Ambient Aquatic Life Water Quality Criteria for Tributyltin


AMBIENT AQUATIC LIFE WATER QUALITY CRITERIA FOR

                  TRIBUTYLTIN

         CAS Registry Number (See Text)
      U.S. ENVIRONMENTAL PROTECTION AGENCY
       OFFICE OF RESEARCH AND DEVELOPMENT
       ENVIRONMENTAL RESEARCH LABORATORIES
                DULUTH, MINNESOTA
           NARRAGANSETT, RHODE  ISLAND
                 Final .March 1991

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                                    NOTICES
This document  has  been reviewed  by Che  Environmental  Research Laboratories,
Duluth, MN  and Narragansett, RI,  Office  of Research and  Development  and the
Health and Ecological Criteria Division, Office  of Science  and Technology, U.S.
Environmental Protection Agency, and approved for publication.

Mention of trade names or commercial products does not constitute endorsement or
recommendation for use.

This  document  is  available  to  the  public  through  the  National  Technical
Information Service (NTIS), 5285 "Port Royal Road, Springfield,  VA  22161.
                                      ii

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                                   FOREWORD
      Section 304(a)(l) of the Clean Water Act of 1977 (P.L.  95-217) requires the
Administrator of the  Environmental  Protection  Agency to publish water quality
criteria that accurately reflect  the latest scientific knowledge on the kind and
extent of all identifiable effects on health and welfare that might be expected
from the presence of  pollutants  in any body of water, including ground water.
This document  is  a  revision of proposed  criteria  based upon consideration of
comments received from other federal agencies,  state  agencies,,  special interest
groups, and individual scientists.  Criteria contained in this  document replace
any previously published  EPA aquatic  life criteria for  the same pollutant(s).

      The  term  "water quality criteria" is used in  two sections of the Clean
Water Act, section  304(a)(l)  and section 303(c)(2).   The term has a different
program  impact  in  each  section.    In section 304,   the  term represents  a
non-regulatory, scientific assessment of ecological effects.  Criteria presented
in  this  document are such  scientific assessments.   If water  quality criteria
associated with specific stream uses  are adopted  by a state as water quality
standards under section 303, they become enforceable maximum acceptable pollutant
concentrations  in ambient:  waters  within that state.   Water quality criteria
adopted  in state water quality standards could have the same numerical values as
criteria developed under  section 304.  However, in many situations states might
want  to  adjust water quality  criteria developed under section 304 to reflect
local environmental conditions and human exposure patterns  before incorporation
into water quality  standards.  It is  not  until their adoption  as part of state
water quality  standards that  criteria become regulatory.

      Guidance  to assist states in the modification of criteria presented in this
document,  in   the  development  of   water  quality  standards,  and  in  other
water-related  programs of this agency have  been developed  by EPA.
                         Tudor T. Davies
                         Director
                         Office of Science and Technology
                                       iii

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                                ACKNOWLEDGMENTS
Larry T. Brooke                         David J.  Hansen
(freshwater author)                  .   (saltwater author)
University of Wisconsin-Superior        Environmental Research Laboratory
Superior, Wisconsin                     Narragansett, Rhode Island
                                        Robert Scott Carr
                                        (saltwater author)
                                        Battelle New  England Laboratory
                                        Duxbury,  Massachusetts
Robert L. Spehar                       David J. Hansen
(document coordinator)                 (saltwater coordinator)
Environmental Research Laboratory      Environmental Research Laboratory
Duluth, Minnesota                      Narragansett, Rhode Island
                                       IV

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                                   CONTENTS








                          i                        •                       Page




 Foreword	 . .	 i i i




 Acknowledgments		  iv




 Tables	•	• • •  vi




 Text  Tables	 vii






 Introduction	   1



 Acute Toxicity to Aquatic Animals	   5




 Chronic Toxicity to Aquatic Animals	   8




 Toxicity  to  Aquatic Plants	  11




• Bioaccumulation	  12




 Other Data	' • • •  *3




 Unused  Data	• •	, 25




• Summary	•	•  27



 National  Criteria	 . . .	  29




 Implementation	  30








 References	  65

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                                    TABLES
                                                                         Page



1.  Acute Toxicity of Tributyltin to Aquatic Animals 	   32




2.  Chronic Toxicity of Tributyltin to Aquatic Animals 	   39




3.  Ranked Genus Mean Acute Values with Species Mean Acute-Chronic



      Ratios	41




A.  Toxicity of Tributyltin to Aquatic Plants  .	46




5.  Bioaccumulation of Tributyltin by Aquatic Organisms  	   48




6.  Other Data on Effects of Tributyltin on Aquatic Organisms  .....   51
                                      VI

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                                 TEXT TABLES

                                                                       Page
1.   Summary of available laboratory and field studies
      relating the extent of imposex of female snails,
      measured by relative penis size (volume female
      penis + male penis «• RPS) and the vas deferens
      sequence index (VDS), as a function of tributyltin
      concentration in water and dry tissue	   17

2.   Summary of laboratory and field data on the effects
      of tributyltin on saltwater organisms at concentrations
      less than the Final Chronic Value of 0.0485 MgA   ........   22
                                       vii

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Introduction
               .*

      Organotins  are compounds  consisting  of one  to  four  organic moieties


attached to a tin atom via carbon-tin covalent bonds.  When there are fewer than


four  carbon-tin  bonds,   the organotin  compound will  be a cation  unless  the


remaining valences of tin are occupied  by  an anion such  as acetate,  carbonate,


chloride,  fluoride,  hydroxide,  oxide,  or  sulfide.    Thus  a  species  such  as


tributyltin  (TBT)  is a cation whose formula Is  (C4H,)jSn*. ' In sea water TBT


exists mainly as  a mixture of the chloride, the hydroxide, the aquo complex, and


the carbonate complex (Laughlin et al.  1986a).


      Several review papers have been written which cover the  production, use,


chemistry, toxicity, fate and hazards of TBT in the aquatic environment  (Clark


et al.  1988; Eisler  1989;  Oceans 86 1986; Oceans  87 1987;  WHO 1990).  The


toxicities of organotin compounds are related to the  number of organic moieties


bonded to the tin atom and to the number of  carbon atoms in the organic moieties.


Toxicity  to  aquatic organisms  generally  increases  as the number  of organic


moieties increases from one  to three and decreases with  the incorporation of a


fourth,  making  triorganotins  more  toxic  than  other  forms.     Within  the


triorganotins, toxicity increases as the number of carbon atoms  in  the organic


moiety increases  from one  to four, then decreases.  Thus the organotin most toxic


to aquatic life is TBT (Hall and Pinkney 1985; Laughlin and Linden 1985; Laughlin


et al.  1985).  TBTs inhibit Ha* and  K* ATPases and are ionophores  controlling


exchange of Cl",  Br~, F* and other ions  across  cell membranes  (Selwyn 1976).


      Organotins are used in several  manufacturing processes,  for example, as an


anti-yellowing agent in clear plastics and  as a catalyst in poly(vinyl chloride)


products  (Fiver  1973).    One of the more  extensive  uses of  organotins  is  as


biocides (fungicides, bactericides, insecticides)  and as preservatives for wood,


textiles, paper,  leather and electrical  equipment.  The  use of TBT in  antifouling

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paints on  ships,  boats,  docks and  cooling  towers probably  contributes  most



significantly  to  direct  release  of organotins  into  the  aquatic  environment



(Clark et al. 1988; Hall and Pinkney 1985; Kinnetic Laboratory 1984).




      The U.S. Navy (1984) proposed application of some paints  containing TBT to




hulls of naval ships.  Such paint formulations have been shown to be an effective




and relatively long-lived deterrent  to adhesion of barnacles and other fouling



organisms.  Encrustations of these organisms on ships' hulls reduce maximum speed



and increase  fuel  consumption.   According to the U.S. Navy (1984),  use of TBT



paints,  relative   to  other  antifouling paints,  would not  only reduce  fuel




consumption by 15% but would also increase time between repainting from less than




5 years  to  5  to 7 years.  Release of TBT to water 'occurs during repainting in



shipyards when old paint is sand-blasted off and new  paint applied.  TBT would




a,lso be released continuously from the hulls  of the painted ships.  Antifouling




paints in current  use  contain copper as the primary biocide, whereas the proposed




TBT paints would contain both copper and TBT.  Interaction between the  toxicities



of TBT and other ingredients in the paint apparently is  negligible (Davidson et




al. 1986a).



      The solubility  of TBT compounds in water is influenced by such  factors as




the  oxidation-reduction potential,  pH,  temperature,   ionic  strength,  and




concentration and composition of the  dissolved organic matter  (Clark et al. 1988;




Corbin 1976).  The solubility of tributyltin oxide in water was  reported to be



750  ug/L «t  pH of 6.6,  31,000  ug/L at pH  of 8.1 and 30,000 ug/L  at pH 2.6



j(Maguire et ml. 1983).  The carbon-tin covalent bond does not hydrolyze in water



.(Maguire et al. 1983,1984), and  the  half-life for photolysis due to sunlight is




greater  than 89 days (Maguire et  al.  1985; Seligman et  al. 1986).  Biodegradation




is the major breakdown  pathway for TBT in water and  sediments with half-lives of




several  days in water to several weeks in sediments (Clark et al. 1988; Lee et

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al. 1987; Maguire and  Tlcaez  1985;  Seligmah et al, 1986, 1988, 1989; Stang and



Seligman  1986).   Breakdown  products include di-  and monobutyltins with some



butylmethy1tins detected.        .




      Some species of algae,  bacteria, and fungi have been  shown  to  degrade TBT



by  sequential  dealkylation,   resulting  in dibutyltin,  then monobutyltin,  and



finally inorganic tin (Barug 1981;  Maguire et al. 1984).  Barug (1981) observed



the biodegradation of  TBT to di- and monobutyltin by bacteria  and fungi only



under aerobic conditions and only when a  secondary carbon  source was supplied.



Inorganic tin  can be methylated by estuarine microorganisms  (Jackson  et al.



1982).  Maguire et al.  (1984) reported that  a  28-day culture of TBT with the



green alga,  Ankistrodesmus  falcatus.. resulted  in 7%  inorganic  tin.   Maguire



(1986) reported that the half-life of TBT exposed  to microbial degradation was



five months under aerobic conditions and 1.5 months under anaerobic  conditions.



TBT is also accumulated and metabolized by Zostera marina (Francios et al. 1989).



The major metabolite  of TBT in saltwater crabs,  fish,  and shrimp was dibutyltin



(Lee 1986).




      TBT readily sorbs to sediments and suspended solids and can persist there



(Cardarelli and Evans 1980).   In some instances, most TBT  in the water column



(70-90%) is associated with  the  dissolved phase  (Valkirs et al.  1986a; Maguire



1986;  Johnson et al.  1987).  The half-life for desorption of TBT from sediments



is reported to be greater than ten months (Maguire and Tkacz 1985).



      Elevated TBT concentrations in fresh and salt waters,  sediments or biota,



are primarily associated with harbors and marinas  (Cleary and Stebbing 1985; Hall



1988;  Hall et  al.  1986; Langston et  al.  1987;  Maguire 1984,1986;  Maguire and



Tkacz 1985;  Maguire et al.. 1982; Quevauviller et al.  1989;  Salazar and Salazar



1985b; Seligman et al.  1986,1989;  Short  and  Sharp 1989;  Stallard et al.  1986.



Stang and Seligman 1986; Unger  et  al. 1986;  Valkirs  et al. 1986b;  Valdock and

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 Miller 1983;  Waldock et al.  1987.).   Lenihan et al.  (1990)  hypothesized that



 changes in faunal composition in hard bottom communities in San Diego Bay were




 related to boat mooring and TBT:  Salazar and Salazar (1988) found an apparent



 relationship between concentrations  of TBT in waters of San Diego Bay and reduced



 growth of mussels.  Organotin concentrations in the low part per trillion range




 have been associated with oyster shell malformations (Alzieu et al. 1989),  In



 some cases the water surface microlayer contained a much higher concentration of



 TBT than the water  column (Cleary and Stebbing 1987; Hall et al. 1986; Valkirs




 et al. 1986a).   Gucinski  (1986) suggested that this enrichment of the surface




 microlayer  might increase  the bioavailability  of TBT.   TBT  accumulates  in



 'sediments with sorption coefficients which may range from  l.lxlO2 to 8.2xl03 L/Kg



 and desorption appears  to be  a two  step process (Unger et al. 1987,1988).  No




••organotlns  were  detected  in  the  muscle  tissue  of  feral chinook salmon,




.Qncorhvnchus  tshawvtscha. caught near Auke Bay,  Alaska,  but concentrations as




 high as 900 ug/kg were reported in muscle tissue of chinook salmon held in pens




 treated with  TBT (Short 1987;  Short and  Thrower  1986a).



       Only  data  generated in toxicity  and bioconcentration  tests  on TBTC




 (tributyltin  chloride; CAS 1461-22-9), TBTF (tributyltin fluoride;  CAS 1983-10-




 4), TBTO  [bis(tributyltin)  oxide;  CAS  56-35-9],  commonly called  "tributyltin



 oxide"  and TBTS [bis(tributyltin)  sulfide;  CAS  4808-30-4],  commonly  called




 "tributyltin  sulfide" were used in  the derivation of the  water quality criteria



 concentrations for  aquatic life presented herein.  All concentrations  from such



* tests  are  expressed as  TBT,  not  as  tin  and  not  as  the  chemical tested.



 Therefore, many  concentrations listed herein are not those in the .reference cited




• but are concentrations  adjusted to  TBT.  A-comprehension of the  "Guidelines  for



 Deriving Numerical  National Water Quality Criteria for the Protection of Aquatic




 Organisms  and Their Uses" (Stephan  et al. 1985), hereinafter referred to  as  the

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Guidelines, and the response to public comment (U.S. EPA 1985a)  is necessary to




understand  the following  text,  tables,  and  calculations.   Results  of such




intermediate calculations as recalculated LCSOs and Species Mean Acute Values are




given  to four  significant figures  to prevent  roundoff error  in  subsequent




calculations, not to reflect the precision of the value.  The Guidelines requires




that all available pertinent laboratory and field information be  used to derive




a criterion consistent with sound scientific evidence.   The  saltwater criterion




for TBT follows this requirement by using data from chronic exposures  of copepods




and molluscs rather than Final Acute Values  and Acute-Chronic Ratios to derive




the Final Chronic Value.  The Federal Insecticide, Fungicide, and Rodenticide Act




(FIFRA) data base of  information from the pesticide industry was searched and



some useful  information was  located for deriving  the criteria.   The latest




comprehensive literature search for information for this document was conducted




in November 1990, some newer information has been included.








Acute Toxicitv to Aquatic Animals




      Data that may be used, according to the Guidelines, in the derivation of




Final Acute Values for TBT are presented in Table 1.   Acute values are available




for thirteen freshwater  species representing twelve genera.  The acute values




range  from 1.14  ug/L for  a hydra,  Hvdra  oli^actis.  to  24,600  ug/L  for  a




freshwater calm, Elliptic  comolanatus.  The relatively low sensitivity of the




freshwater clan to TBT is surprising due to  the mollusicidal qualities of TBT.




The organism likely  closes itself to the environment, minimizing chemical intake,




and is able to tolerate high concentrations of TBT temporarily.




      The most sensitive  freshwater organisms tested are hydras (Table 3).  Three




species were tested and have Species Mean Acute Values (SMAVs) ranging from 1.14




to 1.80 ug/L.   Other invertebrate  species tested are an amphipod,  a cladoceran,
                                       5

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an annelid and a dipceran larvae.  Brooke et al.  (1986) conducted flow-through




measured  tests  with  an amphipod,  Gammarus pseudolimnaeus.  and  an annelid,



Lumbriculus varieyatus. and a  static  measured  test with larvae of a mosquito,




Culex sp.  The 96-hr LCSOs and SMAVs are 3.7, 5.4 and 10.2 ug/L, respectively.



Five tests with the daphnid, Daohnia mapia. were  conducted.  The 48-hr EC50 value



of 66.3 ug/L (Foster 1981) was considerably less sensitive than chose from the



other, tests  which  ranged from  1.58  ug/L  (Lefllanc 1976)  to 11.2  ug/L (ABC



Laboratories, Inc. 1990c).  The SMAV for fi. aaaifl is 4.3  ug/L because, according




to  the Guidelines,  when  test results  are available  from  flow-through and




concentration measured  tests,  these have  precedence over other types of acute




tests.                                                 .



      All the vertebrate species  tested are fish.  The most sensitive species is




the fathead minnow,  Pimephales promelas.  which has  a SMAV of 2.6 ug/L from  a



single 96-hr flow-through  measured test  (Brooke et al.  1986).  Rainbow trout,




Oncorhynchus mvkiss. were tested by four groups with good agreement.  The 96-hr




LCSOs ranged from 3.45 to 7.1 ug/L with a SMAV of 4.571 ug/L for the  three tests



(Brooke et al. 1986;  ABC  Laboratories, Inc.  1990a) which were conducted flow-




through and concentrations were measured.  Bluegill, Lepomis  macrochirus. were




tested by three  groups.   The  value of 227.4 ug/L (Foster 1981) appears high



compared to those of 7.2 ug/L (Buccafusco 1976b) and 8.3 ug/L (ABC Laboratories,




Inc. 1990b).  Only the flow-through measured test can be used, according  to the




Guidelines, to calculate the SMAV of  8.3 ug/L.



       Freshwater Genus mean Acute Values  (GMAVs) are available for twelve  genera




which  vary by more than  21,000  times from  the  least  sensitive  to the most




sensitive.  Removing the least sensitive genera, Elliptic, the remainder  differ




from  one another by a maximum  factor of 8.7  times.    Based upon  the  twelve

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available GMAVs  the  Final  Acute Value (FAV)  for freshwater organisms is 0.9177




ug/L.  The  FAV  is  lower than the lowest freshwater SMAV.




      Tests of  the acute toxicity of TBT  to  resident  North American saltwater




species  that are useful for deriving water quality criteria concentrations have




been performed  with  20 species of invertebrates and six species of fish (Table




1).   The range of acute toxicity to saltwater animals  is a  factor of about 670.




Acute  values range  from 0.42  ug/L for juveniles  of  the  mysid,  Acanthomysis



sculpta  (Davidson  et al.  1986a,1986b) to 282.2 ug/L for adult Pacific oysters,




Crassostrea gigas  (Thain 1983).   The  96-hr LC50s for six saltwater fish species




range  from  1.460  ug/L for juvenile  chinook salmon,  Oncorhvnchus  tshawvtscha




 (Short   and Thrower  1986b)  to  25.9  ug/L  for subadult   sheepshead  minnows,




Cyprinodon  variegatus (Bushong et al. 1988).                       .




      Larval bivalve molluscs and juvenile crustaceans appear  to  be much more




sensitive than adults during acute exposures.  The  96-hr LC50 for larval Pacific




.oysters  was 1.557 ug/L, whereas the value for adults was 282.2 ug/L (Thain 1983).




The 96-hr LCSOs for larval and adult blue mussels, Mvtilus edulis. were 2.238 and




 36.98 ug/L,  respectively (Thain 1983). Juveniles of the crustaceans Acanthomvsis




 sculota  and Metamvsidopsis  elongata  were  slightly more sensitive to  TBT than




 adults  (Davidson et al. 1986a,1986b;  Valkirs et al. 1985;  Salazar and Salazar,




Manuscript).



       Genus Mean Acute Values for 25 saltwater genera range from 0.61 ug/L for




 Acanthomvsis to 204.4 ug/L for Ostrea (Table  3) .  Genus Mean Acute Values for the




 11 most  sensitive genera differ by a  factor of less than four.  Included within



 these genera are three species of molluscs  and eight species of crustaceans.  The




 saltwater Final Acute Value for TBT was calculated to  be 0.7128 ug/L (Table 3),




 which is greater than the lowest saltwater  Species Mean Acute Value  of 0.61 /ig/L.

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Chronic T/ixieitv to Aquatic Anima.s
      The available data that are usable, according to the Guidelines, concerning
the chronic  toxicity  of  TBT are presented  in Table 2.   Brooke  et al.  (1986)
conducted a. 21-day  life-cycle test with a freshwater cladoceran and reported that
the survival of adult Daphnia magna was 40%  at a TBT concentration  of 0.5 ug/L,
and 100% at 0.2 ug/L.  The mean number of young per adult per reproductive day
was reduced 30% by 0.2 ug/L, and was reduced  only 6% by 0.1 ug/L.  The chronic
limits  are 0.1  and  0.2  ug/L based  upon  the  reproductive effects  on adult
daphnids.  The chronic value for Daphnia  magna is calculated to be  0.1414 ug/L,
and the acute-chronic ratio of 30.41 is calculated using the acute  value of 4.3
ug/L from  the same study.
      Daphnia pagna were exposed in a second 21-day  life-cycle tsst to TBT (ABC
Laboratories, Inc.  1990d). Exposure concentrations ranged  from 0.12 to 1.27 ug/L
as TBT.  Survival of adults was significantly  reduced (45%) from the controls at
>0.34 ug/L but not  at 0.19 ug/L.  Mean number of young per adult per  reproductive
day was significantly reduced at the same concentrations affecting survival. The
chronic  limits  are set at 0.19 where no  effects were seen  and 0.34 ug/L where
survival and reproduction were reduced.  The Chronic  Value  is 0.2542 ug/L and the
Acute-Chronic Ratio is 44.06  when  calculated  from the  acute value  of  11.2 ug/L
from the same test.  The Acute-Chronic Ratio  for D..  ma^na. is 36.60  which is the
geometric  mean of  the  two available  Acute-Chronic ratios  (30.41  and 44.06) for
P.. magna.
       In an  early  life-stage  test with the fathead minnow, Pimephales prome;as.
all fish exposed to the highest exposure  concentration  of  2.20 ug/L died during
the test (Brooke et al. 1986).  Survival  was reduced by 2% at the next lower TBT
concentration of 0.92  ug/L, but was higher than in the  controls at 0.45 ug/L and
lower  concentrations.  The mean weight of the surviving fish was reduced  4%  at
                                       8

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0.08 ug/L, 9% at 0.15  ug/L,  26%  at 0.45 ug/L, and 48% at 0.92 ug/L.  Mean length




of fry at the end of the test was significantly reduced at concentrations >0.45



ug/L.  The  mean biomass at  the  end of the test was  higher at the lowest TBT




concentrations (0.08 and O1.15 ug/L) than in the controls, but was reduced by 13




and 52% at TBT concentrations of 0.45 and 0.92  ug/L, respectively.  Because the




reductions in weight were small at  the two lowest concentrations (0.08 and 0.15




ug/L) and the mean biomass increased  at these  same concentrations, the chronic




limits are 0.15 and 0.45 ug/L based upon  growth (length and weight).  Thus the




chronic value  is  0.2598 ug/L and the acute-chronic  ratio  is 10.01 calculated




using the acute value of 2.6 ug/L from the same study.




      Life-cycle toxicity tests  have  been conducted with the saltwater mysid,




Acanthomvsis  sculpta  (Davidson  et  al.  1986a,1986b).    The  effects of  TBT on




survival, growth, and  reproduction of A., sculota were determined in five separate




tests lasting from 28 to 63  days.  The tests separately examined effects of TBT




on survival (1 test),  growth (3 tests) and reproduction (1 test) instead of the




approach of examining  all imdpoints  in one life-cycle test. All tests began with




newly released juveniles and lasted through maturation and spawning, therefore,




are treated as one life-cycle test.  The number of juveniles  released per female




at a TBT  concentration of 0.19 ug/L was 50% of the number released in the control




treatment, whereas  the  number  released at 0.09 ug/L was  higher  than  in the




control  treatment.   Reductions  in  juveniles released resulted from deaths of




embryos  within  broad  pouches   of  individual   females and  not from  reduced




fecundity.  Numbers of  females releasing  viable juveniles was reduced in 0.19




and 0.33 pg/L.  At concentrations of 0.38  ug/L and above, survival and weight of




female mysids were  always reduced;  all mysids  in 0.48 Mg/L died.   The chronic



value is 0.1308 ug/L, and the acute-chronic ratio is 4.664  (Table 2).

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       Two partial  life-cycle  toxi-city tests were conducted using the copepod.



 Eurvtemora  affinis (Hall et  al.  1987;1988a).    Tests  began with egg-carrying



 females and lasted 13 days.  In  the first test,  mean brood size was reduced from




 15.2  neonates/female in  the  control  to 0.2  neonates/female  in  0.479 ug/L.




 Percentage  survival of neonates  relative  to  controls was 21%  in  0.088 ug/L




 (nominal concentration of 0.100 ug/L), and  0%  in 0.479 ug/L.  The chronic value




 is <0.088 ug/L in this test.  In the second test, percentage survival  of neonates



••was significantly  reduced (27%  relative  to  controls) in  0.224 ug/L; brood size



 was unaffected in  any tested concentration (0.018-0.224  ug/L).   Although no




 statistically  significant  effects  were detected  in  <0.100  ug/L,  percentage



 survival of neonates appears reduced;  76% vs 90%  in controls.  The chronic value



 in this test is 0.150 ug/L.  Survival of neonates  in both tests in the 0.100 ug/L



 >nominal concentration (mean measured concentration -  0.094 ug/L) averaged  42%




 -relative to controls.  If this is the best estimate of the upper chronic value.




 and the 0.056  /ig/L treatment  from the second  test is  the  best estimate of  the




 J.ower chronic value, the overall chronic value  for the  two  tests is  0.0725 ug/L.



 The overall acute-chronic ratio is 27.24 when the acute  value of 1.975 ug/L (mean




 of acute values  of 1.4,  2.2 and 2.5 ug/L)  is used.



       The Final Acute-Chronic Ratio of 14.69 was calculated as  the geometric mean




 of  the  acute-chronic ratios  of 36.60 for Daohnia ma?na.  10.01  for PJmephales




 promelas.  4.664  for Acanthomvsis sculota  and  27.24  for  Eurvtemora af, finis,.




 Division  of th«  freshwater  and  saltwater Final Acute Values by 14.69 results in



 -Final Chronic Values for freshwater of 0.0625  ug/L and for saltwater of 0.0485



 ug/L  (Table 3).    Both  of  these Chronic Values are  below the  experimentally




• determined chronic values from life-cycle or early life-stage tests.
                                        10

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Toxieiev co Aauaeie Plants




      Blanck «t  al.  (1984)  reported  the  concentrations  of TBT that prevented




growth of  thirteen freshwater algal  species  (Table  4).   These concentrations




ranged from 56.1 to 1,782 ug/L, but most were between 100 and 250 ug/L.  No data




are available on the effects of TBT on freshwater vascular plants.




      Toxicity tests on TBT have been conducted with five species of saltwater



phytoplankton including  the  green alga,  Dunaliella  tertioleeta:  the diatoms,



Minutorellus polvmorphus, Hitzshic sp., Phaeodactvlum trleornyn^m, Skeletonema




costatum.  and  Thallassiosira  oseudonana:  the  dinoflagellate,   Gvmnodinium




solendens.  the microalga, Pavlova lutheri and the macroalga, Fucus vesiculosus




(Tables  4  and 6).   The  14-day  EC50 of  0.06228 ug/L  for.fi.  costatum (EG&G




Bionomics 1981c) was the lowest value reported, but Thain (1983) reported that




a measured  concentration  of 0.9732 ug/L was algistatic to the same species (Table




4).  The  72-hr ECSOs based on population growth ranged from approximately 0.3 to




< 0.5 ug/L (Table 6).  Lethal concentrations were generally more than an order




of magnitude greater than ECSOs and ranged from 10.24  to  13.82 ug/L.  Identical




tests  conducted on tributyltin  acetate,,  tributyltin   chloride,  tributyltin




fluoride, and tributyltin oxide with  fi. costatum resulted in ECSOs from 0.2346




to 0.4693 ug/L and LCSOs from 10.24 to 13.82 ug/L (Walsh et al. 1985).




      A  Final Plant Value,  as defined  in the Guidelines,  cannot be obtained



because no  test in which Uhe concentrations of TBT were measured and the endpoint




was biologically important has been conducted with  an important aquatic plant




species.  The available  data  do  indicate  that freshwater and saltwater plants




will be protected by TBT concentrations that adequately protect freshwater and




saltwater animals.
                                      11

-------
      Bioaccunulation of TBT has been measured in one species of freshwater fish




(Table 5).   Hartin et  al.  (1989)  determined the  whole  body bioconcentration




factor (BCF) for  rainbow trout to  be 406 after  a  64-day exposure  to 0.513 Mg



TBT/L.  Equilibrium of  the TBT concentration was achieved in the fish in 24 to



48 hrs.   In a separate exposure  to 1.026 MgTBT/L,  rainbow  trout  organs were



assayed for TBT content after a 15-day exposure.  The BCFs ranged from 312 for




muscle to 5,419 for peritoneal fat.  TBT was more highly concentrated than the




metabolites of di- and  monobutyltin or tin.



      The extent to which TBT is accumulated by saltwater animals  from the field




or from laboratory tests lasting 28 days or more' has been investigated with three



species of  bivalve molluscs'and a snail  (Table  5).  . Thain and Waldock (1985)



reported a  BCF of 6,833 for  the soft parts of blue mussel spat exposed to 0.24




ug/L  for 45 days.   In other laboratory exposures of blue mussels,  Salazar and




Salazar (1990a) observed BCFs of 10,400 to 37,500 after 56 to 60 days. BAFs from



field deployments of mussels were similar to BCFs  from laboratory studies; 11,000




to 25,000  (Salazar and  Salazar 1990a) and 5,000  to 60,000  (Salazar  and Salazar




in press).   Laboratory BCFs for the snail Nucella leoillis (11,000  to 38,000)




were  also similar  to field BAFs (17,000)  (Bryan et  al. 1987).  The soft parts of



the Pacific oyster exposed to TBT for 56 days  contained 11,400 times the exposure




concentration  of  Q.146 ug/L  (Waldock  and Thain  1983).   A BCF  of 6,047 was



observed for the  soft parts  of the Pacific oyster exposed to 0.1460 ug/L  for 21



days  (Waldock  et  al. 1983).  The lowest steady-state BCF reported for a bivalve




was 192.3 for the  soft parts of the European  flat oyster, Qstyea eduUs, exposed



 to  a  TBT  concentration of 2.62 ug/L for 45 days (Thain 1986;  Thain and Waldock




 1985).
                                       12

-------
      No U.S.  FDA action level  or other maximum  acceptable concentration in



tissue, as defined in the Guidelines, is available for TBT,  and, therefore, no




Final Residue Value can be calculated.








Other Data




      Additional  data  on the lethal  and sublethal effects  of  TBT on aquatic




species  are  presented  in Table  6.   Wong  et al.  (1982)   exposed  a natural




assemblage of freshwater algae and several pure cultures of various algal species




to TBT  in 4-hr  exposures.    Effects (ECSOs)  were seen  in all cases  on the




production or  reproduction  at concentrations  ranging from  5 to 20 ug/L which




demonstrates a high.sensitivity to TBT.




      Larvae of  the  clam,  Corbicula fluminea. has a  24-hr  EC50 of 1,990 ug/L




which is a high concentration relative to most other species of tested freshwater




organisms.    Another species  of clam,  Elliptic  comolanatus. also  showed low




sensitivity to TBT with a 96-hr LC50 of 24,600 ug/L (Table 1).  Various bivalve




clam  species  may have the  ability to  reduce exposure to  TBT  temporarily by




closing the valves.




      The cladoceran, Daphnia aagna. has 24-hr ECSOs ranging from 3  to  13.6 ug/L




(Bolster and  Halacha 1972;  Vighi  and Calamari 1985).  When a  more  sensitive



endpoint of altered phototaxis was  examined in a longer-term exposure of 8 days,



the effect concentration (0.45 ug/L) was much lower (Meador  1986).  Similarly,




rainbow trout  (Oncorhvnchua? mvkiss) exposed in short-term exposures of 24 to 48




hr have LC50 and EC50 values from 18.9 to 30.8 ug/L  (Table 6).  When the exposure




is increased to  110 days, the LC100 decreased to 4.46 ug/L and a 10%  reduction




in growth is seen at 0.18 ug/L.  The  frog, Rana temporaria.  has a LC50 of 28.2




ug/L  for a 5-day exposure to TBT.           .            •  .
                                       13

-------
      An attempt was made to measure the bioconcentration of  TBT with the green




alga, fnkistrodesmus  falcatus (Maguire et  al.  1984).   The  algae  are  able to




degrade TBT to its di- and monobutyl forms.   As  a  result,  the concentrations of



TBT steadily declined during  the 28-day study.  During the first seven days of



exposure, the concentrations declined from 20 to 5.2 ug/L and  the calculated BCF



was 300 (Table 6).  After 28 days  of  exposure, the TBT concentration had declined



to 1.5 ug/L and the calculated BCF  was 467.




      TBT  has  been  shown  to  produce  the  superimposition  of  male  sexual




characteristics on female neogastropod (stenoglossan) snails (Smith  1981b, GLbbs




and Bryan 1987).  This phenomenon,  termed "imposex," can  result in females with




a penis, a duct leading  to  the vas  deferens, and  a*  convolution of the normally



straight oviduct (Smith 1981a). Other anatomical changes associated with imposex




are detailed  in Gibbs et al. (1988)  and Gibbs  and  Bryan (1987).   Severity of




imposex is quantified using relative penis  size (RPS;  ratio of female to male




penis volume) and  the  six developmental stages of the vas deferens sequence (VDS)



(Bryan et al.  1986;  Gibbs et al. 1987).  TBT  has been shown to  impact populations



of  the Atlantic  dogwhinkle  (dogwhelk),  Nucella  lapillus.  which  has  direct




development.  In neoglossian snails  with indirect development  through planktonic




larval  stages, the  impacts  of  TBT  are less  certain because recruitment is



facilitated.  Natural  pseudohemaphiodism in neoglossans occurs (Salazar and Champ



1988)  and may be  caused by  other  organotin compounds  (Bryan et  al. 1988a) .



However, increased global incidence and severity of imposex has been associated



with  areas  of high boating activity  and high concentrations of TBT in water.




sediment or  snails  and other biota (Alvarez and Ellis 1990; Bailey and Davies




19,88a,1988b; Bryan et al. 1986,1987,;  Davies,et.al. 1987, Durchon 1982;  Ellis and




Pattisima 1990; Gibbs and Bryan 1986,1987;  Gibbs et al;.  1987; Langston et al.




1990; Short et al.  1989; Smith 1981a,1981b; Spence  et al.  1990).




                                       14

-------
      Although imposex has been observed in 45 species of snails worldwide (Ellis




and  Pattisia* 1990,  Jenner  1979),  definitive  laboratory and  field studies




implicating TBT as the cause have focused on three North American or cosmopolitan




species; the Atlantic dogvhinkle (Nucella  laoillus). file dogwhinkle  (JJ. lima)




and the eastern mud  snail  rilvanassa (Nassarius)  obsoletal.   Imposex has been




associated with reduced reproductive potential and altered density and population




structure in  field populations  of JJ. lapillus  (Spence  et  al.  1990).   This is




related to blockage  of  the oviduct by  the vas  deferens,  hence, prevention of




release of egg capsules-, sterilization of the female or change into an apparently




fuctional male (Bryan et  al.  1986; Gibbs et  al.  1987,1988; Gibbs  and Bryan




1986,1987).  TBT may reduce  populations of fi.  lima as snails were absent from




marinas in Auke Bay,  AK.  At intermediate distances  from marinas, about 25 were




caught per hour of sampling and 250 per hour were caught at sites distant from




marinas  (Short et  al.  1989).   Snails from  intermediate sites had blocked




oviducts.  Reduced proportions  of female X. obsoleta  in  Sarah Creek, VA also



suggests population  impacts (Bryan  et  al. 1989).   However,  other  causes  may




explain this as oviducts were not blocked and indirect development facilitating




recruitment may limit impacts.




      Several  field  studies have used transplantations of snails between sites




or snails painted with TBT  paints to  investigate the role of TBT  or proximity to



marinas  in  the  development  of  imposex without  defining  actual  exposure




concentrations of TBT.  Short et al. (1989)  painted Nucellus lima with TBT-based




paint,  copper paints or unpainted controls.   For  21  females  painted with TBT




paint, seven developed penises within one month,  whereas penises were absent from




35  females  from other  treatments.   Smith (1981a) transplanted I.  obsoletus




between marinas and  "clean"  locations and found that incidence of imposex was




unchanged after 19 weeks in snails  kept at clean locations  or marinas, increased




                                      15

-------
in snails  transplanted from clean sites  to  marinas  and decreased somewhat in




transplants from marinas  to clean sites.   Snails exposed in the laboratory to



TBT-based paints in two' separate' experiments  developed imposex within one month




with maximum  impact within 6 to 12 months (Smith 1981a).   Snails painted with




non-TBT paints were unaffected.



      Concentration-response data demonstrate  a similarity in the response of




snails to TBT in controlled laboratory and field studies  (Text Table 1) .  Eastern




mud snails, Illvanassa obsoleta.  collected from the York River, VA near Sarah



Creek had no incidence of  imposex  (Bryan et al.  1989) and contained no detectable



TBT,  (<0.020  ug/g  dry weight).   The  average TBT concentrations of York River




water was  0.0016 ug/L.  In contrast,  the average TBT concentrations from four



locations  in  Sarah Creek, VA were from  0.010  to 0.023 ug/L, snails contained




about 0.1  to  0.73  ug/g and there  was  a 40 to 100%  incidence  of  imposex.  Short




,et al. (1989) collected file dogwinkle snails,  Nucella  Hffia., from Auke Bay, AK




and did not detect imposex or TBT in snails from sites far from marinas.  Snails



from  locations  near marinas all  exhibited imposex and contained 0.03 to 0.16




ug/g.



      The  effects  of  TBT  on the development of imposex has been studied most in




the Atlantic  dogwhinkle,  Nucella lapillus.    Bryan  et  al.  1987 exposed adult




snails for two years  to  0.0036  (control), 0.0083,  0.046  and 0.26 ug/L in the




laboratory and compared responses to  a  field control.   Imposex  was present in




laboratory "control"  snails exposed  to  0.0036  ug/L and extent  of penis and vas



deferens  development  increased significantly  with  increase in TBT exposure;



sterility occurred in some  snails exposed to. 0.26 ug/L.  In a similar  laboratory




experiment that began with snail egg capsules and lasted two years (Gibbs  et al.




1988), imposex development was  more severe.  Field controls spawned and females




were  normal in 
-------








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 to 0.0036 ug/L were sterile and 160 egg cases were produced.  At >0.0093 ug/L all




 females  wer«  -sterile  with  only  two  undersized  egg  capsules   produced.



 Concentration* of  TBT  in females  were  0.19  ug/g  in  the  field,  0.58 ug/g  in the




 0.0036  ug/L treatment  and from  1.39  to  7.71 ug/g in  >0.0093 ug/L.   Similar




 concentrations of TBT (9.7 ug/g)  were found  in snails which became sterile after




 they  were  placed in the  Dart Estuary,  UK where  TBT concentrations range  from




 0.022 to 0.046 ug/L.   Gibbs and Bryan (1986) and  Gibbs  et al.  (1987)  report



 imposex  and reproductive failures at other marine sites where TBT concentrations




 in female  snails range from 0.32  to 1.54  ug/g.



       In summary,  in both field and laboratory studies, concentrations of  TBT  in



 water of about 0.001 ug/L or less and in tissues of about 0.2 ug/g or less  appear



 to not cause  imposex in fi. lapillus.  Imposex begins to occur,  and cause  some




-reproductive failure at about 0.004 ug/L with complete sterility occurring after




 x:hronic  exposure of sensitive early life-stages at >0.009  ug/L and  for  less




 sensitive  stages  at 0.02 ug/L in  some  studies  and greater  than 0.2 ug/L  in



.others.  If K-  laoillus or similarly sensitive species are ecologically important




.at  specific sites, TBT concentrations < 0..001  ug/L may  be required  to limit




 development of imposex.



       Reproductive abnormalities have also been observed  in the European  flat




 oyster (Thain  1986).  After exposure for 75 days  to  a TBT concentration of 0.24



 ug/L, a  retardation in the sex change from male to female was observed and larval




 production w*« completely inhibited.  A TBT concentration of 2.6 ug/L prevented




 .development of gonads.



       Survival and growth  of several  commercially  important  saltwater bivalve




 molluscs  have been  studied  during  acute  and  long-term exposures  to  TBT.




 Mortality of larval blue mussels. Mvcilus  edulis. exposed to 0.0973 ug/L was  51%;




 survivors were moribund and stunted (Beaumont and Budd  1984).  Similarly, Dixon




                                       18
-------
and Prosser (1986) observed 79% mortality of mussel larva after 4 days exposure




to 0.1 ug/L.  Growth of Juvenile blue mussels was  significantly reduced after.7




to 66 days  at 0.31  to  0.3893  ug/L (Stromgren and Bongard 1987;  Valkirs et al.




1985).  Growth rates of mussels  transplanted  into San Diego Harbor were impacted




at sites where TBT concentrations exceeded 0.2 ug/L (Salazar and Salazar 1990b).




At locations  where concentrations  were less  than 0.1 ug/L, the  presence of




optimum environmental conditions for growth appear to limit or mask the effects




of TBT.   Less  than optimum conditions for growth may permit the effect of TBT on




growth to be expressed.  Salazar et al.  (1987) observed that 0.157 ug/L reduced




growth of  mussels after 36 days exposure in  the  laboratory;  a concentration




within  less  than a factor of two  of that  reducing  growth  in the  field.




Similarly, Salazar and  SaLazar (1987) observed reduced growth of mussels exposed




to 0.070 ug/L for 196 days tn  the laboratory.  The 66-day LC50 for 2.5 to 4.1 cm




blue mussels  was  0.97  ug/L (Valkirs  et al.  1985,1987).   Alzieu et al. (1980)




reported 30% mortality  and abnormal shell thickening among Pacific oyster larvae




exposed  to  0.2  ug/L for 113 days.   Abnormal development was also  observed in




exposures  of  embryos for 24 hrs or less to  TBT  concentrations >  0.8604 ug/L




(Robert  and His 1981).  Waldock and Thain  (1983)  observed reduced growth and




thickening of the upper shell valve of Pacific oyster spat exposed to 0.1460 ug/L




for 56 days.  Shell thickening  in Crassostrea  gigas was associated with tissue



concentrations  of >0.2 ag/kg (Davies et al.  1988).   Abnormal shell development




was observed in an exposure to 0.77 ug/L that began with embryos of the eastern



oyster,  Craaaoatrea virginica.  and lasted for 48 hours (Roberts,  Manuscript).




Adult eastern oysters  were  also sensitive to TBT with reductions  in condition




index after exposure for 37 days to  > 0.1 ug/L (Henderson 1986; Valkirs et al.




1985).   Salazar et  al. .(1987)  found  no  effect  on growth after 56 days exposure




to 0.157 ug/L of  oysters £. virginica.  Ostrea  edulis and Q.  lurida.  Condition




                                      19
-------
of  adult  clams,  Macoma nasuta.  and scallops, Hinmites  multirufosus  were noc




affected after 110 days exposure to 0.204 ug/L (Salazar et al. 1987).




      Long-tent  exposures  have  been  conducted with  a  number  of  saltwater




crustacean species.  Johansen and Kohlenberg (1987)  exposed adult Acartia tonsa




for five days to TBT and observed impaired egg production on days  3. 4 and 5 in




0.1 ug/L and only on  day 5 in 0.01 and 0.05 ug/L.  For the five  days, overall egg



production  was reduced  markedly  (25%)  only in  0.1 ug/L.   Davidson  et al.



(1986a,1986b), Laughlin  et  al. (1983.1984b), and Salazar and Salazar  (1985a)



reported that TBT acts slowly on crustaceans and that behavior  might be affected




several days before  mortality occurs.   Survival of larval amphipods, Camflrys




oceanicus.  was significantly reduced after  eight, weeks  of  exposure  to TBT




concentrations > 0.2816 ug/L (Laughlin  et al.  1984b).   Hall et al.  (1988b)




observed no effect of 0.579 ug/L on fffmmfrns ;   after 24 days.  Developmental




rates and growth of larval mud crabs, Rhithrop.   oeus harrisii. were reduced by



a 15-day exposure  to > 14.60 ug/L.  R. harrisii might accumulate more TBT via



ingested food than directly from water (Evans and Laughlin 1984).  TBTF, TBTO,




and TBTS  were about equally  toxic  to amphipods and  crabs  (Laughlin  et al.



1982,1983,1984a).  Laughlin and French (1989) observed LC50 values for  larval




developmental stages of  13  ug/L  for crabs (£.  nauris)  from California vs 33.6




ug/L for  crabs from Florida.   Limb malformations  and  reduced burrowing were



observed  in fiddler crabs  exposed to 0.5  ug/L  (Weis  and Kim  1988;  Weis and



Perlnutter 1987).  Am regeneration was reduced in brittle stars exposed to 0.1



ug/L (Walsh «t al. 1986a).   Exposure to >0.1 ug/L during  settlement of  fouling



organisms  reduced number  of species  and  species  diversity of communities




(Henderson  1986).    The hierarchy  of sensitivities  of phyla  in  this test was




similar to that  of single species  tests.
                                       20
-------
      Exposure of embryos of the California grunion,  Leuresthes  tenuis. for ten



days to 74 ug/L caused a  50% reduction in hatching success (Newton et al. 1985).



At TBT concentrations between 0.14 and 1.72 ug/L,  growth, hatching success, and



survival were significantly enhanced.  In contrast, growth of inland silverside



larvae was reduced  after 28 days exposure  to  0.093  ug/L (Hall  et al. 1988b).



Juvenile Atlantic menhaden, Brevoortia tyrannus.  avoided a TBT concentration of



5,437 ug/L and juvenile striped bass, Morone saxatilis.  avoided  24.9 ug/L (Hall



et al. 1984).   BCFs were 4,300 for liver, 1,300 for brain,  and 200 for muscle



tissue of chinook salmon, Oncorhvnchus tshawvtscha. exposed to 1,490 ug/L for 96



hours (Short and Thrower  1986a,1986c).



      TBT concentrations  less  than the Final Chronic Value of 0.0485 Mg/L from



Table 3 have been shown to affect the  growth of early life-stages  of commercially



important bivalve mollusc.'   id survival of ecologically important  copepods (Table



6; Text  Table 2).   Surv..  1  of  the copepod Acartia  tonsa  was  significantly



reduced in three tests in 0.029, 0.023 and 0.024 Mg/L;  30, 27 and  51 percent of



control survival  (Bushong et al. 1990).   Survival decreased with increase in



exposure concentration but was not significantly affected in 0.012 Mg/L.



      Laughlin et al.  (1987,  1988) observed a significant decrease in growth of



hard claa (Mercenaria mercinaria) larvae  exposed for 14 days  to >0.01 Mg/L (Text



Table 2).  Growth rate  (increase in valve length) was 75% of controls in 0.01



Mg/L, 63% in 0.025 Mg/L,  59% in 0.05  Mg/L, *5% in 0.1 Mg/L, 29% in  0.25 Mg/L and



2.2% in 0.5 M8/L.  A  five-day  exposure followed by nine days in TBT-free water



produced similar responses and little evidence of recovery.



      Pacific oyster (Crassostrea gigas)  spat exhibited shell thickening in 0.01



and 0.05 Mg/L and reduced valve lengths  in >0.02  Mg/L (Lawler and  Aldrich 1987;



Text Table 2).   Increase in valve length was  101% of control  lengths in 0.01



      72%  in  0.02 Mg/L,  17% in 0.05  Mg/L, 35% in 0.1 Mg/L and 0% in 0.2 Mg/L.




                                      21
-------





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Shell chickening was also observed, in this species  exposed to >0.02 Mg/L for 49




days  (Thain  et  al.  1987).  They predicted  from these data that approximately




0.008  jig/L  would be  the  maximum  TBT  concentration  permitting culture 'of




commercially  acceptable  adults.  Their  field  studies  agreed  with laboratory



results showing "acceptable" shell thickness where TBT concentrations averaged




0.011 and  0.015  MgA but not ac higher  concentrations.   Decreased weights of



oyster meats  were associated with  locations where there was shell thickening.




Survival of  Crassostrea  gigas larvae exposed for  21  days was  reduced in 0.02



pg/L  (Springborn  Bionomics  1984a) .   No larvae survived  in >0.050 Mg/L-



      Growth  of spat of the European  oyster  (Ostrea edulis) was  reduced at >0.02




/ig/L  (Thain and Waldock 1985; Text Table 2).  Spit exposed  to TBT in static tests




were  82%  of  control  lengths  and 75%  of  control  weights; extent  of impact




increased  with  increased exposure.   In these static and  flow- through tests at




.exposures  at  about 0.02 /ig/L. weight gain was identical; i.e.,  35%  of  controls.



Growth of  larger  spat  was marginally reduced by 0.2392  pg/L (Thain 1986; Thain




and Ualdock  1985) .



      The  National Guidelines (Stephan et al. 1985; pp 18  and 54) requires  that




,the  criterion be lowered if sound scientific evidence indicates  that  adverse




effects might be  expected on important species.   The above data demonstrate that




reductions in growth occur in commercially or  ecologically important  saltwater




species at concentrations of TBT less than the Final Chronic Value of 0.0485 MgA



derived using  Final  Acute  Values   and  Acute-Chronic  Ratios  from  Table  3.



Therefore, EPA believes  the Final Chronic Value should be lowered to  0.01 /jg/L



to  limit  unacceptable  impacts  on  Acartia  tons a.  ttercen.arja. mercenaria.



Crassostrea  gifas and Ostrea  edulis observed at 0.02 ngfL.  At  this criteria




 concentration,  imposex would be expected in Uyanassa  obsojeta. NucelU
 and similarly sensitive neogastropods ;  populations  of N.. JjaiUsia and similarly




                                       24
-------
 sensitive  snails  with  direct development  might be  impacted  and growth  of


 Mercenaria mereenaria might be somewhat lowered.





 Unused Data


       Some data concerning the effects of TBT on aquatic organisms were not used


 because the  tests were  conducted with species that are not  resident  in North


 America (e.g.,  Allen et ill.  1980; Carney and Paulini  1964; Danil'chenko 1982;


 Deschiens  and Floch 1968; Deschiens et al. 1964,1966a, 1966b;  de Sous a and Paulini


 1970;  Frick and DeJimenez 1964; Hopf and Muller 1962; Kubo et al. 1984; Nishuichi


 and Yoshida  1972; Ritchie et  al.  1964;  Seiffer and Schoof 1967; Shiff  et al.


 1975;  Smith et  al.  1979;  Tsuda  et al.  1986;  Upatham 1975;  Upatham  et  al.


 1980a,1980b; Vebbe and Sturrock 1964).


       Alzieu  (1986),  Cardarelli and Evans (1980),  Cardwell  and Sheldon (1986),


 Cardwell and Vogue (1986), Champ (1986), Chau (1986), Eisler  (1989), Envirosphere


.Company (1986), Gibbs and Bryan (1987), Good et al. (1980),  Guard et al. (1982).


 Hall (1988), Hall and Pinkney  (1985), Hodge et al. (1979),  International Joint


 Commission (1976), Jensen (1977), Kimbrough.(1976), Kumpulainen and Koivistoinen


 (1977),  Laughlin (1986), Laughlin and Linden  (1985),  Laughliri  et al.  (1984a),


 McCullough et al. (1980), Mcmaghan et al. (1980),  North Carolina Department of


 Natural Resources and Community Development (1983,1985) , Rexrode (1987) , Seligman


 et al. (1986),  Slesinger and  Dressier  (1978),  Stebbing (1985),  Thayer (1984),


 Thompson et »1.  (1985),  U.S. EPA (1975,1985b), U.S. Navy (1984), Valkirs et al.
                      •«r

 (1985),  von Ruaker  «t al.   (1974), Walsh (1986) and  Zuckerman et  al.  (1978)


 compiled.data from other sources.   Studies by Gibbs et al.  (1987) were not used


 because data were from the first year of a two-year experiment reported in Gibbs


 et al. (1988).                             .
                                       25
-------
      Results were not used when the test procedures, test material,  or  results




were not adequately  described (e.g.,  Bruno  and Ellis 1988;  Cardwell  and Stuart




1988;  Chau «t  al.  1983;, Danil'chenko  and Buzinova 1982; de  la Court  1980;




Deschiens  1968;  EG&G Bionomics 1981b;  Filenko and Isakova 1980; Holwerda and



Herwig  1986; Kelly et  al.  1990; Kolosova  et al.  1980;  Laughlin  1983;  Lee  1985;



Nosov and  Kolosova  1979;  Smith 1981c; Stroganov et  al. 1972,1977).   The  96-hr




LC50 of 0.01466  /ig/L reported by  Becerra-Huencho (1984) for post larvae of the



hard clam, Mercenaria mercenaria. was not used because results of other  studies




with embryos,  larvae,  and post  larvae  of  the  hard clam  where  acutely lethal




concentrations  range from 0.6 to 4.0 pg/L  (Tables 1 and  6) cast doubt  on this




LC50 value.  Data from the life-cycle test with sheepshead minnows (Ward et al..




1981) were not used  because  ratios  of measured and nominal concentrations were




•inconsistent within and between  tests  suggesting problems in  delivering TBT,




•analytical chemistry or both.  Results of  some  laboratory  tests were not used



because the tests were  conducted in distilled or deionized water without addition




•of  appropriate  salts (e.g.,  Gras  and Rioux  1965; Kumar Das et al. 1984).  The




concentration of dissolved oxygen was too low in tests reported by EG&G Bionomics




(1981a).   Douglas  et al.  (1986)  did  not observe  sufficient  mortalities  to




calculate  a useful LC50.



      Data were not used when TBT  was  a component  of a  formulation, mixture,



paint,  or sediment (Boike and Rathburn 1973; Cardarelli 1978; Deschiens and Floch



1970; Goss et al. 1979; Laughlin et al.  1982; Maguire and Tkacz 1985;  Mattiessen



and Thain 1989; North  Carolina  Department of Natural Resources and Community



Development  1983; Pope  1981; Quick  and  Cardarelli 1977; Salazar and  Sal.azar




1985a,'.1985b;  Santos  et al.  1977;  Sherman 1983;. Sherman and Hoang 1981;  Sherman



and Jackson 1981; Walker 1977; Weisfeld  1970), unless data  were available to show




that the  toxicity was  the same as for TBT alone.




                                       26
-------
      Data were not used  when the test organisms were infested with tapeworms



(e.g., Hnath 1970).  Mottley  (1978) and Mottley and Griffiths  (1977) conducted




tests with a Mutant fora of an alga.  Results of tests  in which  enzymes, excised




or homogenized tissue, or cell cultures were exposed to  the test material were




not used (e.g.,  Blair  et al. 1982; Josephson et al. 1989).  Tests conducted with



too few test organisms  were  not used  (e.g., EG&G Bionomics  1976;  Good et al.




1979).   High control mortalities occurred in  tests  reported by  Salazar and



Salazar (Manuscript) and Valkirs et al. (1985).  Some data were  not used because




of  problems with  the concentration  of the  test material  (e.g.,  Springbom




Bionomics 1984b; Stephens on et al. 1986;  Ward et al. 1981). BCFs were not used




when the concentration of  TBT  in  the test solution was  not measured (Laiighlin et




al. 1986b;  Paul and Davies 1986) or were highly variable  (Laughlin and French




1988) .  Reports of the concentrations in wild aquatic animals were not used if




concentrations in water were  unavailable or excessively variable (Davies et al.




1987; Davies and McKie 1987;  Hall 1988; Han and Weber 1988; Wade et al. 1988.
      The acute toxicity values for thirteen freshwater animal species range from



1.14 ug/L  for a hydra  (Hxi£a.  oltgactis)  to 24,600 ug/L  for a clam (Elliptic



c enrol ana tus^ .  There was no apparent trend in sensitivities with taxonomy; fish



were nearly as sensitive JLS the most sensitive invertebrates  and more sensitive



than others.  Vhen the much less sensitive  clam was not considered,  the remaining



species sensitivities varied by a maximum of 8.7 times.  Three chronic toxicity



tests have  been conducted with   freshwater animals.   Reproduction of Daohnia



magna was reduced by 0.2 ug/L,  but not by  0.1 ug/L,  and the Acute - Chronic Ratio



is 30.41.   In another test with £_.  magna  reproduction and survival was reduced



at 0.34 ug/L  but not at 0.19, and the  Acute - Chronic Ratio is 44.06.  Weight of




                                       27
-------
fathead minnows was reduced by 0.-45 ug/L, but not by 0.15 ug/L, and the acute-




chronic ratio for this species was 10.01.  Bioconcentration of TUT was measured




in rainbow trout, Oncorhvnchus invkiss.  at 406 times  the water concentration for




the whole body.  Growth of thirteen species of freshwater algae was inhibited by




concentrations ranging from 56.1  to 1,782 ug/L.



       Acute values for 27 species of saltwater animals range from 0.61 ug/L for




the nysid, Aeanthotnvsis sculpta.  to 204.4 ug/L for adult  European flat oysters,



Qstrea edulis.   Acute values for the  twelve most  sensitive genera, including




molluscs, crustaceans, and fishes, differ by less than a factor of 4.  Larvae and



Juveniles appear to be more sensitive  than adults.  A life-cycle toxicity test



has been conducted with the  saltwater mysid,  Acanthomysis sculpta.   The chronic




value for A.,  sculota was 0.1308 ug/L based on  reduced reproduction and the acute-



chronic ratio was 4.664.  Bioconcentration factors  for  three  species of bivalve




molluscs range from 192.3 for soft parts  of  the European flat oyster to 11.400




for soft parts of the Pacific oyster,  Crassostrea sigas..  Tributyltin  chronically



affects certain saltwater copepods,  gastropods, and pelecypods at concentrations



less  than those predicted  from  "standard"   acute and  chronic  toxicity tests.




Survival of  the  copepod Acartia  tonsa was reduced  in  £0.023 A»g/L.   Growth of




larvae or spat of two species of oysters,  Crassostrea fci^ai and Qstrea edujjg was



reduced in about 0.02 j*g/L;  some £. glgas larvae died in 0.025  j*g/L.  Generally



concentrations <0.01  Mg/L have not  been  demonstrated to  affect  sensitive  life-



stages of saltwater organisms.   This  above data demonstrate  that reductions in




growth  occur  in coaaercially or ecologically  important saltwater  species at




concentrations of  TBT less  than  the Final Chronic Value  of  0.0485 /*g/L derived




using Final  Acute Values and Acute-Chronic"Ratios from Table 3.   Therefore, EPA




believes, the  Final' Chronic  Value  should be lowered  to 0.01  /ig/L to  limit



unacceptable impacts  on Acartia  tonsa. Mercenaria mercenaria. Cr??$ostrea gjga?




                                       28
-------
and Ostrea edulls observed at 0.02 /ig/L. At this criteria concentration, imposex



would be expected in Ilyanassa obsoleta. Nucella laoillus and similarly sensitive



neogastropods; populations of fi.  lauillus and similarly sensitive snails with



direct development might be impacted and  growth of Mereenaria mercenaria might



be somewhat lowered.








National Criteria




      The procedures described in the "Guidelines for Deriving Numerical National



Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses"



indicate  that,  except  possibly  where a locally important species  is  very



sensitive, freshwater aquatic organisms and their uses should not be affected



unacceptably if the four-clay average concentration of tributyltin does not exceed



0.063 Mg/L more than once every three  years on the average and if the one-hour



average concentration does not exceed 0.46 Mg/L more than once every three years



on the average.




      The procedures described in the "Guidelines for Deriving Numerical National



Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses"



indicate  that,  except  possibly  where a locally important species  is  very



sensitive, saltwater  aquatic  organisms and their uses  should  not be affected



unacceptably if the four-clay average concentration of tributyltin does not exceed



0.010 /ig/L more than oncci every three  years on the average and if the one-hour



average concentration does not exceed 0.36 pg/L more than once every three years



on the average.
                                      29
-------
Implementation




      As discussed in the Water Quality Standards Regulation (U.S.  EPA 1983a) and



the Foreword of this docunent,  a  water quality criterion for aquatic  life has




regulatory impact only if  it has been adopted in a  state water quality standard.




Such a standard specifies  a criterion for  a pollutant that is consistent with a




particular  designated use.   With  the concurrence  of  the  U.S. EPA,  states



designate one or more uses for each body of water or segment thereof and adopt



criteria that are  consistent  with the use(s)  (U.S. EPA  1983b,1987).   In each




standard a  state  may adopt  the  national criterion,  if one  exists,  or,  if




adequately justified,  a site-specific criterion.    (If the  site  is an entire



state, the site-specific criterion is also a s'tate-specific criterion.)



      Site-specific  criteria  may include  not only  site-specific  criterion



concentrations  (U.S.   EPA 1983b),   but   also  site-specific,   and  possibly




pollutant-specific, durations  of  averaging periods  and frequencies  of allowed




excursions (U.S. EPA 1991). The averaging periods of "one hour" and "four days"



were selected by  the  U.S.  EPA on the basis  of data concerning  the  speed with



which some aquatic species can react  to increases  in the  concentrations of some




aquatic pollutants, and "three years" is the Agency's best scientific judgment



of  the  average  amount  of  time aquatic ecosystems  should be  provided between




excursions (Stephan et al. 1985; U.S. EPA 1991).  However, various species and




ecosystems react and recover at greatly differing rates.   Therefore, if adequate



justification    i«   provided,   site-specific    and/or   pollutant-specific



concentration*, durations, and frequencies may be  higher or  lower  than those




given in national  water quality criteria for aquatic life.



      Use of criteria, which have  been adopted  in state water quality standards,




for  developing water  quality-based  permit  limits and for designing waste




treatment facilities requires  selection of an  appropriate wasteload allocation




                                       30
-------
model.   Although dynamic models  are, preferred  for the  application  of these



criteria (U.S. EPA 1991), limited data or other considerations might require the



use of a steady-state model (U.S. EPA 1986).



      Guidance on  mixing xones and the design of  monitoring programs is also



available (U.S. EPA 1985b).
                                       31
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