kg (Klein, 1983).

Samples of fish meal from different areas of the world, collected in
1985, were analysed for the presence of PCBs. In 323 samples, the PCB
contents varied between 0.006 and 0.055 mg/kg dry weight. The PCB
congeners numbers 28, 138, and 153 were present in the highest
quantities (DFG, 1988).

5.3 Levels in human food

5.3.1 General

Two general reviews of PCB residues in food, animal feed, human milk,
plants, soils, and packaging materials have been published by Khan et
al. (1976) and Sawhney & Hankin (1985).

The PCB contents of a variety of foods on the Swedish market has been
measured by Westöö & Norén (1970a) and Westöö et al. (1971). Less than
0.1 mg/kg was found in samples of butter, margarine, vegetable oils,
eggs, beef, lamb, chicken, bread, biscuits, and baby food; one sample
of pork out of more than 100 had a PCB content of <0.5 mg/kg.

In the period 1980-81, 5270 food samples were drawn at wholesale or
production levels or at the site of importation including: butter,
cheese, eggs, kidneys from pigs and cattle, and fat of poultry. Levels
in Danish butter (99.4% of the samples) were below 0.05 mg/kg and
those in imported butter (100%), below 0.125 mg/kg; Danish cheese
(100% of the samples) levels were below 0.05 mg/kg and, in imported
cheese, 82.4% of samples had levels below 0.125 mg/kg and the other
17.6%, below 0.2 mg/kg; 100% of eggs had levels below 0.05 mg/kg, and
100% of kidneys of pigs and cattle were below 0.15 mg/kg; 96% of
poultry fat samples had levels below 0.15, and 4%, below 0.20 mg/kg,
on a fat basis (not stated) (Statens Levnedsmiddelinstitut, Danmark,
undated).

Mes et al. (1989b) studied the presence of specific isomers of PCB
congeners in fatty foods of the Canadian diet. A total of 93 food
composites from the cities of Ottawa and Halifax were analysed for 34
PCB isomers, as part of a revised total diet programme. Each market
basket comprised approximately 200 different food types collected from
each of 4 major supermarkets in Ottawa during September 1985 and
January 1986, and, in Halifax, in September 1986. Foods were used
per se, or prepared and cooked in a manner ready for consumption,
then composited to give 112 composites from each market basket.
Thirty-one selected composites, representing the fatty foods were
analysed from each market basket.

PCB isomers 118, 138, 153, and 180 were found in all dairy products,
except skimmed milk. Cheese and butter contained the highest levels of
PCB residues. The residue level of isomer 118 (2,4,5,3',4'-
pentachlorobiphenyl) in butter was the highest e.g., 0.7 µg/kg, of all
PCB isomers found in dairy products. Almost all meat, fish, and
poultry contained PCB isomers 183 and 187. Occasionally, isomers 49,
87, 185, and 189 were also present, but isomer 105 (2,3,4,3'4'-
pentachlorobiphenyl), present in most dairy products, was only found
in some beef samples. Fresh water fish contained most PCB isomers (28
out of 34 selected PCB isomers), at levels considerably higher than
those in any other meat, fish, or poultry samples. The level of isomer
110 in fresh water fish was 3.05 µg/kg. PCB isomers 138, 153, 180, and
187 were present in almost all samples of meat and fish products,
fats, oils, and soups. Cooking fats, salad oils, and margarine
contained relatively low levels of PCB residues. PCB isomers 37, 49,
87, 105, and 185 were not detected in meat and fish products, fats,
oils, or soups.

The calculated sum of all PCB isomer residues found in selected food
commodities (except fish) ranged from 0.03 to 1.98 µg/kg on a wet
basis, and from 0.07 to 10.71 µg/kg on a lipid basis, with mean values
of 0.60 and 3.91 µg/kg, respectively. However, the mean residue levels
of fish and fish products were considerably higher, i.e., 10 and
194 µg/kg on a wet and lipid basis, respectively.

The major PCB isomers in fatty foods were isomers 37, 52, 99, 110,
118, 138, 153, 180, and 187.

The PCB levels obtained in an extensive study by the US Food and Drug
Administration are shown in Table 15. These values are considerably
higher than those reported from Sweden, but they are probably biased,
as they include samples originating from areas previously suspected of
having been subject to local pollution.

In a Canadian survey, PCB levels of less than 0.01 mg/kg were found in
eggs (Mes et al., 1974) and a mean of 0.042 mg/kg was found in
domestic and imported cheese with a maximum of 0.27 mg/kg (Villeneuve
et al., 1973b).

A preliminary study was carried out to estimate the dietary intake of
PCBs in fresh food composites grown in Ontario in 1985. The following
5 food composites: fresh meat and eggs, root vegetables (including
potatoes), fresh fruit, leafy and other above-ground vegetables, and
cow's milk were analysed. The concentrations in the different food
composites were below 0.0005 mg/kg. The annual dietary intake of PCBs
was estimated to be 32.6 µg (Davies, 1988).

In Japan, a similar range of PCB contents has been reported for most
foods; however, some high levels have been reported for rice and
vegetables harvested in fields polluted with PCBs (Environmental
Sanitation Bureau, 1973). The PCB content of most fish on the market
was less than 3 mg/kg.

Table 15. PCB levels in food in the USA^a

Food % Positive Level in positive samples (mg/kg)
(0.1 mg/kg)
Mean Maximum

Cheese 6 0.25 1.0
Milk 7 2.3 27.8
Eggs 29 0.55 3.7
Fish 54 1.87 35.3

^a From: Kolbye (1972).

Cantoni et al. (1988) analysed different food items, in 1985-87, in
Italy, taking 20-60 samples per item. Different types of meat were
analysed and the median concentrations were 0.25-0.50 mg/kg, on a fat
basis. Twenty to 50% of the samples were positive. Poultry contained
0.028 mg/kg, cow's milk 0.05 mg/kg, cream 0.027 mg/kg, butter
0.065 mg/kg and fish 1.105 mg/kg, on a fat basis; 71% of fish samples
contained PCBs.

When the fat of poultry (42 samples) and 44 eggs was analysed, PCB
values were below 0.3 mg/kg (Dutch Agricultural Advisory Commission,
1983).

In the Federal Republic of Germany, wheat was analysed during the
period 1972-82. The mean concentrations for 1972-78 ranged from 10 to
30 µg/kg; in the period 1980-82, the range was < 2.0-18 µg/kg (Klein,
1983). In wheat and rye (total 850 samples), median levels of
0.4-1 µg/kg product were found in 1984 (Codex Alimentarius, 1986). The
concentrations found in other food items are summarized in Table 16.

Samples of canned ham exported from Czechoslovakia to the USA in 1983
contained PCBs levels of up to 4.8 mg/kg (Anon., 1983a,b).

Table 16. PCBs in food (1982) in the Federal Republic of Germany^a

Food Total no. Number of Variation Mean
of samples min-max (µg/kg)
samples below (µg/kg)
detection
limit^a

Milk 854 234 < 2-3000 126.7 (FB)
Beef 76 43 < 10-687 72.4 (FB)
Pork 58 36 < 10-458 58.1 (FB)
Poultry 64 61 < 10-85 7.3 (FB)^c
Meat products 185 86 < 4-2700 114.2 (FB)
Eggs 82 67 < 5-230 9.1 (FW)
Fish (only 70 - 40-87 41.1 (FW)
cod, herring,
plaice)

Food of plant origin

Oil 167 139 < 5-65 7.1 (FB)
Cereals 345 44 < 2-30 6.7 (FW)
Potatoes 106 106 < 2 -

^a From: Klein (1983).
^b Not stated.
^c Only 3 samples.
FB = fat basis
FW = fresh weight.

5.3.2 Drinking-water

Ruoff et al. (1988) examined 83 drinking-water samples from the
Federal Republic of Germany and from 5 other European countries for
their contents of the PCB congeners 28, 52, 101, 138, 153, and 180.
The average total content of the 6 congeners was 0.002 µg/litre water.
The average concentrations of the above-mentioned PCB congeners in the
drinking-water of 6 countries were 0.0001, 0.001, 0.00018, 0.00035,
0.00037, and 0.00042 µg/litre. The variation between the 6 countries
was quite small.

The highest concentration of PCBs reported in domestic tap water was
0.1 µg/litre in the Kyoto area of Japan (Panel on Hazardous Trace
Substances, 1972 cf. WHO/EURO, 1988), but, levels, more likely to be
encountered, should not exceed 0.001 µg/litre.

In the FAO/WHO collaborating centres for the food contamination
monitoring programme, the median levels were:

Cereals below 10 µg/kg
Vegetable fat/oils below 5 µg/kg
Fresh fruit and vegetables 0.5-5 µg/kg
Animal fat (depending on type of
animal and origin) 20-240 µg/kg
Whole fluid cow's milk (depending
on country) 10-200 µg/kg
(on fat basis)
Butter 30-80 µg/kg
Whole dried cow's milk 20-50 µg/kg
Hen eggs < 10 µg/kg
Fresh finfish 10-200 µg/kg

(WHO, 1985b).

The contamination of a drinking-water system in Pickens County, South
Carolina by PCBs discharged from a manufacturing facility was
described by Billings et al. (1978). They observed that PCBs
discharged by a capacitor manufacturing plant resulted in levels as
high as 0.818 µg/litre in finished potable water.

5.3.3 Dairy products

A number of data on food-producing animals have recently become
available within the framework of the Joint FAO/WHO Food Contamination
Monitoring Programme (JFCMP, 1985). All reported median values of PCBs
in animal fat (excluding milk fat) were below the respective limits of
detection, which varied from 0.001 mg/kg in the United Kingdom to a
high of 0.5 mg/kg in Thailand and the USA. Data on PCB levels in cow's
milk fat were supplied by the Federal Republic of Germany, Japan, the
Netherlands, the United Kingdom, and the USA. The United Kingdom and
the USA reported that median concentrations in cow's milk were below
the detection limits of 0.5 µg/kg and 0.5 mg/kg, respectively.

The available data are summarized in Table 17.

From the end of 1982 to the beginning of 1983, high levels of PCBs
were detected in milk from several dairy farms in Switzerland. The
investigations showed that the silo coatings and consequently the
silage from the silos were the origin of the contamination of the
milk. The PCB levels were between 0.80 and 3.80 mg/kg fat. PCB
dissolution in acid juice, mechanical erosion of the coatings, and
volatilization of the coating surface seemed to be the principal
mechanisms explaining the migration of PCBs into the silage
(Alencastro et al., 1984).

Forty-two samples of cow's milk (14 samples in 1976, 14 in 1983, and
14 in 1986) and 41 samples of market milk (10 in 1976, 16 in 1983, and
15 in 1986) were analysed for PCBs, in Israel. During this period, a
change was observed in the PCB distribution in the milk samples. The
percentage of hexachlorobiphenyl decreased with time and the
pentachlorobiphenyl increased (Pines et al., 1988).

The monitoring data for dairy products from all over the world for
1980-83 have been summarized by the Joint FAO/WHO Food Contamination
Monitoring Programme (WHO, 1986a,b).

5.3.4 Fish and shellfish

A summary of the monitoring data on fish from all over the world for
1980-83 has been published by the Joint FAO/WHO Food Contamination
Monitoring Programme (WHO, 1986a,b).

As might be expected, the PCB values found in fish depended on the fat
content and the pollution of the fishing area (Westöö & Norén, 1970a;
Berglund, 1972).

In a collaborative study by 7 national laboratories (International
Council for the Exploration of the Sea, 1974), the PCB contents in the
muscle tissue of fish taken from the North Sea were measured. A mean
of 0.01 mg/kg was found in cod, while herring contained up to
0.48 mg/kg, with most samples in the range of 0.1-0.2 mg/kg; plaice
contained 0.1 mg/kg or less. Similar values were reported by Zitko
(1974) for fish taken from the North Atlantic.

Risebrough & de Lappe (1972) reported levels higher than 3 mg/kg in
fish from New York Sound and Tokyo Bay, both very polluted areas. Even
higher levels of PCBs were found in fish from polluted lakes and
inland waterways, a level of 20 mg/kg being found in fish from Lake
Ontario, and levels over 200 mg/kg in fish from the Hudson River
(Stalling & Mayer, 1972). Similar correlations between pollution and
PCB levels have been reported from the United Kingdom in fish
(Portmann, 1970), and in mussels (Holdgate, 1971).

Table 17. Occurrence of PCBs in dairy products

Country Year Product Number of Mean concentrations Reference
samples in mg/kg on fat basis
(range)

North America

USA 1973-1974 milk (bulk) 198 (9 positive) 1.91 (0.32-4.99) Willett (1980)

Europe

Germany 1982-1986 milk 3279 0.09-0.14^a DFG (1988)
(3 areas) 1983-1986 butter/cheese 2088 0.05-0.11
Westphalian area 1972-1974 butter - 0.38 (0.25-0.54) Claus & Acker (1975)
Northern part 1978-1980 milk - 0.17-0.20 Codex Alimentarius
1984 milk 3510 0.013 (1986)
Northern part - butter 1836 0.0077^c Codex Alimentarius
(1986)
- meat and fat 957 (about 3/4 0.01^b DFG (1988)
positive)
cows entrails 51 0.149^b

Sweden 1972-1977 beef, pork and meat 232 (217 < 0.001-0.01 Vaz et al. (1982)
products (domestic negative) (whole product)
and imported)

Denmark 1981-1982 milk - 0.10-0.13 Jensen (1983b)

Table 17. (cont'd).

Country Year Product Number of Mean concentrations Reference
samples in mg/kg on fat basis
(range)

Netherlands 1975-1977 milk 315 0.16 (0.06-0.33) Gezondheidsraad
1980-1983 milk - 0.07-0.13 (1985)
1978-1984 milk 2319 < 0.1-0.2 Olling (1984)
1977-1981 cattle fat - 0.11^b (< 0.05-0.55) Greve & Wegman
pork - 0.07 (< 0.05-0.66) (1983)
1983 fat of cattle, pork, 40-45 < 0.03^b Dutch Agric. Adv.
calves Comm. (1983)
sheep 22 < 0.03^b

Switzerland - milk 6 0.034-0.144 Rappe et al. (1987)
(6 locations)

^a Major congeners were Nos. 138 and 153.
^b Median value.
^c Arithmetic mean.

Jensen et al. (1969) found PCB levels of 0.27 mg/kg and 0.33 mg/kg,
respectively, in the muscle tissue of herring and cod from the same
area of the Baltic, though the cod is at a higher trophic stage. The 2
species had 4.4 and 0.32% of extractable fat, respectively, and, when
the PCB level was calculated on the fat content, values of 6.8 mg/kg
for the herring and 11 mg/kg for the cod were obtained. Cod liver has
a much higher fat content than cod muscle, and Jensen (1973) reported
the ratio of PCB concentrations in cod liver and muscle to be over
100, the maximum in liver being 59 mg/kg. Jensen et al. (1969)
remarked that the considerable seasonal variation in the fat content
of the herring, rising from 1% in spring to 10% in autumn, influenced
the tissue level of PCBs.

There are many examples of different PCB levels in similar species
collected from areas of high and low pollution. Jensen et al. (1972b)
found 5 times as much PCBs in herrings caught in waters off
industrialized areas near Stockholm, as in herrings from the cleaner
waters of the west coast of Sweden.

Different freshwater and seawater fish were analysed for PCB contents,
during the period 1981-83, in the Netherlands. Eel from different
places over the period 1971-81 contained 0.2-13 mg/kg on a product
basis (in the edible part). The median value was between 1 and
2 mg/kg. Sea fish from the North Sea, such as herring and mackerel,
contained 0.1-0.2 mg/kg, on a fat basis. The same level was found in
shrimps and mussels (Freudenthal & Greve, 1973; Greve & Wegman, 1983;
van der Kolk, personal communication, 1984a).

The mean PCB contents in the liver of cod from the North Sea, North
Atlantic, and Baltic Sea, were 2.1-5.7, 0.48, and 10.4-12.8 mg/kg,
respectively (Klein, 1983).

When fish from the North Atlantic, North Sea, and Baltic Sea, were
collected in 1985, PCB concentrations of 0.098-0.123 mg/kg fillet
weight were found in fish from the North Atlantic and North Sea and
0.338 mg/kg fillet weight in fish from the Baltic Sea. In total, 60
samples were analysed. The PCBs 101, 138, and 153 were the major
congeners (DFG, 1988).

The PCB concentration in freshwater fish of the River Rhine was found
to be more than 2 mg/kg. The mean PCBs levels decreased, however, over
the period 1976-81 from 1.92 to 0.38 mg/kg (fresh weight) (Klein,
1983).

In 1984, PCB concentrations in freshwater fish (59 samples) collected
in the River Rhine ranged from 0.742 to 1.017 mg/kg fillet weight. In
this case, the major congeners were 138 and 153, but numbers 28, 52,
101, 180 were also present. In total, 199 samples of eel were
collected in a number of surface waters and analysed for the presence
of PCBs. The levels ranged from 1.42 to 6.51 mg/kg fresh weight. In
studies reported by DFG (1988), the highest levels of PCBs were found
in the River Rhine.

In the United Kingdom, fish and shellfish were analysed for PCBs
during the period 1982-84 (HMSO, 1986). The results are summarized in
Table 18.

Table 18. PCB levels in marine fish and shellfish^a

Year Product Tissue No. of Range (mg/kg)
samples

1982 Marine fish (from England) liver 381 0.3-4.1
(7 types of fish)

1982 Marine fish (from England) muscle 326 0.03-0.13
(7 types of fish)

1983 Marine fish (imported) muscle 102 nd-0.06
(5 types of fish)

1983 Shellfish (imported) muscle 53 nd-0.06
(4 types of shellfish)

1984 Fish oils 16 0.11-2.3

^a From: HMSO (1986).

Different types of marine fish and shellfish from different areas in
the United Kingdom were analysed during the period 1977-84. Those from
the North Sea coast contained concentrations in the range of 0.04-5.7
and < 0.001-0.058 mg/kg, respectively, while those from the English
channel contained < 0.05-6.9 and < 0.006-0.1 mg/kg, respectively,
and those from the West coast, < 0.002-8.4 and < 0.001-0.25 mg/kg
wet weight. PCB concentrations in fish livers of 0.2 up to 12.9 mg/kg
wet weight were found during this period (Franklin, 1987).

When samples of fish of different species, collected from major USA
watersheds in 1976, were analysed, PCBs were found in 93% of the
samples. Fifty-eight of the samples had levels exceeding 5 mg/kg, on a
whole fish basis. The PCB concentrations ranged from less than 0.3 to
140 mg/kg, on a whole fish basis (Veith et al., 1979).

Maack & Sonzogni (1988) analysed 98 fish (14 species) of different
sizes from Wisconsin waters, for the presence of PCB congeners. Among
the most prominent congeners were numbers 153/132, 138, 66/95, 110,
180, 70/76, 146, 28/31, 149, 118, and 105. The total PCBs (determined
by adding individual congener concentrations) ranged from 0.070 to
7.0 mg/kg. The mean concentration was 1.3 mg/kg.

Blue crabs ( Callinectes sapidus, an important member of the
estuarine food web), collected from Campbell Creek and surroundings in
South Carolina, were analysed for PCBs in 1985. The highest mean total
concentration was 0.861 mg/kg muscle tissue. In 1986, the mean
concentrations in blue crab collected by 8 stations in the same area
ranged from 0.026 to 0.361 mg/kg muscle tissue. Blue crab (15 samples)
collected from the coast of South Carolina, contained concentrations
of < 0.020-0.372 mg/kg tissue (Marcus & Mathews, 1987).

PCBs concentrations in sea fish were determined in 1971-77 in Japan.
In-shore fish (90 samples) showed concentrations of 0.2-0.72 mg/kg
fresh weight and pelagic fish (112 samples), 0.005-0.265 mg/kg fresh
weight (Watanabe et al., 1979).

Data on individual species of fish, submitted by Japan, showed the
following median levels: barracuda, 70 µg/kg; conger eel, 290 µg/kg;
croaker, 200 µg/kg; flounder (yellow-tail), 90 µg/kg; hair-tail,
100 µg/kg; mullet, 84 µg/kg; and seabass, 110 µg/kg. Median levels for
other species of fish, such as cod, mackerel, pacific saury, rockfish,
salmon, and sardines, were below 100 µg/kg (WHO, 1986b).

Using a very sensitive analytical method, Tanabe et al. (1987) found
the toxic non- ortho-substituted coplanar 3,4,3',4'-tetrachloro-,
3,4,5,3',4'-pentachloro-, and 3,4,5,3',4',5'-hexachlorobiphenyl in
finless porpoise, at concentrations of 13.5, 0.89, and 0.64 µg/kg,
respectively.

Blue mussel (Mytilus edulis) was collected from coastal areas near
Osaka and Hokkaido, Japan, in 1984-86. Depending on the site of
collection, the average PCB concentrations (11-13 samples) ranged from
0.56 to 65.0 µg/kg (Miyata et al., 1987).

5.3.5 Influence of food processing

Fifty striped bass (Morone saxatilis) were analysed for the presence
of PCBs in the fish fillets before, and after, boiling, steaming,
baking, frying, microwaving, or poaching, to study the possible
reduction of the PCB residues by these cooking procedures. PCB
contents were reduced by approximately 10%, by all 6 methods of
cooking. No significant reductions were observed with the other
cooking methods (Armbruster et al., 1987).

5.3.6 Food contamination by packaging materials

When Villeneuve et al. (1973a) analysed packaged food in Canada, they
found that 66.7% of the samples contained PCB levels of less than
0.01 mg/kg, 30.7% contained between 0.01 and 1 mg/kg, and 2.6%
contained more than 1 mg/kg. The highest PCB levels were in a rice
sample (2.1 mg/kg), where the packaging material contained 31 mg/kg,
and in a dried fruit sample (4.5 mg/kg), in a container containing
76 mg/kg. In a survey of packaging containers, approximately 80% were
found to contain PCB levels of less than 1 mg/kg, while about 4%
contained levels higher than 10 mg/kg. The most likely source of PCBs
in packaging materials was the recycling of waste paper containing
pressure-sensitive duplicating paper (carbonless copying paper)
(Masuda et al., 1972).

Relatively high PCB levels in some packaged foods in Sweden, mainly of
imported origin, could be attributed to migration from the packaging
material (Westöö et al., 1971). The highest level encountered was
11 mg/kg in a childrens' breakfast cereal; PCB levels of 70 mg/kg and
700 mg/kg were found in the material of the inner bag containing this
product and in the outer cardboard container, respectively. Up to
2000 mg/kg was found in cartons of other samples.

In the United Kingdom, levels in imported waste-paper, which could be
contaminated with PCBs from carbonless copying paper and subsequently
used to manufacture food contact paper and board materials, were found
to be low, compared with the 10 mg/kg limit for PCBs recommended by
the British Paper and Board Industry Federation for food contact
materials (HMSO, 1989).

5.3.7 Appraisal

Foods have become contaminated with PCBs by 3 main routes:

* accumulation of PCBs in the different food-chains in the
environment and consumption of fish, birds, or other animals and
crops;

* direct contamination of food or animal feed by an industrial
accident;

* migration from packaging materials into food.

During the past years, many thousands of samples of different
foodstuffs have been analysed for PCB contamination. The most common
foodstuffs analysed have been fish, meat, and milk. Many fish samples
have been taken in an effort to monitor aquatic pollution. In
addition, samples have been taken, for regulatory or similar purposes,
from sources suspected of being relatively highly contaminated. The
fact that most samples have not been taken at random, jeopardizes the
proper assessment of the exposure of the general population.

5.4 General population exposure

5.4.1 Air

Relatively high levels of PCBs have been detected in indoor air,
especially in kitchens and offices with electric installations
(Jensen, 1983a) (section 3.2.4 and 5.1.1).

Results from the US EPA indicate PCB concentrations in the air ranging
from 1 up to 50 ng/m³; similar results have been reported from Japan
(WHO, 1976). Assuming a level of 5 ng PCBs/m³ in urban air, a
breathing rate of 22 m³/day, retention and absorption of inhaled
particles/vapour of 50%, and a mean residence time of PCBs in the body
of 3 years, air would contribute 0.8 µg/kg to the PCB concentration in
the body. Higher concentrations of PCBs in indoor air could increase
this estimate (WHO/EURO, 1988).

Van der Kolk (1985) calculated air intake through inhalation for the
Dutch population of about 36 ng/day, a quantity approximately 1000
times lower than the intake with food.

During the manufacture, formulation, or use of PCBs, where levels in
the workroom air correspond to exposure limit values, varying between
0.1 mg/m³ and 1 mg/m³, the calculated mean intakes would range
between 1 and 10 mg during an 8-h workshift. In some occupational
situations, much higher concentrations have been measured and the
estimates of intakes would be higher (WHO/EURO, 1987).

5.4.2 Drinking-water

Levels reported in drinking-water are typically between 0.1 and
0.5 ng/litre. Even assuming a PCB level of 2 ng/litre in
drinking-water, consumption of 2 litre/day contributes 0.04 µg/kg body
weight to the PCB concentration in the body. This additional quantity
is negligible in comparison with the intake via food (WHO/EURO, 1988).

5.4.3 Intake by infants through mother's milk

The daily intake of PCBs was calculated in breast-fed infants in the
countries participating in a monitoring study by Slorach & Vaz (1983,
1985) (Table 19).

The intakes in EEC countries were calculated to range from 3 to
11 µg/kg body weight per day, compared with 0.12-0.3 µg/kg body weight
for bottle-fed infants in Denmark (WHO/EURO, 1985).

In Yusho infants with clinical symptoms of poisoning, the daily intake
of PCBs with breast milk was calculated to be 70 µg/kg body weight
(Jensen, 1983b) (see section 9.1.2.2).

5.4.4 Infant and toddler total diet

Johnson et al. (1979) analysed the average diet of 6-month-old infants
and 2-year-old toddlers for the presence of PCBs. Ten market baskets
were collected in 10 cities in the USA. The foods were prepared in the
manner in which they would be prepared and served in the home. Trace
amounts of PCBs were detected in only one infant and one toddler diet.

In the USA, Gartrell et al. (1986b) found a daily intake of 0.011 µg
PCBs/kg body weight in infants consuming infant diets in 1978. In the
years 1979, 1980, and 1981/82, the intake was below the detection
level. The intake by toddlers was 0.099 µg/kg body weight in 1978 and
not detectable in the following 3 years.

Tuinstra et al. (1985a) analysed samples of infant food from the Dutch
market and found average PCB levels of 0.1-0.2 µg/kg food (the maximum
level found was 1.1 µg/kg).

5.4.5 Total intake by adults via food

The oral consumption of contaminated products is presumed to be the
main route of exposure to the PCBs.

Table 19. Calculated daily intakes of PCBs by breast-fed infants (µg/kg body weight)^a

Country/area Year(s) Calculation according Calculation according
to US FDA method^b to national method^c

median maximum median maximum

Belgium,
Brussels 1982 3.6 10.4 NR NR

China,
Beijing 1982 NR NR 0.45^d 0.45^d

Israel,
Jerusalem 1981/82 2.0 9.5 NR NR

Germany,
Hanau 1981 NR NR 9.5 45

Japan,
Osaka 1980/81 1.6 4.4 2.3 6.3

Sweden,
Uppsala 1981 4.4 8.1 5.9 11

USA
22 states 1979 4.5^e 13.5 4.5^e 22.5

Yugoslavia,
Zagreb 1981/82 2.8 7.2 2.8 7.7

^a Assuming a milk consumption of ca 130 g/kg body weight and a milk fat content of
3.5% (w/w). Calculations based on data for all mothers studied. Results for
different methods of PCB analysis shown separately.
From: Slorach & Vaz (1983, 1985); Van der Kolk (1984b).
^b Sawyer method.
^c "Own method".
^d PCB level below limit of detection (0.1 mg/kg fat) in milk samples.
^e PCB level below limit of detection (1 mg/kg fat) in milk samples.
NR = No data on levels in milk reported.

It has been stated that the major part of the human dietary intake of
PCBs is from fish (Berglund, 1972; Hammond, 1972). This may well be
true in areas such as Japan or certain localities near the North
American Great Lakes, where fish from polluted waters may form a
relatively large part of the diet. Several investigators from Japan
have measured the daily intake of PCBs in food; the highest mean value
recorded was 48 µg/day, of which 90% was from fish (Kobayashi, 1972);
the lowest was 8 µg/day (Ushio et al., 1974).

In much of Europe and North America, however, the daily intake of fish
is in the region of 30-40 g, and most of the fish is taken from waters
of low pollution with PCB levels in the fish not exceeding 0.1 mg/kg.
Berglund (1972) has estimated that the daily intake of PCBs from fish
in Sweden is in the region of 1 µg, though if the fish consumed were
solely Baltic herring, the intake would be about 10 µg/person. It is
difficult to make an assessment of the PCB intake from foods other
than fish. Westöö et al. (1971) in their extensive study of the
Swedish diet, reported that most foods contained PCB levels of less
than 0.1 mg/kg; and concluded that this corresponds to a daily intake
of less than 100 µg.

Weekly intakes in the range of 23-889 µg/person have been reported
from the USA (OTA, 1979). The higher range concerns people consuming
more than 12 kg/year of Lake Michigan fish.

The intake of total PCBs by the general adult population depends
greatly on the geographical area and food habits.

5.4.6 Total diet/market-basket studies

Data on total-diet studies of PCBs have been reported from a few
countries. These reported intakes show a wide variation, which can
partially be explained by methodological factors, such as the ways in
which samples below the limits of determination are considered,
especially when noting the different limits of determination.
Considering the available data, an average intake of 5-15 µg/day for
the non-occupationally exposed population in industrialized countries
may be the best available estimation.

These estimates apply to the average diet of an average adult citizen.
In practice, few people are really "average" in their consumption
pattern. Given the widespread nature of the contamination, however, a
higher intake in one food group is more or less balanced by a lower
intake in another food group with an equal calorie intake. Total
intake will certainly be higher for diets with a more than average
calorie content (van der Kolk, 1985).

Gartrell et al. (1985) determined the total intake of PCBs by 16- to
19-year-old males in the USA. The samples represent a typical 14-day
diet. Approximately 120 individual food items (of 12 food groups),
including drinking-water, were collected for each market-basket sample
in 20 cities in the period 1979-80. Only 2 samples of meat, fish, and
poultry contained PCBs with an average concentration of 0.002 mg/kg.

Gartrell et al. (1985, 1986a) reported a daily intake in the USA of
0.016, 0.027, 0.014, 0.008, and 0.003 µg PCBs/kg body weight during
the years 1977, 1978, 1979, 1980, and 1981/82, respectively.

Manske & Johnson (1975) collected 35 market baskets in 32 cities over
the period 1971-72. PCB residues were found in the range of
0.035-0.15 mg/kg in 51 composites. Fish and oils, fats, and
shortenings contained the highest levels. The same authors (Manske &
Johnson, 1977) carried out a market-basket study representing the
basic 2-week diet of a 16- to 19-year-old male. The various foods were
prepared in the manner in which they would normally be served and
eaten. Thirty market-baskets, containing 12 classes of foods (in total
360 composites) were collected in 30 cities in the period 1973-74. A
trace of PCB was found once in whole milk, ground beef, and fish
fillet.

The FDA revised the concept of the Total Diet Study in 1982. As
discussed by Gunderson (1988b), the Total Diet Study conducted before
1982 was based on a "composite sample approach", regardless of the
diet involved. The revised study is based on updated dietary survey
information and allows the "total diet" of the US population to be
represented by a relatively small number of food items for a greater
number of age/sex groups. The daily intake expressed in ng/kg body
weight per day for PCBs (Aroclor 1221, 1242, and 1254) in 1982-84 for
the age groups 6-11 months, 2 years, 14-16-year females, 14-16-year
males, 25-30-year females, 25-30-year males, 60-65-year females and
60-65-year males were: 0.8, 1.2, 0.4, 0.5, 0.5, 0.6, 0.4, and
0.5 ng/kg body weight per day, respectively (Gunderson, 1988b).

Foods, representative of Canadian eating habits, as determined by a
national nutritional survey, were prepared for eating, categorized,
and blended into 11 different composites representing the dietary
intake for 5 cities over the period 1976-78. It concerned 194 samples,
collected in winter and in summer. The average dietary intake was
0.001 µg PCB/kg body weight (McLeod et al., 1980).

Over a period of 2 years, 126 different food items of a market-basket
of 16- to 18-year-old males were purchased every 2 months in the
period 1976-78, in the Netherlands. The foodstuffs were prepared for
eating and were combined in 12 commodity groups. The mean
concentration and range of PCBs in 5 food classes was:

Class Mean concentration Range
(mg/kg on fat basis)

Meat, poultry, and eggs - 0.13-0.17 (2)^a
Fish 0.07 0.04-0.24 (7)
Dairy products - 0.04-0.06 (2)
Sugar and sweets - 0.08 (1)
Drinks, drinking-water - 0.035 (1)

^a In parentheses: number of positive composites.

The authors calculated a daily intake of PCBs of 15 µg/person (a
maximum level was 90 µg/person (de Vos et al., 1984). In the period
May-July 1976, 100 total diets (summer meals) were collected and
besides organochlorine pesticides, PCBs were determined as
decachlorobiphenyl, after perchloration, and calculated as Aroclor
1260. The mean intake of PCBs/person per day was 11.6 µg with a range
of 3-71 µg (Greve & van Hulst, 1977; Greve & Wegman, 1983; van der
Kolk, 1985).

In 1978, another survey was carried out with 100 total diets during
the winter (winter meals). It was estimated that the daily intake was
6 µg/person (range 1-19 µg).

Zimmerli & Marek (1973) studied the total human intake of PCBs from
prepared meals in 1971-72 in Bern, Switzerland. Five typical total
diets were composed and analysed. The intake of PCBs, especially with
daily diets containing cheese, meat, fish, or fat, ranged from 6 to
84 µg.

According to a calculation by Summerman et al. (1978), the average
weekly intake of PCBs in the Federal Republic of Germany was about 215
and 268 µg/week for females and males, respectively. Much lower
figures, 36-44 µg/week, were calculated by Klein (1983).

A survey of the daily PCB intake from the total diet of Japanese women
(number of samples varied from 18 to 60) was performed for the years
1972-76. The daily intake of PCBs averaged approximately 10 µg/person
(range 2.8-21.2 µg). The main source of PCBs in the diet of Japan was
in-shore fish. There was no clear change in daily intake over the
5-year period studied (Watanabe et al., 1979).

Ushio & Doguchi (1977) studied the dietary intake of PCBs in Tokyo.
They found an average daily intake of PCBs of 6.3 µg/person (range,
trace-17 µg/person). It was concluded that the dietary daily intake of
PCBs for the majority of the population of Tokyo rarely exceeded
20 µg/person, when no heavily contaminated fish were consumed.

Yakushiji et al. (1977) found that the PCB daily intake through meals
of unexposed adults living in Osaka prefecture, was 3-20 µg/day.

Data for PCBs in the diets of Canada, Guatamala, Japan, the United
Kingdom, and the USA over the period 1972-83 were summarized by
Gorchev & Jelinek (1985). The mean dietary intake reported was at, or
below, 0.06 µg/kg body weight, the mean intake per person ranged from
< 0.01 to 0.12 µg/kg body weight (Slorach et al., 1982; WHO, 1986b).

5.4.7 Total intake of major congeners by adults via food

In the Federal Republic of Germany, the daily intake of the 3 PCB
congeners numbers 138, 153, and 180, together with the different food
items, was calculated. The intake (µg/day) with meat and meat products
was 0.30; with fish and fish products 0.36; eggs and egg products
0.008; milk and milk products 0.40; cheese 0.11; butter 0.39; fats and
oil 0.098; bread and pastries 0.17; potatoes 0.081; vegetables 0.11
and fruits 0.082 (DFG, 1988).

5.4.8 Time trends in different matrices

Although many countries introduced severe restrictions on the
manufacture, use, and disposal of PCBs many years ago, it is difficult
to discern any marked decline in the levels in human milk fat, from
the published data.

Levels of PCBs were estimated in 1085 samples of different cereals,
collected in the Federal Republic of Germany over the period
1972/74-1984. The levels, which were the highest in 1972/74 0.04 mg/kg
(0.005-0.12 mg/kg), decreased during the years to 0.004-0.005 mg/kg
dry weight in 1984 (DFG, 1988).

Data from the Federal Republic of Germany showed no clear trend in PCB
levels in human milk during 1975-79 (Slorach et al., 1982). The same
was found in the Netherlands over the period 1974-83 (Greve & Wegman,
1984).

Japanese data showed a decline in PCB levels in the fat of whole cow's
milk during the period 1972-79. A decline was also found in PCB levels
in finfish from coastal waters and in total marine fish (Slorach et
al., 1982).

A downward trend was found in human milk from Japan over the period
1972-80. Each year, a large number of samples (361-877 samples/year)
were analysed. In 1972, the median level was about 0.8 mg/kg and, in
1980, 0.5 mg/kg, on a fat basis. A gradual decline was observed
(Slorach & Vaz, 1983).

In Canada, human milk and adipose tissue from Ontario residents were
analysed over the period 1969-74. The values found did not indicate a
trend in this period.

The mean total PCB intakes determined in the FDA Total Diet Study, for
the period 1971-87, for a typical "adult" diet, represented in Fig. 4,
reflect that of a 14- to 16-year-old male during 1982-87. A clear
decline was shown from approximately 7 µg/person per day to less than
0.1 µg/person per day (Gunderson, 1988a).

The daily intake of PCBs, expressed as ng/kg body weight per day, by
6-month-old and 2-year-old children in the years 1980, 1981/82, and
1982/84 did not show a trend, while, in adults, a decrease from 8 to
0.5 ng/kg body weight per day was observed over the same years
(Gunderson, 1988b).

5.5 Concentrations in the body tissues of the general population

The PCB levels in body tissues are a good indication of the overall
and total exposure of the body to PCBs.

Several factors may influence the concentrations of PCBs in body
tissues, including duration and level of exposure, the route and
pattern of exposure, the chemical structure of the PCB (degree and
position of chlorination in the molecule), the amount of adipose
tissue, other simultaneous exposures, as well as other biological
parameters.

5.5.1 Adipose tissue

In general, while highly chlorinated congeners accumulate more easily,
a lower degree of substitution provides more possibilities for
hydroxylation and facilitates excretion. Factors other than the degree
of substitution also affect accumulation, particularly the position
and pattern of substitution (WHO/EURO, 1987).

The available information on the occurrence of PCBs in the body fat of
the general population is summarized in Table 20.

Table 20. Concentrations of PCBs in the body fat of the general population

Country Year Number of samples Mean concentration in mg/kg Reference
on fat basis (range)

North America

USA (18 states) - 637 < 1 (68.9%)^e Yobs (1972)
< 1-2 (25.9%)^e Price & Welch (1972)
> 2 (5.2%)^f

Northeast Louisiana 1980 8 1.04 (0.38-2.33) Holt et al. (1986)
1984 10 1.23 (0.65-1.44)

Texas 1969-1972 88 (15 positive) 1.7 (0.6-9.9) Burns (1974)

New York - 101 (women) 3.4 ± 1.1 Bush et al. (1984)

(urban and rural
vicinity)

Canada - 99 0.94 (0.04-6.8)^a Mes et al. (1982)

Ontario 1976 and 570 2.1-2.2 Frank et al. (1988)
1984

Table 20. (cont'd).

Country Year Number of samples Mean concentration in mg/kg Reference
on fat basis (range)

Asia

Japan (Kochi area) - - 2.86 (maximum 7.5) Nishimoto et al. (1972a,b)

Japan 1971-1982 - 0.5-6.0^a Katsunuma et al. (1985)

Tokyo 1974 30 1.04 (0.38-2.5) Fukano & Doguchi (1977)

Japan - 241 0.30-1.48 Curley et al. (1973b)

New Zealand - 51 0.82 Solly & Shanks (1974)

Africa

South Africa 1982 63 0.15-5.18 van Dijk et al. (1987)

Europe

Austria (Vienna area) - 32 0.3-7.3 Pesendorfer et al. (1973)

Finland - 105 0.2 Mussalo-Rauhamaa et al.
(1984)

Germany, Federal - 20 5.7 Acker & Schulte (1970)
Republic of - 282 8.3 Acker & Schulte (1974)
1982-1983 50^b 0.5-1.5 Niessen et al. (1984)

Italy (Siena) 1983-1984 26 1.75^c (dry weight) Focardi et al. (1986)

Table 20. (cont'd).

Country Year Number of samples Mean concentration in mg/kg Reference
on fat basis (range)

Netherlands 1973-1983 24-78 per year 1.6-2.5^d Greve & van Harten
(1983a);
Greve & Wegman (1983,
1984)

Norway (Oslo) - 40 1.6 Bjerk (1972)

Spain 1985-1987 14 1.68 Camps et al. (1989)

United Kingdom - 201 < 1.0 Abbott et al. (1972)
1976-1977 236 0.7 (nd-10) HMSO (1986)
1982-1983 187 0.9 (0.1-6.9)

^a Wet weight.
^b 34 infants, 14 children, and 2 older children.
^c About 60% included only five congeners: Nos. 118, 138, 153, 170, 180.
^d Median.
^e Percentage of samples.

5.5.1.1 PCBs in the fetus

PCBs are also present in serum and all organs of the body in
proportion to their fat content. PCBs pass more, or less (depending on
structure and chlorination), through the placenta into the fetus.
Since the fetus has little adipose tissue until 7 months of age, PCB
concentrations may be higher in vital organs, such as the adrenal
gland, but available data suggest somewhat lower levels in the brain
(Masuda et al., 1978a; Kodama & Ota, 1980).

Masuda et al. (1978a) found PCB levels of 270-960 µg/kg fat in adipose
tissue samples of fetuses beyond 7 months of gestation. Levels in the
adipose tissue of adult females from the same geographical area ranged
from 270 to 1360 µg/kg fat. The mean concentrations were 470 µg/kg for
fetuses and 780 µg/kg for adult females. However, since the ranges
showed an overlap and the number of samples was small, it is not clear
whether this represents a true difference.

5.5.1.2 Congeners in adipose tissue

Wegman & Berkhoff (1986) investigated the presence of the different
congeners in 24 human fat samples, collected in 1984. The following
congeners were present at the highest levels: 2,4,4'-trichloro-,
2,4,5,2',5'-pentachloro-, 2,4,5,3',4'-pentachloro-, 2,3,4,2',3',4'-
hexachloro, 2,3,4,2',4',5'-hexachloro-, 2,4,5,2',4',5'-hexachloro-,
2,3,4,5,2',4',5'-heptachloro, 2,3,4,5,2',3',4',5'-octachloro, and
2,3,5,6,2',3',5',6'-octachlorobiphenyl.

Focardi & Romei (1987) analysed 30 samples of adipose tissue,
obtained from patients in Siena, Italy, in 1986, for the presence of
19 PCB congeners. The results indicate that the mean PCB (as sum of
the congeners) concentration was 1063 µg/kg dry weight (range
391-1918 mg/kg). The major constituents of the PCBs (about 60%) were
the isomers 99, 138, 153, 170, and 180.

Human adipose tissue was analysed for 3 non- ortho chlorine
substituted coplanar congeners: 3,4,3',4'-tetrachloro-, 3,4,5,3',4'-
pentachloro- and 3,4,5,3',4',5'-hexachlorobiphenyl (Kannan et al.,
1988). Twelve samples, from 7 male and 5 female persons were obtained
from hospitals. The average total PCB concentrations were 1.22 and
1.02 mg/kg (wet weight basis), respectively. The concentrations of the
3 congeners were 94-860, 120-730, and 36-200 ng/kg, on a wet weight
basis, respectively.

5.5.2 Blood of the general population

Finklea et al. (1972) studied human plasma of different races of the
population (723 volunteers with ages ranging up to 60 years) of urban
and rural areas of South Carolina. The average concentration was
5 µg/litre (range 0-29 µg/litre). No age effect was found, but ethnic

differences and ethnic residence interactions were significant. Kreiss
(1985) found mean serum concentrations in the non-occupationally
exposed population in the USA, of between 4 and 8 µg/litre, with 95%
of the individuals having serum PCB concentrations of less than
20 µg/litre. More data are summarized in Table 21.

Maternal blood and fetal cord blood were collected from volunteers
from an urban and rural vicinity in upstate New York. Whole blood
samples were taken from 101 women (26 ± 4 years) entering maternity
facilities. Maternal blood contained 3.4 ± 1.1 µg PCBs/kg and fetal
cord blood contained 2.4 ± 1.0 µg/kg whole blood. The PCB congeners
making up these totals were surprisingly few; 38% of the total residue
in the maternal blood and 21% of the fetal cord blood comprised only 4
components, 2,4,4'-trichlorobiphenyl, 2,4,5,2',4',5'-hexachloro-,
2,3,4,2',4',5'-hexachloro-, and 2,3,5,6,2',3',6'-heptachlorobiphenyl.
The congener 2,5,2',5'-tetrachlorobiphenyl crossed the placenta
preferentially (Bush et al., 1984).

The concentrations of PCBs were determined in blood samples from 120
women hospitalized for miscarriages and 120 full-term pregnancy
controls. The average PCB level was higher in women with miscarriages
than in control women (8.65 µg/litre and 6.89 µg/litre, respectively,
as Fenclor 54 and 14.81 and 14.90 µg/litre, respectively, as
decachlorobiphenyl). The reproductive history of each woman was
assessed together with confounding variables and with environmental
exposure and food intake. Food consumption did not indicate diet as
the main source of PCB intake (Leoni et al., 1989).

A cross section of the population of Michigan was studied following an
accidental exposure in 1978. Five years after the accident, PCB and
PBB residues were measured in adipose tissue and serum. Serum levels
of PCB were measured in 1681 adults and 1462 children. Children (430)
were found to have uniform levels throughout the state (mean
concentration 4 ± 2 µg/litre). In adults, the serum PCB levels were
higher in the area with highest PBB levels. The mean serum PCB level
was 21 µg/litre, compared with control levels for the rest of the
state of 9 µg/litre. No sex difference was found (Wolff et al.,
1982a).

Table 21. Concentrations of PCBs in whole blood of the general population

Country Year Number of samples Mean concentration in Reference
µg/litre (range)

Canada

Ontario area 1975-1976 118 18 Frank et al. (1988)
(patients suspected 1980-1981
of being exposed 1984
dermally)

Japan

- - 3.2 Doguchi & Fukano (1975)

- 28 (women) 2.6 Kuwabara et al. (1978)

(Osaka area) 1976 16 (women) 2.8 (1.7-4.6) Kuwabara et al. (1979)
1972-1977 - 3-4 Yakushiji et al. (1977)

farmers 1978-1983 - trace-21.4^b Katsunuma et al. (1985)

Tokyo 1973 27 3.19 (2.2-5.1) Fukano & Doguchi (1977)
1975 10 2.59 (1.8-3.8)

Table 21. (cont'd).

Country Year Number of samples Mean concentration in Reference
µg/litre (range)

Finland - 3.1-12 Karppanen & Kolho (1973)

Netherlands - 34 (women) 4.5 (nd-11.6) Blok et al. (1984)
31 (men) 4.8 (1.0-17.1)

1978 48-127 3.1^e Greve & Wegman (1983,
1980 samples/year 3.5 1984)
1981 4.4
1982 4.4

North America

South Carolina 1968 723 5 (4.2-5.5)^a Finklea et al. (1972)
(urban and rural
area)

Michigan 1973 1100 56^c Kreiss (1985)
(areas of Lake 1979-1981 17.2-23.6^c
Michigan)

Lake Michigan 1985 196 5.5 ± 3.7 Schwartz et al. (1983)
(high fish
consumption)

Table 21. (cont'd).

Country Year Number of samples Mean concentration in Reference
µg/litre (range)

Yugoslavia 1984-1986 10^f 155 (35-480)^d Jan & Tratnik (1988a)
(residents around 19^g 11 (6-18)
River Krupa; 4^h 5 (2-7)
contamination by a
plant using PCBs)

^a Plasma.
^b Serum.
^c Geometric mean.
^d Arithmetic mean.
^e Median concentration.
^f Living close to plant.
^g Living 1-3 km from plant.
^h Non-exposed other areas.

Specific PCB isomer levels in the blood of 30 children, ages 2-5
years, residing in an area of PCB-contaminated soil in Canada, were
compared with those of 25 children in a non-contaminated area. The sum
of individual PCB isomer levels in the exposed and non-exposed group
were not significantly different, e.g., 0.54 µg/litre (range
0.22-0.99 µg/litre) and 0.88 µg/litre (range 0.28-2.30 µg/litre). The
major component in both groups was 2,4,5,2',4',5'-hexa-chlorobiphenyl
(Mes, 1987).

High levels of PCBs were found in the blood (up to 100 µg/litre) in
patients with severe weight loss (Hesselberg & Scherr, 1974). This was
attributed to the release of PCBs from the mobilization of fat.

Greve & van Harten (1983b) studied the relationship between the levels
of PCBs in the adipose tissue and in the blood of the same persons. A
total of 48 persons were involved in this study. A concentration
factor (concentration in adipose tissue divided by concentration in
blood) of 660 was found.

5.5.3 Human milk

Human milk is the major source of exposure for breast-fed infants. The
amount of human milk secreted varies widely. The composition of the
milk is related to the amount secreted, the stage of lactation, the
timing of withdrawal (early or late in feeding) and to individual
variations among lactating women. The individual variations depend on
maternal age, health, social class, and diet. The concentration of
PCBs depends primarily on the lipid concentration in milk. Wide
variations in published results are caused by inaccuracies inherent in
the analytical methods used for the quantification of lipids, and
whether the milk sample is collected early or late during the feeding
period. The fat content increases during emptying, and the fat content
of milk from the 2 breasts may differ. According to a recent
determination, the fat level in human milk averages 2.6-4.5%
(WHO/EURO, 1988).

Whether the differences in concentration in various countries are
merely a function of the analytical methods used and the type of
samples collected or whether true differences in body burden exist, is
not clear at present. For instance, some countries have reported
levels of PCBs in human milk fat ranging from nondetectable to
14 mg/kg, while, in other countries, the highest levels found have
been around 3 mg/kg. Because of these variations, calculating an
average dose for nursing infants is difficult. The same difficulties
exist when attempts are made to investigate trends over time
(WHO/EURO, 1988).

The results of the older studies have been obtained with a less
sophisticated method using packed column GC. With this method only a
dozen peaks can be separated. The quantitative results are reported as
"total PCB values", though different techniques of quantification and
different types of calculations were used.

In contrast with the situation with many organochlorine insecticides,
the levels of PCBs in human milk fat are higher in European countries,
Japan, and the USA than in China (Slorach & Vaz, 1983, 1985), and are
significant, particularly in the highly industrialized countries.
Results from a large number of countries have been summarized by
Jensen (1983a, 1985, 1987), Acker et al., (1984), Katsunuma et al.
(1985) (especially Japanese data; period 1972-83); and WHO/EURO,
(1987, 1988). The countries concerned are: Argentina, Austria,
Belgium, Canada, Finland, France, Federal Republic of Germany (Klein,
1983), German Democratic Republic, Israel, Japan, the Netherlands,
Norway, Poland, Romania, South Africa, Sweden, Switzerland, Turkey,
United Kingdom, USA, USSR, and Yugoslavia. The average levels of PCBs
in human milk do not appear to differ very much between the
industrialized countries and range between 0.5 and 2 mg/kg milk fat,
except in Czechoslovakia, the Federal Republic of Germany, India,
Denmark and Italy, where levels up to 3 mg/kg milk fat were found
(Jensen, 1983b; Acker et al., 1984) (Table 22).

The variation in residue levels in human milk during lactation was
investigated in 5 women in the Federal Republic of Germany. Month-mix
samples, composed of breast milk samples collected weekly, were
analysed over a lactation period of between 5 and 9 months. The ages
of the women ranged from 23 to 36 years. The PCB concentrations were
between 0.61 and 2.20 mg/kg, on a fat basis. While the concentrations
remained relatively constant, some fluctuations were seen but no trend
was observed over the lactation period investigated (Fooken & Butte,
1987).

Breast milk samples from 16 women in Canada were analysed for PCBs at
8 intervals (7, 14, 28, 42, 56, 70, 84, and 98 days) during the
lactation period. The average PCB concentrations in breast milk varied
between 22.8 and 29.7 µg/kg whole milk. No clear decrease or increase
was observed. The average milk/blood ratio for PCBs was 23 and
remained relatively constant during lactation (Mes et al., 1984).

Wolff (1983) reported the half-life of PCBs (percentage chlorine not
specified) in breast milk to be 5-8 months and found that the
concentration of PCBs in breast milk was 4-10 times that in the
maternal blood. Similar results were reported by Jacobson et al.
(1984b).

Table 22. Concentrations of PCBs in breast milk of the general population

Region Year Number of Mean concentration in Reference
Country samples mg/kg on fat basis (range)

North America

USA (Michigan) 1977-1978 1057 1.5 (maximum 5.1) Wickizer et al. (1981);
Wickizer & Brilliant (1981)

Canada (Quebec) - 154 0.84 (nd-4.34) Dillon et al. (1981)

Ontario 1971-1974 - 1.2 (0.1-3.0) Atkinson (1979)
1978 215 0.6 ± 0.3

Ontario 1975-1985 348 0.023 (0.016-0.033)^a Frank et al. (1988)

Five regions across Canada 1982 210 0.697 Mes et al. (1986)

Regina, Saskatchewan 1979 80 0.0052 (0.001-0.019)^a Qureshi & Robertson (1987)

Table 22. (cont'd).

Region Year Number of Mean concentration in Reference
Country samples mg/kg on fat basis (range)

Asia

Japan (Osaka) 1972-1977 - 0.030-0.040^a Yakushiji et al. (1977)
1969-1976 19-52 each year 1-2 Yakushiji et al. (1979)

India (Ahmedabad) 1981-1982 50 not present Jani et al. (1988)

Hawaii (different islands) 1979-1980 54 0.80 ± 0.43 (0.13-2.2) Takei et al. (1983)

Europe

Germany, since 1970 several thousands 1.0-2.5 (98% of samples Acker et al. (1984);
Federal Republic of between 0.001-7.2) Cetinkaya et al. (1984);
Heeschen et al. (1986);
Lorenz & Neumeier (1983)
- 2709 1.77 Fooken & Butte (1987)

Netherlands 1983 278 0.72 (0.27-2.20)^b Greve et al. (1985);
(11 centres country-wide) Greve & Wegman (1984)
1977-1979, 2649 2.1 Olling (1984)
1981

United Kingdom (Scotland) 1979-1980 30 0.01 (nd-0.04) HMSO (1986)
1983-1984 30 < 0.01 (nd-0.02)

Table 22. (cont'd).

Region Year Number of Mean concentration in Reference
Country samples mg/kg on fat basis (range)

Italy (Rome) 1983-1985 65 0.070 (0.007-0.176)^a,c Dommarco et al. (1987)

Finland (different parts) 1984-1985 183 (165 of 0.57 (0.05-10.7) Mussalo-Rauhamaa et al.
women) (1988)

Sweden (5 regions) - 300^e 1.06-1.18 (four regions) Noren (1983)
1.44 (one region)
1972 227^d 1.05 Noren (1988)
1976 245 0.99
1980 340 0.78
1984-1985 102 0.60

Austria (Vienna) - 22 1.54 (0.58-3.78) Pesendorfer (1975)

Other regions 9 1.29 (0.95-1.57) Pesendorfer (1975)

^a Whole milk.
^b Median concentration.
^c Arithmetic mean.
^d Number of mothers that provided 4-7 samples each (samples were pooled).
^e In each region, 300 mothers gave breast milk 3-5 days after parturition.

In a study by Kuwabara et al. (1978), the relationship was
investigated between breast-feeding and PCB residues in the blood of
children whose mothers were occupationally exposed to PCBs. The
children ingested their mother's milk for periods of < 1 to 3 years.
The age of the children at the time of the study ranged up to 13
years. The data provide evidence that PCBs are retained in the
children's body for many years and that longer intake of mother's milk
tends to increase PCB levels in the blood of the children. The PCB
levels in the blood of the 20 occupationally-exposed women and their
39 children ranged from 8.3 to 84.5 and 0.8 to 93.2 µg/litre,
respectively.

The results suggest that the PCB levels in the blood of children are
much more influenced by the transportation of PCBs through the
mother's milk than through the placenta. Furthermore, it was found
that the gas chromatographic patterns of the blood PCBs of the
children, breast fed for a long time, were different from those of
their mothers. Blood from 16 non-occupationally exposed mothers and
their children (17), showed that, as the length of the breast-feeding
period increased, there was an increase in the PCB levels in the blood
of the children. The mean blood PCB level in mothers was 2.8 ±
0.8 µg/litre; in children, it was 3.8 ± 3.6 µg/litre. In this study,
no clear change in blood PCBs patterns between mothers and children
was observed (Kuwabara et al., 1979).

Samples of maternal blood, milk, and umbilical cord blood were
collected from 43 mothers giving birth to their first or second child;
all the mothers had lived in Oslo during the previous 2 years. Blood
samples were collected immediately after delivery, either by Caesarean
section (16 Norwegians) or normally (20 Norwegians and 7 immigrants).
Subcutaneous fat samples were obtained during the operation. Samples
of colostrum and milk were obtained 3 and 5 days postpartum. PCBs were
found in 135 of the total 168 samples. In the Norwegian women and
infants, PCBs were the major contaminants, whereas only traces of PCBs
were found in the samples of immigrants. The average concentrations in
the maternal serum, cord serum, colostrum, and breast milk of
Norwegian women (Caesarean and normally delivered taken together)
were: 10, 3-5, 18-21, 20-23 µg/kg wet weight (Skaare et al., 1988).

5.5.3.1 Major PCB congeners in human milk

Commercial PCB preparations consist of complex mixtures of
environmentally stable compounds with a wide range of chlorine
contents. PCBs are transferred to breast-fed infants with the fat of
the mother's milk. Thus, infants nurtured on maternal milk are exposed
to relatively high concentrations of the higher chlorinated PCBs in
the short period preceding the full functioning of certain organs,
e.g., the liver (Jensen, 1983b; Slorach & Vaz, 1983; Gezondheidsraad,
1985).

Three major congeners were present in breast milk, e.g., PCB congener
numbers 138, 153, and 180 (DFG, 1988).

Slorach & Vaz (1983) reported that the GC patterns of PCBs in breast
milk samples from different countries were similar. The peaks denoted
146, 174, and 180 were dominant in the gas chromatograms. The total
levels of PCBs and the concentrations of certain congeners in Swedish
human milk, sampled in 1972-89, were studied by Noren et al. (1990).
Minor changes in the distribution of the congeners were found over the
period of study. The most abundant of the non- ortho coplanar PCBs in
Swedish human milk was 3,4,5,3',4'-pentachlorobiphenyl (126), with
levels decreasing from 0.35 µg/kg milk fat (1972) to about 0.10 µg/kg
(1989).

Safe et al. (1985a) analysed a sample of breast milk using the
congener-specific PCB method and found the following major components:
2,4,4'-trichloro-; 2,4,5,4'-tetrachloro-; 2,4,5,2',4'-pentachloro-;
2,4,5,3',4'-pentachloro-; 2,3,4,5,2',5'-hexachloro-;2,4,5,2',4',5'-
hexachloro; 2,3,4,5,2',3',4'-heptachloro-; and 2,3,4,5,2',4',5'-
heptachlorobiphenyls.

The major PCB congeners in the breast milk of Japanese women from the
general population were: 2,4,4'-trichloro-; 2,4,3',4'-tetrachloro-;
2,4,5,3',4'-pentachloro-; 2,3,4,2',3',4'-hexachloro-;2,3,4,5,2',4'-
hexachloro-; and 2,3,4,5,2',4',5'-heptachlorobiphenyls. The congeners
were present in 5% or more samples; a few other congeners were present
in only 1-3% (Gyorkos et al., 1985; Jensen, 1983b).

Sixty-eight breast milk samples collected in the Netherlands were used
to determine the congener distribution. The indicator congeners,
present in the highest concentrations, were: 2,4,4'-trichloro-,
2,4,5,2',5'-pentachloro-, 2,4,5,3',4'-pentachloro-,
2,3,4,2',4',5'-hexachloro-, 2,4,5,2',4',5'-hexachloro-,
2,3,4,5,2',4',5'-heptachlorobiphenyl (Wegman & Berkhoff, 1986).

Schecter et al. (1989a) analysed a total of 17 samples of human milk
from Thailand and Vietnam, for the presence of PCB congeners. The main
congeners that were present were 138, 153, and 180 (each in the range
of 8-31 µg/litre). The other congeners, normally present, were all
below the detection limit of 2 µg/litre.

In a study on pooled human milk samples from a 1982 nation-wide survey
in Canada, Mes & Marchand (1987) compared the relative amounts of 29
selected PCB isomers with amounts in milk samples of unexposed Rhesus
monkeys. In the pooled milk sample, 397 µg PCBs/litre, on a fat basis,
were found and the PCB isomer numbers 74, 99, 118, 138, 153, and 180
were the main contributors. Most of the predominant PCB isomers in
human milk were also observed in monkey's milk, but monkey's milk had
relatively low levels of PCB isomers numbers 74 and 99.

In another study, Davies & Mes (1987) analysed breast milk samples
from Canadian, Indian, and Inuit (Eskimo) mothers in Canada. The 18
samples were received from 5 Indian and Inuit nursing zones. The
combined total PCB isomer level (on a whole-milk basis) of the native
population was comparable with that of the national population. Even
the levels of the 5 largest PCB congeners (Nos. 74, 118, 138, 153, and
180) were comparable.

Individual congeners in the blood of Yusho- and Yu-Cheng patients are
discussed in section 5.6.

5.5.3.2 Factors that influence the intake of PCBs with milk

Present data suggest that the PCB content of human milk varies
considerably from individual to individual.

Many factors affect the level of PCBs and other organochlorine
compounds in breast milk including the fat content of the milk; time
from start of lactation; mother's age; mother's body weight; parity;
number of children previously breast-fed; origin and residence; eating
habits; season; smoking; use of household products; amount of milk;
and exposure at work (WHO/EURO, 1985, 1988).

In a given woman's milk, there are fluctuation in the PCB levels in
whole milk and in milk fat during one nursing session and during the
day (Jensen, 1983b). A decrease of PCB levels in both milk and milk
fat has been found during the lactation period. Furthermore, the PCB
concentration in human whole milk and milk fat increases with the age
of donor. Another confounding factor is that the PCB levels decrease
with increasing numbers of deliveries and lactations (Greve et al.
1985); lactation serves as a period for the biological elimination of
PCBs (Jensen, 1983b). The PCB levels in human milk are higher in
heavily populated and industrialized areas than in rural areas.
Furthermore, in general, the PCB levels in the breast milk of women
from developing countries are lower (Jensen, 1983b).

Cetinkaya et al. (1984) studied the PCB levels in human milk samples
from all over the Federal Republic of Germany. At the same time, data
were collected by means of a detailed questionnaire on residency,
workplace, smoking, drinking and eating habits, and the age of
participating individuals.

The breast milk of 45 women consuming lacto-vegetarian food was
compared with that of 41 women consuming conventional food in the
Federal Republic of Germany in the period 1979-81. The PCB
concentration was comparable, e.g., 2.2 and 2.5 mg/kg, on a fat basis,
respectively (Acker et al., 1984).

Fish consumption was positively correlated with PCB levels in maternal
serum and breast milk. PCB levels in serum increased with age, but
were unrelated to social class, parity, or body weight (Schwartz et
al., 1983).

Eight hundred and one Wisconsin anglers were surveyed for fishing and
consumption habits in 1985. The mean annual number of sport-caught
fish meals was 18 (range 7.1 to 33.3). The mean number of
non-sport-caught fish meals was 24. The median PCB serum congener sum
level for 192 anglers was 1.3 µg/litre (range, nd to 27.1 µg/litre).
Statistically significant positive Spearman correlations were observed
between sport-caught fish meals and PCB levels in serum and between kg
of fish caught and PCB levels in serum (Fiore et al., 1989).

PCBs were measured in maternal serum, cord blood, placenta, and serial
samples of breast milk and colostrum, from 868 women in North Carolina
(USA). Forty-three per cent of the women were primiparous. Breast milk
was collected at 6 weeks, 3 months, and 6 months, and, in a few cases,
up to 18 months postpartum. The median PCB concentration in breast
milk decreased during the sampling period from 1.77 to 1.02 mg/kg, on
a fat basis. The PCB concentration dropped by about 20% over 6 months
and 40% over 18 months. This implies that excretion in milk is a major
factor in lessening the mother's body burden; however, it also implies
substantial exposure of the child. Colostrum contained a median value
of 1.74 mg/kg. PCBs concentrations were higher in milk than in serum
and higher in maternal serum than in the placenta. The levels in cord
blood were almost always below the limit of quantification. Older
women and women who regularly drank alcohol had higher PCB levels in
their milk; blacks had higher levels than whites. In general, women
had higher levels in their first lactation and in the earlier samples
of a given lactation, and levels declined both with time spent
breast-feeding and with number of children nursed (Rogan et al.,
1986a).

Two hundred and forty-two newborn infants of mothers who consumed
moderate quantities of contaminated lake fish and 71 infants whose
mothers did not eat such fish were examined during the immediate post
partum period. PCB exposure was correlated with lower birth weight and
smaller head circumference, and the authors claimed that these effects
were not attributable to any of 37 potential confounding variables,
including socioeconomic status, maternal age, smoking, etc. (Fein et
al., 1984).

The mother's diet may be an important determinant of the PCB levels in
her milk. In some areas of the world, the intake of PCBs from eating
contaminated fish has been claimed to be the most important source of
PCBs in human milk. Dairy products and meat may be contaminated via
natural food or feedstuffs (WHO/EURO, 1988).

In a pilot study on the course of the PCB concentration in human milk
during 6 months of lactation, some PCB determinants were studied in 23
women and their infants. The average PCB concentration in the milk of
14 mothers during a 6-month period amounted to 0.66 ± 0.12 mg/kg, on a
fat basis. In univariate analyses, the PCB concentration on a fat
basis was strongly associated with pre- versus post-pregnancy weight
gain, age, and occupation. After multiple regression analysis, the PCB
concentration on a fat basis remained significantly associated with
changes in weight gain. The pre-pregnancy Quetelet Index of the mother
(height/weight) and the estimated PCB content of the diet (fish) were
correlated with the PCB concentration, on a milk basis (Drijver et
al., 1988).

5.5.4 Other tissues

Schecter et al. (1989b) analysed the tissues of 3 patients from the
North American continent, with no known history of chemical exposure,
for the presence of PCB isomers. The total PCB concentrations in the 9
tissues studied were different. The highest levels were found in
adipose tissue, subcutaneous fat (range 86-423 µg/kg), adrenals
(25-103 µg/kg), liver (3-149 µg/kg), bone marrow (26 µg/kg), kidneys
(2-31 µg/kg); levels in the spleen, lung, and testes were below
12 µg/kg. Congeners present in the highest concentrations were numbers
28, 74, 118, 153, 105, 138, 183, and 180.

5.6 Accidental exposures (Yusho- and Yu-Cheng)

In 1968, a large number of persons in Japan were accidentally poisoned
by the consumption of a batch of rice oil contaminated with Kanechlor
400. A similar accident happened in the Province of Taiwan in 1979,
where the affected persons had also consumed rice-bran oil
contaminated with PCBs. The 2 cases of poisoning were called Yusho and
Yu-Cheng accidents, respectively (see section 9.1.2.1).

The average PCB concentration in the plasma of Yusho children was
6 µg/litre, compared with 3.7 µg/litre in controls. Breast-fed Yusho
children had higher levels than children not breast-fed (Abe et al.,
1975).

The concentrations of PCBs in the adipose tissue, liver, and blood of
Yusho patients, about 5 years after the outbreak, were 1.9 ±
1.4 mg/kg, 0.08 ± 0.06 mg/kg, and 6.7 ± .3 µg/litre, respectively.
These values were only about twice those of controls. The mean blood
PCB level of 278 persons involved in the Yu-Cheng accident was
89.1 µg/litre (range 3-1156 µg/litre). Six months after the exposure,

the concentrations of PCBs in the blood had decreased to
12-50 µg/litre. The mean blood concentration of 165 patients,
9-18 months after the onset of poisoning, was 38 µg/litre (range
10-720 µg/litre) (see section 9.1.2.1). The blood PCB level of some
Yu-Cheng patients (99 ± 163 µg/litre), was much higher than that of
the Taiwanese population (1.2 ± 0.7 µg/litre), one year after the
outbreak of the intoxication.

Chen et al. (1985) analysed the blood of 165 Yu-Cheng patients, 9-18
months after the onset of poisoning, and found 10-720 µg PCBs/litre
with a mean value of 38 µg/litre. The blood of 10 patients, 9-27
months after poisoning, contained 0.02-0.2 µg PCDFs/litre. The
PCDF-congeners found in tissues were the same as those found by Masuda
et al. (1985).

Seven PCB congeners including: 2,4,5,3',4'-pentachloro-; 2,3,4,3',4'-
pentachloro-; 2,4,5,2',4',5'-hexachloro-; 2,3,4,2',4',5'-hexachloro-;
2,3,4,5,3',4'-hexachloro-; 2,3,4,5,2',4',5'-heptachloro-; and
2,3,4,5,2',3',4'-heptachlorobiphenyls, were identified in the blood
and tissues of Yusho, Yu-Cheng patients and controls.

Major PCDF congeners identified in the tissues and blood of Yusho and
Yu-Cheng patients were 2,3,6,8-tetrachloro-; 2,3,7,8-tetrachloro-;
1,2,4,7,8-pentachloro-; 2,3,4,7,8-pentachloro-; and 1,2,3,4,7,8-
hexachlorodibenzofurans. The 2,3,4,7,8-pentachloro-compound was
predominant. The concentrations of PCDFs in the adipose tissue and
liver of Yusho patients were 6-13 µg and 3-25 µg/kg tissue,
respectively. No PCDFs could be detected in the controls. Besides PCBs
and PCDFs, 4-methylthio-2,5,2',5'-tetrachlorobiphenyl (concentrations
ranging from 0.1 to 1.4 µg/kg tissue) and 4-methylsulfone-
2,5,2',5'-tetrachlorobiphenyl (range 0.3-2.5 µg/kg tissue) were also
found (Masuda et al., 1985).

5.7 Occupational exposure

5.7.1 Accidental exposure

Though the volatility of the PCBs is low, they are found in rather
high concentrations in the workroom air in both the long-term open use
of PCBs and in temporary or acute events where evaporation into the
air is possible. The measured air concentrations of PCBs in long-term
exposure situations, such as the manufacturing of transformers or
capacitors, varied from 30 to 1000 µg/m³, depending on the year of
measurement and the factory concerned (Silbergeld, 1983).

In discontinuous work, such as the inspection and repair of
transformers and capacitors, levels of between 0.1 and 60 µg/m³ have
been observed (Wolff, 1985). PCB concentrations in the breathing zone
of workers in transformer repair and maintenance work varied between
0.01 and 24.0 µg/m³ (Moseley et al., 1982).

In the atmosphere of an electroindustrial plant in Bela Krajina,
levels in the manufacturing room, where the autoclave was emptied,
averaged 2000 µg/m³ (range, 1400-3200 µg/m³); an average of
80 µg/m³ (range 40-120 µg/m³) was found in the working environment
in capacitor manufacture (Jan et al., 1988b).

Digernes & Astrup (1982) determined the concentrations of PCBs in the
atmosphere of the workplace of data screen operators, because skin
rashes and eczema had been reported among the workers. The PCB
concentrations in the working atmosphere (3 samples: concentrations
ranging from 0.056 to 0.081 µg/m³) were about 50-80 times higher than
the maximum level of PCBs in 3 samples collected outside the building
(0.0005-0.001 µg/m³). The indoor and outdoor samples also differed
qualitatively. The indoor samples contained only Aroclor 1242, while
outdoor samples contained a mixture of Aroclor 1242 and 1254.

Acute emergency events may cause extremely high concentrations of PCBs
in the air, particularly in cases when PCBs are burnt or heated (fire,
short circuit with electric arcing, burning in welding, etc.). Levels
of up to 10 000-16 000 µg/m³ have been measured. In the case of
extensive leaks of unheated PCBs from capacitors, concentrations of
1900 µg/m³ have been measured in workroom air (Elo et al., 1985;
WHO/EURO, 1987).

In connection with fires and electrical explosions, due to short
circuits, PCBs may be decomposed at elevated temperatures varying from
a few hundred to 2000°C. Soot may be produced in large amounts,
consisting of particles that may contain PCB concentrations up to
5000-8000 mg/kg of soot (Elo et al., 1985; O'Keefe et al., 1985;
WHO/EURO, 1987).

When evaluating PCB exposure, it is important to take into account
skin absorption from surfaces and tools, in addition to exposure via
inhalation. Surface concentrations of PCBs in capacitor factories have
varied between 4 and 60 µg/m², and, where PCB leaks have occurred,
levels of up to 30 mg/m² have been measured. Where PCBs have been
used long-term, contamination levels of 1-2 µg/cm² have been found on
tools and tables.

A transformer was found to have overheated and released an oily mist
containing PCBs and their pyrolysis by-products, in a Department
building in New Mexico. The transformer contained Askarel (87% Aroclor
1260 and 13% of a mixture of tri- and tetrachlorinated benzenes). The
3-storey building was extensively contaminated via the following ways:

* mist entered 2 rooms, adjacent to the basement in which the
transformer was located;

* direct spread of mist and fumes through stairways;

* air drafts created by open windows and exhaust fans, spreading
fumes throughout the building;

* foot traffic by employees and other persons;

* the exhaust vent of the transformer room, located near the intake
vents for the building's air-conditioning system.

Air samples obtained up to 14 h after the incident showed levels of
48 µg/m in the transformer vault and 20 µg/m³ in the room above the
vault. Wipe samples of surfaces showed PCB levels ranging from 30
million µg/m² for grossly contaminated surfaces to 4700 µg/m² for
surfaces without visible contamination.

Five to 7 days later, air and surface samples were analysed for
2,3,7,8-tetrachlorodibenzofuran (TCDF), which was found to be present
in the air at an average level of 48 µg/m³ in most contaminated
areas. In wipe samples, the levels ranged from 5 ng/m² to
41.224 ng/m². 2,3,7,8-Tetrachlorodibenzo- p-dioxin (TCDD) was not
detectable in either air samples (detection limit, 0.5-5.0 pg/m³ air)
or wipe samples (detection limit 180 ng/m²) (Anon., 1985).

Very high concentrations of these toxic chemicals may be found in soot
emitted in connection with fires and explosions in capacitors.

Thus, skin contamination, and the ingestion and inhalation of soot
particles, may result in serious exposure in PCB accidents and
emergencies.

A short-term, follow-up study was performed on 55 workers in a gear
plant, whose work did not involve the use of PCBs. Exposure was to the
total residual PCB left behind by a capacitor company that had
formerly (3 years before) used the site. Air samples contained
< 10 µg/m³ and mean concentrations in wipe samples ranged from 23 to
161 µg/100 cm². The 38 workers had a mean PCB concentration in serum
of 14.4 and the 17 office workers, 4.8 µg/litre. When the PCB
determinations were repeated in the 2 following years, no clear
decrease was observed (Christiani et al., 1986).

5.7.2 Occupational exposure during manufacture and use

Occupational exposure occurs during the manufacture of PCBs as well as
during their use by the electrical industry. It may also be widespread
among mechanics in contact with lubricating oils and hydraulic fluids,
among workers exposed to varnishes and paints, and among office
workers who have contact with pressure-sensitive duplicating paper
(carbonless copying paper), some brands of which readily transferred
PCBs to skin (Kuratsune & Masuda, 1972).

5.7.2.1 Adipose tissue

Levels of PCBs in the adipose tissue of occupationally exposed workers
have been found to vary between 26 and 50 mg/kg (range, 2.2-290 mg/kg).
There is a strong correlation between the blood PCB concentration and
PCB levels in adipose tissue, but the distribution of the various
congeners between plasma and adipose tissue is not the same, as
described above.

Emmett (1985) found the following congeners in the adipose tissue of
present and past transformer workers exposed to Aroclor 1242 and 1254:
2,4,3',4',5'-pentachloro-, 2,3,4,3',4'-pentachloro-, 2,3,4,5,2',4'-
hexachloro-, 2,3,4,6,3',4'-hexachloro-, 2,4,5,3',4',5'-hexachloro-,
2,3,4,5,2',3',4'-heptachloro-, and 2,3,4,5,6,3',4'-hepta-
chlorobiphenyl.

5.7.2.2 Blood

Karppanen & Kolho (1973) analysed the blood of 26 persons, 9
non-exposed, 6 persons handling PCBs, and 11 persons employed for 4
years in a capacitor-manufacturing plant in Finland. In the latter
case, Aroclor 1242 was used. The average concentrations in the blood
of the 3 groups were 7.1 µg/kg (3.1-12 µg/kg), 49.5 µg/kg
(36-63 µg/kg), and 440 µg/kg (70-1900 µg/kg), on a wet weight basis.

More recent results of a Finnish control group of workers indicated
serum PCB levels of 1.2 ± 0.6 µg/litre in an industrial area (Luotamo
et al., 1985; WHO/EURO, 1987). With acute exposure to high
concentrations of PCBs in air (8000-16 000 µg/m³), for a short
period, blood PCB concentrations rose to levels of 30 µg/litre; a
return to the normal level of 3 µg/litre was achieved, 4 weeks after
termination of exposure (Elo et al., 1985; WHO/EURO, 1987).

Similar plasma values were found in workers from Japanese capacitor
factories, but, here, skin lesions were noted (Hasegawa et al.,
1972a). In this same study, it was reported that air levels of PCBs of
10-50 µg/m³ were measured in a factory where KC-300 was used in the
manufacture of electric condensers. PCB levels in the serum of workers
ranged from 100 to 650 µg/litre. One month after the use of PCBs had
been suspended, serum levels remained unchanged (90-740 µg/litre).
However, in another factory making electric condensers, serum levels
decreased from an average of 800 to 300 µg/litre, within 3 months of
the use of PCBs being discontinued (Kitamura et al., 1973). According
to Hara et al. (1974), the half-time of PCBs in the blood of workers,
engaged in the manufacture of electric condensers for less than 5
years, was several months, while that of workers employed for more
than 10 years was 2-3 years.

Kuwabara et al. (1978) reported mean PCB levels of 36.8 µg/litre
(range 8.3-84.5 µg/litre) blood in 20 PCB-workers, 39 children had
blood levels of 14.3 µg/litre (0.8-93.2 µg/litre), and 12 Yusho
patients, 4.2 µg/litre (1.8-8.6 µg/litre).

Fact-finding surveys of 63 workers, who were occupationally exposed to
PCBs (Kanechlor 500) in the production of silk thread or of paint,
were carried out in Japan in 1974-75; some of them and their families
were also surveyed again in 1975-82. Nineteen per cent of them showed
PCB levels higher than 50 µg/litre plasma. These persons did not show
the typical clinical findings of Yusho patients. During 7 years, no
clear decline was observed (Takamatsu et al., 1984).

There is clear evidence that relatively high PCB levels persist in the
blood of workers whose "external" exposure ceased several months or
years previously. The blood PCB concentrations in capacitor
manufacturing workers, who had been exposed for 1-24 years, varied
between 24.4 and 192 µg/litre; this was higher than levels in the
blood of a reference population (0.5-33 µg/litre) (Maroni et al.,
1981a).

In Japan, Yakushiji et al. (1984a) studied the rate of decrease and
the half-life of PCBs in the blood of children (aged 1-13 years) and
their mothers, who were occupationally exposed to PCBs, over a 5-year
period (1975-79). The mean concentration of 121 blood samples from
50 children was 17.4 ± 22.9 µg/litre and that in 65 samples from 29
mothers was 32.3 ± 20.6 µg/litre. The concentrations of PCBs in the
blood of the children varied over a wide range, because of differences
in the duration of breast-feeding. The rate of decrease of the PCB
concentration in the blood in both 18 children and 8 mothers was
relatively constant and independent of the PCB concentrations. A
one-compartment model equation was sufficient to represent the
decrease in the concentration of PCBs in the blood. The mean rate
constant of the decrease for the children was 24.2% per year,
approximately 2.6 times higher than that of the mothers (9.2%),
equivalent to half-lives of 2.8 ± 1.1 and 7.1 ± 2.7 years,
respectively. The dilution effect due to the increase in body weight
was the most important factor that affected the reduction of the PCB
concentrations in the children.

A total of 118 blood samples, mainly from employees in industries
using PCBs, were collected in the period 1975-85. In 64 blood samples,
an average level of 17 µg/litre (range nd-110 µg/litre) was found
(Frank et al., 1988).

Brown & Lawton (1984) studied the partitioning of PCBs between adipose
tissue and serum in a population of 173 capacitor workers, who were
occupationally exposed to Aroclors 1254, 1242, and 1016 for various
periods of time. The serum levels of PCBs were significantly dependent
on the level of lipids in the serum, but not on that in the albumin.
The apparent contribution of cholesterol and its esters to PCB
transport is nearly equal to their contribution to the total serum
neutral lipids. The level of serum lipids PCBs must be equal to the
adipose fat PCBs level.

Yakushiji et al. (1984b) studied the relationship between
breast-feeding and the PCB levels in the blood. The blood samples of
50 children (121 samples) and of 29 occupationally exposed mothers (65
samples) were analysed during the period 1975-79. The PCB levels in
the blood of the children were greatly influenced by the duration of
breast-feeding, but showed little relationship to the PCBs levels in
maternal blood.

6. KINETICS AND METABOLISM

6.1 Absorption

6.1.1 Inhalation

Studies on rats (6 per group) showed that an aerosol containing a PCB
mixture (Pydraul A200: 42% chlorine), particle size 0.5-3.0 µm, at a
concentration of 30.4 ± 3.4 g/m³ for 30 min, was readily absorbed
through the lungs. The PCB concentration in the liver, 15 min after
cessation of exposure, was 50% of the maximum concentration attained
after 2 h (70 mg/kg tissue) (Benthe et al., 1972).

6.1.2 Dermal

Vos & Beems (1971) and Vos & Notenboom-Ram (1972) applied Aroclor 1260
to the shaved backs of rabbits and found systemic effects in the
kidneys, indicating that PCBs can penetrate the skin (see section
8.2.5).

Nishizumi (1976), using tritium-labelled PCBs (40% chlorine), found
evidence for the dermal absorption of PCBs in rats.

In a study of the occupational exposure of electrical workers to PCBs
(Pyralen 3010 and Apirolio, 42% chlorine content), Maroni et al.
(1981a) concluded that absorption of PCBs occurred through the human
skin. Quantitative data were not available.

6.1.3 Oral

When polychlorobiphenyl isomers were administered orally, by gavage,
to rats, at levels of 5, 50, or 100 mg/kg body weight for the lower
chlorinated compounds and up to 5 mg/kg for the higher chlorinated
compounds, 90% of the compounds were rapidly absorbed by the
gastrointestinal tract (Albro & Fishbein, 1972; Berlin et al., 1973;
Melvås & Brandt, 1973).

Using Rhesus monkeys, Allen et al. (1974a,b) determined that > 90% of
a single oral dose of 1.5 or 3.0 g Aroclor 1248/kg body weight was
absorbed over a period of 2 weeks. Drill et al. (1981) and US EPA
(1985) reviewed a number of studies indicating that PCBs are readily
absorbed from the gastrointestinal tract following oral
administration.

Bleavins et al. (1984) found that, over a period of 5 weeks, European
ferrets absorbed 85.4% of a single dose of ¹⁴C-labelled Aroclor 1254
(0.05 mg) given in food.

In contrast to the above studies, Norback et al. (1978) claimed that
59.3-87% of a single oral dose of 2,4,5,2',4',5'-hexachlorobiphenyl
passed unabsorbed through the intestines of monkeys, the first week
after dosing.

6.2 Distribution

6.2.1 Inhalation (rat)

Maximum PCB concentrations in the liver and brain of rats occurred 2
and 24 h, respectively, after a single, 30-min exposure to 30.4 ±
3.4 g/m³ of Pydraul A200 aerosol (42% chlorine content). The
concentrations in these tissues declined, while concentrations in
adipose tissues reached a maximum after 48 h (Benthe et al., 1972).

6.2.2 Oral (rat)

As in the case of other lipophilic substances, the absorption and
distribution of PCBs will, in all probability, take place via the
ly