United States
Environmental Protection
Agency
Technology Transfer
EPA/625/8-89/015
Biomonitoring for
Control cjf Toxicity in
Effluent Discharges to
the Marine Environment
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EPA/G2S/8-89/015
Biomonitoring for
Control of Toxicity in
Effluent Discharges to
the Marine
Environment
September 1989
Center for Environmental Research Information
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, OH 45268
Environmental Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Narragansett, Rl 02882
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This document was written by Pamela DiBona, William
Heyman, and Heidi Schultz of Eastern Research Group, Inc.,
Arlington, Massachusetts. Production assistance was provided
by Susan Edwards. Technical assistance was provided by
Steven Schimmel, William Nelson, and Donald Phelps of the
EPA Environmental Research Laboratory (ERL) in
Narragansett, Rhode Island; Orville Macomber of the EPA
Center for Environmental Research Information; William Peltier,
EPA Region IV, Athens, Georgia; and Rod Parrish, ERL-Gulf
Breeze. Florida.
The principal investigator at ERL-Narragansett was Steven
Schimmel. Assistance was provided in the development of the
Complex Effluent Toxicity Testing and Biomonitoring Programs
at ERL-Narragansett by Donald Phelps, William Nelson,
George Morrison, Suzanne Lussier, Richard Steele, Margarete
Heber (now at EPA Headquarters, Washington, DC), Diane
Nacci, Glen Thursby, Elise Torello, Ruth Gobell-Gutjahr, David
Bengtson, and Walter Berry. Contributions were provided by
Rod Parrish, ERL-Gulf Breeze, relating to toxicity testing of
drilling muds and produced waters. William Peltier and staff at
EPA Region IV (Athens, Georgia) conducted case studies and
provided data and assistance at two Florida locations.
Photographs were provided by ERL-Narragansett. This
document is ERL-Narragansett Contribution Number 1036.
Technical review of the document was provided by Dr. Gary
Chapman, ERL- Newport, Oregon; Margarete Heber, EPA
Office of Water Enforcement and Permits; Peter Nolan, EPA
New England Regional Laboratory, Lexington, Massachusetts;
Rod Parrish, EPA ERL-Gulf Breeze, Florida; William Peltier,
EPA Region IV, Athens, Georgia; and Landon Ross, Florida
Department of Environmental Regulation, Tallahassee, Florida.
This report has been reviewed by the U.S. Environmental
Protection Agency and approved for publication. The process
alternatives, trade names, or commercial products are only
examples and are not endorsed or recommended by the U.S.
Environmental Protection Agency. Other alternatives may exist
or may be developed. In addition, the information in this
document does not necessarily reflect the policy of the
Agency, and no official endorsement should be inferred.
This guidance was published by
U.S. Environmental Protection Agency
Center for Environmental Research Information
Office of Research and Development
Cincinnati, OH 45268
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Foreword
This document describes the use of biological monitoring
as an effective water quality -based approach to
controlling the toxicity in discharges to estuarine and
marine waters. The development of these methods was
conducted primarily at EPA's Environmental Research
Laboratory in Narragansett, Rhode Island (ERL-N). This
document is intended for Federal and State National
Pollutant Discharge Elimination System (NPDES) permit
writers, water quality specialists, managers of pollution
control systems, regulatory decisionmakers, research
scientists, and informed laypersons.
The document is organized into seven sections. Section
1, Introduction, is an overview of the regulations that
mandate a change to water quality-based permit limits
and the testing programs that support that change.
Section 2, Marine/Estuarine Complex Effluent Toxicity
testing Program (CETTP), describes EPA's efforts to
establish the validity of whole-effluent toxicity testing for
marine/estuarine environments. The section describes the
test methods developed to determine the effects of whole
effluents on survival, growth, and reproduction of several
test species. Section 3, In Situ Biomonitoring, presents a
program to determine the biological responses of marine
and estuarine test species to pollutants. Section 4,
Integration of Effluent and In Situ Biomonitoring Programs,
describes how the complex effluent and in situ
biomonitoring programs may be used in tandem to control
toxic discharges to marine waters. Section 5, Application
of the Biomonitoring Strategy: New Bedford Harbor Pilot
Dredging Program, describes how ERL-N employed the
tandem approach described in the previous section to
monitor cleanup options for one of the most polluted
industrial harbors in the country. Section 6, Conclusions
and Future Directions, discusses possible refinements to
existing marine biomonitoring programs and their use in
NPDES permitting. Appendix A summarizes the results of
five case studies in which the CETTP and in situ
biomonitoring programs were evaluated, and Appendix B
is a glossary of terms useful in understanding the
monitoring programs.
in
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Contents
1.
2.
3.
Introduction 1
Regulatory Background 1
Special Considerations for Marine and Estuarine
Waters
1
Complex Effluent Toxicity Testing
Program 2
In Situ Marine and Estuarine Biomonitoring
Program 2
Marine/Estuarine Complex Effluent Toxicity
Testing Program 5
General Introduction to the Marine/Estuarine
Complex Effluent Toxicity Testing
Program 5
Red Algal Sexual Reproduction Test
Method 5
Kelp Sexual Reproduction Test
Method 6
Sea Urchin Fertilization Test Method 8
Mysid Survival, Growth, and Fecundity Test
Method 9
Sheepshead Minnow Embryo/Larval Survival and
Teratogenicity Test 9
Sheepshead Minnow Larval Survival and Growth
Test 9
Inland Silverside Larval Survival and Growth
Test 9
A Special Case: The Toxicity of Drilling Fluids -
Mysid Static, Acute Toxicity Test
Analysis of Test Results 11
Field Verification of Test Methods
Test Precision 12
Relative Sensitivity of the CETTP Tests
Incorporating Test Results into NPDES
Permitting 13
In Situ Biomonitoring 15
Historical Perspective 15
Biological Methods 16
Chemical Methods 18
Validation of Test Methods with Case
Studies 18
11
12
13
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Contents (continued)
Comparisons among Various Levels of Biological
Organization 20
The Integration of Biomonitoring into Marine
Discharge Regulations 20
Current Research 22
Summary
22
4. Integration of Effluent and In Situ Biomonitoring
Programs 23
Complex Effluent Toxicity 23
Receiving Water Toxicity Estimates 23
Integrated Approach for Toxic Discharge
Control 24
5. Application of the Biomonitoring Strategy: The
New Bedford Harbor Pilot Dredging
Project 25
Introduction 25
Dredging Project Description 25
Monitoring Strategy 25
Summary 28
6. Conclusions and Future Directions 33
Complex Effluent Toxicity Testing 33
Future Research and Refinements to Complex
Effluent Toxicity Testing Program 33
In Situ Biomonitoring 33
Future Research and Refinements for In Situ
Biomonitoring 33
Integrated Approach for Toxic Discharge
Control - Integration with Large-Scale Coastal
Monitoring 34
The Integrated Approach - Field Tested 34
7. References
35
Appendix A. Case Studies 39
CETTP Case Study I: Fernandia Beach,
Florida 39
CETTP Case Study II: East Greenwich, Rhode
Island 40
VI
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Contents (continued)
CETTP Case Study Ml: Panama City,
Florida 43
In Situ Biomonitoring Case Study: East Greenwich,
Rhode Island * 44
Appendix B. Glossary
57
vii
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1. Introduction
Regulatory Background
In 1972, the National Pollutant Discharge Elimination
System (NPDES) was established under the Clean Water
Act to control point source pollution to the Nation's
surface waters. Point sources of pollution can be directly
attributed to a specific source, such as a factory or
publicly owned treatment works. Every point source must
apply for and obtain an NPDES permit that limits the
concentrations of pollutants that may be discharged in its
effluent. EPA has delegated permit-setting activities to
States whose permit programs meet applicable EPA
requirements; these States operate their own NPDES
programs with EPA overview.
In 1984, EPA issued a policy statement recommending
effluent biomonitoring as one part of a "water quality-
based approach to controlling toxic pollutants" (1). The
policy directed States to use data from biological testing
of complex effluents discharged from point sources for
setting NPDES permit limits. EPA also prepared a
technical support document (2) that described ways to
implement the policy. This policy made biological testing
an integral part of the NPDES program under the Clean
Water Act. and effectively changed the permit-setting
emphasis from a technology-based approach to a water
quality-based approach. The 1987 amendments to the
Clean Water Act (now called the Water Quality Act)
reinforced this policy by stating that water quality
standards should be chemical-specific numbers when
such numbers have been established, however biological
monitoring should be used in areas where no criteria for
aquatic life exist and to address the effect that the
interaction of chemicals has on toxicity (3). EPA made a
more definite step toward water quality-based control in a
proposed rule published in the Federal Register in
January, 1989 (4), which states that "controls for whole
effluent toxicity are an essential component of EPA's
integrated approach to toxics control."
To control the discharge of toxics under the toxicity-based
approach, EPA recommends that complex effluent limits
be used in conjunction with limits on specific pollutants. In
many cases aquatic toxicity data for specific chemical
compounds are often not available for use in setting
standards. Also, wastewaters usually contain mixtures of
chemicals that are difficult for industries or municipalities
to characterize. Compounding these problems is the fact
that even if pollutants are discharged at toxic
concentrations below analytical detection limits, the
effluent may still be toxic. This toxicity may be due to a
variety of factors: an excess of nonconventional pollutants
(such as total suspended solids); conventional pollutants;
unanticipated interactions between the chemical mixtures
in the effluent; and/or unanticipated interactions between
the effluent and the receiving water. Complex effluent
toxicity testing takes these factors into account and
complements the chemical-by-chemical approach.
Special Considerations for Marine and Estuarine
Waters
Since the 1984 shift to a water quality-based approach,
biomonitoring methods in freshwaters such as lakes and
rivers have been well established. Translating these
methods from freshwater to marine environments,
however, is not a simple task. Differences between the
two systems require alternative testing methods and test
species. In addition, the physical differences between
streams and estuaries, rivers and bays, as well as the
chemical differences between saltwater and freshwater,
require special consideration.
Physical Differences
Many physical conditions affect the mixing of effluents in
marine/estuarine and freshwater systems. Marine and
estuarine systems, however, are much more complex -
water depth and current direction vary with the tides, and
tidal volume varies on a monthly and seasonal basis.
Models that predict effluent mixing must take these and
other factors into account. For example, an outgoing tide
may quickly carry the effluent plume into deeper waters,
but if little vertical or lateral mixing occurs along the
plume, then a concentrated effluent could be delivered to
a critical resource area (oyster beds, spawning grounds,
etc.). In addition, estuaries tend to act as chemical
"sinks." Stratification can trap effluents below horizontal
temperature gradients in the water, and the tide's
oscillation can restrict the plume from significant
horizontal movement (2). The contaminants sit and collect
in what becomes a basin for pollution. Finally, effluent
plumes may mix or disperse less easily in saltwater than
in freshwater because of differences in salinity and
specific gravity between the effluent and the surrounding
waters.
Chemical Differences
The buffering capacity of saltwater is much greater than
that of freshwater. Certain chemicals either break down
more easily or remain in their original toxic form longer in
saltwater than they would in freshwater. Also, the
speciation or form of the individual chemical mixture
components (e.g., solid, liquid, ion, or complexed) varies
depending on whether the effluent is in saltwater or
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freshwater. This is an important factor because some
chemicals are accumulated or degraded by organisms in
one form, but not in another.
Biological Differences
'Marine and estuarine waters, in general, support a greater
diversity of species than do freshwater rivers and lakes.
In addition, unlike rivers and lakes, in which fish may
spawn and live the rest of their lives in one area,
estuaries and bays act as nursery grounds for many
species that later move out into open waters. Exposure to
toxic effluents is more likely to affect these immature,
more sensitive life stages.
Resulting Differences in Approach
These physical, chemical, and biological differences
between marine and freshwater environments require
different effluent and organism sampling methods as well.
Because of a marine or estuarine water body's depth and
expanse, and the possibility of disturbances from violent
storms, marine biomonitoring logistics are more difficult to
plan and carry out. In freshwater biomonitoring, benthic
surveys are frequently used. These surveys involve
cataloging all of the species inhabiting the benthos, and
characterizing the "health" of the water body on that
basis. However, the benthic zone in marine environments
often may not be directly affected by toxic discharges.
The discharge, if released in a freshwater plume, flows up
through the water column to the surface of the water and
usually does not flow along the bottom. Suspended
solids, however, may settle to the bottom. Furthermore,
findings from biological surveys of benthic species may
indicate residual effects from past pollution. Estuaries
tend to serve as depositional areas, therefore many
discharges, particularly municipal discharges, superficially
enrich the benthic sediments with organic carbon. This
enrichment by itself, in the absence of toxic chemical
contamination, can substantially alter the biological
community.
Populations of species within the marine community also
change naturally as they move, die out, or populate the
benthos in a patchy or irregular fashion, making the
detection of adverse effects from pollutants more difficult.
The ambient receiving water toxicity tests used in marine
and estuarine environments measure the effects of the
suspended plume on indigenous species. In situ
biomonitoring methods provide another alternative for
measuring the effects of effluents on marine species. This
permits the effects of any toxicity present in the waters
surrounding the discharge to be measured directly,
without relying on laboratory exposure systems.
The great diversity in species present in the marine
environment affects species selection for toxicity testing
as well. Animal and plant kingdoms both must be
represented in testing. All of these factors were
considered when adapting freshwater methodologies to
marine conditions and species.
Complex Effluent Toxicity Testing Program
The Complex Effluent Toxicity Testing Program (CETTP)
was initiated as a means of testing whole effluents using
aquatic organisms as indicators of effect. The overall goal
of the CETTP was to enable regulators to use the
parameter toxicity data to set NPDES permit limits based
on water quality protection. Researchers first applied the
technology to freshwater systems. From the program's
outset in 1982, EPA has studied the toxic effects of
effluents in several rivers in the United States, with five
goals in mind:
• To verify that complex effluent toxicity tests predict
and quantify adverse impacts to receiving waters.
• To identify complex effluent toxicity test procedures
that support the NPDES regulatory process.
• To demonstrate, through case studies, the
effectiveness of tests in a variety of discharge
scenarios.
• To field test short-term chronic toxicity tests.
• To compare the receiving water toxicity with the
effluent toxicity observed at known concentrations
(5).
The freshwater CETTP resulted in several biological tests
that established the validity of whole effluent toxicity
testing.
EPA's marine CETTP set out to achieve the same five
goals described above for the freshwater portion of the
CETTP. To adequately address the differences between
freshwater and marine/estuarine water systems, the
number of test species was expanded, and ambient
receiving water toxicity tests were used rather than
biosurveys. The biomonitoring tests described here were
designed for the measurement of toxicity to the
environment where relatively small ratios of effluent
volume to receiving water volume exist. Freshwater
receiving streams may contain more than 6>0 percent
effluent after discharge, but most estuaries and bays, by
virtue of their potentially greater volume, generally provide
higher effluent dilution.
In Situ Marine and Estuarine Biomonitoring
Program
The in situ marine biomonitoring program focuses on
determining not only contaminant concentrations, but also
contaminant effects. This program extends the CETTP
policy of emphasizing specific effects rather than specific
chemicals. The five objectives of the program are:
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I
• To identify general problem areas or conditions of
the Nation's bays and estuaries and other near-
coastal waters.
• To track dischargers' response to regulations by
monitoring point source discharges and nonpoint
sources of pollution.
• To identify links between chemical pollutants and
biological responses.
« To identify links between impacts at different levels
of biological organization.
• To predict population- and community-level effects
from the specific effects of chemicals on individual
organisms.
In situ biomonitoring measures biological response in
individual organisms. Eventually, the tests will monitor
biological parameters at four levels of biological
organization: cellular, tissue/organ, individual, and
population.
The biomonitoring strategy utilizes a stepwise progression
of sensitive tests, from simple to more complex and time-
consuming; the first tests are performed on sensitive
species, and provide information that enables the
investigator to determine if other, more complex and
expensive tests are needed. The strategy brings together
regulators and scientists in a single decisionmaking
process. Together, they decide whether the problem
justifies additional testing, and if so, which tests should be
used (6).
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2. Marine/Estuarine Complex
Program
General Introduction to the Marine/Estuarine
Complex Effluent Toxicity Testing Program
The 1984 amendments to the Clean Water Act
recommend the use of biological test methods to help
.prevent the "discharge of toxic materials in toxic
amounts." The 1987 Water Quality Act further supports
EPA's increased use of biological test methods in setting
National Pollutant Discharge Elimination System (NPDES)
permit limits to protect our Nation's surface waters for
their designated uses. In response to the 1984
amendments, researchers at EPA's Environmental
Research Laboratory in Narragansett, Rhode Island (ERL-
,N) developed the marine component of the Complex
'Effluent Toxicity Tesjing Program (CETTP). The program
includes tests that estimate "safe" or "no effect"
concentrations for toxic effluents and thus predict
allowable effluent concentrations that should provide for
the normal propagation of fish and other aquatic life in
marine receiving waters.
U.S. EPA's Environmental Research Laboratory in Narragansett,
Rl.
Historically, effluent toxicity tests measured only acute or
short-term (generally less than 96 hours) effects with
lethality as the test endpoint. Such methods and their
application are provided in "Methods for Measuring the
Acute Toxicity of Effluent to Freshwater and Marine
Organisms" (7). Since actual field exposures often occur
at sublethal concentrations, EPA realized the need to
directly test the effects of chronic (longer-term, lower
concentration) exposures. The 1985 acute methods
manual provides a means to protect for chronic effects by
Effluent Toxicity Testing
extrapolation from the acute toxicity determinations.
Previously definitive chronic toxicity testing required a
minimum of 30 days up to a full year to measure the
actual effects of low toxicant concentrations on the entire
life cycle of an organism. These full life cycle tests were
developed with several fish species and with mysids but
virtually no other species. Since the observed toxicity in
these tests was most strongly manifested during the
organisms early and most vulnerable life stages, the tests
were abbreviated to include only these life stages. In an
effort to produce comparable, reliable results within much
shorter time periods and at reasonable costs, tests were
developed that focused on several other organism's
vulnerable life stages. Test endpoints were targeted to
survival, growth, fecundity, and reproduction. Eventually,
abbreviated marine chronic tests were developed that
estimate long-term toxicity in less than 9 days. The
CETTP tests evaluate the biological (toxicity) effects of
complex effluents or receiving waters. The tests,
therefore, include the additive effects of complex effluents
on several important life functions of exposed organisms.
Researchers developed tests for six species in four
phylogenetic groups under the CETTP (Table 2-1). The
kelp toxicity test has just recently been developed for the
CETTP. All of these tests may be used to develop permit
limits and assess compliance for toxic discharges into
marine/estuarine environments. The techniques in these
tests are designed to minimize equipment and sample
volume so that they can be conducted either on site, in a
mobile laboratory, or at an offsite lab, by shipping minimal
volume samples of effluents or receiving waters.
The following sections briefly describe these tests.
Detailed descriptions of the test methods are provided in
"Short-term Methods for Estimating Chronic Toxicity of
Effluents and Receiving Waters to Marine and Estuarine
Organisms" (8) and "User's Guide to the Conduct and
Interpretation of Complex Effluent Toxicity Tests at
Estuarine/Marine Sites" (9). General guidelines for
conducting short-term chronic toxicity tests are presented
in Table 2-2. EPA also is currently developing tests with
additional marine plant and animal species from other
phylogenetic groups and other geographic locations.
Red Algal (Champia parvula) Sexual
Reproduction Test Method
This test method estimates the chronic toxicity of
effluents and receiving waters on the sexual reproduction
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Table 2-1. Test Species, Response Parameters, and Test Duration of the
Marine/Estuarine Short-term Complex Effluent Toxicity Tests (8)
Species Test Duration Test Endpoint(s)
Champia parvula
(Red macroalgae)
7 to 9 days total;
2-day exposure, 5- to
7-day incubation
Cystocarp
production
Laminaria saccharina
(Kelp)
Arbacia punctulata
(Sea urchin)
5 to 9 days total;
2-day exposure, 3- to
7-day incubation
1.33 hours total;
1 -hour sperm exposure,
20-minutes fertilization
Sporophyte
production
Egg
fertilization
Mysidopsis bahia
(Mysid)
Cyprinodon variegatus
(Sheepshead minnow)
Cyprinodon variegatus
(Sheepshead minnow)
Menidia beryllina
(Inland silversides)
7 days .total;
7-day exposure
9 days total;
9-day exposure
7 days total;
7-day exposure
7 days total;
7-day exposure,
Survival, growth, and
egg production
Survival without
terata
Survival and
growth
Survival and
growth
ERL-N's mobile laboratories.
of the marine macroalga, Champia parvula. Seaweeds
had previously been considered far less sensitive to
toxicity than aquatic animals, and thus not useful for
toxicity testing. However, this conclusion was based on
algal growth as the test endpoint. Algal reproduction tests,
including those with Champia, can sometimes be more
sensitive than tests with aquatic animals.
Unialgal stock cultures of Champia parvula should be
maintained, so that mature plants are available year round
for toxicity testing. There are three macroscopic stages in
the life cycle of Champia, and all three look the same
superficially (Figure 2-1); however, only the male and
female plants are used in toxicity testing. The female
production of cystocarps, which are the products of red
algal sexual reproduction, is used as the endpoint.
The toxicity test with Champia consists of first combining
male and female plants and exposing them for 48 hours,
during which time fertilization should occur, and then
allowing cystocarps to develop over a 5- to 7-day
recovery period. Five female branches and one male
branch are placed together in each test chamber, with
three or four replicate chambers for each treatment. After
exposure to the effluent, females are removed and placed
into clean seawater. At the end of the recovery period,
the number of cystocarps per plant are counted. The
results are compared to a control in a series of statistical
tests to calculate the "no effect" concentrations of the
effluent.
Kelp (Laminaria saccharina) Sexual
Reproduction Test Method
ERL-N recently developed this test method for use in the
CETTP (10); therefore, the method is not included in the
original test manual (8). The method of estimating the
chronic toxicity of whole effluents and receiving waters on
the sexual reproduction of the brown macroalga
Laminaria saccharina is similar to that of the Champia
test. This test is an excellent complement to the Champia
test because, while Champia is found in warm waters,
kelps are primarily found in cold waters throughout the
west coast of North America and in New England on the
east coast.
ERL-N maintains unialgal stock cultures of male and
female Laminaria that are available for toxicity testing. Six
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Table 2-2. General Guidelines for Short-term Marine Complex Effluent Toxicity Test Methods3 (8,9)
Guideline
Description of Guidelines
Health and safety
Quality assurance
Facilities and equipment
Effluent samples
Appropriate measures should be
effluents or receiving waters.
aken to protect against possible health hazards derived from
Dilution water
All tests should be performed by aquatic toxicologists or under their supervision. Tests require high-
quality test conditions to ensure confidence in test results. Quality considerations include: animal
health, sample collection and han lling methods, instrument condition and calibration, adequate
controls and replication, accurate recordkeeping, use of reference toxicants, and test acceptability
criteria.
Tests can be conducted in an appropriately equipped stationary or mobile laboratory.
Water samples should be collected just prior to the start of tests. Water samples are generally
collected as 24-hour composite samples,b which are appropriate when the components of an
effluent remain relatively constant in volume and abundance. (Note, however, that this sampling
method may not adequately represent short-term peaks in toxicity. This type of toxicity must be
addressed using acute tests on grab samples.) Samples of whole effluents are collected at the
discharge site, while receiving wa ers are collected at a series of predetermined sampling stations,
representative of varying degrees of impact.
A range of test solutions is create i with varying effluent concentrations to determine the
concentration of effluent that is toxic to the organism. Five different effluent concentrations and a
control containing no effluent are selected for each test. Careful steps are taken to ensure that
salinities in all samples are identical, and remain constant throughout the experiment. Salinity is
adjusted with the aid of either artificial seawater (for some but not all species) or with hypersaline
brine, made from evaporating higr
quality, filtered seawater. Seawater of desired salinity is created
by mixing a proportion of brine wit n deionized water.
Endpoints for chronic toxicity Endpoints in these tests include survival, growth, fecundity, and reproduction.
tests I
Statistical analysis of test The statistical analysis of these tests should be conducted by or with the aid of a qualified
results statistician. Through a series of statistical tests, two different estimates of the toxicity of an effluent
are made: the NOEC (No-Obserjed-Effect Concentration), an estimate of the highest concentration
that will not have a significant effect on the organism; and the EC (Effective Concentration), an
estimate of the proportion of orgarjisms that will show effects at any given concentration.
aThese guidelines refer to the conduct of all tests mentioned in this section.
bComposite sampling is generally conducted automatically with a co lection apparatus placed adjacent to a discharge. A 5-liter water
sample is often collected over a 24-hour period by collecting small volumes of water at regular intervals.
days prior to testing, males and females are placed
separately in a standard blender and are blended to
provide smaller cell clusters, filtered to remove remaining
large clusters, and diluted. Males are placed in petri
dishes and females on small glass slides, 100 to 150
females per slide. The cultures are maintained for 5 days
in a high iron medium to allow gametogenesis to occur
(Figure 2-2). If significant pregametic cells are identified
after 6 days, the water in male dishes is replaced either
with treatment (various effluent dilutions) or with control
seawater. One slide of female parts is then placed in
each of the male dishes. Exposure is for 48 hours, after
which the females are removed and placed into clean
seawater. After an additional 4 to 7 days, the slides are
examined for sporophytes, the products of Laminaria
sexual reproduction. The results are expressed as the
number of sporophytes in test conditions compared to the
number of sporophytes that develop in control seawater.
The results are statistically analyzed and the "no effect"
concentrations of the effluents are derived.
Robust sample of the marine macroalga Champia parvula.
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Tetrasporangia^
Spermatia
Fertilization
Tetrasporophyte
-Cystocarp
Figure 2.1. Sexual reproduction in the Red Marine Macroalga
Champ/a parvula (8).
Cystooarps in a mature female Champ/a parvula.
Sea Urchin (Arbacia punctulata) Fertilization Test
Method
This test method, adapted from a method by Dinnel et al.
(11), estimates the chronic toxicity of effluents or
receiving waters to the gametes of the sea urchin
(Arbacia punctulata). Arbacia adults are hardy and can be
held in aquaria for long periods. This species can be
induced to spawn year round, making it especially
convenient to work with. Because this test can be
conducted rapidly {1 day), it can be used to quickly test a
multitude of effluent types and locations.
Adult A. punctulata can be collected easily along the
Atlantic coast or purchased from commercial suppliers.
Once obtained, they are separated into male and female
cultures. Male gametes are released by stimulating the
Laminaria saccharina
Microscopic
Antheridia
Male
eggs
Figure 2-2. Sexual reproduction in the Kelp Laminaria
saccharina (10).
Adult sea urchins Arbacia punctulata.
organism with a 12-volt electrode. The sperm are
collected and iced for not more than 1 hour. Female
gametes are collected similarly and may be maintained at
room temperature for several hours before use.
Urchin sperm are exposed to effluents or receiving waters
for 1 hour. Then eggs are added to the sperm and
allowed to incubate. After 20 minutes of incubation, the
test is terminated by the addition of a preservative. The
test is conducted with three or four replicates of each
treatment, each containing 2,000 eggs and about 5 million
sperm.
Percentage fertilization is calculated after examining a
subsample from each test vial under a dissecting
microscope. The test results are recorded as the
percentage of eggs fertilized in test solutions compared to
the percent fertilized under control conditions. The results
-------
are used in a series of statistical tests to calculate the
; "no effect" concentrations of the effluent.
Mysid (Mysidopsis bahia) Survival, Growth, and
Fecundity Test Method
This test method estimates the chronic toxicity of
complex effluents and receiving waters to the vulnerable
, period of egg development of the mysid (Mysidopsis
bahia). Maturing juveniles are exposed to various
concentrations of effluents or receiving waters in a 7-day
test. Each day, new sample water is collected and
prepared with appropriate effluent concentrations,
salinities, and food concentrations, and maintained under
appropriate environmental controls. Eight replicates of
each concentration are maintained with five mysids in
each treatment to ensure the statistical significance of
test results. The test is completed after the mysids are
exposed to the test solutions for 7 days.
At the termination of the test all mysids are transferred to
clean seawater for observation. The sexes are determined
and the number of immature mysids, mature males,
mature females with eggs, and mature females without
eggs are recorded (Figure 2-3). Dry weights are
; determined to calculate growth. The test results compare
mysid survival, growth (weight), and fecundity (the
percentage of females carrying eggs) in test solutions to
i mysid survival, growth, and fecundity in clean seawater
controls. Statistical tests are used to calculate the "no
effect" concentrations of the effluent.
Sheepshead Minnow (Cyprinodon variegatus)
Embryo/Larval Survival and Teratogenicity Test
This test method estimates the chronic toxicity of whole
effluents and receiving waters to the early developmental
stage of the sheepshead minnow (Cyprinodon
variegatus). (This method was developed at EPA Region
,VI Laboratory, Houston, TX.) Newly fertilized (less than 24
hours old) C. variegatus embryos are exposed to a range
of effluent concentrations or receiving waters from shortly
after fertilization, through hatching, and for 4 days into the
larval period. Each day, new sample water is collected
and prepared with appropriate salinity, effluent
concentrations, and other environmental controls. No food
is required in this test because the extremely young
; larvae can survive on the nutrients provided in their yolk
sacs. Three or four replicates of each concentration are
maintained with 10 to 15 embryos in each treatment to
increase the significance of test results. The test is
completed, after 9 days of total exposure or 4 days of
post-hatch exposure, whichever comes first.
At test termination, all live fish larvae without obvious
terata are counted to determine survival in each test
chamber. Survival is counted as those healthy larvae
without deformity (terata) at the completion of the
exposure period. There is no measurement of growth in
this test. The test results are acceptable if the larvae in
control cultures show survival rates greater than 80
percent. A series of statistical tests again are used to
calculate "no effect" concentrations of the effluent from
the results of survival and terata determinations.
Sheepshead Minnow (Cyprinodon variegatus)
Larval Survival and Growth Test
This test method estimates the chronic toxicity of whole
effluents and receiving waters to the sheepshead minnow
(Cyprinodon variegatus). By focusing on the vulnerable
early larval stage, the test estimates the impact on the
entire life cycle of the fish.
In the test, newly hatched (less than 24 hours old) fish
larvae are exposed to various concentrations of effluents
or receiving waters for 7 days (Figure 2-4). The salinity
range suitable for conducting the test is 20 to 32 parts
per thousand (ppt); an appropriate salinity is selected and
is held constant throughout the experiment. Each day,
new sample water is collected and prepared with
appropriate dilutions, salinity, dissolved oxygen (DO), and
temperature. Three or four replicates of each
concentration are maintained with 10 to 15 larvae in each
treatment. After the new water is added each day, the
larvae are fed to excess with newly hatched larval brine
shrimp, Artemia sp. The test is completed after the
minnows are exposed to the test solutions for 7 days.
At test termination, all healthy larvae are first counted to
determine survival, then sacrificed, dried, and weighed to
determine growth. The test results compare the larval
growth or survival in test solutions to the larval growth
and survival in clean seawater controls. The test results
are acceptable if the larvae in the control cultures (no
effluent) show survival rates greater than 80 percent and
acceptable dry weights (8). A series of statistical tests are
used to calculate the "no effect" concentrations of the
effluent.
Inland Silverside (Menidia beryllina) Larval
Survival and Growth Test
This test method estimates the chronic toxicity of whole
effluents and receiving waters on the vulnerable larval
stage of the inland silverside (Menidia beryllina) (Figure 2-
5). This fish species is an important dietary component of
many commercially harvested species including mackerel,
bluefish, and striped bass. These minnows are tolerant of
a large salinity range (<5 to 32 ppt) and can therefore be
useful for effluent testing in estuaries where salinities may
be lower than in open ocean environments.
Test methods for this species are nearly identical to those
for the sheepshead minnow survival and growth tests
described previously. Seven- to 11 -day-old larvae are
exposed to various concentrations of effluents or
receiving waters for 7 days. Results of this test are
recorded as the comparison between larval growth
-------
Eyestalk
Antennule
Carapace
Developing Brood Sac
Oviducts with Developing Ova
Figure 2-3. Mature female mysld Mysidopsis bahia with eggs in the brood sac (8).
Cyprinodon variegatus
Sheepshead Minnow
Conducting the sheepshead minnow toxicity test in ERL-N's
mobile laboratory.
(weight) or survival in test solutions and larval growth and
survival in clean seawater controls. Test results are
acceptable if larvae in control conditions exhibit survival
rates greater than 80 percent and acceptable dry weights
7-Days Old
Figure 2-4. Early development of the juvenile sheepshead
minnow Cyprinodon variegatus (8).
(8). The results of survival and growth determinations are
put to a series of statistical tests and the "no effect"
concentrations of the effluent are derived.
10
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Menidia beryllina
Inland Silverside
1 Figure 2-5. Inland silverside minnow Menidia beryllina (12).
A Special Case: The Toxicity of Drilling Fluids -
Mysid (Mysidopsis bahia) Static, Acute Toxicity
Test
Oil and gas drilling operations produce effluents that are
: frequently discharged to marine/estuarine environments
(Figure 2-6). Drilling fluids, an added constituent used to
lubricate the drill bit, contain primarily clays and minerals
(for example barium sulfate, bentonite, lignite, and
1 lignosulfonate), but also contain materials used as
dispersants and thinners, fluid loss reducers, lubricants
and emulsifiers, defoamers, biocides, corrosion inhibitors,
and pH control/stability additives (13). Most commonly,
the toxicity of drilling fluids is assessed by determining
, the acute toxicity of the fluids to mysids, Mysidopsis
bahia, according to the method developed at EPA's
Environmental Research Laboratory in Gulf Breeze,
' Florida (ERL-GB).
In the 96-hour test, 3- to 6-day-old mysids are exposed to
various concentrations of drilling fluids. Samples are
prepared by mixing the drilling fluid with seawater (1 part
fluid: 9 parts seawater, v:v), allowing the sample to settle
for 1 hour, and then decanting only the suspended
paniculate phase (SPP) for toxicity testing. If necessary,
the pH is adjusted to the seawater's original alkalinity and
the dissolved oxygen concentration is increased by
; aeration to at least 60 percent of saturation. A range-
finding test should be conducted with samples of
unknown toxicity to determine the range of concentrations
for the definitive test. The definitive test uses mysids from
a single hatch that are randomly distributed into five
concentrations of the SPP plus a control. Three replicates
of each concentration are maintained with 20 mysids in
each. All treatments are gently aerated for the duration of
the test. Mysids are fed Artemia at a rate of 50 Artemia
per mysid per day. Temperature, dissolved oxygen, and
pH are measured and recorded after 0, 24, 48, 72, and
96 hours.
After 96 hours of exposure, the test is terminated and the
number of live mysids in each treatment is recorded. The
test results are acceptable if at least 90 percent of the
mysids in control conditions survived. A series of
statistical tests are then used to estimate the median
toxic concentration of the drilling fluid.
The toxicity of drilling fluids is not routinely assessed with
the chronic toxicity tests thus far described because of
the physical characteristics of the fluids which may
contain materials of varying densities that separate upon
settling. A special test apparatus that maintains particles
in suspension could be adapted so that the CETTP tests
could be used to assess potential impacts of drilling
fluids.
Analysis of Test Results
The results of the CETTP tests are used to estimate the
interval between the highest "no effect" concentrations of
the effluents and the lowest effect concentrations of the
effluents tested. The statistical methods employed are
rather complex and should be handled by, or with the aid
of, a statistician. U.S. EPA has prepared a complete
explanation of these methods (8). The environmentally
"safe" concentration of an effluent is the highest
concentration of that effluent that will not have any
adverse effects on the organisms in a receiving water.
This "safe" or "no effect" concentration of an effluent is
estimated with toxicity testing. By using statistical
methods on the same set of data from a single test, two
different expressions of toxicity are calculated.
The first method of statistical evaluation uses a threshold
model of toxicity. The method can be used to derive the
No-Observed-Effect Concentration (NOEC) and the
Lowest-Observed-Effect Concentration (LOEG) of the
effluent. The use of NOECs and LOECs assumes both of
the following premises. First, the model assumes that a
true threshold concentration exists, below which there is
no adverse effect to the organism and above which there
is an adverse effect. In addition, the model assumes that
adverse effects that are not statistically observable are
not significant from a biological standpoint. If a threshold
model of toxicity is accepted, NOECs will express the
highest concentration of the effluent that will not
adversely affect the organism tested.
The second method of statistical analysis is based on a
continuous or concentration-dependent model of toxicity.
Alternatively to (but not inconsistent with) the threshold
model, the concentration-dependent model is used to
derive Effective Concentrations (ECs). (If the endpoint
being tested is mortality, EC is called LC or Lethal
Concentration). By definition, any EC or LC value is an
expression of some amount of adverse effect. For
example, EC50 is the effluent concentration that would
affect 50 percent of the organisms tested. Interpretation
of EC values, therefore, requires the judgment of a
biologist. By calculating both NOEC/LOEC and EC values
11
-------
''.' • ' ••*," '•*'..'•.' ' '- ' •
— Convective Descent' ^
•'• :.'•'. '.••.;•'! :.-.• Enco
•• . V:!" .""..', ' Neutr
unter
al
ancy
Diffusive Spreading
Greater Than
Dynamic Spreading
-•• ' •-•...'•";• ». , ??.- "**•• •-*" — •-••- — -'^*' ° * * — """"' '"•• "*"
Figure 2-6. Effluent discharge plume from an oil and gas drilling operation in a marine/estuarinc environment (14).
from the same results of each test, biologists can provide
regulators with sound, reproducible data and a solid basis
for informed decisionmaking.
Biologists should use caution in interpreting test results in
cases where the'site-specific effects of pH, salinity,
temperature, and other natural conditions may affect the
bioavailability of the toxicants in question. For instance
the bioavailability of free ammonia is largely dependent on
the pH of the ambient water, and thus pH can affect the
apparently "safe" or "no effect" concentration of the
effluent, and skew test results. An example of interactive
effects between an effluent and a receiving water was
encountered during the evaluation of the toxicity of the
effluent discharged from a pulp and paper mill in
Fernandina Beach, Florida (see Appendix A).
Field Verification of Test Methods
Validation of short-term toxicity tests is a necessary first
step before they are incorporated into the environmental
regulatory process. ERL-N validated these new methods
first with extensive field testing. This was accomplished
by comparing effluent toxicity to receiving water toxicity at
comparable effluent concentrations. Scientists conducted
11 such field tests at a variety of sites with an array of
discharges from chemical plants, sewage treatment
plants, shipyards, and an aircraft refurbishing facility. The
results of these tests indicated a good correlation
between the toxicity of receiving waters and comparable
effluent concentrations in the laboratory. Tests that were
most sensitive to effluent concentrations in the laboratory
were correspondingly most sensitive to the appropriate
receiving waters in almost all cases. Case studies
illustrating this correlation are provided in Appendix A.
Test Precision
Precision tests were performed within a single laboratory
(intralab, ERL-N) to analyze the variability introduced by
the methods and by individual species. Precision
(closeness of repeated measures) is a measure of test
validity. Accuracy (closeness to a standard) for toxicity
tests, however, cannot be determined. The results of
precision tests were compared to the precision results of
similar toxicity tests for freshwater organisms; the
precision of the marine tests were found to fall within the
commonly accepted range (15,16).
Researchers calculated EC50 values from each toxicity
test that regulators would use for setting NPDES permits.
EC50 values are more useful for precision comparisons
than NOECs because they are not expressed only with
reference to the effluent concentrations in the test; rather,
they are derived from a continuous, concentration-
dependent response curve. ECSOs, therefore, are
discrete estimates of toxicity and can be expressed as a
single number for each replicate of a given toxicity test.
Morrison et al. (17) examined EC values for five replicates
of each short-term chronic test and statistically compared
them for variation. Precision test results indicated that
while some tests are more variable than others, the
overall precision of the tests is within the commonly
12
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I
accepted range for similar toxicity tests with freshwater
species. The red algal (Champia parvula) reproduction
test and the sea urchin (Arbacia punctulata) sperm cell
test gave the most variable (yet acceptable) results, while
the sheepshead minnow (Cyprinodon variegatus) results
were most reproducible. For reliable results, EPA
recommends that toxicologists use a representative of the
fish, invertebrate, and plant tests when testing for toxicity.
i The exact testing requirements for a discharger, however,
are specified on a case-by- case basis in NPDES permits.
An intralaboratory precision test of the mysid, drilling fluid
acute toxicity test was conducted at the ERL-GB (18).
The researchers found that the sources of variability
included the condition of the sample itself, the manner in
which the SPP is prepared, the condition of the test
animals, and the skill and experience of the people
conducting the test. They also found that the
interlaboratory variability of the test conducted at 10
commercial laboratories was comparable to the variability
. of interlaboratory acute tests conducted with single
chemicals.
Both single-lab and multilab precision tests were
completed during the validation of the marine and
;freshwater acute toxicity tests (2). Similarly, an interlab
precision test of the sheepshead minnow short-term
chronic toxicity test has been completed from a total of
seven tests by four participating laboratories. The labs
reported very similar results indicating good interlab
precision. Similar "round robin" precision tests of the
remaining marine CETTP tests will be completed by
spring, 1990.
Relative Sensitivity of the CETTP Tests
In adapting the CETTP from the freshwater to the
, marine/estuarine environment, scientists considered the
greater diversity of plant and animal life in
marine/estuarine waters. In an attempt to account for the
extreme natural variation in biological communities, as
well as constantly changing physical and chemical
environments, marine/estuarine toxicity tests should
encompass a wide variety of species to be representative
of the environment. The complex effluents discharged
can react and change and, therefore, are almost
. impossible to quantify or regulate through analytical
chemistry alone. To take into account all of these factors,
,the marine CETTP uses biological toxicity tests with a
variety of species types and test endpoints.
The relative sensitivity of the tests is difficult to assess
because different combinations of conditions may affect
each species differently. All test results, however, can be
used to calculate NOEC and EC values that are used
directly to set NPDES effluent permit limits. While species
and endpoints differ, NOEC and EC values are directly
Comparable, indicating the relative sensitivity of test
species. No individual test has proven consistently the
most sensitive. NPDES permit limits, however, should be
dictated by the results of the most sensitive species
tested, i.e., the lowest NOEC from all the tests
conducted. Therefore, the more species tested, the more
protective the test results will be for the receiving waters.
The sea urchin sperm cell test is often the most sensitive
and can be conducted most rapidly. Therefore, regulators
are encouraged to use this test for rapid screening in
conjunction with two other tests.
Incorporating Test Results into NPDES
Permitting
Delegated States have the responsibility for issuing
NPDES effluent discharge permits. The States are
provided considerable leeway in their use of Federally
approved test methods for this purpose. The "Permit
Writer's Guide for Water Quality-based Permitting for
Toxic Pollutants" (19) supports EPA's shift in emphasis
from the solely chemical-specific approach to one
integrating both the chemical-specffic and whole effluent
techniques as a basis for developing NPDES permits.
Presently, regulators developing NPDES permits can refer
to the permit writers guide, "Methods for Measuring the
Acute Toxicity of Effluents to Freshwater and Marine
Organisms" (7), and "Short-term Methods for Estimating
Chronic Toxicity of Effluents and Receiving Waters to
Marine and Estuarine Organisms" (8). EPA strongly
recommends that States include both acute and chronic
toxicity tests as bases for NPDES permits.
Inherent in the design of many water quality-based
discharge permits is the recognition of the mixing that
occurs after the effluent is released. The "design flow" of
effluents into freshwater is very well understood and can
be described with a variety of computer models (2). While
the mixing of effluents in marine/estuarine waters is not
nearly as well understood, some models are discussed in
"Initial Mixing Characteristics of Municipal Ocean
Discharges" (20). EPA recommends using "mixing
zones" for water quality-based NPDES permits, if
appropriate. Acute toxicity criteria are applied within the
estimated mixing zone while chronic toxicity criteria are
applied at the edge of this zone with consideration of
critical mixing conditions. Treatment systems should be
designed to meet the more stringent requirements (acute
or chronic) in each site-specific case. While toxicity tests
are water-quality based, the permits are based on
allowable allocations of priority pollutants and effluent
volumes, both instantaneous and averaged over various
time periods. A complete description of these permitting
techniques is provided in the "Permit Writer's Guide to
Water Quality-based Permitting for Toxic Pollutants" (19),
and case studies of their application in marine/estuarine
systems are provided in Appendix A.
13
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3. In Situ Biomonitoring
The in situ biomonitoring program at ERL-N includes
toxicity tests with aquatic plants and animals as well as
biological and chemical tests of bioaccumulation. During
the design of the program, it was recognized that
measurements of water chemistry alone (without a direct
biological comparison) may only allow inferences from
' chemical concentrations observed in the field to actual
. biological effects. ERL-N is, therefore, developing
techniques to assess biological effects and chemical
uptake simultaneously. The goal is to provide regulators
with a suite of biological and chemical tests that can be
integrated for an effective assessment of the receiving
water effects of complex effluents discharged into
marine/estuarine environments.
Many biological methods for toxicity testing are
comparable to chemical methods in terms of time and
resources. A recommended strategy for toxicity
evaluations, therefore, is to conduct simple biological
assessments first, and then, if necessary, conduct more
: complex biological and chemical tests for a complete
assessment. Finally, the monitoring data must be
; effectively incorporated into the regulatory process.
Chemical-specific assessment should additionally be used
when pollutant criteria are available. Most traditional
biomonitoring efforts have been implemented as follow-
ups to environmental enforcement measures, and thus
the data collected are frequently not used to their
potential. A key goal of the in situ biomonitoring program
is to link scientists and regulators during the regulatory
effort. This will ensure that data are gathered in concert
with the needs of the regulators and are available soon
enough for inclusion in the decisionmaking process.
Historical Perspective
Narragansett Bay, Rhode Island, has suffered from the
coastal dumping of industrial pollutants longer than any
other area in the United States. Contamination first began
in 1793 when the Slater Saw Mill was built upstream of
jthe upper bay. Over the past two centuries, as the
Narragansett Bay area became more and more
industrialized, textile mills, machine tool activities, jewelry
and plating operations, and human waste disposal
systems all contributed effluents to the bay.
This long history of effluent discharge, combined with the
continual tidal flushing and riverine currents of the area,
have created a contamination gradient throughout the
bay. The upper reaches of Narragansett Bay (especially
sediments) have been heavily impacted by pollutants
while the lower portions of the bay are relatively clean.
This situation has established a relatively stable
contaminant gradient which has provided an ideal location
for field testing marine/estuarine biomonitoring
measurements.
Initial bioaccumulation studies first showed the link
between contaminant concentrations in mussel tissues
and proximity to industrial discharges. For regulatory
purposes, however, stronger inferences had to be made
directly linking the observed mussel tissue contaminant
residues with actual biological toxicity. In the early 1970s,
the Coastal Environmental Assessment Station (CEAS)
program was developed at ERL-N as a multipurpose
program to assess the relative environmental health of
marine coastal waters. The CEAS program used the blue
mussel Mytilus edulis as a biological indicator species.
The blue mussel Mytilis edulis has been selected for ERL-N's in
situ biomonitoring program (21).
The mussels M. edulis and M. californicus are indigenous
to and abundant in bays and estuaries around much of
the coastline of the United States. Mussels feed by
filtering particulates out of surrounding seawater, so they
are excellent indicators of ambient water quality. In
addition, mussels feed continuously throughout the year
(without a dormant period), making them a particularly
convenient species for physiological study. Finally,
mussels are excellent for transplanting because
large.healthy populations of animals can be found and
distributed easily into field cages at various sampling
sites. Another species, the hard clam Mercenaria sp.,
was explored as a test species but found not to be
amenable to year-round in situ testing.
15
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While the mussel has many advantages, its primary
disadvantage for short-term toxicity testing is its
reproductive habits. Species that reproduce frequently
provide results that are useful for extrapolation to
population level effects. Mussels spawn sporadically and
infrequently and, therefore, are not useful for assessing
effects on reproduction. Other species are presently
under consideration as future test organisms to directly
predict the population effects of toxic marine discharges
through reproductive tests.
The CEAS program is an intensive effort to relate
biological effects in mussels to the concentrations of
contaminants observed in their tissues. An initial step in
the program involved transplanting mussels on a transect
along a pollution gradient in Narragansett Bay (Figure 3-
1). Physiological measurements on transplanted mussels
indicated a correlation between biological effect and
contaminant concentration along the Narragansett Bay
Transect. Subsequent testing with a variety of biological^
tests supports these initial findings, thus providing a
strong basis for use of Narragansett Bay as a "field
laboratory" for testing various biomonitoring
methodologies.
Narragansett Bay
Rhode Island
Providence
River
Greenwich
Bay
Figure 3-1. Narragansett Bay, Rhode Island. Transect stations
1-4 (22).
Currently, the in situ biomonitoring program at ERL-N is
evaluating biological effects at various levels of biological
organization. Recent results obtained from extensive field
tests indicate that the biomonitoring program may provide
valuable techniques to use as regulatory tools in the near
future by predicting contaminant-induced perturbations in
marine environments at the population level.
Biological Methods
A wide variety of biological methods have been tested at
ERL-N for potential use in the biomonitoring program. The
methods vary in complexity, cost, stage of development,
and the level of biological organization tested (Table 3-1).
Of all the biological methods explored, the scope for
growth (SFG) index in M, edulis was found to be one of
the most effective methods evaluated and now forms the
primary component of ERL-N's in situ biomonitoring
program. SFG (a measure of the energy budget of an
animal) has been used to measure the physiological
condition of mussels transplanted to polluted areas
compared to those transplanted to "cleaner" conditions
(23). Similarly, SFG measurements on mussels exposed
to contaminated sediments in a lab exposure system have
been compared favorably to those of mussels exposed to
control conditions (see Section on Validation of Test
Methods with Case Studies). Additionally, shell growth
measurements on mussels can be used to estimate
biological effects of contamination. After either field or lab
exposures, chemical analyses of a subset of mussels are
examined for correlation with biological effects.
Scope for Growth Procedures
Calculation of the SFG index for M. edulis requires the
measurement of three parameters: clearance rate,
assimilation efficiency, and respiration rate.
Measurements are completed under standardized
conditions that match exposure conditions as closely as
possible, with respect to temperature and salinity. An
algal food concentration of 0.5 mg/L is used because it
provides sufficient energy for the organism during the
tests. All SFG measurements are completed within 24
hours of collection from field or lab exposures to ensure
that results are indicative of experimental conditions.A
detailed description of these methods is provided by
Nelson et al. (24); however, a brief summary of each
procedure is provided here.
Clearance rate (CR). Mussels are placed into individual
chambers through which filtered seawater, with a set
concentration of algae, flows at a constant rate. After a 1 -
hour acclimation period, CR is determined at three hourly
intervals for each mussel by measuring incoming and
outgoing algal concentrations with an electronic particle
counter.
Respiration rate. Respiration rates are determined by
isolating each mussel in a glass respirometer vessel fitted
with an oxygen electrode. The electrode is connected to
16
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Table 3-1. Methods in Use and Under Development for In
Situ E
vel of
Biomonitoring (8)
Method
Level of
Complexity
Lev
Bioogical
Organization
Tested
Description of Method
Methods Discarded
Adenylate Energy Exchange (AEC) High
Blood chemistry Medium
Methods Requiring Refinement
Biomarkers including metallothionein High
induction, sister chromatid
exchange, immunosuppression
Histopathology Low
Tissue microbiological contamination Medium
Demographic analysis High;
Quantitative benthic sampling High
Methods that Can Be Used
Immediately
Growth, survival, and fecundity tests Low
Scope for growth (SFG) Medium
Gill respirometry Medium
O
an oxygen meter which, in turn, is connected to a strip
chart recorder. The decline in dissolved oxygen is
monitored on the strip chart recorder for approximately 30
minutes. Seawater containing algae (0.5 mg/L) is pumped
into the vessel during an acclimation period at a rate of .
80 mL/min to ensure that food is present in the chamber
and that routine metabolic rate is measured.
Assimilation efficiency. Mussels are allowed to feed
overnight in the individual chambers used to measure
clearance rate. The fecal material is removed from each
Cellular The amount of energy available to an organism from the
pool of adenine nucleotides (ATP, ADP, and AMP) can
sometimes be affected by stress. AEC is not
significantly affected by contaminant concentration and
has been discarded.
Cellular Changes in the relative abundance of several normal
blood components might have indicated pollution effects.
Ce lular There exist several potential biomarkers or sublethal
biochemical indices which might be used as early
biological sentinels of pollution effects.
Organ Visual examination of sensitive tissues such as gills, liver,
and reproductive organs may indicate structural tissue
damage in response to contaminant exposure.
Orgarj, system Various bacteria might be used as tags for contaminated
sediments.
Population Population dynamics can be inferred from the computer
compilation of individual responses to contamination
along a gradient.
Population Community structure and contaminant effects can be
determined by estimating benthic species diversity and
relative abundance.
Ind vidual See Section 2 of this document.
Ind vidual The energy left for growth and reproduction after routine
metabolic costs. SFG decreases with increased
contaminant exposure. (See full explanation in this
section.)
•gan Health of animals determined by the respiration rate of
excised gill tissues. (Produces results which are
redundant to SFG so the test is not used.)
chamber the following morning. Fecal material is dried,
weighed, ashed, and reweighed to determine the ash-free
dry weight to dry weight ratio. A similar procedure is
completed with the cultured algae to obtain the ash-free
dry weight to dry weight ratio of the food. Assimilation
efficiencies are calculated for each treatment (25).
Scope for Growth Calculations
After completion of the physiological measurements, the
length and volume of each mussel is measured and the
tissue excised, dried, and weighed. The clearance rates
17
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and respiration rates are standardized to the mean weight
of all the mussels used in the treatment.
The weight-standardized values for each mussel are then
used to calculate the SFG of each individual by
substitution into the following equation:
Scope for Growth = (C x A) - (R)
where
C = energy consumed (clearance rate x
surrounding food concentration x energy of
food).
A = assimilation efficiency.
R = energy lost through respiration.
Shell Growth
To determine if shell growth is correlated with exposure to
toxic contaminants, measurements of mussel growth are
taken concurrent with field and lab exposures. The shell
length of 15 individually marked mussels per cage is
recorded before and after field or lab exposure. Growth is
calculated as the change in shell length per unit time.
Field Exposure
Prior to a field deployment, mussels are collected with a
scallop dredge 1 to 2 days before introduction into test
waters. The collection site should be similar to the site of
deployment with respect to temperature (<4°C) and
salinity (< 5 parts per 1000) and exhibit good water
quality. The mussels are sorted to obtain a narrow size
range and are distributed into polyethylene field cages
(Figure 3-2). The cages are then transplanted to field
sites along a gradient of contaminant concentrations, as
shown for the Narragansett Bay sampling stations 1
through 4 (Figure 3-1). This technique of transplanting
along a transect is the basis of in situ biomonitoring at
ERL-N. After a specified deployment period, usually 7 to
30 days, the cages are retrieved and the mussels
returned to the lab. A subset of mussels from each cage
is used for biological testing, while the remainder of the
mussels are frozen for chemical analysis.
Lab Exposure System
In addition to in situ biomonitoring, a laboratory exposure
system was developed for the mussel which will provide
further information for evaluating NPDES permits. Initial
exposures with the system indicate that the sensitivity of
the mussel to copper, a standard contaminant, is
comparable to other species currently used in the
CETTP.
Chemical Methods
The two methods used most commonly for determining
the concentrations of specific pollutants in animal tissues
Surface
Plastic Float
Polypropylene
Line
Mussel
Baskets
Anchor
(Sub-Surface)
Approx.
1 m
Figure 3-2.
CEAS Station
Mussel field cages used to transplant mussels
along transects (22).
are atomic absorption spectrophotometry (AA) and gas
chromatography/mass spectrophotometry (GC/MS). AA is
used to measure trace metal concentrations, while
GC/MS, together with electron capture gas
chromatography, is used to determine the concentrations
of volatile and semi-volatile organic chemicals like
polychlorinated biphenyls (PCBs) and^olycyclic aromatic
hydrocarbons (PAHs) (26).
Validation of Test Methods with Case Studies
The sensitivity testing and validation of response
parameters used within the in situ biomonitoring program
have been completed through a series of case studies.
One such study to field verify these methods was
conducted in conjunction with the U.S. Army Corps of
Engineers' Field Verification Program at Black Rock
Harbor (BRH), Bridgeport, Connecticut. The approach
taken was to evaluate the relationship between
contaminant exposure, mussel tissue residue
concentration, and subsequent biological effect. Mussels
were exposed to various concentrations of contaminated
sediments in the laboratory, followed by measurements of
the mussels' contaminant tissue residues, SFG, and shell
growth. Growth measurements as well as SFG were
highly correlated with the concentration of contaminated
BRH sediments to which test organisms were subjected
(Figure 3-3).
A second study was initiated to determine the exposure
duration period required to assess toxicity in the field. The
SFG results of 1-week exposures were compared to SFG
results of 1-month exposures along a contamination
gradient in East Greenwich Cove (see Appendix A).
Figure 3-4 shows that SFG measurements decreased
18
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5-
§ 4-
-------
with increasing proximity to the outfall. While 30-day
exposures induced more pronounced SFG declines than
1-week exposures, the short exposure period produced
significant and usable results, illustrating the utility of 7-
day exposures. This study also showed that both
clearance rate (CR) and respiration rate (RR) were
significantly affected by exposure to toxics.
One might assume from Figure 3-3 that SFG results are
redundant with simple growth measurements; however,
the two endpoints can reflect different effects. SFG is a
measure of all the energy available for growth, both
somatic (including tissue and shell) and reproduction.
Therefore, increases in shell growth reflect only a portion
of the energy available to the organism. During certain
periods of the year, most of the organism's energy may
be allocated to reproduction and very little to shell growth.
However, even during periods of slow shell growth, the
organism must still maintain the physiological functions
necessary for survival. Because of this, the SFG index is
more reflective of the overall condition of an organism at
any time of the year. In addition, the findings in East
Greenwich indicate that both mussel CR and RR can be
significantly affected by toxic exposure. Therefore, the
SFG measurements illustrate not only the extent of the
effect, but may also provide an indication of the biological
mechanism behind the observed effect.
Comparisons among Various Levels of
Biological Organization
Currently, research is being focused on developing tests
for assessing impacts at many levels of biological
organization to ensure that indicated toxicity is of real
ecological significance. As a goal, tests at any level of
biological organization (i.e., cellular, tissue, individual, or
population) should predict the community/population
effects of a discharge. The SFG index on individual
mussels is indirectly indicative of population effects of
toxic pollutants. Contaminant exposure has been shown
to decrease SFG, thus lowering energy available for
somatic and reproductive growth. This lowered energy
may impact reproductive output and cause an effect at
the population level.
The more scientists know about the mechanisms behind
the observed effects, the better they will be able to
interpret individual, tissue, and cellular responses and
predict population responses. For example, showing that
a toxicant can inhibit an organism's ability to reproduce is
a way to predict the population effects of that toxicant.
Future research will concentrate on developing test
methods focusing on all levels of biological organization to
gain a comprehensive understanding of marine
ecosystems and thus better understand contaminant-
induced perturbations within these environments. Tests at
the subcellular level may produce results, but those
results may be difficult to interpret and extrapolate to
community level effects. Changes at the community level
are difficult to detect because of the often extreme natural
variations that can occur in marine ecosystems. Even if
detected, community level changes are extremely difficult
to interpret, especially for regulatory purposes. For
instance, if a change is observed in community structure
or species diversity, it may only reflect natural variations
within a complex and changing ecosystem, and not
necessarily indicate contaminant effects. A
comprehensive understanding of the biology of marine
ecosystems will allow better predictions of contaminant-
induced perturbations within these environments, thus
providing regulators with a more solid base from which to
make informed decisions concerning marine discharges.
The Integration of Biomonitoring into Marine
Discharge Regulations
The implementation of marine toxicity evaluations requires
the combined expertise of both marine scientists and
regulators. Past monitoring efforts have been conducted
without a formal means for regulators to use the
monitoring data generated, so the data have rarely been
effectively incorporated into decisionmaking. Because of
the lack of a framework within which to interact with.
regulators during the implementation of toxicity
evaluations, a conceptual framework for regulator/scientist
interaction was developed. The framework follows EPA's
principle of selecting the simplest reliable indicator of
biological impact. The framework provides two
opportunities for regulator/scientist interaction. In the first
phase, regulators pose questions that scientists and
regulators together design experiments to answer. They
can select tests from the array of methods in Table 3-2.
The sensitivity of the test chosen should be in
accordance with the hierarchy of relative sensitivity, cost,
complexity, and environmental contamination. Rapid and
inexpensive biological tests with a low rate of false
negative results should be used for initial site
assessment, and more sophisticated biological, chemical,
and physical tests should be used when indicated by
toxicity.
When test methods have been selected and monitoring is
in progress, the second phase of regulator/scientist
interaction is invoked. A set of criteria are developed by
the same team of regulators and scientists who designed
the initial monitoring strategy. The decision criteria
provide the team with the means to interpret the data
being gathered, reevaluate the techniques and strategies
employed in monitoring, and formulate decisions
regarding changes and additions to the monitoring effort.
The flowchart in Figure 3-5 depicts steps that might be
followed during the risk assessment of a marine/estuarine
discharge. The decisions made at each juncture are
guided by the decision criteria developed by the
regulator/scientist team at the studies initiation. Using
decision criteria to ensure that proper data are gathered
within appropriate time frames, scientists and regulators
can beneficially interact, streamlining risk assessment and
20
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Table 3-2. Classification of Biological, Chemical, and Physi
the Proposed Monitoring Strategy (28)
Level 1
:al Analyses into Four Categories that Are Subsequently Used in
Level
I 2
Level 3
Level 4
Analysis
Survival
Growth
Fecundity
Chemistry - inorganic
water analysis
Technical sophistication Low
Time required to obtain results Hours to days
Scope for growth
Gill respirometry
Chemistry j- organic water
analysis
Intermedia e-high
Days to weeks
Population dynamics
Chemistry - tissue and
inorganic analyses
High
Weeks/Months
Chemistry - Tissue and
sediment analysis
Physical
Bathymetry
REMOTS interface
Side scan sonar
High
Weeks/Months
Archive Tissue Samples
for Chemical |Analyses
at Low Priority
Area Receives Low
Priority as ar
Environ-
mental Problem
Area Receives Medium
Priority as an Environ-
mental Problem
Implement Pollution Abatement Measures
Figure 3-5. Flow chart depicting the proposed use of decision criteria in risk assessment and coastal monitoring (28)
21
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drastically reducing the cost of marine/estuarine
biomonitoring. Decision criteria were effectively used
during the toxicity evaluation of New Bedford Harbor,
Massachusetts, described in Section 5.
Current Research
Ongoing research is currently focused on evaluating and
developing several new techniques for the in situ
biomonitoring program. Of particular importance is the
field verification of lab results with mussels. If field
verification is successful, scientists will be able to conduct
reliable, rapid, and inexpensive lab studies of mussel
exposure. The methods may then be suitable for
incorporation into the appropriate NPDES permits. Future
research will focus on:
• Tests with other species from other geographic
locations - to increase the range and species
representation of the tests.
• Tests with other levels of biological organization - to
further demonstrate the relationship between
different levels and thus provide insight into
population and community level effects.
Toward these two research goals, ERL-N is currently
evaluating another species, the slipper shell snail
Crepidula fornicata, for in situ transplanting. C. fornicata
are found on the east coast from Maine to the Caribbean
and on the west coast from California to Baha. Like
mussels, these filter feeders are easily transplanted and
convenient for physiological studies. Unlike mussels, C.
fornicata reproduce frequently and prolifically, which may
be extremely useful for reproductive studies. The results
of tests with this species may be useful for extrapolating
to population level effects of toxic marine discharges.
The slipper shell snail Crepidula fornicata is under
consideration as another species for in situ biomonitoring (21).
Summary
Upon completion of laboratory-to-field comparisons of
these tests, short-term laboratory tests will be
complemented by in situ assessment of receiving water
conditions. In contrast to extrapolation of laboratory
estimates of toxicity, in situ tests will measure the actual
physiological effects of field exposure to a variety of
point-source discharges with mussels, thus providing an
important regulatory tool for marine toxicity evaluations.
22
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4. Integration of Effluent and In Situ Biomonitoring
Programs
Complex Effluent Toxicity
To develop NPDES permits for the control of toxic
effluent discharges to marine/estuarine environments,
regulators require a quick, reliable, and inexpensive
procedure to estimate effluent toxicity. To accomplish this
requires an estimate of both the effluent's absolute and
relative toxicity. Absolute toxicity is defined as the
laboratory-determined toxic concentration of the complex
effluent, while relative toxicity is defined as the toxicity of
the effluent after mixing with the receiving water.
Absolute toxicity is estimated with short-term CETTP
tests. Relative toxicity can be estimated in two practical
ways. First, relative toxicity can be estimated by sampling
the receiving water at various sites and testing the
samples with the same CETTP tests used to assess
absolute toxicity. Alternatively, relative toxicity can be
assessed with in situ biomonitoring by determining the
uptake of various toxicants and the corresponding
biological effects on organisms that are transplanted into
the receiving water.
Receiving Water Toxicity Estimates
Use of Dye Studies to Estimate Effluent
Concentrations in Receiving Waters
While laboratory toxicity tests on various concentrations
of effluent provide an estimate of "end-of-pipe" toxicity of
a discharge, for regulatory purposes it is essential that the
toxic concentrations of complex effluents determined in
the laboratory correlate with the actual effluent
concentrations observed in the field. The concentration of
effluent in receiving waters at field sites, however, is
difficult to predict due to the complex mixing and
dispersion of the effluent after it leaves the pipe.
To estimate actual receiving water concentrations of
effluents, dye studies are performed. Prior to sampling
any receiving water, the effluent is saturated with
rhodamine WT dye by continually adding the dye to the
effluent for 3 days or until an initial steady state is
reached. Subsequently, when sampling occurs at various
receiving water sites, the dye should be in equilibrium
with the effluent. The concentration of dye in the effluent
can be measured with fluorometry (fluorescence), and
used along with the dye infusion rate and the effluent
discharge rate to calculate the concentration of the
effluent at each sampling site. In relatively stable
environments, dye studies can be a powerful tool for
comparing known, effluent concentrations in the lab with
effluent concentrations at various field sites. In certain
site-specific conditions, however, dye studies can be
drastically altered by local environmental conditions. For
example, biologists should use caution when interpreting
the results of a dye study conducted in shallow or wind-
driven estuaries.
Researcher collecting a water sample during a dye study.
CETTP Tests of Receiving Waters (Relative Toxicity
Estimates)
Once dye studies have been conducted and the effluent
concentrations at field sites have been estimated, short-
term chronic toxicity tests are conducted with receiving
waters. Sampling sites vary in their proximity to the
discharge site. Toxicity tests on receiving waters are
conducted with the same set of test organisms used with
complex effluent samples alone, and the results are again
used to estimate the toxic concentration of the effluent.
ERL-N found that the toxic concentration of the effluent
measured in the lab is almost always the same as the
toxic concentration of effluent in parallel receiving waters.
The high correlation between lab and field results enables
regulators to incorporate these tests into NPDES permits
with a high degree of confidence. While the tests are
attractive because they can be conducted rapidly, in situ
biomonitoring might provide supplemental information on
23
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longer-term effects. One of the longer-term in situ chronic
toxicity tests is described in the following paragraphs.
Use of In Situ Biomonitoring with Mussels to
Complement CETTP Tests
The efficacy of marine/estuarine complex effluent tests in
rapidly estimating an effluent's toxicity makes them an
important regulatory tool. However, there is concern that
the very speed of the short-term tests may compromise
their long-term accuracy. This becomes increasingly
important during implementation because Regions and
States exhibit a wide latitude in their monitoring
requirements. For example, if all samples are collected
during 1 intensive week, but the effluent discharge
continues throughout the year, this "grab" sample
technique may not account for:
• Temperature and seasonal effects, and
noncontinuous discharge rates from effluent tanks or
reservoirs.
• Differences in effluent components over time.
To ensure that short-term toxicity evaluations are
consistent with long-term trends, it is recommended that
long-term in situ biomonitoring supplement the current
CETTP tests in certain site-specific discharge situations.
Through periodic deployments throughout the year,
mussels can be used to monitor long-term trends
(Section 3). In situ toxicity tests with mussels may
eventually complement CETTP tests, either by monitoring
a specific discharge or as a general monitor of receiving
water health.
Integrated Approach for Toxic Discharge
Control
The integrated water quality-based approach of using in
situ monitoring with existing CETTP methods for toxic
discharge control may be a powerful tool for regulators.
This approach will determine ambient water quality by
measuring SFG on mussels and thereby assess the
toxicity of an effluent (Section 3). The first step in this
approach will be to compare SFG response of mussels
from the laboratory exposure system with field-exposed
mussels, over 7 days. A second step will include a
comparison of the sensitivity of 7-day lab exposure with
mussels to the sensitivity of short-term CETTP tests.
Point-Source Evaluations and Large-Scale Coastal
Monitoring
The integrated water quality-based approach, combining
toxicity testing of effluents and receiving waters with
ambient toxicity testing, could be further incorporated into
large-scale coastal monitoring systems. The National
Oceanic and Atmospheric Administration's (NOAA)
Mussel Watch Program already uses in situ biomonitoring,
with mussels as regional indicators of marine ecosystem
health. This monitoring program periodically collects
mussels at about 150 coastal and estuarine sampling
sites, and measures mussel tissue residues of tracer
contaminants including PCBs, PAHs, and several heavy
metals (29). High tissue residues might indicate potential
toxicity problems within a region. In this way, Mussel
Watch can identify problem areas that require additional
testing to pinpoint discharge sites or upstream runoff
problems. EPA's new Environmental Monitoring and
Assessment Program (EMAP) will be especially important
in addressing ecological status and trends. EMAP will be
phased into implementation in 1992 and will integrate the
existing ecological monitoring programs concerning the
Nation's forest, wetland, near-coastal, inland surface
water, and agricultural areas. The near-coastal
component of EMAP might be an extremely useful tool for
integrating the results of localized marine toxicity
evaluations and large-scale marine monitoring efforts.
Combining the integrated approach with large-scale near-
coastal monitoring programs may allow the tracking of
pollutants from "end-of-pipe" through receiving waters to
estuaries and open bays. This integration has the
potential to vastly increase the ability to assess the
results of existing pollution problems. Baseline data of
this kind also will help to assess the field effectiveness of
remediation efforts.
Hierarchical Framework for Biomonitoring Tests
To further increase the efficiency of marine/estuarine
biomonitoring, it is proposed that all biomonitoring
techniques be placed into a hierarchical framework of test
complexity. Each test will, therefore, be incorporated into
toxicity evaluation, according to the severity of the
problem. Section 3 describes the hierarchy of existing
tests, and the framework within which researchers and
managers can interact to best use their limited resources
to select from the array of tests and control pollution of
our Nation's estuaries and oceans.
The most effective approach to controlling toxic
discharges is for scientists and regulators to work
together closely to design appropriate control strategies.
Biological tests that both rapidly estimate toxicity and
monitor the long-term trends of point-source discharges
can be used in a framework within which managers and
regulators can evaluate the data and make decisions
rapidly. This will ensure that scientists provide the
answers most needed by regulators, in a timely fashion.
In addition, a tandem approach for localized or point-
source toxic control can be integrated into large-scale
coastal monitoring programs. The integration of these
monitoring programs could provide a dynamic picture of
the local effects of toxic discharges and subsequent
remediation measures, as well as the larger-scale effects
in near-coastal waters.
24
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5. Application of the Biomonitoring
Bedford Harbor Pilot Dredging
Introduction
New Bedford Harbor is located between the city of New
Bedford and the town of Fairhaven in Buzzards Bay,
Massachusetts (see Figure 5-1). Since the 1940s,
electronics and manufacturing companies in the area
have discharged PCB-laden effluent into the Acushnet
River and the harbor. High PCB concentrations in the
river and upper harbor sediments were first documented
in 1974 (30). Over the past 15 years, nearly 18,000 acres
of PCB- and heavy metals-contaminated sediment have
been described, with PCB concentrations as high as
1,000 parts per million (ppm) in some areas of the harbor
(31). In 1982, the site was added to EPA's National
Priorities List of hazardous waste sites slated for cleanup
under the Superfund law.
EPA's 1984 feasibility study proposed several alternatives
for the harbor cleanup. One of the alternatives included
dredging contaminated sediments and disposing of them
in either a land-based, confined disposal facility (CDF), or
an aquatic, confined aquatic disposal (CAD) area.
Federal, State, and local officials, as well as the public,
expressed concern over these alternatives. Many thought
that the sediments stirred up during the dredging would
affect the plants and animals that inhabit the harbor's
waters. Others cited potential pollution problems from
contaminated water (leachate) leaking from the proposed
disposal site (32). To answer these concerns, EPA
decided to pre-test the dredging and disposal options.
Working with the U.S. Army Corps of Engineers (COE),
EPA designed a pilot study to determine the best
remediation option for the Superfund site.
Dredging Project Description
The COE examined and compared the effects of three
hydraulic dredges and two disposal methods during this
project. The dredges were selected based on their ability
to remove sediment with the least amount of
resuspension and to operate in shallow water. The latter
was important because the dredging area was only 4 to 5
feet deep at high tide. The first dredge was a cutterhead
dredge; the second, a horizontal auger dredge called a
"mudcat"; and the third, a specially constructed dredge
called a "matchbox" (34). The two disposal methods
investigated were: 1) a confined disposal facility (CDF),
which consisted of a containment dike partially in water
Strategy: The New
Project
and partially on land; and 2) a confined aquatic disposal
cell (CAD), which was an in situ underwater disposal
method.
The COE dredged two areas in the pilot study cove
(Figure 5-2). Contaminated sediments from dredge area 1
were disposed of into the CDF (Figure 5-3). The
underlying "clean" sediment was also dredged to the
CDF to "cap" the contaminated material. The resulting
depression in area 1 became the CAD cell (Figure 5-4).
The dredged material from area 2 then was pumped into
the CAD cell, and a second "clean" layer of underlying
sediment from area 2 was used to cap the underwater
disposal area.
Monitoring Strategy
The pilot project was an opportunity to test the
biomonitoring research and management strategy
described in Section 3. Before the COE began any
operations, baseline physical, chemical, and biological
measurements were completed. The biological
measurements were used to assess the effects of
existing water quality on plant and animal survival, growth,
and reproduction. These tests served as a benchmark
against which increased contamination and/or toxicity
associated with the operational phases of the study were
compared.
The determination of whether the operation caused an
unacceptable effect was complicated by the fact that
State and Federal water quality standards for PCBs and
certain heavy metals were already exceeded under
preoperational baseline conditions. In addition, the U.S.
Food and Drug Administration (FDA) action level for PCBs
in seafood was exceeded before dredging started.
In order to assess the effects of the dredging project,
therefore, it was necessary to develop a set of site-
specific numerical values, called Decision Criteria (Table
5-1). These criteria were compiled in a Decision Criteria
Document (8), and were established for measured
physical, chemical, and biological parameters, based on
the preoperational data. During the dredging and disposal
operations, data on the same physical, chemical, and
biological parameters were collected and compared to the
Decision Criteria values. A Decision Criteria Committee,
25
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comprised of representatives from ERL-N, EPA Region I,
the COE, and the State of Massachusetts, evaluated the
data on a daily basis. If the Decision Criteria values were
exceeded, the committee could require engineering
corrections to the dredging operation before work was
resumed. In this way, the marine environment and its
resources were protected during the pilot study activities.
Water Quality Monitoring
Four stations were selected for water quality monitoring in
NBH, in addition to a reference station located in
Buzzards Bay (Figure 5-5). Station NBH-1 was located
north of the dredge site; Station NBH-7 was adjacent to
the cove where dredging occurred; Station NBH-2 was at
the Coggeshall St. Bridge, the transition point between
Wood St.
Canada
Upper Harbor
Pilot Study
Area
Coggeshall St.
Rte. 195
Lower Harbor
Atlantic Ocean
Cap Code
Buzzards Bay
New Bedford Harbor
/ Long Island Sound
• Butler Flats
Lighthouse
Upper Buzzards Bay
'3,000 ' 6,000 feet
Figure 5-1. New Bedford Harbor, in Buzzards Bay, Massachusetts (33).
26
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Aerial view of the New Bedford Harbor confined disposal facility
(CDF). The silt curtain surrounding the study area extends from
the far edge of the COF dike.
the more severely polluted upper harbor and the lower
harbor; and Station NBH-4 was at the hurricane barrier,
the transition point between NBH proper and Buzzards
Bay. The reference station for all water quality monitoring
was NBH-5, located at West Island in Buzzards Bay.
Mussels were deployed at stations NBH-2, NBH-3, NBH-
4, and NBH-5. Station NBH-6 was used as a mussel
reference station for the first deployment only;
subsequent deployments utilized NBH-5 to facilitate
mussel collections.
Seawater was collected separately for the ebb and flood
tide at each NBH station. This method allowed estimation
of net transport of PCBs and metals over time, especially
at NBH-2. Each water sample was analyzed for various
physical (total suspended solids) and chemical
constituents {PCBs, copper, cadmium, and lead). In
addition, a suite of biological tests were conducted to
assess acute effects (in fish, mysids, mussels, sea urchin
sperm cells, and a red alga) and chronic effects (fish
growth, mysid growth and reproduction, mussel scope for
growth, algal reproduction). A typical 7-day monitoring
cycle is presented in Figure 5-6.
Once the operational phases of the Pilot Project began,
water samples were collected in a manner similar to that
of the preoperational phase. However, during the
operational phase the ebb samples were returned to ERL-
N immediately and chemical analyses and acute biological
tests were completed overnight. These results were
transmitted to the Decision Criteria Committee prior to the
start of that day's dredging. The committee would then
assess any adverse environmental impacts and make any
necessary adjustments to the operation. This "real-time"
monitoring allowed the project managers to make timely
decisions based on actual data, thus ensuring the best
degree of environmental protection possible.
Summary of Results - Operational Phase
Physical measurements. The temperature and salinity
data were similar among the NBH stations during this
study. Within the harbor, temperature decreased slightly
with depth, and salinity increased minimally with depth,
reflecting input from the Acushnet River at the head of
the estuary. Total suspended sediment (TSS)
measurements at the sampling stations indicated that the
dredging operation resulted in little or no increase in TSS
concentrations over background. On one occasion an
increase in TSS concentration occurred at NBH-2. This
was caused by the opening of the silt screen around the
operation site which normally worked to contain
suspended particles within the study area. This resulted in
the discharge of a plume from the construction site. The
problem was quickly rectified and never recurred.
Water Chemistry. The PCB concentrations in NBH
exceeded the Decision Criteria values on only four
occasions (at NBH-2) during the operational phase of the
project. Three of these instances could be directly
attributed to a specific operational event. In all cases,
modification of the operation resulted in lowered PCB
concentrations in the water the following day. The only
other exceedence was caused by a meteorological event:
50 mph winds at low tide caused resuspension of
contaminated sediment from the upper harbor. At no time
did the concentrations of copper, cadmium, and lead
exceed the Decision Criteria values as a result of any
operational event.
Mussel Chemistry. Mussels deployed in NBH showed a
distinct spatial and temporal pattern with respect to PCB
uptake. The PCB tissue residues from mussels deployed
before dredging indicated that mussels located in the
upper harbor exhibited the highest PCB concentrations in
their tissues; the concentration decreased moving down
the harbor (Figure 5-7). This pattern was consistent
regardless of length of field exposure. The data indicated
that PCB tissue residue concentrations also increased
with length of exposure. These preoperational data
demonstrated the bioavailability of PCBs in NBH under
background conditions. Analysis of mussel tissues
exposed during operational phases of the project showed
the same pattern.as the preoperational exposures.
Therefore, no additional increase in PCB bioavailability
could be attributed to the dredging operation.
Biological Tests. The sea urchin (Arbacia punctulata)
sperm cell test and the growth and survival tests with the
sheepshead minnow (Cyprinodon variegatus) indicated no
toxicity at any of the stations. The preoperational red
algae (Champia parvula) reproductive tests proved
inconclusive because of problems with stock cultures.
Toxicity to this species was observed on several
occasions during operation. However, it was not
correlated with measured PCB concentrations and may
have been caused by elevated copper concentrations in
the water from a source other than the dredging
-------
operation, as the red macroalgae are particularly sensitive
to copper. Growth and reproduction in the mysid
(Mysidopsis bahia) indicated no effects as a result of the
dredging operation.
Mussel SFG and shell growth values from the first
preoperational deployment demonstrated a significant
inverse relationship with PCS tissue residue
concentrations in NBH (Figures 5-7 and 5-8). In addition,
SFG values on day 7 were reflective of results obtained
on day 28. This pattern was not evident during a second
deployment and was attributed to the fact that mussels
were at the peak of the gametogenic cycle at this time.
During gametogenesis, the mussels would expend extra
amounts of energy and the SFG measurement would be
affected. The SFG values of mussels collected during the
operational phase of the project never exceeded the
Decision Criteria values; therefore, no adverse impact
could be attributed to the operation;
Summary
The New Bedford Harbor Pilot Dredging Project was a
unique opportunity to use biomonitoring techniques,
ranging from the CETTP tests to in situ exposures, to
evaluate a potentially environmentally damaging dredging
operation on a "real-time" basis. A set of site-specific
criteria was developed and utilized by environmental
managers to assess, on a day-to-day basis, the effects of
this dredging operation on water quality in NBH. The
results indicated that the dredging operation had a
minimal effect on existing water quality. On those
occasions when elevated PCB concentrations were
detected, they could be directly attributed to a specific
operational procedure and modified before environmental
CAD
Top of Bank
EL.+ 6'
Dredge Area 2
CDF Discharge
MLW
Scale: 1" - 400'
Datum: Mean Low Water (MLW)
Figure 5-2. Dredging areas and disposal facilities at New Bedford Harbor pilot study area (35).
28
-------
Secondary Cell
90' f
Scale: 1" = 70' Horizontal
1" = 14' Vertical
Datum: Mean Low Water
CDF Discharge
Primary Cell
- 380'
EL. + 10'
Secondary
Settling
Cell
Primary Settling
Cell
/
Top of D
I
Figure 5-3. Cross section and top view of confined disposal facility (CDF), showing the two settling cells (35,34).
damage was done. The result was that no acute or
chronic biological effects were observed as a result of
this operation. While the approach to real-time monitoring
can be labor intensive, its application is recommended
when an operation, such as dredging a Superfund site,
can potentially have severe environmental impacts. The
Corps of Engineers is presently reviewing the data
gathered during the operation, and will decide on the best
dredging and disposal option for the more contaminated
portion of New Bedford Harbor
29
-------
-0.5'
-0.5'
Datum: Mean Low Water g$ Clean Sediment
Tide Range: 3.7' :->: PCB-Contaminated Sediment
Figure 5-4. Cross section of confined aquatic disposal
(CAD) cell, unfilled and filled (35).
Table 5-1. Numerical Decision Criteria for New Bedford Harbor (36)
Endpoint
Time to
Obtain
First
Value
New Bedford Harbor Station Location
Coggeshall St. Bridge
(NBH-2)
Hurricane Barrier
(NBH-4)
Water Chemistry
PCB (total)
Cd
Pb
Cu
Mussel PCB Tissue
Residues
7-day exposure
28-day exposure
Biological Responses
Acute (% Survival)
Fish
Mysids
Mussels
Sea Urchins (% fort)
Red Algae
Chronic Effects
Fish (dry wt, mg)
Mystds (dry wt, mg)
(reproduction)
Mussels:
scope for growth (J/h)
shell growth (mm)
Sea Urchin (% fert)
Red Algae (# cystocarps)
24 hr
24 hr
24 hr
24 hr
10 days
30 days
All tests
monitored
daily
8 days
8 days
8 days
28 days
1 day
8 days
Ebb (itg/L)
1.4
9.3"
7.2
13.0
Net Transport (kg)
0.71/cycle
80
160
Ebb
Mortality > 20% of control value for two
species or > 50% for one species
Ebb
0.44
9.3*
15.0
6.0
19.0
23.0
Ebb"*
20% (40%)
20% (40%)
50% (100%)
7.5 J/h (15 J/h)
50% (100%)
25% (50%)
50% (100%)
"U.S. EPA water quality criterion; criteria-continuous concentration.
This value represents a statistically and biologically significant reduction from control values. A significant reduction in two endpoints, or a
twofold reduction in one endpoint (in parentheses), requires that the decision criteria committee evaluater the monitoring dataand other
information related to the operation prior to the resumption of operations on the next day.
30
-------
Figure 5-5. New Bedford Harbor sampling station locations for
biological, physical, and chemical tests during
preoperational and operational monitoring (37).
Day 1 Day 2 Day 3 Day 4 Day 5 Day 6 Day 7
(x)
A
A
®
A
A
®
A
A
0
A
Key: x - Chemistry Samples
Suspended Solids
Sperm Cell Test
(x) - Same as Above With 24 Hour Data Turnaround.
fxl - Same as Above With Data Available on the First
— Day of the Next Operation.
A . Water Samples for Seven-Day Chronic Toxicity
*-* Tests.
A - Water Samples for Seven-Day Chronic Toxicity
—• Tests 'Plus Champia.
Figure 5-6. Sampling scheme for a generic 7-day sampling
period during the New Bedford Harbor pilot
Study (38)
Q.
m
100 -
80-
60-
40 -
20-
NBH-2
NBH-3 NBH-4
Station Location
NBH-6
Figure 5-7. PCB concentrations in mussel tissues over time
versus total PCB concentration as determined
from receiving water analysis (39).
31
-------
15
10
0 •
-10
_*Day7
*-* Day 28
NBH-2 NBH-3 NBH-4 NBH-6
Station Location
1.5
0.5
• Day 28
NBH-2 NBH-3 NBH-4
Station Location
NBH-5
Figure 5-8. Mussel scope for growth (SFG) in New Bedford
Harbor at 7 and 28 days, mussel shell growth at 28
days. Note that 7-day SFG tests are predictive of
the 28-day trends in SFG and shell growth (37).
32
-------
6. Conclusions and Future Directions
The Clean Water Act emphasizes the use of biological
monitoring as the basis for maintaining national water
quality standards. Inherent in the process of
environmental protection is the definition of the baseline
health of the ecosystem under scrutiny. Only when such
a baseline is established can trends be identified and
subsequently correlated with natural or anthropogenic
causes.
A Complex Effluent Toxicity Testing Program has been
well established for freshwater systems, but modifications
were necessary to adapt the methods to marine/estuarine
environments. EPA has designed state-of-the-art marine
toxicity test methods to serve as the marine counterpart
to the freshwater program, as presented in previous
sections.
In developing the toxicity test methods, two distinct, yet
related, goals were defined: 1) to develop tests that
rapidly estimate the toxicity of effluent discharges so that
regulatory actions can proceed and 2) to enhance the
comprehensive understanding of ecological systems and
trends and thereby contribute to long-term environmental
protection.
Complex Effluent Toxicity Testing
The Marine/Estuarine Complex Effluent Toxicity Testing
Program is currently used to help develop NPDES
permits. Test species include sheepshead minnows;
inland silversides; mysids; a sea urchin; and two
macroalga! species, Champia and Laminaria. These
species were selected for the following reasons:
• Laboratory cultures can be easily .cultivated.
• Gametes or larvae can be obtained year round for
shipping to onsite locations.
• Gametes or larvae require small volumes of
exposure water.
• Tests are inexpensive to conduct and can be easily
taught.
Future Research and Refinements to Complex
Effluent Toxicity Testing Program
To increase the range and accuracy of the
marine/estuarine CETTP, new species from other
phylogenetic groups and other geographic areas will be
added to the program (specifically, from the west coast).
In addition, future research will focus on special
environments, including brackish (low salinity) and tropical
waters. Biomonitoring techniques for brackish water
would be useful for discharges at estuary headwaters
including many areas within the Chesapeake Bay and the
Gulf of Mexico. Tropical species are needed to test
effluent discharges in the Hawaiian Islands, Virgin Islands,
Puerto Rico, Guam, and other U.S. protectorates. This is
particularly important because most States do not allow
the introduction of nonindigenous species.
In Situ Biomonitoring
ERL-N has developed in situ biomonitoring methods to
enhance and augment the short-term test methods and
monitor long-term trends of marine discharges. The in situ
program relies on mussels transplanted along dilution
gradients that start near the discharge site and end in
clean water. By correlating the contaminant levels in the
mussel tissues and subsequent biological effects, over
time, scientists can determine environmental trends near
to and at various distances from a discharge site.
Additionally, a newly developed laboratory exposure
system can help to field verify the laboratory CETTP
tests.
Future Research and Refinements for In Situ
Biomonitoring
The most successful biomonitoring tests to date have
been conducted at the individual organism level. At higher
levels of biological organization, the complexities of
biological systems increase tremendously. Monitoring at
the cellular and subcellular levels may simplify marine
monitoring because, rather than predicting toxic effects
on only one species, indicators at the cellular and
subcellular levels might predict effects across many
species. However, before this can be done, relationships
among cellular, tissue, individual, and population effects
must be understood, to justify extrapolation from one level
to the next and to explain the mechanisms behind whole
organism effects. Several promising techniques are
currently being investigated to expand the marine in situ
biomonitoring program.
For example, the slipper shell snail Crepidula fornicata is
under consideration as a likely candidate for in situ
33
-------
biomonitoring. C. fornicata survive well in field cages and
reproduce frequently and prolifically with easily observable
reproductive behavior. They can, therefore, be used for
transplanting and for reproductive tests, directly indicating
the onsite population level effects of a marine discharge.
Biomarkers (sublethal biochemical measures of effect)
may serve as subcellular indicators of exposure to
toxicants. In this context, biomarkers are defined as
"molecular biological techniques that may directly link
specific chemicals or classes of chemical compounds to
observed biological effects." Two such biomarkers are
measurements of increased incidence of sister chromatid
exchange and increased synthesis of metallothionein.
Sister chromatid exchange is the exchange of pieces of
DNA between the two arms of a chromosome. The rate
of this exchange has been shown to increase in the
presence of genotoxic agents (28). Even chemicals that
do not directly affect DNA (nongenotoxic agents) may still
affect the complicated mechanism of sister chromatid
exchange, and, therefore, may change the exchange rate.
This method will be explored further as an indicator of
pollution effects on invertebrate and fish species.
Metallothionein induction is another measure of toxic
exposure. Metallothionein is an intracellular metal-binding
protein. In mammals and fish, the binding action induced
by this protein may be a mechanism for tolerance to trace
metals. Synthesis of the protein is induced by the
presence of metals such as copper and cadmium. This
and other "stress proteins" show promise as indicators of
exposure to specific chemicals (28,6).
The winter flounder Pseudopleuronectes americanus is
the first teleost organism studied in this program, and
another candidate for population-level monitoring. This
benthic fish species returns annually to its nursery
grounds for spawning, making it a suitable indicator of
adverse population effects due to pollutant exposure. In
winter flounder, simple morphometric measures may
indicate exposure to certain pollutants. Moore and
Stegeman (40) have found liver tumors and severe
nonneoplastic changes in the liver, and suggest that
these are the result of exposure to aromatic hydrocarbons
and chlorinated pesticides.
Integrated Approach for Toxic Discharge
Control - Integration with Large-Scale Coastal
Monitoring
The integrated approach of combining short-term CETTP
tests with in situ biomonitoring not only promotes toxic
discharge control efforts, but also can determine
environmental trends. This integrated approach, therefore,
might be integrated with large-scale coastal monitoring
systems. A variety of Federally funded, regional
monitoring systems are compiling and integrating vast
amounts of chemical and biological environmental data to
better define present environments and document trends
that occur within them. NOAA's Statusi and Trends
Program and EPA's Environmental Monitoring and
Assessment Program (EMAP) are both designed for this
purpose.
The Integrated Approach - Field Tested
An example of the combined use of CETTP and in situ
biomonitoring tests is the New Bedford Harbor Pilot
Dredging Project (Section 5). In the late 1970s, the
Mussel Watch Program identified the harbor as a
"hotspot" of PCB contamination. During the pilot project,
biomonitoring was employed to monitor potentially toxic
releases from the dredging and disposal operation.
The Superfund study also demonstrated the effectiveness
of the real-time monitoring approach. In routine monitoring
efforts, the time gap from data collection to analysis
hinders progress, or renders the data meaningless when
operations proceed before data analysis. In the New
Bedford Harbor study, decision criteria were developed to
help managers interpret and act on the monitoring data.
Immediate decisions were made based on measurements
that were taken less than 24 hours earlier.
Finally, the New Bedford Harbor pilot study showed the
efficiency of the testing hierarchy - from mussel tissue
residues to in situ tests to extensive lab and field
biomonitoring. The hierarchy ensured that resources were
conserved while the harbor characterization was
completed. The case study in New Bedford Harbor stands
as a model of the integrated approach for toxic discharge
control.
34
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7. References
(1) U.S. EPA. 1984. Development of water quality-
based permit limitations for toxic pollutants:
national policy. Fed. Reg. 49:9016-9019.
(2) U.S. EPA. 1985. Technical support document for
water quality-based toxics control. U.S. EPA Office
of Water, Washington, DC.
(3) U.S. EPA. 1987. Water Quality Act. PL100-4.
(4) U.S. EPA. 1989. National Pollutant Discharge
Elimination System; Surface Water Toxics Control
Program. Proposed Rule. Fed. Reg. 54:1299-1323.
(5) Mount, D.I., N. Thomas, T. Norberg, M. Barbour,
T. Roush, and W. Brandes. 1984. Effluent and
ambient toxicity testing and instream community
response on the Ottawa River, Lima, Ohio.
Environmental Research Laboratory, Duluth, MN,
and Office of Water Enforcement and Permits, U.S.
EPA. EPA 600/3-84/080.
(6) Phelps, D.K. 1988. Marine/Estuarine biomonitoring:
a conceptual approach and future applications.
ERL-N Contribution No. 937. EPA 600/X-88/244.
(7) Peltier, W.H. and C.I. Weber, eds. 1985. Methods
> for measuring the acute toxicity of effluents to
freshwater and marine organisms. EPA 600/4-
85/013. U.S. EPA Environmental Monitoring and
Support Laboratory - Cincinnati, OH, Office of
Research and Development.
(8) U.S. EPA. 1988. Short-term methods for estimating
chronic toxicity of effluents and receiving waters to
marine and estuarine organisms. EPA 600/4-
87/028. U.S. EPA Environmental Research
Laboratory, Cincinnati, OH.
(9) Schimmel, S.C., ed. 1987. Users' guide to the
conduct and interpretation of complex effluent
toxicity tests at estuarine/marine sites. ERL-N
Contribution No. 796.
(10) Steele, R.L and G.B. Thursby. 1987. Draft
guidance manual for conducting sexual
reproduction tests with the marine macroalga,
Laminaria saccharina, for use in testing complex
effluents. Draft.
(11) Dinnel, P., Q. Stober, J. Link, M. Letourneau/W.
Roberts, S. Felton, and R. Nakatani. 1983.
Methodology and validation of a sperm cell toxicity
test for testing toxic substances in marine waters.
FRI-UW-83. University of Washington Sea Grant
Program, U.S. EPA.
(12) Breder, Charles M., Jr. 1948. Field Book of Marine
Fishes of the Atlantic Coast from Labrador to
Texas. G.P Putnam's Sons, New York and
London, p. 110.
(13) Perricone, C. 1980. In: Symposium on Research
on Environmental Fate and Effect of Drilling Fluids
and Cuttings. Volume 1: Proceedings. American
Petroleum Institute, Washington, DC., pp. 15-29.
(14) Brandsma, M.G., L.R. Davis, R.C. Ayers, Jr., and
T.C. Sauer, Jr. 1980. A computer model to predict
the short term fate of drilling discharges in the
marine environment. Abstract. In: Symposium on
Research on Environmental Fate and Effects of
Drilling Fluids and Cuttings. Volume 1:
Proceedings. American Petroleum Institute,
Washington, DC., p. 588.
(15) ASTM. 1987. Standard guide for conducting life
cycle toxicity tests with salt water mysids. ASTM
E-1191-87. American Society for Testing and
Materials, Philadelphia, PA.
(16) Grothe, D.R. and R.A. Kimerle. 1985. Inter- and
intralaboratory variability in Daphnia magna effluent
toxicity test results. In: Environmental Toxicology
and Chemistry. Vol. 4, pp. 189-192.
(17) Morrison, G.E., E. Torello, R. Walsh, A. Kuhn, R.
Burgess, M. Tagliabue, and W. Greene. 1988.
Precision of marine short-term chronic toxicity
tests. Draft.
(18) Parrish, P.R. and T.W. Duke. 1988. Variability of
the acute toxicity of drilling fluids to mysids
(Mysidopsis bahia). American Society for Testing
and Materials Special Technical Publication 976.
Philadelphia, PA.
(19) U.S. EPA. 1987. Permit writer's guide to water
quality-based permitting for toxic pollutants. U.S.
EPA Office of Water, Washington, DC.
35
-------
(20) U.S. EPA. 1985. Initial mixing characteristics of
municipal ocean discharges. Volumes I and II. EPA
600/3-85/073 a and b.
(21) Abbott, R.T. 1974. American Seashells: The
Marine Mollusca of the Atlantic and Pacific Coasts
of North America. Van Nostrand Reinhold, New
York and London.
(22) Phelps, O.K. and W.B. Galloway. 1980. A report on
the coastal environmental assessment stations
(CEAS) program. In: Proceedings of the
International Council for Ocean Exploration
(Rapport et proces verbaux des reunions, counseil
international pour /'exploration de la mer), Vol. 179,
pp. 76-81.
(23) Nelson, W.G. et al. 1987. Effects of Black Rock
Harbor dredged material on the scope for growth of
the blue mussel, Mytilus edulis, after laboratory
and field exposure. Technical Report D-87-7.
Prepared by U.S. Environmental Research
Laboratory, Narragansett, Rl, for U.S. Army Corps
of Engineers Waterways Experiment Station,
Vicksburg, MS.
(24) Nelson, W.Q., D. Black, and D. Phelps. 1985.
Utility of the scope for growth index to assess the
physiological impact of Black Rock Harbor
suspended sediment on the blue mussel Mytilus
edulis: a laboratory evaluation. Technical Report D-
85-6. Prepared by U.S. EPA Environmental
Research Laboratory, Narragansett, Rl, for U.S.
Army Corps of Engineers Waterways Experiment
Station, Vicksburg, MS.
(25) Conover, R.J. 1966. Assimilation of organic matter
by zooplankton. In: Limnology and Oceanography.
Vol. 4, pp. 338-354.
(26) Rogerson, P.P., S.C. Schimmel, and G. Hoffman.
1985. Chemical and biological characterization of
Black Rock Harbor dredged material. Technical
Report D-85-9. Prepared by U.S. EPA,
Environmental Research Laboratory, Narragansett,
Rl, for U.S. Army Corps of Engineers Waterways
Experiment Station, Vicksburg, MS.
(27) Nelson, W.G. 1988. Report on the comparison of
short-term and long- term exposures in East
Greenwich on the scope for growth of Mytilus
edulis. ERL-N Contribution No. 959.
(28) Phelps, O.K., C.H. Katz, K.J. Scott, and B.H.
Reynolds. 1987. Coastal monitoring: evaluation of
monitoring methods in Narragansett Bay, Long
Island Sound and New York Bight, and a general
monitoring strategy. In: T.P. Boyle, ed. New
Approaches to Monitoring Aquatic Ecosystems.
ASTM STP 940. American Society for Testing and
Materials, Philadelphia, PA, pp. 107-124.
(29) NOAA. 1988. National status and trends program
for marine environmental quality, progress report: a
summary of selected data on chemical
contaminants in sediments collected during 1984,
1985, 1986, and 1987. NOAA Technical
Memorandum NOS/OMA/44.
(30) Connolly, J.P. and J.P. St. John. 1988. Application
of a mathematical food chain model to evaluate
remedial alternatives for PCB-contaminated
sediments in New Bedford Harbor, Massachusetts.
In: Superfund '88, Proceedings ;of the 9th National
Conference. The Hazardous Materials Control
Research Institute, pp. 359-362.
(31) Ciavattieri, F.J. and S.L. Stockinger. 1988. New .
Bedford Harbor project management case study.
In: Superfund '88, Proceedings of the 9th National
Conference. Hazardous Materials Control
Research Institute, pp. 343-346.
(32) Averett, D.E. and N.R. Francingues, Jr. 1988. A
case study: dredging as a remedial action
alternative for New Bedford Harbor,
Massachusetts. In: Superfund '88, Proceedings of
the 9th National Conference. The Hazardous
Materials Control Research Institute, pp. 348-352.
(33) Allen, D.C. and AJ. Ikalainen. 1988. Selection and
evaluation of treatment technologies for the New
Bedford Harbor (MA) Superfund project. In:
Superfund '88, Proceedings of the 9th National
Conference. The Hazardous Materials Control
Research Institute, p. 330.
(34) Otis, M. J. 1987. Pilot study of dredging and
dredged material disposal alternatives. Draft. U.S.
Army Corps of Engineers, New England Division,
Waltham, MA.
(35) Otis, M.J. and D.E. Averett. 1988. Pilot study of
dredging and dredged material disposal methods,
New Bedford, Massachusetts Superfund Site. In:
Superfund '88, Proceedings of the 9th National
Conference. The Hazardous Materials Control
Research Institute, pp. 347-352.
(36) U.S. EPA. 1988. Decision criteria: New Bedford
Harbor pilot study. Prepared by U.S. EPA for U.S.
Army Corps of Engineers, New England Division,
Waltham, MA.
(37) U.S. EPA. 1988. New Bedford Harbor Pilot Study;
pre-operational monitoring progress report:
summary of ambient water quality conditions.
Prepared by U.S. EPA Environmental Research
36
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Laboratory, Narragansett, Rl, for U.S. Army Corps
of Engineers, New England Division, Waltham, MA.
(38) Phelps, O.K., D.J. Hansen, K.H. Scott, and A.S.
Fowler. 1988. Monitoring program in support of the
pilot study of dredging and dredged material
disposal methods, New Bedford, Massachusetts
Superfund Site. In: Superfund '88, Proceedings of
the 9th National Conference. The Hazardous
Materials Control Research Institute, pp. 335-337.
(39) Nelson et al. No date. Operational monitoring, the
New Bedford Harbor pilot dredging project, (in
progress)
(40) Moore, M.J. and J.J. Stegeman. 1988.
Histogenesis and molecular pathology in winter
flounder from Boston Harbor. Abstract. Woods
Hole Oceanographic Institute, Woods Hole, MA.
(41) Heber, M. 1986. Report on toxicity test results of
effluent/receiving water samples from [a pulp and
paper mill plant], Fernandina Beach, FL, May 14-
21, 1986. Memorandum to S. Schimmel. U.S.
EPA. July 15.
(42) Schimmel, S.C., G.B. Thursby, M.A. Heber, and
M.J. Chammas. No date. Case study of a marine
discharge: comparison of effluent and receiving
water toxicity. ASTM Proceedings, (in press)
(43) Dettmann, E.H., J.F. Paul, J.S. Rosen, and C.J.
Strobel. 1988. Transport, fate, and toxic effects of
a sewage treatment plant effluent in a Rhode
Island estuary. ERL-N Contribution No. 876. EPA
600/X-87/366.
(44) U.S. EPA. 1988. Dilution study and toxicity testing,
Bay County wastewater discharge: St. Andrew
Bay, Panama City, Florida. U.S. EPA
Environmental Services Division, Athens, GA.
37
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Appendix A.
Case Studies
CETTP Case Study I: Fernandia Beach, Florida
Toxicity testing of a complex effluent discharged into the
Amelia River north of Jacksonville, Florida, was a
cooperative effort between EPA Region IV (Athens, GA);
the EPA Environmental Research Laboratory 'in Duluth,
Minnesota (ERL-D); and ERL-N. The focus of the study
was a point-source discharge from a pulp and paper mill
in Fernandia Beach, Florida. The mill discharged
approximately 16 million gallons (60 million liters) of
effluent each day into the Amelia River with the outgoing
tide. The case study was a week-long analysis of the
effluent and its effects on the surrounding ecosystem.
The Amelia River, a high-salinity tidal basin, proved to be
a good location for a field evaluation of the
Marine/Estuarine CETTP test methods. The goals of this
study were:
• To determine the absolute toxicity of the pulp and
paper mill effluent discharged at the site.
• To determine the relationship between effluent and
receiving water toxicity.
• To generate data for deriving an NPDES discharge
permit for the mill.
Dye studies, toxicity tests, and chemical analysis of the
effluent were used to determine the source of the
observed effects of the pulp-mill effluent.
Effluent Tests
Effluent samples were collected daily from the pulp mill's
effluent lagoon adjacent to the discharge point. The
samples were composited over a 24-hour period and held
on ice until they were used in the toxicity tests. The
following test methods were performed using the
procedures specified in the EPA methods manual (8):
• A 7-day red macroalga (Champia parvula)
reproductive test.
• A 7-day reproductive, growth, and survival test with
the mysid Mysidopsis bahia.
• Tests using the inland silverside (Menidia beryllina)
and sheepshead minnow (Cyprinodon variegatus) to
measure larval fish growth and survival (41).
The effluent was toxic to all test species at concentrations
higher than 32 percent; however, the receiving water
concentrations of effluent in the river did not exceed 2
percent. The red macroalga and mysids were most
sensitive to the effluent; the NOEC for the alga was 1
percent effluent, and the mysid's NOEC was 3.2 percent
effluent. Fish larval growth was not significantly affected
at any effluent dilutions. Survival, however, was reduced
at 32 percent effluent for the sheepshead minnow, and at
10 percent for the inland silverside (see Table A-1).
Ambient Receiving Water Tests
Water samples were collected daily from 7 sampling
stations in the river (except 1 day, when 21 stations were
sampled; Figure A-1) during ebb tide, to coincide with the
discharge. The following toxicity test methods were
performed on the receiving waters (8):
• 7- to 9-day reproductive tests (performed on all 21
samples) with the red macroalga Champia parvula.
• A 7-day reproduction, growth, and survival test with
the mysid (Mysidopsis bahia).
• A 7-day growth and survival test with the
sheepshead minnow (Cyprinodon variegatus).
Using dye studies performed by EPA Region IV,
estimates were made of the effluent concentrations in the
receiving water samples collected (see Table A- 2). The
receiving water concentrations of effluent in the river did
not exceed 2 percent effluent, as was predicted by the
dye studies. Only the red macroalga and mysids indicated
toxicity at several monitoring stations. The alga was
especially sensitive to several receiving water stations,
and cystocarp production was reduced to zero or slightly
above zero (indicating that the species was incapable of
reproduction). The mysid receiving water tests were
inconclusive because receiving water toxicity could not be
correlated with estimated effluent toxicity based on the
dye study. Finally, the sheepshead minnow showed no
evidence of toxicity at concentrations as high as 10
percent effluent. The macroalga and sheepshead minnow
tests showed a correlation between laboratory and field
results (42).
Probable Causes of Toxicity
Using fractionation methods, EPA scientists at ERL-
Duluth and Region IV were able to identify ammonia as
the toxic constituent in the discharge. The freshwater
crustacean Ceriodaphnia dubia showed toxic response to
39
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Table A-1. Effects of Pulp Mill Effluent on Test Species (41)
Percent
Effluent
Control
0.5
1.0
2.5
3.2
5.0
10.0
32
50
Red Algae
Mean No.
Cystooarps
11+21
12 ± 3
13 ± 1
0.2'
0*
0*
Mysid
(%) Survival
100
100
88
41*
0*
0*
Inland silverside
(%) Survival
87
80
58"
33*
0*
0*
Sheepshead minnow
(%) Survival
98
98
93
11*
0*
"Significantly different from control.
"High variability between three replicates, but not significantly different from the control.
the ammonia in the effluent. Other tests, using the red
macroalga in effluent and ammonia mixtures containing^
the same concentration of the chemical, produced a very
similar toxic response (see Figure A-2). The calculated
concentrations of un-ionized ammonia (NHp) in the
effluent, along with the results from unpublished
laboratory studies also point to ammonia as the cause of
toxicity to the mysid and inland silverside (see Table A-3)
(42). Ammonia also exhibits a characteristic concentration
curve as ionization is affected by pH, temperature, and
salinity.
Setting Permit Limits
Numerical, nonbinding permit limits were calculated using
acute end-of-pipe toxicity data from the discharger's
records, along with the NOECs determined during the
study. Since the scientists were able to trace the toxicity
to a specific chemical, and the toxicity was found at the
very edge of the mixing zone set by the State of Florida,
it was relatively simple to direct the discharger on how to
reduce the effluent's toxicity. The State of Florida granted
the pulp and paper mill a 3-year variance to the water
quality standard, allowing them time to bring the effluent
within permit limits. The EPA Region IV office will follow
up on the toxicity testing at the end of the 3-year waiver.
CETTP Case Study II: East Greenwich, Rhode
island
Greenwich Cove is an estuarine cove located in
Narragansett Bay between East Greenwich, Rhode Island,
and Goddard State Park in Rhode Island (see Figure A-3).
A Publicly Owned Treatment Works (POTW) in East
Greenwich discharges^an average of 0.7 million gallons
per day (MGD) into the cove. At the time of the first
phase of this case study, the POTW was designed to
treat only 0.5 MGD. The Town of East Greenwich is
currently upgrading the plant, and after modification, the
plant will be capable of providing advanced secondary
treatment for 1.2 MGD of sewage. A variety of
marine/estuarine CETTP tests were used to determine
the effluent and receiving water toxicity for the cove
during four field studies performed over a 1-year time
period (December 1985, July, September, and November
1986).
After the POTW is upgraded, additional toxicity tests will
be conducted, providing a "before and after" picture of
the effluent's toxicity and its effects on the cove.
Chemical and Physical Effluent Characterization
Chemical Characterization. The POTW effluent was
analyzed for dissolved and particulate trace metals
(cadmium, chromium, copper, iron, lead, manganese,
nickel, zinc); organic compounds such as PCBs and
PAHs; and nutrients (e.g., ammonia, nitrogen, and
phosphorous). Total suspended solids and total residual
chlorine also were measured. Effluent samples were
collected for fractionation studies to determine what
class(es) of compounds caused the observed toxicity. As
part of the fractionation procedure, toxicity tests were
conducted. These tests showed that chlorinated effluent
was more toxic than unchlorinated, and that the majority
of the toxicity could be attributed to a pompound in the
dissolved portion rather than one associated with the
particulates. When further separation of the effluent was
performed, it was determined that the toxicity was
probably due to moderately polar constituents (43).
Physical Characterization. Mixing in the cove was
determined by dye study techniques. The dye studies
also provided information on the age of the effluent at the
different sampling stations in the cove (see Figure A-4).
At the center of the boil area (the area above the
submerged outfall, or where the effluent actually enters
the water from the discharge pipe) effluent concentration
was measured at as high as 40 percent. The
concentration dropped to less than 1 percent, however, at
all other sampling stations. Old effluent (effluent remaining
40
-------
Figure A-1. Amelia River, in Fernandina Beach, Florida (42).
in the cove after one tidal cycle) and new were found to
coexist in the cove - the mixture near the outfall was
primarily fresh effluent (Station 2); older effluent was
found near the mouth of the cove (Stations 4 and 5); and
the oldest effluent was found at the head, or
southernmost extent of the cove (Station 1). Figure A-5
shows the mixing characteristics as determined
studies.
using dye
No permanent (more than 2-day) stratification in
temperature, salinity, or dye concentrations was detected
in the study, although a salinity gradient was measured
that showed increased salinity from the interior of the
41
-------
Table A-2.
Effect of Receiving Waters from the Amelia River, Jacksonville, Florida, on Test Species. Effluent percents
were calculated based on a dye study (42).
Red Algae
Mean No. of Cystocarps
Mysid
Station
Number
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
Percent
Effluent
2.0
control (0%)
- (no data)
1.5
-
0.05
1.4
0.08
1.4
0.5
0.2
0.9
0.7
0.5
0.5
-
0.6
0.5
0.4
0.4
0.5
May 17 May 18
0.5" 0.8*
14.7 19.0
1.0* 3.0*
1.0.4
2.3* 0.6*
1.1*
3.4
8.3 21.0
1.1*
15.2 13.5
12.8 19.9
7.2
5.3*
5.2*
7.9
11.7
9.7
10.4
10.2
9.6
(%) Females
(%) Survival with Eggs
90 64
93 71
88 77
98 47
93 14*"
93 66
93 31
Sheepshead Minnow
(5) Survival
98
98
100
100
(
95
100
' 100
"Significantly different from control.
"Significantly different from control, but not attributed to effluent toxicity.
cove (where the freshwater Maskerchugg River
contributes to the cove) to the outer cove, toward
Greenwich Bay. Table A-4 shows the flow rates in the
cove. Tidal range and currents, as well as the discharge
flow rate of both the POTW and the Maskerchugg River,
all affect the mixing and stratification characteristics in the
cove (43).
Complex Effluent Toxicity Tests
The chlorinated effluent samples were diluted to effluent
concentrations ranging from 32 to 3.2 percent with
relatively clean Narragansett Bay seawater, and the
salinity was adjusted to a standard of 30 parts per
thousand with Narragansett Bay brine (8). The following
tests were performed on the effluent samples:
• Sperm cell toxicity tests with the sea urchin Arbacia
punctulata.
• Reproductive effect tests on the red macroalga
Champia parvula.
• 7-day reproduction, survival, and growth tests using
Mysidopsis bahia, the mysid.
• Tests with sheepshead minnows (Cyprinidon
variegatus) and the inland silverside (Menidia
beryllina) to indicate effects on fish larvae growth
and survival.
The effluent toxicity varied from day to day. Averaging the
toxicity over all of the test periods, a concentration of
0.24 percent of the effluent affected sea urchin
reproductive capacity, while greater than 10 percent
effluent concentration was toxic to the fish species tested
(see Table A-5).
Ambient Receiving Water Tests
The scientists used the same methods to test the toxicity
of the effluent as were used in the complex effluent
analysis, and an additional test was performed, the kelp
reproduction test. This test uses the sea kelp Laminaria
saccharina (see Section 2) to evaluate the reproductive
effects of the receiving water.
All samples from the boil station showed similar effects on
the test species. The other stations gave variable results.
Percent effluent at each station was determined using dye
studies (Table A-6). The species' toxic responses differed
from one test period to the next, indicating a highly
42
-------
120 —i
• NH4CI
D Effluent Dilution
2345
% Effluent
Figure A-2. Number of cystocarps (as percent of the number
on control-exposed red microalga) versus percent
effluent. The effluent data are averaged over 2
days. A stock solution of 26.4 mg NH4CI/100 mL
was used to match the effluent with an assumed
concentration of 70 mg NH4-N/L of effluent (42).
variable effluent. The red macroalga Champia and the
sea urchin were most sensitive to the contaminated
receiving waters, (see Table A-7).
Conclusion
The East Greenwich Cove Marine CETTP was generally
, successful in its attempt to field-validate the laboratory
effluent toxicity tests. The investigators found that effluent
"residence time" and toxicity half-life (or rate of toxicity
reduction) are important factors to build into estimations
, of lab to field results. In this case, the variable receiving
water toxicity could be explained by the variability in the
effluent coming from the POTW, and by toxicity reduction
or degradation with time. Although the toxicity could not
be traced to a specific chemical, it was concluded that
the total toxicity observed in the field generally could be
attributed to the POTW discharge, although occasional
toxicity observed in the more sensitive tests distant from
the discharge could be from other sources (e.g., boats
moored in the cove, runoff, and contaminants in the
Maskerchugg River, etc.).
CETTP Case Study 111: Panama City, Florida
. This study in Saint Andrew Bay (see Figure A-6) focused
on the effluent from the Bay County Wastewater
Treatment Plant. The plant discharges about 30 million
gallons of waste per day into the bay. Scientists from
EPA Region IV (Athens, Georgia) used dye studies,
physical and chemical characterization of the effluent, and
CETTP testing methods to determine the discharge's
effect on the bay's ecosystem.
A 25-hour dye tracer study was conducted to establish
the mixing and dilution characteristics of the effluent. The
release was continuous with the discharge and spanned
one tidal cycle. This initial measurement provided
concentrations for the receiving water tests, and further
measurements over five additional tidal cycles provided
information on the effluent buildup in the bay. Table A-8
shows the dilutions in the bay at the time of the study.
Figure A-7 shows the dilutions averaged over six tidal
cycles. The scientists concluded that those areas east of
the boil receive effluent during flood tides, while those
west of the boil receive maximum exposure during ebb
tide (44). They also found that the effluent's concentration
was maximized in the upper 1 to 3 meters of the water
column because of its lower density.
Effluent Characterization and Toxicity Testing
Measures of biological oxygen demand (BOD5), total
suspended solids (TSS), and fecal coliform counts were
performed on the discharged wastewater, and the results
were within the POTW's reported values.
The toxicity of the freshwater effluent was determined
using a larval survival and growth test with the freshwater
species Pimephales promelas, or fathead minnow, and a
survival and reproduction test using Ceriodaphnia dubia,
a freshwater crustacean. The effluent was made saline
(using commercial salt mix or hypersaline brine) for the
following marine organism tests:
• Static acute toxicity test using the mysid Mysidopsis
bahia.
• Sperm cell toxicity test with the sea urchin Arbacia
punctulata.
• Reproduction and toxicity test using the red
macroalga Champia parvula.
The results of these tests are summarized in Table A-9.
The freshwater tests with 100 percent effluent showed no
effect on daphnid or larval fathead minnow survival and
growth. Reproduction in the daphnid was affected,
however, at 100 percent effluent. The No-Observed-
Effect Concentration (NOEC) in this case was 30 percent.
The effluent was acutely toxic to the mysids, with a LC50
concentration of 70.7 percent. Sea urchin fertilization was
unaffected at 70 percent effluent, the highest
concentration tested. The NOEC for the red macroalgal
reproduction was 12 percent effluent.
Receiving Water Toxicity Tests
Receiving water samples were collected in June and July,
1988. In June, four samples were collected at each
station (boil, 200E, and 200W) in a consecutive ebb,
43
-------
Table A-3. Toxicity Comparisons of Calculated Un-ionized Ammonia in Effluent (Average Concentration was Calculated
from Total Ammonia, which was 3.66 mg/L) and Separate Ammonia Tests (42)
Effluent Tests
Species
Champ/a parvula
(red macroalga)
Mysidopsis bahia
(mysid)
Menidia beryllina
(inland silverside)
Lowest Effect Concentration
(% Effluent)
2.5
3.2
10.0
Calculated Chronic Value"
for Un-ionized Ammonia (NH),
(mg/L)
0.06
0.07
0.20
Ammonia
Tests-
Chronic
Value
(mg/L)
0.04
0.12
0.06
•Chronic value is the geometric mean of the highest no-effect concentration and the lowest effect concentration for 2-day tests (C. ;
parvula) and 7-day tests (M. bahia and M. beryllina). ,
-Unpublished ERL-N data; chronic numbers refer to calculated un-ionized ammonia concentrations from 2-day M. bahia tests, and 28-
day M. beryilina tests.
A survey of the effluent's effects was performed on the
bay's seagrass community. That study is described in
Dilution Study and Toxicity Testing, Bay County
Wastewater Discharge (44).
In Situ Biomonitoring Case Study: East
Greenwich, Rhode Island
An in situ biomonitoring study was conducted in
Greenwich Cove in Greenwich Bay, Rhode Island, during
October and November, 1986, to compare the results of
scope for growth tests on mussels (Mytilus edulis)
conducted over long and short time periods. The short-
term test (7 days) was conducted within the time period
of the long-term, 30-day test. The observed toxicity could
be attributed to the East Greenwich POTW.
Method
Three mussel deployment stations were located around
the cove: one was in the cove near the outfall (Station 3);
the second was in the mixing zone between the cove and
Greenwich Bay (Station 6); and the third was at the
mouth of Greenwich Bay (Station 10; see Figure A-10). A
reference station (REF) was chosen at the mouth of
Narragansett Bay. As described in Section 3,
polypropylene baskets containing mussels were anchored
at each station, suspended 1 meter above the bottom.
Clearance rate, respiration rate, and food assimilation
efficiency were measured as described in Section 3 of
this document. Average scope for growth was calculated
for each group of mussels, and the 7- and 30-day values
were compared for each station (27).
Results
There was a graded response in scope for growth of the
mussels exposed for both 7 days and 30 days. Table A-
12 shows the SFG and individual parameter results at
each station, and compares the 7-day and 30-day
exposure times. The SFG of the mussels at the reference
Greenwich Cove in winter (ice in foreground).
slack low, flood, and slack high tide. Grab samples were
collected near the boil, Shell Island, and station 1E for the
July test period (Figure A-8). The same marine species
tests were performed as for the effluent tests. ERL-N
assisted Region IV, performing the sea urchin and algal
tests in the Narragansett lab. At the site, dissolved
oxygen, salinity, and temperature measurements were
taken, as well as light transmission readings. Figure A-9
shows the light transmission at four stations; Table A-10
gives the salinity, temperature, and dissolved oxygen
measurements for the same time period.
The mysids showed no response to the receiving water at
any of the stations. The red macroalga reproductive tests
were inconclusive; the varied results suggested that the
St. Andrew Bay waters are in themselves unable to
support Champia parvula reproduction (see Table A-11).
44
-------
Maskerchugg
0 0.5 1.0 River
Figure A-3. Greenwich Cove, Rhode Island (43).
station was almost unchanged from 7 to 30 days. SFG at
the remaining stations, however, decreased moving
toward the East Greenwich POTW outfall with both time
periods. This trend in SFG was more pronounced after 30
days of exposure (27). The more dramatic decrease
observed after 30 days may indicate continued
physiological decline with prolonged exposure to the
effluent, but since actual effluent exposure was not
measured concurrently, the physiological decline cannot
be tied conclusively to the effluent's presence in the
cove. Tissue residue analysis is being conducted, and will
provide more information about the conditions of the cove
waters during the exposure period.
The results of this study imply that the effects of 7-day
exposure are indicative of longer-term effects, and can be
used to predict biological response to effluent toxicity.
45
-------
East Greenwich POTW, viewed from Greenwich Cove (December
1985).
41°40'N -
41 °38'45"N
Boil Area
East
Greenwich
POTW
Arbitrary Definition
of Greenwich Cove
Boundary
500m
71°27'30"W 71°26'30"W
Figure A-4. Sampling stations in Greenwich Cove (43).
71 °25'30"W
46
-------
December 9
Low Tide
(1100)
December 9
Low Tide + 4 Hr
(1500)
December 10
Low Tide
(1155)
December 11
Low Tide
(1248)
December 13
Low Tide
(1430)
Figure A-5. Dye distribution in Greenwich Cove during December,
per billion (ppb) (43).
985. Numbers represent dye concentration isopleths in parts
47
-------
Table A-4. Flow Rates into Greenwich Cove (43)
Flow Rate
Source
POTW
Maskerchugg R.
Mean Tidal Flow
(m3/day)
2,650
1,420 to 138,000
2,300,000
(million gallons/day)
0.70
0.38 to 36.46
607.70
Table A-5. Rank Order of Species Sensitivity to Greenwich
Cove POTW Effluent (43)
Study
Period
12/85
7/86
9/86
11/86
Species
Arbacia punctulata (sea urchin)
Mysidopsis bahia (mysid)
Champia parvula (red algae)
Cyprinodon variegatus (sheepshead)
Menidia beryll'tna (inland silverside)
Arbacia punctulata (sea urchin)
Mysidopsis bahia (mysid)
Champia parvula (red algae)
Menidia beryllina (inland silverside)
Arbacia punctulata (sea urchin)
Champia parvula (red algae)
Mysidopsis bahia (mysid)
Menidia beryllina (inland silverside)
Arbacia punctulata (sea urchin)
Champia parvula (red algae)
Mysidopsis bahia (mysid)
Menidia beryllina (inland silverside)
NOEC"
(% Effluent)
1.0
1.0
3.0
10.0
10.0
<0.3
3.2
5.4
10.0
<0.24
4.25
10.0
10.0
<0.6
2.5
3.2
10.0
"NOEC = No-Observed-Effect Concentration.
Table A-6. Mean Percent Effluent from the East Greenwich POTW Found in Surface Water Samples from
Greenwich Cove and Greenwich Bay, Rhode Island During Three 1986 Sampling Periods: July 16 to 22,
September 8 to 15, and November 9 to 18. All sampling was performed at slack low tide (43)
Receiving Water Stations '
Date
July
September
November
Overall
MEAN
Overall
MAXIMUM
Overall
MINIMUM
1
0.16
0.13
0.23
0.17
0.33
0.031
2
0.15
0.15
0.28
0.18
0.73 ,
0.010
outfall
26
29
17
24
40
6.3
3
0.28
0.30
0.38
0.33
0.66
0.082
4
0.35
0.30
0.24
0.30
0.68
0.14
6
0.12
0.12
0.27
0.16
0.86
0.039
7
0.064
; 0.054
0.083
0.063
0.14
0.013
48
-------
Bay Count'
Aerated Lagoon
Figure A-6. St. Andrew Bay, Panama City, Florida (44).
49
-------
Table A-7 Results of Toxicity Tests Conducted on Surface Waters from Greenwich Cove and Greenwich Bay, Rhode
Island, During Three 1986 Sampling Periods: July 16 to 22, September 9 to 15, and November 11 to 18. Toxicity
is expressed in the numerator of each fraction as the number of occasions that the sample collected from a
specific station was significantly different from the control. The denominator denotes the number ol receiving
water samples evaluated. Overall toxicity is the total number toxic of all samples collected from the three test
periods; the number in parentheses denotes the percent samples that were toxic of all tested (43).
Species/Date
Receiving Water Stations
outfall
C. parvula
(red algae)
July
September
November
Overall Toxicity
A. punctulata
(sea urchin)
July
September
November
Overall Toxicity
M. bahia
(mysid)
July
September
November
Overall Toxicity
M. beryllina
(inland silverstde)
July
September
November
Overall Toxicity
0/6
0/5
1/1
1/12(8%)
1/7
4/6
2/6
7/19 (39%)
0/1
0/1
0/1
0/3 (0%)
0/1
0/1
1/1
1/3 (33%)
0/6
0/5
1/1
1/12 (8%)
5/7 .
3/6
1/6
9/19 (47%)
0/1
0/1
0/1
0/3 (0%)
0/1
0/1
0/1
0/3 (0%)
6/6
5/5
1/1
12/12 (100%)
7/7
6/6
6/6
19/19 (100%)
1/1
1/1
1/1
3/3 (100%)
1/1
1/1
1/1
3/3 (100%)
0/6
0/5
1/1
1/12 (8%)
3/7
3/6
2/6
8/19 (42%)
0/1
0/1
0/1
0/3(0%)
0/1
0/1
0/1
0/3(0%)
1/6
0/5
1/1
2/12(17%)
4/7
1/6
2/6
7/19 (37%)
0/1
0/1
0/1
0/3 (0%)
0/1
0/1
0/1
0/3 (0%)
1/6
0/5
0/0
1/12 (8%)
3/7
1/6
2/6
6/19 (32%)
0/1
0/1
0/1 .
0/3 (0%)
0/1 ;
0/1 ;
0/1
0/3 (0%)
1/6
0/5
1/1
2/12 (17%)
4/7
1/6
2/6
8/1 9 (42%)
0/1
0/1
0/1
0/3 (0%)
0/1
0/1
0/1
0/3 (0%)
50
-------
Table A-8. Dilution Summary, St. Andrew Bay, Florida, June,
Low Tide
988 (49)
High Tide
Station
10E
8E
6E
4E
2E
200E
Boil
200W
2W
4W
6W
8W
10W
12W
4S
Ultimate Tracer
Cone."
0.9
1.3
2.1
2.1
2.5
10.8
45.5
15.0
4.3
4.0
4.1
3.2
2.7
0.8
1.4
Wastewater
Dilution
800:1
550:1
340:1
340:1
285:1
66:1
16:1
48:1
165:1
180:1
175:1
225:1
265:1
900:1
515:1
Ultimate Tracer
Cone."
0.6
2.0
2.1
2.2
5.3
18.3
40.0
16:2
3.4
2.2
1.4
0.9
0.2
0.6
0.2
Wastewater
Dilution
1200:1
360:1
340:1
325:1
135:1
39:1
18:1
44:1
210:1
325:1
515:1
800:1
3600:1
1200:1
3600:1
— Average
Conditions
Wastewater Dilution
1000:1
455:1
340:1
335:1
210:1
53:1
17:1
46:1
190:1
255:1
345:1
510:1
1930:1
1050:1
2050:1
*ppb
12W
11050:1
North
Figure A-7. Steady-state dilution, average conditions, June 1988 (49)
51
-------
Table A-9. Toxicity of Effluent from Bay County Wastewater Treatment Plant, Panama City, Florida (44)
Daphnid
Percent
Effluent
0*
1
3
6
10
12
25
30
50
70
100
(%)
Survival
100
100
100
100
100
90
Mean No.
Offspring/Female
26.47
17.63
28.60
30.43
28.83
5.00*
Mysid (%)
Survival1
100
100
100
80
80
20*
Fathead
minnow (%) Sea urchin (%)
Survival Fertilization
100 95.5
100
100
100
100 !
89.6 i
100
Red Algae
Mean No.
Cystocarps
9.0
6.6
5.4*
2.4*
1.0*
0*
'The calculated UC50 for Mysidopsis bahia was 70.7 percent.
^Control sample collected from Narragansett Bay.
"Significantly different from control.
North
Shell Island
Figure A-8. St. Andrew Bay receiving water sampling stations (44).
52
-------
100
H-
o>
8 10 12 14
16 18 20 22 24
Dep'th (Feet)
26 28
30
__.__. Boi| ._....(.— 2QOW
Figure A-9. Light transmission, as measured by marine photometer (44).
200E
Table A-10. Salinity, Terr perature, and Dissolved
Oxygen at the Boil, 200E, and 200W
Stations in June, 1988 (44)
Boil
Depth
(ft.)
1
5
9
15
19
25
Sal.
(PFj>t)
32
32
32
33
34
35
3
9
8
6
9
0
Temp.
(•C)
26.9
27.6
27.2
26.6
26.8
26.6
D.O.
(ppm)
6.0
6.2
6.5
6.4
6.5
6.5
B-200E 1
5
9
15
19
B-200W 1
5
9
15
19
25
32
32
33
34
35
32
9
9
5
6
2
5
32.5
33.B
334
34.
34.
1
4
28.0
27.2
27.3
27.0
26.4
Avg. =
28.1
28.0
27.4
27.0
26.4
26.4
Avg. =
6.2
6.0
6:3
6.3
6.1
6.2
6.9
7.0
6.9
6.8
6.3
6.4
6.6
Avg. =
6.4
53
-------
Table A-11. Results of Mysidopsis bahia (mysid), Arbacia punctulata (see urchin)
Champia parvula (red algae) Tests on F^jceiying Wajers^rn St. ^ngrej
Sample Sample Sample (%) (%) Mean No.
Date Location Time Survival Fertilization Cystocarps
and
6/14-15/88
7/15/88
6/15/88
7/19/88
Boil
SOW
100W
200W
Boil
SOW
50E
100E
Boil
50E
100E
200E
Boil
100W
200W
Control
Boil
Shell I.
1E
Control
1130 100
1325 100
1220 100
1305 90
2050
2110
2145
2130
0800
0815
0830
0840
1100
1110
1125
Treatment1
1321
1307
1347
Treatment1
92.8
89.4
87.9
88.8
81.3
86.9
84.5
86.5
86.0
82.8
93.5*
89.9
87.7
83.2
86.2
95.5
96.0
94.7
93.4
97.4
2.9
1.7*
4.5*
2.7
13.9 '
6.8
1.1*
3.3
9.0
"Significant difference from the boil based on Analysis of Variance and tests.
'Sample collected from Narragansett Bay.
Figure A-10. Greenwich Cove, Rhode Island, with monitoring
stations (27).
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Table A-12. Results of the Physiological Measurements and Scope for Growth Index
of Mussels from 7- and 30-Day Exposure (Values in Parenthesis are
Standard Errors) (27)
7-Day Exposure
Station
Ref
10
6
3
Station
Ref
10
6
3
Clearance
Rate (L/h)
5.11 (0.22) AB
6.07 (0.23) A
4.48 (0.28) B
4.30 (0.39) B
Clearance
Rate (L/h)
5.88 (0.58) A
5.02 (0.46) AB
4.41 (0.42) AB
3.05 (0.54) B
Absorption
Efficiently %
85 (ik A
91 (0) B
89 (2)
86 (0)
Absorf
Efficien
84(5
95(0
96(1
89(1)
AB
AB
30-Day
ition
:y%
A
B
B
AB
Respiration
Rate (mL-02/h)
0.29 (0.03) A
0.48 (0.04) B
0.50 (0.02) B
0.50 (0.03) B
Exposure
Respiration
Rate (mL-02/h)
0.39 (0.04) A
0.53 (0.02) AB
0.68 (0.08) B
0.55 (0.06) AB
Scope for
Growth (J/h)
17.7(1.1) A
17.7 (0.9) A
13.0 (1.7) AB
11.4(1.3)6
Scope for
Growth (J/h)
17.2(2.4) A
15.0 (1.7) AB
10.7 (2.0) AB
6.5 (3.2) B
55
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Appendix B
Glossary
Absolute Toxicity. Laboratory-determined toxicity
(usually determined using CETTP methods) of a complex
effluent without comparison to receiving water toxicity.
Acute. Occurring within a short period of time. In toxicity
tests, a response observed within 96 hours or less
typically is considered an acute effect.
Acute to Chronic Ratio (ACR). The ratio of the acute
toxicity of a substance (expressed as an LC50) to its
chronic toxicity (expressed as a NOEC, LOEC, or EC50).
ACRs can be used to estimate chronic toxicity on the
basis of acute toxicity data, helping in the determination
of NPDES permit limits.
Ambient Toxicity. Toxicity produced by a sample of
water taken from a larger water body.
Benthic Zone. The ocean floor.
Bioaccumulation. The uptake and retention of
substances by an organism from its surrounding media
and food.
Bioavailability. The degree to which a substance is or
becomes available to the tissues after exposure. A
substance's bioavailability determines its effect on
exposed organisms.
Biochemical Oxygen Demand (BOD). The quantity of
oxygen utilized in the biochemical oxidation of organic
matter in a specified time and at a specified temperature.
It is not related to the oxygen requirements in chemical
combustion, being determined entirely by the availability
of the material as a biological food and by the amount of
oxygen utilized by the microorganisms during oxidation.
Biomarker. A molecular biological indicator that may
directly link specific chemicals or classes of chemicals to
observed biological effects.
Biomonitoring. Any method of testing that includes the
use of aquatic plant or animal species to measure acute
or chronic toxicity, and any biological and/or chemical
measure of bioaccumulation.
Biosurvey. A method for determining the variety and
numbers of species in a biological community.
Chronic. Occurring after a prolonged period of time. A
chronic effect is one that is observed after long-term
(usually low-level) exposure.
Chronic Value (ChV). The geometric mean of the
Lowest-Observed-Effect concentration (LOEC) and the
No-Observed-Effect Concentration (NOEC).
Clearance Rate (CR). The volume of water that a
mussel can completely "clean" (remove particles larger
than 3 microns) within a specific time period.
Cystocarp. The spore case of the red marine alga
Champia parvula. Cystocarps are evidence of sexual
reproduction in the algae, and their formation is used as a
toxicity test endpoint.
Diversity. The number of species in a specific location.
Direct Discharge. Release of a pollutant from a pipe or
outfall directly into a water body.
Effective Concentration (EC). The effluent
concentration that has an effect on a given percentage of
test organisms.
Effective Concentration/50 (EC50). The effluent
concentration that has an effect on 50 percent of the test
organisms.
Effluent. Any pollutant or mixture of pollutants that is
discharged into surface waters or treatment works.
Effluent Limitation. A restriction imposed by State or
Federal authorities on the quantities, rates, and
concentrations of pollutants that are discharged into
surface waters.
Estuary. The lower end of a river where freshwater
meets and mixes with seawater.
Fecundity. An organism's capacity for producing
offspring.
Fluorometry. An instrumental method for measuring the
amount of fluorescence in a water sample. Used during
dye studies to estimate the concentration of an effluent in
the receiving water.
Fraotionation. A method for separating and identifying
the components in a chemical mixture and each part's
effects on test organisms. For example, an effluent may
be gravity filtered, and the filtered solution used in a
toxicity test.
Gas Chromatography (GC). An instrumental method for
identifying the component parts (especially volatile
57
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organics) in a chemical mixture. The solvent extract of the
aqueous sample is injected into a column containing a.
stationary phase (resin) and a carrier phase (gas). The
vaporized volatile components leach off the column
according to their relative affinity for the stationary and
carrier phases, and show peaks on a chart proportional to
their presence in the mixture.
Hypersaline. A solution whose salinity is greater than
normal seawater.
Larva. An immature form that is unlike the adult
organism.
Lowest-Observed-Effect Concentration (LOEC). The
lowest concentration of a substance at which an effect
was observed in test organisms.
No-Observed-Effect Concentration (NOEC). The
highest concentration of a substance at which no effect
was observed in test organisms.
Nonpoint Source. A pollution discharge source that is
not released from a distinct source, such as stormwater
runoff that is collected and channeled by humans.
National Pollutant Discharge Elimination System
(NPDES). The National program for administering permits
for the discharge of pollutants into the waters of the
United States under the Clean Water Act.
Outfall. The discharge point from any discernible
conveyance for effluent discharge (i.e., a pipe, ditch, or
channel).
Phylogenetic. Pertaining to ancestral development.
Point Source. A single source of effluent, such as a
manufacturing facility, that discharges into surface waters.
Precision. When repeated test results do not deviate
significantly among themselves. Compare this to
accuracy, in which the test results do not vary
significantly from a standard.
Relative Toxicity. The toxicity of an effluent after it is
mixed in a receiving water - determined with either
CETTP receiving water tests or in situ biomonitoring
methods.
Scope for Growth. A measure of the energy available to
an organism for tissue production, both somatic and ;
reproductive, after routine metabolic costs are accounted
for.
Superfund. Federal authority under the Comprehensive
Environmental Response, Compensation, and Liability Act
(CERCLA, ratified in 1980) to respond directly to releases
or threats of releases of hazardous substances. ;
"Superfund" refers to the Federal funds that are available
to help pay for hazardous waste cleanups.
Teratogenicity. A substance's ability to produce
anomalies of formation or development.
Wasteload Allocation. The portion of a receiving water's
total maximum allowable daily pollution load that is
allocated to one of its existing or future point sources of
pollution.
Whole Effluent Toxicity. The aggregate toxic effect of
an effluent (usually measured directly with a toxicity test).
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