&EPA
United States
Environmental Protection
Agency
Office of Water
4304T
EPA-822-B-02-001
December 2002
Ambient Aquatic Life
Water Quality Criteria
for Tributyltin (TBT) - Draft
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ADDENDUM
The Environmental Protection Agency (EPA) issued a draft ambient water quality criteria
document for tributyltin (TBT) on August 7, 1997. This document was issued to the public for
scientific and technical input through a notice of availability published in the Federal Register
(62 FR 42554). After consideration of peer review, scientific and technical input, and additional
data which became available after the draft was published, EPA is issuing new ambient water
quality criteria for TBT for scientific and technical input.
A comprehensive literature search for toxicity information on TBT was conducted before the
draft criteria document for TBT was issued in 1997. In preparing the new TBT criteria
document, more recent aquatic life toxicity data have been considered (* see new references at
end of Addendum). The major effect of inclusion of this new information on TBT is the
lowering of the draft saltwater four day average, once in three year exceedence, chronic criterion
of 0.01 ug/1 to a new chronic criterion of 0.001 ug/1.
EPA's Office of Pesticide Programs (OPP) has recently updated its Environmental Risk
Characterization for TBT. EPA's Office of Water (OW) has coordinated closely with OPP in
preparing the new ambient water quality criteria document for TBT. This collaboration has
enabled OW to access more recent information on the toxicity of TBT. Consideration of the
more recent data available for TBT leads to the following conclusions:
* TBT is an immunosuppressing agent and an endocrine disrupter
* TBT biomagnifies through the food chain and has been found in tissues of marine mammals
* TBT causes adverse reproductive and developmental effects in aquatic organisms at very low
concentrations
* TBT degrades much more slowly in sediment than earlier studies had indicated and is likely
to persist in sediments at concentrations which cause adverse biological effects
After considering peer review, scientific and technical input from the public, and more
recent data, EPA has set the new saltwater chronic criterion for TBT at 0.001 ug/1.
References:
* Fisher, W.S., L.M. Oliver, W.W. Walker, C.S. Manning, T.F. Lytle. 1999. Decreased
resistance of eastern oysters (Crassostrea Virginicd) to a protozoan pathogen (Perkinsus
marinus) after sublethal exposure to tributyltin oxide. Marine Environ. Res. 47: 185-201.
* Matthiessen, P. and P.E. Gibbs. 1998. Critical appraisal of the evidence for tributyltin-
mediated endocrine disruption in mollusks. Environ. Toxicol. Chem. 17: 37-43.
* Other references are available, but were not relied upon for derivation of the criteria
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EXECUTIVE SUMMARY
BACKGROUND:
Tributyltin (TBT)is a highly toxic biocide that has been used extensively
to protect the hulls of large ships. It is a problem in the aquatic
environment because it is extremely toxic to non-target organisms, is linked
to imposex and immuno-supression in snails and bivalves, and is very
persistent. EPA is developing ambient water quality criteria for TBT through
its authority under Section 304(a) of the Clean Water Act (CWA). These water
quality criteria may be used by States and Tribes to establish water quality
standards for TBT.
CRITERIA:
Freshwater:
For TBT, the criterion to protect freshwater aquatic life from chronic
toxic effects is 0.063 ug/L. This criterion is implemented as a four-day
average, not to be exceeded more than once every three years on the average.
The criterion to protect freshwater aquatic life from acute toxic effects is
0.46 ug/L. This criterion is implemented as a one-hour average, not to be
exceeded more than once every three years on the average.
Saltwater:
For TBT, the criterion to protect saltwater aquatic life from chronic
toxic effects is 0.001 ug/L. This criterion is implemented as a four-day
average, not to be exceeded more than once every three years on the average.
The criterion to protect saltwater aquatic life from acute toxic effects is
0.38 ug/L. This criterion is implemented as a one-hour average, not to be
exceeded more than once every three years on the average.
The saltwater chronic criterion for TBT differs significantly from the
criterion that was originally proposed for public review (0.010 ug/L). The
development of the saltwater chronic criterion for TBT considers four lines of
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evidence:
(1) the traditional endpoints of adverse effects on survival, growth, and
reproduction as demonstrated in numerous laboratory studies;
(2) the endocrine disrupting capability of TBT as observed in the
production of imposex in field studies
(3) that TBT bioaccumulates in commercially and recreationally important
freshwater and saltwater species
(4) that an important commercial organism already known to be vulnerable to
a prevalent pathogen was made even more vulnerable by prior exposure to TBT.
For these reasons, the criterion to protect saltwater aquatic life from
chronic toxic effects is set at 0.001 ug/L.
This document provides guidance to States and Tribes authorized to
establish water quality standards under the Clean Water Act (CWA) to protect
aquatic life from acute and chronic effects of TBT. Under the CWA, States and
Tribes are to establish water quality criteria to protect designated uses.
While this document constitutes U.S. EPA's scientific recommendations
regarding ambient concentrations of TBT, this document does not substitute for
the CWA or U.S. EPA's regulations; nor is it a regulation itself. Thus, it
cannot impose legally binding requirements on U.S. EPA, States, Tribes, or the
regulated community, and it might not apply to a particular situation based
upon the circumstances. State and Tribal decision-makers retain the
discretion to adopt approaches on a case-by-case basis that differ from this
guidance when appropriate. U.S. EPA may change this guidance in the future.
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AMBIENT AQUATIC LIFE WATER QUALITY CRITERIA FOR
TRIBUTYLTIN
CAS Registry Number (See Text)
U.S. ENVIRONMENTAL PROTECTION AGENCY
OFFICE OF WATER
OFFICE OF SCIENCE AND TECHNOLOGY
HEALTH AND ECOLOGICAL CRITERIA DIVISION
WASHINGTON D.C.
OFFICE OF RESEARCH AND DEVELOPMENT
MID-CONTINENT ECOLOGY DIVISION
DULUTH, MINNESOTA
ATLANTIC ECOLOGY DIVISION
NARRAGANSETT, RHODE ISLAND
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NOTICES
This document has been reviewed by the Health and Ecological Criteria Division
(HECD), Office of Science and Technology, Office of Water, U.S. Environmental
Protection Agency, and approved for publication.
Mention of trade names or commercial products does not constitute endorsement
or recommendation for use.
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FOREWORD
Under Section 304(a)of the Clean Water Act (CWA) of 1977 (P.L. 95-217), the
U.S. Environmental Protection Agency (EPA)is to periodically revise water
quality criteria to accurately reflect the latest scientific knowledge. This
document is a revision of previous criteria based upon consideration of
scientific and technical input received from other federal agencies, state
agencies, special interest groups, and individual scientists. Criteria
contained in this document replace any previously published U.S. EPA aquatic
life criteria for tributyltin (TBT).
This document provides guidance to States and Tribes authorized to establish
water quality standards under the CWA to protect aquatic life from toxic
effects of TBT. Under the CWA, States and Tribes are to establish water
quality standards to protect designated uses. While this document constitutes
the U.S. EPA's scientific recommendations regarding ambient concentrations of
TBT, this document does not substitute for the CWA or the U.S. EPA's
regulations, nor is it a regulation itself. Thus, it cannot impose legally
binding requirements on the U.S. EPA, States, Tribes or the regulated
community, and might not apply to a particular situation based upon the
circumstances. The U.S. EPA may change this guidance in the future.
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ACKNOWLEDGMENTS
Great Lakes Environmental Center (GLEC), Traverse City, MI produced this
document under U.S. EPA Contract Number 68-C6-0038, Work Assignment B-04.
people listed on this page contributed to this document in the stated
capacities.
The
AUTHORS
Larry T. Brooke (freshwater)
University of Wisconsin-Superior
Superior, Wisconsin and Great
Lakes Environmental Center,
Traverse City, Michigan
David J. Hansen (saltwater)
U.S. EPA
Atlantic Ecology Division
Office of Research and Development
Narragansett, Rhode Island and
Great Lakes Environmental Center,
Traverse City, Michigan
Frank Gostomski (document coordinator)
U.S. EPA
Health and Ecological Criteria Division
Office of Science and Technology
Office of Water
Washington, DC
Great Lakes Environmental Center, Inc. Work Assignment Leader: Dennis McCauley
TECHNICAL ASSISTANCE AND PEER REVIEW
Herbert E. Allen
University of Delaware
Newark, Delaware
Peter M. Chapman
EVS Environmental Consultants
North Vancouver, British Columbia
Michael H. Salazar
Applied Biomonitoring
Kirkland, Washington
Robert L. Spehar
U.S. EPA
Mid-Continent Ecology Division
Duluth, Minnesota
Rick D. Cardwell
Parametrix, Inc.,
Redmond, Washington
Lenwood W. Hall
University of Maryland
Queenstown, Maryland
Peter F. Seligman
U.S. Navy SSC SD
San Diego, California
Glen Thursby
U.S. EPA
Atlantic Ecology Division
Narragansett, Rhode Island
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CONTENTS
Page
Foreword iii
Acknowledgments iv
Tables vii
Text Tables viii
Introduction 1
Acute Toxicity to Aquatic Animals 7
Chronic Toxicity to Aquatic Animals 9
Toxicity to Aquatic Plants 13
Bioaccumulation 14
Other Data 15
Unused Data 34
Summary 37
National Criteria 43
Implementation 43
References 89
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TABLES
Page
1. Acute Toxicity of Tributyltin to Aquatic Animals 44-A
2. Chronic Toxicity of Tributyltin to Aquatic Animals 51
3. Ranked Genus Mean Acute Values with Species Mean Acute-Chronic
Ratios 53
4. Toxicity of Tributyltin to Aquatic Plants 57
5. Bioaccumulation of Tributyltin by Aquatic Organisms 59
6. Other Data on Effects of Tributyltin on Aquatic Organisms 63
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TEXT TABLES
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 = RPSI) and the vas deferens
sequence index (VDSI), as a function of tributyltin
concentration in water and dry tissue 25
2. Summary of laboratory and field data on the effects
of tributyltin on saltwater organisms at concentrations
less than the Final Chronic Value of 0.0605 ug/L 33
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Introduction
Organotins are compounds consisting of one to four organic components
attached to a tin atom via carbon-tin covalent bonds. When there are fewer
than four carbon-tin bonds, the organotin cation can combine with 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
(C4H9)3Sn+. 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) .
The principal use of organotins is as a stabilizer in the manufacturing
of plastic products, for example, as an anti-yellowing agent in clear plastics
and as a catalyst in poly(vinyl chloride) products (Piver 1973). Another and
less extensive use of organotins is as a biocide (fungicide, bactericide,
insecticide) and as a preservative for wood, textiles, paper, leather and
electrical equipment. Total world-wide production of organotin compounds is
estimated at 50,000 tons per year with between 15 and 20% of the production
used in the biologically active triorganotins (Bennett 1996).
A large market exists for organotins in antifouling paint for the wet
bottom of ship hulls. The most common organometallics used in these paints
are TBT oxide and TBT methacrylate. Protection from fouling with these paints
lasts more than two years and is superior to copper- and mercury-based paints.
These paints have an additional advantage over other antifouling paints, such
as copper sulfate based paint, by not promoting bimetallic corrosion. The
earliest paints containing TBT were "free association" paints that contained a
free suspension of TBT and caused high concentrations of TBT to be leached to
the aquatic environment when the paint application was new. A later
refinement was the "ablative" paint that shed the outer layer when in contact
with water but at a slower rate than the free association paint. Further
development of organometallic antifouling paints have been in the production
of paints containing copolymers that control the release of the organotins and
result in longer useful life of the paint as an antifoulant (Bennett 1996;
Champ and Seligman 1996; Kirk-Othmer 1981). The U.S. Navy (1984) proposed
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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
on ships' hulls by these organisms 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. Interaction between the toxicities of TBT and other ingredients in
the paint apparently is negligible, but needs further study (Davidson et al.
1986a). The use of TBT in antifouling paints on ships, boats, nets, docks,
and water cooling towers probably contributes most to direct release of
organotins into the aquatic environment (Clark et al. 1988; Hall and Pinkney
1985; Kinnetic Laboratory 1984) .
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 at pH of 6.6, 31,000 ug/L at pH of 8.1 and 30,000 ug/L at pH
2.6 (Maguire et al. 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 months or more than a year in
sediments (Clark et al. 1988; de Mora et al. 1989; Lee et al. 1987; Maguire
and Tkacz 1985; Seligman et al. 1986, 1988, 1989; Stang and Seligman 1986).
Breakdown products include di-, monobutyltins and tin with some methyltins
detected (Yonezawa et al. 1994) when sulfate reducing conditions were present.
Porous sediments with aerobic conditions decrease degradation time (Watanabe
et al. 1995) .
Several review papers have been written which cover the production,
use, chemistry, toxicity, fate and hazards of TBT in the aquatic environment
(Alzieu 1996; Batley 1996; Clark et al. 1988; Eisler 1989; Gibbs and Bryan
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1996b; Hall and Bushong 1996; Laughlin 1996; Laughlin et al. 1996; Maguire
1996; Waldock et al. 1996; WHO 1990). The toxicities of organotin compounds
are related to the number of organic components bonded to the tin atom and to
the number of carbon atoms in the organic components. Toxicity to aquatic
organisms generally increases as the number of organic components 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 Na+ and K+ ATPases and are ionophores controlling exchange of Cl~,
Br", F" and other ions across cell membranes (Selwyn 1976).
Metabolism of TBT has been studied in several species. 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 and demethylated 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 an eel
grass, Zostera marina (Francios et al. 1989) . Chiles et al. (1989) found that
much of the TBT accumulated on the surface of saltwater algae and bacteria as
well as within the cell. The major metabolite of TBT in saltwater crabs,
fish, and shrimp was dibutyltin (Lee 1985, 1986) . A review of the metabolism
of TBT by marine aquatic organisms has been provided by Lee (1996).
TBT is an endocrine-disrupting chemical (Matthiessen and Gibbs 1998) .
The chemical causes masculinization of certain female gastropods. It is
likely the best studied example of endocrine-disrupting effect. The metabolic
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mechanism is thought to be due to elevating testosterone titers in the animals
and over-riding the effects of estrogen. There are several theories of how
TBT accomplishes the buildup of testosterone and evidence suggests that
competitive inhibition of cytochrome P450-dependent aromatase is probably
occurring in TBT exposed gastropods (Matthiessen and Gibbs 1998). TBT may
interfer with sulfur conjugation of testosterone and its phase I metabolites
and their excretion resulting in a build-up of pharmacologically active
androgens in some animal tissues (Ronis and Mason 1996).
TBT has been measured in the water column and found highly (70-90%)
associated with the dissolved phase (Johnson et al. 1987; Maguire 1986;
Valkirs et al. 1986a). However, TBT readily sorbs to sediments and suspended
solids and can persist there (Cardarelli and Evans 1980; Harris et al. 1996;
Seligman et al. 1996). TBT accumulates in sediments with sorption
coefficients which range from l.lxlO2 to 8.2xl03 L/Kg and desorption appears
to be a two step process (Unger et al. 1987,1988). At environmentally
realistic concentrations of 10 ng/L, TBT partitioning coefficients were closer
to 2.5 xlO4 (Langston and Pope, 1995). In a modeling and risk assessment
study of TBT in a freshwater lake, Traas et al. (1996) predicted that TBT
concentrations in the water and suspended matter would decrease rapidly and
TBT concentrations in sediment and benthic organisms would decrease at a much
slower rate.
The water surface microlayer contains a much higher concentration of
TBT than the water column (deary and Stebbing 1987; Hall et al. 1986; Maguire
1986; Valkirs et al. 1986a). Gucinski (1986) suggested that this enrichment
of the surface microlayer could increase the bioavailability of TBT to
organisms in contact with this layer.
Elevated TBT concentrations in fresh and salt waters, sediments, and
biota are primarily associated with harbors and marinas (Cleary and Stebbing
1985; Espourteille et al. 1993; Gibbs and Bryan 1996a; Grovhoug et al. 1996;
Hall 1988; Hall et al. 1986; Langston et al. 1987; Maguire 1984,1986; Maguire
and Tkacz 1985; Maguire et al. 1982; Minchin and Minchin 1997; Peven et al.
1996; Prouse and Ellis 1997; Quevauviller et al. 1989; Salazar and Salazar
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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; Waite et al.
1996; Waldock and Miller 1983; Waldock et al. 1987). Several studies have
been conducted in harbors to measure the effects of TBT on biota. 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. No organotins
were detected in the muscle tissue of feral Chinook salmon, Oncorhynchus
tshawytscha, caught near Auke Bay, Alaska, but concentrations as high as 900
ug/kg were reported in muscle tissue of Chinook salmon held in shallow-water
pens treated with TBT (Short 1987; Short and Thrower 1986a). Organotin
concentrations in European coastal waters in the low part per trillion range
have been associated with oyster shell malformations (Alzieu et al. 1989;
Minchin et al. 1987). Reevaluation of harbors in the United Kingdom revealed
that since the 1987 restrictions which banned the retail sale and use of TBT
paints for small boats or mariculture purposes, oyster culture has returned in
the harbor areas where boat traffic is low and water exchange is good (Dyrynda
1992; Evans et al. 1996; Minchin et al. 1996,1997; Page and Widdows 1991;
Waite et al. 1991). Tissue concentrations of TBT in oysters have decreased in
most of the sites sampled in the Gulf of Mexico since the introduction of
restrictions (1988-1989) on its use (Wade et al. 1991). Canada restricted the
use of TBT-containing boat-hull paints in 1989 and there has been a reduction
in female snail reproductive deformities (imposex) in many Canadian west coast
sampling sites (Tester et al. 1996). In a four-year (1987-1990) monitoring
study for butyltins in mussel tissue on the two U.S. coasts, a general
decrease in tissue concentrations was measured on the west coast and east
coast sites showed mixed responses (Uhler et al. 1989,1993). Some small ports
in France have not seen a decline in imposex since the ban on TBT in boat hull
paints (Huet et al. 1996). Suspicions are that the legislation banning the
paints is being ignored. Several freshwater ecosystems were studied since the
ban on antifouling paints in Switzerland in 1988. By 1993 TBT concentrations
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were decreasing in the water, but declines were not seen in the sediment or in
the zebra mussel, Dreissena polymorpha (Becker-van Slooten and Tarradellas
1995; Pent and Hunn 1995).
Because of the assumption that certain anions do not contribute to TBT
toxicity, only data generated in toxicity and bioconcentration tests on TBTC1
(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. The conversion factors are 0.8911 for TBTC1, 0.9385 for
TBTF, 0.9477 for TBTO, 0.9005 for TBTS, and 2.444 for Tin (Sn). 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 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 require 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 January
1997 for fresh- and saltwater organisms. Some more recent data have been
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included in the document.
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. For
freshwater Species Mean Acute Values, 23% were <2.0 ug/L, 38% were <4.0 ug/L,
69% were <8.0 ug/L, and 92% were <12.73 ug/L. A freshwater clam, Elliptio
complanatus, had an LC50 of 24,600 U9/L. The relatively low sensitivity of
the freshwater clam 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 temporarily tolerate high concentrations of
TBT.
The most sensitive freshwater organisms tested are from the phylum
Coelenterata (Table 3). Three species of hydras were tested and have Species
Mean Acute Values (SMAVs) ranging from 1.14 to 1.80 U9/L. Other invertebrate
species tested in flow-through measured tests include an amphipod, Gammarus
pseudolimnaeus, and an annelid, Lumbriculus variegatus, and in a static
measured test, larvae of a mosquito, Culex sp (Brooke et al. 1986). The 96-hr
LCSOs and SMAVs are 3.7, 5.4 and 10.2 ug/L, respectively. Six tests were
conducted with the cladoceran, Daphnia magna,. The 48-hr EC50 value of 66.3
ug/L (Foster 1981) was considerably less sensitive than those from the other
tests which ranged from 1.58 U9/L (LeBlanc 1976) to 18 U9/L (Crisinel et al.
1994). The SMAV for D. magna is 4.3 U9/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. The freshwater clam, Elliptio camplanatus, had an unusually high LC50
value of 24,600 ug/L-
All the vertebrate species tested are fish. The most sensitive species
is the fathead minnow, Pimephales promelas, which has a SMAV of 2.6 U9/L from
a single 96-hr flow-through measured test (Brooke et al. 1986) . Rainbow
trout, Oncorhynchus mykiss, were tested by four groups with good agreement
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between them. 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; Martin et al. 1989; ABC
Laboratories, Inc. 1990a), which were conducted using flow-through conditions
and measured concentrations. Juvenile catfish, Ictalurus punctatus, were
exposed to TBT in a flow-through and measured concentration test and resulted
in a 96-hr LC50 of 5.5 ug/L which is in good agreement with the other tested
freshwater fish species. 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 (ABC Laboratories, Inc. 1990b) 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, Elliptio, the remainder
differ from one another by a maximum factor of 11.2 times. Based upon the
twelve available GMAVs the Final Acute Value (FAV) for freshwater organisms is
0.9177 ug/L. The FAV is lower than the lowest freshwater SMAV of 1.14 ug/L.
The freshwater Criterion Maximum Concentration is 0.4589 U9/L which is
calculated by dividing the FAV by two.
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 26 species of invertebrates and seven
species of fish (Table 1). The range of acute toxicity to saltwater animals
is a factor of about 1,176. Acute values range from 0.24 ug/L for juveniles
of the copepod, Acatia tonsa (Kusk and Petersen 1997) to 282.2 ug/L for adult
Pacific oysters, Crassostrea gigas (Thain 1983). The 96-hr LCSOs for six
saltwater fish species range from 1.460 U9/L for juvenile Chinook salmon,
Oncorhynchus tshawytscha (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, Crassostrea gigas, was 1.557 ug/L, whereas the value for
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adults was 282.2 ug/L (Thain 1983). The 96-hr LCSOs for larval and adult blue
mussels, Mytilus edulis, were 2.238 and 36.98 ug/L, respectively (Thain 1983).
The 96-hr LC50 of 0.01466 ug/L reported by Becerra-Huencho (1984) for post
larvae of the hard clam, M. 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 ug/L (Tables 1 and 6) cast doubt
on this LC50 value. Juveniles of the crustaceans Acanthomysis sculpta and
Metamysidopsis elongata were slightly more sensitive to TBT than adults
(Davidson et al. 1986a,1986b; Valkirs et al. 1985; Salazar and Salazar 1989).
Four genera of amphipods were tested and sensitivity to TBT ranged from 1.3 to
22.8 ug/L. As with bivalve molluscs and other crustaceans, one genus
(Gammarus) demonstrated greater sensitivity to TBT at the younger life-stage
(Bushong et al. 1988).
Genus Mean Acute Values for 30 saltwater genera range from 0.61 ug/L
for Acanthomysis to 204.4 ug/L for Ostrea (Table 3). Genus Mean Acute Values
for the 12 most sensitive genera differ by a factor of less than four.
Included within these genera are four species of molluscs, eight species of
crustaceans, and one species of fish. The saltwater Final Acute Value (FAV)
for TBT was calculated to be 0.7673 ug/L (Table 3), which is greater than the
lowest saltwater Species Mean Acute Value of 0.61 ug/L. The saltwater
Criterion Maximum Concentation is 0.3836 ug/L and is calculated by dividing
the FAV by two.
Chronic Toxicitv to Aquatic Animals
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 D. 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 reproductive effects
on adult daphnids. The chronic value for D. magna is 0.1414 ug/L (geometric
mean of the chronic limits), and the acute-chronic ratio of 30.41 is
-------
calculated using the acute value of 4.3 ug/L from the same study.
Daphnia magna were exposed in a second 21-day life-cycle test 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 0.19 ug/L 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. rnagna is 36.60 which is the geometric mean of the two available Acute-
Chronic ratios (30.41 and 44.06) for this species.
In an early life-stage test (32-day duration) with the fathead minnow,
Pimephales promelas, all fish exposed to the highest exposure concentration of
2.20 ug/L died during the test (Brooke et al. 1986). Survival was not reduced
at 0.92 ug/L or any of the lower TBT concentrations. The mean weight of the
surviving fish was reduced 4% at 0.08 ug/L, 9% at 0.15 ug/L, 26% at 0.45 ug/L,
and 48% at 0.92 ug/L when compared to the control fish. Mean length of fry at
the end of the test was significantly (p«€.05) reduced at concentrations >.0.45
ug/L. The mean biomass at the end of the test was higher at the two lowest
TBT concentrations (0.08 and 0.15 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 of individual fish 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.
A partial life cycle test (began with egg capsules and ended before
egg capsules were produced by the F1 generation) was conducted with the
stenoglossan snail Nucella lapillus (Harding et al. 1996) . The study was
conducted for one year with observations of egg capsule production, survival,
10
-------
and growth. The study by Harding et al. (1996) was a continuation of a study
by Bailey et al. (1991) during which they exposed eggs and juvenile snails for
one year to TBT concentrations similar to those used by Bailey et al. (1991).
The study by Harding et al. (1996) began with egg capsules produced by adults
at the end of the study by Bailey et al. (1991) . Negative effects due to TBT
were only observed in egg capsule production from the adults of the previous
study. Females that had not been exposed for one year to TBT produced 14.42
egg capsules per female. Females that had been exposed to <0.0027, 0.0077,
0.0334, and 0.1246 ugTBT/L for one year in the previous study (Bailey et al.
1991), produced 135.6, 104.6, 44.8, and 23.4% as many egg capsules as the
controls for the respective TBT concentrations. The chronic value is based
upon reproductive effects and is the geometric mean of the lowest observed
effect concentration (LOEC) of 0.0334 ug/L and the no observed effect
concentration (NOEC) of 0.0077 ug/L which is 0.0153 ug/L. Survival and growth
were not affected at any TBT tested concentration. An acute-chronic ratio of
4,752 can be calculated using the acute value from this test of 72.7 ug/L.
The acute-chronic ratio for N. lapillus is about 108 times higher than the
next lower acute-chronic ratio for D. magna (36.60). It is not used to
calculate a final acute-chronic ratio because it is more than ten times higher
than any other ratio.
Two partial life-cycle toxicity tests were conducted using the
copepod, Eurytemora 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 after three days. Percentage survival of neonates was 79% less
than control survival in the lowest tested TBT concentration (0.088 ug/L), and
0% in 0.479 ug/L. The chronic value is <0.088 ug/L in this test.
In the second copepod test, percentage survival of neonates was
significantly reduced (73% less than control survival) in 0.224 ug/L; brood
size was unaffected in any tested concentration (0.018-0.224 ug/L). No
statistically significant effects were detected in concentrations '€.094 ug/L.
The chronic value in this test is 0.145 ug/L. It is calculated as the
11
-------
geometric mean of the NOEC (0.094 ug/L)and the LOEC (0.224 ug/L). The acute-
chronic ratio is 15.17 when the acute value of 2.2 ug/L from this test is
used.
Life-cycle toxicity tests were conducted with the saltwater mysid,
Acanthomysis sculpta (Davidson et al. 1986a, 1986b). The effects of TBT on
survival, growth, and reproduction of A. sculpta 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 endpoints in one life-cycle
test. All tests began with newly released juveniles and lasted through
maturation and spawning; therefore, they 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 the next lower TBT concentration (0.09 ug/L) was not
significantly different from the control treatment. Reductions in juveniles
released resulted from deaths of embryos within brood pouches of individual
females and not from reduced fecundity. Numbers of females releasing viable
juveniles was reduced in 0.19 and 0.33 ug/L. At concentrations of 0.38 ug/L
and above, survival and weight of female mysids were reduced; all mysids in
0.48 ug/L died. The chronic value (0.1308 ug/L) is the geometric mean of 0.09
ug/L and 0.19 ug/L and is based upon reproductive effects. The acute-chronic
ratio is 4.664 when an acute value of 0.61 ug/L reported by Valkirs et al.
(1985) is used (Table 2). The acute and chronic tests were conducted in the
same laboratory.
The Final Acute-Chronic Ratio of 12.69 was calculated as the geometric
mean of the acute-chronic ratios of 36.60 for D. magna, 10.01 for £. promelas,
4.664 for A. sculpta and 15.17 for E_. af finis. Division of the freshwater and
saltwater Final Acute Values by 12.69 results in Final Chronic Values for
freshwater of 0.0723 ug/L and for saltwater of 0.0605 ug/L (Table 3). Both of
these Chronic Values are below the experimentally determined chronic values
from life-cycle or early life-stage tests (0.1414 ug/L for D. magna and 0.1308
ug/L for A. sculpta) . The close agreement between the saltwater Final Chronic
12
-------
Value and the freshwater Final Chronic Value suggests that salinity has little
if any affect on the toxicity of TBT.
Toxicitv to Aquatic Plants
The various plant species tested are highly variable in sensitivity to
TBT. Twenty-one species of algae and diatoms were tested in fresh and salt
water. The saltwater species are more sensitive to TBT than the freshwater
species for which data are available. No explanation is apparent.
Blanck et 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.
Fargasova and Kizlink (1996), Huang et al. (1993), and Miana et al. (1993)
measured severe reduction in growth of several green alga species at TBT
concentrations ranging from 1 to 12.4 ug/L. Several green alga species appear
to be as sensitive to TBT as many animal species (compare Table 4 with Table
1) .
Toxicity tests on TBT have been conducted with five species of
saltwater phytoplankton including the diatoms, Skeletonema costatum, Nitzshia
sp., flagellate green alga, Dunaliella tertiolecta, D. salina, and D. viridis.
The 14-day ECSO's (reduction in growth) for S_. costatum of >0.12 but <0.24
ug/L in one test and 0.06 ug/L in a second test (EG&G Bionomics 1981c) were
the lowest values reported for algal species. Thain (1983) reported that
measured concentrations from 0.97 to 17 ug/L were algistatic to the same
species in five-day exposures. The results for algal toxicity tests with the
same species varied between laboratories by more than an order of magnitude.
A diatom, Nitzschia sp., and two flagellate green alga of the genus Dunaliella
sp. were less sensitive to TBT than Skeletonema costatum, but they were more
sensitive than most species of freshwater algae. No data are available on the
effects of TBT on fresh or saltwater vascular plants.
A Final Plant Value, as defined in the Guidelines, cannot be obtained
because no test in which the concentrations of TBT were measured and the
endpoint was biologically important has been conducted with an important
13
-------
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.
Bioaccumulation
Bioaccumulation of TBT has been measured in one species of freshwater
mollusc and four species of freshwater fish (Table 5). Adults of the zebra
mussel, Dreissena polvmorpha, were placed in cages at a marina and at an
uncontaminated site in a lake for 105 days (Becker-van Slooten and Tarradellas
1994). Subsamples of the organisms were periodically monitored for TBT tissue
concentrations. They reached steady-state concentrations after 35 days. The
BCF/BAF was 17,483 when adjusted for wet weight and lipid normalized to 1% for
TBT at an average water exposure concentration of 0.0703 ug/L. Growth of the
TBT-exposed organisms may have been slightly reduced. Martin et al. (1989)
determined the whole body bioconcentration factor (BCF) for rainbow trout,
Oncorhyncus mykiss to be 406 after a 64-day exposure to 0.513 ug TBT/L.
Equilibrium of the TBT concentration was achieved in the fish in 24 to 48 hrs.
In a separate exposure to 1.026 ugTBT/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. Carp, Cyprinus carpio, and guppy,
Poecilia reticulatus, demonstrated plateau BCF's in 14 days. BCFs were 501.2
and 460, respectively. Goldfish, Carassius auratus, reached a much higher BCF
(1,976) in the whole body than the other fish species tested.
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, two species of snails, and a fish
(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 (1987) observed BCFs
of 10,400 to 37,500 after 56 days of exposure. BAFs from field deployments of
mussels were similar to BCFs from laboratory studies; 11,000 to 25,000
14
-------
(Salazar and Salazar 1990a) and 5,000 to 60,000 (Salazar and Salazar 1991).
In a study by Bryan et al. (1987a), laboratory BCFs for the snail Nucella
lapillus (11,000 to 38,000) also were similar to field BAFs (17,000). Year-
long laboratory studies by Bailey et al. (1991) and Harding et al. (1996)
produced similar BAFs in the snail N. lapillus ranging from 6,172 to 21,964.
In these tests, TBT concentrations ranged from 0.00257 to 0.125 ug/L, but
there was no increase in BAFs with increased water concentration of TBT.
The soft parts of the Pacific oyster, Crassostrea gigas, exposed to
TBT for 56 days contained 11,400 times the exposure concentration of 0.146
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, Ostrea edulis, exposed to a TBT
concentration of 2.62 ug/L for 45 days (Thain and Waldock 1985; Thain 1986).
Other tests with the same species (Table 5) resulted in BCFs ranging from 397
to 1,167. One fish species, Poecilia reticulatus, was exposed in salt water
to 0.28 ug/L for 14 days and a plateau BCF of 240 was demonstrated (Tsuda et
al. 1990b). The BCF agrees reasonably well with the freshwater BCF (460) with
the same species.
In a field study conducted in the Icelandic harbor of Reykjavik with
the blue mussel, M. edulis, and the Atlantic dogwhelk, N. lapillus, seasonal
fluctuations were seen in body burdens of TBT and DBT (Skarphedinsdottir et
al. 1996). They did not report the water concentrations for TBT, and
speculated that because shipping did not vary seasonally, the fluctuations in
body burdens were due to seasonal feeding and resting activities. They
demonstrated that body burdens of TBT and DBT were highest at high latitudes
during late summer or early autumn.
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
15
-------
Some data (Table 6) were located on the lethal and sublethal effects
of TBT on aquatic species that were insufficient to meet the criteria for
inclusion in the tables for acute toxicity, chronic toxicity, plant toxicity,
or bioconcentration in this document. These data are potentially useful and
sometimes support data in other tables. Sometimes the data are unique and
useful to evaluate TBT affects on aquatic organisms.
Several studies report the effects of TBT on natural groups of
organisms in laboratory microcosms. In most of these studies, the effects of
TBT administered to the water were rapid. Two microcosm studies were
conducted with freshwater organisms (Delupis and Miniero, 1989; Miniero and
Delupis, 1991) in which single dose effects were measured on natural
assemblages of organisms. In both studies, the effects were immediate. D.
rnagna disappeared soon after an 80 ug/L dose of TBT, ostracods increased, and
algal species increased immediately then gradually disappeared during the 55-
day study. In the second study (Miniero and Delupis, 1991), metabolism was
monitored by measuring oxygen consumption and again the effects were rapid.
Doses of TBT (4.7 and 14.9 ug/L) were administered once and metabolism was
reduced at 2.5 days and returned to normal in 14.1 days in the lower exposure.
In the higher exposure, metabolism was reduced in one day and returned to
normal in 16 days. Kelly et al. (1990a) observed a similar response in a
seagrass bed at 22.21 ug/L of TBT. The primary herbivor, Cymadusa compta,
declined and the sea grass increased in biomass. Saltwater microbial
populations were exposed for one hour to TBT concentrations of 4.454 and 89.07
ug/L then incubated for 10 days (Jonas et al. 1990). At the lower
concentration, metabolism of nutrient substrates was reduced and at the higher
concentration, 50 percent mortality of microbes occurred.
Several fresh and saltwater algal species were exposed to TBT for
various time intervals and several endpoints determined. Toxicity (EC50) in
freshwater species ranged from 5 ug/L for a natural assemblage to 20 ug/L for
the green alga Ankistrodesmus falcatus (Wong et al. 1982). Several salt water
alga, a green alga, Dunaliella tertiolecta; the diatoms, Minutocellus
polymorphus, Nitzshia sp., Phaeodactylum tricornutum, Skeletonema costatum.
16
-------
and Thailassiosira pseudonana; the dinoflagellate, Gymnodinium splendens, the
microalga, Pavlova lutheri and the macroalga, Fucus vesiculosus were tested
for growth endpoints. 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 with tributyltin acetate,
tributyltin chloride, tributyltin fluoride, and tributyltin oxide exposures
with S. 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).
The freshwater invertebrates, a rotifer (Brachionus calyciflorus) and
a coelentrate (Hydra sp.), showed widely differing sensitivites to TBT. Hydra
sp. were affected at 0.5 ug/L resulting in deformed tentacles, but the rotifer
did not show an effect on hatching success until the exposure concentration
reached 72 iag/L. The cladoceran, D. magna, has 24-hr ECSOs ranging from 3
(Polster and Halacha 1972) to 13.6 ug/L (Vighi and Calamari 1985). When the
endpoint of altered phototaxis was examined in a longer-term exposure of 8
days, a much lower effect concentration of 0.45 U9/L was measured (Meador
1986) .
Saltwater invertebrates (exclusive of molluscs) had reduced survival
at concentrations as low as 0.500 ug/L for the polychaete worm, Neanthes
arenaceodentata in a 10 week exposure to TBT (Moore et al. 1991) and 0.003
ug/L in a copepod, Acartia tonsa, in a eight-day exposure. Other
invertebrates were more hardy including an amphipod, Orchestia traskiana, that
had an LC80 and an LC90 of 9.7 ug/L for nine day exposures to TBTO and TBTF,
respectively. Larvae of the mud crab, Rhithropanopeus harrisii, tolerated
high concentrations of TBT with one test resulting in an LC50 of 33.6 ug/L for
a 40 day exposure (Laughlin and French 1989) .
A number of studies showed that TBT exposure resulted in developmental
problems for non-mollusc invertebrates. For example, the copepod, A. tonsa,
had slower rate of development from nauplii to copepodid stage at 0.003 ug/L
(Kusk and Petersen 1997); the grass shrimp, Palaemonetes pugio, had retarded
telson regeneration at 0.1 ug/L (Khan et al. 1993); the mud crab, R. harrisii,
17
-------
had reduced developmental rate at 14.60 ug/L (Laughlin et al. 1983); retarded
limb regeneration in the fiddler crab, Uca pugilator, at 0.5 ug/L (Weis et al.
1987a); and retarded arm regeneration in the brittle star, Brevoortia
tyrannus, at •€.! ug/L (Walsh et al. 1986a). Lapota et al. (1993) reported
reduced shell growth in the blue mussel, Mytilus edulis, at 0.050 ug/L and no
reduction of shell development at 0.006 ug/L in a 33-d study. The test had
exposure solutions renewed every third or fourth day during which time TBT
concentrations declined 33 to 90%.
Vertebrates are as sensitive to TBT as invertebrates when the
exposures are of sufficient duration. Rainbow trout, 0. mykiss, 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 (Seinen et
al. 1981), the LC100 decreased to 4.46 ug/L and a 20% reduction in growth was
seen at 0.18 ug/L. De Vries et al. (1991) measured a similar response in
rainbow trout growth in another 110 day exposure. They demonstrated decreased
survival and growth at 0.200 ug/L but not at 0.040 ug/L. Triebskorn et al.
(1994) found reduced growth and behavior changes in the fish at 21 days when
exposed to 0.5 ug/L. Hall et al. (1988b) observed reduced growth in the
inland silverside, Menidia beryllina, at 0.093 ug/L in a 28 day exposure. The
frog, Rana temporaria, has a LC50 of 28.2 ug/L for a 5-day exposure to TBT.
An attempt was made to measure the bioconcentration of TBT with the
green alga, Ankistrodesmus 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.
Several studies reported BCFs for fish but failed to demonstrate plateau
concentrations in the organism. In these studies, rainbow trout BCFs ranged
from 990 (Triebskorn et al. 1994) to 3,833 (Schwaiger et al. 1992). Goldfish
achieved a BCF of 1,230 (Tsuda et al. 1988b) in a 14-day exposure and carp
achieved a BCF of 295 in the muscle tissue in 7 days (Tsuda et al. 1987) .
18
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TBT has been shown to produce the superimposition of male sexual
characteristics on female neogastropod (stenoglossan) snails (Smith 1981b,
Gibbs and Bryan 1987) and one freshwater gastropod (Prosobranchia), Marisa
cornuarietis (Schulte-Oehlmann et al. 1995). 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 (RPSI; ratio of female to male penis volume3 x 100) and
the six developmental stages of the vas deferens sequence index (VDSI) (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 (planktonic
larval stages), the impacts of TBT are less certain because recruitment from
distant stocks of organisms can occur. Natural pseudohemaphroditism 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 low to moderate concentrations (low parts per trillion) of TBT in
water, sediment or snails and other biota (Alvarez and Ellis 1990; Bailey and
Davies 1988a, 1988b; Bryan et al. 1986, 1987a; Davies et al. 1987, Durchon
1982; Ellis and Pottisina 1990; Gibbs and Bryan 1986, 1987; Gibbs et al. 1987;
Langston et al. 1990; Short et al. 1989; Smith 1981a, 1981b; Spence et al.
1990a). Imposex has been observed (12% of the females) in common whelk,
Buccinum undatum, in the North Sea as far as 110 nautical miles from land (Ide
et al. 1997). The sample from this site averaged 1.4 ugTBT/kg of wet weight
soft tissues. Other samples of organisms collected nearer to shore in
various places in the North Sea generally had higher TBT concentrations.
Although imposex has been observed in 45 species of snails worldwide
(Ellis and Pattisima 1990, Jenner 1979), definitive laboratory and field
studies implicating TBT as the cause have focused on seven North American or
cosmopolitan species; the Atlantic dogwhinkle (N. lapillus), file dogwhinkle
19
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(N. lima), eastern mud snail [Ilyanassa (Nassarius) obsoleta], a snail (Hinia
reticulta), whelks (Thais orbita and T. clavigera), and the European sting
winkle (Ocenebra erinacea). Imposex has been associated with reduced
reproductive potential and altered density and population structure in field
populations of N. lapillus (Harding et al. 1997; Spence et al. 1990a). 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 N. lima because 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 I_.
obsoleta in Sarah Creek, VA also suggests population impacts (Bryan et al.
1989a). However, other causes may explain this as oviducts were not blocked
and indirect development (plantonic larvae) facilitating recruitment from
other areas 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 N. 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_.
obsoleta 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 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
20
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of snails to TBT in controlled laboratory and field studies (Text Table 1).
Eastern mud snails, I_. obsoleta, collected from the York River, VA near Sarah
Creek had no incidence of imposex (Bryan et al. 1989a) and contained no
21
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Text Table 1. Summary of Available Laboratory and Field Studies relating the Extent of Imposex
of Female Snails, Measured by Relative Penis Size (Volume3 Female Penis/Male Penis = RPSI)
and the Vas Deferens Sequence Index (VDSI), as a Function of Tributyltin
Concentration in Water and Dry Tissue
TBT Concentration
Imposex
Spec i es
Eastern mud
sna i I ,
I I yanassa
obso I eta
Sna i I ,
H i n i a
ret i cu I ata
Whelk,
Thais orbita
Experi mental
Desiqn Water
uq/L
Field-York 0.0016
River, UK 0.01-
-Sarah Creek 0.023
Field-32 sites
N and NW
France
Field-Queens-
cl iff , UK
-Sandringham
-Brighton
-Portar I i ngton
-Morn i ngton
-Will iamstown
-Martha Point
-Ki rk Poi nt
-Cape Schanck
-Cape Schanck
-Barwon Heads
-Barwon Heads
Sna i I
Ti ssue
ug/g dry
<0.02
-0.1 -0.73
<1 .5
>1 .5
0.365*
0.224*
<0.002*
0.255*
0.045*
<0.002*
0.031*
0.011*
0.108*
0.095*
ND
0.071*
RPS I
-
-
<10
>30
19.55
12.16
7.34
3.67
2.55
1 .25
0.03
0.02
0
0
0
0
VDS I Comments
No i mposex
40-100% incidence
<3.0 Low i mposex incidence
>3.0 High i mposex incidence
100% incidence
100% incidence
100% incidence
92 . 3% i nci dence
100% incidence
100% incidence
25% incidence
35.7% incidence
0% i nc i dence
0% i nc i dence
0% i nc i dence
0% i nc i dence
Reference
Bryan et
a I . 1989a
Stroben et
a I . 1992a
Foale 1993
-------
Text Table 1. Continued
TBT Concentration
Imposex
Spec i es
Fi le
dogwh i nkle,
Nucel la I i ma
At I ant i c
dogwh i nkl e ,
(adu Its) ,
Nuce I I a
I api I I us
Exper i menta I
Des i qn
Field
-Auke Bay, AK
-Auke Bay, AK
Crook lets
Beach, UK
Laboratory : 2
year exposure
Water
ua/L
-
-
<0.0012
*
0.0036*
0.0083*
0.046*
0.26*
Sna i I
Ti ssue
uq/q dry
ND(<0.01)
0.03-0.16
0.14-0.25*
0.41*
0.74*
4.5*
8.5*
RPSI
0.0
14-34
2-65
10/14.2
43.8
56.4
63.3
VDS I Comments
0.0 0% incidence
2.2-4.3 100% incidence, reduced
abundance
2.9
3.7/3.7
3.9
4.0
4.1 Some steri I ization
Reference
Short et
a I . 1989
Bryan et
a I . 1987a
At I ant i c
dogwh i nkle,
NucelI a
Iapi I I us
At I ant i c
dogwh i nkle,
(egg capsuI
to aduIt),
NucelI a
lap! I I us
Laboratory,
spi res
pa i nted, 8 mo.
Crook Iets
Beach, UK
Laboratory; 2
year exposure
<0.0012
0.0036
0.0093
0.049
0.24
0.19
0.58
1 .4
4.1
7.7
3.7
48.4
96.6
109
90.4
3.2
4.4
5.1
5.0
5.0
NormaI femaIes
1/3 sterile, 160 capsules
AI I steri le, 2 capsules
AI I steri le, 0 capsules
AI I steri le, 0 capsules
Bryan et
al . 1987b
Gibbs et
a I . 1988
23
-------
Text Table 1. Continued
TBT Concentration
Imposex
Exper i menta I
Species Desiqn Water
ua/L
European Field -19 sites
st i ng wi nkle, SW UK
Ocenebra
er i nacea
-
-
-
-
-
-
-
-
-
-
-
-
-
-
-
Sna i I
Ti ssue
ug/g dry
0.185
<0.024
0.187
0.773
2.313
0.976
1 .057
1 .200
0.303
0.122
0.703
0.764
0.527
0.488
0.366
0.253
0.832
1 .010
0.510
RPSI
0
0
16.3
66.9
88.2
71 .1
53.4
84.2
7.4
7.0
36.0
52.7
46.5
42.3
0.04
33.9
58.0
79.3
59.7
VDS I Comments
No i mposex
No i mposex
Fema I es somewhat deformed
Females highly deformed
Females highly deformed
Females highly deformed
Females highly deformed
Females highly deformed
Fema I es somewhat deformed
Fema I es somewhat deformed
Females highly deformed
Females highly deformed
Females highly deformed
Females highly deformed
Fema I es somewhat deformed
Females highly deformed
Females highly deformed
Females highly deformed
Females highly deformed
Reference
Gibbs et
a I . 1990
24
-------
At 1 ant i c
dogwh i nkl e ,
Nuce 1 1 a
1 api 1 1 us
Atlantic
dogwh inkle,
Nucel la
1 ap i 1 1 us
At 1 ant i c
dogwh inkle,
Nuce 1 la
1 ap i 1 1 us
At 1 ant i c
dogwh inkle,
Nuce 1 la
1 ap i 1 1 us
Field, S.W.
UK
Port Joke, UK
Crook 1 ets
Beach
Meadfoot
Renney Rocks
Batten Bay
Laboratory,
f 1 ow-through ,
one year
Laboratory,
f 1 ow-through ,
one year
0.002-
0.005*
• 6.010
• 6.017-
0.025
-
-
-
-
-
<0.0015
<0.0015
<0.0027
0.0077
0.0334
0.1246
<0.0015
<0.0015
0.0026
0.0074
0.0278
0.1077
<0.5*
0.5-1 .0*
<1 .0*
0.11*
0.21*
0.32*
0.43*
1 .54*
<0.10*
<0.10*
0.35*
1 .10*
3.05*
4.85*
<0.10
<0.10
<0.10
0.38
1 .12
3.32
• ?0-60
• 80-70
• 80-
100
0.0
2.0
30.6
38.9
22.9
0.10
0.04
5.33
20.84
42.08
63.40
0.07
0.04
64.04
88.57
90.96
117.70
• 2.0-
4.5
• 4.5-
6.0
• 4.5-
6.0
-
-
-
-
-
1 .06
0.70
3.15
3.97
4.33
4.25
1 .28
1 .14
3.98
4.96
5.00
4.99
Li m i ted ster i I i ty
• 50% steri le
A I I steri le
0% aborted egg capsules
0% aborted egg capsules
15% aborted egg capsules
38% aborted egg capsules
79% aborted egg capsules
Control, 37.1% i mposex
Solvent control , 24.3%
i mposex
5.3% reduced growth,
92 . 3% i mposex
11.0% reduced growth,
1 00% i mposex
17.1% reduced growth,
1 00% i mposex
18.9% reduced growth,
1 00% i mposex
Control, 42.2% i mposex
Solvent control, 37.5%
i mposex
98.9% i mposex
98.8% i mposex
1 00% i mposex
98 . 7% i mposex
Gibbs et
a I . 1987
Gibbs and
Bryan 1986;
Gibbs et
a I . 1987
Ba i I ey et
a I . 1991
Harding et
a I . 1996
Concentrations changed from ug Sn/L or ug Sn/g to ug TBT/L or ug TBT/g dry weight.
of wet we ight.
Dry weight estimated as
25
-------
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, N. lima, 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- undersized egg capsules produced.
Concentrations 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 two studies
conducted concurrently with N. lapillus for one year each, imposex was
observed. In the first study (Bailey et al. 1991), imposex (••stage 2) was
observed in »92.3% of the females exposed to TBT at '6.0027 ug/L at the end of
the study. Harding et al. (1996) exposed the offspring from parents exposed
the study by Bailey et al. (1991) for one year to similar TBT concentrations.
In the second generation of TBT-treated snails, body burdens of TBT were lower
in the second generation at similar treatment concentrations used in the first
generation. Also, the RPSI and VDSI values were higher for the same
treatments in the second generation. Harding et al. (1996) found «98.7%
imposex in females at TBT concentrations '€.0026 ug/L.
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 N. lapillus. Imposex begins to occur, and
cause some reproductive failure at about 0.004 ug/L with complete sterility
occurring after chronic 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
26
-------
than 0.2 ug/L in others. If N. lapillus 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 was completely inhibited. A TBT concentration of 2.6
ug/L prevented development of gonads. Salazar et al. (1987) found no
negative effects in the same species at 0.157 ug/L, but Thain and Waldock
(1985) and Thain (1986) measured reduced growth at 0.2392 ug/L and reduced
survival (30%) at 2.6 ug/L.
Four species of snails (Hinia reticulata, Thais orbita, T. clavigera,
Ocenebra erinacea) not resident to North America also demonstrated imposex
effects when exposed to TBT in field studies (Text Table 1). The snail H.
reticulata is less sensitive to TBT than other snails having higher body
burdens (>1.5 ug/g) before showing affects of imposex. Thais sp. showed high
imposex incidence at tissue concentrations as low as 0.005 ug/g and no imposex
at other locations with tissue concentrations of 0.108 ug/g- Ocenebra
erinacea did not show imposex in a field study at body burdens as high as
0.185 ug/g- but females were deformed at all higher concentrations.
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, Mytilus edulis, exposed to 0.0973 ug/L for
15 days was 51%; survivors were moribund and stunted (Beaumont and Budd 1984).
Similarly, Dixon 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
27
-------
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 56 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 in 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 mg/kg (Davies et al. 1988). Abnormal shell development was observed in
an exposure to 0.77 U9/L that began with embryos of the eastern oyster,
Crassostrea virginica, and lasted for 48 hours (Roberts, 1987) . Adult eastern
oysters were also sensitive to TBT with reductions in condition index after
exposure for 57 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 to the oysters C. virginica, Ostrea edulis and 0. lurida.
Condition of adult clams, Macoma nasuta, and scallops, Hinnites multirugosus
were not affected after 110 days exposure to 0.204 ug/L (Salazar et al. 1987).
Long-term exposures have been conducted with a number of saltwater
crustacean species. Johansen and Mohlenberg (1987) exposed adult A. tonsa for
five days to TBT and observed impaired (25% reduction) egg production on days
3, 4 and 5 in 0.1 ug/L. Impaired egg production to a lessor amount was
observed on day 5 in 0.01 and 0.05 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, Gammarus
oceanicus, was significantly reduced after eight weeks of exposure to TBT
concentrations ^0.2816 ug/L (Laughlin et al. 1984b). Hall et al. (1988b)
28
-------
observed no effect of 0.579 ug/L on Gammarus sp. after 24 days. Developmental
rates and growth of larval mud crabs, Rhithropanopeus 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 (R. Harrisii) 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 Perlmutter 1987). Arm regeneration was reduced in brittle stars exposed
to 0.1 ug/L (Walsh et 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.
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, Oncorhynchus tshawytscha,
exposed to 1,490 ug/L for 96 hours (Short and Thrower 1986a,1986c).
TBT concentrations less than the Final Chronic Value of 0.0605 ug/L
from Table 3 have been shown to affect the growth of early life-stages of
commercially important bivalve molluscs and survival of ecologically important
copepods (Table 6; Text Table 2). Survival of the copepod A. tonsa was
significantly reduced in three tests in 0.029, 0.023 and 0.024 ug/L; 30, 27
29
-------
and 51 percent of control survival, respectively (Bushong et al. 1990).
Survival decreased with increase in exposure concentration but was not
significantly affected in the 0.012 ug/L exposure concentration.
Laughlin et al. (1987, 1988) observed a significant decrease in growth
of hard clam (Mercenaria mercinaria) larvae exposed for 14 days to >.0.01 ug/L
(Text Table 2). Growth rate (increase in valve length) was 75% of controls in
0.01 ug/L, 63% in 0.025 ug/L, 59% in 0.05 ug/L, 45% in 0.1 ug/L, 29% in 0.25
ug/L and 2.2% in 0.5 ug/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 ug/L and reduced valve lengths in >.0.02 ug/L (Lawler and Aldrich
1987; Text Table 2). Increase in valve length was 101% of control lengths in
0.01 ug/L, 72% in 0.02 ug/L, 17% in 0.05 ug/L, 35% in 0.1 ug/L and 0% in 0.2
ug/L. Shell thickening was also observed in this species exposed to >.0.02 ug/L
for 49 days (Thain et al. 1987). They predicted from these data that
approximately 0.008 ug/L would be the maximum TBT concentration permitting
30
-------
Text Table 2. Summary of laboratory and field data on the effects of tributyltin on saltwater
organisms at concentrations less than the Final Chronic Value of 0.0605 ug/L
Species
Copepod (nauplii-
adult),
Acartia tonsa
Hard clam (4 hr
larvae -
metamorphosis),
Mercenaria
mercenaria
Experimental Design3
#1: F,M, 9-day
duration, •10
copepods/replicate,
4 replicates
#2: F,M, 6-day
duration,
• 10
copepods/replicate,
4 replicates
#3: F,M, 6-day
duration,
• 10
copepods/replicate,
4 replicates
R,M, 14-day
duration,
<150
larvae/replicate,
three replicates.
Measured = 80-100%
nominal at t = 0.4
hr;
20-30% at t = 24 hr
Concentration (uq/L)
Measured
control
0.029
0.05-0.5
control
0.007-0.012
0.023
0.048-0.102
control
0.006-0.010
0.024
0.051-0.115
Nominal
control
0.01-0.5
Response
77% survival
23% survival13
0-2% survival13
71% survival
32% survival
19% survival13
0-14% survival13
59% survival
44-46% survival
30% survival13
2-35% survival13
100% Growth
(Valve length)
~75%-22% Growth
(Value length)13
Reference
Bushong et al.
1990
Laughlin et al
1987,1988
Nominal
-------
Pacific oyster
(spat),
Crassostrea gigas
R,N, 48-day
duration,
20 spat/treatment
control
0.01-0.05
control
0.01-0.2
0.02-0.2
Shell thickening
100% Growth
(Valve length)
101% Growth (Value
length)
0-72% Growth (Value
length)13
Lawler and
Aldrich 1987
Text Table 2.
(Continued)
Species
Pacific oyster
(larvae
and spat),
Crassostrea gigas
Pacific oyster
(spat),
Crassostrea gigas
European oyster
(spat),
Ostrea edulis
Experimental Design3
Field
R,M/N, 21-day
duration,
75,000
larvae/replicate
R,M, 4 week
duration, 4
replicates, 30 spat
each
R,N, 20-day
duration,
50 spat/treatment
Concentration (ug/L)
Measured
0.011-0.015
-0.018-0.060
0.24,0.29,0.69
Measured
Control,0.005,0.010,0.0
15,0.020
Nominal
control, 0.1, 0.05,
0.025
Nominal
control
0.02-2.0
control
0.02-2.0
Response
No shell thickening
Shell thickening
and decreased meat
weight
Mortality 100% by
day 1
Growth decreased
79% in 0.005, 78%
in 0.010, 78% in
0.015, 84% in 0.020
Mortality 100% in
0.05 and 1.0 ug/L;
86% in 0.025 ug/L
100 length
76-81% lengthb
202% weight gain
151-50% weight gain
Reference
Springborn
Bionomics, Inc.
1984a
Nell and Chvojka
1992
Thain and
Waldock 1985
-------
European oyster
(adult),
Ostrea edulis
R,N, 96-hr duration
Nominal
0.010
12% decrease of Axiak et al.
height of digestive 1995a
cells
a R = renewal; F = flow-through, N = nominal, M = measured.
b Significantly different from controls.
33
-------
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 ug/L but not at 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 0.025 ug/L (Springborn Bionomics 1984a) . No larvae
survived in >.0.050 ug/L.
Growth of spat of the European oyster (Ostrea edulis) was reduced at
^0.02 ug/L (Thain and Waldock 1985; Text Table 2). Spat 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 ug/L, weight gain was identical; i.e., 35% of
controls. Growth of larger spat was marginally reduced by 0.2392 ug/L (Thain
1986; Thain and Waldock 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 the reductions in growth occur in commercially or
ecologically important saltwater species at concentrations of TBT less than
the Final Chronic Value of 0.0605 ug/L derived using Final Acute Values and
Acute-Chronic Ratios from Table 3. Therefore, EPA believes the Final Chronic
Value should be lowered to 0.001 ug/L to limit unacceptable impacts on A.
tonsa, Mercenaria mercenaria, Crassostrea gigas and Ostrea edulis observed at
0.02 ug/L. At this criteria concentration, imposex would not be expected in
Ilyanassa obsoleta, N. lapillus and similarly sensitive neogastropods;
populations of N. lapillus and similarly sensitive snails with direct
development would not be impacted and growth of M. mercenaria would not be
lowered.
Unused Data
Some data concerning the effects of TBT on aquatic organisms were not
34
-------
used because the tests were conducted with species that are not resident in
North America (e.g., Ali et al. 1990; Allen et al. 1980; Axiak et al. 1995b;
Batley et al. 1989,1992; Burridge et al. 1995; Carney and Paulini 1964;
Danil'chenko 1982; Deschiens and Floch 1968; Deschiens et al. 1964,1966a,
1966b; de Sousa and Paulini 1970; Pent 1991, 1992; Pent and Hunn 1993; Pent
and Meier 1992; Frick and DeJimenez 1964; Girard et al. 1996; Helmstetter and
Alden 1995; Hopf and Muller 1962; Jantataeme 1991; Karande and Ganti 1994;
Karande et al. 1993; Kubo et al. 1984; Langston and Burt 1991; Lewis et al.
1995; Nagabhushanam et al. 1991; Nagase et al. 1991; Nias et al. 1993;
Nishuichi and Yoshida 1972; Oehlmann et al. 1996; Reddy et al. 1992; Ringwood
1992; Ritchie et al. 1964; Ruiz et al. 1994a, 1994b, 1995a, 1995b, 1995c,
1997; Sarojini et al. 1991, 1992; Scadding 1990; Scammell et al. 1991; Seiffer
and Schoof 1967; Shiff et al. 1975; Shimizu and Kimura 1992; Smith et al.
1979; Spence et al. 1990b; Stebbing et al. 1990; Sujatha et al. 1996; Tsuda et
al. 1986, 1991a; Upatham 1975; Upatham et al. 1980a, 1980b; Vitturi et al.
1992; Webbe and Sturrock 1964; Yamada et al. 1994; Yla-Mononen 1989).
Alzieu (1986) , Cardarelli and Evans (1980) , Cardwell and Sheldon
(1986), Cardwell and Vogue (1986), Champ (1986), Chau (1986), Eisler (1989),
Envirosphere Company (1986), Evans and Leksono (1995), Gibbs and Bryan (1987),
Gibbs et al. (1991a), Good et al. (1980), Guard et al. (1982), Hall (1988,
1991), Hall and Pinkney (1985), Hall et al. (1991), Hodge et al. (1979),
International Joint Commission (1976), Jensen (1977), Kimbrough (1976),
Kumpulainen and Koivistoinen (1977), Lau (1991), Laughlin (1986), Laughlin and
Linden (1985), Laughlin et al. (1984a), McCullough et al. (1980), Monaghan et
al. (1980), North Carolina Department of Natural Resources and Community
Development (1983,1985), Rexrode (1987), Salazar (1989), Seligman et al.
(1986) , Slesinger and Dressier (1978) , Stebbing (1985) , Thayer (1984) ,
Thompson et al. (1985), U.S. EPA (1975,1985b), U.S. Navy (1984), von Rumker et
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).
Results were not used when the test procedures, test material, or
35
-------
results were not adequately described (e.g., Bruno and Ellis 1988; Cardwell
and Stuart 1988; Chau et 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. 1990b; Kolosova et al. 1980; Laughlin 1983;
Mercier et al. 1994; Nosov and Kolosova 1979; Smith 1981c; Stroganov et al.
1972,1977). 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; Henderson and Salazar 1996; Mattiessen and Thain
1989; North Carolina Department of Natural Resources and Community Development
1983; Pope 1981; Quick and Cardarelli 1977; Salazar and Salazar 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. Data were not used when the organisms
were exposed to TBT by injection or gavage (e.g., Pent and Stegeman 1991,
1993; Horiguchi et al. 1997; Rice et al. 1995; Rice and Weeks 1990; Rouleau et
al. 1995). Caricchia et al. (1991), Salazar and Chadwick (1991), and Steinert
and Pickwell (1993), did not identify the organism exposed to TBT. Some
studies did not report toxic effects of TBT (e.g., Balls 1987; Gibbs 1993;
Meador et al. 1984; Page 1995; Salazar 1986; Salazar and Champ 1988).
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 form of an alga. Results of tests in which enzymes,
excised or homogenized tissue, or cell cultures were exposed to the test
36
-------
material were not used (e.g., Avery et al. 1993; Blair et al. 1982;
Bruschweiler et al. 1996; Falcioni et al. 1996; Pent and Bucheli 1994; Pent
and Stegeman 1991; Fisher et al. 1990; Josephson et al. 1989; Joshi and Gupta
1990; Pickwell and Steinert 1988; Reader et al. 1994, 1996; Rice and Weeks
1991; Virkki and Nikinmaa 1993; Wishkovsky et al. 1989; Zucker et al. 1992).
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 Rhea et al. (1995), Salazar and Salazar (1989) and Valkirs et al.
(1985). Some data were not used because of problems with the concentration of
the test material (e.g., Springborn Bionomics 1984b; Stephenson et al. 1986;
Ward et al. 1981) or low survival in the exposure organisms (Chagot et al.
1990; Pent and Looser 1995). BCFs were not used when the concentration of TBT
in the test solution was not measured (Davies et al. 1986; Laughlin et al.
1986b; Paul and Davies 1986) or were highly variable (Becker et al. 1992;
Laughlin and French 1988). Reports of the concentrations in wild aquatic
animals were not used if concentrations in water were unavailable or
excessively variable (e.g., Curtis and Barse 1990; Davies et al. 1987, 1988;
Davies and McKie 1987; Gibbs et al. 1991b; Hall 1988; Han and Weber 1988;
Kannan et al. 1996; Oehlmann et al. 1991; Stab et al. 1995; Thrower and Short
1991; Wade et al. 1988; Zuolian and Jensen 1989).
Summary
Freshwater Acute Toxicity. The acute toxicity values for twelve
freshwater animal species range from 1.14 ug/L for a hydra, Hydra oligactis,
37
-------
to 12.73 ug/L for the lake trout, Salvelinus naymavcush. A thirteenth
species, a clam (Elliptio complanatus), had an exceptionally high toxicity
value of 24,600 ug/L. There was no apparent trend in sensitivities with
taxonomy; fish were nearly as sensitive as the most sensitive invertebrates
and more sensitive than others. When the much less sensitive clam was not
considered, remaining species sensitivities varied by a maximum of 11.2 times.
Plants were about as sensitive as animals to TBT.
Freshwater Chronic Toxicity. Three chronic toxicity tests have been
conducted with freshwater animals. Reproduction of D. 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 D. magna reproduction and survival was reduced at 0.34 ug/L
but not at 0.19, and the Acute-Chronic Ratio is 44.06. The species-mean
Acute-Chronic Ratio for D. magna is 36.60, which is the geometric mean of the
two available Acute-Chronic Ratios (30.41 and 44.06) for this species. Weight
of fathead minnows (£. promelas) 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 TBT was measured in zebra mussels, Dreissena polymorpha,
at 180,427 times the water concentration for the soft parts and in rainbow
trout, Oncorhyncus mykiss, 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.
Saltwater Acute Toxicity. Acute values for 33 species of saltwater
animals range from 0.61 ug/L for the mysid, Acanthomysis sculpta, to 204.4
ug/L for adult European flat oysters, Ostrea edulis. Acute values for the
twelve most sensitive genera, including molluscs, crustaceans, and fishes,
differ by less than a factor of four. Larvae and juveniles appear to be more
acutely sensitive to TBT than adults.
Saltwater Chronic Toxicity. A partial life-cycle test of one-year
duration was conducted with the snail, Nucella lapillus. TBT reduced egg
capsule production. The chronic value for this species was 0.0153 ug/L. No
Acute-Chronic Ratio is available for this species. A life-cycle test was
conducted with the copepod, Eurytemora affinis. The chronic value is based
38
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upon neonate survival and is 0.145 ug/L and the Acute/Chronic Ratio is 15.17.
A life-cycle toxicity test was conducted with the saltwater mysid,
Acanthomysis sculpta. The chronic value for A. sculpta 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 gigas.
The Final Acute-Chronic Ratio of 12.69 was calculated as the geometric
mean of the Acute-Chronic Ratios of 36.60 for D. magna, 10.01 for £. promelas
(the two freshwater species), and 4.664 for A. sculpta and 15.17 for E_.
affinis (the two saltwater species). Division of the freshwater and saltwater
Final Acute Values by 12.69 results in Final Chronic Values for freshwater of
0.0723 ug/L and for saltwater of 0.0605ug/L (Table 3). Both of these Chronic
Values are below the experimentally determined chronic values from life-cycle
or early life-stage tests (0.144 ug/L for D. magna and 0.1308 ug/L for A.
sculpta).
Tributyltin chronically affects certain saltwater copepods, gastropods,
and pelecypods at concentrations less than those predicted from "standard"
acute and chronic toxicity tests. The data show that reductions in growth
occur in commercially or ecologically important saltwater species at
concentrations of TBT less than the Final Chronic Value of 0.0605 ug/L derived
using Final Acute Values and Acute-Chronic Ratios from Table 3. Survival of
the copepod A. tonsa was reduced in >.0.023 ug/L. Growth of larvae or spat of
two species of oysters, Crassostrea gigas and Ostrea edulis was reduced in
about 0.02 ug/L; some C_. gigas larvae died in 0.025 ug/L. Shell thickening
and reduced meat weights was observed in the C. gigas at 0.01 ug/L. Since
these levels were ones at which an effect was seen, a protective level for
these commercially important species is, therefore, below 0.01 ug/L.
Weight of Evidence Considerations. The National Guidelines (Stephan
et.al. 1985) require that the criterion be lowered if sound scientific
evidence indicates that adverse effects might be expected on important
species. The above data demonstrate that the reductions in growth occur in
39
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commercially or ecologically important saltwater species at concentrations of
TBT less than the final Chronic Value of 0.0605 ug/L derived using Final Acute
Values and Acute-Chronic Ratios from Table 3. Consistent with the Guidelines
directive to consider other relevant data when establishing criteria, EPA
believes the final Chronic Value should be lowered to 0.001 ug/L.
Organometallics, particularly TBT and methyl mercury, have been shown
to impair the environment in multiple ways.
A major concern with TBT is its ability to cause imposex (the
superimposition of male anatomical characteristics on females) in a variety of
species. Imposex has been observed in 45 species of snails worldwide, with
definitive laboratory and field studies implicating TBT as the cause in seven
North American or cosmopolitan species. As listed on Table 6, adult
dogwhinkle, Nucella lapillus, exposed to 0.05 ug/L TBT for 120 days showed 41%
of the organisms evidencing imposex. A six month study of the same species in
1992 with a concentration of 0.012 ug/L TBT also showed imposex in the
organisms. Other studies showed more than 92% of the female N. Lapillus
exposed to TBT at 0.0027 ug/L exhibiting imposex; a followup study of
offspring showed almost 99% imposex in females at TBT concentrations of 0.0026
ug/L. Thus, numerous studies show imposex effects at doses well below the
calculated Final Chronic Value of 0.0605 ug/L. Many of the studies did not
produce a No Observed Adverse Effect Level because significant effects were
observed at the lowest concentration tested. The imposex effect may partially
explain the results of the studies in Tables 2 and 6 which show abnormal
growth patterns seen in other studies, including reduced growth, shell
thickening, and deformities. Imposex has also been linked with population
declines of snails in Canada (Tester et. al. 1996) and oysters in the United
Kingdom (Dyrynda 1992 and others); these declines were reversed after
restrictions on TBT use went into effect.
Another factor causing increased concern is the very high
bioaccumulation and bioconcentration factors associated with TBT. For some
40
-------
species, these factors reach into the thousands and tens of thousands. Data
are summarized in Table 5. They show BCF/BAF factors in the thousands for
rainbow trout, Oncorhynchus mykiss, where TBT concentrations were
approximately 1.0 ug/L, and in goldfish, Carassius auratus, where TBT
concentrations were approximately 0.1 ug/L. For saltwater species, field
studies of blue mussels, Mytilus edulis, at TBT concentrations of <0.1 ug/L,
showed BCF or BAF concentrations up to 60,000 (Salazar 1990 and 1991); the
American oyster, Crassostrea virginica, exhibited factors of 15,000 in TBT
concentrations of <0.3 ug/L (Roberts et.al. 1996); and the Pacific oyster,
Crassostrea gigas, had factors in the thousands when exposed to TBT in
concentrations of from 0.24 to 1.5 ug/L.
The National Research Council (NRC) conducted a four year study to
"...review critically the literature on hormonally-active agents in the
environment..." and "...identify the known and suspected toxicologic
mechanisms and impacts on fish, wildlife and humans...." The report, entitled
Hormonally Active Agents in the Environment (National Research Council, 1999),
cited Bettin et. al. (1996) who reported that TBT "is thought to cause penis
growth in female molluscs by affecting steroid metabolism."[p. 102]
Immunologic effects have been observed in eastern oysters exposed to 0.03 ug/1
TBT which resulted in increased infection intensity and mortality when later
exposed to "Dermo", a protozoan pathogen. TBT is widely assumed to enhance the
impairment caused by Dermo. However, data are currently insufficient to
determine which levels of Dermo and of TBT result in this heightened
interaction. Because levels of both Dermo and TBT are known to fluctuate
widely, it is prudent in the face of this uncertainty regarding impact on a
commercially important species to be conservative when establishing acceptable
levels.
Conclusion. The development of a chronic criterion for TBT in saltwater
considers four lines of evidence. The first line of evidence is the
traditional endpoints of adverse effects on survival, growth, and reproduction
41
-------
as demonstrated in numerous laboratory studies, recognizes that a number of
these studies have unbounded LOAELs at or near 0.01 ug/L, and recognizes
further that only one study included levels below 0.01 ug/L and that study (on
Acartia tonsa at 0.003ug/L) showed inhibition of development.
The next three lines of evidence are additional factors. These are: 1)
the production of imposex in field studies and the impact of imposex on
commercially significant species population levels, 2) the accumulation and/or
concentration of TBT in commercially and recreationally important freshwater
and saltwater species, and 3) the potential immunological effects of TBT, as
well as the finding that an important commercial organism (Eastern oyster)
already known to be vulnerable to the prevalent pathogen Dermo was made even
more vulnerable by prior exposure to TBT.
Considering only the traditional endpoints of adverse effects on
survival, growth, and reproduction, and the criteria calculation procedures
described in the National Guidelines, the Final Chronic Value would be set at
0.06 ug/L. However, the Agency believes that this level would not be
adequately protective because of the additional factors cited above. These
types of effects are unusual and seem to be characteristic of TBT's ability to
produce toxicity through multiple mechanisms.
The Agency is faced with the uncertainty created by the lack of
understanding of the relationship of these multiple factors. TBT does not lend
itself to the ordinary application of the existing criteria calculation
procedures described in the National Guidelines. Therefore, considering the
low levels at which adverse effects have been observed, the lack of data
showing no effect below these levels, and the importance of the species
affected, a lower criterion must be established for TBT.
The National Guidelines require that a criterion be consistent with
sound scientific evidence, based on all available pertinent laboratory and
field information. The available information for TBT indicates that it causes
imposex to occur in saltwater snails at concentrations less than 0.003 ug/L.
42
-------
Considering that less than 0.003 ug/L is an effect level and the weight of
evidence for multiple adverse effects, EPA believes that a Final Chronic Value
for TBT in saltwater of 0.001 ug/L is likely to be protective in most
situations.
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-day average concentration of tributyltin
does not exceed 0.063 • g/L more than once every three years on the average and
if the one-hour average concentration does not exceed 0.46 •g/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-day average concentration of tributyltin
does not exceed 0.001 •g/L more than once every three years on the average and
if the one-hour average concentration does not exceed 0.38 •g/L more than once
every three years on the average.
Implementation
As discussed in the Water Quality Standards Regulation (U.S. EPA 1983)
and the Foreword of this document, 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
1987,1994). Water quality criteria adopted in state water quality standards
43
-------
could have the same numerical values as criteria developed under Section 304,
of the Clean Water Act. 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. Alternatively, states
may use different data and assumptions than EPA in deriving numeric criteria
that are scientifically defensible and protective of designated uses. State
water quality standards include both numeric and narrative criteria. A state
may adopt a numeric criterion within its water quality standards and apply it
either state-wide to all waters designated for the use the criterion is
designed to protect or to a specific site. A state may use an indicator
parameter or the national criterion, supplemented with other relevant
information, to interpret its narrative criteria within its water quality
standards when developing NPDES effluent limitations under 40 CFR
122.44(d)(1)(vi).2
Site-specific criteria may include not only site-specific criterion
concentrations (U.S. EPA 1994), 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 is provided, site-specific and/or
pollutant-specific concentrations, 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 model. Although dynamic models are preferred for the application
of these criteria (U.S. EPA 1991), limited data or other considerations might
44
-------
require the use of a steady-state model (U.S. EPA 1986).
Guidance on mixing zones and the design of monitoring programs is also
available (U.S. EPA 1987, 1991) .
45
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Table 1. Acute Toxicity of Tributyltin to Aquatic Animals
Species
Hydra,
Hydra littoralis
Hydra,
Hydra littoralis
Hydra,
Hydra oliqactis
Hydra,
Chlorohydra viridissmia
Annelid (9 mg) ,
Lumbriculus varieqatus
Freshwater clam,
(113 mm TL; 153 g)
Elliptio complanatus
Cladoceran,
Daphnia maqna
Cladoceran (adult) ,
Daphnia maqna
Cladoceran (<24 hr) ,
Daphnia maqna
Cladoceran (<24 hr) ,
Daphnia maqna
Cladoceran (<24 hr) ,
Daphnia maqna
Cladoceran (<24 hr) ,
Daphnia maqna
Amphipod,
Gammarus pseudolimnaeus
Hardness
(mg/L as
Method8 Chemical13 CaCOJ
FRESHWATER SPECIES
S,M TBTO 100
(97.5%)
S,M TBTO 120
(97.5%)
S,M TBTO 100
(97.5%)
S,M TBTO 120
(97.5%)
F,M TBTO 51.8
(96%)
S,U TBTO
(95%)
S,U TBTO
S,U TBTC1
S,U TBTO
(95%)
R,M TBTO 172
(97.5%)
F,M TBTO 51.5
(96%)
S,U TBTC1 250
F,M TBTO 51.8
(96%)
LC50
or EC50
(uq/L)c
1. 11
1.30
1. 14
1. 80
5.4
24,600
66 .3
5.26
1.58
11.2
4.3
18
3.7
Species mean
Acute value
(uq/L)
1.201
1.14
1.80
5.4
24,600
-
-
_
-
4 .3
-
3 .7
References
TAI Environmental
Sciences, Inc. 1989a
TAI Environmental
Sciences, Inc. 1989b
TAI Environmental
Sciences, Inc. 1989a
TAI Environmental
Sciences, Inc. 1989b
Brooke et al. 1986
Buccafusco 1976a
Foster 1981
Meador 1986
LeBlanc 1976
ABC Laboratories,
Inc. 1990c
Brooke et al. 1986
Crisinel et al. 1994
Brooke et al. 1986
-------
Mosquito (larva), S,M TBTO 51.5 10.2 10.2 Brooke et al. 1986
Culex sp. (96%)
-------
Table 1. (continued)
Species
Rainbow trout,
(45 mm TL; 0.68 g)
Oncorhynchus mykiss
Rainbow trout
(juvenile) ,
Oncorhynchus mykiss
Rainbow trout (1.47 g) ,
Oncorhynchus mykiss
Rainbow trout (1.4 g) ,
Oncorhynchus mykiss
Lake trout (5.94 g) ,
Salvelinus naymaycush
Fathead minnow
(juvenile) ,
Pimephales promelas
Channel catfish,
(65 mm TL; 1.9 g)
Ictalurus punctatus
Channel catfish
(juvenile) ,
Ictalurus punctatus
Bluegill,
Lepomis macrochirus
Bluegill,
(36 mm TL: 0.67 g) ,
Lepomis macrochirus
Bluegill (1.01 g) ,
Lepomis macrochirus
Hardness LC50
(mg/L as or EC50
Method3 Chemical13 CaCO,) C«t/L)c
S,U TBTO - 6.5
(95%)
F,M TBTO 50.6 3.9
(96%)
F,M TBTO 135 3.45
(97%)
F,M TBTO 44 7.1
(97.5%)
F,M TBTO 135 12.73
(97%)
F,M TBTO 51.5 2.6
(96%)
S,U TBTO - 11.4
(95%)
F,M TBTO 51.8 5.5
(96%)
S,U TBTO - 227.4
S,U TBTO - 7.2
(95%)
F,M TBTO 44 8.3
(97.5%)
Species Mean
Acute Value
(•ft(/L) References
Buccafusco et al.
1978
Brooke et al. 1986
Martin et al. 1989
4.571 ABC Laboratories,
Inc. 1990a
12.73 Martin et al. 1989
2.6 Brooke et al. 1986
Buccafusco 1976a
5.5 Brooke et al. 1986
Foster 1981
Buccafusco 1976b
8.3 ABC Laboratories,
Inc. 1990b
SALTWATER SPECIES
-------
Walsh et al. 1986b
Lugworm (larva), S,U TBTO 28C «S-4
Arenicola cristata
Lugworm (larva), S,U TBTA 28 «5-10
Arenicola cristata
Table 1. (Continued)
LC50
Salinity or EC50
Species Method3 Chemical13 (q/kq) (uq/L) c
Polychaete (juvenile), S,U TBTO 33-34 6.812
Neanthes arenaceodentata
Polychaete (adult), S,U TBTO 33-34 21.418
Neanthes arenaceodentata
Polychaete (adult), R,M TBTC1 28.5 25
Armandia brevis (96%)
Blue mussel (larva), R, - TATO - 2.238
Mytilus edulis
Blue mussel (adult), R, - TBTO - 36.98e
Mytilus edulis
Blue mussel (adult), S,U TBTO 33-34 34 . 06e
Mytilus edulis
Pacific oyster (larva), R, - TBTO - 1.557
Crassostrea qiqas
Pacific oyster (adult), R, - TBTO - 282. 2e
Crassostrea qiqas
Eastern oyster (embryo), S,U TBTO 22 0.8759
Crassostrea virqinica
Eastern oyster (embryo), R,U TBTC1 18-22 1.30
Crassostrea virqinica
Eastern oyster (embryo), R,U TBTC1 18-22 0.71
• 5.03 Walsh et al .
Species Mean
Acute Value
(uq/L) Reference
Salazar and
1989
6.812 Salazar and
1989
25 Meador 1997
Thain 1983
Thain 1983
2 .238 Salazar and
1989
Thain 1983
1.557 Thain 1983
1986b
Salazar
Salazar
Salazar
EG&G Bionomics
1976a, 1977
Roberts 1987
Roberts 1987
Crassostrea virqinica
-------
Eastern oyster,
Crassostrea virqinica
European flat oyster
(adult),
Ostrea edulis
Atlantic dogwhinkle
(<24 hr-old),
Nucella lapillus
R,U
R, -
R,M
TBTC1
TBTO
TBTO
18-22
34-35
3.96e
204.4
72 .7
0.9316
204.4
72.7
Roberts 1987
Thain 1983
Harding et al. 1996
Table 1. (Continued)
Species
Hard clam
(post larva) ,
Mercenaria mercenaria
Hard clam (embryo) ,
Mercenaria mercenaria
Hard clam (larva) ,
Mercenaria mercenaria
Copepod (juvenile) ,
Eurytemora af finis
Copepod (subadult) ,
Eurytemora affinis
Copepod (subadult) ,
Eurytemora affinis
Copepod (adult) ,
Acartia tonsa
Copepod (subadult) ,
Acartia tonsa
Copepod (10-12-d-old) ,
Acartia tonsa
Methoda
S,U
R,U
R,U
F,M
F,M
F,M
R,U
F,M
S,U
Salinity
Chemical" (q/kq)
TBTC1
TBTC1 18-22
TBTC1 18-22
TBTC1 10.6
TBT 10
TBT 10
TBTO
(95%)
TBT 10
TBTC1 18
(99.3%)
LC50
or EC50
0.014661
1. 13
1.65
2.2
2.5
1.4
0.6326
1.1
0.47
Species Mean
Acute Value
(uq/L) References
Becerra-Huencho 1
Roberts 1987
1.365 Roberts 1987
Hall et al. 1988a
Bushong et al .
1987;1988
1.975 Bushong et al .
1987;1988
U'ren 1983
1 . 1 Bushong et al .
1987;1988
Kusk and Petersen
1997
-------
Copepod (10-12-d-old) ,
Acartia tonsa
Copepod (adult) ,
Nitocra spinipes
Copepod (adult) ,
Nitocra spinipes
Mysid (juvenile) ,
Acanthomysis sculpta
Mysid (adult) ,
Acanthomysis sculpta
Mysid (juvenile),
Acanthomysis sculpta
Mysid (juvenile) ,
Metamysidopsis elonqata
Table 1. (Continued)
Species
Mysid (subadult),
Metamysidopsis elonqata
Mysid (adult) ,
Metamysidopsis elonqata
Mysid (adult) ,
Metamysidopsis elonqata
Mysid (<1 day) ,
Mysidopsis bahia
Mysid (5 day) ,
Mysidopsis bahia
Mysid (10 day) ,
Mysidopsis bahia
Amphipod (subadult) ,
Gammarus sp .
Amphipod (adult) ,
Gammarus sp .
S,U TBTC1 28 0.24
(99.3%)
S,U TBTF 7 1.877
S,U TBTO 7 1.946
R,M 9 - 0.42
F,M 9 - 1.68e
F,M 9 - 0.61
S,U TBTO 33-34 <0.9732
LC50
Salinity or EC50
Method3 Chemical13 (q/kq) (uq/L) c
S,U TBTO 33-34 1.9468
S,U TBTO 33-34 2.433e
S,U TBTO 33-34 6.812e
F,M TBTC1 19-22 1.1
F,M TBTC1 19-22 2.0
F,M TBTC1 19-22 2.2
F,M TBT 10 1.3
F,M TBT 10 5.3e
Kusk and Petersen
1997
Linden et al. 1979
1.911 Linden et al. 1979
Davidson et al .
1986a, 1986b
Valkirs et al . 1985
0.61 Valkirs et al . 1985
Salazar and Salazar
1989
Species Mean
Acute Value
(uq/L) Reference
Salazar and Salazar
1989
Salazar and Salazar
1989
<0.9732 Salazar and Salazar
1989
Goodman et al . 1988
Goodman et al . 1988
1.692 Goodman et al . 1988
Bushong et al . 1988
1.3 Bushong et al . 1988
-------
Amphipod (adult), R,M
Orchestia traskiana
Amphipod (adults), R,M
Rhepoxynius abronius
Amphipod (3-5 mm; 2-5 R,M
mg) ,
Eohaustorius estuarius
Amphipod (adult) , R,M
Eohaustorius
Washington! anus
Grass shrimp (adult), F,U
Palaemonetes pugio
Grass shrimp (subadult) , F,M
Palaemonetes sp .
Grass shrimp (larvae), R,U
Palaemonetes sp .
Grass shrimp (adult), R,U
Palaemonetes sp .
Table 1. (Continued)
Species Methoda
American lobster R,U
(larva) ,
Homarus americanus
Shore crab (larva) , R, -
Carcinus maenas
Mud crab (larva), R,U
Rhi thropanopeus harrisii
Mud crab (larva), R,U
Rhi thropanopeus harrisii
Shore crab (larva), R,U
Hemigrapsus nudus
Amphioxus, F,U
TBTO 30 >14.60h
TBTC1 32.3 108
(96%)
TBTC1 28.8-29.5 10
(96%)
TBTC1 32.7 9
(96%)
TBTO - 20
TBT 10 >31
TBTO 20 4.07
TBTO 20 31.41e
LC50
Salinity or EC50
Chemical6 (g/kg) (ug/L) c
TBTO 32 1.74511
TBTO - 9.732
TBTS 15 >24.3h
TBTO 15 34.9011
TBTO 32 83.2811
TBTO - <10
>14.60 Laughlin et al . 1982
108 Meador 1997
10 Meador 1993; Meador
et al . 1993; Meador
1997
9 Meador 1997
20 Clark et al . 1987
>31 Bushong et al . 1988
Kahn et al . 1993
4.07 Kahn et al . 1993
Species Mean
Acute Value
(ug/L Reference
1.745 Laughlin and
1980
9.732 Thain 1983
Laughlin et
34.90 Laughlin et
83.28 Laughlin and
1980
<10 Clark et al .
French
al. 1983
al. 1983
French
1987
Branchiostoma caribaeum
-------
Chinook salmon
(juvenile),
Oncorhynchus tshawytscha
S,M
TBTO
28
1.460
1.460
Short and Thrower
1986b;1987
Atlantic menhaden
(juvenile) ,
Brevoortia tyrannus
Atlantic menhaden
(juvenile) ,
Brevoortia tyrannus
Sheepshead minnow
(juvenile) ,
Cyprinodon varieqatus
Sheepshead minnow
(juvenile) ,
Cyprinodon varieqatus
Sheepshead minnow
(juvenile) ,
Cyprinodon varieqatus
Sheepshead minnow
(33-49 mm) ,
Cyprinodon varieqatus
Table 1. (Continued)
Species
Sheepshead minnow
(juvenile) ,
Cyprinodon varieqatus
Sheepshead minnow
(subadult) ,
Cyprinodon varieqatus
Mummichog (adult) ,
Fundulus heteroclitus
Mumichog (juvenile) ,
F,M TBT 10 4.7 - Bushong et al .
1987;1988
F,M TBT 10 5.2 4.944 Bushong et al .
1987;1988
S,U TBTO 20 16.54 - EG&G Bionomics
S,U TBTO 20 16.54 - EG&G Bionomics
S,U TBTO 20 12.65 - EG&G Bionomics
F,M TBTO 28-32 2.31511 - EG&G Bionomics
LC50 Species Mean
Salinity or EC50 Acute Value
Method3 Chemical13 (q/kq) (uq/L) c (uq/L) Reference
F,M TBTO 15 12.31 - Walker 1989a
F,M TBT 10 25.9 9.037 Bushong et al .
S,U TBTO 25 23.36 - EG&G Bionomics
(95%)
F,M TBTO 2 17.2 - Pinkney et al .
1979
1979
1979
1981d
1988
1976a
1989
Fundulus heteroclitus
-------
Mummichog (larval),
Fundulus heteroclitus
F,M
TBT
10
23 .4
Bushong et al. 1988
Mummichog (subadult),
Fundulus heteroclitus
F,M
TBT
10
21.34
Bushong et al. 1988
Inland silverside
(larva) ,
Menidia beryllina
Atlantic silverside,
Menidia menidia
Starry flounder
(
-------
Table 2. Chronic Toxicity of Tributyltin to Aquatic Animals.
Species Testa
Cladoceran, LC
Daphnia maqna
Cladoceran, LC
Daphnia maqna
Fathead minnow, ELS
Pimephales promelas
Atlantic dogwhinkle, ELS
Nucella lapillus
Copepod, LC
Eurytemora af finis
Copepod, LC
Eurytemora af finis
Mysid, LC
Acanthomysis sculpta
Chemical13
TBTO
(96%)
TBTO
(100%)
TBTO
(96%)
TBTO
(97%)
TBTC1
TBTC1
d
Hardness Chronic
(mg/L as Limits Chronic Value
CaCO,) (uq/L) c (uq/L) References
FRESHWATER SPECIES
51.5 0.1-0.2 0.1414 Brooke et al . 1986
160-174 0.19-0.34 0.2542 ABC Laboratories, Inc.
1990d
51.5 0.15-0.45 0.2598 Brooke et al . 1986
SALTWATER SPECIES
34-35 0.0077- 0.0153 Harding et al . 1996
0. 0334£
10. 3e <0.088 <0.088 Hall et al. 1987;1988a
14. 6e 0.094-0.224 0.145 Hall et al. 1987;1988a
0.09-0.19 0.1308 Davidson et al.
1986a, 1986b
a LC = Life-cycle or partial life-cycle; ELS = early life-stage.
b TBTO = tributyltin oxide; TBTC1 = tributyltin chloride. Percent purity is given in parentheses when available.
c Measured concentrations of the tributyltin cation.
a The test organisms were exposed to leachate from panels coated with antifouling paint containing a tributyltin polymer and
cuprous oxide. Concentrations of TBT were measured and the authors provided data to demonstrate the similar
toxicity of a pure TBT compound and the TBT from the paint formulation.
e Salinity (g/kg).
£ TBT concentrations are those reported by Bailey et al. (1991) . See text for explanation.
-------
Table 2. (continued)
Acute-Chronic Ratios
Hardness
(mg/L as
Species CaCO,)
Cladoceran, 51.5
Daphnia maqna
Cladoceran, 160-174
Daphnia maqna
Fathead minnow, 51.5
Pimephales promelas
Copepod,
Eurytemora affinis
Copepod,
Eurytemora affinis
Mysid,
Acanthomysis sculpta
Acute Value
(uq/L)
4.3
11.2
2.6
2.2
2.2
0.61a
Chronic Value
(uq/L)
0.1414
0.2542
0.2598
<0.088
0.145
0.1308
Ratio
30.41
44. 06
10. 01
>25.00
15. 17
4.664
Snail,
Nucella lapillus
34-35b
72. 7
a Reported by Valkirs et al. (1985)
b Salinity (g/kg).
0.0153
4752
-------
Table 3. Ranked Genus Mean Acute Values with Species Mean Acute-Chronic Ratios
Rank3
Genus Mean
Acute Value
(uq/L)
Species
Species Mean
Acute Value
(uq/L)b
Species Mean
Acute-Chronic
Ratioc
FRESHWATER SPECIES
12
11
10
24,600
12.73
10.2
5.5
5.4
4.571
4.3
3.7
1.80
1.170
Freshwater clam,
Elliptio camplanatus
Lake trout,
Salvelinus naymaycush
Mosquito,
Culex sp.
Bluegill,
Lepomis macrochirus
Channel catfish,
Ictalurus punctatus
Annelid,
Lumbriculus varieqatus
Rainbow trout,
Oncorhyncus mykiss
Cladoceran,
Daphnia maqna
Amphipod,
Gammarus pseudolimnaeus
Fathead minnow,
Pimephales promelas
Hydra,
Chlorohydra viridissmia
Hydra,
Hydra littoralis
Hydra,
Hydra oliqactis
24,600
12.73
10.2
5.5
5.4
4.571
4.3
3.7
1.80
1.201
1.14
36.60
10.01
-------
Table 3. (Continued)
Rank3
Genus Mean
Acute Value
(uq/L)
Species
Species Mean
Acute Value
(uq/L)b
Species Mean
Acute-Chronic
Ratioc
SALTWATER SPECIES
30
29
28
27
26
25
24
23
22
21
20
19
18
204.4
108
83.28
72 . 7
34.90
25
24.90
21.34
>14.60
10.1
9.732
1.534
European flat oyster,
Ostrea edulis
Amphipod,
Rhepoxynius abronius
Shore crab,
Hemiqrapsus nudus
Atlantic dogwhinkle,
Nucella lapillus
Mud crab,
Rhithropanopeus harrisii
Polychaete,
Armandia brevis
Grass shrimp,
Palaemonetes puqio
Grass shrimp,
Palaemonetes sp.
Mummichog,
Fundulus heteroclitus
Amphipod,
Orchestia traskiana
Starry flounder,
Platichthys stellatus
Amphioxus
Branchiostoma caribaeum
Shore crab,
Carcinus maenas
Amphipod,
Eohaustorius estuarius
204.4
108
83.28
72 . 7
34.90
25
20
21.34
>14.60
10. 1
9.732
10. 1
4752
-------
Table 3.
(Continued)
Amphipod,
Eohaustorius
washinqtonianus
Rank3
17
16
15
14
13
12
11
10
Genus Mean
Acute Value
(uq/L)
6.812
5.167
•5.0
4.944
2.238
1.975
1.911
1.745
1.692
1.460
1.365
1.3
Species
Sheepshead minnow,
Cyprinodon varieqatus
Polychaete,
Neanthes arenacedentata
Inland silverside,
Menidia beryllina
Atlantic silverside,
Menidia menidia
Lugworm,
Arenicola cristata
Atlantic manhaden,
Brevoortia tyrannus
Blue mussel,
Mytilus edulis
Copepod,
Eurytemora affinis
Copepod,
Nitocra spinipes
American lobster,
Homarus americanus
Mysid,
Mysidopsis bahia
Chinook salmon,
Oncorhynchus tshawytscha
Hard clam,
Mercenaria mercenaria
Amphipod,
Gammarus sp.
Species Mean
Acute Value
(uq/L)b
Species Mean
Acute-Chronic
Ratio"
6.812
3.0
• 5.0
4.944
2.238
1.975
1.911
1.745
1.692
1.460
1.365
1.3
15.17
-------
4 1.204 Pacific oyster, 1.557
Crassostrea qiqas
Eastern oyster, 0.9316
Crassostrea virqinica
3 1.1 Copepod, 1.1
Acartia tonsa
Table 3.
(Continued)
Genus Mean Species Mean Species Mean
Acute Value Acute Value Acute-Chronic
Ranka (uq/L) Species (uq/L)b Ratio0
2 <0.9732 Mysid, <0.9732a
Metamysidopsis elonqata
1 0.61 Mysid, 0.61 4.664
Acanthomysis sculpta
Ranked from most resistant to most sensitive based on Genus Mean Acute Value. Inclusion of "greater than" value does not
necessarily imply a true ranking, but does allow use of all genera for which data are available so that the Final Acute
Value is not unnecessarily lowered.
From Table 1.
From Table 2.
This was used as a quantitative value, not as a "less than" value in the calculation of the Final Acute Value.
-------
Fresh Water
Final Acute Value = 0.9177 • g/L
Criterion Maximum Concentration = (0.9177 • g/L)/2 = 0.4589 • g/L
Final Acute-Chronic Ratio = 12.69 (see text)
Final Chronic Value = (0.9177 «g/L)/12.69 = 0.0723 ug/L
Salt Water
Final Acute Value = 0.7673 ug/L
Criterion Maximum Concentration = (0.7673 ug/L)/2 = 0.3836 ug/L
Final Acute-Chronic Ratio = 12.69 (see text)
Final Chronic Value = (0.7673 ug/L)/12.69 = 0.0605 ug/L
Final Chronic Value = 0.010 ug/L (lowered to protect growth of commercially important molluscs and survival of
the ecologically important copepod Acartia tonsa; see text)
-------
Table 4. Toxicity of Tributyltin to Aquatic Plants
Alga,
Bumilleriopsis
filiformis
Alga,
Klebsormidium marinum
Alga,
Monodus subterraneus
Alga,
Raphidonema lonqiseta
Alga,
Tribonema aequale
Blue-green alga,
Oscillatoria sp.
Blue-green alga,
Synechococcus
leopoliensis
Green alga,
Chlamydomonas dysosmas
Green alga,
Chlorella emersonii
Chemical3
TBTC1
TBTC1
TBTC1
TBTC1
TBTC1
TBTC1
TBTC1
TBTC1
TBTC1
Hardness
(mg/L as Duration
CaCO,) (days) Effect
FRESHWATER SPECIES
14 No growth
14 No growth
14 No growth
14 No growth
14 No growth
14 No growth
14 No growth
14 No growth
14 No growth
Concentration
(uq/L)b Reference
111.4 Blanck
Blanck
1984
222.8 Blanck
Blanck
1984
1,782.2 Blanck
Blanck
1984
56.1 Blanck
Blanck
1984
111.4 Blanck
Blanck
1984
222.8 Blanck
Blanck
1984
111.4 Blanck
Blanck
1984
111.4 Blanck
Blanck
1984
445.5 Blanck
Blanck
1984
1986;
et al
1986;
et al
1986;
et al
1986;
et al
1986;
et al
1986;
et al
1986;
et al
1986;
et al
1986;
et al
-------
Green alga, TBTC1
Kirchneriella contorta
Green alga, TBTC1
Monoraphidium pusilium
Green alga, TBTC1
Scenedesmus
obtusiusculus
Green alga, TBTC1
Scenedesmus
quadricauda
Green alga, TBTO
Scenedesmus
quadricauda
Table 4. (Continued)
14
14
14
12
No growth
No growth
No growth
Reduced
growth
(87.6%)
EC50
chlorophyll
production
111.4
111.4
445.5
5.0
Blanck 1986;
Blanck et al.
1984
Blanck 1986;
Blanck et al.
1984
Blanck 1986;
Blanck et al.
1984
Fargasova and
Kizlink 1996
Fargasova 1996
Hardness
(mg/L as
Species Chemical3 CaCO, )
Green alga, TBTO 0.67
Scenedesmus
quadricauda
Green alga, TBT 72.7
Scenedesmus obliquus
Green alga, TBTC1
Selenastrum
capricornutum
Green alga, TBTC1
Selenastrum
capricornutum
Duration
(days) Effect
12 Reduced
growth
87.6%
95.9%
100%
4 EC50 (reduced
growth)
14 No growth
4 EC50
Concentration
(uq/L)b
1
10
100
3.4
111.4
12.4
Reference
Fargasova and
Kizlink 1996
Huang et al . 1993
Blanck 1986;
Blanck et al .
1984
Miana et al . 1993
SALTWATER SPECIES
-------
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Nitzschia sp .
Flagellate alga,
Dunaliella tertiolecta
Mixed algae,
Dunaliella salina and
D. viridis
TBTO
TBTO 30C
(BioMet
Red)
TBTO 30C
TBTO
TBTO
TBT
5 Algistatic
Algicidal
14 EC50 (dry
cell weight)
14 EC50 (dry
cell weight)
8 EC50 (reduced
growth)
8 EC50 (reduced
growth)
4 EC50 (reduced
growth)
0.9732-17.52
>17.52
>0.1216; <0.2433
0.06228
1.19
4.53
0.68
Thain 1983
EG&G Bionomics
1981c
EG&G Bionomics
1981c
Delupis et al .
1987
Delupis et al .
1987
Huang et al . 1993
a TBTC1 = tributyltin chloride; TBTO = tributyltin oxide. Percent purity is given in parentheses when available.
b Concentration of the tributyltin cation, not the chemical. If the concentrations were not measured and the published
results were not reported to be adjusted for purity, the published results were multiplied by the purity if it was
reported to be less than 95%.
c Salinity (g/kg).
-------
Table 5. Bioaccumulation of Tributyltin by Aquatic Organisms
Chemical3
Zebra mussel
(1.76±0.094 cm),
Dreissena
polymorpha
Rainbow trout (13.8
9>,
Oncorhynchus mykiss
Rainbow trout (32.7
9>,
Oncorhynchus mykiss
Carp (9.5-11.5 cm;
20.0-27.5 g);
Cyprinus carpio
Carp (8.5-9.5 cm;
16.5-22.1 g) ;
Cyprinus carpio
TBT
TBTO
(97%)
TBTO
(97%)
TBTO
TBTO
Hardness
(mg/L as Concentration Duration
(CaCo,) in Water (days) Tissue
FRESHWATER SPECIES
0.0703 105 Soft parts
135 0.513 64 Whole body
135 1.026 15 Liver
Gall
bladder/bil
e
Kidney
Carcass
Peritoneal
fat
Gill
Blood
Gut
Muscle
2.1 14 Muscle
34.5-39.0 1.8 (pH = 14 Whole body
6.0)
1.6 (pH =
BCF or
BAFC Reference
17,483d Becker-van
Slooten and
Tarradellas 1994
406 Martin et al .
1989
1, 179 Martin et al .
1989
331
2,242
1,345
5,419
1, 014
653
487
312
501.2 Tsuda et al.
1988a
• 1190 Tsuda et al.
•1523 1990a
• £250
6.8)
1.7 (pH
7.8)
-------
Goldfish (3.5-4.0
cm; 1.6-2.9 g) ;
Carassius auratus
Guppy (2.4-2.7 cm;
0.41-0.55 g);
Poecilia
reticulatus
TBTC1
TBTO
(95%)
36
0.13
0.54
28
14
Whole body
Whole body
1, 976
460
Tsuda et al.
1991b
Tsuda et al.
1990b
Table 5.
(Continued)
Snail (adults),
Littorina littorina
Atlantic dogwhinkle
(female),
Nucella lapillus
Atlantic dogwhinkle
(female),
Nucella lapillus
Atlantic dog
whinkle (18-22 mm),
Nucella lapillus
Atlantic dogwhinkle
(1 year-old),
Nucella lapillus
Atlantic dogwhinkle
(1 year-old),
Nucella lapillus
Chemical3
TBTC1
TBT
Field
TBTC1
TBTO
TBTO
Salinity
(q/kq)
35
34-35
34-35
Concentration Duration
in Water (days)
(uq/L)b
SALTWATER SPECIES
Tissue
BCF or
BAFC
Reference
0.488
0.976
0.0038 to
0.268
182
182
249 to 408
Soft parts
Soft parts
Soft parts
1,420
1, 020
11,000 to
38, 000
Bauer
1997
Bryan
1987a
et al
et al
0
0.
0.
0.
0.
0.
0.
0.
0.
0.
. 070
0205
0027
0077
0334
1246
0026
0074
0278
1077
529 to 634
49
365
365
365
365
365
365
365
365
Soft
Soft
Soft
Soft
Soft
Soft
Soft
Soft
Soft
Soft
parts
parts
parts
parts
parts
parts
parts
parts
parts
parts
17
30
18
21
16
7
<
10
8
6
, 000
, 000
, 727
, 964
, 756
,625
7782
, 121
, 088
, 172
Bryan et
1987a
Bryan et
1989b
al.
al.
Bailey et al .
1991
Harding
1996
et al
-------
Blue mussel (spat) ,
Mytilus edulis
Blue mussel Field
(adult) ,
Mytilus edulis
Blue mussel Field
(juvenile) ,
Mytilus edulis
Blue mussel,
Mytilus edulis
Blue mussel Field
(juvenile) ,
Mytilus edulis
Table 5 .
(Continued)
Species Chemical8
Blue mussel TBTC1
(3.0 - 3.5 cm) ,
Mytilus edulis
Pacific oyster, TBTO
Crassostrea qiqas
American Oyster (6-
9 cm length) ,
Crassostrea
virqinica
28.5-34.2 0.24 45 Soft parts
<0.1 60
<0.1 60
0.452 56 Soft parts
0.204
0.204
0.079
<0.105 84 Soft parts
Salinity Concentration Duration
(q/kq) in Water (days) Tissue
(uq/L)b
25.1-26.3 0.020 60 Muscle and
mantle
Muscle and
mantle
28-31.5 1.216 21 Soft parts
18±1 0.283 28 Soft parts
6 , 833£ Thain and
Waldock 1985;
Thain 1986
11,000 Salazar and
Salazar 1990a
25,000 Salazar and
Salazar 1990a
23,000 Salazar et al .
27,000 1987
10,400
37, 500
5,000- Salazar and
60,000 Salazar, 1991
BCF or
BAFC Reference
7,700 Guolan and Yong
1995
11, 000
1, 874£ Waldock et al .
1983
15,460 Roberts et al .
1996
Pacific oyster,
Crassostrea qiqas
TBTO
28-31.5
0.1460
21
Soft parts 6,047£
Waldock et al.
1983
-------
Pacific oyster,
Crassostrea qiqas
Pacific oyster,
Crassostrea qiqas
Pacific oyster,
Crassostrea qiqas
Pacific oyster
(spat) ,
Crassostrea qiqas
European flat
oyster,
Ostrea edulis
European flat
oyster,
Ostrea edulis
European flat
oyster,
Ostrea edulis
European flat
oyster,
Ostrea edulis
European flat
oyster,
Ostrea edulis
Guppy ( • • 2.4-2.7
cm; 0.41-0.55 g) ;
Poecilia
reticulatus
28.5-34.2 0.24
TBTO 29-32 1.557
TBTO 29-32 0.1460
TBTO - 0.29
0.92
2.83
TBTO 28-31.5 1.216
TBTO 28-34.2 0.24
TBTO 28-34.2 2.62
28.5-34.2 0.24
d
28.5-34.2 2.62
TBTC - 0.28
(95%)
45 Soft parts 7,292£ Thain and
Waldock 1985;
Thain 1986
56 Soft parts 2,300 Waldock and
Thain 1983
56 Soft parts 11,400 Waldock and
Thain 1983
30 Soft parts 2275 Osada et al.
1369 1993
621
21 Soft parts 960f Waldock et al .
1983
75 Soft parts 875£ Waldock et al .
1983
75 Soft parts 397f Thain 1986
45 Soft parts 1,167£ Thain and
Waldock 1985;
Thain 1986
45 Soft parts 192. 3£ Thain and
Waldock 1985;
Thain 1986
14 Whole body 240 Tsuda et al.
1990b
a TBTO = tributyltin oxide; Field = field study. Percent purity is given in parentheses when available.
b Measured concentration of the tributyltin cation.
c Bioconcentration factors (BCFs) and bioaccumulation factors (BAFs) are based on measured concentrations of TBT in water
and tissue.
a BCF normalized to 1% lipid concentration and converted to wet weight estimate based upon 85% moisture.
e Test organisms were exposed to leachate from panels coated with antifouling paint containing tributyltin.
£ BCFs were calculated based on the increase above the concentration of TBT in control organisms.
-------
Table 6. Other Data on Effects of Tributyltin on Aquatic Organisms
Chemical8
Hardness
(mg/L as
CaCO,)
Duration Effect
Concentrati
on Reference
(uq/L)b
Microcosm natural
assemblage
Microcosm natural
assemblage
TBTO
TBTO
FRESHWATER SPECIES
55 days Daphnia maqna
disappeared; Ostracoda
increased; algae
increased immediately
then gradually
disappeared
24 days Metabolism reduced (2.5
days)
Metabolism returned to
normal (14.1 days)
Metabolism reduced (1
day)
Metabolism returned to
normal (16 days)
80
4.7
14. 9
Alga,
Natural assemblage
Blue-green alga,
Anabaena flos- aquae
Green alga,
Ankistrodesmus falcatus
Green alga, TBTO
Ankistrodesmus falcatus (97%)
Green alga,
Scenedesmus quadricauda
4 hr
4 hr
4 hr
7 days
14 days
21 days
2 8 days
4 hr
EC50
(production)
EC50
(production)
EC50
(production)
(reproduction)
BCF = 300
BCF - 253
BCF =448
BCF = 467
EC50
(production)
5
13
20
5
5.2
4.7
2.1
1.5
16
Delupis and
Miniero 1989
Miniero and
Delupis 1991
Wong et al. 1982
Wong et al. 1982
Wong et al. 1982
Maguire et al.
1984
Wong et al. 1982
-------
Hydra,
Hydra sp.
Rotifer,
Brachionus calyciflorus
Asiatic clam (larva),
Corbicula fluminea
Table 6. (Continued)
Cladoceran,
Daphnia maqna
Cladoceran (<24 hr),
Daphnia maqna
Cladoceran (<24 hr),
Daphnia maqna
Cladoceran (adult),
Daphnia maqna
Cladoceran (14-d-old),
Daphnia maqna
Cladoceran (<24-h old),
Daphnia maqna
Fairy shrimp (cysts),
Streptocephalus texanus
Rainbow trout
(yearling),
Oncorhynchus mykiss
Rainbow trout,
Oncorhynchus mykiss
TBTO 51.0
(96%)
TBTC1
TBTO
Hardness
(mg/L as
Chemical8 CaCCO
TBTO
TBTC
TBTO
TBTC1
TBTC1
TBTC1
TBTC1
TBTO
TBTO
200
200
150
312. I
250
96 hr EC50
(clubbed tentacles)
24 hr EC50 (hatching)
0 . 5 Brooke et al.
1986
24 hr
24 hr
24 hr
7 days
48 hr
24 hr
24 hr
48 hr
24 hr
EC50
Duration Effect
LC50
EC50
(mobility)
24 hr EC50
(mobility)
8 days Altered phototaxis
Altered behavior
Reproductive failure
Digestive storage cells
reduced
EC50
(mobility)
EC50 (hatching)
LC50
EC50
(rheotaxis)
72
Crisinel et al.
1994
1,990 Foster 1981
Concentrati
on Reference
(uq/L)b
3
11.6
13 .6
0.45
1
1
5
9.8
15
25.2
18 .9
Polster and
Halacha 1972
Vighi and
Calamari 1985
Vighi and
Calamari 1985
Meador 1986
Bodar et al .
1990
Miana et al .
1993
Crisinel et al.
1994
Alabaster 1969
Chliamovitch and
Kuhn 1977
-------
Rainbow trout
(embryo, larva),
Oncorhynchus mykiss
Rainbow trout (fry),
Oncorhynchus mykiss
Ra inbow t rout,
Oncorhynchus mykiss
TBTC1
94-102
TBTC1
TBTO
96-105
Table 6. (Continued)
Chemical3
Hardness
(mg/L as
CaCO3)
110 days 20% reduction in growth
23% reduction in
growth; 6.6% mortality
100% mortality
110 days NOEC (mortality and
growth)
LOEC (mortality and
growth)
28 days BCF = 3833 (whole body)
BCF = 2850 (whole body)
BCF = 2700 (whole body)
BCF = 1850 (whole body)
Cell necrosis within
gill lamellae
Duration Effect
Seinen et al.
1981
0. 18
0. 89
4.46
0.040
0.200
0.6
1.0
2.0
4.0
4.0
Concentrati
on Reference
(uq/L)b
de Vries et al.
1991
Schwaiger et al.
1992
Rainbow trout,
Oncorhynchus mykiss
TBTO
Rainbow trout (3 week)
Oncorhynchus mykiss
TBTO
(98%)
Goldfish (2.8-3.5 cm;
0.9-1.7 g),
Carassius auratus
TBTO
(reagent
grade)
28 days BCF = 3833 (whole body)
BCF = 2850 (whole body)
BCF = 2700 (whole body)
BCF = 1850 (whole body)
Cell necrosis within
gill lamellae
400 21 days Reduced growth
Reduced avoidance
BCF =540 (no head; no
plateau)
BCF = 990 (no head; no
plateau
14 days BCF = 1230 (no plateau)
4.0
0.5
2.0
Schwaiger et al.
1992
Triebskorn et
al. 1994
Tsuda et al.
1988b
-------
Carp (10.0-11.0 cm;
22.9-30.4 g),
Cyprinus carpio
Guppy (3-4 wk),
Poecilia reticulata
TBTO
TBTO
7 days BCF in muscle = 295
Half-life = 1.67 days
209 3 mo Thymus atrophy
Hyperplasia of kidney
hemopoietic tissue
Marked liver
vacuolation
1. 80
0.32
1.0
1.0
Tsuda et al.
1987
Wester and
Canton 1987
Guppy (4 wk),
Poecilia reticulata
Frog (embryo, larva)
Rana temporaria
TBTO
TBTO
TBTF
TBTO
TBTF
Hyperplasia of corneal
epithelium
1 mo NOEC
3 mo NOEC
5 days LC40
LC50
Loss of body water
Loss of body water
10.0
1.0
0.32
28 .4
28 .2
28 .4
28 .2
Wester and
Canton 1991
Laughlin and
Linden 1982
Table 6. (Continued)
Chemical3
Salinity
(q/kq)
Duration
Effect
Concentrati
on Reference
(uq/L)b
SALTWATER SPECIES
Natural microbial
populations
Natural microbial
populations
TBTC1
TBTC1
2 and 17
2 and 17
1 hr
(incubated
10 days)
1 hr
(incubated
10 days)
Significant decrease in 4.454
metabolism of nutrient
substrates
50% mortality 89.07
Jonas et al.
1984
Jonas et al.
1984
-------
Fouling communities
Fouling communities
Microcosm (seagrass
bed)
Microcosm (seagrass
bed)
TBT
TBTC1
Periphyton communities TBTC1
Periphyton communities TBTO
Green alga, TBTO
Dunaliella tertiolecta
Green alga, TBTO
Dunaliella sp.
Green alga, TBTO
Dunaliella sp.
Green alga, TBTO
Dunaliella tertolecta
Table 6. (Continued)
33-36 2 months Reduced species and
diversity; no effect at
0.04 ug/L
126 days No effect
21.5- 6 wks Fate of TBT
28.9 Sediments 81-88%
Plants 9-17%
Animals 2-4%
6 wks Reduced plant material
loss; loss of amphipod
Cymadusa compta
15 min EC50 (reduced
photosynthesis
15 min EC50 (reduced
photosynthesis
34-40 18 days Population growth
72 hr Approx. EC50 (growth)
72 hr 100% mortality
days EC50
0.1
Henderson 1986
0.204 Salazar et al.
1987
0.2-20 Levine et al.
1990
22.21 Kelly et al.
1990a
28.7 Blanck and Dahl
1996
27.9 Blanck and Dahl
1996
1. 0 Beaumont and
Newman 1986
1.460 Salazar 1985
2.920 Salazar 1985
4.53 Delupis et al.
1987
Species
Salinity
Chemical8 (q/kq) Duration Effect
Concentrati
on Reference
(uq/L)b
Diatom,
Phaeodoctylum
tricornutum
Diatom,
Nitzschia sp.
TBTO
TBTO
72 hr No effect on growth
days EC50
1.460-5.839 Salazar 1985
1.19 Delupis et al.
1987
-------
Diatom,
Nitzschia sp .
Diatom,
Nitzschia closterium
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Diatom,
Skeletonema costatum
Table 6 . (Continued)
Species
TBTO - 8 days EC50
TBTC1 - 7 days EC50 (growth)
TBTA 30 72 hr EC50
(population growth)
TBTA 30 72 hr LC50
TBTO 34-40 12-18 days Population growth
TBTO 30 72 hr EC50
(population growth)
TBTO 30 72 hr LC50
TBTC1 30 72 hr EC50
(population growth)
TBTC1 30 72 hr LC50
TBTF 30 72 hr EC50
(population growth)
TBTF 30 72 hr LC50
TBTC1 30.5 96 hr NOEC
TBTC1 - 7 days EC50 (growth)
Salinity
Chemical3 (q/kq) Duration Effect
1.19
1.16
0.3097
12.65
1.0
0.3212
13.82
0.3207
10.24
>0.2346,
>0.4693
11.17
3.48
Delupis et al.
1987
Nakagawa and
Saeki 1992
Walsh et al.
1985;1987
Walsh et al.
1985;1987
Beaumont and
Newman 1986
Walsh et al.
1985;1987
Walsh et al.
1985
Walsh et al.
1985;1987
Walsh et al.
1985;1987
Walsh et al.
1985;1987
Walsh et al.
1985
Reader and
Pelletier 1992
Nakagawa and
Saeki 1992
Concentrati
on Reference
(uq/L)b
-------
Diatom, TBTC1
Chaetoceros debilis
Diatom, TBTC1
Chattonella antiqua
Diatom, TBTC1
Tetraselmis tetrathele
Diatom, TBTO
Minutocellus
polymorphus
Diatom, TCTC1
Minutocellus
polymorphus
Diatom, TBTA
Thalassiosira
pseudonana
Diatom, TBTO
Thalassiosira
pseudonana
Alga, TBTO
Pavlova lutheri
Alga, TBTO
Pavlova lutheri
Dinoflagellate, TBTO
Gymnodinium splendens
Macroalgae, TBT
Fucus vesiculosus
Giant kelp (zoospores), TBT
Macrocystis pvrifera
Polychaete worm TBTC1
(juvenile), (96%)
Neanthes
arenaceodentata
Polychaete worm TBTC1
(adult) , (96%)
Armandia brevis
32-33
30
28.5
7 days EC50 (growth)
7 days EC50 (growth)
7 days EC50 (growth)
48 hr EC50
1.16
2.05
48 hr
EC50
30 72 hr EC50
(population growth)
30 72 hr EC50
(population growth)
34-40 12-26 days Population growth
16 days NOEC
LOEC
72 hr 100% mortality
7 days Photosynthesis and
nutrient uptake reduced
48 hr EC50 (germination)
EC50 (growth)
10 wks NOEC (survival)
LOEC (survival)
10 days BCF = 5,100 (no
plateau)
S40
• S30
1.101
1.002
1.0
5.36
21.46
1.460
0.6
11.256
13.629
0.100
0.500
233
Nakagawa and
Saeki 1992
Nakagawa and
Saeki 1992
Nakagawa and
Saeki 1992
Walsh et al.
1988
Walsh et al.
1988
Walsh et al.
1985
Walsh et al.
1985;1987
Beaumont and
Newman 1986
Saint-Louis et
al. 1994
Salazar 1985
Lindblad et al.
1989
Brix et al.
1994a
Moore et al.
1991
Meador 1997
-------
Rotifer (neonates),
Brachionus plicatilis
Table 6. (Continued)
TBT
15
30 min Induction of the stress
protein gene SP58
20-30
Cochrane et al.
1991
Hydroid,
Campanularia flexuosa
Pale sea anemone
(1-2 cm oral disc),
Aiptasia pallida
Sand dollar (sperm),
Dendraster excentricus
Starfish (79 g),
Leptasterias polaris
Dogwhinkle (adult),
Nucella lapillus
Dogwhinkle (adult),
Nucella lapillus
Dogwhinkle (subadult)
Nucella lapillus
Mussel (juvenile),
Mvtilus sp.
Chemical8
TBTF
TBT
TBT
TBTC1
TBTC1
TBTC1
Field
Salinity
(q/kq)
35
32-33
25.!
35
35
Duration Effect
11 days Colony growth
stimulation; no growth
28 days Reduced (90.4%)
symbiotic zooxanthellae
populations; incresed
bacterial aggregates;
fewer undischarged
nematocysts
80 min EC50 (mortality)
48 hr BCF = 41,374 (whole
body)
120 days 41% Imposex
(superimposition of
male anatomical
characteristics on
females)
6 months Imposex induced
Concentrati
on Reference
22
12 weeks
BCF = •
), 000
NOEC tissue cone.
growth = 0.5 ug/g
LOEC tissue cone.
growth = 1.5 ug/g
NOEC (growth)
LOEC (growth)
BAF = 5,000-100,000
(uq/L)b
0 . 01 Stebbing
1.0
0 . 05 Mercier
1997
0.465 Brix et
1994b
0.072 Rouleau
1995
0. 05 Bryan et
1986
• C. 012 Stroben
1992b
0.019 Bryan et
1993
Salazar
Salazar
1996
0.025
0.100
<0.105
1981
et al
al.
et al
al.
et al
al.
and
1990b
-------
Blue mussel (larva),
Mytilus edulis
Blue mussel (larva),
Mytilus edulis
Blue mussel (spat),
Mytilus edulis
Blue mussel (larva),
Mytilus edulis
Table 6. (Continued)
TBTO
TBTO
TBTO
28.5-
34.2
33
24 hr No effect on sister 1.0
chromatid exchange
4 days Reduced survival ^>0.1
45 days 100% mortality 2.6
15 days 51% mortality; reduced 0.0973
growth
Dixon and
Prosser 1986
Dixon and
Prosser 1986
Thain and
Waldock 1985;
Thain 1986
Beaumont and
Budd 1984
Salinity
Chemical3 (q/kq) Duration Effect
Concentrati
on Reference
(uq/L)b
Blue mussel (larva),
Mytilus edulis
Blue mussel (juvenile), TBTO
Mytilus edulis
Blue mussel (juvenile), TBT
Mytilus edulis (field)
Blue mussel (juvenile), TBT
Mytilus edulis (field)
45 days Reduced growth 0.24
33.7 7 days Significant reduction 0.3893
in growth
1-2 wk Reduced growth; at <0.2 0.2
ug/L environmental
factors most important
12 wks Reduced growth ^0.2
Thain and
Waldock 1986
Stromgren and
Bongard 1987
Salazar and
Salazar 1990b
Salazar and
Salazar 1988
Blue mussel (juvenile)
Mytilus edulis
Blue mussel (juvenile)
Mytilus edulis
Blue mussel (juvenile)
Mytilus edulis
Blue mussel (juvenile)
Mvtilus edulis
TBT
(field
12 wks Reduced growth at
tissue cone, of 2.0
ug/g
56 days Reduced condition
196 days Reduced growth (no
effect at day 56 of 0.2
ug/L)
56 days No effect on growth
Salazar and
Salazar 1988
0.157 Salazar et al.
1987
0.070 Salazar and
Salazar 1987
0.160 Salazar and
Salazar 1987
-------
Blue mussel
(2.5 to 4.1 cm) ,
Mytilus edulis
Blue mussel
(2.5 to 4.1 cm),
Mytilus edulis
Blue mussel
(juveniles and adults)
Mytilus sp.
Blue mussel (3.0-3.5
cm) ,
Mytilus edulis
TBT
(field)
TBT
66 days LC50
66 days Significant decrease in
shell growth
84 days BCF
2 days Reduced ability to
survive anoxia
0.97 Valkirs et al.
1985;1987
0.31
3,000-
100,000
Valkirs et al.
1985
Salazar 1996
Wang et al. 1992
Table 6. (Continued)
Salinity
Chemical8 (q/kq) Duration Effect
Concentrati
on Reference
(uq/L)b
Blue mussel (4 cm)
Mytilus edulis
Blue mussel (8-d-old
larvae),
Mytilus edulis
Scallop (adult),
Hinnites multiruqosus
Pacific oyster (larva)
Crassostrea qiqas
Pacific oyster (larva)
Crassostrea qiqas
Pacific oyster (spat),
Crassostrea qiqas
TBTC1
TBT
TBTO
2.5 days Increased respiration
0.15 ug/g tissue
Reduced food absorption
efficiency 10 ug/g
33 days NOEC (growth) 0.006
LOEC (growth) 0.050
110 days No effect on condition 0.204
30 days 100% mortality 2.0
113 days 30% mortality and 0.2
abnormal development
48 days Reduced growth 0.020
Widdows and Page
1993
Lapota et al.
1993
Salazar et al.
1987
Alzieu et al.
1980
Alzieu et al.
1980
Lawler and
Aldrich 1987
-------
Pacific oyster (spat), TBTO
Crassostrea qiqas
Pacific oyster (spat), 28.5-
Crassostrea qiqas 34.2
Pacific oyster (spat), 28.5-
Crassostrea qiqas 34.2
Pacific oyster (spat),
Crassostrea qiqas
Pacific oyster (spat), TBT
Crassostrea qiqas
Pacific oyster (spat), TBTO 29-32
Crassostrea qiqas
Pacific oyster (spat), TBTO 29-32
Crassostrea qiqas
Pacific oyster (adult), TBT
Crassostrea qiqas (field)
14 days Reduced oxygen
consumption and feeding
rates
45 days 40% mortality; reduced
growth
45 days 90% mortality
45 days Reduced growth
49 days Shell thickening
56 days No growth
56 days Reduced growth
Shell thickening
0.050 Lawler and
Aldrich 1987
0.24 Thain and
Waldock 1985
2 . 6 Thain and
Waldock 1985
0.24 Thain and
Waldock 1986
0.020 Thain et al.
1987
1.557 Waldock and
Thain 1983
0.1460 Waldock and
Thain 1983
_>0.014 Wolniakowski et
al. 1987
Table 6. (Continued)
Salinity
Chemical8 (q/kq) Duration Effect
Concentrati
on Reference
(uq/L)b
Pacific oyster (larva), TBTF 18-21 21 days
Crassostrea qiqas
Pacific oyster (larva), TBTF 18-21
Crassostrea qiqas
Pacific oyster TBTA 28 24 hr
(embryo),
Crassostrea qiqas
Reduced number of
normally developed
larvae
15 days 100% mortality
Abnormal development;
30-40% mortality
0.02346
0.04692
4.304
Springborn
Bionomics, Inc.
1984a
Springborn
Bionomics, Inc.
1984a
His and Robert
1980
-------
Pacific oyster TBTA
(embryo),
Crassostrea qiqas
Pacific oyster (larva), TBTA
Crassostrea qiqas
Pacific oyster (larva), TBTA
Crassostrea qiqas
Pacific oyster
(150-300 mg) ,
Crassostrea qiqas
Pacific oyster TBT
(3.5-25 mm), (field)
Crassostrea qiqas
Pacific oyster TBTO
(fertilized eggs),
Crassostrea qiqas
Pacific oyster TBTO
(straight-hinge
larvae),
Crassostrea qiqas
Pacific oyster TBTO
(spat) ,
Crassostrea qiqas
Pacific oyster TBTA
(24-h-old),
Crassostrea qiqas
24 hr Abnormal development
24 hr Abnormal development
48 hr 100% mortality
56 days No effect on growth
2-5 mo Reduced growth rate
Normal growth rate
24 hr LC50
Delayed development
24 hr LC50
48 hr LC50
12 days LC50
0.8604
Robert and His
1981
^0.9 Robert and His
1981
2.581 Robert and His
1981
0.157 Salazar et al.
1987
0.040 Stephanson 1991
0.010
7.0 Osada et al.
1.8 1993
15 . 0 Osada et al.
1993
35. 0 Osada et al.
1993
0.04 His 1996
Table 6. (Continued)
Salinity
Chemical8 (q/kq) Duration Effect
Concentrati
on Reference
(uq/L)b
Eastern oyster
(2.7-5.3 cm),
Crassostrea virqinica
67 days Decrease in condition
index (body weight)
0.73 Valkirs et al.
1985
-------
Eastern oyster
(2.7-5.3 cm),
Crassostrea virqinica
Eastern oyster (adult)
Crassostrea virqinica
Eastern oyster (adult)
Crassostrea virqinica
Eastern oyster
(embryo),
Crassostrea virqinica
Eastern oyster
(juvenile),
Crassostrea virqinica
Eastern oyster (adult)
Crassostrea virqinica
Eastern oyster (adult)
Crassostrea virqinica
European flat oyster
(spat) ,
Ostrea edulis
European flat oyster
(spat) ,
Ostrea edulis
European flat oyster
(spat) ,
Ostrea edulis
European flat oyster
(spat) ,
Ostrea edulis
European flat oyster
(adult),
Ostrea edulis
TBTC1
TBTO
TBT
TBTO
67 days No effect on survival
33-36 57 days Decrease in condition
index
33-36 30 days LC50
18-22 48 hr Abnormal shell
development
11-12 96 hr EC50; shell growth
8 wks No affect on sexual
development,
fertilization
21 wks Immune response not
weakened
20 days Significant reduction
in growth
30
28.5-
34.2
28.5-
34.2
45 days Decreased growth
45 days 70% mortality
20 days Reduced growth
28-34 75 days Complete inhibition of
larval production
1.89
0.1
2.5
0.77
0.31
1.142
0.1
0.01946
0.2392
2.6
0.02
0.24
Valkirs et al.
1985
Henderson 1986
Henderson 1986
Roberts 1987
Walker 1989b
Roberts et al.
1987
Anderson et al.
1996
Thain and
Waldock 1985
Thain and
Waldock 1985;
Thain 1986
Thain and
Waldock 1985;
Thain 1986
Thain and
Waldock 1986
Thain 1986
Table 6. (Continued)
-------
Chemical8
Salinity
(q/kq)
Duration Effect
Concentrati
on Reference
(uq/L)b
European flat oyster
(adult),
Ostrea edulis
European flat oyster
(adult),
Ostrea edulis
28-34 75 days Retardation of sex
change from male to
female
28-34 75 days Prevented gonadal
development
0.24 Thain 1986
2.6 Thain 1986
European flat oyster
(140-280 mg),
Ostrea edulis
Native Pacific oyster
(100-300 mg),
Ostrea luricla
Quahog clam
(embryo, larva),
Mercenaria mercenaria
Clam (adult),
Macoma nasuta
Quahog clam (veligers)
Mercenaria mercenaria
Quahog clam (post
larva),
Mercenaria mercenaria
Quahog clam (larva),
Mercenaria mercenaria
Common Pacific
Littleneck (adult),
Protothaca stamina
Copepod (subadult),
Eurytemora affinis
TBTO
TBTO
TBTO
TBTC1
TBTO
TBT
56 days No effect on growth
56 days No effect on growth
14 days Reduced growth
110 days No effect on condition
8 days Approx. 35% dead;
reduced growth; _>! . 0
ug/L 100 mortality
25 days 100% mortality
18-22 48 hr Delayed development
33-34 96 hr 100% survival
10
72 hr LC50
0.157 Salazar et al.
1987
0.157 Salazar et al.
1987
2.0.010 Laughlin et al.
1987;1988
0.204 Salazar et al.
1987
0.6 Laughlin et al.
1987;1989
10 Laughlin et al.
1987;1989
0.77 Roberts 1987
2.2.920 Salazar and
Salazar 1989
0.5 Bushong et al.
1988
-------
Copepod (subadult),
Eurytemora affinis
TBT
10
72 hr LC50
0.6 Bushong et al.
1988
Copepod,
Acartia tonsa
Table 6. (Continued)
Species
Copepod (adult) ,
Acartia tonsa
Copepod (nauplii) ,
Acartia tonsa
Copepod (nauplii) ,
Acartia tonsa
Copepod (nauplii) ,
Acartia tonsa
Copepod (nauplii) ,
Acartia tonsa
Amphipod (larva,
juvenile) ,
Gammarus oceanus
Amphipod (larva,
juvenile) ,
Gammarus oceanus
Amphipod (larva,
juvenile) ,
Gammarus oceanus
Amphipod (larva,
juvenile) ,
Gammarus oceanus
Amphipod,
Gammarus sp .
TBTO - 6 days
Salinity
Chemical8 (q/kq) Duration
TBTO 28 5 days
TBTC1 10-12 9 days
TBTC1 10-12 6 days
TBTC1 10-12 6 days
TBTC1 18 8 days
TBTO 7 8 wk
TBTF 7 8 wk
TBTO 7 8 wk
TBTF 7 8 wk
TBTC1 10 24 days
EC50
Effect
Reduced egg production
Reduced survival
Reduced survival; no
effect 0.012 ug/L
Reduced survival; no
effect 0.010 ug/L
Inhibition of
development
EC50 (survival)
100% mortality
100% mortality
Reduced survival and
growth
Reduced survival and
increased growth
No effect
0.3893
Concentrati
on
(uq/L)b
0.010
^0.029
0.023
0.024
0.003
0.015-0. 020
2 .920
2 .816
0.2920
0.2816
0.579
U'ren 1983
Reference
Johansen and
Mohlenberg 1987
Bushong et al .
1990
Bushong et al .
1990
Bushong et al .
1990
Kusk and
Peterson 1997
Laughlin et al.
1984b
Laughlin et al.
1984b
Laughlin et al.
1984b
Laughlin et al.
1984b
Hall et al.
1988b
-------
Amphipod (adult) ,
Orchestia traskiana
Amphipod (adult) ,
Orchestia traskiana
Amphipod (adult) ,
Eohaustorius estuarius
Amphipod (adult) ,
Eohaustorius
Washington! anus
Amphipod (adult) ,
Rhepoxynius abronius
Table 6 . (Continued)
Species
Grass shrimp,
Palaemonetes puqio
Grass shrimp,
Palaemonetes puqio
Mud crab (larva) ,
Rhi thropanopeus
harrisii
Mud crab (larva) ,
Rhi thropanopeus
harrisii
Mud crab (larva) ,
Rhi thropanopeus
harrisii
Mud crab (larva) ,
Rhi thropanopeus
harrisii
Mud crab (zoea) ,
Rhitropanopeus harrisii
TBTO 3 0
TBTF 3 0
TBTC1 28.8-
(96%) 29.5
TBTC1 32.7
(96%)
TBTC1 32.3
(96%)
Salinity
Chemical3 (q/kq)
TBTO 9.9-11.2
(95%)
TBTO 2 0
TBTO 15
TBTS 15
TBTO 15
TBTS 15
TBTO 15
9 days Approx. 80% mortality
9 days Approx. 90% mortality
10 days BCF = 41,200 (no
plateau)
10 days BCF = 60,300 (no
plateau)
10 days BCF = 1,700 (no
plateau)
Duration Effect
20 min No avoidance
14 days Telson regeneration
retarded; molting
retarded
15 days Reduced developmental
rate and growth
15 days Reduced developmental
rate and growth
15 days 63% mortality
15 days 74% mortality
20 days LC50
9.732 Laughlin et al.
1982
9.732 Laughlin et al.
1982
0.48 Meador et al .
1993
109 Meador 1997
660 Meador 1997
Concentrati
on Reference
(uq/L)b
30 Pinkney et al .
1985
0.1 Khan et al. 1993
14.60 Laughlin et al.
1983
18.95 Laughlin et al.
1983
>24.33 Laughlin et al.
1983
28.43 Laughlin et al.
1983
13.0 Laughlin and
French 1989
-------
Mud crab (zoea) ,
Rhi thropanopeus
harrisii
Mud crab,
Rhi thropanopeus
harrisii
Mud crab,
Rhi thropanopeus
harrisii
Mud crab,
Rhi thropanopeus
harrisii
Mud crab,
Rhi thropanopeus
harrisii
Mud crab,
Rhi thropanopeus
harrisii
Fiddler crab,
Uca puqilator
Fiddler crab,
Uca puqilator
Table 6 . (Continued)
Species
Fiddler crab,
Uca puqilator
TBTO 15 40 days LC50 33.6 Laughlin and
French 1989
TBTO 15 6 days BCF=24 for carapace 5.937 Evans and
Laughlin 1984
TBTO 15 6 days BCF=6 for 5.937 Evans and
hepatopancreas Laughlin 1984
TBTO 15 6 days BCF=0.6 for testes 5.937 Evans and
Laughlin 1984
TBTO 15 6 days BCF=41 for gill tissue 5.937 Evans and
Laughlin 1984
TBTO 15 6 days BCF=1.5 for chelae 5.937 Evans and
muscle Laughlin 1984
TBTO 25 <24 days Retarded limb 0.5 Weis et al.
regeneration and 1987a
molting
TBTO 25 3 weeks Reduced burrowing 0.5 Weis and
Perlmutter 1987
Salinity Concentrati
Chemical3 (q/kq) Duration Effect on Reference
(uq/L)b
TBTO 25 7 days Limb malformation 0.5 Weis and Kim
1988; Weis et
al. 1987a
Blue crab (6-8-day-old
embryos),
Callinectes sapidus
TBT
28
4 days EC50 (hatching)
0.047
Lee et al. 1996
-------
Brittle star,
Ophioderma brevispina
Atlantic menhaden
(juvenile) ,
Brevoortia tyrannus
Atlantic menhaden
(juvenile) ,
Brevoortia tyrannus
Chinook salmon (adult) ,
Oncorhynchus
tshawytscha
Chinook salmon (adult) ,
Oncorhynchus
tshawytscha
Chinook salmon (adult) ,
Oncorhynchus
tshawytscha
Mummichog (juvenile) ,
Fundulus heteroclitus
Mummichog,
Fundulus heteroclitus
Inland silverside
(larva) ,
Menidia beryllina
Mummichog (embryo) ,
Fundulus heteroclitus
Mummichog (5.3 cm; 1 . 8
g>,
Fundulus heteroclitus
Three-spined
stickleback (45-60 mm) ,
Gasterosteus aculeatus
TBTO 18-22
TBTC1 10
TBTO 9-11
TBTO 2 8
TBTO 2 8
TBTO 2 8
TBTO 2
TBTO 9.9-11.2
TBTC1 10
TBTO 2 5
TBTO 15
(95%) 16-19.5
TBTO 15-35
(painted
panels)
4 wks Retarded arm
regeneration
28 days No effect
Avoidance
96 hr BCF=4300 for liver
96 hr BCF=1300 for brain
96 hr BCF=200 for muscle
6 wks Gill pathology
20 min Avoidance
28 days Reduced growth
10 days Teratotogy
96 hr LC50
6 wks NOEC
7.5 mo 80% mortality (2
months)
Histological effects
• C.I Walsh et al
1986a
0.490 Hall et al.
1988b
5.437 Hall et al.
1.49 Short and
Thrower
1986a, 1986c
1.49 Short and
Thrower
1986a, 1986c
1.49 Short and
Thrower
1986a, 1986c
1984
17.2 Pinkney 1988;
Pinkney et
1989a
3 . 7 Pinkney et
1985
0.093 Hall et al.
1988b
30 Weis et al.
1987b
17 .2 Pinkney et
2.000 1989a
10 Holm et al.
2.5
al.
al.
al.
1991
Table 6. (Continued)
-------
Chemical8
Salinity
(q/kq)
Duration
Effect
Concentrati
on Reference
(uq/L)b
California grunion
(gamete through
embryo),
Leuresthes tenuis
California grunion
(gamete through
embryo),
Leuresthes tenuis
California grunion
(gamete through
embryo),
Leuresthes tenuis
California grunion
(embryo),
Leuresthes tenuis
10 days Significantly enhanced
growth and hatching
success
10 days Significantly enhanced
growth and hatching
success
10 days 50% reduction in
hatching success
10 days
California grunion
(larva) ,
Leuresthes tenuis
Striped bass
(juvenile) ,
Morone saxatilis
Striped bass
(juvenile) ,
Morone saxatilis
Striped bass
(juvenile) ,
Morone saxatilis
Speckled sanddab
(adult) ,
Citharichthys stiqmaeus
Stripped mullet (3.2
g);
G
TBTO 9-11
(95%)
TBT 13.0-
(painted 15.0
panels)
TBT 1.1-3.0
(painted 1.9-3.0
panels) 12.2-
14.5
TBTO 33-34
TBTO
(96%)
7 days
-
14
6 days
7 days
7 days
96 hr
8 wks
No adverse effect on
hatching success or
growth
Survival increased as
concentration increased
Avoidance
NOEC (serum ion
concentrations and
enzyme activity)
NOEC 0.067; LOEC 0.766
NOEC 0.444; LOEC 1.498
LOEC >0.514
LC50
BCF 3000 (no plateau)
BCF 3600 (no plateau)
0.14-1.71
0.14-1.72
0.14-1.72
0.14-1.72
24. 9
1. 09
18.5
0.122
0.106
Newton et al.
1985
Newton et al.
1985
74 Newton et al.
1985
Newton et al.
1985
Newton et al.
1985
Hall et al. 1984
Pinkney et al.
1989b
Pinkney et al.
1990
Salazar and
Salazar 1989
Yamada and
Takayanagi 1992
Muqil cephalus
-------
Starry flounder TBTC1 30.2 10 days BCF 8,700 (no plateau) 194 Meador 1997
(
-------
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