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

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

    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

         The World Health Organization welcomes requests for permission
    to reproduce or translate its publications, in part or in full.
    Applications and enquiries should be addressed to the Office of
    Publications, World Health Organization, Geneva, Switzerland, which
    will be glad to provide the latest information on any changes made
    to the text, plans for new editions, and reprints and translations
    already available.

    (c) World Health Organization 1990

         Publications of the World Health Organization enjoy copyright
    protection in accordance with the provisions of Protocol 2 of the
    Universal Copyright Convention. All rights reserved.

         The designations employed and the presentation of the material
    in this publication do not imply the expression of any opinion
    whatsoever on the part of the Secretariat of the World Health
    Organization concerning the legal status of any country, territory,
    city or area or of its authorities, or concerning the delimitation
    of its frontiers or boundaries.

         The mention of specific companies or of certain manufacturers'
    products does not imply that they are endorsed or recommended by the
    World Health Organization in preference to others of a similar
    nature that are not mentioned. Errors and omissions excepted, the
    names of proprietary products are distinguished by initial capital



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 


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


   3.1. Uses                  
   3.2. Production            
   3.3. Regulations           


   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   


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


   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  


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


   8.1. Aquatic plants        

   8.2. Aquatic invertebrates 
          8.2.1. Trematode parasites of man    
          8.2.2. Freshwater molluscs   
           Acute toxicity  
           Short- and long-term toxicity   
           Factors affecting toxicity  
          8.2.3. Marine molluscs   
           Acute toxicity  
           Short- and long-term toxicity   
           Reproductive effects    
           Effects on growth   
           Shell thickening    
          8.2.4. Crustaceans   
           Acute effects
           Short- and long-term toxicity   
           Reproductive effects    
           Limb regeneration   
           Behavioural effects 
          8.2.5. Other aquatic invertebrates   
           Acute effects   
           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  


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


   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


   11.1. 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  
          Skin irritation 
          Eye irritation  
          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 


   13.1. 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        






MEDIO AMBIENTE                        




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,

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

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

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


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

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

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


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

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


    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).


    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.


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


1.1.  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 Pow)  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-

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

    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

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

    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

    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

    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

    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

    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

    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

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.


2.1.  Identity of tributyltin compounds

    Tributyltins  compounds are organic derivatives of tin
(SnIV)    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-

    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-

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

    Bu3-Sn-O-Sn-Bu3+ HO  -»  2Bu3-Sn-OH
    Bu3-Sn-OH-H+  -»  Bu3SnOH2+
    Bu3-Sn-OH + CO32-  -»  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

    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

Relative       596            411         325         309         568.7         374.7         ca.500

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)

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

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  Pow)
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-

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

    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)
---------------------------------------------------------------------------------------------------------  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).  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-
*   formation  of  hydrides  with  the  general  structure
    RnSnH4-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.  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.  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-

    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

    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.


3.1.  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-

    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.



     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

     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

    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

    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:

R3SnX -----> (R3Sn)2O -----> (R3Sn)2CO3
                                  |  UV or microorganisms
                                  |  UV or microorganisms
                                  |  UV or 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

    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-

    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 Pow >  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.



     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   deptha    (µg/litre)    Formb   limit    Reference
                        (metres)                        (µg/litre)
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     OTo              ICES (1987)

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

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)

Oslo fjord                         <0.01        TBTt   0.01      NIVA (1986)

Coastal waters                     ND-0.04       TBTt             Bjorklund (1987b)

United Kingdom
Essex coast       1982   0.1-0.2   <0.03-0.9    TBTt   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)
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   TBTt             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   deptha    (µg/litre)    Formb   limit    Reference
                         (metres)                       (µg/litre)
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)
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.
    TBTt = sample analysed for TBT and expressed as tin.
    tin = total tin expressed as tin.
    OT = total organic tin expressed as tin.
    OTo = 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
0.58 µg/litre    in May, but by July the levels had fallen
again to 0.028 µg/litre.    At the same time TBT levels in
sediment rose from 830 µg/kg   in May to  2700 µg/kg    in
June  and then to  3700 µg/kg   in July.   Similarly, when
samples were taken on Wannsee in Berlin, levels were found
to  be 0.02 µg/litre    when there  were no  boats on  the
water, but at Tegel, Berlin, TBT levels were 0.25 µg   per
litre  when most of the boats were in the water. A coastal
marina  at Kiel, on the  Baltic, showed levels of  0.35 µg
per  litre in April  when only half  of the moorings  were

    Shiff  et al. (1975)  monitored water and  mud samples
6.5 months  after  the  application of  controlled-release
BioMet SRM pellets (rubber formulation containing TBTO) in

Zimbabwe.  The pellets were applied  at a rate of  20 g/m2
for  the control of freshwater  snails, intermediate hosts
of the schistosomiasis parasite. Highest levels of organo-
tin  were found in the  mud immediately under the  pellets
(up  to 5 mg/kg). Levels  in the mud  dropped off  rapidly
further  away from the  pellets; at 2 cm  organotin levels
were < 0.6 mg/kg. The organotin level in surface water was
< 0.01 mg/litre and in background mud < 0.06 mg/kg.
Table 4.  Concentrations of tributyltin in fresh water
                          Sample      Concentration        Detection
Location            Year  depth       (µg/litre)    Forma  limit     Reference
                          (metres)                         (µg/litre)
Ontario lakes &
 rivers                               0.01-2.91     TBTt   0.01      Maguire et al. (1982)
Ontario lakes &           surface
 rivers                   microlayer  0.15-60.7     TBTt   0.01      Maguire et al. (1982)
St Clair River,           surface
 Ontario                  microlayer  ND-0.03       TBTt   0.01      Maguire et al. (1985)
Canadian waterways        0.5         <0.01-2.34   TBTt   0.01      Maguire et al. (1986)
Ontario waterways         surface     1.9-473       TBTt   1.0       Maguire & Tkacz (1987)
Ontario waterways         0.5         <0.01-1.72   TBTt   0.01      Maguire & Tkacz (1987)
Quebec waterways          surface     5.5 & 15.2    TBTt   1.0       Maguire & Tkacz (1987)
Quebec waterways          0.5         <0.01-0.03   TBTt   0.01      Maguire & Tkacz (1987)
British Columbian
 coast                                up to 0.078   TBT              Humphrey & Hope (1987)
Federal Republic of Germany
Lake Constance      1987-             up to 0.58    TBT              Kalbfus (1988)
 marinas            1988
Lake Zurich &       1985  surface     0.007-0.015   TBTc   0.001     Muller (1987b)
 rivers                   water
Harbours            1983-
                    1984              0.005-1.636   TBT    d
Rivers              1983-
                    1985              0.001-0.016   TBT    d
United Kingdom
Wroxham Broad,
 Norfolk            1987              up to 0.9b    TBT              Waldock et al. (1987a)
River Thames        1987                  0.064c    TBT              Waldock et al. (1987a)
River Bure          1986-
                    1987  0.1         ND-1.54       TBT    0.001     Waldock (1989)
River Yare          1986-
                    1987  0.1         <0.001-3.26  TBT    0.001     Waldock (1989)
New York State            surface     2.0-23.8      TBTt   1.0       Maguire & Tkacz (1987)
 waterways                microlayer
a   TBT = sample analysed for TBT and expressed as TBT.  TBTt = sample analysed for TBT 
    and expressed as tin. TBTc = sample analysed for TBT and expressed as tributyltin chloride.
b   Samples from local boatyards contained up to 1.5 µg/litre.
c   Samples from marinas contained up to 1.3 µg/litre.
d   Personal communication from M.D. Muller to IPCS.
Table 5.  Concentrations of tributyltin in sediment
                          Sample   Concentrationa      Detection
Location            Year  depth    (µg/kg)       Formc   limit   Reference
                          (metres)                       (µg/kg)
Ontario lakes &
 rivers                   0.02     30.9-110      TBT     5       Maguire (1984)
Canadian waterways        0.02     <10-10 780   TBTt    10      Maguire et al. (1986)
British Columbian                  up to
 coast                             17 000        TBT             Humphrey & Hope (1987)

Canada & USA
Detroit & St Clair
 rivers                   0.02     ND-70         TBTt    5       Maguire et al. (1985)

Eems-Dollard                       < 25b        TBTt    25      TWG (1988c)
Various locations                  <50-8800     TBTt    50      TWG (1988c)

Lake Zurich         1880-d
                    1985  120      ND            TBTc    0.01    Muller (1987b)
Lake Zurich         1980-
                    1984  120      280           TBTc    0.01    Muller (1987b)
Lake Zurich &
 Boden              1984           2.0-3550      TBT             e

United Kingdom
Poole Harbour,
 Dorset             1986           20-520        TBTt            Langston et al. (1987)

Californian coast   1986  0.1      <2.0-23      TBT     1.0-    Stallard et al. (1987)
San Diego Bay       1983  0.35     <2.0-300     TBT             Stang & Seligman (1986)
USA harbours &
 estuaries                         1.4-178       OT              Grovhoug et al. (1986)
Californian coast                  15-527        TBT             Stephenson et al.
Virginian coast           0.02     23-290        TBT             Rice et al. (1987)
Great Bay estuary         0.02     12-44         TBTt            Weber et al. (1986)
a   Concentrations given as µg/kg dry weight unless stated otherwise.
b   Wet weight value.
c   TBT = sample analysed for TBT and expressed as TBT.
    TBTt = sample analysed for TBT and expressed as tin.
    TBTc = sample analysed for TBT and expressed as tributyltin chloride.
    OT = total organic tin expressed as tin.
d   Museum core from the nineteenth century.
e   Personal communication from M.D. Muller to IPCS.
5.3.  Sewage treatment

    The  mono-, di-, and  tri-butyltin  content  of  waste
water entering a sewage treatment plant in Switzerland was
measured  and its fate  was monitored through  the various
processes  of settlement, digestion, and filtration of the
sewage (Fent, 1989a; Fent et al., 1989). Concentrations of
MBT,  DBT,  and  TBT  were  170,  152,  and  155 ng/litre,
respectively,  in the incoming  raw waste water,  averaged
over  three days of monitoring. About 90% of the organotin
was  associated with particulate matter, 10% being in sol-
ution  (Table 6).   A  substantial amount  of the incoming
butyltin  compounds was lost from the effluent during pri-
mary settlement. The removal of particulate matter at this
stage took 74% of the incoming organotin. In the secondary
effluent,  after activated sludge  digestion, MBT and  DBT
were  found at  levels similar  to those  in  the  primary
effluent;  TBT  concentrations were  reduced to 6 ng/litre
and  found only on  the particulate matter.  In the  final
effluent  from the plant, after filtration, concentrations
were  4, 3, and 4 ng/litre for MBT, DBT, and TBT, respect-
ively.  Thus, 98% of the butyltin was removed  from  waste
water  in the sewage plant. The authors point out that not
all  treatment plants have filtration; in these cases only
87% would be removed and effluent concentrations  of  9-70
ng/litre found.  Levels of butyltin in the  sewage  sludge
(which is removed from the plant and used as fertiliser on
farm  land) were 0.36, 0.38, and 0.34 mg/kg dry weight for
MBT,  DBT, and TBT,  respectively, in the  raw sludge  and
0.62,  1.23,  and 1.12 mg/kg  dry  weight in  the digested
sludge  after 35 days of anaerobic conditions. The authors
point out that 900 kg/year of butyltin could be  added  to
Swiss  soils via sewage sludge. The source of TBT detected
in  the sludge  was not  identified or  specified  in  the
Table 6.  Levels of organotin compounds in municipal waste watera
                           MBT                    DBT                     TBT
Date              water  particles  %    water  particles  %     water  particles  %
23 February 1988   34      216     86     14       113     89     14       178     93
23 February 1988   25      181     88     10       163     94     14       158     92
28 February 1988   28      114     80     11       180     94     27       129     83

Mean               29      170     85     12       152     93     18       155     90
a   Levels in ng/litre are calculated as ions and corrected for recovery (55-70%); the
    percentage of organotins associated with particles is also given. From Fent (1989a).
5.4.  Biota

    Concentrations of TBT in biota are given in Table 7.

    Alzieu (1981) analysed the Pacific oyster  (Crassostrea
 gigas) for  total  tin  levels following  problems  in the

French  oyster  industry in  the  late 1970s  (see section
10.1).  He reported that most of the tin  accumulated  was
in the digestive gland and in the gills.  Highest residues
were found in oysters from the Bay of  Arcachon  (residues
in  digestive  gland and  gill were up  to 7.03 and  17.37
mg/kg,  respectively), an area with large numbers of small
pleasure boats. Tin levels were stated to be influenced by
tidal  flushing;  both the  Bay  of Arcachon  and Marennes
Oleron were used by a large number of boats,  but  residue
levels in oysters collected from the latter site had lower
tin  levels (the Bay of  Arcachon has poor tidal  flushing
compared  to  Marennes  Oleron).  Alzieu  &  Heral  (1984)
reported  that  the greatest  accumulation  of tin  was in
close  proximity to a  marina. Oysters transferred  to the
marina  site accumulated a  total tin level  of  110 mg/kg
(dry weight) within 80 days, whereas oysters maintained as
controls in a local river or in the laboratory accumulated
< 1 mg/kg  over the same  period. Waldock &  Miller (1983)
analysed  oysters  from  the Essex  coast, United Kingdom,
and,  although  both Pacific  and European oysters  (Ostrea
 edulis) contained  similar  residues  of  total  tin,  the
Pacific oyster residues had a higher percentage of TBT.

    There  are seasonal differences  in the levels  of TBT
(and  DBT) found in mussels  (Mytilus edulis) in the field.
It  has been suggested that, while these are predominantly
due  to changes in  boating activity affecting  the avail-
ability of TBT to the organisms, physiological differences
in  the animals at different times of year may also partly
explain the results. The relative amounts of TBT  and  DBT
in mussels are thought to reflect the rate of input to the
animal.   A high ratio of DBT to TBT residues reflects low
input rates, and vice versa (Page, 1989).

Table 7.  Concentrations of tributyltin in biota
                                                            Concentrationc         limit    
Organism                 Year   Locationa           Organb     (mg/kg)     Formf  (mg/kg)   Reference
European oyster                 French coast        DG      0.54-7.03      tin              Alzieu
  (Ostrea edulis)                French coast        gill    <0.5-17.37    tin              (1981)
                         1982   Essex coast, UK     DG      <0.23-2.05    TBTo   0.075     Waldock &
                         1982   Essex coast, UK     rest    <0.4-1.99     TBTo   0.075     Miller 
Pacific oyster                  French coast        DG      <0.5-2.5      tin              Alzieu
  (Crassostrea gigas)            French coast        gill    <0.5-3.5      tin              (1981)
                         1982   Essex coast, UK     DG      4.05-8.64      TBTo   0.075     Waldock & 
                         1982   Essex coast, UK     rest    3.5-7.5        TBTo   0.075     Miller (1983)
                                Coos Bay, USA               0.05-0.189     TBT              Wolniakowski 
                                                                                             et al. (1987)
Eastern oyster                  Virginia, USA       WB      0.59-1.57      TBT              Rice et al. (1987)
  (Crassostrea virginica)         USA coast           WB      <0.12-3.9     TBT              Wade et al. (1988)

Common mussel            1985-  Japan                       ND-0.28g       TBTo   0.05      EAJ (1988)
  (Mytilus edulis)        1987   USA coast           WB      0.25-3.85      TBT              Wade et al. 
Asiatic mussel           1985-  Japan                       0.3-0.48g      TBTo   0.05      EAJ (1988)
Mussel                          Californian coast,          0.107-6.39     TBT              Stephenson et 
                                USA                                                         al. (1987)

Shellfish                       USA coast                   0.23-7.35      OT               Grovhoug et al. (1986)
                                B.C., Canada                up to 1.8      TBT              Humphrey & Hope (1987)
                                Netherlands                 <0.025-0.22   TBTt   0.025     TWG (1988c)

Dogwhelk                        Fal estuary, UK             0.023-0.786    TBTt             Bryan et al. (1987)
  (Nucella lapillus)             South-west coast,           0.036-0.633    TBTt             Gibbs et al. 
                                UK                                                          (1987)

Various snail species    1988   Finnish harbours    SP      0.04-0.1g      TBT    0.01      Yla-Mononen 

Table 7.  (contd.)
                                                             Concentrationc        limit    
Organism                 Year   Locationa           Organb      (mg/kg)    Formf  (mg/kg)   Reference
Herring                  1984   Vancouver harbour,  WB      0.24           TBTt   0.01      Maguire et 
  (Culpea harengus)              Canada                                                      al. (1986)
Finfish                         B.C., Canada        DM      up to 11       TBT              Humphrey & 
                                                                                             Hope (1987)
Salmon species                  USA                 MT      ND-0.2d        TBT              Short & 
                                USA                         0.28-0.9e      TBT              Thrower (1986)

Various fish species     1982-  Jordan harbour,     WB      <0.01-0.02g   TBTt   0.01      Maguire et 
                         1983   Canada                                                      al. (1986)
Various fish species            Japan               MT      ND-0.31d       TBTc             Hada (1986)
Various fish species     1985-  Japan                       ND-1.7g        TBTo   0.05      EAJ (1988)
Various fish species            Netherlands                 <0.025-0.26   tin    0.025     TWG (1988c)
Various fish species     1988   Finnish harbours    WB      <0.01-0.1g    TBT    0.01      Yla-Mononen 
Oystercatcher            1986   Exe estuary, UK     liver   TR-0.08        TBTt   0.02      Osborn & 
  (Haematopus ostralegus) 1986   Exe estuary, UK     MT      0.01-0.19      TBTt   0.02      Leach (1987)
Grey starling            1985-  Japan                       ND (< 0.05)g  TBTo   0.05      EAJ (1988)
Black-tailed gull        1985-  Japan                       ND (< 0.05)g  TBTo   0.05      EAJ (1988)
Seals                                               BL      ND             TBT    0.01      h
a   UK = United Kingdom; B.C. = British Columbia.
b   DG = digestive gland; rest = tissues other than digestive gland; DM = dorsal muscle; MT = muscle 
    tissue; BL = blubber; WB = whole body; SP = soft parts.
c   Concentrations measured as mg/kg dry weight unless stated otherwise; TR = trace; ND = not detectable.
d   Fish collected from local fish markets.
e   Salmon raised in TBT-treated sea pens.
f   TBT = sample analysed for TBT and expressed as TBT; TBTt = sample analysed for TBT and expressed as 
    tin; TBTc = sample analysed for TBT and expressed as tributyltin chloride; OT = total organic tin 
    expressed as tin; tin = total tin expressed as tin; TBTo = sample analysed for TBT and expressed 
    as TBTO.
g   Wet weight value.
h   Personal communication from M.J. Waldock to IPCS.
    Gibbs  et  al.  (1987)  found  highest  levels  of TBT
(0.132-0.633 mg  tin/kg dry weight) in  dogwhelks from the
"enclosed"  waters of Plymouth Sound  and Torbay, United
Kingdom,  whereas levels were less than 0.113 mg tin/kg on
the  North Cornish coast.   Bryan et al.  (1987)  reported
residues  between 0.374 and 0.786 mg tin/kg (TBT fraction)
for dogwhelks from the Fal estuary, in the  south-west  of
England, whereas dogwhelks from around the Isle  of  Mull,
off the Scottish mainland, contained levels of  less  than
0.03 mg/kg.  The  Environment  Agency of  Japan  monitored
various fish and shellfish species from different areas of
Japan  between 1985 and  1987. The lowest  levels of  TBTO
(< 0.05 mg/kg wet weight)  were found off the  open  coast
of Japan, higher levels being found in bays and estuaries.
The highest levels reported were in sea bass from the Seto
Inland  Sea (up to  1.7 mg/kg). The level  of TBTO in  the
biota  did  not  change significantly  during the sampling
period (EAJ, 1988).

    Since 1987 only vessels of > 25 m have been allowed to
use  TBT  antifouling paints  in  the United  Kingdom (see
section 3.3).  Bailey  & Davies  (1988a) analysed dogwhelk
and scallop from an area around an oil terminal frequented
by  large  ships at  Sullom  Voe, Shetland.   Elevated tin
levels were found in both dogwhelk (up to  0.16 mg  tin/kg
wet  weight)  and scallops  (up  to 0.23 mg/kg  in gonadal
tissue)  within Sullom Voe  (especially in areas  close to
the  oil terminal) compared  to those collected  from  the
surrounding area (< 0.03 mg/kg).

    Increased  levels  of TBT  in  biota have  been  found
associated with fish nets and cages. Davies et al. (1987b)
found  that residues of total tin in dogwhelks were higher
near fish cages in Loch Laxford (< 0.01-0.33 mg/kg), a sea
loch  in Scotland, and in the harbour areas of Loch Crinan
(< 0.01-0.17 mg/kg) than outside the sea lochs (< 0.02 mg/kg).

    Short & Thrower (1986) found TBT residues  of  between
0.28    and   0.9 mg   tin/kg    in   salmon  (Oncorhynchus
 tshawytscha) maintained  in TBT-treated sea pens  for 3 to
19 months.   The authors also monitored salmon for sale in
American  fish markets and found TBT residues of up to 0.2
mg/kg. They also found that cooking does  not  effectively
destroy or remove TBT from salmon tissues.



     Tributyltin is absorbed from the gut (20-50%  depending  on
 the vehicle) and via the skin of mammals (about 10%),  and  can
 be transferred across the blood-brain barrier and from the pla-
 centa  to the fetus.  Absorbed material is  rapidly and  widely
 distributed amongst tissues (principally liver and kidney).

     Metabolism  in mammals is rapid; metabolites are detectable
 in  blood within 3 h of TBT administration.  TBT is a substrate
 for  mixed-function  oxidases in vitro,   but these  enzymes are
 inhibited  by TBT in vitro  at very high concentrations. Rate of
 loss  differs with different tissues and estimates for biologi-
 cal half-lives in mammals range from 23 to about 30 days.

     Metabolism  occurs in lower  organisms but is  slower, par-
 ticularly  in  molluscs.  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 the
 calcification of the shell of oysters (Crassostrea  species).

6.1.  Metabolism of TBT in mammals

    A  number  of  workers have  studied  the  absorption,
metabolism,  and  elimination of  organotin derivatives in
various  animals  species,  especially  in  mammals.  Some
studies  were conducted  in vivo and  others  in vitro using
isolated liver microsomes.

    The behaviour of organotin compounds depends partly on
their  chemical structure and partly  on speciation.  How-
ever, the following statements generally apply:

*   The distribution of TBT in organisms is usually rapid.
    In  a number of  species (rat, mouse,  rabbit, guinea-
    pig),  it  is found  preferentially  in the  liver and
    kidney  and, to a lesser  extent, in the spleen,  fat,
    lungs, brain, and muscle.
*   Excretion is via the bile rather than the urine.
*   In tissues, particularly the liver, there is a process
    of   biotransformation  characterized  by  progressive
    de-alkylation  leading  to breakdown  to inorganic tin
    (Cremer, 1957; Bridges et al., 1967).

    In  an  in vivo study, Brown et al. (1977) administered
113Sn-labelled   TBTO to mice by ip injection.   They  re-
ported  an initial rapid elimination, followed by a slower
phase,  in the faeces. Part of the radiolabel was retained
in the tissues but turn-over occurred, with  a  biological
half-life for elimination of 23 to 29 days.

    Evans  et al. (1979), under similar conditions, admin-
istered 14C-labelled   TBTO to mice in the drinking-water,
at  low doses continuously  for up to  30 days. There  was
absorption  from  the  intestine and  accumulation  in the
liver,  spleen, kidney, and fat (and to a lesser degree in
muscle, lung, brain, and blood). In a second  study,  mice
were  similarly dosed for 31 days.  On cessation of dosing
with 14C-labelled   TBTO, examination of the animals for a
further  15 days  demonstrated  loss of  TBTO  retained in
these  tissues; the  loss reached  97% in  liver,  73%  in
kidney,  and 30%  in fat,  and the  TBTO  had  disappeared
completely  from the blood.   Studies in metabolism  cages
indicated  that the principal  route of loss  was via  the
faeces; limited amounts of labelled CO2 were exhaled.

    Iwai  et  al.  (1980)  studied  the  distribution  and
accumulation  of tributyltin and its  metabolites in areas
of  the brain of rabbits.  After a single oral dose of TBT
chloride, high concentrations of tributyltin were found in
the  frontal  and temporal  lobes  and in  the  cerebellum
initially.   Thereafter, there was a rapid decrease in TBT
residues  and an increase in levels of monobutyltin, which
persisted  for much longer. Persistence occurred preferen-
tially  in the grey matter  rather than the white  matter.
The  authors'  interpretation  was that  TBT, which passes
readily  through  the  blood-brain barrier,  is mainly de-
alkylated  in  the  grey  matter  and  that  the metabolic
product remains there.

    Humpel et al. (1986) administered 113Sn-labelled  TBTO
orally  to  rats  and  found  that  the  absorption varied
between 20% and 55% depending on the vehicle  used.   High
residues of tin were found in the liver and kidney (1 to 3
days  after  dosing) of  which  only approximately  5% was
unchanged  TBT.  Other tissues showed lower concentrations
of the label but the fraction of unchanged TBT was higher.
The  exact nature of the  metabolites could not be  ident-
ified  by  the  analytical  method  used  (HPLC),  but the
pattern was indicative of progressive debutylation.  Daily
administration  of TBTO for  14 days resulted in  steadily
increasing concentrations of label in all tissues. Steady-
state  levels were estimated  to be reached  after 3 to  4
weeks.  When  Snoeij  et  al.  (1987)  administered   14C-
labelled  TBT acetate as a single oral dose to rats, about
20% absorption occurred.  The presence of DBT and  MBT  in
plasma  (after  TLC  separation) was  demonstrated 3 h and
27 h after dosing.

    TBT  may cross  the placenta  to some  extent, as  was
shown  by  the presence  of label in  rat fetuses after  a
single oral dose to the mother at day 18 of pregnancy. The
concentration  in fetal tissue  was comparable to  that of
the mother's muscle tissue (Humpel et al., 1986).

    After  administration of neat 113Sn-labelled   TBTO to
the  intact skin of baboons  for 7 h, 10 to  15% was esti-
mated  to reach the  systemic circulation (Humpel  et al.,

    Metabolism of tributyltin derivatives has been clearly
demonstrated in  in vitro studies. Casida et al. (1971) and
Fish et al. (1975, 1976) studied the  possible  metabolism
of TBT acetate using rat hepatic microsomes in  the  pres-
ence of NADPH. They demonstrated hydroxylation by monooxy-
genases  of the principal carbon-hydrogen bonds  ( alpha and
 beta to   the tin atom) of 24% (at the  alpha position) and
50%  (at the  beta position). The hydroxylated  alpha metab-
olite is unstable and rapidly splits to form  the  dibutyl
derivative, followed by 1-butanol and then butane. Accord-
ing  to Kimmel et  al. (1977), the  same type of  reaction
occurs in microsome preparations from mice.

    Uhl  (1986) dissolved TBTO (9.88 or 5.54 mg) in a mix-
ture  of 3 ml cherry brandy  and 7 ml ethanol and  gave it
orally to a volunteer. TBTO and its  degradation  products
were determined in urine by gas chromatography after reac-
tion with methyl magnesium bromide. Only 5.1% to  5.4%  of
the  dose was  found in  the urine,  mainly as  dibutyltin
metabolites.   Butyltin  levels  in  the  urine  decreased
rapidly  during the first days after administration. After
dermal  application of 20 µl   (23.4 mg) of undiluted TBTO
on  the arm of a volunteer, approximately 0.2% of the dose
was excreted in the urine, of which about 20% was found to
be tributyltin.

6.2.  Metabolism of TBTO in other organisms

    Lee  (1985, 1986) examined  the capacity of  organisms
from  various aquatic trophic  levels to metabolize  TBTO.
He  used  the  blue crab  (Callinectes  sapidus), the brown
shrimp  (Penaeus  aztecus), a  fish  (the  spot,  Leiostomus
 xanthurus), and     the     Eastern    oyster  (Crassostrea
 virginica). The  organisms  were  exposed to  14C-labelled
TBTO  via the water (6 µg/litre   for the crab and shrimp;
2 µg/litre    for the fish  and the oyster)  and via  food
(shrimp containing about 20 mg/kg) in the case of the crab
and fish. In all test species he reported a  rapid  uptake
of 14C-labelled   TBTO into various organs.  In crabs and
shrimps,  he  observed,  after 3 days,  the  appearance of
various  metabolites in the hepatopancreas (dibutyl, mono-
butyl,  and polar derivatives).  In the fish, the same was
seen in the liver. In oysters, the process was much slower
and  metabolites appear only at low concentrations after 4
days.    In vitro studies, conducted with  liver microsomes
from fish and stomach microsomes from crabs, confirmed the
presence  of a route of  metabolism comparable to that  in
mammals.  Within microsomes, a  cytochrome-P-450-dependent
oxygenase  acts in  the presence  of NADPH  and oxygen  to
allow  progressive  degradation of  TBTO. Such biochemical
mechanisms are apparent in a number of  species.  However,

their  activity  is  limited in  molluscs, particularly in
bivalves;  thus  the  capacity of  molluscs  to metabolize
xenobiotics is generally weak.  Tsuda et al.  (1988)  fol-
lowed  the  metabolism  of tributyltin  oxide  in  various
tissues  of  the  carp  Cyrpinus carpio over  14 days.   In
muscle,  there  was  little evidence  of  metabolites  and
almost all of the tin present was in the form of tributyl-
tin. In the kidney, liver, and gall bladder, large amounts
of  monobutyltin  were  evident.   Little  dibutyltin  was
present  in any of  the tissues, suggesting  that  further
metabolism of the intermediate to the monobutyl  form  was
rapid.  Ebdon et al. (1989) could not positively determine
whether  dibutyltin and monobutyltin present  in adult and
seed oysters in British estuaries derived from  intake  of
the  metabolites  or  from metabolism  within the oysters.
However,  they observed that  peak seasonal levels  of the
metabolites  occurred approximately 1 month after peaks of
tributyltin.   They  concluded that  metabolism within the
oysters  was responsible for the DBT and MBT present. This
also suggests that metabolism in oysters is slow.

6.3.  General mechanisms of toxicity of TBTO

    Different  mechanisms of action have  been advanced to
explain  the biological effects and toxicity of TBTO. Some
of  the mechanisms are  present in all  living  organisms,
others only in certain species.

6.3.1.  General toxic mechanisms

    Several  studies  (Aldridge, 1958;  Aldridge & Street,
1964,  1970) have demonstrated  that the trialkyl  deriva-
tives  of  tin,  and notably  tributyltin  compounds,  are
inhibitors  of  oxidative phosphorylation  in mitochondria
and  are,  therefore,  responsible for  inhibiting  energy
transfer. This inhibition results from various phenomena:

*   disturbance of synthesis of ATP;
*   action on mitochondrial membranes causing swelling and
*   alteration in ion transport across lipid membranes.

    Rosenberg  et al. (1980,  1981, 1984) and  Rosenberg &
Drummond (1983) showed TBTO inhibition of cytochrome P-450
activity  in  cells  from various  tissues (liver, kidney,
small  intestine mucosa) after dosing  in vitro or  in vivo.
Evans   et al. (1979) demonstrated inhibition of oxidative
phosphorylation due to formation of complexes between tri-
alkyltin derivatives and proteins or certain  alpha or  beta
amino   acids.  They most notably form chemical links with
nitrogen and sulfur atoms in protein chains (see chapter 2).

6.3.2.  Toxic mechanisms in bivalve molluscs

    In bivalve molluscs, notably in oysters, one sublethal
effect  of TBTO involves  abnormal calcification. This  is

shown  particularly  in  Crassostrea gigas, the  Pacific or
Eastern  oyster,  in  areas contaminated  with  TBTO.  The
effect  is reproducible in experiments  where healthy oys-
ters  are  transferred  to contaminated  areas,  and  also
reversible in transfers from contaminated to clean areas.

    The  abnormal calcification leads to distortion of the
shells;  layers are formed successively of calcium carbon-
ate,  flaking and open  space (Alzieu et  al., 1982),  and
result  partly  from  interference with  synthesis  of the
organic matter (gel) (which allows calcium deposition) and
partly  from interference with crystallization  of calcium
carbonate.  Krampitz et al. (1976, 1983) showed  that  the
protein  constituents  of the  interlamellar gel assisting
deposition  of calcium were  deficient in the  amino acids
necessary  for calcium fixation (serine, alanine, glycine,
glutamic  acid, aspartic acid); these amino acids are com-
plexed by TBTO.

    The work on mammals showing effects of TBTO on oxidat-
ive  phosphorylation  (Aldridge  & Street,  1971) suggests
another possible effect, since ATP plays an important role
in the crystallization of calcium carbonate, as  shown  in
the following schematic diagram.




     Tributyltin  is toxic to microorganisms and is used commer-
 cially  to control bacteria and fungi. Concentrations producing
 toxic effects are very variable between species.   The  primary
 productivity  of a natural  community of freshwater  microalgae
 was reduced by 50% at a TBTO concentration of 3 µg/litre.   Re-
 cently  established  no-observed-effect-levels for  two species
 are  18 and 32 µg/litre.   Toxicity to sea-water microorganisms
 is  similarly variable between species and between studies; no-
 observed-effect-levels are difficult to set but lie below 0.1 µg
 per litre for some species. Most toxicity tests  with  microor-
 ganisms use batch cultures, and in these systems concentrations
 of  TBT in solution may decline rapidly. Such tests may, there-
 fore, underestimate true toxicity.

7.1.  Bacteria and fungi

    Bokranz & Plum (1975) presented data on the effective-
ness  of TBT compounds against bacteria, algae, and fungi.
TBT  is  more toxic  to  gram-positive bacteria  (such  as
 Staphylococcus   aureus with a minimal  inhibitory concen-
tration of between 0.2 and 0.8 mg/litre in  culture)  than
to  gram-negative bacteria (such  as  Escherichia coli with
an MIC of 3.1 mg/litre) in serial dilution tests for vari-
ous  TBT compounds.  MIC values for four species of fungus
in    culture  (Botrytis,   Penicillium,   Aspergillus, and
 Rhizopodium) varied   from  0.5  to 1.0 mg/litre  for  TBT
acetate.  Impregnation of textiles with TBTO, TBT sulfide,
or  TBT fluoride at  0.01%, 0.05%, and  0.2% showed  clear
inhibition  of fungal growth on agar nutrient medium.  The
TBT compounds were resistant to leaching from the textiles
when  washed  before culture  at  the two  highest concen-
trations.  Fungicidal action was also retained in textiles
treated  with  TBTO and  subsequently  buried in  soil. In
tests  with three species  of fungus, limiting  values for
complete  inhibition of effects  on wood were  determined.
Values for TBTO ranged from 0.058 to 0.704 kg/m3   without
prior  leaching of the compound and from 0.055 to 2.178 kg
per  m3   after prior  leaching. For TBT  fluoride, values
ranged  from 0.055 to 0.88 kg/m3    without prior leaching
and from 0.135 to 0.886 kg/m3 after leaching.

    Soracco  & Pope (1983) investigated the action of TBTO
on various physiological and biochemical activities of the
bacterium  Legionella   pneumophila, the  organism  causing
Legionnaires disease, which commonly lives in the water of
cooling  systems. The minimal concentration of TBTO having
any  effect on  Legionella growth was  about 0.02 mg/litre.
At  concentrations between 0.5 and  1.1 mg/litre, TBTO re-
duced  the growth rate initially and subsequently caused a
further  reduction in growth rate. At 1.1 mg/litre, growth
was almost static, while at higher concentrations TBTO was

bactericidal,  causing a reduction in  the optical density
of  the cultures. The effect  of TBTO on the  cultures was
dependent  on cell density; its  effectiveness was reduced
at  high cell densities.  Between 69% and 88% of the added
TBTO was found to be associated with the cells rather than
in  free solution. There was  a dose-response relationship
between TBTO concentration per unit biomass and the effect
on   growth.   The  bacteriostatic  concentration  of  1.1
mg/litre  did not kill the  cells; transfer of cells  from
this culture to fresh nutrient medium established that all
cells  were still viable. A concentration of 2.24 mg/litre
was  similarly lacking in  bactericidal action and  only a
dose  of 11.2 mg/litre successfully killed  cultures.  The
most  marked and immediate  effect of TBTO  was on  intra-
cellular  ATP levels and  on "energy charge"  (the ratio
between  ATP and AMP in the cell). Three concentrations of
TBTO  (0.112,  1.12,  and 11.2 mg/litre)  were  tested and
produced  reductions of intracellular ATP to 45%, 18%, and
15%,  respectively, each within  1 min of addition  of the
TBTO.  The effect persisted at the lowest  exposure  level
for at least 3 h. Dramatic and immediate falls  were  also
seen in energy charge. The authors consider this to be the
major  effect of the  TBTO.  Concomitant falls  in nucleic
acid  synthesis, synthesis of macromolecules, and  CO2 pro-
duction were assumed to follow from the basic action.  The
wide range of concentrations producing graded growth inhi-
bition,  compared to the very small additional increase in
exposure  required to cause  cell death, suggested  to the
authors  that there were  two separate mechanisms  for the
growth inhibition and lethality of TBTO.

    Argaman et al. (1984) investigated the toxic effect of
TBTO  on activated sludge from  municipal sewage treatment
plants. Sludge challenged with a single dose of  TBTO  was
inhibited   (Warburg   respirometer   oxygen   consumption
measurements)  by concentrations of 25 µg/litre   or more.
However,  sludge pre-treated with TBTO at levels of 200 or
1000 µg/litre  adapted to the TBTO and no effect was found
on the ability of sludge organisms to break  down  organic

7.2.  Freshwater algae

    The  MIC  for  cultures of  the  green  alga  Chlorella
 pyrenoidosa with  TBTO  was 0.5 mg/litre  (Bokranz & Plum,
1975).  Floch et al. (1964) reported a no-observed-effect-
level on the growth of a freshwater green alga (desmid) of
0.25 mg  TBTO/litre or 0.15 mg  TBT acetate/litre over  an
exposure  period of 10 days. No growth occurred at concen-
trations of 0.5 mg TBTO/litre or 0.3 mg TBT acetate/litre.
Deschiens  &  Floch  (1968) reported  an  LC100 value  for
 Chlorella over 10 to 20 days of 0.5 mg TBTO/litre.

    More  recent studies have suggested that aquatic algae
are  much more sensitive to TBT than earlier reports indi-

    Wong  et  al.  (1982) determined  the  IC50   (concen-
tration  required to produce a 50% inhibition) for primary
productivity  (uptake  of  14C-labelled    carbonate)  and
reproduction of pure cultures of algae and for  a  natural
phytoplankton  community  from  Lake Ontario,  Canada. The
natural  community was the most sensitive to TBTO, with an
IC50 for  primary productivity of 3 µg/litre.     Ankistro-
 desmus  falcatus showed similar patterns for the effect of
TBTO on primary productivity and on reproduction (growth),
though  the  latter was  slightly  more sensitive  with an
IC50    of 5 µg/litre   compared to 20 µg/litre   for pro-
ductivity.  The  green alga  Scenedesmus  quadricaudata and
the  cyanobacterium (blue-green alga)  Anabaena  flos-aquae
gave   IC50   values for  primary productivity of  16  and
13 µg/litre,    respectively.   RIVM (1989)  reported 96-h
EC50   values for  Chlorella and  Scenedesmus  pannonicus of
42  and 64 µg   TBTO/litre, respectively, and no-observed-
effect-levels of 18 and 32 µg/litre, respectively.

7.3.  Estuarine and marine algae

    Many  diatoms are highly  resistant to the  effects of
organometallic compounds. Thomas & Robinson (1987) studied
the  tolerance of the diatom  Amphora  coffeaeformis to TBT
fluoride  and found that its growth was unaffected at con-
centrations  of less than 10-7mol/litre   when the initial
culture cell density was 10 x 104 cells/ml.  At the end of
the incubation, when the diatom had stopped growing, there
was no nitrate left in the medium. There was a significant
effect of TBT fluoride at 10-7mol/litre   on  growth  when
the  diatom was grown in nitrate-deficient medium.  Growth
was  also affected, to  a lesser degree,  by reducing  the
silicate  in the medium in  the presence of TBT  fluoride.
The authors concluded that TBTO tolerance in the diatom is
not due to the exclusion of the organotin but to detoxifi-
cation  mechanisms requiring increased uptake  of nitrate.
Recovery  of the organisms  after 24 h exposure  supported
this  theory.  After short-term exposure to sublethal, but
inhibitory,  concentrations  of TBT  fluoride,  Amphora re-
covered  within 24 h (Thomas & Robinson, 1986).  Uptake of
nitrate was inhibited initially but recovered after 24 h.

    Salazar  (1985) exposed three species of marine phyto-
plankton  (Gymnodinium   splendens,   Dunaliella sp.,   and
 Phaeodactylum  tricornutum) to TBTO concentrations of 1.5,
3,  and 6 µg/litre   for a  period of 72 h. At  the lowest
concentration,  all of the  G. splendens cells  were killed
and  growth of  Dunliella sp. was inhibited.  The growth of
 Dunliella sp.   was  completely  inhibited at  both  3 and
6 µg/litre,    whereas  no  effect  on  the  growth  of  P.
 tricornutum was  observed  at  any  of  the  test  concen-
trations.  Beaumont  &  Newman (1986)  cultured the marine
algae  Pavlova    lutheri,    Dunaliella   tertiolecta, and
 Skeletonema  costatum with TBTO at  0.1, 1.0, and   5.0 µg
per litre. All algae exposed to 5.0 µg/litre   died within
2 days. A comparison of the slope of the growth curve with

maximum  increase in cell  density in the  culture  showed
that all of the algae were significantly inhibited by TBTO
at the lowest concentration tested. This concentration was
not  algicidal. Thain (1983)  gave the algistatic  concen-
tration  for  TBTO  against  Tetraselmis suecica as  560 to
1000 µg/litre   and against  Skeletonema costatum as 1.0 to
18 µg/litre,    within 5 days of  exposure.  Corresponding
algicidal concentrations were > 1000 µg/litre   for  Tetra-
 selmis  and > 18 µg/litre   for  Skeletonema.  Walsh et al.
(1985)  calculated the EC50   values for growth inhibition
of  the  marine alga  Skeletonema  costatum by TBT acetate,
TBTO,  TBT chloride, and  TBT fluoride to  be 0.36,  0.33,
0.36,  and 0.25-0.50 µg/litre,   respectively, for  a 72-h
exposure  period. The EC50   values  for growth inhibition
of  Thalassiosira  pseudonanna, another marine species,  by
TBT  acetate and TBTO were  1.28 and 1.03 µg   per  litre,
respectively.   The LC50   for  Skeletonema was 14.7, 14.2,
11.5,  and  11.9 µg/litre    for TBT  acetate,  TBTO,  TBT
chloride,  and TBT fluoride,  respectively. Algae did  not
adapt  to  the presence  of  TBTO after  exposure  through
12 serial  transfers over 12 weeks; EC50   values were the
same  for  previously exposed  cells  as for  naive cells.
Dojmi  Di  Delupis  et  al.  (1987)  calculated  the 8-day
EC50   for growth inhibition of the marine  algal  species
 Dunaliella    tertiolecta and  Nitzschia sp.,   exposed  to
TBTO,  to be 4.53 µg/litre   and 1.19 µg/litre,   respect-

    His et al. (1986) conducted bioassays to  measure  the
susceptibility of algae that are food organisms  for  oys-
ters  to  TBT-containing  antifouling paints.  Four  algal
species  were  used  in  the  studies:  Isochrysis  galbana
(Prymnesiophyceae);  Chaetocerus    calcitrans (Bacillario-
phyceae);  Tetraselmis   (Platymonas)  suecica (Prasinophy-
ceae);  and  Phaeodactylum tricornutum (Bacillariophyceae).
Cultures  were  maintained in  filtered  sea water  with a
salinity  of 27o/oo     and at a temperature of 20 °C. TBT
exposure  was either  to pure  TBT acetate  or  as  plates
painted with "International TBT antifouling" with a sur-
face  area of between 0.01  and 1.0 cm2   in a  culture of
2 litres.  Culture density was estimated every 3 to 4 days
using a Coulter counter.  Exposure to the pure TBT acetate
at  1 µg/litre   had no  effect on any  of the algal  cul-
tures.  Isochrysis growth  was totally inhibited by painted
panels  of  1.0, 0.25,  and  0.125 cm2   within  2 days of
culture. Smaller panels were then used to find  the  limit
of the effect. A 0.02-cm2 panel  was also totally toxic to
growth  of the alga;  panels of 0.01 cm2    allowed growth
comparable  to a control  culture over the  first week  of
culture but then inhibited growth. By the 21st day of cul-
ture, the number of cells was reduced to 1.8 x 106   cells
per ml, compared to a control density of 3.2 x 106   cells
per  ml. For  Chaetocerus, panels of 0.02 cm2    were toxic
to  the alga  from the  beginning of  the culture  period;
numbers  of cells were  reduced, indicating that  not only
growth  but also viability  was affected.  With  panels of

0.01 cm2,    there was complete inhibition  of development
over the first 4 days, followed by a decrease in cell num-
bers  from the original value. The two other algal species
were less sensitive.  Phaeodactylum was inhibited by panels
of 1.0 and 0.5 cm2 from  the outset of culture. Inhibition
also  occurred with panels  of 0.25 and  0.125 cm2,    but
only  after several days.  Tetraselmis grew in the presence
of  panels of all  sizes up to  1.0 cm2,   though at  this
exposure, growth was reduced.  Exposure to panels  of  0.5
cm2 had  little effect. The authors also tested the effect
on  algal growth of fresh water from the river feeding the
area of interest and also the effect of river sediment. In
both cases, growth of the algae was greater than growth of
controls,  suggesting a greater availability of nutrients.
However, it should be noted that no analysis was  made  of
actual  water concentrations of  TBT in this  study; since
the precise exposure levels are unknown, the  results  are
difficult to interpret or evaluate.


8.1.  Aquatic plants


     Few  studies have been carried out on the effects of TBT on
 aquatic  plants.   The lowest  effect  level was  observed  for
Enteromorpha   intestinalis.  Motile spores were  inhibited from
 settling  by TBTO, the  EC50   being 1 ng/litre;  newly settled
 spores  increased in resistance  with time. Results  should  be
 interpreted  with  care  because TBT  concentrations  were  not
 measured and the experimental protocols were incomplete.

     Reduction  in the growth of freshwater species was observed
 at concentrations down to 0.06 mg/litre.

    Davies et al. (1984) studied the effect of various TBT
compounds  on spore development in the marine green macro-
alga,  Enteromorpha  intestinalis. The 5-day EC50    values
for    newly-settled  Enteromorpha spores    ranged    from
0.027 µg/litre  (TBT benzoate) to 8.6 µg/litre  (TBT acry-
late). The authors stated that the toxicity appeared to be
influenced by the type of anion. Using TBTO it  was  found
that  sensitivity  decreased  with  increasing  settlement
time; the 5-day EC50   values for spore development ranged
from  0.22 µg/litre,    when  exposure began  30 min after
settlement,  to  10 µg/litre,    when exposure  began 72 h
after  settlement.  Motile spores were  the most sensitive
(5-day EC50 = 0.001 µg/litre).

    The  marine  angiosperm  Zostera marina showed  reduced
growth  at TBT concentrations in sediment of 1.0 mg/kg but
no  effect  at  0.1 mg/kg (Personal  communication by M.J.
Waldock to IPCS, 1989).

    Floch  et al. (1964) exposed freshwater aquatic plants
to  TBTO or TBT  acetate, at water  concentrations between
0.03  and  1.2 mg/litre,  for 10 days.  The duckweed  Lemna
 media and  Canadian  pondweed  Elodea sp. both  showed some
growth  at TBTO concentrations of 0.03 mg/litre.  Duckweed
maintained  itself, without significant growth, at concen-
trations between 0.06 and 0.25 mg/litre, but died  at  0.5
mg/litre.  Elodea showed degeneration between 0.06 and 0.25
mg/litre  and died at 0.5 mg/litre. Degeneration of  Elodea
was  evident at 0.15 mg TBT acetate/litre; growth occurred
at  0.03 mg/litre and death at 0.3 mg/litre.  Lemna grew in
0.15 mg  TBT acetate/litre and died at 1.2 mg/litre; there
was maintenance without growth at 0.6 mg/litre.

    L.A.  Boorman  (personal  communication to  the  IPCS,
1989)  grew  plants  of  two  salt  marsh   species,  Aster
 tripolium and  Limonium  vulgare, in  mud with  added TBTO.
Plants  of Aster were  killed  by sediment  TBTO levels in
excess  of 10 µg/kg   (dry weight), while  Limonium was not
significantly affected by levels of up to 150 µg/kg.

    Chu  (1976) found that the  aquatic weed  Ceratophyllum
died   within  2 months  exposure to  a controlled release
rubber  formulation  containing 5%  TBTO (5 mg/litre); the
exposure period being 24 h every 3 to 5 days.

8.2.  Aquatic invertebrates

    The  acute toxicity of tributyltin  to aquatic invert-
ebrates  is  summarized in  Tables 8,  9, and  10.  Larval
stages are considerably more sensitive to TBT than adults;
the  LC50 for  the larval  Pacific oyster is  1.6 µg   per
litre, over 48 h, whereas that for adults is 1800 µg   per
litre  (Thain, 1983).  Other species  show similar differ-
ences between life stages. The 96-h LC50 values  for crus-
taceans  range between 1.0  and 41 µg/litre.    There  are
fewer  data on freshwater  species; these relate  to  just
three  species  other  than target  organisms. Various TBT
salts give a range of 48-h LC50 values  for  Daphnia of 2.3
to  70 µg/litre    and  for  Tubifex of 5.5  to 33 µg   per
litre. The 24-h LC50 for  the Asiatic clam is 2100 µg  per
litre, and that for target snail adults in schistosomiasis
control is 30 to 400 µg/litre.

8.2.1.  Trematode parasites of man

    Some  organotin  molluscicides  have been  shown to be
toxic to schistosome larvae in the aquatic stages. Ritchie
et  al. (1974) found  that TBTO concentrations  of 10  and
100 µg/litre   rendered  Schistosoma mansoni cercariae, the
infective  stage released from  the secondary host  (water
snails), incapable of progressive movement, following a 5-
min  exposure.  Infectivity of  the cercariae to  mice was
completely suppressed. A 30-min exposure to concentrations
of  1 µg/litre   or less  had relatively little  effect on
motility  of  cercariae  and on  subsequent infectivity of
mice.  The  authors  also exposed  S.  mansoni miracidia to
TBTO and found that 10 µg/litre  immobilized the miracidia
after  a  40-min  exposure and  completely  suppressed the
infectivity  to snails  (Biomphalaria glabrata). However, a
concentration  of 1 µg/litre   had  no effect on  motility
and infectivity after an exposure period of 120 min.

Table 8.  Toxicity of tributyltin to marine invertebrates
Organism                  Size/   Stat/  Temper-  Salin-  pH    TBT salt   Dura-   LC50c     Reference
                          age     flowa  ature    ity                      tion  (ug/litre)
                                           (°C)     (o/oo)                 (h)
Eastern oyster            embryo  statb  28               7.1   chloride   48      1.3       Roberts 
 (Crassostrea virginica)                                                         (0.78-1.38)d (1987)
                          larva                                            48      3.96             
European oyster           adult   statb                         oxide      48     >300      Thain 
 (Ostrea edulis)                                                            96      210       (1983)

Pacific oyster            larva   statb                         oxide      48      1.6       Thain 
 (Crassostrea gigas)       adult                                            48      1800      (1983)
                          adult                                            96      290

Mussel                    larva   statb                         oxide      48      23        Thain 
 (Mytilus edulis)          adult                                            48      300       (1983)
                          adult                                            96      38

Hard clam                 embryo  statb  28               7.1   chloride   48      1.13      Roberts 
 (Mercenaria mercenaria)                                                         (0.72-1.31)d (1987)
                          larva                                            48      1.65d

Brown shrimp              larva   statb                         oxide      48      6.5       Thain 
 (Crangon crangon)         larva                                            96      1.5       (1983)
                          adult                                            48      73
                          adult                                            96      41

Grass shrimp              sub-    flow   19.4-    9.8-    8.15- oxide      96      20d       Walsh (1986)
 (Palaemonetes pugio)      adult          21.3     12.1    8.31  chloride   96     > 31d,f   Bushong 
                                                                                             et al. (1988)
Mysid shrimp            < 1 day   flow  24-26    19-     7.98- chloride   96      1.1       Goodman et 
 (Mysidopsis bahia)                                22.3    8.01                  (0.68-1.4)d  al. (1988)
                          5 day                                                    2           
                          10 day                                                   2.2 

Shore crab                larva   statb                         oxide      48      110       Thain (1983)
 (Carcinus maenus)                                                          96      10

Table 8.  (contd.)
Organism                  Size/   Stat/  Temper-  Salin-  pH    TBT salt   Dura-   LC50c     Reference
                          age     flowa  ature    ity                      tion  (ug/litre)
                                           (°C)     (o/oo)                 (h)

Harpacticoid copepod      adult   stat   20-22    7       7.8   fluoride   96      2         Linden et 
 (Nitocra spinipes)                                                               (1-2)e      al. (1979)
                                                                oxide      96      2       
Copepod                   sub-    flow   20       10            chloride   72      0.6       Bushong et 
 (Eurytemora affinis)      adult                                                 (0.1-2.0)d,f al. (1987)
                          sub-    flow   20       10            chloride   48      1.4 
                          adult                                                 (0.8-2.3)d,f
                          sub-    stat   19.4-    10.1-   8.17- chloride   48      2.2       Hall et al. 
                          adult          20.3     11.2    8.32                  (0.2-7.3)d,f (1988b)
                          sub-    stat   19.4-    10.1-   8.17- chloride   72      0.6 
                          adult          20.3     11.2    8.32                  (0-3.3)d,f
Copepod                   sub-    flow   20       10            chloride   48      1.1       Bushong et 
 (Acartia tonsa)           adult                                                 (0.7-2.2)d,f al. (1987)
                          adult   statb  19.5-                  oxide      96      1.0       U'Ren (1983)
                                         20.5                                   (0.8-1.2)d
Amphipod                  young   flow   19.4-    9.8-    8.15- chloride   96      1.3d,f    Bushong et 
 (Gammarus sp.)            adult          21.3     12.1    8.31  chloride   96      5.3d,f    al. (1988)
a   stat = static conditions (water unchanged for the duration of the test unless stated otherwise); flow = flow-through conditions (TBT concentration in
    water continuously maintained).
b   Static renewal conditions (water changed periodically).
c   95% confidence limits are given in brackets.    
d   Measured concentration.
e   Nominal concentration.                          
f   Concentration expressed as TBT.

Table 9.  Toxicity of tributyltin to freshwater invertebrates
Organism           Size/     Stat/  Temper-  Hard-    pH   TBT salt  Dura-  LC50d             Reference
                   age       flowa   ature   nessc                   tion  (ug/litre)
                                     (°C)  (mg/litre)                 (h)
Asiatic clam       larva     stat     20                   oxide     24    2100e              Foster 
 (Corbicula fluminea)                                                                          (1981)

Water flea         adult     statb    21                   chloride  96    5.9 (3.7-9.4)e     Meador
 (Daphnia magna)    adult     statb    21                   chloride  120   3.4 (1.3-8.8)e     (1986)
                                                           acetate   48    3.3 (1.5-6.0)e,f   Polster & 
                                                           oleate    48    8.5 (4.2-10.5)e,f  Halacka 
                                                           benzoate  48    4.3 (1.8-9.5)e,f   (1971)
                                                           chloride  48    4.5 (1.2-9.3)e,f
                                                           laurate   48    4.7 (1.2-9.3)e,f
                                                           oxide     48    2.3 (1.2-5.2)e,f
                   juvenile  stat     20     200      7.5  chloride  24    13e                Vighi & 
                   juvenile  stat     20     200      7.5  oxide     24    14e                Calamari 
                   juvenile  stat     19              8.2  oxide     48    4.7                RIVM 
                             stat     20                   oxide     48    70e                Foster 
Tubifex worm                                               acetate   48    8.0 (2.8-10.3)e,f  Polster & 
 (Tubifex tubifex)                                          oleate    48   17.0 (10.1-30.0)e,f Halacka 
                                                           benzoate  48   16.0 (10.1-27.0)e,f (1971)
                                                           chloride  48   15.0 (10.1-23.0)e,f
                                                           laurate   48   33.0 (10.6-75.0)e,f
                                                           oxide     48    5.5 (1.6-10.3)e,f
a   stat = static conditions (water unchanged for the duration of the test unless stated otherwise); 
    flow = flow-through conditions (TBT concentration in water continuously maintained).
b   Static renewal conditions (water changed periodically).
c   Hardness expressed as mg/litre CaCO3).
d   95% confidence limits are given in brackets; concentrations expressed as the TBT salt used 
    unless otherwise stated.
e   Nominal concentration.
f   Concentration expressed as TBT.
    Viyanant  et  al.  (1982) exposed  Schistosoma  mansoni
cercariae   to TBT fluoride  and calculated an  LC50    of
16.8 µg/litre   and an LC90   of 21.7 µg/litre   for a 1-h
exposure.   Following a 1-h exposure  of  S. mansoni to TBT
fluoride,  cercarial infectivity of mice was observed over
a  period of 30 min. A 100% suppression of infectivity was
found  at  6 µg/litre;    effective doses  (99%-100%) were
between 2 and 6 µg/litre.

8.2.2.  Freshwater molluscs


     The  LC50    values for target  fresh water snail adults  in
 schistosomiasis  control range from 30 to 400 µg/litre.    This
 indicates  very low selectivity of TBT as a Bilharzia  mollusci-
 cide, and a high risk to sensitive non-target aquatic species.

     The  lethal concentrations of TBT to adult Bilharzia  snails
 are also expected to inhibit motility and infectivity  of  both
 cercariae and miracidiae in the contaminated water, as the sup-
 pressive  levels were found to  be 10 µg/litre   for both  cer-
 cariae  and  miracidiae  of Schistosoma mansoni  after  5 and 40
 min, respectively. The 1-h LC 50  for  cercariae is 16.8 µg   per

     Generally, the toxicity of TBT to freshwater snails depends
 on  the species, stage, age, temperature, pH, time of exposure,
 time  of observation, suspended  matter, and type  of structure
 and formulation.

     Pellets or matrices of natural rubber or synthetic polymers
 impregnated  with TBT produce a slow-release effective level of
 the  molluscicides,  which results  in  low acute  toxicity but
 extended long-term toxicity.

     Exposure  of adult snails to  TBT levels as low  as 0.01 to
 0.001 µg/litre    reduced egg laying, inhibited hatchability of
 the exposed eggs, and retarded the development of the surviving

     The  no-observed-effect  level for  adult freshwater snails
(Limnaea stagnalis)  was 0.32 µg/litre in long-term tests.

     TBT  compounds are strongly  adsorbed on to  suspended clay
 and  organic particles.  These particulates are ingested by the
 snails and provide one of the inputs for toxicity.  Studies  of
 the impact of type and amount of suspended matter on  TBT  tox-
 icity to snails vary in their conclusions.

     Increasing  the pH was  found to enhance  the molluscicidal
 toxicity of slow-release TBT formulations.

    The acute toxicity of TBT to adult  freshwater  snails
is  summarized in Table 10. The  data are based mainly  on
species that are intermediate hosts in the  life-cycle  of

the  parasite causing Bilharzia (schistosomiasis)  in man,
TBT being used as a molluscicide to control  the  disease.
The  toxic action of TBT is slow, the mortality at the end
of  a 24-h exposure is often low, and for a more realistic
result a post-exposure observation period is required.

    Rubber  impregnated with TBT to produce a slow-release
molluscicide  that  maintains  a low,  but  toxic, concen-
tration over a long period was developed by Cardarelli and
colleagues,  and was shown  to be very  effective  against
snail  pest  species such  as  Biomphalaria glabrata and B.
 globosus (Berrios-Duran & Ritchie, 1968).

    Molluscicidal  activity has been  demonstrated against
 Bulinus spp.,  Biomphalaria spp.,   and  certain operculate
freshwater  molluscs, but organotin compounds  have proved
to  be not as  toxic against the  amphibious  oncomelaniid
snails (McCullough et al., 1980).  Acute toxicity

    Webbe  (1963)  found  that young  snails  (Biomphalaria
 sudanica and  Bulinus  nasutus) were  more  sensitive  than
adults to TBT acetate, the 24-h LC50   values being 14 and
15 µg/litre    for  the  two snail  species, respectively.
When eggs from the same two species were exposed, the 24-h
LC50    values ranged from 100 to 1000 µg/litre.   Paulini
(1964)  found embryos of  Biomphalaria glabrata to  be more
susceptible  than  adults to  TBT  acetate. The  24-h   LC50
values  for embryos ranged from 26 µg/litre   at 5-12 h of
age  to 46 µg/litre   at 77-87 h of age, whereas the value
for adults was 170 µg/litre.

Table 10.  Acute toxicity of tributyltin to freshwater snails
Species          TBT salt  Exposure  Post-exposure    LC50     Reference
                           duration  observation    (µg/litre)
                             (h)         (h)
 Biomphalaria     oxide        6          72          410       Seiffer & Schoof (1967)
  glabrata        acetate      6          72          290       Seiffer & Schoof (1967)
                 oxide        6          24          370       Ritchie et al. (1964)
                 oxide       24          24           40       Ritchie et al. (1964)
                 acetate      6          24          190       Ritchie et al. (1964)
                 acetate     24          24           85       Ritchie et al. (1964)
                 acetate     24                     100-300    Hopf et al. (1967)
                 phenate     24                     100-400    Hopf et al. (1967)
                 oxide       24                      50-100    Hopf et al. (1967)
                 oxide       24                       30       Deschiens et al. (1966)
                 acetate     24          72          170       Paulini (1964)

 Biomphalaria     oxide       24                       30       Deschiens et al. (1966)

 Biomphalaria     acetate     24          48           34       Webbe (1963)

 Bulinus nasutus  acetate     24          48           32       Webbe (1963)

 Bulinus          oxide       17          24           10       de Villiers &
  tropicus                                                      MacKenzie (1963)

 Limnaea          oxide       24          96           60       Temmink & Everts (1987)
  stagnalis       oxide       96          96           24       Temmink & Everts (1987)
                 oxide       96                       42       RIVM (1989)
---------------------------------------------------------------------------------------  Short- and long-term toxicity

    Ritchie  et  al.  (1974)  found  that  egg  laying  in
 Biomphalaria   glabrata was completely inhibited  by   10 µg
TBTO/litre,  all the  snails being  killed within  2 to  5
days.   Egg laying was  reduced, over a  period of 2  to 3
weeks,  by  more than  90% at 1 µg/litre    and by 50%  at
0.1 µg/litre.  At 0.01 µg/litre,   egg laying was unaffec-
ted.   Newly-laid eggs were  exposed to TBTO  for  34 days
followed  by  50 days in clean water. Eggs exposed to 10 µg
per  litre did  not hatch,  even after  50 days  in  clean
water.   Of those exposed to 1 µg/litre,   only 3% hatched
during  the exposure  and 35%  of the  rest hatched  after
transfer  to  clean water  (but  with a  delayed  hatching
time). When newly-hatched snails were exposed to TBTO, 95%
of  those exposed to  1 µg/litre   from hatching  died and
those  that survived failed to  lay eggs for 85 days.   At
0.1 µg/litre,   60% of the snails died and egg  laying  in
the  survivors was reduced  by 80%.  Egg  laying was  also
significantly reduced at both 0.01 and 0.001 µg/litre.

    Upatham  et al. (1980) studied the toxicity to  Bulinus
 abyssinicus of  various controlled-release organotin  mol-
luscicides,  i.e.  BioMet  SRM rubber  pellets  (6% TBTO),
CBL-9B rubber pellets (20% TBT fluoride), and EC-13 float-
ing  ethylene propylene co-polymer pellets (30% TBT fluor-
ide).   The organotin compounds  killed all of  the snails
within 1 to 2 days at 100 mg/litre (active ingredient) and
within  5 to 7 days at 1 mg/litre, there being no signifi-
cant  difference between the compounds.  Changing the test
water  daily had no  significant effect. At  lower concen-
trations  the molluscicides required 36 to 40 days to kill
all  the snails at  an active ingredient  concentration of
0.03 mg/litre,  and at 0.3 mg/litre  9 to 10 days  was re-
quired for BioMet SRM and CBL-9B, and 22 days for EC-13.

    In  long-term exposure tests, toxic  effects on fresh-
water  snails have  been found  at very  low  TBT  concen-
trations.   Cardarelli (1973) quoted a  120-day LC100   of
7 µg  TBTO/litre for  Biomphalaria glabrata. At 0.7 µg  per
litre, 26% of the snails died within 120 days. RIVM (1989)
reported a NOEL for the freshwater snail  Limnaea stagnalis
of 0.32 µg/litre in long-term tests.  Factors affecting toxicity

    Paulini  &  de  Souza  (1970)  studied  the  effect of
various  factors on the  molluscicidal activity of  TBT to
freshly  laid  eggs  of  Biomphalaria  glabrata. A  concen-
tration  of  colloidal clay  (the  proportion of  clay  to
molluscicide was 1000:1) of 1000 mg/litre reduced the 24-h
LC50    (with a 7-day recovery period included in the mor-
tality count) by 30% for TBT acetate and by 7%  for  TBTO.
The  addition  of  a yeast ( Saccharomyces sp.)  suspension
(1 g/litre)  reduced the 24-h LC50   (with a 24-h recovery
period)  by 95% for TBTO and 72% for TBT acetate. The pro-
portion of yeast to molluscicide was 1000:1 for  TBTO  and
100:1 for TBT fluoride.

    Cardarelli  & Evans (1980) studied the effect of vari-
ous  factors  on the  toxicity  to snails  of  controlled-
release  organotin molluscicides, i.e. BioMet SRM (6% TBTO
in natural rubber) and CBL-9B (20% TBT fluoride in natural
rubber). Using 100-mg/kg pellets, they found that increas-
ing the pH from 6 to 8 increased the toxicity (as measured
by  both LT50   and LT100)    of BioMet SRM but  decreased
the  toxicity of CBL-9B to  both  Biomphalaria glabrata and
 Bulinus  globosus. The authors found no effect of yeast or
humic  acid (1 to  100 mg/litre) on the  toxicity  (LT100)
of  these  controlled-release  molluscicides, neither  did
they  find an effect  of suspended colloidal  clays.  They
concluded  that,  although  the  organotin  compounds  are
adsorbed, they are still toxic because the  snails  ingest
the  added materials.  Therefore, snails are still exposed
to TBT even after it is adsorbed to surfaces. In an exper-
iment  to  compare  adsorbed uptake  of TBT, soil-browsing
snails  and isolated snails were  compared during exposure

to  slow-release organotin pellets. The molluscicides were
found  to be more  toxic when the  snails were allowed  to
browse  on soil, and, at a distance of 60 cm from the pel-
lets,  isolated snails were  producing egg masses  whereas
those  browsing on soil  were not. The  authors also  con-
cluded that the organotin molecules that come into contact
with soil particles are adsorbed and slowly  form  ligands
of a nontoxic nature, so that soil only remains  toxic  as
long as it is freshly exposed to organotin. Upatham et al.
(1980)  found that  the presence  of mud  or  plant  life,
compared  to exposure in water alone, had no effect on the
toxicity  (LT100)   of either CBL-9B  or EC-13 at 1  or 10
mg/litre  (active ingredient) to  Bulinus abyssinicus. When
the snails were exposed to BioMet SRM, no effect of mud or
plants  was  found at  10 mg/litre,  but at  1 mg/litre  a
slightly  shorter time was required to kill all the snails
in water alone.  Chu (1976) noted that  organic  materials
such as mud and weeds reduced the  molluscicidal  activity
of  TBTO on  Bulinus rohlfsi, when exposed to a rubber for-
mulation containing 5% TBTO (5 mg/litre TBTO), with a 24-h
exposure  period every 3 to 5 days for 70 days and twice a
month for a further 7 months.

    Macklad  et  al. (1983)  found  that the  toxicity  of
controlled-release  molluscicides containing TBT  fluoride
was dependent on the pre-exposure soaking time.  The  48-h
LC50    for  Biomphalaria alexandrina of 10 mg/litre  total
available  toxicant was achieved  after 7, 3,  and  1 days
soaking period for the formulations EC27 (10% TBT fluoride
in  a plastic polymer), EC1320  (20% TBT fluoride in  rub-
ber),  and EC1330 (30%  TBT fluoride in  a  polypropylene/
polyethylene  mixture),  respectively. EC1330  produced no
mortality following a 3-h soaking period and  exposure  to
concentrations  ranging from 1  to 100 mg/litre for  48-h;
the  authors suggested that the polypropylene/polyethylene
mixture  (EC1330) needs to  be sufficiently wet  before it
begins  to release TBT  fluoride.  A second  experiment to
test  the aging of slow-release  molluscicides was carried
out.  EC27 gave 89% snail mortality during a 48-h exposure
period to 50 mg/litre after a soaking time of 24-h.  After
being  left to age for 1 month the molluscicide was tested
again and had lost 80% of its original toxicity.

8.2.3.  Marine molluscs


     A large body of data exists on the effects of TBT on marine
 molluscs  and in particular on commercially important bivalves.
 Sublethal effects occur at very low concentrations. It has been
 shown experimentally that TBT affects shell deposition of grow-
 ing  oysters, gonad development  and gender of  adult  oysters,
 settlement,  growth,  and mortality  of  larval oysters  and of
 other  bivalves, and causes  imposex (the development  of  male
 characteristics)  in  female  gastropods. The  NOEL  for  shell
 thickening  of the most sensitive  oyster species (C. gigas)  is

 about  20 ng/litre.  Embryo-larval stages  are more susceptible
 than  adults; adverse effects  on larval development  have been
 demonstrated  at concentrations as low as 50 ng/litre. The NOEL
 is 20 ng/litre.

     The  authors  of  recent work  on  imposex  in Nucella  have
 determined  threshold concentrations by extrapolation below the
 limit  of reliable detection. While the circumstantial evidence
 in  support is substantial (see  chapter 10), the determination
 of toxicological thresholds below the limits of chemical detec-
 tion is not practicable.  It is generally agreed  that  imposex
 is not a specific index of TBT contamination, in that it can be
 induced  by other factors. The wide distribution of low concen-
 trations  of TBT and the incidence of imposex at levels similar
 to  or  below the  analytical  detection limits  make long-term
 controlled  experiments difficult.  The establishment  of NOELs
 will have to await the development of better  analytical  tech-
 niques.  Acute toxicity

    Waldock & Thain (1985) calculated the 24-h  EC50 (mor-
tality plus moribundity) of two organotin-containing fish-
net  antifouling  preparations  for larval  oysters  to be
12 µg/litre    for  `Norimp  200' and  320 µg/litre    for
`Flexgard'. These  compare with a value for TBTO of 1.7 µg
per  litre. Thain (1983)  pointed out that  adult bivalves
tend  to appear more  resistant to pollutants  in standard
short-term tests since they can close their shell over the
test period and thus reduce exposure. Larval stages appear
to be much more sensitive in these tests.  Short- and long-term toxicity

    Alzieu et al. (1982) kept Pacific oysters  (Crassostrea
 gigas) in  150-litre  tanks that  were successively filled
and  drained according to the tidal period. To these tanks
were  added panels coated on  one side with TBT  fluoride,
giving estimated concentrations of 0.2 and 2.0 µg  TBT per
litre.  All oysters died within 30 days in a tank contain-
ing  the larger panel. In the tank with a panel surface of
50 cm2,   30% of oysters died after 110 days  of  exposure
and all oysters within 170 days.

    His  & Robert (1985) experimentally  tested hypotheses
regarding the poor performance of Pacific oysters  in  the
Bay  of Arcachon, France.  Larvae were maintained  in  the
laboratory at TBT acetate concentrations ranging from 0.02
to  100 µg/litre.    Growth  was affected  by  all concen-
trations  except  0.02 µg/litre.   The  next concentration
tested  (0.05 µg/litre)   reduced growth, led to mortality
within 10 days, and interrupted normal feeding by  day  8.
No other effects were noted at 0.02 µg/litre  and this was
regarded as the NOEL for larvae (see Table 11).

Table 11.  Effects of TBT acetate on  Crassostrea gigas larvae at
various water concentrationsa
Water concentration               Effect
   100          inhibition of fertilization
    50          inhibition of segmentation
    25          partial inhibition of segmentation (40%)
    10          no formation of trochophores
  3 to 5        no veligers; malformed trochophores
     1          abnormal veligers; total mortality within 6 days
    0.5         numbers of abnormal larvae; total mortality within 8 days;
                perturbation of feeding regime, particularly from 4 to 8 days
                after exposure; growth greatly reduced
    0.2         percentage of D larvae showing abnormalities less elevated;
                perturbation of feeding regime from day 4; progressive
                mortality; total by day 12; weak growth
    0.1         majority of D larvae normal; marked perturbation of feeding
                regime from day 6; weak growth until day 6; some survivors
                after 12 days
   0.05         normal D larvae; perturbation of feeding regime, marked on day
                8; significant mortality beginning at day 10; reduced growth
   0.02         normal D larvae; little mortality; good growth; no effect of
a   From: His & Robert (1985).

    Thain  & Waldock (1985)  exposed various bivalve  spat
(common  European  oyster,  Ostrea edulis; Pacific  oyster,
 Crassostrea    gigas; common  mussel,  Mytilus  edulis; and
carpet   shells,  Venerupis  decussata and  Venerupis  semi-
 decussata) to  TBT leachate by maintaining them in flowing
sea-water  tanks  containing  house  slates  painted  with
`Micron 25R' (containing TBT methacrylate co-polymer). The
water concentrations of TBT were 0.24 or 2.6 µg/litre.  At
2.6 µg/litre,    growth  was  completely inhibited  in all
groups and mortality was high (except for  V. semidecussata)
within  the 45-day exposure period; all mussels  had  died
within  14 days.  At  the  lower  exposure  concentration,
growth was significantly inhibited in  C. gigas, M. edulis,
and  V.  decussata but not in the other two species.  In  a
second study, the authors reported a severe  reduction  in
growth  rate  of  recently  metamorphosed   oyster  (Ostrea
 edulis) spat after exposure to TBT leachate (0.06 µg   per
litre)  for 10 days under static  conditions.  Growth rate
was reduced between 0 and 10 days of exposure  to  0.02 µg
per  litre, but there was only slight reduction in growth,
relative to controls, between days 10 and 20.

    Growth  curves for oysters  (Ostrea  edulis), 2-3 mm in
size,  exposed  to  different concentrations  of  TBTO are
given in Fig. 3.


    Valkirs  et al. (1987) calculated the 66-day LC50   of
TBT  chloride for the mussel  Mytilus  edulis to be 0.97 µg
per  litre under flow-through conditions.  Beaumont & Budd
(1984)  kept larvae of the  common mussel,  Mytilus edulis,
in   filtered sea water containing  TBTO concentrations of
0.1, 1.0, or 10.0 µg/litre.  No larvae survived for longer
than  5 days at 10 µg/litre,  or 10 days at   1.0 µg/litre.
Approximately half of the mussel larvae, at  0.1 µg    per
litre,  had died within 15 days.  The survivors were mori-
bund,  and their growth was significantly slower than that
of controls.

    Laughlin  et  al.  (1987) exposed  the hard-shell clam
 Mercenaria   mercenaria to various concentrations of TBTO,
under  static  renewal  conditions  for  larval  (veliger)
stages  and in flowing sea water for juveniles. They found
the post-larval settlement stages to be the  least  sensi-
tive.  In a 25-day exposure, only juvenile  clams  exposed
at  10 µg/litre    suffered  100% mortality,  while  those
exposed  to  7.5 µg/litre    or less  showed mortality not
significantly  different  from  that  of  controls.   When
veligers were exposed, none survived TBTO levels  of  1 µg
per  litre or more for longer than 7 days, mortality being
100% within 2 days at 2.5, 5.0, and 7.5 µg/litre.   At the
end  of  the 8-day  experiment,  all controls  had  become
pediveligers. Clams exposed to 0.6 µg/litre  showed a sur-
vival level of approximately 40% of the control level, but
survivors  achieved  little  growth and  metamorphosis  to
pediveligers  did  not occur.  In  another set  of studies
(Laughlin  et al., 1987,  1988), clams were  exposed, from
fertilization to metamorphosis (approximately 14 days), to

TBTO  concentrations of between 10 and 500 ng/litre. Clams
were  also exposed for the  first 5 days and then  kept in
clean  water for a further 9 days.  Survival was found not
to  be exposure  dependent and  a recovery  period had  no
effect.   Growth was reduced at all concentrations, higher
exposure  causing greater growth depression.  At TBTO con-
centrations  above 100 ng/litre, veligers failed  to meta-
morphose   to  pediveligers  within  the  14-day  exposure
period. Although the 9-day recovery period caused a slight
increase  in  growth,  the animals  were not significantly
larger than clams exposed continuously.

    Pickwell  &  Steinert  (1988)  exposed  adult  mussels
 (Mytilus  edulis) and oysters  (Crassostrea virginica) in a
flowing sea-water system contaminated with TBT from panels
painted  with antifouling paint. The TBT concentration was
0.7 µg/litre   and exposure lasted for 60 days. Haemolymph
was  collected from the  exposed animals and  its  protein
content measured. Mussel haemolymph protein content at the
end of the exposure period was 462 mg/litre, compared with
44 mg/litre  in controls. Measurement of  haemolymph lyso-
zyme  activity  and  DNA content  revealed  no  difference
between  controls  and  treated mussels.  This showed that
there  had been no lysis of the haemocytes and no increase
in their numbers. Protein was not, therefore, derived from
blood  cells.   Oysters showed  no  similar effect  of TBT
exposure,  although  haemolymph  protein content  was much
higher than in mussels. Over the course of the experiment,
there  was approximately 50%  mortality in mussels  but no
deaths among the oysters.  Reproductive effects

    Thain & Waldock (1986) reported the results of studies
on  the reproduction of  the European flat  oyster  (Ostrea
 edulis) exposed  to TBT antifouling paints.  Three holding
tanks  were  set up  each  with 50 adult  oysters weighing
between  50 and 70 g. One tank held controls and the other
two  were exposed to TBT leaching from painted panels in a
mixing tank. A flow rate of 1 litre/min was maintained and
the  two treatment tanks showed TBT concentrations of 0.24
and  2.6 µg/litre   measured at the  outflow. Only control
oysters  released larvae during  the course of  the 75-day
experiment;   about  five  million  larvae  were  released
representing  probably between four and  six spawnings. At
the end of the experiment, the gonads of oysters from each
treatment  were  examined  histologically. There  were  no
females  in either of the  treated groups of oysters  (20%
females in the control group). At 2.6 µg/litre,    18  out
of 25 oysters examined were undifferentiated, 7 were male,
and none were female. There were 3 undifferentiated gonads
out  of 27 at the lower exposure level.  Gonadal thickness
was reduced in a dose-related manner.  Mortality  was  low
in all groups (0 in controls and 3 and 5 in the two treat-
ment  groups). Shell growth  was reduced by  TBT exposure;
25 oysters  showed  growth in  the  control group,  18  at

0.24 µg/litre,    and 4 at 2.62 µg/litre).   There were no
significant  differences  between  the various  groups  of
oysters using two measures of condition (dry  meat  weight
compared with internal shell volume or compared  with  wet
meat  weight).  Final body burdens of TBT were 0.19, 0.40,
and  1.23 mg/kg wet meat weight for control, low-dose, and
high-dose groups, respectively.

    Roberts   et  al.  (1987)  maintained   adult  oysters
 (Crassostrea  virginica) in TBT solutions containing 0.05,
0.1, 0.5, or 1.0 µg/litre   for up to 8 weeks. The oysters
were  brought  into  reproductive condition  by increasing
water  temperature.  There were  no deaths except  at  the
highest exposure concentration, where 20% to 30% mortality
occurred  between the second and fourth weeks of exposure.
Gametes  stripped from the exposed oysters were fertilized
by  gametes from a reference population. There was no evi-
dence that TBT exposure had any effect on the  ability  of
gametes to be fertilized, and there was  no  statistically
significant effect of treatment on the gender  of  exposed
oysters over the 8-week exposure period.  Effects on growth

    Waldock  & Thain (1983) maintained spat of the Pacific
oyster  (Crassostrea gigas) in experimental tanks, contain-
ing  either TBTO, TBT and  marine sediment, or just  sedi-
ment, for 56 days. Weekly growth measurements (measured as
wet weight) showed enhanced weight gain in oysters exposed
to sediment alone (50 or 100 mg/litre). Low levels of TBTO
(0.15 µg/litre)    inhibited growth and  showed pronounced
thickening  of the upper  valve, and severe  inhibition of
growth was noted at 1.6 µg/litre.    Addition of sediment,
along with the TBT, slightly reduced the adverse effect on
oyster  growth. This experiment  was conducted to  counter
arguments  that  sediment  caused the  effects observed in
oysters in the field.

    Thain   (1986)  exposed  Eastern   oyster  (Crassostrea
 virginica) spat  to TBTO concentrations  of 0.02, 0.2,  or
2.0 µg/litre    under static renewal test conditions for 5
weeks. The percentage increase in growth was substantially
reduced at 2.0 µg/litre,  whereas at the other two concen-
trations growth rate was similar to that of  controls.  No
deaths  occurred and there was no evidence of shell thick-
ening or deformity in any of the treated animals.

    Lawler & Aldrich (1987) exposed Pacific oyster spat to
TBTO  concentrations of 0.01,  0.02, 0.05, 0.1,  or   0.2 µg
per  litre and monitored the  average rate of oxygen  con-
sumption before and after exposure. A significant negative
correlation  was found between TBTO concentration and oxy-
gen consumption. There was also a significant relationship
between TBTO concentration in the water and feeding rates.
Feeding   rates  were  measured  by  transmittance,  which
increases as particulate food matter is removed  from  the

water.  Change in transmittance was monitored after 1 h of
feeding. As the TBTO concentration in the water increased,
feeding  rate  decreased.  Neither oxygen  consumption nor
feeding  was  significantly affected  below   0.05 µg/litre.
Increasing the TBTO level also progressively decreased the
ability  of the oysters  to compensate for  hypoxia;  this
effect was significant down to 0.01 µg/litre.  Growth rate
(measured as average increase in valve length)  was  moni-
tored  over a 48-day exposure period. There was a signifi-
cant  negative correlation between increasing TBTO concen-
tration  and growth. Growth was not significantly affected
at  the lowest concentration of  0.01 µg   TBTO/litre.  At
levels  of 0.05 µg/litre   or more, there was an increased
incidence of shell thickening.  However, reservations must
be  expressed  regarding  the experimental  design of this
work.  The authors did  not analyse TBT  concentrations in
the  test solutions. In addition, no analyses were carried
out  on the dilution water,  which may have been  contami-
nated by ambient TBT, typically in excess  of  10 ng/litre
(the lowest threshold determined).  Furthermore, the ratio
of  biomass  to water  was such that  TBT would have  been
rapidly  removed  from  solution, so  that nominal concen-
trations  would have been maintained for only a few hours.
Although  experimental solutions were replaced  daily, the
rate of loss invalidates the threshold values cited. Simi-
lar criticisms can be made of other work using  static  or
static renewal systems.

    Valkirs  et  al. (1987)  exposed  adults of  both  the
common    mussel  (Mytilus   edulis) and   Eastern   oyster
 (Crassostrea   virginica) to  TBT concentrations  of 0.04,
0.13, 0.31, 0.73, or 1.89 µg/litre,   for a 66-day period,
under  flowing sea-water conditions. The  TBT consisted of
leachate  from  plastic  panels painted  with  antifouling
paint. Growth effects were measured by shell length, shell
width, and whole body wet weight (soft tissues and shell).
Since  there was high mussel  mortality at a TBT  level of
1.89 µg/litre,    growth effects were examined  in animals
exposed to TBT concentrations of up to 0.73 µg/litre.   No
significant  effect on mussel  shell width or  whole  body
weight  was  found, but  a  significant decrease  in shell
length was observed at 0.31 and 0.73 µg/litre.   More than
90% of all oysters tested within each  concentration  sur-
vived  the test period.   Statistical analysis of  length,
width,  and weight  of oysters  could not  be carried  out
since  controls were significantly different  from exposed
groups with respect to initial length of  individuals.   A
condition  index  (ratio of  wet  body weight  to internal
shell volume) was calculated for both mussel  and  oyster.
No  significant difference was  found at any  test concen-
tration for mussels. For oysters, the mean condition indi-
ces were significantly lower at concentrations of 0.73 and
1.89 µg/litre,    compared both with the  indices at lower
TBT concentrations and with controls.

    Salazar  &  Salazar  (1987)  exposed  juvenile  common
mussels  (Mytilus  edulis), in  flowing  sea water,  to TBT
concentrations  of 70, 80,  or 200 ng/litre in  a  196-day
test and 40, 50, or 160 ng/litre in a 56-day test.  Mussel
growth  (measured as wet weight  and length) was not  sig-
nificantly  affected up to 56 days in either test. From 63
days to the end of the experiment, there was a significant
reduction  in  growth  at all  exposure concentrations. No
significant mortality was reported in either experiment. A
group  of  controls  kept under  "field"  conditions had
growth  rates four times  that of the  laboratory controls
over  a 56-day period, suggesting that bioassay conditions
were stressful for the mussels.

    Stromgren  &  Bongard (1987)  exposed juvenile mussels
 (Mytilus    edulis) to  TBTO  concentrations  of   0.1  to
10 µg/litre    in  flowing  sea water  and  measured shell
growth  (length) at intervals of 24 to 48 h for 7 days. No
effect  was observed at the lowest exposure concentration,
but  at  0.4 µg/litre   or  more  there was  a significant
reduction  in shell growth rate.  The relationship between
TBTO  concentration and growth response  was approximately
hyperbolic.  After 7 days of exposure,  all groups treated
with 0.4 µg/litre  or more showed growth rates of approxi-
mately  25% (or less)  of the control  value. The  highest
exposure  concentration reduced growth to approximately 5%
of the control value.  Shell thickening

    Alzieu  et  al.  (1982) reported  that  adult  oysters
 (Crassostrea   gigas) developed  gel centres  in the shell
when  they were exposed to TBT fluoride at a concentration
of 0.2 µg/litre.

    Thain et al. (1987) exposed spat of the Pacific oyster
 (Crassostrea   gigas) to  TBT  concentrations of  2 to 200
ng/litre  for 49 days. No  effect on  shell  thickness was
observed in controls or at 2 ng TBT/litre.  Between 20 and
200 ng/litre,  there was a dose-related  increase in shell
thickness. Severe "balling" occurred at both 100 and 200
ng/litre (where the overall appearance of the oyster shell
is spherical rather than having the flattened  profile  of
one valve of a normal oyster) (Fig. 4).  Imposex

    The  phenomenon of "imposex"  was first observed  in
the  field  (see section 10.2),  the  term being  used  to
describe the development of male characteristics by female
gastropods.   The females develop  a penis and  ultimately
become infertile. Stages in the development of imposex are
illustrated in Fig. 5.

    Smith  (1981a) collected female  mud snails  (Nassarius
 obsoletus) from  three  localities  designated  "dirty",
"intermediate", and "clean" on the basis of the degree
of  imposex noted in the field (see section 10.2). A piece
of  filter paper with  1.5 to 1.8 g  of dried  antifouling
paint  containing  TBT  and lead  arsenate (Alumacide) was
placed in 110-litre tanks, and the snails were exposed for
75 days.   Halfway through the exposure period, the filter
paper  was removed because the snails became lethargic. At
the  end of the exposure period, all snails exposed to the
Alumacide  had developed significantly more  intense impo-
sex.  Snails from the "clean" area showed an increase in
imposex  incidence  from 0%  to  14.3%, whereas  levels of
imposex  remained  constant  in snails  from  "dirty" or
"intermediate" areas (> 95% incidence). Penis expression
in   snails  from  "dirty"  and  "intermediate"  areas
regressed  significantly  when  they were  transferred  to
clean  water.  However, the  extent to which  the  effects
observed  in  this  study can  be  attributed  to TBTO  is
unclear,  since the antifoulant contained two biocides and
no analytical measurements were made.


    Feral  & Le Gall  (1983) attempted to  identify  which
part of the neuroendocrine system of a  marine  gastropod,
 Ocenebra    erinacea, was   primarily   affected  in   the
induction, by TBT, of the development of a penis in female
snails.  A biological assay was established using isolated

female  pedal ganglia or  complete nervous system  (intact
complex  of pedal ganglia  and cerebropleural ganglia)  of
 Ocenebra   erinacea and isolated presumptive penis-forming
areas  of a second species  Crepidula fornicata. When pedal
ganglia  were cultured with the  presumptive penis-forming
area in a medium based on either clean or "polluted" sea
water  (from areas known to have the imposex phenomenon in
the field) there was no penis development.  Culturing  the
whole  nervous system in a medium based on clean sea water
also  resulted in no penis growth.  However, culturing the
complete  nervous  system with  a  medium based  on "pol-
luted" sea water caused the growth of a penis.  Culturing
in artificial sea water with added TBT (0.2 µg  per litre)
also induced penis development. The authors concluded that
the primary effect of TBT is on the cerebropleural ganglia
of the snails (Fig. 6).

    In  studies  by  Bryan  et  al.  (1986),  the dogwhelk
 Nucella   lapillus was  exposed  to TBT  concentrations of
0.02 µg    tin/litre in tidal tanks, the TBT being leached
from  a  co-polymer  antifouling paint.   Within 4 months,
animals  of  both sexes  had  accumulated 1 mg  tin/kg (as
TBT), and females showed a high degree of  imposex,  which
was still increasing. The authors stated that  the  exper-
iment tended to underestimate the exposure, since the diet
of  barnacles was initially uncontaminated but would later
have  contributed TBT to the dogwhelks via an extra route.
Gibbs et al. (1987) found that after 12 months of exposure
to 18.7 ng tin/litre, penis size in female  dogwhelks  was
increased,  and that there  was very little  difference in
size  between the sexes. The "control" dogwhelks in this
study  were actually exposed to TBT at 1.5 ng/litre, which
was  the background concentration  in the sea  water used.
These  control females showed  a penis bulk  of 10-14%  of
that  of control males.  Field observations on populations
living  in < 0.5 ng/litre showed little  penis development
(between  2% and 5% of male penis bulk).  The NOEL for the
development  of  imposex  is, therefore,  less than 1.0 ng

    Gibbs et al. (1988) reared dogwhelks in the laboratory
for 2 years from hatching in various concentrations of TBT
leached from antifouling paints. TBT concentrations in the
water were monitored at 1-2, 3-5, 20, and 100 ng tin/litre
and   tissue  concentrations  in  the   whelks  were  also
measured. All exposed females were affected at all concen-
trations, developing a penis. Penis development (expressed
as  a relative size index compared to males exposed to the
same TBT concentrations) in females was between 50 and 60%
of male penis bulk after exposure for 1 and 2 years to TBT
at a level of 1-2 ng tin/litre. The female  penis  reached
comparable  size to that  of the male  on exposure to  TBT
levels  of  3-5 ng tin/litre  or  more. At  increasing TBT
exposure   concentrations,  further  male  characteristics
developed   and   further   female  characteristics   were

repressed.  At TBT concentrations in  the water of 1-2  ng
tin/litre,  some  females  retained the  capacity to breed
although  others were sterilized by  oviduct blockage.  At
3-5 ng/litre,  virtually  all females  were sterilized but
oogenesis was apparently normal. At 10 ng/litre, oogenesis
was suppressed, oocytes were resorbed, and spermatogenesis
was  initiated.  At  20 ng/litre, there  was  a functional
testis  in the "females", with  ripe sperm in the  most-
affected animals.


    Bryan et al. (1988) investigated the capacity of vari-
ous organotin compounds to induce imposex in the dogwhelk.
Whelks,  already slightly affected by  imposex, were taken
from the wild and exposed to 200 ng/litre tin, in the form
of  TBT chloride, tri- n- propyl   tin (TPrT), tetrabutyltin
(TTBT),  dibutyltin (DBT), or triphenyltin  (TPhT), for 14
days  before being returned to the shore. Penis size, as a
percentage of male penis size, increased to 44% in females
exposed  to TBT chloride,  compared with 6%  in  controls.
TPrT  increased relative penis  size to only  14%, and  no
other  compound had any effect. In an attempt to eliminate
differences  due to differential uptake  of the compounds,

the organotin compounds were injected in a  second  exper-
iment.   The females were maintained in the laboratory for
up to 105 days. TBT again induced increased penis size but
TTBT  also showed  an effect  (19.5% compared  to the  34%
shown  after TBT injection). Other compounds were ineffec-
tive. The authors believed the effect of TTBT to be caused
by  contamination by TBT and conversion to TBT in the tis-
sues.  However, the increased imposex after treatment with
TPrT  could not be explained  in this way and  the imposex
effect  is considered to be  not totally specific to  tri-
butyltin.  Genotoxicity

    Dixon & Prosser (1986) found that TBTO was  not  geno-
toxic, at concentrations of 0.05 to 5.0 µg   tin/litre, to
the  larvae  of  the mussel  Mytilus  edulis. Results  were
based   on   chromosome  analysis   and  sister  chromatid
exchange  (SCE). In 4-day acute toxicity studies, TBTO was
found to cause a dose-dependent reduction in  both  larval
survival and development of mussels.  Survival ranged from
46%  of controls at 0.05 µg/litre  to 1.4% at  5 µg/litre.
The  percentage of animals  reaching the D-shell  stage of
development was 8.7% of controls at the lowest  dose,  but
none  reached this stage at either 1 or 5 µg/litre.   How-
ever,  when  mussel  larvae  were  exposed  to  a standard
mutagen  (mitomycin C) or crude  oil, in the  presence  of
TBTO,  the SCE frequency increased  to approximately twice
that  found when larvae were exposed to either toxicant in
isolation (Dixon & McFadzen, 1987).

8.2.4.  Crustaceans


     Most  of the available  data for freshwater  organisms  are
 derived  from acute exposure tests with Daphnia. LC 50  data  for
 similar exposure times are variable (up to 2 orders  of  magni-
 tude)  probably  due  to  age  differences  in  the populations
 studied. The NOEL for Daphnia  has been estimated to  be  around
 0.5 µg/litre, behaviour being the most sensitive parameter.

     A range of marine species from copepods to crabs  and  lob-
 sters  has been studied,  the lowest NOEL  being 0.09 µg    per
 litre for reproductive effects in mysid shrimp.

     The  presence of sediment  in test aquaria  greatly reduces
 the toxic effect of TBTO on estuarine crustaceans.  Acute effects

    Using TBTO and TBT acetate, Floch et al. (1964) calcu-
lated LC100 and  LC0 values  for various species of fresh-
water invertebrates.  The lethal concentrations of TBTO to
 Daphnia   magna over  24  and 72 h  were 0.12 mg/litre and

0.06 mg/litre,  respectively.  Another aquatic crustacean,
 Cypridopsis  hartwigi, was less sensitive with lethal con-
centrations  of 4 mg/litre for a 24-h exposure, 2 mg/litre
for   a  48-h  exposure,  and  0.12 mg/litre  for  a  96-h
exposure.  NOELs  were 0.03  and 0.06 mg/litre for  Daphnia
and  Cypridopsis, respectively.    For  TBT  acetate,   the
LC100    was 0.15 mg/litre for a  72-h exposure of  Daphnia
and  0.15 mg/litre for a 96-h exposure of  Cypridopsis. The
LC0 was 0.075 mg/litre for both species.

    However, more recent work has found  Daphnia to be con-
siderably  more sensitive to TBT. Meador (1986) reported a
96-h  LC50   for  Daphnia magna of 5.9 µg/litre   and noted
that   TBT  is  a  slow-acting  toxicant  for  Daphnia with
effects only being shown after 96 h or more  of  exposure.
Polster  & Halacka (1971)  quoted 48-h LC50    values  for
 Daphnia   magna using various TBT salts, which ranged from
2.2 µg TBTO/litre to 8.5 µg TBT oleate/litre.

    RIVM  (1989)  reported  a 48-h  EC50    for  Daphnia of
4.7 µg/litre   and a NOEL, over the same time  period,  of
0.56 µg/litre.

    Davidson et al. (1986) calculated the 96-h  LC50    to
be   0.42 µg/litre    after  exposing  the   mysid  shrimp
 Acanthomysis sculpta to a leachate of TBT.

    When   Walsh  (1986)  exposed  the  mole  crab  Emerita
 talpoida to  concentrations  of 10 µg    TBTO/litre of sea
water  or 4500 µg/kg   of sand, no effect on crab survival
was  observed after 7 days of exposure. In continuous-flow
bioassays,  10 000 µg   TBTO/kg of  sediment did not  kill
grass shrimp after a 96-h exposure.  Short- and long-term toxicity

    U'Ren  (1983)  maintained  the marine  copepod  Acartia
 tonsa in  solutions  containing  TBTO, under  static  con-
ditions,  and calculated a 144-day LC50   of 0.55 µg   per
litre;  by  combining  moribundity and  mortality  as end-
points, the 144-day EC50 was found to be 0.4 µg/litre.

    Laughlin  et  al.  (1983) maintained  mud  crab larvae
 (Rhithropanopeus   harrisii), from hatching, in  solutions
containing  either  TBTO at  0.5  to 25 µg/litre    or TBT
sulfide at 0.5 to 50 µg/litre,   under static renewal pro-
cedures.   The survival of the  zoeae, up to 15 days,  was
unaffected  by TBTO at levels up to 10 µg/litre.  At   15 µg
per litre, 84% successfully moulted to the megalopa stage,
but  at 5 µg   TBTO/litre  only 37% survived.  Zoeal  sur-
vival was unaffected by concentrations of TBT  sulfide  up
to 5 µg/litre.    Survival at 20, 30, and 50 µg/litre  was
78%, 26%, and 4%, respectively. The development rate, over
the same time period, decreased with increasing  TBT  con-
centration,  although this was not  statistically signifi-
cant below 10 µg  TBTO/litre or 20 µg   TBT sulfide/litre.

At the highest exposure concentrations (10 µg   TBTO/litre
and  20 µg   TBT sulfide/litre), metamorphosis was delayed
by approximately 2 days in the case of TBTO and 6 days for
TBT  sulfide.  Growth (measured  as mean wet  weight)  was
significantly  reduced  at  TBTO concentrations  of 15 and
25 µg/litre   or TBT sulfide concentrations of 20, 30, and
50 µg/litre,  and showed dose dependency. Daily growth was
monitored  for up to  12 days at concentrations  (of  both
TBTO and TBT sulfide) of 0.5, 1.0, and 5.0 µg   per litre.
Although  the final weights were not significantly differ-
ent,  all  TBT treatments  caused  an initial  growth  lag
during the first three days of exposure. Laughlin  et  al.
(1985)  found  the  LC50    for  exposure  of   zoeae   of
 Rhithropanopeus   harrisii to  TBTO during  the 12 days of
zoeal development to be 55 nmol/litre.

    Laughlin   &   French   (1980)  exposed   shore  crabs
 (Hemigrapsus   nudus), 2 to 3 days after hatching, to TBTO
(as "Biomet") for up to 14 days under static conditions.
At  the  highest  concentrations (500  and   1000 µg/litre),
all  the zoeae died within 2 days. Survival time increased
as the concentration of TBTO decreased from 100  to  25 µg
per  litre; most larvae died  within 8 days even at  25 µg
per litre. The estimated values of LT50   for 100, 75, 50,
and  25 µg/litre    were  3.4,  4.8,  5.8,  and  6.2 days,
respectively.   Lobster  larvae  (Homarus  americanus) were
much  more sensitive; 100%  being killed within  24 h  and
2 days by 20 and 15 µg   TBTO/litre, respectively. Concen-
trations of 10 and 5 µg/litre   killed all  larvae  within
5 to 6 days. In the group exposed to  1 µg/litre,    there
was a similar mortality pattern to the controls  and  high
mortality at the first ecdysis (3 to  5 days  post-hatch).
However, in this group only a single  larva  metamorphosed
successfully, compared with 43% in the control group.

    Davidson  et  al.  (1986) kept  juvenile mysid shrimps
 (Acanthomysis   sculpta), newly-released from the  female,
in  TBT concentrations of  between 0.03 and  0.48 µg   per
litre  for a 63-day period  under flow-through conditions.
The  TBT source was leachate from panels coated with anti-
fouling  paint.   All animals  died  within 7 days  at the
higher exposure level. There was no significant difference
in  shrimp mortality between those exposed at levels up to
0.38 µg/litre   and the controls, either at 22 or 41 days.
From  day 41 to  the end of  the test (63 days),  survival
decreased at 0.38 µg/litre  and only 22.5% survived to the
end  of the experiment, compared  with 60% in the  control
group.   The authors stated that this indicated a lowering
of the NOEL for the entire life cycle  of  A.  sculpta from
0.38 µg/litre   to 0.25 µg/litre  at 41 days. The increase
in  mortality coincided with  the release of  juveniles by
the  females, which indicated a sensitive time in the life
cycle.  Both mean length and  weight of females, at  a TBT
level of 0.38 µg/litre,   were significantly reduced after
63 days of exposure, but no effect of  TBT  concentrations
up to and including 0.38 µg/litre   was observed in males.

A  similar result was found in another study after 28 days
at 0.49 µg/litre;  again no effect on length or weight was
found  in males.  There  was a significant  effect on  the
length  of developing juveniles  and sub-adults. After  14
days of exposure to either 0.19 or 0.33 µg/litre,   mysids
were  significantly  shorter; this  was  also the  case at
0.2 µg/litre    after 27 days. At TBT concentrations up to
and  including 0.33 µg/litre,   there was no effect on the
number  of juveniles released  per individual female,  the
number of individuals in unhatched broods, or  the  number
of  days from hatching of  a female to the  release of its
juveniles.  However, there was a  significant reduction in
the  number of viable  juveniles released at  both   0.19 µg
per litre and 0.33 µg/litre.    The authors suggest a NOEL
of  0.09 µg/litre    for reproduction,  the most sensitive
parameter found in the study.

    Laughlin  et  al.  (1984) exposed  the Baltic amphipod
 Gammarus  oceanicus to TBTO or TBT fluoride concentrations
of  0.3 or 3.0 µg/litre    for 8 weeks, under  48-h static
renewal  conditions.  They  also exposed  the amphipods to
leachates  from  TBT-containing  antifouling  paints,   by
placing a 1 cm2   plexiglass plate, which was painted with
either "Micron 25" or "Interracing", in the  tank  for
5 weeks.  Again  the water  was  changed every  48 h.  TBT
levels  in  the  experiment with  pure compounds, remained
constant,  but  the  TBT concentration  (measured as TBTO)
from  the  leachates  gradually increased  over  each 48-h
exposure   period  to  approximately   5.5 µg/litre    for
"Interracing"   and  0.5 µg/litre    for  "Micron 25",
reflecting  the different leaching  rates of the  two com-
pounds.  At the highest concentration of both TBTO and TBT
fluoride, 50% of the adults died within 10 to  12 days  of
the  exposure and all had died within 16 days (TBTO) or 33
days (TBT fluoride). The TBT paint "Interracing"  caused
100%  mortality within 1 week.  The lower concentration of
the exposures to pure compounds and "Micron 25" resulted
in  mortality patterns that were not dependent on TBT con-
centration  but on senility. The exposures to 0.3 µg   per
litre  and  "Micron 25"  had significantly  reduced  the
number of surviving larvae by the end of  the  experiment.
Slight  decreases  in  larval growth  were  observed after
exposure to TBTO and "Micron 25" but, generally, concen-
trations of less than 1 µg   TBT/litre had  little  effect
on growth and no effect on whole animal oxygen consumption

    Clark  et  al. (1987)  monitored  the survival  of the
grass shrimp  (Palaemonetes pugio) in water and water/sedi-
ment test aquaria after the addition of TBTO.  When  added
to  water and tested in the absence of sediment, TBTO gave
96-h LC50 values  comparable to other published results at
20 µg/litre.    However,  when  the TBTO  was admixed with
sediment rather than water, the LC50   could not be deter-
mined.   TBT levels in sediment  at 1 mg/kg in static  and
10 mg/kg in flow-through tests showed no effects on shrimp

survival. Similar tests using  Amphioxus gave an LC50   for
sediment  containing TBTO of between 1 and 10 mg/kg over 4
and  10 days. The animals were  killed by 10 µg   TBT  per
litre in water.  Reproductive effects

    Hall et al. (1988b) maintained egg-carrying females of
the  copepod  Eurytemora affinis in TBT chloride  at levels
of  0.1 or 0.5 µg/litre   for up to 13 days.  After 3 days
of  exposure,  the higher  concentration had significantly
reduced  the mean brood size to 0.2, compared with 15.2 in
controls.  Neonate  survival was  significantly reduced at
0.1 µg/litre   after 6 days (22% of control survival).  No
offspring survived the higher exposure concentration. In a
second  experiment,  mean  brood  size,  after  2 days  of
exposure, was not significantly affected at concentrations
of between 12.5 and 200 ng/litre.  Neonate survival, after
13 days,  was  unaffected  by  concentrations  up  to  and
including  50 ng/litre;  survival at  100 ng/litre was 76%
(compared to 22% survival at the same exposure  level,  in
the first experiment, over 6 days) and was further reduced
to 24% at 200 ng TBT/litre.

    When  Johansen & Mohlenberg (1987)  exposed fertilized
mature  female copepods  (Acartia tonsa) to TBTO,  egg pro-
duction  was significantly reduced at the highest exposure
level  of  0.1 µg/litre    after 72 h.  After  120 h,  all
exposure  concentrations  had  significantly  reduced  egg
production by 18%, 19%, and 37% relative to  controls,  at
TBTO concentrations of 0.01, 0.05, and 0.1 µg   per litre,
respectively.  Limb regeneration

    Weis  et al. (1987a,b)  exposed the fiddler  crab  (Uca
 pugilator) to  TBTO concentrations of  0.5, 5.0, or    50 µg
per  litre  under static  renewal conditions. Regeneration
(autotomy)  of 1 chela and  5 walking legs was  induced by
pinching  them  off at  the  merus. Although  some  growth
retardation was observed, the most striking effect was the
development  in  regenerated  limbs of  deformities, e.g.,
backward  curling or complete absence of the dactyl of the
claw,  chelae or walking legs, stunted, unjointed, or bent
in the wrong direction. The percentage of males exhibiting
deformities was 17% at 0.5 µg/litre,   24% at  5 µg    per
litre, and 67% at 50 µg/litre,   the test  solution  being
changed  twice weekly. In  a second experiment,  where the
test solution was changed three times weekly, the percent-
age deformities were as follows: males, 100% at both 5 and
50 µg/litre;    females, 29% at  5 µg/litre   and 100%  at
50 µg/litre.     There were no deformities  in the control
groups during the experiments.  Behavioural effects

    When  Pinkney  et al.  (1985)  gave the  grass  shrimp
 (Palaemonetes   pugio) a choice between  TBTO-contaminated
water  and clean water, it did not avoid total organic tin
concentrations of 2.3 to 30 µg/litre.   The response data,
at both 2.3 and 30 µg/litre, were very similar.

    Meador  (1986) reported the effects of TBT chloride on
the  photobehaviour  of  water fleas  (Daphnia  magna). The
normal  response  of  Daphnia to  a  unidirectional   light
source is to swim away from the light; this is an adaptive
response  to  avoid  predators. The  author  reported that
 Daphnia exposed  to TBT chloride showed a reversal of this
behaviour  and swam towards the light source (usually with
considerably  increased swimming intensity). The threshold
concentration  of  TBT  chloride causing  this behavioural
reversal was 0.5 µg/litre   (the LC50   over the same time
period was between 3.5 and 6 µg/litre).

8.2.5.  Other aquatic invertebrates


     Few  studies have been carried  out on other species.   The
 lowest  effect  level  in  annelids  was  observed   for Nereis
 diversicolor; mortality and behavioural effects were seen after
 chronic exposure to 2 µg/litre.  Limb regeneration was signifi-
 cantly  inhibited in a  brittle star after  exposure to    0.1 µg
 per  litre.  Experiments have been  carried out on four  insect
 species.  The  most  sensitive was Notonectes,  the LC0    being
 0.03 mg/litre and the LC50  0.06 mg/litre.  Acute effects

    Walsh  et al. (1986a)  exposed larvae of  the  lugworm
 Arenicola   cristata to either TBTO or TBT acetate, for 96
or  168 h,  respectively.  The concentrations  that killed
100%  of the animals  were: 4 µg/litre   (96 h)  for TBTO;
10 µg/litre  (96 h) and 5 µg/litre   (168 h) for TBT acet-
ate. At 5 µg   TBT acetate/litre, all larvae were abnormal
after  96 h of exposure.  No deaths or  abnormalities  re-
sulted from exposure to 2 µg   TBTO/litre for up to 168 h.
When  Beaumont  et  al. (1987)  exposed  adult polychaetes
 (Nereis   diversicolor) to TBT in flowing sea water for up
to  22 days, there was 100%  mortality by 22 days at  4 µg
per litre and 55% mortality at 2 µg/litre    (20%  control
deaths).  Eversion of the proboscis was a  more  sensitive
indication  of toxicity.  No controls  showed this effect,
whereas  75%  of  animals exposed  to  2 µg/litre   showed
everted  proboscis  after  22 days (50%  after  20 days of

    Dragonfly  larvae  (Aeschna) sp. showed a 48-h  LC100 of
0.25 mg  TBTO/litre  and  an LC0 of   0.12 mg  TBTO/litre.
Corresponding  values  for  TBT acetate  were 0.5 mg/litre
(72-h LC100)   and 0.025 mg/litre (LC0).  Notonectes    sp.

yielded LC0 values  of 0.03 mg/litre for both TBTO and TBT
acetate  and 72-h LC100   values of 0.06 and 0.15 mg/litre
for TBTO and TBT acetate, respectively. Chironomid (midge)
larvae revealed a NOEL of 0.075 mg/litre for  TBT  acetate
and a 48-h LC100 of 0.15 mg/litre (Floch et al., 1964).

    Cardarelli  (1978) studied the efficacy of controlled-
release  organotin  compounds as  mosquito larvicides. The
toxicity  (LT100)   of BioMet (6% TBTO in natural rubber),
CBL-9B   (20%  TBT  fluoride   in  natural  rubber),   and
ECOPRO-1230  and ECOPRO-1330 (both are  ethylene propylene
polymers  containing 30% TBT fluoride) were tested against
larvae  of the mosquito  Culex quinquefasciatus. For BioMet
and  CBL-9B, LT100   values  ranged from 4  to 9 days  for
toxicant concentrations of 0.1 to 10 mg/kg of pellet added
to  the  water,  the toxicity  increasing  with increasing
toxicant concentration. A degradation product, TBT carbon-
ate,  was found to be less toxic, the LT100   being 9 days
after  exposure  to  10 mg/litre.  When  the polymers were
tested at toxicant concentrations ranging from 0.9 to 32.6
mg/kg,  toxicity  was  found to  increase  with increasing
toxicant  concentration (LT100s   range from  2 days to 16
days).  Although initially toxicity tended  to decrease as
the  soaking  time  of  the  pellet  increased,  prior  to
exposure,  between days 100  and 500 of  soaking  toxicity
remained  relatively unchanged. The authors  observed that
the   organotin  compounds  dramatically  slowed  or  even
prevented morphogenesis.  Limb regeneration

    In  a study on limb regeneration, Walsh et al. (1986b)
maintained  the brittle star  Ophioderma brevispina in TBTO
concentrations of 0.01, 0.1, or 0.5 µg/litre   under flow-
through  conditions. On the  first day of  the experiment,
autotomy was induced in two arms, at opposite sides of the
disc,  by pinching midway between  disc and arm tip.   The
animals were then exposed for four weeks to  TBTO.   There
were  no deaths and no  effect on disc diameter.  Both 0.1
and 0.5 µg/litre   significantly inhibited regeneration of
arms as measured by length; both groups showed average and
median lengths less than those in the  lowest-dose  group,
but variability precluded statistical significance.  Aver-
age  weights of  limbs were  also reduced  in  the  groups
exposed to 0.1 and 0.5 µg/litre,  but statistical analysis
was not carried out because pooled weights did not provide
enough data.

8.3.  Fish


     The  acute toxicity of tributyltin to marine and freshwater
 fish  is highly  variable, LC 50     values ranging  from 1.5  to
 240 µg/litre.     It is unclear whether this is due to inherent
 differences  in  sensitivity  or  to  differences  in  route of

 exposure.  Many of the acute toxicity studies need to be inter-
 preted  with care, since the TBT concentrations cited are often
 nominal  and  biologically  available concentrations  were  not

     In long-term toxicity tests, the NOEL for  general  toxico-
 logical  parameters (growth and  behaviour) for trout  yolk-sac
 fry  was greater  than 0.2 µg/litre,    and for  medaka it  was
 3.2 µg/litre.  Using histopathological parameters, the NOEL for
 medaka was found to be 0.32 µg/litre,  vacuolation of the reti-
 nal epithelium being the most sensitive parameter.

     Few embryotoxicity tests have been performed and it has not
 been possible to establish a NOEL.

8.3.1.  Acute effects

    The 96-h LC50   of TBTO for marine fish ranges between
1.5  and 36 µg/litre   (Table 12).  Larvae seem to be more
sensitive  than adults in those few studies examining dif-
ferent  life stages in the same test.  Fewer LC50   values
have been published for freshwater fish; they  range  from
13 to 240 µg/litre.

8.3.2.  Short- and long-term toxicity

    Matthiessen (1974) measured a lethal threshold concen-
tration  of TBTO for  Tilapia  mossambica of between 8  and
16 µg/litre  (24-h LC50,   28 µg/litre;  LC5,   24 µg  per
litre;  LC90,   33 µg/litre)   as a preliminary to studies
at  sublethal concentrations.  A temporary  opacity of the
surface of the eyes developed at concentrations below this
threshold  (> 8 µg/litre)    but  disappeared after  a few
days.  Other symptoms included sluggishness and difficult-
ies  with balance.  Melanophores in the skin were found to
be  constricted,  giving  treated fish  a paler appearance
than  controls.  The  fish did  not  produce behaviourally
related  display  patterns,  an activity  important in the
species.  As with the  eye opacity, these  other  symptoms
disappeared after a few days or weeks, suggesting that the
fish  develop some tolerance to  the TBTO.  The growth  of
 Tilapia exposed  to 0, 5, or 8 µg   TBTO/litre  was  moni-
tored  over a 5-week period.  These concentrations of TBTO
were  estimated, on the basis  of release rates, in  water
after  the TBTO had been  used to control mollusc  vectors
for  schistosomiasis.  There was  no difference in  growth
between fish exposed to 5 µg/litre   and controls.  Tilapia
exposed   to 8 µg/litre   showed negative  growth and lost
about  6% of body weight over the experimental period. Eye
opacity was only found in the fish exposed to  8 µg    per
litre,  but this  effect did  not seem  to affect  feeding
behaviour.  The only other  change in the  high-dose group
was  an increase in  aggressive encounters between  males.
There  was an increased reluctance  amongst attacked males
to avoid conflict, which led to the death of several indi-
viduals. Fatal engagements between males are normally very
rare in this species.

    When Seinen et al. (1981) exposed rainbow trout  (Salmo
 gairdneri) yolk-sac  fry to TBT chloride concentrations of
0.2, 1, or 5 µg/litre   for up to 110 days, all  fry  died
within  10 to 12 days (at the transition from yolk-sac fry
to the swimming fry stage) at 5 µg/litre,   but there were
no  deaths  at  lower doses.   The major histopathological
change in fish that died after exposure to 5 µg  per litre
was hydropic degeneration of tubule segments of  the  pro-
nephros. At 0.2 and 1 µg/litre,   there was a dose-related
retardation of growth, resulting in a 44% decrease in body
weight  (relative to controls)  in the 1 µg/litre    group
after 110 days. At the end of the experiment, both remain-
ing  groups  showed  significant reductions  in the haemo-
globin  titre in the blood and in body weight. Only at the
higher  concentration was there a  significant decrease in
blood cell number. Relative liver weight was significantly
increased  at both concentrations but  relative numbers of
thymus  cells were unaffected.  The authors also  measured
the relative area distribution of the various hepatic com-
partments in liver sections. The area occupied  by  nuclei
significantly  increased  at both  concentrations, whereas
glycogen  storage area decreased, but this was significant
only at the highest exposure level. Cytoplasmic  area  was
unaffected, as were all non-parenchymal compartments.

Table 12.  Toxicity of tributyltin to fish
Organism                 Size/      Stat/  Temper- Salinity  pH     TBT salt  Dura- LC50c    Reference
                         age        flowa  ature    (o/oo)                    tion  (µg/      
                                           (°C)                               (h)   litre)
Marine and estuarine species

Sheepshead minnow        sub-adult  flow   19.4-   9.8-      8.15-  chloride  48   >31d,f   Bushong 
 (Cyprinodon variegatus)                    21.3    12.1      8.31             72   28.1      et al. (1988)
                                                                              96   25.9 

Bleak                    8 cm       stat   10      7         7.8    fluoride  96     6-8e    Linden et 
 (Alburnus alburnus)                                                 oxide     96     15      al. (1979)

Inland silverside        larva      flow   20      10               chloride  48     7.7     Bushong et
 (Menidia beryllina)                                                                                (5.5-10.9)d,f al. (1987)
                                                                              72     4.6 

Atlantic silverside      sub-adult  flow   20      10               chloride  48     12.7    Bushong et
 (Menidia menidia)                                                               (7.8-15.2)d,f al. (1987)
                                                                              72     9.3 
                                                                              96     8.9 
Atlantic menhaden        juvenile   flow   20      10               chloride  48     6.8     Bushong et al. (1987)
 (Brevoortia tyrannus)                                                           (4.1-infinite)d,f    
                                                                              72     5.2 
                                                                              96     4.5 

Sole                     larva      statb                           oxide     48     8.5     Thain (1983)
 (Solea solea)            larva                                                96     2.1
                         adult                                                48     88

Armed bullhead           adult      statb                           oxide     48     26      Thain (1983)
 (Agonus cataphractus)                                                         96     16

Table 12.  (contd.)
Organism                 Size/      Stat/  Temper- Hard-     pH     TBT salt  Dura-  LC50c   Reference
                         age        flowa  ature   nessg                      tion   (µg/      
                                           (°C)  (mg/litre)                   (h)    litre)

Girella                  2.4 g      statb  19.9-             7.8-   oxide     48     5.2d    Kakuno & 
 (Girella punctata)                         20.5              8.1              96     3.2d    Kimura (1987)

Saltwater goby                      statb  20.2-   32.5-     8.1-   oxide     24     12d     Shimizu & 
 (Chasmichthys dolichognathus)              21.0    32.8      8.3              48     9d      Kimura (1987)
                                                                                96     4d

Chinook salmon           juvenile   stat   3-5     28               oxide     6      54d     Short & 
 (Oncorhynchus tshawytscha)                                                    12     20d     Thrower (1987)
                                                                              96     1.5d

Mummichog                larvae     flow   19.4-   9.8-      8.15-  chloride  48    >32.2d,f Bushong et 
 (Fundulus heteroclitus)  larvae            21.3    12.1      8.31             72     28.2     al. (1988)
                         larvae                                               96     23.4 
                         sub-adult                                            48     >32.2d,f
                         sub-adult                                            72     28.3 
                         sub-adult                                            96     23.8 

Freshwater species

Rainbow trout            yearling   stat   18      250              oxide     24     28e     Alabaster 
 (Salmo gairdneri)        yearling   statb  18      250              oxide     48     21e     (1969)


Table 12.  (contd.)
Organism                 Size/      Stat/  Temper- Hard-     pH     TBT salt  Dura-  LC50c   Reference
                         age        flowa  ature   nessg                      tion   (µg/      
                                           (°C)  (mg/litre)                   (h)    litre)
Guppy                    3-4 weeks  statb  23                       oxide     96     21      RIVM 
 (Lebistes reticulatus)                                              chloride  168    21      (1989)
                                                                                  (16-29)e,f Polster & 
                                                                    oleate    168    33      Halacka 
                                                                                  (19-60)e,f (1971)
                                                                    benzoate  168    25 
                                                                    laurate   168    30 
                                                                    acetate   168    28 
                                                                    oxide     168    39 

Medaka                   4-5 weeks  statb  23                       oxide     96     17      RIVM 
 (Oryzias latipes)                                                                            (1989)

Stickleback              4-5 weeks  statb  19                       oxide     96     13      RIVM 
 (Gasterosteus aculeatus)                                                                     (1989)

Carp                                                                oxide     24     75      Temmink & 
 (Cyprinus carpio)                                                   oxide     96     32      Everts 
Golden orfe                                                         oxide     48     50e     Plum 
 (Leuciscus idus melanotus)                                          napht-    48     70e     (1981)
Bluegill sunfish                    stat   20                       oxide     96     240e    Foster 
 (Lepomis macrochirus)                                                                        (1981)
a   stat = static conditions (water unchanged for the duration of the test unless stated otherwise);
    flow = flow-through conditions (TBT concentration in water continuously maintained).
b   Static renewal conditions (water changed periodically).
c   95% confidence limits are given in brackets.
d   Measured concentration.
e   Nominal concentration.
f   Concentration expressed as TBT.
g   Hardness expressed as mg CaCO3/litre.
    Pinkney  et  al.  (1988)  exposed  13- and  16-day-old
larvae  of striped bass  (Morone saxatilis) to varying con-
centrations  of TBT from methacrylate-painted panels for 6
to  7 days.  They found significant reductions in survival
at measured concentrations of 0.766 µg  TBT/litre or more.
At lower exposure concentrations (0.067 µg/litre),  growth
parameters  changed  in  the 13-day-old  larvae  only.  No
changes  occurred in the  growth parameters of  16-day-old
larvae exposed to 0.444 µg/litre  or less. The authors did
not  know whether this apparent  difference in sensitivity
between  larvae of  different ages  was a  true effect  or
simply  the  result  of differences  between  the  various
batches of larvae.

    Wester  et al. (1988)  reported NOELs for  the  medaka
 (Oryzias  latipes) of 3.2 µg  TBTO/litre for general toxi-
cological  parameters (mortality, growth,  general appear-
ance,  abnormal behaviour) and 0.32 µg/litre    for histo-
pathological  effects during 1 month of exposure. The most
sensitive  histopathological effect was the development of
vacuolation  in the retinal  epithelium of the  eye. There
was a dose-related increase in hepatocellular vacuolation,
with  swelling in pronounced cases.  The vacuoles appeared
to  be glycogen deposits, although, at higher doses, lipid
vacuoles  were also noted.  The NOEL for liver effects was
1 µg    TBTO/litre.   Kidney  effects mostly  involved the
tubule and included dilation, epithelial atrophy, degener-
ation  and regeneration, and proteinaceous casts including
cellular  debris  (tubulonephrosis).  Severe  cases   also
showed  glomerular effects. Similar lesions  were reported
after  3 months of exposure, with, in addition, effects on
the  swim bladder, skin, oral cavity, and thyroid gland at
the  highest concentration tested (10 µg/litre).    Thymus
atrophy,  reported by the same  authors in the guppy,  was
not  found in the  medaka, indicating some  species speci-
ficity regarding the effects of TBT. Increased glycogen in
liver  and  muscle  after TBTO  treatment was demonstrated
analytically  in  the  medaka,  confirming   histochemical
suggestions of increased glycogen in the guppy.

    Shimizu  &  Kimura  (1987)  exposed  the  marine  goby
 Chasmichthys   dolichognathus to TBTO in short (4 day) and
long  (12 week) experiments, and reported a 96-h LC50   of
4 µg/litre.    Long-term exposure to 2.1 µg/litre   during
the season of gonadal recrudescence led to  a  significant
depression of the gonadosomatic index in male fish.  There
was  no effect on  female fish in  terms of  gonadosomatic
index or ovarian histology.

8.3.3.  Embryotoxicity

    Newton et al. (1985) exposed eggs, embryos, and larvae
of  the  California grunion  (Leuresthes  tenuis) to plates
painted with 9.4% TBT methacrylate (0.5% TBTO;  44.7%  cu-
prous  oxide; 45.5% inert ingredient) and aged for 30 days

in  flowing sea water.  The authors claimed  that the  ex-
posure to concentrations of between 0.14 and 1.72 µg   TBT
per  litre had no adverse  effect on hatching, growth,  or
development. In fact the presence of TBT at  such  concen-
trations  enhanced  both  hatching success  and stimulated
growth. A concentration of 10 µg   TBT/litre had no effect
on hatchability, but at 74 µg/litre   hatching success was
reduced  by  approximately  50%. The  authors  observed no
effect  on  the  survival  of  larvae,  hatched  from eggs
exposed to concentrations of 0.14 to 1.72 µg/litre,  up to
7 days post-hatch. The presence of copper, however, should
be taken into consideration in the interpretation of these

    Weis et al. (1987a) reported considerable variation in
the   response  of  embryos  of   the  killifish  (Fundulus
 heteroclitus) to  concentrations of TBTO ranging between 3
and 30 µg/litre.   Two batches of embryos developed abnor-
malities  at  the  highest concentration  of  TBTO  tested
(30 µg/litre)  and showed some mortality. Other batches of
embryos  showed  embryotoxicity  for  all  tested  concen-
trations (down to 3 µg/litre)  but no abnormalities in the
embryos. Owing to this difference in sensitivity  of  dif-
ferent batches of eggs and the separation  of  embryotoxic
and  teratogenic effects, it  is difficult to  establish a
NOEL for this species.

    Fent (1989b) exposed fertilized eggs of the freshwater
minnow  Phoxinus  phoxinus to  two  TBT  concentrations  in
petri  dishes in an environment chamber. The eggs were ex-
posed from the blastula stage (20 to 24 h after spawning),
and  the water was renewed daily. Analytical determination
of  TBT showed that  the actual concentration  had reduced
over the 24-h exposure period from 1.5 to 0.5 µg/litre  in
the low-dose group and from 8.4 to 7.9 µg/litre    in  the
high-dose  group.  There  were no  differences in survival
between the control and low-dose groups. Hatching was nor-
mal but some fish showed vertebral malformations.  At  the
high dose, hatching was delayed or reduced and all hatched
larvae  had  severe  vertebral  malformations;  many  were
unable  to uncurl. These larvae remained motionless at the
bottom of the dish. Some had oedema in the region  of  the
heart. Within 3 to 4 days of hatching, all of  the  larvae
in  the high-dose  group had  died. A  further  series  of
experiments  similarly showed malformation at  an exposure
level of 4.5 µg TBT/litre.

8.3.4.  Behavioural effects

    The  avoidance  response  of  the  mummichog  (Fundulus
 heteroclitus) to TBTO has been studied by Pinkney  et  al.
(1985).  When given a choice between TBTO-contaminated and
clean  water, 4 out  of 6 groups of  fish avoided a  total
organic tin level of 1 µg/litre.    There were significant
avoidance responses for all test groups at  total  organic
tin levels of 3.7, 8, and 13.8 µg/litre.     However,  the

higher concentrations of tin did not result in an increase
in avoidance response. The authors found that,  under  the
same  conditions, no avoidance  response was shown  by the
grass shrimp (see section This shrimp is a major
food item of the mummichog in tidal marshes. Hall  et  al.
(1984)  studied the avoidance  response of two  species of
juvenile   estuarine   fish,   the  striped   bass  (Morone
 saxatilis) and Atlantic menhaden  (Brevoortia tyrannus). An
avoidance  response to  TBTO was  exhibited by  bass at  a
total  organic tin concentration  of 24.9 µg   per  litre.
Atlantic  menhaden were more sensitive, showing a "mild"
avoidance  at 5.5 µg/litre   and a "strong" avoidance at
9.1 µg/litre.

    Chliamovitch  & Kuhn (1977) measured an EC50   for the
loss of positive rheotaxis of 30.8 µg/litre  for the rain-
bow  trout  (Salmo  gairdneri) and  53.2 µg/litre    for  a
tilapia  (Tilapia rendalli).

8.4.  Amphibians


     There are few data on the effects of tributyltin on amphib-
 ians.   Survival  of frog  larvae was at  least 80% of  control
 levels after exposure to 3 µg TBT/litre.

    Floch et al. (1964) reported a NOEL for  mortality  of
tadpoles  of two species of amphibians (Rana and Alytes sp.)
of 30 µg/litre.   For both species, the 24-h  LC100    for
TBTO was 75 µg/litre,  the 48-h LC100   50 µg/litre,   and
the 48-h LC100 for TBT acetate was 75 µg/litre.

    Laughlin & Linden (1982) exposed eggs of the frog  Rana
 temporaria to  concentrations of either TBTO or TBT fluor-
ide of 0.3, 3, or 30 µg/litre,   for 5 days from the post-
gastrula  stage of development, under  static renewal con-
ditions.   All surviving larvae hatched on either day 4 or
5 of exposure. Survival was at least 80% of control levels
after exposure to 0.3 or 3 µg/litre.   Only at the highest
concentration (30 µg/litre)   did TBT affect survival (40%
mortality  with TBT fluoride and 50% mortality with TBTO).
The  authors collected all surviving tadpoles and measured
wet  and dry weight.  Wet weights were significantly lower
in  tadpoles  exposed  to 30 µg/litre    but unaffected at
lower  exposure  concentrations.  Tadpoles  exposed to TBT
always  showed  higher  mean  dry  weight  than  controls,
although  increases were not consistently  dose dependent.
The  percentage of body  water declined from  86-88.5% for
controls  to 73.8% after  exposure at the  highest concen-
tration  of TBT.  The changes resulting from TBT treatment
were  significant, but not the differences produced by the
different TBT compounds.

8.5.  Multispecies studies


     The  few available microcosm studies indicate that the most
 sensitive  organism(s) within the microcosm are affected by TBT
 at  concentrations  greater  than 0.05 µg/litre.    Recovery of
 these  organisms occurred a  few months after  exposure. Simul-
 taneous exposure of snails and fish yields conflicting results,
 snails appearing somewhat more sensitive than fish.

    Beaumont  et al. (1987) exposed sandy-substrate micro-
cosms  of flowing  sea water  to TBT  derived  from  slate
panels painted with "Micron 25" antifouling paint. Three
replicate  microcosms  were  exposed for  4 months at high
(1-3 µg/litre)  and at low (0.06-0.17 µg/litre)   TBT con-
centrations  in water, together with  three control micro-
cosms. Flow rates were maintained by gravity from a header
tank at approximately 1 litre/min; the painted panels were
located  in the header  tank where water  was passed  over
them  using a circulating pump. The substrate sediment was
sieved  after collection (to 2 mm) to remove larger organ-
isms and allowed to settle in the microcosms  for  3 weeks
before  the addition of  animals. The sediment  was  10 cm
deep and the overlying water 15 cm deep in each microcosm.
At the beginning of the trial, 50 specimens of  a  bivalve
 (Cerastoderma  edule), a crustacean  (Corophium volutator),
and   two  polychaetes  (Nereis diversicolor and  Cirratulus
 cirratus) were  introduced. After 4 weeks, 12 specimens of
the  gastropod  Littorina littorea were added and,  after 6
weeks,  26 specimens  of  a second  bivalve  (Scrobicularia
 plana) were  added. Other species were found at the end of
the  trial derived from  small (< 2 mm) juveniles  in  the
sediment and inflowing sea water. Each day, 10 litres of a
suspension of the microalga  Pavlova lutheri (5 x 106 cells/ml)
were added to the header tank as additional  food  supply.
The  most  sensitive of  the  introduced species  was  the
cockle  (C.  edule); 100% died within  2 weeks at the  high
concentration of TBT and there was cumulative mortality at
the low TBT concentration over 17 weeks (14%  of  controls
died).  Two-way analysis of variance indicated significant
differences  between all three  groups (i.e. control,  and
low and high concentrations of TBT; p < 0.05).  The  other
bivalve  (S.  plana) showed  high mortality  at the highest
exposure  level (100% after 10 weeks). At the low TBT con-
centration  there  were  no time-related  deaths  in  this
species  and  no  significant  difference  from  controls.
Among  the  polychaetes,  high death  rates  were recorded
for  N.  diversicolor in  all  microcosms,  including   the
control,  possibly because adults introduced were ripe and
died  after  spawning  during the  experiment.   The other
polychaete  (C. cirratus) survived well even after exposure
to  the high concentration of TBT.  Only one gastropod  (L.
 littorea) died in any of the microcosms. Juvenile bivalves
were  the  most common  species  found; their  numbers and
diversity  were lower in  the low-dose microcosms  (66  to

109)  than in control  microcosms (137 to  179), and  they
were virtually absent from the high-dose microcosms (0, 1,
and  1 for the 3 replicates). The size of juvenile mussels
 (Mytilus   edulis) was  also significantly  reduced at the
low  TBT concentration, relative to  controls; other self-
introduced  bivalve species were  not affected by  the low
TBT  exposure. Measurements of  chlorophyll  a in  sediment
cores, as a measure of algal biomass, indicated a signifi-
cant  rise after  exposure to  high TBT  levels.  This  is
explained  by the relative  insensitivity of algae  to the
toxic  effects  of TBT  and  by reduced  (or non-existent)
animal life available to consume the algae.   The  authors
emphasize the sensitivity of some species, for  which  TBT
is toxic at levels of 1 to 2 µg/litre    under  conditions
approximating  natural  exposure.  They  further emphasize
the  great variation in  sensitivity between species;  for
example, while  Mytilus edulis and  Cerastoderma  edule were
clearly  affected at very low levels of TBT, other related
bivalves  were  largely unchecked.  Differences in feeding
behaviour, leading to different effective exposure levels,
is  offered as one possible explanation for differences in
response.  Differential  absorption  or  subsequent  loss,
observed  in  laboratory  studies with  other molluscs, is
also proposed as a possible mechanism.

    Henderson   (1986)  conducted  long-term  flow-through
microcosm  studies on communities of marine organisms from
Pearl  Harbour,  Hawaii,  where  the  organisms  had  been
exposed  to leaching TBT from  naval ships for some  time.
Panels (20 x 25 cm) of roughened plexiglass were suspended
in  the harbour for  14 weeks prior to  the experiment  to
allow settlement and growth of organisms, and 30 different
species were found on the panels. Prior to  exposure,  the
panels were kept for a further 5 weeks in the experimental
tanks.  These tanks had a capacity of 155 litre  and  were
supplied with sea water pumped directly from  the  harbour
and through the tanks at a rate of 4 litres/min.  The  sea
water  was passed  through a  1 cm mesh  with  no  further
filtration;  organisms  could, therefore,  enter the tanks
easily  and settle on  the panels. The  water was rich  in
plankton  consisting  largely  of diatoms,  copepods,  and
chaetognaths.  Exposure to TBT derived from painted panels
treated  with antifouling paint containing  9.4% TBT meth-
acrylate,  0.5% TBTO, 44.7% cuprous oxide, and 45.4% inert
ingredients.   Nominal concentrations of TBT were 0, 0.05,
0.13,  0.31,  0.78,  and 1.95 µg/litre,    but actual mean
concentrations  were  0.01,  0.04, 0.10,  0.54,  1.77, and
2.52 µg/litre,   respectively.  The measured copper levels
in  the  same samples  were 1.0, 1.2,  1.2, 3.1, 4.4,  and
5 µg/litre,    respectively.   Changes in  coverage of the
panels was assessed using photographs taken with an under-
water  camera at weekly  or biweekly intervals  throughout
the   experimental  period  of  approximately  60 days  of
exposure and a further 60 days of recovery. One week after
the  start of exposure, a further plastic panel was intro-
duced  to  monitor new  colonization.   There was  a clear

decline  in both total  number of species  and in  species
diversity  (reductions of 55%, 60%,  and 80% at the  three
highest  exposure  levels,  respectively)  on   pre-fouled
panels  exposed to TBT at 0.5, 1.8, and 2.5 µg/litre   but
no  effect at lower concentrations. The mortality of indi-
vidual species, in relation to TBT exposure,  showed  con-
siderable variation.  The most sensitive of the  six  most
common  species  on  the panels  was  Botrylloides spp,  an
orange-coloured  colonial tunicate, which showed 100% mor-
tality at TBT concentrations of 0.1 µg/litre   or more. An
encrusting bryozoan,  Schizoporella errata, showed 49% mor-
tality at 0.1 µg/litre  and 80 to 100% mortality at higher
concentrations.  Specimens of a second  colonial tunicate,
 Didemnus  candidum, and the saddle oyster ( Anomis nobilis)
were   all killed at  concentrations of 0.5 µg/litre    or
more.  A  tube  worm  (Hydroides  elegans) and  a  solitary
tunicate    (Ascidia spp.)   survived  even   the  highest
exposure  level, though with considerable mortality at 1.8
and  2.5 µg/litre.    Recovery of populations was complete
60 days  after cessation of treatment.   Settlement on the
panels not previously exposed was reduced by  TBT  concen-
trations of 0.1 µg/litre  or more, but not by 0.04 µg  per
litre.   Algal settlement and growth  on the walls of  the
tanks was not affected by exposure to TBT at  any  concen-
tration;  coralline algae increased their coverage of pre-
fouled   panels.   A   total  of   18 oysters  (Crassostrea
 virginica) placed  in  each  tank prior  to  treatment was
monitored  after 60 days.  There  was 50% mortality  after
TBT exposure at 1.8 µg/litre  but no significant reduction
in  survival  at  lower concentrations.  However, the con-
dition  index of the oysters  was affected by exposure  to
0.1 µg/litre    or more, but  had returned to  near normal
after  2 months  recovery  in  clean  water.  The   author
regarded 0.05 µg/litre   as a reliable NOEL for  the  most
sensitive organisms exposed.

    When   Salazar  &  Salazar  (1985)   exposed  copepods
 (Acartia   tonsa), mysids  (Acanthomysis sculpta), and fish
 (Citharichthys  stigmaeus) to suspended sediment at a TBTO
concentration  of 0.49 µg/litre   water for  96 h, no sig-
nificant  mortality  was  found. In  a  second experiment,
mysids  (A. sculpta), worms  (Neanthes arenaceodentata), and
clams  (Macoma  nasuta) were  exposed to  TBTO-contaminated
sediment with overlying water for 10 days in the  case  of
mysids and 20 days in the case of clams and worms.  Levels
of TBTO in the sediment varied from 155 to 610 µg  per kg,
falling  during the exposure period, while measured levels
in  the overlying water were about 0.2 µg/litre.   No mor-
tality was observed during the exposure period.

    Cardarelli    (1973)    exposed   snails  (Biomphalaria
 glabrata) and    fish  (Lebistes    reticulatus) to   slow-
releasing  molluscicides  containing  either TBTO  or  TBT
fluoride. At a daily release rate of  35 µg    TBTO/litre,
100% mortality was achieved in snails within  60 days  but
all  of the fish also  died within this period.  At a TBTO

release  rate of 7 µg/litre   per day, snail mortality was
100% after 120 days exposure and fish mortality  only  2%.
However,  a second experiment gave a much higher fish mor-
tality of 46% after 120 days exposure to only 3.5 µg  TBTO
per  litre per day. Exposure to TBT fluoride at 7 µg   per
litre per day killed all snails and fish  within  60 days,
and  52% of fish and  100% of snails were  killed after 90
days of exposure to 3.5 µg/litre per day.

    In  order to simulate the effect of molluscicides con-
taining  6% TBTO on  biota, Jordan (1985)  set up a  model
ecosystem  as follows: filtered river water was allowed to
flow  into a  tray containing  washed mud  (to simulate  a
marsh), and overflow from this upper tank was  allowed  to
flow into a lower tank containing washed coarse  sand  (to
simulate   a  river).   Snails  (Biomphalaria  glabrata and
 Potamopyrgus    coronatus) and  guppies  (Lebistes  reticu-
 latus) were  added  to  the  "marsh",  while  snails  (B.
 glabrata and  Pomacea       glauca), shrimps  (Macrobrachium
 faustinum), and  L.    reticulatus were   added   to    the
"river".  The molluscicide was added to the "marsh" as
pellets  at levels of 2, 5, 10, or 20 g/m2   of marsh sur-
face.   The percentage of both  B.  glagrata and  L. reticu-
 latus surviving  14 days  of  exposure was  recorded  (see
Table 13).  The authors concluded that, in this model eco-
system,  there was higher mortality of  B. glabrata than of
 L.   reticulatus at low doses of slow-release TBTO, i.e. 2
and 5 g of pellet per m2 of "marsh".

Table 13.  Percentage of  B. glabrata and  L. reticulatus surviving 
14 days exposure to slow-release pellets containing 6% TBTO added 
to the "marsh" in a "marsh" or "river" model ecosystema
Dose of               B. glabrata             L. reticulatus
               "marsh"      "river"       "marsh"      "river"
20 g/m2          00.0*        00.0*         00.0*        00.0*
               (0/240)      (0/240)       (0/240)      (0/240)

10 g/m2          00.0*        00.0*         00.0*         2.5*
               (0/240)      (0/240)       (0/240)      (6/240)

 5 g/m2          00.0*        16.3**        23.3**       60.0+
               (0/240)      (39/240)     (56/240)      (144/240)

 2 g/m2          52.9+        75.4++        75.8++       78.8
              (127/240)     (181/240)    (182/240)     (189/240)
a   From: Jordan (1985).
    The percentages represent percent survival at various doses 
    for all weeks of exposure in replicate experiments. The 
    values marked with the same symbol are not statistically 
    different from each other at the 0.05  probability level. The 
    numbers in parentheses represent the number of survivors  
    compared to the number of animals exposed. 



     Exposure  of terrestrial organisms to TBT derives primarily
 from  its use as a  wood preservative.  However, little  infor-
 mation is available. TBTO has proved toxic to honey-bees coming
 into close contact with TBTO-treated wood.  TBT  compounds  are
 toxic  to insects exposed  either topically or  via feeding  on
 treated  wood. The acute toxicity to wild small mammals is mod-
 erate  (between 37 and 240 mg/kg per day). The toxicity of TBTO
 to  bats is probable but not proven. There is no information on
 other species.

9.1.  Microcosm studies

    Gile  et al. (1982) introduced TBTO into a terrestrial
microcosm  on pine posts,  each microcosm containing  four
posts (3.3 x 2.6 x 14 cm) treated with 14C-labelled   TBTO
(167 mg/cm3).    The  microcosms  consisted  of  soil  and
endogenous  soil organisms, ryegrass, earthworms, pillbugs
(woodlice),  mealworms,  crickets,  garden snails,  and  a
gravid female gray-tailed vole. There were no  effects  on
any  of the organisms over a period of 77 days.  About 95%
of  the TBTO remained  in the posts  for the whole  of the
exposure  period. A similar  system set up  to investigate
cricket  mortality, with and without  predation, showed no
effect of TBTO (Gillett et al., 1983).

9.2.  Terrestrial insects

    When  Gardiner & Poller  (1964) exposed larvae  of the
common clothes moth  (Tineola bisselliella) to wool treated
with  a  TBTO concentration  of 1% by  weight, all of  the
exposed larvae were killed within the exposure  period  of
14 days.  The action  appeared to  be that  of  a  contact
insecticide because none of the cloth was eaten. Phenyltin
compounds were not as toxic as TBTO. Baker & Taylor (1967)
found that the action of TBTO more closely  resembled  the
slow  toxicity  of  a stomach  insecticide  after exposing
wood-boring beetles  (Lyctus brunneus) to impregnated wood.
Contact insecticides such as  lindane ( gamma-hexachlorocyclo-
hexane)  and dieldrin were 100 times more toxic than TBTO.
The LD50 of  TBTO for another wood-boring species,  Anobium
 punctatum, was 0.254 kg/m3 (application rate to wood).

    Saxena  & Crowe (1988) applied TBTO, TBT chloride, and
TBT  linoleate  topically  to the  thorax of newly-emerged
insects  of three species.  The LD50   values  ranged from
0.48%  to 0.72% (dilutions with acetone) for the house fly
 Musca    domestica, 0.29%   to  0.69%   for  the  mosquito
 Anophelese   stephensi, and 0.52% to 0.87%  for the cotton
stainer  Dysdercus   cingulatus. TBT  compounds  were  more
toxic  than the other organotin  compounds tested (triphe-
nyltin,   tricyclohexyltin,  dimethyltin,  phenyltin,  and

diethyltin).   The authors pointed out  that TBT compounds
are considerably less toxic than trimethyltin compounds to
 Musca sp.

    In  studies  by  Kalnins  &  Detroy  (1984),  wood was
treated  with TBTO (1.9 kg/m3)    after sawing and  before
use  in the construction of beehives. Five hives made from
the  treated wood were stocked with bees.  Tin residues of
3.24 mg/kg were found in bees during the first summer, and
residues of 8.67 mg/kg were found in the wax of the combs.
However, there were no detectable tin residues  in  honey.
There  was high mortality in the bee colonies over winter,
only one colony surviving. No control colonies  died  over
winter.   Residues in bees and wax in the second year (1.3
and  4.6 mg/kg, respectively) were lower than in the first
year in the one surviving colony.

9.3.  Terrestrial mammals

    Racey   &   Swift   (1986)  housed   pipistrelle  bats
 (Pipistrellus   pipistrellus) in  roosting  cages  treated
with TBTO. The bats were pregnant females  collected  from
nursery roosts, and they were trained to feed on mealworms
before transfer to the experimental cages.  The cages were
made of metal, lined with plywood, and were  painted  with
TBTO as a 1% solution in white spirit  (the  manufacturers
recommended rate for use of TBTO as a fungicide). The wood
was  treated 2 months before the bats were introduced into
the cages.  During the course of the  142-day  experiment,
seven of the ten bats died. Median survival time  was  100
days,  with a range between 49 and 142 days. Two deaths in
the  white spirit control group and three in the untreated
control group meant that results for TBTO were  not  stat-
istically significant.

    Schafer  & Bowles (1985) conducted toxicity and repel-
lency  tests  on  deer  mice  (Peromyscus  maniculatus) and
house  mice  (Mus musculus) using various TBT salts. Treat-
ment of feed seeds with 2% (by mass) produced clear repel-
lent  effects  of TBT  in  both species.  The  approximate
lethal  dose (estimated by  increasing the TBT  dose until
the  test animals died) for  TBT acetate and fluoride  was
320 mg/kg. The estimated dietary LC50,   based on consump-
tion  of treated seed used in the repellency tests, varied
between 37.5 mg/kg per day for TBT acetate  and  238 mg/kg
per day for fluoride and sulfate.  The LC50   for TBTO was
200 mg/kg per day.



     Tributyltin compounds have had a wide range of  uses.  Most
 concern  has focussed on their  use in the marine  environment,
 where  TBT has been  associated with mortality  and failure  of
 settlement of bivalve larvae, reduced growth, shell thickening,
 and other malformations in oysters, imposex (the development of
 male  reproductive appendages in female animals) in mud snails,
 and  imposex concurrent with  population decline in  dogwhelks.
 Controls  on the use  of TBT in  antifouling paints has  led to
 recovery  of economically important shellfish growth and repro-
 duction. Water concentrations of TBT are still high  enough  in
 some areas to affect marine gastropods.
     Field  and laboratory results  for marine molluscs  are  in
 good agreement.  Both imposex in dogwhelks and shell growth and
 chambering  in Pacific oysters  are effective biological  indi-
 cators of TBT contamination.
     There have been few studies into the effects  on  organisms
 of TBT in sediment. There are indications that TBT is bioavail-
 able  to burrowing  organisms and  can cause  mortality in  the
     Gross toxic effects and histopathological changes have been
 reported in farmed marine fish exposed to TBT through  the  use
 of antifouling paints on retaining nets.
     Although  TBT has been detected in fresh water at high con-
 centrations in some areas, there have been few studies  on  its
 effects. Spills of large amounts of TBT from  timber  treatment
 plants  have  caused  ecological damage,  but recovery occurred
 within 9 months.
     Field  testing  of  tributyltin derivatives,  mainly  slow-
 release formulations of TBTO, has shown that it is difficult to
 apply  TBT  without  damaging  non-target  organisms.  Recovery
 occurs through recolonization.

10.1.  Effects on bivalves

    In   the  1970s,  the   French  oyster  industry   was
undergoing  a  crisis. During  the  1977, 1978,  and  1979
seasons,  very poor spatfalls were reported, together with
increasing  reports of poor shell growth and shell malfor-
mations.  An extensive survey of the occurrence and inten-
sity of malformations related to metal residues in oysters
suggested, for the first time, a connection with organotin
compounds  (Alzieu, 1981).  The shell thickening was found
to  be  due to  the appearance of  chambers in the  oyster
shell  and interlamellar gel  formation in these  cavities
(Alzieu,  1981; Alzieu et al., 1982). The authors of these
reports  described in detail  the shell abnormalities  and

the  process  of calcification  of  the oyster  shell, and
suggested mechanisms of TBT action.  TBT is known  to  in-
hibit  oxidative phosphorylation and it has been suggested
that this forms the basis of its action on the  shell.  It
is also known to complex amino acids. The effect on calci-
fication  derives from inadequate calcium  addition to the
organic  matrix (a process dependent on ATP) and incorrect
deposition of this matrix. Alzieu et al. (1982) found good
correlation between the occurrence of shell thickening and
the proximity of ports where large numbers of  boats  were
usually moored. These field observations were corroborated
by  the finding of similar shell abnormalities in the lab-
oratory  when  oysters were  exposed  to TBT  fluoride, an
organotin  compound present in  antifouling paint used  on
the  boats. Oysters placed in flowing sea-water tanks con-
taining  plates  coated  with TBT  antifouling paints died
after  a  30-day exposure  to  an estimated  water concen-
tration  of  2 µg/litre    (organotin leachate)  and shell
thickening was found to occur at water  concentrations  of
0.2 µg/litre.

    Alzieu  et al. (1982)  assessed the shell  quality  of
18-month-old   Pacific   oysters ( Crassostrea  gigas; also
known  as the Japanese  oyster) sampled along  the eastern
coast of Oleron Island and in the vicinity of La Rochelle,
France.   They  concluded  that the  proximity of pleasure
craft ports or a commercial harbour could badly affect the
quality and growth of oyster shells. However, the presence
of other chemicals, along with TBT in the sea  water  made
direct  assessment  difficult.  The  authors,   therefore,
carried   out  a  set  of  experiments  to  confirm  their
conclusions.  Groups of oysters  from areas with  no shell
abnormalities  were distributed to other locations, i.e. a
marina,  a  river, and  the  laboratory (three  groups:  a
control group; a group exposed to 50-cm2    panels  coated
with  TBT fluoride; and a  group exposed to coated  panels
500 cm2   in area).  All oysters died within 30 days after
exposure  to the larger panel in the laboratory and within
170 days  after transfer to the marina. Oysters exposed to
the  smaller panels showed a  mortality rate of 30%  after
110 days  of exposure. Oysters  transferred to the  marina
site and those exposed to 50-cm2 TBT-coated  panels devel-
oped  gel-filled  shell  cavities after  100 and 110 days,
respectively, during the period of shell growth.  Analysis
of  the oysters for total  tin content revealed levels  of
< 1 mg/kg (dry weight) in "unpolluted" groups, 110 mg/kg
after 80 days at the marina site, and up to 25 mg/kg after
exposure to 50-cm2 panels in the laboratory.

    Maurer et al. (1985) found TBT levels to be related to
inhibition of settlement of Pacific oyster in the  Bay  of
Arcachon and the Gironde estuary, France. The authors used
arrays  of settling tubes  mounted around a  central  tube
which  was painted with "International  TBT Antifouling"
at a rate of 6.4 g of paint on a surface of  8 dm2.    The
settling  tubes were mounted at varying distances (between

4.5  and 13 cm) from the central painted tube.  In control
arrays,  the central tube  was unpainted.  The  arrays  of
tubes  were placed out in the two study areas in July 1982
and  observations on settlement  and growth were  made  in
September  and  November 1982.  In  the Gironde  region, a
second  method was also used; slates were painted with TBT
paint  (21 g  of  "International TBT  Antifouling"  on a
surface  of 26 dm2)   and mounted 10 cm apart, alternating
unpainted  with painted slates on  a rod.  The spacing  of
tubes  in the  arrays was  up to  25 cm from  the  central
painted  tube. In the Gironde estuary, settlement was com-
parable  with controls except  on the painted  tube; tubes
5 cm or more away from the TBT paint had  similar  numbers
of  settling larvae as the control.  However, deaths among
the settled larvae were high; 100% on tubes 15 cm away (or
less),  99.3% on the tube  at 20 cm distant, and  78.3% on
the  tube 25 cm distant from the paint. Slates showed high
settlement  rates and a mortality  of 82.5% at 10 cm  away
from the paint. In the Bay of Arcachon, there was  a  much
lower settlement of larvae; in the Villa  Algerienne  area
there  were  178 settlements/dm2    on the  treated  tubes
compared  with 440 on controls,  and in the Comprian  area
81  or 83 settlements/dm2   on treated tubes compared with
323  on controls.  Barnacles, which were less sensitive to
the  paint, were also found  to settle on the  tubes.  The
settlement period in this area was longer than that in the
Gironde  estuary; some larvae settling later in the season
showed reduced growth compared with controls.

    Thain  & Waldock (1986) reported that the Pacific oys-
ter was introduced into the United Kingdom during the mid-
1970s.  At this time, growth trials were performed at sev-
eral  coastal locations.  The  oysters grew well  at  most
sites, but at some east coast sites, such as  the  estuary
of the River Crouch, they exhibited poor  growth,  reduced
meat  yield, and shell thickening. At other sites, such as
the  north Norfolk coast,  there were good  growth results
with  none of the  deformities found at  the Crouch.   The
cause  of the poor growth results in some areas was inves-
tigated,  and  Key et  al.  (1976) found  good correlation
between  poor  growth and  high  levels of  fine suspended
particles.  Areas of poor  performance in the  trials were
not  used  for  the cultivation  of  the  newly-introduced
oyster  species.  The types of shell abnormality exhibited
by oysters in France were very similar to  those  observed
in  oysters from the east coast of England, which had been
attributed  to sediment.  The sediment in the marinas used
by Key et al. (1976) for resettlement studies was probably
contaminated  with  TBTO  (Personal communication  by M.J.
Waldock to IPCS). In 1982 it was decided to  reassess  the
causes of poor oyster shell growth in Britain.  Levels  of
TBT  were measured in the  estuaries of the Rivers  Crouch
and  Blackwater, on the  east coast of  England, and  were
found regularly to exceed 0.2 µg/litre,   a level shown by
the  French studies to be  harmful to the oyster  (Thain &
Waldock,  1986). In a  laboratory study, Waldock  &  Thain

(1983)  investigated the effect of both suspended sediment
and  TBT  on growth  and shell thickening  in spat of  the
Pacific oyster. They found that TBT levels of 0.16 µg  per
litre  inhibited and 1.6 µg/litre   stopped  growth. Shell
thickening  was observed in  the oysters exposed  to  TBT.
Exposure  to sediment in "clean"  water (i.e. containing
no TBT) actually enhanced growth. Thain &  Waldock  (1986)
reported  that, during 1983,  the laboratory finding  that
TBT  and not suspended  sediment was affecting  growth and
causing  shell abnormalities in oyster was corroborated in
the field by studying oysters from different sites.

    Alzieu  & Portmann (1984)  reported that although  the
findings clearly implicated TBT as a major cause of growth
problems   and  abnormalities,  they  did  not  completely
exclude the possibility that other chemicals present might
have  caused similar effects. They reviewed the effects of
a 1982 ban by French authorities on use of TBT  paints  on
boats  shorter than  25 m. In  the Bay  of  Arcachon  area
(where, in 1980 and 1981, 95% to 100% of oysters had shown
deformities), the 1982 figures for deformities were 70% to
80%  and by 1983 deformities  had declined to 45%  to 50%.
The number of oysters from the same area  showing  deform-
ities in both upper and lower shells was between  70%  and
90%,  in 1980 and 1981,  and zero by 1983.   The spatfall,
which  in  both  1980  and  1981  failed  completely, sub-
sequently  recovered; it was described as good in 1982 and
excellent  by 1983.  The  authors also reported,  however,
that  these  results were  not  reflected in  another area
(La Rochelle Bay). This was attributed both to  its  close
proximity  to a major commercial harbour and the fact that
the ban on the use of TBT antifouling paints  on  pleasure
craft in this area had not been as strictly observed as in
the Bay of Arcachon.

    His  et  al.  (1986) reported  bioassays  conducted on
oysters  (Crassostrea gigas) using sea water collected from
the Bay of Arcachon both before and after  the  imposition
of  a ban on the use of TBT paints on small boats. The sea
water  collected in 1981  caused abnormalities in  40%  of
larval  oysters over 12 days of observation (compared with
4%  in controls), whereas the sea water from 1982 produced
only 12% abnormalities over the same period. Growth of the
oyster  larvae was also  improved relative to  the  period
when  TBT paints  were still  being used.  In  1981,  mean
growth  of larval shells over 12 days was 133 µm,   in Bay
of Arcachon sea water, compared with 142 µm   in controls.
In 1982, mean growth was 162 µm,  as opposed to 168 µm  in
controls.  Alzieu et al.  (1986) stated that  between 1980
and 1982 about 90% of oysters displayed anomalies in shell
calcification.  Each  year  anomalies became  apparent  in
April  and reached maximum intensity during June and July.
During  the period 1983 to 1985, the percentage of oysters
displaying  shell anomalies fell steadily. By 1985 none of
the  oysters examined had malformations in both valves and
less  than 40% had  shell anomalies in  one of the  valves

(usually  the upper valve). Alzieu et al. (1989) monitored
oysters  in 1986 and 1987 and found that, although oysters
with deformities in both valves had stabilized  at  < 10%,
those  with deformities in  at least one  valve had  risen
again to a peak of 70% for both years.

    Effects  on oysters have  also been reported  near  to
fish  farms containing nets treated with TBT antifoulants.
Davies  et  al.  (1987a) maintained  caged Pacific oysters
(obtained from a nursery unit distant from any significant
sources  of TBT) at varying  distances from fish farms  at
Loch  Sween, Scotland. They found that significant accumu-
lation of tin was restricted to within 200 m of  the  fish
farms  and  that effects  on  oyster shell  structure were
observed at a distance of 1000 m but not at 5000 m. A com-
parison  of  shell  thickness index  and  tin accumulation
showed  significant tin accumulation in those oysters with
the most severe shell thickening.

    Studies  in the USA  and Japan have  revealed  similar
effects. Stephenson et al. (1986) transplanted culched and
culchless   Pacific  oysters  (Crassostrea  gigas) and  two
species      of     mussel  (Mytilus     edulis and Mytilus
 californianus) to areas of San Diego Bay, California, USA,
along  a  gradient  of known  sea-water TBT concentrations
(0.01-0.93 µg/litre).    Reduced shell growth was observed
in  all  three  species in  areas  where  TBT levels  were
highest.   Oyster and Californian mussel  samples (but not
 M.  edulis ) showed marked trends  of reduced growth  with
increasing  TBT  levels.  Shell thickening  in the oysters
also  correlated with increasing TBT levels.  Wolniakowski
et al. (1987) found Pacific oysters in Coos  bay,  Oregon,
USA,  to have thickened and ball-shaped shells. When these
oysters  were analysed, they  were found to  contain  high
levels  of  TBT (49.7  to  189 µg/kg).   The  most  marked
deformities  occurred in animals collected in a marina and
near  to where boats  were painted. Okoshi  et al.  (1987)
transplanted  spat  of  two different  strains  of Pacific
oysters  (Crassostrea  gigas) to two different experimental
field sites in northern Japan for 41 weeks.  The number of
chambers in the oyster shells increased gradually  in  the
Miyagi  strain at one site. In contrast, few chambers were
observed  in  the Hiroshima  strain  at either  site.  The
authors  concluded  that  both genetic  and  environmental
factors  were involved in the formation of shell chambers.
 Crassostrea   gigas is  a  non-indigenous species  in  the
United Kingdom and France. Stocks for breeding were intro-
duced into both countries in the late 1960s and originated
from the Miyagi region of Japan (Walne & Helm, 1979).

    When  Paul & Davies (1986) maintained scallops  (Pecten
maximus) and  Pacific oysters  (Crassostrea gigas) in  nets
coated with a TBTO-based paint, the mortality  of  scallop
spat  was 24%, compared with  a control mortality of  less
than 6%, over the 31-week exposure period.  Adult  scallop
mortality  was   less than that of controls in the treated

group,  and no effect was  observed on growth. No  oysters
died.  Scallop spat in TBT-treated nets grew significantly
less  rapidly  than the  control  spat. Oyster  growth, as
measured  by mean shell length,  was significantly reduced
and  had largely ceased after 10 weeks of exposure. A sig-
nificant  thickening  of  the oyster  shell  was  observed
within  10 weeks of exposure and  continued throughout the
31-week exposure. It was maintained even after transfer to
untreated nets for 10 weeks.

    Minchin  et al. (1987) monitored bivalve settlement in
Mulroy  Bay, Ireland, between 1979 and 1986 using settling
panels  placed in the sea.   They found that, during  this
period,  settlement either failed or  was reduced. Scallop
numbers fell from an average of > 1000 per panel  in  1979
to  zero in 1983.  In 1984, no  settlement of any  bivalve
species was recorded on the panels. This  reduced  settle-
ment  corresponded to the introduction  of organotin fish-
net  dips in local  salmonid farms. The  first use of  TBT
paints appears to have been in 1981. This use of organotin
compounds  ceased  after 1985.  In  1986, in  all bivalves
monitored  (except  flame  shells), there  was  again good
settlement.   Scallops  (Pecten  maximus) and flame  shells
 (Lima    hians and  Chlamys  varia) (all  members   of  the
Pectinacea)  were  found  to be  particularly sensitive to
organotin compounds. It is not known whether the effect of
the TBT was on the reproduction of the bivalves or a toxic
effect on the larvae.

10.2.  Effects on gastropods: imposex

    Between the years 1972 and 1976, Smith (1981c)
sampled mud   snails  (Nassarius  obsoletus) from four  
locations bordering  Long  Island  Sound in  Westport and
Fairfield, Connecticut,  USA.  The snails  were examined
for  imposex (see  section, and the results were
quantified by measuring the percentage of snails in a
sample showing any imposex  and estimating the degree of
imposex within indi- vidual  snails. In two of  the
areas, adjacent to  a yacht yard in Southport Harbour and
at the mouth of the harbour, 95%  to 100% of snails  had
some degree of  imposex.  Both showed  significantly 
more  imposex than  the  other  two sites,  an  area used
to moor a few old boats and an area protected  from human
interference. The mud snails in this last area showed no
imposex. The area with a few old boats showed imposex
levels of 30% to 50%, much less than in the first two
areas. Smith (1981a) collected mud  snails  from three of
the  locations, the  yacht yard  in the  harbour (where
all female snails were abnormal and degree of imposex 
greatest), the mouth of the harbour (where almost all
snails were abnormal but the degree of imposex changes
was intermediate),  and the area  protected from human 
impact (at least 3.5 km from the nearest marina, where

all female snails  were normal). The  author labelled
these  areas as "dirty", "intermediate",  and "clean". 
Snails  transferred from  the  "clean"  area  to  the 
"dirty"  area developed imposex. In those transferred
from the "dirty" area  to the  "clean" area  the degree 
of  imposex  was reduced.  An analysis of  chemicals
present in  water from the  "dirty" area was then carried
out and snails  were exposed  to  some  of the 
contaminants individually, i.e.  marina   disinfectants, 
detergents,  copper   antifouling paints,  leaded 
gasoline,  combustion emissions,  and two types  of
TBT-based antifouling paints.  Only the tin-containing 
antifouling paints increased  the level of  penis
expression in female snails.

    In  a  survey of  the dogwhelk  Nucella lapillus around
the  south-west  of England,  Bryan  et al.  (1986)  found
imposex  to be widespread.  The south coast of England was
the  most  severely  affected.   Populations  showing  the
highest  incidence and highest  intensity of imposex  were
close to areas of boating or shipping activity. The degree
of imposex had increased markedly between 1969 and 1985 in
Plymouth  Sound (an  area on  the south  coast with  large
numbers  of small boats  and ships), coinciding  with  the
introduction  and  increasing use  of TBT-containing anti-
fouling  paints  in  the  area.  Imposex  correlated  with
concentrations  of sea-water tin (TBT  fraction) and resi-
dues of tin in the dogwhelks (Fig. 6). Transferring whelks
from  an  area with  little  boating activity  to Plymouth
marina  resulted in  a marked  increase in  the degree  of
imposex.  Bailey & Davies (1988b) found an increase in the
degree  of imposex and  also a higher  incidence of  penis
development in female dogwhelks collected in 1987  in  the
Firth  of Forth, Scotland,  compared with those  caught in

    One of the effects of imposex on the  female  dogwhelk
is  the blocking of  the pallial oviduct,  preventing  the
release  of egg capsules and rendering the female sterile.
A  high incidence of females carrying aborted capsules was
found  in declining populations  close to sources  of TBT.
The build-up of aborted capsules seemed, eventually, to be
lethal  to  the  female;  there  were  fewer  females than
expected in affected areas (Gibbs & Bryan, 1986). The same
authors  reported  that  the gross  morphological  changes
occurring in late imposex in the dogwhelk seem to be irre-
versible,  since  animals  transferred from  a moderately-
contaminated site to a "clean" site showed no resorption
of  the penis. Gibbs & Bryan (1987) stated that imposex in
the dogwhelk was seen in sea water with TBT concentrations
of less than 1 ng tin/litre. They reported that the repro-
ductive failure, along with a lack of recruitment, had led
to population declines, almost to the point of extinction,
in areas of heavy TBT contamination. Gibbs et  al.  (1988)
experimentally  transferred  dogwhelks from  an uncontami-
nated  area to one showing water concentrations of 9-19 ng

tin/litre  (TBT fraction) and demonstrated the development
of  imposex within 18 months  at the new  location.  These
transferred  adults were able  to spawn for  much of  this
period.  The authors compared  these results with  results
for  juveniles, which developed  imposex earlier and  were
sterile  before  reaching  maturity.  They  discussed  the
implications  for the recolonization of areas where repro-
duction in the dogwhelk has been eliminated.   Adults  are
irreversibly   affected  by  imposex.   Recolonization  is
unlikely  until the TBT levels in sea water fall to around
2 ng  tin/litre,  a  concentration at  which the juveniles
are  not sterilized before they reach sexual maturity. The
extent of recovery of populations would, therefore, depend
on  the  success  of  control  measures  and   the   water
concentrations resulting from bigger ships exempt from the
ban on TBT-containing antifouling paints.

    Bryan et al. (1987) stated that while the evidence did
not show conclusively that TBT is solely  responsible  for
imposex  in the dogwhelk, circumstantial evidence was con-
siderable.   Imposex is related  to the level  of  boating
activity.  There  is  a significant  relationship  between
imposex  and the body  residue of organotin.   Populations
that are in decline show the highest levels of TBT. Female
populations  are declining faster than males and also have
the  higher levels  of TBT.   The rise  in the  degree  of
imposex coincided with the introduction and use  of  anti-
fouling  paints containing TBT.  Imposex has been shown to
be  caused by TBT in other species of stenoglossan snails,
i.e.  mud snails (Smith, 1981b).  A significant decline in
imposex  in monitored juvenile  Nucella (Personal  communi-
cation  by P.E. Gibbs &  G.W. Bryan to IPCS)  followed the
introduction of TBT legislation in the United Kingdom.



10.3.  Effects on farmed fish

    Wooten  et al. (1986) reported the effects of exposure
to  TBT antifouling paints  applied to retaining  nets for
farmed  salmon.  Fish exposed  to newly-treated cage  nets
were reported to be blind. The authors  examined  diseased
post-smolt  salmon from cages treated with TBT at the time
of smolt transfer. Elevated tin levels were found  in  all
salmon  exposed to the  TBT paint; liver  residues  ranged
between  1.01 and 1.62 mg  tin/kg wet weight.  Residues in
the  liver of blind fish  ranged between 1.01 and  1.07 mg
tin/kg wet weight.  Blindness had been caused by rupturing
of the eye; the eye lens was missing. Eye tissues appeared
normal  histologically apart from the  rupture; there were
no histological lesions. The kidney, liver, stomach, heart
and  muscle of the  blind fish appeared  normal histologi-
cally.  The  intestine  showed mucosal  sloughing, vasodi-
lation,  and  pyknotic nuclei.  The  spleen had  an "open
structure" with increased numbers of erythrocytes.  There
was  thickening  of  secondary  gill  filaments  and  some
necrosis and capillary separation in the epithelium. Other
fish  in  the affected  population  were reported  to have
swellings  along the lateral line. These were shown histo-
logically to be thick folds of epidermis under which there
was  cellular infiltration of necrotic collagen forming an
abscess.   Pancreas disease was also  reported in affected

fish. In the liver there was a breakdown of lobular struc-
ture with lack of cohesion between cells. The effects were
thought to be consistent with poisoning by organotin.

10.4.  Effects of TBT-contaminated sediment

    Matthiessen  & Thain (1989) studied the recolonization
by  marine organisms in the field of sediment contaminated
with  TBT-containing  paint.  Sediment was  collected from
mudflats,   and  macroorganisms  present  were  killed  by
repeated  freezing and thawing. TBT was added to the sedi-
ment  in the form of abraded paint.  The paint was abraded
with  scourers used on yachts  to produce material of  the
kind  likely to contaminate sediments  normally.  Sediment
samples  containing  0.1, 1,  10,  and 100 mg  TBT/kg were
prepared.  The contaminated sediment  was returned to  the
mudflats  and  placed  in excavated  trenches (3 m x 30 cm
wide x 20 cm deep)  lined  with  polyethylene  mesh  (5-mm
apertures).  Recolonization of the sediment  could, there-
fore,  take  place both  by  settlement of  organisms from
above  and by lateral transfer of organisms from adjoining
mud. The sites were revisited five times during  the  next
160 days  when samples were  taken both to  count recolon-
izing  organisms  and to  measure  TBT levels  at  various
depths.   Surface  sediment decontaminated  rapidly; water
movement  would have removed  sediment and deposited  new.
The  subsurface  TBT  levels  remained  reasonably  stable
throughout  the experiment. Burrowing  activity (estimated
by  counting  casts  on  the  surface)  of  the polychaete
 Arenicola   marina was  reduced  at all  concentrations of
TBT,  though the effect of 0.1 mg TBT/kg disappeared after
3 months.  A dose-related reduction in  populations of the
burrowing   polychaete  Scoloplos   armiger and   burrowing
amphipod  Urothoe  poseidonis was  observed over  the whole
concentration  range. There were no clear effects on other
species,  including molluscs. The authors pointed out that
some  unaffected groups were  associated with the  surface
sediment,  where TBT was lost  rapidly.  Feeding behaviour
of  these  surface  dwellers, such  as the cockle  (Cardium
 edule), would lead to little TBT exposure, since they fil-
ter  overlying sea water and ingest little or no sediment.
Species  associated  with  deeper layers  of  sediment and
feeding  on  fine  particles, such  as  Urothoe poseidonis,
showed   greater  effects  of TBT,  resulting, presumably,
from  greater exposure rather than greater inherent sensi-
tivity.  The authors also noted  the wide range of  sensi-
tivity  of different species  in laboratory tests  without
sediment. The bioavailability of TBT associated with sedi-
ment  would probably be  different from that  dissolved in

10.5.  Effects of freshwater molluscicides

    Following  earlier laboratory trials on the use of TBT
as a molluscicide for schistosomiasis control (Deschiens &
Floch,  1962), Deschiens et al.  (1966) applied TBTO to  a
pond  used for fish culture, in Cameroon, of approximately

1 metre in depth and with an area of  50 m2 (approximately
50 tonnes  of  water) and  a  water temperature  of 23 °C.
Caged  pond snails  (Biomphalaria [Australorbis]  glabratus
and  Bulinus   contortus), 10 to a cage supplied with food,
were  placed in the pond in different areas and at differ-
ent depths.  Renewal of cages was performed to  study  the
persistence  of the compound.  The pond was  treated  with
TBTO,  as  "Biomet"  (50% TBTO  and  50% toxicologically
inert  dispersant), at rates giving TBTO concentrations in
water  of  0.015, 0.03,  or  0.045 mg/litre, and  the side
effects on fish and plankton were investigated.  At  0.045
mg/litre,  TBTO killed 100% of  the snails and 70%  of the
fish within 24 to 48 h. At 0.03 mg/litre, TBTO killed 100%
of  snails within 2 to  3 days, but no effect  on the fish
was  observed within 15 days.  At 0.015 mg/litre, the TBTO
killed  all of the snails within 4 to 7 days and again had
no  observable effect on the fish.  The authors considered
this  to  be a  practical  concentration to  achieve  full
effect  on  snails  without killing  non-target organisms.
Various  planktonic  organisms  (desmids,  daphnids,   and
aquatic insect larvae) were killed by TBTO at 0.5 mg/litre
experimentally.  However, no effect  was found on  any  of
these  species  in  the pond  treated  at  0.015 mg/litre,
although  there  was some  inhibition  (but no  deaths) of
these organisms at 0.03 and 0.045 mg/litre.

    Gilbert  et  al.  (1973) applied  slow-release TBTO in
underseal (quick-drying asphalt-rubber-asbestos-clay paste)
to four sites, in Brazil, at concentrations of 15, 15, 50,
or  300 g  TBTO/m2,   to  control  Biomphalaria tenagophila
and  B.   straminea. Dead fish were  observed at the  sites
treated with 50 or 300 g/m2   and also at one of  the  two
sites  treated with 15 g/m2,    a static man-made  pit. At
the  other site, with a flow rate of 100 to 200 litre/min,
there  was a 100%  reduction in the  snail  population  (B.
 tenagophila) maintained  for  up to  12 months, while both
fish  and aquatic insect  life were apparently  normal one
month  after treatment. At this site abundant natural veg-
etation flourished in the immediate vicinity of  the  mol-
luscicide.  Toledo et al. (1976) found that  15 g  TBTO/m2
eliminated  snails from 76%  of treated sites  for 1 year.
Guppies, which are often present in highly polluted sites,
normally  suffered some mortality at the beginning of mol-
luscicidal  treatment  but  in general  they  tolerated it
well.  The authors  found no  lasting effect  on plant  or
insect life, and concluded that, because of the ability of
an  unidentified fungus to grow profusely, microbiological
life continued in treated areas.

    Shiff et al. (1975) applied BioMet SRM  pellets  (con-
taining TBTO) at concentrations of 20 and 30 g/m2    to  a
night-storage  dam in Zimbabwe.  At the lower  application
rate,  pellets remained active against the snails (various
 Biomphalaria spp.)  for up to one year. At 30 g/m2,   100%
mortality  of snails was achieved within 2 weeks of appli-
cation  and was maintained up until the last sampling time

2 months   after  application.  Prior  to   treatment,  80
specimens  of  Tilapia fish sp. and 1 of  Clarias gariepinus
were   caught.  Four weeks after treatment 26  Tilapia were
caught  and one dead  Clarias was found.  Eight weeks after
treatment   400  Tilapia were  caught,  of   which  80 were
examined and all appeared normal.

    Ayala  et al. (1980) applied slow-release molluscicide
(rubber  impregnated with 11% TBTO) at a rate of between 5
and 15 g/m2 to  a site in Brazil and studied the effect on
non-target  organisms. The TBTO molluscicide was initially
herbicidal  towards floating vegetation but did not affect
either  plants rooted in the  mud or marginal plants.  The
chemical  was initially repellent to aquatic animals, some
of which were killed by the molluscicide. Within 3 months,
populations  of  non-target  snails  such  as  Pomacea sp.,
 Drepanotrema sp.,  and a  Physa sp. had returned to normal.
Although  no fish had  returned within 3 months,  after  5
months numbers had returned to normal. In  fact,  5 months
after  application all but  Spirogyra sp. were  normal, and
the  molluscicidal activity of  the TBTO was  retained for
more than one year.  The authors concluded that after 3 to
5 months the action of the TBTO molluscicide  is  confined
to the bottom mud where  Biomphalaria glabrata snails spend
much of their time.

10.6.  Effects from spills

    According to Waldock et al. (1987a), major  inputs  of
TBT  into the environment may  have arisen from spills  of
timber-treatment  products, often containing  dieldrin and
pentachlorophenol  as well as the organotin. They reported
that a detailed study was carried out after a  spill  con-
taminated a section of the Newmill Channel,  Kent,  United
Kingdom. There was a major kill of fish and  all  macroin-
vertebrates  (except oligochaetes, chironomid  larvae, and
elmid beetles) were also killed.  Concentrations of TBT in
the water shortly after the spill were 540 µg   per litre,
falling  to 10 µg/litre   after 3 weeks  and 0.75 µg   per
litre  after  6 weeks.   Nine months  after  the incident,
macroinvertebrate   populations  were  reported   to  have
recovered.  The authors stated that five such major spills
from  timber  treatment plants  had  been reported  in the
United Kingdom within 2 years.

10.7.  The use of indicator species for monitoring the environment

    Both  shell  growth  and shell  chambering  in Pacific
oysters  and imposex in dogwhelks  have been used as  bio-
logical  indicators  of  TBT contamination.  Smith  et al.
(1987)  transferred juvenile Pacific oysters to major bays
and harbours of California, USA, known to contain elevated
levels  of TBT. They observed a graded increase of stunted
growth and/or shell deformation, which was associated with
poor flushing and the proximity of large numbers of boats.
Gibbs  et al. (1987)  used the dogwhelk  to monitor  water

around the south-west peninsula of England. They found the
species  to be a  very sensitive indicator,  especially if
the  individuals were about  1 year old and  in the  early
stages of ovarian development. Davies et al. (1987a) found
imposex  in dogwhelks to be a far more sensitive indicator
than  either shell thickening  in oysters or  tin  accumu-
lation.  The dogwhelk has been used to identify  areas  of
contamination associated with seasonal small boat activity
and  salmon farm cages  in Scottish sea  lochs (Davies  et
al.,  1987b), in areas of pleasure craft activity, fishing
harbours  and a boat yard in the Firth of Forth (Scotland)
(Bailey & Davies, 1988b), and to investigate contamination
from  an oil terminal  in Sullom Voe,  Shetland (Bailey  &
Davies, 1988a).


11.1.  Single exposure


     The acute toxic effects of the various TBT  compounds  that
 have  been  studied  are comparable  and are characteristically
 delayed for several days. Oral LD 50    values in laboratory ani-
 mals  range  from approximately  40  to 250 mg/kg  body weight.
 These  compounds exhibit greater lethal potential when adminis-
 tered  parenterally,  relative  to  the  oral  route,  probably
 because they are only partially absorbed from the  gut.   Acute
 toxicity via the dermal route is low.
     TBT  compounds are potent  skin irritants and  extreme  eye
 irritants.  Dermal  exposure to  TBTO  appears to  have  little
 sensitization potential.
     Aerosols  of TBT compounds  are highly toxic.  However, TBT
 vapour/air mixtures at room temperature produce no effect, even
 at saturation.
     Other  effects of acute exposure may include alterations 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  compounds,  is questionable  and  the  cause of  death
 remains unknown.

11.1.1.  Oral and parenteral administration

    The  acute toxicity of tributyltin  to laboratory mam-
mals  by various routes of administration is summarized in
Table 14. The acute oral LD50   for the rat ranges between
94  and 234 mg/kg body weight and for the mouse between 44
and  230 mg/kg body weight.  Truhaut et al. (1976) pointed
to the delayed toxicity of tributyltin and, therefore, the
necessity  to continue the observation of animals, after a
single acute dose, for several days. The period  of  post-
dosing observation is therefore indicated in Table 14.

    LD50   values for ip and iv administrations of TBT are
very  much lower than  in the case  of the oral  route (10
mg/kg for rat ip and 6 mg/kg for mouse iv).

Table 14.  Acute toxicity of tributyltin to laboratory mammals
Species TBT          Route  Observation  LD50 (mg/kg   Reference
                            perioda      body weight)
Rat     oxide        oral       7      194 (165-227)b  Elsea & Paynter (1958)
        oxide        oral       7      148 (113-195)c  Elsea & Paynter (1958)
        oxide        oral       7      180 (132-228)   Truhaut et al. (1976)
        oxide        oral      21      197 (137-273)   Funahashi et al. (1980)
        oxide        oral                   127        Schweinfurth (1985)
        oxide        ip        14        20 (18-21)    Poitou et al. (1978)
        oxide        oral                   234        Sheldon (1975)
        fluoride     oral                   200        Sheldon (1975)
        fluoride     oral      14            94        Schweinfurth (1985)
        chloride     oral      14           122        Schweinfurth (1985)
        acetate      oral                  113.5       Klimmer (1969)
        benzoate     oral                   141        Klimmer (1969)
        benzoate     oral      14          99/203      Schweinfurth (1985)
        oleate       oral                   225        Klimmer (1969)
        linoleate    oral      14           190        Schweinfurth (1985)
        abietate     oral      14           158        Schweinfurth (1985)
        naphthenate  oral      14           224        Schweinfurth (1985)

Mouse   oxide        oral       7       85 (52-130)    Polster & Halacka (1971)
        acetate      oral       7        46 (25-85)    Pelikan & Cerny (1968)
        oleate       oral       7      230 (175-301)   Pelikan & Cerny (1968)
        benzoate     oral       7       108 (74-156)   Pelikan & Cerny (1968)
        chloride     oral       7       117 (80-170)   Pelikan & Cerny (1968)
        laurate      oral       7      180 (136-237)   Pelikan & Cerny (1968)
        oxide        sc         7      200 (140-270)   Polster & Halacka (1971)
        oxide        iv         7       6 (5.5-6.5)    Truhaut et al. (1976)
        oxide        ip        14        16 (15-17)    Poitou et al. (1978)

Rabbit  fluoride     dermal                 680        Sheldon (1975)
a Following a single application of TBT, the observation period was as indicated.
b Application as aqueous suspension.
c Application in corn oil.
    Pelikan  & Cerny (1968) administered TBT (as the acet-
ate, benzoate, chloride, laurate, or oleate) to white mice
(body  weight 25 g) in  a single oral  gavage dose of  500
mg/kg body weight, dissolved in sunflower oil. Ten animals
were  used for each TBT  compound. The mice were  observed
for 8 h before being killed, and the viscera were examined
histologically.  Four  hours after  treatment, all animals
showed signs of intoxication with the exception  of  those
given  the laurate; after  8 h all showed  toxic symptoms.
Gross damage was seen in the digestive tract,  liver,  and
spleen.  Histological findings included a steatosis of the
liver  cells  in all  animals  (but to  varying  degrees),
traces  of lipid in renal  tubule cells (in those  animals
receiving  oleate or laurate) and many haemorrhages in the
digestive  tract and kidneys.   Effects on the  liver  and
spleen  were also noted  after dermal absorption  of  TBTO

(Pelikan  & Cerny, 1969). However, no real conclusions can
be drawn from these studies as there was no clear reported
evidence that steatosis was caused either by the TBT or by
the fatty acids.

    In  studies by Funahashi et al. (1980), Sprague-Dawley
rats  were given a  single dose, in  olive oil, of  100 mg
TBTO/kg body weight by intubation and were  examined  over
the   following  21 days.   Adrenal  weight  was  slightly
increased  12 h after treatment  and reached a  maximum on
the second day.  Histological changes in the  adrenal  had
returned to normal by day 14. The thyroid follicles showed
signs  similar to those  produced by hypophysectomy,  i.e.
distension  with  colloid,  flat epithelial  cells.  These
changes were severe after 72 h, but had returned to normal
within  14 days.  Absolute pituitary weight  was increased
slightly  (and  not  significantly)  1  and  2 days  after
dosing,  while  relative  weight increased  significantly.
There was atrophy of pituitary adrenocorticotrophin (ACTH)
cells 6 h after dosing. After 72 h, the staining  of  ACTH
cells  became more intense.  There  were marked reductions
in  the  circulating  levels of  both  thyroid stimulating
hormone (TSH) and thyroxine (T4)  during the 72 h follow-
ing dosing, serum titres falling to one half and one sixth
of control values for the two hormones, respectively.  The
authors  concluded that  TBTO had  both a  direct  and  an
indirect  effect on the thyroid, since initially there was
a simultaneous rise in T4 and  fall in TSH. Serum cortisol
levels  increased to twice  the control level  96 h  after
treatment with TBTO at a level of 100 mg/kg  body  weight.
Intramuscular  ACTH administration, 8 h after  dosing, led
to  increased cortisol in  the blood of  both treated  and
control  rats.  After  16 h, controls  showed  release  of
cortisol after ACTH stimulation, but treated rats showed a
decrease  in circulating levels of  cortisol after similar

    Matsui  et al. (1982) reported the effects of a single
dose  of  TBT fluoride  (100 mg/kg  body weight)  given by
gastric intubation to Japanese white rabbits. A reversible
but  pronounced hyperlipidaemia was observed, particularly
involving  triglycerides and total  cholesterol. Ultracen-
trifugation  showed a marked  increase in the  chylomicron
plus  VLDL (very low  density lipoprotein) fraction.   The
activity of lipoprotein lipase (LPL) in plasma was reduced
to  about 50% of control  levels. Fasting levels of  blood
glucose  were elevated, and the response to iv infusion of
glucose   (given  3 days  after  the   TBT  fluoride)  was
inhibited (insulin release reduced). The authors suggested
that  the hyperlipidaemia was  the result of  reduced  LPL
activity, which in turn was brought about by inhibition of
insulin release.

    Calley et al. (1967) investigated the effects  of  TBT
acetate  on liver function in rabbits.  Following a single
oral  dose of TBT acetate at 50 mg/kg body weight (half of
the  measured LD50   in rabbits), only the serum glutamic-
pyruvate  transaminase (SGPT) activity was  affected. This

activity  was not elevated  after 48 h but  was  increased
within  144 h of dosing. Prothrombin  time, alkaline phos-
phatase,   and  thymol  turbidity  showed  no  significant
changes  from control values.  SGPT is a  highly  specific
indicator  of liver damage, directly reflecting liver cell
injury.  Tributyltin acetate also  significantly increased
the  hexabarbitol-induced  sleeping  times of  mice;  a 50
mg/kg  body weight oral  dose increased the  sleeping time
from  29  to  43 min following  a  standard  dose  of  the

    When Aldridge et al. (1977) administered TBTO (and its
gamma-keto,  gamma-hydroxy, and delta-hydroxy metabolites)
intraperitoneally to mice on two consecutive days (at dose
levels  of 12.5, 25,  and 50 µmol/kg   body  weight), they
found  no evidence  of cerebral  oedema after  any of  the

    Crofton et al. (1989) gave rat pups a single  dose  of
TBTO  on day 5 following birth by  gastric intubation with
0,  40, 50, or  60 mg/kg body weight.   The monitoring  of
behavioural parameters up to 62 days post-dosing showed no
persistent  effect on motor  activity or acoustic  startle

    When  Barnes & Stoner (1958) fed rats with tributyltin
acetate  during the first 3 months after birth, they found
increased  brain  and spinal  cord  water content  at  the
highest  dose tested (100 mg/kg diet).   This increase was
not  sufficiently great to  be seen in  histological  sec-
tions.  Bouldin et al. (1981) found no light  or  electron
microscopic evidence of neuronal damage in the hippocampus
or  the pyriform cortex of  neonate or adult rats  exposed
daily  to tributyltin acetate  at 10 mg/kg body  weight by
gavage for up to 30 days.

    O'Callaghan  & Miller (1988) reported the effects of a
single  ip injection of  TBTO on neonatal  rat brain.  The
rats  were injected when 5 days old with 2, 3, or 4 mg/kg.
They  were sacrificed at  13, 22, or  60 days of age,  and
various  proteins  in  homogenates of  brain  tissue  were
measured  by radioimmunoassay. Brain weight was reduced in
a   dose-dependent  manner,  the  cerebellum   being  most
affected. There was no evidence of altered brain histology
under  the  light  microscope. Dose-dependent  and region-
dependent  decreases were found in  P-38 (a synaptic-vesi-
cle-associated   protein)  and  myelin  basic  protein  (a
protein  associated with oligodendroganglia and the myelin
sheath); there were decreases in both total  (per  tissue)
and concentration (per mg) levels of these proteins in the
cerebellum  and forebrain but  not the hippocampus.  These
effects were seen at a dose level that did not affect body
weight.  However, in contrast to the neurotoxic effects of
trimethyltin compounds, which are irreversible and persist
into  adulthood,  these  neurotoxic effects  of  TBTO were
transitory at dose levels that did not affect body weight.

    Robinson  (1969) monitored tissue levels  of catechol-
amines  after ip administration of  TBTO, in corn oil,  to
rats  at  a  level of  10 mg/kg  body  weight.  This  dose
exceeded  the 6-day LD50,   determined by the same authors
to  be 7.21 mg/kg. Brain noradrenalin levels were signifi-
cantly  reduced at 2, 24,  and 48 h after dosing,  whereas
5-hydroxytryptamine  levels  were  only  reduced   (though
markedly) after 48 h.  Noradrenalin levels in heart tissue
were  also significantly reduced  after 2, 24,  and  48 h.
Adrenal  adrenalin  and  noradrenalin levels  were reduced
after 24 and 48 h, but only adrenalin was  affected  after
2 h.

11.1.2.  Dermal administration

    The acute LD50   of TBTO administered to  rabbits  via
the  dermal route is  very high, i.e.  approximately  9000
mg/kg  body weight (Table 14).   Although there is  dermal
absorption of TBT (see later), the degree of absorption is
not  great enough  to lead  to acute  toxic  effects  sys-
temically, except at very high exposure levels.

11.1.3.  Administration by inhalation

    In  a "nose only"  inhalation exposure (to  minimize
dermal  exposure  and exposure  through ingestion) lasting
4 h, the acute LC50   of TBTO for the rat was estimated to
be  77 mg/m3   by measurement  of airborne droplets.  This
value decreased to 65 mg/m3 when  "inhalable" particles,
10 µm   or less in diameter, were considered.   There  was
evidence  of  lung irritation  and  oedema in  this  study
(Schweinfurth, 1985). When groups of 10 male and 10 female
rats  were each exposed  to atmospheres containing  almost
saturated  vapours of TBTO, TBT benzoate, or TBT naphthen-
ate once for 7 h, no deaths occurred during exposure or in
a 14-day observation period following exposure. Only minor
clinical  signs were noted occasionally  during the exper-
iment, such as slight nasal discharge (Schweinfurth, 1985).

    Truhaut  et al. (1979) exposed  mice to an aerosol  of
TBTO in olive oil, for either a single 1-h period or seven
1-h  periods on successive days, using TBTO concentrations
in  air  ranging  between  0.05  and  0.4 mg/litre  (50 to
400 mg/m3).    Exploratory behaviour was scored over 5-min
periods 2 h after the single exposure was complete or 24 h
after  the last of the  seven exposure periods. The  lower
two  exposure doses caused significant increases in explo-
ratory behaviour (17% and 5% for 42 and 84 mg/kg, respect-
ively) while the higher exposure doses reduced exploratory
behaviour  (-18% and -38% for  170 and 340 mg/kg, respect-
ively). Truhaut et al. (1981) found a median survival time
for  mice of  22 min and  for guinea-pigs  of 9 min  after
exposure  to an aerosol of  tributyltin (concentration not
stated).  They reported that only tissues in the respirat-
ory  system showed significant lesions.  There was diffuse
congestion of the pulmonary blood vessels extending to the

septal  capillary  beds.   There  were  also  inflammatory
responses  in the trachea and bronchii, secretion of mucus
in the bronchii and bronchioles, and distension  and  rup-
ture of the alveoli.

    Anger et al. (1976) exposed guinea-pigs to aerosols of
TBTO ranging between 0.1 and 1 mg/litre air for  1 h.  All
males  exposed to  0.2 mg/litre died,  and 12  out  of  15
females exposed to 1 mg/litre died.  No particular lesions
were  observed in those animals that died, with the excep-
tion  of a general  congestion of the  lungs. Exposure  of
either males or females to 0.17 mg/litre caused no deaths,
though  nasal irritation was  observed; all these  animals
survived a further 7 days of observation.

11.1.4.  Irritation and sensitization  Skin irritation

    In a study by Elsea & Paynter (1958),  undiluted  TBTO
was applied to the closely-clipped skin of  the  abdominal
area of rabbits, which was then covered with rubber dental
damming,  gauze,  and  adhesive tape.  After  an  exposure
period of 24 h, the covering was removed, and the TBTO was
washed  off, as  far as  possible, by  sponging with  warm
water.  This single application of doses up  to  11.7 g/kg
body  weight caused some  of the rabbits  to die from  the
effects  of TBTO absorbed into  the body. There was,  how-
ever,  only  moderate dermal  irritation, characterized by
reddening,  oedema, atonia, blanched  areas, and areas  of
brown  discolouration.  Examination of the skin at autopsy
showed  some subcutaneous oedema.  Repeated  dermal appli-
cation of TBTO, impregnated into paper at  8 mg/kg,  daily
for  5 days, only affected one  animal out of four.   This
rabbit  showed  slight  erythema and  oedema following the
first  and second applications.  Later in the  experiment,
the skin appeared normal. No detailed histological examin-
ation of the skin was carried out.

    Pelikan & Cerny (1969) applied to the shaved  skin  of
rats  two preparations of TBTO,  i.e. "Lastanax T" (con-
taining 20% TBTO plus a medium of water, alcohols of short
chain   length,   and  n -alkyl-polyethylene    oxide)   and
"Lastanax P" (containing 15% TBTO plus a medium also con-
taining  bis-(5-chloro-2-hydroxyphenyl)-methane).    Actual
doses  applied, corresponding to water  dilutions of 100%,
33%,  10%, and  1% of  the preparations,  were 185,  61.5,
18.5, and 1.85 mg TBTO/kg body weight for  Lastanax T  and
145,  47,  14, and  1.4 mg  TBTO for  Lastanax P. Controls
received  either  water or  the  medium without  TBTO. The
experiment  was  carried out  in  duplicate. In  the first
series,  clinical observations were  made twice daily  for
60 days,  whereas in the second series rats were killed on
day 10 and  the skin was examined  histologically. Control
rats  showed no clinical  or histological effects  on  the
skin.  However,  clinical  and histological  changes  were

found  in all  treated rats,  even with  the 1%  dilution,
though severity increased with exposure dose. On the first
and  second days after application, there was reddening of
the  skin and oedema  developed. From day 3,  haemorrhagic
eschars  developed with well defined borders; there was no
inflammation  of  the  surrounding areas  of skin. Between
these  foci, numerous papules and pustules developed, some
with bleeding; the apices of some papules  were  necrosed.
Later  the  eschars  joined  to  create  larger  areas  of
ulceration.   Healing of the  areas began between  the 9th
and  12th days; necroses disappeared by day 20, pustules 3
to 4 days before, and papules by the 30th to 33rd day. All
signs had disappeared after 35-38 days (exposure to 1% and
10%  solutions) or after  45-50 days (exposure to  33%  or
100% solutions). Changes with Lastanax P were similar, but
less  severe, than with  Lastanax T.  In the  histological
study,  numerous large bullae, filled  with leucocytes and
coagulated   exudate  (proteins),  were  found  under  the
stratum corneum of the epidermis after 10 days. Akanthosis
and  vacuolation  of  epidermal cells  occurred  following
exposure  to the 10%  and 33% solutions  and, to a  lesser
extent,  the 1% solution.  Small  haemorrhages were found.
Epidermal changes were dispersed and alternated with areas
of skin appearing normal. The authors suggested  that  the
TBTO NOEL for human skin should be set at 0.005% to 0.01%.

    In a study by Middleton & Pratt (1977),  TBT  chloride
was  applied  to  a shaved  area  of  dorsal skin  of male
Alderley  Park (Wistar-derived) rats, four-weeks old (body
weight  50-80 g),  in absolute  ethanol  as a  solution of
10 mmol/litre.   This dose was equivalent  to   67 nmol/cm2.
The  TBT  had produced  microscopic  changes in  the  skin
within  2 h  of application.  Polymorphonuclear leucocytes
accumulated  in capillaries and in dermal tissues. Epider-
mal  cells showed progressive degenerative changes between
2  and 8 h after  application until, by  8 h after  appli-
cation,  separation  of  the  epidermis  and  dermis   had
occurred  and  fluid  collected in  this separation.  Many
inflammatory cells were present in both dermis and epider-
mis.  There was widespread epidermal necrosis within 12 to
24 h, and separation of necrotic epidermis from the dermis
was almost complete by this time. The vesicles  formed  by
this  separation were frequently packed  with inflammatory
cells  as well as  exudative fluid.  Regeneration  of  the
epidermal layer was observed after 18 to 24 h  and  dermal
inflammatory  infiltration  was  regressing by  this time.
Erythema,  visually assessed and scored, reached a maximum
within 5 h of application and remained at this  level  for
48 h. The erythema had subsided by 72 h after application.
Vascular   permeability  of  the  skin   was  assessed  by
injecting  rats with a  dye, tryphan blue,  45 min  before
sacrifice.  There was a biphasic response.  At  2 h  after
application of TBT chloride, there was an initial increase
in permeability of the skin to 135% of control  levels.  A
second  peak began  at  12 h and persisted  for more  than
25 h after application. The effect was still evident after

48 h  and had occurred in  both the treated and  untreated
flanks  of the animal,  indicating absorption of  the com-
pound and a systemic effect. The water content of the skin
was  increased by TBT chloride, reaching a peak within 2 h
of  application;  by  10 h this  effect  had  disappeared.
Middleton  & Pratt (1978) reported that TBT produced focal
epidermal  necroses and dermal  inflammation at levels  as
low  as 33 nmol/cm2,   fairly extensive epidermal necroses
and dermal inflammation at 67 nmol/cm2,   and almost total
epidermal necrosis at 167 nmol/cm2.  Eye irritation

    Pelikan  (1969) applied TBTO, as  "Lastanax T or P",
in  a single dose of  0.03 ml to the left  eye of rabbits.
Lastanax T  is 20% TBTO in alcohols of short chain length,
non-ionic surface active substances    ( n -alkyl-polyethyl-
ene oxide) and water. Lastanax P is 15% TBTO in  the  same
vehicle  with the addition of   bis-(5-chloro-2-hydroxyphe-
nyl)  methane. A 10% and a 1% solution in water were used.
The  10% solution represented an  actual TBTO dose of  6.1
and  4.6 mg/kg  body  weight  for  the  two  formulations,
respectively.   With the 10% solution,  both rabbits died,
11 and 12 days after application. Severe ulceration of the
eye  preceded  death.   Histopathological  examination  of
internal  organs  after  death revealed  hyperaemia of the
brain  and medulla oblongata and  hyperplasia of reticulo-
endothelial  cells of the  spleen.  Necrotic changes  were
seen on the cornea of eyes treated with the  1%  solutions
within  3 h of application, and symptoms worsened over the
next 2 to 5 days. Recovery was incomplete 100 days later.  Skin sensitization

    Poitou et al. (1978) investigated the skin-sensitizing
potential  of  TBTO  in guinea-pigs  using  the Magnussen-
Kligman method.  The concentrations used for sensitization
were  1% (intradermal phase) and 5% (topical phase). Using
challenge concentrations of 0.25% and 0.1%, no sensitizing
action was demonstrated in the 20 test animals.

11.1.5.   In vitro studies

    Johnson  & Knowles (1983) demonstrated that incubation
of  rat  blood platelets  with  TBT chloride  or  TBTO,  in
 vitro, inhibited their ability to take up 14C-labelled  5-
hydroxytryptamine   (5-HT).  The  organotin   also  caused
release  of 14C-labelled   5-HT, taken up before exposure,
along  with  endogenous 5-HT.  Treating  rats ip  with TBT
chloride  at 5.0 mg/kg body weight also led to reduced up-
take  of 5-HT (37% inhibition) 30 min after treatment. The
level of 5-HT in platelets, however, was unaffected by the
 in vivo treatment with TBT  chloride.   knowles &  Johnson
(1986)  reported  that exposure  of  rat platelets  to TBT

chloride  inhibited aggregation induced with ADP or colla-
gen.  The inhibition of ADP-induced aggregation was depen-
dent  on the dose of TBT and on the exposure time prior to
ADP addition. Exposure of the platelets to 5 µmol  TBT per
litre  for 5 min or  10 µmol/litre   for 0.1 min  produced
inhibition.  However, exposure to 1 µmol/litre   for 5 min
was  ineffective. With an  incubation time of  1 min, TBTO
was  effective in lengthening  the time taken  for  aggre-
gation after collagen induction at 0.625 µmol/litre,   but
not at 0.5 µmol/litre.

    The  trialkyltins,  including  tributyltin, have  been
shown  to inhibit oxidative  phosphorylation in rat  mito-
chondria   (Aldridge  &  Street,  1964).   TBT  stimulated
adenosinetriphosphatase   (ATPase)  activity  and   caused
limited swelling of rat liver mitochondria. These last two
effects occurred at similar concentrations of TBT  in  the
medium. The authors postulated that the two  effects  were
linked; the ATPase activity increased over a concentration
range  of TBT and was always mirrored by a decrease in the
mitochondrial   swelling.   This  relationship   persisted
despite a complex response to TBT over  the  concentration
range; at higher concentrations there was a  sudden,  much
smaller  mitochondrial swelling and a  concomitant rise in
ATPase  activity. TBT also inhibited the hydrolysis of ATP
by  rat  brain  microsomes,  although  the  concentrations
required  were much higher than those affecting phosphory-
lation.  The authors postulated that a combination of tri-
alkyltin  compounds  with  negatively  charged  lipids  is
involved in their biological activity.

    Elferink  et  al.  (1986)  treated   polymorphonuclear
leucocytes  (PMNs), obtained from the peritoneal cavity of
rabbits,  with an unspecified  tributyltin   (1 µmol/litre).
Uptake  of opsonized  zymosan was  used as  a  measure  of
phagocytosis,  and enzyme (lysozyme) release to the super-
natant  during incubation was  also monitored.  There  was
almost   complete   inhibition   of  both   parameters  at
10-6 mol  TBT/litre but no effect at 10-7 mol/litre.   In-
hibition  of phagocytosis was exactly  paralleled by inhi-
bition  of  enzyme  release. At  concentrations of between
10-6 and  10-5 mol/litre,  tributyltin caused lysis of the
cells with release of LDH, suggesting damage to the plasma
membrane. Exocytosis, induced by FMLP in the  presence  of
"cytocolasin B", was also inhibited by the TBT  at  con-
centrations  of 10-6 mol/litre   or more. There was little
effect  of  the compound  on ATP levels  in PMNs, and  the
authors  suggested  that interference  with ATP production
was not the basis for the effect of the TBT. Activation of
PMNs  is  accompanied by  an  increase in  plasma membrane
permeability to Ca2+;   this was strongly inhibited by the
TBT.  The  authors proposed  two alternative explanations.
Either the Ca2+   permeability change is directly affected
or  an earlier step is  inhibited which, in turn,  affects
calcium  permeability. The observation that the exocytosis
effect  could be counteracted by the addition of sulfydryl
compounds  led the authors  to conclude that  the earliest

stages  of activation of the PMNs were affected. Arakawa &
Wada  (1984)  reported  a suppression  of  the chemotactic
response  of rabbit leucocytes (neutrophils)  towards for-
myl-methionyl-leucyl-phenylalanine by tributyltin chloride
at concentration between 0.1 and 10 µmol/litre  in vitro.

    Reinhardt  et  al.  (1982) used  cell  detachment  and
cloning  efficiency of baby  hamster kidney cells  (BHK-21
C13)  to quantify the cytotoxicity of organotin compounds.
The  two  parameters are  independent,  but each  covers a
range  of cellular damage. Cell detachment indicates irre-
versible  damage to the  cytoskeleton over a  4-h  period,
while  cloning  efficiency  covers influence  on growth of
single  cells over a  6-day period and  includes cell  re-
attachment  followed by clone formation. The  IC50 (concen-
tration  at which  there was  a 50%  reduction in  cloning
efficiency) for TBTO was 5 x 10-7 mol/litre  (0.3 mg/litre
medium) and for TBT chloride was 1.4 x 10-6 mol  per litre
(0.5 mg/litre). The CD50   (50% effect on cell detachment)
was  less sensitive to TBT,  i.e. 3 x 10-5 mol/litre   for
TBTO  (18 mg/litre)  and  4.3 x 10-5 mol/litre    for  TBT
chloride (14 mg/litre).

    When  Snoeij  et  al. (1986a)  incubated  isolated rat
thymocytes  with TBT chloride,  there was membrane  damage
and  disintegration of the cells  at concentrations higher
than  1 µmol/litre.   Various parameters of  cell function
were investigated. The TBT chloride increased the consump-
tion of glucose in a dose-related manner at concentrations
in  the culture medium higher  than 0.1 µmol/litre;   pro-
duction of lactate was also increased in parallel. The ef-
fect peaked at 1 µmol   TBT/litre and thereafter declined;
damage to cell integrity occurred at these concentrations.
Over  the same effective  range of concentrations  (0.1 to
1 µmol/litre),    TBT decreased ATP levels. The ATP dimin-
ished very rapidly, within 2.5 min, and remained  low  for
at  least 3 h.  The incorporation of radiolabelled nucleo-
tide precursors into DNA and RNA and of amino  acids  into
protein  was  also affected.  Thymidine incorporation into
DNA was reduced, relative to TBT dose, to a minimum of 20%
of  control incorporation at 1 µmol    TBT/litre.  Uridine
incorporation was similarly reduced, but less effectively.
Inhibition  of amino acid incorporation was substantial at
0.25 µmol    TBT chloride/litre and protein  synthesis was
virtually  totally inhibited at 1 µmol/litre.   The median
inhibitory concentrations (IC50)   for thymidine, uridine,
proline,  and leucine were 0.32 ± 0.04,  0.95 ± 0.06, 0.38
± 0.04,  and  0.36 ± 0.03 µmol/litre,   respectively.  TBT
chloride  concentrations of 0.1 µmol/litre   or more mark-
edly  reduced  the  production  of  cyclic  AMP   by   the
thymocytes  under stimulation from prostaglandin E1.   TBT
chloride,  therefore,  has  marked effects  on  the energy
metabolism  of isolated thymocytes at  concentrations well
below  those  affecting  membrane  integrity,  and,  as  a
result,  the  authors stated  that  TBT chloride  is  best
characterized as an energy poison.

    Snoeij  et al. (1988a)  similarly isolated rat  thymo-
cytes,  but separated them  into fractions based  on size.
The  same parameters of cell function were investigated in
each  of three fractions.  Fraction 1 consisted of  small,
non-proliferating  thymocytes.  Fractions 2  and 3  showed
some overlap in size, but fraction 2 was enriched in cells
actively synthesizing macromolecules, while fraction 3 was
enriched  in dividing cells  and showed the  greatest pro-
liferative  capacity. Resting ATP  levels were highest  in
the  bigger  cells,  steady state  levels  increasing with
fraction  number.  TBT chloride reduced ATP levels in each
subfraction proportionally to between 62 and 64%  of  con-
trol  values. Thus, TBT reduces  intracellular ATP concen-
trations  irrespective of cell  volume or number  of mito-
chondria.  As with unfractionated thymocytes, TBT chloride
at  both  0.25  and 1.0 µmol/litre    inhibited  thymidine
incorporation   (to  around  40%  of   control  values  at
0.25 µmol/litre   and between 21 and 37% of control values
at 1 µmol/litre)  and leucine incorporation (to between 13
and  29% of controls at  0.25 µmol/litre   and 1 to  3% of
controls at 1 µmol/litre)   in all fractions.  In the case
of uridine incorporation, although there was inhibition in
most cases, 0.25 µmol   TBT chloride/litre caused a stimu-
lation to 184% of control levels in fraction 1,  the  non-
proliferating cells.

11.2.  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

     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
 per 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, weak-
 ness, 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  pitu-
 itary and thyroid, have been noted in both short- and long-term
 studies.   Changes  in  circulating hormone  concentrations 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/m 3    produced high mor-
 tality,  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/m 3 )  at room  temperature produced no

     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 hepatocytes was noted  in dogs fed a  dose of 10
 mg/kg  body  weight for  8 to 9 weeks.  These changes were  oc-
 casionally  accompanied by increased liver weight and increased
 serum activities of liver enzymes.
     Decreases in haemoglobin concentration and erythrocyte vol-
 ume  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  alterations  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  investigations
 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  func-
 tion  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  lymphocyte
 counts,  and  total  serum immunoglobulin  concentrations) 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 tributyl-
 tin  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 re-
 sistance studies. Decreased clearance of Listeria monocytogenes
 was   seen after exposure to  a dietary 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  ani-
 mals, 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 sys-
 tem, there is no information on host resistance.

     It  would be prudent  to base assessment  of the  potential
 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  sig-
 nificance  of these data for  human risk assessment is  contro-
 versial.  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.

11.2.1.  Oral dosing: general body effects

    Iwasaki  et  al.  (1976)  reported  some  oedema   and
destruction of nerve axons following the administration of
0.01 ml  TBTO/kg directly to the stomach of rats every day
for  28 days.  No details  of  procedures or  results were
presented in this report.

    Schweinfurth  (1985) found no evidence of brain oedema
after  dosing rats with TBTO  orally in arachis oil.  Even
doses producing marked toxic effects on other  organs  (25
mg/kg  body  weight)  failed to  produce  noticeable brain
oedema.   However,  triethyltin  chloride  produces  brain
oedema at a dose of 1.5 mg/kg body weight.

    Krajnc  et  al.  (1984)  investigated  the  short-term
effects of  bis -tributyltin   oxide in the rat.  In  exper-
iments  lasting  4 weeks,  Wistar rats  were fed technical
TBTO  at levels of  0, 5, 20,  80, or 320 mg/kg  diet. All
animals  survived the dosing period.   Various symptoms of
poisoning were seen within 1 week of dosing in  the  group
fed  320 mg/kg diet, i.e. weakness,  emaciation, roughened
fur,  and blood-tinged discharge around the eyes and nose.
Some of these signs were seen at the lower doses  after  4
weeks.  Body weight gain was not affected by doses  up  to
20 mg/kg  diet  in males  and  80 mg/kg in  females. Males
showed  reduced  weight  gain (96 g  compared  to  control
weight gain of 117 g) over 4 weeks at 80 mg  TBTO/kg.   At
320 mg/kg  diet, TBTO caused a reduction in body weight of

13 g  in males and  21 g in females.   Almost all of  this
weight  loss occurred in  the first week  of dosing,  body
weight  remaining constant after that.  A 50% reduction in
food consumption was seen at this dose rate,  compared  to
control  animals.  Urine analysis, during  the fourth week
of dosing, revealed no differences in urine  volume,  pro-
tein  concentration,  or  creatinine clearance  related to

    Funahashi  et  al. (1980)  conducted histopathological
and  biochemical studies on Sprague-Dawley rats given bis-
tributyltin  oxide (TBTO) dissolved in olive oil. The rats
were  dosed by  intubation 5 times  weekly for  13  or  26
weeks.  Body weight was reduced after 13 weeks  dosing  at
6 mg/kg  body weight, but  not significantly so.  After 26
weeks,  body weight was significantly  reduced relative to
controls (382 g compared to the control value  of  439 g).
Dosing  at 12 mg/kg reduced  body weight significantly  at
both  13 and 26 weeks  (to 311 and  356 g,  respectively).
There  was a slight decrease, but unrelated to either dose
or time, in spleen weight. This decrease was just signifi-
cant  at 6 mg/kg over 13 weeks and 12 mg/kg over 26 weeks,
but at no other time or dose.

    When Mushak et al. (1982) dosed neonatal rats with TBT
acetate  at levels of 1, 3, or 10 mg/kg per day from day 2
to  day 29 of age, all rats given 1 mg/kg per day survived
with no apparent gross or histopathological effect. Totals
of  9 out of  24 and 17  out of 24  rats survived at  dose
levels  of 10 and 3 mg/kg  per day, respectively, but  all
showed reduced body weight.  There were liver  effects  in

11.2.2.  Inhalation studies

    A  "nose only" inhalation study  (lasting 4-5 weeks)
by Schweinfurth (1985) with rats exposed for  4 h/day  (on
weekdays only; 21 to 24 exposure periods) to an aerosol of
TBTO  (2.8 mg/m3)   produced mortality  (50% of males  and
60%  of females), apathy,  and respiratory distress.  Food
consumption  and body weight gain were reduced. There were
inflammatory  reactions  within the  respiratory tract and
lymphotoxic  effects  (depletion  of  lymphocytes  in  the
thymic  cortex, atrophy of  the thymus, and  lymph nodes).
Inhalation of TBTO vapour/air mixtures produced no observ-
able  effect. A concentration of 0.16 mg/m3   in the inha-
lation  chamber,  which  corresponds  to  the  equilibrium
vapour  pressure  of TBTO  at  room temperature,  was con-
sidered to be the NOEL for rats.

    When Gohlke et al. (1969) exposed rats to TBT chloride
in a 4-month inhalation study at nominal concentrations of
4 to 6 mg/m3,   all animals survived the exposure. Towards
the  end  of the  exposure, in the  final month, the  rats
showed  minor irritation of the eye and nose. There was an

initial increase in relative liver weight but  a  signifi-
cant  reduction over the  whole experimental period.   Fat
droplets were seen in the livers at autopsy, together with
a  diffuse  oedema of  the  brain. However,  controls also
showed  oedema. The oedema disappeared  slowly as recovery
time  after exposure increased. There  were also inflamma-
tory changes in the respiratory tract of exposed animals.

    Crofton  et al. (1989) exposed pregnant female rats to
TBTO by intubation at 0 to 10 mg/kg per day on  days 6  to
20 of gestation. There was no effect on the age  at  which
the  testes of male offspring  descended. However, females
showed a delay of approximately 2 days in vaginal opening,
compared  to controls, after their mothers were exposed to
10 mg/kg per day.

11.2.3.  Histopathological effects

    Snoeij et al. (1985) reported that weanling  rats  fed
tributyltin chloride at a level of 150 mg/kg  diet  showed
marked reductions in body weight (treated rats, 87 g; con-
trol  rats, 119 g) and brain weight (treated rats, 1.54 g;
control  rats,  1.67 g)  associated with  a  reduced  food
intake  of 25%. Thymus  weight was reduced  to 39% of  the
control value over the 2-week feeding period.

    Krajnc  et al. (1984) found  histopathological changes
in  male and female Wistar  rats exposed to 5,  20, 80, or
320 mg  TBTO/kg  diet  for 4 weeks.  No  treatment-related
changes  were found in the brain, heart, kidney, pancreas,
adrenals, popliteal lymph nodes, intestinal tract, or bone
marrow.  Slight atrophy of hepatocytes (reduction in size)
in  the centrilobular region was  noted in some livers  at
80 mg/kg  and marked atrophy in 16 out of 20 livers at 320
mg/kg.  The  authors reported  dystrophic calcification in
the liver at the highest dose level.  Three animals showed
multiple-focus  inflammation  with  necrosis  of   hepatic
parenchyma,  which  was  associated with  mononuclear  and
polymorphonuclear  infiltration, fibrosis, and hyperplasia
of  the intrahepatic bile  duct. In one  animal there  was
inflammation  of  the common  bile  duct. No  bacteria  or
viruses  were  found  in  the  lesions.  Similar  necrotic
lesions were observed, in two animals, in a  repeat  study
at 320 mg TBTO/kg diet.

    Mori et al. (1984) painted the shaved dorsal  skin  of
guinea-pigs  daily for 50 days with  an ethanolic solution
of TBTO at 10 or 40 mg/kg body weight. Monitoring of urine
and  blood  electrolyte levels  during  the course  of the
treatment showed increased loss of sodium, chloride, phos-
phate,  glucose, and amino acids in the urine. There was a
concomitant  loss of electrolytes  from serum. The  effect
was  most marked between 40 and 50 days of treatment. His-
tological lesions of the kidney tubules were observed when
the animals were killed after the 50th day.

11.2.4.  Haematological and biochemical effects

    Measurements  of blood biochemical parameters  of rats
fed  TBTO for 4 weeks indicated few significant effects at
dietary  dose levels below  320 mg/kg diet. Alanine  amino
transferase  (ALAT) activity was  significantly increased,
in  a dose-related manner,  at 20 mg/kg or  more, in  both
males  and  females.  The highest  dose  significantly de-
creased  blood glucose in males, aspartate amino transfer-
ase (ASAT) in both males and females, and  liver  glycogen
in  both sexes. Serum triglycerides, alkaline phosphatase,
and  creatinine kinase activities  were unaffected at  any
dose  level, as  were blood  lactate and  pyruvate. At  80
mg/kg  diet, TBTO significantly reduced  blood haemoglobin
in  both sexes and  haematocrit in females.  Mean erythro-
cytes volume was reduced in both sexes, as was  mean  cor-
puscular  haemoglobin content (mass), but mean corpuscular
haemoglobin concentration and erythrocyte numbers were not
affected  (Krajnc et al., 1984). A 6-week study at dietary
doses of 5, 20, and 80 mg/kg showed significant reductions
in haematocrit at 20 and 80 mg/kg and in blood haemoglobin
levels  at 80 mg/kg.  Iron concentration was reduced at 80
mg/kg, as was the isocitrate dehydrogenase (ICDH) activity
of  erythrocytes.  Numbers of  erythrocytes, thrombocytes,
and  reticulocytes were not significantly  affected at any
dose  level of  TBTO, though  there was  a  trend  towards
increased  numbers of reticulocytes. The authors suggested
an  effect on haemoglobin synthesis  and other indicators,
implying  either  reduced  iron uptake  or  increased iron
loss.  The enhanced ICDH activity and increasing reticulo-
cyte  numbers indicated the presence  of immature erythro-
cytes, and the authors could not exclude  the  possibility
of  an  in vivo haemolytic action of TBTO comparable to the
reported  in vitro haemolysis (Krajnc et al., 1984).

    Rosenberg  et al. (1984) measured the activity of haem
oxygenase  in mucosal cell fractions from control mice and
mice  dosed  by  intubation  with  TBTO  at  60 mg/kg body
weight.  The  activity of  the  enzyme, monitored  by  the
bilirubin absorbance spectrum, was substantially elevated,
compared  to  controls,  16 h after  administration of the
TBTO.  The activity of the same enzyme in liver and kidney
microsomes  was not affected by this dose of TBTO given by
gavage, but was elevated when TBTO was given parenterally.
The  activities  of  cytochrome P-450  and  benzo(a)pyrene
hydroxylase were substantially reduced in the intestine of
TBTO-treated  mice, this reduced activity  being statisti-
cally significant for the latter but not for  the  former.
Similar results were obtained in liver fractions when TBTO
was applied parenterally.

11.2.5.  Effects on lymphoid organs and immune function

    Funahashi  et al. (1980) reported the effects of feed-
ing  TBTO in olive oil  (0, 3, 6, or  12 mg/kg body weight
per  day) for 13  or 26 weeks by  gavage to groups  of ten
male Sprague-Dawley rats, aged 5 weeks at the beginning of

the  study.   No  analyses of  haematology, immunoglobulin
levels, or specific aspects of immune function  were  per-
formed.  A marked dose-related  reduction in absolute  and
relative  thymus weight was seen following dosing for both
periods.   Relative thymus weights were 682, 629, 340, and
313 mg/kg  body weight after  13 weeks of dosing  and 449,
313,  278,  and 248 mg/kg  body  weight after  26 weeks of
dosing  in the rats given  0, 3, 6, and  12 mg/kg per day,
respectively. All results, with the exception of those for
the  group given 3 mg/kg body weight per day for 13 weeks,
were  statistically significant (p < 0.003).   Despite the
considerable  reductions in thymus weight, the only histo-
logical observation was a slight reduction in the width of
the  thymic cortex. A  dose-related reduction in  relative
spleen weight was seen after 26 weeks, which was statisti-
cally significant (p < 0.05) at the dose level of 12 mg/kg
body weight per day.  Reduced body weight and increases in
relative pituitary and relative adrenal weights were stat-
istically  significant (p < 0.01) at 12 mg/kg  body weight
per day for 13 and 26 weeks and at 6 mg/kg body weight per
day for 26 weeks, and there was a  significant  (p < 0.05)
increase in relative adrenal weight at 6 mg/kg body weight
per  day for 13 weeks.  The relative pituitary  weight was
also  increased  following  26 weeks of  dosing at 3 mg/kg
body  weight per day  (p < 0.01). Based on  thymus weight,
the NOEL was the most sensitive end-point in  this  study,
3 mg/kg  body weight per  day for 13 weeks.   However,  an
effect  was  seen at  this dose when  it was given  for 26

    Funahashi  et  al.  (1980)  also  reported  that   the
reduction  in relative thymus weight, present 3 days after
a  single dose  of TBTO  in olive  oil of  100 mg/kg  body
weight  to  5-week-old  male Sprague-Dawley  rats,  showed
signs  of reversal at  day 8.  The reversibility  of  TBT-
induced  thymus atrophy was also demonstrated by Snoeij et
al. (1988b). Groups of 4 or 5 young (4-5 weeks  old)  male
Wistar  rats received a single gavage dose of TBT chloride
(0 or 16 mg/kg body weight) in corn oil. A  30%  reduction
in relative thymus weight was evident in animals killed on
days 2, 3, or 4, but there was complete recovery by day 7.
A  similar effect and recovery was seen in the thymus cell
counts.  An equimolar dose of dibutyltin chloride produced
more  pronounced effects, the  recovery period being  pro-
longed until day 9.

    Krajnc et al. (1984) reported the effects  of  feeding
TBTO-containing diets (0, 5, 20, 80, or 320 mg/kg diet) to
groups  (10 males and 10 females  per group) of  young SPF
Wistar  rats for 4 weeks. These dose levels are equivalent
to approximately 0, 0.5, 2, 8, or 32 mg/kg body weight per
day,   respectively,  based  on  actual  food  consumption
measurements.  Total leucocyte and  circulating lymphocyte
counts  were significantly reduced  in males receiving  80
mg/kg  diet  (p < 0.01) and  in  both sexes  receiving 320

mg/kg  diet  (p < 0.001). Eosinophil  counts were signifi-
cantly reduced (p < 0.05) in males receiving 5, 80, or 320
mg/kg diet and in females given the highest dose. Monocyte
counts  were  increased significantly  (p < 0.05) in males
receiving  20, 80, or 320 mg/kg  diet but not in  females.
Total immunoglobulin levels were significantly affected at
80  and 320 mg/kg diet.  At 80 mg/kg diet,  TBTO  signifi-
cantly  increased serum IgM  to 132% (p < 0.01)  and  145%
(p < 0.001)  of  control  levels for  males  and  females,
respectively. At the same dose, IgG in males  was  reduced
significantly  (p < 0.05), but in females IgG was unaffec-
ted.  A significant reduction (p < 0.001) in serum IgG was
found in both sexes fed TBTO at 320 mg/kg diet. Thymus and
relative thymus weights were significantly reduced in both
sexes  (p < 0.001) at 80 and 320 mg/kg diet and in females
given  20 mg/kg  diet  (p < 0.05). Relative  spleen weight
was  significantly increased in males  receiving 320 mg/kg
diet but this was probably related to the  decreased  body
weight  of these animals. Histological  examination showed
that  all  the animals  given  the highest  dose exhibited
lymphocyte  depletion from the thymic cortex, resulting in
an  indistinct cortico-medullary junction and  an increase
in  ceroid/lipofuschin-loaded macrophages.  Slight atrophy
of  the thymic cortex was  seen in two males  fed 80 mg/kg
diet.  Diffuse atrophy of the white pulp of the spleen was
seen  in all animals at 320 mg/kg, the periarteriolar lym-
phocyte  sheaths (PALS) being particularly affected. At 80
mg/kg diet, one male and two females showed slight splenic
atrophy.   Depletion of T lymphocytes, determined by pan-T
immuno-staining, was seen in the PALS at  320 mg/kg  diet.
There was an increase in the incidence of mesenteric lymph
nodes atrophy, observable in some animals fed 20 mg/kg and
increasing with dose to affect all animals given 320 mg/kg
diet.  Both the paracortex and medulla of lymph nodes were
reduced  in size and cellularity, and the numbers and size
of  follicles were reduced by the highest TBTO dose level.
Again,  the total number  of T lymphocytes  was strikingly
reduced  in the paracortex by  TBTO at 320 mg/kg diet,  as
determined by immuno-staining. Although thymic involvement
was  marked, it was  not only thymus-dependent  areas that
were affected.  In rats exposed to the highest dose level,
which  was overtly toxic,  B lymphocyte areas also  showed
low  level activity, as  indicated by fewer  follicles and
inconspicuous  germinal centres in lymph nodes and spleen.
An  increase in  the number  of animals  with rosettes  in
sinuses  in the mesenteric lymph nodes, composed of eryth-
rocytes  surrounding mononuclear cells, was seen. This was
dose related: half of the animals dosed at 5 mg/kg and all
the animals treated with 80 or 320 mg/kg  showed  erythro-
cyte   rosettes.  The  biological  significance  of  these
rosettes  is unclear.  Specific aspects of immune function
were  not  tested.  In further  studies (personal communi-
cation by E.I. Krajnc to IPCS, 1989; Wester, in press), no
increase in rosette formation was observed at 5 mg/kg diet
or  50 mg/kg diet. Signs of general toxicity were observed
in  animals treated with 320 mg/kg diet, i.e. reduced body

weight  (p < 0.001) and increased serum  ALAT activity. At
80 mg/kg,  there  was reduced  body  weight gain  in males
(p < 0.05)  and increased ALAT, and  at 20 mg/kg increased
ALAT  in males only. The  liver was the only  non-lymphoid
organ displaying histological changes: centrilobular hepa-
tocyte  atrophy,  reduced glycogen  retention, parenchymal
necrosis,  and hyperplasia of  the intrahepatic bile  duct
were seen in some animals treated at 320 mg/kg diet. Three
animals  from the group given 80 mg/kg showed slight hepa-
tocyte  atrophy. This study  indicated a NOEL  of  5 mg/kg
diet (approximately 0.5 mg/kg body weight per day).

    Vos  et al. (1984) investigated  functional aspects of
the immunological effects reported by Krajnc et al. (1984)
in  in vivo  and  ex vivo experiments using weanling (3 to 4
weeks old) male SPF Wistar rats fed diets containing TBTO.
Haematological  parameters  were  not studied.  Levels  of
total  circulating  immunoglobulins  (IgG  and  IgM)  were
determined  in animals fed  diets containing 80  or 320 mg
TBTO/kg.   At   320 mg/kg  diet,   IgM  was  significantly
increased  (p < 0.05) and IgG was  significantly decreased
(p < 0.001)  on day 28 (no assays were performed on day 42
for this group). At 80 mg/kg diet, IgM was elevated by 30%
on  day 42 (p < 0.01), while IgG  was decreased by 30%  on
days 28 and 42.

    The  effects  of TBTO  on  lymphocyte counts  and cell
viability in lymphoid organs was investigated, and several
specific  tests of immune  system function were  also per-
formed.  Numbers of B and T lymphocytes were significantly
reduced in the spleens (by 15% and 48%,  respectively)  of
animals fed diets containing 80 mg/kg diet. The  ratio  of
total  T:B  lymphocytes  was decreased  in  a dose-related
manner following 9 weeks of exposure.

    A  significant  reduction  occurred in  the numbers of
viable  cells  (trypan  blue exclusion)  obtained from the
thymus,  spleen, and bone marrow of rats treated with TBTO
at 80 or 320 mg/kg diet.  At 320 mg/kg diet, viability and
cell  numbers  were  significantly (p < 0.05)  reduced for
both  thymus and spleen following 3, 8, and 20 days of ex-
posure,  and bone marrow counts were significantly reduced
on days 8 and 20. Doses of 80 mg/kg diet produced signifi-
cant  reductions in thymus cell count and viability on day
8 and thymus, spleen, and bone marrow counts on day 20.

    The  antibody  response  to sheep  erythrocytes  (ip),
tetanus  toxoid (iv), ovalbumin (sc into the foot pad, in-
cluded  H37Ra adjuvant), and the worm  Trichinella spiralis
(oral)  was assessed in animals receiving 20 or 80 mg TBTO
per  kg diet for  6 weeks. The primary  response to  sheep
erythrocytes (0.5 ml of a 20% suspension, ip)  was  deter-
mined  10 days  post  inoculation (pi)  and  the secondary
response  was  determined on  day  20 pi following  an  iv
booster inoculation on day 15 pi. The primary response was
unaffected, but there was a dose-related decrease  in  the

secondary  response seen both in untreated and 2-mercapto-
ethanol-treated (IgM-inactivated) sera, reaching statisti-
cal  significance  (p < 0.05)  in treated  sera  from  the
highest-dose  group.  The response (IgG and IgM titres) to
tetanus  toxoid inoculation measured only on day 21 pi was
equivocal,  and that to ovalbumin measured on days 15, 21,
and  28 pi was unaffected by TBTO. IgG response to oral  T.
 spiralis infection  was significantly (p < 0.05) increased
on  day 21 pi at 20 mg/kg diet, but not on day 42 or at 80
mg/kg  diet at either  time. IgM response  was unaffected.
IgE  response, possibly the most relevant to resistance to
parasitic  infection, was decreased in a dose-related man-
ner on days 21 and 42 following  T. spiralis infection; log
titres  were 3.8, 2.5  (p < 0.05), and 1.8  (p < 0.001) on
day 21 and 4.4, 3.6, and 3.4 (p < 0.05) on day 42  in  the
control, 20-, and 80-mg/kg groups, respectively.  Delayed-
type  hypersensitivity  was  determined as  change in skin
thickness following a challenge of ovalbumin to  the  skin
of  the  ear or  tuberculin challenge to  the skin of  the
flank, made 3 and 4 weeks, respectively, following initial
immunization  (sc into a  footpad) 6 weeks after  starting
dietary  dosing of  TBTO at  20 or  80 mg/kg. Compared  to
controls  similarly  injected  intradermally with  medium,
significant  reductions in response to ovalbumin challenge
were  found 24, 48, and 72 h after dosing at 20 mg/kg (all
p < 0.05),  and 24 h (p < 0.01) and  48 h (p < 0.05) after
dosing at 80 mg/kg. With tuberculin challenge, significant
effects  were found in the group given 80 mg/kg diet after
24 h  (p < 0.01), 48 h (p < 0.001), and  72 h (p < 0.001),
but  only after 72 h (p < 0.05)  at the lower dose  level.
Reduced  responses were seen at all three times after both

    Thymus or spleen cells, obtained from rats fed TBTO at
20  or  80 mg/kg  diet,  were  cultured  with  and without
mitogens  (PHA = phytohaemagglutinin; Con A = concanavalin
A;  PWM = pokeweed mitogen) for  24 h before the  addition
of  3H-thymidine   to monitor DNA synthesis. PHA and Con A
are both T cell mitogens and PWM is a mitogen for  both  T
and  B cells. Due to  a reduction in the  number of viable
cells in thymic cultures from the high-dose group, respon-
siveness  to mitogens was  expressed per culture  and  per
thymus.  Significant  reductions in  3H-thymidine   uptake
(expressed  per  thymus)  were found  in unstimulated (48%
reduction),  PHA-treated  (64%), Con A-treated  (62%), and
PWM-treated  (50%)   cultures  derived from  the high-dose
group;  PHA (48%) and  PWM (28%) were  the only  responses
reduced  at 20 mg/kg diet.   Similar effects were  seen on
splenic  cultures, with a  reduced number of  viable cells
per   spleen   and  significant   reductions  in  response
(expressed  per spleen) to  PHA (50% reduction)  and Con A
(45%) in cultures from the high-dose group;  responses  to
PWM  and  E.  coli lipopolysaccharide  (a B  cell  mitogen)
were  significantly increased on  a per culture  basis but

not  a per spleen basis.  Cultures from the low-dose group
showed reduced response to PHA (25% reduction)  and  Con A
(20%), but these were not statistically significant.

    TBTO,  at  both  20  and  80 mg/kg  diet,  reduced the
resistance of rats to infection by  T. spiralis. The number
of  worm  larvae in  muscle  significantly increased  in a
dose-dependent manner (74 000 in controls; 106 000 in rats
fed  20 mg TBTO/kg, p < 0.01; 198 000 in rats fed 80 mg/kg
p < 0.001)  after  standard infection  with 1000 larvae by
mouth.  The number of adult  worms in the small  intestine
also  increased  significantly  after  10  (p < 0.05),  12
(p < 0.001),  and  14  (p < 0.01) days  in  the  high-dose
group,  and after 12 (p < 0.05) and 14 (p < 0.001) days in
the  low-dose groups. In a study of the inflammatory reac-
tion  around larva-containing muscle cells  in the tongues
of  animals killed 14 days  after infection, the  response
(mononuclear  cells  and eosinophilic  granulocytes) after
treatment  with  TBTO at  20 mg/kg  diet was  described as
"slightly   reduced".   However,  there   was  a  marked
reduction after treatment at 80 mg/kg diet.

    The   clearance  of  Listeria  monocytogenes from   the
spleen  (a measure of  host resistance) was  monitored  in
rats  fed 20, 80, or 320 mg TBTO/kg diet for 6 or 7 weeks.
Statistically  significant  increases  in  the  number  of
viable  bacteria were seen  2 days after an  iv  injection
into  animals treated with 320 mg/kg diet for 6 (p < 0.01)
or  7 (p < 0.001) weeks and with 80 mg/kg diet for 6 weeks
(p < 0.001).  A dose-related reduction, statistically sig-
nificant  (p < 0.05) at 80 mg/kg diet,  in viable bacteria
was  seen 1 day post injection in the group receiving TBTO
for  7 weeks, but this was  not seen in the  6-week study.
The  ex  vivo phagocytosis  of  L. monocytogenes by  spleen-
and  peritoneal-derived macrophages from TBTO  treated (20
or  80 mg/kg diet) animals was not significantly affected,
although   a  dose-related  reduction  in  the  phagocytic
activity of splenic macrophages was seen.

    Cells  derived from the  spleen and peritoneal  cavity
were  tested  for  spontaneous cell-mediated  cytotoxicity
against   murine   YAC  lymphoma   target  cells  labelled
with  51Cr.   "Specific release"  of the chromium  label
(release  in experimental minus spontaneous release in the
controls)  was used as the  end-point of the assay.  Stat-
istically  significant reductions in specific release were
found  with spleen cells  from rats fed  TBTO at  80 mg/kg
diet (p < 0.05), but not in spleen cells from those fed 20
mg/kg  diet.  Statistically significant effects (p < 0.05)
were  found  in  peritoneal macrophages  derived from rats
receiving 20 and 80 mg/kg diet, but there were no signifi-
cant  effects  on  non-adherent  cells  ("natural  killer

    These  studies  were  designed to  investigate  immune
function  rather  than to  determine a no-observed-effect-
level, and the lowest dose used (20 mg/kg  diet)  produced
statistically  significant  effects  (in  particular,  de-
pression of host resistance to  T.  spiralis  and   L. mono-
 cytogenes ). No evidence of general  toxicity was seen  at
this  dose  level.   The  only  sign  of  general toxicity
recorded was a significant reduction in body  weight  gain
seen  after exposure to  320 mg/kg diet for  3, 8, and  20
days and to 80 mg/kg diet for 20 days.

    A study of TBTO (0, 0.5, and 50 mg/kg  diet)  adminis-
tered to groups of weanling Wistar rats (five  animals  of
each  sex per group) for  2 years was briefly reported  by
Vos  et  al. (1985)  and Wester (in  press). It should  be
noted that the intakes in this study were lower, on a body
weight  basis,  than  in the  shorter-term  studies, being
equivalent  to 0, 0.025,  0.25, and 2.5 mg/kg  body weight
per day for the controls, and low, medium, and high doses,
respectively  (based  on  actual intake  measurements).  A
significant  decrease in peripheral lymphocyte count and a
statistically  significant increase in platelets were seen
in the females fed 50 mg/kg diet for  1 year.  Circulating
levels of total IgM and IgA were  significantly  increased
at  4-6  and 16-18 months,  in  rats given  50 mg/kg diet,
while  significant reductions in IgG levels were recorded,
particularly in females.

    General  effects  on  the  lymphoid  organs  were  not
recorded, though several specific tests of immune function
were performed using the methods of Vos et al.  (1984).  A
dose-related  decrease in resistance to  T. spiralis infec-
tion  was seen at  5 and 16 months,  achieving statistical
significance (p < 0.05) at 5 and 50 mg/kg diet. IgE titres
were  reduced, but IgA levels  increased approximately 50-
fold at the highest dose level. No effects on delayed-type
hypersensitivity  were seen, in contrast to results in the
short-term  study by Vos et al. (1984). Host resistance to
 L.   monocytogenes (as  measured  by the  number of viable
organisms  in the spleen)  was significantly reduced  at 5
and  17 months in the  50-mg/kg group, but  a  significant
increase  was  seen at  17 months  in the  5-mg/kg  group.
Natural  killer cell activity  against YAC lymphoma  cells
was  reduced significantly at the highest dose level after
15-17 months. It will not be possible to fully assess this
study  until the final  report is published,  although  it
would  appear that 0.5 mg/kg diet (0.025 mg/kg body weight
per day) was the NOEL.

    A preliminary report on a study where diets containing
TBTO (0, 0.5, 5, or 50 mg/kg diet) were fed to Wistar rats
(12 months  old) for 6 months,  has been issued  (personal
communication  by  E.I.  Krajnc to  IPCS, 1989).  Specific
tests of immune function were measured after  5 months  of
dosing.  Significant (p < 0.05) reductions in host resist-
ance to  T.  spiralis  and   L. monocytogenes were  seen  at

50 mg/kg  only. These results  suggest that aged  rats are
less sensitive to TBTO in the diet than weanlings animals,
although  this may be due  to a lower intake  on a per  kg
body weight basis. Full assessment of this study will need
to  await the completion  of the statistical  analyses and
final report.

    A series of studies have been performed to investigate
certain  aspects  of  immune function (Schering, 1989a,b,c,d).
When  groups  (10 animals of  each sex) of  young (4 to  5
weeks  old) Sprague-Dawley rats were  fed diets containing
TBTO (0, 0.5, 2, 5, or 50 mg/kg diet) for 4 weeks, no sig-
nificant  effects were seen on total or differential white
blood  cell counts. Serum  immunoglobulin levels were  not
measured. A statistically significant decrease in absolute
and  relative thymus weight  (p < 0.01) was seen  in males
fed 50 mg/kg. A decrease in absolute and  relative  spleen
weight was also seen in this group, though it was not sig-
nificant  statistically. The viability of  cultured spleen
cells (monitored with trypan blue exclusion) obtained from
females  fed  5 mg/kg  was reduced  (p < 0.05),  but other
groups  were unaffected.  A significant  decrease in total
thymus  cell count was seen in preparations from males fed
50 mg/kg   only  (p < 0.05).  A  slight   but  significant
(p < 0.05) reduction in the thickness of the thymic cortex
was also seen in these males, this being the only signifi-
cant  histological  finding.  Specific  measures of immune
function  were not performed in this study but in parallel
studies.    The  NOEL  in  this  study  was  5 mg/kg  diet
(approximately  0.6 mg/kg  body  weight per  day, based on
measured intake) (Schering 1989a).

    An assay of plaque-forming cells, a measure of humoral
immunity,  revealed no effects due  to TBTO (at levels  of
0.5,  2, 5, or 50 mg/kg diet) fed to groups (10 animals of
each  sex) of young (4 to 5 weeks old) Sprague-Dawley rats
for  5 weeks. The response was measured on day 36, follow-
ing an iv inoculation of sheep erythrocytes. The only sign
of  toxicity was a reduction  in body weight in  males fed
50 mg/kg,  but this was not  statistically significant. No
investigation  of general toxicity was performed. The NOEL
was 50 mg/kg diet (approximately 5.6 mg/kg body weight per
day) in this study (Schering, 1989b).

    To  assess the effects of  TBTO on host resistance  to
infection,  the  number  of viable  Listeria  monocytogenes
cells   in  the  spleen,  4 days  after  inoculation  with
1.4 x 106 cells,  was counted. No effects were produced in
groups  (10 animals of each  sex) of young  Sprague-Dawley
rats  fed diets containing TBTO at 0.5, 2, or 5 mg/kg diet
for  34 days. However, statistically significant increases
in numbers of viable bacteria were seen in the  spleen  of
males  (p = 0.055)  and  females (p < 0.01)  fed  50 mg/kg
diet.  No  investigations  of general  toxicity  were per-
formed. The NOEL was 5 mg/kg diet (approximately 0.6 mg/kg
body weight per day) (Schering, 1989c).

    The  effects of TBTO on  delayed-type hypersensitivity
reactions,  a  measure  of  cell-mediated  immunity,  were
assessed  in groups (10 animals of  each sex) of 4- to  5-
week-old  Sprague-Dawley rats fed diets containing 0, 0.5,
2,  5, or 50 mg/kg diet for 37 days. A sensitizing dose of
100 µg    bovine serum ablumin  (BSA) mixed with  Freund's
complete adjuvant was given on day 29, followed by a chal-
lenge dose of heat-inactivated BSA injected into  a  hind-
foot pad on day 37. No difference in  response  (increased
footpad  thickness)  was  seen between  test  and  control
groups.  Body  weight gain  was  unaffected, but  no other
investigations  of  general  toxicity were  performed. The
NOEL  in this study  was 50 mg/kg diet  (approximately 5.8
mg/kg body weight per day) (Schering 1989d).

    The effects of inhaled TBTO were studied in groups (10
animals of each sex) of young (initial  weights  86-131 g)
SPF  Wistar rats exposed  to 0, 0.03,  0.16, or    2.8 mg/m3
for 4 h/day, 5 days per week, for 21 to 24 exposures.  The
two  lower doses were provided by filtered vapour, and the
highest dose was provided by an aerosol with over  90%  of
the particles < 5 µm  in diameter (Schering 1983). Lympho-
cyte and total leucocyte counts, measured at 2 or 4 weeks,
were  unaffected  by  TBTO. Some  inconsistent  changes in
reticulocyte counts were found, but these appeared  to  be
primarily  related to use  of the respiration  chamber and
not  to TBTO exposure. Immunoglobulins  were not measured.
Thymolysis  and lymphocyte depletion of  the thymus-depen-
dent  areas of the spleen and lymph nodes were reported in
the 11 animals (five males, six females) from the highest-
dose  group that died  during the study.  No such  lesions
were  detected in the  survivors, although in  three  sur-
vivors  from  the highest-dose  group  an increase  in the
number  of macrophages containing nuclear  debris was seen
in  the thymic cortex.  No  significant (p > 0.05) changes
were seen in absolute or relative weights of  the  thymus,
spleen,  or  iliac lymph  node  in animals  surviving  the
study, but no organ weights were recorded for  those  ani-
mals  dying or sacrificed  during the study.  No  specific
aspects  of  immune  function were  studied.  Histological
signs of general toxicity were limited to those consistent
with  inflammatory reactions within the  respiratory tract
and,  with one exception, were confined to animals exposed
to  the aerosol.  Food intake  and body  weight gain  were
reduced  in both sexes, the  reduction being statistically
significant  (p < 0.01) in male rats exposed to   2.8 mg/m3.
The cause of death in the 11 animals from the highest-dose
group was not ascertained (Schering, 1983).

    When TBTO (0, 4, 20, 80, or 200 mg/kg diet) was fed to
groups  (10 animals of each sex) of CD-1 mice, aged 6 to 7
weeks at commencement of the study, for  3 months,  leuco-
cyte  counts were increased  at 80 and  200 mg/kg in  both
sexes,  although this increase  did not reach  statistical
significance    (p > 0.05).   Immunoglobulins   were   not
measured.   Thymus weight was reduced in both sexes at 200

mg/kg,   but   this  was   not  statistically  significant
(p > 0.05).   Spleen weights were increased  in both sexes
at  80  and  200 mg/kg (reaching  statistical significance
(p < 0.05) at 200 mg/kg), possibly secondary to effects on
erythrocytes.  No specific immune function tests were per-
formed.  Histological changes were  seen in the  livers of
both  sexes at 80  and 200 mg/kg. Dose-related,  statisti-
cally  significant (p < 0.01) increases in  absolute liver
weights were seen at 80 and 200 mg/kg, and adrenal weights
were  significantly (p < 0.01) increased in  the male rats
fed  200 mg/kg. The NOEL in  this study was 20 mg/kg  diet
(approximately  4 mg/kg body weight per day) (Biodynamics,

    In studies by Schering (1989e), groups (two animals of
each sex) of beagle dogs received variable doses  of  TBTO
in arachis oil by oral gavage:

    Group 1: controls;

    Group 2:  0.1 mg/kg body weight  per day for  5 weeks,
    0.2 mg/kg body weight per day for 4 to  5 weeks,  then
    10 mg/kg body weight per day for 8 to 9 weeks;

    Group 3:  0.5 mg/kg  body  weight per  day for 5 weeks
    then 1 mg/kg body weight per day for 13 to 14 weeks;

    Group 4:  2.5 mg/kg  body  weight per  day for 5 weeks
    then 5 mg/kg body weight per day for 13 to 14 weeks.

All males in groups 2 and 4 died as a result of mis-dosing
to  the lungs.  Increased leucocyte  and neutrophil counts
were  seen in group 4  at weeks 9 and  18 (p < 0.05),  and
there  was a statistically significant  increase in leuco-
cyte  count  recorded  at 13 weeks  in group 2 (p < 0.05).
Non-statistically  significant  reductions  in  leucocyte,
neutrophil, and lymphocyte counts were found in group 2 at
18 weeks.   Specific  immunoglobulins  were not  measured.
Thymus  weights were reduced in the two survivors of group
2 (0.9 ± 0.1 g compared with 4.0 ± 0.6 g in controls), and
slight increases in thymus weights were seen  in  groups 3
and  4.  Iliac  and  mesenteric  lymph  node  weights were
reduced,  though spleen weights were increased, in the two
survivors  from group 2.  Histological changes in lymphoid
organs  were confined to group 2 where a reduction of lym-
phocyte numbers in the thymus, spleen (particularly PALS),
and lymph nodes was seen. A dose-related increase in rela-
tive  liver weight (33.1,  41.7, 49.9, and  57.6 g/kg body
weight  in  groups  1, 3,  4,  and  2,  respectively)  was
accompanied  by cytoplasmic vacuolation of  hepatocytes in
group 2 (Schering, 1989e).

11.2.6.  Mechanism of immunotoxicity

    The  precise mechanism of  the immunotoxic effects  of
TBTO  is not yet clear. However, a hypothesis has been put
forward, based on the work of Snoeij (1987) and Pieters et
al. (1989).

a)  Absorbed TBTO is present as the TBT+ cation  or a salt
    (chloride  or  carbonate). Studies  using TBT chloride
    are, therefore, relevant to TBTO immunotoxicity.

b)  As dibutyltin chloride (DBT chloride), a metabolite of
    TBT  chloride, is, mole for mole, more potent that TBT
    chloride at producing effects in the thymus and thymic
    cells, it is probable that DBT chloride or another DBT
    salt is the active species in TBTO toxicity.

c)  The primary action of DBT is to suspend the maturation
    of  immature  thymocytes  by inhibiting  their  inter-
    action/binding with thymic epithelial cells. The turn-
    over period of thymocytes is 3 to  4 days.  Therefore,
    as  cell proliferation/maturation is  inhibited, rapid
    depletion of thymocyte numbers without cytotoxicity is
    expected,  followed by a rapid proliferation (observed
    by  Snoeij et al. (1988b) and Funahashi et al. (1980))
    on removal of DBT.

Other evidence indicating that TBT compounds have  a  par-
ticular effect on the thymus is provided by  the  in vitro
studies  of Snoeij (1987), which show TBT chloride  to  be
cytotoxic to thymocytes.

    The demonstration of reduced thymus weights in certain
fish  species (e.g., the freshwater  guppy) indicates that
TBT  may  have  immunotoxic effects  on  a  wide range  of

11.2.7.  Effects on the endocrine system

    The weights of both the adrenal and  pituitary  glands
of  Sprague-Dawley rats were significantly increased after
exposure to 6 mg TBTO/kg body weight daily, by intubation,
for  26 weeks.  Pituitary  weight was  also  significantly
increased  by  3 mg TBTO/kg  body  weight given  daily for
26 weeks (Funahashi et al., 1980).

    After 4 weeks, the serum insulin concentration of rats
was not significantly affected by dosing with TBTO  at  up
to 80 mg/kg diet, but was undetectable (< 2 milliIU/litre)
in  rats fed 320 mg/kg (control levels of insulin in serum
were 111 and 74 milliIU/litre for males and  females,  re-
spectively) (Krajnc et al., 1984). In a further study, the
same  authors monitored endocrine  changes in male  Wistar
rats  exposed to TBTO in the diet (0, 20, or 80 mg/kg) for
6 weeks.   Serum  thyroxine,  thyroid stimulating  hormone
(TSH),  and insulin levels  were reduced significantly  at
the  higher  dose  level, but  serum  follicle stimulating

hormone  (FSH) and corticosterone levels  were unaffected.
Only  insulin was significantly affected at 20 mg TBTO/kg.
The  luteinizing hormone (LH)  concentration  in serum was
significantly  increased  by  TBTO at  80 mg/kg,  but  was
unaffected by 20 mg/kg. The authors further examined endo-
crine  function  by  monitoring hormone  release following
physiological  stimulus.  Insulin  release,  following  iv
administration  of  glucose,  was unaffected  by  a 6-week
exposure to TBTO at either 20 or 80 mg/kg diet. The effect
of  TBTO  on insulin  was attributed by  the authors to  a
decreased food intake. The release of TSH, after iv admin-
istration  of thyrotrophin releasing hormone (TRH), showed
a  tendency (p < 0.1) to  be inhibited in  rats fed at  80
mg/kg.  The  titre of  circulating  TSH, 20 min  after TRH
administration,  was  significantly reduced  compared with
controls.   Release of  both LH  and FSH,  in response  to
luteinizing  hormone  releasing  factor  stimulation,  was
enhanced  in rats fed TBTO  at both 20 and  80 mg/kg diet,
but  only significantly so at the higher dose rate. Histo-
logical  examination  of  endocrine organs  after a 6-week
exposure  to TBTO revealed  some changes.  No  differences
were  observed  in  either insulin- or  glucagon-producing
cells in the pancreas. Some flattening of  the  epithelial
lining  of  the  thyroid  follicles  was  observed   after
exposure  of rats to 80 mg/kg but not to 20 mg/kg. Immuno-
cytochemical  staining  of the  pituitary gland identified
each cell type producing the different pituitary hormones.
There was a dose-related decrease in both the intensity of
staining  of TSH cells  and the number  of cells  stained.
Conversely,  there  was  a dose-related  increase  in  the
staining  intensity of LH cells.  No effects were found on
FSH,  growth hormone, or adrenocorticotrophin cells in the
pituitary (Krajnc et al., 1984).

11.3.  Long-term toxicity

    Wester  (in  press)  carried out  a  106-week toxicity
and  carcinogenicity  study  with  groups  of  50 weanling
Wistar  rats of each sex.  An additional group of  10 rats
was  used for an interim sacrifice after 1 year.  TBTO was
fed  at 0,  0.5, 5,  or 50 mg/kg  diet (equivalent  to  0,
0.025,  0.25,  or  2.5 mg/kg body  weight). Increased food
consumption  occurred  in  all treated  males (not clearly
dose-related),  and there was increased  water consumption
in  males at 5 and  50 mg/kg. During the second  year, the
body  weight of the  highest-dose group was  significantly
lower  than that of controls.   Excess mortality, compared
with  that of controls, was confined to the 50 mg/kg group
towards  the  end  of the  experiment  (see section 11.6).
Haematological changes (anaemia, lymphopenia, and thrombo-
cytosis)  and increases in plasma enzyme activities (ALAT,
ASAT, and AP)  were noted mainly at the  high-dose  level.
Serum  IgM and IgA levels  increased, while the IgG  level
decreased (females). No effect was observed on circulating
concentrations  of T4,   free T4,   TSH, LH, FSH, or insu-
lin; only the free T4:T4     ratio was  decreased.   Organ

weight  changes  consisted  of  increased  liver,  kidney,
adrenal,  and  pituitary  weights  and  decreased  thyroid
weight.  Non-neoplastic histological alterations consisted
of  a decrease in cell height of the thyroid follicles (at
50 mg/kg  diet after 1  and 2 years), decrease  in splenic
iron content (at 5 and 50 mg/kg after 1 year only), slight
bile  duct proliferation (at 50 mg/kg  after 1 year only),
and  vacuolation of kidney proximal tubular epithelium and
nephrosis (at 50 mg/kg after 2 years only).

11.4.  Genotoxicity


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

    Davis et al. (1987) conducted a comprehensive study of
the  genetic effects of TBTO  using a wide range  of tech-
niques in order to assess possible hazards in the  use  of
the compound as a molluscicide for the control of schisto-
somiasis. TBTO did not produce gene (point)  mutations  in
Salmonella  typhimurium  strains  TA1530, TA1535,  TA1538,
TA97, TA98, or TA100, either in the presence or absence of
an  exogenous metabolic activation system  (rat liver S9).
The  compound did give some  evidence of gene mutation  in
Salmonella  typhimurium TA100 using the fluctuation method
in the presence of S9, but no  dose-response  relationship
was seen; negative results were seen in the absence of S9.
TBTO  did  not  induce  point  mutations  in   the   yeast
 Schizosaccharomyces  pombe. Negative results were obtained
when  TBTO was tested  in the sex-linked  recessive lethal
assay  using  Drosophila  melanogaster, the compound  being
given in food and by injection, indicating that  TBTO  did
not produce gene mutations in  Drosophila. Negative results
were also obtained when the ability of TBTO to produce DNA
damage  in  Bacillus subtilis (recombination assay)  or the
yeast  Saccharomyces  cerevisiae (mitotic  gene conversion)
was tested.

    The ability of TBTO to produce gene mutations in  Sal-
monella  has also been studied  by Reimann & Lang  (1987).
Negative    results   were   obtained   using   Salmonella
typhimurium  strains  TA1535,  TA1537, TA1538,  TA98,  and
TA100, both in the presence and absence of rat  S9.  Simi-
larly,  negative results were obtained with six TBT esters
(abietate,  borate, linoleate, naphthenate, phosphate, and
tallate).  Further negative results were obtained when the
ability  of  TBTO to  produce  mitotic gene  conversion in
 Saccharomyces cerevisiae was tested.

    The  ability  of TBTO  to  produce gene  mutations  in
mammalian  cells  in vitro was extensively  investigated by
Davis  et al. (1987),  and negative results  were consist-
ently  obtained. TBTO did not induce gene mutations in V79

Chinese  hamster cells (using resistance  to 8-azaguanine,
ovabain, or 6-thioguanine as markers) in the  presence  of
rat liver S9. Negative results were also obtained  in  the
V79  cell assays when epidermal  cells of mice and  humans
(primary  cultures) were used  as the source  of metabolic
activation in cell-mediated assays.

    The  in vitro clastogenic  potential of  TBTO has been
investigated  in mammalian cells  in two sets  of studies.
When  Davis et al. (1987) used Chinese hamster ovary (CHO)
cells, harvested at 8, 15, and 24 h, an increase in struc-
tural   aberrations  (mainly  deletions),   together  with
endoreduplication,  was  seen,  but only  at  the  highest
concentration  tested (5 µg/ml   in the presence of S9 and
1.5 µg/ml   in its absence).  The increase in the presence
of  S9 was seen  only after 15 h,  "toxicity" precluding
any analysis of results at this concentration  after  8 h.
The  increase in the absence  of S9 was seen  only at 8 h,
there being no increase at either 15 or 24 h.  The results
of  these studies are  difficult to interpret,  since  the
effect  was limited to concentrations associated with high
toxicity  and no data  on mitotic index  was reported.  No
increase  in  sister chromatid  exchange  was seen  at any

    Reimann & Lang (1987) used human lymphocytes  to  test
the  clastogenic potential of  TBTO and obtained  negative
results  both in the  presence and absence  of S9. In  the
latter  case, TBTO was added 22 h after stimulation of the
cultures  with  phytohaemagglutinin,  and cells  were har-
vested 31 h later. In the former case, TBTO was added with
S9 after 26 h; 3 h later the S9 was removed, and the cells
were  harvested 22 h later. Parallel studies on blood cul-
tures that were differentially stained with BUdR indicated
that  almost all cells analysed for chromosome aberrations
were  in the first mitotic stage. The highest TBTO concen-
trations  used  (0.1 µg/ml    in  the  absence  of  S9 and
1 µg/ml   in its presence) were associated with  a  marked
reduction in mitotic index (57% and 58%, respectively). No
increase in aberrations was seen at any dose  level.  This
study, therefore, failed to confirm the suggestion of some
clastogenic potential in the studies using CHO cells.

    The  ability of TBTO to  produce chromosomal damage  in
 vivo has  been investigated in two  separate studies using
the  micronucleus test. In one  study, four doses of  TBTO
(31.25, 62.5, 125, or 250 mg/kg body weight) were given by
gavage,  in arachis  oil, to  NMRI mice,  and bone  marrow
cells  were  analysed  for  micronuclei  in  polychromatic
erythrocytes 24, 48, and 72 h after treatment  (Reimann  &
Lang,  1987). The highest  dose level resulted  in  marked
lethality  (16  out  of 36 animals  died),  precluding any
analysis of the results. The 126-mg/kg dose level was also
associated  with some deaths (4 mice died), but 5000 poly-
chromatic  erythrocytes were analysed  from five male  and
five  female mice at each harvest interval. Similar analy-
ses  were carried out at the two lower dose levels.  There

was no increase in micronuclei at any dose level  or  har-
vest  time.  This study  provided no evidence  to indicate
that  TBTO produces chromosomal  damage in bone  marrow  in
 vivo. A  second  micronucleus  study (Davis  et al., 1987)
used Balb/c mice.  Groups of 10 male and 10 female animals
were given 30 or 60 mg TBTO/kg as a solution in olive oil.
Bone  marrow cells were harvested from five males and five
females after 30 h and 48 h, and 1000 polychromatic eryth-
rocytes  were  analysed  from  each  for  micronuclei.  An
increase  in micronuclei was  seen only in  the male  mice
after 48 h and at the highest dose level.   These  results
conflict  with those of  Reinmann & Lang  (1987). Further-
more, there was an unusually high spontaneous incidence of
micronuclei  (up  to  5.8 per  1000 polychromatic erythro-
cytes) in the study of Davis et al. (1987).  These factors
prompted  a re-analysis of the  slides from this study  by
the  Institute  of Occupational  Health, Helsinki, Finland
(Schering  1986). This re-analysis, again using 1000 poly-
chromatic  erythrocytes per animal, failed  to confirm the
increase  in  micronuclei  seen  after  48 h  in  the male
animals  given 60 mg TBTO/kg. A  slight, but statistically
significant,  increase  in  micronuclei was  seen  in  the
female  mice on re-analysis,  but this was  thought to  be
biologically  non-significant due to the  high variability
in  the  control  data. The  re-analysis  highlighted  the
problem  of interpreting studies in  which only relatively
few  polychromatic  erythrocytes  are analysed  (1000) and
there is marked variability in the control data  (with  an
incidence outside the normal range at some  time  points).
Thus,  no conclusions can  be drawn from  the micronucleus
study of Davis et al. (1987).

    The  ability of TBTO to  inhibit metabolic cooperation
between  V79  Chinese hamster  6-thioguanine-resistant and
sensitive  cells  has been  investigated  by Davis  et al.
(1987).  This assay  has been  suggested as  a  model  for
tumour  promoter activity. Negative results were obtained.
The significance of the assay is, however, unclear.

    In  summary,  the genotoxicity  of  TBTO has  been the
subject  of extensive investigation. Negative results were
obtained  in the vast majority of studies, and there is no
convincing evidence that TBTO has any mutagenic potential.

11.5.  Reproductive toxicity


     The  potential embryotoxicity of TBTO has been evaluated 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
 regards  to  embryotoxicity  and  fetotoxicity  for  all  three
 species, was 1 mg/kg body weight.

11.5.1.  In vivo

    Reproductive  toxicity has been studied  in NMRI mice.
The highest dose used (35 mg/kg body weight) was chosen to
give  minimal maternal mortality  based on acute  toxicity
tests. An increase in cleft palate was seen in the fetuses
of  mice treated orally  with 11.7 mg TBTO/kg  body weight
(7%  cleft palate), 23.4 mg/kg (24%),  and 35 mg/kg (48%),
compared  to the incidence in controls (0.7%). However, 11
out of a total of 15 affected mice were clustered  in  one
of the 18 litters and 15 litters contained none.  The  two
highest  doses  of TBTO  also  increased the  frequency of
irregular  ossification centres of sternabrae and of minor
abnormalities, such as fusion of the bases of  os  occipi-
talis. The strain of mice used in the study has a tendency
to produce these particular abnormalities as a  result  of
non-specific  stress on the mother. The authors considered
that TBTO has a very low teratogenic  potential;  electron
microscopy  26 and 48 h after treatment showed no evidence
of  damage to the embryos  but considerable damage to  the
maternal  liver. At a level  of 6 mg/kg body weight,  TBTO
produced no increase in fetal abnormalities (Davis et al.,

    Nemec  (1987) investigated the  maternal, embryotoxic,
and  teratogenic  effects of  TBTO  in New  Zealand  white
rabbits.  The TBTO was administered in corn oil by gavage,
once a day from day 6 of gestation to day 18 inclusive, at
doses  of 0.2, 1, and  2.5 mg/kg per day body  weight in a
volume  of 0.5 ml. Twenty  female rabbits were  dosed with
corn oil as controls and 20 rabbits were used at each dose
level.   The  rabbits  were artificially  inseminated  and
injected iv with human chorionic gonadotrophin immediately
afterwards to ensure ovulation.  With the exception  of  a
single female dosed at 1 mg/kg, all treated  rabbits  sur-
vived  to  day 29 of  gestation  when the  experiment  was
terminated.  A  total  of 12 animals  aborted  during  the
dosing  period: 3, 1, 1,  and 7 from the control,  0.2, 1,
and 2.5 mg/kg groups, respectively.  Females aborting were
killed  the same day and immediate postmortem examinations
carried  out. The increased occurrence of abortions in the
highest-dose group was considered to be a secondary effect
of maternal toxicity. No clinical findings in  the  groups
given 0.2 or 1 mg/kg were considered to be the  result  of
TBTO treatment. There was a statistically significant mean
body  weight loss in  females dosed at  2.5 mg/kg per  day
compared  to controls between  days 6 and 18 of  gestation
(the  period of actual dosing). Doses of 0.2 and 1.0 mg/kg
per  day had no effect  on growth or survival  of fetuses.
There   was   a  slight   (statistically  non-significant)
decrease  in mean fetal weight in the group dosed with 2.5
mg/kg  per  day.  This may  represent  minor fetotoxicity.
There  were no differences  in the types  or frequency  of
fetal  malformations  related  to treatment  and  the con-
clusion  was  that  TBTO was  not  teratogenic. Postmortem

examination  of  the  mother rabbits  indicated no changes
associated  with treatment, and  1 mg/kg per day  was con-
sidered  to be the NOEL for maternal toxicity and toxicity
to the fetus.

    When Crofton et al. (1989) treated pregnant Long-Evans
rats  with TBTO by gastric intubation at 0 to 16 mg/kg per
day  body weight from day 6 to day 20 of gestation, litter
size and pup weight were significantly reduced by doses of
10,  12,  or 16 mg/kg  per day but  no effect was  seen at
doses of 2.5 and 5 mg/kg per day. Maternal weight gain was
affected by the same dose levels that affected litter size
and  pup weight. Pup survival  was further reduced in  the
first  3 days after birth. Litter size was reduced by 50%,
73%,  and 96% at dose  levels of 10, 12,  and 16 mg/kg per
day, respectively, on day 1 post partum and by  63%,  88%,
and 100% on day 3 post partum.  Pup weight on  day 1  post
partum  was reduced by 45%, 45%, and 68% at the three dose
levels. Two out of 71 pups born had cleft palate,  but  no
controls  showed  any  abnormalities. These  two pups were
both born dead.  The five pups born to rats given 16 mg/kg
per  day showed no  malformations, but all  died within  3
days.   The authors concluded  that it was  impossible  to
distinguish  between  possible  fetotoxicity and  maternal
toxicity,  since all effects on offspring occurred at TBTO
dose  rates that also affected  the weight of the  mother.
They  also showed that pregnant female rats were more sen-
sitive to TBTO than non-pregnant females; the MTD (maximum
tolerated  dose) for non-pregnant females was 16 mg/kg per
day, whereas pregnant females showed an MTD of  between  5
and 10 mg/kg per day. Most effects on pups were transitory
in that survivors to adulthood showed only  reductions  in
body  weight and brain weight compared to controls even at
the  highest dose rates.   Motor activity of offspring was
monitored  in a maze fitted  with photoelectric detectors.
During  the  pre-weaning  period there  was  a significant
age/treatment  interaction, but only on  post-natal day 14
was there a significant response per individual  day.  All
doses  of TBTO produced a significant decrease in activity
on that day.  Post-weaning activity was reduced  on  post-
natal  days 47 and 62  (p < 0.01) but only  at a level  of
10 mg/kg  per day. There was  no clear effect on  acoustic
startle response.

    In a two-generation reproduction study, TBTO was given
to  rats at dietary  concentrations of 0,  0.5, 5, and  50
mg/kg.  The  parental  (F0)   generation  (30 males and 30
females per group) was exposed for 10 weeks before mating,
whereas the pre-mating treatment period for the  F1 adults
(30 of  each  sex  per  group)  was  15 weeks.  Culling of
litters  was performed on day 4 in the F1 and  F2   gener-
ations.   Preliminary  data  (Biodynamics, 1989b)  on mor-
tality, body weight development, fertility indices, litter
data, and organ weights have been reported. There  was  no
evidence  of  compound-related mortality,  and body weight
development was normal in the F0   generation. At 50 mg/kg

diet,  the pup weights were decreased on days 14 and 21 in
the  F1   generation  and on  days 7,  14, and  21 in  the
F2 generation.   Among the F1 parents,  at 50 mg/kg, lower
body weights were noted in males throughout the pre-mating
period  and in females during  the first 3 weeks only.  No
effects  on  mating,  pregnancy, and  fertility rates were
noted  in either generation,  and number of  pups,  litter
size, and pup survival were not affected by  treatment  in
either  the F1   or  the F2   generations.   Relative  and
absolute  thymus weights were  decreased in both  sexes at
the dietary concentration of 50 mg/kg.

11.5.2.  In vitro

    Krowke et al. (1986) demonstrated an effect of TBTO on
the  development of limb  buds of mice  in organ  culture.
Clear-cut  interference  with differentiation  was seen at
the  lowest dose tested  (0.03 mg/litre), while at  0.1 mg
per  litre drastic impairment and  abnormal development of
the paw skeleton were recorded. The effect was  more  pro-
nounced  still at the highest  dose tested (0.3 mg/litre).
In  view of their failure to show effects on the embryo  in
 vivo, Davis  et al. (1987)  suggested that there  is  only
very  limited  movement of  TBTO  across the  placenta  of

11.6.  Carcinogenicity


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

    Wester (in press) reported the results of  a  106-week
study  on carcinogenicty in  Wistar rats at  dietary  TBTO
doses  of  0, 0.5,  5, and 50 mg/kg  (0, 0.025, 0.25,  and
2.5 mg/kg body weight). At the highest dose level, general
toxicological effects were present (see section 11.3). The
incidence  of benign tumours of the pituitary (mainly pro-
lactinomas) was elevated at 0.5 and 50 mg/kg, but  not  at
5 mg/kg diet, for both sexes. At 50 mg/kg,  a  significant
increase was noted in the incidence of  adrenal  medullary
tumours  (pheochromocytomas) in both sexes and of parathy-
roid  adenomas  in male  animals,  while the  incidence of
adrenal  cortical  tumours was  significantly decreased at
0.5 and 50 mg/kg diet in males only.  Isolated  occurrence
of  pancreatic carcinoma was found in treated female rats.
These  were not considered  to be compound  related  since
there  was no dose dependency and the incidence rates were



     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 skin response.

12.1.  Ingestion

    There  have been no  reported cases of  poisoning from
ingestion of TBTO or other TBT salts.

12.2.  Inhalation

    Seventy  percent of the  workers in a  rubber  factory
using  TBTO in the vulcanizing process reported irritation
of the upper respiratory tract (and eyes).  About 20% also
experienced  lower chest symptoms  (irritation, tightness,
and pain), but in all cases pulmonary function  was  unaf-
fected.  The  extent  of  the  exposure  was  not recorded
(WHO/FAO, 1985).

12.3.  Dermal exposure

    Lyle  (1958) described skin  lesions in workers  occu-
pationally  exposed to dibutyl and tributyl tin compounds.
Skin  burns were most commonly caused by small splashes of
liquid  dibutyl or tributyl tin chlorides. Since the irri-
tancy  of  these  compounds was  not  immediately apparent
(taking  at least an hour to be perceived), small splashes
were frequently ignored by workers. More severe lesions on
the hands were caused by leaking gloves or failure to wear
hand  protection.  Some  maintenance staff  also  suffered
burns  after kneeling or rubbing against surfaces wet with
the  chlorides.  Clothes wet with these compounds produced
burns  on the ventral skin and burns were seen immediately
above  the level of  protective boots on  the calf of  the
leg.  Two kinds of lesion were seen in workers.  The first
was  an acute burn,  which healed relatively  quickly, and
the  second  a more  diffuse  dermatitis (seen  in workers
wearing  contaminated  clothes  in close  contact with the
skin), which persisted.  Treatment of the back of the hand
of  volunteers with various undiluted TBT compounds estab-
lished  that the chloride, acetate, laurate, and oxide all
produced  acute burns. Burns were not caused by dibutyltin
esters or oxide or by tetrabutyltin, but were  mostly  due
to  dibutyltin chloride or the tributyltin compounds. Red-
dening  of the area treated was seen 2-3 h after treatment
with  the compounds.  Inflammation  of the hair  follicles
was  the most obvious  symptom, effects being  confined to
the  treated area. On the second day after treatment, min-
ute  pustules appeared, which remained discrete and disap-
peared  on the third or fourth day. After a week, all that
remained was a faint erythema. The burns were not reported

to be painful, either in the case of experimental or occu-
pational  exposure.   Sufferers complained  of itching and
"stickiness"  of the skin causing  adherence of clothes.
Proper  use of protective  clothing, rapid washing  of the
skin  after exposure,  and the  use of  aprons to  prevent
wetting  of overalls were  found to be  effective in  pre-
venting burns, both acute and diffuse.

    Baaijens (1987) described cases of accidental exposure
to  TBTO  during  the manufacture  of organotin compounds.
Severe  dermatitis  developed  only where  splashes of the
material had been retained on the skin for  long  periods.
In  one case, a worker had been splashed over the face and
neck. He left the work area after the splash and showered.
An  area behind one ear had not been washed and the derma-
titis  had developed in this one area. There were no symp-
toms  other  than  dermatitis.  Another  worker  had  been
splashed on the arm. He washed his skin but did not change
his  overalls.  Contact was  extended and a  large blister
developed  on the arm. Monitoring of the urine tin content
showed  no difference from normal  in these two cases.   A
third  worker, who complained of an intense smell of TBTO,
suffered  nausea  and  vomiting after  10 min of exposure.
Urine  tin levels in this  case were elevated for  several
days. In all cases the symptoms disappeared within  a  few
days.  The delayed irritancy of tributyltin was emphasized
by the author. It tends to lead to extended exposure since
the  affected person does not perceive the effect for sev-
eral  hours.  The  author found  no  relationship  between
clinical  or  haematological  parameters and  normal occu-
pational exposure to organotin compounds.

    Goh (1985) reported irritant effects and dermatitis in
painters  applying TBTO formulations (0.6% TBTO in acrylic
resin-based  paints) to buildings. The  painters developed
rashes 8 to 10 h after exposure.  When required  to  paint
ceilings,  the men developed  more severe symptoms  on the
face,  neck,  and  trunk, associated  with dripping paint.
Severe  itching,  redness,  swelling, and  blistering were
recorded.  Hospital examination showed extensive vesiculo-
bullous lesions, erythema, and oedema of the  face,  neck,
trunk, arms, and thighs. Although only two  patients  were
examined in detail, most of the workers on the site devel-
oped dermatitis. Patch testing of the two  reported  cases
showed erosion 48 and 96 h after patch tests using aqueous
solutions of TBTO (0.1 and 0.5 g/litre). Five other volun-
teer  controls were patch-treated  with solutions of  0.01
and 0.1 g/litre; these also showed a similar  reaction  to
the  patients.  No allergenic  reaction  was found  in the
patch  tests.  Replacement of the paint with a preparation
having  similar  constituents  other than  TBTO  prevented
further problems amongst the painters.

    Lewis & Emmett (1987) describe contact dermatitis in a
shipwright  resulting  from  exposure  to  TBTO-containing
antifouling  paint. The man had been spray-painting blocks

of  wood and his skin had been exposed to the spray. There
was no immediate sensation, but some irritation  was  evi-
dent  within about an hour. Erythema and ulceration of the
exposed  areas were noted  on the second  day. There  were
also  some mild pustular lesions on the mucous membrane of
the  lips, presumed to be  the result of wiping  the mouth
with  paint-contaminated arms. The authors conducted patch
testing  with  TBTO  at aqueous  concentrations  of  0.1%,
0.01%, and 0.0015% using the A1 test tape method.  At  the
highest  dose tested, there  were large bullae  after 48 h
and crusting after 96 h. The other two doses  produced  no
adverse  effects.  The  delayed sensation  and effect were
highlighted  by the authors  as being particular  problems
with  TBTO. Great care to prevent any exposure to products
containing TBTO at high concentrations was recommended.

    Molin  & Wahlberg (1975)  investigated an outbreak  of
dermatitis  on the feet  and ankles of  trainee  soldiers.
Seventy  soldiers  reported itching  and sometimes painful
erythematous,  vesiculous  to  bullous,  and  haemorrhagic
lesions  after long  marches on  a hot  day.  Fifty  other
soldiers   reported   less  severe   symptoms  on  another
occasion. The skin lesions disappeared within 1 or 2 weeks
in most cases after treatment with saline  compresses  and
topical steroid creams. In two cases, the skin was red and
slightly tender more than 6 months later. The outbreak was
traced  to a single batch of socks that had been soaked in
7 times  the  recommended concentration  of a disinfectant
solution containing TBTO. Recommended use of the disinfec-
tant  was  calculated to  give  about 0.001%,  whereas the
concentration  after  the  accidental over-use  would have
been  about 0.01%. Patch  testing of eczema  patients  not
previously  exposed  to  TBTO suggested  that  the primary
irritant  concentration for the  compound was between  0.1
and  0.01%.  Soldiers  patch-tested 2 months  after  their
dermatitis  had fully healed gave negative results to both
0.001%  and  0.01% TBTO.  The  authors suggested  that the
effective  concentration for TBTO as a disinfectant is too
close  to the primary  irritancy concentration to  justify
the use of TBTO for disinfecting textiles.

    Zedler  (1961), on the basis of industrial experience,
stated  that a concentration of 0.05% TBTO was not harmful
to the human skin.

12.4.  Miscellaneous effects

    Women  using a latex spray paint containing TBTO as an
additive  showed  immediate  irritant effects  on  the eye
(profuse  lacrimation,  eye  inflammation) and  the  nasal
mucosa.  The symptoms worsened  over 14 days of  spraying,
but  subsided at the  weekends and disappeared  completely
when addition of TBTO to the paint was  discontinued.  The
extent of exposure in this case was not recorded (WHO/FAO,

    Akatsuka et al. (1959) reported a case of occupational
poisoning  with butyltin compounds where, along with symp-
toms  of lassitude, slight occipital headaches, and stiff-
ness in the shoulders, there was a marked  disturbance  of
the  sense of smell. The authors conducted studies on cats
to  test whether this effect could be confirmed experimen-
tally. Exposure of the cats to a vapour mixture containing
mainly tributyltin bromide revealed a marked loss  of  the
sense of smell.


13.1.  Evaluation of human health risks

    Exposure  of  workers  occurs principally  during  the
manufacture  and formulation of tributyltin  compounds, in
the  application and removal of  TBT paints, and from  the
use of TBT in wood preservatives.  Exposure of the general
public may come from the contamination of  food,  particu-
larly fish and shellfish, and from domestic application of
wood preservatives.

    On  the basis of both  animal tests and direct  obser-
vations  on humans,  it is  clear that  TBT compounds  are
irritant to the skin and eyes and that inhalation of aero-
sols leads to respiratory irritation.

    The  handling of treated wood poses no dermal irritant
hazard  once the wood has dried.  However, aerosols of TBT
are  very  hazardous and  re-entry  to the  treatment area
should be prohibited until the wood has thoroughly dried.

    Acute  systemic poisoning has never  been reported and
clearance of TBT from the body is expected to occur within
a few days.  Acute toxicity from handling TBT products is,
therefore, unlikely if proper precautions are taken.

    Short- and  long-term effects on  experimental animals
have  been reported in  the liver and  haematological  and
endocrine  systems.  The effects  of TBT compounds  on the
immune  system, and particularly on  host resistance, have
proved  the most sensitive  parameter of toxicity  in  the
rat,  the most sensitive species tested.  The no-observed-
effect  level (NOEL), using the  Trichinella spiralis host-
resistance  model,  lies  between 0.5  and  5.0 mg/kg diet
(0.025 and 0.25 mg/kg body weight), whereas using measures
of immune function it is 0.6 mg/kg body weight.

    Owing to wide variation in the consumption of fish and
shellfish and local differences in residues of TBT in sea-
food,  only  illustrative estimates  relating exposure and
NOEL values can be made.  It needs to be  emphasized  that
local  measurements of residues,  local estimates of  sea-
food consumption, and local decisions on acceptable safety
margins  must be made  to assess potential  risk of  these

    Using  fish consumption figures of 15 and 150 g/day, a
value  of 1 mg/kg  for residues  in fish,  and an  average
human body weight of 60 kg, the following  safety  margins
based on different immune endpoints are obtained.

   Fish         Estimated           Safety margin
consumption    daily intake     T. spiralis     Other
  (g/day)        of TBT          model         immune
                (µg/kg)                        parameters
    15           0.25          100-1000        2500
   150           2.5            10-100          250

    Indiscriminate  and irresponsible use of TBT compounds
and a failure to follow the recommendations,  outlined  in
this monograph, to reduce exposure of humans could lead to
intake  of  levels of  TBT  compounds hazardous  to  human

    Teratogenic effects have only occurred in experimental
animals at doses that caused overt maternal toxicity.  The
teratogenic  potential of TBT is, therefore, considered to
be very low.

    Based  on  the  results of  comprehensive mutagenicity
studies,  tributyltin compounds are not considered to have
mutagenic  potential.  In a carcinogenicity  study on rats
with  TBTO, an increased incidence was noted for endocrine
tumours  that occur spontaneously  at a high  and variable
incidence.   Therefore,  the  available evidence  does not
clearly demonstrate a carcinogenic hazard of TBT compounds
for humans.

13.2.  Evaluation of effects on the environment

    Diffuse  input of tributyltin (TBT)  into the environ-
ment occurs predominantly from the use of TBT in antifoul-
ing paint. It could also occur if it were used as  a  mol-
luscicide.   Point source contamination  occurs if TBT  is
used  as  a  biocide  in  cooling  systems,  wood pulping,
leather  processing, wood preservation processes, and tex-
tile treatment.

    Due  to  their  physico-chemical properties,  TBT com-
pounds  concentrate in the surface microlayer and in sedi-
ments.   Abiotic degradation does not appear to be a major
mechanism  of  removal  under  environmental   conditions.
Although TBTO is biodegradable in the water  column,  this
process is not rapid enough to prevent the  occurrence  of
elevated TBT levels in some areas.  Bioaccumulation occurs
in  most  aquatic  organisms, but  in  laboratory mammals,
metabolic degradation is a more efficient process.

    TBT  is extremely hazardous to  some aquatic organisms
because it is toxic at very low concentrations  in  water.
Such concentrations have been found in some areas. Adverse
effects  on  non-target  invertebrates, particularly  mol-
luscs, have been reported in field studies, and these have

been  sufficiently severe to lead  to reproductive failure
and population decline.  Adverse effects on the commercial
production of shellfish have been successfully reversed by
restrictions  on  the use  of  antifouling paints  in some
areas,  and  these restrictions  are  also leading  to the
reversal of imposex effects in gastropod populations.  The
effects on farmed fish indicate that TBT-containing paints
should not be used on restraining nets.

    The  general hazard to the  terrestrial environment is
likely to be low.  TBT-treated wood could pose a hazard to
terrestrial organisms living in close contact with it.

    The  enhancement of TBT concentrations  in the surface
microlayer  may  present  a hazard  to littoral organisms,
neustonic species (including benthic invertebrate and fish
larvae)   and  surface-feeding  sea-birds   and  wildfowl.
Accumulation  and low biodegradation  of TBT in  sediments
may  present a hazard to aquatic organisms when these pol-
luted  sediments  are  disturbed by  natural  processes or
dredging activities.


14.1.  Recommendations for protecting human and environmental health

a)  Member countries that have not yet regulated  the  use
    of TBT compounds should be encouraged to do so.

b)  There  is  a need  for  evaluation and,  if necessary,
    regulation  of organotin input to the environment from
    sources  other than antifouling paints.   For example,
    this  would include evaluation  of the potential  risk
    from the application of TBT-contaminated sewage sludge
    to soil.

c)  Improved  methods  for the  safe application, removal,
    and disposal of organotin paints should be developed.

14.2.  Research needs

a)  Methods  of detection and analysis need to be improved
    to provide rapid and accurate measurements of butyltin
    species  in  pg/litre  concentrations. One  reason for
    this  recommendation is that a biological effect, i.e.
    imposex  in gastropods, may occur at levels lower than
    present detection limits.

b)  There  is a need  for research into  mechanisms  which
    concentrate  rather than disperse TBT and which retard
    degradation,  with particular attention to  the funda-
    mental  chemistry of TBT and its interaction with bio-
    logical molecules.  More study is needed on the uptake
    of TBT at all trophic levels.

c)  A study of the toxicity of TBT in aquatic organisms is
    required.   This  work should  investigate metabolism,
    endocrine  effects, and immunological  toxicity, where

d)  A  search for other sensitive  bioindicator species in
    other   groups,   including  freshwater   species,  is

e)  Models for the assessment of immunotoxicity in mammals
    need to be validated and no-effect levels for relevant
    parameters need to be defined more accurately.

f)  A  long-term  toxicity  study in  a  second  mammalian
    species should be undertaken.

g)  A  tumorigenicity study in a  second mammalian species
    should be undertaken.

h)  Information  on  butyltin  residue levels  in fish and
    shellfish  for  human  consumption  using   speciating
    methods is needed.


(1987) Assessment and regulatory actions for TBT in the UK. In:  Proceedings
 of the  Organotin Symposium,  Oceans '87  Conference, Halifax, Nova Scotia,
 Canada, 28 September-1 October, 1987, New York, The Institute of Electrical
and Electronics Engineers, Inc., Vol. 4, pp. 1314-1319.

NAGASAKI, T.,  KOTANI, Y.,  MATSUTANI, W.,  FUKUDA, I.,  & IYO,  T.  (1959)
[Experimental studies  on disturbance  of sense  of smell  due to  butyltin
compounds.]  J. Tokyo med. Coll., 17: 1393-1402 (in Japanese).

ALABASTER, J.S.  (1969) Survival  of fish  in 164 herbicides, insecticides,
fungicides, wetting agents and miscellaneous substances.  Int. Pest Control,
11: 29-35.

ALDRIDGE, W.N. (1958) The biochemistry of organotin compounds. Trialkyltins
and oxidation phosphorylation.  Biochem. J., 69: 367-376.

ALDRIDGE, W.N. & STREET, B.W. (1964) Oxidative phosphorylation: Biochemical
effects and properties of trialkyltins.  Biochem. J., 91: 287-297.

ALDRIDGE,  W.N.  &  STREET,  B.W.  (1970)  Oxidative  phosphorylation:  The
specific binding of trimethyltin and triethyltin to rat liver mitochondria.
 Biochem. J., 118: 171-179.

ALDRIDGE,  W.N.  &  STREET,  B.W.  (1971)  Oxidative  phosphorylation:  The
relation between  the specific  binding of  trimethyltin and triethyltin to
mitochondria and their effects on various mitochondrial functions.  Biochem.
 J., 124: 221-234.

(1977) Action on mitochondria and toxicity of metabolites of tri-n-butyltin
derivatives.  Biochem. Pharmacol., 26: 1997-2000.

ALLEN, A.J., QUITTER, B.M., & RADICK, C.M. (1980) The biocidal mechanism of
controlled release bis (tri-n-butyltin) oxide in  Biomphalaria glabrata. In:
Baker, R., ed.  Controlled release of bioactive materials, New York, London,
Academic Press, p. 399.

ALZIEU, C.P.  (1981)  Evaluation  des risques  dus à  l'emploi des peintures
 anti-salissures dans les zones conchylicoles, Nantes, Institut scientifique
et technique des Pêches maritimes, 84 pp.

ALZIEU, C.  &  HERAL,  M.  (1984)  Ecotoxicological  effects  of  organotin
compounds on oyster culture.  Ecotoxicol. Test. mar. Environ., 2: 187-196.

ALZIEU, C.  & PORTMANN,  J.E. (1984)   The effect  of tributyl  tin  on  the
 culture of  C. gigas.  Proceedings of the 15th Annual Shellfish Conference,
 15-16 May, 1984, London, The Shellfish Association of Great Britain, 17 pp.

Influence des  peintures antisalissures  à base  d'organostanniques sur  la
calcification de  la coquille  de l'huitre   Crassostrea gigas.  Rev.  Trav.
 Inst. Pêches Marit., 45: 101-116.

ALZIEU,  C.,   SANJUAN,  J.,   DELTREIL,  J.P.,  &  BOREL,  M.  (1986)  Tin
contamination in  Arcachon Bay:  Effects on  oyster shell  anomalies.   Mar.
 Pollut. Bull., 17: 494-498.

ALZIEU, C.,  SANJUAN, J.,  MICHEL, P.,  BOREL, M.,  &  DRENO,  J.P.  (1989)
Monitoring and  assessment of  butyltins in  Atlantic coastal  waters.  Mar.
 Pollut. Bull., 20: 22-26.

DRIESSCHE, J.  (1976) Effets chez le cobaye, d'un aérosol à base d'oxyde de
tributylétain (OTBE).  Eur. J. Toxicol., 9: 339-346.

ARAKAWA, Y.  & WADA,  O. (1984)  Inhibition  of  neutrophil  chemotaxis  by
organotin compounds.  Biochem. biophys. Res. Commun., 123: 543-548.

ARGAMAN, Y., HUCKS, C.E., & SHELBY, S.E. (1984) The effects of organotin on
the activated sludge process.  Water Res., 18: 535-542.

 effect of  hexabutyldistannoxane (TBTO)  and other  molluscicides  on  non-
 target species,  Geneva,  World  Health  Organization,  Parasitic  Diseases
Programme (Unpublished report).

BAAIJENS, P.A. (1987) Health effect screening and biological monitoring for
workers in  organotin industries.  In:   Toxicology  and  analytics  of  the
 tributyltins: The  present  status.  Proceedings  of  an  ORTEPA  workshop,
 Berlin,  15-16   May,  1986,   Vlissingen-Oost,  The   Netherlands,  ORTEP-
Association, pp. 191-208

BACCI, E.  & GAGGI,  C. (1989)  Organotin compounds  in harbour  and marina
waters from the Northern Tyrrhenian Sea.  Mar. Pollut. Bull., 20: 290-292.

BAHR, G.  & PAWLENKO,  S. (1978)  Organic  tin  compounds.  In:  Bahr,  G.,
Kalinowski, H.-O., & Pawlenko, S., ed.  Organometallic compounds, germanium,
 tin, Stuttgart,  Georg Thieme  Verlag,  pp.  512-515  (Methods  in  Organic
Chemistry series).

BAILEY, S.K. & DAVIES, I.M. (1988a) Tributyltin contamination around an oil
terminal in Sullom Voe (Shetland).  Environ. Pollut., 55: 161-172.

BAILEY, S.K.  & DAVIES, I.M. (1988b) Tributyltin contamination in the Firth
of Forth (1975-87).  Sci. total Environ., 76: 185-192.

BAKER, J.M.  & TAYLOR, J.M. (1967) The toxicity of tributyltin oxide to the
wood-boring beetles   Lyctus brunneus  Steph. and   Anobium punctatum (Deg.).
 Ann. appl. Biol., 60: 181-190.

BALLS, P.W.  (1987) Tributyltin  (TBT) in the waters of a Scottish sea loch
arising from  the use  of antifoulant  treated  netting  by  salmon  farms.
 Aquaculture, 65: 227-237.

BARNES, J.M.  & STONER,  H.B. (1958)  Toxic properties  of some dialkyl and
trialkyl tin salts.  Br. J. ind. Med., 15: 15-22.

BARUG,  D.   (1981)  Microbial  degradation  of  bis  (tributyltin)  oxide.
 Chemosphere, 10: 1145-1154.

BARUG,  D.   &  VONK,  J.W.  (1980)  Studies  on  the  degradation  of  bis
(tributyltin) oxide in soil.  Pestic. Sci., 11: 77-82.

BEAUMONT, A.R.  & BUDD,  M.D. (1984)  High mortality  of the  larvae of the
common mussel at low concentrations of tributyltin.  Mar. Pollut. Bull., 15:

BEAUMONT, A.R.  & NEWMAN,  P.B. (1986)  Low levels  of  tributyltin  reduce
growth of marine micro-algae.  Mar. Pollut. Bull., 17: 457-461.

BEAUMONT, A.R.,  NEWMAN,  P.B.,  &  WALDOCK,  M.J.  (1987)   Sand  substrate
 microcosm studies on tributyl tin (TBT) toxicity to marine organisms, 24 pp
(Final report to the UK Department of the Environment, London. Contract No.
PECD 7/8/73).

BERRIOS-DURAN, L.A.  &  RITCHIE,  L.S.  (1968)  Molluscicidal  activity  of
bis(tri-n-butyltin) oxide  formulated in rubber.  World Health Organ. Bull.,
39: 310-312.

BIODYNAMICS (1989a)  A three month oral range-finding toxicity study in mice
 with  bis  (tri-n-butyltin)  oxide  (TBTO),  East  Millstone,  New  Jersey,
Biodynamics (Final  report to Aceto Chemical Corporation, M. + T. Chemicals
Inc., and Schering Berlin Inc.) (Unpublished).

BIODYNAMICS (1989b)   A two-generation reproduction study in rats with TBTO,
East Millstone,  New Jersey,  Biodynamics (Interim report to Aceto Chemical
Corporation,  M.   +  T.   Chemicals  Inc.,   and  Schering   Berlin  Inc.)

BJORKLUND, I.  (1987a)  [Environmental  aspects of  ship's  bottom  paints],
Stockholm, Swedish  Chemicals Inspectorate, Investigation Department, 16 pp
(in Swedish).

BJORKLUND,  I.   (1987b)  [Environmental  effect  of  antifouling  paints],
Stockholm, Swedish National Environment Protection Board, 17 pp.

BLAIR,  W.R.,   OLSON,  G.J.,  BRINCKMAN,  F.E.,  &  IVERSON,  W.P.  (1982)
Accumulation and  fate of  tri-n-butyltin  cation  in  estuarine  bacteria.
 Microbiol. Ecol., 8: 241-251.

BLAIR, W.R.,  OLSON, G.J., & BRINCKMAN, F.E. (1986) Speciation measurements
of butyltins:  Application to  controlled release  rate  determination  and
production  of  reference  standards.  In:   Proceedings  of  the  Organotin
 Symposium, Oceans  '86 Conference,  Washington, DC,  USA, 23-25  September,
 1986, New  York, The  Institute of  Electrical and  Electronics  Engineers,
Inc., Vol. 4, pp. 1141-1145.

BLUNDEN, S.J.  & CHAPMAN, A. (1986) Organotin compounds in the environment.
In: Craigh,  P.J., ed.  Organometallic compounds in the environment, Harlow,
Essex, Longman Group Ltd, pp. 111-159.

BLUNDEN,  S.J.,  HOBBS,  L.A.,  &  SMITH,  P.J.  (1984)  The  environmental
chemistry of organotin compounds. In: Bowen, H.J.M, Blunden, S.J., Colbeck,
I., Harrison,  R.M., Hobbs,  L.A., Katz,  S.A., Simkiss, K., Smith, P.J., &
Taylor,  M.G.,  ed.   Environmental  chemistry,  London,  Royal  Society  of
Chemistry, Vol. 3, pp. 49-77.

BOKRANZ, A.  & PLUM,  H. (1975)  Industrial manufacture and use of organotin
 compounds, Bergkamen, Federal Republic of Germany, Schering AG, 33 pp.

BOORMAN, L.A.  (1989) The  effects of  TBT on two salt marsh plant species,
 Aster tripolium and  Limonium vulgare. In Preparation.

BOULDIN,  T.W.,  GOINES,  N.D.,  BAGNELL,  C.R.,  &  KRIGMAN,  M.R.  (1981)
Pathogenesis  of   trimethyltin  neuronal   toxicity.  Ultrastructural  and
cytochemical observations.  Am. J. Pathol., 104: 237-249.

BRAMAN, R.S.  & TOMPKINS,  M.A.  (1979)  Separation  and  determination  of
nanogram amounts  of inorganic  and methyltin compounds in the environment.
 Anal. Chem., 51: 12-19.

BRESSA, G.,  CIMA, L., CANOVA, F., & CARAVELLO, G.U. (1984) Bioaccumulation
of  tin  in  the  fish  tissues   (Liza  aurata).  In:  Proceedings  of  the
 International Conference  on Environmental Contamination, London, July 1984
- Geneva,  International Register  of Potentially  Toxic Chemicals,  United
Nations Environment Programme, pp. 812-815.

BRIDGES, J.W.,  DAVIES, D.S.,  & WILLIAMS, R.T. (1967) The fate of ethyltin
and diethyltin derivatives in the rat.  Biochem. J., 105: 1261-1266.

BRINCKMAN, F.E. (1981) Environmental organotin chemistry today: Experiences
in the field and laboratory.  J. organomet. Chem. Libr., 12: 343-376.

(1983) Ultratrace  speciation and biogenesis of methyltin transport species
in estuarine waters. In: Trace metals in sea water, New York, Plenum Press,
pp. 39-72.

(1977) A  comparison of  the half-life  of inorganic and organic tin in the
mouse.  Environ. Res., 13: 56-61.

BRYAN, G.W.,  GIBBS, P.E.,  HUMMERSTONE, L.G.,  &  BURT,  G.R.  (1986)  The
decline of  the  gastropod   Nucella  lapillus  around  south-west  England:
Evidence for  the effect  of tributyltin  from antifouling  paints.  J. Mar.
 Biol. Assoc. UK, 66: 611-640.

BRYAN, G.W.,  GIBBS, P.E.,  HUMMERSTONE, L.G.,  & BURT, G.R. (1987) Copper,
zinc, and  organotin as  long-term factors  governing the  distribution  of
organisms in the Fal estuary in southwest England.  Estuaries, 10: 208-219.

BRYAN, G.W.,  GIBBS,  P.E.,  &  BURT,  G.R.  (1988)  A  comparison  of  the
effectiveness of tri-n-butyltin chloride and five other organotin compounds
in promoting  the development of imposex in the dog-whelk  Nucella lapillus.
 J. Mar. Biol. Assoc. UK, 68: 733-744.

BUSHONG, S.J.,  HALL, W.S.,  JOHNSON, W.E., & HALL, L.W. (1987) Toxicity of
tributyltin to  selected Chesapeake  Bay  biota.  In:   Proceedings  of  the
 Organotin Symposium,  Oceans '87  Conference, Halifax, Nova Scotia, Canada,
 28 September-1  October, 1987,  New York,  The Institute  of Electrical and
Electronics Engineers, Inc., Vol. 4, pp. 1494-1503.

Acute  toxicity   of  tributyltin  to  selected  Chesapeake  Bay  fish  and
invertebrates.  Water Res., 22: 1027-1032.

CALLEY, D.J.,  GUESS, W.L.,  & AUTIAN, J. (1967) Hepatotoxicity of a series
of organotin esters.  J. pharm. Sci., 56: 240-243.

CARDARELLI,  N.F.  (1973)   Effects  of  ultralow  molluscicide  dosages  on
 Biomphalaria glabrata and Lebistes reticulatus in micro ecological systems:
 Report I, Akron, Ohio, University of Akron, 15 pp.

CARDARELLI,  N.F.   (1978)  Controlled   release  organotins   as  mosquito
larvicides.  Mosq. News, 38: 328-333.

CARDARELLI,  N.F.  &  EVANS,  W.  (1980)  Chemodynamics  and  environmental
toxicology of  controlled release organotin molluscicides. In: Baker, R.W.,
ed.  Controlled  release of  bioactive materials.  Proceedings  of  the  6th
 International Meeting  of the Controlled Release Society, New York, London,
Academic Press, pp. 357-385.

CASIDA, J.E.,  KIMMEL, E.C.,  HOLM, B.,  &  WIDMARK,  G.  (1971)  Oxidative
dealkylation of tetra-, tri-, and dialkyltin and tetra- and trialkyleads by
liver microsomes.  Acta chem. Scand., 25: 1497-1499.

CHAMP,  M.A.   &  PUGH,   W.L.  (1987)   Tributyltin  antifouling   paints:
Introduction and  overview. In:   Proceedings of  the  Organotin  Symposium,
 Oceans  '87  Conference,  Halifax,  Nova  Scotia,  Canada,  28  September-1
 October, 1987,  New York,  The  Institute  of  Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1296-1308.

CHAPMAN, A.H.  & PRICE,  J.W. (1972) Degradation of triphenyltin acetate by
ultraviolet light.  Int. Pest Control, 14: 11-12.

CHENG, Z.  & JENSEN, A. (1989) Accumulation of organic and inorganic tin in
blue mussel,  Mytilus edulis, under natural conditions.  Mar. Pollut. Bull.,
20: 281-286.

CHLIAMOVITCH, Y.-P.  & KUHN,  C.  (1977)  Behavioural,  haematological  and
histological studies  on acute  toxicity of  bis(tri- n-butyltin)  oxide  on
 Salmo gairdneri  Richardson and   Tilapia rendalli Boulenger.  J. Fish Biol.,
10: 575-585.

CHU, K.Y.  (1976) Effects  of environmental  factors on  the  molluscicidal
activities of  slow-release hexabutyldistannoxane and copper sulfate.  Bull.
 World Health Organ., 54: 417-420.

CLARK, J.R.,  PATRICK, J.M.,  MOORE, J.C.,  & LORES, E.M. (1987) Waterborne
and sediment-source  toxicities of  six organic  chemicals to  grass shrimp
 (Palaemonetes  pugio)   and  amphioxus    (Branchiostoma  caribaeum).  Arch.
 environ. Contam. Toxicol., 16: 401-407.

CLEARY, J.J.  & STEBBING,  A.R.D. (1985) Organotin and total tin in coastal
waters of southwest England.  Mar. Pollut. Bull., 16: 350-355.

CLEARY, J.J.  & STEBBING, A.R.D. (1987) Organotin in the surface microlayer
and subsurface  waters of  Southwest England.   Mar. Pollut. Bull., 18: 238-

CREMER, J.E.  (1957) The  metabolism  in vitro of  tissue slices  from rats
given triethyltin compounds.  Biochem. J., 67: 87-96.

CHERNOFF, N.,  & REITER,  L.W. (1989)  Prenatal or  postnatal  exposure  to
bis(tri- n-butyltin)oxide in the rat: Postnatal evaluation of teratology and
behavior.  Toxicol. appl. Pharmacol., 97: 113-123.

DAVIDSON, B.M.,  VALKIRS, A.O.,  & SELIGMAN,  P.F. (1986) Acute and chronic
effects of  tributyltin  on  the  mysid   Acanthomysis  sculpta  (Crustacea,
Mysidacea).  In:    Proceedings  of  the  Organotin  Symposium,  Oceans  '86
 Conference, Washington,  DC, USA,  23-25 September,  1986,  New  York,  The
Institute of Electrical and Electronics Engineers, Inc., pp. 1219-1225.

DAVIES, R.J.,  FLETCHER, R.L.,  & FURTADO,  S.E.J. (1984)  The  effects  of
tributyltin compounds  on spore  development in the green alga  Enteromorpha
 intestinalis  (L) Link. In:  Proceedings of the 6th International Congress on
 Marine Corrosion and Fouling, Athens, September 1984, pp. 557-565.

DAVIES, I.M.,  DRINKWATER, J.,  MCKIE, J.C., & BALLS, P. (1987a) Effects of
the use  of tributyltin antifoulants in mariculture. In:  Proceedings of the
 Organotin Symposium,  Oceans '87  Conference, Halifax, Nova Scotia, Canada,
 28 September-1  October, 1987,  New York,  The Institute  of Electrical and
Electronics Engineers, Inc., pp. 1477-1481.

DAVIES, I.M.,  BAILEY, S.K.,  & MOORE, D.C. (1987b) Tributyltin in Scottish
sea Lochs,  as indicated  by degree  of imposex  in the  dogwhelk,   Nucella
 lapillus (L.).  Mar. Pollut. Bull., 18: 400-404.

BARTSCH, H.  (1987) Evaluation  of the  genetic and  embryotoxic effects of
bis(tri- n-butyltin) oxide  (TBTO), a  broad-spectrum pesticide, in multiple
 in vivo and  in vitro short-term tests.  Mutat. Res., 188: 65-95.

DESCHIENS, R.  & FLOCH,  H. (1962) Les propriétés molluscicides du chlorure
et de  l'acétate de  triphénylétain dans  le cadre  de la  prophylaxie  des
bilharzioses.  C. R. Acad. Sci. Paris, 255: 1236-1237.

DESCHIENS,  R.   &  FLOCH,  H.  (1968)  Action  biologique  comparée  de  6
molluscicides chimique  dans le  cadre de  la prophylaxie des bilharzioses.
Conclusions pratiques.  Bull. Soc. Pathol. Exot., 61: 640-650.

DESCHIENS, R.,  BROTTES, H., & MVOGO, L. (1966) Application sur le terrain,
au Cameroun,  dans la prophylaxie des bilharzioses de l'action molluscicide
de l'oxyde de tributylétain.  Bull. Soc. Pathol. Exot., 59: 968-973.

DE VILLIERS,  J.P.  &  MACKENZIE,  J.G.  (1963)  Organotin  and  organolead
molluscicides,  Geneva,   World  Health  Organization,  Parasitic  Diseases
Programme, p. 63 (Unpublished report Mol/Inf/13).

DIXON, D.R.  &  MCFADZEN,  I.  (1987)  Bis(tributyltin)  oxide  (TBTO),  an
antifouling compound,  promotes SCE  induction in  the larvae of the common
mussel,  Mytilus edulis. Mutagenesis, 2: 312.

DIXON, D.R.  & PROSSER, H. (1986) An investigation of the genotoxic effects
of an  organotin  antifouling  compound  (bis(tributyltin)  oxide)  on  the
chromosomes of  the edible mussel,  Mytilus edulis. Aquat. Toxicol., 8: 185-

DOJMI DI DELUPIS, G., GUCCI, P.M.B., & VOLTERRA, L. (1987) Toxic effects of
bis-tributyltinoxide on phytoplancton.  Main Group Metal Chem., 10: 77-82.

DONARD, O.,  RAPSOMANIKIS, S., & WEBER, J.H. (1986) Speciation of inorganic
tin  and   alkyltin  compounds  by  atomic  absorption  spectrometry  using
electrothermal quartz  furnace after  hydride generation.   Anal. Chem., 58:

EAJ (1988)   Outline of TBT compounds monitored in Japan, Tokyo, Environment
Agency of Japan (OECD Clearing House Project on Organotins).

EBDON, L., EVANS, K., & HILL, S. (1988) The variation of tributyltin levels
with time  in selected  estuaries prior  to the introduction of regulations
governing the  use of  tributyltin-based anti-fouling  paints.   Sci.  total
 Environ., 68: 207-223.

EBDON, L.,  EVANS, K.,  & HILL, S. (1989) The accumulation of organotins in
adult and seed oysters from selected estuaries prior to the introduction of
U.K. regulations governing the use of tributyltin-based antifouling paints.
 Sci. total Environ., 83: 63-84.

ELFERINK, J.G.R.,  DEIERKAUF, M.,  & VAN  STEVENINCK, J. (1986) Toxicity of
organotin  compounds   for  polymorphonuclear  leukocytes:  the  effect  on
phagocyctosis and exocytosis.  Biochem. Pharmacol., 35: 3727-3732.

ELSEA, J.R.  & PAYNTER,  O.E. (1958)  Toxicological studies  on  bis(tri-n-
butyltin) oxide.  Am. Med. Assoc. Arch. Ind. Health, 18: 214-217.

EVANS, C.J.  & KARPEL,  S. (1985) Organotin compounds in modern technology.
 J. organomet. Chem. Libr., 16: 178-217.

EVANS, D.W  & LAUGHLIN, R.B. (1984) Accumulation of bis (tributyltin) oxide
by the mud crab,  Rhitropanopeus harrisii. Chemosphere, 13: 213-219.

EVANS, W.H.,  CARDARELLI, N.F.,  &  SMITH,  D.J.  (1979)  Accumulation  and
excretion of  [1-14C] bis  (tri- n-butyltin)  oxide  in  mice.   J.  Toxicol.
 environ. Health, 5: 871-877.

FENT, K.  (1989a) Organotin  speciation in  municipal wastewater and sewage
slude: Ecotoxicological consequences.  Mar. environ. Res., 28: 477-483.

FENT, K.  (1989b)  Teratogenic  effects of  tributyltin on  embryos  of  the
 freshwater fish  Phoxinus phoxinus.  Presented  at  the  OECD  Workshop  on
Tributyltin: Activities  related to  Field Studies  at Sea, Centre IFREMER,
Nantes, 27-29 June, 1989 (Unpublished report).

FENT, K.,  FASSBIND, R.,  & SIEGRIST,  H. (1989)   Organotins in a municipal
 wastewater treatment  plant. Proceedings  of the 1st European Conference on
 Ecotoxicology, Copenhagen,  Denmark, 17-19  October, 1988, Lyngby, Denmark,
The Technical  University of  Denmark, Laboratory of Environmental Sciences
and Ecology.

FERAL, C.  & LE  GALL, S.  (1983)  The  influence  of  a  pollutant  factor
(tributyltin)  on   the  neuroendocrine   mechanism  responsible   for  the
occurrence of  a penis in the females of  Ocenebra erinacea. In: Lever, J. &
Boer,  H.H.,   ed.   Molluscan   neuroendocrinology.  Proceedings   of   the
International Minisymposium  on Molluscan  Endocrinology,  Amsterdam, North-
Holland Publishing Company, pp. 173-175.

FISH, R.H.,  KIMMEL, E.C.,  & CASIDA,  J.E. (1975)  Bioorganotin chemistry:
Biological oxidation  of tributyltin  derivatives.  J. organomet. Chem., 93:

FISH, R.H.,  KIMMEL, E.C.,  & CASIDA,  J.E. (1976)  Bioorganotin chemistry:
Biological oxidation  of organotin  compounds.  In:  Zuckerman,  J.J.,  ed.
 Organotin  compounds:  New  chemistry  and  applications,  Washington,  DC,
American Chemical  Society, pp.  197-203 (Advances  in Chemistry Series No.

FLOCH,  H.,   DESCHIENS,  R.,   &  FLOCH,  T.  (1964)  Sur  les  propriétés
molluscicides de  l'oxyde et de l'acétate de tributylétain (prophylaxie des
bilharzioses).  Bull. Soc. Pathol. Exot., 57: 454-465.

FOSTER,  R.B.   (1981)  Use  of  Asiatic  clam  larvae  in  aquatic  hazard
evaluations. In:  Bates, J.M.  & Weber, C.I., ed.  Ecological assessments of
 effluent  impacts   on  communities   of  indigenous   aquatic   organisms,
Philadelphia, American Society for Testing and Materials, pp. 280-288 (ASTM
STP No. 730).

FUNAHASHI, N., IWASAKI, I., & IDE, G. (1980) Effects of bis(tri-n-butyltin)
oxide on endocrine and lymphoid organs of male rats.  Acta pathol. Jpn., 30:

GARDINER, B.G. & POLLER, R.C. (1964) Insecticidal activity and structure of
some organotin compounds.  Bull. entomol. Res., 55: 17-21.

GIBBS, P.E. & BRYAN, G.W. (1986) Reproductive failure in populations of the
dog-whelk,  Nucella  lapillus, caused by imposex induced by tributyltin from
antifouling paints.  J. Mar. Biol. Assoc. UK, 66: 767-777.

GIBBS, P.E.  & BRYAN,  G.W. (1987)  TBT paints  and the  demise of the dog-
whelk, Nucella  lapillus (Gastropoda).  In:  Proceedings  of  the  Organotin
 Symposium,  Oceans   '87  Conference,  Halifax,  Nova  Scotia,  Canada,  28
 September-1 October,  1987, New  York,  The  Institute  of  Electrical  and
Electronics Engineers, Inc., Vol. 4, pp. 1482-1487.

GIBBS, P.E.,  BRYAN, G.W., PASCOE, P.L., & BURT, G.R. (1987) The use of the
dog-whelk,   Nucella   lapillus,  as   an  indicator  of  tributyltin  (TBT)
contamination.  J. Mar. Biol. Assoc. UK, 67: 507-523.

GIBBS, P.E.,  PASCOE, P.L.,  & BURT,  G.R. (1988)  Sex change in the female
dogwhelk,  Nucella lapillus, induced by tributyltin from antifouling paints.
 J. Mar. Biol. Assoc. UK, 68: 715-731.

C. (1973)  Field tests  of  hexabutyldistannoxane  (TBTO)  in  slow-release
formulations against   Biomphalaria spp. Bull. World Health Organ., 49: 633-

GILE, J.D.,  COLLINS, J.C.,  & GILLETT, J.W. (1982) Fate and impact of wood
preservatives in  a terrestrial  microcosm.  J.  agric. food Chem., 30: 295-

GILLETT, J.W.,  GILE, J.D.,  & RUSSELL,  L.K. (1983)  Predator-prey  (vole-
cricket) interactions: the effects of wood preservatives.  Environ. Toxicol.
 Chem., 2: 83-93.

GOH, C.L.  (1985) Irritant  dermatitis from tri-n-butyl tin oxide in paint.
 Contact dermatitis, 12: 161-163.

GOHLKE, R., LEWA, W., STRACHOVSKY, A., & KOHLER, R. (1969) [Investi-gations
in experimental  animals of  the inhalatory effects of tributyltin chloride
in a  subchronic experiment.]   Z. gesamte  Hyg. Grenzgeb.,  15: 97-104  (in

GOODMAN, L.R.,  CRIPE, G.M.,  MOODY, P.H.,  & HALSELL,  D.G.  (1988)  Acute
toxicity of  malathion, tetrabromobisphenol-A,  and tributyltin chloride to
mysids  (Mysidopsis  bahia) of  three ages.  Bull. environ. Contam. Toxicol.,
41: 746-753.

GROVHOUG, J.G.,  SELIGMAN, P.F., VAFA, G., & FRANSHAM, R.L. (1986) Baseline
measurements of  butyltin in U.S. harbors and estuaries. In:  Proceedings of
 the Organotin  Symposium, Oceans '86 Conference, Washington, DC, USA, 23-25
 September, 1986,  New York,  The Institute  of Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1283-1288.

GUARD,  H.E.,   COGBET,  A.B.,   &  COLEMAN,  W.M.  (1981)  Methylation  of
trimethyltin compounds by estuarine sediments.  Science, 213: 770-771.

HADA, N.  (1986) [Studies  on the  kinetics of tributyltin compounds in the
body: With special reference to their biological half times in goldfish and
rats as  estimated by  newly developed  analytical method.]   Nichidai Igaku
 Zasshi, 45: 1005-1013 (in Japanese).

(1984) Behavioral  responses to two estuarine fish species subjected to bis
(tri- n-butyltin) oxide.  Water Resour. Bull., 20: 235-239.

(1986) Monitoring organotin concentrations in Maryland waters of Chesapeake
Bay. In:   Proceedings of  the Organotin  Symposium, Oceans  '86 Conference,
 Washington, DC,  USA, 23-25  September, 1986,  New York,  The Institute  of
Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1275-1279.

HALL, L.W.,  BUSHONG, S.J., JOHNSON, W.E., & HALL, W.S. (1988a) Spacial and
temporal distribution  of butyltin  compounds in  a Northern Chesapeake Bay
marina and river system.  Environ. Monit. Assess., 10: 229-244.

HALL, L.W.,  BUSHONG, S.J.,  HALL, W.S.,  & JOHNSON, W.E. (1988b) Acute and
chronic effects  of tributyltin  on  a  Chesapeake  Bay  copepod.   Environ.
 Toxicol. Chem., 7: 41-46.

HALLAS, L.E.,  MEANS, J.C.,  & COONEY,  J.J. (1982)  Methylation of  tin by
estuarine microorganisms.  Science, 215: 1505-1507.

HARRIS, J.R.W.  &  CLEARY,  J.J.  (1987)  Particle-water  partitioning  and
organotin dispersal  in  an  estuary.  In:   Proceedings  of  the  Organotin
 Symposium,  Oceans   '87  Conference,  Halifax,  Nova  Scotia,  Canada,  28
 September-1 October,  1987, New  York,  The  Institute  of  Electrical  and
Electronics Engineers, Inc., Vol. 4, pp. 1370-1374.

HENDERSON, R.S.  (1986) Effects of organotin antifouling paint leachates on
Pearl  Harbor   organisms:  a   site  specific  flowthrough  bioassay.  In:
 Proceedings of  the Organotin Symposium, Oceans '86 Conference, Washington,
 DC, USA,  23-25 September,  1986, New York, The Institute of Electrical and
Electronics Engineers, Inc., Vol. 4, pp. 1226-1233.

(1978) The  permanence of  tributyltin oxide  in timber.  In:  Record of the
 1978  Annual   Convention  of  the  British  Wood  Preserving  Association,
 Cambridge, 27-30  June, 1978,  London, British Wood Preserving Association,
pp. 19-29.

HINGA, K.R.,  ADELMAN, D.,  & PILSON,  M.E.Q. (1987) Radiolabeled butyl tin
studies in  the Merl  enclosed ecosystems. In:  Proceedings of the Organotin
 Symposium,  Oceans   '87  Conference,  Halifax,  Nova  Scotia,  Canada,  28
 September-1 October,  1987, New  York,  The  Institute  of  Electrical  and
Electronics Engineers, Inc., Vol. 4, pp. 1416-1420.

HIS, E. & ROBERT, R. (1980)  Action d'un sel organo-métallique, l'acétate de
 tributylétain sur  les oeufs et les larves de Crassostrea gigas (Thunberg),
Copenhagen, International  Council for  the Exploration  of the  Sea (ICES)
Mariculture Commission, 27 pp (CM 1980/F).

HIS, E.  & ROBERT,  R. (1985)  Développement des  véligères de   Crassostrea
 gigas dans  le Bassin d'Arcachon, études sur les mortalités larvaires.  Rev.
 Trav. Inst. Pêches Marit., 47: 63-88.

HIS, E.,  MAURER, D.,  & ROBERT, R. (1986) Observations complémentaires sur
les causes  possibles des anomalies de la reproduction de Crassostrea gigas
(Thunberg) dans  le Basin  d'Arcachon.  Rev.  Trav. Inst. Pêches Marit., 48:

HODGE, V.F.,  SEIDEL, S.L.,  & GOLDBERG,  E.D. (1979)  Determination of tin
(IV) and  organotin compounds  in natural  waters, coastal  sediments,  and
macro algae by atomic absorption spectrometry.  Anal. Chem., 51: 1256-1259.

(1967) Molluscicidal  properties of  organotin  and  organolead  compounds.
 World Health Organ. Bull., 36: 955-961.

HUGGETT,  R.J.,   UNGER,  M.A.,   &  WESTBROOK,   D.J.   (1986)   Organotin
concentrations in  the southern  Chesapeake Bay.  In:   Proceedings  of  the
 Organotin Symposium,  Oceans '86  Conference, Washington,  DC,  USA,  23-25
 September, 1986,  New York,  The Institute  of Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1262-1265.

HUMPEL, M.,  KUHNE, G.,  TAUBER, U.,  & SCHULZE, P.E. (1986) Studies on the
kinetics of  bis (tris- n -butyl-113tin)  oxide (TBTO).  In:   Toxicology  and
 analytics of the tributyltins: The present status. Proceedings of an ORTEPA
 workshop,  Berlin,  15-16  May,  1986,  Vlissingen-Oost,  The  Netherlands,
ORTEP-Association, pp. 122-142.

HUMPHREY, B.  & HOPE,  D. (1987) Analysis of water, sediments and biota for
organotin compounds. In:  Proceedings of the Organotin Symposium, Oceans '87
 Conference, Halifax, Nova Scotia, Canada, 28 September-1 October, 1987, New
York, The  Institute of Electrical and Electronics Engineers, Inc., Vol. 4,
pp. 1348-1351.

ICES (1987)   Concentrations of  organotin and  total tin in open Danish sea
 areas and  in pleasure craft marinas along the sound, International Council
for the Exploration of the Sea (ICES) (TWG 14/9/3-E).

IWAI, H.,  MANABE, M.,  MATSUI, H.,  ONO, T.,  & WADA, O. (1980) Effects of
tributyltin and  its metabolites  on brain  function.  J.  toxicol. Sci., 5:

IWASAKI, I., FUNABASHI, N., TOIZUMI, S., & IDE, G. (1976) Histopathological
studies on  rat's nervous  system by  oral  administration  of  bis-(tri-n-
butyltin) oxide. (I).  J. toxicol. Sci., 1: 99 (abstract).

JENSEN, A.  & CHENG,  Z. (1987)  Total tin  and organotin  in seawater from
pleasure craft  marinas along Danish coast of the Sound. In:  Proceedings of
 the 15th  Conference of  Baltic  Oceanographers  (CBO),  Copenhagen,  1986,
Charlottenlung, Denmark, Marine Pollution Laboratory, Vol. 1, pp. 289-298.

JEWETT,  K.L.  &  BRINCKMAN,  F.E.  (1981)  Speciation  of  trace  di-  and
triorganotins in  water by ion exchange HPLC-GFAA.  J. chromatogr. Sci., 19:

JOHANSEN, K.  & MOHLENBERG,  F. (1987)  Impairment  of  egg  production  in
 Acartia tonsa exposed to tributyltin oxide.  Ophelia, 27: 137-141.

JOHNSON,  T.L.  &  KNOWLES,  C.O.  (1983)  Effects  of  organotins  on  rat
platelets.  Toxicology, 29: 39-48.

JORDAN, P.  (1985)  Schistosomiasis  -  The  St  Lucia  Project,  Cambridge,
Cambridge University Press, 442 pp.

KAKUNO, A.  & KIMURA,  S. (1987) [Acute toxicity of bis (tributyltin) oxide
to girella   (Girella punctata).] Bull. Tokai Reg. Fish. Res. Lab., 123: 41-
44 (in Japanese).

KALBFUS, W.  (1988)  TBT-burden from antifouling paints in different waters.
Presented at  the  OECD  Workshop  on  Monitoring,  Chemical  Analysis  and
Leaching  Rates  of  TBT,  Paris,  30  November-2  December,  1988,  11  pp
(Unpublished report).

KALNINS, M.A.  & DETROY,  B.F. (1984) Effect of wood preservative treatment
of beehives  on honey  bees and  hive products.   J. agric.  food Chem., 32:

KEY, D., NUNNY, R.S., DAVIDSON, P.E., & LEONARD, M.A. (1976)  Abnormal shell
 growth in  the Pacific  oyster Crassostrea  gigas.  Some preliminary results
 from experiments  undertaken in 1975, Copenhagen, International Council for
the Exploration of the Sea (ICES), 12 pp. (C. M. 1976/K/11).

KIMMEL, E.C.,  FISH, R.H.,  & CASIDA,  J.E. (1977)  Bioorganotin chemistry:
Metabolism of organotin compounds in microsomal monooxygenase system and in
mammals.  J. agric. food Chem., 25: 1-8.

KLIMMER, O.R.  (1969) [The  use of organic tin compounds from the viewpoint
of  experimental   toxicology.]   Arzneimittelforschung,   19:  934-939  (in

KNOWLES, C.O.  & JOHNSON,  T.L.  (1986)  Influence  of  organotins  on  rat
platelet aggregation mechanism.  Environ. Res., 39: 172-179.

KRAJNC, E.I.  (1989) Presentation  to OECD  working  group  on  TBT  immune
effects.  In Preparation.

VAESSEN, H.A.M.G.,  & VAN  DER HEIJDEN,  C.A. (1984) Toxicity of bis(tri- n-
butyltin)oxide in  the rat. 1. Short-term effects on general parameters and
on the  endocrine and lymphoid systems.  Toxicol. appl. Pharmacol., 75: 363-

KRAMPITZ, G., ENGELS, J., & CAZAUX, C. (1976) Biochemical studies on water-
soluble proteins  and related  components of gastropods shells. In: Watabe,
N.  &   Wilbur,  K.M.,   ed.   The   mechanisms  of  mineralization  in  the
 invertebrates and plants, Columbia, University of South Carolina Press, pp.

KRAMPITZ, G.,  DROLSHAGEN, H.,  &  DELTREIL,  J.P.  (1983)  Soluble  matrix
components in malformed oyster shells.  Experientia (Basel), 39: 1105-1106.

KROWKE, R.,  BLUTH, U.,  & NEUBERT,  D. (1986)   in vitro  studies  on  the
embryotoxic potential  of (bis[tri- n-butyltin])oxide  in a limb bud culture
system.  Arch. Toxicol., 58: 125-129.

LANGSTON, W.J.,  BURT, G.R.,  & ZHOU, M. (1987) Tin and organotin in water,
sediments, and  benthic organisms of Poole Harbour.  Mar. Pollut. Bull., 18:

LAUGHLIN, R.B.  & FRENCH,  W.J.  (1980)  Comparative  study  of  the  acute
toxicity of  a homologous  series of  trialkyltins to  larval shore  crabs,
 Hemigrapsus nudus,  and lobster,  Homarus americanus. Bull. environ. Contam.
 Toxicol., 25: 802-809.

LAUGHLIN, R. & LINDEN, O. (1982) Sublethal responses of the tadpoles of the
European frog   Rana temporaria to two tributyltin compounds.  Bull. environ.
 Contam. Toxicol., 28: 494-499.

LAUGHLIN, R., FRENCH, W., & GUARD, H.E. (1983) Acute and sublethal toxicity
of  tributyltin  oxide  (TBTO)  and  its  putative  environmental  product,
tributyltin sulfide  (TBTS) to  zoeal mud  crabs,  Rhithropanopeus harrisii.
 Water Air Soil Pollut., 20: 69-79.

LAUGHLIN, R.,  NORDLUND, K.,  & LINDEN,  O.  (1984)  Long-term  effects  of
tributyltin compounds  on the  baltic amphipod,   Gammarus  oceanicus.  Mar.
 environ. Res., 12: 243-271.

(1985) Structure-activity  relationships for  organotin compounds.  Environ.
 Toxicol. Chem., 4: 343-351.

LAUGHLIN, R.B.,  GUARD,  H.E.,  &  COLEMAN,  W.M.  (1986a)  Tributyltin  in
seawater: Speciation and octanol-water partition coefficient.  Environ. Sci.
 Technol., 20: 201-204.

LAUGHLIN, R.B.,  FRENCH, W.,  & GUARD,  H.E. (1986b)  Accumulation  of  bis
(tributyltin) oxide  by the  marine mussel   Mytilus edulis.  Environ.  Sci.
 Technol., 20: 884-890.

LAUGHLIN, R.B., PENDOLEY, P., & GUSTAFSON, R.G. (1987) Sublethal effects of
tributyltin on  the hard shell clam, Mercenaria mercenaria. In:  Proceedings
 of the  Organotin Symposium,  Oceans '87  Conference, Halifax, Nova Scotia,
 Canada, 28 September-1 October, 1987, New York, The Institute of Electrical
and Electronics Engineers, Inc., Vol. 4, pp. 1494-1498.

LAUGHLIN, R.B.,  GUSTAFSON, R., & PENDOLEY, P. (1988) Chronic embryo-larval
toxicity of tributyltin (TBT) to the hard shell clam  Mercenaria mercenaria.
 Mar. Ecol. Prog. Ser., 48: 29-36.

LAWLER, I.F.  &  ALDRICH,  J.C.  (1987)  Sublethal  effects  of  bis(tri- n-
butyltin)oxide on  Crassostrea gigas spat. Mar. Pollut. Bull., 18: 274-278.

LEE, R.F.  (1985) Metabolism  of tributyltin  oxide by  crabs, oysters  and
fish.  Mar. environ. Res., 17: 145-148.

LEE, R.F.  (1986)  Metabolism  of  bis  (tributyltin)  oxide  by  estuarine
animals. In:  Proceedings of the Organotin Symposium, Oceans '86 Conference,
 Washington, DC,  USA, 23-25  September, 1986,  New York,  The Institute  of
Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1182-1188.

LEE, R.F.,  VALKIRS, A.O.,  & SELIGMAN,  P. (1987)  Fate of  tributyltin in
estuarine waters.  In:  Proceedings  of the  Organotin Symposium, Oceans '87
 Conference, Halifax, Nova Scotia, Canada, 28 September-1 October, 1987, New
York, The  Institute of Electrical and Electronics Engineers, Inc., Vol. 4,
pp. 1411-1415.

LEWIS, P.G.  & EMMETT,  E.A. (1987)  Irritant dermatitis  from tributyl tin
oxide and  contact allergy  from chlorocresol.  Contact dermatitis, 17: 129-

LINDEN, O.  (1987) The  scope of  the organotin  issue in  Scandinavia. In:
 Proceedings of  the Organotin  Symposium, Oceans  '87 Conference,  Halifax,
 Nova Scotia,  Canada, 28 September-1 October, 1987, New York, The Institute
of Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1320-1323.

acute toxicity  of 78  chemicals and  pesticide  formulations  against  two
brackish  water   organisms,  the   bleak   (Alburnus   alburnus)  and   the
harpacticoid  Nitocra spinipes. Chemosphere, 8: 843-851.

LYLE, W.H.  (1958) Lesions of the skin in process workers caused by contact
with butyl tin compounds.  Br. J. ind. Med., 15: 193-196.

MCCULLOUGH,  F.S.,   GAYRAL,  P.,  DUNCAN,  J.,  &  CHRISTIE,  J.D.  (1980)
Molluscicides in  schistosomiasis control.   World Health  Organ. Bull., 58:

MACKLAD, F.,  TAMAN, F.,  & EL  SEBAE, A.K.  (1983) Toxicity of tributyltin
fluoride and  copper sulfate  in slow  release formulation  to  Biomphalaria
 alexandrina snails.  Bull. High Inst. Public Health, 14: 1-12.

MAFF/HSE (1988)   Pesticides 1988:  Pesticides approved under the control of
 pesticides regulations  1986,  London, UK Ministry of Agriculture, Fisheries
and Food (MAFF) and UK Health and Safety Executive (HSE), 399 pp (Reference
Book HMSO 500).

MAFF/HSE (1989)  Pesticides 1989:  Pesticides approved under the control of
pesticides regulations  1986, London, UK Ministry of Agriculture, Fisheries
and Food (MAFF) and UK Health and Safety Executive (HSE), 407 pp (Reference
Book HMSO 500).

MAGUIRE, R.J.  (1984) Butyltin  compounds and inorganic tin in sediments in
Ontario.  Environ. Sci. Technol., 18: 291-294.

MAGUIRE, R.J.  (1987) Review:  Environmental aspects  of tributyltin. Appl.
organomet.  Chem., 1: 475-498.

MAGUIRE, R.J.  & HUNEAULT,  H. (1981)  Determination of butyltin species in
water  by   gas  chromatography   with  flame   photometric  detection.   J.
 Chromatogr., 209: 458-462.

MAGUIRE, R.J.  & TKACZ,  R.J. (1983)  Analysis of butyltin compounds by gas
chromatography. Comparison  of  flame  photometric  and  atomic  absorption
spectrophotometric detectors.  J. Chromatogr., 268: 99-101.

MAGUIRE, R.J.  & TKACZ,  R.J.  (1985)  Degradation  of  the  tri-n-butyltin
species in  water and  sediment from  Toronto Harbor.  J. agric. food Chem.,
33: 947-953.

MAGUIRE, R.J.  & TKACZ,  R.J. (1987)  Concentration of  tributyltin in  the
microlayer of natural waters.  Water Pollut. Res. J. Can., 22: 227-233.

MAGUIRE, R.J.,  CHAU, Y.K.,  BENGERT, G.A.,  HALE, E.J.,  WONG,  P.T.S.,  &
KRAMAR, O.  (1982) Occurrence  of organotin  compounds in Ontario lakes and
rivers.  Environ. Sci. Technol., 16: 698-702.

MAGUIRE, R.J.,  CAREY, J.H.,  & HALE, E.J. (1983) Degradation of the tri-n-
butyltin species in water.  J. agric. food Chem., 31: 1060-1065.

MAGUIRE, R.J.,  WONG,  P.T.S.,  &  RHAMEY,  J.S.  (1984)  Accumulation  and
metabolism  of  tri-n-butyltin  cation  by  a  green  alga,   Ankistrodesmus
 falcatus. Can. J. Fish. aquat. Sci., 41: 537-540.

MAGUIRE, R.J.,  TKACZ, R.J.,  & SARTOR,  D.L. (1985)  Butyltin species  and
inorganic tin in water and sediment of the Detroit and St. Clair rivers.  J.
 Great Lakes Res., 11: 320-327.

(1986) Occurrence  of organotin  compounds in water and sediment in Canada.
 Chemosphere, 15: 253-274.

MATSUI, H.,  WADA, O., MANABE, S., ONO, T., IWAI, H., & FUJIKURA, T. (1982)
[Properties  and   mechanism  of   hyperlipidemia  induced  in  rabbits  by
tributyltin fluoride.]  Jpn. J. ind. Health, 24: 163-171 (in Japanese).

MATTHIAS, C.L.,  BELLAMA, J.M.,  & BRINCKMAN, F.E. (1986a) Determination of
ultra-trace concentrations  of butyltin  compounds in water by simultaneous
hydridization/extraction with  GC/FPD detection.  In:   Proceedings  of  the
 Organotin Symposium,  Oceans '86  Conference, Washington,  DC,  USA,  23-25
 September, 1986,  New York,  The Institute  of Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1146-1151.

MATTHIAS, C.L.,  OLSON, G.J.,  BELLAMA, J.M.,  &  BRINCKMAN,  F.E.  (1986b)
Comprehensive  method   for   determination   of   aquatic   butyltin   and
butylmethyltin   species    at   ultratrace   levels   using   simultaneous
hydridization/extraction with GC/FPD detection.  Environ. Sci. Technol., 20:

MATTHIESSEN, P.  (1974)  Some  effects of slow-release bis (tri-n-butyl tin)
 oxide on  the tropical  fish, Tilapia mossambica Peters. Controlled Release
 Pesticide Symposium,  Akron, Ohio, University of Akron, Community Technical
College, Engineering and Science Division, pp. 25.1-25.16.

MATTHIESSEN, P.  & THAIN,  J.E. (in press) A method for studying the impact
of polluted  marine sediments  on colonising  organisms; tests with diesel-
based drilling mud and tributyltin antifouling paint.  Hydrobiologia.

MAURER, D., HERAL, M., HIS, E., & RAZET, D. (1985) Influence d'une peinture
antisalissure à base de sels organométalliques de l'étain sur le captage en
milieu naturel  de l'huitre   Crassostrea gigas.  Rev.  Trav.  Inst.  Pêches
Marit., 47: 241-250.

MEADOR, J.P.  (1986) An  analysis of photobehavior of Daphnia magna exposed
to tributyltin.  In:  Proceedings  of the  Organotin Symposium,  Oceans  '86
 Conference, Washington,  DC, USA,  23-25 September,  1986,  New  York,  The
Institute of  Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1213-

(1978) Determination  of trace  amounts of  butyltin compounds  in  aqueous
systems by  gas chromatography/mass  spectrometry.  Environ.  Sci. Technol.,
12: 288-293.

MICHEL, P. (1987) Automatization of a hydride generation/A.A.S. system - an
improvement for  organotin  analysis.  In:   Proceedings  of  the  Organotin
 Symposium,  Oceans   '87  Conference,  Halifax,  Nova  Scotia,  Canada,  28
 September-1 October,  1987, New  York,  The  Institute  of  Electrical  and
Electronics Engineers, Inc., Vol. 4, pp. 1340-1343.

MIDDLETON, M.C.  & PRATT,  I. (1977)  Skin water  content as a quantitative
index of  the vascular and histologic changes produced in rat skin by di- n-
butyltin and tri- n-butyltin.  J. invest. Dermatol., 68: 379-384.

MIDDLETON,  M.C.   &  PRATT,   I.  (1978)   Changes  in   incorporation  of
[3H]thymidine into  DNA of  rat skin  following  cutaneous  application  of
dibutyltin,   tributyltin    and   1-chloro-2:4-dinitrobenzene    and   the
relationship of these changes to a morphological assessment of the cellular
damage.  J. invest. Dermatol., 71: 305-310.

MINCHIN, D., DUGGAN, C.B., & KING, W. (1987) Possible effects of organotins
on scallop recruitment.  Mar. Pollut. Bull., 18: 604-608.

MOLIN,  L.   &  WAHLBERG,  J.E.  (1975)  Toxic  skin  reactions  caused  by
tributyltin oxide (TBTO) in socks.  Berufs-dermatosen, 4: 138-142.

Y.,  WAKASHIM,   Y.,  WAKASHIN,  M.,  &  OKUDA,  K.  (1984)  Renal  tubular
disturbances induced  by tributyl-tin  oxide in  guinea pigs:  a  secondary
Fanconi syndrome.  Clin. Nephrol., 21: 118-125.

MULLER,  H.A.  (1987)  Determination  of  tributyltin  oxide  in  air.  In:
 Toxicology  and   analytics  of   the  tributyltins:  The  present  status.
 Proceedings of  an ORTEPA  workshop, Berlin,  15-16 May,  1986, Vlissingen-
Oost, The Netherlands, ORTEP Association, pp. 162-172.

MULLER, M.D.  (1984) Tributyltin  detection at  trace levels  in water  and
sediments using GC with flame-photometric detection and GC-MS.  Fresenius Z.
 anal. Chem., 317: 32-36.

MULLER, M.D.  (1987b) Comprehensive  trace level determination of organotin
compounds in environmental samples using high-resolution gas chromatography
with flame photometric detection.  Anal. Chem., 59: 617-623.

MUSHAK, P.,  KRIGMAN, M.R.,  & MAILMAN,  R.B. (1982)  Comparative organotin
toxicity in  the developing  rat: somatic  and  morphological  changes  and
relationship to  accumulation of  total tin.  Neurobehav. Toxicol. Teratol.,
4: 209-215.

NEMEC, M.D.  (1987)  A  teratology study in rabbits with TBTO: Final Report,
Wil Research Laboratories Inc., 210 pp (Report No. WIL-B0002) (Confidential
report to the Tributyl Tin Oxide Consortium).

 Effects on  the growth  and survival  of eggs and embryos of the California
 grunion (Leuresthes  tenuis)  exposed  to trace  levels of  tributyltin, San
Diego, California,  Naval Ocean Systems Center, 15 pp (Technical Report No.

NIVA (1986)   [Organic  tin  compounds  in  fjord  areas],  Oslo,  Norwegian
Institute for Water Research (NIVA) (in Norwegian).

O'CALLAGHAN, J.P.  & MILLER, D.B. (1988) Acute exposure of the neonatal rat
to  tributyltin   results  in   decreases  in   biochemical  indicators  of
synaptogenesis and myelinogenesis.  Pharmacol. exp. Ther., 246: 394-402.

OKOSHI, K.,  MORI, K., & NONURA, T. (1987) Characteristics of shell chamber
formation between  the two  races in  Japanese oyster,   Crassostrea  gigas.
 Aquaculture, 67: 313-320.

OLSON, G.J.  & BRINCKMAN,  F.E. (1986)  Biodegradation  of  tributyltin  by
Chesapeake Bay  microorganisms. In:  Proceedings of the Organotin Symposium,
 Oceans '86  Conference, Washington,  DC, USA,  23-25 September,  1986,  New
York, The  Institute of Electrical and Electronics Engineers, Inc., Vol. 4,
pp. 1196-1201.

OSBORN,  D.   &  LEACH,  D.V.  (1987)   Organotin  in  birds:  Pilot  study,
Huntingdon, Institute of Terrestrial Ecology, 15 pp (Final report to the UK
Department of the Environment. Contract No. F3CR/27/D4/01).

PAGE, D.S.  (1989) An  analytical method for butyltin species in shellfish.
 Mar. Pollut. Bull., 20: 129-133.

PAUL, J.D.  & DAVIES,  I.M. (1986)  Effects of  copper- and tin-based anti-
fouling compounds  on the  growth of  scallops  (Pecten maximus) and oysters
 (Crassostrea gigas). Aquaculture, 54: 191-203.

PAULINI, E.  (1964)  Laboratory  experiments with  some organotin compounds,
Geneva, World  Health Organization,  Parasitic Diseases  Programme, pp. 1-3
(Unpublished report Mol/Inf/16).

PAULINI, E. & DE SOUZA, C.P. (1970)  Influence of different suspended solids
 in the water upon molluscicide activity, Geneva, World Health Organization,
Parasitic Diseases Programme, pp. 1-10 (Unpublished report PD/Mol/70.12).

PELIKAN, Z.  (1969) Effects  of bis(tri- n-butyltin)  oxide on  the eyes  of
rabbits.  Br. J. ind. Med., 26: 165-170.

PELIKAN, Z.  & CERNY,  E.  (1968)  [The  toxic  effects  of  tri- n-butyltin
compounds on white mice.]  Arch. Toxikol., 23: 283-292 (in German).

PELIKAN, Z.  & CERNY,  E. (1969)  Toxic effects  of bis (tributyltin) oxide
(TBTO) on the skin of rats.  Berufs-dermatosen, 17: 305-316.

PICKWELL,  G.V.  &  STEINERT,  S.A.  (1988)  Accumulation  and  effects  of
organotin  compounds   in  oysters  and  mussels:  correlation  with  serum
biochemical and cytological factors and tissue burdens.  Mar. environ. Res.,
24: 215-218.

A.H., & SEINEN, W. (1989) Organotin-induced thymus atrophy concerns the OX-
44+  immature   thymocytes:  Relation  to  the  interaction  between  early
thymocytes and thymic epithelial cells?  Thymus, 14(1-3): 79-88.

(1985) Comparison  of avoidance  responses of  an estuarine  fish,  Fundulus
 heteroclitus, and  crustacean,  Palaemonetes  pugio, to bis (tri-n-butyltin)
oxide.  Water Air Soil Pollut., 25: 33-40.

PINKNEY, A.E.,  MATTESON, L.L.,  & WRIGHT,  D.A.  (1988)  Effects  of  tri-
butyltin on  survival, growth,  morphometry and  RNA-DNA  ratio  of  larval
striped bass, Morone saxatilis. In:  Proceedings of the Organotin Symposium,
 Oceans  '88   Conference,  New   York,  The  Institute  of  Electrical  and
Electronics Engineers, Inc., Vol. 4, pp. 987-991.

PLUM, H.  (1981) Comportement  des composés  organostanniques vis-à-vis  de
l'environnement.  Inf. chim., 220: 135-139.

POITOU, P.,  MARIGNAC, B.,  CERTIN, C.,  & GRADISKI,  D.  (1978)  Etude  de
l'effet sur  le système  nerveux central  et du  pouvoir  sensibilisant  de
l'oxyde de tributylétain.  Ann. Pharm. fr., 36: 569-572.

POLSTER, M.  & HALACKA,  K. (1971)  [9. Contributions  to the  health-toxic
problems   of    some   anti-microbially    used   organo-tin   compounds.]
 Ernährungsforschung, 16: 527-535 (in German).

RACEY, P.A.  & SWIFT,  S.M. (1986)  The residual effects of remedial timber
treatments on bats.  Biol. Conserv., 35: 205-214.

RANDALL, L.,  DONARD, O.F.X., & WEBER, J.H. (1986) Speciation of n-butyltin
compounds by  atomic  absorption  spectrophotometry  using  electro-thermal
quartz furnace after hydride generation.  Anal. Chim. Acta., 184: 197-203.

REIMANN, R.  &  LANG,  R.  (1987)  Mutagenicity  studies  with  tributyltin
compounds. In:   Toxicology and  analytics of  the tributyltins: The present
 status. Proceedings  of  an  ORTEPA  workshop,  Berlin,  15-16  May,  1986,
Vlissingen-Oost, The Netherlands, ORTEP Association, pp. 66-90.

REINHARDT, C.A.,  SCHAWALDER, H.,  & ZBINDEN, G. (1982) Cell detachment and
cloning efficiency as parameters for cytotoxicity.  Toxicology, 25: 47-52.

RICE,  C.D.,  ESPOURTEILLE,  F.A.,  &  HUGGETT,  R.J.  (1987)  Analysis  of
tributyltin  in   estuarine  sediments   and  oyster   tissue,   Crassostrea
 virginica. Appl. organomet. Chem., 1: 541-544.

RITCHIE,  L.S.,   BERRIOS-DURAN,  L.A.,  FRICK,  L.P.,  &  FOX,  I.  (1964)
Molluscicidal time-concentration  relationships  of  organo-tin  compounds.
 World Health Organ. Bull., 31: 147-149.

RITCHIE, L.S.,  LOPEZ, V.A.,  & CORA, J.M. (1974) Prolonged applications of
an organotin  against  Biomphalaria  glabrata and   Schistosoma mansoni.  In:
Molluscicides in schistosomiasis control, New York, London, Academic Press,
pp. 77-88.

RIVM (1989) In: Mathijssen-Spiekman, E.A.M., Canton, J.H., & Roghair, C.J.,
ed.  [Investigation  into the  toxicity of  TBTO for  a number of freshwater
 organisms],  Bilthoven,   National   Institute   of   Public   Health   and
Environmental Hygiene, 48 pp (Report No. 668118.001) (in Dutch).

ROBERTS, M.H.  (1987) Acute toxicity of tributyltin chloride to embryos and
larvae of  two  bivalve  mollusks,   Crassostrea  virginica  and   Mercenaria
 mercenaria. Bull. environ. Contam. Toxicol., 39: 1012-1019.

(1987) Sex  ratio and  gamete production  in American  oysters  exposed  to
tributyltin in  the laboratory. In:  Proceedings of the Organotin Symposium,
 Oceans  '87  Conference,  Halifax,  Nova  Scotia,  Canada,  28  September-1
 October, 1987,  New York,  The  Institute  of  Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1471-1476.

ROBINSON, I.N.  (1969) Effects  of some organotin compounds on tissue amine
levels in rats. Food Cosmet.  Toxicol., 7: 47-52.

ROSENBERG, D.W.  & DRUMMOND,  G.S. (1983)  Direct  in vitro effects  of bis
(tributyl) tin  oxide on hepatic cytochrome P-450.  Biochem. Pharmacol., 32:

ROSENBERG, D.W.,  DRUMMOND,  G.S.,  CORNISH,  H.C.,  &  KAPPAS,  A.  (1980)
Prolonged induction  of hepatic  haem oxygenase and decreases in cytochrome
P-450 content by organotin compounds.  Biochem. J., 190: 465-468.

ROSENBERG, D.W.,  DRUMMOND, G.S.,  & KAPPAS,  A. (1981)  The  influence  of
organometals on  heme  metabolism.   in vivo   and   in vitro  studies  with
organotins.  Mol. Pharmacol., 21: 150-158.

ROSENBERG, D.W.,  ANDERSON, K.E.,  & KAPPAS, A. (1984) The potent induction
of intestinal heme oxygenase by the organotin compound, bis(tri- n-butyltin)
oxide.  Biochem. biophys. Res. Commun., 119: 1022-1027.

SALAZAR, S.M.  (1985)  The  effects of  bis(tri-n-butyltin) oxide  on  three
 species of marine phytoplankton, San Diego, California, Naval Ocean Systems
Center, 16 pp (Technical Report No. 1039).

SALAZAR,  M.H.   &  SALAZAR,   S.M.   (1985)    Ecological   evaluation   of
 organotincontaminated sediment,  San Diego, California, Naval Ocean Systems
Center, 21 pp (Technical Report No. 1050).

SALAZAR, M.H. & SALAZAR, S.M. (1987) Tributyltin effects on juvenile mussel
growth. In:   Proceedings of the Organotin Symposium, Oceans '87 Conference,
 Halifax, Nova  Scotia, Canada,  28 September-1 October, 1987, New York, The
Institute of  Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1504-

SCHULTE, K.J.  (1987) Effects  of TBT on marine organisms: Field assessment
of a  new site-specific  bioassay system.  In:  Proceedings of the Organotin
 Symposium,  Oceans   '87  Conference,  Halifax,  Nova  Scotia,  Canada,  28
 September-1 October,  1987, New  York,  The  Institute  of  Electrical  and
Electronics Engineers, Inc., Vol. 4, pp. 1461-1470.

SAXENA, P.N.  & CROWE,  A.J. (1988)  An investigation  of the  efficacy  of
organotin compounds  for the  control  of  the  cotton  stainer,   Dysdercus
 cingulatus, the  mosquito,  Anopheles  stephensi, and  the common house fly,
 Musca domestica. Appl. organomet. Chem., 2: 185-187.

SCHAFER, E.W.  & BOWLES,  W.A. (1985) Acute oral toxicity and repellency of
933 chemicals  to house and deer mice.  Arch. environ. Contam. Toxicol., 14:

SCHERING (1983)   Repeated dose  inhalation study of ZK21.995 in the rat for
 29-32 days  (21-24 exposures),  Bergkamen,  Federal  Republic  of  Germany,
Schering Inc. (Report No. IC 1/83. Study No. TX81.177).

SCHERING (1986)   Re-evaluation of  a "mouse  micronucleus test  on TBTO",
Bergkamen, Federal  Republic of Germany, Schering Inc. (Confidential report
No. IC  7186 prepared  by the  Institute of  Occupational Health, Helsinki,
Finland, September 1984).

SCHERING (1989a) TBTO -  4 week oral (dietary administration) toxicity study
 in the  rat, Bergkamen,  Federal Republic of Germany, Schering Inc. (Report
No. 280118 by Hazelton, France. Study No. 14/502).

SCHERING (1989b)  TBTO -   Plaque forming  assay following  a  5  week  oral
 toxicity study in the rat, Bergkamen, Federal Republic of Germany, Schering
Inc. (Report No. 283118 by Hazelton, France. Study No. 14/503).

SCHERING (1989c)  TBTO -   Resistance  to  Listeria  monocytogene  infection
 following a  34 day  oral toxicity  study in  the rat,  Bergkamen,  Federal
Republic of  Germany, Schering Inc. (Report No. 282118 by Hazelton, France.
Study No. 14/505).

SCHERING (1989d)  TBTO -   Delayed type hypersensitivity test following a 37
 day oral toxicity study in the rat, Bergkamen, Federal Republic of Germany,
Schering Inc. (Report No. 281118 by Hazelton, France).

SCHERING (1989e)  TBTO -   Systemic toxicity study in beagle dogs with daily
 oral (intragastric)  administration over a total of 18-19 weeks, Bergkamen,
Federal Republic of Germany, Schering Inc. (Report No. IC6/88).

SCHWEINFURTH, H. (1985) Toxicology of tributyltin compounds.  Tin Uses, 143:

SEIFFER, E.A.  & SCHOOF, H.F. (1967) Tests of 15 experimental molluscicides
against  Australorbis glabratus. Public Health Rep., 82: 833-839.

Short term  toxicity of  tri- n-butyltinchloride  in  rainbow  trout   (Salmo
 gairdneri Richardson) yolk sac fry.  Sci. total Environ., 19: 155-166.

SELIGMAN,  P.F.,   VALKIRS,  A.O.,  &  LEE,  R.F.  (1986a)  Degradation  of
tributyltin  in  marine  and  estuarine  waters.  In:   Proceedings  of  the
 Organotin Symposium,  Oceans '86  Conference, Washington,  DC,  USA,  23-25
 September, 1986,  New York,  The Institute  of Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1189-1195.

SELIGMAN, P.F.,  GROVHOUG, J.G.  & RICHTER,  K.E.  (1986b)  Measurement  of
butyltins in  San Diego  Bay, CA: A monitoring strategy. In:  Proceedings of
 the Organotin  Symposium, Oceans '86 Conference, Washington, DC, USA, 23-25
 September, 1986,  New York,  The Institute  of Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1289-1296.

STALLARD, M.O.,  DAVIDSON, B.,  & LEE, R.F. (1989) Distribution and fate of
tributyltin in  the United  States  marine  environment.   Appl.  organomet.
 Chem., (in press).

SHELDON, A.W.  (1975) Effects of organotin anti-fouling coatings on man and
his environment.  J. Paint Technol., 47: 54-58.

SHERMAN, L.R.  & CARLSON,  T.L. (1980) A modified phenylfluorone method for
determining organotin  compounds in  the ppb  and sub-ppb  range.  J.  anal.
 Toxicol., 4: 31-33.

SHIFF, C.J.,  YIANNAKIS, C.,  & EVANS, A.C. (1975) Further trials with TBTO
and other  slow release  molluscicides in  Rhodesia. In:  Proceedings of the
 Controlled Release  Pesticide Symposium,  Wright State  University, Dayton,
 Ohio, 8-10 September, 1975, pp. 177-188.

SHIMIZU, A.  & KIMURA,  S. (1987)  [Effect of  bis (tributyltin)  oxide  on
gonadal development  of a  salt-water  goby,   Chasmichthys  dolichognathus:
Exposure during  maturing period.]   Bull. Tokai  Reg. Fish. Res. Lab., 123:
45-49 (in Japanese).

SHORT, J.W.  & THROWER,  F.P. (1986)  Accumulation of  butyltins in  muscle
tissue of  chinook salmon  reared in  sea pens treated with tri- n-butyltin.
In:   Proceedings   of  the  Organotin  Symposium,  Oceans  '86  Conference,
 Washington, DC,  USA, 23-25  September, 1986,  New York,  The Institute  of
Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1177-1181.

SHORT, J.W.  & THROWER,  F.P. (1987) Toxicity of tri- n-butyl-tin to chinook
salmon,  Oncorhynchus  tshawytscha, adapted  to seawater.   Aquaculture,  61:

SLESINGER, A.E.  & DRESSER,  I. (1978) The environmental chemistry of three
organotin chemicals.  In: Good,  M., ed.   Report of the Organotin Workshop,
 New Orleans, Louisiana, 17-19 February, 1978, pp. 115-162.

SMITH, B.S.  (1981a) Male  characteristics on  female mud  snails caused by
antifouling paints.  J. appl. Toxicol., 1: 22-25.

SMITH, B.S.  (1981b) Tributyltin  compounds induce  male characteristics on
female mud  snails  Nassarius  obsoletus  =  Ilyanassa  obsoleta.  J.  appl.
 Toxicol., 1: 141-144.

SMITH, B.S.  (1981c) Male  characteristics in  female Nassarius  obsoletus:
Variations related to locality, season and year.  Veliger, 23: 212-216.

(1987) The  use of  transplanted juvenile  oysters  to  monitor  the  toxic
effects of  tributyltin  in  California  waters.  In:   Proceedings  of  the
 Organotin Symposium,  Oceans '87  Conference, Halifax, Nova Scotia, Canada,
 28 September-1  October, 1987,  New York,  The Institute  of Electrical and
Electronics Engineers, Inc., Vol. 4, pp. 1511-1516.

Study by  119m  Sn  Mossbauer  spectroscopy  of  bis(tri- n-butyltin)  oxide
adsorbed on cellulosic materials.  Chem. Ind., 23: 874-875.

SNOEIJ, N.J.  (1987)  Triorganotin  compounds in  immunotoxicology and  bio-
 chemistry, Utrecht,  The Netherlands,  University of Utrecht, 170 pp (Ph.D.

SNOEIJ, N.J.,  VAN IERSEL,  A.A.J., PENNINKS,  A.H., &  SEINEN,  W.  (1985)
Toxicity of  triorganotin compounds:  Comparative  in vivo  studies  with  a
series of  trialkyltin compounds  and triphenyltin  chloride in  male rats.
 Toxicol. appl. Pharmacol., 81: 274-286.

SNOEIJ, N.J.,  PUNT, P.M.,  PENNINKS, A.H., & SEINEN, W. (1986a) Effects of
tri- n-butyltin chloride  on energy  metabolism,  macromolecular  synthesis,
precursor uptake  and cyclic  AMP production  in isolated  rat  thymocytes.
 Biochim. Biophys. Acta, 852: 234-243.

SNOEIJ, N.J.,  VAN IERSEL,  A.A.J., PENNINKS,  A.H., &  SEINEN, W.  (1986b)
Triorganotin-induced cytotoxicity  to rat thymus, bone marrow and red blood
cells as determined by several  in vitro assays.  Toxicology, 39: 71-83.

SNOEIJ, N.J., PIETERS, R.H.H., PENNINKS, A.H., & SEINEN, W. (1987) Toxicity
of triorganotin compounds: Orally administered tri- n-butyltin compounds are
rapidly dealkylated in the rat. In:  Triorganotincompounds in immunology and
 biochemistry, Utrecht,  The Netherlands,  University of Utrecht, Chapter 4,
pp. 73-93 (Snoeij, N.J., Ph.D. Thesis).

Differential effects of tri- n-butyltin chloride on macromolecular synthesis
and ATP  levels of  rat thymocyte  subpopulations obtained  by  centrifugal
elutriation.  Int. J. Immunopharmacol., 10: 29-37.

SNOEIJ,  N.J.,   PENNINKS,  A.H.,  &  SEINEN,  W.  (1988b)  Dibutyltin  and
tributyltin compounds  induce thymus  atrophy in  rats due  to a  selective
action on thymic lymphoblasts.  Int. J. Immunopharmacol., 10: 891-899.

SORACCO, R.J.  & POPE, D.H. (1983) Bacteriostatic and bactericidal modes of
action of bis (tributlytin) oxide on  Legionella pneumophila. Appl. environ.
 Microbiol., 45: 48-57.

STALLARD, M.,  HODGE, V., & GOLDBERG, E.D. (1987) TBT in California coastal
waters: Monitoring and assessment.  Environ. Monit. Assess., 9: 195-220.

STANG, P.M.  & SELIGMAN,  P.F. (1986)  Distribution and  fate  of  butyltin
compounds in  the sediment  of  San  Diego  Bay.  In:   Proceedings  of  the
 Organotin Symposium,  Oceans '86  Conference, Washington,  DC,  USA,  23-25
 September, 1986,  New York,  The Institute  of Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1256-1261.

STANG, P.M.  & SELIGMAN,  P.F. (1987)  In situ adsorption and desorption of
butyltin compounds  from Pearl Harbor, Hawaii, sediment. In:  Proceedings of
 the Organotin  Symposium, Oceans  '87  Conference,  Halifax,  Nova  Scotia,
 Canada, 28 September-1 October, 1987, New York, The Institute of Electrical
and Electronics Engineers, Inc., Vol. 4, pp. 1386-1391.

STEIN, T.  & KUSTER,  K. (1982)  Der biologische abbau von tributylzinnoxid
(TBTO) in  einer belebtschlammanlage.   Z. Wasser Abwasser Forsch., 15: 178-

(1986) Growth  abnormalities in  mussels and  oysters from  areas with high
levels of  tributyltin in  San Diego  Bay. In:  Proceedings of the Organotin
 Symposium, Oceans  '86 Conference,  Washington, DC,  USA, 23-25  September,
 1986, New  York, The  Institute of  Electrical and  Electronics  Engineers,
Inc., Vol. 4, pp. 1246-1251.

international intercomparison  of butyltin  determinations in mussel tissue
and sediments.  In:  Proceedings  of the  Organotin  Symposium,  Oceans  '87
 Conference, Halifax, Nova Scotia, Canada, 28 September-1 October, 1987, New
York, The  Institute of Electrical and Electronics Engineers, Inc., Vol. 4,
pp. 1334-1338.

STROMGREN, T.  & BONGARD,  T. (1987)  The effect  of tributyltin  oxide  on
growth of  Mytilus edulis. Mar. Pollut. Bull., 18: 30-31.

TEMMINK, J.H.M.  & EVERTS, J.W. (1987) Comparative toxicity of tributyltin-
oxide (TBTO)  for fish  and snail.  In:  Proceedings  of the  Seventh  World
 Meeting of  the ORTEP-Association,  Amsterdam, 7-8  May, 1987,  Vlissingen-
Oost, The Netherlands, ORTEP-Association, pp. 6-20.

THAIN, J.E.  (1983)  The  acute toxicity  of bis (tributyl tin) oxide to the
 adults and  larvae of  some  marine  organisms,  Copenhagen,  International
Council for  the Exploration  of the  Sea (ICES),  5 pp  (Report No.  C. M.

THAIN, J.E.  (1986) Toxicity  of TBT  to bivalves: Effects on reproduction,
growth and survival. In:  Proceedings of the Organotin Symposium, Oceans '86
 Conference, Washington,  DC, USA,  23-25 September,  1986,  New  York,  The
Institute of  Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1306-

THAIN, J.E.  & WALDOCK,  M.J. (1985)   The growth of bivalve spat exposed to
 organotin leachates  from  antifouling  paints,  Copenhagen,  International
Council for  the Exploration  of the  Sea (ICES),  10 pp  (Report No. C. M.

THAIN, J.E.  & WALDOCK,  M.J. (1986)  The  impact  of  tributyl  tin  (TBT)
antifouling paints  on molluscan  fisheries.  Water  Sci. Technol., 18: 193-

THAIN, J.E.,  WALDOCK, M.J.,  & WAITE, M.E. (1987) Toxicity and degradation
studies  of   tributyltin  (TBT)   and  dibutyltin  (DBT)  in  the  aquatic
environment.  In:   Proceedings  of  the  Organotin  Symposium,  Oceans  '87
 Conference, Halifax, Nova Scotia, Canada, 28 September-1 October, 1987, New
York, The  Institute of Electrical and Electronics Engineers, Inc., Vol. 4,
pp. 1398-1404.

THOMAS, T.E.  & ROBINSON,  M.G. (1986)  The physiological  effects  of  the
leachates from  a self-polishing  organotin  antifouling  paint  on  marine
diatoms.  Mar. environ. Res., 18: 215-229.

THOMAS, T.E.  & ROBINSON,  M.G.  (1987)  Initial  characterization  of  the
mechanisms responsible for the tolerance of Amphora coffeaeformis to copper
and tributyltin.  Bot. mar., 30: 47-53.

CULLEN, W.R.,  & MAGUIRE,  R.J. (1985)   Organotin compounds  in the aquatic
 environment:  Scientific   criteria  for   assessing   their   effects   on
 environmental quality,  Ottawa, National  Research Council  Canada, 284  pp
(NRCC No. 22494).

SILVA NETTO,  J.A., &  GILBERT, B.  (1976) Snail  control in urban sites in
Brazil with slow-release hexabutyldistannoxane and pentachlorophenol.  World
 Health Organ. Bull., 54: 421-425.

GUESNIER, L.R.,  & MORIN, N. (1976) Contribution à l'étude toxicologique et
pharmacologique de  l'oxyde de  tributylétain (OTBE).   Eur. J. Toxicol., 9:

J. (1979)  Influence de  l'oxyde de  tributylétain (OTBE) en aérosol sur le
comportement exploratoire chez la souris.  Toxicol. Eur. Res., 11: 181-186.

Y. (1981)  Dégradation thermique  de l'oxyde  de  tributylétain  (OTBE)  et
toxicité pulmonaire des produits de combustion chez la souris et le cobaye.
 Toxicol. Eur. Res., 111: 35-44.

TSUDA,  T.,   NAKANISHI,  H.,   AOKI,  S.,   &   TAKEBAYASHI,   J.   (1986)
Bioconcentration of  butyltin compounds  by round  crucian  carp.   Toxicol.
 environ. Chem., 12: 137-143.

TSUDA,  T.,   NAKANISHI,  H.,   AOKI,  S.,   &   TAKEBAYASHI,   J.   (1987)
Bioconcentration and  metabolism of  phenyl tin  chlorides in  carp.   Water
 Res., 21: 949-953.

TSUDA,  T.,   NAKANISHI,  H.,   AOKI,  S.,   &   TAKEBAYASHI,   J.   (1988)
Bioconcentration and  metabolism of butyltin compounds in carp.  Water Res.,
22: 647-651.

TWG (1988a)   Uses and  production of  organotin compounds. Presented by the
Federal Republic  of Germany at the Convention for the Prevention of Marine
Pollution from  Land-based Sources  (15th Meeting  of the Technical Working
Group), Brussels, 7-11 March, 1988, 6 pp (Report No. TWG 15/5/1-E).

TWG (1988b)   Tributyl tin compounds in antifouling paints. Presented by the
Federal Republic  of Germany at the Convention for the Prevention of Marine
Pollution from  Land-based Sources  (15th Meeting  of the Technical Working
Group, Brussels, 7-11 March, 1988, 2 pp, (Report No. TWG 15/5/2-E).

TWG (1988c)   Organotins in the Netherlands. Presented by the Netherlands at
the Convention  for the  Prevention of  Marine  Pollution  from  Land-based
Sources (15th Meeting of the Technical Working Group, Brussels, 7-11 March,
1988, 11 pp (Report No. TWG 15/5/4-E).

UHL,  S.   (1986)   [Population   exposure  to  the  environmental  chemical
 pentachloro-ophenol (PCP)  and bis  (tri-n- butyltin)oxide (TBTO)],  Zurich,
Maus Offsetdruck  Konstanz, 145  pp (ETH  Thesis No.  8014,  University  of
Gottingen) (in German).

UNEP (1989)  Mediterranean Action Plan. Assessment of organotin compounds as
 marine pollutants  in the Mediterranean, Athens, United Nations Environment
Programme (MAP Technical Report Series, No. 33).

 antifouling paints  environmental considerations,  London, UK Department of
the Environment, 82 pp (Pollution Paper No. 25).

 environment, London,  UK Department of the Environment, p. 26 (Circular No.

UPATHAM, E.S.,  KOURA, M.,  AHMED, M.D.,  & AWAD,  A.H.  (1980)  Laboratory
trials of  controlled release  molluscicides on   Bulinus (Ph.) abyssinicus,
the intermediate  host of   Schistosoma haematobium  in Somalia.  In: Baker,
R.W., ed.  Controlled release of bioactive materials. Proceedings of the 6th
 International Meeting  of the controlled Release Society, New York, London,
Academic Press, pp. 461-469.

U'REN, S.C.  (1983) Acute  toxicity of  bis(tributyltin) oxide  to a marine
copepod.  Mar. Pollut. Bull., 14: 303-306.

VAFA, G.,  & DOOLEY,  C.A. (1986)  Measurement of butyltin compounds in San
Diego Bay.  Mar. Pollut. Bull., 17: 319-324.

VALKIRS, A.O.,  DAVIDSON, B.M.,  & SELIGMAN,  P.F. (1987)  Sublethal growth
effects and  mortality  to  marine  bivalves  from  long-term  exposure  to
tributyltin.  Chemosphere, 16: 201-220.

VIGHI, M.  & CALAMARI,  D. (1985)  QSARs for organotin compounds on  Daphnia
 magna. Chemosphere, 14: 1925-1932.

VIYANANT, V.,  THIRACHANTRA, S.,  &  SORNMANI,  S.  (1982)  The  effect  of
controlled  release   copper  sulphate  and  tributyltin  fluoride  on  the
mortality and infectivity of  Schistosoma mansoni cercariae. J. Helminthol.,
56: 85-92.

ROZING, J.  (1984) Toxicity  of bis(tri- n-butyltin)oxide  in the  rat.  II.
Suppression of  thymus-dependent immune responses and of parameters of non-
specific resistance  after short-term  exposure.  Toxicol. appl. Pharmacol.,
75: 387-408.

VOS, J.G.,  KRAJNC, E.I., & WESTER, P.W. (1985) Immunotoxicity of bis (tri-
 n-butyltin)   oxide.    In:   Dean,    J.,   ed.     Immunotoxicology    and
 immunopharmacology, New York, Raven Press, pp. 327-340.

WADE,  T.L.,   GARCIA-ROMERO,  B.,   &  BROOKS,   J.M.  (1988)  Tributyltin
contamination in  bivalves  from  U.S.  coastal  estuaries.   Environ.  Sci.
 Technol., 22: 1488-1493.

WALDOCK, M.J.  (1989) Organotin concentrations in the Rivers Bure and Yare,
Norfolk Broads, England  (Unpublished report of the Ministry of Agriculture,
 Fisheries and Food, Fisheries Laboratory, Burnham-on-Crouch, Essex, UK).

WALDOCK, M.J.  & MILLER,  D. (1983)  The determination of total and tributyl
 tin in  seawater and  oysters in  areas of  high pleasure  craft  activity,
Copenhagen, International Council for the Exploration of the Sea (ICES), 18
pp (Report No. C. M. 1983/E.12).

WALDOCK, M.J.  & THAIN,  J.E. (1983) Shell thickening in Crassostrea gigas:
Organotin antifouling or sediment induced?  Mar. Pollut. Bull., 14: 411-415.

WALDOCK, M.J. & THAIN, J.E. (1985)  The comparative leach rates and toxicity
 of two fish net antifouling preparations, Copenhagen, International Council
for the Exploration of the Sea (ICES), 5 pp (Report No. C. M. 1985/E.29).

WALDOCK, M.J.,  THAIN, J.E.  &  MILLER,  D.  (1983)   The  accumulation  and
 depuration of bis (tributyl tin) oxide in oysters: A comparison between the
 Pacific oyster  (Crassostrea gigas)  and the  European flat  oyster (Ostrea
edulis), Copenhagen,  International Council  for the Exploration of the Sea
(ICES), 6 pp (Report No. C. M. 1983/E.52).

WALDOCK, M.J., WAITE, M.E., & THAIN, J.E. (1987a) Changes in concentrations
of organotins  in U.K.  rivers and estuaries following legislation in 1986.
In:  Proceedings of the Organotin Symposium, Oceans '87 Conference, Halifax,
 Nova Scotia,  Canada, 28 September-1 October, 1987, New York, The Institute
of Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1352-1356.

WALDOCK, M.J.,  THAIN, J.E.,  & WAITE,  M.E. (1987b)  The distribution  and
potential  toxic  effects  of  TBT  in  UK  estuaries  during  1986.   Appl.
 organomet. Chem., 1: 287-301.

WALDOCK, M.J.,  WAITE, M.E.,  & THAIN,  J.E. (1988)  Inputs of  TBT to  the
marine environment  from shipping  activity in  the U.K.   Environ. Technol.
 Lett., 9: 999-1010.

WALNE, P.R.  & HELM, M.M. (1979) Introduction of Crassostrea gigas into the
United Kingdom.  In: Mann,  R., ed.   Proceedings of  a Symposium  on Exotic
 Species in  Marinculture: Case histories of the Japanese Oyster Crassostrea
gigas  (Thunberg),  with  implications  for  other  fisheries,  Woods  Hole,
 Massachusetts, USA,  18-20 September,  1978, Cambridge,  London, MIT Press,
pp. 83-105.

WALSH, G.E.  (1986) Organotin  toxicity  studies  conducted  with  selected
marine organisms  at EPA's  environmental research laboratory, Gulf Breeze,
Florida. In:  Proceedings of the Organotin Symposium, Oceans '86 Conference,
 Washington, DC,  USA, 23-25  September, 1986,  New York,  The Institute  of
Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1210-1212.

(1985) Effects  of organotins on growth and survival of two marine diatoms,
 Skeletonema costatum  and  Thalassiosira  pseudonana. Chemosphere,  14: 383-

WALSH, G.E.,  LOUIE, M.K.,  MCLAUGHLIN, L.L., & LORES, E.M. (1986a) Lugworm
 (Arenicola cristata)  larvae in  toxicity tests:  Survival and  development
when exposed to organotins.  Environ. Toxicol. Chem., 5: 749-754.

(1986b)  Inhibition   of  arm   regeneration   by    Ophioderma   brevispina
(Echinodermata, Ophiuroidea)  by tributyltin  oxide and triphenyltin oxide.
 Ecotoxicol. environ. Saf., 12: 95-100.

WALTON, R.,  ADEMA, C.M.,  & SELIGMAN, P.F. (1986) Mathematical modeling of
the transport  and fate  of organotin  in harbors.  In:  Proceedings  of the
 Organotin Symposium,  Oceans '86  Conference, Washington,  DC,  USA,  23-25
 September, 1986,  New York,  The Institute  of Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1297-1301.

(1981) Bioaccumulation  and  chronic  toxicity  of  bis(tributyltin)  oxide
(TBTO): Tests with a saltwater fish. In: Branson, D.R. & Dickson, K.L., ed.
 Proceedings of  the Fourth  Conference on  Aquatic  Toxicology  and  Hazard
 Assessment, Philadelphia,  American Society  for Testing and Materials, pp.
183-200 (ASTM STP No. 737).

WEBBE, G.  (1963) Laboratory  tests of  some new  molluscicides  (organotin
compounds),  Geneva,   World  Health   Organization,   Parasitic   Diseases
Programme, pp. 1-2 (Unpublished report Mol/Inf/14).

WEBER, J.H.,  DONARD, O.F.X., RANDALL, I., & HAN, J.S. (1986) Speciation of
methyl and  butyltin  compounds  in  the  Great  Bay  estuary  (N.H.).  In:
 Proceedings of  the Organotin Symposium, Oceans '86 Conference, Washington,
 DC, USA,  23-25 September,  1986, New York, The Institute of Electrical and
Electronics Engineers, Inc., Vol. 4, pp. 1280-1282.

WEIS,  J.S.,  WEIS,  P.,  &  WANG,  F.  (1987a)  Developmental  effects  of
tributyltin on the fiddler crab,  Uca pugilator, and the killifish,  Fundulus
 heteroclitus. In:   Proceedings  of  the  Organotin  Symposium,  Oceans  '87
 Conference, Halifax, Nova Scotia, Canada, 28 September-1 October, 1987, New
York, The  Institute of Electrical and Electronics Engineers, Inc., Vol. 4,
pp. 1456-1460.

WEIS, J.S.,  GOTTLIEB, J.,  & KWIATKOWSKI,  J. (1987b)  Tributyltin retards
regeneration and  produces deformities  of limbs  in the  fiddler crab,  Uca
 pugilator. Arch. environ. Contam. Toxicol., 16: 321-326.

H.A.M.G. (1988)   The toxicity  of bis(tri-n- butyltin)oxide (TBTO) and di-n-
 butyltindichloride  (DBTC)  in  the  small  fish  species  Orysias  latipes
(medaka)   and   Poecilia  reticulata   (guppy),  Utrecht,  The  Netherlands,
University of Utrecht (Wester, P.W., Ph.D. Thesis).

WESTER, P.W. (in press) Chronic toxicity and carcinogenicity of bis (tri-n-
butyltin) oxide in the rat.  Food Chem. Toxicol.

WHO/FAO (1984)   Data sheet  on pesticides  No. 65:  Bis (tributyltin)oxide,
Geneva, World Health Organization (VBC/PDS/DS/85.65).

WOLNIAKOWSKI, K.U.,  STEPHENSON, M.D.,  & ICHIKAWA, G.S. (1987) Tributyltin
concentrations and  Pacific oyster  deformations in  Coos Bay,  Oregon. In:
 Proceedings of  the Organotin  Symposium, Oceans  '87 Conference,  Halifax,
 Nova Scotia,  Canada, 28 September-1 October, 1987, New York, The Institute
of Electrical and Electronics Engineers, Inc., Vol. 4, pp. 1438-1442.

WONG, P.T.S.,  CHAU, Y.K.,  KRAMAR, O.,  & BENGERT,  G.A. (1982) Structure-
toxicity relationship of tin compounds on algae.  Can. J. Fish. aquat. Sci.,
39: 483-488.

WOOTEN, R.,  DAVIES, I.M.,  MCKIE, J.C.,  & BRUNO, D.W. (1986)  Chemical and
 histopathological changes  in  salmon  exposed  to  low  concentrations  of
 tributyltin in  seawater, Aberdeen, Department of Agriculture and Fisheries
for Scotland, 3 pp (Working Paper No. 15/86).

YLA-MONONEN, L.  (1988)  Finnish  studies on the use, occurrence and effects
 of organic  tin compounds.  Presented at  the OECD  Workshop on Monitoring,
Chemical Analysis and Leaching Rates of TBT, Paris, 30 November-2 December,
1988, 9 pp (Unpublished report).

LANDY, R.B.  (1986)  Summary  report  -  Interagency  workshop  on  aquatic
sampling and  analysis for  organotin compounds.  In:   Proceedings  of  the
 Organotin Symposium,  Oceans '86  Conference, Washington,  DC,  USA,  23-25
 September, 1986,  New York,  The Institute  of Electrical  and  Electronics
Engineers, Inc., Vol. 4, pp. 1135-1140.

ZEDLER, R.J.  (1961) Organotin as industrial biochemicals.  Tin Uses, 
53: 7-11.

ZIMMERLI, B. & ZIMMERMANN, H. (1980) Gas-chromatische bestimmung von spuren
von n-butylzinnverbindungen  (tetra-, tri-, di-).  Fresenius Z. anal. Chem.,
304: 23-27.

Organotins in  biology and  the environment.  In: Brinkman, F.E. & Bellama,
J.M., ed.   Organometals and  organometalloids, occurrence  and fate  in the
 environment, Washington,  DC, American  Chemical Society,  pp. 388-424 (ACS
Symposium Series No. 82).


1.  Identité, propriétés physiques et chimiques

    Les  dérivés du tributylétain  (TBT) sont des  dérivés
organiques  de l'étain tétravalent.  Ils  se caractérisent
par la présence de liaisons covalentes entre des atomes de
carbone et un atome d'étain et leur formule  générale  est
la suivante: (n-C4H9)3       Sn-X (dans laquelle X désigne
un anion). La pureté de l'oxyde de tributylétain  du  com-
merce  est généralement supérieure à  96%; les principales
impuretés  sont le dibutylétain et dans une moindre mesure
le tétrabutylétain et d'autres trialkylétains.  L'oxyde de
tributylétain  est un liquide incolore d'odeur caractéris-
tique  et de densité comprise  entre 1,17 et 1,18.  Il est
peu  soluble dans l'eau  (solubilité comprise entre  moins
de 1,0 et plus de 100 mg/litre selon le pH, la température
et les anions présents dans l'eau qui déterminent l'espèce
chimique  en cause). Dans  l'eau de mer  et dans les  con-
ditions normales, on rencontre trois dérivés du tributylé-
tain: l'hydroxyde, le chlorure et le carbonate,  qui  sont
en  équilibre. Aux pH inférieurs à 7,0 les formes prédomi-
nantes  sont Bu3SnOH2+     et  Bu3SnCl;   à pH 8  ce  sont
Bu3SnCl,    Bu3SnOH   et Bu3SnCO3-;     au-delà  de 10, ce
sont Bu3SnOH et Bu3SnCO3- qui prédominent.

    Le  coefficient  de  partage octanol/eau  (log de   Pow)
est  compris entre 3,18 et 3,84 pour l'eau distillée et il
est égal à 3,54 pour l'eau de mer. L'oxyde  de  tributylé-
tain  s'adsorbe  fortement  aux  matières   particulaires,
puisque  les  coefficients  d'absorption indiqués  dans la
littérature vont de 110 à 55 000. La tension de vapeur est
faible  mais les valeurs publiées  sont extrêmement varia-
bles. On n'a constaté aucune perte d'oxyde  de  tributylé-
tain  à partir d'une  solution de 1 mg/litre  en 62 jours,
toutefois 20% de l'eau avait disparu par évaporation.

2.  Méthodes d'analyse

    On  utilise  plusieurs  méthodes pour  le  dosage  des
dérivés  du tributylétain dans l'eau, les sédiments ou les
biotes.  La plus communément utilisée est la spectrométrie
d'absorption   atomique.   La  spectrométrie  d'absorption
atomique   avec  flamme  a  une  limite  de  détection  de
0,1 mg/litre.   Sans flamme, avec atomisation dans un four
électrique  à  graphite, elle  est  plus sensible  et  ses
limites  de  détection  varient entre  0  et    1,0 µg/litre
d'eau.   Il existe plusieurs  méthodes d'extraction et  de
préparation  de  dérivés  volatils.  La  séparation de ces
dérivés  s'effectue  habituellement  par piégeage  ou  par
chromatographie en phase gazeuse. Les limites de détection
se  situent entre 0,5 et 5,0 µg/kg   dans le cas des sédi-
ments et des biotes.

3.  Sources de pollution de l'environnement

    Les  dérivés  du  tributylétain sont  homologués comme
molluscicides,  comme  produits  antisalissures  pour   la
préservation  des coques de bateaux, des appontements, des
bouées,  des casiers à  crabes, des filets  et des  cages,
comme enduits de protection du bois, comme alguicides dans
le  bâtiment, comme désinfectants  et comme biocides  dans
les  systèmes de réfrigération, les tours de réfrigération
des  centrales électriques, les  usines de pâte  à papier,
les brasseries, les tanneries et les usines textiles.  Les
premières  peintures  antisalissures  à base  de  TBT con-
tenaient  ce produit sous une  forme qui en permettait  la
libération sans entrave. Plus récemment, sont apparues des
peintures  dans lesquelles l'incorporation du TBT dans une
matrice  en copolymère permet d'en  limiter la libération.
On a également mis au point des  matrices  caoutchouteuses
qui permettent une libération lente et durable et assurent
aux  peintures  antisalissures  et aux  molluscicides  une
efficacité  prolongée. Le TBT  n'est pas utilisé  en agri-
culture en raison de sa forte phytotoxicité.

4.  Réglementation

    De  nombreux  pays  ont  restreint  l'utilisation  des
peintures antisalissures à base de TBT du fait de l'action
de  cette substance sur les fruits de mer.  Les détails de
la  réglementation  varient d'un  pays  à l'autre  mais la
plupart  interdisent l'emploi de  peintures à base  de TBT
sur les navires de moins de 25 mètres. Dans certains pays,
les  navires à  coque d'aluminium  ne sont  pas visés  par
cette  interdiction.  En outre,  certaines réglementations
limitent  la teneur des peintures en TBT ou la lixiviation
de  cette substance à partir des peintures qui en contien-
nent (4 à 5 µg/cm2 par jour sur une longue période).

5.  Concentrations dans l'environnement

    On  a  trouvé de  fortes  concentrations de  TBT  dans
l'eau, les sédiments et les biotes à proximité de zones de
plaisance,  plus particulièrement de marinas, de chantiers
navals  et  de bassins  de radoub, de  filets et de  cages
traités  au moyen de  peintures antisalissures et  de sys-
tèmes  de réfrigération. Ces concentrations  de TBT dépen-
dent  également  de la  submersion par la  marée et de  la
turbidité de l'eau.

    On a observé que les concentrations de  TBT  pouvaient
atteindre 1,58 µg/litre   dans l'eau de mer et les estuai-
res,  7,1 µg/litre   dans l'eau douce, 26 300 µg/kg   dans
les  sédiments littoraux, 3700 µg/kg   dans  les sédiments
d'eau douce, 6,39 mg/kg dans les bivalves, 1,92 mg/kg dans
les  gastéropodes et 11 mg/kg dans le poisson.  Il ne faut
pas  considérer  cependant  ces  concentrations  maximales
comme  caractéristiques car un certain  nombre de facteurs
peuvent  donner  lieu  à des  teneurs anormalement élevées

(par  exemple la présence  de particules de  peinture dans
les échantillons d'eau et de sédiments). On a constaté que
les  concentrations de TBT dans la micro-couche de surface
des  eaux douces et  des eaux de  mer étaient jusqu'à  100
fois plus élevées que celles qu'on pouvait  mesurer  juste
en dessous de la surface.  Toutefois, il convient de noter
que  la  concentration  en  TBT  dans  la  micro-couche de
surface  peut dépendre dans  une très large  mesure de  la
technique d'échantillonnage.

    Il  se peut que  les données anciennes  ne soient  pas
comparables  aux  données  récentes en  raison des amélio-
rations  apportées  aux méthodes  de  dosage du  TBT  dans
l'eau, les sédiments et les tissus.

6.  Transport et transformation dans l'environnement

    Du fait de sa faible solubilité dans l'eau et  de  son
caractère  lipophile, le TBT s'adsorbe facilement aux par-
ticules. On estime que 10 à 95% de l'oxyde  de  tributylé-
tain  qui pénètrent dans  l'eau s'adsorbent ainsi  sur les
particules.   La  disparition  progressive du  TBT absorbé
n'est pas due à sa désorption mais à sa  dégradation.   Le
degré  d'adsorption dépend de la salinité, de la nature et
de  la taille des particules en suspension, de la quantité
de  matières en  suspension, de  la température  et de  la
présence de matières organiques dissoutes.

    La  dégradation de l'oxyde de tributylétain s'effectue
par  rupture de la  liaison carbone-étain.  Celle-ci  peut
résulter de divers mécanismes qui se produisent simultané-
ment  dans  l'environnement  et notamment  des  mécanismes
physico-chimiques  (hydrolyse  et  photodécomposition)  ou
biologiques  (dégradation  par  des  micro-organismes   et
métabolisation par des organismes supérieurs). L'hydrolyse
des  dérivés  organostanniques  se produit  à  des valeurs
extrêmes du pH mais n'apparaît guère dans  les  conditions
qui  règnent  normalement dans  l'environnement. La photo-
décomposition  se produit par exposition en laboratoire de
solutions à un rayonnement ultra-violet de 300 nm (et à un
moindre degré, à un rayonnement de 350 nm). Dans le milieu
naturel, la photolyse est limitée par la  longueur  d'onde
du rayonnement solaire et par la pénétration  du  rayonne-
ment  ultra-violet dans l'eau.  La  présence de substances
photosensibilisatrices  peut accélérer la  photodécomposi-
tion.  La biodégradation dépend de l'état du milieu, et de
caractéristiques  telles que sa température,  son oxygéna-
tion,  son pH, sa teneur en éléments minéraux, la présence
de substances organiques facilement biodégradables pouvant
subir  une  co-métabolisation ainsi  que  la nature  de la
micro-flore  et  sa capacité  à  s'adapter. Elle  ne  peut
également avoir lieu que si la concentration en  oxyde  de
tributylétain  est inférieure à la concentration létale ou
inhibitrice  pour les bactéries. Comme  dans le cas de  la
décomposition  abiotique, la dégradation biologique du TBT

comporte  une débutylation oxydante progressive  avec rup-
ture de la liaison carbone-étain.  Il se forme des dérivés
dibutylés dont la dégradation est plus facile que celle du
tributylétain.  Les monobutylétains sont lentement minéra-
lisés.   Il se produit également une dégradation anaérobie
mais  son importance reste discutée.   Certains chercheurs
estiment que la dégradation en anaérobiose est lente alors
que  d'autres  la jugent  plus  rapide que  la dégradation
aérobie. On a identifié des espèces de bactéries, d'algues
et de champignons attaquant le bois qui sont  capables  de
dégrader  l'oxyde de tributylétain.  Les estimations de la
demi-vie   du   TBT  dans   l'environnement  varient  dans
d'importantes proportions.

    Le  TBT s'accumule dans les  organismes du fait de  sa
solubilité  dans les graisses.  Des recherches en  labora-
toire  portant  sur des  mollusques  et des  poissons  ont
donné,  pour les facteurs de bioconcentration, des valeurs
allant  jusqu'à 7000 mais des valeurs  encore plus élevées
ont  été observées lors d'études sur le terrain. L'absorp-
tion  à partir de la  nourriture est plus importante  qu'à
partir de l'eau. Les facteurs de concentration plus élevés
observés  chez les micro-organismes (entre  100 et 30 000)
peuvent  s'expliquer par une adsorption plutôt que par une
absorption  intracellulaire.  Rien  n'indique que  le  TBT
puisse  passer dans les organismes terrestres par l'inter-
médiaire de la chaîne alimentaire.

7.  Cinétique et métabolisme

    Le  tributylétain est absorbé au niveau intestinal (20
à  50% selon le  véhicule) ainsi que  par voie  percutanée
chez  les mammifères (dans  la proportion d'environ  10%).
Il peut traverser la barrière hémo-méningée et  passer  du
placenta  dans le foetus. Une fois absorbé, il est rapide-
ment et largement diffusé dans l'organisme (principalement
au niveau du foie et des reins).

    Chez  les  mammifères,  la métabolisation  du  TBT est
rapide;  on peut déceler les métabolites dans le sang dans
les 3 heures suivant l'administration. Des études in vitro
ont   montré que le TBT servait de substrat aux oxydases à
fonction  mixte  mais qu'il  inhibait  ces enzymes  à très
forte concentration.

    L'élimination  du TBT s'effectue plus ou moins rapide-
ment  selon la nature  du tissu et  les estimations de  la
demi-vie  biologique chez les  mammifères varient de  23 à
environ 30 jours.

    Les  organismes  inférieurs métabolisent  également le
TBT  mais le processus est plus lent - en particulier chez
les  mollusques - que chez les mammifères.  La capacité de
bioaccumulation est donc beaucoup plus importante que chez
les mammifères.

    Les  tributylétains  inhibent la  phosphorylation oxy-
dative  et modifient la structure et la fonction des mito-
chondries.  Le TBT empêche la calcification de la coquille
des huîtres (espèces du genre  Crassostrea ).

8.  Effets sur les micro-organismes

    Le  TBT est toxique pour les micro-organismes et on le
vend  comme bactéricide et alguicide.   Les concentrations
toxiques  varient considérablement selon les  espèces.  Le
TBT  est  plus  toxique pour  les bactéries gram-positives
avec  une concentration minimale inhibitrice  (CMI) allant
de  0,2 à 0,8 mg/litre  que pour les  bactéries gram-néga-
tives  (CMI des 3 mg/litre).  La  CMI de l'acétate de  TBT
pour les champignons est de 0,5 à 1 mg/litre et  celle  de
l'oxyde  de tributylétain est de 0,5 mg/litre pour l'algue
verte   Chlorella   pyrenoidosa. La  productivité  primaire
d'une  communauté  naturelle  d'algues d'eau  douce  a été
réduite  de 5% par une concentration d'oxyde de tributylé-
tain  de 3 µg/litre.    On a récemment établi la dose sans
effet   observable  pour  2 espèces  d'algues;   elle  est
respectivement  de 18 et 32 µg/litre.     La toxicité pour
les  microorganismes  marins  varie  également  selon  les
espèces et selon les études;  il est  difficile  d'établir
la valeur de la dose sans effet observable mais  on  pense
qu'elle  est  inférieure  à 0,1 µg/litre    pour certaines
espèces.   Les concentrations alguicides vont  de moins de
1,5 µg/litre à plus de 1000 µg/litre selon les espèces.

9.  Effets sur les organismes aquatiques

9.1  Effets sur les organismes marins et estuariels

    La  Figure 1  donne  un  diagramme  récapitulatif  des
effets létaux et sublétaux que peuvent produire  les  con-
centrations  de TBT relevées en mer et dans les estuaires.
Des concentrations supérieures à celles qui produisent les
effets létaux aigus ont été observées en différents points
du  globe, notamment là où  se déroulent des activités  de

    Ce  sont les spores  mobiles d'une algue  verte géante
qui  se sont révélés les plus sensibles au TBT (CE50   à 5
jours:  0,001 µg/litre).    On a constaté une réduction de
la  croissance d'un angiosperme marin à des concentrations
de TBT de 1 mg/kg de sédiments, aucun effet n'étant noté à
0,1 mg/kg.


    Le  tributylétain est très toxique pour les mollusques
marins.  On a montré expérimentalement qu'il perturbait la
formation  de la coquille, le développement des gonades et
la  différenciation  sexuelle  des huîtres  adultes,  leur
fixation et leur croissance; en outre on a noté une morta-
lité   des  larves  d'huîtres  et   d'autres  bivalves  et
l'apparition  de  caractères  mâles chez  les gastéropodes
femelles.   La  dose  sans  effet  observable  serait   de
20 ng/litre  pour le naissain de l'espèce d'huître la plus
sensible, l'huître japonaise  (Crassostrea gigas). Chez les
adultes,  il se produit  également une déformation  de  la
coquille qui est liée à la dose. Expérimentalement, on n'a
pas  observé d'effets sur la morphologie coquillière à des
concentrations de TBT de 2 ng/litre.  La dose  sans  effet
observable  correspondant  à  l'apparition  de  caractères
mâles  chez  les  mollusques femelles  du  genre  Thais est
inférieure  à  1,5 ng/litre.   Les formes  larvaires  sont
généralement plus sensibles que les adultes; la différence
est particulièrement marquée dans le cas des huîtres.

    Les  copépodes  sont  plus sensibles  que  les  autres
crustacés aux effets létaux aigus du TBT, avec des valeurs
de  la CL50   pour  des périodes allant  jusqu'à 96 heures
comprises  entre 0,6 et 2,2 µg/litre.     Ces valeurs sont
comparables  à celles qui s'appliquent aux larves les plus
sensibles  des autres groupes de crustacés.  Le TBT réduit

la  capacité de reproduction,  la survie néonatale  et  la
vitesse  de croissance des crustacés.   Dans le cas de  la
crevette  Acanthomysis  sculpta, un  mysidé,  la dose  sans
effet   observable   sur   la   reproduction   serait   de
0,09 µg/litre.  La crevette ne cherche pas à éviter le TBT
au-dessous de 30 µg/litre.

    La  toxicité du tributylétain pour les poissons de mer
est  très variable, les valeurs  de la CL50   à  96 heures
allant  de 1,5 à 36 µg/litre.    Les stades larvaires sont
plus  sensibles que les adultes  (Figure 1). Il semblerait
que les poissons de mer évitent l'oxyde de tributylétain à
partir de 1 µg/litre.

9.2  Effets sur les organismes d'eau douce

    Un  diagramme  récapitulatif  concernant  les   effets
létaux  et  sublétaux  des concentrations  de TBT mesurées
dans l'eau douce est donné à la Figure 2.  On a observé la
présence   de  concentrations  supérieures  à  celles  qui
produisent  des effets sublétaux, notamment dans les zones
de navigation de plaisance.

    Une   concentration   d'oxyde   de  tributylétain   de
0,5 mg/litre  a été mortelle  pour des angiospermes  d'eau
douce et leur croissance était inhibée dès 0,06 mg/litre.

    On ne dispose que de peu de données sur  les  inverté-
brés  d'eau douce, tout au  plus sur trois espèces  autres
que les organismes cibles.  Pour divers sels de tributylé-
tain on a obtenu des valeurs de la CL50   à  48 heures  de
2,3-70 µg/litre    pour la daphnie  et de 5,5-33 µg    par
litre pour les vers de vase. La dose sans effet observable
pour la daphnie est évaluée à 0,5 µg/litre,    le  critère
choisi étant la réapparition d'une réaction normale  à  la
lumière. En ce qui concerne la palourde d'Asie, on indique
une  CL50    à  24 heures de  2100 µg/litre,   les valeurs
corrrespondantes  allant de 30  à 400 µg/litre   pour  les
mollusques  adultes que l'on  cherche à détruire  dans les
opérations de lutte contre la schistosomiase.

    On  a montré que  le tributylétain était  toxique pour
les  larves  de schistosome  à  leur stade  aquatique;  la
CL50   du fluorure de tributylétain est de  16,8 µg    par
litre  pour une exposition d'une  heure.  Une dose de  TBT
comprise en 2 et 6 µg/litre   supprime à hauteur de  99  à
100% l'infectiosité des cercaires pour la souris.

    La  sensibilité  des  mollusques au  TBT  diminue avec
l'âge mais les oeufs sont plus résistants que  les  jeunes
ou les adultes.  La ponte est notablement affectée  à  une
concentration en oxyde de tributylétainde 0,001 µg/litre.

    La toxicité aiguë du TBT pour les poissons d'eau douce
s'est située, pour des périodes allant jusqu'à 168 heures,

dans les limites de 13 à 240 µg/litre,  valeurs correspon-
dant à la CL50.   Dans le cas du guppy, la dose sans effet
histopathologique observable a été estimée à 0,01 µg/litre.

    Après  exposition de grenouilles  Rana temporaria à des
concentrations  inférieures ou égales 3 µg/litre,   on n'a
observé aucun effet sur la survie des oeufs ni des larves;
en revanche, à la concentration de 30 µg/litre   on a noté
une mortalité sensible.


9.3  Etudes de microcosme

    En  vue de la modélisation des écosystèmes marins on a
effectué  des études de microcosme consistant à introduire
dans ces milieux certains organismes et en se plaçant dans
des  conditions où un  apport d'eau de  mer permettait  la
colonisation du milieu par d'autres biotes.  On a constaté
qu'à des concentrations d'oxyde de tributylétain comprises
entre 0,06 et 3 µg/litre,   il y avait réduction du nombre
d'individus et une moindre diversité des espèces.

    Les  résultats obtenus par  modélisation d'écosystèmes
d'eau  douce montrent  que les  doses qui  tuent les  mol-
lusques  sont  également  nocives pour  d'autres  espèces,
notamment les poissons.

10.  Effets sur les organismes terrestres

    L'exposition   des   organismes   terrestres  au   TBT
découlent  essentiellement  de l'utilisation  de ces subs-
tances pour la protection du bois. L'oxyde  de  tributylé-
tain  est toxique pour  la population des  ruches dont  le

bois  de construction  a été  traité au  TBT.  Une  étude,
d'ailleurs  unique, a montré que le TBT était toxique pour
les  chauves-souris mais le résultat ne peut pas être con-
sidéré  comme statistiquement significatif en raison de la
forte mortalité des témoins.  Les dérivés du tributylétain
sont  toxiques pour les insectes exposés, soit topiquement
soit  par suite de  xylophagie.  Pour les  souris de  type
sauvage, la toxicité aiguë du TBT est moyenne; les valeurs
de  la CL50   par  voie alimentaire, calculées  d'après la
consommation  de  semences  traitées  et  soumises  à  des
épreuves de répulsivité, vont de 37 à 240 mg/kg par jour.

11.  Effets sur l'aquaculture

    Les  observations effectuées dans des zones d'aquacul-
ture ont permis d'attribuer à la présence de  fortes  con-
centrations  de tributylétain, un certain  nombre d'effets
nocifs  observés  sur  des bivalves  tels  que: mortalité,
incapacité  à se fixer, moindre croissance, épaississement
de  la coquille et autres malformations notamment chez les
huîtres,  apparition de caractères mâles chez les gastéro-
podes  (avec diminution simultanée des  populations) ainsi
que  chez les mollusques  du genre  Thais. C'est en  France
que pour la première fois on a attribué à la  présence  de
TBT dans l'eau, l'anéantissement total de parcs à huîtres,
observations  qui  ont  été faites  ensuite  dans d'autres
pays. Les effets étaient particulièrement marqués dans les
secteurs  proches  de  marinas destinées  aux  navires  de
plaisance.   La  réglementation  de l'usage  des peintures
antisalissures  à base TBT sur les petits navires a permis
aux  huîtres de retrouver leur capacité de reproduction et
de  croissance.   Toutefois  la concentration  du TBT dans
l'eau reste suffisamment forte dans certains secteurs pour
nuire  aux gastéropodes marins.  On  utilise la croissance
de la coquille et son gaufrage chez les huîtres japonaises
ainsi  que  l'apparition  de  caractères  mâles  chez  les
mollusques  du  genre  Thais comme indicateurs  biologiques
d'une contamination par du tributylétain.

    Peu  d'études  ont été  consacrées  aux effets  du TBT
sédimentaire  sur la faune marine mais on pense qu'il peut
être  absorbé par les  organismes fouisseurs et  provoquer
une certaine mortalité.

    Des  effets toxiques macroscopiques et des altérations
histopathologiques  ont été observés dans  des élevages de
poissons de mer, les filets délimitant les  bassins  ayant
été traités au moyen de peintures aintisalissures  à  base
de TBT.

    On  a  proposé  d'utiliser du  TBT comme molluscicides
pour  détruire les mollusques d'eau douce qui transmettent
la  bilharziose (bilharzies).  Un certain  nombre d'essais
ont été effectués sur le terrain, dont il ressort  que  le
TBT  est  difficile à  utiliser  sans préjudice  pour  les
organismes non visés.

12.  Toxicité pour les mammifères de laboratoire

12.1  Toxicité aiguë

    Le  tributylétain est moyennement à  fortement toxique
pour les mammifères de laboratoire, avec des valeurs de la
DL50    par voie orale allant  de 94 à 234 mg/kg  de poids
corporel  chez  le  rat et  de  44  à 230 mg/kg  de  poids
corporel  chez la souris.  Pour le cobaye et le lapin, les
valeurs  sont du même  ordre de grandeur.   Les variations
enregistrées  sont dues aux différents anions entrant dans
la  composition  du  sel de  tributylétain.   Ces composés
entraînent   une  mortalité  plus  forte  lorsqu'ils  sont
administrés  par  voie  parentérale plutôt  que  par  voie
orale,  probablement du fait qu'ils ne sont que partielle-
ment absorbés au niveau intestinal.

    Parmi les autres effets toxiques aigus on  peut  citer
des anomalies concernant les taux de lipide  sanguins,  le
système  endocrinien,  le  foie,  la  rate  et  un déficit
passager  dans le développement cérébral.  La portée toxi-
cologique  réelle de ces  effets, qui n'ont  été  observés
qu'après  administration de doses  uniques élevées de  ces
composés,  reste discutable et  la cause effective  de  la
mort n'est pas véritablement connue.

    Par  voie percutanée, la toxicité aiguë est faible, la
DL50 étant  supérieure à 9000 mg/kg de poids corporel chez
le  lapin.  Chez le rat, après inhalation purement nasale,
la DL50 à  4 heures se situait à 77 mg/m3   (65 mg/m3   si
l'on  ne tient compte que des particules respirables). Des
mélanges d'air et de vapeurs de tributylétain  ne  produi-
sent pas d'effets toxiques observables, même à saturation.

Toutefois le TBT est très dangereux sous  forme  d'aérosol
lorsqu'il est inhalé et il produit alors une irritation et
un oedème des poumons.

    Le TBT est très irritant pour la peau  et  extrêmement
irritant  pour l'oeil.  L'oxyde  de tributylétain n'a  pas
d'effet sensibilisateur cutané.

12.2  Toxicité à court terme

    Les  composés  du  tributylétain ont  été très étudiés
chez  le rat (toutes  les données présentées  dans ce  qui
suit concernent cet animal, sauf indication contraire).

    On a observé un fort taux de mortalité après une durée
d'exposition de plus de quatre semaines à des  doses  dans
l'alimentation  de  320 mg/kg  (environ 25 mg/kg  de poids
corporel).   Acune mortalité n'a été observée à la dose de
100 mg/kg  de nourriture (10 mg/kg  de poids corporel)  ni
après l'administration par gavage d'une dose correspondant
à  12 mg de TBT/kg de  poids corporel.  Administrée à  des
ratons  peu après leur naissance,  une dose de 3 mg/kg  de

poids  corporel a augmenté  la mortalité.  Les  principaux
symptômes observés après administration de doses mortelles
consistaient en perte d'appétit, faiblesse et émaciation.

    On  a observé des  effets marginaux sur  la croissance
aux  doses  de 50 mg/kg  de  nourriture (6 mg/kg  de poids
corporel)  et de 6 mg/kg  de poids corporel  (administrées
par  gavage).  Les souris sont moins sensibles, ces effets
n'étant observés qu'à partir de 150 à 200 mg/kg  de  nour-
riture (22 à 29 mg/kg de poids corporel).

    Des  effets  structuraux  ont  été  observés  sur  les
organes  endocrines,  essentiellement  l'hypophyse  et  la
thyroïde, lors d'études à court et à long terme.  Lors des
études à court terme, on a observé des anomalies  dans  la
concentration  des hormones circulantes ainsi  que dans la
réponse  aux  stimuli  physiologiques (trophines  hypophy-
saires);  toutefois après une exposition  de longue durée,
la  plupart de ces anomalies avaient disparu. Le mécanisme
sous-jacent n'est pas connu.

    L'exposition  à un aérosol d'oxyde  de tributylétain à
la  dose de 2,8 mg/m3    a déterminé une  forte mortalité,
une  détresse respiratoire, une réaction  inflammatoire au
niveau  des  voies  respiratoires et  des anomalies histo-
pathologiques  des  organes  lymphatiques.   En  revanche,
l'exposition  à  des vapeurs  à  la concentration  la plus
forte  possible (0,16 mg/m3)   à la  température ambiante,
n'a produit aucun effet.

    On  a signalé des effets toxiques au niveau du foie et
des canaux biliaires chez ces trois espèces de mammifères.
Ainsi une nécrose des cellules hépatiques et  des  altéra-
tions  inflammatoires du canal  cholédoque ont été  obser-
vées  chez  des  rats  qui  avaient  reçu  de  l'oxyde  de
tributylétain  à raison de  320 mg/kg de nourriture  (soit
approximativement  25 mg/kg  de  poids  corporel)  pendant
quatre semaines et des souris qui en avaient reçu 80 mg/kg
de  nourriture  (soit approximativement  12 mg/kg de poids
corporel) pendant trois mois. Chez des chiens soumis à une
dose  de 10 mg/kg de  poids corporel pendant  huit à  neuf
semaines,  on a observé une  vacuolisation des hépatocytes
de la région périportale. Ces altérations s'accompagnaient
occasionnellement  d'un accroissement du poids  du foie et
d'une  augmentation  de  l'activité  sérique  des  enzymes

    La  réduction de la concentration  d'hémoglobine et de
l'hématocrite  chez le rat, à la suite de l'administration
d'une  dose correspondant à  80 mg/kg de nourriture  (soit
8 mg/kg  de poids corporel), montre qu'il y a un effet sur
la  synthèse  de l'hémoglobine  qui  conduit à  une anémie
hypochrome  microcytaire. La réduction des  taux d'hémosi-
dérine splénique donne à penser qu'il y a action au niveau
des  réserves  martiales.   Une  anémie  a  également  été
observée chez la souris.

    La formation de rosettes érythrocytaires au niveau des
ganglions  lymphatiques mésentériques a été  observée lors
de certaines études à court terme mais pas lors d'études à
long  terme.   La  portée biologique  de cette observation
(vraisemblablement passagère) reste obscure.

    L'oxyde  de  tributylétain  exerce  un  effet  toxique
caractéristique sur le système immunitaire; du fait de son
action  sur le thymus,  il perturbe les  fonctions immuni-
taires  à médiation cellulaire.  Le  mode d'action demeure
inconnu mais il pourrait y avoir conversion métabolique en
dibutylétain.   La résistance non spécifique est également

    Un certain nombre d'études sur rats et chiens  mais  à
l'exclusion des souris, ont été effectuées avec de l'oxyde
de   tributylétain  et  ont  révélé  l'existence  d'effets
généraux  sur le système immunitaire (par exemple poids et
morphologie  des tissus lymphoïdes, numération des lympho-
cytes  périphériques, concentration totale  des immunoglo-
bulines  sériques),  à  des doses  largement toxiques (des
effets  ont  été observés  chez la souris  à des doses  de
150 mg/kg  de  chlorure  de tributylétain).   Seul  le rat
manifeste  des effets généraux sur  le système immunitaire
sans  autres signes patents  de toxicité et  il se  révèle
indiscutablement  être  l'espèce  la plus  sensible.   Des
études à court terme sur le rat ont permis  d'établir  que
la  dose sans effet observable  était de 5 mg/kg de  nour-
riture (soit 0,6 mg/kg de poids corporel). Lors des études
portant   sur le chlorure  de tributylétain, on  a observé
des  effets  analogues au  niveau  du thymus.  Ces  effets
disparaissaient rapidement lorsqu'on cessait d'administrer
la substance.  On a montré que l'oxyde  de  tributylétain,
étudié  dans le  cadre de  travaux sur  la  résistance  de
l'hôte,  perturbait les fonctions immunitaires spécifiques
du  rat.  L'organisme de cet animal présentait une moindre
aptitude   à  éliminer  les  Listeria   monocytogenes après
exposition à une dose de 50 mg/kg de nourriture (dose sans
effet  observable: 5 mg/kg et  par jour), et  une  moindre
résistance à  Trichinella spiralis a été observée aux doses
respectives  de  50 et  5 mg/kg  de nourriture,  cet effet
disparaissant à la dose de 0,5 mg/kg de nourriture (ce qui
correspond  à des doses  quotidiennes respectives de  2,5,
0,25  et  0,025 mg/kg  de  poids  corporel).   Des  effets
analogues  ont été observés chez des animaux âgés mais ils
étaient moins marqués.

    Dans  l'état actuel des connaissances,  les effets sur
la résistance de l'hôte sont probablement très utiles pour
évaluer  les risques pour la  santé humaine, mais l'on  ne
possède  pas  une  expérience suffisante  de  ces systèmes
d'épreuve  pour  tirer  des  conclusions  définitives  des
résultats  obtenus.   Il  reste que  l'observation de rats
glabres athymiques soumis à une inoculation de  T. spiralis
virulentes    a  permis  l'interprétation   des  résultats
obtenus  sur  le  modèle  T. spiralis. En effet,  l'absence

totale d'immunité thymo-dépendante a multiplié par 10 à 20
le nombre de larves présentes dans les muscles; par contre
l'exposition à des concentrations d'oxyde de tributylétain
respectivement  égales  à 5  et  50 mg/kg de  nourriture a
multiplié  par 2  et 4  respectivement le  nombre  de  ces

    Il  existe quelques données  concernant les effets  du
tributylétain  sur le système immunitaire en développement
mais sans aucune information sur la résistance de l'hôte.

    Pour  évaluer les dangers potentiels  pour l'homme, il
serait plus prudent de prendre en considération les effets
produits  sur l'espèce la plus sensible.  On a observé des
effets  sur  la  résistance de  l'hôte à  T. spiralis à des
concentrations   dans  l'alimentation  ne   dépassant  pas
5 mg/kg (soit l'équivalent de 0,25 mg/kg de poids corporel
par  jour), la dose  sans effet observable  étant égale  à
0,25 mg/kg  (ce  qui  correspond à  0,025 mg/kg par jour).
Toutefois,  l'interprétation de ces résultats en vue d'une
évaluation  du  risque  pour l'homme  reste  controversée.
Toutes  les études qui  ont été effectuées  font ressortir
que  la dose quotidienne sans effet observable pour ce qui
est  des  effets généraux  ou  spécifiques sur  le système
immunitaire  se  situe  à  5 mg/kg  de  nourriture   (soit
l'équivalent  de  0,5 mg/kg  de poids  corporel  selon les
études à court terme).

12.3  Toxicité à long terme

    Une  étude de longue durée chez le rat a montré que le
tributylétain  a  une  effet marginal  sur  les paramètres
toxicologiques  généraux (sans grande  signification toxi-
cologique)  à  la  dose  de  5 mg/kg  de  nourriture (soit
0,25 mg/kg de poids corporel).

12.4  Génotoxicité

    La  génotoxicité  de  l'oxyde de  tributylétain a fait
l'objet  d'études approfondies.  La plupart  de ces études
ont  donné  des résultats  négatifs  et rien  n'indique de
façon  convaincante  que  ce produit  puisse  présenter le
moindre risque mutagène.

12.5  Toxicité vis-à-vis de la fonction de reproduction

    On a évalué l'embryotoxicité potentielle de l'oxyde de
tributylétain sur trois espèces de mammifères (souris, rat
et lapin), après administration de la substance  par  voie
orale à la mère.  La principale malformation observée chez
les  foetus de rat et de souris était une fissure congéni-
tale  du palais  osseux mais  il est  vrai qu'elle  ne  se
produisait  qu'à des doses manifestement  toxiques pour la
mère.   On  ne  peut  pas  considérer  que  ces  résultats
témoignent d'une activité tératogène aux doses inférieures

à celles qui sont toxiques pour la mère.  La dose minimale
sans  effet observable pour ce qui est de l'embryotoxicité
et de la foetotoxicité chez ces trois espèces se situait à
1,0 mg/kg de poids corporel.

12.6  Cancérogénicité

    Une étude de cancérogénicité a été effectuée  sur  des
rats  et on a observé  à cette occasion des  modifications
néoplasiques au niveau des organes endocrines à la dose de
50 mg/kg  de  poids  corporel.  Les  tumeurs hypophysaires
observées à la dose de 0,5 mg/kg de nourriture ne semblent
pas  avoir d'importance  biologique car  il n'y  a pas  de
relation  précise  dose-réponse.   Ces  types  de  tumeurs
apparaissent  en général chez les témoins à des fréquences
élevées  et  variables  et leur  signification  reste donc
discutable.   Une  autre  étude de  cancérogénicité est en
cours sur des souris.

13.  Effets sur l'homme

    On   a   constaté  que   l'exposition  professionnelle
d'ouvriers  à des dérivés du  tributylétain produisait une
irritation  des  voies  respiratoires  supérieures.    Ces
substances sont dangereuses pour l'homme lorsqu'elles sont
sous  la forme d'aérosols.   L'oxyde de tributylétain  est
irritant pour la peau et les muqueuses oculaires et  il  a
été fait état de dermatites graves à la suite d'un contact
direct avec la peau. Le risque est d'autant plus grave que
la réaction cutanée n'est pas immédiate.


1.  Evaluation des risques pour l'homme

    Les  travailleurs sont principalement exposés  lors de
la  préparation et de la  formulation des dérivés du  tri-
butylétain,   de  l'application  ou  de  l'élimination  de
peintures à base de TBT et de l'utilisation de produits de
protection du bois à base de ces substances.  Quant  à  la
population dans son ensemble, elle peut être  exposée  par
contamination   de  la  nourriture,  en   particulier  des
poissons et des coquillages, et lors de  l'utilisation  de
produits de protection du bois à usage ménager.

    Si  l'on  s'appuie  sur l'expérimentation  animale  et
l'observation  directe sur l'homme,  il est clair  que les
dérivés  du tributylétain sont  irritants pour la  peau et
les  yeux  et que  l'inhalation  d'aérosols conduit  à une
irritation des voies respiratoires.

    La  manipulation  de  bois traités  n'entraîne  aucune
irritation  de l'épiderme une fois que le produit a séché.
Cependant  les aérosols de TBT  sont extrêmement dangereux
et il ne faut pas revenir dans le local où le bois  a  été
traité avant séchage complet.

    On n'a jamais fait état d'intoxications aiguës généra-
lisées  par le  TBT qui  en principe  doit  s'éliminer  de
l'organisme en l'espace de quelques jours. Il est donc peu
probable  que  l'on  puisse s'exposer  à des intoxications
aiguës  par la  manipulation de  produits à  base de  tri-
butylétain si l'on prend les précautions voulues.

    On a fait état d'effets à court et à long  terme  chez
des  animaux de laboratoire, sur  le foie, le sang  et les
glandes endocrines.  Chez le rat, qui est l'espèce la plus
sensible,  c'est l'effet au niveau  du système immunitaire
et  en  particulier  sur  la  résistance  de  l'hôte,  qui
constitue  le paramètre le plus sensible de la toxicité de
ces produits. En utilisant comme modèle la  résistance  de
l'hôte   à   l'infestation  par  Trichinella   spiralis, on
obtient  une dose sans effet observable qui se situe entre
0,5 et 5 mg/kg de nourriture (c'est-à-dire 0,025  et  0,25
mg/kg de poids corporel) alors que si l'on se  rapporte  à
l'action  sur la fonction immunitaire, la dose est égale à
0,6 mg/kg de poids corporel.

    En  raison des grandes variations dans la consommation
de poissons et de coquillages ainsi que dans  les  teneurs
locales  des fruits de mer  en résidus de TBT,  on ne peut
donner  que quelques exemples pour  illustrer l'exposition
résultante et les valeurs des doses sans effet observable.

Il  importe  de souligner  que  pour déterminer  le risque
résultant de la présence de ces composés, il faut procéder
sur  place à  des mesures  de résidus  dans  les  denrées,
évaluer la consommation locale de fruits de mer et définir
une marge de sécurité acceptable.

    En  retenant  pour  la  consommation  de  poisson  les
chiffres  de 15 et 150 grammes  par jour, de 1 mg/kg  pour
les  résidus de TBT dans  le poisson et un  poids corporel
moyen  de  60 kg,  on  obtient  pour  l'homme  les  marges
suivantes  de sécurité selon les divers paramètres immuni-

Consommation   Apport estimatif            Marge de sécurité
 de poisson     journalier de        Modèle        Autres paramètres
  (g/jour)       TBT (µg/kg)       T. Spiralis        immunitaires
     15             0,25            100-1000             2500
    150             2,5              10-100               250

    Utiliser  des dérivés du TBT à tort et à travers et de
manière  irresponsable sans suivre les recommandations qui
figurent dans la présente monographie pour réduire l'expo-
sition humaine peut conduire à l'ingestion de quantités de
TBT dangereuses pour la santé humaine.

    On  a noté jusqu'ici d'effets tératogènes, chez l'ani-
mal  de laboratoire, qu'à des doses manifestement toxiques
pour  la  mère.  On  peut  donc considérer  que l'activité
tératogène du TBT est très faible.

    En   s'appuyant   sur  les   résultats  d'études  très
complètes  de mutagénicité, on  estime que les  dérivés du
tributylétain  n'ont aucun pouvoir mutagène.  Une étude de
cancérogénicité  chez  le  rat,  portant  sur  l'oxyde  de
tributylétain,  a fait apparaître une  incidence accrue de
tumeurs  endocriniennes;  mais  il s'agissait  de  tumeurs
spontanées dont l'incidence, généralement élevée, est très
variable.   Ces  résultats  ne  constituent  donc  pas  un
argument bien net en faveur d'un risque  cancérogène  pour

2.  Evaluation du risque écologique

    La  pénétration  diffuse  du tributylétain  (TBT) dans
l'environnement  a  principalement  pour origine  l'utili-
sation  de  peintures  antisalissures qui  en contiennent.
Une  contamination  ponctuelle peut  se produire lorsqu'on
utilise du TBT comme biocide dans les systèmes  de  réfri-
gération,  lors du  traitement de  la pulpe  de  bois,  le
tannage  du cuir, la préservation du bois et le traitement
des textiles.

    Du  fait  de leurs  priopriétés physico-chimiques, les
dérivés  du  TBT se  concentrent  dans la  micro-couche de
surface  ainsi que dans les  sédiments.  Il ne semble  pas
que le principal mécanisme d'élimination de ces substances
dans  les conditions naturelles soit une dégradation abio-
tique.   L'oxyde  de tributylétain  est biodégradable dans
l'eau mais le processus n'est pas suffisamment rapide pour
empêcher  la  présence  dans certaines  zones  de  concen-
trations  élevées en TBT.  Une  bioaccumulation se produit
dans  la plupart des  organismes aquatiques mais  chez les
mammifères  de laboratoire, la dégradation métabolique est
plus efficace.

    Le  TBT  est  extrêment dangereux  pour certains orga-
nismes aquatiques, du fait de sa toxicité à  très  faibles
concentrations dans l'eau. On le rencontre à  ces  concen-
trations dans certaines zones.  On a signalé  la  présence
d'effets nocifs sur des invertébrés non visés,  en  parti-
culier  des  mollusques  et ces  effets  sont suffisamment
graves pour bloquer la reproduction et entraîner un déclin
des  populations de mollusques.  Ces  effets néfastes pour
la conchyliculture ont pu être combattus avec succès grâce
à  des restrictions imposées à l'utilisation des peintures
antisalissures  dans  certains secteurs,  restrictions qui
ont  également permis d'éviter l'apparition  de caractères
mâles  chez les gastéropodes femelles.  En ce qui concerne
le pisciculture, il convient de ne pas utiliser  de  pein-
tures  à  base  de TBT  sur  les  filets qui  limitent les

    D'une manière générale, le risque pour l'environnement
terrestre  est  vraisemblablement  faible.   Cependant  le
traitement  du bois par ces substances pourrait se révéler
dangereux pour les organismes qui vivent à son contact.

    L'augmentation  de la teneur en TBT de la micro-couche
superficielle pourrait se révéler dangereuse pour la faune
côtière,  pour  les  neustons (y  compris  les invertébrés
benthiques et les larves de poissons) ainsi que  pour  les
oiseaux  de mer et le  gibier d'eau qui se  nourrissent en
surface.  L'accumulation  et  le faible  taux de biodégra-
dation  du TBT dans les sédiments peut présenter un risque
pour  les  organismes  aquatiques  lorsque  des  sédiments
pollués  sont soulevés par  des processus naturels  ou des
activités de dragage.


1.  Recommandations pour la protection de la santé humaine
et de l'environnement

a)  Les  pays membres qui  jusqu'ici n'ont pas  réglementé
    l'utilisation   des  dérivés  du  TBT  devraient  être
    invités à le faire.

b)  Il est nécessaire d'évaluer la pénétration des dérivés
    organo-stanniques  dans  l'environnement  à partir  de
    sources autres que les peintures antisalissures et, si
    besoin  est, d'édicter une réglementation  à ce sujet.
    Il  faut en particulier évaluer le risque résultant du
    déversement  sur le sol  de boues d'égouts  contaminés
    par du tributylétain.

c)  Il faudrait améliorer, du point de vue de la sécurité,
    les  techniques d'application, d'élimination et d'éva-
    cuation des peintures à base d'organo-stanniques.

2.  Recherches à effectuer

a)  Il  faut  améliorer les  méthodes  de recherche  et de
    dosage  du  butylétain  afin  d'avoir  une  estimation
    rapide  et  exacte  des concentrations  de  l'ordre de
    pg/litre.  Une  des  raisons de  cette  recommandation
    tient  à un effet biologique, à savoir l'apparition de
    caractères  mâles chez les gastéropodes  femelles, qui
    est  susceptible de se  produire à des  concentrations
    plus basses que les limites actuelles de détection.

b)  Il faut effectuer des recherches sur les mécanimes par
    lesquels le TBT se concentre au lieu de  se  disperser
    et  qui  retardent sa  décomposition;  à cet  effet on
    étudiera  avec  une  attention particulière  la chimie
    fondamentale du tributylétain et ses interactions avec
    les  molécules biologiques.  Il faut étudier davantage
    la  fixation du TBT  à tous les  niveaux de la  chaîne

c)  Il  est  nécessaire  d'entreprendre des  études sur la
    toxicité  du TBT pour  les organismes aquatiques.   On
    étudiera  en  particulier  le métabolisme,  les effets
    endocriniens et immunologiques, selon le cas.

d)  Il  est nécessaire de rechercher d'autres espèces sen-
    sibles  capables de servir  d'indicateurs biologiques,
    notamment parmi les espèces d'eau douce.

e)  Il faut valider les modèles permettant l'évaluation de
    l'immunotoxicité  chez  les  mammifères et  définir de
    façon  plus  précise  les  doses  sans  effet  toxique
    correspondant aux paramètres pertinents.

f)  Etude  de  toxicité  chronique à  entreprendre sur une
    deuxième espèce de mammifères.

g)  Etude   de  tumorigénicité  à  entreprendre   sur  une
    deuxième espèce de mammifères.

h)  Données  sur les résidus de butylétain dans le poisson
    et  les coquillages destinés à la consommation humaine
    en distinguant les différentes espèces.


1.  Propiedades físicas y químicas

    Los  compuestos de tributilestaño (TBE)  son derivados
orgánicos  del estaño tetravalente.  Caracterizados por la
presencia  de enlaces covalentes entre átomos de carbono y
un átomo de estaño, tienen la siguiente  fórmula  general:
(n-C4H9)3       Sn-X, en la que X es un anión. En general,
la  pureza  del  óxido de  tributilestaño (OTBE) comercial
pasa del 96%; las principales impurezas están constituidas
por  derivados del dibutilestaño  y, en menor  grado,  por
compuestos  de  tetrabutilestaño  y otros  compuestos tri-
alquílicos de ese elemento. El OTBE es un líquido incoloro
con  olor característico y una densidad relativa de 1,17 a
1,18.  La solubilidad en el agua es baja,  variando  entre
< 1,0  y > 100 mg/litro según el pH, la temperatura, y los
aniones  presentes en el agua (que determinan la especifi-
cidad).   En el agua del mar y en condiciones normales, el
TBE  aparece en tres formas o especies (hidróxido, cloruro
y  carbonato), que se mantienen en equilibrio.  En valores
de  pH  inferiores  a 7,0, las  formas  predominantes  son
Bu3SnOH2+ y  Bu3SnCl, a pH 8  son Bu3SnCl,   Bu3SnOH
y  Bu3SnCO3-,     mientras  que cuando  el pH  pasa de  10
predominan Bu3SnOH y Bu3SnCO3-.

    El  coeficiente  de  partición octanol/agua  (log   Poa)
varía  entre 3,19 y  3,84 para el agua  destilada y es  de
3,54 para el agua del mar.  El OTBE  adsorbe  intensamente
las partículas, con coeficientes de adsorción comprendidos
entre  110  y 55 000  según  los informes  publicados.  La
presión  de  vapor es  baja,  pero los  valores publicados
acusan  variaciones considerables.  En  una solución de  1
mg/litro no se observó disminución alguna del OTBE durante
62 días, pero el 20% del agua se perdió por evaporación.

2.  Métodos analíticos

    Se  utilizan diversos métodos para medir los derivados
tributilestánnicos  en el agua,  en el sedimento  o en  la
flora  y la fauna  (biota). El más  usado es la  espectro-
metría de absorción atómica (AA).  La espectrometría de AA
con llama permite alcanzar un límite de detección  de  0,1
mg/litro.  Resulta más sensible la AA sin llama, basada en
la  atomización en un  horno eléctrico con  grafito, cuyos
límites  de  detección  están  comprendidos  entre  0,1  y
1,0 µg/litro    de  agua.   Existen diferentes  métodos de
extracción  y  para  formar derivados  volátiles. La sepa-
ración  de  estos derivados  suele  hacerse por  "purga y
captura"  o cromatografía de gases. Los límites de detec-
ción  son de  0,5 y  5,0 µg/kg   para  el sedimento  y  la

3.  Fuentes de contaminación ambiental

    Se  han  registrado diversos  compuestos de tributile-
staño  como  molusquicidas, productos  antiincrustantes en

botes  y  otras  embarcaciones, muelles,  boyas, jaulas de
langostas, redes de pesca y nasas, como  conservadores  de
la  madera, como "productos anti-cieno"  en los trabajos
de albañilería, como desinfectantes y como biocidas en los
sistemas  de refrigeración, las torres de refrigeración de
las  centrales electrógenas, las  fábricas de pulpa  y  de
papel,  las cervecerías, las  industrias del cuero  y  los
telares.   En  las  pinturas antiincrustantes,  el  TBE se
comercializó al principio bajo una forma que  permitía  la
liberación sin trabas del compuesto. En fecha más reciente
se han puesto a la venta pinturas de liberación controlada
en  las que el TBE se incorpora a una matriz copolimérica.
También  se han ideado matrices  de caucho para hacer  más
lenta  y  retrasar la  liberación  del compuesto,  con  la
consiguiente  prolongación de la eficacia  de las pinturas
antiincrustantes  y de los  molusquicidas.  El TBE  no  se
utiliza en la agricultura por su elevada fitotoxicidad.

4.  Reglamentación del empleo

    Muchos  países han restringido  el empleo de  pinturas
antiincrustantes con TBE por los efectos de éste sobre los
mariscos. Los reglamentos varían en detalle de unos países
a  otros, pero casi siempre prohíben el uso de pinturas de
TBE en las embarcaciones de 24 metros de longitud o menos.
Algunos  países  han excluido  de  esta prohibición  a las
embarcaciones  con casco de aluminio.   Además, en algunos
reglamentos  se  restringe  el  contenido  de  TBE  en las
pinturas  o el ritmo con  que se libera este  compuesto de
las mismas (a 4 o 5 µg/cm2 por día, a largo plazo).

5.  Concentraciones en el medio ambiente

    Se  han encontrado concentraciones elevadas  de TBE en
el agua, los sedimentos y la biota próximos a las zonas de
navegación  de  recreo,  especialmente  en  "marinas"  o
fondeaderos de yates, embarcaderos, diques secos, redes de
pesca  y nasas tratadas con pinturas antiincrustantes, así
como en los sistemas de refrigeración.  La  intensidad  de
las  mareas y la turbiedad  del agua influyen en  las con-
centraciones de TBE.

    Se  ha  visto que  las  concentraciones de  TBE pueden
llegar a 1,58 µg/litro   en el agua del mar y en los estu-
arios, a 7,1 µ/litro  en el agua dulce, a 26 300 µg/kg  en
los  sedimentos costeros, a 3700 µg/kg   en los sedimentos
de  agua dulce, a 6,39 mg/kg en los bivalvos, a 1,92 mg/kg
en los gasterópodos y a 11 mg/kg en los peces. Ahora bien,
no hay que considerar como representativas esas concentra-
ciones  máximas  de TBE,  toda  vez que  diversos factores
pueden  dar  lugar  a valores  anormalmente  elevados (por
ejemplo, las partículas de pintura en las muestras de agua
y de sedimentos). Se ha comprobado que en  las  concentra-
ciones  de TBE en la microcapa superficial del agua, tanto
dulce  como salada, pueden ser dos veces más altas que las
medidas  inmediatamente por debajo de  la superficie.  Sin

embargo, conviene tener en cuenta que los  valores  regis-
trados de TBE en las microcapas superficiales pueden verse
muy afectados por el método de muestreo.

    Es  posible que los datos antiguos no sean comparables
con los más recientes a causa del perfeccionamiento de los
métodos  analíticos disponibles para determinar  el TBE en
el agua, los sedimentos y los tejidos.

6.  Transporte y transformación en el medio ambiente

    A  consecuencia de su  baja hidrosolubilidad y  de  su
carácter lipofílico, el TBE se adsorbe fácilmente  en  las
partículas.  Se calcula que entre el 10% y el 95% del OTBE
introducido  en  el  agua  se  adsorbe  de  ese  modo.  La
desaparición  progresiva del TBE adsorbido no se debe a la
desorción  sino a la  degradación.  El grado  de adsorción
depende de la salinidad, de la naturaleza y el  tamaño  de
las partículas en suspensión, de la cantidad  de  material
suspendido, de la temperatura y de la presencia de materia
orgánica disuelta.

    La degradación del OTBE entraña la escisión del enlace
carbono-estaño.   Este  fenómeno puede  deberse a diversos
mecanismos  que  intervienen  simultáneamente en  el medio
ambiente,  unos fisicoquímicos (hidrólisis  y fotodegrada-
ción)  y otros biológicos (degradación por microorganismos
y metabolización por organismos superiores).  Mientras que
en condiciones extremas de pH se produce una hidrólisis de
los compuestos de estaño orgánico, ésta es apenas evidente
en condiciones ambientales normales.  En el laboratorio se
observa fotodegradación cuando se exponen soluciones a una
irradiación  ultravioleta de 300 nm (y, en menor grado, de
350 nm).   En  condiciones  naturales, la  fotólisis  está
limitada por la gama de longitudes de onda de la luz solar
y por la escasa penetración de la luz ultravioleta  en  el
agua.   La  presencia  de sustancias  fotosensibilizadoras
puede  acelerar  la  fotodegradación.   La  biodegradación
depende  de condiciones ambientales tales como la tempera-
tura,  la oxigenación, el pH, el nivel de elementos miner-
ales,  la  presencia  de sustancias  orgánicas  fácilmente
biodegradables  a efectos de cometabolismo y la naturaleza
de  la microflora y  su capacidad de  adaptación.  También
depende de que la concentración de OTBE sea más  baja  que
el umbral letal o inhibitorio para las bacterias.  Como en
la  degradación abiótica, la ruptura biótica del TBE es un
proceso  progresivo de debutilización oxidativa fundado en
la escisión del enlace carbono-estaño, en el que se forman
derivados  dibutílicos que se degradan  más fácilmente que
el  tributilestaño.   Los monobutilestaños  se mineralizan
lentamente.  Aunque existe degradación anaerobia, no se ha
llegado a un acuerdo sobre su importancia. Algunos autores
estiman  que es lenta, mientras  que otros piensan que  es
más rápida que la degradación aerobia. Se han identificado
varias  especies de bacterias, algas y hongos nocivos para

la madera que pueden degradar el OTBE.   Las  estimaciones
de la semivida del TBE en el medio ambiente acusan grandes

    El TBE se bioacumula en los organismos a causa  de  su
solubilidad  en las grasas. En  investigaciones de labora-
torio con moluscos y peces se han visto que  los  factores
de  bioconcentración pueden llegar a  un valor de 7000,  y
aun  se han  obtenido valores  más altos  en los  estudios
sobre  el terreno.  La absorción a partir de los alimentos
es  más importante que  la que se  efectúa directamente  a
partir del agua. La presencia de factores de concentración
más  altos  en los  microorganismos  (entre 100  y 30 000)
puede deberse más a adsorción que a  absorción  intracelu-
lar. No hay ninguna indicación de que el TBE se transfiera
a los microorganismos terrestres a través de  las  cadenas

7.  Cinética y metabolismo

    El  tributilestaño se absorbe  a través del  intestino
(20-50%,  según el vehículo) y de la piel en los mamíferos
(10%  aproximadamente),  pudiendo  atravesar  la   barrera
hematoencefálica  y  pasar  de  la  placenta  al  feto. El
material  absorbido se distribuye rápida y ampliamente por
los tejidos (principalmente el hígado y el riñón).

    En  los mamíferos, el metabolismo del TBE es rápido: a
las 3 h de administrar TBE pueden ya  descubrirse  metabo-
litos  en la sangre.  Los  estudios  in vitro han  revelado
que el TBE es un sustrato para ciertas oxidasas de función
mixta, pero las concentraciones muy altas de  TBE  inhiben
esas enzimas.

    La velocidad de desaparición del TBE difiere  de  unos
tejidos a otros; las estimaciones de la semivida biológica
en los mamíferos van desde 23 hasta unos 30 días.

    La  metabolización del TBE  se observa también  en los
organismos  inferiores, pero es más lenta (particularmente
en  los moluscos) que en  los mamíferos.  La capacidad  de
bioacumulación,  por consiguiente, es  mucho mayor que  en

    Los  compuestos  de  TBE  inhiben  la  fosforilización
oxidativa y alteran la estructura y las funciones  de  las
mitocondrias.  El TBE interfiere en la calcificación de la
concha de las ostras ( Crassostrea spp).

8.  Efectos en los microorganismos

    El  TBE es  tóxico para  los microorganismos  y se  ha
utilizado  comercialmente  como  bactericida y  alguicida.
Las  concentraciones  que provocan  efectos tóxicos varían
mucho de unas especies a otras.  El TBE es más tóxico para
las  bacterias  gram-positivas (concentración  inhibitoria

mínima (CIM): entre 0,2 y 0,8 mg/litro) que para las gram-
negativas  (CIM: 3 mg/litro).  La  CIM del acetato  de TBE
para  los hongos es  de 0,5-1 mg/litro y  la CIM del  OTBE
para   el  alga  verde  Chlorella  pyrenoidosa es   de  0,5
mg/litro.   La  productividad  primaria de  una  comunidad
natural de algas de agua dulce se redujo en un 50% con una
concentración  de OTBE de 3 µg/litro.    En fecha reciente
se  han determinado en  dos especies de  algas niveles  de
efecto no observado (NENO) de 18 y 32 µg/litro.   De igual
modo,  la toxicidad para los microorganismos marinos varía
de unas especies a otras y de unos estudios a  otros;  los
NENO   son  difíciles  de  establecer  pero  no  llegan  a
0,1 µg/litro    en algunas especies.   Las concentraciones
alguicidas varían entre < 1,5 µg/litro   y > 1000 µg   por
litro según las especies.

9.  Efectos en los organismos acuáticos

9.1  Efectos en los organismos de agua salada (mar y estuarios)

    En  la figura 1 se  resumen en un  diagrama las  rela-
ciones entre los efectos letales y subletales y  las  con-
centraciones de TBE medidas en diversos organismos del mar
y de los estuarios. En todo el mundo, especialmente en los
lugares relacionados con actividades náuticas recreativas,
se  han  encontrado  concentraciones superiores  a las que
producen efectos letales.


    El  desarrollo de las esporas móviles de una macroalga
verde se ha revelado como el índice evolutivo más sensible
al TBE (CE50 de  5 días: 0,001 µg/litro).   El crecimiento
de  una angiosperma marina se redujo en concentraciones de
TBE  de un mg/kg de  sedimento, pero no se  observó ningún
efecto a 0,1 mg/kg.

    El   tributilestaño  es  sumamente  tóxico   para  los
moluscos  marinos.  Experimentalmente se ha demostrado que
altera la formación de la concha en las ostras  en  creci-
miento,  así como el desarrollo  gonadal y el sexo  de las
ostras adultas, la formación de colonias, el crecimiento y
la  mortalidad de las ostras  y otros bivalvos en  la fase
larvaria, y que causa "imposex" (desarrollo de caracter-
ísticas  masculinas) en los  gasterópodos hembras.  Se  ha
obtenido  un valor NENO de 20 ng/litro para la freza de la
especie  de ostra más sensible  (Crassostrea gigas). El TBE
provoca una deformación de la concha de las ostras adultas
más  o menos acentuada según la dosis.  No se ha observado
experimentalmente  ningún efecto sobre la morfología de la
concha con concentraciones de TBE de 2 ng/litro.   En  las
hembras de buccino, el valor NENO para el imposex no llega
a 1,5 ng/litro. En general, las formas larvarias  son  más
sensibles   que   los  adultos,   siendo  esta  diferencia
especialmente marcada en el caso de las ostras.

    Los  copépodos son más  sensibles que otros  grupos de
crustáceos al efecto letal agudo del TBE; los  valores  de
CL50    para periodos de exposición  de hasta 96 h van  de
0,6  a 2,2 µg/litro.    Estas cifras son comparables a las
obtenidas en las larvas más sensibles de otros  grupos  de
crustáceos.   El TBE reduce el rendimiento reproductor, la
supervivencia  de los recién nacidos  y la tasa de  creci-
miento  juvenil en los  crustáceos.  En el  camarón mísido
 Acanthomysis   sculpta el valor NENO para  la reproducción
parece  ser de 0,09 µg/litro.    Otro  camarón, el llamado
por los anglosajones "grass shrimp", no evita el  TBE  a
concentraciones que pueden llegar hasta 30 µg/litro.

    La toxicidad del tributilestaño para los peces marinos
es sumamente variable; los valores de la CL50   a 96 h van
desde 1,5 hasta 36 µg/litro.   Las fases larvarias son más
sensibles  que los adultos  (fig. 1). Hay indicios  de que
los  peces marinos evitan  las concentraciones de  OTBE de
1 µg/litro o más.

9.2  Efectos en los organismos de agua dulce

    En  la figura 2 se  resumen en un  diagrama las  rela-
ciones  entre  los  efectos  letales  y  subletales  y las
concentraciones  de  TBE medidas  en  agua dulce.   Se han
observado  concentraciones  mayores  que las  que provocan
efectos  subletales  en diversos  sitios, especialmente en
relación con actividades náuticas recreativas.

    Una  concentración  OTBE  de 0,5 mg/litro  produjo  la
muerte de las angioespermas de agua dulce, mientras que el
crecimiento se inhibió a 0,06 mg/litro o más.


    Los  datos sobre los  invertebrados de agua  dulce son
escasos y no comprenden más que tres especies,  además  de
los organismos tomados como objetivo. Con diferentes sales
de  TBE  se han  obtenido valores de  CL50   en 48 h  para
Daphnia de    2,3-70 µg/litro    y  para  Tubifex de   5,5-
33 µg/litro.     Basándose en la reversión de la respuesta
normal  a la luz, se ha calculado que el NENO para  Daphnia
es  de 0,5 µg/litro.    La CL50   de 24 h para  la  almeja
asiática parece ser, según se ha señalado, de 2100 µg  por
litro, mientras que para los moluscos adultos  contra  los
que  se dirige la lucha  antiesquistosomiásica los valores
correspondientes son de 30-400 µg/litro.

    Se  ha demostrado que el tributilestaño es tóxico para
las  larvas  de esquistosoma  en  las fases  acuáticas; la
CL50    (fluoruro de TBE)  parece ser, según  los cálculos
realizados, de 16,8 µg/litro   en el caso de  una  exposi-
ción de 1 h.  La dosis de TBE que suprime el 99-100% de la
infectividad  de las cercarias para el ratón está compren-
dida entre 2 y 6 µg/litro.

    La sensibilidad de los moluscos al TBE  disminuye  con
la  edad, pero  los huevos  son más  resistentes  que  los
individuos jóvenes y adultos.  La puesta de huevos  se  ve
considerablemente  afectada a una concentración de OTBE de
0,001 µg/litro.

    En  las pruebas de  CL50   con exposiciones  de  hasta
168 h, la toxicidad aguda del TBE para los peces  de  agua
dulce varía entre 13 y 240 µg/litro.    Basándose  en  los
efectos  histopatológicos,  se  ha calculado  que el valor
NENO para el "guppy" es de 0,01 µg/litro.

    No  se  ha observado  ningún  efecto sobre  la  super-
vivencia  tras la exposición de huevos y larvas de la rana
 Rana   temporaria a concentraciones de  TBE de 3 µg    por
litro  o menos; en cambio, la mortalidad fue significativa
a 30 µg/litro.

9.3  Estudios microcósmicos

    Se  han  realizado algunos  "estudios microcósmicos"
utilizando como modelo ecosistemas marinos en los  que  se
habían introducido organismos y en los que la  entrada  de
agua  de mar facilitaba  la colonización por  otros organ-
ismos.   Los  resultados  obtenidos muestran  que tanto el
número  de individuos como  la diversidad de  las especies
disminuyen cuando la concentración de OTBE en el  agua  se
sitúa entre 0,06 y 3 µg/litro.

    Los resultados obtenidos con modelos de ecosistemas de
agua  dulce hacen pensar que las dosis que matan los cara-
coles  afectan también a otras especies, en particular los

10.  Efectos en los organismos terrestres

    La exposición de organismos terrestres al TBE proviene
sobre todo del uso de este compuesto como  conservador  de
la madera.  El OTBE es tóxico para las abejas que viven en
colmenas  de madera tratadas  con TBE.  Este  producto  se
mostró  tóxico para los murciélagos en un estudio aislado,
pero  la elevada mortalidad de los testigos resta signifi-
cación  estadística a este  resultado.  Los compuestos  de
TBE  son tóxicos para los insectos que están en contacto o
se  alimentan con madera tratada.  El TBE tiene una  toxi-
cidad  aguda  moderada para  los  ratones en  libertad  y,
basándose  en  el  consumo  de  semillas  tratadas  en las
pruebas  de  repelentes, se  calcula  que los  valores  de
CL50 en la dieta se sitúan entre 37 y 240 mg/kg al día.

11.  Efectos en organismos estudiados sobre el terreno

    Las  observaciones  sobre  el  terreno  han  permitido
establecer una relación entre las concentraciones elevadas
de  tributilestaño  y  la  mortalidad  de  las  larvas  de
bivalvos,  la  incapacidad  de las  mismas para constituir
colonias,  el retraso del crecimiento, el engrosamiento de
la   concha  y  otras  malformaciones  de  las  ostras  en
desarrollo,   el  imposex  en  los   caracoles  de  suelos
cenagosos  y el imposex (con disminución concurrente de la
población)  en el buccino.  En Francia inicialmente, y más
tarde en otros países, se ha identificado y  atribuido  la

destrucción  completa  de los  criaderos  de ostras  a  la
presencia de TBE en el agua.  Los efectos son más acusados
en  las zonas próximas a los centros de deportes náuticos.
Al  dejar de emplear  pinturas antiincrustantes a  base de
TBE  en  las  pequeñas embarcaciones  se restablecieron la
reproducción y el crecimiento de las ostras.  Sin embargo,
en  algunas zonas las  concentraciones de TBE  en el  agua
siguen  siendo  bastante  elevadas  para  afectar  a   los
gasterópodos marinos.

    Tanto  el desarrollo y formación  de la concha en  las
ostras del Pacífico como el imposex en el buccino  se  han
utilizado como indicadores biológicos de contaminación por

    Aunque se han hecho pocos estudios sobre  los  efectos
en  los organismos del TBE presente en los sedimentos, hay
indicios de que este compuesto puede llegar a los animales
que viven en madrigueras y producir mortalidad  en  condi-
ciones prácticas.

    Se  han  observado  efectos  tóxicos  macroscópicos  y
alteraciones  histopatológicas  en los  criaderos de pesca
marítima expuestos al TBE por el uso de  pinturas  antiin-
crustantes en las redes de contención.

    Se ha propuesto la utilización de TBE  como  molusqui-
cida  contra los caracoles de agua dulce que transmiten la
esquistosomiasis (bilharziasis).  Varios ensayos prácticos
han  demostrado que  es difícil  aplicar el  TBE  sin  que
resulten  perjudicados  otros organismos  distintos de los
tomados como objetivo.

12.  Toxicidad para los mamíferos de laboratorio

12.1  Toxicidad aguda

    El tributilestaño tiene una toxicidad entre moderada y
alta  para los mamíferos de laboratorio; los valores de la
DL50    oral aguda  varían desde  94 a  234 mg/kg de  peso
corporal para la rata y desde 44 hasta 230 mg/kg  de  peso
corporal para el ratón.  La toxicidad aguda para el cobayo
y el conejo es del mismo orden.  La variación se  debe  al
componente  aniónico de la  sal de tributilestaño.   Estos
compuestos  tienen un potencial letal mayor por vía paren-
teral  que por vía oral, probablemente porque la absorción
intestinal es incompleta.

    Entre  otros efectos de la exposición aguda cabe citar
alteraciones  de las cifras de lípidos sanguíneos, el sis-
tema endocrino, el hígado y el bazo, así como  un  déficit
transitorio  del desarrollo cerebral. La importancia toxi-
cológica  de estos efectos, observados tras la administra-
ción  de una fuerte dosis única del compuesto, es discuti-
ble y sigue ignorándose cuál es la causa de la muerte.

    La  toxicidad  aguda  por  vía  dérmica  es  baja;  la
DL50    es > 9000 mg/kg de  peso corporal para  el conejo.
La  DL50   por inhalación exclusivamente nasal (4 h) es de
77 mg/m3   (65 mg/m3   cuando sólo se tienen en cuenta las
partículas  inhalables) para la rata.  Las mezclas de aire
y  vapores de TBE no producen efectos tóxicos observables,
ni siquiera en el punto de saturación. Sin embargo, el TBE
es  muy peligroso cuando  se inhala en  forma de  aerosol,
produciendo irritación y edema de los pulmones.

    El TBE es muy irritante para la piel y sumamente irri-
tante para los ojos.  El OTBE no  produce  sensibilización

12.2  Toxicidad a corto plazo

    Los  compuestos de TBE han  sido muy estudiados en  la
rata, hasta el punto de que todos los datos  expuestos  en
esta  sección se  refieren a  ese animal  a menos  que  se
indique otra cosa.

    Con   dosis  de  320 mg/kg  (unos   25 mg/kg  de  peso
corporal)  en la  dieta se  han obtenido  altas  tasas  de
mortalidad  cuando  la exposición  se  prolonga más  de  4
semanas.   No se observaron  muertes con 100 mg/kg  en  la
dieta (10 mg/kg de peso corporal) ni tras  la  administra-
ción  forzada de 12 mg/kg de peso corporal al día.  En las
ratas tratadas en sus primeros días de vida, la mortalidad
aumentó  con una dosis de  3 mg/kg de peso corporal.   Los
principales síntomas causados por las dosis letales fueron
inapetencia, debilidad y emaciación.

    Se  han observado efectos "limítrofes"  en el creci-
miento de la rata con dosis de 50 mg/kg (6 mg/kg  de  peso
corporal)  en la dieta  y de 6 mg/kg  de peso corporal  en
estudios   de  administración  forzada.   Los  ratones  se
muestran  menos  sensibles,  observándose los  efectos del
compuesto con dosis de 150 a 200 mg/kg (22 a  29 mg/kg  de
peso corporal) en la dieta.

    Tanto  los estudios  a corto  plazo como  en los  pro-
longados  se han observado efectos estructurales sobre los
órganos  endocrinos,  principalmente  la  hipófisis  y  el
tiroides. Las pruebas a corto plazo pusieron de manifiesto
alteraciones  en la concentración  de las hormonas  circu-
lantes  y  en la  respuesta  a los  estímulos fisiológicos
(hormonas  tróficas hipofisarias), pero tras la exposición
prolongada  desaparecieron  casi todas  esas alteraciones.
El mecanismo de acción es desconocido.

    La  exposición  a un  aerosol de OTBE  a razón de  2,8
mg/m3    produjo  un aumento  de  la mortalidad,  así como
dificultades   respiratorias,   inflamación   del   tracto
respiratorio   y  alteraciones  histopatológicas   de  los
órganos  linfáticos.  En cambio, la exposición a la máxima

concentración  de  vapor  accesible  (0,16 mg/m3)    a  la
temperatura del local no produjo ningún efecto.

    En  tres especies de mamíferos se han señalado efectos
tóxicos en el hígado y los conductos biliares.  En ratas a
las  que  se administró  OTBE en la  dieta a razón  de 320
mg/kg (unos 25 mg/kg de peso corporal) durante 4 semanas y
en  ratones tratados con  80 mg/kg (unos 12 mg/kg  de peso
corporal)  en  la dieta  durante  90 días se  produjo  una
necrosis  hepatocelular con alteraciones inflamatorias del
árbol  biliar.  En perros que  recibieron una dosis de  10
mg/kg  de peso corporal durante  8 o 9 semanas se  observó
una  vacuolización de los hepatocitos periportales.  Estos
cambios  se acompañaban a veces de un aumento del peso del
hígado y de la actividad sérica de las enzimas hepáticas.

    El descenso de la concentración de hemoglobina  y  del
volumen eritrocítico en la rata, causado por  la  adminis-
tración  de  80 mg/kg (8 mg/kg  de  peso corporal)  en  la
dieta,  traduce un efecto en la síntesis de la hemoglobina
que  da lugar a  una anemia hipocrómica  microcítica.   El
descenso  de las cifras de  hemosiderina esplénica sugiere
una  alteración del equilibrio del hierro.  En los ratones
se ha observado también anemia.

    En ciertas investigaciones a corto plazo, pero  no  en
los  estudios  prolongados,  se ha  observado formación de
rosetas  de  hematíes  en  los  ganglios  linfáticos   del
mesenterio.   No está clara la  significación biológica de
este fenómeno (posiblemente transitorio).

    El  efecto tóxico característico del  OTBE tiene lugar
en el sistema inmunitario:  a causa de los efectos  en  el
timo,  la  inmunidad  celular  se  menoscaba.   Aunque  se
desconoce  el  mecanismo de  acción,  es posible  que esté
relacionado  con la conversión metabólica en compuestos de
dibutilestaño.   También  resulta afectada  la resistencia

    Asimismo  se  han  observado efectos  generales  en el
sistema   inmunitario  (por  ejemplo,  en  el  peso  y  la
morfología  de  los  tejidos linfoides,  los  recuentos de
linfocitos  periféricos  y las  concentraciones totales de
inmunoglobulinas  séricas) en diferentes estudios con OTBE
realizados en ratas y perros, pero no en ratones, mediante
dosis  claramente tóxicas (en  el ratón se  han registrado
efectos  con  dosis de  cloruro  de tributilestaño  de 150
mg/kg).  Solamente la rata presenta signos de toxicidad y,
evidentemente, es la especie más sensible. En los estudios
a  corto plazo, el NENO  para esta especie fue  de 5 mg/kg
(0,6 mg/kg de peso corporal) en la dieta.  En los estudios
con  cloruro  de  tributilestaño se  han observado efectos
análogos  en el timo,  rápidamente reversibles tan  pronto
como  se  interrumpía  la  administración.  Según  se   ha
observado  en estudios  in vivo de resistencia del huésped,
el  OTBE pone en peligro la función inmunitaria específica

en  la rata.  Tras la exposición a un nivel de 50 mg/kg en
la  dieta  (siendo el  valor NENO de  5 mg/kg al día),  se
observó  una disminución del  "aclaramiento" de  Listeria
monocytogenes, mientras  que con 50 y 5 mg/kg en la dieta,
pero  no con 0,5 mg/kg  (2,5, 0,25 y  0,025 mg/kg de  peso
corporal  al día, respectivamente) se  produjo un descenso
de  la  resistencia  a  Trichinella  spiralis. Los   mismos
efectos,  aunque  menos  pronunciados,  se  observaron  en
animales de más edad.

    Según  los conocimientos actuales,  los efectos en  la
resistencia  del  huésped  se deben  probablemente  a  una
evaluación  más adecuada de  los posibles riesgos  para el
hombre;  sin embargo, no  se tiene suficiente  experiencia
con  los  sistemas  de  prueba  para  apreciar   bien   su
importancia.   En  cualquier  caso, las  observaciones  en
ratas  atímicas  "desnudas",  estimuladas en  las condi-
ciones   ordinarias,   proporcionan  algunos   datos  para
interpretar  el modelo de  T. spiralis. En  estos estudios,
la  ausencia completa de inmunidad timodependiente aumentó
de 10 a 20 veces los recuentos de larvas en el músculo; en
cambio,  la exposición a concentraciones de OTBE de 5 y 50
mg/kg  en la dieta duplicó y cuadruplicó, respectivamente,
la cifra inicial.

    Aunque  ahora se dispone  de algunos datos  sobre  los
efectos   de  los  compuestos  de   tributilestaño  en  el
desarrollo  del sistema inmunitario, carecemos de informa-
ción sobre la resistencia del huésped.

    Lo prudente sería evaluar los posibles riesgos para el
hombre en función de los datos obtenidos en la especie más
sensible.   Los efectos sobre la resistencia del huésped a
 T.  spiralis se han observado ya con niveles en  la  dieta
de  5 mg/kg (equivalentes a 0,25 mg/kg de peso corporal al
día),  por lo que el  valor NENO es de  0,5 mg/kg (equiva-
lente  a  0,025 mg/kg al  día).   Sin embargo,  no  existe
acuerdo sobre el significado de estos datos  para  evaluar
los  riesgos en el hombre.   En todos los demás  estudios,
una concentración de 5 mg/kg al día en la  dieta  (equiva-
lente a 0,5 mg/kg de peso corporal, sobre la base  de  los
estudios  a corto plazo)  correspondía al valor  NENO  con
respecto  a los efectos  en el sistema  inmunitario, tanto
generales como específicos.

12.3  Toxicidad a largo plazo

    Un  estudio a  largo plazo  en las  ratas  sugiere  un
efecto  marginal del TBE  en los parámetros  toxicológicos
generales  (cuya significación toxicológica es limitada) a
una concentración de 5 mg/kg (0,25 mg/kg de peso corporal)
en la dieta.

12.4  Genotoxicidad

    La  genotoxicidad del OTBE ha sido objeto de detenidas
investigaciones.   En la gran  mayoría de los  estudios se
han obtenido resultados negativos, y no hay ninguna prueba
convincente de que el OTBE tenga propiedades mutagénicas.

12.5  Toxicidad en el sistema reproductor

    En tres especies de mamíferos (ratón, rata  y  conejo)
se  ha evaluado la  posible embriotoxicidad del  OTBE tras
administrarlo  por  vía oral  a  la madre.   La  principal
malformación observada en los fetos de rata y de ratón fue
la  fisura palatina, pero este efecto sólo se registró con
dosis  claramente tóxicas para las  madres.  Tal resultado
no se ha considerado como indicio de efectos teratogénicos
del  OTBE en dosis inferiores a las que producen toxicidad
materna.  El NENO más bajo en lo relativo a la embriotoxi-
cidad  y la fetotoxicidad para  todos las especies fue  de
1,0 mg/kg de peso corporal.

12.6  Carcinogenicidad

    En  las ratas se ha  realizado un estudio de  carcino-
genicidad en el que se obtuvieron alteraciones neoplásicas
de  los órganos endocrinos con  50 mg/kg en la dieta.   En
cuanto  a  los  tumores hipofisarios  registrados  con 0,5
mg/kg  en la dieta,  se consideró que  no tenían  signifi-
cación  biológica por no existir relación dosis-respuesta.
Estos  tipos de tumores suelen aparecer con una incidencia
general  elevada y variable,  por lo que  su significación
parece  discutible. Está en  curso un estudio  de carcino-
genicidad en el ratón.

13.  Efectos en el ser humano

    Se ha observado que la exposición profesional  de  los
trabajadores  al tributilestaño provoca irritación  de las
vías  respiratorias superiores.  En  forma de aerosol,  el
TBE entraña un riesgo para las personas.  El  OTBE  ejerce
un efecto irritante en la piel y en los  ojos,  habiéndose
observado  casos  de  dermatitis grave  a consecuencia del
contacto directo con la piel.  El problema se  agrava  por
la falta de una respuesta cutánea inmediata.


1.  Evaluación del riesgo para las personas

    La  exposición  de  los trabajadores  se produce sobre
todo  en las actividades  de fabricación y  formulación de
compuestos  de tributilestaño, durante la  aplicación y la
eliminación de pinturas a base de TBE y a consecuencia del
empleo de TBE como conservador de la madera. La exposición
del público en general puede deberse a la contaminación de
los  alimentos, particularmente el pescado y los mariscos,
así como a la aplicación doméstica de productos de protec-
ción de la madera.

    A  juzgar por las pruebas realizadas en animales y por
la  observación directa de  las personas, parece  evidente
que los compuestos de TBE ejercen efectos irritantes en la
piel  y en los ojos  y que la inhalación  de los aerosoles
provoca irritación respiratoria.

    La  manipulación de madera tratada  no entraña riesgos
de  irritación dérmica si  se ha secado  el material.   En
cambio,  los aerosoles de TBE son muy peligrosos y no debe
permitirse  que la madera  vuelva a entrar  en la zona  de
tratamiento hasta que esté perfectamente seca.

    No  se ha señalado  ningún caso de  intoxicación  sis-
témica  aguda  y, probablemente,  el  TBE se  elimina  del
organismo  al cabo de pocos días.  Por consiguiente, no es
de temer que el manejo de productos de TBE entrañe peligro
de   toxicidad   aguda   si  se   toman  las  precauciones

    En  los  animales  de  laboratorio  se  han  observado
efectos  a  corto y  a largo plazo  en el hígado  y en los
sistemas  hematológico  y  endocrino.  Los  efectos de los
compuestos de TBE en el sistema inmunitario,  y  especial-
mente  en la resistencia del huésped, han resultado ser el
parámetro  más sensible de toxicidad en la rata, que es la
especie más sensible de todas las estudiadas.   Cuando  se
utiliza  Trichinella  spiralis como  modelo de  resistencia
del  huésped, el nivel  de efecto no  observado (NENO)  se
sitúa  entre 0,5 y 5,0 mg/kg  (0,025 y 0,25 mg/kg de  peso
corporal) en la dieta, mientras que si se utilizan índices
de función inmunitaria es de 0,6 mg/kg de peso corporal.

    Debido  a  las grandes  variaciones  en el  consumo de
pescado  y mariscos  y a  las diferencias  locales  de  la
cantidad  de residuos de TBE presentes en la fauna marina,
sólo  a título de orientación  pueden hacerse estimaciones
de los valores de exposición y del NENO.   Conviene  tener
en  cuenta que para evaluar  el riesgo potencial de  estos
compuestos hay que proceder en el plano local a determinar
los  residuos, calcular el consumo de pescado y mariscos y
fijar los límites aceptables de seguridad.

    Partiendo  de cifras de consumo de pescado de 15 y 150
g/día,  de un valor de residuos en el pescado de 1 mg/kg y
de  un peso  medio de  las personas  de 60 kilos,  se  han
obtenido  los siguientes márgenes de  seguridad basados en
diferentes puntos finales inmunológicos.

Consumo de     Ingestión diaria           Margen de seguridad
 pescado       estimada de TBE      Modelo de         Otros parámetros
 (g/día)           (µg/kg)          T. spiralis         inmunológicos
     15             0,25            100-1000               2500
    150             2,5              10-100                 250

    El  empleo indiscriminado e irresponsable de compuesto
de TBE y la observancia de las  recomendaciones  esbozadas
en  la presente monografía  para reducir la  exposición de
las  personas puede dar lugar a la ingestión de concentra-
ciones  de  compuesto  de  TBE  peligrosas  para  la salud

    En  los  animales  de  experimentación  sólo  de   han
observado  efectos  teratogénicos  con  dosis  que  causan
signos  patentes de toxicidad materna.   Por consiguiente,
se considera que el potencial teratogénico de TBE  es  muy

    Basándose  en los resultados de detallados estudios de
mutagenicidad,   se   considera  que   los  compuestos  de
tributilestaño  no  tienen  potencial mutagénico.   En  un
estudio de carcinogenicidad realizada en la rata con OTBE,
se  observó un aumento de la incidencia de ciertos tumores
endocrinos que aparecen espontáneamente con una incidencia
elevada   y   variable.    Por  consiguiente,   los  datos
disponibles  no indican claramente  que los compuestos  de
TBE entrañen riesgos carcinogénicos para las personas.

2.  Evaluación de los riesgos ambientales

    La  difusión  del  tributilestaño (TBE)  en  el  medio
ambiente  se debe sobre todo al empleo de ese compuesto en
las  pinturas antiincrustantes.  También puede  deberse al
empleo de TBE como molusquicida.  La contaminación  en  la
fuente  se  produce a  consecuencia  de utilizar  TBE como
biocida  en los sistemas de  refrigeración, la fabricación
de  pulpa de  madera, el  tratamiento de  los cueros,  los
procesos  de conservación de la madera y el tratamiento de
los productos textiles.

    Debido  a sus propiedades fisicoquímicas, los compues-
tos  de TBE se concentran en la microcapa superficial y en
los  sedimentos.  La degradación abiótica no parece ser un
mecanismo importante de eliminación del TBE en condiciones
ambientales. Aunque el OTBE es biodegradable en la columna
de  agua, este proceso  no es suficientemente  rápido para

impedir la aparición de concentraciones elevadas de TBE en
algunos  sitios.   En la  mayor  parte de  los  organismos
acuáticos  se  produce bioacumulación  pero la degradación
metabólica  es un proceso más  eficaz en los mamíferos  de

    El  TBE es sumamente peligroso para algunos organismos
acuáticos  por  ser  tóxico incluso  a concentraciones muy
bajas en el agua.  Tales concentraciones se  han  señalado
en  diversos lugares.  En los estudios sobre el terreno se
han   observado  efectos  adversos   sobre  invertebrados,
especialmente  moluscos, a los que no se pretendía atacar,
y esos efectos resultan suficientemente graves para inter-
ferir  en la  reproducción y  provocar un  descenso de  la
población.  En algunos sitios ha sido posible  anular  los
efectos  adversos en la  producción comercial de  mariscos
restringuiendo el empleo de pinturas antiincrustantes, con
lo que también se ha logrado suprimir el efecto de imposex
en  las poblaciones de  gasterópodos.  Los efectos  en las
piscifactorías  hacen  pensar  que  no  conviene  utilizar
pinturas que contengan TBE en las redes de contención.

    Para  el medio terrestre, el riesgo general parece ser
bajo.  La madera tratada con TBE podría ser peligrosa para
los  organismos terrestres que viven  en estrecho contacto
con ella.

    El  aumento de las concentraciones de TBE en la micro-
capa  superficial puede ser peligroso  para los organismos
del  litoral, las especies neustónicas (inclusive inverte-
brados bénticos y larvas de peces) y las aves  salvajes  y
marinas  que se alimentan en  la superficie del agua.   La
acumulación  y la baja biodegradación del TBE en los sedi-
mentos  puede representar un  peligro para los  organismos
acuáticos cuando esos sedimentos contaminados se movilizan
por procesos naturales u operaciones de dragado.


1.  Recomendaciones para proteger la salud humana y la
higiene del medio

a)  Hay  que instar a los  países Miembros que todavía  no
    hayan reglamentado el uso de compuestos de TBE  a  que
    lo hagan.

b)  Es  necesario  evaluar  y, si  procede reglamentar, el
    ingreso en el medio ambiente de estaño  orgánico  pro-
    cedente  de fuentes distintas de  las pinturas antiin-
    crustantes.  Por ejemplo, convendría evaluar el riesgo
    potencial  que entraña la aplicación al suelo de lodos
    de alcantarillado contaminados con TBE.

c)  Habrá  que mejorar los métodos  utilizables para apli-
    car,  eliminar y evacuar  en condiciones de  seguridad
    las pinturas de estaño orgánico.

2.  Investigaciones necesarias

a)  Habrá  que mejorar los métodos de detección y análisis
    a  fin de poder determinar con rapidez y precisión los
    compuestos de butilestaño presentes en concentraciones
    de  pg/litro.  Una de  las razones que  justifican esa
    recomendación  es que ciertos efectos  biológicos (por
    ejemplo,  el imposex en los  gasterópodos) pueden pro-
    ducirse  con concentraciones inferiores a los actuales
    límites de detección.

b)  Esnecesario  estudiar los mecanismos que concentran en
    vez de dispersar el TBE y que retrasan la degradación,
    prestando especial atención a los fundamentos químicos
    de ese compuesto y a su interaccción con las moléculas
    biológicas.  Se necesitan más estudios sobre la absor-
    ción del TBE en todos los niveles tróficos.

c)  Habrá  que estudiar la toxicidad del TBE en los organ-
    ismos  acuáticos.   Estos estudios  deberán versar, si
    procede,  sobre el metabolismo, los efectos endocrinos
    y la toxicidad inmunológica.

d)  Habrá  que buscar otras especies  sensibles (en parti-
    cular  de agua dulce)  que sirvan de  bioindicador  en
    otros grupos.

e)  Habrá que confirmar la validez de los  modelos  utili-
    zados para evaluar la inmunotoxicidad en los mamíferos
    y  definir con más precisión los niveles sin efecto de
    los parámetros pertinentes.

f)  Convendría  emprender un estudio de  toxicidad crónica
    en una segunda especie de mamífero.

g)  Convendría  emprender un estudio de tumorigenicidad en
    una segunda especie de mamífero.

h)  Hay  que obtener datos mediante métodos de especiación
    sobre  los niveles de  residuos de butilestaño  en los
    peces y mariscos destinados al consumo humano.

    See Also:
       Toxicological Abbreviations
       Tributyltin compounds (PIM G018)