EPA 903-R-98-023
CBP/TRS 216/98
October 1998
A Probabilistic Ecological Risk
Assessment of Tributyltin in the
Chesapeake Bay Watershed
Chesapeake Bay Program
EPA Report Collection
Regional Center for Environmental Information
U.S. EPA Region III
Philadelphia, PA 19103
Printed on Recycled Paper for EPA by CBP
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itr' [or Environmental
19103
A Probabilistic Ecological Risk Assessment
of Tributyltin in the
Chesapeake Bay Watershed
October 1998
Chesapeake Bay Program
410 Severn Avenue, Suite 109
Annapolis, Maryland 21403
1-800-YOUR-BAY
http://www.chesapeakebay.net/bayprogram
Printed by the U.S. Environmental Protection Agency for the Chesapeake Bay Program
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October 1998
Final Report
A Probabilistic Ecological Risk Assessment of Tributyltin in the Chesapeake Bay Watershed
Lenwood W. Hall, Jr.
Mark C. Scott
William D. Killen
University of Maryland
Agricultural Experiment Station
Wye Research and Education Center
P. O. Box 169
Queenstown, Maryland 21658
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ABSTRACT
The goal of this study was to conduct a probabilistic ecological risk assessment for
tributyltin (TBT) in esruarine areas of the Chesapeake Bay watershed by using the following distinct
phases: problem formulation, analysis and risk characterization. This probabilistic ecological risk
assessment characterized risk by comparing the probability distributions of environmental saltwater
exposure concentrations with the probability distributions of species response data determined from
laboratory studies. The overlap of these distributions was a measure of risk to aquatic life.
Comparative risk from TBT exposure was determined for various basins (tributaries) in the
Chesapeake Bay watershed.
Tributyltin saltwater exposure data from the Chesapeake Bay watershed were available from
over 3,600 water column samples from 41 stations in nine basins from 1985 through 1996. Most of
the stations were located in the Virginia waters of Chesapeake Bay, primarily the James, Elizabeth
and York Rivers. In Maryland waters of the Bay, various marina, harbor and river systems were also
sampled. As expected, the highest environmental concentrations of tributyltin (based on 90th
percentiles) were reported in and near marina areas. The sources of TBT causing these high
concentrations were primarily boat hulls and painting/hull cleaning operations. Lower concentrations
of TBT were reported in open water areas ,such as the Potomac River, Choptank River and C and
D Canal, where the density of boats was minimal. Temporal data from a ten year data base (1986-
1996) from two areas in Virginia showed that TBT water column concentrations have declined since
1987 legislation prohibited the use of TBT paints on recreational boats (<25m).
Acute saltwater and freshwater TBT toxicity data were available for 43 and 23 species,
respectively. Acute effects for saltwater species were reported for concentrations exceeding 420
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ng/L; the lowest acute value for a freshwater species was 1,110 ng/L. The acute 10th percentiles for
all saltwater and freshwater species were 320 and 103 ng/L, respectively. The order of sensitivity
from most to least sensitive for saltwater trophic groups and corresponding acute 10th percentiles
were as follows: zooplankton (5 ng/L), phytoplankton (124 ng/L), benthos (312 ng/L) and fish (1,009
ng/L). For freshwater species, the order of sensitivity from most to least sensitive trophic groups and
corresponding acute 10th percentiles were: benthos (44 ng/L), zooplankton (400 ng/L), and fish (849
ng/L). Chronic data for both saltwater and freshwater species were limited to a few species in each
water type. Based on these limited data, the saltwater and freshwater chronic 10th percentiles were
5 and 102 ng/L, respectively. Limited mesocosm and microcosm studies in saltwater suggested that
TBT concentrations less than 50 ng/L did not impact the structure and function of biological
communities.
The saltwater acute (320 ng/L) and chronic (5 ng/L) 10th percentiles were used to determine
potential ecological risk because all exposure data were from saltwater areas of the Chesapeake Bay
watershed. Highest ecological risk was reported for marina areas in Maryland waters of Chesapeake
Bay and for areas in Virginia such as the Elizabeth River, Hampton Creek and Sarah Creek. Low
ecological risk was reported for areas such as the Potomac River, Choptank River, C and D Canal
and Norfolk Harbor. Regulation of TBT on recreational watercraft in 1987 has successfully reduced
water column concentrations of this organometallic compound. However, various studies have shown
that TBT may remain in the sediment for years and continue to be source for water column
exposures.
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ACKNOWLEDGMENTS
We would like to acknowledge the U. S. Environmental Protection Agency's Chesapeake Bay
Program Office for funding this study through grant number CB993589010. The "Toxics of Concern
Workgroup" of EPA's Toxics Subcommittee is also acknowledged for their support. The following
individuals are acknowledged for providing data: Michael Unger (Virginia Institute of Marine
Science), Peter Seligman (U. S. Navy) and Aldis Valkirs (Computer Sciences Corporation).
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TABLE OF CONTENTS
Page
ABSTRACT i
ACKNOWLEDGMENTS iii
TABLE OF CONTENTS iv
LIST OF TABLES vi
LIST OF FIGURES viii
1. INTRODUCTION 1
1.1 Problem Formulation 3
1.1.1 Stressor Characteristics 3
1.1.2 Ecosystems at Risk 4
1.1.3 Ecological Effects 4
1.1.4 Endpoints 5
1.1.5 Cumulative Stressors Potentially Impacting Aquatic Communities 6
1.1.6 Conceptual Model 6
2. EXPOSURE CHARACTERIZATION 8
2.1 Introduction 8
2.2 Tributyltin Loadings in the Chesapeake Bay Watershed 8
2.3 Chemical Properties of Tributyltin 9
2.4 Measured Concentrations of Tributyltin in the Chesapeake Bay Watershed 11
2.4.1 Data Sources and Sampling Regimes 11
2.4.2 Methods of Tributyltin Analysis 12
2.4.3 Methods of Data Analysis 12
2.5 Measured Concentrations by Basin 13
2.6 Temporal Trends 13
2.7 Summary of Exposure Data 14
3. ECOLOGICAL EFFECTS 15
3.1 Mode of Toxicity 15
3.2 Methods of Toxicity Data Analysis 15
3.3 Effects of Tributyltin from Laboratory Toxicity Tests 16
3.3.1 Acute Toxicity of Tributyltin 17
3.3.2 Chronic Toxicity of Tributyltin 17
3.4 Mesocosm/Microcosm Studies 18
3.5 Summary of Effects Data 19
iv
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Table of Contents (continued) Page
4. RISK CHARACTERIZATION 20
4.1 Characterizating Risks 20
4.2 Risk Characterization of Tributyltin in the Chesapeake Bay Watershed 21
4.3 Uncertainty in Ecological Risk Assessment 22
4.3.1 Uncertainty Associated with Exposure Characterization 22
4.3.2 Uncertainty Associated with Ecological Effects Data 23
4.3.3 Uncertainty Associated with Risk Characterization 24
5. CONCLUSIONS AND RESEARCH NEEDS 25
6. REFERENCES 27
TABLES 39
FIGURES 56
APPENDICES
Appendix A - Tributyltin risk characterization by basin and station
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LIST OF TABLES
Page
Table 1. Summary of three tributyltin data sources used for this risk assessment 40
Table 2. Summary of tributyltin exposure data for all basins and stations.
Maximum concentrations and 90th percentile values (minimum
of four detected concentrations) are presented by station and basin 41
Table 3. Summary of TBT sample preparation procedures and analytical methods for
the various monitoring studies 43
Table 4. Saltwater acute TBT toxicity data measured as TBT (ng/L). Concentrations marked
with an asterik * were converted from reported compounds to TBT. The abbreviations
used are: N=nominal, M=measured, S=static, R=renewal, FT=flowthrough,
LC=life cycle, ELS=early life stage and NR=not reported 44
Table 5. The 10th percentile intercepts for freshwater and saltwater tributyltin toxicity
data by test species and trophic group. These values represent protection of
90% of the test species 48
Table 6. Freshwater acute TBT toxicity data measured as TBT (ng/L). Concentrations
marked by an asterisk* were converted from reported compounds to TBT.
The abbreviations used are: N=nominal, M=measured, S=static, R=renewal,
FT=flowthrough, LC=life cycle, ELS=early life stage and NR=not reported 49
Table 7. Saltwater chronic TBT toxicity data measured as TBT (ng/L). Concentrations
marked with an asterisk* were converted from reported compounds to TBT.
The abbreviations used are: N=nominal, M=measured, S=static, R=renewal
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Page
FT=flowthrough, LC=life cycle, ELS=early life stage and NR=not reported 52
Table 8. Freshwater chronic TBT data measured as TBT (ng/L).Concentrations
marked with an asterisk* were converted from reported compounds to TBT.
The abbreviations used are: N=nominal, M=measured, S=static, R=renewal
FT=flowthrough, LC=life cycle, ELS=early life stage and NR=not reported 53
Table 9. The percent probability of exceeding the TBT acute and chronic saltwater
10th percentiles for all species 54
Vll
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LIST OF FIGURES
Page
Figure 1. Ecological risk assessment approach 57
Figure 2. Location of the 41 stations where TBT was measured from 1985 to 1996.
See key to map where stations are described 58
Figure 3. Temporal trend of 90th percentile concentrations of TBT for Sarah Creek
from 1986 to 1996 60
Figure 4. Temporal trend of 90th percentile concentrations of TBT for Hampton Creek
from 1986 to 1996.: 61
Figure 5. Distribution of TBT acute saltwater toxicity data 62
Figure 6. Distribution of TBT acute freshwater toxicity data 63
Figure 7. Distribution of TBT chronic saltwater toxicity data 64
Figure 8. Distribution of TBT chronic freshwater toxicity data 65
Vlll
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SECTION 1
INTRODUCTION
Improvement and maintenance of water quality were identified as the most critical elements
for the restoration and protection of the Chesapeake Bay in the 1987 Chesapeake Bay Agreement
(Chesapeake Executive Council, 1988). Another goal of this Agreement was the development and
adoption of a Chesapeake Bay Basinwide Toxics Reduction Strategy in order to achieve a reduction
of toxic substances consistent with the Water Quality Act of 1987. The Chesapeake Bay Basinwide
Toxics Reduction Strategy contained various commitments in areas such as research, monitoring and
toxic substance management that were directed to overall reduction of toxic chemicals in the
Chesapeake Bay watershed (Chesapeake Bay Executive Council, 1988). One commitment specified
for the creation of a Toxics of Concern List (TOC) for the Chesapeake Bay. This TOC list was
designed to prioritize over 1000 chemicals that may be impacting aquatic life or human health in
Chesapeake Bay by using a risk based ranking system and direct future research efforts and
management.
The first TOC list was completed in 1990 and was revised in 1996 (U. S. EPA, 1991; U. S.
EPA, 1996). The revised list as currently proposed is currently under review. The proposed revised
TOC list was developed using a chemical ranking system that incorporates sources, fate, exposure
and effects of chemicals on living resources and human health in Chesapeake Bay (Battelle, 1989).
The TOC list contains both a list of primary toxics of concern as well as a secondary list (chemicals
of potential concern). For both the 1990 and 1996 TOC lists, tributyltin (TBT) was identified as a
primary toxic of concern. Tributyltin enters the aquatic environment primarily as an antifouling paint
additive used on boat hulls although loading from drydocks during painting and hull cleaning
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operations can also occur. Antifouling paints containing TBT prevent the growth of fouling organisms
such as tube worms and barnacles on boat hulls. Tributyltin is one of the most effective biocides ever
used in antifouling paints but unfortunately, toxic qualities of TBT that make it effective in
controlling fouling organisms (target species) also poses a risk to non-target species in the aquatic
environment. Imposex (development of male characteristics in female snails) in female dog welks at
low ng/L concentrations and shell thickening and reduced productivity of oysters at 100 to 500 ng/L
concentrations have raised concern because these concentrations have been reported in the aquatic
environment of the Chesapeake Bay watershed (Huggett et al., 1996).
Although TBT has been identified as a toxic of concern in the Chesapeake Bay watershed,
a quantitative probabilistic ecological risk assessment has not been conducted for this organometallic
compound. The objective of this study was to apply EPA's Ecological Risk Assessment paradigm
for assessing ecological risk of TBT in the Chesapeake Bay watershed. Procedures described in the
following documents were used for this assessment: Report of the Aquatic Risk Assessment and
i
Mitigation Dialogue Group (SET AC, 1994), the EPA Framework for Ecological Risk Assessment
(U. S. EPA, 1992) and a recent paper entitled "An ecological risk assessment ofatrazine in North
-*,
American surface waters" (Solomon et al., 1996). This probabilistic risk assessment characterizes
risk by comparing probability distributions of environmental saltwater exposure concentrations with
the probability distributions of species response data (determined from laboratory studies). The
overlap of these distributions is a measure of potential risk to aquatic life in Chesapeake Bay. This
approach has a number of advantages over a quotient method (comparing the most sensitive species
with the highest environmental concentrations) because it allows, if not exact quantification, a least
a strong sense for the magnitude and likelihood of potential ecosystem effects of TBT in Chesapeake
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Bay. An implied assumption of this approach is that protecting a large percentage of species will also
preserve ecosystem structure and function. Various studies in basic ecology (Tillman, 1996; Tillman
et al., 1996) and of ecological effects of pesticides conducted in aquatic mesocosms (Solomon et al.,
1996) support the concept that in ecological risk assessment, some effects can be allowed at the
population level provided that these do not impair ecosystem structure and function and keystone
species are not impacted. The final result of the risk characterization is expressed as the probability
that exposure concentrations of TBT (within a defined spatial and temporal range) will exceed
concentrations deemed protective of aquatic life in the Chesapeake Bay watershed.
1.1 Problem Formulation
This ecological risk assessment has the following distinct phases: Problem Formulation,
Analysis and Risk Characterization (Figure 1). The problem formulation phase involves the
identification of major issues to be considered in the risk assessment. The analysis phase reviews
existing data on exposure (environmental monitoring of saltwater TBT concentrations) and ecological
effects ( primarily laboratory toxicity studies). The risk characterization phase involves estimation of
the probability of adverse effects on aquatic populations and communities in potentially impacted
areas of the Chesapeake Bay watershed.
The problem formulation phase of this risk assessment identified the following major issues
to be addressed: stressor characteristics, ecosystems at risk, ecological effects, endpoints, stressors
impacting aquatic communities, and a conceptual model for risk assessment.
1.1.1 Stressor Characteristics
The chemical and physical properties of TBT are described in detail in the Exposure section
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of this report. In the problem formulation phase of this risk assessment, the solubility, degradation
persistence in water and sediment, metabolism and bioconcentration potential of TBT were
considered important.
Solubility of TBT in the water column is influenced by such factors as the oxidation-reduction
potential, pH, temperature, ionic strength and concentration and composition of dissolved organic
matter (Clark et al., 1988). Maguire et al. (1983) has reported that TBT solubility ranges from 750
to 31,000 ug/L at pH of 2.6 to 8.1. Microbial degradation of TBT to the less toxic di- and
monobutyltin compounds has been reported as the most important process limiting pesistence of TBT
in the environment (U. S. Navy, 1997). Degradation half lives ranging from 4 to 19 days in seawater
have been reported while sediment half lives on the order of months to years have been observed (U.
S. Navy, 1997, De Mora et al., 1989). Significant TBT metabolism potential exists in fish and
crustaceans with minimal metabolic potential in mollusks. Mollusks also exhibit the highest
bioaccumulation factors and highest tissue burdens while fish and crustaceans generally accumulate
lower burdens of TBT.
1.1.2 Ecosystems at Risk
The aquatic ecosystem addressed in this risk assessment was the estuarine portion of the
Chesapeake Bay watershed. Most of the exposure data for TBT were reported for the Virginia
waters of Chesapeake Bay, primarily the James, Elizabeth and York Rivers. Various marina, harbor
and river systems were also sampled in the Maryland waters of Chesapeake Bay. Exposure data were
available for over 3,600 samples collected at 41 stations between 1985 to 1996.
1.1.3 Ecological Effects
A comprehensive review and synthesis of the TBT aquatic toxicity literature was conducted
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by using literature searches (AQUIRE etc.), various regulatory review documents such as the U. S.
EPA water quality criteria report (U. S. EPA, 1997) and other TBT review documents (e. g. Hall
and Pinkney; 1985, Champ and Seligman, 1996 and U. S. Navy 1997, among others). The ecological
effects data used in this risk assessment were derived from 43 saltwater species tested in acute
studies and 4 species tested in chronic toxicity tests (where chronic values were reported). Only
saltwater toxicity data were used for assessing risk because exposure data were only available from
saltwater areas in the Chesapeake Bay watershed. The acute saltwater 10th percentile (protection
of 90% of the species) was 320 ng/L for all species. The 10th percentile by trophic group from most
sensitive to least sensitive was: zooplankton (5 ng/L), phytoplankton (124 ng/L), benthos (312 ng/L),
and fish (1,009 ng/L). A 10th percentile of 5 ng/L was determined from the limited saltwater chronic
data (same value reported for zooplankton tested in acute tests). Limited microcosm and mesocosm
studies showed that TBT concentrations less than 50 ug/L generally did not impact the structure and
function of biological communities.
1.1.4 Endpoints
The selection of appropriate endpoints is a basic element of the risk assessment process. In
ecological risk assessment, it is recognized that individual organisms are part of the food wed and are
therefore somewhat expendable as they are either consumed or being consumed. Therefore, the focus
of ecological risk assessment is the protection of population, community or ecosystem function, rather
than individuals. This acknowledges the fact that a population is less sensitive than its most sensitive
member and likewise that communities and ecoystems are less sensitive that their most sensitive
populations. A consensus of recent ecological risk assessments has lead to an important conclusion-
some effects at the organism and population level can be allowed if these effects are restricted in
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space and time and keystone species are not impacted (Solomon et al., 1996; Giddings et al., 1997).
The Framework for Ecological Risk Assessment has defined two types of endpoints:
assessment endpoints and measurement endpoints (U. S. EPA, 1992). Assessment endpoints are the
actual environmental values that are to be protected (e.g. fish or shellfish populations). Measurement
endpoints are the measured responses to a stressor that can be correlated with or used to protect
assessment endpoints (Suter, 1990). With each higher level of testing, measurement endpoints differ
while assessment endpoints remain the same.
The assessment endpoints for this risk assessment are the long term viability of aquatic
communities in the Chesapeake Bay (fish, invertebrates etc.). Specifically, the protection of at least
90% of the species 90% of the time (10th percentile from species susceptibility distributions) from
acute TBT expsoures is the defined assessment endpoint. Measurement endpoints include all acute
TBT toxicity data (survival, growth and reproduction) generated from saltwater laboratory toxicity
studies.
1.1.5 Cumulative Stressors Potentially Impacting Aquatic Communities
When assessing the potential impact of TBT on aquatic communities in the Chesapeake Bay
watershed, it is important to remember that both biotic (food quality and quantity, disease) and abiotic
factors (water quality, other contaminants, physical habitat alteration) influence the status of
biological communities. As discussed above, individuals within the various biological communities
are more sensitive to contaminant stress than the community as a whole. Therefore, individual losses
due to a stressor such as TBT may or may not impact the viability (persistence, abundance,
distribution) of the population depending on all the factors influencing the population.
1.1.6 Conceptual Model
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Problem formulation is completed with the development of a conceptual model where a
preliminary analysis of the ecosystem at risk, stressor characteristics and ecological effects are used
to define the possible exposure and effects scenarios. The goal is to develop a working hypothesis
to determine how the stressor might affect exposed ecosystems. The conceptual model is based on
information about the ecosystem at risk and the relationship between the measurement and assessment
endpoints. Professional judgement is used in the selection of risk hypotheses. The conceptual model
describes the approach that will be used for the analysis phase and the types of data and analytical
tools that will be needed. Specific data gaps and areas of uncertainty will be described later in this
report.
The hypothesis considered in this risk assessment was:
• TBT may cause permanent reductions at the population and community level for fish,
benthos, zooplankton or phytoplankton in the Chesapeake Bay watershed and these
reductions may adversely impact community structure and function.
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SECTION 2
EXPOSURE CHARACTERIZATION
2.1 Introduction
The potential for exposure of aquatic organisms to TBT is an important component of a
probabilistic ecological risk assessment. Exposure data are used in conjunction with effects data (see
next section) to conduct a risk characterization. The exposure analysis for TBT considers use rates,
sources, loadings, chemical properties and spatial/temporal scale of measured concentrations (data
sources, sampling regimes, analytical methods and data analysis).
2.2 Tributyltin Loading in the Chesapeake Bay Watershed
The major source of TBT to Chesapeake Bay is from the use of antifouling paint on
watercraft hulls. Loading of TBT into the aquatic environment from either industrial or sewage
treatment plant effluents was reported to be minimal (Huggett et al., 1996). Based on this
information, h is not surprising that the highest concentrations of TBT have been measured in areas
with the greatest number of watercraft. Highest concentrations were generally found in marinas that
had a high density of boats painted with TBT and a low flushing rate. Two types of watercraft are
generally considered when determining loading of TBT- recreational and commercial. Estimates
conducted for the State of Virginia reported that ~ 70% of the TBT entering Virginia waters came
from recreational watercraft while ~ 27% was from large commercial vessels such as freighters and
tankers (Huggett et al., 1996). The remaining 3% came from miscellaneous sources such as the
military.
The logical focus for controlling TBT input to the Chesaepeake Bay's aquatic environment
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was to restrict use on recreational watercraft. With this consideration in mind, both Virginia and
Maryland passed legislation to restrict use of TBT on vessels <25m in length in 1987. Longer vessels
and aluminum craft hulls were exempted but could only use paints that release TBT at a rate of <5
ug cm"2 d"1. A number of other states followed the actions of Virginia and Maryland and federal
legislation was established in 1988.
2.3 Chemical Properties of Tributyltin
Tributyltin compounds used in antifouling paints consist of a tin atom covalently bonded to
three butyl moieties and an associated anion such as chloride or oxide or copolymers such as
methacrylate/methyl methacrylate. Copolymer paints allow manufacturers to formulate paints that
have better controlled leach rates in seawater (e. g. tributyltin methacrylate). Due to the hydrophobic
nature of the TBT antifouling coating, seawater interacts with the copolymer at the surface which
initiates a hydrolysis reaction that cleaves TBT from the copolymer backbone and releases it into the
water.
The most important process limiting the persistence of TBT in the aquatic environment is
microbial degradation to the less toxic dibutyltin (DBT) and monobutyltin (MET) compounds.
Degradation half-lives in various seawater experiments were reported to vary from 4 to 19 days (U.
S. Navy, 1997). Lee et al. (1989) also reported that phytoplankton were active in degrading TBT to
the less toxic DBT and MBT; both 2 and 6 day half-lives were reported. Photolysis and chemical
degradation were reported to be insignificant in the degradation of TBT in seawater (U. S. Navy,
1997). Degradation of TBT in sediments is much slower than in the water. Various studies have
reported half-life values in sediments to range from months to years in anaerobic sediments (Stang
and Seligman, 1986; De Mora et al. 1989; Dawson et al. 1993). These data suggest that that
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sediments may remain a source for TBT after limiting water column concentrations through
regulation.
Due to the hydrophobic nature of TBT it partitions rapidly to particulate material in the water
column and bottom sediments. Partitioning rates are dependent on TBT concentration, suspended
sediment, load, pH, salinity and organic carbon (U. S. Navy, 1997). Langston and Pope (1995) have
reported a partitioning coefficent of 25,000 L/kg at a water column concentration of 10 ng/L (a
realistic environmental concentration). Various investigators have reported that partitioning
coefficients can vary between 340 to 39,000 (Valkirs et al. 1986,1987; Harris et al. 1996).
Laughlin (1996) has reported that TBT is accumulated by nearly all taxa that have been
evaluated. Mollusks were reported to have the highest bioaccumulation factors and highest tissue
burdens. Fish and crustaceans generally accumulate lower burdens of TBT because they possess the
active cytochrome P-450 enzyme system that oxides TBT to less toxic components (Lee, 1996).
Bioaccumulation factors (BCFs) range from about 200 in some fish tissues to 100,000 in American
oysters and mussels (U. S. Navy, 1997). Salazar et al. (1987) reported that bioconcentration factors
showed an inverse relationship to concentrations in the field, with lower exposures leading to higher
bioconcentration factors in bivalves.
Although the potential for sediment-bound TBT to cause risk to sediment dwelling aquatic
biota exists, the focus of this risk assessment was an evaluation of risk to aquatic biota from
exposures to surface water concentrations. Probabilistic risk assessment techniques for assessing risk
of aquatic species to sediment exposures is still developmental and contains a higher degree of
uncertainty than water column exposures. By using surface water concentrations in this risk
assessment, the results can be more closely related to regulatory issues such as the proposed U. S.
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Environmental Protection Agency's chronic marine water quality criteria of 10 ng/L TBT ( U. S.
EPA, 1997).
2.4 Measured Concentrations of Tributyltin in the Chesapeake Bay Watershed
2.4.1 Data Sources and Sampling Regimes
Tributyltin exposure data (seawater measurements) were available from three primary data
sources from 1985 to 1996 for over 3,600 samples at 41 stations (Figure 2, Tables 1 and 2).
Approximately 92% of the measurements were from Virginia waters of Chesapeake Bay. The
remaining samples were measured from marinas, harbors and river systems in Maryland waters of the
Chesapeake Bay. The data sources are briefly described below:
Hall et alData (Hall et al. 1987, 1988, 1989r and 1992)
These data were collected from 1985 to 1989. During the 1985-86 effort, samples were
collected monthly from July through June at eight stations located in four small and large marinas,
a large harbor, two major river systems and a heavily used shipping channel (Hall et al., 1987). For
the other three studies, samples were collected bi-weekly from June-September of 1986, 1988 and
1989 at six stations in or near marinas in Back Creek (Annapolis, MD) and one location in the Severn
River near the confluence of Back Creek and the Severn River (Hall et al., 1988, 1989, and 1992).
Navy Data (Valkirs et al.r 1995)
Samples were collected during the summer of 1986 and/or quarterly from 1988 to 1992 at
12 stations in the Elizabeth River, 10 stations in Norfolk Harbor and one station in Hampton River
(see Figure 2 and Table 2).
VIMS Data (Unger, personal communication)
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These data were collected monthly at five stations in Hampton River and four stations in
Sarah Creek (York River basin) from 1986 through 1996 (See Figure 2 and Table 2).
2.4.2 Methods of Tributyltin Analysis
A summary of preparation procedures and analytical methods for the various TBT monitoring
studies is presented in Table 3. Detection limits were generally less than 5 ng/L for all studies (except
the Hall et al. 1987 - Hall et al 1985 data). In all cases, the water samples (grab samples) collected
for analysis were unfiltered and iced for preservation. The analytical method used for all studies
except the Navy monitoring was Gas Chromatograph - Flame Photometric Detector (GC-FPD)
(Unger et al., 1986). In all cases, TBT was speciated from the degradation products MET and DBT.
2.4.3 Methods of Data Analysis
Approaches for handling values below the detection limits include assigning these values as
zero, one-half the detection limit or the detection limit (MacBean and Rovers, 1984; Giddings et al.,
1997). For this risk assessment, TBT values below the detection limit were assumed to be log-
normally distributed. The distribution of exposure data was calculated based on the measured values
and the concentrations of the non-detects were assumed to be distributed along a lower extension of
this distribution. For example, if 80 out of 100 were reported as non-detects, the 20 measured values
were assigned ranks from 81 to 100 and the frequency distribution was calculated from these 20
values. For the very few cases where more than one value was available at the same time and
station, the highest value was used in the frequency distribution.
For data sets arranged by basin or station with four or more values above the detection limit,
log-normal distributions of exposure concentration were determined as follows. The observations in
each data set were ranked by concentration and for each observation the percentile ranking was
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calculated as n/(N+l) where n is the rank sum of the observation and N is the total number of
observations including the non-detects. Percentile rankings were converted to probabilities and a
linear regression was performed using the logarithm of concentration as the independent variable and
normalized rank percentile as the dependent variable. Although non-detected observations were not
included in the regression analysis, they were included in the calculation of the observation ranks.
The 90th percentile concentrations (exceedence of a given value only 10% of the time) were
calculated for sampling stations (or basins) based on the calculated log-normal concentration
distributions.
2.5 Measured Concentrations by Basin
A summary of maximum concentrations and 90th percentiles of individual stations and pooled
stations by basin or drainage is presented in Table 2. Maximum TBT concentrations by basin ranged
from below detection limit in the Choptank River and C and D Canal to 1801 ng/L in the Back Creek
marina area in Maryland. The 90th percentile values by basin for locations with at least 4 detected
concentrations, ranged from 4.1 ng/L in Norfolk Harbor to 387 ng/L in Pier 1 Marina in Maryland.
As expected, the highest 90th percentiles (138 to 387 ng/L) were reported in marina areas with high
densities of boats using TBT paints; much lower values were reported in open water areas such as
the Potomac and Choptank Rivers.
2.6 Temporal Trends
Tributyltin saltwater monitoring data were available from 1986 to 1996 in Sarah Creek and
Hampton Creek in Virginia to assess temporal trends in TBT (M. A Unger, personal communication).
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These data showed a clear trend of decreasing concentrations in the water column after the 1987
legislation (effective in 1988) restricted the use of TBT on recreational watercraft in Chesapeake Bay
watershed (Figure 3 and 4). For Sarah Creek, 90th percentile values dropped from approximately
40 ng/L in 1987 to approximately 9 ng/L in 1996 (Figure 3). The reduction of 90th percentiles values
in Hampton Creek was from 160 ng/L in 1987 to approximately 15 ng/L in 1996 (Figure 4).
2.7 Summary of Exposure Data
Highest environmental concentrations of TBT (based on 90th pereentiles) in the Chesapeake
Bay watershed were reported in and near marina areas. Sources of TBT responsible for these high
exposures were boat hulls and/or painting and hull cleaning operations. As expected, lower
concentrations of TBT were reported in open water areas such as rivers where the density of boats
was minimal. Temporal trends analysed from a 10 year data base (1986 to 1996) in two areas in
Virginia showed that TBT water column concentrations have declined after 1987 legislation
prohibited the use of TBT paints on recreational watercraft (<25m). Due to the long half-life of TBT
concentrations in sediment and the equilibrium shift occurring with lower water column
concentration, sediments will continue to be a source for TBT.
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SECTION 3
ECOLOGICAL EFFECTS
3.1 Modes of Toxicity
Both aquatic plants and animals have enzyme systems that can metabolize TBT to less toxic
derivatives. Plants such as eel grass, diatoms and dinoflagellates have been shown to metabolize TBT
(Lee, 1996). Diatoms produce a series of hydroxylated derivatives and it is likely that the algal
dioxygenase system is involved. For animal species, crustaceans, annelids and fish have enzyme
systems that rapidly metabolize TBT. The hydroxylation of TBT that occurs in aquatic vertebrates
and invertebrates is controlled by the microsomal cytochrome P-450 systems present in the hepatic,
intestinal and kidney tissues of these organisms. These hydroxylated derivatives are conjugated to
sulfate or carbohydrate by phase two enzyme systems, which facilitates the elimination of TBT (Lee,
1996). Mollusks have low cytochrome P-450 content and mixed function oxygenase activity and
therefore exhibit TBT accumulation with slow depuration rates. Various TBT effects in mollusks
include shell thickening, reduced growth rates and imposex (Champ and Seligman, 1996).
3.2 Methods of Toxicity Data Analysis
The primary toxicity benchmark used for this risk assessment was the 10th percentile of
species sensitivity (protection of 90% of the species) from acute and chronic exposures. The implied
assumption when using this benchmark is that protecting a large percentage of the species assemblage
will preserve ecosystem structure and function. This level of species protection is not universally
accepted, especially if the unprotected 10% are keystone species and have commercial or recreational
significance. However, protection of 90% of the species 90% of the time (10th percentile) has been
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recommended by the Society of Environmental Toxicology and Chemistry (SET AC, 1994) and others
(Solomon et al., 1996; Giddings et al. 1997). Recent mesocosm studies have reported that this level
of protection is conservative (Solomon et al.,1996; Giddings, 1992).
Tributyltin toxicity data were analyzed as a distribution on the assumption that the data
represented the universe of species. An approximation was made since it is not possible to test all
species in the universe. This approximation assumes that the number of species tested (N) is one less
than the number in the universe. To obtain graphical distributions for smaller data sets that are
symmetrical (normal distributions) percentages were calculated from the formula (100 x n/(N + 1))
where n is the rank number of the datum point and N is the total number of data points in the data
set (Parkhurst et al., 1994). This formula compensates for the size of the data sets as small
(uncertain) data sets will give a flatter distribution with more chance of overlap than larger (more
certain) data sets. In cases where there were multiple data points for a given species, the lowest value
was used in the regression analysis of the distribution. When data were available for multiple life
stages of a species the lowest values were generally reported for early life stages. Using the lowest
value therefore provides a conservative approach for protecting the most sensitive life stage of a
species. Data were plotted using Sigma Plot (Jandel Corporation, 1992) .
3.3 Effects of Tributyltin from Laboratory Toxicity Tests
Acute and chronic TBT toxicity data used in this risk assessment were obtained from the
AQUIRE database through 1995, U. S. EPA water quality criteria documents (U. S. EPA, 1997),
literature review documents (Hall and Pinkney, 1985; Hall and Bushong, 1996; Champ and Seligman,
1996) and manual searches of grey literature from academia, industry and government sources.
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Only data that met the various criteria established by the U. S. EPA for use in water quality criteria
development were used in the analysis (acceptable control survival, complete description of test
methods etc.). Tributyltin acute and chronic toxicity data by water type (saltwater and freshwater )
are discussed below. However, only the saltwater acute and chronic data were used for risk
characterization since all exposure data were from Chesapeake Bay saltwater environments.
3.3.1 Acute Toxicity of Tributyltin
Acute saltwater TBT toxicity data were available for 43 species, including five algal species,
five zooplanlcton species, 24 benthic species and nine fish species (Table 4, Figure 5) The range of
acute toxicity values was 420 ng/L for the mysid, Acanthomysis sculpta to 330,000 ng/L for an algal
species. The acute 10th percentile for all saltwater species was 320 ng/L (Table 5). A breakdown of
10th percentiles by trophic group from most to least sensitive was as follows: zooplanlcton (5 ng/L),
phytoplankton (124 ng/L), benthos (312 ng/L) and fish (1,009 ng/L) (Table 5).
Acute freshwater toxicity data were available for 23 species (Table 6, Figure 6). The data base
included three zooplanlcton species, five fish species and 11 benthic species. The range of acute
toxicity for freshwater species was 1,110 ng/L for the hydra, Hydra littoralis to greater than
114,000,000 ng/L for the adult clam, Elliptic complanata. The high value for the clam likely resulted
from shell closure during the acute exposure. The acute 10th percentile of all freshwater species was
103 ng/L (Table 5). The 10th percentiles by trophic group from most to least sensitive were as
follows: benthos (44 ng/L), zooplanlcton (400 ng/L) and fish (849 ng/L).
3.3.2 Chronic Toxicity of Tributyltin
Chronic saltwater TBT toxicity data, where chronic values were reported, were limited to four
invertebrate species (Table 7, Figure 7). These chronic values ranged from 14 to 131 ng/L. The 10th
17
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percentile for these four saltwater species was 5 ng/L (Table 5).
The freshwater chronic TBT toxicity data, where chronic values or no observed effect level
(NOEL) were reported, were also limited to only three species (Table 8, Figure 8). These chronic
values ranged from 137 to 253 ng/L for the cladoceran and fish species. The chronic 10th percentile
for the freshwater species was 102 ng/L (Table 5).
3.4 Mesocosm/Microcosm Studies
Saltwater mesocosm and microcosm studies with TBT were limited (Henderson, 1985, 1986,
1988). A three month mesocosm study by Henderson (1985,1986) with TBT concentrations ranging
from 500 to 1,800 ng/L showed the following: (1) annelid worms, crustaceans and fish were
insensitive to the exposures; (2) larval stages of various animals species such as corals, anemones,
echinoderms and mollusks were most sensitive. In later three month microcosm studies with TBT
concentrations ranging from 40 to 2,500 ng/L, Henderson (1988) reported the following: (1)
significant declines in pre-established fouling communities occurred at 500 ng/L and higher, (2)
significant reductions in number of species and species diversity of larval forms and reductions in the
condition index of the American oyster occurred at 100 ng/L and (3) oyster condition index, species
diversity and mortality did not occur at concentrations of 40 ng/L. The conclusion from these studies
is that TBT concentrations in the environment should be less than 50 ng/L to avoid effects on aquatic
communities.
Two microcosm studies with durations of 24 to 55 days were conducted in freshwater (
Delupis and Miniero, 1989; Miniero and Delupis, 1991). In both studies effects were immediate at
80,000 ng/L dose as Daphnia magna disappeared, ostracods increased and algal species increased
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immediately and then gradually dissappeared. The lowest concentration (4,700 ng/L) to even suggest
an effect in these studies caused a temporary reduction in metobolism (oxyen consumption). This
suggested effect concentration of 4,700 ng/L is much higher that concentrations typically measured
in the environment.
3.5 Summary of Effects Data
*
Acute effects with saltwater species were generally reported at concentrations greater than
or equal to 420 ng/L. The 10th percentile for all species derived from the acute TBT saltwater
toxicity database was 320 ng/L. The order of sensitivity from most to least sensitive trophic group
and corresponding 10th percentiles were as follows: zooplankton (5 ng/L), phytoplankton (124 ng/L),
benthps (312 ng/L) and fish (1,009 ng/L). For freshwater acute TBT studies, effects were reported
at concentrations at or above 1,110 ng/L. The acute freshwater 10th percentile for all species was
103 ng/L. The 10th percentiles by trophic group from most to least sensitive were as follows: benthos
(44 ng/L), zooplankton (400 ng/L) and fish (849 ng/L).
Chronic data from both saltwater and freshwater studies were limited to a few species. The
saltwater and freshwater chronic 10th percentiles for all species were 5 and 102 ng/L, respectively.
Limited mesocosm/microcosm studies in saltwater demonstrated that TBT concentrations less than
50 ng/L generally did not impact the structure and of biological communities. Freshwater microcosm
studies suggested that concentrations as high as 4,700 ng/L only caused temporary reductions in
biological community metabolism.
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SECTION 4
RISK CHARACTERIZATION
4.1 Characterizating Risks
The report of the Aquatic Risk Assessment Dialogue Group (SETAC, 1994) recommends
using tiers when assessing the risk of pesticides in the aquatic environment. The first tier is a
simple and commonly used risk quotient. Risk quotients are simple ratios of exposure and effects
concentrations where the susceptibility of the most sensitive species is compared with the highest
environmental exposures. If the exposure concentration equals or exceeds the effects concentration
an ecological risk is suspected. The quotient method is a valuable first tier assessment that allows a
determination of a worst case effects and exposure scenario for a particular contaminant. However,
some of the major limitations of the quotient method for ecological risk assessment are that it fails
to consider variability of exposures among individuals in a population, ranges of sensitivity among
species in the aquatic ecosystem and the ecological function of these individual species. The quotient
method also assumes that that there is a 100% probability of co-occurrence of the stressor and the
most sensitive organism and that the most sensitive organism is a keystone organism in the
environment. The probabilistic approach addresses these various concerns as it expresses the results
of an exposure or effects characterization as a distribution of values rather than a single point
estimate. Quantitative expressions of risks to aquatic communities are therefore determined by using
all relevant single species toxicity data in conjunction with exposure distributions. A detailed
presentation of the principles used in a probabilistic ecological risk assessment are presented by
Solomon et al. (1996) and Hall et al. (in press).
The following section will summarize the results of the risk characterization phase of this
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probabilistic ecological risk assessment of TBT in the Chesapeake Bay watershed. The toxicity
benchmarks used for the risk characterization will be both the saltwater acute and chronic 10th
percentiles for all species since all exposure data were reported in saltwater environments. Both the
acute 10th percentile for all species (320 ng/L) and the chronic 10th percentile for all species (5 ng/L)
are similar to the proposed U. S. EPA acute (370 ng/L) and chronic (10 ng/L) TBT criteria (U. S.
EPA, 1997). It is also important to note that the chronic 10th percentile of 5 ng/L also equals the
acute 10th percentile of the most sensitive trophic group (zooplankton) resulting from acute
laboratory toxcity tests. Therefore, the most sensitive trophic group (zooplankton) will be protected
in this risk characterization.
4.2 Risk Characterization of Tributyltin in the Chesapeake Bay Watershed
Potential ecological risk from TBT exposure was characterized by using both acute and
chronic saltwater effects data (10th percentiles for all species) since all exposure data were collected
in saltwater areas of the Chesapeake Bay. Using the acute 10th percentile as a toxicity benchmark
generally results in low risk for all areas sampled (Table 9). The greatest risk (12% exceedence) at
Back Creek marina in the Severn River would still be considered in the low risk range. The use of
the chronic benchmark of 5 ng/L results in some significant ecological risk in all marinas and
Baltimore harbor (> 97% exceedence), Hampton River (73% exceedence), Sarah Creek (52%
exceedence) and the Elizabeth River (33% exceedence). These data suggest that TBT may be posing
a risk to aquatic biota in specific areas of Chesapeake Bay associated with boating activity. The low
ecological risk reported in the Potomac River, Choptank River, C and D Canal and Norfolk Harbor
using the chronic 10th percentile indicates that potential ecologial risk is not present throughout the
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entire Chesapeake Bay.
4.3. Uncertainty in Ecological Risk Assessment
Uncertainty plays a particularly important role in ecological risk assessment as it impacts
problem formulation, analysis of exposure and effects data and risk characterization. Uncertainty in
ecological risk assessment has three basic sources: (1) lack of knowledge in areas that should be
known; (2) systematic errors resulting from human or analytical error and (3) non-systematic errors
resulting from the random nature of the ecosystem ( e.g. Chesapeake Bay watershed). The following
sections will address specific uncertainty from the above three sources as associated with exposure
data, effects data and risk characterization.
4.3.1 Uncertainty Associated with Exposure Characterization
The tributyltin exposure data used for this risk assessment were obtained from three different
data sources from 1985 to 1996 as described in Section 2. The spatial scale of these data (41
stations in 10 basins, primarily in Virginia) was somewhat limited considering that there are at least
50 major rivers that discharge into the Chesapeake Bay and numerous marinas where exposures are
unknown. Extensive exposure data from basins in Maryland waters of Chesapeake Bay were
particularly limited on a temporal scale. Uncertainty also existed because the exposure data used in
this risk assessment were not collected from random sampling designs compatible with unbiased
statistical analysis of frequency distributions. In fact, these data were biased toward high values
because many of the stations were located in areas where concentrations were expected to be high
(near marinas and harbors) and samples were collected during the high use season (summer) for
antifouling coatings such as TBT.
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Analytical techniques differed among three the laboratories - a GC-FPD method was used
by Hall et al and VIMS while the Navy used an AAS technique. Detection limits among the
laboratories were less than 5 ng/L for all measurements except the Hall et al 1985 data (< 20 ng/L).
Sample preparation techniques were consistent among all the laboratories; therefore, this area of
uncertainty was somewhat reduced.
4.3.2 Uncertainty Associated with Ecological Effects Data
Due to the relatively small number of species that can be routinely cultured and tested in
laboratory toxicity studies, there is uncertainty when extrapolating these toxicity data to responses
of natural taxa found in the Chesapeake Bay watershed. In the case of TBT in the Chesapeake Bay
watershed, saltwater acute and chronic toxicity data were available for 43 and 4 species, respectively,
for use in the calculation of the 10th percentile. Although the acute data provide a reasonable
representation of species in the Bay environment, the chronic data were limited and should be
expanded to reduce uncertainty in this risk assessment.
Variability in the results of toxicity tests for a given species tested in different experiments or
by different authors is a potential source of random and systematic errors. In this assessment, the
most conservative (lowest) effect value was used when multiple data points were available for a given
species. The range of toxicity data among trophic groups differed for the acute saltwater toxicity data
as the 10th percentiles ranged from 5 ng/L for zooplankton to 1,005 ng/L for fish. Using the
distribution of susceptibility accounts for this range of data points. Distributions will be flatter, with
greater chance of overlap with exposure distributions, when the range is large.
Acute and chronic saltwater TBT toxicity data were used in the risk characterization as
previously discussed. The use of acute (and chronic) data for predicting ecosystem effects is often
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questioned and assumed to be an area of significant uncertainty. However, Slooffet al. (1986) in
their review of single species and ecosystem toxicity for various chemical compounds, have reported
that there is no solid evidence that predictions of ecosystem level effects from acute tests are
unreliable. The result of Slooffet al. (1986) coupled with the use of a distribution of acute toxicity
data reduces some of the uncertainty associated with using acute data.
4.3.3 Uncertainty Associated with Risk Characterization
Many of the uncertainties associated with the variability in the exposure and effects
characterizations discussed above are incorporated in the probabilistic approach used in this risk
assessment (SETAC, 1994). Quantitative estimation of risks are analyzed as a distribution of
exposure and effects data.
Ecological uncertainty includes the effects of confounding stressors such as other
contaminants and the ecological redundancy of the functions of affected species. In the Chesapeake
Bay watershed, numerous contaminants other than TBT may be present simultaneously in the same
aquatic habitats near marinas and harbors; therefore, "joint toxicity" may occur. The concurrent
presence of various contaminants along with TBT makes it difficult to determine the risk of TBT in
isolation.
Ecological redundancy is known to occur in aquatic systems. Field studies have shown that
resistant taxa tend to replace more sensitive species under stressful environmental conditions
(Solomon et al., 1996; Giddings et al., 1992) The resistant species may replace the sensitive species
if it is functionally equivalent in the aquatic ecosystem and the impact on overall ecosystem function
is reduced by these species shifts. For this risk assessment, information on the ecological interactions
among species would help to reduce this area of uncertainty.
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SECTION 5
CONCLUSIONS AND RESEARCH NEEDS
Potential ecological risk from TBT exposure was reported for various areas in Chesapeake
Bay that are in close proximity to boating or shipping activity. Highest risk was reported for marinas
in Maryland but significant potential ecological risk was also reported for areas in Virginia such as
the Elizabeth River, Hampton Creek, and Sarah Creek. Low ecological risk was reported for areas
such as the Potomac River, Choptank River, C and D Canal and Norfolk Harbor. Temporal exposure
data has demonstrated that TBT concentrations in the water column have declined since 1987
regulation of this compound on recreational watercraft in Chesapeake Bay. Although these declining
water column concentrations are beneficial in reducing risks to aquatic biota, various studies have
showed that the TBT in the sediment may last for years and continue to remain a source for TBT
exposure in the water column. The use of TBT by commercial watercraft (>25m) and associated
painting and hull cleaning activities also continues to contribute to TBT loading in Chesapeake Bay.
The following research is recommended to supplement existing data for assessing the
ecological risks of TBT in the Chesapeake Bay watershed:
(1) Post-regulation exposure assessments for TBT in water and sediment for Sarah Creek and
Hampton Creek in Virginia should be continued as a long term monitoring effort since a 10 year data
base has already been established in these areas. These data will allow an analysis of TBT trends.
Similar type post-regulation monitoring activities should also be conducted in the Severn River/Back
Creek areas of Maryland where exposure data have also been collected.
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(2) Chronic TBT toxicity experiments are recommended for various keystone Chesapeake Bay
species (sensitive bivalve species) to determine the lowest observed effect concentrations (LOEC)
for these valuable species. In the current chronic saltwater toxicity data base, chronic values were
only available for four species.
(3) Biological communities such as fish and benthos (community metric approaches) should be
evaluated in areas where potential ecological risk of TBT is reported to be the greatest (near marinas
and harbors). Imposex in gastropods should also be assessed in these areas. These data would provide
a validation step for this ecological risk assessment.
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SECTION 6
REFERENCES
Alabaster, J.S. 1969. Survival offish in 164 herbicides, insecticides, fimgicides, wetting agents and
miscellaneous substances. Int Pest Control 11:29-35.
Battelle. 1989. Implementation of a chemical ranking system. Draft report to the U.S.
Environmental Protection Agency Criteria and Standards Division, Washington, DC.
Brix, K.V., P.P. Sweeney and R.D. Cardwell. 1994. Procedures for Conducting Acute Toxicity
Tests Using the Echinoderm Sperm Cell Test to Determine the Acute Toxicity of
Bis(Tributyltin) Oxide. Elf Atochem North America, Inc. Philadelphia, Pennsylvania.
Brooke, L.T., DJ. Call, S.H. Poirer, T.P. Markee, C.A. Lindberg, DJ. McCauley and P.O.
Simonsoa 1986a. Acute Toxicity and Chronic Effects of Bis(tri-n-Butyltin) Oxide to Several
Species of Freshwater Organisms. Report to Battelle Memorial Research Institute,
Columbus, Ohio, 20 pp.
Brooke, L.T., DJ. Call, S.H. Poirer, T.P. Markee, C.A. Lindberg, DJ. McCauley and P.O.
Simonsoa 1986b. Acute Toxicity and Chronic Effects of Bis(tri-n-Butyltin) Oxide to Several
Species of Freshwater Organisms. Center for Lake Superior Environmental Studies,
University of Wisconsin-Superior, Superior, WI. 20 p.
Buccafusco, R_, 1976b. Acute Toxicity of Tri-n-butyhin oxide to channel catfish Ictaluruspunctatus,
the freshwater clam Elliptic complanatus, the common mummichog Fundulus heteroclitus
and the American oyster Crassostrea virginica. US EPA-OPP Registration Standard.
Buccafiasco, R., 1976a. Acute Toxicity of Tri-n-butyhin oxide to bluegill Lepomis macrochirus. U.S.
EPA-OPP Registration Standard.
27
-------
Bushong, S.J., L.W. Hall, Jr., W.S. Hall, W.E. Johnson and R. L. Herman. 1988. Acute toxicity of
tributyltin to selected Chesapeake Bay fish and invertebrates. Water Res., 22:1027-1032.
Bushong, S.J., M.C. Ziegenfuss, M.A. Unger and L.W. Hall, Jr. 1990. Chronic tributyltin toxicity
experiments with the Chesapeake Bay copepod Acartia tonsa. Environ. Toxicol. Chem.
9:359-366.
Champ, M. A and P. F. Seligman. 1996. Organotin Environmental Fate and Effects. Chapman and
Hall, London. 623 pp.
Chesapeake Bay Executive Council. 1988. Chesapeake Bay Toxics Reduction Strategy.
Chesapeake Bay Agreement Report. Chesapeake Bay Liaison, Annapolis, MD.
Clark, J.R., J.M. Patrick, Jr., J.C. Moore and E.M. Lores. 1987. Waterborne and sediment source
toxicities of six organic chemicals to grass shrimp Palaemonetes pugio and amphioxus
Branchiostoma caribaewn. Arch. Environ. Con. Tox., 16:401-407.
Clark, E. A., K M. Sterritt and J. N. Lester. 1988. The fate of tributyltin in aquatic environment.
Environ. Sci. Tech. 22: 600-604.
Cramm, G. 1979. Acute and Chronic Toxicity of Tributyltin oxide (TBTO) to Sheepshead Minnows
(Cyprinodon variegatus). U.S. EPA-OPP Registration Standard
Das, V.G.K., L.Y. Kuan, K.I. Suderuddin, C.K. Chang, V. Thomas, C.K. Yap, M.K. Lo, C.C. Ong.
1984. The Toxic Effects of Triorganotin (IV) Compounds on the Culicine Mosquito Aedes
Aegypti (L). Toxicology 32:57-66.
Davidson, B.M., AO. Valkirs and P.F. Seligman. 1986a. Acute and chronic effects of tributyltin on
the mysid Acanthomysis sculpta (Crustacea, Mysidacea). NOSC-TR-1116 or AD-A175-294-
8. National Technical Information Service, Springfield, VA
28
-------
Davidson, B.M., A.O. Valkirs and P.F. Seligman. 1986b\ Acute and chronic effects of tributyltin on
the mysid Acanihomysis sculpta (Crustacea, Mysidacea). In: Oceans 86, Vol 4. Proceedings
International Organotin Symposium. Marine Technology Society, Washington, DC. pp. 1219-
1225.
Dawson, P. H. et al. 1993. Depositional profiles and relationships between organotin compounds in
freshwater and estuarine sediment cores. Environ. Monitor. Assess. 28: 145-161.
Delupis, G. D. D. and R. Miniero. 1989. Preliminary studies on the TBTO effects on freshwater biotic
communities. Riv. Idrobiol. 28: 1-2.
De Mora, S. J. et al. 1989. Tributyltin and total tin in marine sediments: profiles and the apparent rate
of degradation. Environ. Tech. 10: 901-908.
Floch, H. R. Deschiens and T. Floch. 1964. Sur les proprietes molluscicides de F oxyde et de F
acetate de tributyl-etain (prophylaxie des bilharzioses). Bull. Soc. Pathol. Exotique, 57:454-
465.
Foster, R.B. 1981. Use of Asiatic Clam Larvae in Aquatic Hazard Evaluation. In: Ecological
Assessments of Effluent Impacts on Communities of Indigenous Aquatic Organisms, J.M.
Bates and C.I. Weber (Eds). American Society for Testing and Materials, Philadelphia,
pp.280-288.
Giddings, J. M., L. W. Hall, Jr., K. R. Solomon, W. Adams, D. Vogel, L. Davis and R. Smith. 1997.
An ecological risk assessment of diazinon in the Sacramento and San Joaquin basins.
Technical Report 11/97. Novartis Crop Protection, Greensboro, NC.
Giddings, J.M. 1992. Aquatic mesocosm test for environmental data and ecological effects of
diazinon. Report. No. 92-3-4155, Springborn Laboratories, Inc., Wareham, MA.
29
-------
Goodman, L.R., G.M. Cripe, P.H. Moody and D. G. Halsell. 1988. Acute toxicity of malathion,
tetrabromobisphenol-A and tributyltin chloride to mysids (Mysidopsis bahia) of three ages.
Bull. Environ. Contam. Toxicol. 41:746-753.
Hall, L.W. Jr. and S.J. Bushong. 1996. A Review of Acuute Effects of Tributyltin Compounds on
Aquatic Biota In: Organotin Environmental Fate and Effects M. A Champ and P.P. Seligman
(Eds.) Chapman & Hall, London, pp. 157-190.
Hall, L.W. Jr., S.J. Bushong, W.S. Hall and W.E. Johnson. 1988. Acute and chronic effects of
tributyltin on a Chesapeake bay copepod. Environ. Toxicol.. Chem. 7:41-46.
Hall, L. W. Jr., S. J. Bushong, W. D. Johnson, and W. S. Hall. 1988. Spatial and temporal distribution
of butyltin compounds in a northern Chesapeake Bay marina and river system. Environ.
Monitor. Assess. 10: 229-244.
Hall. L. W. Jr., M. J. Lenkevich, W. S. Hall, A. E. Pinkney and S. J. Bushong. 1987. Evaluation of
butyltin compounds in Maryland waters of Chesapeake Bay. Mar. Pollut. Bull. 18: 78-83.
Hall, L.W. Jr., and A.E. Pinkney. 1985. Acute and Sublethal Effects of Organotin Compounds on
Aquatic Biota: An Interpretive Literature Evaluation. CRC Critical Reviews in Toxicology
14:159-209.
Hall, L. W., M. C. Scott and W. A. Killen. in press. An ecological risk assessment of copper and
cadmium in surface waters of the Chesapeake Bay watershed. Environ. Tox. Chem.
Hall, L. W. Jr., M. A. linger, S. J. Bushong and M. C. Ziegenfuss. 1989. Butyltin monitoring in
Northern Chesapeake Bay marina and river system in 1988: An assessment of tributyltin
legislation. Report. The Johns Hopkins University, Applied Physics Laboratory, Aquatic
Ecology Section, Shady Side, MD.
30
-------
Hall, L. W. Jr., M. A. Unger, M. C. Ziegenfuss, J. A. Sullivan and S. J. Bushong. 1992. Butyltin and
copper monitoring in a northern Chesapeake Bay marina and river system in 1989: An
assessment of tributyltin legislation. Environmental Monitoring and Assessment 22: 15-38.
Harris, J. R. W., J. J. Cleary and A. 0. Valkirs. 1996. Particle water partitioning and the role of
sediments as a sink and secondary source of TBT. In: Organotin Environmental Fate and
Effects, M. A. Champ and P. F. Seligman (eds.) Chapman and Hall, London, pp 459-473.
Henderson, R S. 1985. Effects of tributyltin antifouling paint leachates on Peral Harbor organisms.
Naval Ocean Systems Center Technical Report #1079, San Deigo, CA.
Henderson, R. S. 1986. Effects of organotin paint leachates on Pearl Harbor organisms: A site
specific flowthrough bioassay. In: Proceeding, Oceans 86 Conference - Organtin Symposium,
Washington, D. C. pp 1126-1233.
Henderson, R. S. 1988. Marine micrososm experiments on effects of copper and tributyltin-based
antifouling paint leachates. Naval Oceans System Center Technical Report #1060, San Deigo,
CA.
Ho, S.K. 1984. Evaluation of Carbon-14 Uptake Algal Toxicity Assay and its Application in Field
Assessment of Tributyltin Chloride and Chlorinated Sewage Toxicities. Masters Thesis, The
College of William and Mary, Williamsburg, Virginia, 108pp.
Hooftman, R.N., D.M.N. Adema and J, Kauffinan-Van Bommel. 1989. Developing a Set of Test
Methods for the Toxicological Analysis of the Pollution Degree of Waterbottoms Rep. No.
16105, Netherlands Organization for Applied Scientific Research:68p. (DUT)
Huggett, R. J., D. A. Evans, W. G. Maclntyre, M. Unger, P. F. Seligman and L. W. Hall, Jr. 1996.
Tributyltin concentrations in waters of Chesapeake Bay. In: Organotin Environmental Fate
31
-------
and Effects. M. A Champ and P. F. Seligman (Eds.) Chapman and Hall, London, pp. 485-501.
Jandel Corporation. 1992. SigmaPlot Scientific Graphing System, Version 5. San Rafael, CA.
Khan, A.T., J.S. Weis, C.E. Saharig and A.E. Polo. 1993. Effect of Tributyltin on Mortality and
Telson Regeneration of Grass Shrimp, Palaemonetespugio. Bull. Environ. Contam Toxicol.
50:152-157.
Langston, W. J. andN. D. Pope. 1995. Determinants of TBT adsorption and desorption in estuarine
sediments. Mar. Pollut Bull. 31: 32-43.
Lapota, D., D.E. Rosenberger, M.F. Platter-Rieger and P.P. Seligman. 1993. Growth and Survival
of Mytilus edulis larvae exposed to low levels of dibutyltin and tributyltin. Mar. Biol.
115:413-419.
Laughlin, R. B. 1996. Bioaccumlation of TBT by aquatic organisms. In: Organotin Environmental
Fate and Effects. M. A. Champ and P. F. Seligman (eds), Chapman and Hall, London, pp 331-
355.
Laughlin, R.B., O. Linden and H.E. Guard. 1982. Acute toxicity of tributyltins and tributyltin
leachates from marine antifouling paints. Bull. Liaison Comite Int. Permanent Rech.
Preservation Matoriaux Milieu Mar. (COIPM), 13:3-20.
Leblanc, G., 1976. Acute Toxicity of Tributyltin oxide to Daphnia magna U.S. EPA-OPP
Registration Standard.
Lee, R. F., A O. Valkirs and P. F. Seligman. 1989. Importance of microalgae in the biodegradation
of tributyltin in esturaine waters. Environ. Sci. Technol. 23: 1515-1518.
Lee, R. F. 1996. Metabolsim of tributyltin by aquatic organisms. In: Organotin Environemental Fate
and Effects. M. A Champ and P. F. Seligman (eds), Chapman and Hall, London, pp 369- 382.
32
-------
MacBean, E. A. and F. A. Rovers. 1984. Alternatives for handling detection limit data in impact
assessments. Groundwater Monitoring Review, Spring 1984.
Maguire, R. J., J. H. Carey and E. J. Hale. 1983. Degradation of the tri-n-butylitn species in water.
J. Agric. Food Chem. 31: 1060-1065.
M & T Chemicals Company. 1981. Comparative Toxicity of Tri-n-butyltin Oxide TBTO) Produced
by two Different Chemical Processes to Pink Shrimp Penaeus duorarum. Report No. BP-81-
4-55 submitted by EG & G Bionomics to M & T Chemicals Co., Rahway, New Jersey.
Matthijssen- Spiekman, E.A.M., J.H. Canton and C.J. Roghair. 1989. Research after the Toxicity of
TBTO for a Number of Freshwater Organisms Rep. No. 668118-00\ Natl. List. Public
Health and Environ. Hyg.:48p. (DUT).
Meador, J.P. 1993. The Effect of Laboratory Holding on the Toxicity Response of Marine Infaunal
Amphipods to Cadmium and Tributyltin. J. Exp. Mar. Biol. Ecol. 174:227-242.
Meador, J.P., U. Varanasi and C.A. Krone. 1993. Differential Sensitivity of Marine Infaunal
Amphipods to Tributyltin. Mar. Biol. 116:231-239.
Miniero, R. and G. D. D. Delupis. 1991. Effects of TBT (tributyltin) on aquatic microcosms. Toxicol.
Environ. Chem. 31: 425-431.
Naval Oceans Systems Center (NOSC), 1981. unpublished data, San Diego, CA.
Parkhurst, B.R., W. Warren-Hicks, T. Etchison, J.B. Butcher, R.D. Cardwell and J. Volison. 1994.
Methodology for aquatic risk assessment. Draft final report RP91-AER-1 prepared for the
Water Environment Research Foundation, Alexandria, VA.
Pinkney, A.E., D.A. Wright and G.M. Hughes. 1989. A morphometric study of the effects of
tributyltin compounds on the gills of the mummichog Fundulus heteroclitus. J. Fish. Biol.
33
-------
34:665-677.
Robert, R. and E. His. 1981. Action de L'acetate de Tributyl-etain sur les oeufs et les larves de deux
mollusques d'interet commercial: Crassostrea gigas (Thunberg) et Mytilus gaUoprovincialis
(Lamarck). Internal. Council Explor. Sea Paper, CM1981:F41, 16 pp.
Roberts, M.H. Jr. 1987, Acute toxicity of tributyltin chloride to embryos and larvae of two bivalve
molluscs, Crassostrea virginica and Mercenaria mercenaria. Bull, Environ. Contam.
Toxicol., 39:1012-1019.
Salazar, S. M., B. M. Davidson, M. H. Salazar, P. M. Stang and K. Meyers-Schulte. 1987. Effects
of TBT on marine organisms: Field assessment of a new site-specific bioassay system. In:
Proceedings, Oceans 1987 - Organotin Symposium. Halifax, Nova Scotia, Canada, pp 1461-
1470.
Salazar, M.H. and S.M. Salazar. 1989. Acute effects of (Bis)tributyltin Oxide on Marine Organisms,
Summary of Work Performed 1981-1983. Draft Report, Naval Ocean Systems Center, San
Diego, California, 88 pp.
Seinen, W., T. Helder, H. Vermij, A. Penninks and P. Leeuwangh. 1981. Short-term toxicity of tri-n-
butyltinchloride in rainbow trout (Salmo gairdneri Richardson) yolk sac fry. Sci. Total
Environ., 19:155-166.
Short, J.W. and P.P. Thrower. 1986. Tri-n-butyltin caused mortality of Chinook Salmon,
Oncorhynchus tshawytscha, on transfer to a TBT treated marine net pen. In: Proceedings of
the Qceans '86 Organotin Symposium, Vol. 4, Washington DC. pp. 1201-1205.
Sloof, W., J.A.M. van Oers and D. De Zwart. 1986. Margins of uncertainty in ecotoxicological
hazard assessment. Environ. Toxicol. Chem. 5:841-852
34
-------
Smith, P.J., A.J. Crowe, V.G. Kumar Das and J. Duncan. 1979. Structure-activity relationships for
some organotin molluscicides. Pesticide Sci., 10:419-422.
Snell, T.W., B.D. Moffat, C. Janssen and G. Persoone. 1991a. Acute Toxicity Tests Using Rotifers
IV. Effects of Cyst Age, Temperature and Salinity on the Sensitivity of Brachionus
cafyciflorus. Ecotoxicol. Environ. Saf. 21:308-317.
Snell, T.W., B.D. Moffat, C. Janssen and G. Persoone. 1991b. Acute Toxicity Tests Using Rotifers
III. Effects of Temperature Strain and Exposure Time on the Sensitivity of Brachionus
plicatitis. Environ. Toxicol. Water Qual. 6:63-75.
Society of Environmental Toxicology and Chemistry (SETAC). 1994. Aquatic Risk Assessment
and Mitigation Dialogue Group. Final Report. SETAC Foundation for Environmental
Education, Inc., Pensacola, FL
Solomon, K.R. et al. 1996. Ecological risk assessment of atrazine in North American surface
waters. Environ. Tox. Chem. 15:31-76.
Stang, P. M. and P. F. Seligman. 1986. Distribution and fate of butyltin compounds in the sediment
of San Diego Bay. In: Ocean 86 Conference Record: Science-Engineering-Adventure.
Volume 4 Organotin Symposium. Oceans 86. Washington, D. C., pp 1256-1261.
Stephan, C.E., D. I. Mount, D. J. Hansen, J. H. Gentile, G. A. Chapman and W. A. Brungs. 1985.
Guidelines for deriving numerical national water quality criteria for the protection of aquatic
organisms and their uses. Report # PB85-227049, National Technical Information Service,
Springfield, VA.
Suter, G.W. n. 1990. Endpoints for regional ecological risk assessment. Environ. Mange. 14:19-
23.
35
-------
TAI Environmental Sciences Inc. 1989a. Toxicity of Tri-butyl tin oxide to two species of Hydra.
Mobile, AL, 15 August. 31pp.
TAI Environmental Sciences Inc. 1989b. Toxicity of Tri-butyl tin oxide to Hydra littoralis and
Chlorohydra viridissima. Mobile, AL, 13 September. 25pp.
Thain, J.E. 1983. The Acute Toxicity of Bis (Tributyltin) Oxide to the Adults and Larvae of Some
Marine Organisms, Intdernational Council for the Exploration of the Sea, Copenhagen, Paper
CM1983/E:13, 5pp.
Tillman, D. 1996. Biodiversity: Population versus ecosystem stability. Ecology 77: 350-363.
Tillman, D., D. Wedlin and J. Knops. 1996. Productivity and sustainability influenced by biodiversity
in grassland ecoystems. Nature 379: 718-720.
Unger, M. A. Personal communication. Raw data from VIMS TBT monitoring.
Unger, M A., W. G. Maclntyre, J. Greaves and R. J. Huggett. 1986. GC determination of butyltins
in natural waters by flame photometric detection of hexyl derivatives with mass spectrometric
confirmation. Chemosphere 15: 461-470.
ITren, S.C. 1983. Acute toxicity of bis(tributyltin)oxide to a marine copepod. Marine Pollut. Bull.,
14:303-306.
U.S. Environmental Protection Agency (U.S. EPA). 1991. Chesapeake Bay Toxics of Concern List
Information Sheets. Report prepared by the Chesapeake Bay Programs Toxics Subcommittee
Living Resources Subcommittee's Joint Criteria and Standards Workgroup. Annapolis, MD.
U.S. Environmental Protection Agency (U.S. EPA). 1992. Framework for ecological risk
assessment. Risk Assessment Forum, EPA 630/R92/001. Washington, DC.
U.S. Environmental Protection Agency (U.S. EPA). 19%. Chesapeake Bay Toxics of Concern List.
36
-------
Prepared by the Chesapeake Bay Program Toxics Subcommittee's Living Resources
Subcommittee's Joint Criteria and Standards Workgroup. Annapolis, MD.
U.S. Environmental Protection Agency. 1997. Ambient Aquatic Life Water Quality Criteria for
Tributyltin. USEPA Office of Research and Development, Environmental Research
Laboratories. Duluth, Minnesota Narragansett, Rhode Island
U. S. Navy. 1997. Navy program to monitor ecological effects of organotin. Report to congress.
Naval Command, Control and Ocean Surveillance Center, San Diego, CA.
Valkirs, A. O. et al., 1995. U. S Navy monitoring of tributyltin in selected U. S harbors. Final
report/March 1995, Naval Command, Control and Ocean Surveillance Center, San Diego,
CA.
Valkirs, A. O., P. F. Seligman, and R. F. Lee. 1986. Butyhin partitioning in marine waters and
j
sediment. In: Proceedings, Oceans 86 Conference - Qrganotin Symposium, Washington, D.
C., pp 1165-1170.
Valkirs, A. O., M. O. Stallard and P. F. Seligman. 1987. Butyltin partitioning in marine waters. In:
Proceedings, Ocean 86 Conference - Organotin Symposium, Nova Scotia, Canada, pp 1375-
1380.
Walsh, G.E., L.L. McLaughlan, E.M. Lores, M.K. Louie and C.H. Deans. 1985. Effects of organotins
on growth and survival of two marine diatoms, Skeletonema costatum and Thalassiosira
pseudonana. Chemosphere, 14:383-392.
Walsh, G.E., M.K. Louie, L.L. McLaughlan and E.M. Lores. 1986. Lugworm Arenicola cristata
larvae in toxicity tests: survival and development when exposed to organotins. Environ.
Toxicol. Chem., 5:749-754.
37
-------
Walsh, G.E., L.L. McLaughlan, N.J. Yoder, P.H. Moody, E.M. Lores, J. Forestet and P.B.
Wessinger-Duval. 1988. Minutocelluspolymorphus; A new marine diatom for use in algal
toxicity tests. Environ. Toxicol Chem. 7:925-929.
Webbe, G. and R.F. Sturrock. 1964. Laboratory tests of some new molluscicides in Tanganyika.
Ann. Trop. Med. Parasitol., 58:234-239.
Wong, P.T.S., Y.K. Chau, O. Kramer and G.A Bengert. 1982. Structure-toxicity relationship of tin
compounds on algae. Can. J. Fish. Aquat. Sci. 39:483-488.
38
-------
TABLES
39
-------
Table 1. Summary of three data sources used for this risk assessment.
Reference
Data ID Total # Sample Period
samples
Detection Limit
(ng/L)
Hall etal., 1987
Hall etal., 1988
Hall etal., 1989
Hall etal., 1992
VaDcirs etal., 1995
MA. linger, pers. comm.
Hall Data 85 % monthly July 1985-June 1986
Hall Data 86 63 June-September 1986
Hall Data 88 63 June-September 1988
Hall Data 89 63 June-September 1989
Navy 1,027 summer 1986; quarterly 1988-1992
VIMS 2,307 monthly 1986-1996
20
5
2
2
0.2
1
40
-------
Table 2. Summary of TBT exposure data for all basins and stations. Maximum concentrations and 90th percentile values
(minimum of four detected concentrations) are presented by station and basin.
Drainage
Data ID
James Basin
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
TBT Concentration (ng/L)
Station
.Elizabeth River
Elizabeth River Station 1 5
Elizabeth River Station 17A
Elizabeth River Station 19
Elizabeth River Station 2 1
Elizabeth River Station 32
Elizabeth River Station 13 A
Elizabeth River Station 1 1
Elizabeth River Station 10
Lafayette River Station 37
Naval Station 9
Naval Station 4
Naval Station 3
Elizabeth River all stations combined
James Basin
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
James mainstem/Norfolk Harbor
Hampton Roads Station 29
Hampton Roads Station 35
Hampton Roads Station 23
Hampton Roads Station 3A
Hampton Roads Station 34
Hampton Roads Station 1
Hampton Roads Station 25
Hampton Roads Station 25 A
Hampton Roads Station 25B
Hampton Roads Station 36
Norfolk Harbor all stations combined
James Basin
VMS
VMS
NAVY
VMS
VMS
VMS
. Hampton River
OPC
HRM2
Station 33
HRM1
HYC
CD
Hampton River all stations combined
York Basin.
VMS
VMS
VMS
VMS
Sarah Creek
Potomac
Hall Data
Choptank
Hall Data
West
Sarah Creek
A
B
C
D
all stations combined
Potomac River
Choptank River
Hartge Marina
# Samples
7
10
82
10
85
7
84
82
6
79
80
22
611
79
4
5
74
3
79
78
14
14
A
354
258
258
86
257
256
251
1372
256
255
256
254
1021
12
12
12
# Detections
4
5
80
8
84
6
79
78
0
72
74
22
562
71
0
0
70
0
70
69
13
13
SL
306
252
257
84
257
256
257
1363
251
255
253
122
881
2
0
10
Maximum
8.9
10.7
41
13.4
29.4
9.8
45.4
14.3
BLD
8.9
9.8
7-2
45.4
9.8
BLD
BLD
10.7
BLD
6.1
5.1
5.3
3.2
BLD
10.7
42
1,300
38
180
340
25
1,300
120
72
23
16
120
24
BLD
186
90th percentile
11.1
10.2
19.7
28.4
18.6
12.8
16.0
10.0
-
6.4
5.7
5.5
14.2
4.9
-
-
3.3
-
4.5
5.4
4.5
3.0
_;
4.1
7.2
337
10
41
54
26
77
27
31
14
16.
23
-
-
138
41
-------
Drainage
Data ID
Severn
Mid-Bay
Chester
Patapsco
C&D Canal
TBT Concentration (ng/L)
Station
six Back Creek stations
Severn River
Pier 1 Marina
Piney Narrows Marina
Baltimore Harbor
C&D Canal
# Samples
174
27
12
12
12
12
# Detections
174
27
10
11
8
0
Maximum
1,801
89
307
338
112
BLD
90* percentile
351
53
387
354
129
.
42
-------
Table 3. Summary of TBT sample preparation procedures and analytical methods for the various monitoring
studies.
Database ID Detection Limits
ng/L
Hall Data 85
Hall Data 86
Hall Data 88
Hall Data 89
Navy
VIMS
20
5
2
2
0.2
1
Filtered/
Unfiltered
unfiltered
unfiltered
unfiltered
unfiltered
unfiltered
unfiltered
Preservation Sample Type Analytical Method
iced
iced
iced
iced
iced
iced
grab
grab
grab
grab
grab
grab
GC-FPD
GC-FPD
GC-FPD
GC-FPD
AAS
GC-FPD
43
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47
-------
Table 5. The 10* percentile intercepts for freshwater and saltwater tributyltin toxicity data by test
duration and trophic group. These values represent protection of 90% of the test species.
Water type
Test type
Trophic Group
n 10th Percentile (ng/L)
Freshwater
Freshwater
Saltwater
Saltwater
acute
chronic
acute
chronic
All species
zooplankton
benthos
fish
All species
All species
phytoplankton/algae
zooplankton
benthos
fish
All species
23
3
11
5
3
43
5
5
24
9
4
103
400
44
849
102
320
124
5
312
1,009
5
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-------
Table 9. The percent probability of exceeding the TBT acute and chronic saltwater 10th percentiles for all species.
Drainage
James Basin. Elizabeth
Elizabeth River
Station
River
Elizabeth River Station 15
Elizabeth River Station 17A
Elizabeth River Station 1 9
Elizabeth River Station 21
Elizabeth River Station 32
Elizabeth River Station 13A
Elizabeth River Station 1 1
Elizabeth River Station 10
Lafayette River Station 37
Naval Station 9
Naval Station 4
Naval Station 3
all stations combined
acute saltwater tests -
all species (320 ng/L)
0.01
O.01
O.01
0.12
O.01
O.01
O.01
O.01
-
- O.01
O.01
<0.01
<0.01
chronic saltwater
tests (5 ng/L)
30
74
61
49
70
60
50
35
-
15
13
12
33
James Basin. Jai^ies jnainstem/Norfolk Harbor
Norfolk Harbor
James Basin. Hampton
Hampton River
Hampton Roads Station 29
Hampton Roads Station 35
Hampton Roads Station 23
Hampton Roads Station 3A
Hampton Roads Station 34
Hampton Roads Station 1
Hampton Roads Station 25
Hampton Roads Station 25A
Hampton Roads Station 25B
Hampton Roads Station 36
all stations combined
River
OPC
HRM2
Station 33
HRM1
HYC
m
all stations combined
<0.01
-
-
O.01
-
<0.01
O.01
0.01
O.01
-
O.01
O.01
12.0
O.01
0.02
0.13
0.01
1.12
10
-
-
3
-
8
11
8
3
.
7
26
99
26
83
85
8Q
73
York Basin. Sarah Creek
Sarah Creek
Potomac
Choptank River
West
A
B
C
U
all stations combined
Potomac River
Choptank River
Hartge Marina
O.01
O.01
O.01
0.01
0.08
-
-
0.99
69
88
59
12
52
-
-
97
54
-------
Drainage
Severn
Mid-Bay
Chester
Patapsco
C&D Canal
Station
six Back Creek stations
Severn River
Pier 1 Marina
Piney Narrows Marina
Baltimore Harbor
C&D Canal
acute saltwater tests
- all species
(320 ng/L)
12.0
0.01
10.0
8.40
0.12
m
chronic saltwater
tests (5 ng/L)
>99
97
97
99
>99
—
55
-------
FIGURES
56
-------
Figure 1. Ecological risk assessment approach
PROBLEM FORMULATION
• Stressor Characteristics
• Ecosystems at Risk
• Endpoints
• Cumulative Stressors Impacting Aquatic Communities
• Conceptual Model
I
ANALYSIS
Characterization of Exposure: Water column monitoring data on tributyltin
in the Chesapeake Bay watershed.
Characterization of Ecological Effects: Laboratory toxicity studies.
RISK CHARACTERIZATION
Probabilistic comparison of species sensitivity and
surface water exposure distributions.
57
-------
Figure 2. Location of 41 stations where TBT was measured from
1985 to 1996. See key to map where stations are
^J *\ r* *% •**« V^ yt 4*4
described.
BALTIMORE
41
9
N
1-8
58
-------
Figure!, continued.
Station number
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
Drainage Basin
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
James
York
York
York
York
Potomac
Choptank
West
Severn
Severn
Mid-Bay Mainstem
Chester
Patapsco
C&D Canal
Dataset
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
NAVY
VIMS
VIMS
NAVY
VIMS
VIMS
VIMS
VIMS
VIMS
VIMS
VIMS
Hall Data
Hall Data
Hall Data
Hall Data
Hall Data
Hall Data
Hall Data
Hall Data
Hall Data
Station
Elizabeth River station 15
Elizabeth River station 17A
Elizabeth River station 19
Elizabeth River station 21
Elizabeth River station 32
Elizabeth River station 13A
Elizabeth River station 1 1
Elizabeth River station 10
Lafayette River station 37
Naval Station 9
Naval Station 4
Naval Station 3
Hampton Roads station 29
Hampton Roads station 35
Hampton Roads station 23
Hampton Roads station 3A
Hampton Roads station 34
Hampton Roads station 1
James River station 25
James River station 25A
James River station 25B
James River station 36
Hampton River station OPC
Hampton River station HRM2
Hampton River station 33
Hampton River station HRM1
Hampton River station HYC
Hampton River station CD
Sarah Creek station D
Sarah Creek station A
Sarah Creek station C
Sarah Creek station B
Potomac River
Choptank River
Hartge Marina
Back Creek (6 stations)
Severn River
Pier 1 Marina
Piney Narrows Marina
Baltimore Harbor
C&D Canal
Latitude
36.7755
36.7973
36.8031
36.8216
36.8323
36.8398
36.8472
36.8700
36.9065
36.9159
36.9491
36.9611
36.9453
36.9610
36.9613
36.9781
36.9849
36.9928
36.9988
37.0051
37.0142
37.0242
37.0005
37.0163
37.0164
37.0170
37.0205
37.0228
37.2458
37.2553
37.2597
37.2628
Longitude
76.2953
76.2938
76.2949
76.2920
76.2961
76.2757
76.3000
76.3295
76.3072
76.3416
76.3343
76.3322
76.3913
76.4365
76.4108
76.3497
76.3750
76.3017
76.4744
76.4925
76.4785
76.5252
76.3138
76.3442
76.3411
76.3417
76.3442
76.3438
76.5005
76.4797
76.4675
76.4850
see map for remaining locations
59
-------
Figure 3. Temporal trend of 90th percentile concentrations of
TBT for Sarah Creek from 1986-1996.
Sarah Creek 90th percentiles by year
CD
0
8
8
40
20 -
1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996
Year
60
-------
Figure 4. Temporal trend of 90th percentile concentrations of
TBT for Hampton Creek from 1986-1996.
Hampton Creek 90th percentiles by year
1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996
Year
61
-------
Figure 5. Distribution of TBT acute saltwater toxicity data.
TBT - saltwater acute effects
99 -r
90 -
I 70
.*
i 50
or
I
(D
0- 10
1 -_
10°
1Q1 1Q2 103 104 10s 106
TBT concentration (ng/L)
107
108
n = 43
b[0] = -4.3203754973
b[1] = 1.2130417922
• benthos
o fish
v phytoplankton
v zooplankton
regression line
62
-------
Figure 6. Distribution of TBT acute freshwater toxicity data.
TBT - freshwater acute effects
99 -f
102
103
10« 105 106
TBT concentration (ng/L)
10s
n = 23
b[0] =-2.583918915
b[1 ] = 0.6472125449
• amphibians
O cyanobacteria
T benthos
v fish
• phytoplankton
a zooplankton
regression line
63
-------
Figure 7. Distribution of TBT chronic saltwater toxicity data.
TBT - saltwater chronic effects
99 H
~ 90'
'to
® 70 -
i 50 -
cn
-------
Figure 8. Distribution of TBT chronic freshwater toxicity data.
TBT - freshwater chronic effects
99 -f
90 -
70 -
50 -
'55
®
CO
«
a:
I
Q- 10 -
1 -I
10
100
TBT concentration (ng/L)
1000
n = 3
b[0] =-11.3749917057
b[11 = 5.0255763015
• zooplankton
O fish
regression line
65
-------
APPENDIX A
Tributyltin risk characterization by basin and station
-------
Elizabeth River basin
CO
§
CO
o
Q.
X
0)
T3
-------
James River
Norfolk Harbor
99.9 -f
CO
O
CL
X
0)
•o
0)
o
-------
Hampton River drainage
§
CO
O
Q.
X
0)
•o
Q.
99.9 -
99 -
90 -
70 -
50 -
30 -
10 -
1 -
0.1
Probability of
exceedence = 73%
Probability of
exceedence
= 1%
i 1—i—n-
1 10 100
TBT Concentration (ng/L)
1000
A-3
-------
co
2
D
CO
O
O.
X
0)
0.
o
-------
Hartge Marina
99
CO
9?
3
CO
O
Q.
X
0
T3
Q)
^
fi
M—
O
0
Q.
90 -
70 -
50 -
30 -
10 -
Probability of
exceedence = 97%
Probability of
exceedence
= 0.99%
10 100
TBT Concentration (ng/L)
1000
A-5
-------
Six Back Creek Stations
CO
2
13
CO
o
Q.
X
99%
Probability of
exceedence =12%
10
100
TBT Concentration (ng/L)
1000
A.-6
-------
Severn River
99 -f
o
Q.
X
0)
T3
0)
90 -
70 -
C 50 -
(0
*0 30
0)
c
-------
Pier 1 Marina
CO
o
Q.
X
(D
•D
0)
J^
I
(D
Q.
Probability of
exceedence= 10%
Probability of
exceedence = 97%
10 100
TBT Concentration (ng/L)
1000
A-8
-------
Piney Narrows Marina
99
to
o
Q.
X
(D
-------
Baltimore Harbor
99
to
§
CO
o
Q.
T3
(U
90 -
70 -
to 50 -
*o
J 30 H
I
0 m
n iU
Probability of
exceedence = >99%
'o
8.
M
"co
_»
1
u
'c
I
u
(A
_g>
"Q
8.
tf)
"5
0)
o
0)
•s
u
CD
10 100
TBT Concentration (ng/L)
1000
A-10
------- |