CBP/TRS 82/92
November 1992
Year Two Report
A Pilot Study for
Ambient Toxicity Testing
in Chesapeake Bay
Ch esapeake Bay Program
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Year Two Report
A Pilot Study for
Ambient Toxicity Testing
in Chesapeake Bay
Produced under contract to the U.S. Environmental Protection Agency
Contract No. 6S-WO-00-43
Printed by the U.S. Environmental Protection Agency for the Chesapeake Bay Program
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Year 2 Report
A Pilot Study for Ambient
Toxicity Testing in Chesapeake Bay
Lenwood W. Hall, Jr.
Michael C. Ziegenfuss
Steven A. Fischer
Ronald D. Anderson
William D. Killen
University of Maryland System
Maryland Institute for Agriculture and Natural Resources
Agricultural Experiment Station
Wye Research and Education Center
Box 169
Queenstown, Maryland 21658
Raymond W. Alden, III
Emily Deaver
Old Dominion University
College of Sciences
Applied Marine Research Laboratory
Norfolk, Virginia 23529-0456
Jay W. Gooch
Nikki Shaw
University of Maryland System
Chesapeake Biological Laboratory
Box 38
Solomons, Maryland 20688
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FOREWORD
This pilot study was designed to evaluate ambient toxicity in
the Chesapeake Bay watershed by using a battery of water column,
sediment and suborganismal tests. A team of scientists from three
Chesapeake Bay research laboratories worked jointly to complete
this goal. Water column toxicity studies were directed by Lenwood
W. Hall, Jr. of the University of Maryland System's Agricultural
Experiment Station, sediment toxicity tests were managed by Raymond
W. Alden, III of Old Dominion University Applied Marine Research
Laboratory, and suborganismal tests were directed by Jay Gooch of
the University of Maryland System's Chesapeake Biological
Laboratory. This report summarizes data from the second year of
this two-year study. The following government agencies were
responsible for supporting and/or managing this research: Maryland
Department of Natural Resources (CB-90-001-005), Maryland
Department of Environment (182-C-MDE-91) and the U.S. Environmental
Protection Agency (X-003554-01).
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ABSTRACT
Contaminants and adverse water quality conditions (pollution)
in the Chesapeake Bay watershed have been postulated as factors
responsible for declining resources. Conclusive research designed
to establish a link between pollution in the ambient Bay
environment and declining resources can only be accomplished by
direct measurement of biological responses. In 1990, we initiated
a pilot study to broadly assess ambient toxicity of living resource
habitats for the purpose of identifying defined regions where
ambient toxicity levels warrant further investigations (Hall et
al., 1991). The primary goal of the second year (1991) of this
ambient toxicity testing pilot study was to identify toxic areas in
living resource habitats of the Chesapeake Bay watershed by using
a battery of standardized, directly modified or recently developed
water column, sediment and suborganismal toxicity tests. Tests
were conducted twice at the following stations: Potomac River-
Morgan'town, Potomac River-Dahlgren, Patapsco River and Wye River.
A suite of inorganic and organic contaminants was evaluated in the
water column and sediment during these tests. Standard water
quality conditions were also evaluated in water and sediment from
all stations.
The following water column tests were conducted in saltwater:
8-d sheepshead minnow, Cyprinodon varieaatus. survival and growth
test; 8-d larval grass shrimp, Palaemonetes pugio. survival and
growth test; 8-d Eurvtemora affinis life cycle test and 8-d mysid
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shrimp. Mvsidopsis bahia. survival, growth and reproduction test.
Sediment toxicity tests conducted in saltwater areas at the four
stations were: 2 0-d amphipod, Lepidactvlus dvtiscus. survival,
growth and reburial test; 20-d grass shrimp, Palaemonetes pugio.
survival and growth test; 20-d polychaete worm, Streblospio
benedicti. survival and growth test and 20-d amphipod, Hvalella
azteca, survival and growth test. Suborganismal tests were
conducted to evaluate EROD activity in feral fish collected
adjacent to the four "sample sites; determine the holding time
required for Fundulus heteroclitus from St. Johns Creek and the Wye
River to reach baseline EROD levels and measure EROD responses in
Fundulus from different geographic areas following treatment with
a known EROD inducer.
Significant biological effects (statistically different from
controls) were demonstrated from water column tests during at least
one sampling period for all stations except the Patapsco River.
The most persistent biological effects in the water column were
reported from the Wye River station as significant mortality from
two different test species was reported from both the first and
second test. Sediment tests demonstrated significant biological
effects for both tests at the Dahlgren, Morgantown, and Patapsco
River stations. Significant biological effects were reported in
sediment during the first Wye River test but not the second.
Results from suborganismal evaluations of EROD activity in
feral fish (Fundulus and spot) suggested the presence of planar
aromatic hydrocarbons in St. John's Creek and the Patapsco River.
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Fundulus collected from St. John's Creek required three to four
weeks to reach baseline EROD levels. EROD levels in Fundulu.
collected from the Wye River showed initial decreases, followed by
a rapid increase (»6 fold) from week 1 to week 2 and then returned
to collection levels during week 3. After 10 weeks, EROD activity
declined to levels similar to those in the St. John's Creek fish.
EROD activity at collection was similar in Fundulus collected from
the Wye River and St. John's Creek.
A general ranking of sensitivity among wat«r column test
species demonstrated that the sheepshead minnow larval test was the
most sensitive test. However, a specific example (Wye River Test
2) was reported where the sheepshead test did not show biological
effects but the E. affinis test did demonstrate effects. These
data emphasize the importance of multispecies testing. The
amphipod L. dytiscus was the most sensitive test species for
sediment. However, there were still cases where this amphipod did
not demonstrate effects while other species did show significant
effects. As with the water column tests, sediment toxicity data
also demonstrated the importance of multispecies testing.
Water quality and contaminant evaluations conducted in the
water and sediment during this pilot study provided supportive
information on possible causes of biological effects but these data
were not intended to provide conclusive data on specific "cause and
effect" relationships. Comparisons of biological effects data with
the limited contaminants data for water column tests in 1991
demonstrated both good and poor linkages between these types of
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data. A good correlation was reported for the first Wye River test
as significant mortality was observed for sheepshead minnow larvae
concurrently with the presence of two potentially toxic metals
(copper and nickel). The presence of a potentially toxic
Gvmnodinium dinoflagellate was also suspected as a factor
contributing to the mortality of this larval fish species. The
second Wye River test is an example of a poor correlation between
biological effects and the presence of those toxic contaminants
which were, measured. Significant mortality of E. affinis was
reported but concentrations of the toxic substances analyzed for
were not detected.
Data collected from the Patapsco River provided a strong case
for the link between biological effects in sediment and the
presence of potentially toxic concentrations of chromium, lead and
zinc. Three of the four sediment test species demonstrated
biological effects concurrently with the presence of the metals
mentioned above. Biological effects from two tests were also
reported in sediment at the Morgantown station concurrently with
the presence of potentially toxic concentrations of 4,4' - DDT.
The Dahlgren station (Test 1 and 2) and Wye River (Test 1) provided
examples of biological effects without the detection of potentially
toxic contaminant conditions in sediment.
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ACKNOWLEDGEMENTS
We would like -to acknowledge the following individuals or
organizations for assisting in this study: Pete Adolphson and Ted
Turner for technical assistance; Walter Maddox for the use of his
boat; Versar, Inc. for contaminant analysis, and Maryland
Department of Health and Mental Hygiene for triazine herbicide
analysis. Ian Hartwell, Ron Klauda, Mary Jo Garreis, Deirdre
Murphy and Rich Batiuk are acknowledged for their comments on the
study design. Special consideration is extended to Mary Hancock
for typing the report.
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TABLE OF CONTENTS
Page
Foreword 1
Abstract ii
Acknowledgements vi
Table of Contents vii
1. Introduction 1
2. Objectives^ -4
3. Methods 7
3.1 Study Areas 7
3.1.1 Potomac River - Morgantown 7
3.1.2 Potomac River - Dahlgren 9
3.1.3 Patapsco River 9
3.1.4 Wye River 10
3.2 Water Column Tests 11
3.2.1 General Description 11
3.2.2 Culture and Maintenance for Test Species 11
3.2.2.1 Mysid 11
3.2.3 Reference Toxicant Tests 12
3.2.4 Sample Collection, Handling and Storage 13
3.2.5 Test Procedures 14
3.2.5.1 Mysid 14
3.2.6 Statistical Analysis .15
3.2.7 Contaminant and Water Quality
Evaluations 16
3.3 Sediment Tests 19
3.3.1 General Description 19
3.3.2 Culture and Maintenance for Test Species 22
3.3.2.1 Hvalella azteca 22
3.3.3 Reference Toxicant Tests 25
3.3.4 Sample Collection, Handling and Storage 26
3.3.5 Test Procedures 26
3.3.5.1 Hyalella azteca 27
3.3.6 Statistical Analysis 27
3.3.7 Contaminant and Sediment Quality
Evaluations 30
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Table of Contents - continued
Page
3.4 Suborganismal Tests 35
3.4.1 General Description 35
3.4.2 Collection of Fish 36
3.4.3 EROD Evaluations in Feral Fish 37
3.4.4 Holding Time for Fundulus to Reach
Baseline EROD Levels 38
3.4.5 Evaluations of EROD in Fundulus from
Wye River versus St. John's Creek 39
3.4.6 Supplemental Studies on Copper
Concentrations in Livers of White
Perch from the Potomac River" 40
4. Results 42
4.1 Water Column Tests 42
4.1.1 Toxicity Data 42
4.1.2 Contaminants Data 49
4.1.3 Water Quality Data 53
4.1.4 Reference Toxicant Data 53
4.2 Sediment Tests 56
4.2.1 Toxicity Data 56
4.2.2 Contaminants Data 74
4.2.3 Pore Water Data 83
4.2.4 Reference Toxicant Data 83
4.3 Suborganismal Tests 86
4.3.1 EROD Activity in Feral Fish 86
4.3.1.1 Fundulus heteroclitus 86
4.3.1.2 White perch 89
4.3.1.3 Spot 90
4.3.2 Holding Time for Fundulus to Reach
Baseline EROD Levels 91
4.3.3 Evaluations of EROD in Fundulus from
Wye River versus St. John's Creek 94
4.3.4 Supplemental Studies on Copper
Concentrations in Livers of White
Perch from the Potomac River 98
5. Discussion 103
5.1 Potomac River - Morgantovn 103
5.2 Potomac River - Dahlgren 105
5.3 Patapsco River 107
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Table of Contents - continued
Page
5.4 Wye River 112
5.5 Laboratory Studies on Hepatic EROD Activity
in Fundulus 117
Conclusions After Two Years of Ambient Toxicity
Testing 119
References 139
Appendices-
Appendix A - The Percent of Females Carrying Eggs
After 8-d Exposures in the Eurytemora
affinis Tests
Appendix B - Water Quality Conditions Reported
in Test Beakers During all Water
Column Tests
Appendix C - Summary of Water Quality Monitored
During Sediment Toxicity Tests
Appendix D - Hyalella azteca Reference Sediment
Tests
Appendix E - Organics
Appendix F - Pesticides
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SECTION 1
INTRODUCTION
The Chesapeake Bay watershed is a large dynamic aquatic
ecosystem that provides habitat for numerous aquatic species. In
recent years, there has been concern for this estuary due to the
decline of various living resources such as submerged aquatic
vegetation, anadromous fish and the American oyster (Majumdar et
al., 1987). Various factors such as nutrient enrichment, fishing
pressure, pollution and diseases are often postulated as possible
causes of these declining resources. The link between contaminants
and biological effects has been of concern in critical Chesapeake
Bay habitat areas in recent years. Information derived from the
loading of toxic chemicals and/or chemical monitoring studies are
not adequate for assessing the biological effects resulting from
numerous sources such as multiple point source effluents, nonpoint
source runoff from agriculture, silviculture and urban sites,
atmospheric deposition, groundwater contamination, and release of
toxic chemicals from sediments. The most realistic approach for
evaluating the adverse effects of toxic conditions on living
resources is by direct measurement of biological responses in the
ambient environment (aquatic areas located outside of mixing zones
of point source discharges).
Various state and federal agencies in the Chesapeake Bay
watershed have made strong commitments to support research efforts
designed to address the link between contaminants and adverse
effects on living aquatic resources. For example, the Chesapeake
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Bay Basinwide Toxics Reduction Strategy has a commitment to develop
and implement a plan for Baywide assessment and monitoring of tht
effects of toxic substances, within natural habitats, on selected
commercially, recreationally and ecologically important species of
living resources (CEC, 1988a) . This commitment is consistent with
the recommendations of the Chesapeake Bay Living Resource
Monitoring Plan (CEC, 1988b).
An Ambient Toxicity Assessment Workshop was held in Annapolis,
Maryland in July of 1989 to provide a forum on how to use
biological indicators to monitor the effects of toxic contaminants
in Chesapeake Bay on living resources. Recommendations from this
workshop were used to develop a two year pilot study in 1990 and
1991. Objectives from the first year of this effort in 1990 have
been completed and a report was submitted to the sponsors (Hall et
al., 1991). Results from our first year of this two year study
demonstrated that toxic ambient conditions were present in the
Elizabeth River and Patapsco River based on water column, sediment
and suborganismal tests. Data from sediment and suborganismal
tests also suggested that toxic conditions were present at the
proposed reference site in the Wye River; water column tests did
not demonstrate the presence of toxic conditions at this reference
site. Several ambient stations in the Potomac River also had toxic
conditions based on water column and sediment tests. The need for
multispecies testing was supported by the water column tests as no
significant ranking of sensitivity among species was reported.
Results from the sediment tests showed that the amphipod test was
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most sensitive, followed by the polychaete worm test and the grass
shrimp test. The need for integrated water column, sediment and
suborganismal testing was confirmed during our first year of
testing. A spectrum of tests was needed to maximize our ability to
identify toxic conditions in the ambient environment of the
Chesapeake Bay watershed.
The purpose of this report is to present data from the second
year of testing and summarize all information collected over the
two year period. Many of the test procedures described in the
first year report were used for the second year of testing;
therefore, the first year report by Hall et al. (1991) should be
used to provide details on specific procedures. Objectives for
the second year of testing were as follows: (1) develop a survey
program to broadly assess ambient toxicity of living resource
habitats for the purpose of identifying defined regions where
ambient toxicity levels warrant further investigation of effects on
living resources; (2) assess the feasibility of such a program
through a pilot study; (3) field test existing standardized,
directly modified or recently developed water column, sediment and
suborganismal test methods for use in ambient toxicity testing and
determine the relative sensitivity of these three test methods and
(4) identify any additional test methods development or follow up
survey design testing needs (if any) to support Baywide assessment
of ambient toxicity. Data from both the first and second year of
ambient toxicity testing will be discussed in this report.
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SECTION 2
OBJECTIVES
The broad objective of this pilot study was to identify toxic
ambient areas in the Chesapeake Bay watershed by using a battery of
standardized, directly modified or recently developed water column,
sediment and suborganismal tests. Results from this study were
synthesized to identify additional test methods development or
follow up survey .design needs to "support additional Baywide
assessment of ambient toxicity. A battery of ambient tests were
conducted twice during 1991 in the Potomac River - Morgantown,
Potomac River - Dahlgren, Patapsco River and Wye River stations to
identify "short term" variability in results. The specific
objectives of the water column, sediment and suborganismal tests
are presented below.
2.1 Water Column
Objectives of the water column tests were as follows:
• determine the toxicity of ambient estuarine water twice
at the four sites by using the following estuarine tests:
•8-d sheepshead minnow, Cvprinodon variegatus, survival
and growth test; 8-d larval grass shrimp, Palaemonetes
pugio. survival and growth test; 8-d Eurvtemora affinis.
life cycle test and 8-d mysid shrimp, Mysidopsis bahia.
survival, growth and reproduction test.
• measure inorganic contaminants, organic contaminants and
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water quality conditions in ambient water concurrently
during toxicity tests to determine "possible causes" of
toxicity.
• determine the relative sensitivity of the four test
species and compare these data with the sediment and
suborganismal data.
2.2 Sediment Tests
Objectives of the sediment toxicity tests were as follows:
• evaluate the toxicity of ambient sediment twice at the
four sites by using the following estuarine tests:
20-d amphipod, Lepidactvlus dvtiscus. survival, growth,
and reburial test; 20-d grass shrimp, Palaemonetes pugio.
survival and growth test; 20-d polychaete worm,
Streblospio benedicti. survival and growth test and 20-d
amphipod, Hvalella azteca. survival and growth test.
• measure inorganic contaminants, organic contaminants and
pore water quality conditions in sediment concurrently
with toxicity testing to determine "possible causes" of
toxicity.
• determine the relative sensitivity of the four test
species and compare these data with water column and
suborganismal data.
2.3 Suborganismal Tests
Objectives of the suborganismal tests were as follows:
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evaluate EROD activity in feral fish collected adjacent
to the four sample sites.
determine the holding time required for Fundulus from St.
Johns Creek and the Wye River to reach baseline EROD
levels.
measure EROD response in Fundulus from different
geographic areas (Wye River and St. Johns Creek).
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SECTION 3
METHODS
3.1 Study Areas
Ambient toxicity tests were conducted during August and
September-October in 1991 at the Potomac River - Horgantown,
Potomac River - Dahlgren, Patapsco River and Wye River sites
(Figure 3.1). The Potomac River stations were selected because
they are located in a critical habitat area with potentially toxic
conditions. The Potomac River, a lotic system shared by three
states, also provides gradients of salinity, contamination and
habitat similar to other areas of the Chesapeake Bay watershed.
The Patapsco River was selected as a polluted area to allow field
testing of existing test methods for ambient toxicity testing.
Adverse biological effects expected at this site provided a means
to "ground truth" the test methods. The Wye River was selected to
represent a relatively "clean" reference area with no point source
pollution. However, non-point source inputs exist in this area.
Biological effects were reported from all four of these sites with
at least one test during the 1990 study (Hall et al., 1991). A
brief description of each site and rationale for selection are
included below.
3.1.1 Potomac River - Horaantown
This station was located on the Maryland side of the Potomac
River near the Harry Nice Memorial Bridge (38° 21' 48" N x 77°
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Palapsco River
Wye River
Morgantowu
Dahlgren
Chesapeake
Bay
Four stations sampled for ambient toxicity testing
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59' 30" W) . Samples were collected approximately 200 m from shore;
depth of the site was 3 m. Tributyltin (TBT) concentrations of 20
and 24 ng/L were detected at this station during a monitoring study
in 1985 and 1986 (Hall et al., 1987b). These TBT concentrations
exceeded the U.S. EPA provisional water quality criteria of 10 ng/L
(U. S. EPA, 198.7b) . There are no other background contaminants
data for this station. During our 1990 ambient testing, we
reported significant biological effects at this site with both
water column and sediment tests; suborganismal tests suggested
liver abnormalities from white perch collected from this site (Hall
et al., 1991).
3.1.2 Potomac River - Dahlgren
The Dahlgren station was located in a typical ambient
mesohaline area on the Virginia side of the Potomac River at the
mouth of Machodoc Creek near the Naval Surface Warfare Center at
Dahlgren, Virginia (38° 19' 00" N x 77° 2* 00" W) . Samples were
collected approximately 300 m from shore; depth of the site was 2
m. The Dahlgren facility has six discharge points in this area.
Results from our 1990 ambient tests at this site showed biological
effects from one water column and one sediment test (Hall et al.,
1991).
3.1.3 Patapsco River
This station was selected to allow field testing of our
methods in an area where toxic conditions were suspected. The
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station was located adjacent to the Bethlehem Steel Corporation a*-
Sparrows Point on the Patapsco River ne"ar Bear Creek (39° 14* 0,
N x 76° 29' 42" W) . .. Samples were collected approximately 75 m from
shore; depth of the site was 2 m. This site was located in a
mesohaline area outside the mixing zone of Bethlehem Steel's
largest discharge. Effluent toxicity tests conducted by the
University of Maryland's Agricultural Experiment Station have
reported chronic effects of Bethlehem Steel's effluent from both
Ceriodaphnia and fathead minnow tests (Dan Fisher,- personal
communication). Various potentially toxic contaminants such as
zinc, chromium, lead-free cyanide and tetrachloroethane have also
been reported in effluents from the Bethlehem Steel Facility.
Biological effects were reported at this site from all three types
of tests conducted in 1990 (Hall et al., 1991).
3.1.4 Wye River
The Wye River was selected as a reference area where toxic
conditions were not suspected. This station was located in a
mesohaline area at Wye Narrows above the Manor house (38° 53' 12"
N x 76° 01' 54" W) . Samples were collected approximately 75 m from
shore; depth of the site was 2 m. There are no known point sources
of contaminants in this area although agricultural activity, a
potential non-point source, exists. Results from ambient toxicity
testing in 1990 demonstrated the presence of toxic conditions from
both sediment and suborganismal tests (Hall et al., 1991).
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3.2 Water Column Tests
3.2.1 General Description
The following estuarine tests were conducted with water
collected at the Potomac River - Morgantown, Potomac River -
Dahlgren, Patapsco River and Wye River sites: 8-d sheepshead
minnow, Cyprinodon varieqatus. survival, and growth test; 8-d
larval grass shrimp, Palaemonetes pugio. survival and growth test;
8-d Eurvtemora affinis. life cycle test and 8-d mysid, Mvsidopsis
bahia. survival, growth and reproduction test. Experiments were
conducted at each station twice during 1991. Testing periods were
August 13 - 21 and September 24 - October 2. Reference toxicant
tests described later in this methods section were conducted at
least once for each test species.
3.2.2 Culture and Maintenance for Test Species
Culture and maintenance procedures for grass shrimp,
Eurvtemora and sheepshead minnows described in Hall et al. (1991)
were used. The mysid was not tested in year 1 of this study;
therefore, culture and maintenance procedures for this species are
described below.
3.2.2.1 Mvsid
Mysids used for experiment 1 were obtained from Sea
Plantations Inc. in Salem, Massachusetts. The test procedures for
this mysid shrimp test requires that tests are initiated with 7-d
old mysids. A larger than expected size range was reported in the
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shipment we received, thus suggesting that some of these shrimp
were older than 7 days. Due to these size and possible a
differences, we did not use growth or reproduction as endpoints in
this experiment. Survival was the only endpoint used and all test
organisms were randomly selected for the various test conditions.
Mysids used for experiment 2 were obtained from Chesapeake
Cultures in Hayes, Virginia. The size range for these mysids was
small suggesting that the age for all test species was 7 days.
Survival, growth and reproduction were evaluated.
3.2.3 Reference Toxicant Tests
A 48-h static reference toxicant test was conducted with each
test species in 1991. A reference toxicant is used to establish
the relative health and sensitivity of the test organisms. Cadmium
chloride was selected as the reference toxicant because we have
3 year data base with this chemical for all of the test species
except Eurvtemora. A complete description of the rationale and use
of reference toxicant tests is described in Fisher et al. 1988 and
Hall et al. 1991. Cadmium concentrations were made from a common
stock solution of cadmium chloride (Sigma Chemical). A 1.0 mg/L
cadmium stock solution was created by dissolving 407.7 mg CdCl2 in
250 mis deionized water. The stock solution was diluted to final
concentrations by pipetting stock solutions into 2 L glass beakers
containing filtered artificial seawater (15 ppt).
Survival of test organisms was monitored at each test
condition at both 24 and 48 hours (except E. affinis at 24 h). An
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LC50 value and 95% confidence intervals were calculated at 24 and
48 h using the moving-average angle and probit methods (Stephan,
1978). Temperature, salinity, dissolved oxygen and pH were
measured in each beaker at both 24 and 48 hours (except E.
affinish.
3.2.4 Sample Collection. Handling and Storage
The procedures for sample collection, handling and storage
described in Hall et al. (1991) for the first year of this study
were used. Ambient water was collected from the four sites and
transported back to our laboratory for testing at the Wye Research
and Education Center. Grab samples collected from each station
were a composite of two water column depths (" 1 m for the top and
~ 1 m from the bottom). A mid-depth sample was taken at stations
with a 2 m depth. A metering pump (12 V DC Little Giant Utility
Pump, Model PPS-12) with polyethylene line was used to collect
samples in 11.25 L glass containers.
Samples were collected on day 0, 3 and 6 during the 8-day
tests. The time lapsed from the collection of a grab sample and
the initiation or renewal of the test did not exceed 72 h. All
samples were chilled after collection and maintained at 4 C until
used. The temperature of the ambient water used for testing was 25
C. Salinity adjustments (increases) were performed on all aqueous
samples to obtain a standard test salinity of 15 ppt. These
temperature and salinity conditions were used during the first year
of testing.
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3.2.5 Test Procedures
The test procedures for grass ¦ shrimp, Eurvtemora an-
sheepshead minnows described in detail in Hall et al., (1991) were
used. The mysid shrimp was not tested in the first year of this
study; therefore, testing procedures are described below.
3.2.5.1 Mvsid
The test was initiated with 7-d old mysids exposed to the
following test conditions: Potomac River - Morgantown, Potomac
River - Dahlgren, Patapsco River, Wye River and a contaminant-free
seawater water control (15 ppt). Eight replicate 250 ml beakers,
each containing five mysids, were used for each condition. Due to
insufficient numbers of test organisms for experiment 1, we could
not use equal numbers of test species for all test conditions. The
Morgantown and Wye conditions had eight replicates while the
control and Dahlgren conditions had five replicates. Test beakers
were held in biological incubators to control temperature (25 C)
and photoperiod (16 h L: 8 h D) . A one liter sample was taken
daily from each of the field samples being stored at 4 C. Field
samples were warmed to 25 C and one-half the water volume in each
test beaker was removed and replenished with the fresh sample.
Temperature, pH, salinity and dissolved oxygen were monitored daily
in selected beakers from all test conditions. Survival of mysids
was recorded daily for eight days.
Mysids were fed twice daily (before and after water renewals)
with newly hatched Artemia nauplii. Each beaker received 0.05 ml
Artemia suspension per feeding («150 Artemia per mysid per day) for
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day 0 through day 2. On day 3 through 8, each beaker received
0.075 ml Artemia per feeding. The proportion of male, female and
immature mysids was determined in each test beaker after the test
was terminated on day 8. Twenty mysids from each test condition
were sampled for growth determination. Dry weights for pooled
samples were recorded following 24-h drying at 78 C. Mean
individual weights were calculated from the pooled samples.
Survival proportions were transformed using the arc sine
(square root (Y)) transformation. Steeles many-one rank test and
Wilcoxon rank sum test were used to compare results. Growth
parameters were compared by using ANOVA and Dunnetts tests.
Differences between means were considered significant at the p <
0.05 level.
3.2.6 Statistical Analysis
Statistical tests described in Fisher et al. (1988), Hall et
al., (1988a) and Hall et al., (1991) were used for each test
species when appropriate. The goal of this study was not to
generate typical LC50 data with various dilutions of ambient water.
For each test species response (survival, growth etc), a control
was compared with one test condition (100% ambient water). A
statistical difference between the response of a species exposed to
a control condition and an ambient condition was used to determine
toxicity. Analysis of Variance or Dunnetts Procedures were used in
cases where comparisons of a species response on a spatial or
temporal scale was needed.
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3.2.7 Contaminant and Water Quality Evaluations
A limited number of contaminants were evaluated concurrent
with toxicity tests. It was not our intention to suggest that the
analyses for inorganic and organic contaminants would provide an
absolute "cause and effect" relationship between contaminants and
biological effects if effects were reported. Information on
suspected contaminants in the study areas does, however, provide
valuable insights if high potentially toxic concentrations of
contaminants were reported in conjunction with biological effects.
Aqueous samples for analysis of organic and inorganic
contaminants listed in Table 3.1 were collected during the ambient
toxicity tests. Analytical procedures and references for analysis
of these samples are presented in Table 3.2. Analysis for triazine
herbicides (simazine, atrazine, alachlor, metachlor and cyanazine)
was conducted during this study but not in 1990. Total inorgani
contaminant analysis was conducted on filtered samples using 0.40
um polycarbonate filters.
Four liter whole water samples were collected for organic
contaminants analysis (Table 3.1). Organic contaminants other than
those identified in Table 3.1 (non-target organics) were measured
if GC/MS peaks were identified. Detailed procedures for preparing
samples for inorganic and organic analyses are described in Hall et
al. (1988a). Analysis for both organic and inorganic contaminants
was conducted at least one time on aqueous samples collected from
each station. If toxicity was found during the experiments, then
the suite of contaminants was analyzed on at least 2 separate
16
-------
Table 3.1 Concentrations of the following organic and inorganic
contaminants were evaluated in the water column.
Contaminant Detection Limit ( ug/L )
Simazine
<0.26
Atrazine
<0.32
Alachlor
<2.5
Metolachlor
<3.35
Cyanazine
<0.55
Aroclor 1248
0.050
Aroclor 1254
0.050
Aroclor 126.0
0. 050
DDE
0.02
Toxaphene
0.2
Chlordane
0. 02
Perylene
0.70
Fluorene
0.90
Phenanthene
0.70
Anthracene
0.70
Fluoranthrene
1.1
Pyrene
1.0
Benz(a)anthracene
1.7
Chrysene
0.7
Aluminum
3.0
Arsenic
3.0
Cadmium
2.0
Chromium, total
3 . 0
Copper
2 . 0
Lead
2 . 0
Mercury
0.2
Nickel
5.0
Selenium
3.0
Tin
5.0
Zinc
5.0
17
-------
Table 3.2 Analytical methods used for organic and inorganic
analysis. The following abbreviations were used: GC-I
(Gas Chromatography - Electron Capture), GC -MS (Gas
Chromatography - Mass Spectrometry), GC-NP (Gas
Chromatography - Nitrogen Phosphorous Detector), Atomic
Emission - ICP (AE-ICP), AA-H (Atomic Absorption -
Hydride), AA- F (Atomic Absorption - Furnace) and AA-DA
(Atomic Absorption - Direct Aspiration) and AA-CV
(Atomic Absorption - Cold Vapor).
Contaminant Method Method # Reference
Halogenated Hydro-
carbon Pesticides
GC-EC
608
U. S. EPA,
1984
Polychlorinated GC-EC
Biphenyls
Base-Neutral GC-MS
Extractable Organic
Compounds
Triazine Herbicides GC-NP
(Simazine, Atrazine,
Alachlor, Metachlor,
Cyanazine)
Aluminum AE-ICP
Arsenic AA-H
Cadmium AA-F
Chromium, Total AA-F
Copper AA-F
Lead AA-F
Mercury AA-CV
Nickel AA-F
Selenium AA-H
Tin AA-F
Zinc AA-DA
608
625
507
200.7
206.3
213.2
218.2
220.2
239.2
245.1
249.2
270.3
282.2
289.1
U. S. EPA,
1984
U. S. EPA,
1984
U. S. EPA,
1989
U. S.
1984
U. S.
1979
U. S.
1979
U. S.
1979
U. S.
1979
U. S.
1979
U. S.
1979
U. S.
1979
U. S.
1979
U. S.
1979
U. S.
1979
EPA,
EPA,
EPA,
EPA
EPA
EPA
EPA
EPA,
EPA,
EPA,
EPA,
18
-------
samples taken during these tests. For example, if mortality was
observed on day 5, the samples taken from day 0 and 3 were
analyzed. Versar, Inc. was responsible for all organic and
inorganic analyses except triazine herbicides, which were analyzed
by Maryland's Department of Health and Mental Hygiene.
Standard water quality conditions of temperature, salinity,
dissolved oxygen, pH and conductivity were evaluated at each site
after sample collection. These parameters were also evaluated
every 24 h at all test conditions.
3.3 Sediment Tests
3.3.1 General Description
All tests and analyses were conducted according to the
Standard Operating Procedures and Quality Assurance plans
previously submitted to the sponsors. The methods described in
this report are general summaries of those protocols. Sediment
samples (100% ambient sediment samples) from four stations were
tested using three field collected species: the grass shrimp
Palaemonetes puqio. the amphipod Lepidactvlus dvtiscus. the
polychaete worm Streblospio benedicti (hereafter referred to as
worm), and one laboratory cultured species, the amphipod Hvallela
azteca. All tests were conducted for 10 days at 25C and monitored
daily. At the end of 10 days, mortalities were recorded, and the
animals were returned to the original test containers. The
organisms were then monitored daily for an additional 10 days.
Numbers of live animals were recorded on day 20, and any living
19
-------
organisms were preserved for length/ weight measurements. The
sediment samples were collected from one site in the Patapst
River, one site in the Wye River and two sites in the Potomac River
(Morgantown and Dahlgren) . The salinity at all sites was 10-12
parts per thousand (ppt) at sampling, except for the Patapsco River
site, which was 9 ppt at sampling. All samples were adjusted to 15
ppt prior to testing by sieving with 15 ppt control water made from
deionized water and artificial sea salts. Control sediments for
each -species consisted of native sediments from the area.in which
the test organisms were collected or sediments in which they were
cultured. Control and reference sediments (see below) were tested
with each set of experiments.
Sediment for performing particle size analysis was collected
from each of the test stations several weeks prior to initiation of
the toxicity tests, in order to select a reference sediment for
each set of test samples. Control sediment tests were performed to
assess inherent species effects (mortality, growth, etc.) in native
sediment. Uncontaminated reference sediment tests were performed
to assess particle size effects. It was determined from the
initial sediment collection and particle analysis that all test
sites were greater than or equal to 72% sand (see Table 3.3). The
sediment sample from Dahlgren was 72% sand; Wye River sample 79%
sand, Morgantown 89% sand, and Patapsco 98.6% sand. The initial
particle analysis for S. benedicti control sediment from Lynnhaven
Inlet indicated that the sediment was 68% sand. The sediment used
as the control for L. dvtiscus from the Lynnhaven Beach site was
20
-------
Table 3.3 Initial particle size analysis of sediments from fotir
stations and controls used in toxicity tests. Morgantown
and Dahlgren are Potomac River stations. Samples were
collected 8/1/91.
Site
% Sand
% Silt
% Clav
Morgantown
89.0
6.3
4.7
Dahlgren
72.0
12.7
15.3
Patapsco River
98.6
0.9
0.5
Wve River
79.2
12.9
7.9
Lynnhaven Beach
100
0
0
Lynnhaven Inlet
67.8
23.8
8.4
21
-------
100% sand. These two control sediments bracket the particle size
of all test samples and were therefore considered suitable a:
reference sediments as well. Therefore, the Lynnhaven Inlet
control sediment for S. benedicti and H. azteca was used as the
reference sediment for the p. pugio and L. dytiscus and the
Lynnhaven Beach control sediment for £. pugio and L. dytiscus was
used as the reference sediment for the S. benedicti and H. azteca
(see Tables 3.4 and 3.5). The actual test sediment samples were
collected and again analyzed for sand, silt,, and clay content.
Although the particle size/composition of the test sediments were
slightly different from the initial samples (see Table 3.4), all
samples were still primarily sand, with the lowest concentration
occurring at the Wye River site (82.2% sand).
3.3.2 Culture and Maintenance of Test Species
Culture and maintenance procedures used for grass shrimp,
worms and the amphipod Lepidactvlus dytiscus are as described in
Hall et al. (1991) . Hvalella azteca was not used in year 1 of this
study, therefore, culture and maintenance procedures for this
organism are described below.
3.3.2.1 Hvalella azteca
Hvalella azteca was cultured at the Applied Marine Research
Laboratory for approximately a year and a half prior to the
experiments. Animals are cultured in 20 gallon tanks filled with
freshwater (calcium, magnesium, potassium and sodium salts added to
deionized water (ASTM, 1990)) in a 16 Light:8 Dark cycle. The tanks
22
-------
Table 3.4
Site
Particle size analysis of sediments from four
stations and controls used in toxicity tests.
Morgantown and Dahlgren are Potomac River stations,
Set 01 was collected 8/14/91,
collected 9/25/91,
% Sand
% Silt
Set #2 was
% Clay
Set #1:
Morgantown
Dahlgren
Patapsco River
Wye River
Lynnhaven Beach
Lynnhaven Inlet
Set #2:
Morgantown
Dahlgren
Patapsco River
Wye River
Lynnhaven Beach
Lynnhaven Inlet
94.0
91.2
98.8
93.7
100
68.6
94,
88,
98.
82.
100
63,
2.9
6.2
0.8
3.9
0
25.4
2.7
4.8
0,
9,
0
26,
3,
2,
0,
2,
0
6.0
1
6
4
4
2.8
6.7
0.7
8.8
0
10.1
23
-------
Table 3.5 Control and reference test sediment
allocation by species.
Sedimfent Sediment Sediment
Site Description Function
SPECIES
dvtiscus
Lynnhaven
Lynnhaven
Beach
Inlet
100% Sand
65-68% Sand
Control
Reference
P.
Duaio
Lynnhaven
Lynnhaven
Beach
Inlet
100% Sand
65-68% Sand
Control
Reference
H.
azteca
Lynnhaven
Lynnhaven
Beach
Inlet
100% Sand
65-68% Sand
Reference
Control
benedicti
Lynnhaven
Lynnhaven
Beach
Inlet
100% Sand
65-68% Sand
Reference
Control
24
-------
are gently aerated, and approximately 30 percent of the overlying
water is replaced weekly. Dried maple leaves, presoaked and rinsed
in deionized water, are added weekly as the primary substrate and
food. Ground rabbit chow is added twice a week as a supplementary
food. Two Hvalella azteca culture tanks were acclimated to 15 ppt
over a seven day period. . Animals were held at 15 ppt for an
additional seven days prior to initiation of the tests. Test
organisms were obtained from the population tanks by sieving with
a 710 micron sieve stacked on a 425 micron sieve. Those animals
that passed through the 710 sieve, but were retained on the 425
micron sieve were used as test organisms. Juvenile Hvalella azteca
were sieved from the population tank no more than 24 hours prior to
initiation of the toxicity test.
3.3.3 Reference Toxicant Tests
The relative sensitivities of each set of test organisms were
evaluated by performing reference toxicant tests. These tests were
designed to assess changes in toxicity tolerances of the organisms
that may have occurred due to disease, or stress resulting from
handling and acclimation. The same reference toxicant test methods
and procedures were employed in the 1991 study (see Hall et al.
1991 for details) that were used in the 1990 project.
Hyalella azteca was tested in a 96-hour, water only, cadmium
chloride reference toxicant test using 200 ml of solution with 10
animals per container. Five concentrations and a control
(artificial seawater only) were prepared by serial dilution. Tests
25
-------
were conducted at 15 ppt salinity with two replicates for each
concentration. A data base of Hvalella azteca reference toxicc
data exists at the AKRL laboratory for comparison.
Survival of test organisms was monitored every 24 hours. A
96-h LC50 value and 95% confidence intervals were calculated using
Probit Analysis or the Moving Average Angle Method, as was
appropriate for each set of data.
3.3.4 Sample Collection. Handling and Storage
Sediment sample collection, handling, and storage procedures
described in Hall et al. 1991 were used in this study. Sediment
samples were collected at each site by AMRL personnel and returned
to the laboratory for testing. The first set of sediments was
collected August 14, 1991, by petite ponar grab. The second set of
sediment samples was collected by petite ponar grab on Septembt
25, 1991 at the same sites.
Approximately 8 gallons of sediment were collected at each
site, homogenized and distributed to sample containers. All samples
were transported on ice, out of direct sunlight. Samples were held
in refrigerators at 4C until initiation of toxicity tests. Samples
for chemical analysis were frozen and stored until tested. All
samples were analyzed within EPA recommended holding times.
3.3.5 Test Procedures
The test procedures for Palaemonetes puqio. Streblospio
benedicti, and Lepidactvlus dytiscus described in detail in Hall et
26
-------
al. (1991) were used. Hyalella azteca was not tested in the first
year of this study, therefore, these procedures are described
below.
3.3.5.1 Hvalella azteca
A series of test containers was set up according to the
methods outlined in the ASTM "New Standard Guide for Conducting
Sediment Toxicity Tests with Freshwater Invertebrates" (ASTM,
1990). Two centimeters of sediment were placed into each of five
replicate 1 liter test containers with 700 ml of overlying water.
Twenty animals were added to each test vessel and monitored for 10
days. Control sediment consisted of Lynnhaven Inlet muddy sand.
A subset of the test animal population was selected for initial
length/weight measurements. All length measurements were conducted
using the Optimas Image Analysis system. Test containers were
monitored daily for oxygen, temperature, and pH. Number of animals
emerged from the sediment was also recorded. The amphipods were
fed 14 mg of ground rabbit chow per test container every three days
throughout the duration of the test. At the end of ten days,
animals were sieved from test containers and mortality was
recorded. Surviving animals were then returned to the test
containers for an additional 10-day period. On day twenty, animals
were again sieved from the containers and mortality recorded. Live
animals were preserved for growth measurements.
3.3.6 Statistical Analysis
The goal of this study was not to generate LC50 data from
27
-------
dilution series tests. The main objective was to evaluate for each
test species, the response (mortality, growth, etc) when tested in
100% ambient sediment, as compared to a control. Statistica.
differences between the responses of species exposed to control and
ambient sediments were used to determine the toxicity. Evaluations
relative to particle size effects were made based on the response
seen in the reference sediments.
The same statistical approaches that were employed in the
first year of the study (Hall et al.^1991) were utilized in the
second year. Basically, the analyses consisted of analysis of
variance (ANOVA) models with a. priori tests of each treatment
contrasted to the controls. Arcsine transformations were used for
the percent mortality data. For growth data comparisons, the
length and weight measurements were tested in a multivariate
analysis of variance design (MANOVA) to remove the effect of inter-
variance correlation on the statistical comparisons. The length
measurements at day 20 were also compared to the initial (pre-
experimental) lengths by the ANOVA/contrast tests.
During the Set 1 experiments, it became apparent that
Streblospio benedicti displayed mortalities if placed in very sandy
sediments (> 90% sand) without supplementary feeding (Figure 3.2).
A feeding rate for Set 2 experiments of 0.25 mg/replicate offered
every third day was employed to limit this mortality. This "worm
food" consisted to dried ground Ulva spp. and finely ground (<500
um) Tetramin in a 50:50 ratio. Supplementary experiments with
clean sediments of different particle sizes demonstrated a
28
-------
Fig. 3.2 Survival of Streblfflypio benedicti versus percent
silt/clay.
1201
hO
VO
100
i 80
& 60
in
40 -
20
0
0
10
15
20
25
30
% SILT/CLAY
-------
regression relationship between percent survival (at day 10) and
percent sand:
% survival = 135.2 - 0.60627 x % sand
This relationship was employed to determine the portion of
mortalities observed during the Set 1 experiments which were
associated with particle size (and starvation) effects. These
predicted mortalities were taken into account mathematically before
treatment effects were statistically evaluated. This sort of
procedure has been used for "adjusting" for particle-size effects
in the amphipod Rhepoxynius abronius (DeWitt, 1988).
Much of the particle-size mortalities observed during Set 1
were thus eliminated by supplementary feeding in Set 2. Additional
studies with this species have suggested an optimal feeding regime
which will be employed in the year 3 studies.
3.3.7 Contaminant and Sediment Quality Evaluations
Contaminants were evaluated concurrently with toxicity tests.
It was not our intention to suggest that the presence of inorganic
and organic contaminants demonstrate an absolute "cause and effect"
relationship between contaminants and any observed biological
effects. Information on suspected contaminants does however,
provide valuable insights, if high concentrations of potentially
toxic contaminants were reported in conjunction with biological
effects.
Sediment samples for organic contaminants analysis were
collected in conjunction with bioassay sediment samples. The
30
-------
contaminants assayed are listed in Tables 3.6 and 3.7. Organic
analytical procedures used were in accordance with USEPA methods
3550 and 8270 (USEPA, 1986) and are detailed in Hall et al. (1991).
Samples were analyzed for organochlorine pesticides (OCP) as veil
as polychlorinated biphenyls (PCBs) in accordance with USEPA
Methods 3550 and 8080 (Table 3.7). Atrazine was not analyzed in
year l, but was included in the sediment analysis for year 2. The
methods for atrazine detection are the same as those listed above.
Organic analysis was conducted at least once per site. The Wye
River sediments from both samplings (Set 1 and Set 2) were analyzed
for organic contaminants.
All sediment samples were analyzed for Acid Volatile Sulfides
(AVS) and Total Organic Carbon (TOC). Samples were frozen until
analysis, at which time they were thawed and homogenized by gently
stirring. Sediment samples were analyzed in duplicate for AVS
using the method of DiToro et al. , (1990). Details of the
analytical procedures for both AVS and TOC are described in Hall et
al., 1991.
Pore water samples were removed from all sediment samples by
squeezing with a nitrogen press. All pore water samples were
filtered then frozen until analyses of ammonia, nitrite and
sulfides were conducted. These analyses were conducted on all
samples. Details of the methods are described in Hall et al.,
1991.
All sediment samples were analyzed for the following bulk
metals: aluminum, cadmium, chromium, copper, lead, nickel, tin and
31
-------
Table 3.6 Semi-volatile organic compounds analyzed, utilizing
user-created calibration library. Sediment method
detection limits (MDL) are reported in /ig/kg dry weight.
CAS NO.
COMPOUND SEDIMENT ]
65-53-3
Aniline
14.5
95-57-8
2-chlorophenol
13.2
111-44-4
Bis(2-chloroethyl)ether
11.2
108-95-2
Phenol
11.9
541-73-1
1,3-dichlorobenzene
11.9
106-46-7
1,4-dichlorobenzene
12.5
95-50-1
1,2-dichlorobenzene
11.9
100-51-6
Benzyl alcohol
27.1
39638-32-9
Bis(2-chloroisopropyl)ether
-5.9
95-48-7
2-methylphenol
15. 8
91-57-6
2-methylnaphthalene
9.2
67-72-1
Hexachloroethane
21.8
621-64-7
n-nitroso-di-n-propylamine
13.2
106-44-5
4-methylphenol
13.9
98-95-3
Nitrobenzene
11.2
78-59-1
Isophorone
6.6
88-75-7
2-nitrophenol
27. 1
65-85-0
Benzoic acid
18. 5
105-67-9
2,4-dimethylphenol
15.8
111-91-1
Bis (2-chloroethoxy)methane
9.9
120-83-2
2,4-dichlorophenol
21.8
120-82-1
1,2,4-trichlorobenzene
15.2
91-20-3
Naphthalene
4.6
106-47-8
4-chloroaniline
26. 4
87-68-3
Hexachlorobutadiene
22.4
59-50-7
4-chloro-3-methylphenol
20.5
77-47-4
Hexachlorocyclopentadiene
25.7
88-06-2
2,4,6-trichlorophenol
37.0
95-95-4
2,4,5-trichlorophenol
44 . 9
88-74-4
2-nitroaniline
37.6
91-58-7
2-chloronaphthalene
9.9
208-96-B
Acenaphthalene
5.9
84-66-2
Dimethylphthalate
9.9
606-20-2
2,6-dinitrotoluene
48.2
99-09-2
3-nitroaniline
247
83-32-9
Acenaphthene
9.9
51-28-5
2,4-dinitrophenol
262
132-64-5
Dibenzofuran
7.9
100-02-7
4-nitrophenol
268
121-14-2
2,4-dinitrophenol
43.6
86-73-7
Fluorene
9.9
7005-72-3
4-chlorophenylphenylether
20.5
84-66-2
Diethylphthalate
9.9
100-01-6
4-nitroaniline
279
534-52-1
4,6,-dinitro-2-methylphenol
122
32
-------
Table 3.6 (Continued)
CAS NO.
COMPOUND
SEDIMENT 1
86-30-6
n-nitrosodiphenylamine
19.1
101-55-3
4-bromophenylphenylether
41.6
85-01-8
Phenanthrene
9.2
118-74-1
Hexachlorobenzene
37 . 6
87-86-5
Pentachlorophenol
136
120-12-7
Anthracene
9.9
84-74-2
Di-n-butylphthalate
5.9
206-44-0
Fluoranthene
10.6
129-00-0
Pyrene
10. 6
85-68-7
Butylbenzylphthalate
17.8
56-55-3
Benzo(a)anthracene
17.8
218-01-9
Chrysene
14 .5
91-94-1
3,3' -dichlorobenziciine
101
117-81-7
Bis(2-ethylhexyl)phthalate
12.5
117-84-0
Di-n-octylphthalate
7 . 3
205-99-2
Benzo(b)fluoranthene
13.9
207-08-9
Benzo(k)fluoranthene
13.9
50-32-8
Benzo(a)pyrene
15. 2
193-39-5
Indeno(1,2,3-cd)pyrene
16.5
53-70-3
Dibenz(a,h)anthracene
17.8
191-24-2
Benzo(ghi)perylene
16.5
103-33-3
Azobenzene
7 . 3
92-87-5
Benzidine
24 .4
33
-------
Table 3.7 Method detection limits for organochlorine pesticides
and PCBs. Detection limits for sediment are reported i-
jig/kg dry weight.
CAS NO.
COMPOUND
SEDIMENT MDL
391-84-6
q-BHC
0.714
301-85-7
0-BHC
0.559
391-86-8
6-BHC
1.062
58-89-9
Lindane
0.616
76-44-8
Heptachlor
0.819
309-00-2
Aldrin
0. 608
1024-57-3
Heptachlor epoxide
0.570
959-98-8
Endosulfan I
0.859
60-57-1
Dieldrin
0.898
72-55-9
4,4'-DDE
0.528
33213-65-9
Endosulfan II
0.74 5
72-20-8
Endrin
1.240
72-54-8
4,4'-DDD
0.469
1031-07-8
Endosulfan sulfate
1.500
50-29-3
4,4'-DDT
3.420
72-43-5
Methoxychlor
5.0
57-74-5
Chlordane
5.0
80001-35-2
Toxaphene
10.0
2385-85-5
Mirex
1. 000
7421-93-4
Endrin aldehyde
2 . 410
12574-11-2
Aroclor 1016
16. 6
11104-28-2
Aroclor 1221
16. 6
11141-16-5
Aroclor 1232
16.6
53469-21-9
Aroclor 1242
16.6
12672-29-6
Aroclor 1248
16.6
11097-69-1
Aroclor 1254
16.6
11096-82-5
Aroclor 1260
16.6
34
-------
zinc, using an inductively coupled plasma atomic emission
spectrometer (ICP) following USEPA/SW-846, Method 6010 (see Hall et
al., 1991). Detection limits are given in Table 4.18. Arsenic and
selenium were predigested using EPA SW-846, Method 3050, and
analyzed by hydride generation following Standard Methods for the
Examination of Water and Wastewater (1989), while mercury was
analyzed using EPA SW-846, Method 7471. In addition, a
Simultaneously Extractable Metals (SEM) analysis was conducted on
all samples to use with the AVS data to determine the potential
toxicity of the sediment due to metals. The sample for the SEM
analysis was obtained from a step in the AVS procedure. The AVS
method was detailed in Hall et al. 1991. The SEM sample was the
sediment suspension remaining in the generation flask after the
cold acid extraction was completed. The sediment suspension was
filtered through a 0.2 micron membrane filter into a 250 ml
volumetric flask. The sample was then diluted to volume with
deionized water. The concentrations of the SEM were determined by
EPA-600/4-79-020 Methods for Chemical Analysis of Water and Wastes.
Cadmium, lead, copper, nickel, and zinc were determined by ICP
following USEPA Method No. 200.7. Mercury was determined by cold
vapor generation following USEPA Method No. 245.1. The
concentrations were then converted to micromoles per gram dry
sediment and were added together to provide a total SEM estimate.
3.4 Suborganismal Tests
3.4.1. General Description
35
-------
Based on the experience from Year 1 testing, the scope of tf
Year 2 effort was refined to include a reduced subset of tasK.
which were consistent with data production rates of the other tasks
for the pilot testing program. There were three broad goals for
Year 2: (1) Continuation of the evaluation of Ethoxyresorufin-O-
deethylase (EROD) activity in feral fish collected in proximity of
the test stations; (2) Determination of the holding time necessary
to reduce EROD activity in feral Fundulus heteroclitus collected
from the Wye River station and from the St. John's Creek (Patuxent
River tributary) reference population used for Year 1 exposures to
test sediment and water; and (3) To determine differences (if any)
in the induction of EROD activity in the two populations of
Fundulus following exposure to a model EROD inducer
(Benzo(a)pyrene).
3.4.2. Collection of Fish
Fundulus heteroclitus were collected from the Wye River and
St. John' s Creek (a tributary of the Patuxent River ca. 1 mile from
Solomons, MD) . Fish were collected using standard minnow traps.
Collections of Wye River fish were conducted on August 8, 1991,
while St. John's Creek fish were collected on July 25, 1991.
Spot, Leiostomus xanthurus. were collected from the vicinity
of the Patapsco River Station via hook and line on August 15 and
September 26, 1991. We surveyed the shoreline areas of Bear Creek
up to the vicinity of the railway bridge looking for Fundulus
and/or suitable Fundulus habitat. We set traps at locations we
36
-------
felt may hold Fundulus. but were unable to collect any fish from
any locations. We did not attempt to find fish from areas further
upstream due to the lack of any proximity to the sampling station.
White perch, Morone americana. were collected from the
Dahlgren and Morgantown, Potomac River stations on August 16 and
September 27, 1991. Fish were collected via a standard beach
seine. As in 1990, we seined widely along these beaches in an
effort to collect Fundulus. but were unable to do so. We did not
set minnow traps in these areas in 1991. As with the Patapsco
River area, there was not suitable Fundulus habitat proximate to
the sampling stations.
3.4.3 EROD Evaluations in Feral Fish
As fish were collected, they were placed in containers filled
with site water and held until processing. During processing,
livers from individual fish were removed and immediately frozen in
liquid nitrogen. The remainder of the fish was packaged in whirl-
pak bags and transported to the laboratory on ice for measurements
of body size (wt) , sex determination and evaluation of
gonadosomatic indices (gonad wt./body weight).
In the laboratory, livers were thawed and microsomes were
prepared using methods described by Stegeman et al. (1987).
Briefly, these methods involve homogenizing the liver in buffer and
using differential ultracentrifugation to isolate the microsomal
(endoplasmic reticulum) fraction. Microsomes were then stored
frozen in liquid nitrogen until biochemical assays were performed.
37
-------
Ethoxyresorufin-O-deethylase (EROD) assays were conducted
using the fluorometric technique described by Burke and Maye
(1974). This enzymatic activity is catalyzed by the cytochrome
P450IA isoform in fishes (Stegeman, 1989) and is the activity
induced by aromatic hydrocarbon exposure.
In Year Two, EROD evaluations were performed on Fundulus from
the Wye River and St. John's Creek. EROD activity was also
measured in white perch from the Dahlgren and Morgantown stations
on the Potomac River. Finally, EROD was measured in spot from the
Patapsco River station and from a population (presumed to be
reference) sampled in the Patuxent River near Sheridan Point.
3.4.4 Holding Time for Fundulus to Reach Baseline EROD Levels
In Year 1 we discovered that Fundulus sampled from various
locations had elevated EROD activity in comparison to fish that
were used for sediment and water column exposures. The fish that
had been used for the lab exposures were from a reference
population of Fundulus that had been sampled from St. John's Creek
in the Patuxent River and held in the lab for several months prior
to their use in the sediment and water exposures. This holding
time was used originally to allow for the possibility that these
reference fish could in fact have elevated EROD activity caused by
the presence of aromatic hydrocarbon inducing substances present,
but undocumented, in the sampling area.
The results from Year 1, where the lab held fish had low EROD
activity as compared to all freshly sampled fish (from the Wye
38
-------
River as well as the more highly contaminated Elizabeth and
Patapsco Rivers), indeed suggested that the lab holding time had
reduced activity as compared to freshly sampled fish.
In order to more thoroughly document the decline in EROD
activity following sampling, we sampled groups of Fundulus from the
Wye River and from St. John's Creek. A sample of fish was
dissected on the day of collection and livers were frozen and
subsequently processed for EROD measurements. The rest of the fish
were held in separate tanks at the Chesapeake Biological Laboratory
in flowing ambient, unfiltered, Patuxent River water (temp = 28-31
C and salinity = 14-16 ppt) . Fish were fed Tetramin® fish food
flakes once a day.
Additional samplings of fish were conducted on weeks 1, 2
3, 10 (Wye River) and 12 (St. John's Creek) post collection. In
all cases, approximately 12 fish were sacrificed and mircrosomes
prepared from pooled (same sex) liver samples from 2-3 fish.
Samples sizes (n) varied from 4 to 7. Though additional discussion
will be found below, it is important to note that these experiments
were offset by a two week period and thus the actual dates which
correspond to the sampling weeks are not the same. Wye River fish
experiments were initiated on August 8, 1991, while St. John's
Creek experiments were initiated on July 25, 1991 (two weeks
earlier).
3.4.5 Evaluation of EROD Induction in Fundulus from the Wye
River Versus St. John' s Creek.
39
-------
Fundulus from the two separate populations were held through
the winter in flowing ambient Patuxent River water. Approximate,
one month prior to the initiation of the induction experiments,
fish were slowly acclimated to flowing, filtered, 20C, Patuxent
River water (salinity = 14-15 ppt)..
To test whether the two different populations would respond
similarly to a known aromatic hydrocarbon inducer, fish were
injected intraperitoneally with corn oil solutions of
benzo(a)pyrene (>99% pure, Aldrich). Fish were given 2 "ul
injections/g fish, equivalent to doses of 0, 2 and 10 mg B(a)P/kg.
Fish were held following treatment in 38 liter glass aquaria with
recirculating, charcoal filtered, Patuxent River water (temp = 21
C and salinity = 14-15 ppt) . Ten fish were injected in each
treatment group. Liver samples were pooled (by sex) at sacrifice
to yield 5 samples per treatment group (i.e., n = 5 for al!
groups). Fish were sacrificed 96 hours following treatment. This
time has been shown in previous investigations (Kloepper-Sams,
1989) to produce strong induction (6-fold) in Fundulus
heteroclitus. Fish were sacrificed and livers removed and frozen
in liquid nitrogen in the same fashion as field samples.
3.4.6 Supplemental Studies on Copper Concentrations in Livers
of White Perch .from the Potomac River
As described subjectively in the Year 1 report, we observed
that the liver in many of the white perch sampled from the Potomac
River were exceptionally dark in color (like an old penny) . We
40
-------
suggested that this observation was consistent with previous
studies on white perch which have described a propensity for white
perch to store copper at high levels in their livers (Bunton et
al., 1987; Bunton and Baksi, 1988).
In order to corroborate these findings, we subsampled livers
from a selection of white perch from the Dahlgren and Morgantown
stations. Samples of liver were preserved in formalin in the
field. During sampling, we made no formal attempt at random
sampling. In fact, we purposely preserved some livers Vhich
appeared to be less affected (i.e., not dark brown) in order to get
a range of concentrations.
Small pieces of liver were digested using concentrated nitric
acid and hydrogen peroxide. This method is essentially that
described by Roesijadi (1980) with the addition of hydrogen
peroxide for final digestion and color removal. Levels of copper
in the digestates were determined by flame atomic absorption
spectrophotometry using external standards. Formalin buffer blanks
were also analyzed. All concentrations are reported as wet weight
values.
41
-------
SECTION 4
RESULTS
4.1 Water Column Tests
The following results from water column tests are presented
below: toxicity data, contaminants data, water quality data and
toxicity data from reference toxicant tests.
-4.1.1 Toxicity Data
Survival, growth and reproduction data from the four estuarine
tests conducted from 8/14/91 to 8/21/91 in the Wye River, Patapsco
River, Potomac River - Morgantown and Potomac River - Dahlgren are
presented in Tables 4.1 and 4.2. Survival and reproduction data
from the Eurvtemora tests did not suggest the presence of toxic
conditions (statistical difference between control and test
condition) at any of the ambient stations although percent survival
at the Wye River station was 23% lower than the controls. Mean
brood size from the controls (ground water with sea salt) was
significantly lower than all other stations except the Patapsco
River. We encountered this problem periodically during the 1990
studies and attributed the low reproduction in the controls to
reduced levels of phytoplankton and nutrients. Both of these
constituents are found naturally in the ambient water in addition
to the concentrations added during feeding. During the second test
discussed later in this section, we used another source of control
water from a pristine area in the Indian River in Delaware in
42
-------
Table 4.1 Sn; ival data from E. affinis. sheepshead minnow larvae, grass shrimp and mysid
si' -p 8-d tests conducted in the Wye River, Patapsco River, Potomac River-
M'.m ;.\ntown and Potomac River-Dahlgren from 8/13/91 - 8/21/91.
Species
Station
1
2
Cumulative %
3 4
Survival
5
Per Day
6
7
8
Eurvtemora
Control
90
Wye
67
Patapsco
97
Morgantown
87
Dahlgren
83
Sheepshead
Control
100
100
100
100
100
100
100
100
minnows
Wye**
38
35
33
33
33
33
30
25*
Patapsco
100
100
100
100
100
100
100
100
Morgantown
100
98
98
98
98
98
95
88
Dahlgren
100
100
100
100
100
100
100
100
Grass
Control
100
100
100
98
93
78
68
60
shrimp
Wye
100
100
90
88
88
88
85
85
Patapsco
100
100
100
98
95
93
88
85
Morgantown
100
100
100
100
95
83
75
70
Dahlgren
100
98
93
93
90
85
78
78
Mysid
Control
96
93
93
89
89
85
85
82
shrimp
Wye
95
90
90
90
88
88
88
88
Patapsco
100
100
93
85
85
85
85
85
Morgantown
93
86
86
86
86
83
83
83
Dahlgren
88
80
80
80
80
76
76
76
* Indicates value is significantly different than the control, p <0.05. Kruskall-Wallis
Rank Test.
** A second set of beakers were set up with 10 larvae each with water collected on 8-14-91.
Survival was >95% after 8 d.
-------
Table 4.2 Reproduction (brood size) and growth data from E.
aff inis. grass shrimp larvae and sheepshead rcinnr
larvae 8-d tests conducted at the four stations fr
8/14/91 - 8/21/91.
E.affinis brood size comnarisions following 8-d exposures
Station
n
Mean Nauplii Produced
S.E.
Control
5
2.6
2.4
Wye
5
51.8*
12.7
Patapsco
5
27.4
7.2
Morgantown
5
47 .8*
13 .7
-Dahlgren
5
54.2*
10.3-
* Indicates value is significantly different than the controls
using a Dunnett's Test, p<.05.
Grass shrimp
and Sheepshead minnow crrowth comparisons
Grass Shrimp
Drv Weiqht (Mean
weight at day 0 =
0.45 ;
mg)
Station
n
Mean (ma)
S.E.
Control
20
0.70
0. 02
Wye
20
0.82
'0.05
Patapsco
20
0.74
0.03
Morgantown
20
0.81
0. 03
Dahlgren
20
0.76
0.04
Sheepshead Larvae Drv Weiaht
(Mean weight at day
0 = 1
0.03 mg)
Station
n
Mean fmai
S.E.
Control
25
0.82
0. 07
Wye
10**
0.91
0.13
Patapsco
25
0.81
0.04
Morgantown
25
0.55*
0.06
Dahlgren
25
0.51*
0.05
* Indicates value is significantly different than control,
Dunnett's Test or Bonferroni's T-test, p<0.05.
** Larvae from the 8/13/91 test were used.
hU
-------
addition to our initial control water. Brood size of this copepod
increased significantly in the Indian River control water. The
percent of females carrying eggs after 8-d exposures in the E.
affinis tests are presented in Appendix A. There were no
statistical differences between the controls and any test
condition.
Survival of sheepshead minnow larvae in the Wye River was
significantly lower (25% after 8 days) when compared with the
controls or other stations. Most of the mortality occurred within
24 h thus indicating a toxic condition in water collected on 8-13-
91. A second set of beakers with 10 larvae each was set up using
Wye River water collected one day later (8-14-91) to determine if
the toxic conditions were persistent. Survival was greater than
95% in these two beakers after 8 day exposures. The growth of
surviving larvae at the Wye Station (8-13-91) was not lower than
the controls or other test conditions. Growth of sheepshead minnow
larvae at both the Horgantown and Dahlgren stations was
significantly lower than the controls or the other stations.
Survival and growth of grass shrimp larvae and survival of
mysids was not significantly reduced in ambient water from any
station when compared with the controls. Survival for grass shrimp
in the controls was lower than the ambient water conditions. Due
to problems with different age mysids, we could not conduct growth
reproduction evaluations with this test species during the first
experiment (see Section 3.2.2.1).
Survival, growth and reproduction data from the second set of
45
-------
experiments conducted from 9/24/91 to 10/2/91 are presented i"
Tables 4.3 and 4.4. A significant reduction in survival w
reported for Eurvtemora exposed to Wye River water vhen compared
with the H-W controls. Survival of Eurvtemora at the Patapsco
River station was also 36% lower than controls; however, this
difference was not reported to be significant. There was no
statistical difference between the number of egg carrying females
after 8 d of exposure in the controls (H-W and I-R controls) when
compared with any test condition (Appendix A).
Survival and growth of both sheepshead minnow larvae and grass
shrimp larvae were not significantly lower in any of the ambient
test conditions when compared with the controls. Mysid shrimp
survival was not significantly reduced at any of the ambient
conditions when compared with the controls. However, growth of
mysids in both the Wye and Patapsco River stations was
significantly increased when compared with the controls. The
reproductive endpoint used for the mysid shrimp test was the
proportion of females without eggs in the control compared with the
test conditions. All surviving females had eggs; therefore, we
detected no effects using this endpoint. In recent years, various
investigators have used the proportion of immature mysids as an
endpoint to identify stressful conditions (Susan Lussier, personal
communication, U.S. Environmental Protection Agency; Anwar et al.,
in press). Using this endpoint, we reported a significant
reduction in the proportion of immature mysids (and increase in
mature mysids) at the Patapsco River station when compared with the
46
-------
Table 4.3 Survival data from E. affinis. sheepshead minnow larvae, grass shrimp and mysid
shrimp 8-d tests conducted in the Wye River, Patapsco River, Potomac River-
Morgantown and Potomac River-Dahlgren from 9/24/91 - 10/2/91. E. affinis control
water was H-W {groundwater and seasalt) and I-R (Indian River, Delaware).
Species
Station
Cumlative % Survival Per Day
3 4 5 6
8
Eurvtemora
H-W Control
I-R Control
Wye
Patapsco
Morgantown
Dahlgren
93
90
32*
57
96
93
4S
Sheepshead
larvae
Grass
shrimp
Mysid
shrimp
Control
Wye
Patapsco
Morgantown
Dahlgren
Control
Wye
Patapsco
Morgantown
Dahlgren
Control
Wye
Patapsco
Morgantown
Dahlgren
100
100
100
100
100
100
100
98
100
100
100
100,
100
100
100
100
100
100
100
100 '
100
100
100
93
98
98
98
95
95
95
95
100
100
100
100
100
100
100
100
100
100
100
100
98
93
90
88
100
100
98
95
95
95
95
95
100
98
98
98.
98
85
85
83
100
100
100
100
100
98
98
95
85
85
85
85
85
85
83
83
100
100
100
100 .
100
100
100
98
100
100
100
100
100
100
100
100
100
100
100
98
98
98
98
98
100
100
100
100
100
100
100
100
100
98
98
98
98
98
98
98
* Indicates value is significantly different than the H-W control> Bonferroni T-Test, p<0.05.
For the I-R control, p is not exactly <0.05 as the T-statistic = 2.67 and the critical
value is 2.69. Therefore, p is just slightly greater than 0.05.
-------
Table 4.4 Reproduction and growth data from E. affinis. gras<-
shrimp larvae, sheepshead minnow larvae and mysid shrix
8-d tests conducted at the four stations from 9/24/91
10/2/91. E. affinis control water was H-W (groundwater
plus sea salts) and I-R (Indian River, Delaware).
E.affinis brood size comparisions following 8-d exposures
Station
s
Mean Nauplii Produced
S.E.
H-W Control
5
36.6
12 . 3
I-R Control
5
63. 6
6.8
Wye
4
61.3
1.1
Patapsco
4
38.3
15.7
Morgantown
5
63.4
8 . 6
Dahlgren
5
77.6
9.8
Comparisons were made with H-W control and I-R control and no
signnificant differences were reported using the Bonferroni T-test
(p<0.05) .
Grass Shrimp
Dry Weiaht
Station
n
Mean fmal
S.E.
Control
25
1.05
0.10
Wye
25
1.08
0.11
Patapsco
25
1. 08
0. 08
Morgantown
25
1.24
0.06
Dahlgren
25
1.28
0.09
Sheepshead Larvae Dry Weiaht
Station
n
Mean (ma)
S.E.
Control
23
1.25
0.17
Wye
20
1.68
0.22
Patapsco
24
1.16
0. 08
Morgantown
25
1.20
0.06
Dahlgren
21
1.49
0.09
Mvsid Shrimp Dry Weiaht
Control
25
0.23
0.01
Wye
20
0.32*
0. 01
Patapsco
20
0.30*
0.01
Morgantown
20
0.26
0.01
Dahlgren
20
0.25
0. 02
* Indicates value
is significantly
different than
control,
Dunnetts Test,
p<0.05.
48
-------
controls or other ambient conditions (Table 4.5). This result does
not necessarily mean that an increase in the proportion of mature
niysids suggests an adverse condition was present but it does
reflect a difference from the control that is defined as normal.
4.1.2 Contaminants Data
Inorganic contaminants (trace metals) data from the four
stations for both test 1 and 2 are presented in Table 4.6,
Concentrations exceeding U. S. EPA marine water quality criteria
re underlined in the table (U. S. EPA, 1987a). Total filtered
aluminum concentrations were detected at all stations but the
reported values were likely non-toxic. Arsenic, cadmium and
chromium were below marine water quality criteria at all stations.
Arsenic was not detected at any of the stations while cadmium was
detected at all stations except the Wye River. Chromium
concentrations were detected at all stations during at least one of
the tests and this metal was detected during both tests at the Wye
River (reference site). Copper concentrations exceeded marine water
quality criteria at all stations during at least one test.
Lead was not detected at any of the stations except the
Patapsco River during test 2 (2.1 ug/L). Concentrations of mercury
were not detected at any of the stations during testing. Nickel
concentrations of 9.3 ug/L exceeded chronic water quality criteria
during the first test at the Wye River station. Concentrations of
this trace metal were only 0.1 of a ug/L below the water quality
criteria at the Patapsco Station during the second test. Selenium,
49
-------
Table 4.5 Proportion of immature mysid shrimp after 8 d of
exposure to ambient water from the four stations.
Site
n
Proportion of Immature
S.E.
Control
8
0.23
0.08
Wye
8
0.05
0.03
Patapsco
8
0.03*
0.03
Morgantovm
8
0.20
0.05
Dahlgren
8
0.35
0.05
* Indicates value is significantly different than control, Dunnetts
Test, p<0.05.
50
-------
Table 4.6 Inorganic contaminants data for test 1 (8/14/91 to 8/21/91) and test 2 (9/24/91
to 10/2/91) for the following four stations, Dahlgren (DG), Morgantown (MT), Wye
River and Patapsco River (PA). All underlined values exceed U.S. EPA chronic
marine water quality criteria (WQC), (U.S. EPA, 1987a).
Metals
(ug/L)
WQC
DG—1
DG-2
Stations
MT-1
and Test
MT-2
WYE-1
WYE-2
PA-1
PA-2
A1
, ,
23
13 .2
34.5
7.8
44.7
30.4
12.5
18. 2
10. 4
18 .9
18.0
17.1
32.7
As
—
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
Cd
9.3
.68
1.8
.82
. 59
ND
ND
ND
. 62
.71
1.0
ND
ND
1.4
Cr
50
3 . 1
ND
8.2
ND
7.3
3.1
ND
ND
ND
ND
3 . 3
3.4
3 . 9
Cu
2.9
3.4
4 . 5
5.5
ND
ND
ND
3.7
1+1
3 . 7
4.2
2.9
2.6
2.7
Pb
5.6
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
2.1
Hg 0
. 025
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
Ni
8.3
ND
ND
7.4
ND
9.3
ND
5.2
8.2
ND
ND
ND
ND
6.5
Se
54
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
-------
Table 4.6 continued
Stations and Test
Metals WQC DG-1 DG-2 MT-1 MT-2 WYE-1 WYE-2 PA-1 PA-2
(ug/L)
Sn - ND ND ND ND ND ND ND ND
ND ND ND ND ND
Zn 86 ND ND ND ND ND ND ND ND
ND ND ND ND ND
Ln
ho
-------
tin and zinc were below detection limits at all stations during
both tests.
None of the organic contaminants listed in Table 3.1 were
measured above detection limits in any of the samples collected
from the four stations during either test. A minimum of one sample
was analyzed from each station for each test.
4.1.3 Water Quality Data
Water quality parameters reported from grab samples collected
during both tests are presented in Table 4.7. The temperature and
salinity of all ambient water collected from all stations was
adjusted to 25 C and 15 ppt before testing. Ambient water quality
conditions appeared adequate for survival of test species with the
possible exception of dissolved oxygen concentrations of 4.0 mg/L
at the Morgantown station on 8/13/91. Water quality conditions
reported in test containers during testing are reported in Appendix
B. All parameters appeared adequate for survival of test species.
4.1.4 Reference Toxicant Data
Twenty-four and 48-h LC50 values for the test species exposed
to cadmium chloride during reference toxicant tests are presented
in Table 4.8. These LC50 values were compared with mean 48-h
values for cadmium chloride collected during the first year of the
study. Mysids were not tested in the first year. Forty-eight h
LC50s for three of the test species in the present study were
generally similar to previous values. These data, in conjunction
53
-------
Table 4.7 Water quality parameters reported in the field during
collection of water samples from both tests during I9r
Stations were Patapsco River (PR), Wye River (WR),
Dahlgren (DG) and Morgantown (MT).
Date
Station
Temp
Salinity
Cond
DO
P«
(C)
(PPt)
umhos/cm
(ppm)
8/13
PR
26.0
9.9
16900
8.8
8.44
WR
29.0
10.0
17000
10.0
8.60
DG
26.0
10.0
16000
5.2
7.60
MT
27.0
10.0
16500
4.0
7.40
8/16
PR
25.0
8.5
13900
5.7
8.15
WR
27.3
10.6
18900
6.5
8.01
DG
26.0
10.0
16500
5.8
7.48
MT
26.5
9.5
16000
5.2
7.66
8/19
PR
26.0
9.8
17000
6.5
8.06
WR
27.0
11.0
19000
5.7
7.86
DG
27. 0
10.5
17000
6.4
7.56
MT
27. 0
11.0
18000
5.4
7.77
9/24
PR
20.8
11.2
17100
7.6
7.78
WR
21.6
12.2
19000
7.5
7.9"
DG
22.0
11.0
15500
7.6
7.
MT
23.0
11.5
17000
7.2
7.'/
9/27
PR
20.0
9.7
15000
8.9
8.29
WR
21.0
11.3
18100
11.5
8.30
DG
20.0
10. 0
14500
6.8
7.85
MT
19. 0
11.0
16100
6.2
7.79
9/30
PR
18.2
11.2
16100
10.0
8.27
WR
19.1
12.3
18000
10.9
8.41
DG
19.5
11.0
16000
7.2
7.91
MT
21.0
11.0
16500
6.4
7.96
54
-------
Table 4.8 Toxicity data (LC50s mg/L) from reference toxicant tests
conducted with cadmium chloride for the four test
species. Two reference toxicant tests were conducted
with mysid shrimp because this species was obtained from
two different sources. The mean 48-h values from Year
1 were included.
Date
Oraanism
24 h
48 h 48
-h Values-Year 1
9/20/91
Grass Shrimp
0.932
0.231
0.502
9/12/91
Sheepshead Larvae
1.554
0.510
8/30/91
Eurvtemora
0.095
0.021
9/12/91
Mysids1
0.122
0.009
-
10/22/91
Mysids2
0.026
0.005
-
1 Mysids were obtained from S.P., Inc. (Salem, MA).
2 Mysids were obtained from Chesapeake Cultures (Haynes, VA).
55
-------
with acceptable control survival, demonstrate that these species
were healthy and test results were valid. The 48-h LC50 values f
mysids (0.005 and 0.009 ug/L) were lower than a 3 year mean valu^
of 0.077 ug/L obtained from an effluent toxicity testing program
conducted by our laboratory. These lower LC50 values may be
related to the. test water used. Our tests were conducted in
reconstituted water with sea salts. The 3 year mean value from the
effluent testing program was obtained using natural seawater where
cadmium was likely less bioavailable^ Survival and growth data
with mysids indicated the test species were healthy.
Two reference toxicant tests were conducted with mysids
because the sources of these test species were different. Both 48-
h values were similar (0.009 and 0.005 mg/L).
4.2 Sediment Tests
The following results from sediment toxicity tests an
presented below: toxicity data, contaminants data, and data from
reference toxicant tests. A summary of the water quality
parameters monitored daily in each treatment, and the range of each
parameter is included in Appendix C.
4.2.1 Toxicity Data
Survival results from toxicity tests of the four estuarine
sediments from the Wye, Patapsco, and Potomac River (Morgantown and
Dahlgren) with grass shrimp, amphipods and worms are included in
Tables 4.9 and 4.10. Those stations that were significantly
56
-------
Table 4.9 Cumulative survival data from grass shrimp, amphipod and
worm tests conducted for the four stations during Set 1
experiments. Tests were conducted from August 16, 1991
to September 5, 1991. (LB = Lynnhaven Beach; LI =
Lynnhaven Inlet).
% Survival
Sioecies
Site1
Dav 10
Dav 2C
Grass shrimp
Control (LB)
99
99
Morgantown
99
97
Dahlgren
100
99
Patapsco River
100
97
Wye River
100
97
Reference (LI)
100
97
Amphipod
Control (LB)
96
72
(Lepidactvlus)
Morgantown
92
66
Dahlgren
93
57
Patapsco River
71*
32*
Wye River
92
77
Reference (LI)
91
63
Amphipod
Control (LI)
92
76
fHvalella}
Morgantown
51*
37*
Dahlgren
64*
18*
Patapsco River
20*
2*
Wye River
72
54
Reference (LB)
90
73
Adjusted1
Worm
Control (LI)
99
100
90
Morgantown
80*
92
55*
Dahlgren
68*
82*
61*
Patapsco River
80*
92
77
Wye River
53*
67*
47*
Reference (LB)
70*
91
17*
* Significantly less than the controls (p<0.05).
Notes:
1 Adjusted worm survival is percent survival adjusted for
predicted particle size effects.
57
-------
Table 4.10 Cumulative survival data from grass shrimp, amphipod
and worm tests conducted for the four stations during
Set 2 experiments. Tests
were conducted from Septemb'
27,
1991 to October 17,
1991. (LB
= Lynnhaven Beac.
LI
= Lynnhaven Inlet).
%
Survival
Species
Site1
Dav 10
Dav 20
Grass shrimp
Control (LB)
98
97
Morgantovn
96
93
Dahlgren
97
96
Patapsco River
98
98
Wye River
100
97
Reference (LI)
100
98
Amphipod
Control (LB)
95
69
(Leoidactvlus)
Morgantovn
84
50*
Dahlgren
63*
22*
Patapsco River
46*
8*
Wye River
89
80
Reference (LI)
95
68
Amphipod
Control (LI)
86
62
(Hvalella)
Morgantovn
72
49
Dahlgren
49*
36*
Patapsco River
9*
2*
Wye River
75
62
Reference (LB)
53*
15*
Worm
Control (63%s)
98
87
Morgantovn
89
89
Dahlgren
93
93
Patapsco River
80
70
Wye River
96
91
Reference (100%s) 78*
72*
* Significantly less than the controls (p<0.05).
58
-------
different (a=0.05) from the controls are indicated by an asterisk.
Growth data (mean length and dry weight) for grass shrimp,
amphipods and worms eif.ter 20-day exposure to sediments are included
in Tables 4.11 and 4.12. Any length measurements significantly
greater than the initial length are indicated by a + mark. The
weight measurements were not statistically compared to the initial
values, because the initial weight consisted of 4 0 organisms
grouped on one weigh pan to obtain an average weight per organism.
Amphipod reburial data are shown in Table 4.13. Emergence data for
both amphipods, Hvalella azteca and Lepidactvlus dvtiscus. are
presented in Figures 4.1 - 4.4. In Figures 4.1 - 4.4, the
predicted number alive was determined by linear interpolation based
upon the number of live individuals on day 0, 10 and 20.
Set 1;
Survival and growth data from toxicity tests conducted August
16, 1991 through September 5, 1991 on the first set of sediments
from the four stations are presented in Tables 4.9 and 4.11.
Survival and growth data from the grass shrimp test did not suggest
the presence of toxic conditions (statistical difference between
control and test conditions) in any of the test sediments, and
yielded greater than 97% survival in sediment from all sites.
Growth of animals in the Dahlgren, Wye River and Reference
sediments was significantly different from the initial organisms.
The shrimp in the controls, Morgantown and Patapsco increased in
size over the test period, but the average increases were not
59
-------
Table 4.11 Growth data (dry weight and length) for grass shrimp,
amphipods and worms after 20-day exposure to Set 1
sediments. Initial weight and length represent the
average of 40 animals for each species at the start l.
the test. Data for each replicate are the mean of the
surviving animals. Tests were conducted August 16
through September 5, 1991 (LB = Lynnhaven Beach: LI =
Lynnhaven Inlet).
# of
True
Weight
Length
Site ReDlicates
(mq)
S.E.
^ iron!
S.E.
Grass shrimp:
Initial
NA
14.47
0.00
23.10
0.45
Control(LB)
5
20.18
0.73
24 .78
0.23
Morgantown
5
19.27
0.96
24.16
0. 36
Dahlgren
5
23. 16
1.01
25.72+
0.48
Patapsco River
5
21.96
1.67
25.02.
0.52
Wye River
5
23.09
0.96
25.47+
0.24
Reference (LI)
5
23.19
1.07
26.15+
0.48
Amphipod CLeoidactvlus)
•
•
Initial
NA
0.95
0. 00
6.96
0.19
Control (LB)
5
0.97
0. 09
7.41
0.29
Morgantown
5
1.00
0.05
7.77
0.40
Dahlgren
5
1.05
0.05
7.78
0.13
Patapsco River
5
1.01
0.03
7 . 78
0.31
Wye River
5
1.03
0.03
7.84
0.20
Reference (LI)
5
0.90
0. 03
7 .70
0.06
Amphipod (Hyalella'i :
Initial
NA
0.14
0.00
3 . 35
0.41
Control (LI)
5
0.16
0.07
5.10 +
0. 07
Morgantown
5
0.14
0.01
4.75+
0.24
Dahlgren
5
0.14
0.01
4 .95 +
0.07
Patapsco River
5
0.16
0.00
4 .90+
0. 64
Wye River
5
0.17
0.01
4 .71 +
0. 05
Reference (LB)
5
0.07
0.02
2.80
0.06
Worm:
Initial
NA
0.014
0.000
2.92
0.12
Control(LI)
5
0. 05
0.010
7 . 63 +
0.32
Morgantown
5
0.035*
0. 005
5.04*+
0.23
Dahlgren
5
0.032*
0.003
5.40*+
0.25
Patapsco River
5
0.056
0. 006
6.02*+
0.12
Wye River
5
0.041*
0.008
5.53*+
0.17
Reference (LB)
5
0.047*
0.004
3.84*+
0.15
* Significantly less than the controls (p<0.05).
+ Lengths significantly greater than the initial (p< 0.05).
60
-------
Table 4.12 Growth data (dry weight and length) for grass shrimp,
amphipods and worms after 20-day exposure to Set 2
sediments. Initial weight and length represent the
average of 40 animals for each species at the start of
the test. Data for each replicate are the mean of the
surviving animals. Tests were conducted September 27
through October 17, 1991 (LB = Lynnhaven Beach; LI =
Lynnhaven Inlet).
# of True
Site Reolicates
Weight
fmcM
s.p.
Length
Cmm)
S,E.
Grass shrimp:
Initial
NA
14.4 5
0.00
-
-
Control (LB)
5
12.33
1.00
20.20
0.55
Morgantown
5
13.18
0.69
19.96
0.34
Dahlgren
5
13.99
0.72
20.29
0.30
Patapsco River
5
12.98
0.56
18.77
0.28
Wye River
5
14.51
0.72
20.90
0.34
Reference (LI)
5
12.89
0.26
20. 45
0.12
AitiDhicod (Leoidactvlus)
•
•
Initial
NA
0.84
0. 00
6.57
0.15
Control(LB)
5
1.02
0.09
8.13+
0.23
Morgantown
5
1.00
0.06
7.53 +
0.24
Dahlgren
5
1.05
0.16
7 .12*
0.42
Patapsco River
5
1.15
0.09
6.90*
0.28
Wye River
5
0.93
0.05
7.98 +
0.16
Reference (LI)
5
0.96
0.07
8.06+
0.14
AmDhiDod (Hvalella):
Initial
NA
0. 02
0. 00
2.73
0.06
Control (LI)
5
0.12
0.02
4 .20+
0.24
Morgantown
5
0.12
0.01
4 .16+
0.20
Dahlgren
5
0.14
0.04
3.80+
0.14
Patapsco River
5
0.13
0.10
3.21*+
0.26
Wye River
5
0.13
0.06
4.48+
0.08
Reference (LB)
5
0.07
0.02
2 .80*
0.06
Worm:
Initial NA 0.03 0.00 5.53 0.25
Control (LI) 5 0.04 0.00 4.35 0.15
Morgantown 5 0.07 0.01 6.45 0.51
Dahlgren 5 0.10 0.01 7.74+ 0.14
Patapsco River 5 0.05 0.01 5.24 0.14
Wye River 5 0.09 0.01 6.98+ 0.34
Reference (LB) 5 0.07 0.02 4.22 0.06
* Significantly less than the controls (p<0.05).
+ Lengths significantly greater than the initial (p< 0.05).
6]
-------
Table 4.13 Amphipod fLepidactvlus dvtiscus) reburial data after
10 day exposure to sediments. Table shows percent o
surviving animals able to rebury within one hour (LE
Lynnhaven Beach; LI = Lynnhaven Inlet).
Site % Reburial S.E.
Set 1:
Control (LB) 100 0.0
Morgantovn 99.0 1.0
Dahlgren 99.0 1.0
Patapsco River 89.7* 2.8
Wye River 100 0.0
Reference (LI) 100 0.0
Set 2:
Control (LB) 100 0.0
Morgantown 98.7 1.3
Dahlgren 98.0 2.0
Patapsco River 82.9* 5.5
Wye River 100 0.0
Reference (LI) 99.0 1.0
* Significantly less than control (p<0.05)
62
-------
Fig. 4.1 Emergence of L. dytlscus Set 1 (Aug. '91). Number emerged expressed
as a percent, corrected for predicted no. alive at each day.
Control
to% -
so*-
¦o
£40* -\
E
^ 30% -
10* •
/v^
0 1 2 3 4 5 6 7 1
-A—*-
0 20 21 12 13 14 25 16 17 18 19 20
Div
Reference
«J to*
"¦ 30*
> 20*
2 i 4 S « 7 t 9 10 11 12 13 14 U 14 11 II M »
D.y
Dahlgren
60*
*40*
^ 30*
£ 20*
1 2 3 4 J 6 7 8 9 10 II 22 13 M J5 16 17 18 19 20
D.y
Morgantown
60*
£40*
W 30*
v 20*
k
t 10 11 12 13 14 13 16 17 II 19 20
D.)
Patapsco
Wye River
0 1 I J 4 3 6 1 8 5 10 11 12 13 J4 1J 16 37 It 19 20
-------
Fig.4.2 Emergence of H. azteca Set 1 (Aug. '91). Number emerged expressed
as a percent, corrected for predicted no. alive at each day.
Control
Reference
so*
2 3 4 5 6 7 1
9 10 11 12 13 M 15 16 17 IS 19 20
D.y
t 30*
K 20*
0 1 2 J 4 J 6 7
t 9 10 11 12 13 14 13 16 17 It 19 2D
D.y
Dahlgren
50"%
9 10 11 12 13 14 13 16 17 18 19 20
Morgan town
50*
9 10 U 12 13 14 13 16 17 It 19 20
Patapsco
Wye River
-------
Fig. 4.3 Emergence of L. dytiscus Set 2 (Sep. '91). Number emerged expressed
as a percent, corrected for predicted no. alive at each day.
Control Reference
C 1 2 J 4 5 6 7 i 9 10 11 1J 13 1< 1) 16 17 It 19 »
0*
0 1 : 3 4 J 6 7 t 9 10 11 12 13 14 15 16 17 19 19 20
Diy D«y
Dahlgren
100*
«>*-
t 60* -
E
u:
t 40*
20*-
9 10 11 12 13 14 1! 16 17 It 19 20
Diy
Morgantown
1001
9 10 11 12 13 14 1) 16 17 II It 20
D.)
Patapsco
Wye River
100*
M* -
t 60* -
E
ui
40* -
20* -
i: 13 14 13 16 17 It 19 20
100*
JO* -
60*
E
u
20*-
65
3 3 4 3 6 7 t 9 10 11 12 13 14 15 It 17 l( 19 Jo
D«v
-------
Fig. 4.4 Emergence of H. azteca Set 2 (Sep. '91). Number emerged expressed
as a percent, corrected for predicted no. alive at each day.
Control
20*
10*-
2 3 4 5 6 7ft
9 10 11 12 13 14 15 16 17 It 19 20
D.y
Reference
70* ¦
60* -
¦O 50* -
| 40* H
II
E 30* -
e
20* H
10*-
-A.
012345678
9 10 11 12 13 14 15 16 17 18 19 20
D«y
Dahlgren
70*
60*-
0* -
9 10 11 12 13 M 15 16 17 1! ]9 20
Day
Morgantown
70*
60* -
= 30* -
6- 20*
10* -
2 3 4 5 6 7 I 9 10 11 i: 13 14 15 16 17 IB 19 20
D«v
Patapsco
Wye River
-------
significant, when compared to the initial measurements.
Lepidactvlus dvtiscus showed significantly reduced survival at
days 10 and 2 0 in the Patapsco test sediments. Amphipod reburial
data are presented in Table 4.13. Exposure to Patapsco sediments
had a significant negative effect on amphipod reburial in clean
sediment at the end of the test. Another indication of possible
toxicity is the number of animals emerged throughout the test
period. The normal behavior of the anphipod is to remain buried in
the sediment, due to a negative phototactic response which also
aids in avoidance of predators. Figure 4.1 illustrates the percent
of Lepidactvlus dytiscus emerged per day from each test sediment.
The graphs are presented as the number of animals emerged per day
adjusted by the predicted number of animals alive on that day. The
predicted values were derived from regression analysis of the
survival data at days 0, 10 and 20. The number of Lepidactvlus
dvtiscus emerged in the Patapsco sediments was much greater than
those in other test sediments (see Figure 4.1), with an average of
approximately 25% emerged for the duration of the test. The
increased emergence would have ecological significance, in that
those animals swimming in the water column are exposed to increased
predation pressure. If a high percentage of animals are exposed,
due to this aberrant behavior, the increased predation may
significantly reduce the amphipod population. No significant
effects for survival or growth were seen with any other sediments
in the first set of tests with Lepidactvlus dvtiscus. There were
no significant changes in length in any of the treatments when
67
-------
compared to the initial lengths.
Hvalella azteca showed significantly reduced survival at days
10 and 20 in the Patapsco, Morgantown and Dahlgren sediments.
Patapsco was highly toxic, with only 20% of the amphipods surviving
after 10 days, and only 2% survival at 20 days. The Wye River
sediment test had reduced survival (72% and 54% at day 10 and 20,
respectively) although the reduction was not statistically
significant. No statistically significant growth effects were seen.
Animals in all treatments except the reference (100% sand) grew
significantly from the initial length. The emergence data (see
Figure 4.2) show a high percentage of Hvalella azteca in the
Patapsco treatments out of the sediment for the first 10 days of
the test. The lack of emergence during the second 10 days is
probably a function of the high mortality; only 2% of the organismr
survived to 20 days.
Streblospio benedicti survival was significantly reduced in
all test and reference sediments at day 10, and all sediments
except Patapsco at day 20. All sediment tests also showed reduced
growth (length and weight) as compared to the controls. The worms
had very poor survival and growth in the reference sediment at 20
days. The reference sediment for the worms was 100% sand with very
little organic matter. The worms are naturally found in somewhat
muddy sediment (60-70% sand) and exhibit some particle size induced
mortalities. A regression was run on data from tests conducted
with clean sediments of varying particle size, to enable the worm
survival data to be adjusted for sediment effects and re-evaluated
68
-------
(see Section 3.3.6). The sediments from Dahlgren and the Wye
produced significantly reduced survival as compared to the controls
at day 10 of set 1 tests. The day 20 data were not included in the
assessment of ambient toxicity, because it appears that the worms
may have starved to death during the extended test period of 20
days. The worms in the reference sediment (100% sand) had only 17%
survival at 20 days, and only a slight increase in size as compared
to the lengths of individuals from the initial subset.
Approximately 70-75% survival would be expected at day 10 for
Streblospio benedicti in clean 100% sand. The reference sediment
had a low organic content (1.38%) and, therefore, provided little
or no food for the organisms. To alleviate this problem, all worm
treatments for the second set of toxicity tests were fed every
third day for the duration of the test.
Set 2:
Survival and growth data from toxicity tests conducted
September 27, 1991 through October 17, 1991 on Set 2 sediments from
the four stations are presented in Tables 4.10 and 4.12. Survival
and growth data from the grass shrimp test did not suggest
statistical differences between controls and any of the test
sediments. The initial length measurement for the grass shrimp was
lost due to a calibration error with the image analysis system.
The results of the Lepidactvlus dytiscus sediment tests showed
statistically reduced survival at 10 days in the Dahlgren and
Patapsco River sediments. At day 20 there was significantly
69
-------
reduced survival in the Morgantown, Dahlgren and Patapsco sedime*
samples. The Patapsco appeared highly toxic, with only 8% of ti._
amphipods surviving to day 20. Both the Patapsco and Dahlgren
sediments produced statistically reduced growth of Lepidactvlus
dvtiscus. The surviving amphipods in Patapsco and Dahlgren
sediment tests did not grow significantly in length compared to the
initial measurements, another indicator of a population under
stress. All other test populations displayed significant increases
in length from the initial values. Patapsco sediments also had a
significantly negative effect on the amphipod during the reburial
portion of the test. The emergence observations from set 2
sediments indicated that a large percentage of animals in the
Patapsco sediments were swimming in the overlying water throughout
the test. This is considered an avoidance response for the
amphipod, indicating adverse sediment conditions, and ha.
potentially important ecological implications. Animals exposed and
swimming in the water column risk increased predation, thereby
reducing local amphipod populations. Microcosm experiments have
demonstrated that amphipods are capable of leaving contaminated
sediments and preferentially seeking out clean sediments, even when
it means migrating through the water column and circulating systems
of microcosm tanks (Alden and Butt, 1988). Therefore, sediments
contaminated enough to produce an avoidance response of this sort
would be expected to display low benthic standing stocks and
productivity. We are not aware of any direct experimental evidence
from the field to support the idea that contaminated sediments are
70
-------
avoided by field populations of amphipods. In fact, we believe
that the experimental design for a study to provide such evidence
would be very difficult and plagued with confounding effects.
However, Dr. Dan Dauer at ODU has made field observations that
amphipods routinely migrate out of the sediments, possibly to seek
new s.ubstrates.
Lepidactvlus dvtiscus is naturally found in very sandy
sediments, and exhibits some mortality due to particle size
effects. Lepidactvlus dvtiscus exhibits increasing mortality with
decreasing sediment sand content. Extensive testing has been
conducted at the AMRL with Lepidactvlus dvtiscus in a range of
sediment types. These data have been compiled to generate a
regression that predicts expected mortality due to particle size
(see Figure 4.5). In sediments greater than 80% sand, less than 5%
of the mortalities would be due to particle size effects. The
Reference sediment was the muddiest sediment tested (63% sand).
From the Lepidactvlus dvtiscus regression, survival should be 85%
or better in a 10-day test in this sediment. Lepidactvlus dvtiscus
survival in the reference sediment was 91% (set 1) and 95% (set 2) .
Therefore, the mortalities seen in all test sediments can be
attributed to toxicity, rather than particle size effects.
Hvalella azteca tests indicated statistically reduced survival
in the Patapsco, Dahlgren and Reference sediments at days 10 and
20. The Patapsco sediment was highly toxic with only 9% of the
test organisms surviving at day 10, and only 2% surviving at day
20. Both the Patapsco and the Reference sediment tests produced
71
-------
Fig. 4.5 Survival of L. dytiscus versus percent silt/clay
100
•vj
ho
I
I
P
w
15 25 35 45 55 65 75 85
^SILT/CLAY
-------
statistically significant reductions in growth. All amphipods
increased in size significantly from the initial subset, except the
animals in the reference sediment. The emergence data for Hvalella
azteca revealed that a large percentage of the organisms in the
Patapsco sediment were emerged compared to emergence rates in other
sediment samples. The low survival and growth observed in the
reference sediment for the second set of tests was of concern. The
first set of tests conducted with Hvalella azteca in the same
reference sediment (100% sand) produced 90% survival at day 10 and
73% survival at day 20. The results from the second set of tests
were radically different: 53% survival at day 10 and only 15%
survival at day 20. Additional testing of Hyalella azteca in the
reference sediment was conducted to investigate the unexpectedly
high mortalities observed in the Set 2 reference sediments (August
through September test). Samples of reference and control
sediments were treated identically to the sediments used in the
ambient toxicity tests. A series of ten day Hvalella azteca tests
were conducted and the results evaluated (see Appendix D). The
data from these tests indicated greater than or equal to 90%
survival of Hvalella azteca in all of the 100% sand treatments.
This is consistent with the results from the reference sediment
used with Set 1 sediment samples. Therefore, it is probable that
the results from the reference sediment used for the second set of
the ambient sediment toxicity tests were aberrant. The population
of Hvalella azteca used for the second set of tests may have been
less healthy than those used for the first set, since the control
73
-------
results were slightly lower for Set 2. The reference toxicant
tests conducted on the Set 2 amphipods gave valid results, althou'
the data did indicate a slightly more sensitive population than the
first set of test organisms.
The Streblospio benedicti tests for Set 2 indicated no
statistically reduced survival or growth. All treatments were fed
0.25 mg of "worm food", a combination of dried algae (Ulva) and
tetramin ground to a fine powder (<500 um) , every third day.
Feeding the worms increased survival at day 20 in the reference
sediment from 17% (Set 1) to 72% (Set 2). Worm survival in clean
100% sand is predicted by the particle size effects regression to
be 70-75%. The survival in the reference sediment at day 20 was
72%. Therefore, the addition of a worm feeding protocol appears to
have successfully eliminated the problem of starvation. The worm
particle size regression was then used to estimate the predicted
effects due to particle size and adjust the survival data. The
reference sediment adjusted survival was not statistically
different from the controls. In addition, no test sediments
produced statistically significant effects with the worms in Set 2.
4.2.2 Contaminants Data
Toxicity of chemicals in sediments is influenced by the extent
to which chemicals bind to the sediments. There are many factors
that influence the binding capabilities of a particular sediment.
The toxicity of non-ionic organic chemicals is related to the
organic content of the sediments, and it appears that the
74
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bioavailability of sediment-associated metals is related to the
concentration of Acid Volatile Sulfides (AVS) present in the
sediment (DiToro, 1990). Sediment samples from the four stations
and the controls were analyzed for Total Organic Carbon (TOC) and
Acid Volatile Sulfides (AVS). The results are shown in Tables 4.14
and 4.15. At present, there is no readily accessible data base for
comparison of TOC normalized data, therefore the TOC analysis from
this study was included to allow for future comparisons. The AVS
approach to sediment contaminants evaluation is still developmental
and has been published only recently (DiToro, 1990) . To
appropriately interpret the AVS data, simultaneously extractable
metals (SEM) must also be analyzed. The data for SEM are presented
in Table 4.16. In evaluating the AVS values, a ratio of the sum of
the SEM to the total AVS is calculated. If the SEMrAVS ratio
produces a value less than one, it is assumed that there is
sufficient AVS present in the sediment to bind with the metals,
rendering them biologically less available and therefore less
toxic. If the ratio is greater than one (1), toxicity due to
metals is less predictable, but generally expected to be more
likely. An evaluation of the SEM to AVS ratio is included in Table
4.17. All stations had ratios much less than one, except the
Patapsco River. The sediments collected in August had a ratio of
0.946, and the sediments collected in September produced an SEM:AVS
ratio of 1.113. These ratios indicate the potential for toxicity
due to metals. The SEM analysis (Table 4.16) indicated a very high
concentration of zinc in both Patapsco samples.
75
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Table 4.14 Chemical data (TOC) for sediment samples from the
four stations and the controls. All data is or
dry weight basis. Set 1 was tested August
through September 5, 1991. Set 2 was testt.-
September 27 through October 17, 1991.
Station Total Organic Carbon
Set 1:
Lynnhaven Beach <0.37
Morgantovn 0.41
Dahlgren 0.42
Patapsco River <0.37
Wye River 0.38
Lynnhaven Inlet 1.38
Set 2:
Lynnhaven Beach <0.37
Morgantovm <0.37
Dahlgren 0.58
Patapsco River <0.37
Wye River 0.4 0
Lynnhaven Inlet 1.78
76
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Table 4.15 Chemical data (AVS) for sediment samples from the four
stations and the controls. All data is on a dry weight
basis. Set 1 was tested August 16 through September 5,
1991. Set 2 was tested September 27 through October
17, 1991.
Site AVS
-------
Table 4.16
SEM analysis for sediments for Set 1 (Aug. '91) and Set,2 (Sept. *91). Concentrations
for each metal are expressed in umol per gram of sediment (LB = Lynnhaven Beach; LI =
Lynnhaven Inlet).
Set 1
Cadmium Lead Copper Nickel zinc Sum
Mean SE Mean SE Mean SE Mean SE Mean SE Mean SE
Site
Morgantown
0.011
0.0005
0.031
0.0012
0.077
0.0005
0.059
0.0274
0.302
0.0035
0. 479
0.0288
Dahlgren
0. 018
0.0003
0.057
0.0015
0. 066
0.0038
0.146
0.0473
0.643
0.0111
0.930
0.0378
Patapsco
0. 022
0.0003
0.119
0.0012
0.161
0.0023
0.090
0.0225
5.669
0.0355
6.062
0.0394
Wye River
0. 001
0.0002
0. 007
0.0011
0. 012
0.0034
0.020
0.0103
0.068
0.0039
0.109
0.0115
LB
0.000
0.0003
0.001
0.0013
0. 003
0.0020
0.068
0.0301
0.025
0.0033
0.098
0.0297
LI
0. 006
0.0010
0.037
0.0018
0.041
0.0023
0.029
0.0063
0.663
0.0073
0.776
0.0107
, Set 2
Cadmium
Lead
Copper
Nickel
Zinc
Sum
Mean
SE
Mean
SE
Mean
SE
Mean
SE
Mean
SE
Mean
SE
Site
Morgantown
0.010
0.0009
0.032
0.0015
0.060
0.0009
0.110
0.0381
0.256
0.0079
0.468
0.0297
Dahlgren
0. 016
0.0005
0.049
0.0009
0.062
0.0025
0.105
0.01,91
0.581
0.0057
0. 812
0.0192
Patapsco
0. 025
0.0013
0.123
0.0010
0.174
0.0047
0.081
0.0184
5.803
0.1074
6. 206
0.1175
Wye River
0.003
0.0010
0.008
0.0013
0.021
0.0068
0.052
0.0252
0.080
0.0020
0.164
0.0306
LB
0. 000
0.0000
0.002
0.0020
0.008
0.0025
0.133
0.0476
0.023
0.0012
0.166
0.0491
LI
0.005
0.0003
0.041
0.0023
0.033
0.0007
0.126
0.0607
0.621
0.0035
0.826
0.0555
Detection
Limits 0.0003 0.005 0.0006 0.0004 0.0005
NOTE: All Mercury values were below detection limits of 0.00005 in both sets for all sites.
-------
Table 4.17 Average SEM and AVS values and the SEM:AVS ratio for
sediment samples tested in 1991.
Site
Set 1:
Lynnhaven Beach
Morgantovn
Dahlgren
Patapsco River
Wye River
Lynnhaven Inlet
Mean AVS
<1.0
<1.0
1.78
6.41
1.48
3.25
Mean SEM
0.098
0.479
0.930
6.062
0.109
0.776
Ratio
0.524
0.946
0.074
0.239
Set 2:
Lynnhaven Beach
Morgantovn
Dahlgren
Patapsco River
Wye River
Lynnhaven Inlet
<1.0
1.67
1.64
5.58
0.73
3.94
0.166
0.468
0.812
6.206
0.164
0.826
0.280
0. 497
1.113
0.226
0.2099
79
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It should be noted that the zinc sulfide solubility parameter
is greater than those of all the other metals measured except f
nickel (DiToro et al. , 1990). Therefore, nickel and then zinc
would be predicted to be the first two metals to be released to a
dissolved free ion form (i.e., the toxic form of most metals) when
the binding capacity of the AVS is exceeded by the total SEM
concentrations. Therefore, toxicity may be associated with these
two metals (particularly zinc, which is found in high
concentrations in these sediments, see Table 4.18 below). On the
other hand, the AVS-metals dynamics are not well understood and
other metals not measured in the current study may play a role in
the observed toxicity.
Inorganic contaminants data from the stations are presented in
Table 4.18. All test sites had concentrations above the detection
limits for ten of the eleven metals analyzed. The eleventh metal
tin, was below detection limit at all sites except the Patapsco
Fiver, which had 42 ug/g and 56 ug/g from the first and second
sets, respectively. At the Wye River station, mercury
concentrations were below detection limit in the sample collected
on August 14th. The amphipod control sediment tested with the
second set of sediments had concentrations of mercury, selenium,
and tin below detection limits. The amphipod control sediment used
in the first set of tests had concentrations of mercury, selenium,
tin and nickel below detection limits. Sediment-sorbed
contaminants have been extensively studied by Long and Morgan
(1990). They have established a table of concentrations at which
80
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Table 4.18 Inorganic contaminants for sediment samples from the four stations and the
controls. (Note: single underlined values represent concentrations exceeding
"Effects Range-Low", and double underlined values represent concentrations
exceeding "Effects Range-Median" levels as defined in Long and Morgan, 1990 (LB = Lynnhaven
Beach; LI = Lynnhaven Inlet).
Contaminant fua/a)
Site
Al
As
Cd
Cr
CU
Hq
Ni
Pb
Se
Sn
Zn
Set 1:
LB
173
2.24
0. 015
0.6
0.25
<0.020
<0.2
0.7
<0.01
< 2
2.0
Morgantown
7462
1.33
0.466
24.6
5.62
0.011
4.6
7.7
0.10
< 2
33.7
Dahlgren
6619
2.74
0.156
12.4
4.41
0.027
6.6
10.6
0.09
< 2
47 .9
Patapsco
674
3 .77
1.646
107.8
14.22
0.011
4.2
23.0
0.37
42
382.6
Wye
2918
0.52
0.074
4.6
0. 57
<0.006
1.1
25.4
0.07
< 2
8.6
LI
10462
4.46
0. 288
17.7
8.35
<0.006
8.8
11.7
0.25
< 2
56.4
DO
"Set 2:
LB
274
0.35
0. 017
0.8
0.70
<0.006
0.3
1.0
<0.01
< 2
3.2
Morgantown
7889
2.85
0.298
25.5
7.86
0.027
6.,1
7.6
0.04
< 2
36.5
Dahlgren
7315
2.46
0.071
13.0
8.39
0.036
7.4
10.6
0.05
< 2
4 6.1
Patapsco
732
3.72
1.254
130.6
16.79
0.024
4.8
37,8
0.24
56
455. 3
Wye
4443
0.95
0. 053
6.4
1.77
0.008
2.1
3.8
0. 05
< 2
12.6
LI
10309
3.95
0.085
18.3
9.13
0.036
9.5
13.4
0.01
< 2
55.7
Detection
Limit
25
0.01
0.005
0.1
0.05
0.006
0.2
0.5
0.01
2
0.1
Effects range:
LOW
—
33
5
80
70
0.15
30
35
—
NA
120
MEDIAN
85
9
145
390
1.3
50
110
—
NA
270
-------
biological effects would be expected if these contaminants wer'
present in the sediment. The lower ten percentile of their date
was established as the "Effects Range-Low" (ER-L) and median
concentrations were identified as the "Effects Range-Median" (ER-
M) . Comparisons can be made between sites with respect to the
potential for adverse biological effects. This is accomplished by
comparing the level of toxicants observed through chemical analysis
with the ER-L or ER-M values. Those contaminants with levels
exceeding the ER-L are in the "possible" effects range for toxic
effects. The contaminant levels above the ER-M fall in the range
of "probable" toxic effects.
The Patapsco River station had concentrations of zinc (382.6
ug/g for set 1 and 455.3 ug/g for set 2) that exceeded the ER-M of
270 ug/g as defined by Long and Morgan (1990). Therefore,
evaluating the results according to these criteria leads to similar
conclusions to that of the SEM/AVS evaluation: that zinc may be a
problem in these sediments. The Patapsco River site also had
concentrations of chromium and lead (Table 4.17) exceeding the ER-L
for chromium (80.0 ug/g) and lead (35 ug/g). Thus, other
contaminants, both measured and unmeasured may also be responsible
for the observed toxicity.
The results of semi-volatile organic compounds and pesticides
analyses in sediment samples are presented in Appendices E and F.
The Morgantown sediment was the only sample with compounds listed
in the Long and Morgan (1990) reference that exceeded the possible
effects ranges. The concentration of 4,4'-DDT was above the ER-L
82
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value (Appendix F).
4.2.3 Pore Water Data
Sediment pore water was analyzed for sulfide, ammonia and
nitrite for all stations and the controls and references. High
ammonia values were recorded for Lynnhaven Inlet (63% mud) and
Lynnhaven Beach (100% sand) for Set 1 sediments and the reference
for Set 2 sediments (Table 4.19). No relationship existed between
test sediments that produced significantly reduced survivals and
those sediments with high ammonia levels.
4.2.4 Reference Toxicant Data
The relative sensitivities of each set of test organisms was
evaluated with reference toxicant tests. The results of each
reference toxicant test conducted with each batch of grass shrimp,
amphipods and worms are shown in Table 4.20. The grass shrimp was
tested using SDS (sodium dodecyl sulfate); all other organisms were
tested using cadmium chloride (CdCl2). All test LC50s were within
the range of the previous reference toxicant tests conducted,
indicating healthy and sensitive populations of test organisms.
The grass shrimp reference toxicant tests results were similar
to previous tests conducted at the AMRL, as well as results
published in the literature. Tatem et al. (1976) found median
lethal toxicity values ranging from 52 to 98 mg/L SDS for grass
shrimp held in the laboratory less than a month. Reference
toxicant tests from 1990 gave values ranging from 56.7 to 67.7 mg/L
83
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Table 4.19 Chemical data for pore water samples from the four
stations and the controls.
Site
Set 1:
Lynnhaven Beach
Morgantown
Dahlgren
Patapsco River
Wye River
Lynnhaven Inlet
Set 2:
Lynnhaven Beach
Morgantown
Dahlgren
Patapsco River
Wye River
Lynnhaven Inlet
Ammonia Nitrite Sulfide
(rcq/L) fmq/L) (mq/L)
13.209
2.637
3.559
4.542
3.190
19.639
0.0022
0.0451
0.0303
0.0069
0.0041
0.0013
0.036
0.031
0. 010
0.115
0.011
0.043
0.64 3
1.996
3.989
4.266
2.729
20.407
1.2117
0.0061
0.0025
0.0041
0.0016
0.0013
0.036
0. 014
<0.008
0.040
0.019
0.031
84
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Table 4.20
Reference toxicant data results from 96-hr, water only,
reference toxicant tests for the second year of the ambient
toxicity project. Sodium dodecyl sulfate (SDS) was used with
the grass shrimp and cadmium chloride (CdCl2) was used for all
other organisms.
Organism
P. puqio
Test Set i
1
2
Chemical LC50 & CIs (mq/L)
SDS* 53.04 (43.74 - 64.29)
SDS 52.32 (38.70 - 70.44)
Historical
Mean
53.01
L. dytiscus
1
2
CdC12 * * 6.55 (4.54 - 9.45)
CdCl2 7.13 (5.23 - 10.36)
5 .22
S. benedicti
1
2
CdC12
CdCl2
4.12 (2.21 - 8.83)
5 . 27 (4 . 19 - 6 . 56)
5.79
azteca
1
2
CdC12
CdC12
0.045 (0.034 - 0.059)
0.038 (0.029 - 0.050)
0. 028
85
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SDS. The reference toxicant tests on grass shrimp used for this
study produced comparable LC50s of 52 and 53 mg/L SDS.
Lepidactvlus dvtiscus and Streblospio benedicti reference
toxicant responses were within the range of those from previous
tests, and were similar to the 1990 reference toxicant data.
Since Hvalella azteca is a freshwater organism, most of the
reference toxicant data available in the literature are from tests
conducted at 0 ppt. However, multiple tests have been conducted at
the AMRL with Hvalella azteca in 15 ppt water.in a effort to build
a data base for comparison of future tests. Previous reference
toxicant tests conducted with Hvalella azteca in 15 ppt water
produced LC50's ranging from 0.25 to 0.47 mg/L CdCl2. The
reference toxicant test results for this project fall within that
range.
4.3 Suborqanismal Tests
4.3.1 EROD Activity in Feral Fish
EROD activity was measured in feral Fundulus (Wye River and
St. John's Creek), white perch (Morgantown and Dahlgren) and spot
(Patapsco River and Patuxent River) on sampling dates in August and
September (when water and sediment were also being sampled for
bioassays) (Table 4.21).
4.3.1.1 Fundulus heteroclitis
EROD activity in Fundulus from the Wye River was variable and
not significantly different in September versus August. EROD
activity in Fundulus from St. John's Creek was greater than (0.05<
86
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Table 4.21 Hepatic ethoxyresorufin-O-deethylase (EROD) activity
and in feral fish. August and September sampling
dates correspond to dates for water column and
sediment tests.
Species N EROD Activity
(pmol/min/mg)
Fundulus
Wye River
August
September
St. John's Creek
July
White Perch
Dahlgren
August
September
Morgantown
August
September
Spot
Patapsco River
August
September
Patuxent River*
November
7
5
10
6
10
6
8
10
8
623.4
970.1
440.7
136.4
1580.9 1051.6a
41.5
55.2
41.2
49.5
15.0
17.3
11.4
9.6
2706.0 1068.0
1214.5 337.2
126.1
88.0
All data as mean std. ± deviation.
*Spot collected from Sheridan Pt. in the Patuxent River between
November 19 and 25, 1991.
ASignificantly different from August Wye River sampling (0.10 < p
< 0.05).
87
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p < 0.10) Fundulus sampled from the Wye River in August and similar
to values found in the Wye River in Year 1. Values for EP
activity in Wye River Fundulus in Year 2 were less than those founa
in year 1 (ca. two-fold).
Elevations in EROD activity in Fundulus (Elskus and Stegeman,
1989), and other fish species (Stegeman and Kloepper-Sams, 1987)
are associated With planar aromatic hydrocarbon exposure. Thus,
where activities are high, it is inferred, in the absence of
corroborating chemical information, that exposure to these
compounds has occurred. Thus, the EROD activities suggest that Wye
River Fundulus were less exposed to planar aromatic hydrocarbons
than those from St. John's Creek. Since the area from St. John's
Creek where these fish were collected is a relatively quiescent
area with modest tidal exchange, and the Creek in general
experiences a significant amount of motor boat activity (an^
attendant fuel pollution) , this observation is perhaps no>.
surprising.
However, EROD activity in feral Fundulus from the Wye River in
Year l was much higher and similar to that seen in the St. John's
Creek Fundulus and also similar to values seen in Fundulus from the
heavily creosote contaminated areas of the Elizabeth River. In the
absence of any corroborating chemical information, we cannot be
certain as to the cause of these elevations in Wye River fish,
though the data suggest that aromatic hydrocarbon exposure is
occurring at levels high enough to elicit this cellular response.
Alternately, other, uncharacterized, chemical (natural dietary) or
88
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physiological factors could be influencing EROD activity. The
large number of studies which have documented clear relationships
between EROD activity find planar aromatic hydrocarbon exposure (See
Stegeman and Kloepper-Saros, 1987 for review), however, argue
against this alternate hypothesis. The lab studies which are
described below also support this discussion.
4.3.1.2 White Perch
EROD activity in white perch from both Potomac River stations,
at both sampling times., was low and similar among stations. The
mean activity values are approximately 6-7 fold less than the
values measured in Year 1.
The reason for the differences in mean activity values between
years 1 and 2 are not immediately apparent. One explanation would
be that fish in year 1 were actually experiencing a higher exposure
level and the higher activity levels are a reflection of this.
However, since laboratory studies have not documented the degree of
responsiveness of white perch to PAH' s, we cannot conclude with any
certainty that this explanation is tenable.
One possible confounding factor could be related to the high
concentration of copper that has been observed in livers from this
species (See data below and Bunton and Baksi, 1988; Bunton et al.,
1987). High concentrations of some trace metals are known to
decrease the activity of the monooxygenase system in general
(George, 1989) and the high levels of copper in livers of this
species may obscure the ability to detect any environmental
induction of EROD activity. Until laboratory studies are conducted
89
-------
to examine this hypothesis, white perch (in the Chesapeake Bay, in
particular) may not be a suitable species for field studies
monooxygenase induction.
4.3.1.3 Efoot
EROD activity in feral spot collected from the Patapsco River
in August was approximately two-fold higher than spot collected in
September. The values for EROD activity were 3-5 fold less than
the levels found in spot collected from this area in year 1. In
comparison, EROD activity in spot collected in the Patuxent River
in November was 10-20 fold less than the values seen in spot from
the Patapsco River.
However, there was a significant difference in the
hepatosomatic index (liver wt./body wt.) in the Patuxent River
spot. The livers of spot collected in the Patuxent River in
November were nearly three-fold larger (mean HSI 3.3 versus 1.2
than those collected in the Patapsco. There is no immediate
explanation for this observation and thus we cannot be certain if
there is any relationship between factors causing this increase in
liver size and changes in EROD activity, or responsiveness to
aromatic hydrocarbon inducers. We suspect that the increased liver
size may be due to changes in hepatic physiology (increased lipid
storage) as the animals experience cooling water temperatures and
approach the winter season (Hazel, 1979). Our casual observations
at the time of dissection noted that the livers were friable and
appeared to be carrying high levels of lipid. Previous experiments
with Fundulus (Stegeman, 1979) have shown that cold acclimation can
90
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inhibit the ability to induce benzo(a)pyrene hydroxylase (analogous
to EKOD) over short time frames. However, there was only a modest
influence of temperature on basal monooxygenase activity.
We suggest, therefore, that the high levels of EROD activity
in spot from the Patapsco are consistent with the aromatic
hydrocarbon pollution in that region as compared to the Patuxent
River. Van Veld fet al. (1990) have recently provided convincing
evidence that spot liver EROD activity can be used as a reliable
reflection of the aromatic hydrocarbon contamination gradient
present in the Elizabeth River, Virginia. The activities (high and
low) we have detected are similar to those found in the Van Veld et
al. (1990) study. Spot appear to be an excellent choice for
further surveys of EROD activity in feral fishes around the
Chesapeake Bay, though a further investigation of the increased
liver size in late season spot is warranted.
4.3.2 Holding Time for Fundulus to Reach Baseline EROD Levels
Fundulus were collected from the Wye River and St. John's
Creek during late July and early August. Experiments with the St.
John's Creek Fundulus began on July 25, while experiments with Wye
River fish began on August 8. Results are presented in Figure 4.6.
Fundulus from St.. Johns' Creek exhibited a rapid decline in
EROD activity from the time of collection through week 3. Also
associated with this decline was a corresponding decrease in the
variance (as shown by the narrowing error bars). Levels of EROD
activity after three weeks were similar to those found after twelve
91
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Wye River Fundulus
o 2000
E 1000
a.
cc
1 0
St John's Creek Fundulus
E 2000
Q.
O)
| 1000
Q.
0
10
Weeks Held in Laboratory
Fig. 4.6 EROD activity in Fundulus heteroclitus following collection
in St. John's Creek and the Wye River.
92
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weeks, indicating that 3 weeks to one month is the time necessary
for EROD activity to reach baseline levels in St. John's Creek
Fundulus.
EROD activity in Fundulus from the Wye River declined during
the first week, increased (nearly 6-fold) from week 1 to 2 and then
returned to collection levels at week 3. After 10 weeks, EROD
activity had declined to levels similar to those in the St. John's
Creek fish. There are several things to note from this data.
First, Wye River fish in year 2 did not have the high levels of
EROD activity at collection that we found in year 1. Thus the
decline in activity at 10 weeks was not as great as with St. John's
fish. In fact, the Wye River fish do not appear to have highly
elevated EROD activity at all in comparison to control fish from
treatments used in year 1. We will note below, however, that in
the lab experiments using a model PAH inducer, the absolute
activity levels seen in Fundulus from Wye River are similar to
those seen in lab Fundulus treated with benzo(a)pyrene.
Secondly, we have no immediate explanation for the spike in
activity between week 1 and 3 in Wye River fish as compared to the
more idealized behavior of the St. John's Creek fish. As mentioned
earlier, the weeks when these activities were measured do not
correspond directly (i.e., the week 2 sample date for Wye River was
August 22, while the fish from St. John's Creek were in between the
3 and 12 week sampling period on August 22). Since these fish were
being held in ambient Patuxent River water, it is possible that a
pulse of aromatic hydrocarbon contaminated water went through these
93
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tanks and the Fundulus responded accordingly. We cannot be sure
that the St. John's Fundulus did not also exhibit this spike
activity because we did not sample them during this time. In this
case, the EROD activity may have accurately reflected
contamination.
An alternate- hypothesis could be that the spike in activity
was due to the mobilization of some stored contaminant in the Wye
River Fundulus and subsequent induction of EROD. VJe suggest that
this is unlikely since there do not appear to be substantial
sources of persistent contaminants, such as PCBs, Dioxins, etc.,
near this site which could have caused this.
It is important to note that these fish ultimately reached
similar baseline levels of EROD as the St. John's Creek fish. It
should also be noted from the following section that EROD activity
in all of these fish did continue to decline to low levels over t
next six months, prior to the initiation of the laboratory
induction studies.
4.3.3 Evaluations of EROD Induction in Fundulus from the Wye
River Versus St. John's Creek
Fundulus from the Wye River and St. John1 s Creek were treated
with 0 (corn oil), 2 and 10 mg/kg of benzo(a)pyrene and sacrificed
after 96 hours. EROD activity from these fish is shown in Figure
4.7.
EROD activity was induced to a similar extent in Fundulus from
the Wye River and St. John's Creek. That is, there was no
94
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wye River Fundulus
~ St. John's Creek Fundulus
1 0
Dose (mg/kg)
Fig. 4.7 Induction of EROD activity in Fundulus heteroclitus from
two different populations, if = significantly different
from control at p< 0. 10. * = significantly different from
control at p < 0.05.
95
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significant difference in the magnitude of the response between the
populations. Mean EROD activity was induced approximately two-i
at the 2 mg/kg dose in both populations, though the variability of
the response of the St. John's fish prevented a statistical
difference. The mean EROD activity in fish treated with 10 mg/kg
was similar and both populations were statistically elevated
(approximately 5-fold) over the controls.
We note here that the absolute levels of EROD activity were
low in the controls, the lowest levels we have seen in the two
years of study. The activity in the fish induced with 10 mg/kg,
while 5-fold higher than the control fish, was similar to the
levels seen in freshly collected fish from the Wye River in year 2,
but still less than half that seen in fish from the Wye in year 1
and from the Elizabeth River in year 1.
Since the submission of the year 1 report, we have al
conducted an additional dose-response experiment using ip
injections of benzo(a)pyrene and fish from St. John's Creek (Figure
4.8). In this experiment, we used a broader range of treatment
levels and detected significant induction only with 10 mg/kg of
B(a)P. The response seen with the 10 mg/kg treatment was
substantially higher than observed in the experiment described
above. The data also suggest a rather steep dose-response between
1 mg/kg and 10 mg/kg-, which is why the 2 mg/kg treatment was used
in the more recent experiment.
Ke also reported data on the sex of the fish, since the fish
were developing gonads at the time the experiment was conducted
96
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4000
3000 -
2000 -
W-
W
1000 -
Control 0.05
0.1
0.5
1 0
Dose (mg/kg)
Fig. A.8 Dose-response ER0D activity in Fund'alus heteroclitus
treated vith benzo(a)pyrene. * = significantly different
from control at p C 0.05.
97
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(some females actively extruded eggs with gentle abdominal
pressure) and sexual maturity has been shown to effect ER
inducibility in female fish (Gray et al., 1991). As shown in Figure
4.9, induction was strong in both males and females, with the
magnitude of induction similar among males and females, but the
absolute rates less in the female controls and 10 mg/kg treatment
group. While this observation is different from that seen with
some female teleosts (little or no induction in gravid females), it
is consistent with other recent experiments with Fundulus
heteroclitus (Kloepper-Sams, 1989). This suggests that Fundulus
will show environmental induction in the absence of an experimental
design which includes segregation of animals by sex during active
gonadal development.
4.3.4 Supplemental Studies on Copper Concentrations in Liver;
of White Perch from the Potomac River
In Year 1 we observed, during dissection, that the livers of
many of the white perch sampled from the Dahlgren and Morgantown
stations were abnormally dark brown in color. Data in the
literature on other white perch from around the Chesapeake Bay
(Bunton and Baksi, 1988; Bunton et al., 1987) suggested that this
was due to high levels of copper. A phone conversation with Dr.
Tracy Bunton of Johns Hopkins University confirmed this observation
qualitatively.
In Year 2, we examined a set of livers from white perch from
our August sampling at these stations for copper to confirm and
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n Controls+non-responsive doses
H 10mg/kg
Soli*
Male Female
Fig. 4.9 Sex differences in EROD activity in Fundulus heteroclitus
treated with benzo(a)pyrene.
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quantitate this observation (Table 4.22). The concentrations found
in the livers of these white perch are similar to those found
Bunton previously. The mean concentrations were not significantly
different between the two stations. In general, Bunton et al.
(1987) reported that concentrations increased with increasing size
of the fish, and that general observation is apparent here.
However, in some of the largest fish, copper concentrations are not
necessarily as high as in some smaller fish. At the time we
collected these sampled we noted that'"some livers Were not as dark
as others and selected a few fish specifically for this reason.
Therefore, these data may contain some bias and no special
significance should be attached to the fact that some large fish
have low hepatic copper concentrations. This appears to be part of
the normal variability that is encountered in these fish.
Bunton et al. (1987) have also suggested that this observatic
does not appear to be related to generally high levels of copper in
Chesapeake Bay waters, as this observation has been made on fish
from many of the tributaries of the Chesapeake. In addition,
hepatic copper concentrations in the closely related striped bass
(Morone saxatilis^, and other fish sampled from the same locations
as white perch, do not show this anomaly. White perch from the
Chesapeake appear to accumulate copper via some novel biochemical
mechanism which has yet to be described. We do not know if this
observation extends to white perch from other locations. It is
also unknown what effect, if any, this high hepatic copper has on
the ability to use white perch for monooxygenase monitoring or on
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Table 4.22 Copper concentrations in livers from white perch
sampled from the Korgantown and Dahlgren stations.
Station
Fish Wt. (g)
Copper Concentration
(ug/g wet weight)
Dahlgren
Mean ± s.d.
Korgantown
Mean 1 s.d,
55
59
65
54
57
52
63
51
57 ± 5
121
101
55
77
142
55
56
52
44
t 35
1148.3
915.7
1602.0
2162.7
1570.3
671.3
2072.3
514 .1
1332.1 ± 618.8
552,
3366,
1281,
2705,
290,
1096,
1249 ,
1822 ,
1021,
78
1487.1 ± 993.8
Copper concentrations were determined by nitric acid digestion of
tissue samples and flame atomic absorption spectrophotometry.
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other physiological processes in white perch such as growth and
reproduction.
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SECTION 5
DISCUSSION
5.1 Potomac River - Moraantown
The Morgantown station is located in a critical mesohaline
habitat of the Potomac River. Previous studies have suggested that
toxic concentrations of tributyltin (TBT) (20 to 24 ng/L) were
present at this station in 1985 and 1986 (Hall et al., 1987b).
Results from our first year of water column testing at this station
suggested the presence of toxic conditions as survival of
sheepshead minnows was significantly reduced (Hall et al., 1991).
Results from the first test conducted in 1991 demonstrated
biological effects using a different endpoint with the same fish
species. Growth of larval sheepshead minnows was significantly
reduced after 8 d of exposure.
Contaminants data from this station during the first test did
not provide absolute data for establishing "cause and effect"
relationships. Copper was reported at potentially stressful
concentrations of 5.5 ug/L during the first test and 4.2 ug/L
during the second test. Both values exceeded the U.S. EPA marine
water quality criteria for this metal (U. S. EPA, 1987a). These
copper concentrations were higher than values reported from this
station in 1990 (Hall et al., 1991). A nickel concentration of 7.4
ug/L was also reported during the first test in 1991. This
concentration was similar to the 8.4 ug/L concentration reported at
this station during the 1990 study. Although the value reported in
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1991 was approximately 1 ug/L lower than the chronic water quality
criteria of 8.3 ug/L, it may have been stressful. Comparison c
other contaminants data between the two years showed the absence of
mercury in 1991 in contrast to the reported value of 0.29 ug/L in
1990. Neither organic contaminants or water quality conditions
appeared to be stressful in either year for species used for water
column testing.
Data from the 1990 sediment toxicity tests indicated low
levels of toxicity (Kali et al., 1991). Grass shrimp and worms
showed no significant difference in survival or growth compared to
their control groups. However, amphipods exhibited a statistically
significant reduction in survival (70% survival at day 10 and 20%
survival at day 20) . The Morgantown sediment was approximately 92%
sand, and was therefore very similar to the native amphipod
sediment, so differences in mortality cannot be attributed tc
particle size effects (see Figure 4.5). No toxic levels of
inorganic contaminants or unionized ammonia were found in this
sediment. Organic analysis was not conducted in 1990 due to
funding limitations.
The results of the 1991 sediment toxicity tests again gave an
indication of low ambient toxicity levels. Although there were no
significant effects with the grass shrimp, worm or Lecidactvlus
dytiscus tests, Hvalella azteca showed significantly reduced
survival in Set 1 tests in the Morgantown sediments. The organic
contaminant analysis indicated levels of the pesticide 4,4'-DDT
exceeding the Long and Morgan ER-L for expected toxic effects. No
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significant concentrations of inorganic contaminants were
identified.
The suborganismal (biomarker) component of the two year
ambient toxicity pilot program has not shed additional light on the
condition of the Potomac River Morgantown Station, other than the
high hepatic copper concentrations documented in white perch from
the 1991 sampling program. As stated previously, the copper
observation in white perch is not unique to Morgantown or the
Potomac River in general. Fundulus heteroclitus were not available
at this station due to unsuitable habitat, and therefore
monooxygenase monitoring with Fundulus while useful at other
stations, was not fruitful here. We also did not encounter any
spot during our trawling or beach seining activities at this
station.
5.2 Potomac River - Dahlaren
This station is located downstream from the Morgantown station
and represents a typical mesohaline area in the Potomac River.
Data from our 1990 water column testing demonstrated reduced
survival of larval sheepshead minnows after 8 day exposures.
Results from the 1991 study also demonstrated effects with this
larval fish; a significant reduction in growth was reported during
the first test. Contaminants data provided minimal insight as to
possible causes of biological effects during the first test.
Copper concentrations of 3.4 ug/L exceeded the marine water quality
criteria for this metal. Low concentrations of both cadmium (0.68
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cadmium (0.68 ug/L) and chromium (3.1 ug/L) were also reported but
values were well below the water quality criteria for each metal.
Metals data from 1990 showed the presence of lead (3.6 and 5.6
ug/L) at this station; concentrations of this metal were not
detected in the 1991 study. Neither organic contaminants or water
quality parameters were reported at potentially toxic
concentrations during the 1990 or 1991 water column testing.
Results of the 1990 sediment toxicity tests indicate a
significant reduction in amphipod survival (22% survival), but no
significant differences in shrimp and worm survival or growth (Hall
et al., 1991). Contaminant analysis showed no evidence of toxic
inorganic contaminants. Pore water analysis did show relatively
high levels of unionized ammonia (0.159 mg/L), just below the EPA
acute criteria of 0.233 mg/L for a one hour average of unionized
ammonia. Organic analysis identified one pesticide, two semi-
volatiles, and 11 tentatively identified semi-volatile compounds
present in the sample.
The 1991 sediment toxicity tests resulted in reduced survival
of Hvalella azteca in Set 1 tests at 10 and 20 days, and
significantly reduced survival of the worms in the set 1 sediments
after a 10 day exposure. Since the worms are sensitive to particle
size composition, the percent survival was adjusted to account for
sediment particle size effects. The worm survival in Dahlgren
sediment was still significantly lower (82% adjusted survival) than
survival in the controls (100% adjusted survival). In Set 2
sediment tests, both Hvalella azteca and Leoidactvlus dvtiscus
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showed significant mortalities, and Lepidactvlus dvtiscus growth
was significantly reduced. There were no significant differences
in the grass shrimp survival or growth in either set of sediment
tests. No high concentrations were observed for the sets of
organic or inorganic chemical contaminants which were analyzed.
The same general observations made above for the Morgantown
station also applied to suborganissai data at the Dahlgren station.
White perch was the most abundant species available and did not
show any apparent contaminant related effects on EROD activity.
5.3 Patapsco River
The Patapsco River station was selected to represent an area
where toxic conditions were suspected based on effluent toxicity
data from a local industrial facility. Selection of this station
prior to the 1990 study was not based on previous data
demonstrating biological effects in this ambient area. Results
from our first year of water column testing showed biological
effects from our larval grass shrimp test. Biological effects were
not reported for the other two species (E. affinis and sheepshead
minnows). Toxicity data from the 1991 study were similar to the
data from these two tests. Results from four water column tests
did not suggest the presence of toxic conditions at this station.
Concentrations of metals measured from this station do not
suggest the presence of potentially toxic conditions with the
possible exception of copper and nickel. Copper concentrations of
3.7 ug/L reported from both 1991 tests exceeded the marine water
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criteria quality (U.S. EPA, 1987a). The value was also higher than
concentrations (< 2.0 ug/L) reported in 1990. Nickel
concentrations ranging from 5.2 to 8.2 ug/L were detected in all
three samples collected during the 1991 study. The value of 8.2
ug/L is only slightly less than the water quality criteria of 8.3
ug/L (U. S. EPA, 1987a). This value is also similar to the
concentration of 10.8 ug/L reported in 1990 (Hall et al., 1991).
These data suggest that nickel was consistently present at this
station in 1990 and 1991. Organic contaminants and water quality
parameters were not reported at potentially toxic levels for
species in water column tests during the 1991 study. A similar
finding was also reported in 1990.
The 1990 sediment toxicity tests demonstrated toxic conditions
at this site (Hall et al., 1991). Survival of two of the sediment
test species, Lepidactvlus dvtiscus and the worm, was significantly
reduced after 10 and 20 days of exposure to sediments collected
from this site. Survival of the grass shrimp was not reduced
compared to the controls. Test organisms exposed to Patapsco River
sediments exhibited no significant reduction in growth. Although
the Lepidactvlus dvtiscus showed no significant difference in
growth at this station, they did show a significant difference in
reburial in clean sediments at the end of the 20-day test. Ninety-
nine percent of the control animals reburied, while only 75% of the
surviving amphipods from the Patapsco River sediments reburied when
placed into clean sediment, indicating a level of toxicity capable
of producing behavioral effects. In addition, observations during
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the test indicated that a high percentage of the organisms were
emerged from the sediments or swimming in the overlying water
throughout the test period. Elevated levels of zinc, chromium, and
lead were observed in the Patapsco River sediments. Chromium
(157.0 ug/g) and zinc (603.1 ug/g) were found in levels exceeding
the ER-M, while lead (55.8 ug/g) was observed at concentrations
which exceeded the ER-L level. The high levels of these metals is
consistent with the various toxic responses exhibited at this site.
Results of the 1991 sediment tests confirm the findings of the
1990 study. Significantly reduced survival was seen with both
Hvalella azteca and Lepidactvlus dvtiscus at 10 and 20 days for
both Set 1 and Set 2 tests. In addition, both amphipod species
tested in Set 2 exhibited significantly reduced growth in those
animals that survived. There were no significant effects seen with
the grass shrimp exposed to Patapsco River sediments in set 1 or
set 2 tests. Lepidactvlus dvtiscus demonstrated significantly
reduced reburial for both Set 1 and Set 2 tests. One hundred
percent of the control animals reburied, while only 89.7% (Set 1)
and 82.9% (Set 2) of the surviving amphipods from the Patapsco
River sediments reburied when placed into clean sediment,
indicating behavioral effects on the surviving organisms. In
addition, observations during the test indicated that a high
percentage of Lepidactvlus dvtiscus and Hvalella azteca were
emerged from the sediments or swimming in the overlying water
throughout the test period. This high percentage of emergence is
abnormal for both species, and indicates an avoidance response. It
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is possible that fewer mortalities may have occurred because the
animals were swimming in the water column and not actually in
contact with the sediment throughout much of the test period.
Elevated levels of zinc, chromium, and lead were again
observed in the Patapsco River sediments during 1991. Chromium
(107.8 ug/g Set 1 and 130.6 ug/g Set 2), while lead (37.8 ug/g)
were found in levels exceeding the ER-L, and zinc (328.6 and 455.3
ug/g) concentrations exceeded the ER-M level. The high levels of
these metals are consistent with the various toxic responses
exhibited at this site. The high percentage of sand in the
sediments found at this and several other sites (Table 3.4),
generally coincides with lower AVS and TOC values for that
sediment. Lower AVS and TOC tend to increase toxic responses to
metals over sites with similar metal contamination, but lesser
sediment sand fractions. The bioavailability of the metals
frequently becomes diminished with increasing AVS and TOC. The
SEM:AVS calculations indicated a ratio greater than 1.0, which
predicts the potential for metal toxicity. The high level of zinc
(5.7 umoles/g and 5.8 umoles/g, Sets 1 and 2 respectively) in the
extractable metal phase and its rather high sulfide solubility
parameter (DiToro et al., 1991) corroborates the speculation that
zinc may be responsible for a large percentage of the toxicity
exhibited. No toxic levels of organic contaminants were
identified.
Suborganismal data collected from spot at the Patapsco station
during 1991 continued to show evidence of induced EROD activity,
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particularly in comparison to a reference sample taken in the
Patuxent River in November. Hepatic EROD activities, however, were
not as high as levels found in 1990. Mean EROD activities were
approximately two-fold higher in August than in September. These
activities are presumed to reflect the planar aromatic hydrocarbon
contamination in the Patapsco (Helz and Huggett, 1987). However,
we do not have data to ascribe differences between August and
September to differences in environmental PAH concentrations.
There has been considerable debate as to whether elevations in
EROD activity reflect a "toxic" response. This is due, in part, to
the fact that measures at the cellular level are often made outside
the context of toxic effects measured as depressed growth and/or
increased mortality. What can be stated with confidence is that
increases in cellular responses which can mechanistically be
related to specific classes of contaminants (such as EROD and
planar aromatic hydrocarbons), reflect an exposure which is high
enough for cell or organ system response. That is, if you consider
the cell, or in this case hepatic EROD activity, as a biosensor
which will only respond if exposure exceeds some threshold level,
then these fish have been exposed to concentrations which exceed
this threshold. Furthermore, if cellular responses in fact precede
overt toxic effects, along some dose response curve, then
elevations can be considered a "toxic" effect. Clearly, whether
this toxic effect has any relationship to effects which may be
propagated as compromises in population vitality (presumably
through effects on growth and reproduction) requires further
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research.
5.4 Wve River
The Wye River was selected at the beginning of this two year
study to represent a "relatively clean" pristine reference area
without point source discharges. This station was located in a
rural area where the major land-use is dominated by agricultural
activity. Results from our water column testing in 1990 did not
demonstrate the presence of toxic conditions in this area although
both sediment and suborganismal testing did suggest adverse
conditions (Hall et al., 1991). In contrast to our 1990 study,
water column data from the present study demonstrated the presence
of toxic conditions during both tests using two different species.
A significant reduction in survival (25%) was reported for
sheepshead minnow larvae during the first test. This finding was
surprising since the sheepshead minnow is considered to be a
moderately resistant species. The ambient water collected on the
first day of Test 1 (8/13/91) had a pungent foul odor and a darker
hue when compared with samples collected from other ambient areas.
Survival of sheepshead minnow larvae was reduced to 38% after 24 h
of exposure in this water. We examined the ambient water sample
used during the first day of testing under a microscope and found
dominant presence of a Gvmnodinium dinoflagellate when compared
with other plankton. Another water sample was collected from the
Wye River the next day (8/14/91) and another test was initiated.
This sample did not have a pungent odor and dominant presence of
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Gvmnodinium were not reported. Less than 5% mortality was reported
from this test thus suggesting that some constituent in the Wye
River water sample collected on 8/13/91 was responsible for the
significant reduction in survival of sheepshead minnow larvae. One
possible explanation was the presence of the Gvmnodinium
dinoflagellate. Other investigators have reported that small
Gvmnodinium-like species are responsible for mortality of various
fish species (striped bass, spot and summer flounder, among others)
held in the laboratory (Jo Ann Burkholder, personal communication,
North Carolina State University). This dinoflagellate causes
adverse effects in fish more readily than invertebrates. Based on
conversations with Dr. Burkholder, it appeared that the Gvmnodinium
species in the Wye River sample was similar to the species she has
reported. Current research by Dr. Burkholder is designed to
identify the trigger for Gvmnodinium blooms in the environment.
She suspects nutrient loading (likely phosphorous) as the primary
factor responsible for the blooms. These data have important
implications for ambient testing in the Wye River which is heavily
influenced by nutrient loading.
Results from the second test at this station in 1991 also
demonstrated significant biological effects with a different test
species. Significant reductions in survival were reported for E.
affinis after 8 d of exposure in Wye River water. Gvmnodinium was
not suspected as a cause of mortality with this species as
dominant numbers of this dinoflagellate were not observed in water
collected during this test.
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Copper (2.9 ug/L) and nickel (9.3 ug/L) concentrations
exceeding U. S. EPA marine water quality criteria during the first
test in 1991 nay have been stressful to sheepshead minnow larvae
and produced a "joint toxicity" response concurrently with the
Gvmnodinium species (U.S. EPA, 1987a). Both of these metals also
exceeded water quality criteria during our 1990 testing at this
station thus suggesting their continued presence over a one year
period. None of the metals measured during the second test
exceeded the U.S. EPA water quality criteria although detectable
concentrations of both chromium (2.9 ug/L) and copper (2.6 ug/L)
were reported. Potentially toxic organic contaminants and water
quality conditions were not reported at stressful concentrations in
the water column for tests conducted during the 1990 or 1991 study.
However, it is important to note that the semi-volatile organic
contaminant 4 methyl phenol was reported in the sediment at this
station during the first test in 1991. Concentrations of this
organic contaminant were below values expected to be toxic but it
may have been an indicator of other potentially toxic conditions.
Due to the polar nature of this soluble organic, it would be
expected to be more of a potential toxicant in water and not
sediment (see discussion below).
In the 1990 sediment toxicity tests, survival of the grass
shrimp was > 99% for both the 10- and 20-day tests, and showed no
significant differences in growth as compared to the controls (Hall
et al., 1991). The worm and Lepidactvlus dvtiscus tests, however,
exhibited significantly lower survival. The amphipod had only 57%
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survival at the Wye Station after 10 days of exposure and 24%
survival after 20 days. Amphipod growth (mean length) was
significantly lower than the controls. The results of the worm
bioassays were similar to those of the amphipod tests. There was
statistically reduced survival of the worms in the Wye River
sediments after 10 and 20-day exposures. The worm showed no
significant growth effects. Since two of the three species tested
(the amphipod and the worm) showed significant reduction in
survival at the Wye River station, it is possible that there was
some undetected contaminant present that contributed to the toxic
response. Organic contaminants were not analyzed in the Wye
sediments in 1990 due to funding limitations.
The sediment samples collected at the Wye River in 1991 were
very different sediment from those tested, in 1990. The sample
tested in 1990 was 80% silt/clay. The sediment samples tested in
1991 were 93.7% sand (6.3% silt/clay) and 82.2% sand (17.8%
silt/clay) for sets 1 and 2, respectively. In addition, the 1991
samples had only about 0.4% organic carbon as compared to 7.8% TOC
in 1990. No significant effects were observed in the grass shrimp
or amphipod tests in the Wye River sediments. The 1991 sediment
did produce a significant reduction in survival of the worms for
set 1 test sediments. It is suggested that this increased
mortality may have been due to starvation rather than toxicity of
the sediment. After adjustment of the data to compensate for
particle size induced mortalities, there was still a significant
reduction in survival (67%) as compared to the controls (100%). No
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other significant responses of sediment test organisms were
recorded.
Very low levels of inorganic contaminants were found in the
sediment in 1991. Organic analysis indicated the presence of 4-
methylphenol, although not in significant concentrations. It
should be noted, however, that this contaminant is considerably
more polar and soluble than most of the semi-volatile organic
contaminants analyzed, so would not be expected in high
concentrations in the sediments. However, its presence may be
indicative of organic contaminants in the water, which may have
been associated with transient events. This organic compound is
associated with wood treatment and preservatives, and may have been
introduced during the construction of a new bridge slightly
downstream from the tidally influenced sampling site. Field
technicians noted the strong smell of creosote when passing under
this structure in transit to the collection site.
Hepatic EROD activities in Fundulus from the Wye River in 1991
were less that those found in 1990 and less than levels found in
the St. John's Creek on the Patuxent River. Levels of EROD
activity tended to be higher in September than in August, though
variability masked any statistical differences. EROD activities in
Wye River Fundulus were also higher than EROD activities in
Fundulus from the same location held in the laboratory over several
months. Based on the two years of data, it appears as though the
Wye River station may contain enough planar aromatic hydrocarbons
(probably PAH) to cause elevations in hepatic EROD activities.
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EROD activity is highly induced by components of creosote. The
observations made above regarding 4-methylphenol and creosote
suggest that these Fundulus may in fact be responding to these
contaminants.
5.5 Laboratory Studies on Hepatic EROD Activity in Fundulus
In 1991, laboratory studies were conducted to determine the
time necessary for EROD activities, which are elevated in field
sampled fish, to decrease to baseline levels. In addition,
laboratory experiments were conducted to determine if two different
populations of Fundulus would respond similarly to challenge with
a known planar aromatic hydrocarbon inducer. These experiments
were conducted with Fundulus from the Wye River and from St.
John* s Creek in the Patuxent River.
It took approximately one month for hepatic EROD activities to
fall approximately six-fold in Fundulus from St. John's Creek.
EROD activity in Fundulus from the Wye River decreased the first
week, but then elevated 2-3 fold from week one to two. Levels
returned to initial collection levels at week three and continued
to decline thereafter. While there is no immediate explanation for
this behavior, we have hypothesized that it may have been due to a
contamination event which came in through the Chesapeake Biological
Laboratory Patuxent River intake lines. Wye River Fundulus were
not sampled on the same dates as those from St. John's Creek and
did not show this spike in activity.
There was no significant difference in the response of the two
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different Fundulus populations to challenge with benzo(a)pyrene.
Both were modestly induced 96 hours after a 2 mg/kg treatment and
significantly induced by 10 mg/kg. Additional studies demonstrated
that gonadal maturity of female Fundulus did not prevent the
hepatic EROD induction response following treatment with B(a)P.
This indicates that studies using Fundulus EROD activity as a
measure of planar aromatic hydrocarbon exposure will not be
compromised by the use of females during the reproductive season.
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SECTION 6
CONCLUSIONS AFTER TWO YEARS OF TESTING
This two year pilot study was designed to use a suite of water
column, sediment and suborganismal tests on a limited spatial and
temporal scale to identify toxic ambient areas in the Chesapeake
Bay watershed. In 1990, the first year of testing, eight stations
from the following areas were tested once: Elizabeth River,
Patapsco River, Wye River, Potomac River-Indian Head, Potomac
River-Freestone Point, Potomac River-Possum Point, Potomac River-
Morgantown and Potomac River-Dahlgren. Results from these three
types of tests demonstrated the presence of toxic conditions in
various environmental media in suspected contaminated areas such as
the Elizabeth River and the Patapsco River, and also in critical
habitat areas in the Potomac River (Table 6.1, Fig. 6.1) (Hall et
al., 1991). One of the more surprising results from the first year
of testing was the data reported from sediment and suborganismal
tests in the suspected toxic-free area of the Wye River. Both
sediment and suborganismal tests suggested the presence of toxic
conditions in this area while water column tests failed to
demonstrate any adverse biological effects during the 1990 testing
(Table 6.1). Results from sediment tests at the Freestone Point and
Possum Point Stations in the Potomac River also suggested the
presence of toxic conditions. Neither water column or
suborganismal tests showed the presence of toxic conditions at
these stations. The data collected in 1990 demonstrate the
119
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Table 6.1 Toxicity data from water column, sediment and suborganismal tests conducted
in 1990. Effects were noted if significantly different than controls for water
column and sediment tests. The following abbreviations were used for various
effects: NE «= No Effect; M « Mortality; RG = Reduced Growth; RR = Reduced
Reproduction; RB = Reduced Reburial; IE = Induced EROD Activity; RE = Reduced EROD
Activity and LA = Liver Abnormality. Station abbreviations were as follows: ER
= Elizabeth River; PR = Patapsco River; WR = Wye River; IH = Indian Head: FP
= Freestone Pt; PP = Possum Pt; MT = Morgantown and DA = Dahlgren.
Station
EB EE WE IH FP PP MT DA
Test Type
Water Column
E. affinis
M
NE^
NE^
NE^
NE^
NE^
NE^
NE^
£. puqio
M
M
NE
NE
NE
NE
NE
NE
C. varieaatus
NE
NE
NE
-
-
-
M
M
CeriodaDhnia so.
—
-
-
M, NE
NE
NE
—
-
P. Dromelas
NE
NE
NE
•
Sediment
L. dvtiscus
M
M, RB
M, RG
M
M
M
M
M
S. benedicti
M
M
M
NE
NE
NE
NE
NE
£. pyqio
M
NE
NE
NE
NE
RG
NE
NE
>uboraanismal^
Fundulus - water
M
RE
NE
—
_
—
NE
NE
Fundulus - sediment
M
NE
NE
NE
NE
NE
NE
NE
Feral Fish
IE
IE
IE
NE
NE
NE
LA
NE
~Low control reproduction limits the use of this endpoint; therefore, NE is based on
survival.
~~Qualitative determinations.
-------
Indian Head
1990 Ambient Toxicity Study
Patapsco River
Water Column Sediment
Water Cflhimn Sediment
Wye River
Freestone Point
Water fiftiiwrm Sediment
Water Column Sediment
Possum Point
Water Column Sediment
Water Cnhimn Sediment
Elizabeth River
Morgantown
Water Cnhimn Sediment
Water Column Sediment
Pie Chart Key
Toxic
Effect
No Effect
Fig. 6.1 Toxic effects reported from water column and sediment toxicity
tests in 1990. The shaded portion of the pie—charts indicate
the percent of the total tests in which the endpoints were
significantly different from the controls (p C 0.05). There
were at least 2 endpoints for each species.
121
-------
importance of using a battery of tests when attempting to identify
toxic conditions in the ambient environment.
Ambient toxicity testing in 1991 (year 2) was expanded on a
temporal scale to include tests at two different time periods
(summer and fall) but reduced on a spatial scale to only four of
the eight stations tested in 1990 (Patapsco River, Wye River,
Potomac River-Morgantown and Potomac River-Dahlgren). Biological
effects were demonstrated from water column tests during at least
one sampling period for all stations except the Patapsco River
(Table 6.2, Fig. 6.2). The most persistent biological effects were
reported from the Wye River station, as significant mortality from
two different water column test species was reported from both the
\
first (summer) and second (fall) tests. Sediment tests
demonstrated significant biological effects for both summer and
fall tests at the Dahlgren, Morgantown and Patapsco River stations
(Table 6.2, Fig. 6.2). Significant biological effects were
reported in sediment during the first test at the Wye River but not
the second. Suborganismal tests suggested that the Wye River and
the Patapsco River stations contained concentrations of planar
aromatic hydrocarbon contaminants (e.g. PAH) sufficient to cause
elevations of hepatic EROD activity in Fundulus and spot,
respectively. Laboratory tests with Fundulus from the Wye River
and St. John's Creek demonstrated that there were no population
differences in the ability of Fundulus to respond to aromatic
hydrocarbon EROD inducers. White perch from the Morgantown and
Dahlgren stations contained high concentrations of copper in their
122
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Table 6.2 Toxicity data from water column and sediment tests conducted in 1991. Test 1 and
Test 2 were conducted August 13-21 and September 24-October 2, respectively.
Effects were noted if significantly different than controls. The following
abbreviations were used: NE = No Effect, M «= Mortality, RG = Reduced Growth, RB
= Reduced Reburial. Stations were DG = Dahlgren; MT = Morgantown; Wye = Wye, and
PA = Patapsco.
Test Type
DG
Test 1
Test 2
Station
MT
Test 1 Test 2
and Test
Wye
Test 1
Test 2
Test 1
PA
Test 2
Water Column
E. affinis
NE
NE
NE
NE
NE
M
NE
NE
Sheepshead
EG
NE
EG
NE
M
NE
NE
NE
Grass Shrimp
NE
NE
NE
NE
NE
NE
NE
NE
Mysid Shrimp
NE*
NE
NE*
NE
NE*
NE
NE*
NE
Sediment
L. dvtiscus
NE
M, RG
NE
M
NE
NE
M
M,RG,RB
P. Ducrio
NE
NE
NE
NE
NE
NE
NE
NE
H. azteca
M
M
M
NE
NE
NE
M
M, RG
S. benedicti
M,RG
NE
M, RG
NE
M, RG
NE
M,RG
M
~Mortality was
the only
parameter
evaluated
-------
1991 Ambient Toxicity Study
Morgan town
Water column Sediment
Water mhimn Sediment
CMMONO
Pie Chart Key
Toxic
Effect
No Effect
Patapsco River
Water column
_ Sediment
Wye River
Water column
Srrllmrnt
\
v )
Fig. 6.2 Toxic effects reported from water column and sediment toxicity
tests in 1991. The shaded portion of the pie-charts indicate
the percent of the total number of tests in which the endpoints
were significantly different from controls (p 0.05). There
were at least 2 endpoints for each species.
124
-------
livers.
A comparison of water column and sediment data from 1991
testing shows general agreement at both Potomac River stations
during test l. One out of four test species showed a toxic effect
with water column testing; two out of four tests showed a toxic
effect with sediment. Water column and sediment data for test 2 at
both Potomac River stations are less similar. No effects were
reported at either station for water column tests. For sediment
tests, significant toxic effects for two test species were reported
at Dahlgren and one test species at Morgantown. Water column and
sediment data at the Wye River station for test 1 showed a general
agreement as a significant biological effect was reported from one
test species for each test type. Water column tests showed
significant biological effects from one test species for the second
test while sediment tests showed no effects for any of the four
test species. Water column and sediment tests demonstrated
drastically different results at the Patapsco River station. Ho
effects were reported for water column tests during either test;
sediment tests (3 out of 4) demonstrated significant effects during
both test 1 and 2.
The same result should not necessarily be expected with
concurrent water column and sediment tests nor should samples
collected at different times give the same result. Toxic
conditions in water tend to be more transient and may be related to
periodic discharges of toxic materials or episodic events (i.e.,
runoff of contaminants after precipitation events). Conversely,
125
-------
toxic conditions in sediments are more persistent, as this media
serves as a sink for contamination, while water column toxicity may
represent current problems. Of course, sediments can be a non-
point source of environmental contamination and can serve as a
source for toxic conditions to over-lying water. Toxicity data
generated from concurrent water column and sediment testing provide
insight into areas that need additional investigation.
In general, the results from suborganismal hepatic EROD
activity in feral fish from the Wye River and the Patapsco River
agreed most closely with the sediment test results. The fish
species we used are both benthos feeding fishes and would
presumably receive most of their contaminant exposure through this
route. Thus, while we are not suggesting that in all cases these
two types of tests should concur, it is not unreasonable to suggest
that sediment toxicity tests and sublethal measures in benthos
feeding fishes might agree qualitatively, when the contaminants
causing effects are primarily particle associated.
The need for multispecies testing within each type of test is
clearly demonstrated for both years of testing. During the first
year of testing, two water column tests (E. affinis and sheepshead
minnow larvae) conducted at the Patapsco River station showed no
significant effects. However, the third test species (larval grass
shrimp) demonstrated reduced survival at this station (Table 6.1).
The same scenario existed for sediment tests conducted at all
Potomac River stations in 1990. The worm tests showed no effects
at any of the 5 stations and the grass shrimp demonstrated no
126
-------
effects at 4 of the stations. In contrast, reduced survival of
amphipods was observed at all 5 stations.
Water column data from 1991 demonstrated significant effects
(reduced survival) from the sheepshead minnow larval test during
test 1 at both Dahlgren and Morgantown stations. The other three
test species showed no effects. Two different test species
(sheepshead minnows and E. affinis) showed effects at each of the
two Wye River tests. Sediment tests generally demonstrated effects
from more than one species when effects were reported. For
example, adverse biological effects were reported from at least two
test species during Dahlgren Test 1 and 2, Morgantown Test 1 and
Patapsco River Test 1 and 2. A biological response from sediment
tests was only reported from one test species during Morgantown
Test 2 and Wye River Test 1.
A summary of species responses from water column tests in 1990
and 1991 is presented in Table 6.3. The sheepshead minnow larval
test was the most sensitive estuarine water column test followed by
the E. affinis and the larval grass shrimp test. The Ceriodaphnia
test was more sensitive than the fathead minnow test based on a
limited number of freshwater water column tests. Although these
data suggest a possible ranking of sensitivity among estuarine
water column tests, it is important to note that if only the most
sensitive species (sheepshead minnow) was used, biological effects
would not have been reported at the Elizabeth River station during
1990 and the Wye River station (test 2) during 1991. These data
strongly support the need for multispecies testing.
127
-------
Table 6.3 Summary of water column test species responses from 1990
and 1991 ambient toxicity testing.
Test Species
Number of
Tests
Number of
Tests With
Significant
Effects
% of Tests with
Significant
Effects
E. affinis
16
2
12.5
Sheepshead Minnow
16
5
31
Grass Shrimp
16
2
12.5
Mysid Shrimp
8
0
0
CeriodaDhnia sd
(freshwater)
4
1
25
P. promelas
(freshwater)
3
0
0
128
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A summary of species responses from sediment tests in 1990 and
1991 is presented in Table 6.4. Sediment tests were generally
found to be more sensitive in detecting toxic conditions than water
column tests (See Table 6.3). The amphipod (L. dvtiscus) was the
most sensitive species followed by H. azteca (freshwater), S.
benedicti. and P. puaio (juvenile grass shrimp). The rationale for
multispecies testing with sediment was not as obvious during the
1990 testing as significant effects were found at all eight
stations with the most sensitive test species (L. dvtiscus).
However, 1991 data illustrate three examples (Dahlgren Test 1,
Morgantown Test 1 and Wye River Test 1) where L dvtiscus detected
no biological effects while other test species did show significant
effects. These data demonstrate the importance of multispecies
testing with sediment.
Suborganismal testing provided several important pieces of
information during 1990 and 1991. Hepatic monooxygenase activity
(EROD) in feral Fundulus and spot provided a useful biological
confirmation of what is already documented regarding PAH
contamination of the Elizabeth and Patapsco River systems.
Surprisingly, EROD activity in feral Fundulus from the Wye River
suggested a level of contamination not previously suspected. These
two species appear to be suitable for further surveys of this type.
EROD activity in feral white perch did not appear to provide
the same type of information as spot and Fundulus. We caution,
however, that we were not able to sample all three species at each
station and therefore do not have adequate comparative information
129
-------
Table 6.4 Summary of sediment test species responses from 1990
and 1991 ambient toxicity testing.
Test Species
Number of
Tests
Number of
Tests With
Significant
Effects
% of Tests with
Significant
Effects
L. dvtiscus
16
12
75
S. benedicti
16
8
50
P. puaio
16
2
12.5
H. azteca
(freshwater)
8
5
62.5
130
-------
(one of the difficulties any survey of fishes from broad areas of
the Chesapeake will encounter, is the lack of overlap in suitable
habitat for different species). We did note that white perch had
generally low hepatic EROD activities, hepatic concentrations of
copper were extremely high. Until further laboratory studies are
conducted with white perch to demonstrate their responsiveness to
aromatic hydrocarbon monooxygenase inducers in the presence and
absence of high hepatic copper, they are not suitable for surveys
of the type attempted here.
Evaluation of hepatic monooxygenase activity in Fundulus
exposed to test water and sediment in 1990 did not necessarily
reflect what was seen in the feral fish. We therefore omitted
these tests in 1991. Lack of concordance between measurements of
EROD activity in feral Fundulus and of activity in Fundulus exposed
to sediments from these same areas may not be surprising. If
sediment contaminants are the main source of exposure, either
through direct ingestion of sediments or through consumption of
organisms that have been contaminated by these sediments, then the
way these tests were conducted may have been inappropriate.
Fundulus were exposed to test sediment in aquaria only after
indigenous organisms had been removed. The fish were fed
exogenously using a commercial feed. Therefore, the only exposure
they would have received would have been via desorption of
contaminants from the sediments into the overlying water column.
These aqueous concentrations were apparently not high enough to
cause induction to the levels seen in feral fish, though they were
131
-------
high enough to cause mortality using Elizabeth River sediments. In
this scenario, the use of feral fish is clearly supported.
Additional observations from laboratory studies suggest that
Fundulus from different geographic areas should respond similarly
to aromatic hydrocarbon inducers, regardless of stage of sexual
maturity. Animals collected from the field should be held at least
one month in order to reduce the influence of "background"
elevation of hepatic EROD activity.
The "apparent temporal pattern" in ambient toxicity can be
evaluated on a very limited basis using water column and sediment
data from four stations tested once in 1990 and twice in 1991
(Tables 6.1 and 6.2). It is not possible to state conclusively
that "apparent temporal patterns" are beyond the environmental
"year to year" variability. Water column data from the Wye River
showed a drastic change in the presence of toxic conditions between
the two years. No biological effects were reported from any of the
three test species in 1990 but significant effects were found
during both tests in 1991. Water column data from the Patapsco
River suggested contrasting results. Biological effects were
reported from the grass shrimp test in 1990 but no effects were
reported from either test 1 or test 2 in 1991 using any of the four
test species. A comparison of data generated from water column
tests at Morgantown and Dahlgren stations over a one year period
demonstrated different results each year. A severe biological
effect was reported in 1990 with sheepshead minnow larvae
(significant mortality), a less severe sublethal effect was
132
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reported with this species during test 1 of 1991 (reduced growth)
and no effects with any of the test species were found during the
second test in 1991.
Apparent temporal patterns in sediment toxicity data over the
three testing periods during the two years (1990-1991) showed a
decrease in toxicity at the Wye River station. Adverse biological
effects from two species were reported in 1990, effects from one
species were reported during the first test in 1991 and no effects
were reported from the second test in 1991. Sediment toxicity data
from the Patapsco River demonstrated persistent toxic conditions
over the two year period. Two tests detected toxic conditions in
1990 while three tests suggested toxic conditions in 1991.
Sediment toxicity data from both the Morgantown and Dahlgren
stations also demonstrated persistent biological effects. Data
from either one or two test species suggested the presence of toxic
conditions from the three sampling periods in 1990 through 1991.
There were distinct temporal differences in EROD activity in
feral fish. Fundulus from the Wye River had lower EROD activity in
1991 than 1990 and activity was similar during August and September
of 1991. Spot from the Patapsco River had lower EROD activities in
1991 than 1990, with the lowest activity found in September of
1991. In general, this suggests that exposures to contaminants
were less during 1991 at these two stations.
Water quality and contaminants evaluations conducted in water
and sediment during this pilot study provided supportive
information on possible causes of biological effects but these data
133
-------
were not intended to provide conclusive data on specific Mcause and
effect" relationships. In 1990, the highest number of potentially
toxic contaminants in both water and sediment were reported at the
station (Elizabeth River) where biological effects from the
greatest number of tests were reported. A similar scenario existed
for the second most toxic area tested in 1990 (Patapsco River).
The relationship between contaminants data and biological data at
the Wye River station was not positively correlated. Both sediment
and suborganismal tests suggested toxic conditions while water
column tests did not. Toxic inorganic contaminants were not
detected in the Wye River sediment and organic evaluations were not
conducted in this media. In contrast, two potentially toxic metals
were reported in the water column at the Wye River station.
Contaminants data from the Potomac River stations in 1990
provided various examples of biological effects with and without
the identification of potentially toxic contaminants (Hall et al.,
1991). One potentially toxic concentration of a metal (cadmium)
was reported in the water column at the Indian Head and Freestone
Point station and four to five potentially toxic metals
concentrations (cadmium, mercury, nickel, lead and zinc) were
measured in the sediment at each of these stations (Hall et al.,
1991). Results from one water column test and one sediment test
demonstrated toxic conditions at Indian Head. However, only one
sediment test showed toxic conditions at Freestone Point.
Potentially toxic concentrations of metals were not found in the
water column at Possum Point but five potentially toxic metals were
134
-------
found in the sediment (cadmium, mercury, nickel, lead and zinc).
The amphipod test (sediment exposure) demonstrated reduced survival
while no significant effects were reported during water column or
suborganismal tests. Potentially toxic contaminants were not
reported in sediment from the Morgantown or Dahlgren station. Two
metals exceeded U.S. EPA water quality criteria at Morgantown
(mercury and nickel) and no potentially toxic contaminants were
reported in the water column at Dahlgren. Significant effects were
reported from water column and sediment tests at both Morgantown
and Dahlgren.
A comparison of a biological effects data with the limited
contaminant evaluations for water column tests in 1991 demonstrated
examples of both good and poor linkages between these types of
data. A good correlation was reported for the first Wye River test
as significant mortality was observed for sheepshead minnow larvae
(August 13 test) concurrently with the presence of two potentially
toxic metals (copper and nickel). As previously discussed in
Section 5.4, the presence of a potentially toxic Gvmnodinium
dinoflagellate may have also contributed to the mortality of the
test species. The second Wye River test is an example of a poor
correlation between biological effects and the presence of toxic
contaminants. Significant mortality of E. affinis was reported but
potentially toxic contaminant conditions were not identified.
Data collected from the Patapsco River station in 1991
provided a strong case for the link between biological effects in
sediment and the presence of potentially toxic contaminants.
135
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Chromium, lead and zinc were reported at potentially toxic
concentrations in sediment concurrently with biological effects
from three of the four sediment test species. Data from the
Morgantown station also provided an example of biological effects
in sediment (adverse effects for two tests) concurrently with the
presence of potentially toxic concentrations of 4,4 •- DDT. The
Dahlgren station (Test 1 and 2) and Wye River (Test 1) provided
examples of biological effects without the identification of
potentially toxic contaminant conditions.
One of the primary conclusions from this two year pilot study
is that integrated water column and sediment toxicity testing is a
feasible approach for assessing toxic conditions in ambient areas
of the Chesapeake Bay watershed. Suborganismal testing provided
additional important insights, though it did not reveal any toxic
conditions where water column and/or sediment tests were negative.
Technical constraints on the rate at which suborganismal testing
can provide data per unit cost, compared to the water column and
sediment tests, lead us to conclude that this testing framework may
not be optimal for the inclusion of suborganismal testing. While
sublethal effects measurements in indigenous organisms may clearly
play a role in determining the conditions of a habitat with respect
to certain forms of contamination, they require an experimental
design separate from the design of water column and sediment tests.
The pilot phase has allowed us to determine the most feasible
tests for accomplishing our goals and discard those tests that are
less promising. Detecting toxic conditions in areas suspected to
136
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be toxic such as the Elizabeth River was expected but this step was
necessary to "ground truth" our test approach. One of our initial
concerns using this test approach was the sensitivity of the test
methods in detecting toxic conditions in areas that may exhibit low
levels of contamination. Results from both the Wye River and
Potomac River stations demonstrated that our test methods (species)
were sensitive enough to detect low levels of contamination. The
next step in the developmental process of this ambient toxicity
test approach is to determine the variability of test results on
both a spatial and temporal scale and assess the relative toxicity
of ambient water and sediment using statistical methods. The third
year of this research effort has been designed to address the
question of spatial and temporal variability on a limited scale
using a battery of sensitive water column and sediment toxicity
tests. Statistical analysis will also be conducted to determine
the relative toxicity of water and sediment from ambient areas and
provide a degree of confidence to observed differences between
ambient areas and reference areas (controls).
The final developmental phase of this test approach will be to
consolidate and evaluate data from all three years of the pilot
study and then assess the nature and magnitude of any remaining
uncertainties in the methodology. The third year of the pilot
study began in July 1992. If the ambient toxicity methodology is
determined to be a useful approach, the Bay program or any
interested state agency could use the methodology to design a
multi-year Bay-wide or regional assessment of ambient toxicity in
137
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important living resource habitats. The ambient toxicity
methodology developed during the three year pilot study could also
be added to an array of multi-metric assessment tools that the
Chesapeake Bay Program is beginning to develop and use in targeting
tributaries and watersheds for nonpoint source monitoring and
remediation. The goals of this targeting effort will be to
determine where management-based habitat improvement programs
should be focused, based on the status of biological communities
and other environmental indicators. Follow-up studies could then
be designed for any areas of the Bay watershed where these multi-
metric assessments (including the ambient toxicity approach)
demonstrate evidence of toxic conditions to purse "cause and
effect" relationships, identify sources of toxic conditions, assess
the degree of impact to living resources communities and develop
control strategies.
138
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SECTION 7
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145
-------
APPENDIX A
The Percent of Females Carrying Eggs
After 8-d Exposures in the
Eurvtemera affinis Tests
-------
TEST #1
TEST *2
(8/21/91)
H-W
Control
i
X % + S.E.
A
5/10
50
57.9 ± 4.8
B
8/12
66.7
C
4/7
57. 1
Dahlcrren
A
4/11
36.4
B
6/7
85.7
62.9 ± 14.4
C
6/9
66.7
Pataosco
A
8/13
61.5
B
4/11
36.4
49.3 ± 7.3
C
5/10
50
Wve
A
5/8
62.5
B
3/8
37.5
42.9 ± 10.1
C
2/7
28.6
Moraantown
A
3/11
27.3
B
6/10
60
41.6 ± 9.7
C
3/8
37.5
(10/2/91)
H-W
Control
i
X % + SE
A
4/8
50
B
1/11
9.1
39.7 + 15,
C
6/10
60
IR Control
A
3/13
23.1
B
4/7
57.1
46.7 ± 11
C
9/15
60
PataDsco
A
5/7
71.4
B
4/13
30.8
34.1+20
C
0/1
0
Dahlcrren
A
9/16
56.3
B
6/8
75
57.1+10
C
4/10
40
Wye
A
2/5
50
B
6/8
75
46.7 + 3
C
4/10
40
Moraantown
A
B
6/11
7/11
54.5
63.6
59.1 + 4
-------
APPENDIX B
Water Quality Conditions Reported in Test
Beakers During all Water Column
Tests
-------
Experiments conducted with Palaemonetes puaio (pg), Cvprinodon
varieaatus (Cv), Eurvtemora affinis (Ea) and Mvsidopsis bahia (Mb)
using water samples collected from the Wye River (WR), Patapsco
River (PR) , Morgantown (MT) and Dahlgren (DG) from 8/13/91 to
8/21/91 and 9/24/91 to 10/2/91.
Date
8/14/91
8/15/91
Test
Station
Temp
Sal
DO
vH
Species
(ppt)
(mg/L)
Pp
WR
26.2
14
9.0
8.6
PR
26.3
15
8.8
8.4
MT
25.3
16
5.6
8.1
DG
25.2
17
6.9
8.0
Control
26.9
15
6.2
8.2
Cv
WR
25.0
14
5.8
8.0
PR
24.8
15
7.8
8.3
MT
24.7
16
6.5
8.2
DG
24 .9
17
6.2
8.2
Control
24.8
15
6.5
8.2
Ea
WR
26.0
14
5.2
8.1
PR
26.0
15
7.1
8.5
MT
26.0
15
7.2
8.3
DG
16.0
17
6.4
8.2
Control
25.8
15
7.3
8.3
Mb
WR
25.0
14
6.0
7.8
PR
25.0
15
6.6
7.9
MT
25.0
16
5.9
8.1
DG
25.2
17
5.2
7.8
Control
24.8
15
5.1
8.1
Pp
WR
24.9
15
6.1
8.0
PR
24.9
20
8.9
8.4
MT
25.1
17
6.7
8.0
DG
25.1
19
6.6
8.0
Control
24.5
18
7.3
8.1
Cv
WR
25.4
14
5.2
7.9
PR
25.3
16
7.7
8.2
MT
25.2
16
6.6
7.9
DG
25.3
18
6.6
7.9
Control
25.7
15
6.2
8.0
Ea
WR
25.3
14
5.5
7.8
PR
25.3
16
7.5
8.2
MT
25.6
16
6.9
8.1
DG
25.4
18
7.4
8.2
Control
25.8
15
7.9
8.3
B-l
-------
Date Test Station Temp Sal 00 pH
Species (ppt) (mg/L)
8/15/91
8/16/91
8/17/91
Mb
WR
25.8
14
4.3
7.5
PR
25.0
16
6.6
7.7
MT
25.4
16
5.3
7.8
DG
25.7
18
5.0
7.7
Control
25.6
15
5.5
7.9
Pp
WR
24.9
15
4.4
7.7
PR
25.6
18
6.9
8.2
MT
24.7
18
6.2
7.7
DG
25.0
19
5.2
7.7
Control
25.4
16
5.4
8.0
Cv
WR
25.0
14
4.7
7.7
PR
25.0
16
6.7
8.2
MT
24.7
16
5.1
7.8
DG
25.0
18
5.3
7.7
Control
25.1
15
4.8
8.1
Ea
WR
25.8
14
4.6
7.6
PR
25.1
16
6.7
8.1
MT
25.2
17
7.4
7.9
DG
25.0
18
7.4
8.0
Control
25.2
15
6.4
8.2
Mb
WR
24.9
14
4.3
7.6
PR
24.9
16
5.0
7.8
MT
24.6
16
4.8
7.5
DG
24.7
18
4.4
7.6
Control
25.1
15
4.8
8.0
Pp
WR
25.7
16
8.8
8.3
PR
25.9
17
6.5
8.1
MT
25.6
15
5.8
7.9
DG
25.0
17
5.6
7.9
Control
25.4
16
5.6
8.1
Cv
WR
25.0
15
4.7
7.9
PR
25.0
16
5.6
8.1
MT
25.0
15
5.6
7.8
DG
25.1
16
5.8
7.9
Control
25.0
15
5.1
8.0
Ea
WR
24.8
15
5.4
7.9
PR
24.9
16
7.0
8.1
MT
24.9
15
7.1
8.1
DG
24.8
16
7.1
8.1
Control
25.0
15
7.7
8.2
B-2
-------
Date Test Station Temp Sal DO pH
Species (ppt) (mg/L)
8/17/91
Mb
WR
PR
MT
DG
Control
25.1
25.0
25.0
25.1
25.1
15
15
15
16
15
4.1
4.7
4.4
4.3
5.2
7.7
7.7
8.0
7.6
7.9
8/18/91
Pp
WR
PR
MT
DG
Control
25.1
25.1
25.3
25.0
25.1
16
17
16
17
16
7.5
6.9
6.1
5.7
6.8
8.1
8.1
7.9
7.9
8.2
Cv
WR
PR
MT
DG
Control
25.0
25.0
24.9
25.0
25.0
15
16
15
16
15
5.1
4.7
5.6
5.8
4.7
7.7
7.8
7.8
7.8
8.0
Ea
WR
PR
MT
DG
Control
25.0
25.2
24.9
25.0
25.1
15
16
15
16
15
6.1
6.8
6.8
6.8
7.0
8.0
8.2
8.1
8.1
8.2
Mb
WR
PR
MT
DG
Control
24.8
25.0
25.0
24.8
25.0
15
16
15
16
15
4.0
5.2
5.4
4.8
5.0
7.5
7.6
7.8
7.6
8.0
8/19/91
Pp
WR
PR
MT
DG
Control
25.0
26.0
25.0
25.0
25.5
15
16
15
15
15
6.6
8.4
7.7
6.6
9.7
7.9
8.3
6.3
7.7
8.7
Cv
WR
PR
MT
DG
Control
25.0
25.2
25.1
25.0
25.3
14
15
14
15
15
6.0
6.5
6.4
5.9
5.8
7.9
8.1
8.0
7.9
8.2
Ea
WR
PR
MT
DG
Control
25.3
26.0
25.4
25.6
14
14
15
15
6.2
6.5
6.8
7.9
8.0
8.0
8.5
B-3
-------
Date Test Station Temp Sal DO pH
Species (ppt) (mg/L)
8/19/91
Mb
WR
PR
MT
DG
Control
25.0
25.0
24.9
24.9
25.5
14
15
14
15
15
4.9
5.0
5.5
4.8
5.1
7.7
7.7
7.8
7.7
8.0
8/20/91
Pp
WR
PR
MT
DG
Control
25.3
25.4
26.8
25.2
25.2
15
15
14
14
15
6.7
9.0
6.3
6.7
9.7
8.2
8.6
7.9
8.1
8.7
Cv
WR
PR
MT
DG
Control
25.0
24.9
25.0
24 .9
25.0
14
15
14
14
15
6.6
6.4
7.0
6.9
5.5
8.0
8.0
8.0
8.0
7.9
Ea
WR
PR
MT
DG
Control
24.9
25.0
24.9
24.8
25.3
14
14
14
14
15
6.6
6.9
6.6
6.9
7.5
8.0
8.3
8.0
8.0
8.4
Mb
WR
PR
MT
DG
Control
25. l
25.1
25.1
25.2
25.2
14
15
14
14
15
5.3
4.9
6.1
5.8
6.0
7.8
7.7
8.0
7.7
8.1
8/21/91
Pp
WR
PR
MT
DG
Control
25.4
26.0
25.8
24.9
25.0
13
14
14
13
15
6.9
7.0
7.4
7.7
5.9
8.2
8.5
8.0
8.2
8.7
Cv
WR
PR
MT
DG
Control
24
24
24
24
24
14
14
13
13
15
6.9
7.0
7.4
7.7
5.9
7.9
8.0
8.0
8.2
8.2
Ea
WR
PR
MT
DG
Control
25.5
25.0
26.0
25.0
25.0
14
14
12
13
15
6.5
6.8
6.2
6.7
7.9
7.9
7.9
7.9
7.9
8.4
B-A
-------
Date Test Station Temp Sal DO pH
Species (ppt) (mg/L)
8/21/91
9/25/91
9/26/91
Mb
WR
26.0
14
10.8
8.5
PR
25.8
14
6.9
8.0
MT
24.9
13
7.1
8.2
DG
25.4
13
8.2
8.2
Control
25.0
15
11.0
8.7
Pp
WR
24.5
15
6.3
7.9
PR
24.5
15
6.1
7.8
MT
24.5
12
6.3
7.8
DG
25.0
13
6.5
7.9
Control
25.0
15
6.2
8.3
Cv
WR
25.0
15
6.8
8.1
PR
25.0
16
7.4
8.1
MT
25.0
12
6.4
7.8
DG
25.0
12
6.5
8.0
Control
25.0
15
6.6
8.3
Ea
WR
25.5
15
7.3
8.0
PR
25.0
15
7.5
8.0
MT
25.0
13
7.2
8.0
DG
25.0
14
7.2
8.0
H-W
Control
25.0
15
7.1
8.3
I-R
Control
25.5
15
7.0
8.0
Mb
WR
25.0
15
6.7
8.0
PR
24.5
16
6.8
7.9
MT
24.5
12
6.0
7.8
DG
25.0
12
6.4
7.8
Control
25.0
15
5.9
8.3
Pp
WR
24.5
15
6.2
7.9
PR
24.0
16
6.3
7.8
MT
24.0
14
5.9
7.8
DG
24.0
12
6.0
7.9
Control
24.0
15
5.8
8.3
Cv
WR
25.0
15
6.0
8.0
PR
25.0
15
7.5
8.2
MT
25.0
14
5.8
7.8
DG
25.0
13
5.6
7.9
Control
25.0
15
5.5
8.3
Ea
WR
25.0
15
5.8
7.9
PR
25.5
16
6.7
8.0
MT
24.0
14
6.3
8.0
DG
25.0
13
6.1
8.0
H-W
Control
25.5
15
6.8
8.3
I-R
Control
25.0
15
6.8
7.9
B-5
-------
Date Test Station Temp Sal DO pH
Species (ppt) (mg/L)
9/26/91 Mb WR 24.0 15 5.7 7.9
PR 24.5 16 6.2 7.8
MT 24.5 14 5.9 7.8
DG 25.0 14 5.8 8.1
Control 24.0 15 5.1 8.2
9/27/91 Pp WR 23.0 15 6.7 8.0
PR 23.5 17 6.4 7.9
MT 23.5 15 6.8 7.9
DG 24.0 15 6.8 7.9
Control 24.0 16 6.8 8.4
Cv WR 24.0 15 6.5 7.8
PR 24.5 15 8.8 8.1
MT 24.5 15 5.4 7.7
DG 25.0 14 5.6 7.8
Control 24.0 15 6.1 8.2
Ea WR 24.0 15 7.2 7.8
PR 24.0 15 7.8 7.8
MT 25.0 15 7.2 7.8
DG 25.0 14 7.4 8.0
H-W Control 24.0 15 8.0 8.3
I-R Control 24.0 15 7.7 7.8
Mb WR 24.0 15 6.1 7.6
PR 24.0 15 6.5 7.7
MT 24.0 14 5.7 7.7
DG 24.0 14 6.1 7.8
Control 24.0 15 5.8 8.1
9/28/91 Pp WR 22.0 15 6.4 7.9
PR 22.0 16 6.6 7.8
MT 22.5 15 6.9 7.9
DG 22.5 15 6.9 7.9
Control 22.0 15 6.8 8.4
CV WR 24.5 15 4.1 7.8
PR 25.0 15 5.5 7.9
MT 24.0 15 4.9 7.6
DG 24.5 14 4.6 7.6
Control 24.0 15 4.4 8.2
Ea WR 25.0 15 6.5 8.0
PR 25.0 15 6.6 8.0
MT 24.5 15 6.8 8.0
DG 24.5 14 7.0 8.0
H-W Control 24.5 15 7.5 8.5
I-R Control 24.5 15 7.5 8.1
B-6
-------
Date Test Station Temp Sal DO pH
Species (ppt) (mg/L)
9/28/91
9/29/91
Mb
WR
24.0
15
4.4
7.6
PR
24.0
16
5.3
7.6
MT
25.0
15
4.2
7.5
DG
24.0
14
5.0
7.7
Control
24.0
15
4.7
8.2
Pp
WR
24 . 0
15
6.9
8.0
PR
24.0
16
6.7
7 . 9
MT
24.0
16
7.0
7.9
DG
24.0
15
7.0
8.0
Control
24.0
16
7.2
8.4
Cv
WR
24.5
15
4.2
7.6
PR
24.5
15
5.3
7.7
MT
24.0
14
4.6
7.5
DG
24.0
14
4.2
7.6
Control
25.0
15
4.5
8.2
Ea
WR
24.5
14
7.0
8.0
PR
24.5
14
6.9
8.0
MT
24.5
15
7.2
8.1
DG
24.5
15
7.1
8.0
H-W
Control
24.0
15
7.5
8.4
I-R
Control
24.0
15
7.3
8.1
Mb WR 24.0 15 4.1 7.6
PR 24.0 15 4.6 7.6
MT 24.5 15 4.4 7.5
DG 24.0 14 4.8 7.6
Control 24.0 15 4.4 8.1
9/30/91 Pp
Cv
Ea
WR
25.0
16
6.9
8.1
PR
25.0
17
7.2
8.1
MT
26.5
16
7.1
8.0
DG
26.0
15
7.5
8.1
Control
25.0
17
6.8
8.4
WR
24.5
14
4.3
7.6
PR
24.0
14
6.1
7.7
MT
24.5
14
4.4
7.5
DG
24.5
14
5.9
7.7
Control
24.0
15
5.2
8.2
WR
24.0
15
7.3
8.1
PR
24.0
15
7.5
8.1
MT
24.5
14
7.4
8.0
DG
24.5
15
7.3
8.0
Control
23.0
15
7.9
8.5
Control
23.5
15
7.9
8.0
B-7
-------
Date Test Station Temp Sal DO pH
Species (ppt) (og/L)
9/30/91
10/1/91
10/2/91
Mb
WR
25.0
15
4.4
7.6
PR
24.0
14
4.9
7.6
MT
24 .0
14
4.4
7.5
DG
24 .0
14
5.6
7.7
Control
24.0
15
4.-3
8.1
Pp
WR
25.0
16
5.4
7.9
PR
25.1
16
5.9
7.9
MT
25.8
16
6.0
7.9
DG
25.6
16
6.1
8.0
Control
25.0
16
6.2
8.4
Cv
WR
25.8
15
7.3
8.3
PR
25.5
15
5.4
7.9
MT
25.8
15
4 . 1
7.6
DG
25.6
15
4.9
7.7
Control
25.7
15
4.4
8.2
Ea
WR
26.0
15
8.2
6.2
PR
25.9
15
6.2
8.1
MT
26.0
15
6.2
8.1
DG
25.7
15
6.4
8.1
H-W
Control
25.9
15
6.6
8.5
I-R
Control
25.9
15
6.9
8.0
Mb
WR
26.0
15
4.1
7.6
PR
25.6
15
3.9
7.5
MT
25.1
15
3.7
7.5
DG
25.2
15
4.2
7.5
Control
25.8
15
4.0
8.1
Pp
WR
25.0
16
6.5
8.0
PR
25.0
16
6.9
8.0
MT
25.0
16
7.0
8.1
DG
25.0
16
7.0
8.0
Control
25.0
16
6.6
8.4
Cv
WR
-
-
-
-
PR
25.5
15
6.2
7.8
MT
25.5
15
7.1
8.0
DG
25.5
15
8.0
8.2
Control
26.0
15
6.2
8.3
Ea
WR
25.0
15
6.5
o
•
CO
PR
-
-
—
-
MT
25.5
15
7.1
8.0
DG
25.0
15
7.4
8.5
H-W
Control
25.0
15
7.4
8.5
I-R
Control
26.5
15
7.5
7.8
B-8
-------
Date Test Station Temp Sal DO pH
Species (ppc) (mg/L)
10/2/91 Mb WR 25.0 15 6.0 7.8
PR 25.5 15 4.7 7.7
MT 25.0 15 5.9 7.7
DG 25.5 15 7.1 7.8
Control 25.0 15 6.2 8.3
B-9
-------
APPENDIX C
Summary of Water Quality Monitored
During Sediment Toxicity Tests
-------
Summary of water chemistry monitored during sediment toxicity tests, Set 1
Set *1
n,
i
Species
Grass shrimp
P. puaio
Amphipod
£. dvsticus
Amphipod
H. azteca
Worm
£. benedicti
Water Chemistry Range fmin-max)
Temp
(•C)
<>2
(mg/1)
PH
Day
Day
Day
Day
Day
Day
Site
1-10
10-20
1-10
10-20
1-10
10-20
Control
24-25
23-24
5.8-7.4
4.4-7.2
7.8-8.
. 1
7.
,5-8.2
Morgantown
24-25
23-24
6.5-7.0
5.5-7.3
7.8-8.
. 1
7.
.6-8.1
Oahlgren
24-25
23-25
5.3-6.9
3.4-7.2
7.8-8.
. 1
7.
.6-8.1
Patapsco
24-25
23-24
6.2-6.8
2.6-6.9
7.7-7.
,9
7.
, 3-8.0
Wye
24-25
24-24
6.3-7.0
5.3-7.1
7.8-8.
.0
7 .
.7-8.0
Reference
24-25
24-24
6.3-7.0
4.9-7.2
7.9-8.
. 1
7,
.6-8.1
Control
23-25
25
7.0-7.5
6.5-7.8
8.3-8.
.0
8.
.3-8.1
Morgantown
23-25
24-25
7.0-7.6
6.7-7.7
8.2-8.
.0
8.
.3-8.1
Dahlgren
22-25
24-25
6.8-7.6
6.7-7.7
8.3-8,
.0
8.
.4-8.1
Patapsco
23-25
24-25
6.5-7.5
6.5-7.6
7.8-8,
.0
7,
.7-8.0
Wye
22-25
24-25
7.0-7.6
5.3-7.6
7.8-8.
.2
7.
.7-8.1
Reference
23-25
24-25
6.5-7.8
6.7-7.8
8.0-8.
.2
7.
.9-8.1
Control
24-25
24-24
6.6-7.4
6.8-7.5
8.1-8.
.2
7.
.9-8.4
Morgantown
24-25
24-25
6.9-7.4
5.4-7.8
8.0-8.
. 3
8.
.0-8.3
Dahlgren
24-25
24-24
6.7-7.4
6.8-7.9
8.0-8.
.3
8.
.0-8.3
Patapsco
24-25
24-25
6.8-7.4
6.8-7.9
7.8-8.
.2
8.
.0-8.2
Wye
24-25
24-25
5.8-7.4
7.0-7.9
7.9-8.
, 1
8.
,0-8.3
Reference
24-25
24-25
7.0-7.5
6.2-7.9
8.0-8.
, 3
8.
. 1-8.3
Control
24-25
24-24
6.6-7.4
6.8-7.5
8.1-8.
,2
7.
.9-8.4
Morgantown
24-25
24-25
6.9-7.4
5.4-7.8
8.0-8.
.3
8.
.0-8.3
Oahlgren
24-25
24-24
6.9-7.4
6.8-7.9
8.0-8.
,3
8.
,0-8.3
Patapsco
24-25
24-25
6.8-7.4
6.8-7.9
7.8-8.
. 2
8.
,0-8.2
Wye
24-25
24-25
5.8-7.4
7.0-7.9
7.9-8.
, 1
8.
,0-8.3
Reference
24-25
24-25
7.0-7.5
6.2-7.9
8.0-8.
3
8.
1-8.3
-------
Summary of water chemistry monitored during sediment toxicity tests. Set 2.
Set #2
Water Chemistry Range (min-max)
Temp
CC)
o2
(mg/1)
PH
Day
Day
Day
Day
Day
Day
SDecies
Site
1-10
10-20
1-10
10-20
1-10
10-20
Grass shrimp
Control
24-25
23-24
5.6-7.8
5.5-7.4
7.7-8.2
7.7-8.1
P. ouqio
Morgantown
24-25
24-25
6.2-7.2
6.2-6.9
7.7-8.2
7.7-8.0
Dahlgren
24-25
24
6.2-7.0
4.9-6.8
7.9-8.2
7.7-8.1
Patapsco
24-25
24
5.5-6.6
4.2-7.3
7.6-8.1
7.6-8.1
Wye
24-25
24
6.3-7.0
6.3-7 . 1
7.8-8.2
7.7-8.0
Reference
24-25
24-25
5.5-7.2
6.2-7.2
7.8-8.2
7.8-8.1
Amphipod
Control
25-27
25
6.9-7.3
6.9-7.8
7.9-8.3
7.8-8.3
L. dvsticus
Morgantown
25-27
25
6.8-7.3
6.9-7.4
7.9-8.3
7.8-8.3
Dahlgren
25-27
25
5.5-7.0
6.9-7.2
7.9-8.4
7.6-8.2
Patapsco
25-27
25
6.7-7.2
6.5-7.3
7.7-8.1
7.6-8.1
Wye
25-27
25
5.7-7.3
6.7-7.4
7.7-8.2
7.8-8.1
Reference
25-27
25
6.7-7.1
6.5-7.3
7.9-8.2
7.7-8.2
Amphipod
Control
24
24
6.4-7.6
6.6-7.6
7.9-8.4
7.8-8.2
H. azteca
Morgantown
24
24
6.4-7.9
6.0-7.4
7.9-8.3
7.8-8.3
Dahlgren
24
24
6.8-7.5
1.9-7.6
8.1-8.6
7.6-8.2
Patapsco
24
24
6.7-7.4
6.3-7.3
7.9-8.2
7.7-8.1
Wye
24
24
6.8-7.6
6.6-7.6
7.8-8.4
7.8-8.1
Reference
24
24
6.7-7.5
5.3-7.8
7.9-8.3
7.7-8.2
Worm
Control
24
24
6.4-7.6
6.6-7.6
7.9-8.4
7.8-8.2
Morgantown
24
24
6.4-7.9
6.0-7.4
8.0-8.3
7.8-8.3
Dahlgren
24
24
6.8-7.5
1.9-7.6
8.0-8.3
7.6-8.2
Patapsco
24
24
6.7-7.4
6.3-7.3
7.8-8.2
7.7-8.1
Wye
24
24
6.8-7.6
6.6-7.6
7.9-8.1
7.8-8.1
Reference
24
24
6.7-7.5
5.3-7.8
8.0-8.3
7.7-8.2
-------
APPENDIX D
Hvalella azteca Reference
Sediment Tests
-------
BIOASSAY DATA SUMMARY
Laboratory: AMRL
Sample No.: NA
Sample Type: Sediment
Collection Date: 10-21-91
Time: 1030
Dilution Water Used: D.I. Water & Sea Salts
Test Temperature: 25°C +/- 2° C
Test Mode: Static Acute Non-Renewal
Test Duration: 10 and 20 days
Contractor: AMRL/In house
Effluent/Sediment Source: Lynnhaven
Test Organism: Amphipod
Species: Hvalella azteca
Age: Juveniles
Start of Test: Date: 11-5-91 Time: 1400
End of Test: Date: 11-15-91 Time: 1400
Test Technicians: PB, RT
Data Released By: E. Deaver
RESULTS
Water Chemistry and Toxicity Analyses
(Range- max/min)
Concentration Temp. Dissolved Oxygen pH S Percent
% (°C) (mg/L) (ppt) Survival
Sand A
Sand B
Sand C
Sand D
Sand E
CTRL (mud)
25/24
25/24
25/24
25/24
25/24
25/24
7.8/6.7
7.7/6.4
7.8/6.5
7.6/6.4.,
7.8/6.4
7.5/6.4
8.1/7.8
8.0/7.7
8.1/7.5
8.1/7.6
8.0/7.4
8.2/7.8
15
15
15
15
15
15
92.5
86.5
81.3
92.5
88.8
91.0
KEY: All test treatments were 100% sand. "Sterilized" refers to heat sterilized and dried
to kill predators. "Aged" means the sand was placed into a tank of 15 ppt water(DI and
sea salts) & aerated for 2 weeks. Algae (isochrysis) and artemia were added 3x/week.
Sand A- Sterilized, then aged (All 100% sand for the Amb. Tox. study was treated this way).
Sand B= Sterilized, aged, sterilized
Sand C- Sterilized, no aging
Sand D= No sterilization. Field collected, then aged
Sand E= Aged, then sterilized
CTRL* Control sediment. Lynnhaven mud, 68% sand (same sediment used as Control for Amb Tox)
-------
APPENDIX E
Organics
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
SEMI-VOLATILE COMPOUNDS
Laboratory: Organics
Project ID: Ambient Toxicity
Sample ID: Morgantown
Dates:
Collected:
Received:
Method:
Analyst:
8/14/91
8/15/91
3550/8270
RJM
Matrix: Sediment
Sample w/v: 30.04
Contractor: MD DNR
Sample No.: 37400
Extracted: 8/16/91
Analyzed: 9/4/91
Instrument: INCOS 50
Data Released By: M. Helmstetter
Units: Mg/kg dry
% Moisture: 27%
CAS No.
Compound
Detection
Cone. Tag Limit
No compounds detected.
E-l
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
SEMI-VOLATILE COMPOUNDS
Laboratory:
Project ID:
Sample ID:
Dates:
Collected:
Received:
Method:
Analyst:
Organics
Ambient Toxicity
Dahlgren
8/14/91
8/15/91
3550/8270
RJM
Matrix: Sediment
Sample w/v: 30.02
Contractor: MD DNR
Sample No.: 37401
Extracted: 8/16/91
Analyzed: 9/4/91
Instrument: INCOS 50
Data Released By: M. Helmstetter
Units: pg/kg dry
% Moisture: 29%
CAS No.
Compound
Detection
Cone. Tag Limit
No compounds detected.
E-2
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
SEMI-VOLATILE COMPOUNDS
Laboratory:
Project ID:
Sample ID:
Dates:
Collected:
Received:
Method:
Analyst:
Organics
Ambient Toxicity
Patapsco
8/14/91
8/15/91
3550/8270
RJM
Matrix: Sediment
Sample w/v: 30.07
Contractor: MD DNR
Sample No.: 37402
Extracted: 8/16/91
Analyzed: 9/4/91
Instrument: INCOS 50
Data Released By: M. Helmstetter
Units: ng/kg dry
% Moisture: 26%
CAS No.
Compound
Detection
Cone. Tag Limit
No compounds detected.
E-3
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
SEMI-VOLATILE COMPOUNDS
Laboratory:
Project ID:
Sample ID:
Organics
Ambient Toxicity
Wye
Contractor: MD DNR
Sample No.: 37403
Dates.:
Collected:
Received:
8/14/91
8/15/91
Extracted: 8/16/91
Analyzed: 9/4/91
Method:
Analyst:
3550/8270
RJM
Instrument: INCOS 50
Data Released By: M. Helmstetter
Matrix:
Sample w/v:
Sediment
30.05
Units: /*g/kg dry
% Moisture: 24%
CAS No.
Compound
Detection
Cone. Tag Limit
106-44-5
4-methylphenol
9.9 J 13.9
J = Compound detected below the calculated method detection limit.
E-4
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
SEMI-VOLATILE COMPOUNDS
Laboratory:
Organics
Contractor: MD DNR
Project ID:
Ambient Toxicity
Sample No.: 37522
Sample ID:
Reference
Pates:
Collected:
8/14/91
Extracted: 8/16/91
Received:
8/15/91
Analyzed: 9/4/91
Method:
3550/8270
Instrument: INCOS 50
Analyst:
RJM
Data Released By: M. Helmstetter
Matrix:
Sediment
Units: PgAg dry
Sample w/v:
29.98
% Moisture: 35%
Detection
CAS No. Compound Cone. Tag Limit
No compounds detected.
E-5
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
SEMI-VOLATILE COMPOUNDS
Laboratory:
Organics
Contractor: MD DNR
Project ID:
Ambient Toxicity
Sample Nc.: 37523
Sample ID:
Control
Dates:
Collected:
8/14/91
Extracted: 8/16/91
Received:
8/15/91
Analyzed: 9/4/91
Method:
3550/8270
Instrument: INCOS 50
Analyst:
RJM
Data Released By: M. Helmstetter
Matrix:
Sediment
Units: ^g/kg dry
Sample w/v:
30.00
% Moisture: 15%
Detection
CAS No. Compound Cone. Tag Limit
117-81-7 bis(2-ethylhexyl)phthalate 63.9 12.5
E-6
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
SEMI-VOLATILE COMPOUNDS
Laboratory:
Organics
Contractor: MD DNR
Project ID:
Ambient Toxicity
Sample No.: 37841
Sample ID:
Wye River
Dates:
Collected:
9/25/91
Extracted: 10/24/91
Received:
9/26/91
Analyzed: 11/11/91
Method:
3550/8270
Instrument: INCOS 50
Analyst:
RJM
Data Released By: M. Helmstetter
Matrix:
Sediment
Units: PgAg dry
Sample w/v:
30.12 g
% Moisture: 15%
Detection
CAS No.
Compound
Cone. Tag
Limit
(^g/kg)
541-73-1
1,3-Dichlorobenzene
14.0 B
11.9
10544-50-0
Sulfur, Mol. (S8)
6.01X103'
10.0
B = Compound detected in the QC blank.
* = Tentative identification. Concentration based on 1:1 response with internal standard.
E-7
-------
APPENDIX F
Pesticides
-------
Table 1. Semi-volatile organics (GC/MS) and pesticide/Aroclor
(GC/ECD) compounds evaluated in sediments employing a user-created
calibration library. Sediment detection limits are reported in
jig/kg dry weight.
Sediment
CAS #
ComDound
MDL
62-53-3
aniline
14.5
95-57-8
2-chlorophenol
13.2
111-44-4
bis(2-chloroethyl)ether
11.2
108-95-2
phenol
11.9
541-73-1
1,3-dichlorobenzene
11.9
106-46-7
1,4-dichlorobenzene
12.5
95-50-1
1,2-dichlorobenzene
11.9
100-51-6
benzyl alcohol
27.1
39638-32-9
bis(2-chloroisopropyl)ether
5.9
95-48-7
2-methylphenol
15.8
91-57-6
2-methylnaphthalene
9.2
67-72-1
hexachloroethane
21.8
621-64-7
n-nitroso-di-n-propylamine
13.2
106-44-5
4-methylphenol
13.9
98-95-3
nitrobenzene
11.2
78-59-1
isophorone
6.6
88-75-5
2-nitrophenol
27.1
65-85-0
benzoic acid
18.5
105-67-90
2,4-dimethylphenol
15.8
111-91-1
bis(2-chloroethoxy)methane
9.9
120-83-2
2,4-dichlorophenol
21.8
120-82-1
1,2,4-trichlorobenzene
15.2
91-20-3
naphthalene
4.6
106-47-8
4-chloroaniline
26.4
87-68-3
hexachlorobutadiene
22.4
59-50-7
4-chloro-3-methylphenol
20.5
77-47-4
hexachlorocyclopentadiene
25.7
88-06-2
2,4,6-trichlorophenol
37.0
95-95-4
2,4,5-trichlorophenol
44.9
91-58-7
2-chloronaphthalene
9.9
88-74-4
2-nitroaniline
37.6
208-96-8
acenaphthalene
5.9
84-66-2
dimethylphthalate
9.9
606-20-2
2,6-dinitrotoluene
48.2
99-09-2
3-nitroaniline
247
83-32-9
acenaphthene
9.9
51-28-5
2,4-dinitrophenol
262
132-64-5
dibenzofuran
7.9
100-02-7
4-nitrophenol
268
121-14-2
2,4-dinitrophenol
43.6
86-73-7
fluorene
9.9
7005-72-3
4-chlorophenylphenylether
20.5
84-66-2
diethylphthalate
9.9
100-01-6
4-nitroaniline
279
534-52-1
4,6-dinitro-2-methylphenol
122
86-30-6
n-nitrosodiphenylamine
19.1
101-55-3
4-bromophenylphenylether
41.6
F-l
-------
Table 1 (Cont'd):
CAS «
ComDound
MDL
85-01-8
phenanthrene
9.2
118-74-1
hexachlorobenzene
37.6
87-86-5
pentachlorophenol
136
120-12-7
anthracene
9.9
84-74-2
di-n-butylphthalate
5.9
206-44-0
fluoranthene
10.6
129-00-0
pyrene
10.6
85-68-7
butylbenzylphthalate
17.8
56-55-3
benzo(a)anthracene
17.8
218-01-9
chrysene
14.5
91-94-1
3,3'-dichlorobenzene
101
117-81-7
bis(2-ethylhexy)phthalate
12.5
117-84-0
di-n-octylphthalate
7.3
205-99-2
benzo(b)fluoranthene
13.9
207-08-9
benzo(k)fluoranthene
13.9
50-32-8
benzo(a)pyrene
15.2
193-39-5
indeno(1,2,3-cd)pyrene
16. 5
53-70-3
dibenz(a, h)anthracene
17.8
191-24-2
benzo(g,h,i)perylene
16. 5
103-33-3
azobenzene
7.3
92-87-5
benzidine
24.4
1912-24-9
atrazine
0.100
391-84-6
a-BHC
0.714
391-85-7
B-BHC
0. 559
391-86-8
6-BHC
1.052
58-89-9
lindane
0.616
76-44-8
heptachlor
0.819
309-00-2
aldrin
0.608
1024-57-3
heptachlor epoxide
0. 570
959-98-8
endosulfan I
0.859
60-57-1
dieldrin
0.898
72-55-9
4,4'-DDE
0.528
33213-65-9
endosulfan 11
0.745
72-20-8
endrin
1.240
72-54-8
4,4'-DDD
0.469
1031-07-8
endosulfan sulfate
1.500
50-29-3
4,4'-DDT
3.420
72-43-5
methoxychlor
5.0
57-74-5
chlordane
5.0
80001-35-2
toxaphene
10.0
2385-85-5
mi rex
1.000
7421-93-4
endrin aldehyde
2.410
12574-11-2
Aroclor 1016
16.6
11104-28-2
Aroclor 1221
16.6
11141-16-5
Aroclor 1232
16.6
53469-21-9
Aroclor 1242
16.6
12672-29-6
Aroclor 1248
16.6
11097-69-1
Aroclor 1254
16.6
11096-82-5
Aroclor 1260
16.6
37324-23-5
Aroclor 1262
16.6
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
PESTICIDE AND AROCLOR COMPOUNDS
Laboratory: Organics
Project ID: Ambient Toxicity
Sample ID: Morgantown
Dates:
Collected:
Received:
Method:
Analyst:
8/14/91
8/20/91
3550/8080
RJM
Data Released By: M. Helmstetter
Matrix: Sediment
Contractor: MD DNR
Sample No.: 37400
Extracted: 8/15/91
Analyzed: 9/12/91
Instrument: PE 9000
Calib/MIDPT Date: 8/23/91
Units:
pg/kg dry
CAS No.
Compound
Cone.
Detection
Tag Limit
(^g/kg)
391-84-6
a-BHC
0.793
0.714
391-85-7
fi-BHC
133
0.559
58-89-9
Lindane
1.38
0.616
76-44-8
Heptachlor
1.18
0.819
60-57-1
Dieldrin
1.40
0.898
1031-07-8
Endosulfan sulfate
2.32
1.50
50-29-3
4,4' -DDT
7.01
3.420
F-3
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
PESTICIDE AND AROCLOR COMPOUNDS
Laboratory:
Organics
Contractor: MD DNR
Project ID:
Ambient Toxicity
Sample No.: 37401
Sample ID:
Dahlgren
Dates:
Collected:
8/14/91
Extracted: 8/15/91
Received:
8/20/91
Analyzed: 9/12/91
Method:
3550/8080
Instrument: PE 9000
Analyst:
RJM
Calib/MIDPT Date: 8/23/91
Data Released By: M. Helmstetter
Matrix:
Sediment
Units: Mg/kg dry
Detection
CAS No. Compound Cone. Tag Limit
(Mg/kg)
391-84-6
391-85-7
cr-BHC
B-BHC
1.00
0.702
0.714
0.559
F-A
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
PESTICIDE AND AROCLOR COMPOUNDS
Organics
Ambient Toxicity
Patapsco
R ¦' ] -1 /QI
8/20/91
3550/8080
RJM
Laboratory:
Project ID:
Sample ID:
Cciicciec:
Received:
Method:
Analyst:
Data Released By: M. Helmstetter
Matrix: Sediment
Contractor: MD DNR
Sample No.: 37402
Extracted: 8/15/91
Analyzed: 9/12/91
Instrument: PE 9000
Calib/MIDPT Date: 8/23/91
Units:
Mg/kg dry
CAS No.
Compound
Detection
Cone. Tag Limit
(Mg/kg)
No compounds detected.
F-5
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
PESTICIDE AND AROCLOR COMPOUNDS
Laboratory:
Project ID:
Sample ID:
Dates:
Collected:
Received:
Organics
Ambient Toxicity
Wye
8/14/91
8/20/91
Method: 3550/8080
Analyst: RJM
Data Released By: M. Helmstetter
Contractor: MD DNR
Sample No.: 37403
Extracted: 8/15/91
Analyzed: 9/12/91
Instrument: PE 9000
Calib/MIDPT Date: 8/23/91
Matrix:
Sediment
Units:
Mg/kg dry
CAS No.
Compound
Detection
Cone. Tag Limit
(Mg/kg)
391-84-6
391-86-8
a-BHC
6-BHC
0.793
33.5
B
0.714
1.05
B = Compound detected in the QC blank.
F-6
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
PESTICIDE AND AROCLOR COMPOUNDS
Organics
Ambient Toxicity
Reference
R/8/91
8/20/91
3550/8080
RJM
Laboratory:
Project ID:
Sample ID:
Dates:
Coiiec:;^-
Pvcceivec:
Method:
Analyst:
Data Released By: M. Helmstetter
Matrix: Sediment
Contractor: MD DNR
Sample No.: 37522
Exrracted:
An airbed:
8/15/91
9/12/91
Instrument: PE 9000
Calib/MIDPT Date: 8/23/91
Units:
pg/kg dry
CAS No.
Compound
Detection
Cone. Tag Limit
(/*gAg)
391-84-6
391-86-8
58-89-9
76-44-8
cr-BHC
6-BHC
Lindane
Heptachlor
0.876
1.48
0.817
0.880
B
0.714
1.05
0.616
0.819
B = Compound detected in the QC blank.
F-7
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
PESTICIDE AND AROCLOR COMPOUNDS
Laboratory:
Project ID:
Sample ID:
Organics
Ambient Toxicity
Control
Contractor: MD DNR
Sample No.: 37523
Dates:
Collected:
Received:
8/6/91
8/20/91
Extracted: 8/15/91
Analyzed: 9/12/91
Method: 3550/8080
Analyst: RJM
Data Released By: M. Helmstetter
Instrument: PE 9000
Calib/MIDPT Date: 8/23/91
Matrix:
Sediment
Units: /'gAg dry
CAS No.
Compound
Detection
Cone. Tag Limit
(**g/kg)
72-54-8
4,4'-DDD
0.559 0.469
F-8
-------
AMRL
ORGANIC ANALYSIS DATA SHEET
PESTICIDE AND AROCLOR COMPOUNDS
Laboratory:
Project ID:
Sample ID:
Dates:
Collected:
Received:
Method:
Analyst:
Data Released By: M. Helmstetter
Matrix: Sediment
Organics
Ambient Toxicity
Wye River
9/25/91
9/26/91
3550/8080
RJM
Contracctor: MD DNR
Sample No.: 37841
Extracted: 10/24/91
Analyzed: 11/11/91
Instrument: PE 9000
Calib/MIDPT Date: 11/1/91
Units:
/ig/kg dry
Detection
CAS No. Compound Cone. Tag Limit
(fg/kg)
33213-65-9 Endosulfan II 1.32 0.745
F-9
-------
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