INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY
ENVIRONMENTAL HEALTH CRITERIA 124
LINDANE
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.
Published under the joint sponsorship of
the United Nations Environment Programme,
the International Labour Organisation,
and the World Health Organization
First draft prepared by Dr M. Herbst, International Centre
for the Study of Lindane, Lyon, France and
Dr G.J. Van Esch, Bilthoven, The Netherlands
World Health Orgnization
Geneva, 1991
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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
Lindane.
(Environmental health criteria ; 124)
1.Benzene hexachloride - adverse effects 2.Benzene hexachloride
- toxicity 3.Environmental exposure 4.Environmental poluutants
I.Series
ISBN 92 4 157124 1 (NLM Classification: WA 240)
ISSN 0250-863X
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CONTENTS
1. SUMMARY AND EVALUATION; CONCLUSIONS; RECOMMENDATIONS
1.1. Summary and evaluation
1.1.1. General properties
1.1.2. Environmental transport, distribution and
transformation
1.1.3. Environmental levels and human exposure
1.1.4. Kinetics and metabolism
1.1.5. Effects on organisms in the environment
1.1.6. Effects on experimental animals and in vitro
1.1.7. Effects on humans
1.2. Conclusions
1.2.1. General population
1.2.2. Subpopulations at special risk
1.2.3. Occupational exposure
1.2.4. Environmental effects
1.3. Recommendations
2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, ANALYTICAL
METHODS
2.1. Identity
2.1.1. Primary constituent
2.1.2. Technical product
2.2. Physical and chemical properties
2.3. Conversion factors
2.4. Analytical methods
2.4.1. Sampling
2.4.2. Analytical methods
3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE
3.1. Natural occurrence
3.2. Man-made sources
3.2.1. Production levels and processes
3.2.1.1 Manufacturing process
3.2.1.2 World-wide production figures
3.2.2. Emissions
3.2.3. Uses
3.2.4. Extent of use
3.2.5. Formulations
4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
4.1. Transport and distribution between media
4.1.1. Volatilization
4.1.2. Precipitation
4.1.3. Movement in soils
4.1.4. Uptake and translocation in plants
4.2. Biotransformation
4.2.1. Degradation
4.2.1.2 Degradation under humid conditions
4.2.1.2 Degradation under submerged conditions
4.2.2. Degradation under field conditions
4.2.3. Hydrolytic degradation
4.2.4. Photolytic degradation (laboratory studies)
4.2.5. Biodegradation in water
4.2.6. Microbial degradation (field studies)
4.2.7. Bioaccumulation/Biomagnification
4.2.7.1 n-Octanol/water partition coefficient
4.2.7.2 Aquatic environment
4.2.7.3 Terrestrial environment
4.2.7.4 Bioconcentration in humans
4.2.7.5 Field studies
5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE
5.1. Environmental levels
5.1.1. Air
5.1.2. Water
5.1.2.1 Rain and snow
5.1.2.2 Fresh water
5.1.2.3 Sea water
5.1.3. Soil
5.1.3.1 Sediment
5.1.3.2 Dumping grounds and sewage sludge
5.1.4. Drinking-water, food and feed
5.1.4.1 Drinking-water
5.1.4.2 Cereals, fruits, pulses, vegetables,
and vegetable oil
5.1.4.3 Meat, fat, milk, and eggs
5.1.4.4 Animal feed
5.1.4.5 Miscellaneous products
5.1.5. Terrestrial and aquatic organisms
5.1.5.1 Plants
5.1.5.2 Aquatic organisms
5.1.5.3 Terrestrial organisms
5.2. Exposure of the general population
5.2.1. Total-diet studies
5.2.2. Intake with drinking-water and air
5.2.3. Concentrations in human samples
5.2.3.1 Blood
5.2.3.2 Adipose tissue
5.2.3.3 Breast milk
6. KINETICS AND METABOLISM
6.1. Absorption
6.1.1. Oral administration - experimental animals
6.1.2. Dermal application - experimental animals
6.1.3. Other routes - experimental animals
6.2. Distribution
6.2.1. Oral administration - experimental animals
6.2.2. Inhalation - experimental animals
6.2.3. Other routes
6.3. Metabolic transformation
6.3.1. Enzymatic involvement
6.3.2. Identification of metabolites
6.3.3. Metabolites identified in humans
6.4. Elimination and excretion in expired air, faeces, and
urine
6.4.1. Oral administration
6.4.1.1 Rat
6.4.1.2 Rabbit
6.4.2. Other routes
6.4.2.1 Mouse
6.4.2.2 Rat
6.4.2.3 Human
6.5. Retention and turnover (experimental animals)
6.6. Biotransformation
6.6.1. Plants
6.6.2. Microorganisms
6.6.2.1 Anaerobic conditions
6.6.2.2 Aerobic conditions
6.7. Isomerization
7. EFFECTS ON LABORATORY MAMMALS AND IN IN-VITRO TEST SYSTEMS
7.1. Single exposure
7.1.1. Oral
7.1.2. Intraperitoneal and intramuscular
7.1.3. Inhalation
7.1.4. Dermal
7.2. Short-term exposure
7.2.1. Oral
7.2.1.1 Mouse
7.2.1.2 Rat
7.2.1.3 Dog
7.2.1.4 Pig
7.2.2. Inhalation
7.2.2.1 Mouse
7.2.2.2 Rat
7.2.3. Dermal
7.3. Skin and eye irritation; sensitization
7.3.1. Primary skin irritation
7.3.2. Primary eye irritation
7.3.3. Sensitization
7.4. Long-term exposure
7.4.1. Oral
7.4.2. Appraisal of acute and short- and long-term
studies
7.5. Reproduction, embryotoxicity, and teratogenicity
7.5.1. Reproduction
7.5.2. Embryotoxicity and teratogenicity
7.5.2.1 Oral administration
7.5.2.2 Subcutaneous injection
7.5.3. Reproductive behaviour
7.5.4. Appraisal of reproductive toxicology
7.6. Mutagenicity and related end-points
7.6.1. DNA damage
7.6.2. Mutation
7.6.3. Chromosomal effects
7.6.4. Miscellaneous tests
7.6.5. Appraisal of mutagenicity and related end-
points
7.7. Carcinogenicity
7.7.1. Mouse
7.7.2. Rat
7.7.3. Initiationpromotion
7.7.4. Mode of action
7.7.5. Appraisal of carcinogenicity
7.8. Special studies
7.8.1. Immunosuppression
7.8.2. Behavioural studies
7.8.3. Neurotoxicity
7.8.3.1 Dose-response studies using intact
animals
7.8.3.2 Studies on mechanism
7.8.3.3 Summary
7.9. Factors that modify toxicity; toxicity of metabolites
8. EFFECTS ON HUMANS
8.1. Exposure of the general population
8.1.1. Acute toxicity, poisoning incidents
8.1.2. Effects of short- and long-term exposures -
controlled human studies
8.1.2.1 Oral administration
8.1.2.2 Dermal application
8.1.3. Epidemiological studies (general population)
8.2. Occupational exposure
8.2.1. Toxic effects
8.2.2. Irritation and sensitization
9. EFFECTS ON OTHER ORGANISMS IN THE LABORATORY AND FIELD
9.1. Microorganisms
9.1.1. Bacteria
9.1.2. Algae
9.1.2.1 Blue-green algae
9.1.2.2 Freshwater algae
9.1.2.3 Marine algae
9.1.3. Dinoflagellates, flagellates, and ciliates
9.2. Aquatic organisms
9.2.1. Invertebrates
9.2.1.1 Crustacea
9.2.1.2 Aquatic arthropods
9.2.1.3 Molluscs
9.2.2. Fish
9.2.2.1 Acute toxicity
9.2.2.2 Short- and long-term toxicity
9.2.2.3 Reproduction
9.2.3. Amphibia
9.2.3.1 Acute toxicity
9.2.3.2 Effects on hatching and larval
development
9.3. Terrestrial organisms
9.3.1. Honey-bees
9.3.2. Birds
9.3.2.1 Acute toxicity
9.3.2.2 Short-term toxicity
9.3.2.3 Reproduction
9.3.3. Mammals
9.4. Appraisal
10. PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES
APPENDIX I
REFERENCES
RESUMÉ
RESUMEN
WHO TASK GROUP MEETING ON ENVIRONMENTAL HEALTH CRITERIA FOR LINDANE
Members
Dr S. Dobson, Pollution and Ecotoxicology Section, Institute of
Terrestrial Ecology, Monkswood Experimental Station, Abbots
Ripton, Huntingdon, United Kingdom
Dr G.J. van Esch, Bilthoven, the Netherlands (Joint Rapporteur)
Dr M. Herbst, Biological Research, ASTA Pharma AG, Frankfurt,
Germany (Joint Rapporteur)
Professor J.S. Kagan, Department of General Toxicology and
Experimental Pathology, All-Union Scientific Research Instiute
of Hygiene and Toxicology of Pesticides, Polymers, and
Plastics, Kiev, USSR (Vice-Chairman)
Dr S.G.A. Magwood, Pesticides Division, Environmental Health Centre,
Health and Welfare Canada, Tunney's Pasture, Ottawa, Ontario,
Canada
Professor W.-O. Phoon, National Institute of Occupational Health and
Safety, University of Sydney, Sydney, Australia (Chairman)
Dr J.F. Risher, US Environmental Protection Agency, Environmental
Criteria and Assessment Office, Cincinnati, Ohio, USA
Dr Y. Saito, Division of Foods, National Institute of Hygienic
Sciences, Setagaya-ku, Tokyo, Japan
Dr V. Turusov, Laboratory of Carcinogenic Substances, All-Union
Cancer Research Centre, Moscow, USSR
Representatives of Non-Governmental Organizations
Dr P.G. Pontal, International Group of National Associations of
Manufacturers of Agrochemical Products (GIFAP), Brussels,
Belgium
Observers
Dr A.V. Bolotny, All-Union Scientific Research Institute of Hygiene
and Toxicology of Pesticides, Polymers, and Plastics, Kiev,
USSR
Dr D. Demozay, International Centre for the Study of Lindane (CIEL),
Rhône-Poulenc Agrochimie, Lyon, Franch.
Secretariat
Dr G.J. Burin, International Programme on Chemical Safety, World
Health Organization, Geneva, Switzerland
Dr K.W. Jager, International Programme on Chemical Safety, World
Health Organization, Geneva, Switzerland (Secretary)
Dr V.A. Rezepov, Centre for International Projects, USSR State
Committee for Environmental Protection, Moscow, USSR
NOTE TO READERS OF THE CRITERIA DOCUMENTS
Every effort has been made to present information in the
criteria documents as accurately as possible without unduly delaying
their publication. In the interest of all users of the environmental
health criteria documents, readers are kindly requested to
communicate any errors that may have occurred to the Manager of the
International Programme on Chemical Safety, World Health
Organization, Geneva, Switzerland, in order that they may be
included in corrigenda, which will appear in subsequent volumes.
* * *
A detailed data profile and a legal file can be obtained from
the International Register of Potentially Toxic Chemicals, Palais
des Nations, 1211 Geneva 10, Switzerland (Telephone No. 7988400 or
7985850)
* * *
The proprietary information contained in this document cannot
replace documentation for registration purposes, because the latter
has to be closely linked to the source, the manufacturing route, and
the purity/impurities of the substance to be registered. The data
should be used in accordance with paragraphs 82-84 and
recommendations paragraph 90 of the Second FAO Government
Consultation (1982).
ENVIRONMENTAL HEALTH CRITERIA FOR LINDANE
The WHO Task Group on environmental health criteria for lindane
met in Moscow, USSR, on 20-24 November 1989. The meeting was
convened with the financial assistance of the United Nations
Environment Programme (UNEP) and was hosted by the Centre for
International Projects (CIP), USSR State Committee for Environmental
Protection. On behalf of the CIP, Dr V.A. Rezepov opened the meeting
and welcomed the participants. Dr K.W. Jager welcomed the
participants on behalf of the three cooperating organizations of the
IPCS (UNEP, ILO, WHO). The Task Group reviewed and revised the draft
document and made an evaluation of the risks to human health and the
environment from exposure to lindane.
The first drafts of this monograph were prepared by Dr M.
Herbst (on behalf of the International Centre for the Study of
Lindane (CIEL)) and Dr G.J. van Esch (on behalf of the IPCS). The
second draft was prepared by Dr G.J. van Esch, incorporating
comments received following circulation of the first draft to the
IPCS contact points for Environmental Health Criteria publications.
The help of the CIEL in making available their proprietary
toxicological information on lindane to the IPCS and the Task Group
is gratefully acknowledged. This enabled the Task Group to make its
evaluation on the basis of more complete data than would otherwise
have been possible.
The efforts of all who helped in the preparation and
finalization of the document are also gratefully acknowledged. Dr
K.W. Jager of the IPCS Central Unit was responsible for the
technical development of this monograph and Mrs E. Heseltine of St
Léon-sur-Vézère, France, for the editing.
1. SUMMARY AND EVALUATION; CONCLUSIONS; RECOMMENDATIONS
1.1 Summary and evaluation
1.1.1 General properties
Technical-grade hexachlorocyclohexane (HCH) consists of 65-70%
alpha-HCH, 7-10% beta-HCH, 14-15% gamma-HCH, and approximately 10%
of other isomers and compounds. Lindane contains > 99% gamma-HCH.
It is a solid, with a low vapour pressure, and is poorly soluble in
water but very soluble in organic solvents, such as acetone, and in
aromatic and chlorinated solvents. The n-octanol/water partition
coefficient (log Pow) is 3.2-3.7.
Lindane can be determined separately from the other isomers of
HCH after extraction by liquid/liquid partition, column
chromatography and detection by gas chromatography with electron
capture. As these analytical methods are highly sensitive, residues
of lindane can be identified at a level of nanograms per kilogram or
per litre.
Lindane has been used as a broad-spectrum insecticide since the
early 1950s for agricultural and nonagricultural purposes, which
include treatment of seeds and soil, application on trees, timber
and stored materials, treatment of animals against ectoparasites and
in public health.
1.1.2 Environmental transport, distribution and transformation
Lindane is strongly adsorbed on soils that contain a large
amount of organic matter; furthermore, it can move downward through
the soil with water from rainfall or artificial irrigation.
Volatilization appears to be an important route of its dissipation
under the high-temperature conditions of tropical regions.
Lindane undergoes rapid degradation (dechlorination) in the
presence of ultra-violet irradiation, to form
pentachlorocyclohexenes (PCCHs) and tetrachlorocyclohexenes (TCCHs).
When lindane undergoes environmental degradation under humid or
submerged conditions and in field conditions, its half-time varies
from a few days to three years, depending on type of soil, climate,
depth of application and other factors. In agricultural soils common
in Europe, its half-time is 40-70 days. The biodegradation of
lindane is much faster in unsterilized than in sterilized soils.
Anaerobic conditions are the most favourable for its microbial
metabolization. Lindane present in water is degraded mostly by
microorganisms in sediments to form the same degradation products.
Limited amounts of lindane and gamma-PCCHs are taken up by and
translocated into plants, especially in soils with a high content of
organic matter. Residues are found mainly in the roots of plants,
and little, if any, is translocated into stems, leaves or fruits.
Rapid bioconcentration takes place in microorganisms, invertebrates,
fish, birds and humans, but biotransformation and elimination are
relatively rapid when exposure is discontinued. In aquatic
organisms, uptake from water is more important than uptake from
food. The bioconcentration factors in aquatic organisms under
laboratory conditions ranged from approximately 10 up to 6000; under
field conditions, the bioconcentration factors ranged from 10 to
2600.
1.1.3 Environmental levels and human exposure
Lindane has been found in the air above the oceans at
concentrations of 0.039-0.68 ng/m3 and has been measured at up to
11 ng/m3 in the air in some countries. The estimated
concentrations in surface water in a number of European countries
were mainly below 0.1 µg/litre. The concentration in the River Rhine
and its tributaries in 1969-74 varied between 0.01 and 0.4 µg/litre;
after 1974, it was below 0.1 µg/litre. Levels of 0.001-0.02 µg/litre
have been reported in seawater. The concentrations of lindane in
soil are generally low - in the range 0.001-0.01 mg/kg, except in
areas where waste is disposed of.
Fish and shellfish have been found to contain gamma-HCH at
concentrations ranging from none detected up to 2.5 mg/kg on a fat
basis, depending on whether they live in fresh or seawater and
whether they have a low or high fat content. Levels of about 330 and
440 µg/kg (wet weight) were found in adipose tissue of polar bears
in 1982 and 1984, respectively. The concentration of lindane in the
livers of birds of prey varied between 0.01 and 0.1 mg/kg. Eggs of
sparrow-hawks collected in 1972-73 in the Federal Republic of
Germany contained levels of 0.6 up to 11.1 mg/kg (on a fat basis).
The concentration of lindane in the livers of predatory birds
varied between 0.01 and 0.1 mg/kg. Eggs of sparrow- hawks collected
in 1972-73 in the Federal Republic of Germany contained levels of
0.6 up to 11.1 mg/kg (on a fat basis). The concentrations of lindane
in drinking-water are generally below 0.001 µg/litre, and in
industrialized countries more than 90% of human intake of lindane
originates from food. Over the last 25 years, selected food items
have been analysed for lindane in a large number of countries. The
concentrations found in cereals, fruits, vegetables, pulses, and
vegetable oils ranged from not detected up to 0.5 mg/kg product, and
those in milk, fat, meat, and eggs from not detected up to 1.0 mg/kg
product (on a fat basis). In only a few instances were higher
concentrations found. The concentrations in fish were generally far
lower than 0.05 mg/kg product (on a fat basis). In total-diet and
market-basket studies to estimate daily human intake of lindane, a
clear difference was observed with time: intake in the period around
1970 was up to 0.05 µg/kg body weight per day, whereas by 1980
intake had decreased to 0.003 µg/kg body weight per day or lower. In
the USA, the daily intake of gamma-HCH between 1976 and 1979
decreased from 0.005 to 0.001 µg/kg body weight per day for infants
and from 0.01 to 0.006 µg/kg body weight per day for toddlers.
Determinations of the lindane content in body tissues in the
general population have been made in a number of countries. The
content in blood in the Netherlands was in the order of < 0.1-0.2
µg/litre, but much higher concentrations were found in several
countries where technical-grade HCH was used. The mean
concentrations in human adipose tissue in various countries ranged
from < 0.01 up to 0.2 mg/kg on a fat basis. The concentrations of
lindane in human milk are generally rather low, at average
concentrations of < 0.001 up to 0.1 mg/kg on a fat basis; however,
there has been a clear decrease over time.
Lindane is thus distributed all over the world and can be
detected in air, water, soil, sediment, aquatic and terrestrial
organisms, and food, although the concentrations in these different
compartments are generally low and are gradually decreasing. Humans
are exposed daily via food, and lindane has been found in blood,
adipose tissue, and breast milk; the levels of intake, however, are
also decreasing.
1.1.4 Kinetics and metabolism
In rats, lindane is absorbed rapidly from the gastrointestinal
tract and distributed to all organs and tissues within a few hours.
The highest concentrations are found in adipose tissues and skin; in
various studies, the fat:blood ratio was about 150-200, the
liver:blood ratio, 5.3-9.6 and the brain:blood ratio, 4-6.5. The
same fat:blood ratio was found in rats exposed by inhalation. These
ratios vary with sex, being higher in females. Uptake of lindane
through the skin after dermal application is slow and occurs to a
very limited extent; this may explain the low toxicity of lindane
after dermal exposure.
Lindane is metabolized mainly in the liver by four enzymatic
reactions: dehydrogenation to gamma-HCH, dehydrochlorination to
gamma-PCCH, dechlorination to gamma-TCCH and hydroxylation to
hexachlorocyclohexanol. The end-products of biotransformation are
di-, tri-, tetra-, penta-, and hexachloro- compounds. These
metabolites are excreted mainly via the urine in the free form or
conjugated with glucuronic acid, sulfuric acid or N-acetylcystein.
The elimination is relatively fast, with half-times in rats of 3-4
days. Bacteria and fungi metabolize lindane to TCCH and PCCH. The
rate of metabolic transformation in plants is low, and the main
degradation pathway proceeds via PCCH to tri- and tetrachlorophenol
and conjugates with beta-glucose and other, unknown compounds. There
is no evidence that lindane is isomerized to alpha-HCH.
1.1.5 Effects on organisms in the environment
Lindane is not very toxic for bacteria, algae, or protozoa: 1
mg/litre was generally the no-observed effect level (NOEL). Its
action on fungi is variable, with NOELs varying from 1 to 30
mg/litre depending on the species. It is moderately toxic for
invertebrates and fish, the L(E)C50 values for these organisms being
20-90 µg/litre. In short-term and long-term studies with three
species of fish, the NOEL was 9 µg/litre; no effect on reproduction
was seen with levels of 2.1-23.4 µg/litre. The LC50 values for both
freshwater and marine crustacea varied between 1 and 1100 µg/litre.
Reproduction in Daphnia magna was depressed in a dose-dependent
fashion; the NOEL was in the range 11-19 µg/litre. Reproduction of
molluscs was not adversely effected by a dose of 1 mg/litre.
The LD50 for honey-bees was 0.56 µg/bee.
Acute oral LD50 values for a number of bird species were
between 100 and 1000 mg/kg body weight. In short-term studies with
birds, doses of 4-10 mg/kg diet had no effect, even on egg-shell
quality. Laying ducks treated with doses of lindane up to 20 mg/kg
body weight, however, had decreased egg production.
Bats exposed to wood shavings that initially contained 10-866
mg/m2 lindane, resulting from application at the recommended rate,
all died within 17 days. No effect on mortality or reproductive
success was seen in small field mammals given 20 mg/kg diet (the
highest dose tested). No data were available on effects on
populations and ecosystems.
1.1.6 Effects on experimental animals and in vitro
The acute oral toxicity of lindane is moderate: the LD50 for
mice and rats is in the range 60-250 mg/kg body weight, depending on
the vehicle used. The dermal LD50 for rats is approximately 900
mg/kg body weight. Toxicity was manifested by signs of central
nervous system stimulation.
Lindane does not irritate or sensitize the skin; it is slightly
irritating to the eye.
In a 90-day study in rats, the NOEL was 10 mg/kg diet
(equivalent to 0.5 mg/kg body weight). At 50 and 250 mg/kg diet, the
weights of the liver, kidneys, and thyroid were increased; at 250
mg/kg diet, an increase was seen in liver enzyme activity. This
increase in enzyme activity accelerates the breakdown of both
lindane and other compounds. In another 90-day study in rats, 4
mg/kg diet (equivalent to 0.2 mg/kg body weight) was considered to
be the no-adverse-effect level (NOAEL); renal and hepatic toxicity
were observed at concentrations of 20 mg/kg diet and higher. No
neurological effect was observed in a 30-day feeding study in rats
with 240 mg/kg diet (equivalent to 12 mg/kg body weight); however,
when this dose was given by gavage, neurological effects were seen.
A short-term toxicity study in mice was considered to be inadequate
to establish a NOEL.
Administration of lindane to dogs at 15 mg/kg in the diet
(equivalent to 0.6 mg/kg body weight) for 63 weeks had no toxic
effect. In a two-year study of the toxicity of this compound in
dogs, in which a large number of parameters were measured, no
treatment-related abnormality was apparent at doses of 50 mg/kg diet
(equivalent to 2 mg/kg body weight) and lower. In the group given
100 mg/kg diet, however, levels of alkaline phosphatase were
increased; and with 200 mg/kg diet, abnormalities in
electroencephalogram tracings indicative of non-specific neuronal
irritation were observed.
In rats exposed by inhalation to lindane at 0.02-4.54 mg/m3
for 6 h/day for 3 months, the highest dose induced increases in
hepatic cytochrome P450 values; the NOAEL was found to be 0.6
mg/m3. In two long-term studies in rats, carried out many years
ago, doses of 10-1600 mg/kg diet were tested. In one of these
studies, 50 mg/kg diet (equivalent to 2.5 mg/kg body weight) was
found to be the NOAEL. At 100 mg/kg diet, an increase in liver
weight, hepatocellular hypertrophy, fatty degeneration, and necrosis
were found. In the other study, 25 mg/kg diet (equivalent to 1.25
mg/kg body weight) had no effect, but hepatocellular hypertrophy and
fatty degeneration were seen with 50 mg/kg diet.
Lindane has been investigated for its effects on all aspects of
reproduction (in rats over three generations) and for its
embryotoxicity and teratogenicity after oral, subcutaneous and
intraperitoneal administration in mice, rats, dogs, and pigs. It had
no teratogenic effect after oral or parenteral administration (extra
ribs were regarded as variations). Fetotoxic and/or maternal toxic
effects were observed with doses of 10 mg/kg body weight and above
given by oral gavage; 5 mg/kg body weight is considered to be the
NOAEL. Lindane had no effect on reproduction or maturation in the
three-generation study in rats at doses of up to 100 mg/kg diet; but
with 50 mg/kg diet, morphological changes in the liver indicating
enzyme induction occurred in the offspring of the third generation.
The NOEL in this test was 25 mg/kg diet (equivalent to 1.25 mg/kg
body weight).
The NOEL for neurotoxicity in a 22-day study in rats was 2.5
mg/kg body weight.
The mutagenicity of lindane has been studied adequately. In
extensive investigations of its ability to induce gene mutations in
bacteria and in mammalian cells, and for its capacity to induce
sex-linked recessive lethal mutations in Drosophila melanogaster,
negative results were obtained consistently. Lindane also gave
negative results in tests for chromosomal damage and sister
chromatid exchange in mammalian cells in vitro and in vivo . The
results of assays for DNA damage in bacteria and for covalent
binding to DNA in the liver of rats and mice in vivo following
oral administration were also negative. In the very few studies in
which positive results were obtained, either the study design was
invalid or the purity of the compound tested was not reported.
Overall, however, lindane appears to have no mutagenic potential.
Studies to define the carcinogenic potential of lindane have
been carried out in mice and rats using dose levels of up to 600
mg/kg diet in mice and up to 1600 mg/kg diet in rats. Hyperplastic
nodules and/or hepatocellular adenomas were found in mice given
doses of 160 mg/kg diet or more; in some studies, the dose levels
exceeded the maximum tolerated dose. Two studies in mice with dose
levels of up to 160 mg/kg diet and one in rats with 640 mg/kg diet
showed no increase in the incidence of tumours.
The results of studies on initiation-promotion of
carcinogenicity, on the mode of action, and on mutagenicity indicate
that the tumorigenic response observed with gamma-HCH in mice is
mediated by a nongenetic mechanism.
1.1.7 Effects on humans
Several cases of fatal poisoning and of non-fatal illness
caused by lindane have been reported, which were either accidental,
intentional (suicide), or due to gross neglect of safety precautions
or improper uses of medical products containing lindane. Symptoms
included nausea, restlessness, headache, vomiting, tremor, ataxia,
and tonic-clonic convulsions and/or changes in the
electroencephalographic pattern. These effects were reversible after
discontinuation of exposure or symptomatic treatment.
Notwithstanding extensive use over 40 years, very few cases of
poisoning in the occupational setting have been reported. In workers
exposed for long periods during either manufacture or application of
lindane, the only sign found was increased activity of drug
metabolizing enzymes in the liver. There is no evidence for the
relationship suggested in some publications between exposure to
lindane and the occurrence of blood dyscrasias. A few acute and
short-term studies in humans indicate that a dose of approximately
1.0 mg/kg body weight does not induce poisoning; however, a dose of
15-17 mg/kg body weight resulted in severe toxic symptoms.
Approximately 10% of a dermally applied dose is absorbed,
although more passes through damaged skin.
1.2 Conclusions
1.2.1 General population
Lindane is circulating in the environment and is present in
food chains, so that humans will continue to be exposed. The daily
intake and total exposure of the general population are decreasing
gradually, however; they are clearly below the advised acceptable
daily intake and are of no concern to public health.
1.2.2 Subpopulations at special risk
The presence of lindane in breast milk results in exposure of
breast-fed babies to levels that are generally below the acceptable
daily intake and therefore of no concern to health. Although lower
levels of exposure would be preferred, the present levels are not a
limiting factor for the practice of natural breast-feeding.
Prescriptions should be followed strictly with regard to the
therapeutic use of lindane against scabies and to control body lice.
1.2.3 Occupational exposure
As long as the recommended precautions to minimize exposure are
observed, lindane can be handled safely.
1.2.4 Environmental effects
Lindane is toxic to bats that roost in close contact with wood
treated according to the recommended conditions of application.
Apart from the results of studies of spills into the aquatic
environment, there is no evidence to suggest that the presence of
lindane in the environment poses a significant hazard to populations
of other organisms.
1.3 Recommendations
1. In order to minimize environmental pollution by other isomers
of HCH, lindane (> 99% gamma-HCH) must be used instead of
technical-grade HCH.
2. In order to avoid environmental pollution, by-products of and
effluents from the manufacture of lindane should be disposed of
in an appropriate way.
3. In disposing of lindane, care should be taken to avoid
contamination of natural waters and soil.
4. As for other pesticides, proper instructions about application
procedures and safety precautions should be given to people who
handle lindane.
5. Long-term carcinogenicity tests conducted according to
present-day standards should be conducted.
6. Monitoring of the daily intake of lindane by the general
population should continue.
2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, ANALYTICAL
METHODS
2.1 Identity
2.1.1 Primary constituent
Common name: Lindane
Chemical structure:1
Fig. 1. Chemical structure of lindane
Chemical formula: C6H6Cl6
Relative molecular mass: 290.8 (290.9)
CAS chemical name: 1alpha,2alpha,3ß,5alpha,6ß-
hexachlorocyclohexane
CAS registry number: 58-89-9
RTECS registry number: GV4900000
Synonym: Hexachlorocyclohexane (gamma-isomer)
According to IUPAC rules, the designation 'benzene
hexachloride' is incorrect; nevertheless, it is still widely used,
especially in the form of its abbreviation, BHC. This is therefore
another common name approved by the ISO. The compound is called
gamma-HCH by the WHO, but gamma-BHC by the FAO (FAO, 1973). The
synonym hexachlorocyclohexane (gamma isomer) is used by the
Environmental Protection Agency and the American Conference of
Governmental Industrial Hygienists in the USA. The definitions of
these different appellations are given in Table 1.
1 See Appendix I
Table 1. Definitions of appellations of lindane
Name Definition Remarks
Lindane product containing not less ISO-AFNOR name for a product
than 99% gamma-HCH (not yet recognized by BSI)
Lindane = gamma-HCH Common name used for
gamma-HCH in the USSR only
gamma-HCH gamma isomer of 1,2,3,4,5,6- ISO-AFNOR common name
hexachlorocyclohexane
gamma-BHC gamma isomer of 1,2,3,4,5,6- ISO BSI common name in
benzene hexachloride English-speaking countries
(recognized by ISO as
synonym of gamma-HCH)
2.1.2 Technical product
Common trade names: A great number of products containing lindane
are on the market; no attempt has been made to
list the hundreds of trade names here (see
Hudson et al., 1984; Hill & Camardese, 1986;
International Register for Potentially Toxic
Chemicals, 1989).
Purity: The FAO (1973) requires that lindane "... shall
consist, essentially, of gamma-BHC as white or
nearly white granules, flakes or powder, free
from extraneous impurities or added modifying
agents and with not more than a faint odour."
The FAO further requires that it contain not
less than 99.0% gamma-HCH and that the
melting-point be at least 112 °C, which is not
depressed when the sample is mixed with an equal
amount of pure gamma-HCH.
In some processes for manufacturing lindane, low levels of
dioxin may be formed (US Environmental Protection Agency, 1985).
Under appropriate manufacturing conditions, however, no
2,3,7,8-tetrachlorodibenzodioxin or 2,3,7,8-tetrachlorodibenzofuran
is detected in HCH, lindane, trichlorobenzene, industrial liquid or
gaseous effluents at the analytical limit of detection of 1 µg/kg
letter from D. Demosay, Rhône-Poulenc, to IPCS dated 17 November
1989.
2.2 Physical and chemical properties
Lindane is a colourless, crystalline solid with either a faint
or no smell (the characteristic smell of technical-grade HCH is
attributed to impurities, particularly heptachlorocyclohexane).
Melting-point: 112.8 °C
Boiling-point: 288 °C
Vapour pressure: 0.434 x 10-5 kPa (3.26 x 10-5 mmHg) at 20 °C
60.6 x 10-5 kPa (45.6 x 10-5 mmHg) at 40 °C
Density: 1.85
Solubility: nearly insoluble in water at 20 °C (10
mg/litre); moderately soluble in ethanol (6.7%);
slightly soluble in mineral oils;
soluble in acetone and in aromatic and
chlorinated solvents
Stability: stable to light, air, heat, carbon dioxide, and
strong acids; dehydrochlorinates in the presence
of alkali or on prolonged exposure to heat with
the formation of trichlorobenzenes, phosgene,
and hydrochloric acid. It is incompatible with
strong bases and powdered metals, such as iron,
zinc, and aluminium, and with oxidizing agents;
can undergo oxidation when in contact with
ozone.
Corrosivity: corrosive to aluminium
Inflammability: not inflammable
n-Octanol/water 3.2-3.7 (see section 4.2.7.1) (Demozay &
partition Marechal, 1972; Dutch Chemical Industry
coefficient Association 1980; American Conference of
(log Pow): Governmental Industrial Hygienists, 1986;
Rhône-Poulenc Agrochimie, 1986)
2.3 Conversion factors
1 ppm = 12.1 mg/m3
1 mg/m3 = 0.083 ppm
2.4 Analytical methods
2.4.1 Sampling
Sampling procedures and methods for preparing samples of
formulations and for analysing residues have been described by
Mestres (1974), the Deutsche Forschungsgemeinschaft (1979), the
Association of Official Analytical Chemists (1980), and Hildebrandt
et al. (1986).
2.4.2 Analytical methods
Products are analysed by a cryoscopic method (Raw, 1970; FAO,
1973; WHO, 1985). Formulated products can be analysed by determining
hydrolysable chlorine (Raw, 1970; FAO, 1973). Since the latter
method is not specific, other methods, such as gas chromatography,
are used to obtain sufficient separation of the HCH isomers.
Residues in food and in soil can be determined after adequate
clean-up by gas chromatography and other chromatographic methods
(Nash et al., 1973; Eichler, 1977; Association of Official
Analytical Chemists, 1980; DeutscheForschungsgemeinschaft, 1983).
The principle of the method is extraction of a sample with organic
solvents (acetonitrile, hexane/acetone, acetone, and others). Fat is
extracted from fatty foods and partitioned between petroleum ether
and acetonitrile by extracting aliquots or an entire solution of
acetonitrile into petroleum ether. Residues are purified by
chromatography on a Florisil colum, and eluted with a mixture of
petroleum ether and ethylether. Concentrated residues are measured
by gas chromatography with electron capture detection.
The method described by the Deutsche Forschungsgemeinschaft
(1979) for fruits and vegetables is based on extraction of samples
with acetone and extraction of the aliquot with dichloromethane. The
residue obtained after evaporation of the solvent is cleaned by
co-distillation, and the distillate is analysed by gas
chromatography with electron capture detection. The limit of
determination depends on the method, the substrate and the sample
size; the lower limit of determination is 0.001-0.01 mg/kg.
Palmer & Kolmodin-Hedman (1972) analysed air samples by gas
chromatography with electron capture detection, and alpha-, beta-,
and gamma-HCH were determined in serum by gas chromatography after a
deproteinization extraction step (Palmer & Kolmodin-Hedman, 1972;
Angerer & Barchet, 1983).
Wittlinger & Ballschmiter (1987) provided an extensive
description of analytical methods for HCHs in air, involving
sampling by adsorption, extraction and preseparation, and
determination by high-resolution gas chromatography. Sampling was
performed by pumping air through a glass-fibre filter and then
through a silica-gel layer, using an internal standard. The sample
was extracted with dichloromethane and the extract evaporated. The
preseparation was done on silica gel, and the aliquot was eluted in
a mixture of hexane and dichloromethane. High-resolution capillary
gas chromatography, electron capture detection and a mass selective
detector were used for determination.
Eder et al. (1987) described three analytical methods for the
determination of HCHs in sediments: moist samples are extracted with
a solvent or a mixture of solvents, concentrated or fractionated and
determined by gas chromatography and electron capture detection.
Greve (1972) described a method for the determination of
organochlorine pesticides in water based on gas chromatography of a
petroleum ether extract after clean-up over Florisil or silica gel.
The limit of detection for lindane is 0.01 µg/litre.
Methods used for the determination of lindane in samples of
soil, animal, and vegetable products in the USSR are described by
Izmerov (1983). These methods are based on extraction with organic
solvents, purification and concentration of the extracts and
determination by gas-liquid chromatography with electron-capture
detection.
3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE
3.1 Natural occurrence
Lindane is not known to occur as a natural product.
3.2 Man-made sources
3.2.1 Production levels and processes
3.2.1.1 Manufacturing process
HCH was discovered in 1825, but its insecticidal properties
were first patented only in the 1940s. It has been produced
commercially since 1949.
Technical-grade HCH is synthesized from benzene and chlorine in
the presence of ultra-violet light and comprises 65-70% alpha-HCH,
7-10% beta-HCH, 14-15% lindane (gamma-HCH), approximately 7%
delta-HCH, 1-2% epsilon-HCH, and 1-2% other components. By-products
can be minimized by careful control of the reaction conditions.
Lindane (> 99% gamma-HCH) can be purified by multiple extractions
with methanol.
The extraction of lindane from HCH produces 85%
non-insecticidal HCH isomers, which can be used as intermediates in
the production of trichlorobenzene and hydrochloric acid after
cracking in an integrated installation. Trichlorobenzene is used in
the synthesis of other chemicals (van Velsen, 1986; Rhône-Poulenc
Agrochimie, 1986).
3.2.1.2 World-wide production figures
Lindane is produced in Austria, France, and Spain and in China,
India, Turkey, and the USSR. Before 1984, lindane was also
manufactured in the German Democratic Republic, Poland, Yugoslavia,
Romania, and Hungary; since then, all production has been stopped in
Germany, Japan, the Netherlands, the United Kingdom, and the USA.
Although in most developed countries use of technical-grade HCH
has been prohibited, it is still used elsewhere on a large scale:
total consumption of technical-grade HCH in India in 1986-87 was
approximately 27 000 tonnes (International Atomic Energy Agency,
1988).
3.2.2 Emissions
According to De Bruijn (1979), approximately 0.1% of the
lindane processed reaches the waste-water of a formulating plant.
Treatment of the waste-water, however, leads to solid waste, which
should be incinerated. In the past, it was often dumped in the
environment and could be dispersed from (open) chemical dumping
grounds to more remote soils by the wind.
Lindane enters the environment following application of
lindane-containing pesticides. Emissions can cross national
boundaries in water and air. For instance, the total trans-frontier
flux of lindane into the Netherlands via the surface water of the
River Rhine was approximately 1.8 tonnes per year (average for
1980-83) and that via the River Meuse, 0.2 tonnes per year (Slooff &
Matthijsen, 1988).
3.2.3 Uses
Lindane is a broad-spectrum insecticide, which has been used
since 1949 for agricultural as well as non-agricultural purposes.
Approximately 80% of the total production is used in agriculture
(Demozay & Marechal, 1972), mostly for seed and soil treatment. Wood
and timber protection is the major non-agricultural use. Lindane is
also used against ectoparasites in veterinary and pharmaceutical
products (Rhône-Poulenc Agrochimie, 1986).
3.2.4 Extent of use
Lindane is used worldwide, with the major exception of Japan,
where all uses of HCH were cancelled in 1971 mainly because of
environmental pollution with alpha- and beta-HCH resulting from
extensive use of technical-grade HCH. At that time, no clear
difference was made between the risks presented by the individual
HCH isomers, and lindane was banned as well. In almost all other
countries, lindane is registered for one or more applications,
although the use pattern differs from one country to another.
In 1979, the US Department of Agriculture and the Environmental
Protection Agency summarized the percentage uses of lindane in the
USA as follows: seed treatment 48%, hardwood lumber 23%, livestock
16%, pets 3%, pecans 3%, pineapples 2%, ornamentals 2%, household
1%, cucurbits 1%, forestry 0.5%, and structures 1%. In France and
Germany, 70-80% of all lindane used agriculturally is for soil
treatment, to protect maize and sugar beets, and 15-20% is used for
seed treatment. De Bruijn (1979) reported an estimate of the pattern
of use of lindane in the European Economic Community.
3.2.5 Formulations
Formulation facilities exist in many countries. Lindane is made
in numerous forms, the most important of which are: wettable powders
(up to 90% active ingredient); emulsifiable concentrates (not more
than 20% active ingredient); flowable suspensions (in water);
solutions in organic solvents (up to 50% active ingredient); dusts
and powders (0.5-2% active ingredient); granules and coarse dusts
(3-4% active ingredient); ready-for-use baits; aerosols; and special
formulations for use in human and veterinary medicine (Demozay &
Marechal, 1972).
Lindane dissolved in organic solvents may be used in 'thermal
foggers' in glasshouses or atomized in open areas; such solutions
are appropriate for aerial application (5-10 litres/ha of
formulations containing 5-10% active ingredient). Concentrated
solutions containing an anti-vaporization component have been
applied using an ultra-low volume method at 0.5-1 litre/ha. Various
fumigation preparations for indoor use have been sold, including
fumigation strips, tablets, and smoke generators. These contained
virtually pure lindane to which a small quantity of binding material
was added. Because of its versatility and relatively low acute
toxicity, lindane is often used in mixed formulations with other
insecticides and fungicides (Demozay & Marechal, 1972).
4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
4.1 Transport and distribution between media
4.1.1 Volatilization
Some of the active ingredient of lindane volatilizes after it
has been applied to control insect pests, especially on leaves.
Starr & Johnson (1968) demonstrated that 20% of an applied dose had
evaporated 96 h after bean plants had been sprayed with lindane at
16 °C. The evaporation was dependent on temperature and on the
humidity of the air.
Some of the lindane that reaches the soil may also vaporize as
degradation products. Cliath & Spencer (1972) showed the presence of
vapours of the metabolite PCCH, which has a vapour pressure
approximately 14 times higher than that of lindane.
In a model test, four soil types, ranging from a loamy sand to
a clay, were treated with 14C-lindane to give a concentration of
10 mg/kg soil; water was then added to the samples, they were
air-dried at 33 °C and at 55 °C, and volatilization was measured by
trapping the vapours. Three cycles of about 14 days each were
followed. Lindane volatilized from the soils only with water, and no
further volatilization occurred after the soils had reached the dry
state. The four soil types were associated with different
volatilization rates: the highest occurred in loamy sand. In the
analysis of vaporized material, unchanged lindane and its
degradation products were not differentiated; however, considerable
degradation of lindane was found in the soils, and PCCH was
identified as a metabolite. At least some of the vaporized material
may therefore have consisted of degradation products (Guenzi &
Beard, 1970).
4.1.2 Precipitation
Evaporation and adsorption to solid particles are important
processes in the distribution of lindane. Reverse processes such as
deposition from the air and remobilization from silt and sediment
also play a part. Buscher et al. (1964) demonstrated that aeration
of aqueous solutions of lindane resulted in a loss of 10% over three
days, which was ascribed to a co-distillation process as it was
greater than could be explained by evaporation alone. MacKay &
Walkoff (1973) confirmed that evaporation is an important process in
the loss of HCH. Lichtenstein & Schulz (1970) found that 16.5% HCH
was lost from a non-aerated aqueous solution in 24 h at 30 °C.
The amount of lindane that is distributed by dry deposition
depends on the nature of the surface above which the organic
components are present. The half-time for dry deposition of HCHs
(height of mixing layer, 1000 m) in the Netherlands was calculated
to be 2-8 days. On this basis, a rough estimate of the annual flux
to soil and water in that country would be 0.5-1.5 tonnes from an
outdoor air concentration of 0.4 ng/m3 (Slooff & Matthijsen,
1988).
4.1.3 Movement in soils
Movement of a substance through the soil profile depends on its
adsorption-desorption characteristics in soil/water systems and, to
some extent, on its volatility in the soil pores and its diffusion.
The adsorption-desorption characteristics of lindane have been the
object of a number of studies (Kay & Elrick, 1967; Mills & Biggar,
1969; Baluja et al., 1975; Portmann, 1979; Wahid & Sethunathan,
1979, 1980; Wirth, 1985), all of which showed that lindane is
strongly adsorbed to organic soil material and weakly adsorbed to
inorganic matter. In the absence of organic matter, the clay content
and free iron oxide are implicated in the sorption of lindane (Wahid
& Sethunathan, 1979). It can be concluded that the mobility of
lindane is very low in soils with a high content of organic matter
but might be higher in soils containing little organic matter.
No consensus has been reached in the literature about the
possibility that lindane can be remobilized by desorption from
polluted soil. Generally, HCH isomers are strongly adsorbed. Under
certain conditions - high concentrations of lindane in highly
permeable soils with a low organic carbon content (< 0.1%) - a
small percentage of the compound may be washed out and reach the
groundwater. Nevertheless, the low rate of transport of lindane
makes the probability that it will reach groundwater low or very low
(Slooff & Matthijsen, 1988).
The diffusion of lindane through soil was investigated by
Ehlers et al. (1969a,b) and by Shearer et al. (1973). Diffusion was
strongly influenced by the water content of the soil, by the bulk
density and by temperature. The diffusion coefficient is nearly zero
in soil containing 1% water, but with a water content of 3%, lindane
is displaced from the adsorbing surface so that the diffusion
coefficient becomes maximal; a further increase in water content
reduces the diffusion coefficient. The diffusion of lindane in soils
can thus vary between a 'vapour' and a 'non-vapour' phase, depending
on the concentration of lindane, the length of time and the water
content of the soil.
Leaching of three formulations of lindane was investigated in a
series of model studies by Heupt (1974) in different soil types. The
test system consisted of 30-cm columns filled with soil to which
lindane formulations were applied at application rates corresponding
to 6 kg of active ingredient per hectare. Rainfall was simulated at
a rate of 200 mm within two days. No lindane was found in the eluate
at the limit of detection of 1 µg/litre. In field tests by Cliath &
Spencer (1971, 1972), lindane was worked into topsoil of two plots
of sandy loam and two of silty clay to a depth of 0-7.5 or 7.5-15
cm, corresponding to an application rate of 21 kg/ha. One of the
plots of each soil type received additional irrigation. Almost no
movement of lindane was found in the dry plots at the end of the
two-year observation period. In the irrigated plots, a broadening of
the lindane-containing zone and downward movement to a depth of 60
cm were observed, especially in sandy loam; in clay soil, lindane
had almost no mobility.
In a series of dissipation studies with 14C-labelled lindane
in soils, coordinated by FAO and the International Atomic Energy
Agency, it was demonstrated that persistent pesticides such as
lindane dissipate much faster in the tropics than in temperate
climates, probably owing to a large extent to volatilization
(International Atomic Energy Agency, 1988), as had been found by
Edwards (1973a,b, 1977). Table 2 summarizes the results of these
studies.
Table 2. Field half-times of gamma-HCH in soils (0-10-cm depth)a
Country Half-time (days)b Time required for
initial loss of 50%
Overall First phase Second phase of radioactivity (days)
India 138 (124-147) 41 (30-50) 188 (83-362) 30-45
Ecuador 150-171 54-60 120-160 40-50
Kenya 5-8 - - 3-4
a Adapted from International Atomic Energy Agency (1988)
b In temperate soils, the mean half-time was 438 (401-1022) days
(Edwards, 1973a, b)
4.1.4 Uptake and translocation in plants
One of the first investigations on the absorption of lindane by
various seeds was reported by Bradbury & Whitaker (1956). Lindane
was taken up from a nutrient solution by the roots of wheat
seedlings at a rate of up to 100 mg/kg (fresh weight) within seven
days. Subsequent investigations demonstrated that uptake by plants
is dependent on a variety of factors.
The influence of soil type was investigated by Bradbury
(1963). Seedlings grown from seed dressed with lindane and planted
in sand had residue levels about two-fold higher than those of
plants grown in compost. A further study was reported of the fate of
14C-lindane in loam and sandy soil and in oat plants grown in
these soils for 13 days. The loam soil was treated with about 7.3
mg/kg active ingredient and the sandy soil with about 3 mg/kg.
Residues were found to be more persistent in loam (53.5% of
radiolabel) than in sandy soil (33.8%), but oats grown in sandy soil
took up more residues than those grown in loam (loam soil: roots,
0.5%; tops, 0.3%; sandy soil: roots, 2.5%; tops, 1.2%).
14C-Lindane was the major constituent of the soil residues soluble
in organic solvents. A major metabolite, which was probably
gamma-PCCH, represented 11% of the organic-soluble radiolabelled
residues in loam soil; 2,4,5-trichlorophenol accounted for 2.7% of
these residues. The authors concluded that the three major factors
that determine the environmental fate of 14C-lindane and other
insecticides are the insecticide itself, its solubility in water and
the type of soil to which it is applied. Compounds with greater
aqueous solubility are more mobile, are taken up by plants to a
greater extent, and appear to be more susceptible to degradation
than compounds less soluble in water. In soils with little organic
matter, insecticide residues are more mobile and hence more
susceptible to volatilization, uptake by plants, and degradation
than in more adsorbent soils such as loam (Fuhremann & Lichtenstein,
1980).
The rate of application to soil was found to be a further
important factor in determining residue levels. Transfer of lindane
from soil into rice plants was almost proportional to the level of
contamination of the soil (Kawahara, 1972), but only at low levels
of contamination. Charnetski & Lichtenstein (1973) reported a good
correlation between the content of 14C-lindane in sand (at up to 6
mg/kg, which is about 12 times the concentration expected after
normal application) and that in pea plants grown for six days; at
concentrations greater than 10 mg/kg of soil, there was no further
increase in the residue levels.
Uptake of lindane after application to leaves is lower than
that resulting from application to soil. In lettuce and endives
treated with 14C-lindane and grown for 21 and 37 days,
respectively, only 4.5-13.9% of the applied radioactivity was found
at the time of harvest, and most of the lindane had evaporated into
the atmosphere (Kohli et al., 1976a).
Differences in residue levels are also dependent on the plant
species. Of a series of edible crops grown in soil containing
lindane at an initial concentration of 5 mg/kg (about 10 times the
normal application rate), carrots had higher levels than beans,
tomatoes, or potatoes (San Antonio, 1959). More lindane was absorbed
from soil with an initial concentration of 2.6 mg/kg by radishes,
turnips, and spinach than by Chinese cabbage (Kawahara et al.,
1971). The amounts of residues of HCH isomers in turnips were
proportional to the initial concentrations of the isomers in the
soil (0.05, 0.1, 0.5, 1, 5, 10, or 50 mg/kg soil). The soil:plant
residue ratios were in the range 10-20:1 (Kawahara & Nakamura,
1972).
The translocation of lindane and its metabolites in plants has
also been investigated in detail (San Antonio, 1959; Bradbury, 1963;
Itokawa et al., 1970; Kawahara, 1972; Kawahara & Nakamura, 1972;
Charnetski & Lichtenstein, 1973; Balba & Saha, 1974; Eichler, 1980;
Korte, 1980). Neither lindane taken up from soil nor its metabolites
were evenly distributed within the plants: Comparatively high
residue levels were always detected in the roots, whereas only small
amounts were translocated into stems, leaves, and fruits. Paasivirta
et al. (1988) showed that in water-plants, lindane concentrations
are similar in roots and leaves.
4.2 Biotransformation
4.2.1 Degradation
4.2.1.1 Degradation under humid conditions
The half-times of lindane found by different investigators vary
considerably, depending on the type of soil to which it is applied
and, possibly, temperature. Lindane incubated in a sandy-loam soil
with a water capacity of 28% and 60% saturation at room temperature
had a half-time of approximately 40 days (Heeschen et al., 1980).
The half-times of lindane in model tests were 4-6 weeks in humid
sand with a high content of organic matter and 30 weeks in sandy
loam (Heupt, 1979). The half-times in aerobic and anaerobic
conditions ranged from 12 to 174 days and 100 to 720 days,
respectively; in aerobic field conditions, the half-time was 88-1146
days (Edwards, 1966; Kohnen et al., 1975; Kampe, 1980; Rao &
Davidson, 1982; MacRae et al., 1984).
Assuming that lindane is not washed out below the level of
ploughed furrows (approximately 20 cm), a half-life of 350 days will
result in persistence of 50% of a dose one year after application
(Slooff & Matthijsen, 1988). One month after double treatment of
potato, beet, and maize crops with lindane, the gamma-HCH content in
sandy loam soil was 0.32 mg/kg in the field occupied by maize and
0.68-0.70 mg/kg in the fields with potatoes and beet. After nine
months, the lindane content in the beet fields had decreased 14
times and that in the maize fields by only 1.3 times (Kovaleva &
Talanov, 1973; see Izmerov, 1983).
4.2.1.2 Degradation under submerged conditions
Half-time values for lindane of a few to about 120 h were
determined after incubation in various submerged soil samples. More
rapid degradation occurred in soils with a high amino acid content,
and the rate also clearly depended on the number of degrading
microorganisms present (Ohisa & Yamaguchi, 1979). The rapidity with
which lindane was degraded under flooded conditions varied in soil
samples from different locations in Japan. Enrichment of the soils
with peptone and exclusion of oxygen increased the degradation rate
(Ohisa & Yamaguchi, 1978a).
Half-time values of 10-30 days were observed in a comparison of
four Philippine rice soils under flooded conditions at a temperature
of 30 °C. Lindane was degraded faster at higher temperatures
(Yoshida & Castro, 1970). In a similar study with five Indian rice
soils at 28 °C, 14C-labelled lindane was degraded at half-times of
between 10 days and more than 41 days. Addition of rice straw
enhanced the degradation (Siddaramappa & Sethunathan, 1975).
Tsukano (1973) found a half-time for lindane of 10-14 days in
soil samples mixed with water. The degradation was almost completely
inhibited after addition of sodium azide to the soils, indicating
that the degradation observed in non-sterilized soils was due to
microbial activity.
4.2.2 Degradation under field conditions
Nash (1983) used a microagroecosystem in which moist fallow
sandy loam was placed in a glass chamber at a depth of 15 cm, plants
were grown in the chamber and lindane was applied to the surface. A
half-time of 1-4 days was found for dissipation of lindane in the
soil.
In April 1954, formulations containing lindane were applied to
a sandy loam soil at rates of 2.25 and 4.5 kg/ha on field plots in
the Rhine valley, and loss of active ingredient was followed during
the subsequent 1.5 years using a biological test method. The
insecticidal activity disappeared rapidly during the following
vegetation period but remained almost constant in winter; further
degradation was observed during the second vegetation period. At the
end of the observation period, 3.5-5.5% of the lindane applied at
2.25 kg/ha remained, and 17-19.5% of that applied at 4.5 kg/ha: the
speed of degradation was therefore greater at the lower application
rate. Degradation was virtually identical when the lindane was
worked into the soil to a depth of 1-2 cm and when it was introduced
to a depth of 10 cm (Schmitt, 1956).
In a field test in Miami, Florida, USA, on silt loam and muck
soils, lindane was applied at the extremely high rates of 11.2 or
112.1 kg/ha. The initial half-time at the lower rate was 15.5 months
in muck soil and 4.75 months in loam soil. Degradation was slower at
the higher rate: the initial half-times were 28.8 months in muck
soil and 11.1 months in loam soil (Lichtenstein & Schulz, 1958a). In
an earlier study on the same field plots with the same application
rates, however, Lichtenstein & Schulz (1958b) found that most of the
material detected chemically was inactive in the bioassay and
therefore did not represent lindane. They concluded that the
breakdown of lindane is faster than it appeared to be using their
analytical method.
In an extensive study, sandy loam, silt loam, and muck soils on
plots in three midwestern states of the USA were treated with
lindane in 1954 at application rates of 1.1, 11.2, and 112.1 kg/ha
to a depth of 15.2 cm. After a 4.5-year follow-up, no lindane was
detected on plots treated with 1.1 kg/ha; but after application at
the higher rates (far in excess of normal rates), about 36% of the
applied dose remained. Two major factors that affect the persistence
of lindane in soils appear to be the amount of organic matter in the
soil and the climatic conditions of the area (Lichtenstein et al.,
1960).
The rates of loss of lindane were calculated by Wheatley (1965)
in 10 long-term field studies in the United Kingdom. When lindane
was applied to the soil surface, there was a 50% loss within 4-6
weeks and a 90% loss within 30-40 weeks. When the lindane was mixed
into the soil, a 50% loss was observed after 15-20 weeks and 90%
within 2-3 years. No lindane was recovered 13 years after the last
application of lindane to a loam soil in Nova Scotia at a rate of
0.84-1.7 kg/ha (Stewart & Fox, 1971). Cliath & Spencer (1971)
treated two test plots in California, USA, with 21 kg/ha, which is
an application rate about 20 times above normal. A half-time of 8
months was found in sandy loam and 10 months in silty clay.
After application of lindane on three test plots of light sandy
soil in the Netherlands for 15 years, to give total amounts of 6.5,
13.0, and 24.3 kg/ha, only 3, 4, and 8% of the applied amount,
respectively, was recovered in layers to a depth of 20 cm (Voerman &
Besemer, 1970). A further follow-up of these plots for four years
showed rapid disappearance on the two locations with the lower
application rates; slower degradation was seen on the plot that had
received the highest application, where lindane was found to a depth
of 40 cm (Voerman & Besemer, 1975). Admixture of a 5% lindane dust
to the top 15-cm layer of a test plot at a rate of 10 kg of active
ingredient per hectare in India led to an initial concentration of
3.2 mg/kg soil. After an observation period of 180 days, 97.7% of
the applied lindane had disappeared. The initial half-time of
lindane in this study was about 30 days (Agnihotri et al., 1977).
The degradation of gamma-HCH was also determined in a variety
of studies in which technical-grade HCH was applied to soils. In
most of these investigations, the application rates were extremely
high, and in some, applications were made once a year for several
years (Lichtenstein & Polivka, 1959; Stewart & Chisholm, 1971;
Shiota & Kanda, 1972; Nash et al., 1973; Jackson et al., 1974;
Suzuki et al., 1975). Under these conditions, gamma-HCH disappeared
slowly from the soils and sometimes persisted for long periods.
The distribution of HCHs was studied in soil treated with
BHC-20 (containing 70% alpha-HCH, 6.5% beta-HCH, 13.5% gamma-HCH,
and 5% delta-HCH) in an agricultural area. The concentrations
changed with time after application; the mean value for gamma-HCH
was 16 µg/kg. The organic carbon content of the soil appeared to be
of primary importance, and the significant decrease in isomer
concentration observed with greater soil moisture was attributed to
microbial degradation, which is favoured by these conditions
(Chessells et al., 1988).
Kathpal et al. (1988) studied the behaviour of a formulation
consisting of a mixture of five HCH isomers in paddy soils under
subtropical conditions in India. The recommended application rate of
2.5 kg active ingredient per hectare and a rate of 5.0 kg/ha were
used. Gamma-HCH had dissipated by 50-63% within three months under
paddy, and average residues in soil at harvest were 0.3-0.34 mg/kg.
Dissipation after nine months (two crop seasons) was 98%. The
persistence under paddy in this study was fairly high, probably
owing to the anerobic conditions, which slow microbial degradation.
The paddy plants absorbed gamma-HCH from the soil: the residues at
harvest were about 1.0 mg/kg in plants and 0.03-0.06 mg/kg in seeds.
4.2.3 Hydrolytic degradation
Determination of the hydrolytic stability of a substance
provides an indication of whether this process can contribute to the
disappearance of the substance from the aquatic environment and, to
a certain extent, from soil. In a model experiment, the half-time of
lindane at 22 °C was 47.9 h at pH 9 and 100.7 h at pH 7; no
measurable hydrolysis occurred at pH 5 (Heupt, 1983).
4.2.4 Photolytic degradation (laboratory studies)
As lindane has measurable volatility and can be found at low
levels in air, its degradation in sunlight has been studied.
Carbon dioxide was formed after 14C-lindane was adsorbed onto
silica-gel plates at a concentration of 33 µg/kg and irradiated with
artificial sunlight (> 290 nm) in the presence of air; 6.4% of the
carbon was oxidized within 17 h. This photo-induced oxidation was
enhanced when the lindane was exposed to pure oxygen during
irradiation (Kotzias et al., 1981). No measurable degradation (less
than 0.5%) was observed 2000 h after exposure of lindane to the
light of a Xenon lamp in a Xenotest 150 on the wall of a quartz
vessel (solid phase). When the irradiation was performed in aqueous
solution, about 4% of the applied lindane was degraded after 2000 h.
The main degradation product was PCCH (Gardais & Scherrer, 1979).
Irradiation of lindane with ultra-violet light (254 nm) is
obviously more effective for degradation of the compound than
irradiation with light of longer wavelengths. Hamada et al. (1981,
1982) found rapid degradation of lindane in both the crystalline
state and in solution with 2-propanol under these conditions, with
PCCHs and TCCHs as reaction products. Eichler (1977) also found
rapid degradation of lindane in the solid or gaseous form and in
aqueous solution in the presence of ultra-violet irradiation, with
half-times of 12-24 h for the first two phases and 1-2 days for the
latter two.
4.2.5 Biodegradation in water
In a study of the degradation of lindane in a biological
purification plant, 75% of the compound was degraded within 6 h
(Eichler et al., 1976).
Newland et al. (1969) investigated the degradation of gamma-HCH
in simulated lake impoundments. Sediments from Lake Tomahawk,
Wisconsin, USA, were added to solutions of 5 mg/litre 14C-labelled
lindane and equilibrated for 18 h, and aerobic and anaerobic tests
were run for approximately 88 days. Initially, about 45% of the
applied lindane was adsorbed to the sediment (200 g per 3-litre
solution). Under aerobic conditions, about 16% of the added lindane
was degraded by the end of the observation period, whereas more than
97% was degraded under anaerobic conditions. When lindane
degradation was tested in samples of surface water from two
different regions for periods of 3, 6, or 12 weeks, decreases of up
to 90% of the initial concentration were found. Most of the lindane
was metabolized by microorganisms in the sediments: In samples of
sediment and water autoclaved prior to treatment and incubation, up
to 95% of the applied lindane was still present (Oloffs et al.,
1973).
In a field test in rice fields in the Camargue, France, a
formulation containing lindane was applied at a rate that resulted
in an initial concentration in water of 54.8 mg/m3. Rapid
disappearance was observed, for a half-time of about 1.5 days, and
within 10 days the concentration had dropped to the background value
of 0.08 mg/m3 (Podlejski & Dervieux, 1978).
The degradation of lindane was also tested in the water of a
drainage canal in the Holland Marsh, Ontario, Canada, in distilled
water, and in both water types after sterilization. The half-time of
lindane in the natural water was about six weeks, but a very low
disappearance rate was seen in the distilled and sterilized water,
indicating the importance of microbial action for degradation of
lindane in water (Sharom et al., 1980).
An aquatic model ecosystem, with pond water, sludge, aquatic
plants, and fish, was used to study the decomposition and migration
of lindane. In water without hydrobionts, the half-time was 50 days.
When sludge and aquatic plants were present, the half-time was 34
days, and that in the presence of fish was 2 days (Vrochinsky, 1973;
see Izmerov, 1983).
4.2.6 Microbial degradation
A variety of experiments on the degradation of lindane was
performed with mixed populations of the microorganisms that occur in
different types of soil, in aquatic sediments (Newland et al., 1969;
Benezet & Matsumura, 1973), and in other types of soil under
aerated, submerged, and strictly anaerobic conditions (Macrae et
al., 1967; Yule et al., 1967; Kohnen et al., 1975; Mathur & Saha,
1975, 1977; Tu, 1975; Haider, 1979). The fact that lindane was
removed faster from non-sterile than from autoclaved soil
demonstrated that its degradation in soil is due to microbial
activity (Macrae et al., 1967; Kohnen et al., 1975).
The microorganisms shown by screening experiments to be capable
of metabolizing and degrading lindane are as follows (Tu, 1976;
Jagnow et al., 1977):
Bacteria Fungi Algae
Arthrobacter sp. Penicillium sp. Chlamydomonas sp.
Bacillus sp. Rhizopus sp. Chlorella sp.
Citrobacter sp.
Clostridium sp.
Enterobacter sp.
Micromonospora sp.
Pseudomonas sp.
Thermoactinomycetes sp
In addition, lindane was metabolized in cell-free preparations of
Clostridium sp. in vitro (Heritage & Macrae, 1977a; Ohisa et
al., 1980).
Lindane is degraded by soil microorganisms under aerobic as
well as under anaerobic conditions, but anaerobic conditions are the
most favourable for its metabolism (Newland et al., 1969; Haider &
Jagnow, 1975; Vonk & Quirijns, 1979). In an anaerobically grown
culture of Clostridium sphenoides supplemented with lindane at 5
mg/litre, none was found, even after 2 h (Heritage & Macrae, 1979).
Several species of soil bacteria that have been shown to degrade
lindane effectively are described in detail in section 6.6.2.
In field studies in which gamma-HCH was applied at excessive
doses, it was degraded more slowly than at doses closer to those
used for normal agricultural applications. Introduction of HCH at up
to 224 kg/ha, corresponding to 33.6 kg gamma-HCH per hectare,
exceeded the degradation capacity of soil microorganisms for a long
period (Nash et al., 1973). In addition, the analytical methods used
might have resulted in an overestimation of the actual gamma-HCH
concentration, as concluded by Lichtenstein & Schulz (1958b).
Therefore, studies in which technical-grade HCH is applied,
especially at excessive rates, cannot be used to evaluate the
degradability of lindane in soil.
4.2.7 Bioaccumulation/Biomagnification
4.2.7.1 n-Octanol/water partition coefficient
The n-octanol/water partition coefficient (Pow) of lindane
was determined in several studies, with good agreement, covering the
narrow range of log Pow = 3.29-3.72 (Kurihara et al., 1973;
Platford, 1981; Darskus, 1982; Geyer et al., 1982; Hermens &
Leeuwangh, 1982; Geyer et al., 1984). These values indicate that
lindane can become enriched in lipid-containing biological
compartments.
4.2.7.2 Aquatic environment
The bioconcentration factor for lindane was found to be
dependent on the concentration to which the organisms, such as
algae, crustaceae, and fish, were exposed (Bauer, 1972; Ernst, 1975;
Schimmel et al., 1977; Trautmann & Streit, 1979; Marcelle & Thome,
1983): The highest bioconcentration factors were seen with the
lowest exposure concentrations. For example, Marcelle & Thome (1983)
determined the residues of lindane in brain, liver, and muscle of
the gudgeon (Gobio gobio) after exposure to concentrations of
0.22-142 µg/litre lindane in water. At the lowest concentration, the
bioconcentration factors in brain, liver, and muscle were about 600,
200, and 100, respectively, but they decreased to values of less
than 50 at higher concentrations.
Mouvet (1985) transplanted the freshwater aquatic moss
Cinclidotus danubicus from an uncontaminated area to a river that
received the effluent from an insecticide factory and determined
gamma-HCH concentrations in water and moss 13, 24, and 51 days after
the transplant. A three-fold increase in the gamma-HCH level was
found, with a bioconcentration factor of 294.
In a variety of aquatic organisms exposed to contaminated
water, the bioconcentration factor for lindane ranged from 13 to
1000 on a wet weight basis (Table 3).
Table 3. Bioconcentration factors of lindane in laboratory
experiments; test organisms were exposed to contaminated
water for the specified time
System Exposure Exposure Bioconcentration Reference
time concentration factora
(µg/litre)
Algae
Cladophora sp. up to 80.0 180 (d) Bauer (1972)
48 h 3.9 341 (d)
Nitzschia 24 h 6.1 1500-4700 (v) Trautmann & Streit
actinastroides 4400-12 400 (d) (1979)
Molluscs
Aplysia punctata 3-6 days 9000 201-436 (w) Chabert & Vicente
(1978)
Mya arenaria 5 days 5 40 Butler (1971)
Mercenaria 5 days 13
mercenaria
Mytilus edulis ns 2.61 74 (w) Ernst (1975)
0.02 242 (w)
Mytilus edulis ns 2-5 139 (w) Ernst (1979)
Venerupis japonica 3 days 1 121 (ns) Yamato et al.
(1983)
Annelidae
Lanice conchilega ns 2-5 1240 (w) Ernst (1979)
Crustacea
Penaeus duorarum 96 h 0.23 143 (ns) Schimmel et al.
Palaemonetes pugio 96 h 1.0 80 (ns) (1977)
Table 3 (contd)
System Exposure Exposure Bioconcentration Reference
time concentration factora
(µg/litre)
Insects
Sigara striata and 1 day 10 70-100 Kopf & Schwoerbel
Sigara lateralis (1980)
Fish
Lagodon 96 h 23.0 287 (ns) Schimmel et al.
rhomboides (1977)
Cyprinodon 96 h 108.7 727 (ns)
variegatus
Leuciscus idus ns 10-500 765 (ns) Sugiura et al.
Cyprinus carpio 281 (ns) (1979)
Salmo truttafario 442 (ns)
Poecilia 938 (ns)
reticulata
Poecilia 4 days 1 697 (ns) Yamato et al.
reticulata (1983)
Salmo gairdneri 27 days 30-290 319 Ramamoorthy
(1985)
a Calculated on the basis of: wet weight (w), dry weight (d), volume (v); ns, not specified
Another approach to the study of the bioconcentration of
lindane is the use of systems that simulate natural conditions,
taking into account sedimentary absorption processes and the
influence of contaminated food. The bioconcentration factors for
brine shrimp, mosquito larvae, and the brook silverside
(Haludesthes sicculus sicculus) exposed to lindane applied to the
sand of a test aquarium were 95, 220-383, and 600-1613,
respectively, depending on the food chains used (Matsumura &
Benezet, 1973). Marcelle & Thome (1984) investigated the
bioconcentration of lindane in the gudgeon (Gobio gobio) in
relation to the route of exposure. Fish were exposed either to
contaminated water alone or additionally to contaminated food. After
18 days, the group fed contaminated food had a 2.5-fold higher level
of lindane residues in liver than the group exposed to contaminated
water alone. Within three days after cessation of exposure, 98.4% of
the lindane residues had been excreted.
The uptake, transport, and bioconcentration of lindane were
also studied in a freshwater food chain, which consisted of
Chlorella sp., Daphnia magna, and Gasterosteus aculeatus
(algae-crustacea-fish). Uptake from water was more rapid than uptake
from food and depended on the duration of the experiment and the
feeding rate. The increase in lindane residues in the last link of
the food chain (fish) was not directly proportional to the
concentration found in the primary links (Hansen, 1980).
4.2.7.3 Terrestrial environment
The bioconcentration of lindane was investigated in a
terrestrial food chain, which consisted of soil, barley,
caterpillar, and quail. Doses up to 400 times the standard
agricultural dose (50 and 200 mg/kg soil) were applied to the soil.
Although lindane was found in all of the links of the food chain,
the concentrations decreased progressively (Dugast, 1980).
Feeding hens diets containing lindane at 0.05, 0.15, or 0.45
mg/kg for 20 weeks resulted in constant values of 0.01, 0.03, and
0.09 mg/kg of eggs, demonstrating a dose-related accumulation of
lindane (Cummings et al., 1966).
Several studies are available on the bioconcentration of
lindane in rats. After seven rats had received daily doses of 2 or 4
mg/kg body weight for up to 12 weeks, gamma-HCH was found at a
concentration of about 8 mg/kg in adipose tissues of the group that
had recived the high dose (Jacobs et al., 1974). In another
experiment, four generations of rats were fed a diet containing 20%
fat and a mixture of insecticides including lindane at levels of
0.07-0.8 mg/kg. Even in the F3 generation, the levels of gamma-HCH
residues in adipose tissues were of the same order of magnitude
(< 0.05-0.56 mg/kg) as those of lindane in the diet (Adams et al.,
1974). No accumulation occurred, even in four consecutive
generations.
Accumulation factors have been determined from feeding studies
in rats (Fitzhugh et al., 1950; Oshiba, 1972; Baron et al., 1975;
Suter et al., 1983). In comparison to the concentration of lindane
in the diet, the highest reported bioconcentration factor was about
2 for adipose tissue. The average bioconcentration factor for
adipose tissues in rats derived from all these studies is 1; the
bioconcentration factors for other tissues are considerably lower.
4.2.7.4 Bioconcentration in humans
Geyer et al. (1986) examined data on environmental chemicals
detectable in adipose tissue and/or breast milk of
non-occupationally exposed humans and concluded that, in
industrialized countries, more than 90% of human exposure to HCHs
originates from food. Mean concentrations of gamma-HCH in human
adipose tissue in Czechoslovakia, the Federal Republic of Germany,
and the Netherlands were 0.086, 0.024-0.061, and 0.01-0.02 mg/kg,
respectively, on a fat basis. The mean bioconcentration factor,
calculated on the basis of the concentration in the diet (2.3, 5.0,
and 0.62-0.9 µg/kg, respectively) and levels in adipose tissue, was
18.6 ± 9 on a lipid basis (range, 10.4-32.5). Greve & Wegman (1985)
found an accumulation factor (adipose tissue/blood) of 70 for
gamma-HCH in humans.
4.2.7.5 Field studies
The bioconcentration of lindane was investigated by
environmental monitoring in aquatic ecosystems. The residue levels
found in different organisms were related to the environmental
background levels, and these data were used to calculate the
bioconcentration factors.
The bioconcentration factor for gamma-HCH in sea water and
bladder wrack (Fucus vesiculosus) in the Husum estuary and the
adjacent North Friesian Wadden Sea in the Netherlands was about 150
(Herrmann et al., 1984). On the basis of the data given in section
5.1.5.2 on the occurrence of gamma-HCH in muscle and fat of bream
collected in the River Elbe, a bioconcentration factor of 10 000 to
50 000 was calculated (Arbeitgemeinschaft für die Reinhaltung der
Elbe, 1982).
Frisque et al. (1983) studied the accumulation of lindane by
bryophytes (Cinclidotus danubicus and C. nigricans) in the Meuse
River and found a concentration factor of 300-350. The average level
in the river was 0.067 µg/litre. Hartley & Johnston (1983) found a
bioconcentration factor for the freshwater clam Corbicula
manilensis of 2610 on a lipid basis; and Cosson Mannevy & Marchand
(1980) found a mean factor of 26 198 (on a dry-weight basis) in
Mytilus edulis.
On the basis of the concentrations of gamma-HCH in sea water,
sediments, and fish from the Mediterranean Sea, El-Dib & Badawy
(1985) calculated a bioconcentration factor of about 1000 (on a
lipid basis). Tanabe et al. (1984) reported bioconcentration factors
for total HCHs in a trophic chain in the western North Pacific. As
the contribution of gamma-HCH to the residue levels was determined,
the bioconcentration factors for this isomer can be estimated to be
about 40, 40, 100, and 1850 for zooplankton, myctophid, squid, and
dolphin, respectively.
5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE
5.1 Environmental levels
5.1.1 Air
An average of 0.23 ng/m3 (0.039-0.68 ng/m3) gamma-HCH was
found in 24 samples of air taken from over the western Pacific, the
eastern Indian Ocean, and the Antarctic Ocean (Tanabe et al., 1982).
Levels of gamma-HCH in the air of various regions of the USA
were within a similar range (US Environmental Protection Agency,
1976). The levels were below 1 ng/m3 in most samples, and values
up to 16.2 ng/m3 were found in only two regions.
Gamma-HCH was found at an average concentration of 0.14 ng/m3
in the air of unpolluted areas in the Federal Republic of Germany in
1972; in polluted areas (the Ruhr), a level of 0.8 ng/m3 was found
in 1976/77 (Deutsche Forschungsgemeinschaft, 1983; Hildebrandt et
al., 1986). It occurred at 0.52-11 ng/m3 in a location with heavy
traffic near Ulm in the Federal Republic of Germany and at 0.18-1.1
ng/m3 in a rural area. The authors concluded that the
concentrations in the lower troposphere under different
meteorological conditions reflect regional input and long-range
transport (Wittlinger & Ballschmiter, 1987).
The average concentration of gamma-HCH in 55 air samples taken
near Delft, the Netherlands, in 1979-80 was 0.36 ng/m3 (maximum,
3.4 ng/m3); in three other locations in the Netherlands, average
levels of 0.2-0.9 ng/m3 were found. In six houses built on former
dumping grounds, the average concentration of gamma-HCH was 6
ng/m3 (range, 1-14 ng/m3), whereas in the space beneath the
floor the level was below the detection limit (1 ng/m3). Outdoor
concentrations in this area were 0.3-0.4 ng/m3. In another study,
the concentrations of gamma-HCH in the space beneath the floor of
houses were 90 ng/m3. Much higher levels were found in houses
treated with lindane-containing products for the control of woodworm
or of long-horned beetle. Peak levels of 51-61 µg/m3 were found
four weeks after application; these decreased gradually to 8-24
µg/m3 after 10 weeks. After indoor application of lindane for wood
preservation, levels of 50 µg/m3 were common, with peak levels of
up to 100 mg/m3 (Sloof & Matthijsen, 1988).
5.1.2 Water
5.1.2.1 Rain and snow
Levels of 0.001-0.005 µg/litre were found in rain-water
analysed in the Federal Republic of Germany in 1970-72 (Mestres,
1974); in 1983, gamma-HCH was found at an average of 0.06 (range,
0.01-0.18) µg/litre in rain-water near de Bilt, the Netherlands
(Slooff & Matthijsen, 1988).
Strachan et al. (1980) found traces of gamma-HCH in 17 samples
of snow collected from the Canadian side of the Great Lakes in 1976
and 5-12 ng/litre in 81 samples of rain-water collected in 1976 and
1977.
5.1.2.2 Fresh water
Water samples from selected rivers in Yorkshire, United
Kingdom, analysed for gamma-BHC in 1966 contained levels of
0.001-0.18 µg/litre; in 1968, however, the highest value was 0.622
µg/litre. Water samples from six other rivers, also analysed in
1968, contained mean values of 0.011-0.030 µg/litre, and the highest
levels found were 0.020-0.098 µg/litre (Lowden et al., 1969).
River water samples analysed in 1969-72 in Belgium, France, the
Federal Republic of Germany, the Netherlands, and Italy contained
less than 0.1 µg/litre and usually less than 0.05 µg/litre. In 1826
water samples taken at 99 sites in the Netherlands in 1966-77, the
highest concentrations of gamma-HCH were found in those from the
River Rhine and its tributaries. The concentrations of gamma-HCH
over the period 1969-74 varied between 0.01 and 0.4 µg/litre, but in
1974-77, the concentrations were all below 0.1 µg/litre (Mestres,
1974). Gamma-HCH concentrations have been measured in the Rivers
Rhine, Meuse, and West-Scheldt and in other surface waters in the
Netherlands since 1969. Since 1974-75, the levels have been below
0.05 µg/litre in the Rhine and about 0.05 µg/litre in the
West-Scheldt; in the Meuse, the concentrations were more variable
and ranged from 0.01 to 1.0 µg/litre. In agricultural and
horticultural areas, the levels were 0.01-1.0 µg/litre, with
incidental peaks up to 0.5 µg/litre, probably due to use of lindane.
The average concentration of dissolved gamma-HCH in the Meuse-Rhine
estuary in 1974 was 20 ng/litre and that of suspended gamma-HCH
between 1 and 20 ng/litre. In coastal waters of the Netherlands, the
concentration of dissolved gamma-HCH was 0.9-4.6 ng/litre and that
of bound gamma-HCH, 3.1-8.7 ng/litre (Sloof & Matthijsen, 1988).
A sampling trip along the River Rhine, from Rheinfelden in
Switzerland to Rotterdam in the Netherlands, proved that the source
of alpha-, beta-, and gamma-HCH was located in the upper reaches of
the River. In the Meuse, lindane levels in 1969-77 were all below
0.1 µg/litre (Wegman & Greve, 1980). In an extensive programme in
1982 to determine pollution in Dutch surface waters at 45 locations,
gamma-HCH concentrations were generally between 0.01 and 0.1
µg/litre (Wammes et al., 1983).
The mean concentration of gamma-HCH in the River Elbe, from
Schnackenburg to the North Sea, in 1981-82 was 0.021
(< 0.001-0.051) µg/litre; during February-November 1988, the
concentrations were 0.005-0.044 µg/litre (Arbeitsgemeinschaft für
die Reinhaltung der Elbe, 1988). More figures for Germany are given
by Wirth (1985). Gamma-HCH was found at three locations in the River
Rhine at 0.02 µg/litre and in six side-rivers at 0.01-0.06 µg/litre.
These levels had decreased markedly since 1975 (Landesamt für Wasser
und Abfall, 1988).
5.1.2.3 Sea water
Atlas & Giam (1981), Bidleman & Leonard (1982), Oehme & Stray
(1982), and Oehme & Mano (1984) analysed water from such widely
differing areas as the Eniwetok Atoll in the North Pacific, the
Arabian Sea, the Persian Gulf, the Red Sea, Lillestrøm, Norway, Bear
Island, and Spitzbergen in the Arctic Ocean. The gamma-HCH
concentrations were in the range 0.01-0.05 ng/litre, except in the
Arabian Sea, the Persian Gulf, and at Lillestrøm, where levels up to
0.67 ng/litre were found (Slooff & Matthijsen, 1988). Levels of
0.0001-0.004 µg/litre gamma-HCH were measured in the Western
Pacific, the Eastern Indian, and Antarctic Oceans (Tanabe et al.,
1982). No gamma-HCH was found in 60 water samples from the Japan Sea
and Pacific Ocean (detection limit, 0.1 µg/litre) (A. Hamada, letter
to M. Mercier, dated 28 July 1989; T. Onishi, letter to M. Mercier,
dated 24 July 1989). The levels detected in water from the North Sea
and the Arctic Sea are of the order of 0.001-0.02 µg/litre (Deutsche
Forschungsgemeinschaft, 1983). The maximal level of gamma-HCH in
North Sea water in 1972 was 0.028 µg/litre; 5-10% of the samples
contained gamma-HCH (Mestres, 1974). The level of gamma-HCH in
surface-water of the North Sea in June-July 1986 ranged from 1.0 to
4.0 ng/litre. The highest concentrations were found close to the
coast (Umweltbundesamt, 1988-89).
5.1.3 Soil
Traces of gamma-HCH are transmitted to soil by precipitation;
the resulting contamination is generally below the limit of
detection (0.0001-0.001 mg/kg). Application of lindane in
agricultural areas can result in higher concentrations: levels in
some German districts were mainly in the range 0.001-0.01 mg/kg, but
in certain fields up to 0.6 mg/kg was found (Fricke, 1972).
Edelman (1984) analysed 96 samples of the upper 10 cm of soil
from 38 natural reserves in the Netherlands for gamma-HCH: 59
samples contained < 1 µg/kg, 21 contained 1-10, 9 had 10-20, and 7
had 20-80 µg/kg (Slooff & Matthijsen, 1988). In the National Soils
Monitoring Program of the US Environmental Protection Agency (Carey
et al., 1979), several thousand samples from cropland sites were
analysed for residues; no gamma-HCH residues were detected in more
than 99% of the samples. In the Ukraine, however, 36 of 136 soil
samples taken at various locations contained lindane at levels of
0.1-5 mg/kg (Talanov, 1977; see Izmerov, 1983).
In a study on the application of lindane dust by aircraft on
mosquito breeding sites at 1.3 kg/ha, the gamma-HCH content of the
soil was 1 mg/kg; after one year, the level was 0.01 mg/kg
(Vroschinsky, 1973; see Izmerov, 1983).
5.1.3.1 Sediment
Gamma-HCH was present in three of six samples of sediment taken
from Nyumba Ya Mungu Lake in the United Republic of Tanzania in
1986, at a concentration of 1-4 µg/kg dry weight (Paasivirta et al.,
1988).
Martin & Hartmann (1985) found gamma-HCH at levels above the
detection limit (5 µg/litre) in less than 4% of 117 samples of
sediment taken in 1980-82 from riverine and pothole wetlands in
north-central USA. In less than 4% of the samples, gamma-HCH was
present at above the detection level of 5 µg/kg.
In Japan, gamma-HCH was found in 9 out of 60 samples of
sediment at a concentration of 10 µg/kg in 1974 (A. Hamada,letter to
M. Mercier, dated 28 July 1989; T. Onoshi, letter to M. Mercier,
dated 24 July 1989).
The median levels of gamma-HCH in sediments from eight rivers,
harbours, and sites close to dumping places in the Netherlands were
15-342 µg/kg dry matter (Slooff & Matthijsen, 1988).
5.1.3.2 Dumping grounds and sewage sludge
The soil at various locations in the Netherlands is polluted
with HCHs as a result of spillage during production, storage, and
handling of this chemical during the 1950s. The concentrations found
range up to a few thousand milligrams of HCHs per kilogram of dry
soil. Further pollution has been caused by the dumping of chemical
waste, sometimes in order to level the ground; this waste can be
dispersed from dumping areas by leaching or wind erosion. In certain
polluted areas, high concentrations of HCHs (mainly alpha- and
beta-HCHs) were found at depths of more than 2 m below ground level.
In 18 locations in the Netherlands, the average concentrations of
gamma-HCH in sewage sludge in 1981 were 8-50 µg/kg dry matter.
Groundwater was also found to be polluted, but this was restricted
to the vicinity of the production areas; horizontal transportation
of HCHs in groundwater appeared to be limited (Slooff & Matthijsen,
1988).
Fieggen (1983) found gamma-HCH in sewage sludge at mean values
of 25 µg/kg dry matter in 1975, 43 µg/kg in 1978, and 12 µg/kg in
1981.
5.1.4 Drinking-water, food and feed
Although in most countries nowadays only lindane is used,
residues of alpha- and beta-HCH can still be found in crops and
animal products originating from regions where technical-grade HCH
(containing all of the HCH isomers) is still in use.
5.1.4.1 Drinking-water
Gamma-HCH was found at 0.0001-0.001 µg/litre in water from 19
lakes in Germany and at levels below 0.001 µg/litre (0.0001-0.0008
µg/litre) in the drinking-water derived from them (Bernhardt &
Ziemons, 1974). In the USA, only 3% of drinking-water samples
examined contained gamma-HCH, in a range of 0.001 to about 0.1
µg/litre (US Environmental Protection Agency, 1976). In Ottawa,
Canada, drinking-water samples collected in 1976 contained 0.4-11
ng/litre (Williams et al., 1978).
5.1.4.2 Cereals, fruits, pulses, vegetables, and vegetable oil
The large body of information on gamma-HCH residue levels in
crops grown and treated with this chemical according to Good
Agricultural Practice has been reviewed comprehensively by the
FAO/WHO Joint Meeting on Pesticide Residues and summarized in
published monographs (FAO/WHO, 1967, 1968, 1969, 1970, 1974, 1975,
1976, 1978, 1980).
In samples of ready-to-eat foods collected from 30 markets in
27 US cities in 1966-67, gamma-HCH levels were 0.003-0.009
(occasionally 0.06) mg/kg in grains and cereals, 0.002-0.027 mg/kg
in garden fruits, 0.001-0.005 mg/kg in potatoes, 0.002-0.007 mg/kg
in leafy vegetables, and 0.004-0.012 mg/kg product in oils, fat, and
shortening (Martin & Duggan, 1968). In 1967-68, residues of
gamma-HCH were found at 0.002-0.006 in leafy and root vegetables, at
0.002-0.003 in garden fruits, and at 0.029-0.085 mg/kg product in
oils, fat, and shortening (Corneliussen, 1969).
In monitoring studies carried out on grain in the Federal
Republic of Germany at one-year intervals since 1975, gamma-HCH
residues in wheat and barley were 0.001 mg/kg or less (Ocker, 1983).
More than 800 samples of cereal and cereal products analysed in
Germany in 1975-78 and 1979-83 contained mean concentrations of
0.0009-0.04, but cereal products had up to 0.11 mg/kg. The mean
concentration of gamma-HCH in 200 samples of wheat and rye collected
in 1986 and 1987 was 0.06 mg/kg, with a maximum of 0.3 mg/kg
(Umweltbundesamt, 1988-89).
Of 281 samples of wheat analysed for the presence of gamma-HCH
in the United Kingdom between October 1978 and April 1979, 71
contained levels in the range 0.002-0.04 mg/kg. Gamma-HCH was also
found in one sample of polished rice from Spain, at a concentration
of 0.008 mg/kg (Steering Group on Food Surveillance, 1982).
Gamma-HCH was found in 16% of samples of imported maize in the
United Kingdom in the range none detected to 0.007 mg/kg, and in 28
samples of different types of pulses at none detected to 0.05 mg/kg.
Of retail cereal products, only bran and wheat contained detectable
levels (0.01 mg/kg product) of gamma-HCH in 1982 (Steering Group on
Food Surveillance, 1986). In 1986-87, 31 of 142 samples of pulses
contained residues; in nine, levels of < 0.01-0.4 mg/kg were found.
Peanut butter and vegetable oils contained 0.01 mg/kg (Steering
Group on Food Surveillance, 1989).
About 80-90% of samples of fruit, potatoes, and other
vegetables analysed in the Federal Republic of Germany contained no
detectable residues of gamma-HCH (Weigert et al., 1983). The
remaining 10-20% had mean levels up to 0.01 mg/kg, with no
significant difference between 360 samples originating from
conventional agriculture and 360 samples from 'alternative'
agriculture (Vetter et al., 1983). In 1976-78 and 1980, the mean
concentrations of gamma-HCH were < 0.001-0.002 mg/kg product in
more than 400 samples of fruit, potatoes, and other vegetables. In
the Netherlands, residues in fruit and vegetables were generally in
the range 0-0.1 mg/kg, although some leafy crops, such as endive,
lettuce, celery, and leek, contained levels up to 5 mg/kg. Samples
of wheat contained only 0-0.05 mg/kg, with a few measurements up to
0.2 mg/kg (FAO/WHO, 1978). In France, gamma-HCH residues were found
in wheat at 0.01-0.02 mg/kg, and at low levels in other commodities,
such as carrots and endives (Laugel, 1981). Engst et al. (1967)
found that the gamma-HCH content of carrots grown from seed treated
with this compound decreased continuously during the first 120 days.
At normal harvesting time, the early varieties contained 3-6 mg/kg
product, the mid-season varieties about 2 mg, and the late
varieties, 0.4 mg/kg. When the carrots were harvested after 200
days, 0.3-0.7 mg/kg was present (independently of variety). Even
after 6 months' storage, low residues were still present.
5.1.4.3 Meat, fat, milk, and eggs
Martin & Duggan (1968) found gamma-HCH at levels of 0.09 mg/kg
in dairy products and at 0.01-0.03 mg/kg (with a peak of 0.374
mg/kg) in samples of meat, fish, and poultry collected from 30
markets in 27 cities in the USA in 1966-67. Residue levels in
samples of meat, fish, and poultry in 1967-68 were 0.003-0.026 mg/kg
(Corneliussen, 1969). No gamma-HCH or levels of 0.01-0.1 mg/kg were
found in 99% of samples of cow's milk and manufactured milk products
from Illinois (USA) (Wedberg et al., 1978). In milk samples
collected during Spring 1983 from 359 bulk transporters,
representing 16 municipalities of Ontario, Canada, gamma-HCH was
found in 68% of the samples at a mean concentration of 4.0 µg/kg
butter fat (Frank et al., 1985). Six samples of cow's milk from six
locations in Switzerland contained 3.0-5.1 mg/kg on a fat basis
(Rappe et al., 1987).
In about 25% of 976 samples of meat and poultry products
(including eggs) collected in the United Kingdom in 1984-86,
gamma-HCH was present at a mean concentration of 0.01-0.02 mg/kg.
The highest level, 3.7 mg/kg, was found in lamb. Processed meat and
poultry products (631 samples collected in 1985-87) contained mean
concentrations of 0.01-0.06 mg/kg product. About half of 849 samples
of retail milk and dairy products collected in 1984-87 contained
gamma-HCH at concentrations of 0.01-0.03 mg/kg; the highest level,
0.7 mg/kg, was found in milk (Steering Group on Food Surveillance,
1989). Imported meat products were also analysed in the United
Kingdom for the presence of alpha-, beta-, and gamma-HCH. No
detectable residue of gamma-HCH was found in beef or pork products:
processed pork contained none detectable to 0.03 mg/kg. Processed
poultry contained none detectable to 0.04 mg/kg (Steering Group on
Food Surveillance, 1986). In 1967-70, in the Ukraine, gamma-HCH was
found in cows' milk at an average concentration of 0.6 mg/litre
(Medvedev & Perepechkina, 1973; see Izmerov, 1983). In the USSR, the
following concentrations were found: milk and milk products, 0.055 ±
0.005 mg/kg; poultry and fish, 0.068 ± 0.021 mg/kg; butter, 0.003 ±
0.002 mg/kg; vegetables and fruits, 0.008 ± 0.003 mg/kg; groats and
flour, 0.005 ± 0.002 mg/kg (Sizova & Bogomolova, 1976; see Izmerov,
1983).
Concentrations of gamma-HCH were measured in 1250 samples of
milk and other dairy products in France in 1970-77 and in 1981. In
the first period, the gamma-HCH concentration was < 0.1 mg/kg of
fat; by 1981, the levels had declined to < 0.03 mg/kg of fat
(Laugel, 1981; Rhône-Poulenc Agrochimie, 1986). Higher levels (mean,
0.85 mg/kg) were found in animal fat, but meat and eggs generally
contained no detectable residue (Laugel, 1981). The mean levels of
gamma-HCH found in a large number of samples of various food items
in Germany (Hildebrandt et al., 1986) are shown in Table 4.
The levels of gamma-HCH in food items analysed in France were
0.006-0.01 mg/kg in 113 samples of vegetables, 0.005-0.04 mg/kg in
192 samples of fish and seafood, 0.005-0.041 mg/kg in 154 samples of
preserved meat, 0.007-0.017 mg/kg in 104 samples of cereal products,
0.007-0.034 mg/kg in 120 samples of butter and cheese, 0.005-0.059
mg/kg in 25 samples of oil and fat, and 0.006-0.021 mg/kg in 26
samples of fruit (Rhône-Poulenc Agrochimie, 1986).
Table 4. Levels of gamma-HCH (mg/kg) in food items in Germany
Food item 1973-78 1979-83 1973-83
Meata 0.004-0.04
Meat productsa 0.006-0.055
(maximum, 0.52)
Animal fata 0.007-0.09
(maximum, 0.5)
Gamea 0.042-4.072
Poultrya 0.01-0.05 0.004-0.046
(maximum, 0.471)
Chicken eggs < 0.001-0.01
Chicken eggsa,c 0.001-0.02
(maximum, 1.9)
Milk and milk productsa 0.05 0.01-0.02
Cow's milka,b 0.03 0.01
Vegetable oil and 0.01-0.02
margarinea
Oil seeds, nuts, pulses 0.001-0.127
Fish and fish products 0.01-0.02 0.002-0.009
Shell-fish and molluscs < 0.001-0.020
a From Hildebrandt et al. (1986); on fat basis
b From Anon. (1984)
c From Koelling (1978)
Skaftason & Johannesson (1979) found a mean value of 13 µg/kg
in 35 samples of butter from Iceland in 1968-70. Of 32 samples
analysed in 1974-78, only five contained gamma-HCH, at a mean value
of 7 ± 2 µg/kg. The mean concentration in meat, poultry and eggs in
the Netherlands in 1976-78 was 0.002 mg/kg (range, 0.001-0.004
mg/kg) (De Vos et al., 1984); the levels in dairy products were
similar.
Fifteen of 105 chicken eggs from seven areas in Kenya had a
median concentration of 0.01 mg/kg (range, 0.01-0.04 mg/kg) (Kahunyo
et al., 1988). Ten samples each from two lots of lamb and beef were
collected randomly from markets in Bagdad, Iraq, in 1983 and
analysed for the presence of gamma-HCH. An average concentration of
0.225 (0.004-1.611) mg/kg was found in lamb, and 0.116 (0.005-0.83)
mg/kg was found in beef (Al-Omar et al., 1985).
5.1.4.4 Animal feed
Of 114 samples of animal feed analysed in the United Kingdom in
1982-85, 49 contained gamma-HCH at concentrations up to 2.3 mg/kg
product (Steering Group on Food Surveillance, 1986).
5.1.4.5 Miscellaneous products
Lanolin produced from crude wool grease may contain gamma-HCH:
a level of 1.2 mg/kg was found in the USA (Anon. 1989); and Meemken
et al. (1982) found average levels of 2.4 and 2.1 mg/kg in 1976 and
1981, respectively, in Germany. Concentrat