Tributyltin compounds (EHC 116, 1990)

INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY

ENVIRONMENTAL HEALTH CRITERIA 116

TRIBUTYLTIN COMPOUNDS

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. S. Dobson,
Institute of Terrestrial Ecology, United Kingdom,
and Dr. R. Cabridenc, Institut National de
Recherche Chimique Appliquée, France

World Health Orgnization
Geneva, 1990

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WHO Library Cataloguing in Publication Data

Tributyltin compounds.

(Environmental health criteria ; 116)

1.Trialkyltin compounds - adverse effects 2.Trialkyltin compounds
-toxicity I.Series

ISBN 92 4 157116 0 (NLM Classification: QV 290)
ISSN 0250-863X

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CONTENTS

ENVIRONMENTAL HEALTH CRITERIA FOR TRIBUTYLTIN COMPOUNDS

1. SUMMARY

1.1. Physical and chemical properties
1.2. Analytical methods
1.3. Sources of environmental pollution
1.4. Regulations on use
1.5. Environmental concentrations
1.6. Transport and transformation in the environment
1.7. Kinetics and metabolism
1.8. Effects on microorganisms
1.9. Effects on aquatic organisms
1.9.1. Effects on marine and estuarine organisms
1.9.2. Effects on freshwater organisms
1.9.3. Microcosm studies
1.10. Effects on terrestrial organisms
1.11. Effects on organisms in the field
1.12. Toxicity to laboratory mammals
1.12.1. Acute toxicity
1.12.2. Short-term toxicity
1.12.3. Long-term toxicity
1.12.4. Genotoxicity
1.12.5. Reproductive toxicity
1.12.6. Carcinogenicity
1.13. Effects on humans

2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, ANALYTICAL METHODS

2.1. Identity of tributyltin compounds
2.2. Physical and chemical properties
2.3. Analytical methods
2.3.1. Measurement of organotin compounds
2.3.1.1 Extraction of tributyltin derivatives
2.3.1.2 Formation of volatile derivatives
2.3.1.3 Separation of organotin derivatives
2.3.1.4 Detection and measurement of different forms
of organotin
2.3.2. Interlaboratory calibrations

3. SOURCES OF ENVIRONMENTAL EXPOSURE

3.1. Uses
3.2. Production
3.3. Regulations

4. ENVIRONMENTAL TRANSPORT AND TRANSFORMATION

4.1. Adsorption onto and desorption from particles
4.2. Abiotic degradation
4.2.1. Hydrolytic cleavage of the tin-carbon bond
4.2.2. Photodegradation

4.3. Biodegradation
4.4. Bioaccumulation and elimination

5. ENVIRONMENTAL CONCENTRATIONS

5.1. Sea water and marine sediment
5.2. Fresh water and sediment
5.3. Sewage treatment
5.4. Biota

6 KINETICS AND METABOLISM

6.1. Metabolism of TBT in mammals
6.2. Metabolism of TBTO in other organisms
6.3. General mechanisms of toxicity of TBTO
6.3.1. General toxic mechanisms
6.3.2. Toxic mechanisms in bivalve molluscs

7. EFFECTS ON ORGANISMS IN THE ENVIRONMENT: MICROORGANISMS

7.1. Bacteria and fungi
7.2. Freshwater algae
7.3. Estuarine and marine algae

8. EFFECTS ON ORGANISMS IN THE ENVIRONMENT: AQUATIC ORGANISMS

8.1. Aquatic plants

8.2. Aquatic invertebrates
8.2.1. Trematode parasites of man
8.2.2. Freshwater molluscs
8.2.2.1 Acute toxicity
8.2.2.2 Short- and long-term toxicity
8.2.2.3 Factors affecting toxicity
8.2.3. Marine molluscs
8.2.3.1 Acute toxicity
8.2.3.2 Short- and long-term toxicity
8.2.3.3 Reproductive effects
8.2.3.4 Effects on growth
8.2.3.5 Shell thickening
8.2.3.6 Imposex
8.2.3.7 Genotoxicity
8.2.4. Crustaceans
8.2.4.1 Acute effects
8.2.4.2 Short- and long-term toxicity
8.2.4.3 Reproductive effects
8.2.4.4 Limb regeneration
8.2.4.5 Behavioural effects
8.2.5. Other aquatic invertebrates
8.2.5.1 Acute effects
8.2.5.2 Limb regeneration
8.3. Fish
8.3.1. Acute effects
8.3.2. Short- and long-term toxicity
8.3.3. Embryotoxicity
8.3.4. Behavioural effects

8.4. Amphibians
8.5. Multispecies studies

9. EFFECTS ON ORGANISMS IN THE ENVIRONMENT: TERRESTRIAL ORGANISMS

9.1. Microcosm studies
9.2. Terrestrial insects
9.3. Terrestrial mammals

10. EFFECTS ON ORGANISMS IN THE ENVIRONMENT: FIELD OBSERVATIONS

10.1. Effects on bivalves
10.2. Effects on gastropods: imposex
10.3. Effects on farmed fish
10.4. Effects of TBT-contaminated sediment
10.5. Effects of freshwater molluscicides
10.6. Effects from spills
10.7. The use of indicator species for monitoring the environment

11. EFFECTS ON EXPERIMENTAL ANIMALS AND IN VITRO TEST SYSTEMS

11.1. Single exposure
11.1.1. Oral and parenteral administration
11.1.2. Dermal administration
11.1.3. Administration by inhalation
11.1.4. Irritation and sensitization
11.1.4.1 Skin irritation
11.1.4.2 Eye irritation
11.1.4.3 Skin sensitization
11.1.5. In vitro studies
11.2. Short-term toxicity
11.2.1. Oral dosing: general body effects
11.2.2. Inhalation studies
11.2.3. Histopathological effects
11.2.4. Haematological and biochemical effects
11.2.5. Effects on lymphoid organs and immune function
11.2.6. Mechanism of immunotoxicity
11.2.7. Effects on the endocrine system
11.3. Long-term toxicity
11.4. Genotoxicity
11.5. Reproductive toxicity
11.5.1. In vivo
11.5.2. in vitro
11.6. Carcinogenicity

12. EFFECTS ON HUMANS

12.1. Ingestion
12.2. Inhalation
12.3. Dermal exposure
12.4. Miscellaneous effects

13. EVALUATION OF HUMAN HEALTH RISKS AND EFFECTS ON THE ENVIRONMENT

13.1. Evaluation of human health risks
13.2. Evaluation of effects on the environment

14. RECOMMENDATIONS

14.1. Recommendations for protecting human and environmental health
14.2. Research needs

REFERENCES

RESUME

EVALUATION DES RISQUES POUR LA SANTE HUMAINE ET EFFETS SUR L'ENVIRONNEMENT

RECOMMANDATIONS

RESUMEN

EVALUACION DE LOS RIESGOS PARA LA SALUD HUMANA Y DE LOS EFECTOS SOBRE EL
MEDIO AMBIENTE

RECOMENDACIONES

WHO TASK GROUP ON TRIBUTYLTIN COMPOUNDS

Members

Dr C. Alzieu, French Institute for Research on Exploi-
tation of the Sea, Nantes, France

Dr I.J. Boyer, Division of Toxicological Review and Evalu-
ation, Food & Drug Administration, Washington, DC, USA

Dr A.H. El-Sabae, Faculty of Agriculture, Alexandria Uni-
versity, Alexandria, Egypt

Dr B. Gilbert, Company for the Development of Technology
Transfer (CODETEC), Cidade Universitaria, Campinas,
Brazil

Dr Y. Hayashi, Biological Safety Research Centre, National
Institute of Hygienic Sciences, Setagaya-ku, Tokyo,
Japan

Dr R. Koch, Institute for Geography & Geoecology, Academy
of Sciences, German Democratic Republic (Chairman)

Dr E.I. Krajnc, National Institute for Public Health and
Environmental Hygiene, Bilthoven, Netherlands

Dr H. Schweinfurth, Schering AG, Chemical Industry,
Bergkamen, Federal Republic of Germany

Mr D. Spatz, Office of Pesticide Programs, US Environmen-
tal Protection Agency, Washington, DC, USA

Dr A.R.D. Stebbing, Natural Environment Research Council,
Plymouth Marine Laboratory, Plymouth, United Kingdom

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

Dr J.E. Thain, Ministry of Agriculture, Fisheries and
Food, Fisheries Laboratory, Burnham-on-Crouch, United
Kingdom

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

Observers

Mr J. Chadwick, Health and Safety Executive, Bootle,
United Kingdom

Dr R.J. Fielder, Department of Health, London, United
Kingdom

Dr R. Lange, Schering AG, Department of Experimental Toxi-
cology, Berlin, Federal Republic of Germany

Secretariat

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

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

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

NOTE TO READERS OF THE CRITERIA DOCUMENTS

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

* * *

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

ENVIRONMENTAL HEALTH CRITERIA FOR TRIBUTYLTIN COMPOUNDS

A WHO Task Group meeting on Environmental Health Cri-
teria for tributyltin compounds was held at the Institute
of Terrestrial Ecology (ITE), Monks Wood, United Kingdom,
from 11 to 15 September 1989. Dr M. Roberts, Director,
ITE, welcomed the participants on behalf of the host
institution and Dr M. Gilbert opened the meeting on behalf
of the three cooperating organizations of the IPCS (ILO,
UNEP, WHO). The Task Group reviewed and revised the draft
criteria document and made an evaluation of the risks for
human health and the environment from exposure to
tributyltin compounds.

The first draft of this document was prepared by Dr S.
Dobson (ITE) and Dr R. Cabridenc (Institut National de
Recherche Chimique Appliquée, France). Dr M. Gilbert and
Dr P.G. Jenkins, both members of the IPCS Central Unit,
were responsible for the technical development and
editing, respectively.

ABBREVIATIONS

AA atomic absorption
BCF bioconcentration factor
DBT dibutyltin
EC₅₀ median effective concentration
EEC European Economic Community
EQT environmental quality target
FAA flameless atomic absorption
FMLP formyl methionyl leucyl phenylalanine
FPD flame photometric detector
GC gas chromatography
GLC gas-liquid chromatography
HPLC high-performance liquid chromatography
IC₅₀ median inhibitory concentration
ip intraperitoneal
IU international unit
iv intravenous
LC₅₀ median lethal concentration
LDH lactate dehydrogenase
LT₅₀ median lethal time
MBT monobutyltin
MIC minimal inhibitory concentration
MS mass spectrometry
ND not detectable
NOEL no-observed-effect level
OECD Organization for Economic Cooperation and Development
PALS periarteriolar lymphocyte sheath
sc subcutaneous
T₄ thyroxine
TBT tributyltin
TBTO tributyltin oxide
TLC thin-layer chromatography
TLV threshold limit value

1. SUMMARY

1.1. Physical and chemical properties

Tributyltin (TBT) compounds are organic derivatives of
tetravalent tin. They are characterized by the presence of
covalent bonds between carbon atoms and a tin atom and
have the general formula (n-C₄H₉)₃ Sn-X (where X is
an anion). The purity of commercial tributyltin oxide
(TBTO) is generally above 96%; the principal impurities
are dibutyltin derivatives and, to a lesser extent, tetra-
butyltin and other trialkyltin compounds. TBTO is a
colourless liquid with a characteristic odour and a rela-
tive density of 1.17 to 1.18. The solubility in water is
low, varying between <1.0 and >100 mg/litre according to
the pH, temperature, and anions present in the water
(which determine speciation). In sea water and under nor-
mal conditions, TBT exists as three species (hydroxide,
chloride, and carbonate), which remain in equilibrium. At
pH values less than 7.0, the predominate forms are
Bu₃SnOH₂⁺ and Bu₃SnCl, at pH 8, they are Bu₃SnCl,
Bu₃SnOH, and Bu₃SnCO₃^-, and at pH values above 10,
Bu₃SnOH and Bu₃SnCO₃^- predominate.

The octanol/water partitioncoefficient (log P_ow) lies
between 3.19 and 3.84 for distilled water and is 3.54 for
sea water. TBTO adsorbs strongly to particulate matter,
the reported adsorption coefficients ranging between 110
and 55 000. Vapour pressure is low but published values
show considerable variation. There was no loss of TBTO
from a solution of 1 mg/litre over 62 days, but 20% of the
water was lost by evaporation.

1.2. Analytical methods

Several methods are used for measuring tributyltin
derivatives in water, sediment, or biota. Atomic absorp-
tion spectrometry (AA) is the most common. AA spectrometry
with a flame allows a detection limit of 0.1 mg/litre.
Flameless AA, using atomization in an electric furnace
with graphite, is more sensitive and allows detection
limits of between 0.1 and 1.0 µg/litre water. There are
several different methods of extraction and for forming
volatile derivatives. Separation of these derivatives is
commonly done using "purge and trap" or gas chromato-
graphy. The detection limits are 0.5 and 5.0 µg/kg for
sediment and biota.

1.3. Sources of environmental pollution

Tributyltin compounds have been registered as mollus-
cicides, as antifoulants on boats, ships, quays, buoys,
crab pots, fish nets, and cages, as wood preservatives, as
slimicides on masonry, as disinfectants, and as biocides
for cooling systems, power station cooling towers, pulp

and paper mills, breweries, leather processing, and tex-
tile mills. TBT in antifouling paints was first marketed
in a form that allowed free release of the compound. More
recently, controlled-release paints, in which the TBT is
incorporated in a co-polymer matrix, have become avail-
able. Rubber matrices have also been developed to give
long-term slow release and lasting effectiveness for anti-
fouling paints and molluscicides. TBT is not used in agri-
culture because of high phytotoxicity.

1.4. Regulations on use

Many countries have restricted the use of TBT anti-
fouling paints as a result of effects on shellfish. The
regulations vary in detail from country to country, but
most ban the use of TBT paints on boats of 25 metres
length or less. Some countries have excluded boats with
aluminium hulls from this ban. In addition, some regu-
lations restrict the TBT content of paints or the leaching
rate of TBT from paints (to 4 or 5 µg/cm² per day, long-
term).

1.5. Environmental concentrations

High levels of TBT in water, sediment, and biota have
been found close to pleasure boating activity, especially
in or near marinas, boat yards, and dry docks, fish nets
and cages treated with antifouling paints, and cooling
systems. The degree of tidal flushing and the turbidity
of the water influence TBT concentrations.

TBT levels have been found to reach 1.58 µg/litre in
sea water and estuaries, 7.1 µg/litre in fresh water,
26 300 µg/kg in coastal sediments, 3700 µg/kg in fresh-
water sediments, 6.39 mg/kg in bivalves, 1.92 mg/kg in
gastropods, and 11 mg/kg in fish. However, these maximum
concentrations of TBT should not be taken as representa-
tive, because a number of factors may give rise to anomal-
ously high values (e.g., paint particles in water and
sediment samples). It has been found that measured TBT
concentrations in the surface microlayer of both fresh
water and sea water are up to two orders of magnitude
above those measured just below the surface. However, it
should be noted that recorded levels of TBT in surface
microlayers may be highly affected by the method of
sampling.

Older data may not be comparable with newer data
because of improvements in the analytical methods avail-
able for measuring TBT in water, sediment, and tissue.

1.6. Transport and transformation in the environment

As a result of its low water solubility and lipophilic
character, TBT adsorbs readily onto particles. Between
10% and 95% of TBTO introduced into water is estimated to

undergo particulate adsorption. Progressive disappearance
of adsorbed TBT is not due to desorption but to degra-
dation. The degree of adsorption depends on the salinity,
nature and size of particles in suspension, amount of sus-
pended matter, temperature, and the presence of dissolved
organic matter.

The degradation of TBTO involves the splitting of the
carbon-tin-bond. This can result from various mechanisms
occurring simultaneously in the environment, including
physico-chemical mechanisms (hydrolysis and photodegra-
dation) and biological mechanisms (degradation by microor-
ganisms and metabolism by higher organisms). Whereas the
hydrolysis of organotin compounds occurs under conditions
of extreme pH, it is barely evident under normal environ-
mental conditions. Photodegradation occurs during labora-
tory exposure of solutions to UV light at 300 nm (and to a
lesser extent at 350 nm). Under natural conditions, pho-
tolysis is limited by the wavelength range of sunlight and
by the limited penetration of UV light into water. The
presence of photosensitizing substances can accelerate
photodegradation. Biodegradation depends on environmental
conditions such as temperature, oxygenation, pH, level of
mineral elements, the presence of easily biodegradable
organic substances for co-metabolism, and the nature of
the microflora and its capacity for adaptation. It also
depends on the TBTO concentration being lower than the
lethal or inhibitory threshold for the bacteria. As with
abiotic degradation, biotic breakdown of TBT is a pro-
gressive oxidative debutylization founded on the splitting
of the carbon-tin bond. Dibutyl derivatives are formed,
which are more readily degraded than tributyltin. Mono-
butyltins are mineralized slowly. Anaerobic degradation
does occur but there is a lack of agreement as to its
importance. Some workers consider that anaerobic degra-
dation is slow, others that it is more rapid than aerobic
degradation. Species of bacteria, algae, and wood-
degrading fungi have been identified that can degrade
TBTO. Estimates of the half-life of TBT in the environ-
ment vary widely.

TBT bioaccumulates in organisms because of its
solubility in fat. Bioconcentration factors of up to 7000
have been reported in laboratory investigations with
molluscs and fish, and higher values have been reported in
field studies. Uptake from food is more important than
uptake directly from the water. Higher concentration
factors in microorganisms (between 100 and 30 000) may
reflect adsorption rather than uptake into cells. There
is no indication that TBT is transferred to terrestrial
organisms via food chains.

1.7. Kinetics and metabolism

Tributyltin is absorbed from the gut (20-50% depending
on the vehicle) and via the skin of mammals (approximately
10%). It can be transferred across the blood-brain barrier

and from the placenta to the fetus. Absorbed material is
rapidly and widely distributed among tissues (principally
the liver and kidney).

TBT metabolism in mammals is rapid; metabolites are
detectable in blood within 3 h of TBT administration. In
in vitro studies, it has been shown that TBT is a
substrate for mixed-function oxidases, but these enzymes
are inhibited by very high concentrations of TBT.

The rate of TBT loss differs with different tissues,
and estimates for biological half-lives in mammals range
from 23 to about 30 days.

TBT metabolism also occurs in lower organisms, but it
is slower, particularly in molluscs, than in mammals. The
capacity for bioaccumulation is, therefore, much greater
than in mammals.

TBT compounds inhibit oxidative phosphorylation and
alter mitochondrial structure and function. TBT interferes
with calcification of the shell of oysters ( Crassostrea
species).

1.8. Effects on microorganisms

TBT is toxic to microorganisms and has been used
commercially as a bactericide and algicide. The concen-
trations that produce toxic effects vary considerably
according to the species. TBT is more toxic to gram-
positive bacteria (minimal inhibitory concentration (MIC)
between 0.2 and 0.8 mg/litre) than to gram-negative bac-
teria (MIC: 3 mg/litre). The TBT acetate MIC for fungi is
0.5-1 mg/litre and the TBTO MIC for the green alga
Chlorella pyrenoidosa is 0.5 mg/litre. The primary pro-
ductivity of a natural community of freshwater algae was
reduced by 50% at a TBTO concentration of 3 µg per litre.
Recently established no-observed-effect level (NOEL)
values for two species of algae are 18 and 32 µg per
litre. Toxicity to marine microorganisms is similarly
variable between species and between studies; NOEL values
are difficult to set but lie below 0.1 µg/litre for some
species. Algicidal concentrations range from <1.5 µg per
litre to >1000 µg/litre for different species.

1.9. Effects on aquatic organisms

1.9.1. Effects on marine and estuarine organisms

A summary diagram relating lethal and sublethal ef-
fects to measured marine and estuarine TBT concentrations
is presented in Fig. 1. Concentrations exceeding those
producing acute lethal effects have been found in many
different worldwide locations, particularly associated
with pleasure boating activity.

The development of the motile spores of a green
macroalga was the stage most sensitive to TBT (5-day
EC₅₀: 0.001 µg/litre). There was reduced growth of a
marine angiosperm at TBT concentrations of 1 mg/kg sedi-
ment but no effect at 0.1 mg/kg.

Tributyltin is highly toxic to marine molluscs. It
has been shown experimentally to affect shell deposition
of growing oysters, gonadal development and gender of
adult oysters, settlement, growth, and mortality of larval
oysters and other bivalves, and to cause imposex (the
development of male characteristics) in female gastropods.
The NOEL for spat of the most sensitive oyster species
(Crassostrea gigas) has been reported to be about
20 ng/litre. TBT causes deformation of the shell of adult
oysters in a dose-related manner. No effect on shell mor-
phology was observed experimentally at TBT concentrations
of 2 ng/litre. The NOEL for the development of imposex in
female dogwhelks is below 1.5 ng/litre. Larval forms are
generally more sensitive than adults; in the case of oys-
ters this difference is particularly marked.

Copepods are more sensitive than other crustacean
groups to the acute lethal effects of TBT, LC₅₀ values
for exposure periods up to 96 h ranging from 0.6 to
2.2 µg/litre. These values are comparable to those of
the more sensitive larvae of other crustacean groups. TBT

reduces reproductive performance, neonate survival, and
juvenile growth rate in crustaceans. The NOEL for repro-
duction in the mysid shrimp Acanthomysis sculpta has been
suggested to be 0.09 µg/litre. There was no avoidance of
TBT by the grass shrimp at concentrations up to 30 µg/litre.

The toxicity of tributyltin to marine fish is highly
variable, 96-h LC₅₀ values ranging between 1.5 and
36 µg/litre. Larval stages are more sensitive than adults
(Fig. 1). There are indications that marine fish avoid
TBTO concentrations of 1 µg/litre or more.

1.9.2. Effects on freshwater organisms

A summary diagram relating lethal and sublethal
effects to measured TBT concentrations in fresh water is
presented in Fig. 2. Concentrations exceeding those pro-
ducing sublethal effects have been found, particularly
associated with pleasure boating activity.

Fresh-water angiosperms were killed by a TBTO concen-
tration of 0.5 mg/litre, and growth was inhibited at
0.06 mg/litre or more.

Data on fresh-water invertebrate species are few, re-
lating to just three species other than target organisms.
Different salts of TBT yield 48-h LC₅₀ values for Daphnia
of 2.3-70 µg/litre and for Tubifex of 5.5-33 µg/litre.
The NOEL for Daphnia has been estimated to be 0.5 µg per
litre, based on reversal of normal response to light. The
24-h LC₅₀ for the Asiatic clam has been reported to be
2100 µg/litre, and for target snail adults in schistoso-
miasis control the corresponding values are 30-400 µg/litre.

Tributyltin has been shown to be toxic to schistosome
larvae in the aquatic stages; the LC₅₀ (TBT fluoride)
was calculated to be 16.8 µg/litre for a 1-h exposure.
The TBT dose causing 99% to 100% suppression of cercarial
infectivity of mice was between 2 and 6 µg/litre.

The sensitivity of snails to TBT decreases with age,
but eggs are more resistant than both young and adults.
Egg laying is significantly effected at a TBTO concen-
tration of 0.001 µg/litre.

The acute toxicity of TBT to freshwater fish in LC₅₀
tests up to 168 h ranges from 13 to 240 µg per litre.
The NOEL for the guppy was estimated to be 0.01 µg per
litre, based on histopathological effects.

No effect on survival was found when eggs and larvae
of the frog Rana temporaria were exposed to TBT concen-
trations of 3 µg/litre or less, but at 30 µg/litre sig-
nificant mortality was observed.

1.9.3. Microcosm studies

Microcosm studies modelling marine ecosystems have
been conducted with introduced organisms and in conditions
where inflowing sea water allowed colonization by other
organisms. Results showed decreases in both numbers of
individuals and in species diversity at TBTO concen-
trations in water between 0.06 and 3 µg/litre.

Results from freshwater model ecosystems suggest that
doses which kill freshwater snails also affect other
species, including fish.

1.10. Effects on terrestrial organisms

The exposure of terrestrial organisms to TBT results
primarily from its use as a wood preservative. TBTO is
toxic to bees housed in hives made from TBT-treated wood.
TBT was toxic to bats in a single study, but this result
was not statistically significant owing to high control
mortality. TBT compounds are toxic to insects exposed
topically or via feeding on treated wood. The acute tox-
icity of TBT to wild mice is moderate; estimated dietary
LC₅₀ values, based on consumption of treated seeds used
in repellency tests, range from 37 to 240 mg/kg per day.

1.11. Effects on organisms in the field

Field observations have related high concentrations of
tributyltin to mortality and settlement failure of larval
bivalves, reduced growth, shell thickening and other mal-
formations in developing oysters, imposex in mud snails,
and imposex (concurrent with population decline) in the
dogwhelk. Complete failure of oyster fisheries was ident-
ified initially in France and afterwards in other

countries and related to water concentrations of TBT. The
effects were most marked in areas close to pleasure boat
marinas. Controlling the use of TBT antifouling paints on
small boats has resulted in recovery of oyster repro-
duction and growth. However, water concentrations of TBT
are still high enough in some areas to affect marine
gastropods.

Both shell growth and chambering in Pacific oysters
and imposex in dogwhelks have been used as biological
indicators of TBT contamination.

There have been few studies of the effects on organ-
isms of TBT in sediment, but there are indications that
the TBT is available to burrowing organisms and can cause
mortality in the field.

Gross toxic effects and histopathological changes have
been reported in farmed marine fish exposed to TBT by the
use of antifouling paints on retaining nets.

The use of TBT as a molluscicide against the fresh-
water snails that carry schistosomiasis (bilharzia) has
been proposed. Some field trials have been conducted which
show that it is difficult to apply TBT without damaging
non-target organisms.

1.12. Toxicity to laboratory mammals

1.12.1. Acute toxicity

Tributyltin is moderately to highly toxic to labora-
tory mammals, acute oral LD₅₀ values ranging from 94 to
234 mg/kg body weight for the rat and from 44 to 230 mg/kg
body weight for the mouse. The acute toxicity to the
guinea-pig and the rabbit fall within the same range. The
variation comes from the "anion" component of the tri-
butyltin salt. These compounds exhibit greater lethal
potential when administered parenterally, as opposed to
orally, probably due to only partial absorption from the
gut.

Other effects of acute exposure may include alter-
ations in blood lipid levels, the endocrine system, liver,
and spleen, and transient deficits in brain development.
The toxicological significance of these effects, reported
after high single doses of the compound, is questionable
and the cause of death remains unknown.

The acute toxicity via the dermal route is low, the
LD₅₀ being >9000 mg/kg body weight for the rabbit.
"Nose only" inhalation LD₅₀ (4 h) for the rat is
77 mg/m³ (65 mg/m³ when only inhalable particles are
considered). TBT vapour/air mixtures produce no observ-
able toxic effects, even at saturation. However, TBT is
very hazardous as an inhaled aerosol, producing lung irri-
tation and oedema.

TBT is severely irritating to the skin and an extreme
irritant to the eye. TBTO is not a skin sensitizer.

1.12.2. Short-term toxicity

TBT compounds have been studied most extensively in
the rat (all the data in this section refer to the rat
unless otherwise indicated).

At dietary doses of 320 mg/kg (approximately 25 mg/kg
body weight), high mortality rates were observed when the
exposure time exceeded 4 weeks. No deaths were noted at
100 mg/kg diet (10 mg/kg body weight) or after daily
administration of 12 mg/kg body weight by gavage. In rats
dosed during early post-natal life, 3 mg/kg body weight
resulted in increased deaths. The main symptoms at lethal
doses were loss of appetite, weakness, and emaciation.

Borderline effects on rat growth were observed at
50 mg/kg diet (6 mg/kg body weight) and 6 mg/kg body
weight (gavage studies). Mice are less sensitive, effects
being observed at 150 to 200 mg/kg diet (22 to 29 mg/kg
body weight).

Structural effects on endocrine organs, mainly the
pituitary and thyroid, have been noted in both short- and
long-term studies. Changes in circulating hormone concen-
trations and altered response to physiological stimuli
(pituitary trophic hormones) were observed in short-term
tests, but after long-term exposure most of these changes
appeared to be absent. The mechanism of action is not
known.

Exposure to TBTO aerosol at 2.8 mg/m³ produced high
mortality, respiratory distress, inflammatory reaction
within the respiratory tract and histopathological changes
of lymphatic organs. However, exposure to the highest
attainable vapour concentration (0.16 mg/m³) at room
temperature produced no effects.

Toxic effects on the liver and bile ducts have been
reported in three mammalian species. Hepatocellular
necrosis and inflammatory changes in the bile duct were
observed in rats fed TBTO at a dietary level of 320 mg/kg
(approximately 25 mg/kg body weight) for 4 weeks and in
mice fed 80 mg/kg diet (approximately 12 mg/kg body
weight) for 90 days. Vacuolization of periportal hepato-
cytes was noted in dogs fed a dose of 10 mg/kg body weight
for 8 to 9 weeks. These changes were occasionally
accompanied by increased liver weight and increased serum
activities of liver enzymes.

Decreases in haemoglobin concentration and erythrocyte
volume in rats, resulting from dosing with 80 mg/kg diet
(8 mg/kg body weight), indicate an effect on haemoglobin
synthesis, leading to microcytic hypochromic anaemia. The

decrease in splenic haemosiderin levels suggests alter-
ations in iron status. Anaemia has also been observed in
mice.

The formation of erythrocyte rosettes in mesenteric
lymph nodes has been observed in certain short-term inves-
tigations but not in long-term studies. The biological
significance of this finding (possibly transient) is
unclear.

The characteristic toxic effect of TBTO is on the
immune system; due to effects on the thymus, the cell-
mediated function is impaired. The mechanism of action is
unknown, but may involve the metabolic conversion to
dibutyltin compounds. Non-specific resistance is also
affected.

General effects on the immune system (e.g., on the
weight and morphology of lymphoid tissues, peripheral lym-
phocyte counts, and total serum immunoglobulin concen-
trations) have been reported in several different studies
with TBTO using rats and dogs, but not mice, at overtly
toxic dose levels (effects in mice have been seen with
tributyltin chloride at 150 mg/kg). Only the rat exhibits
general effects on the immune system without other overt
signs of toxicity and is clearly the most sensitive
species. The NOEL in short-term rat studies was 5 mg/kg
diet (0.6 mg/kg body weight). In studies with tributyltin
chloride, analogous effects on the thymus were seen. These
were readily reversible when dosing ceased. TBTO has been
shown to compromise specific immune function in rat in
vivo host resistance studies. Decreased clearance of
Listeria monocytogenes was seen after exposure to a diet-
ary level of 50 mg/kg (the NOEL being 5 mg/kg per day),
and decreased resistance to Trichinella spiralis was seen
at 50 and 5 mg/kg diet, but not at 0.5 mg/kg diet (2.5,
0.25, and 0.025 mg/kg per day body weight, respectively).
Similar effects were seen in aged animals, but these were
less pronounced.

With present knowledge, the effects on host resistance
are probably of most relevance in assessing the potential
hazard to man, but there is insufficient experience in
these test systems to fully assess their significance.
However, some data on the significance of the T. spiralis
model are provided by findings in athymic nude rats after
the standard challenge. In these studies, the complete
absence of thymus-dependent immunity resulted in a 10- to
20-fold increase in muscle larvae counts; by contrast,
exposure to TBTO concentrations of 5 and 50 mg/kg diet
resulted in a 2-fold and a 4-fold increase, respectively.

Although some data are now available from studies on
the effects of tributyltin compounds on the developing
immune system, there is no information on host resistance.

It would be prudent to base assessment of the poten-
tial hazard to humans on data from the most sensitive
species. Effects on host resistance to T. spiralis have
been seen at dietary levels as low as 5 mg/kg (equivalent
to 0.25 mg/kg per day body weight), the NOEL being
0.5 mg/kg (equivalent to 0.025 mg/kg per day). However,
the interpretation of the significance of these data for
human risk assessment is controversial. In all other
studies a concentration of 5 mg/kg per day in the diet
(equivalent to 0.5 mg/kg body weight, based on the short-
term studies) was the NOEL with respect to general, as
well as specific, effects on the immune system.

1.12.3. Long-term toxicity

A long-term study in rats indicates a marginal effect
of TBT on general toxicological parameters (of limited
toxicological significance) at a level of 5 mg/kg diet
(0.25 mg/kg body weight).

1.12.4. Genotoxicity

The genotoxicity of TBTO has been the subject of
extensive investigations. Negative results were obtained
in the vast majority of studies, and there is no convinc-
ing evidence that TBTO has any mutagenic potential.

1.12.5. Reproductive toxicity

The potential embryotoxicity of TBTO has been evalu-
ated in three mammalian species (mouse, rat, and rabbit)
after oral dosing of the mother. The main malformation
noted in rat and mouse fetuses was cleft palate, but this
occurred at dosages overtly toxic to the mothers. These
results are not considered to be indicative of teratogenic
effects of TBTO at doses below those producing maternal
toxicity. The lowest NOEL, with regard to embryotoxicity
and fetotoxicity for all three species, was 1.0 mg/kg body
weight.

1.12.6. Carcinogenicity

One carcinogenicity study has been carried out on
rats, in which neoplastic changes were observed in endo-
crine organs at 50 mg/kg diet. The pituitary tumours
reported at 0.5 mg/kg diet were considered as having no
biological significance since there was no dose-response
relationship. These tumour types usually appear in high
and variable background incidences, and their significance
is, therefore, questionable. A carcinogenicity study on
mice is in progress.

1.13. Effects on humans

Occupational exposure of workers to tributyltin has
been found to result in irritation of the upper respir-
atory tract. TBT as an aerosol poses a hazard to humans.
TBTO is a skin and eye irritant and severe dermatitis has
been reported after direct contact with the skin. The
potential problem is made worse by the lack of an
immediate response to the skin.

2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, ANALYTICAL METHODS

2.1. Identity of tributyltin compounds

Tributyltins compounds are organic derivatives of tin
(Sn^IV) characterized by the presence of covalent bonds
between three carbon atoms and a tin atom. They conform to
the following general formula ( n-C₄H₉)₃ Sn-X, where X is
an anion or a group linked covalently through a hetero-
atom.

The nature of X influences the physico-chemical
properties, notably the relative solubility in water and
non-polar solvents and the vapour pressure.

These compounds differ from inorganic tin both in
behaviour and effects. An important member of the group is
tributyltin oxide (TBTO; RTECS number, JN8750000). Commer-
cial TBTO has a purity generally above 96%. Principle
impurities are dibutyltin derivatives and, to a lesser
extent, tetrabutyl or dibutylalkyl tin compounds.

Other industrially important tributyltin derivatives
include tributyltin fluoride, tributyltin methacrylate (monomer
or copolymer), tributyltin benzoate, tributyltin linoleate,
tributyltin naphthenate, and tributyltin phosphate.

2.2. Physical and chemical properties

TBTO is flammable but does not form explosive mixtures
with air. It is a mild oxidizing agent. It reacts quanti-
tatively at room temperature with bromide or iodine with
cleavage of the Sn-O bond (a reaction that may be used for
quantitative analysis) (Bahr & Pawlenko, 1978).

In the presence of oxygen, light or heat, slow break-
down occurs with the formation of tetra-n-butyltin, di-
n-butyltin oxide, and eventually tin (IV) oxide by de-
alkylation (Evans & Karpel, 1985). This degradation may be
inhibited by the addition of 0.1-1.0% of stabilizers (such
as lactic or citric acids).

It has been suggested (Maguire et al., 1984; Laughlin
et al., 1986a) that TBTO in aqueous solution dissociates
with the formation of a hydrated tributyltin cation, which
can undergo reaction with anions present. Data are not
available on the equilibrium constants for these reac-
tions.

Laughlin et al. (1986a) showed that TBTO can react
with normal constituents of the sea water in the following
ways:

Bu₃-Sn-O-Sn-Bu₃+ HO -» 2Bu₃-Sn-OH
Bu₃-Sn-OH-H⁺ -» Bu₃SnOH₂⁺
Bu₃-Sn-OH + CO₃^2- -» Bu₃SnCO₃^- + OH^-
Bu₃-Sn-OH₂⁺ + Cl- -» Bu₃-Sn-Cl + H₂O

The predominant forms are Bu₃SnOH₂⁺ and Bu₃SnCl
at pH < 7, Bu₃SnCl, Bu₃SnOH, and Bu₃SnCO₃^- at pH 8,
and Bu₃SnOH and Bu₃SnCO₃^- at pH > 10.

Under normal conditions in sea water, it is considered
that the three species (hydroxide, chloride, and carbon-
ate) remain in equilibrium.

The physical and chemical properties of some commer-
cially available tributyltin compounds are listed in Table
1.

Varying data on the solubility of TBTO in water, which
ranges from < 1.0 to > 100 mg/litre at different tempera-
tures and pH values, may be related to the presence of
different anionic species as described above.

In the same way as described in the reaction between
TBTO and water, the TBT group can be transferred to other
oxygen-, nitrogen-, and sulfur-containing groups. Thus,
anaerobically in sediments, TBTO can be transformed to TBT
sulfide. With amino acids, or their derivatives such as
proteins, reaction can occur on the nitrogen and sulfur
atoms, and, with wood, it has been suggested that the TBT
group may react with hydroxylic groups (Blunden et al.,
1984) or form tributyltin carbonate (Smith et al., 1977).
Thus adsorption on to particulate matter could involve
chemical reaction as well as physical adsorption or sol-
ution. TBTO adsorbs strongly to particulate matter, the
reported adsorption coefficients ranging between 110 and
55 000.

Table 1. Identity and physical and chemical properties of tributyltin compounds
---------------------------------------------------------------------------------------------------------
Oxide Benzoate Chloride Fluoride Linoleate Methacrylate Naphthenate
(TBTO) (TBTB) (TBTCl) (TBTF) (TBTL) (TBTM) (TBTN)
---------------------------------------------------------------------------------------------------------
IUPAC name distannoxane, stannane, stannane, stannane, stannane, stannane, stannane,
hexabutyl (benzyloxy) tributyl- tributyl- tributyl- tributyl- tributyl-
tributyl chloro fluoro (1-oxo-9,12- (2-methyl-1- mono (naph-
octadecadi- oxo-2-propyl) thenoyloxy)
enyl)oxy- oxy- derivatives

CAS name Bis(tributyl- Tributyltin Tributyltin Tributyltin Tributyltin Tributyltin Tributyltin
tin) oxide benzoate chloride fluoride linoleate methacrylate naphthenate

CAS number 56-35-9 4342-36-3 1461-22-9 1983-10-4 24124-25-2 2155-70-6 85409-17-2

Molecular C₂₄H₅₄OSn₂ C₁₉H₃₂O₂Sn C₁₂H₂₇ClSn C₁₂H₂₇FSn C₃₀H₅₈O₂Sn C₁₆H₃₂O₂Sn
formula

Relative 596 411 325 309 568.7 374.7 ca.500
molecular
mass

Boiling 173 ca.135 140 > 350 ca.140 > 300 ca.125
point (°C) (130 Pa) (30 Pa) (1300 Pa) (extrapol) (50 Pa) (extrapol) (50 Pa)

Melting < -45 20 -16 240 < 0 16 < 0
point (°C)

Relative
density 1.17-1.18 ca.1.2 ca.1.2 1.25 1.05 1.14 ca.1.1
(20 °C)

Vapour
pressure (Pa 1 x 10^-3 2 x 10^-4 9 x 10^-2 3 x 10^-2 9 x 10^-5
at 20 °C)

Refractive 1.4880-
index (20 °C) 1.4895
---------------------------------------------------------------------------------------------------------
TBTO is soluble in lipids and very soluble in a number
of organic solvents (ethanol, ether, halogenated hydro-
carbons, etc.).

The octanol/water partition coefficient (log P_ow)
lies between 3.19 and 3.84 for distilled water and is 3.54
for sea water.

As shown in Table 1, the vapour pressures of TBT
compounds are low. The work of Maguire et al. (1983)
confirmed this directly by showing no loss of TBTO from a
1 mg/litre solution after 62 days; 20% of the water was
lost by evaporation.

2.3. Analytical methods

The control levels of contamination of different
environmental compartments (water, sediment, biota) and
the interpretation of laboratory experimental and field
study results regarding levels, fate, biodegradation, and
bioaccumulation of tributyltin compounds require sensitive
analytical techniques to allow identification and quanti-
fication.

2.3.1. Measurement of organotin compounds

These methods, which are summarized in Table 2, have
been applied initially to water and later to sediment and
biota. They must be sufficiently sensitive and specific to
allow monitoring of ng/litre levels, and they need to be
able to distinguish between different forms of organic tin
derivatives present in the environment, i.e. mono-, di-,
tri-, or tetra-butyltins and different species of alkyl
moieties (butyl, methyl). They have also to avoid all
interference from other metals and other organometallic
derivatives.

Generally there are four successive stages to analy-
sis, although some are optional:

* extraction;
* formation of volatile derivatives;
* separation of these derivatives;
* detection, identification, and quantification.

Table 2. Sampling, preparation, and analysis of tributyltin compounds
---------------------------------------------------------------------------------------------------------
Medium Sampling method Sample volume Analytical method Detection limit Reference
---------------------------------------------------------------------------------------------------------
Air adsorption on 50-100 litres derivatization with Zimmerli &
Chromosorb, RMgX; GC/MS or Zimmermann
cation exchange GC/FPD (1980);
resin, or Tenax Muller (1987a)

Water 250 ml NaBH₄ conversion to 0.1-2 ng/litre Hodge et al.
hydride; separation by (1979);
fractional distillation; Michel (1987);
AA Donard et al.
(1986); Braman &
Tompkins (1979);
Valkirs et al.
(1986); Weber
et al. (1986)

Water and extraction with 8 litres (water) derivatization with 1 ng/litre Maguire &
sediments dichloromethane or 1 g (sediment C₅H₁₁ MgBr; GC-FPD (water) Huneault (1981);
dry weight) or GC-FAA or 5 ng/mg Maguire &
(sediment Tkacz (1983,
dry weight) 1985); Maguire
et al. (1986)

Water and acidification, 1 litre derivatization with 10 ng/litre Meinema et al.
biota extraction with CH₃ Mgl; GC-MS or AA (1978); Bjorklund
dichloromethane (1987a)

Water, 200 ml or NaBH₄ conversion to 5 ng/litre or Matthias et
biota, or 16 litres hydride; extraction with 0.2 ng/litre al. (1986a,b);
sediments dichloromethane Humphrey &
Hope (1987)

Water and adsorption on 60 litres extraction with dichloro- 0.07 ng/litre Humphrey &
sediment silica (water) or 10 g methane/tropolone; deriva- (water) Hope (1987)
bonded C18 (sediment) tization with C₅H₇ MgBr; 0.2 mg/kg
GC-MS (sediment)

macroreticular 1 litre extraction with n-pentane < 1 ng/litre Muller
resin (water) diethylether (water) (1984)
adsorption (sediment); derivatization 0.5 mg/kg
with CH₃MgCl; GC-MS (sediment)
---------------------------------------------------------------------------------------------------------
2.3.1.1 Extraction of tributyltin derivatives

Extraction may be independent of or coincident with
the formation of volatile derivatives. It is necessary for
sediments and biological tissues and can also be applied
in the analysis of water samples.

Following acidification, various organic solvents have
been used. The following are most often cited: methyliso-
butylketone, hexane, ethyl acetate, toluene, methanol,
chloroform, dichloromethane, and mixtures of tropolone
(2-hydroxy-2,4,6-cycloheptatrienone) with chloroform, ben-
zene, or dichloromethane.

In the case of water, liquid-liquid extraction may be
replaced by adsorption onto silica gel bonded with C18
aliphatic chains (Matthias et al., 1986a,b; Humphrey &
Hope, 1987).

2.3.1.2 Formation of volatile derivatives

Mono-, di-, and tri-butyltins are not sufficiently
volatile to assure their separation on gas-phase chromato-
graphy; it is, therefore, necessary to prepare more vol-
atile derivatives to allow better separation. Two pro-
cedures have been advocated:

* formation of alkyl derivatives (methyl or pentyl) by
the use of Grignard's reagent (reactive organomag-
nesium);
* formation of hydrides with the general structure
R_nSnH_4-n by reaction with sodium borohydride
(NaBH₄) (Hodge et al., 1979).

These volatile derivatives can then be extracted using
organic solvents, such as dichloromethane, or purged by a
stream of hydrogen.

2.3.1.3 Separation of organotin derivatives

Less sensitive methods for direct separation of mono-,
di-, and tri-butyltins include high performance liquid
chromatography (Jewett & Brinckman, 1981) and thin-layer
chromatography. The latter method is only qualitative and
little used because of its low sensitivity.

2.3.1.4 Detection and measurement of different forms of organotin

Volatile derivatives prepared in the laboratory may be
separated by two procedures:

* separation as a function of boiling point with collec-
tion in a cold trap ("purge and trap" procedure);
* separation by gas chromatography.

After separation by GLC or by the "purge and trap"
procedure, it is possible to detect and quantify, at the
ng/litre level, different forms of organotin using the
following methods:

* a flame photometric detector selective for tin (FPD)
is considered satisfactory;
* a flame atomic absorption (AA) spectrometer or flame-
less atomic absorption (FAA) spectrometer using a
graphite furnace (tin is detected at 286.3 nm or
244.6 nm);
* a mass spectrometer (MS); this is useful for precise
identification of the substance but has limited sen-
sitivity.

There are several methods available for measuring TBT
down to detection limits of 0.2 to 5 ng/litre in water and
5 to 30 µg/kg (in tissues of biota and in sediments).
Some of them can be adapted for routine monitoring pur-
poses. It is necessary, however, to have sophisticated
equipment and the difficulty of the methods requires
experienced laboratories.

His & Robert (1980, 1985) developed a biological assay
based on toxic effects on larvae of the Pacific oyster,
Crassostrea gigas, sensitive only above 20 ng/litre and
nonspecific between organotin and other toxic compounds.
Colorimetric methods (Sherman & Carlson, 1980) have been
based on forming coloured derivatives with phenylfluorone
(nonspecific and with a sensitivity around 0.1 to 4 µg tin).

2.3.2. Interlaboratory calibrations

Interlaboratory comparison of assay methods have been
performed to compare the various proposed methods and to
validate their usefulness as standards.

Young et al. (1986) reported the conclusions of a
workshop held in the USA to examine the problems posed by
the analysis of organotins in water. Nine methods, based
on the principles outlined above, were considered as sat-
isfactory, since the range of results fell within + 15% of
the mean when the TBT concentration was in the order of
ng/litre.

Stephenson et al. (1987) reported the results of
interlaboratory calibrations conducted in 1986-1987 and
carried out on TBT derivatives in mussel tissues and in
sediments. The measurements were made in seven labora-
tories, each using its own technique and using different
extraction conditions, derivative formation, and detec-
tion. A first examination of results showed that they did
not vary by more than a factor of 3. The results were con-
sidered satisfactory.

Blair et al. (1986) took part in an interlaboratory
calibration exercise organised by the National Bureau of
Standards (NBS) in 1984 in the USA and carried out deter-
minations of TBT in water (at a concentration of 1 µg/litre).

Under the auspices of the OECD, it was decided
recently to organize a new worldwide intercalibration to
be carried out on:

* water samples containing 10 ng/litre each of mono-,
di-, and tri-butyltin;
* samples of dried sediment containing the above com-
pounds at a concentration of 100 µg/kg;
* samples of mussel tissue, frozen or freeze-dried, con-
taining the above compounds at 100 µg/kg.

It seems premature to impose a single analytical
method and preferable to allow a certain freedom of choice
between methods to allow sufficient sensitivity to be
attained. However, control of the competence of labora-
tories that carry out such difficult and complex analysis
is required through new calibration procedures.

3. SOURCES OF ENVIRONMENTAL EXPOSURE

3.1. Uses

Dutch scientists first recognized the biocidal proper-
ties of triorganotin compounds in the 1950s; major pro-
duction and use of these substances dates from this
period. It was found that the different triorganotin com-
pounds have different toxicities to different organisms.
Tributyltin compounds were found to be the most toxic of
the triorganotins to gram-positive bacteria and to fungi.
They were also found to have biocidal properties to a wide
spectrum of aquatic organisms.

In the early 1960s, both tributyltin oxide (TBTO) and
TBT fluoride were tested, mainly in Africa, as mollusci-
cides against several freshwater snail species that are
vectors of the disease schistosomiasis, the snails being
the intermediate hosts of the trematode parasite. This use
led to the introduction of TBT, during the mid 1960s, as
an antifouling paint on boats. At the same time TBT com-
pounds were being registered as wood preservatives (the
first registration was in 1958).

Tributyltin compounds have been registered as mollus-
cicides, as antifoulants on boats, ships, quays, buoys,
crabpots, fish nets, and cages, as wood preservatives, as
slimicides on masonry, as disinfectants, and as biocides
for cooling systems, power station cooling towers, pulp
and paper mills, breweries, leather processing, and tex-
tile mills.

When introduced as antifouling paints, TBT paints were
of the "free association" type, where the TBT is physi-
cally incorporated into the paint matrix. In this form it
has a high early release and very short life. Co-polymer
paints were introduced later; in these the TBT moiety is
chemically bonded to a polymer backbone, e.g., those
formed from TBT acrylate or methacrylate and the corre-
sponding acid. The biocide is released by chemical
hydrolysis of the organotin ester linkage. Dissolution is
slow from ships' hulls and a low level of released TBT can
be achieved over a prolonged period. TBT compounds have
also been impregnated into neoprene rubber to produce
elastomeric antifoulant coatings and slow-release mollus-
cicides. In this form, much of the TBT remains in the
matrix of the rubber, though the effectiveness lasts for
several years.

TBT compounds have not been suggested for use in agri-
culture because of their high phytotoxicity.

3.2. Production

The world consumption of tin in 1976 was estimated to
be 200 x 10³ tonnes, of which 28 x 10³ tonnes was
organotin. Approximately 40% of the total was consumed

in the USA (Zuckerman et al., 1978). The United Kingdom
Department of the Environment (1986) reported that the
worldwide use of organotin in 1980 was 30 x 10³ tonnes.
This total was made up as follows:

* PVC stabilizers (dibutyl), approximately 20 x 10³ tonnes;
* wood preservatives (tributyl), 3-4 x 10³ tonnes;
* antifouling paints (tributyl), 2-3 x 10³ tonnes;
* other uses of both di- and tri-butyltin, < 2 x 10³ tonnes.

The annual world production of TBT compounds is esti-
mated to be 4000 to 5000 tonnes (Organotin Environmental
Programme Association (ORTEPA); personal communication to
IPCS, 1989).

The total annual use (production and imports) of
organotin compounds in Canada was reported by Thompson et
al. (1985) to be in excess of 1 x 10³ tonnes. The total
annual production of TBTO in the Federal Republic of
Germany is reported to be 2 x 10³ tonnes, of which 70%
is exported. National usage is as follows: 70% antifouling
paints; 20% timber protection; 10% textile and leather
protection; small amounts are also used as a preservative
in dispersion paints and as a disinfecting agent. Annual
tin emissions are reported to be less than 300 kg (TWG,
1988a). Annual TBT use in the Netherlands in 1985 was
reported to be 1.5 x 10⁴ kg for wood preservation and
10 x 10⁴ kg for antifouling paints (TWG, 1988c). Organo-
tin antifoulant use in Norway was 13.7 x 10⁴ kg in 1986
for the treatment of nets and sea pens at approximately
600 fish farms (Linden, 1987). In Japan, usage was
estimated at 1300 tonnes in 1987, of which two-thirds was
used for antifouling paints on vessels and one-third for
antifouling of nets in fish culture.

A survey of total and retail sales of TBT-containing
paints and antifouling preparations for nets was carried
out in Finland in 1987. Of a total of 42 000 litres,
37 000 litres were sold retail. The concentration of TBT
in the antifouling paints was 4-18%. The previous use of
TBT as a slimicide or fungicide (estimated at 2.1 tonnes
per year during the period 1968-1970) has been discon-
tinued. The estimated sale of wood preservatives contain-
ing TBT was 130 tonnes in 1987; these contained between
0.9 and 1.8% of TBT. Champ & Pugh (1987) reported that
about 300 TBT antifouling paints were registered in the
USA in 1987, but only about 17 paints are now registered
for use (US EPA; personal communication to IPCS, 1989).
MAFF/HSE (1988) listed 345 different wood preservative
formulations, 24 surface biocides and 215 antifouling
paints containing TBT with registration approval for use
in the United Kingdom under the Control of Pesticides
Regulations. In 1989, the number of antifouling paints
containing TBT registered for use in the United Kingdom
had fallen to 148, with the number of wood preservatives
and surface biocides remaining about the same (337 and 26
registered products, respectively) (MAFF/HSE, 1989).

3.3. Regulations

In 1974, the USA set an occupational limit for organo-
tin compounds in air of 0.1 mg tin/m³ (time-weighted
average). In 1979, the American Conference of Governmental
Industrial Hygienists (ACGIH) recommended that the occu-
pational exposure standard for organotin compounds in air
should be set at a threshold limit value (time-weighted
average) of 0.1 mg tin/m³ and a short-term TLV at 0.2 mg
tin/m³. The Federal Republic of Germany was recommended,
in 1979, to adopt an occupational exposure standard for
organotin compounds in air of 0.1 mg tin/m³, specified
as a maximum worksite concentration (MAK). The United
Kingdom has also set a recommended occupational exposure
limit of 0.1 mg tin/m³.

A tentative acceptable daily intake (ADI) of 1.6 µg/kg
per day has been adopted in Japan.

In December 1979, the Japanese Government banned the
use of tributyltin compounds in certain products for
household use, e.g., paint, adhesive, wax, shoe polish,
and textile products.

Following the effects on the oyster industry in France
in the late 1970s, and the subsequent correlation of the
effects with TBT usage, the French government banned the
use of TBT antifouling paints for an initial trial period
of three months, which was later extended. In 1982, paints
containing more than 3% TBT by weight were banned on boats
of < 25 m in length, although boats with aluminium hulls
were excluded. Initially the regulation only covered the
Atlantic coast (January 1982) but was later extended
(September 1982) to the whole French coastline. All use of
organotin compounds in antifouling paints, at any concen-
tration, is now banned in France.

The exception in the regulations for TBT-based anti-
fouling paints that many countries have made for boats
with aluminium hulls is based on the fact that the copper-
based alternative paints react chemically with the alu-
minium.

In January 1986, the United Kingdom enforced regu-
lations that prohibited the retail sale and supply of
antifouling paints with a total tin concentration greater
than 7.5% by weight in co-polymer paints (reduced to 5.5%
in January 1987) or 2.5% in other paints. These regu-
lations were meant to control the use on small pleasure
craft, ban the sale of "free association" paints con-
taining high levels of organotin and set an upper limit on
organotin compounds in co-polymer paints. An ambient water
quality target of 20 ng/litre was set. The United Kingdom
Department of the Environment took steps to determine the
effectiveness of the legislation by setting up a monitor-
ing programme. Based on the results of this monitoring, a

total ban on the use of TBT paints on small boats (< 25 m)
and fish farming equipment was implemented in July 1987
(Abel et al., 1987). An environmental quality standard
(EQS) of 20 ng/litre for fresh water (covering both
potable water and protection of sensitive aquatic biota)
and 2 ng/litre for sea water has been set (United Kingdom
Department of the Environment, 1989).

The paint industry of the Federal Republic of Germany
(FRG) issued a renunciation in 1986 on the use of mono-
meric organotin compounds in antifouling paints and a
restriction to 3.8% TBT in co-polymeric paints. The FRG
has not, as yet, issued any national ban on TBT marine
antifouling paints and is awaiting the outcome of dis-
cussions on an EEC directive (TWG, 1988b). Champ & Pugh
(1987) reported that both Switzerland and the FRG have
banned all uses of TBT in antifouling paints in the fresh-
water environment.

In 1987, the US EPA reviewed TBT usage, weighing risks
to the environment against benefits to users. In the mean-
time, some individual States have passed their own regu-
lations. Both Virginia and Washington State have banned
the use of TBT antifouling paints on boats of < 25 m in
length, excepting those with aluminium hulls. Only paints
that conform to a leaching rate of 5 µg/cm² per day
(steady state) can be used on boats longer than 25 m. Both
states continued to permit the use of TBT paints, with
acceptable leach rates, in 16 oz (0.45 kg) aerosol cans
for use on outboard motors and lower units. Maryland
instituted similar restrictions but set a lower permiss-
ible leaching rate of 1 µg/cm² per day (steady state).
Since 1985, North Carolina, Oregon, and Michigan have
instituted restrictions on TBT use. California, Alaska,
New York, and New Jersey had TBT Bills pending in their
respective legislatures (Champ & Pugh, 1987). In April
1988, both the US House of Representatives and the Senate
passed bills to restrict the use of TBT in antifouling
paints. The legislation was signed by the President on
16th June 1988 and came into effect on 16th December 1988.
This Act established an interim release rate restriction
of 4.0 µg/cm² per day (steady state) and a provision
prohibiting application of TBT antifouling paints to non-
aluminium vessels under 25 m length. Application to larger
vessels was restricted to certified applicators only. The
outboard motor or lower drive unit of a vessel less than
25 m in length was exempted. A limit on sales, delivery,
purchase, and receipt of TBT paints was set in December
1988 and a limit on use in June 1989 for existing stocks
of paint.

A voluntary ban on the use of TBT compounds for nets
in fish culture was imposed in 1987 by the National
Federation of Fisheries Cooperative Association of Japan.
In 1988, the Japanese Ministry of Health and Welfare and
the Japanese Ministry of International Trade and Industry

"designated" eight TBT compounds (and a further five TBT
compounds in 1989) on the basis of persistence, accumu-
lation, and toxicity. "Designated" indicates that no
final decision on regulation has yet been taken but that
the compounds have a recognized hazard. Following this
action, the Japan Paint Manufacturers Association volun-
tarily reduced the upper limit for TBT in paints to < 10%
wet weight for monomers and < 15% wet weight for polymers.
There is current action to monitor release rates from
paint products as the next step in limiting human exposure.

Maguire (1987) reported that tributyltin for the pres-
ervation of fish-farm nets is banned in Canada. In 1987,
the Canadian Department of Agriculture served notice that
antifouling uses of TBT compounds must conform to the
following: a maximum short-term (first 14 days) cumulative
release-rate from paint formulations of 168 µg/cm²; a
long-term average daily release of 4 µg/cm²; and a
minimum hull length of 19.5 m for the use of TBT antifoul-
ing paints on non-aluminium vessels.

In Australia, control measures on the use of TBT-based
paints were introduced in the States of New South Wales
and Victoria. TBT is prohibited for use on boats with a
hull length of less than 25 m, while a leaching rate of
5.0 µg/cm² per day was set for hulls of 25 m or more.
Aluminium vessels are not exempt from the ban.

The Republic of Ireland instituted a by-law banning
the use of organotin compounds on boats and other aquatic
structures in April 1987 (Minchin et al., 1987).

Norway has also prohibited use of TBT in antifouling
paints except for boats longer than 25 m and those with
aluminium hulls; the regulation became effective from
January 1989. There is also prohibition on the sale, manu-
facture, and import of paints containing TBT without a
specific permit from the State Pollution Control auth-
ority. An agreement to prohibit use on nets of fish farms
has been concluded. Under the Helsinki Convention, the
Baltic States have formed an agreement on the banning of
TBT paints on small boats and have set up a joint monitor-
ing programme.

The Commission of the European Communities has made a
proposal to the Council of Ministers concerning restric-
tions on the use of antifouling paints that mirrors
national restrictions in member states (except that there
would be no derogation for boats with aluminium hulls).
This proposal is currently being considered by the
European Parliament and Council.

4. ENVIRONMENTAL TRANSPORT AND TRANSFORMATION

Summary

Due to its physico-chemical properties, TBT introduced into
natural waters will partly adsorb onto particles. The quanti-
tative data show large variation due to differences in exper-
imental conditions such as salinity and concentration and
organic content of particulate matter. Once it is adsorbed,
decrease in TBT concentration takes place mainly by degra-
dation. It is known that TBT degradation rates in sediment are
slower than in the water column, particularly in anaerobic
conditions.

Although abiotic degradation occurs, the process remains
less important than biological action.

Biodegradation of TBTO in soil and water depends on the
environmental conditions and the toxic effect of the available
concentrations to the organisms involved. Hydroxylated inter-
mediates are formed during stepwise debutylation. Aerobic and
anaerobic organisms both cause biodegradation, but the relative
efficiency is not known conclusively. Illumination of the cul-
tures lowers the half-life, indicating the involvement of
photosynthetic organisms.

The lipophilic properties of TBTO contribute to bioaccumu-
lation in aquatic organisms, especially molluscs. Laboratory
and field studies corroborate this, although it is unclear how
adsorption processes complicate the results. Bioaccumulation in
all organisms studied is due, at least in part, to bioconcen-
tration from the water phase. Elimination takes place when
organisms are no longer exposed to tributyltin compounds.

Whether it is directly discharged into the environment or
diffuses progressively (at 1 to 10 µg/cm² per day) from
coatings of the hulls of boats or nets, TBTO enters the aquatic
environment and is subject to transformation resulting from
physico-chemical and biochemical processes. Speciation is out-
lined in chapter 2.

4.1. Adsorption onto and desorption from particles

The effects of TBTO vary in relation to the state in
which the substance is present in the aquatic environment,
in particular whether it is available to organisms in
estuaries or sea shores. It is important to have infor-
mation on its distribution in natural waters likely to
have large amounts of suspended matter of various types.
Several workers (Valkirs et al., 1986; Maguire et al.,
1986; Randall et al., 1986; Harris & Cleary, 1987; Stang &
Seligman, 1987; Hinga et al., 1987) have conducted studies
on adsorption and desorption of TBTO in laboratory exper-
iments, observations in the field, studies conducted in
microcosms, and mathematical modelling.

Mathematical models have been developed to estimate
the distribution of TBT in enclosed or semi-enclosed har-
bours (Walton et al., 1986) and estuaries (Harris &
Cleary, 1987). Good agreement has been found between
measured and estimated concentrations of tin in San Diego
harbour, USA (Walton et al., 1986). The authors considered
the results useful in predicting levels in ecologically
sensitive areas of the bay. The Harris & Cleary (1987)
model was based on the estuary of the River Tamar in
south-west England. This model, still under development,
aimed to reduce inputs in order to allow the model to be
used by non-experts and to be applicable to all estuaries.
Output for the River Tamar suggested that sediment-bound
tin would be distributed up the estuary by tidal influence
leading to increased bound tin further from the open sea.
This effect would be most marked in the summer. Relative
to soluble TBT, this bound fraction does not currently
amount to a significant source of tin for organisms. The
authors point out, however, that this source may become
increasingly important as use of TBT declines and sedi-
ment-bound TBT represents the only available source of the
compound.

The chemical properties of TBT, particularly its lipo-
philic character and poor water solubility, are such that,
when TBTO is introduced into water, repartition will
occur, TBTO leaving the aqueous phase and preferentially
adsorbing onto particles (Hinga et al., 1987). Adsorption
and desorption are dependant on the nature of the sedi-
ment. Little data is available to indicate whether
adsorbed TBT is bioavailable.

If this phenomenon is generally evident, its intensity
varies considerably as a function of the method of study
used and the measurements made. Contradictory results are
apparent in the literature.

Reports from different authors using various con-
ditions have estimated that between 10% and 95% of TBTO
introduced into water is adsorbed onto particles. There
is, however, general agreement that the compound remains
strongly adsorbed. It has been stated that sediments
remain contaminated for at least 10 months; progressive
disappearance of TBTO is not due to desorption but to
degradation.

In an in situ study of Pearl Harbour sediment, the
rate of adsorption of tributyltin derivatives was found to
be 0.57 ng TBT/cm² per day (Stang & Seligman, 1987).
There was, apparently, no desorption of TBTO itself but
dibutyltin derivatives formed by degradation desorbed with
rates varying between 0.16 and 0.55 ng DBT/cm² per day.

Variability in results, more evident in field studies
than laboratory studies, is explained by the fact that
adsorption depends on many different factors, amongst
which are the following:

* salinity;
* nature and size of particles in suspension;
* amount of suspended particles;
* temperature;
* presence of dissolved organic matter.

Uncertainties are also evident in relation to the bio-
availability of TBT adsorbed onto sediment. Salazar et al.
(1987) considered that the effects of adsorbed TBTO were
partially masked, i.e. that the compound was unavailable
to organisms. This conclusion could not be verified
regarding effects on filtering or burrowing organisms
living in the sediment.

It is generally agreed that part of the TBTO accumu-
lates in the surface monolayer of natural waters. This
TBTO will also be adsorbed onto organic matter and lipid
material present on the surface.

4.2. Abiotic degradation

A number of studies have shown that a degradation
pathway for tributyltin compounds exists in the environ-
ment, which involves progressive debutylation. It is
theoretically completed with the liberation into water of
the tin oxide (SnO₂).

R₃SnX -> R₂SnX₂ -> RSnX₃ -> SnX₄

A number of studies have looked for evidence of such
degradation, the cause and mechanisms, and an understand-
ing of the kinetics in different environmental conditions
(Chapman & Price, 1972; Brinckman, 1981; Blunden et al.,
1984; Maguire & Tkacz, 1985; and Seligman et al., 1986a).

Degradation of TBTO proceeds via splitting of the
carbon-tin bond, which can result from various mechanisms
occurring simultaneously in the environment. These include
physico-chemical mechanisms (hydrolysis and photodegra-
dation) and biological mechanisms (degradation by microor-
ganisms and metabolism by higher organisms). While degra-
dation definitely occurs as a result of these different
mechanisms in laboratory studies, it is necessary to
assess the relative importance of these different pathways
to degradation of TBTO in the field.

4.2.1. Hydrolytic cleavage of the tin-carbon bond

Since hydrolysis of the tin-carbon bond of organotin
derivatives occurs only under conditions of extreme pH, it
is barely evident under normal environmental conditions.

Studies were carried out in darkness and a sterile
medium to assess the importance of hydrolysis in the
degradation of TBTO. According to the work of Maguire et
al. (1983) and of Maguire & Tkacz (1985), TBTO remains
stable for 11 months in distilled or natural water at
20 °C, in the dark, and in a sterile medium. Under various
conditions of pH, between 2.9 and 10.3, these authors
found no change in TBTO over 63 days. According to
Seligman et al. (1986a), slight degradation of TBTO was
apparent after 94 days in darkness in the presence of
formalin as a sterilizing agent.

It is, therefore, considered that degradation occurs
either not at all or only very slowly in normal environ-
mental conditions of pH and temperature, when monitored in
the dark and in a sterile medium.

4.2.2. Photodegradation

Photodegradation of TBTO by ultraviolet light is
theoretically possible. UV light with a wavelength longer
than 290 nm possesses an energy of 300 kJ/mol, whereas the
energy required to break the carbon-tin bond is 190-220
kJ/mol. At the same time, TBTO absorbs in the UV region at
300 nm and, less strongly, at 350 nm.

Field and laboratory measurements have shown that this
route of degradation can occur and that it forms deriva-
tives of dibutyltin. These seem to be resistant to pho-
tolysis, since very little monobutyltin is formed (Blunden
& Chapman, 1986). While the phenomenon clearly exists, its
importance varies considerably with different environmen-
tal conditions. Conditions of illumination, conditions of
transmission of light, and the presence of photosensi-
tizing substances (acetone, humic acids, etc.) can con-
siderably accelerate the process.

Results of laboratory studies vary considerably de-
pending on whether experiments are conducted under natural
sunlight or UV light of known wavelength. According to
Slesinger & Dresser (1978), the half-life of TBTO in sea
water subjected to ultraviolet light is 18.5 days. In the
presence of a photosensitizing substance, such as acetone,
the half-life is 3.5 days. Seligman et al. (1986a)
suggested that, under natural conditions, photodegradation
is less important than biological action, the development
of phytoplankton leading to a partial degradation of TBTO.
Their measurements were made at relatively high concen-
trations of TBTO (744 µg/litre). Under these conditions,
light caused no degradation over 144 days. According to
Lee et al. (1987), degradation of low concentrations of
TBTO (less than 5 ng/litre) in estuary water is increased
when the assay is conducted in light. The half-life is
between 6 and 12 days, and the presence of significant
concentrations of phytoplankton increases the speed of

degradation. According to Maguire et al. (1983), photoly-
sis under natural light conditions in distilled or natural
water is limited, leading to a TBTO half-life in excess of
89 days. Under experimental conditions of strong UV light,
degradation is apparent. At 300 nm the half-life of TBTO
is 1.1 days, whereas at 350 nm it is more than 18 days.

In these assays, it is possible to demonstrate the role of
humic acids, particularly fulvic acid, which considerably
augment the speed of photolysis. Under such conditions,
the half-life of TBTO falls to 0.6 days at 300 nm and to
6 days at 350 nm. Under natural conditions in the port of
Toronto, Canada, the degradation after 89 days, remained
less than 50%.

4.3. Biodegradation

A number of studies have been conducted to verify that
microorganisms, notably bacteria, are capable of degrading
TBTO. In practice, physico-chemical mechanisms and bio-
logical mechanisms of degradation overlap. Evidence for
biodegradation constitutes an important element in the
assessment of risk. Published studies of observations made
in the field or the laboratory have shown definite evi-
dence of biological degradation of TBTO. Biodegradation
kinetics depend on environmental conditions such as tem-
perature, oxygenation, pH, the level of mineral elements,
the presence of easily biodegradable organic substances,
and the nature of the microflora and the possibility of
their adaptation. Biodegradation also depends on the con-
centration of TBTO being lower than the lethal or inhibi-
tory threshold for the bacteria.

Biodegradation is based on the formation of intermedi-
ate hydroxylated derivatives, progressive oxidative
debutylization following the splitting of the carbon-tin
bond. Dibutyl derivatives are formed, which appear to be
degraded more rapidly than tributyl derivatives to give
monobutyl derivatives; these, conversely, are mineralized
slowly. The end product may be butene. The quantities of
carbon dioxide formed remain small. The biodegradation of
organotin compounds does not seem to involve the formation
of methyl derivatives of tin. Such methyl derivatives have
been measured in some studies (Braman & Tompkins, 1979;
Guard et al., 1981; Hallas et al., 1982; Brinckman et al.,
1983), but have been shown to be the result of the trans-
methylation of inorganic tin by certain marine bacteria
( Pseudomonas ) frequently found in estuaries.

Sheldon (1975) proposed the following scheme for
degradation involving microorganisms:

A mechanism of biodegradation also exists under
anaerobic conditions (Maguire & Tkacz, 1985). Anaerobic
degradation is considered to be very slow by some workers
and more rapid than aerobic degradation by others.

Slesinger & Dresser (1978) conducted studies in a
Warburg respirometer under aerobic conditions and showed
that microflora derived from activated sludge and soil
were capable of partially degrading TBTO. The half-life
was 70 days, whereas under anaerobic conditions it was
200 days.

Henshaw et al. (1978) showed that pure cultures of
certain wood-degrading fungi, such as Coniophora puteana
and Coriolus polystictus, were capable of slowly biode-
grading TBTO and transforming it to dibutyl and monobutyl
derivatives.

Barug & Vonk (1980) studied the degradation of TBTO in
soil but could show no clear evidence for the action of
microorganisms. Under their experimental conditions, in
sterile or non-sterile medium, the half-life of TBTO
varied between 15 and 20 weeks depending on the soil type.
Barug (1981) was not able to isolate, from sediment or
soils, microorganisms capable of utilizing TBTO as a sole
carbon source. By contrast, in the presence of easily
biodegradable organic matter, biodegradation of TBTO is
apparent with the production of monobutyl derivatives and
smaller quantities of dibutyl derivatives. A number of
species were found to be capable of conducting such degra-
dation aerobically (bacteria: Pseudomonas aeruginosa and
Alcaligenes faecalis ; wood-degrading fungi: Coniophora
puteana, Trametes versicolor, and Chaetomium globosum ).
Under these conditions, they observed 70% degradation in
3 weeks. However, the breakdown of TBTO is not clearly
proved since the authors showed that TBTO accumulates in
the cell walls of bacteria and fungi.

Using water containing natural microflora, Olson &
Brinckman (1986) found no degradation of TBTO at a concen-
tration of 100 µg/litre and a temperature of 5 °C but did
record degradation at 28 °C. Their work also confirmed an
acceleration of degradation when the incubations were con-
ducted under light; the authors explained this acceler-
ation by invoking the role played by photosynthetic micro-
organisms.

Seligman et al. (1986a) also showed evidence for bio-
degradation; in medium polluted by TBTO at 0.5 µg/litre,
the TBTO half-life was 7 days in the dark and 6 days in
the light. In water containing 0.03 µg TBTO/litre, the
half-life was 19 days in the dark and 9 days in the light.
In all cases, dibutyl derivatives were formed and, to a
lesser extent, monobutyl derivatives. In studies with
¹⁴C-labelled TBTO, the measurement of ¹⁴CO₂ production
suggested a half-life of between 50 and 75 days.

Stein & Kuster (1982) demonstrated that TBTO is elim-
inated from waste water passing through sewage treatment
plants by adsorption onto sludge and biodegradation by
sludge organisms, provided that concentrations of TBTO
remain less than 5 mg/litre (see also section 5.3).

According to Maguire et al. (1984), the green alga
Ankistrodesmus falcatus was capable of bioaccumulating
TBTO (with bioconcentration factors of 3 x 10⁴) when it
was cultured in the presence of 20 µg TBTO/litre. When
the cultures were transferred to a non-contaminated
medium, 50% of the TBTO was transformed to dibutyl deriva-
tives or monobutyl derivatives and even to inorganic tin
over the course of 4 weeks. The assays were conducted on
axenic cultures of algae. It may be supposed that a bio-
logical effect was superimposed on physico-chemical degra-
dation mechanisms.

Maguire & Tkacz (1985) have shown that in sediments
there are oligochaetes that are also capable of metab-
olizing TBTO after it has been accumulated. However, the
simultaneous presence of bacteria in the test systems
means that a clear conclusion could not be reached.

According to Maguire et al. (1986), degradation can be
characterized as follows:

* Loss of TBTO by volatilization is very limited with a
half-life of more than 11 months.
* Hydrolysis of TBTO is equally slow with a half-life of
11 months.
* Photodegradation of TBTO plays a more important role
but the half-life of photodegradation is longer than
3 months. This route theoretically takes place but,
under natural conditions of illumination and the poor
penetration of UV light into turbid or coloured water,
it is inefficient.

* Aerobic biodegradation plays a role in water and sedi-
ment. The half-life varies considerably according to
conditions but is in the region of 4 to 5 months.
* Anaerobic degradation plays a role in water and sedi-
ment. The half-life varies considerably but is around
1.5 months.

The kinetics of degradation of dibutyl and monobutyl
tins are less well known. However, the degradation pro-
cesses of TBTO always results in the formation of metab-
olites less toxic than the parent compound.

Hinga et al. (1987) indicated a TBTO half-life of
between 5 and 19 days at 22-24 °C in model ecosystems.
Thain et al. (1987) suggested half-lives of 6 days in
fresh water and 60-90 days at 5 °C in sea water. In water
and sediment of the port of Toronto, the half-life varied
between 4 and 5 months (Maguire & Tkacz, 1985). In estuar-
ine waters of San Diego Bay, USA, the half-life varied
between 7 and 11 days at 12 °C, while in waters of the
Skidaway Estuary, it varied between 5 and 9 days at 28 °C
(Seligman et al., 1986a). Stang & Seligman (1986) using
contaminated sediment from San Diego Bay found that TBT
was degraded to monobutyltin. The degradation kinetic was
lower than in water, the half-life being approximately
162 days. In studies carried out by J.E. Thain & M.J.
Waldock (Personal communication to IPCS, 1989), naturally
contaminated sediments were maintained in the laboratory,
under flow-through conditions, at 12 °C. Degradation of
sediment-bound TBT was found to be a slow process. In
aerobic layers the half-life of TBT was between 4 and 5
months, but in deeper anaerobic layers a half-life value
was not obtained within 500 days.

4.4. Bioaccumulation and elimination

The lipophilic properties of TBTO and its moderately
high octanol-water partition coefficient (log P_ow > 3)
contribute to bioaccumulation in living organisms.

Evidence for such mechanisms and an evaluation of
their importance is highly relevant for hazard assessment,
both for the environment and for humans, since some of the
organisms exposed to TBTO are human food items, e.g.,
bivalve molluscs, crustaceans, and fish. Alzieu et al.
(1980) showed that in contaminated areas tin levels in the
flesh of oysters were 100 times higher than concentrations
in the water.

Laboratory experiments have been conducted under dif-
ferent conditions to demonstrate such bioaccumulation, and
have shown that bioconcentration factors vary considerably
between species.

In estuarine bacteria, Blair et al. (1982) found bio-
concentration factors varying between 100 and 30 000 in
species resistant to concentrations of 20 mg TBTO/litre.
As was indicated earlier, such bioconcentration might
result either from adsorption to the surface of the organ-
isms or from true bioaccumulation into the cells. In
phytoplankton, Maguire et al. (1984) reported a bioconcen-
tration factor of 30 000 in the green alga Ankistrodesmus
falcatus exposed for 1 week to concentrations of 20 µg
TBTO/litre. In the diatom Isochrysis galbano, Laughlin et
al. (1986b) reported a bioconcentration factor of 5500.

Studies on the possibility of bioaccumulation and bio-
magnification in molluscs, particularly bivalve molluscs,
are prominent in the literature because of human consump-
tion of oysters and mussels. Alzieu et al. (1982) showed
that TBTO accumulated in oysters, maintained in tanks with
panels of antifouling paint based on TBTO, to levels of
25 mg/kg (dry weight) of tissue and that this resulted in
problems of cavitation of the shell. Waldock et al.
(1983), in studies of the Pacific oyster Crassostrea gigas
exposed for 22 days to TBTO concentrations of 0.15 µg/litre
and 1.25 µg/litre, reported bioconcentration factors of
6000 and 2000, respectively. In European oysters (Ostrea
edulis) exposed to the same concentrations, they found
concentration factors of 1500 and 1000, respectively. In
both cases, after transfer of the oysters to clean water
there was a 50% fall in TBTO levels due to loss or degra-
dation. Laughlin et al. (1986b) reported bioconcentration
factors between 1000 and 7000 for mussels (Mytilus edulis)
exposed for between 3 and 7 weeks to TBTO concentrations
of 23, 45, 63, 141, and 670 ng/litre. For the higher con-
centrations, a plateau in uptake was reached within
2 weeks, but for lower concentrations, no plateau was
reached within the 7-week experiment. The authors con-
sidered that the mussel would be a good indicator organism
for monitoring marine pollution. Cheng & Jensen (1989)
transferred mussels ( Mytilus edulis ) from an unpolluted
area into net bags suspended in a marina in Denmark. They
monitored tin uptake and water concentrations of tin over
a period of 51 days. Accumulation was found to increase
exponentially with time for both total tin and organic
tin. Bioconcentration factors of 5000 to 60 000, much
higher than those from laboratory experiments, were
reported. Transfer of the mussels to the laboratory after
exposure resulted in a half-time for loss of organic and
total tin of 40 and 25 days, respectively. Laughlin et al.
(1986b) showed that bioaccumulation of TBTO by mussels
was not significantly affected by the presence of humic
acids or kaolin but that the presence of mucins secreted
by bacteria did limit bioaccumulation. It was also shown
that bioaccumulation by mussels was greater if the phyto-
plankton used as a food organism (Isochrysis galbana) was
also contaminated with TBT. Contamination via food organ-
isms was more important than via the water.

When feeding crabs with the brine shrimp (Artemia
salina) containing concentrations of TBTO of 6200 µg/kg
wet weight, Evans & Laughlin (1984) found a concentration
factor of 4400. Allen et al. (1980) reported limited bio-
accumulation (< 50) in a 1-week study using freshwater
gastropods (Biomphalaria glabrata). In crustaceans, par-
ticularly the crab Rhithropanopeus harisii, accumulation
of TBTO from a water concentration of 0.28 µg/litre pro-
duced a moderate bioconcentration factor of 60 over 4 days.

Bioaccumulation of TBTO is equally evident in fish.
After exposure of the sheepshead minnow (Cyprinodon
variegatus) for 58 days to concentrations of TBTO varying
between 0.96 and 2.07 µg/litre, Ward et al. (1981)
reported a whole body concentration factor of 2600. After
returning the fish to clean water, loss of TBTO was rapid
over the first 7 days then slower. After 20 days, the
authors reported a loss of 74% from the muscle and 80%
from the viscera. Detection of dibutyltin, monobutyltin,
and inorganic tin suggested possible metabolism. Bressa et
al. (1984) exposed the mullet Liza aurata for 2 months to
concentrations of 5 µg TBTO/litre and reported biocon-
centration factors of 20 to 30 in the liver and kidneys
but no residues in the muscle. After transfer to clean
water, concentrations of tin fell in all organs. Short &
Thrower (1986) studied bioaccumulation in salmon
(Oncorhynchus tshawytscha) exposed for 96 h to concen-
trations of 1.49 µg/litre and obtained concentration
factors of 4300 in the liver, 1300 in the brain, and 200
in muscle. Tsuda et al. (1987) showed that TBTO was
accumulated by carp (Cyprinus carpio) exposed for 14 days
to concentrations varying between 1.8 and 2.4 ng/litre.
Over 10 days they found a plateau in uptake and a concen-
tration factor of 1000; metabolism was evident. Tsuda et
al. (1986) reported concentration factors ranging between
360 and 3400 for round crucian carp (Carassius carassius
grandoculis) tissues exposed to tributyltin chloride for
7 days.

5. ENVIRONMENTAL CONCENTRATIONS

Summary

Levels of TBT in water, sediment, and biota are elevated
within the proximity of marinas, commercial harbours, cooling
systems, and fish nets and cages treated with TBT-based anti-
foulant paints.

TBT levels have been found to reach 1.58 µg/litre in sea
water and estuaries, 7.1 µg/litre in fresh water, 26 300 µg/kg
in coastal sediments, 3700 µg/kg in fresh water sediments,
6.39 mg/kg in bivalves, 1.92 mg/kg in gastropods, and 11 mg/kg
in fish. However, these maximum concentrations of TBT should
not be taken as representative, because a number of factors may
give rise to anomalously high values (e.g., paint particles in
water and sediment samples).

It has been found that measured TBT concentrations in the
surface microlayer of both sea water and fresh water are up to
two orders of magnitude above those measured just below the
surface. However, it should be noted that the recorded levels
of TBT in surface microlayers may be highly affected by the
method of sampling.

Older data may not be comparable to newer data because of
improvements in the analytical methods available for measuring
TBT in water, sediment, and tissue.

5.1. Sea water and marine sediment

The concentrations of TBT in sea water and sediment
are shown in Tables 3 and 5, respectively. Many papers
have reported an association between increased levels of
TBT in water, sediment, and biota and proximity to
pleasure boating activity (especially marinas) and the use
of antifouling paints on fish nets and cages. The degree
of tidal flushing and turbidity of water also influence
TBT concentrations in particular locations.

Table 3. Concentrations of tributyltin in estuarine and sea water
---------------------------------------------------------------------------------------
Sample Concentration Detection
Location Year depth^a (µg/litre) Form^b limit Reference
(metres) (µg/litre)
---------------------------------------------------------------------------------------
Denmark
Coastal waters 1986 0.1-0.2 <0.04 tin 0.04 Jensen & Cheng (1987)
Marinas 1986 <0.04-1.05 tin 0.04 Jensen & Cheng (1987)
Harbour areas 0.63-2.64 OT_o ICES (1987)

Finland
Harbours 1988 0.2 0.02-0.2 TBT 0.01 Yla-Mononen (1988)

France
Bay of Arcachon 1982 0.1-0.3 OT Alzieu & Heral (1984)
1984 0.7-1.2 tin 0.15 Alzieu et al. (1986)
<0.15-0.5 OT 0.1 Alzieu et al. (1986)
1985 0.3-1.0 tin 0.15 Alzieu et al. (1986)
<0.15 OT 0.1 Alzieu et al. (1986)
Anse de Camaret,
Brest 1987 (1) <0.002-0.004 TBT 0.002 Alzieu et al. (1989)
Auray river 1986-
estuary 1987 (1) 0.009-0.069 TBT 0.002 Alzieu et al. (1989)
La Rochelle 1986-
1987 (1) 0.02-0.119 TBT 0.002 Alzieu et al. (1989)
Oleron Island 1986-
1987 (1) 0.039-1.5 TBT 0.002 Alzieu et al. (1989)
Arcachon Bay 1986-
1987 (1) <0.002-0.089 TBT 0.002 Alzieu et al. (1989)

Norway
Oslo fjord <0.01 TBT_t 0.01 NIVA (1986)

Sweden
Coastal waters ND-0.04 TBT_t Bjorklund (1987b)

United Kingdom
Essex coast 1982 0.1-0.2 <0.03-0.9 TBT_t 0.03 Waldock & Miller (1983)
South-west coast 1984 <0.04-0.35 OT 0.04 Cleary & Stebbing (1985)
South-west coast 1986 surface 0.12-5.34 OT 0.04 Cleary & Stebbing (1987)
water
South-west coast 1986 0.5 <0.04-1.44 OT 0.04 Cleary & Stebbing (1987)
South-west coast 1986 bottom <0.04-2.6 OT 0.04 Cleary & Stebbing (1987)
South-west coast 1985 <0.02-0.68 TBT 0.02 Ebdon et al. (1988)
Poole harbour 1986 0.002-0.646 TBT_t Langston et al. (1987)
Essex coast 1986 0.1 <0.001-0.831 TBT 0.001 Waldock et al. (1987b)
South coast 1986 0.1 <0.001-1.52 TBT 0.001 Waldock et al. (1987b)
South-west coast 1986 0.1 <0.001-1.27 TBT 0.001 Waldock et al. (1987b)
South Wales coast 1986 0.1 <0.001-0.29 TBT 0.001 Waldock et al. (1987b)
North Wales coast 1986 0.1 <0.001-0.012 TBT 0.001 Waldock et al. (1987b)
---------------------------------------------------------------------------------------

Table 3. (contd.)
---------------------------------------------------------------------------------------
Sample Concentration Detection
Location Year depth^a (µg/litre) Form^b limit Reference
(metres) (µg/litre)
---------------------------------------------------------------------------------------
USA
Chesapeake Bay 1985 surface ND-1.171 TBT 0.008- Hall et al. (1986)
microlayer 0.01
Chesapeake Bay
(South) 1986 0.15 ND-0.1 TBT 0.001 Huggett et al. (1986)
San Diego Bay 1986 >0.5 0.005-0.235 TBT 0.005 Seligman et al. (1986b)
Californian coast 1986 <0.002-0.6 TBT 0.001- Stallard et al. (1987)
0.002
San Diego Bay 1983-
1985 0.3-0.6 <0.01-0.93 TBT 0.01 Valkirs et al. (1986)
San Diego Bay 1983-
1985 (0.1) <0.01-0.55 TBT 0.01 Valkirs et al. (1986)
USA harbours &
estuaries (0.5) <0.005-0.35 TBT 0.005 Grovhoug et al. (1986)
Coos Bay, Oregon surface 0.007-0.014 TBT Wolniakowski et al.
water (1987)
---------------------------------------------------------------------------------------
^a Figures in parentheses indicate distance from water bottom.
^b TBT = sample analysed for TBT and expressed as TBT.
TBT_t = sample analysed for TBT and expressed as tin.
tin = total tin expressed as tin.
OT = total organic tin expressed as tin.
OT_o = total organic tin expressed as TBTO.
Alzieu et al. (1986) monitored tin and organotin con-
centrations in both water and oyster tissue from Arcachon
Bay, France, between 1982 and 1985. They found that levels
in oyster tissue decreased by 5 to 10 times over this
sampling period following French Government restrictions
on the use of TBT in antifouling paints. Alzieu et al.
(1989) monitored TBT water levels at various locations on
the French Atlantic coast in 1986 and 1987 (Table 3), and
found that concentrations generally ranged between < 0.002
and 0.1 µg TBT/litre with the exception of a marina on
Oleron Island, which had levels of up to 1.5 µg/litre.
Levels were highest both in marinas and in the autumn,
presumably when boats were being hosed off ready for the
winter. The authors concluded that levels of TBT had gen-
erally decreased since the restrictions on TBT antifouling
paints, but in certain marinas levels were significantly
higher, suggesting continued use of TBT paints in contra-
vention of restrictions.

Waldock & Miller (1983) measured TBT levels in water
samples collected monthly during 1982 at Burnham-on-Crouch
on the east coast of the United Kingdom. They found a rise
in TBT levels in May, at a time when boats were being
freshly painted with TBT antifouling paints. There was a
second rise in TBT water concentrations in August, at a
time when boats were repainted for the major sailing event

of the year. Analysis of water samples from several areas
on the Essex coast showed that the highest levels (up to
2.25 µg TBTO/litre) were associated with the highest
density of pleasure craft. The authors also reported that
a site used by a large number of boats (on the south
coast of the United Kingdom but situated on an open
coastal site and with less turbid water) had relatively
low TBT levels in the sea water (< 0.08 µg TBTO/litre in
early August).

Waldock et al. (1987b) analysed water samples from
nine sites around the United Kingdom coast during 1986
following restrictions placed on the tin content of anti-
fouling paints containing TBT in January 1986. They
sampled from an enclosed bay, an open coastal site, and
seven estuarine sites. Within these general areas,
locations were found which reflected the incoming water
from a river, an area fished for shellfish, and a harbour
or marina. Half of the 250 samples taken during 1986 were
found to equal or to be above the United Kingdom environ-
mental quality target level (EQT; 20 ng/litre). Levels
were barely above the detection limits at the sites
upstream of boats. Harbours and marinas showed the highest
levels with tidal flushing being an important factor in
determining amounts of TBT detected. A marina in Plymouth,
which has poor flushing, had TBT concentrations consist-
ently greater than 1 µg/litre from May to September,
whereas a marina in the estuary of the River Dart, with
good flushing, had levels of less than 0.2 µg/litre. Six
of the nine sites exceeded the EQT by 3 to 4 times; these
were all sites used regularly by yachts. The other three
sites not used by yachts all showed low but often detect-
able levels with just one sample exceeding the EQT. The
authors also found increased levels of TBT close to areas
where boats were hosed down. Other reports confirmed that
the distribution of TBT in water was associated with the
proximity to intense boating activity (Cleary & Stebbing,
1985; Ebdon et al., 1988). Langston et al. (1987) reported
that sediments, likewise, contained more TBT (up to
520 µg tin/kg) near marinas than at the harbour mouth
(20 µg tin/kg) in Poole harbour, United Kingdom. There
was poor flushing in the harbour and sediment was not dis-
tributed; this was reflected in the water levels, which
were 0.002 to 0.139 µg tin/litre in the general harbour
area and 0.234 to 0.646 µg/litre in the marina.

Cleary & Stebbing (1987) surveyed vertical water pro-
files in south-west England at sites already investigated
two years before. They did not find a systematic decline
in concentrations between the two surveys. The concen-
trations in the surface microlayer were 1.9 to 26.9 times
higher than those at 0.5 m below the surface (see Table 3).

Waldock et al. (1988) analysed water samples collected
in 1987 from commercial harbours and anchorages in the
United Kingdom. Significant concentrations were found;

several samples taken in the immediate vicinity of ships
had levels exceeding 0.05 µg TBT/litre. However, the
highest concentrations were found near to centres of
yachting activity, with over 0.6 µg/litre being found at
one site. The highest concentration found close to commer-
cial vessels in harbours was 0.078 µg/litre, but this was
within 2 m of an oil tanker. A concentration of 0.25 µg per
litre was recorded outside a shipyard where a 3000-tonne
vessel was being hosed down on the foreshore, and a con-
centration of 0.137 µg/litre was measured in surface
water close to a vessel at anchor in the River Fal. In
general, however, few samples taken in close proximity to
commercial ships exceeded 0.02 µg/litre.

Bacci & Gaggi (1989) monitored TBT and its degradation
products in harbours, marinas, and the open sea from the
northern Tyrrhenian Sea, Italy. Concentrations of up to
3.93 µg TBT/litre were measured in the various harbours
and marinas, but no organotin compounds were detected in
samples from the open sea. However, considering the detec-
tion limits of the analytical technique used (0.02 µg per
litre for both TBT and DBT), levels higher than the NOEL
(i.e. 0.01 µg/litre, UNEP, 1989) cannot be excluded.
From these preliminary results, it appears that, under
unfavourable meteorological conditions (e.g., moderate
southerly winds), significant quantities of TBT and
related compounds could contaminate open sea sites for a
few days per year.

The highest levels of TBT around the coasts of the USA
and Denmark were also associated with marinas or harbours
used by small pleasure craft, with TBT levels generally
showing a falling trend from the inner part to the
entrance (Grovhoug et al., 1986; Seligman et al., 1986b;
Jensen & Cheng, 1987). Stallard et al. (1987) analysed
both water and sediment from the Californian coast.
Highest TBT levels, up to 0.6 µg/litre water and 23 µg/kg
sediment, were found near marinas. Levels were lower in
other coastal areas and were lowest out in the open sea.
Valkirs et al. (1986) measured TBT in surface water (at a
depth of 0.3 to 0.6 m) and found that, over the period
1983-1985, TBT levels had increased in San Diego Bay, USA.
Seligman et al. (1989) measured TBT in the waters of
several harbours in the USA. Of the samples collected, 75%
contained TBT levels below the detection limit (< 5 ng per
litre). The highest concentrations were found in yacht
harbours and near to vessel repair facilities, with sig-
nificant levels being found near dry docks. The authors
also found a high degree of variability in TBT concen-
trations depending on the tidal movement, the season, and
intermittent point source discharges.

Hall et al. (1988a) measured TBT biweekly for a 4-
month period (June-September 1986) in the Port Annapolis
marina, Mears marina, Back Creek, and the Severn River
area of northern Chesapeake Bay, USA. Maximum concen-
trations of TBT were reported at both Port Annapolis

marina (1.8 µg tin/litre) and Mears marina (1.17 µg tin
per litre) during early June, followed by significant
reductions during late summer and early autumn. The day of
the week (Thursday-Monday) on which samples were taken
during the daily experiments was not found to signifi-
cantly affect TBT concentrations. Peak concentrations were
found to occur during a rising tide.

Balls (1987) reported that TBT levels in water were
initially (immediately after fish cages were treated with
antifoulants) 1 µg/litre (as tin) within fish cages,
falling to 0.1 µg/litre after 2 weeks and 0.005 µg per
litre after 5 months. Initial concentrations were 0.1 µg
tin/litre at a distance of 20 m from the cages, with
concentrations in the main body of the sea loch being
< 0.028 µg/litre.

5.2. Fresh water and sediment

Analysis for TBT compounds in the Great Lakes, N.
America, has revealed levels often comparable, and in many
cases higher (200 times higher in one sample), than those
measured in estuaries (Maguire et al., 1982; Maguire,
1984; Maguire et al., 1985; Maguire et al., 1986; Maguire
& Tkacz, 1987). Levels of TBT in water were found to be
greater in the surface microlayer than in the subsurface
samples. For example, water samples from Ontario lakes and
rivers showed surface levels of 0.15 to 60.7 µg tin/litre
compared to subsurface levels of between 0.01 and 2.91 µg
per litre (Maguire et al., 1982). TBT was found in the
Great Lakes and in rivers at levels up to those causing
effects on trout in the laboratory; Maguire & Tkacz (1987)
reported a level of 66.8 µg tin/litre in the surface
microlayer. In the United Kingdom, samples of fresh water
from near boatyards contained up to 3.2 µg TBT/litre
(Waldock 1989). In Lake Zurich and Swiss rivers, levels
were found to be much lower, i.e. up to 0.015 µg/litre
(Muller, 1987b). Kalbfus (1988) analysed water samples
from marinas on Lake Constance in 1987 and 1988 and found
that TBT levels rose to a peak in May which corresponded
to the boating activity on the lake. For example, at
Goren, TBT levels rose from 0.13 µg/litre in April to
0.5