NOAA Status and Trends Mussel Watch Project

[From the U.S. Government Printing Office, www.gpo.gov]

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NOAA STATUS AND TRENDS
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January 1994

NOAA NATIONAL
STATUS AND TRENDS

Mussel Watch Project

Year 7 Technical Report

Prepared by

The Geochemical and Environmental Research
Group (GERG)
Texas A&M University
833 Graham Road
College Station, Texas 77845

PrOPertY Of CSC Library
Submitted to
t U.S. Department of Commerce
"Z
National Oceanic & Atmospheric Administration
ZS 1305 East-West Hwy.
Silver Spring, M.D 20910 U - S . DEPARTMENT OF COMMERCE NOAA
-January 1994
COASTAL SERVICES CENTER
2234 SOUTH HOBSON AVENUE
CHARLESTON , SC 29405-2413

Table of Contents

Introduction ............................................................................................ 1-1

Reprint 1: The UsejWness of Transplanted Oysters in Biomonitoring
Studies ............................................................................................... 1-11

Reprint 2: Overview of the Mrst Four Years of the NOAA National
Status and Trends Mussel Watch Program ........................................... 1-21

Reprint 3: Trace Organic Contamination in Galveston Bay Oysters:
Results from the NOAA National Status and Trends Mussel Watch
Program .............................................................................................. 1-31

Reprint 4: Indicators of Trace Metal Pollution in Galveston Bay ................ 1-37

Reprint 5: Oysters as Biomonitors of the APEX Barge Oil Spill. ................. 1-45

Reprint 6: Meld Studies Using the Oyster Crassostrea virginica to
Determine Mercury Accumulation and Depuration Rates ....................... 1-53

Reprint 7: Trace Metal Chemistry of Galveston Bay: Water,
Sediment and Biota. ............................................................................ 1-59

Reprint 8: Mercury Bioaccumulation by Shrimp (Penaeus aztecus)
Transplanted to Lavaca Bay, Texas ..................................................... 1-81

Reprint 9: Polynuclear Aromatic Hydrocarbon Contaminants in
Oystersfirom the GLdf of Mexico (1986-1990) ........................................ 1-87

Reprint 10: Butyltin Concentrations in Oysters from the Gulf of
Mexico During 1989-1991 .................................................................... 1-97

Reprint 1 1:The American Oyster (Crassostrea virginica) as a
Bioindicator of Trace Organic Contamination ........................................ 1-109

NOAX S NATIONAL STATUS AND TRENDS (NS&T) MUSSEL WATCH
PROGRAM - GULF OF MEXICO

The purpose of the NOAA National Status and Trends (NS&T)
Mussel Watch Project is to determine the long-term temporal and spatial
trends of selected environmental contaminant concentrations in bays
and estuaries. The key questions in this regard are:
(1) What is the current condition of the nation's coastal zone?
(2) Are these conditions getting better or worse?

This report represents the Year 7 Technical Report from this multi-
year project. These questions have been addressed in detail as evidenced
by the scientific papers and reports that have resulted from the
Geochemical and Environmental Research Group's (GERG)
interpretations of the Gulf Coast data (Table 1). Publications not
included in GERG's previous Technical Reports are contained in this
technical report.

This report is an update on the current condition of the Gulf of
Mexico coastal zone, based on results from Years 1 through 7 of the
NOAA NS&T Mussel Watch Project. Following is a brief sampling survey
of these years:

Year 1 - 49 sites (147 stations) of the original 51 sites were
successfully sampled. Sediments and oysters were
analyzed at triplicate stations from all sites.
Year 2 - 48 sites (144 stations) of the original 51 sites were
successfully sampled. Sediments and oysters were
analyzed at triplicate stations from all sites.
Year 3 - Twenty (20) sites were added to the original list of 51
sites for a total of 71 sites. Sixty-four (64) sites (192
stations) of the 71 sites were sampled (only 19 of the
new sites were sampled). Oysters were analyzed at
triplicate stations from all sites. Sediments were
analyzed at only the new sites (three stations analyzed
per site).
Year 4 - Seven (7) new sites were added (only six of the new sites
were successfully sampled). Sixty-seven (67) sites (201
stations) of the 78 total sites were sampled. Oysters
were analyzed at triplicate stations from all sites.
Sediments were analyzed at only the new sites (three
stations analyzed per site).
Year 5 - Three (3) new sites were added to the sampling project
(only two of these sites were successfully sampled-,
79:MBDR and 80:PBSP). Sixty-eight (68) sites (204
stations) of the 80 total sites were sampled. Oysters
were analyzed at triplicate stations from all sites.

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Sediments were analyzed at only the new sites (three
stations analyzed per site).
Year 6 - Two (2) new sites were added to the sampling project
(81:BHKF in Bahia Honda Key, FL and 63:LPGO in
Lake Pontchartrain, LA). Sixty-four (64) sites (192
stations) were sampled. Oysters were analyzed at
triplicate stations from all sites. Sediments were
analyzed at only the new sites (three stations analyzed
per site).
Year 7 - Five new sites were established including three new
sites in Puerto Rico (Sites 86 to 88) and two new sites
in Choctawhatchee Bay (Sites 84 and 85). Sixty-seven
(67) sites were analyzed. Only one oyster analysis was
conducted at each of the old sites on a composite from
the three stations. Sediments were analyzed at the five
new sites and one site in Florida (PBPH) (three stations
analyzed per site).

Details of the sample collection and location of field sampling sites are
contained in a separate report titled "Field Sampling and Logistics in
Year 7".

The oyster and sediment samples were analyzed for contaminant
concentrations [trace metals, polynuclear aromatic hydrocarbons (PAH),
pesticides and polychlorinated biphenyls (PCBs)], disease incidence and
other parameters that aid in the interpretation of contaminant
distributions (grain size, oyster size, lipid content, etc.). The analytical
procedures used and the QA/QC Project Plan are detailed in a separate
repor-t titled "Analytical Methods". The data that were produced from the
sample analyses for Year 7 are found in a separate report titled
"Analytical Data".

A complete and comprehensive interpretation of the data from the
National Status and Trends Project for oyster data coupled with the
sediment data is an ongoing process. We have begun and are continuing
that process as evidenced by this report and the scientific manuscripts
that we have published or submitted for publication (Table 1). As part of
the data interpretation and dissemination, over 40 presentations of the
NOAA NS&T Gulf Coast Mussel Watch Project were given at national and
international meetings. With seven years of data, the question of
temporal trends of contaminant concentrations has been addressed. A
general conclusion found for most contaminants measured is that the
concentrations have remained relatively constant over the seven-year
sampling period. This general trend, however, is not observed at all sites.
Some sites show significant changes (both increases and decreases)
-among the years. Continued sampling is addressing the frequency and
rates of these changes.

Exceptions to this general trend are found for DDTs and TBT.
When historical data for DDT in bivalves is compared to current NS&T

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data, a decrease in concentration is apparent. Also based on TBT data
collected as part of the NOAA NS&T Mussel Watch Project, a decline in
TBT concentration in oysters is apparent. Both declines may be in
response to regulatory actions.

During Year 3 of this project, 20 new sites were added. These.sites
were chosen to be closer to urban areas, and therefore, to the sources of
contaminant inputs. These new sites were not, however, located near
any known point sources of contaminant input. These sites were added
to better represent the current status of contaminant concentrations in
the Gulf of Mexico. Over the subsequent years of the project (Years 4
through 7) additional sites have been added to increase the
representative coverage of the Gulf of Mexico and U.S. Caribbean
territories.

While sampling sites for this project were specifically chosen to
avoid known point sources of contamination, the detection of coprostanol
in sediment from all sites indicates that the products of man's activities
have reached all of the sites sampled. However, when compared to
known point sources of contamination, all of the contaminant
concentrations reported are, in most cases, many orders of magnitude
lower than obviously contaminated areas. The lower concentrations in
Gulf of Mexico samples most likely reflect the fact that the sites are far
removed from point sources of inputs, a condition which is harder to
achieve in East and West Coast estuaries. In fact, new sites added in
Years 3 through 7 are closer to urban areas and generally had higher
contaminant concentrations. An important conclusion derived from the
extensive NS&T data set is that contamination levels in Gulf Coast near
shore areas remain the same or are getting better, and most areas
removed from point sources are not severely contaminated.

This document represents one of three report products as part of
Year 7 of the NS&T Gulf of Mexico projects. The other two reports are
entitled:

* Analytical Data, Year 7
0 Field Sampling and Logistics, Year 7

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Table 1. GERG/NOAA NS&T PUBLICATIONS Included in
Year Report

Wade, T.L., B. Garcia-Romero and J.M. Brooks (1988)
Tributyltin contamination of bivalves from U.S. coastal
estuaries. Environmental Science and Technology, 22:
1488-1493. IV

Wade, T.L., E.L. Atlas, J.M. Brooks, M.C. Kennicutt 11, R.G. Fox,
J. Sericano, B. Garcia-Romero and D. DeFreitas (1988)
NOAA Gulf of Mexico Status and Trends Program:
Trace organic contaminant distribution in sediments
and oysters. Estuaries, 11: 171-179. IV
Wade, T.L., B. Garcia-Romero and J.M. Brooks (1988) 0
Tributyltin analyses in association with NOAA's
National Status and Trends Mussel Watch Program. In:
OCEANS '88 Conference Proceedings, Baltimore, MD, 31
Oct. - 2 Nov. 1988, pp. 1198-1201. IV

Wade, T.L., M.C. Kennicutt, 11 and J.M. Brooks (1989) Gulf of
Mexico hydrocarbon seep communities: III: Aromatic
hydrocarbon burdens of organisms from oil seep
ecosystems. Marine Environmental Research, 27: 19-30. IV

Wade, T.L. and J.L. Sericano (1989) Trends in organic
contaminant distributions in oysters from the Gulf of
Mexico. In: Proceedings, Oceans '89 Conference, Seattle,
WA, pp. 585-589. IV

Wade, T.L. and B. Garcia-Romero (1989) Status and trends of
tributyltin contamination of oysters and sediments from
the Gulf of Mexico. In: Proceedings, Oceans '89
Conference, Seattle, WA, pp. 550-553. IV

Wade, T.L. and C.S. Giam (1989) Organic contaminants in the
Gulf of Mexico. In: Proceedings, 22nd Waterfor Texas
Conference, Oct. 19-21, 1988, South Shore Harbour Resort
and Conference Center, Lzague City, TX (R. Jensen and C.
Dunagan, Eds.), pp. 25-30. V

Craig, A., E.N. Powell, R.R. Fay and J.M. Brooks (1989)
Distribution of Perkinsus marinus in Gulf coast oyster
populations. Estuaries, 12: 82-91. IV

Presley, B.J., R.J. Taylor and P.N. Boothe (1990) Trace metals
in Gulf of Mexico oysters. The Science of the Total
Environment, 97/98: 551-553. IV

1-4

Sericano, J.L., E.L. Atlas, T.L. Wade and J.M. Brooks (1990)
NOAA's Status and Trends Mussel Watch Program:
Chlorinated pesticides and PCB's in oysters
(Crassostrea virginica) and sediments from the Gulf of
Mexico, 1986-1987. Marine Environmental Research, 29:
161-203. IV

Wade, T.L., B. Garcia-Romero and J.M. Brooks (1990) Butyltins
in sediments and bivalves from U.S. coastal areas.
Chemosphere 20: 647-662. IV

Brooks, J.M., M.C. Kennicutt II, T.L. Wade, A.D. Hart, G.J.
Denoux and T.J. McDonald (1990) Hydrocarbon
distributions around a shallow water multiwell
platform. Environmental Science and Technology, 24:
1079-1085. IV

Sericano, J.L., T.L. Wade, E.L. Atlas and J.M. Brooks (1990)
Historical perspective on the environmental
bioavailability of DDT and its derivatives to Gulf of
Mexico oysters. Environmental Science and Technology,
24: 1541-1548. IV

Wade, T.L., J.L. Sericano, B. Garcia-Romero, J.M. Brooks and
B.J. Presley (1990) Gulf coast NOAA National Status &
Trends Mussel Watch: the first four years. In: MTS'90
Conference Proceedings, Washington, D.C., 26-28
September 1990, pp. 274-280. IV, V

Brooks, J.M., T.L. Wade, B.J. Presley, J.L. Sericano, T.J.
McDonald, T.J. Jackson, D.L. Wilkinson and T.F. Davis
(1991) Toxic contamination of aquatic organisms in
Galveston Bay. In: Proceedings Galveston Bay
Characterization Workshop, February 21-23, pp. 65-67. VI

Wade, T.L. J.M. Brooks, J.L. Sericano, T.J. McDonald, B. Garcia-
Romero, R.R. Fay, and D.L. Wilkinson (1991) Trace
organic contamination in Galveston Bay: Results from
the NOAA National Status and Trends Mussel Watch
Program In: Proceedings Galveston Bay Characterization
Workshop, February 21-23, pp. 68-70. VI

Presley, B.J., R.J. Taylor and P.N. Boothe (1991) Trace metals
in Galveston Bay oysters. In: Proceedings Galveston Bay
Characterization Workshop, February 21-23, pp. 71-73. VI

Sericano, J.L., T.L. Wade and J.M. Brooks (1991) Transplanted
oysters as sentinel organisms in monitoring studies. In:
Proceedings Galveston Bay Characterization Workshop,
February 21-23, pp. 74-75. VI

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McDonald, S.J., J.M. Brooks, D. Wilkinson, T.L. Wade and T.J.
McDonald (1991) The effects of the Apex Barge oil spill
on the fish of Galveston Bay. In: Proceedings Galveston
Bay Characterization Workshop, February 21-23, pp. 85-
86. VI

Wade, T.L., J.M. Brooks, M.C. Kennicutt 11, T.J. McDonald, G.J.
Denoux and T.J. Jackson (1991) Oysters as biomonitors
of oil in the ocean. In: Proceedings 23rd Annual Offshore
Technology Conference, No. 6529, Houston, TX, May 6-
9,, pp. 275-280. V

Brooks, J.M., M.A. Champ, T.L. Wade, and S.J. McDonald
(1991) GEARS: Response strategy for oil and
hazardous spills. SeaTechnology, April 1991,pp.25-32. V

Sericano, J.L., T. L. Wade and J.M. Brooks (1991) Chlorinated
hydrocarbons in Gulf of Mexico oysters: Overview of
the first four years of the NOAA's National Status and
Trends Mussel Watch Program (1986-1989). In: Water
Pollution: Modelling, Measuring and Prediction. Wrobel,
L.C. and Brebbia, C.A. (Eds.), Computational Mechanics
Publications, Southampton, and Elsevier Applied Science,
London, pp. 665-68 1. V, VI

Wade, T.L., B. Garcia-Romero and J.M. Brooks (1991)
Bioavailability of butyltins. In: Organic Geochemistry -
Advances and Applications in the Natural Environment.
Manning, D.A.C. (Ed.), Manchester University Press,
Manchester, pp. 571-573. V

Wilson, E.A., E.N. Powell, M.A. Craig, T.L. Wade and J.M.
Brooks (199 1) The distribution of Perkinsus marinus in
Gulf coast oysters: its relationship with temperature,
reproduction and pollutant body burden. Int. Reuve der
Gesantan Hydrobioligie, 75: 533-550. IV

Sericano, J.L., A.M. El-Husseini and T.L. Wade (1991) Isolation
of planar polychlorinated biphenyls by carbon column
chromatography. Chemosphere, 23(7): 915-924. V, VI

Wade, T.L., B. Garcia-Romero and J.M. Brooks (1991) Oysters
as biomonitors of butyltins in the Gulf of Mexico.
Marine Environmental Research, 32: 233-:@41. IV, V

Wilson, E.A., E.N. Powell, T.L. Wade' R.J. Taylor, B.J. Presley
and J.M. Brooks (1991) Spatial and temporal
distributions of contaminant body burden and disease
in Gulf of Mexico oyster populations: The role of local
and large-scale climatic controls. Helgolander
Meeresunters, 46: 201-235. V, VI

Powell, W.N., J.D. Gauthier, E.A. Wilson, A. Nelson, R.R. Fay
and J.M. Brooks (1992) Oyster disease and climate

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change. Are yearly changes in Perkinsus marinus
parasitism in oysters (Crassostrea virginica) controlled
by climatic cycles in the Gulf of Mexico? PSZNI: Marine
Ecology, 13: 243-270. IV

Hofmann, E.E., E.N. Powell, J.M. Klinck E.A. Wilson (1992)
Modeling oyster populations III. critical feeding
periods, growth and reproduction. J. Shellfish
Research, 2: 399-416. V

Sericano, J.L., T.L. Wade, A.M. El-Husseini and J.M. Brooks
(1992) Environmental significance of the uptake and
depuration of planar PCB congeners by the American
oyster (Crassostrea virginica). Marine Pollution Bulletin,
24: 537-543. VI

Wade, T.L., E.N. Powell, T.J. Jackson and J.M. Brooks (1992)
Processes controlling temporal trends in Gulf of Mexico
Oyster health and contaminant concentrations. In:
Proceedings MTS '92, Marine Technology Society, Oct. 19 -
21, Washington, D.C. pp. 223-229. VI

Tripp, B.W., J.W. Farrington, E.D. Goldberg and J.L. Sericano
(1992) International mussel watch: the initial
implementation phase. Marine Pollution Bulletin, 24:
371-373. VI

Sericano, J.L., T.L. Wade and J.M. Brooks (1993) The
usefulness of transplanted oysters in bionionitoring
studies. In: Proceedings of The Coastal Society Twelfth
International Conference, Oct. 21-24, 1990, San Antonio,
TX, pp. 417-429. V, VU

Wade, T.L., J.L. Sericano, J.M. Brooks and B.J. Presley (1993)
Overview of the first four years of the NOAA National
Status and Trends Mussel Watch Program. In:
Proceedings of The Coastal Society Twelfth International
Conference, Oct. 21-24, 1990, San Antonio, TX, pp. 323-
334. V, VU

Sericano, J.L., T.L. Wade, E.N. Powell and J.M. Brooks (1993)
Concurrent chemical and histological analyses: Are
they compatible? Chemistry and Ecology, 8: 41-47. V, VI

Sericano, J.L., T.L. Wade, J.M. Brooks, E.L. Atlas, R.R. Fay and
D.L. Wilkinson (1993) National Status and Trends
@4ussel Watch Program: chlordane-related compounds
in Gulf of Me3dco oysters: 1986-1990. Environmental
Pollution, 82: 23-32. V, VI

Wade, T.L., T.J. Jackson, J.M. Brooks, J.L. Sericano, B. Garcia-
Romero and D.L. Wilkinson (1993) Trace organic
contamination in Galveston Bay oysters: results from

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the NOAA National Status and Trends Mussel Watch
Program. In: Proceedings, The Second State of the Bay
Symposium, Galveston, TX, February 4-6, pp. 109-111. Vil

Presley, B.J. and K.T. hann (1993) Indicators of trace metal
pollution in Galveston Bay. In: Proceedings, The Second
State of the Bay Symposium, Galveston, TX, February 4-6,
pp. 127-13 1. V101

Wade, T.L., T.J. Jackson, T.J. McDonald, D.L. Wilkinson, and
J.M. Brooks (1993) Oysters as biomonitors of the APEX
barge oil spill. In: Proceedings, 1993 International Oil
Spill Conference, Tampa, FL, March 29-April 1, pp. 127-
131. VI[

Palmer, S.J., B.J. Presley, R.J. Taylor and E.N. Powell (1993)
Field studies using the oyster Crassostrea virginica to
determine mercury accumulation and depuration rates.
Bulletin Environmental Contamination Toxicology 5 1:
464-470. VH

Morse, J.W., B.J. Presley and R.J. Taylor (1993) Trace metal
chemistry of Galveston Bay: water, sediment and biota.
Marine Environmental Research, 36: 1-37. VH

Sericano, J.L. (1993) The American oyster (Crassostrea
virginica) as a bioindicator of trace organic
contamination. Ph.D. Dissertation, Department of
Oceanography, Texas A&M University, 242 p. VI[

Palmer, S.J. and B.J. Presley (1993) Mercury bioaccumulation
by shrimp (Penaeus aztecus) transplanted to Lavaca
Bay, Texas. Marine Pollution Bulletin, 26(10): 564-566. V11

Jackson, T.J., T.L. Wade, T.J. McDonald, D.L. Wilkinson and J.M.
Brooks (1993) Polynuclear aromatic hydrocarbon
contandnants in oysters from the Gulf of Mexico (1986-
1990). Environmental Pollution , 83: 291-298. VI, VII

Garcia-Romero, B., T.L. Wade, G.G. Salata, and J.M. Brooks
(1993) Butyltin concentrations in oysters from the Gulf
of Mexico during 1989-1991. Environmental Pollution,
81: 103-111. V1, V11

Sericano, J.L., T.L. Wade, B. Garcia-Romero and J.M. Brooks
(1994) Environmental accumulation and depuration of
tributyltin by the American Oyster, Crassostrea
virginica. Marine Biology (submitted). IV

Hofmann, E.E., J.M. Klinck, E.N. Powell, S. Boyles, M. Ellis
(1994) Modeling oyster populations 11. Adult size and
reproductive effort. Journal of Shellfish Research (in
press). V

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Ellis, M.S., K.-S. Choi, T.L. Wade, E.N. Powell, T.J. Jackson and
D.H. Lewis (1994) Sources of local variation in
polynuclear aromatic hydrocarbon and pesticide body
burden in oysters (Crassostrea virginica) from Galveston
Bay, Texas. Estuaries (in press). VI

To be included in Year 8 Technical Report:

Kennicutt, M.C. IL T.L. Wade, B.J. Presley, A.G. Requejo, J.M.
Brooks and G.J. Denoux (1994) Sediment contaminants
in Casco Bay, Maine: inventories, sources and potential
for biological effects. Environmental Science and
Technology (in press).

McDonald, S.J., M.C. Kennicutt IL J.L. Sericano, H. Liu, T.L.
Wade and S.H. Safe (1994) Correlation between
bioassay-derived P4501A1 -Induction activity and
chemical analysis of dam (Laternula elliptica) extracts
from McMurdo Sound, Antarctica. M a r i n e
Environmental Research (Submitted).

Sericano, I.L., T.L. Wade and J.M. Brooks (1994) Accumulation
and depuration of organic compounds by the American
oyster (Cassostrea virginica). Science q the Total
Environment (in press).

Sericano, I.L., S.H. Safe, T.L. Wade, and J.M. Brooks (1994)
Toxicological significance of non-, mono-, and di-ortho
substituted polychlorinated biphenyls in oysters from
Galveston and Tampa Bays. Environmental Toxicology
and Chemistry (submitted).

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I Reprint 1
1 The Usefulness of Transplanted Oysters in
Biomonitoring Studies
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I Jos6 L. Sericano, Terry L. Wade,
and James M. Brooks
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M M M M M 1111111IN
The Usefulness or Transplanted Oysters in Biomonitoring Studies

Jos6 L. Scricano, Terry L. Wade, and James M. Brooks
Texas A&M University
The Coastal Society Twelfth International Conference
Abstract

This study was designed to examine, the uptake and depuration of
selected organic contaminants of environmental concern, i.e., polynuclear
aromatic hydrocarbons (PAHs) and polychlorinated biphcnyls (PCBs), by
CONFERENCE transplanted oysters (Crassostrea' virginica) in Galveston Bay, Texas and to
establish the feasibility of using transplanted oysters for biomonitoring the
contamination status in areas were no indigenous bivalves are present.
PROCEEDINGS Transplanted oysters bioaccumulated individual PAHs and low molecular
weight PCBs to concentrations that were not statistically differentiable from
the levels encountered in native oysters within 30 to 48 days. In contrast, high
Our Coastal Experience:
molecular weight PCBs did not reach equilibrium in transplanted oysters;
Assessing the Past, whose high molecular weight PCB concentrations were lower than those
measured in indigenous oysters during the seven-week uptake period. During
Confronting the Future the depuration phase of this study, originally uncontaminated oysters
aster rate than
depurated PAHs and low molecular wight PCBs at a f,
chronically contaminated oysters. Clearance of high molecular weight PCBs
4 was limited in both oyster populations.

Introduction

Contamination of the coastal marine environment by a number of
organic compounds of synthetic or natural origin has received increasing
a
tention over the last several years. Biomonitoring of these compounds in
t
the aquatic environment has been well established and bivalves are generally
preferred for this purpose. The rationale for the "Mussel Watch" approach
using different bivalves, e.g. mussel, oysters and/or clams, has been
summarized by different authors (Goldbcrg et aL, 1978; Farrington et A,
1980; Phillips, 1980; Risebrough et al., 1983) and its concept has been applied
... to many monitoring programs during the last decade (Farrington et al., 1983;
Martin, 1985; Tavares et al., 1988; Wade et al., 1988; Sericano et al., 1990).
r
The National Oceanic and Atmospheric Administration's National
Status and Trends Program (NOAA's NS&T) is designed to monitor the
current status and long-term effects of selected organic and inorganic
contaminants of environmental concern, e.g. polynuclear aromatic
OCTOBER 21-2411990 hydrocarbons (PAHs), chlorinated pesticides, polychlorinated biphenyls
(PCBs), and trace metals, along the coasts of the U.S. by measuring their
St. Anthony Hotel
concentrations in bivalves over a number of years. During the first five years
San Antonio, Texas of this program (1986-1990), the intent was to sample all the locations

417

prescribed by NOAA. However, locations depleted or devoid of living oysters in native oysters represent the time-integrated contaminant concentrations
caused by virtue of diseases, predators, excessive freshwater runoff, harvesting, available to the oysters in solution, adsorbed onto particles, and incorporated
or dredge material burying entire reefs compromised this goal. Therefore, in into food.
some instances, it was not possible to obtain samples. After the first five
years of the NS&T, nearly 20% of the original locations presented some of Initial concentrations of total PAHS in transplanted oysters increased
the above-mentioned sampling problems that left the database with missing from 290 ng/g to a final value of 4360 ng/g. Two- and three-ring PAHs were
values. Transplantation of bivalves to areas where indigenous individuals were detected in low concentrations in transplanted and indigenous oysters. Four-
not originally present or have been lost because of natural or man-induced and five-ring compounds were accumulated to the highest concentrations in
actions could be a potentially useful tool in monitoring environmental Hanna Reef oysters. By the end of the first 48 days, transplanted oysters
pollution. accumulated these PAHs to levels that were not statistically differentiable
from the concentrations measured in native individuals (Figure 2a). The
The present study was designed to examine the uptake and PAHs accumulated to the highest concentrations by transplanted oysters were:
depuration of selected organic contaminants, i.e. polynuclear aromatic pyrene> fluoranthene> chrysene> benzo(e) pyrene> benzo(b)-anthracene
hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs), in oysters (Figure 2b). Clams and mussels exposed to sediments contaminated with high
(Crassostrea virginic ) through transplantation experiments in two locations PAH concentrations accumulated pyrene> bcnzo(e) pyrene> benzo(b)
in Galveston Bay, Texas. fluoranthene> benz(a) anthracene (Obana et al., 1983) and chryscnc>
benzo(b) fluoranthene> fluoranthene> benzo(e)pyrene> benz(a)anthracene
Materials and Methods (Pruell et al., 1986), respectively.
Experimental design Hanna Reef and Ship Channel oysters showed statistically significant
depuration (p<0.05) of four- and five-ring PAHs after relocation to the
Approximately 250 oysters of similar dimensions were collected from Hanna Reef area (Figures 2c and 2d). Depuration of these aromatic
a relatively uncontaminated area in Galveston Bay, Hanna Reef, and compounds by both groups of oysters was approximately exponential. This is
transplanted in 24x7O cm net bags, containing 25-30 individuals per bag, to a indicated in Figure 3, where the concentration of selected PAHs plotted on
new location near the Houston Ship Channel in the upper part of the Bay a semi-log plot approximate straight lines.
(rigure 1). Composite samples of 20 transplanted and. 15 indigenous oysters Kinetics parameters describing uptake and release of PAHs can be
were collected at 0, 3, 7, 17, 30, and 48 days during the First phase of the
transplantation experiment. The remaining Hanna Reef oysters were then calculated assuming the first-order equation
back- transplanted to their original location in Galveston Bay. At the same
time, approximately 150 indigenous oysters from the Ship Channel site were (1) dCt/dt = kuC. - kdCt
also transplanted to the Hanna Reef area. Composite samples of 20 oysters where Ct is the PAH concentration in the transplanted oyster at timc=t, C,,,
from each population were collected at 3, 6, 18, 30, and 50 days after is the PAH concentration in the seawater, and k and k are the uptake and
transplantation. U d
depuration rate constant, respectively. If the C,,, at Hanna Reef is regarded
Analytical method as zero, i,e., CW=O, which is considerably reasonable because of the very low
PAH concentrations measured in indigenous oysters, then equation (1)
The analytical procedures used during this study are modifications of reduces to
previously reported methods (MacLeod et al., 1985) and are fully described (2) dC,/dt -k
elsewhere (Wade et al., 1988; Sericano et al., 1990). A
Results and Discussion or, after integration,
The concentrations of some of the organic contaminants increased (3) log C, log C,-(kd/2.301)t
dramatically during the seven-week exposure period. Comparatively, where CO is the PAH concentration in oysters at the time of their relocation
concentrations of individual PAHs and PCBs in indigenous oysters during the to the Hanna Reef area. Using this equation and the PAH concentrations
first phase of this experiment were fairly constant. The analyte concentrations
NW1 8M M M M M UK 1 9M M M

corresponding t t"oyster rPU9 ns during the depuration period, values study, the biological half-lives of PCBs increased with the number of chlorine
o bo atio I
of kd can be calculated. Statistical analyses, at the a=0.05 level, of the atoms in the biphenyl rings. Langston (1978) also reported that the less
regression lines of the logarithm of the concentrations versus sampling time chlorinated PCB congeners were depurated more rapidly by bivalves
for the depuration period showed significant differences between the slopes, (Cerastoderma edule and Macoma balthica) with half-lives from 5 to 21 days
i.e., depuration rates, measured for Hanna Reef and Ship Channel oysters. for di- to tetrachlorobiphenyls. In contrast, hexachlorobiphenyls, and some of
the pentachlorobiphenyls, did not decrease in concentration during the 21-day
The biological half-life, tj/2, can be derived from equation (3) study. Courtney and Denton (1976) reported that environmentally
contaminated clams and clams exposed to Aroclor 1254 in the laboratory did
(4) t112 = 0.693/kd not depurated PCBs during three months in. control seawater.
The half-lives are reported in Tablel. They ranged from 10A and In summary, PAHs and -low molecular weight PCBs were rapidly
12.4 days for pyrene to 25.6 and 38.5 days for fluoranthene in Hanna Reef and accumulated by transplanted oysters. Apparent steady-state concentrations
Ship Channel oysters, respectively. Most -of the values were, however, were reached after 30 to 48 days. In contrast, high molecular weight PCBs
between 10 and 16 days. did not reached an equilibrium plateau within the sevcn-week exposure to
high PCB concentrations. However, the still- increasing concentrations
Recently, Pruell ct al. (1986) reported the half-lives for selected measured for these PCBs by the end of the exposure period seems to indicate
PAHs in mussels (Mytilus edulis) exposed to environmentally contaminated that, given enough time, equilibrium concentrations will eventually be reached.
sediments. The calculated half-lives compared well with the values measured When back-transplanted to the Hanna Reef area, originally uncontaminated
in this study (Table 1). and chronically exposed oysters depurated PAHs with half-lives ranging from
10.4 to 23.6 days and from 12.4 to 38.5 days, respectively. These rates were
PCB concentrations in transplanted oysters increased from 30 ng/g similar to those calculated for tri- and tetrachlorobiphenyls but faster than
to 850 ng/g after the 48-day exposure period. Pentachlorobiphenyls-were the those estimated for heavier molecular weight PCBs. Despite the limitations
compounds accumulated to the highest concentrations in transplanted and discussed in the text, transplanted oysters are considered valuable
native oysters (Figures 4a and 4b). In comparison, practically no octa-, nona-, bioindicators of environmental contamination by PAHs and PCBs in areas
or decachlorobiphenyls were detected in either oyster group.. Contrasting with lacking indigenous bivalves.
PAHs, not all the PCB bomologs measured in transplanted oysters reached
the concentration encountered in indigenous individuals by the end of the first Acknowledgements
phase of this experiment. While there were no statistically significant
differences in the tri- and tetrachlorobiphenyl concentrations measured in Funding for this research was provided by the National Oceanic and
transplanted and native oysters, significant differences were observed in the Atmospheric Administration Grant Number 50-DGNC-5-00262 (National
total concentrations of penta- and hexachlorobiphenyls. It seems evident that Status and Trends Program).
a longer exposure period is needed for the higher molecular weight PCBs to
reach an steady state concentration (Figure 5).

Hanna Reef and Ship Channel oysters showed statistically significant
depuration (p<0.05) of low molecular weight PCBs when relocated to the
Hanna Reef area (Figures 4c and 4d). Originally uncontaminated oysters
depurated PCBs at a faster rate than chronically contaminated oysters. The
dcpuration rates of high molecular weight PCBs were significantly slower in
both oyster populations. This differential PCB depuration can be observed in
Figures 4b and 4d showing the concentrations of selected PCBs at the end of
the uptake and depuration periods. Biological half-lives for these PCBs in
Hanna Reef and Ship Channel oysters ranged from 21 to 129 days and from
20 days to >year, respectively (Table 1). Pruell et al. (1986) reported
half-lives for tri- to hexacblorobiphenyls in mussels exposed to resuspended
contaminated sediments ranging from 16.3 to 45.6 days. Similar to the present

420 421

References
Wade, T.L. Atlas, J.M. Brooks, M.C. Kennicutt II, R.G. Fox, J.L.
Courtney, W.A.M. and G.R.W. Denton. 1976. Environ. Pollut. 10:55-64 Sericano, B. Garcia-Romero, D. DeFreitas. 1988. Estuaries, 11:171-179.

Farrington, J.W., J. Albaiges, K.A. Burns, B.P. Dunn, P. Eaton, J.L. Laseter,
P.L. Parker, and S. Wise. 1980. In: The International Mussel Watch: Report
of a workshop sponsored by the Environmental Studies Board Commission on
Natural Resources. National Research Council, pp. 7-77.

________, E.D. Goldberg, R.W. Riserbrough, J.H. Martin, and V.T. Bowen.
1983. Environ. Sci. Technol. 17: 490-496.

Goldberg, E.D., V.T. Bowen, J.W. Farrington, G. Harvey, J.H. Martin, P.L.
Parker, W. Risebrough, E. Schneider, and E. Gamble. 1978. Envir. Conserv.
5: 1-25.

Langston, W.J. 1978. Mar. Biol., 46: 35-40.

MacLeod, W.D., D.W. Brown, A.J. Friedman, D.G. Burrows, O. Maynes,
R.W. Pearce, C.A. Wigren, and R.G. Bogar. 1985 In: Standard Analytical
Procedure of the NOAA National Analytical Facility, 1985-1986. Extractable
Toxic Organic Compounds. (2nd Ed.), U.S. Department of Commerce,
NOAA/NMFS. NOAA Tech. Memo. NMFS F/NWC-92.

Martin, M. 1985. Mar. Poll. Bull. 16: 140-146.

Obana, H., S. Hori, A. Nakamura, and T. Kashimoto. 1983. Water Res., 17:
1183-1187.

Phillips, D.J.H. 1980. "Quantitative Biological Indicators, Their Use to
Monitor Trace Metals and Organochlorine Pollution;" Applied Science:
London.

Pruell, R.J., J.L. Lake, W.R. Davis, and J.G. Quinn. 1986. Mar. Biol., 91:
497-507.

Risebrough, R.W., B.W. de Lappe, W. Walker II, B.R.T. Simoneit, J.
Grimalt, J. Albaiges, J.A. Garcia Regueiro, I. Ballester A. Nolla, and M.G.
Marino Fernandez. 1983. Mar. Poll. Bull., 14: 181-187.

Tavares, T.M., V.C. Rocha, C. Porte, D. Barcelo, and J. Albaiges. 1988. Mar.
Poll. Bull., 19: 575-578.

Sericano, J.L., E.L. Atlas, T.L. Wade, J.M. Brooks. 1990. Mar. Environ. Res.
29: 161-203.

422

TABLE 1. Biological half-lives of selected PAIis and PCBs In transplanted and
Indigenous oysters
T E X A S
---------------------- 0-Y-S-T-E-R-S ------------------------
MUSSELS1
HANNA REEF
SHIP CHANNEL

-----------------------------------------------------------
enan
threne -

yrene
Fluoranthene 25.6 38.5 29.8
P 10.4 12.4 LA PO
Benzo(a)alithracene 13.2 15.3 17'.8 -t 'F@
Chrysene 12.3 15.8 14.2
Benzo(e)pyrene Q
11.5 16.1 14.4

PCB#26 21
20
PCB#52 28 VC,
PCB#110 47
45 147
PCB# 118 75 >year 0
PCB# 149 129 >year
EA S
PCB#22 16.3 '291JO,
PCB#101 27.9 SAN L
PCB# 128
36.5
PCB# 153 45.6

------------------------------------------------

TEXAS CITY

GALVESTON

Site 1: Hanna Reef

w
Site 2 : Ship Channel

Figure 1: Galveston Bay transplantation sites.

424
425

END OF UPTAKE PERIOD END OF DEPURATION PERIOD

a Ime"t- C 0 ItFtC"I-
z U fic 07.1 SC OY0...

z
z
b)
z U
0 z
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4-1 NUMBER OF RINOS NUMBZR OF RINGS

END OF UrrAKE PERIOD END OF DEPURATION PERIOD
2@ b d

z
0

Iwo

hi bi
U
0 40
U U
2:
0 AR
1 5 3 4 6 6

ANALYTE MIALYTE

Figure & Total and selected individual polynuclear aromatic hydrocarbon
concentrations (ng/g, dry weight) in Hanna Reef and Ship Channel oysters
after the uptake and depuration periods. The error bars represent one
standard deviation from the mean (n=4).

PYRFNE CIIRYSENE
100007
LOO@ @ 1IR oysters
---0- 11ROyi(cre @ &C oysters
&C Oysters

2 1000 1000,

zoo
z 100
0

UPTAPLZ - DEPuRATION
UPTAKE DEPURATION 10
101
0 10 20 30 40 so so 70 Do 90 100 0 10 20 30 40 so 80 70 so 90 too
rA111'.1111ZIA-101
juLLkN DATE JML" DATE

Eigur
a: Selected PAH concentrations (ng/g. dry weight) in Hanna Reef and Ship
Channel oysters during the uptake and depuration phases of the experiment.

END OF UPTAKE PERIOD END OF DEPURATION PERIOD

C a ial

F3 &C oy.t-. SC 07-t--

0 0

z r
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N) b d
00

z z
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z z
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0
so MM M
0
2.(,) 92M ILO(G) 116(s) 149(a) ls@m 20(31 92M 110(a) &1*($) 14"61 IGTM

ANALYTE ANALYTE

Figure 4 Homolog and selected individual polychlorinated biphenyl concentrations
(ng/g, dry weight) in Hanna Reef and Ship Channel oysters after the uptake
and depuration periods. The error bars represent one standard deviation
from the mean (n=4).

100 PCB 152 1000 PCB #110
HR OyVteM KROystexe
Oysters
SC Oysters

0
to
(D

10
z
94
U
z
0 to
U

VVTAJM DI[PURATION
VPTMM - DEPURATION
0 10 20 30 40 so go 70 so 90 0 10 20 30 40 6.0 so 7 .0 so 90 too
JULIAN DATE JULIAN DATE

D=M-5: Selected PCB concentrations (ng/g, dry Weight)in Hanna Reef and Ship
Channel oysters during the uptake and depuration phases of the experiment.
sc

Reprint 2

Overview of the First Four Years of the NOAA
National Status and Trends
Mussel Watch Program

Terry L. Wade,, Jos6 L. Sericano,, James M.
Brooks, and Bobby J. Presley

1-21

Overview of the First Four Years of the NOAA National Status and Trends
0 Mussel Watch Program
Terry L. Wade, Jose L. Sericano, James M. Brooks, and Bobby J. Presley
Texas A&M University
The Coastal Society Twelfth International Conference
Abstract

oysters are utilized as bioindicator organisms to characterize the
current status and long-term trends for 13 trace elements and 57 organic
contaminants from 75 Gulf of Mexico sampling sites. Sampling sites are
CONFERENCE distributed throughout the U.S. waters of the Gulf of Mexico, away from
known point sources of input, and are sampled yearly in the winter to provide
a geographical description of the chronic contaminant loading of the entire
PROCEEDINGS U.S. Gulf on a regional basis. Three stations at each site are analyzed
individually to assess natural intra-site variability so that significant changes
Our Coastal Experience: can be detected. Extensive intercomparison exercises assure the comparability
of analytical measurements with companion studies on the East and West
Assessing the Past, Coasts. The first four years of data for the Gulf of Mexico represents over
40,000 individual data points. The general trend from this large data set is
Confronting the Future contaminant concentrations that show no changes during the four-year
sampling period. There are, however, certain sites that have experienced
:ZK@ significant changes in contaminant concentration over the last four-year
0 sampling period, including monotonic increases and decreases. Generally, the
concentrations of the various contaminants do not show any significant
relationship to each other. This is probably due to different input sources.
Higher concentrations of most contaminants are associated with proximity to
large urban areas. Two areas that appear to be exceptions to this
k generalization, St. Andrew Bay, FL and Choctawhatchee Bay, FL, are
V1% discussed in more detail.

Introduction

The National Oceanic and Atmospheric Administration (NOAA)
National Status and Trends (NS&T) Mussel Watch Program has sampled and
analyzed bivalves from U.S. coastal areas since 1086. This report summarizes
the first four years of NS&T data for the Gulf Coast of the U.S. Sampling
sites give coverage of the Gulf Coast from southernmost Texas to
southernmost Florida. Portions of the data have been previously discussed
(Wade et al., 1988, 1989, and 1990; Wade and Sericana, 1989; Sericano et al.,
.!7 to 1990a and b) and only an overview is presented here.

Methods
OCTOBER 21-24,1990
The NS&T program utilizes standard operating procedures and a
St. Anthony Hotel strong quality assurance/quality control program for trace element and trace
San Antonio, Texas
323

we have, there is no correlation between As concentration in oysters and
organic analyses. Details of these methods arc found elsewhere (Brooks et phosphate rock occurrence, shipping, or mining. The As distribution does
A, 1989; Wade ct al., 1988). The accuracy and precision of these methods seem to be controlled by local environmental inputs, as do certain other metal
have been established by several intercalibration exercises conducted by the distributions. There seems to be no other explanation for high and low
U.S. National Institute of Standards and Testing and the Canadian National
Research Council. concentrations of trace metals to occur at adjacent sites, often in a given bay,
and to have these patterns consistent from year to year.
Results and Discussion Mercury (Hg) is generally enriched in Florida sites (Figure 3), where
12 of the 25 sites are well above average. The oysters from Old Tampa Bay
Exact sample location and the years in which samples were collected are especially high in Hg, rivaling even those from Lavaca Bay, Texas which
at each site are presented elsewhere (Wade et al., 1990). The geographical are known to be contaminated with Hg and where harvesting of oysters has
distribution for selected @ contaminants or suite of contaminants is shown in been limited because of the potential threat to human health.
Figures 1 to 7. The sites are listed in geographical sequence starting with the
southernmost Texas site and continuing along the coast to the southernmost Silver (Ag) distribution (Figure 4) was more .similar to that of Se than
Florida site. The smaller bars on Figures 1 to 4 represent "plus one standard to As, being somewhat enriched in Texas relative to Florida. The most
deviation". interesting feature of the Ag distribution is the low values in central Louisiana.
Trace metal concentrations in oysters varied considerably from site This same pattern was seen for cadmium (Cd) and is somewhat surprising
to site; in general, these variations were consistent over the four-year period. because central Louisiana Bays have been extensively disturbed by oil
exploration activities and are immediately downstream of the Mississippi River
That is, the same sites showed consistently above or below average I
concentrations each year, The high concentrations, with very few exceptions, outflow. In this area, then, intense activities by man does not seem to be
could not be shown to be associated with known activities of man, such as the influencing trace metal concentrations in oysters.
presence of industry or oil well drilling operations. However, the fact that The regional geographical distribution of the concentration of the sum
high values were often found in only one part of a particular bay (e.g., Tampa of 18 individual polyaromatic hydrocarbons (PAHs) (Wade et al., 1988) is
or Galveston Bay) while at other nearby sites in these same bays the shown in Figure 5. The concentration of PAHs for regional sites are plotted
concentrations were average or below for trace metals, suggests localized as the average. For example, for Galveston Bay 6 sites are averaged.
inputs of these metals.
Regional trends in trace metal concentrations in oysters arc more Two PAHs, fluoranthrene and pyrene, generally account for more
likely to be due to geologic or climatic factors than to activities of man. than 25% of the t *otal amounts detected. The predominance of these
Regional trends can be seen for only a few of the 13 metals assessed and, compounds would suggest the major source of PAHs is probably combustion
even for these, large site-to-site variations are superimposed on rather subtle and not oil seeps or oil spills. In general, higher PAH concentrations are
regional trends. For example, Figure 1 shows the distribution of selenium found at major river mouths where you also generally rind large urban areas
(Se) for the entire Gulf Coast. A gradual decrease in concentration is and associated industrial complexes. This is not surprising, since urban runoff
apparent when concentrations from Texas and Louisiana are compared to and sewage treatment plants are well known chronic sources of PAHs.
those in south Florida, even though some high values are found in northern The Panama City and St. Andrcv/s Bay regions are exceptions to this
Florida. trend. Their is no major river in these locations, yet they have the highest
Arsenic (Figure 2) is usually thought to be chemically similar to Se, PAH concentrations. It is possible that these sites were affected by an
but it shows a distribution pattern almost opposite to that of Se (Figure 1). episodic input of petroleum -(i.e., spill). The hydrocarbon distribution at
Arsenic (As) is much higher in some of the Florida oysters than elsewhere on these sites indicates they may be contaminated by used crank case oil.
the Gulf Coast, yet some Florida oysters, for example those from most sites An extensive interpretation of the chlorinated pesticide and
in Tampa Bay, had very low arsenic concentrations all four years. Only the polychlorinated biphenyl (PCB) data has been published elsewhere (Wade and
Tampa Bay site at Navarez Park near the city of St. Petersburg was Sericano, 1989; Wade et al., 1988 and 1990; Scricano ct.al., 1990a and b).
significantly enriched in As. Even the new site at Knight Airport on the edge T o t a I DDT ( s u in o, f o - p'D D E + p - p'D D E + o -
of the city of Tampa was low in As. It is possible that the extensive phosphate p'DDD+p-pDDD+o-p'DDT+p-p'DDT) regional distribution for oyster
rock deposits in Florida are a source of arsenic, but, based on the limited data
M M M 32M M M M Ml M

samples collected along (lie U.S. Gulf of Mexico coast is shown in Figure 6. REFERENCES
Total DDT is the most abundant chlorinated pesticide found in Gulf of
Mexico oysters. Most of the DDT is present as the metabolites, DDE and Brooks, J. M., T.L. Wade, E.L. Atlas, M.C. Kennicutt 11, BJ. Presley, R.R.
DDD. Less than 10% of the total contaminant load in oysters is the parent Fay, E.N. Powell, and G. Wolff. 1989. Analyses of Bivalves and Sediments for
compound, DDT. Organic Chemicals and Trace Elements from Gulf of Mexico Estuaries.
Annual Report 1989, 678 pp.
The regional distribution of total DDT shows that four of the five
highest concentrations are associated with major river outfalls including the Sericano, J.L., E.L. Atlas, T.L. Wade, and J.M. Brooks, 1990a. NOAA's
Brazos, Mississippi, Mobile, and Choctawatchee Rivers. There were also Status and Trends Mussel Watch Program: Chlorinated pesticides and PCB's
relatively high total DDT concentrations at St. Andrews Bay and Panama in oysters (Crassostrea yLirginica) and sediments from the Gulf of Mexico,
City, although no major rivers are found there. These are the same regions 1986-1987. Marine Environmental Research, 29: 161-203.
where the PAHs were the highest. DDTs associated with soils may be
transported downstream and collect in estuaries. This process provides a T.L. Wade, E.L. Atlas, and J.M. Brooks, 1990b. Historical
plausible explanation of the higher total DDT associated with major river perspective on the environmental bioavailabitity of DDT and its derivatives to
outfalls. The contin ued use of DDT in Mexico and other Latin American Gulf of Mexico oysters. Environmental Science and Technology, 24:
countries and its atmospheric transport and deposition to the sampling areas 1541-1548.
is another possible source.
Wade, T.L., E.L. Atlas, J.M. Brooks, M.C. Kennicutt 11, R G. Fox, J.
The regional distribution of PCBs is shown in Figure 7. PCBs were Sericano, B. Garcia-Romero, and D. Defreitas. 1988. NOAA Gulf of Mexico
detected in all NS&T oyster samples analyzed from Gulf of Mexico waters. Status and Trends Program: trace organic contaminant distribution in
The highest regional concentration was in St. Andrew's Bay. As mentioned sediments and oysters. Estuaries, 11: 171-179.
before for PAH and total DDT, this is an anomalous station and at present
we do not know The reason for the high concentrations at this site. Possible and J.L. Sericano. 1989 Trends in contaminant distribution in
sources of contaminants at this site may be nearby oil storage tanks and a oysters from the Gulf of Mexico. Pages 585-589 In: Proceedings of the
paper/pulp mill. The PCB concentrations do not show much difference on Oceans'89 Conference, Sept. 18-21,1989, Seattle, WA, Organotin Symposium,
a regional basis. All the regions have average concentrations within a factor Vol. 2.
of 5. There are somewhat higher concentrations near areas of higher
population density (i.e., Galveston Bay, Mobile Bay, etc.). -, J.L. Sericano, B. Garcia-Romero, J.M. Brooks, and BJ. Presley,
1990. Gulf Coast NOAA National Status & Trends Mussel Watch: The first
Conclusions four years. Pages 274-280 In: Proceedings, Marine Technology Society '90,
Washington, D.C., 1990.
Most of the contaminants monitored by the Status and Trends
Program have relatively long environmental half-lives. These contaminants,
in general, show no change in environmental concentrations over the first four
years of this study as seen in the standared deviation for trace metals (Figures
1 to 4). There are specific sites that are exceptions to this general trend and
they merit further detailed examination.

Acknowledgements

Funding for this research was provided by the National Oceanic and
Atmospheric Administration Grant Number 50-DGNC-5- 00262 (National
Status and Trends Mussel Watch Program).

327
326

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Copano Bay (1) Mesquite Bay (1)
Mesquite Bay (1) 0
San Antonio Bay (2) San Antonio Bay (2)
Espiritu Santo Bay (2) tZI t:1
Espiritu Santo Bay (2) H 0.4
0 Matagorda Bay (6) (n
Matagorda Bay (6)
Brazos River (1) Brazos River (1)
Galveston Bay (6) Galveston Bay (6)
Sabine Lake (1)
Sabine Lake (1) td (1) 0%% (7-4
Calcacieu Lake 2) 1-3
Calcacieu Lake (2)
Joseph Harbor Bayou (1)
Joseph Harbor Bayou (1)
Vermillon Bay (1) 0 Vermillon Bay (1)
Atchafalaya Bay (1) Atchafalaya Bay (1)
Caillou Lake (1) . Caillou Lake 1)
Terrebone Bay (2) Terrebone Bay (2)
Barataria Bay (3) Barataria Bay (3).
Mississippi River (2) Mississippi River (2)
Breton Sound (2) Breton Sound (2)
Lake Borgne (2) Lake Borgne (2)
Mississippi Sound (3) Mississippi Sound (3)
Mobile Bay (2) Mobile Bay (2)
Pensacola Bay (2) Pensacola Bay 2)
Choctawhatchee Bay (2) 0
Choctawhatchee Bay (2) Panama City (1)
Panama City (1)
San Andrew Bay (1) San Andrew Bay (1)
Apalachicola Bay (2) Apalachicola Bay (2)
Suwanee River (1)
Suwanee River (1) 0
Cedar Key (1) 0 Cedar Key (1)
Tampa Bay (5) Tampa Bay (5)
Charlotte Harbor (2) Charlotte Harbor (2)
Naples Bay (1 ) Naples Bay (1)
Rookery Bay (1) Rookery Bay (1)
Everglades
Everglades (1)

332 333

C@ C>
0 C) 0

Laguna Madre (2)
Corpus Christi (3)
Aransas Say (2)
Copano Bay (1)
Mesquite Bay (1) 0
t4
San Antonio Bay (2)
Espiritu Santo Bay.(2)
.(2)
Matagorda Say (6)
Brazos River (1) Q%%
Galveston Bay (6)
Sabine Lake (1) tZ
Calcacieu Lake (2) C.4
Joseph Harbor Bayou (1)
0
Vermillon Bay (1)
Atchafalaya Bay (1)
Caillou Lake (1)
Terrebone Bay
Barataria Bay (3)
Mississippi River (2)
Breton Sound (2)
Lake Borgne (2)
Mississippi Sound (3)
Mobile Bay (2)
Pensacola Bay (2)
Choctawhatchee Bay (2)
Panama City (1)
San Andrew Bay (1)
Apalachicola Bay (2)
Suwanee River (1) W
Cedar Key (1)
Tampa Bay (5)
Charlotte. Harbor (2)
Naples Bay (1)
Rookery Bay (1)
Everglades (1)

334

Reprint 3

Trace Organic Contamination in Galveston
Bay Oysters: Results from the NOAA National
Status and Trends Mussel Watch Program

Terry L. Wade,, Thomas J. Jackson, James M.
Brooks, Jos6 L. Sericano, Bernardo Garcia-
Romero and Dan L. Wilkinson

1-31

_____________________________________________________________________________________________________________________________

PROCEEDINGS

The Second State of the Bay Symposium
February 4-6, 1993
_____________________________________________________________________________________________________________________________

EDITORS

Richard W. Jensen
Texas Water Resources Institute,
Texas A&M University

Russell W. Kiesling
Galveston Bay National Estuary Program

and

Frank S. Shipley
Galveston Bay National Estuary Program

The Galveston Bay National Estuary Program
`
Publication GBNEP-23
February, 1993

1-33

Trace Organic Contamination in Galveston Bay
Oysters: Results from the NOAA National Status
and Trends Mussel Watch Program

Terry L. Wade, Thomas J. Jackson, James M. Brooks, Josi L. Scricano, Bernardo
Garcia-Romero and Dan L. Wilkinson
Geochemical and Environmental Rcsearcli Group, College of Geosciences and
Maritime Studies, Texas A&M University

It is important to determine the current status of contaminant concentrations in
order to assess the environmental response to management decisions that reduce or
stop the input of selected contaminants. To fill this information gap with high
quality data for U -S. coastal areas, the National Oceanic and Atmospheric
Administration (NOAA) established the National Status and Trends (NS&T) Mussel
Watch Program. As part of the NS&T Program, sediment and oyster samples have
been collected and analyzed from over 70 estuarine sites in the Gulf of Mexico
representing all major Gulf Coast estuaries. Sampling sites were located in areas not
influenced by known point sources of contaminant inputs, including Galveston Bay.
Oysters were employed as sentinel organisms because they are cosmopolitan,
sedentary, bioaccumulate, able to provide an assessment of bioavailability, not
readily capable of metabolizing contaminants, able to sukvive pollution loading,
transplantable, and commercially valuable. Oysters are, therefore, excellent
biomonitors for contamination in estuarine areas.

The Galveston Bay system is one of the largest and most economically important
estuaries along the U.S. Gulf Coast. This area has been the recipient of various
contaminant inputs because of an aggressively growing urban and industrial region.
Houston, Deer Park, Baytown, Texas City and Galveston, surrounding Galveston Bay
to the nort@h''I-and-west, are some of the most heavily industrialized areas in Texas.
Hundreds of ind'ustrial plants, including petrochemical complexes and refineries,
bordering the Galveston Bay estuarine system, as well as runoff, are likely to
introduce significant amounts of organic contaminants into the Bay. In general,
ecological studi6s have suggested that the waters of Galveston Bay contained
contaminants in sublethal amounts which caused stress to organisms resulting in
significant changes in the estuarine community structure. Galveston Bay NOAA
NS&T sampling sites (Figure 1) included the Ship Channel (GBSC), Yacht Club
(GBYC), Todd's Dump (GBTD), Hanna Reef (GBHR), Offats Bayou (GBOB) and
Confederate Reef (GBCR). Samples were collected in the winter starting in January
of 1986 at four sites (GBYC, GBTD, GBHR, GBCR) and in December of 1987 (Year 3) at
two additional sites (dBSC, GBOB). Samples were collected at some of these sites at
other times to provide information on seasonal trends in contaminant
concentrations. Sediments (top 1 cm) and oysters (20) were collected at three stations
at each site and analyzed for polynuclear aromatic hydrocarbons (PAH),
polychlorinated biphenyls (PCB), chlorinated pesticides (e.g., DDT, chlordane), and

1-34

W
W.
S; 9_"Z@ 6
Wa,
1:483
M .7'
L
;;'n 0,;
w.
T'O@ t",xis Mv
jHIP CNAAM

6AYTOWN

00
y
59
1ASSS
04Y
TRINI T Y SA Y

c_
5

CL CAR
-A LAKJ- 5AY 7.
C
DICKINSON 17
SAr
GALVESTON BAY ROLLOVCOC? POSSJ
SAW)
OICKIN 5_00
ire
0 5 A DOLLAR BA)

TEXAS CITY

VAY-
c_ @re GALVEST N

CHOCOL A rC
46A Y of

&Ar

5 b
SAN Luis I J_
PASS

CHRIS rmAS BAY

Figure 1. Galveston Bay sampling sites included the Ship Channel (59), Yacht
Club (15), Todd's Dump (16), Offat's Bayou (58), and Confederate Reef (18).

1-35

tributyltin. Sampling started in the winter of 1985/86 and is continuing with
sampling each winter. Seven years of data are currently available and Year 8
sampling has just been completed. All sample analyses were performed using
standard operating procedures (SOPs) to provide high quality, precise, accurate, and
reproducible data. Data quality was further assured by yearly participation in
NOAA/NIST intercalibration exercises. This allows for direct comparison of NS&T
Gulf Coast data with NS&T data for the East and West coasts.

Contaminant concentration patterns were similar for most contaminants. The
upper bay sites (GBSC, GBYC) had higher concentrations than the mid-bay sites
(GBTD, GBHR) for PAH, DDT, PCB, and butyltins. Sites from the lower bay (GBOB,
GBCR) had intermediate concentrations. This most likely results from proximity to
large urban areas and runoff inputs. The lower contaminant loading in the mid-bay
region probably results from dilution effects. For example, total PAH average
concentrations ranged from 20 to 15,000 ng/g. The higher concentrations were
measured in oysters from the upper portion of Galveston Bay (i.e., GBSC and GBYQ
and near the city of Galveston (i.e., GBCR and GBOB). Oyster s7ampleg from areas
farther away from urban centers (i.e., GBHR and GBTD) had average concentrations
one to two orders of magnitude lower. In general, these concentrations are in good
agreement with those previously encountered during temporal studies in Galveston
Bay. Two PAHs, pyrene and fluoranthene, generally accounted for >25% of the total
PAHs measured. The predominance of these compounds suggests that the maj
Jor
source of PAHs is from combustion products.

Average total PCB and DDT concentrations in Galveston Bay oysters were in the 48-
1100 and 12-240 ng/g ranges, respectively. Most of the DDT residue is present as
metabolites, DDE and DDD. In general, less than 10% of the total contaminant load
in oysters is the parent compound, DDT. Samples from the GBYC and GBSC were
the most contaminated while oysters from GBHR had the lowest residue
concentrations. ---These concentrations agree with the ranges reported earlier for
Galveston Bay bivalves. The median concentrations found in Galveston Bay for
PAH, chlordane, dieldrin, PCB, and butyltins are higher than the median
concentrations found throughout the Gulf of Mexico for the NS&T Program. The
median DDT concentrations found in Galveston Bay are about the same as those
found for the entire Gulf of Mexico. Therefore, compared to the rest of the Gulf of
Mexico the median concentrations of most organic contaminants are generally
higher in Galveston Bay. However, when Galveston Bay sites are compared to all
U.S. NS&T sites none of the concentrations, with the exception of chlordanes at
GBYC and GBSC, are ranked as high on a national scale.

Sample collections at other times of the year indicate some seasonal variability of
contamination concentrations. This may result from the loss of a considerable
amount of contaminants by oysters during spawning. Other studies of Galveston
Bay oysters indicate that body burdens of contaminants can change due to
accumulation and depuration. These preliminary studies indicate. that more
information regarding the use of oysters as bioindicators would provide for better
interpretation of the data from the NS&T program.

I
I
I
I
I
I
I
I Reprint 4
1 Indicators of Trace Metal Pollution In
I Galveston Bay
I Bobby J. Presley and Kuo-Tung Jiann
I
I
I
I
I
I
I

1-37
I

Proceedings
The Second State of the Bay Symposium
February 4-6, 1993

Editors

Ricard W. Jensen
Texas Water Resotirces Instiftite,
Texas A&M University

Russell W. Kiesling
Galveston Bay National Estuanj Prograrn

and

Frank S. Shipley
Galveston Bay National Eshiary Prograrn

The Galveston Bay National Estuary Program

Publication GBNEP-23
February,1993

1-39

Indicators of Trace Metal Pollution in Galveston Bay

Bobby Joe Presley and Kuo-Tting liann
Department, of Occanograpljy, Texas A&M University

Sediments and organisms are usually more reliable and more convenient media for
trace metal analysis than is water. Even polluted bay and estuarine water is very low in
trace metal concentration, making it difficult to analyze reliably. Furthermore,
concentrations in water are subject to rapid changes with changing metal inputs.
Sediments and organisms have higher metal concentrations and they integrate values
over time so less frequent sampling is needed.

Oyster (Crassostrea virginica) and other bivalves have been used as "sentinel" organisms
for assessing the pollution status of marine water bodies for almost twenty years. For
example, Goldberg et al (1983) report data for a USEPA funded "Mussel Watch"
program conducted in 1976-78, and the current NOAA-funded "National Status and
Trends Program" (NS&T) is an outgrowth and extension of the "Mussel Watch"
concept. Bivalves are widely recognized as being responsive to changes in pollution
levels in the environment, good accum, ulators of pollutants, widely distributed along
coasts, and easy to collect and analyze. Sediments also respond to changes in pollutant
trace metal inputs because most pollutant metals are particle reactive; that is, they
readily attach to particles which can then sink to the bottom and become part of the
sediments.

Oysters have been collected at six different sites in Galveston Bay (GB) since 1986 as
part of NS&T. Each site is on an identifiable oyster reef and, for the first 5 years, twenty
oysters were taken from each of three stations, the stations being 100 and 500 m apart.
Currently only one station is sampled at each site. Each site has been sampled once
each year, except two of the sites were not sampled the first two years. The twenty
oysters from each station are combined and analyzed as a single sample ea*.h year. In
most cases, stations are located hundreds of meters to many kilometers away from any
obvious point sources of pollutant inputs in an attempt to characterize large areas of
GB, rather than to identify specific point sources of pollutant input. Similar NS&T
sampling is conducted in all other major bays and estuaries along the U.S. Gulf of
Mexico coastline. The program allows different bays to be compared and pollutant
concentration changes with time at a given bay to be documented.

Data obtained by atomic adsorption spectrophotometry (AAS) after acid digestion of
oysters from the first four years of NS&T have been reported (Presley et al., 1990, 1991).
The samples were an 'alyzed for Ag, As, Cd, Cr, Cu, Fe, Hg, Mn, Ni, Pb, Se, Si, Sn, and
Zn. Flame AAS was used for Cu, Fe, and Zn, which exhibit high concentrations in
oysters, cold vapor AAS for mercury, and graphite furnace AAS for the remaining
elements. Blanks and reference materials were analyzed with the samples. Precision
and accuracy of the data was estimated to be �10%.

1-40

Trace metal concentrations found in oysters collected along the entire Gulf of Mexico
coastline during the first four years of NS&T were generally similar to those reported in
oysters taken from non-contaminated water in other parts of the world (Presley et al.,
1990). Only a few sites showed obvious trace metal pollution and these were restricted
geograph.ically such that nearby sites were usually unaffected. Abnormally high or low
values at a site did, however, usually repeat year after year suggesting local control.
Abnormal sites for most metals were just as likely to be visibly pristine as to be highly
industrialized.

Presley et al. (1991) reported that the oysters collected in Galveston Bay during the first
four years of NS&T were similar in trace metal content to those collected elsewhere
along the Gulf coastline, i.e., there was no indication of generalized trace metal
pollution in GB. The average Ag, Cd, Cr, Fe, Mn, and Pb in GB oysters differed by 10%
or less from the Gulf-wide average. Copper was 13% higher in GB, while Ni was 15%
higher, and Se 16% higher. Zinc, however, was 43% higher. Furthermore, the highest
Zn levels were found along the industrialized west side of Galveston Bay.

The four year NS&T sampling and analysis of oysters from the Gulf Coast discussed by
Presley et al. (1990, 1991) has been-'continued for three more years with at least an
additional three years planned. The basic patterns in concentration variability seen
earlier have not changed significantly. With few exceptions, Galveston Bay oysters
continue to be about average in trace metal content when compared to oysters from
other bays along the Gulf Coast. Furthermore, oysters from near the entrance to the
indand part of the Houston Ship Channel and from the industrialized western shoreline
have about the same metal content as those from pristine areas of East and West Bays.

In non-funded student research designed to further investigat e the relationship between
trace metal concentrations in oysters and proximity of industry, samples were taken at
twelve sites at the end of June and at the end of September, 1992. At most sites, 10-30
individual oysters were taken. They were collected, handled, and analyzed as
described previously (Presley et al., 1990). No oysters were collected in extreme
northern Galveston Bay, but shoreline samples were taken near Eagle Point and the
highly industrialized areas of Texas City. Samples were also taken in central GB along
the open-water part of the Houston Ship Channel and from East and West Bays.

Most trace metal concentrations were lower in oysters collected in September, 1992,
than those collected at the same locations in June, 1992. In many cases, the decrease was
by a factor of two and was, thus, larger than most site to site differences ill the bay' It is
very unlikely that this change was caused by human activity because there is little
correlation betwe.en metal concentrations in oysters and proximity to population or
industry, and even Fe concentrations in the oysters changed by up to a factor of two.
Rather, the change in trace metal concentration must be related to some physiological
change in the oysters. In order to minimize such changes, oysters are always collected
in December for NS&T. The September, 1992, data is similar to the six-year average
NS&T data, so perhaps oysters change in metal content less during fall and winter.

1-41

Silver concentrations are above the Gulf-wide average in several GB samples, but with
no clear relationship to proximity to industry. Very high Ag concentrations were found
in oysters collected at Confederate Reef in years V and VI (1990-1991) of NS&T, but not
in previous years. A site on Deer Island near Confederate Reef was sampled for the
1992 student work. Oysters from it were somewhat higher than average in Ag content,
but no more so than those from other sites in Galveston Bay. It is possible that human
activity is responsible for the silver and zinc enrichments but no specific causative
activity can be identified. In any case, the enrichments are not high enough to harm the
oysters or human health.

Based on the discussion above and other data from our laboratory, oysters seem to
integrate trace metal concentrations in the surrounding environment for one to two
months. For a longer integration period sediments can be analyzed. As part of the
unpublished student work reported here, sediments were collected at nineteen locations
throughout Galveston Bay, including Morgan's Point and other locations along the
industrialized northern and western shoreline, as well as locations far back into East
Bay well away from industry. The sediment was sieved to separate the <63 Prn grain
size fraction, which was analyzed along with an aliquot of the unsieved bulk sample.
Analysis was by AAS after both a partial leach of the sample with 0.5 N HC1 and
complete dissolution using HN03-HCI-HF. Results showed the sediment to be
generally constant in trace metal concentration from place to place when the <63 @m
size fractions were compared and to be similar to sediment from other Texas bays
which were analyzed for NS&T. Average concentrations of metals in the <63 pm
fraction of Galveston Bay sediments and the percentage ofthat metal leachable with 0.5
N HCI are shown in Table 1, along with average values for other Texas bays
(normalized to 100% <63 pm grain size). Data from Morse et al. (in press) on another
set of sediment samples taken from throughout GB confirms the relative constancy of
trace metal concentrations. The most notable exception to sediment trace metal
constancy found in the present work was a sample taken near the end of the Texas City
Dike. It had <0.5% fine material but that fine material was enriched in several metals.
Based on other data from this laboratory, it may well be that the fine fraction of very
sandy sediment is easily enriched in trace metals from human activity.

Table 1. Average cdncentratioiis of trace inetals hi the <63 joit sizefraction of
Galveston Bay and other Texas bay sedhnents.

Fe Ag As Cd CU Ni Pb Zn
(PPM) (PPM) (PPM) (PPM) (PPM) (PPM) (PPM)
GB Avg 2.9 0.164 8.21 0,157 28.7 23.9 24.5 98.8
GB S.D. 0.7 0.040 1.57 0.106 15.5 4.4 4.6 22.7
GB Leach (%) 15 61 19 76 56 21 68 33
TX Avg. 2.12 0.156 7.91 0.253 15.1 17.7 24.5 85.3
TX S.D. 0.83 0.055 3.27 (1171 3.5 4.2 6.0 25.2

1-42

Several species of finfish (flounder, drum, trout, catfish, etc.) as well as blue crabs and
oysters were collected from Galveston Bay in May-September, 1990, for the Galveston
Bay National Estuary Program (GBNEP) (Brooks, 1992). These were analyzed for trace
metals in our laboratory following procedures used for NS&T (Presley et al., 1990). The
samples came from near Morgan's point, Eagle point, Hannah Reef and Carancahua
Reef; thus, from areas of contrasting proximity to population centers and industry. In
spite of the contrasts between the collection sties, no clear differences were found in
trace metal concentrations in the organisms. Furthermore, the GB organisms were
similar in trace metal content to organisms from non-polluted bays elsewhere.

Oysters are better accumulators of trace metals and, being attached to the sediment,
should better characterize a site than the other organisms collected for GBNEP. The
GBNEP oysters were generally similar in trace metal content to NS&T oysters from GB,
but somewhat lower in Zn concentration. Zinc was also less clearly related to
population and industry than in NS&T.

GB fish flesh was much lower in trace metals than oyster flesh and while isolated high
values were found, concentrations were generally similar to those found -in non-
contaminated bays elsewhere. Trace metals in fish showed no relationship to
population or industry. Fish livers proved to have much higher concentrations of trace
metals (except for Hg) than fish flesh and high variability, but again no clear
relationship to population or industry. Blue crabs from GB were generally intermediate
in trace metal concentration between fish flesh and oyster flesh with similar high
variability and lack of correlation with population or industry. The GBNEP data,
therefore, gave no indication of trace metal pollution in Galveston Bay.

Bibliography

Brooks, J.M. 1992. Toxic contaminant characterization of aquatic organisms in
Galveston Bay: a pilot study. The Galveston Bay National Estuary Program
GBNEP-20.

Goldberg, E. D., M. Koide, V. Hodge, A.R. Flegal and J. Martin. 1983. U.S. mussel
watch: 1977-1978 results on trace metals and radionuclides. Estuarine, Coastal
and Shelf Science, 16: 69-93.

Morse, J.W., B.J. Presley, R.J. Taylor, G. Benoit and P. Santschi. (in press). Trace metal
chemistry of Galveston Bay: water, sediments and biota.

Presley, B.J., R.J. Taylor and P.N. Boothe. 1990. Trace Metals in Gulf of Mexico oysters.
The Science of the Total Environment, 97/98, 581-593.

1-43

Presley, B.J., R.J. Talyor and P.N. Boothe. 1991. Trace Metals in Galveston Bay oysters.
In, Proceedings, Galveston Bay characterization workshop, Feb. 21-23, 1991. The
Galveston Bay National Estuary Program Publication GBNEP-6, F.S. SI-lipley and
R.W. Kiesling (eds.).

Texas A&M Geochemical and Environmental Research Group. 1990. NOAA status and
trends mussel watch program for the Gulf of Mexico. Technical Report
submitted to National Oceanic and Atmospheric Administration, Rockville, MD.

1-44

Reprint 5

Oysters as Biomonitors of the
APEX Barge Oil Spill

Terry L. Wade, Thomas J. Jackson, Thomas J.
McDonald, Dan L. Wilkinson,
and James M. Brooks

1-45

Proceedi-n- -as

1993
International Oil Spill
Conference
(Prevention, Preparedness, Response)

March 29-April 1, 1993
Tampa, Florida

*onsored by: United States Coast Guard, American Petroleum Institute,
and U.S. Environmental Protection Agency

USCG API EPA

OIL POLLUTION CONTROL A COOPERATIVE EFFORT

1-47

OYSTERS AS BIOMONITORS
OF THE APEX BARGE OIL SPILL

Terry L. Wade, Thomas J. Jackson, Thomas j. McDonald,
Dan L. Wilkinson, James M. Brooks
Texas A&M University
Geochemical and Environmental Research Group
833 Graham Road
College Station, Texas 77845

ABSTRACT: The collision of the Greek tankership Shinoussa resulted Apex barge oil spill indicated that they efficiently metabolize the PAH
in a spill of an estimated 692,000 gallons of cxalyticfeed stock oil into after the initial insult.` This report discusses the use of oysters as
Galveston Bay on July 28, 1990. Oysters were collectedfrom Galveston biomonitors of the Apex barge oil spill at the historical NS&T Todds
Bay Todds Dump (GBTD) 235 daysprevious to the spill and 6,3 7,132, Dump site and at Redfish Island, a site reported to be impacted by the
and 495 days after the spill. Oysters were also collectedfrom Galveston od Spill.3
BayRedfish lslandfGBRI), asiteknown to be impacted by the spill, 37
and 110 days after the spill. The concentration of the 24 polynuclear
aromatic hydrocarbons (PAH) measuredfor the National Oceanic and Materials and methods
Atmospheric Administration's national status and trends program
(NS& T) site showed a sharp increasefirom about 100 nglg to over 600
nglg one week after the spill compared to concentrations 235 days Oysters (Crassostrea virginica) were collected for analyses from
previous to the spill. The concentration of the 24 NS& T PAH in oysters Galveston Bay Todds Dump and Redfish Wand. Individual stations at
from GBRI ranges from 400 to over 1000 ng1g. Soon after the spill the each site are generally from 100 to 1,000 m apart. An analysis at each
concentration of the 24 NS& T PAH at Todds Dump decreased to levels GBTD, from routine NS&T sampling program, represents a compos-
not statistically different from pre-spill samples. However, analyses of ite of 20 individual oysters, However, samples from GBTD and GBRI
alkylated and sulfur containing aromatic compounds indicate the oys- taken 6, 37, and 110 days after the spill represent from I to 20 oysters
ters were stillcontaminafed with Apex barge odw kast37 and 110 days depending on availability.
after the spill at GBTD and GBRI, respectively. Data from NS&T Tissue extraction followed the method used for NS&V Approx-
sampling at GBTD more than a year after the spill (495 days) indicates imately 15 grams of wet tissue were used for the PA14 analysis. After
thepresence of alkylated aromatic hydrocarbons that may befirom Apex the addition of internal standards (surrogates) and 50 grams of an-
barge oil still in the area. It appears that a sink of Aper barge oil (i.e., in hydrous Na2SO,, the tissue is extracted three times with dichlo-
sediments) may periodically be released by storms or other events into romethane using a tissuernizer. The solvent is concentrated to approx-
the ecosystem near GBTD. Therefore, bioavailable Apex barge oil is imately 20 mL in a flat-bottomed flask equipped with a three-ball
still present and may adversely affect oysters 495 days after the spill. Snyder column condenser. The tissue extract is then transferred to
Kudema-Danish tubes heated in a water bath (6(r C) to concentrate
the extracts to a final volume of 2 mL. During concentration, the
solvent dichloromethane is exchanged for hexane.
Oysters are analyzed as part of NOAA's national status and trends The tissue extracts are fractionated by alumina:sil@ (80 to 100
(NS&T) program' to detemiine the current status and long-term mesh) open-column chromatography. Ile silica gel is activated at 170*
trends of selected contaminant loadings. As part of this program, C for 12 hours and partially deactivated with 3 percent distWed water
pollynticlear aromatic hydrocarbons (PAH), toxic components of oils, (v/w). Twenty grams of silica gel are slurry packed in dichloromethane
are measured. Coastal waters are continually impacted by chronic over 10 grams of alurnina. Alumina is activated at 400* C for four hours
inputs of PAH from wastewater treatment plants, storm water runoff, and partially deactivated with I percent distilled water (v/w). The
atmospheric deposition, and the like. The NS&T program is designed dichloromethane is replaced with pentane by clution. The extract is
to determine the extent of this chronic contamination throughout the then applied to the top of the column. The extract is sequentially eluted
entire U.S. coastal area including the Gulf Coast. However, sporadic frowhe column with 50 mL of pentane (alipiatic fraction) and 200 mL
inputs of PAH into the coastal environment also come from small- and of 1:1 pentane:dichloromethane (aromatic fraction). The aromatic
large-scale oil spills.' The seven years of historical NS&T data along fraction is further purified by high-pcrformance liquid chromatogra-
with more recent U. S. EPA environmental monitoring and assess- phy to remove lipids. The lipids are removed by size exclusion using
ment-near coastal (EMAP-NC) data can be used as the basis for a dichloromethane as an isocratic mobde phase (7 mLJmin) and two 22.5
geochemical and environmental response strategy (GEARS) as de- X 250 mm Phenogel 100 columns.' The purified aromatic fraction is
smibed by Brooks et aL' This approach utilizes available data to collected from 1.5 minutes prior to the clution of 4,4'-dibromo-
provide historical control sites to determine the extent of a spill and octal3uorobiphenyt to 2 minutes after the clution of perylene. Tle
allow for a better estimate of ecosystem exposure. retention times of the two marker peaks are chocked prior to the
On July 28,1990, an estimated 692,000 gallons of catalytic feed stock beginning and at the end of a set of ten samples. The purified aromatic
oil product was spilled into Galveston Bay when a tanker collided with fraction is concentrated to I mL using Kudema-Danish tubes heated in
three Apex barges in the Houston Ship Channel. The spill was within a a water bath at 60* C.
mile of Todds Dump (GBTD), one of the historical NS&T oyster Quality assurance for each set of 20 samples includes a procedural
sampling sites. The spill resulted in the closure of recreational and blank, matfix spike, duplicate, and tissue standard reference material
commercial fisheries for several days. A study of frsh exposed to the 313 (NIST-SRM 1974), which are carried through the entire analytical

1-48

314 1993 OIL SPILL CONFERENCE

scheme. Internal standards (surrogates)are added to the samples prior
to extraction and are used for quantitation. The surrogaters are d6-
naphthalene, d10-acenaphthene, d10-phenanthrene, d12-chrysene, and
d12-perylene. Surrogates are added at a concentration similar to that
expected for the analytes of interest. To monitor the recovery of the
surrogates, chromatography internal standards d10-flourene and d12-
benzo(a)pyrene are added just prior to GC-MS analysis.
Gas chromatography-mass spectrometry (GC-MS). The PAHs were
separated and quantified by GC_MS (HP5890-GC interfaced to a
HP5970-MSD). The samples were injected in the splitless mode onto a
0.25 mm x 30 m (0.32 pm film thickness) DB-5 fused silica capillary
column (J & W Scientific, Inc.) at an initial temperature of 60 c and
temperature programmed at 12 C/min to 300 C and held at the final
temperature for 6 minutes. The mass spectral data were acquired using
selected ions for each of the PAH analytes. The GC-MS was calibrated
and linearity determined by injection of a multicomponent standard at
five concentrations ranging from 0.91 ng/uL to 1 ng/uL. Sample com-
ponent concentrations were calculated from the average response
factor for each analyte. Analyte indentifications were based on correct
retention time of the quantitation ion (molecular ion) for the specific
analyte and confirmed by the ratio of quantitation to confirmation ion.
Calibration check samples are run with each set of samples (begin-
ning, middle, and end), with no more than 6 hours between calibration
checks. The calibration check must maintain an average response
factor within + or - 10 percent for all analytes, with no one analyte greater
than + or - 25 percent of the known concentration. A laboratory reference
oil solution is also analyzed with each set of samples to confirm GC-MS
system performance and peak indentification.

Results and discussion

The location of the Apex barge oil spill is shown in Figure 1. A
detailed account of the spill, including cleanup and bioremediation
activities has been published. 3 the spilled oil was described as a
1. Location of Galveston Bay Todds Dump (GBTD) catalytic feed stock or similar to a No. 5 fuel oil 8 with a density of 0.92
and Galveston Bay Redfish Island (GBRI) collection g/mL. The Apex barge oil is not a typical Gulf Coast oil, but resembles
sites and the Apex barge oil spill a distillate or refined product. This is apparent from the gas chromato-

0.00 8.12 16.25 24.37 32.50 40.63 48.75 56.86 65
RT in minutes

Figure 2. Gas chromatogram of Apex barge oil spilled into Galveston Bay (normal alkanes with 21, 27, and 35)

1-49

FATE AND EFFECTS 315

Table 1. Oyster NS&T PAH, total PAH, and C3-phenanthrene
concentrations before and after the Apex barge oil spill

Davs before C3'
Collection or after NS&T PAH Total PAH phenanthrenes
Site date spill (ngig) (ng1g) (ng/g)
GBTD3 12/6189 -235 141 1,057 10
GBTD 2 12/6189 -235 175 471 10
GBT*D 1 12/6/89 -235 323 1,893 10
Apex Spill 7/28190 0 -1 -1 -1
GBTD 8/3/90 6 705 14,411 2,293
GBTD 9/3/90 37 122 852 77
GBTD I 12n/90 132 120 877 122
GBTD 2 12nlgo 132 150 1,204 178
GBTD 3 12nlgo 132 364 4,313 781
GBTD 12/5191 495 236 1,997 213
GBRI 1 9/3/90 37 790 19,146 2,735
GBRI 2 9/3/90 37 470 12,723 1,636
GBRI 11/15/90 110 1,110 25,213 3,238
1. Not applicable

graph (GQ shown in Figure 2.Tbe GC is a plot of detector response variability within a site for oyster samples collected on the same
versus increasing temperature and time. The area of peaks and their date makes evaluation of input from the Apex barge spill more prob-
retention times are used to determine the identity and concentration of lematic. However, the total PAH in these samples were predominantly
the components. The labeled peaks in Figure 2 are normal alkanes with lower molecular weight alkylated naphthalenes, alkylated fluorene
21, 27, and 35 carbon, respectively. Other peaks represent normal and Cj- and C2-phenanthrenes. No C3-phenanthrenes were detected
alkanes ranging from 15 to 37 carbons. The presence of oaly these (Table 1). Based on the fact that the Apex barge oil that was spilled
higher boiling components is consistent with a distillate or refined contained predominantly higher molecular weight hydrocarbons
product and is not typical of a Gulf Coast oil. (Figure 2), it is not surprising that C3-phenanthrenes might be better
The total oyster concentration of the 24 PAH measured as part of indicators of oyster exposure to the Apex barge oil than NS&T PAH or
NOAA NS&T program, the total of all PAH measured, and the total PAH. Therefore, the concentration Of C3-phenanthrenc was plot-
concentration of aU the phenanthrenes containing 3 carbon (C,-phen- ted versus the days before or after the spill (Figure 4). This log plot
anthrene) are provided in Table 1. Some of these data are also pres- indicates a clear distinction between oysters collected before the spill
ented graphicaUy in Figures 3 and 4. The total NS&T PA14 ranged and those collected after the spill. AD samples collected after the spill
from 120 to 1110 ng1g. The concentrations found at GBTD 235 days had detectable concentrations of C,-phenanthrenes, while none of the
before the spill ranged from 141 to 323. This GBTD data shows an samples collected before the spill do.
increase in concentrations as you move from the western onshore 'Me data for all the PAH individual compounds or groups of com-
station (GBTD-3) to the Offshore station (GBTD-1). The 24 NS&T pounds that are present in the Apex barge oil as well as oysters
PAH were only diagnostic of the spill at higher concentrations (i.e., collected from GBTD 235 days before and 132 days after the spill are
greater than 400 ng/g). Total PAH concentrations in oysters from shown in Figure 5, 6, and 7, respectively. A description of the compo-
GBID and GBRI ranged from 471 to 25,213 ng/g. The total oyster nents that were measured and plotted is fisted in Table 2. These plots
PAH concentrations when plotted versus the number ofdays before or provide a "fingerprint" ofthe Apex barge oil and the PAH distribution
after the spill that the samples were collected (Figure 3) clearly show found in the oysters. Clear differences appear in these fingerprints.
the influence of the Apex barge spill at concentrations above 10,000 ng/ The GBTD sample from NS&T Year 5 (235 days before the spill)
g. Based on total PAH in oysters, the bioavailabibty of Apex barge spill sampling contains mostly peaks in the left side of the plot (Figure 6)
oil is clearly evident for the sample collected at GBTD 6 days after the indicating a predominance of lower molecular weight PAR The
spill and samples at GBRI, located closer to the spill, 37 and 110 days GBTD NS&T Year 6 (132 days after the spill) samphng has mostly
after the spW (Table I and Figure 3). The oyster total PAH concentra- peaks in the midrange of the plot (Figure 7). The fingerprint from the
tion was 852 ng/g at GBTD 37 days after the spill. The total PAH oyster GBTD oysters after the oil spill is similar to the fingerprint for the
concentration 235 days before the spill at the three GBTD sites was Apex barge oil (Figure 5). This indicates that these oysters are still
1057, 471, and 1893, respectively (Table 1). Therefore the concentra- being exposed to Apex barge oil. The NS&T oyster samples collected
tion at GBTD 37 days after the spill is within the range of concentra- in Year 7 (495 days after the spill) also appear to contain a PAH
tions of total PAH found before, the spill (Figure 3). fingerprint consistent with bioaccumulation of Apex barge oil.

100000 10000.

1000,
'ElOODO z
0 0
too.-

z
1000.:
it. W I I -A
0 0 TOTAL GErM 10 C3 PHEN G
0 TOTAL GBRI UP
U
too--
-400 -200 0 200 400 60D -400 .200 0 200 400 600
DAYS BEFORE H OR AFTER (+) THE APEX SPELL DAYS BEFORE OR AFTER (+) THE APEX SPILL
Figure 3. Total PAK concentrations in oysters before and after the Figure 4. CI-phenanthrene concentrationsin oysters before and after
Apex barge oil spill the Apex barge oil spill

1-50

316 1993 OIL SPILL CONFERENCE

800 Table 2. Polynuclear aromatic hydrocarbons (PAH) analyzed

700 ____________________________________________________________________________

600 Abbreviation Analyte
____________________________________________________________________________
500 Naph naphthalene
C1-naphthalenes
400 C2-naphthalenes
C3-naphthalenes
300 C4-naphthalenes
Bi biphenyl
200 acenaphthylene
acenaphthene
100 Fl fluorene
C1-fluorenes
0 C2-fluorenes
Naph Fl P CH BaP BPe C3-fluorenes
Bi DBT FL BbF I DBT dibenzothiophene
C1-dibenzothiophenes
C2-dibenzothiophenes
Analytes C3-dibenzothiophenes
P phenanthrene
Figure 5. Apex barge oil PAH fingerprint (see anthracene
Table 2 for abbreviations) C1-phenanthrene-anthracenes
C2-phenanthrene-anthracenes
C3-phenanthrene-anthracenes
C4-phenanthrene-anthracenes
350 Fl fluoranthene
pyrene
300 C1-fluoranthene-pyrenes
benz (a) anthracene
250 Ch chrysene
C1-chrysenes
200 C2-chrysenes
C3-chrysenes
150 C4-chrysenes
BbF benz (b) fluoranthene
100 benz (k) fluoranthene
benzo (e) pyrene
50 BaP benzo (a) pyrene
pyrene
0 I indeno{1,2,3-cd}pyrene
Naph Fl P CH BaP BPe dibenz{a,h}anthracene
Bi DBT Fl BbF I BPe benzo{g,h,i}perylene

Abalytes 1. Used in Figures 5 through 7

Figure 6. Oysters PAH fingerprint 235 days before
the Apex barge oil spill, NS&T Year 5 (see Table 2
for abbreviations)
Conclusions

The analyses of oyster samples before and after the Apex barge oil
900 spill indicate that oyster do act as biomonitors and that the spilled oil is
bioabailable. Measurement of total PAH was diagnostic of exposure
800 for months after the spill; however the complication of other chronic
sources of input make diagnosis of exposure specific to Apex barge oil
700 more difficult. The use of fingerprinting, using all of the available PAH
data coupled with the historical NS&T data, suggests that Apex barge
600 oil or a similar oil is still present in the vicinity of the GBTD NS&T
sampling site. The Apex barge oil could be trapped in the sediments in
500 the shallow surrounding areas. There is then the potential for periodic
releases during storms or other events that disturb the sediments. A
400 more extensive data set might have provided enough information to
quantitate this possibility better. However, since the decision was made
300 to do a Type A assessment of the damage,2 which is based on a natural
resource damage assessment model for coastal and marine environ-
200 ments (NRDAM/CME) and not on environmental monitoring, no
extensive data set exists. The damage assessment model does not
100 consider the fact that bioavailable PAH from the Apex barge oil spill
may still be present 495 days after the spill. While the use of computer
0 models to assess damage is a politically expedient measure, in the case
Naph FL P CH BaP BPe of the Apex barge spill there is insufficient date to determine it it
Bi DBT Fl BbF I adequately considers long-term damae to the enviromnent. More
Analytes research is needed to address this possible limitation of the model.

Figure 7. Oyster PAH fingerprint 132 days after
the Apex barge oil spill NS&T Year 6 (see Table 2
for abbreviations)

1-51

FATE AND EFFECTS 317

Acknowledgments 4. Jackson. T. J., T. L. Wade, T. J. McDonald, D. L. Wilkinson, and
J. M. Brooks, in press. Polynuclear aromatic hydrocarbon contami-
nants in oysters from the Gulf of Mexico (1996-1990). Environmen-
This research was partially funded by the National Oceanic and tal Pollution
Atmospheric Administration (NOAA), U. S. Department of Com- 5. Krahn. M. M., L. K. Moore, R. G. BogaT, C. A. Wigren,
merce, Grants 50-DGNC-5-00262 and 50-DGNC-0-00047 from the Chan. and D. W. Brown, 1988. High- performance liquid chroma-
Office of Ocean Resources Conservation and Assessment; U.S. Envi- tOgT2phV method for isolating organic contaminants from tissue and
ronmcntal Protection Agency, Galveston Bay National Estuaries Pro- sediinen't extracts. Journal of Chromatography, v 437, ppl6l-175
gram, Contract IAC (90-91) 1145; and the Geochernical and Environ- 6. McDonald, S. J., J. M. Brooks, D. Wilkinson, T. L. Wade, and
mental Research Group, Texas A&M University. T. J. @ IcDonald, 1991. 'nc effects of the Apex barge oil spill on the
fish of Galveston Bay. Proceedings of the Galveston Bay Character-
ization Workshop, Feb. 21-23, 1991, pp85-86
7. McDonald, S. J., T. L. Wade, J. M. Brooks, and T. J. McDonald,
1991. Assessing the exposure of fish to a petroleum spill in Gal-
References veston Bay, Texas. in Water Pollution: Modelling, Measuring and
Prediction, L. C. Worbel and C. A. Brebbia, eds. Computational
I . Brooks, James M., Michael A. Champ, Terry L. Wade, and Sus- Mechanics Publications, Southampton and Elsevier Applied Sci-
anne J. McDonald, 1991. GEARS: Response strategy for oil and ence, London, pp707-718
hazardous spHls. Sea Technology, April 1991, pp25-32 8. Nelson, C. 1., 1990. NOAA Response Report, October 29, 1990.
2. Galveston Bay Oil Spill Plan, 1992. Draft Assessment Plan for the Apex T/B Barges 3417 & 3503 Collision and Oil Spill, Houston Ship
Measurement of Natural Resource Damages. Available from J. H. Channel Buoy 58, Galveston Bay, Texas, 29 July 1990
Jeansonne, Damage Assessment Center, Southeast Region, 9450 9. Wade, T. L., E. L. Atlas, J. M. Brooks, M. C. Kennicutt 11, R. G.
Koger Blvd., St. Petersburg, Florida 33702. 17pp Fox, J. Sericano, B. Garcia-Romero, and D. DeFreitas, 1998.
3. Green, T. C., 1991. 'ne Apex barges spill, Galveston Bay, July NOAA Gulf of Mexico status and trends program: Trace orgarlic
1990. Proceedings of the 1991 Intemational Oil Spill Conference, contaminant distribution in sediments and oysters. Estuaries, v 11,
American Petroleum Institute, Washington, D.C., pp291-297 ppl7l-179

1-52

Reprint 6

Field Studies Using the Oyster Crassostrea
virginica to Determine Mercury accumulation
and Depuration Rates

Sally J. Palmer, Bobby J. Presley,
Robert J. Taylor, and Eric N. Powell

1-53

Bull. Environ. Contam. Toxical. (1993) 51:464-470 Environmental
1993 Springer-Vertag New York Inc. Contamination
and Toxicology

Field Studies Using the Oyster Crassostrea virginica To
Determine Mercury Accumulation and Depuration Rates

Sally J. Palmer,1 Bobby J. Presley,1 Robert J. Taylor,2 and Eric N. Powell1

Department of Oceanography, Texas A&M University, College Station, Texas
77843, USA and 2On-Site Analytical, Houston, Texas 77082, USA

Mercury as an environmental hazard, especially with regard to human health,
has been of concern since the Minamata diaster (Huddle et al. 1975). From
1966 to 1970 a chlor-alkali plant in Point Comfort, texas released mercury-
enriched wastewater (up to 29.9 kgHg/day)into Lavaca Bay (TWQB 1977).
Since 1970 the Texas Department of Health (TDH) has periodically closed
and the re-opened portions of Lavaca Bay to the harvesting of crabs and
finfish based on their levels (<>0.5 ppm Hg wet weight)of mercury. A
1988 closure remains in effect as of this writing (Wiles, 1993). Mercury
contamination in Lavaca Bay organisms thus continues to be a problem 22
years after the chlor-alkali plant ceased releasing mercury into the bay. The
goal of the following research was to better understand the behavior of
mercury in Lavaca Bay.

Oysters have been widely used as indicator species in metal pollution studies
(Goldberg et al 1983). Most such programs have focused on the
concentrations of metals in oysters from different geographic areas. This
study, however, investigated the rate and amount of mercury a "clean" oyster
would accumulate when transplanted to a contaminated estuary and the rate of
mercury depurations by contaminated oysters placed in a clean environment.
The oysters were additionally analyzed bor Ba, Cu, Fe, P, and Zn to test for
the possible involvement of these metals in mercury accumulation and
depuration.

MATERIALS AND METHODS

In August 1991, mature Crassostrea virginica were collected from an
uncontaminated area, Carancahua Reef, in Carancahua Bay, Texas for use in Figure 1. Map of Lavaca Bay showing caging sites and oyster collection
the accumulation study. The oysters were placed in nylon bags with a mesh locations for the accumulation and depuration experiments.
size of 0.5 cm. Each bag contained about 50 oysters. The bags were taken
about 16 km away to Lavaca Bay and a control site, Keller Bay (see Fig.1) an area of North Lavaca Bay known to be contaminated with Hg. These
and placed on plastic grates, which prevented them from sinking into the soft oysters were placed in a contaminated area of lower Lavaca Bay, as a
sediment at the 1 m deep sites. Nine transplanted oysters from each site were control site, and in uncontaminated Keller Bay (see Fig. 1). Deployment
collected on days 0, 7, 14, 21, 36, and 51, placed in plastic bags, and frozen techniques and collection times were the same as those used in the
for later analysis. accumulation study.

The depuration study, also conducted in August 1991, used C, virginica from On each sampling date oysters were removed from the experimental bags in
both the accumulation and depuration experiments and nine individuals
from the natural population of C virginica at the respective sites were
Send reprint requests to Sally Jo Palmer at the above address. collected.

In the lab, oysters were thawed and opened with a stainless steel knife. The
soft tissue was removed with a teflon spatula and plastic forceps, rinsed in

464 465

1500 3000 + Lavaca Bay
Keller Bay

1000 2500

500 + Lavaca Bay 1500
Keller Bay

0 1000
0 10 20 30 40 50 60

Day 500

0
0 10 20 30 40 50 60
Day

Figure 2. Mercury concentrations in Carancahua Reef Figure 3. Mercury concentrations in oysters collected
oysters transplanted to Lavaca and Keller Bays for the 51-d from North Lavaca Bay when transplanted to lower
accumulation experiment. Lavaca and Keller Bays for the 51 day depuration experiment.

Table 1. Certified and experimental data for Natural Bureau of Standards Statistical analyses using SAS Institute, Inc. software (SAS INSTITUTE
reference material 1566a Oyster Tissue and the elemental detection limits INC> 1985) were performed on the data from the three oyster groups,
obtained in this study. accumulation (aa), depuration (DD), and the natural population (NP) to
investigate possible relationships among the variables analyzed. The
Element Hg Ba Cu Zn P Fe Spearmen correlation test and the general linear model (GLM) were used to
(ppb) (ppm) (ppm) (ppm) (ppm) (ppm) find correlations and linear relationships among variables. The dry weight of
the oysters was used as a covariant in the GLM. The Least Square Means
NBS Certified test, LSMEANS, was used to verify changes in Hg with time during the
1566a Oyster Tissue caging experiments.
64.2+ - NC 66.3+ - 830+ - 6230+ - 539+ -
6.7 4.3 57 180 15 RESULT AND DISCUSSION
Experimental
1566a Oyster Tissue Oysters removed from Carancahua Bay readily accumulated mercury when
(n=15) 55.7+ - 1.49+ - 73.8+ - 935+ - 6203+ - 549+ - placed in Lavaca Bay. Mercury accumulaton was rapid through the first 14
2.0 4 4.7 36 260 43 days of exposure and leveled off with time (Fig. 2). The rate of Hg
accumulation between days 0 and 14 averaged 70 ppb Hg per day. From day
Detection 6 0.4 10 90 680 110 15 to 51 mean daily Hg uptake ranged from 0 to 10 ppb. The Hg levels in the
Limit control oysters from Carancahua Bay did not significantly change over the 51-
__________________________________________________________________________________ d experiment when caged in Keller Bay. The GLM procedure showed a
NC=not certified significant (p<0.05) difference in Hg levels between sites and days collected.

The LSMEANS test showed that Hg concentrations in oysters caged in
distilled-deionized water, weighed, freeze dried, and homogenized before Lavaca Bay on days 7, 14, 21, 36, and 51 were all significantly higher than
analysis. All oysters were individually digested according to a modification day 0 at p<0.05, while the Hg levels on days 14, 21, 36, and 51 were not
of the USEPA 245.1 (USEPA 1990) method and analyzed for mercury by significantly different from one another.
cold vapor atomic absorption spectrophotometry (Hatch and Ott 1968).
Samples were additionally analyzed for Ba, Cu, Fe, P, and Zn using a The results of the depurations study showed that contaminated oysters released
modified Applied Research Laboratories, Inc. (ARL) SpectraSpan VI Hg when placed in an uncontaminated environment, Keller Bay (Fig. 3). The
Direct Current Argon Plasma (DCP)Emission Spectrophotometer following average Hg level dropped from 1660 + - 363 ppb on day 0 to 550 + - 416 ppb
ARL's instructions (ARL 1991). Every digest (about 30 samples) included on day 51. The rate of Hg depuration varied throughout the experiment. The
two aliquots of the reference material, 1566a Oyster Tissue, certified by the rate was highest on days 14 to 21, averaging a loss of 72 ppb Hg per day.
National Bureau of Standards. The certified and experimental values from the According to the LSMEANS test, concentrations in oysters caged in Keller
1566a Oyster Tissue and the detection limits for each element are listed in Bay on days 21, 36, and 51 were significantly lower than Hg levels in oysters
Table 1. from North Lavaca Bay transplanted to Keller Bay on day 0 at p<0.05.

466 467

Table 2. Correlations among variables measured in the oyster accumulation showing that Hg contaminated is a continuing problem in Lavaca Bay and
and depuration experiments and the natural populations in Keller (KB) and that the Hg is readily available for bioaccumulation by oysters.
Lavaca Bay (LB). Upper, Spearmans rho; lower, P value

Variables Accumulatons Natural Population Depuration Positive correlatons between Cu and Zn, Zn and P, and Ba and Fe were
KB LB KB KB KB LB found in every experiment (Table 2). Copper and phosphorus were positively
correlated in every oyster group except the natural population from Keller
Cu and Zn 0.83100 0.81849 0.91104 0.87307 0.89350 0.82946 Bay. A positive relationship between P and Fe and a negative correlation
(<0.0001) (<0.0001) (<0.0001) (<0.0001) (<0.0001) (<0.0001) between Hg and oyster dry weight were found in several experiments.
Correlations were also seen between Hg and Cu, Hg and Zn, and Cu and Zn
Zn and P 0.40244 0.42855 0.28371 0.40736 0.32464 0.37663 in oysters from Lavaca Bay, i.c., those collected for the depuration
(0.003) (0.001) (0.04) (0.002) (0.02) (0.005) experiment and the natural population of oysters from Lavaca Bay.

Ba and Fe 0.64172 0.62662 0.60290 0.60290 0.68725 0.49675 Positive relationshiops between Cu and Zn in bivalves such as those found
(0.0001) (<0.0001) (<0.0001) (<0.0001) (<0.0001) (<0.0001) here have been noted in previous studies (Paez-Osuna and Marmolejo-Rivas
1990a, Marcus and Thompson 1986, Wright et al. 1985, and Paez-Osuna and
Cu and P 0.37744 0.41719 NS 0.36245 0.36627 0.49388 Marmolejo-Rivas 1990b). Copper and zinc are both biologically active
(0.005) (0.002) (0.002) (0.007) (<0.0001) elements, but the reason for their strong correlation is not understood.
Previous work (George and Pirie 1980)found that Zn transferred in the
P and Fe 0.35252 NS 0.28380 NS 0.41407 0.33303 plasma of Mytilus cululis was mostly associated with granules that also
(0.01) (0.04) (0.002) (0.01) contained FE, S, P, K, and Ca. Others have indicated that M, edulis
sequesters Zn and Fe in lysosomes in various cell types (Lowe and Moore
Dry Wt. NS -0.40646 -0.48637 NS -0.46336 -0.38954 1979). George et al. (1978) found Zn and Cu were immobilized in individual
and Hg (0.002) (0.0003) (0.0004) (0.004) granular amoebocytes; granular cells whih contained Cu and Zn were present
in Ostrea edulis, O, angasi, and C, gigas. It is believed that oysters
Hg and Cu NS NS NS 0.55418 0.66808 0.29423 concentrate Cu, Zn P and other metals in granules to detoxify and change
(<0.0001) (<0.0001) (<0.0001) them to an excretable form (George and Pirie 1980, george et al. 1978). The
correlatons between Cu and Zn, Cu and P, and Zn and P found in this study
Hg and Zn NS NS NS 0.56905 0.70215 0.43221 could be related to granular formation in C, virginica, but no documentation
(<0.0001) (<0.0001) (<0.0001) of this was obtained.

Cu and Fe NS NS NS 0.26899 0.49198 0.37663 The relationship among the elements Hg, Zn, and Cu may be a result of the
(0.05) (0.0002) (0.005) high Hg concentrations in the water and sediment in Lavaca Bay. Perhaps Cu
and Zn function in protective mechanisms in the detoxification of Hg, in C,
virginica. Another possible explanation is that Lavaca Bay oysters contain
NS = Not significant more metal-binding granules or low molecular weight binding proteins,
metallothioneins, which could detoxify the metals (Lobel and Payne 1984).
The strongest correlation between the three elements was found in oysters in
The average oyster living in the transplant sites in Lavaca and Keller Bays the depuration study, where Cu and Zn closely followed the trend of
contained 2068 + - 676 ppb Hg and 354 + - 124 ppb Hg respectively. The decreasing Hg with time.
Carancahua Bay oysters caged in Lavaca Bay for the accumulation experiment
increased dramatically in Hg concentraton but did not reach the high average Acknowledgments. Research supported in part by Texas A&M University
Hg levels in the natural population of oysters in Lavaca Bay. Similarly, Seagrant Program, Institutional Grant #NA89aa-d-sg-139. The authors
oysters caged in Keller Bay for the depurations experiment decreased in Hg also thank the TAMU trace metal laboratory group for their invaluable help
from 1660 + - 363 to 550 + - 416 but did not acquire the low Hg concentrations with field work.
found in the natural population of oysters in Keller Bay.

The transplant experiments clearly show that C virginica rapidly accumulated REFERENCES
mercury when placed in a contaminated environment, Lavaca Bay.
Furthermore, mercury-contaminated C virginica from Lavaca Bay were ARL (1991) DIRECT CURRENT PLASMA (DCP) optical emission
found to quickly depurate Hg when placed in an uncontaminated spectrometric method for trace elemental analysis of water and wastes
environment, Keller Bay. Although the initial rate of Hg uptake was much method AES0029. Applied Research Laboratories
faster than was the release, the oysters in the accumulation and depuration George SG and Pirie BJS (1980) Metabolism of zinc in the mussel, Mytilus
experiments both changed average Hg levels by about 1000 ppb over the 51-d edulis (L): a combined ultrastructural and biochemical study. J Mar Biol
experiment. The results of this study confirm earlier work (Riegal 1990) Assoc UK 60: 575-590

468 469

George SG, Pirie BJS, Cheyne AR, Coombs TL, and Grant PT (1978)
Detoxification of metals by marine bivalves: an ultrastructural study of the
compartmentation of copper and zinc in the oyster Ostrea edulis. Mar Biol
45: 147-156
Goldberg ED, Koide M, Hodge V, Flegal AR and Martin J (1983) U. S.
Mussel Watch: 1977-1978 results on trace metals and radionuclides,
Estuarine Coastal Shelf Sci 22: 395-402
Hatch WR and Ott WL (1968) Determination of sub-microgram quantities of
mercury by atomic absorption spectrophotometry. Anal Chem 40: 2085-
2087
Huddle N, Reich M, and Stiskin N (1975) Island of dreams: Environmental
crisis in Japan. Autumn Press, New York
Lobel PB and Payne JF (1984) An evaluation of mercury-203 for assessing
the induction of metallothionein-like proteins in mussels exposed to
cadmium. Bull Environ Contam Toxicol 33: 144-152
Lowe DM and Moore MN (1979) The cytochemical distribution of zinc (Zull)
and iron (FE III) in the common mussel, Mytilus edulis, and their
relationship with lysosomes. J Mar Bio Assoc UK 59: 851-858
Marcus JM and Thompson AM (1986) Heavy metals in oysters around three
coastal marinas. Bull Environ Contam Toxicol 36: 587-594
Paez-Osuna F and Marmolejo-Rivas C (1990a) Occurrence and seasonal
variation of heavy metals in the oyster Saccrostrea iridescens. Bull
Environ Contam Toxicol 44: 538-544
Reigel DV (1990) The distribution and behavior of mercury in sediments and
marine organisms of Lavaca Bay, Texas. December 1990
SAS Institute Inc. (1985) SAS User's Guide: Statistics, Version 5 Edition.
SAS Institute Inc.
TWQB Texas Water Quality Board (1977) Water Quality Segment Report for
Segment No's 2453 and 2454 Lavaca and Cox Bays. TWQS-21
USEPA (1990) Contract Laboratory Program Statement of Work for
Inorganics Analysis. Document Number ILM01.0
Wiles, K. (1993) Personal communication on closure of Lavaca Bay to
fishing. Texas Department of Health, Shellfish Sanitation Control
Division. Austin, TX
Wright DA, Mihursky JA, and Phelps HL (1985) Trace metals in Chesapeake
Bay Oysters: Intra-sample variability and its impliations for
biomonitoring. Marine Environ Res 16: 181-197.

Received December 30, 1992; accepted March 1, 1993

470

Reprint 7

Trace Metal Chemistry of Galveston Bay:
Water, Sediment and Blota

John W. Morse, Bob J. Presley,
and Robert J. Taylor

1-59

Marine Environmental Research 36 (1993) 1 37

Trace Metal Chemistry of Galveston Bay:
Water, Sediments and Biota

John W. Morse, Bob J. Presley, Robert J. Taylor

Department of Oceanography, Texas A&M University, College Station
Texas 77843, USA

Gaboury Benoit* & Peter Santschi

Department of Marine Sciences, Texas A&M University at Galveston, Galveston
Texas 77553, USA

(Received 5 Septemter 1991; revised version received 30 January 1992
Accepted 5 February 1992)

ABSTRACT

Galveston Bay is the second largest estuary in Texas. It receives major
urban runoff from the Houston area, its major river drains the Dallas-
Ft Worth Metroplex, and the area surrounding the Bay is intensely
industrialized, with chemical and petroleum production being especially
prominent. Consequently, there are serious concerns about the possible
contamination of the Bay and previous studies have indicated toxic metals
at elevated concentrations (e.g. NOAA, 1989a).
We have conducted an exgtensive investigation of Galveston Bay trace
metals, in which their distribution in the water column, oysters and sedi-
ments were determined. Results of the water column and oyster analyses
indicate that metal levels in open areas of Galveston Bay are currently
similar to those in more pristine bays elsewhere. Industrial metal inputs to
the Bay have not led to greatly increased concentrations in water, sedi-
ments and biota. However, the sediment analyses indicated that such
inputs may have been significant in the past. Total Cu, Zn, Pb, and
Ag concentrations in the waters, determined by state-of-the-art clean

* Present address: Yale School of Forrestry and Environmental Studies, Sage Hall, 205
Prospect St., New Haven, Connecticut 06511, USA

Marine Environ. Res. 0141-1136/93/$06.00 1993 Elsevier Science Publishers Ltd. England
Printed in Great Britain

2 J.W. Morse, R. J. Presley, R. J. Taylor, G. Benoit, P. Santschi Trace metal chemistry of Galveston Bay 3

techniques, are 1. 2-7, 0-3, and 0-0006 pg liter, respectively, and are The combination of large population, high industialization, shallow
mostly regulated by the dynamcs of sediment suspension and settling. depth and restricted water exchange gives Galveston Bay the potential
This leads to a correlation of particulate trace metal concentration with for serious trace metal contamination problems. Such problems have,
the suspended particulate matter (SPM) concentrations, and trace metal however, not been well documented. Hann & Slowey (1972) showed sedi-
enrichment in particles at low SPM concentrations. Forty-four percent of ments in the uper, confined parts of the Houston Ship Channel to be
the individual sediment sampling sites exhibited an anomalous concentra- highly enriched in several trace metals and this has been confirmed by
tion with respect to at least one of the metals studied and about half of yearly sampling by the Texas Water Quality Board and Texas Water
these sites were directly associated with dredge spoils. The study also indi- Commission (TWC) since 1974 (Texas Water Commission, 1987). How-
cated that many of the metals are significantly converted to a coprecipitate ever, sediment from the ship channel where it crosses the open Galveston
with pyrite in the top 10 cm of sediment. Bay did not appear to be contaminated in these early studies. TWC ana-
lyses for trace metals in Houston Ship Channel water show a decline in
all metals between 1974 and 1986 but the quality of the data is in ques-
INTRODUCTION tion, so no firm conclusions can be reached. The only reliable previously
published data for dissolved trace metals in open Galveston Bay waters
Galveston Bay, with a surface area of 1600 km', is one of the largest em- are thought to be those of Tripp (1988) who determined only As and Sb
bayments on the US coastline. The Bay water is, however, very shallow, concentrations. These were both near normal seawater values in the open
averaging only about 2 m in depth and is largely cut off from the Gulf of Bay but As was enriched by up to a factor of five in the Houston Ship
Mexico by the Bolivar Peninsula and Galveston Island. Tides, which av- Channel and Sb was depleted there.
erage about 40 cm in height, thus exchange water primarily through a One of the feared effects of contaminant inputs into coastal embay-
channel between these two land barriers. It is generally accepted that ments is this buildup in water and sediments leading to accumulaton
winds are often more important than tides in Bay circulation and water and negative effects in biota. For this reason, oysters and sedentary
exchange (NOAA, 1989a), but exact current patterns in the Bay and the organisms have been advocated as pollution biomonitors (Goldberg, 1975;
residence time of water in the Bay with respect to exchange with the Gulf Goldberg et al., 1983). Bioaccumulation in benthic fauna can occur from
of Mexico are not well known. A mean residence time for waters in the both dissolved and particulate forms. Generally higher trace metal con-
Bay of about 40 days has been estimated from its average salinity of 15 centration in either water or sediments lead to elevated eoncentrations in
and an average river flow of 12 km year (Armstrong, 1982). the biota, depending on the organism type, trace metal speciation and
The Trinity River supplies about 70% of the freshwater that enters the solid carier phases. Bioabailability of trace metals can be controlled by
Bay, with the remainder coming from the San Jacinto River and from many factors, for example, by the amount and speciation of iron in sedi-
ungauged flows (NOAA, 0989a). The Dallas, Ft. Worth Metroplex with its ments. An inverse correlation between bioaccumulation of Pb and bound
4 million people is located 400 miles up the Trinity River from Galveston Fe concentration in sediments has been found (Luoma & Bryan, 1978).
Bay and is separated from the Bay by a large freshwater lake. The Hous- Therefore, a general overview of trace metal chemistry has to include all
ton metropolitan area with its 3 million people is immediately adjacent to three reservoirs (i.e. water, sediments, and biota) and has to pay special
upper Galveston Bay and drains directly into it through the San Jacinto attention to the chemical forms of Fe and other geochemical indicator
River and the Houston Ship Channel. The Houston Texas City Galves- elements or key phsicochemical variables. Very few such comparisons
ston area is intensively industrialized, especially by the petroleum, petro- which used state-of-the-art techniques for the analysis of all three reser-
chemical and chemical industries. For example, 30 50% of the US voirs are available in the open literature.
chemical production and oil refineries are situated around Galveston In order to obtain such an overview indicating possible heavy metal
Bay. This industrialization and the large population results in Houston contamination in Galveston Bay, we have combined the results of three
being the third largest seaport in the US in terms of total shipping ton- studies. These studies were of trace metal concentrations in both dis-
nage. Galveston Bay receives more than half of the total permitted solved and particulate form in the water column (studies by Benoit and
wastewater discharges for the state of Texas and a total of about 5 km Santschi), concentrations in oysters (studies by Presley and Taylor), and
year of wastewater input. their form and distribution in sediments (studies by Morse).

4 J. W. Morse, R. J. Presley, R. J. Taylor, G. Benoit, P. Santschi Trace metal chemistry of Galveston Bay 5

METHODS

Water column-associated metals

Water samples for trace metal analysis were collected in summer and fall Trinity River
1989, along a transect from the Trinity River near the town of Anahuae, Houston |
through Trinity and Galveston Bays, and out to the Gulf of Mexico |
(Fig. 1). Samples were not collected in the Houston Ship Channel/San N
Jacinto arm of the Bay, but water from this source would have been |
included as an admixture, especially in higher salinity samples. The data |
do not reflect a truly synoptic picture of the Bay system for eigher collec- Trinity Bay
tion for two reasons. First, the size of the Bay made it impossible to
collect across the entire salinity gradient in a single day. Second for the
summer collection, additional samples were collected during a trial run
two weeks before the other samples.
Salinity was monitored using a refractometer, and samples were col- Galveston Bay
lected in salinity intervals of approximately 5%. The refractometer had a
precision of about + or - 1 and reported salinities were measured in the labor- Clear Lake
atory by argentometric litraton (Amer. Public Health Assoc., 1985).
Salinity sometimes changes rapidly and unpredictably over short dis- East Bay
tances, at times decreasing locally in the seaward direction. Likewise, tur-
bidity is very patchy. To help verify that samples collected at a given
station were drawn from a discrete water mass, salinity was checked peri-
odically during sample collection.
For measurements of trace metals in water column samples, ultraclean
techniques were used during all stages of sample collection, transport, O
handling, processing, and analysis (Patterson & Settle, 1976). A separate C
sample was collected by peristaltic pumping through a 0.4 um Nuclepore I
filter for measurement of salinity, suspended particulate matter, alkalin- X
ity, and DOC. E
Immediately after return to shore, water samles were preserved by M
acidification with 2 ml ultrapure HNO, per liter of seawater within a of
portable laminar-flow clean bench. Filters were unloaded from their Galveston
Teflon assemblies and transferred to acid-cleaned 15 ml screw-cap vials. F
Filtered water samples were digested in their original bottles using the L
preservation acid and ultrasonification for 60 min at 60C. Filters were U
digested in their original vials using 10 ml of 0.5% HNO, and the same West Bay G
heand sonification.
Extraction columns consisted of a 2 cm bed of silica-immobilized Fig. 1. Locations of sampling sites in Galveston Bay. Solid circles are water column
8-hydroxyquinoline (Sturgeon et al., 1981: Marshall & Mottola, 1983) sites. Numbers are sediment sites. Leters are oyster sites. A=Galveston Bay Yacht Club
contained in a 1.5 cm diameter glass chromatography column connected (GBYC): B = Todd's dump(GBTD): C = Hanna Reef (GBHR: D = Confederate Reef
to a 500 cm diameter glass reservois. The column was precleaned by passage of (GBCR): E=Ship Channel (GBSC).

6 J.W. Morse, R.J. Presley, R.J. Taylor, G. Benoit, P. Santschi Twice metal chemistry of Galveston Bay 7

approximately 300 ml of a solution 1 M in HCI and 0-1 M in HNO, Alkalinity and salinity were measured by titrtion, DOC by the wet
(Seastar Brand). The concentration of zinc in the effluent was monitored digestion method (persulfate phosphoric acid), and suspended particulate
until it nearly matched the added acid. A sample of the final acid rinse matter (SPM) by gravimetry. Wet digestion may not quantitatively
was collected and measured for a column blank for all metals, though measure dissolved organic carbon (DOC) in open ocean waters
this was usually negligible. The column was rinsed with 10 ml of 18 (Sugimura & Suzuki, 1988), but we believe it gives reliable results
Mohm water to remove traces of acid. Immediately before preconcentra- for qualitatively different DOC found in estuaries. Virtuall all SPM
tion, aliquots were decanted in a clean bench and litrated to determine analyses were measured in duplicate. Here, we only report selected
the amount of ultrapure NII4OH required to adjust the pH to 8-0 + or - 0.5. results as they relate to dediment and oyster data. A full account of these
Approximately 200 ml of seawater sample (the entire 10 ml of digestion data will be given elsewhere (Benoit et al., 1992).
solution was used for filters) was passed through the column, which was
then rinsed with 10 ml ultrapure water to remove sea salts. Metals were Oyster-associated metals
then cluted with 15 ml of 1 M HC/10-1 M HNO and collected in a 15 ml
acid-cleaned plastic vial tht was checked for contamination by rinsing Oysters (Crassostrea virginica) were collected at six different sites in
with the same acid and testing for Zn. Galveston Bay during 1986-90 as part of the National Status and Trends
Lead, Cu, and Ag were measured in duplicate in acid-washed Teflon Program (GERG, 1990). Each site was on an identifiable oyster reef
autosampler cups in the graphite furnace of a Perkin Elmer Zeeman 5100 (Fig. 1) and at each, twenty oysters were taken from each of three
atomic absorption spectrophotometer equipped with Zeeman back- stations, the stations being 100-500 m apart. Each site was sampled once
ground correction, pyrolized furnace tubes and L'vov platforms. Injec- each year, except two of the sites (GBOB and GBSC) were not sampled
tion volumes were 40 pl for Cu and 150 pl (3 x 50 pl) for Ag and Pb. during the first two years. The twenty oysters from each station were
Zine was measured by manual injection of 10 pl since the autosampler combined and analyzed as a single sample each year.
introduced unacceptably high blanks of unknown source. Manual injec- Oysters were usually handpicked from exposed reefs, but in deeper
tions were repeated until reproducible absorbance readings were obtained water they were taken by dredge or tongs. In most cases stations were
and verified (typically 3 to 5 injections.) located hundreds of metres away from obvious point sources of contami-
Column yields were monitered by parallel measurement of certified nant imputs such as industrial discharges in an attempt to characterize
reference seawaters (CASS-2, S=29.2%, or SLEW-1,S=11.6". Research large areas of Galveston Bay, rather than to identify specific point
Council of Canada) for one out of every five columns. Silver column sources. The new sites added in year three were, however, selected to be
yields were determined on spiked seawaters, since the seawater reference closer to industrial areas or population centers than were the original
materials are not certified for silver. Concentrations were calculated by four sites. Stations were reoccupied as closely as possible each year, both
correcting for metals in the elution acid, the column blank, and column in time and space.
yield (typically near 90%). Bottle blanks, determined on distilled water Frozen oysters were returned to the laboratory where they were
collected in the field using the same filtration system, were negligible. brushed to remove mud from the shells and allowed to thaw. They were
Column yields had a standard deviation close to 20% and introduced then opened under clean room conditions using a stainless steel oyster
most of the uncertainty in the final calculated concentrations. Duplicate knife. The tissue was washed sparingly with distilled-deionized water to
samples (20% of collected samples) and replicate laboratory analyses remove any adhering mud and was put into a 500 ml Teflon jar. When
(10% of measurements) agreed with in the calculated uncertainties. the jar was approximately one-third full, or when all twenty oysters from
We believe that our measurements are reliable for several reasons: (1) a given station had been added, three solid Teflon balls of 3.5 cm dia-
state-of-the-aret clean techniques were used throughout: (2) the method meter were added. The jars were tightly closed and were put into plastic
was checked frequently against certified reference seawaters: (3) the con- Ziplock bags. The jars were then loaded into an industrial paint shaker
centration ranges were measured are similar to values found by careful investi- and were shaken vigorously for 15-20 min to completely homogenize the
gators working on other estuaries; and (4) the trends in the data are samples. An aliquot of the combined and homogenized sample was
geochemically reasonable, i.e. they correlate in a simple manner with freeze-dried, re-homogenized by ball milling in plastic, and weighed into
anneillary key physicochemical variables. a digestion vessel.

8 J.W. Morse, R.J. Presley, R.J Taylor, G. Benoit, P. Santschi Trace metal chemistry of Galveston Bay
The digestion vessels were 60 ml capacity screw top "bombs of PF banks are widespread. Because of the great heterogeneity of this environ- Teflon (Savillex Corp., Minnetonka, MN model 561). Digestion of
approximately 200 mg dry weight samples of oyster tissue used 3 ml of ment, anything short of a monumental sampling effort will not yield a
detailed picture of distributions in the Bay. Whether such an effort is
4 to 1 mixture of ultra-pure nitric and perchloric acids. The samples we justifiable is open to question because major storms, including hurri-
first allowed to pre-digest for 2 3 in the mixture on a warm hotpla canes, which are common in this area, cause substantial redistribution of
while the bombs were covered with Teflon watch covers. The bowl the sediments. With this in mind, sites were chosen throughout Galve-

were then tightly closed to a constant torque (2.5 kg-m) with matchin ston Bay that were deemed reasonable representative, Sites associated
screw caps and were placed in an oven at 130'C for 8 h. After remov. with dredged areas and that have a potential for abnormal metal concen-
from the oven and cooling, 20 ml of distilled- deionized water with added trations were given special emphasis. Sampling locations are shown in
The bombs were weighed and from the known empty bomb weight an Fig. 1.
final solution density (1-04 g ml') an exact final solution Volume W; Samples were collected from the RV Roman Empire either by hand
calculated. This and the sample weight were used to calculate a dilutio coring with a precleaned plastic core tube or with an epoxy-coated grab
factor for each sample. sampler in deeper waters. Only the top -10 cm of sediment was used
Two blanks and two refrences materials were digested with every s and it was homogenized upon collection in sealed bags from which air
of 20 40 samples. Reference material used included National Institu was excluded to prevent oxidation. This sampling depth was used
of Standards and Technology (NIST, formerly NBS) 1566 oyst because it represents 'near surface' sediments and the sandy nature of
tissue. National Institute for Enviromental Studies, Japan (NIES) muss, many of the sediments often precluded sampling much deeper. The
tissue, EPA trace metals in fish standard. National Research Counc homogenized bag sample was immediately divided between two pre-
of Canada (NRCC) DOLT-1 dogfish liver tissue and it Texas A&M cleaned plastic bottles which were filled to the top to exclude air, capped
University (TAMU) house stanard oyster tissue. Repealed analysis and scaled with electrical tape. One bottle was quick-frozen on dry ice
these reference materials and participation in several intercalibrtio for sulfide and metal analyses.
exercises organized by Dr Shier Berman of the NRCC give an estimat A portion of sediment was weighed and freeze-dryed to determine water
of 10% or better for both the precision and accuracy of the data report content.Porosity was calculated by using 2.5 as the solid density and correcting for the weight loss of pore water. Sediment grain-size distributions - here. All data reported here were obtained by atomic absorption spectr were determined by wet sieving following the method of Folk (1968).
metry (AAS). Flame AAS was used for Cu. Fe. and Zn which exhib Organic carbon and carbonate carbon were determined with a LECO
high concentrations in oysters, cold vapor AAS for mercury and graphit Carbon analyzer (CR12) in which a finely ground sample is combusted at
furnace AAS (Perkin-Elmer Corp. model 3030 with Zeeman backgroun 1350"C. A portion of the sediment was acidified with 2 ml of 6 NHCI to
correction) for tile remaining elements. Some samples of freeze-drie remove carbonate carbon and dried on a hotplate for 12 at -70'C.
oyster tissue were also analyzed for some elements by neutron activatio This sample was then combusted in the LECO carbon analyzer to yield
analysis (NAA) which required no sample pre- or post-treatment. Agre organic carbon content.The carbonate carbon was then determined from
ment between AAS and NAA was good (+10%) for elements analyzed to the difference between the total and organic carbon. The precision was
both techniques. 1% (1 SD). A LECO carbon standard of 0.98 wt% C was used for
The samples were analyzed for Ag. As. Cd, Cr, Cu. Fe, Hg, Mn. N standardization.
Pb, Se, Si, Sn, and Zn. In addition, temperature. salinity and related en Sedimentary iron sulfide minerals were operationally defined as acid
vironmental parameters were measured as were size, sex, parasite pres volatile sulfide (AVS) mid Pyrite.AVS was extracted from sediments
ence and indicators of health of the oysters (see GERG, 1990). using a cold HCI (6 N)+ SnCl2 (2%) method (Cornwell & Morse, 1987).

Pyrite was determined from the difference between AVS and the total
Sediment-associated metals reduced sulfur (TRS) concentrations. TRS was extracted using a boiling
Cr(11)+ acid method (Zhabina & Volkov, 1978, as modified by Canfield
The sediments of Galveston Bay are highly Variable, with major viria et al., 1987). The Cr(11) + acid method has been demonstrated to
tions in grain size occurring on a small scale. Also. oyster reefs and spoi extract Pyrite sulfur effectively and specifically after removal of

10 J.W. Morse, R.J. Presley, R.J. Taylor, G. Benoit, P. Santschi trace metal chemistry of Galveston Bay 11

metastable iron sulfide minerals by AVS extraction (Canfield et al. 1987). Also, this method has been demonstrated to extract none of the
organic sulfur from sulfur-containing compounds or natural organic matter, thus, eliminating the interference from organic sulfur(see Can-
field et al., 1987). The concentration of hydrogen sulfide evolved by these methods was determined by potentiometric titration with Pb(CIO4)2.
The detection limit of potentiometric titration was- 1 amol g 1, and the precision was 5% (1 SD).
Metels were extracted from the sediments using teaching techniques. Freeze-dried portions of the frozen samples were subjected to a series of
teaching procedures in order to separate sedimentary pyrite from the bulk sediment. Briefly, the sequential extraction procedure involves diges-
tion of the sediment sample with 1 m HCI (reactive fraction). 10 m HF (silicate fraction) and concentrated HNO1 (silicate fraction) and concentrated
HNO1, (pyrite fraction). A more complete explanation of the sequential extraction procedures and the development of the separation method is given
in Huerta-Diaz & MOrse (1990). Additionally, frozen samples were extracted with the 1 m HCI results.
Trace metals (as,Cd,Cr,Cu,Fe,Hg,Mn,Mo,Ni,Pb and Zn) were determined by flame atomic absorption(FAA) using a Perkin-Elmer model 2380 spectrophotometer.
Metals below the detection limit of this instrument were analyzed by direct injection into a Hitachi model 170-70 graphite furnace atomic absorption
(GFAA) spectrophotometer with Zeeman background correction. The analytical precision (relative standard deviation) was normally between 5 and 10% for
flame atomic absorption analyses and between 10 and 15% for graphite furnace atomic absorption analyses. Salt matrix effects for As determination by GFAA
were partially overcome by using a Ni Pd ascorbic acid matrix modifier (Robert Taylor, pers. comm.) Determination of Pb by GFAA in samples containing high
Fe/Pb ratios (>250) was carried out following the procedure developed by Shao & Winefordner (1989). Mercury was determined with a Laboratory Data Control UV
monitor equipped with a 30 cm path length cell, using the well-known cold vapor technique. For samples suspected of having high dissolved organic matter (DOM)
concentrations. Hg was measured after destruction of the DOM with bromine monochloride. Detection limits were calculated as 2.5 times the standard deviation
of the reagent blank (e.g.Kaiser, 1970; Bruland et al., 1979; Kremling, 1983). All reagents used were ACS reagent grade or better. Milli-Q water was used for
the preparation of all aqueous solutions. Acidic working standard solutions were always freshly prepared. All materials were carefully cleaned using established
acid teaching procedures. The use of the sequential extraction procedures precluded comparisons with standard reference materials for which only total metal con-
centrations are generally available.

RESULTS

Because of the large number (1000s) of analyses done as part of this study, original analytical data are generally not given in this paper. These data are available
from the authors upon request.

Trace metals in the water column

Samples were collected over salinities ranging from 0.1% to 27.4%, and particulate matter levels from 2.4 to 48.4 mg liter 1(Fig.2). Particulate matter did not show
the same trend with salinity on the two sampling dates. In August, there was a mid-salinity SPM maximum, while in October, there was an SPM minimum at intermediate
salinities. In fact, the two curves are almost inverses of each other. The variations over time at a given location probably reflect different levels of wind-driven
mixing and turbulence, and subsequent sediment entrainment in the water column of this shallow estuary.
DOC levels in the fresh water samples were 5.5+1.4 and 5.3+0.2 mg C liter', while in the Gulf and member they were 0.1 + 0.3 and 0.29+

SALINITY (%0)
Fig. 2. Suspended particulate matter concentration as a function of saliity. SPM probably reflects local wind stress induced resuspension of bottom sediments.

12 J.W. Morse, R.J. Presley, R.J. Taylor, G. Benoit, P. Santschi Trace metal chemistry of Galveston Bay 13

.14 mg litre'. DOC showed very consistent behaviour on the two dates, decreasing non- conservatively (curve concave-upwards) with salinity in
both cases. Therefore, a sink for DOC in the intermediate salinity range is indicated for Galveston Bay. This is consistent with previously
observed removal of DOC via flocculation with increasing ionic strength in other estuaries (Boyle et al., 1977). Alkalinity was nearly constant at
about 2 meq liter' at all salinities in the estuary.
Results of trace metal measurements are summarized in Table 1 and illustrated in Fig.3 as a function of SPM. In general, trace metal concen-
trations were 5 to 10 times higher than open ocean values (Boyle & Huested, 1983; Bruland & Frank, 1983; Martin et al., 1983. Schaule & Patterson, 1983),
and are similar to measurements from other estuaries, conducted by careful analysts (c.g. Bewers & Yeats, 1978; Windom et al., 1983, 1985; Keeney-
Kennicutt & Presley, 1985; Mart el al., 1985; Shiller & Boyle, 1985; Valenta et al., 1986). Dissolved metals often did not show systematic variations with
salinity, white particulate metal concentrations in the water column (pg liter' showed trends that were broadly similar to those of suspended particulate
matter (mg liter')
Zinc, Pb, and Cu each had similar concentration ranges on the two dates while dissolved Ag levels were significantly lower at the later date. In August
dissolved Ag averaged 5.6 + 2.2 mg liter', while in October it was 1.3 + 0.8 ng liter'. Particulate Ag dropped from 3.6 + 1.8 to 2.2 + 0.8 ng liter', but
this is not a statistically significant change (Bevington, 1969). the change in dissolved Ag could result from high freshwater inputs in spring and summer
with Ag dilution
in freshwater sources, or from a decrease in the organically complexed form of the metal. If the

TABLE 1
Summary of Trace Metal Data in the Water Column of Galveston Bay
Zn Pb Cu Ag

Diss Part Diss Part Diss Part Diss Part
Maximum 4.50 2.56 0.133 0.530 1.41 0.39 8.9 5.9
Minimum 0.30 0.44 0.022 0.026 0.13 0.03 0.2 0.7
Average 1.68 1.04 0.071 0.209 0.86 0.15 3.2 2.8
SD 1.14 0.66 0.029 0.150 0.33 0.10 2.7 1.5
n 17 20 19 19 21 19 20 15

A single particulate sample gave the follwing values in pg liter'. Zn 6.85, Pb 0.36,
Co 0.80, and Ag 21 ng liter'.This sample was assumed contaminated and not included in the compilation.
Diss-dissolved, Part= particulate, concentrations are in pg liter; except Ag which is in ng liter'.

Fig.3. Metal concentrations in suspended particulate matter as a function of SPM concentrations in the water column. Note that metals have similar
concentrations in SPM and in average surficial sediments (dotted lines). Metal concentrations are higher at low SPM levels probably because more line-
grained sediments are suspended at such times.
lattef were true, it seems likely that Cu would change in s similar manner. Systematic contamination with Ag alone seems improbable. Further study of
the seasonal concentration of dissolved Ag could resolve this question.

Trace metals in oysters

Oysters and other bivalves have been used as 'sentinel' organisms for assessing the contamination of coastal marine water bodies for almost

14 J. W. Morse, R. J. Presley, R. J. Taylor, G. Benoit. P. Santschi

15 Table 2

twenty years. For example, Goldberg et at. (1983) report data for a Summary Statistics for Trace Metal Concentrations in Galveston Bay and US Gulf of Mexico Oysters Collected for the NOAA NS&T
USEPA-funded Mussel Watch program conducted in 1976 78 and the Program in 1986-90 and for the EPA Mussel Watch in 1976-78. All Values in ppm dry weight
current NOAA funded National Status and Trends (NS&T) program
(NOAA, 1985, 1989b) is an outgrowth and extension of the Mussel Ag As Cd Cr Cu Fe Hg Mn Ni Pb Se Sn Zn
Watch concept. Bivalves are widely recognized as being responsive to Galveston Bay
changes in contaminant levels in the envoroment, good accumulators of 1986-90 2.77 4.50 4.33 0.53 165 275 0.078 15.9 1.89 0.71 3.42 0.29 3263
metals, widely distributed along coasts and easy to collect and analyze. SD 2.42 1.08 1.58 0.44 61 142 0.066 8.1 0.64 0.45 0.88 0.22 1648
They integrate contaminant levels in the enviroment over weeks to US Gulf of Mexico
months and therefore allow areas to be compared even when sampling is 1986-90 2.24 9.69 4.20 0.55 156 320 0.142 14.8 1.64 0.69 2.99 0.23 2417
limited to a frequency of once or twice per year. SD 1.59 7.00 2.46 0.42 107 243 0.156 8.8 1.29 0.93 1.33 0.16 1753
Trace metal concentrations found inoysters collected along the Gulf GB GOM 1.23 0.46 1.03 0.96 1.06 0.86 0.55 1.08 1.16 1.03 1.14 1.24 135
of Mexico coastline during the first five years of NS&T were generally Significance of t-test
similar to those reported in oysters taken from non-contaminated water of means <0.01 <0.01 0.64 0.63 0.44 0.11 <0.01 0.29 0.08 0.86 <0.01 <0.01 <0.01
in other parts of the world (Presley et al., 1990). Only a few sites showed US Gulf of Mexico
obvious trace metal contamination and these were restricted geographi- 1976-78b 1.8 ---- 4.6 --- 162 ---- ---- ---- 2.7 0.9 ---- ---- 1940
cally such that nearby sites were usually unaffected. Abnormally high or SD 1.5 ---- 2.6 ---- 138 ---- ---- ---- 1.4 0.9 ---- ---- 1480 low values at a site did, however, usually repeat year after year suggest-
ing local influences. Sites giving higher than average values for most
metals were just likely to be far from populatin or industrial centres
as to be near such areas. Crassotrea virginia: for Galveston Bay, n= 78 pooled samples of 20 oysters each; for GOM, n=874 pooled samples of 20 oysters each.
The oysters collected in Galveston Bay for NS&T were similar in trace Crassostrea virginia: mean +1SD;EPA Gulf Mussel Watch: Goldbert et al. (1983).
metal content to those collected elsewhere along the Gulf coastline, i.e.
the oysters give no indication of generalized trace metal contamination in
Galveston Bay. This can be seen by comparing the overall averages
(Table 2) for all years and all sites in Galveston Bay with those for the
entire Gulf. The Galveston data include five sites sampled all four years
and two sites sampled for two years. Three stations were sampled at each
site resulting in almost 1500 Galveston oysters being analyzed over the
five year period. The Gulf data set includes more than 18000 oyster, as
fifty sites were sampled all five years and an additional twenty sites for
three years. Thus, the averaged data are unlikely to be biased by a few
abnormal individuals.
The average Cd, Cr, Cu, Mn, and Pb in NS&T oysters from Galveston
Bay differs by 10% or less from the Gulf-wide average. This difference is
about the same as our analytical precision and is not significant at the
99% confidence level based on a 't-test'. Silver is 23% higher in Galveston
Bay. NI is 16% higher and Se is 14% higher. A 't-test' of the significance of
those figures shows that the Ni averages are not significantly different at the
95% confidence level. Although, the Se average for Galveston Bay oysters
is significantly higher than the Gulf-wide average, it is not significantly
different from the Texas Louisiana average. It differs from the Gulf-wide
average only because of low Se oysters in MIssissippi and Florida.

CIO

16 J.W. Morse, R.J. Presley, R.J. Taylor, G. Benoit, P. Santschi Trace metal chemistry of Galveston Bay 17

The high average Ag in Galveston Bay oyster is caused by a 300% TABLE 3
enrichment found at a site on Confederate Reel in 1990. The enrichment Sediment characteristics C=clay; S=sand; s=silt; C=clay.
appears to be real because it was in all three of the twenty oyster pooled ________________________________________________________________________
samples collected at that site. No cause for the enrichment can be sug- Sample Sand Silt Clay <62pm Class Porosity Organic-C CaCO3
gested, and if these samples are neglected NS&T Galveston Bay oysters (wt%) (wt%) (wt%) (wt%) (%) (wt%) (wt%)
would have average silver content. _________________________________________________________________________
Arsenica and mercury in Galveston Bay oysters are less than one-half GB1-1 25 32 43 75 SsC 79 1-16 0-33
the Gulf-wide average but the Gulf-wide averages are greatly influenced GB1-2 66 17 17 34 S 64 0-41 0-76
by several sites in southern Florida that produce oysters greatly enriched GB1-3 65 11 24 35 S 62 0-35 1-37
in arsenic and mercury and by a site in Lavaca Bay, Texas which is en- GB1-4 43 48 19 57 SsC 63 0-29 0-85
riched in mercury. Oysters from other Texas and Louisiana bays are sim- GB1-5 45 36 19 55 Ss 51 0-24 4-05
ilar in As and Hg content to those in Galveston Bay. GB1-6 44 40 16 56 SsC 61 0-30 1-95
Tin seems to be about 24% higher than Gulf averages in Galveston GB1-7 78 14 8 22 S 51 0-18 2-68
Bay, but Sn values are near the detection limit of the method used and a GB1-8 75 17 8 24 S 52 0-19 1-08
24% difference may not be meaningful. Finally, Zn is 35% higher in GB1-9 36 36 28 65 SsC 60 0-43 3-70
Galveston Bay oysters collected for NS&T than in Gulf-wide average GB1-10 65 30 5 35 S 54 0-22 2-01
oysters. This difference is highly significant as the high Zu found in all oys- GB1-11 93 5 2 7 S 45 0-07 0-70
ters leads to precise analytical determination and few analytical aritifacts. GB1-12 93 4 3 6 S 26 -- --
Trace metals in sediments GB1-13 34 47 19 66 SsC 74 0-52 24-38
GB1-14 34 29 37 66 SsC 69 0-44 1-91
Sediment characteristics vary widely at the different study sites (Table 3). GB1-15 87 9 4 13 S 54 -- --
This reflects the highly heterogencous environment in Galveston Bay and GB1-16 93 4 3 8 S 49 0-15 11-86
the impact of man's activities, such as dredging and trawling for shrimp. GB1-17 87 7 6 12 S 54 0-18 3-95
The sediments are generally poorly sorted mixtures of sand, silt and clay, GB1-18 98 2 0 2 S 47 0-1 4-25
and exhibit major variations in mean grain size. Porosity averages 62 GB1-19 95 2 3 5 S 47 0-05 0-60
(+10)%. GB1-20 36 36 28 64 SsC 51 0-11 7-87
The organic carbon content of these sediments is highly variabe aver- GB1-21 57 12 32 43 CS 53 0-11 13-46
aging 0-42 wt% and ranging from 0-05 to 1-18 wt%. It is positively corre- GB1-22 63 22 15 37 CS 50 0-14 2-46
lated (r2=6-69) with the <62 on suze fractub if tge sediments. The GB1-23 87 5 8 13 S 48 0-10 6-94
CaCO, content of the sediments is also highly variable averaging 3-9 wt% GB1-24 75 8 17 25 S 57 0-22 1-37
and ranging from 0-33 to 24-4 wt%. It is not correlated with the <62 pm GB1-25 40 20 31 51 SsC 67 0-41 1-26
size fraction (r2-0-00) and is dominantly present as shell fragments. GB1-26 53 20 27 47 SsC 58 0-34 2-92
Average trace metal concentrations are given in Table 4. Also included GB1-27 60 13 27 40 CS 60 0-35 17-48
in this table are the range of trace metal concentrations. trace metal con- GB1-28 59 14 27 41 CS 58 0-36 0-95
centrations normalized to the <62 pm grain size fraction, and the percent GB1-29 20 37 43 79 SsC 73 0-67 3-16
of the metal in the total reactive fraction (here defined as 1 N HCI GB1-30 88 10 2 11 S 54 0-18 1-77
extractable + pyrite-associated metal). The average concentrations are GB1-31 56 16 28 44 SsC 65 0-46 2-07
similar to values observed in other estuaries (e.g. Naragansett Bay, GB1-32 8 22 70 92 sC 73 1-11 2-14
Santschi et al., 1984; Baffin Basy, Huerta-Diaz & Morse, 1992) especially GB1-33 27 30 43 272 SsC 76 0-77 2-58
when grain-size normalized values are used. This is further elaborated in GB1-34 29 25 46 71 SsC 76 0-75 2-13
the Discussion section. GB1-35 42 18 40 58 SsC 69 0-61 1-42
GB1-36 5 27 68 95 sC 78 1-18 2-55
GB1-37 36 25 39 64 SsC 80 0-79 6-67
GB1-38 65 22 13 35 sS 57 0-36 1-38
GB1-39 51 27 22 49 SsC 70 0-62 2-48
GB1-40 84 8 8 16 S 69 0-26 1-89
GB1-41 9 28 63 90 sC 77 1-07 2-64
_________________________________________________________________________
These are combined for other classes of sediment (e.g. Cs=Claycy silt) according to
Folk (1968).

18 J.W. Morse, R.J. Presley, R.J. Taylor G Benoit, P. Sanschi Trace metal chemistry of Gulveston Bay 19

TABLE 4 TABLE 5
Summary of Metal Concentrations in Galveston Bay Sediments Data on Reduced Sulphur and Extent of Pyritization
__________________________________________________________________ ______________________________________________________________________________
Metal Average Range in Average % Reactive Sample TRS AVS %AVS Pyrite DOP-FCD DOP__1 N
concentration concentration concentration (pmol g') (pmol g') (pmol g') (pmol g')
__________________________________________________________________ ______________________________________________________________________________
As 56 23.98 125 42 GB1-1 64.5 5-2 8-0 29-7 0-43 0-55
Fe 15 900 1 570 40 200 35 200 30 GB1-2 33-2 0-6 1-7 16-3 0-27 0-45
Ct 37 4 W2 82 37 GB1-3 42-4 0-6 20-0 16-9 0-23 0-35
Cu 8 2 15 18 51 GB1-4 41-1 4-6 11-1 18-3 0-28 0-43
Hg 0-08 0-01 0-28 0-19 98 GB1-5 45-6 2-1 4-6 21-8 0-41 0-52
Mn 605 165 2 365 1 320 42 GB1-6 71-1 1-4 1-9 34-9 0-44 0-68
Mo 41 25 79 95 29 GB1-7 22-7 1-4 6-1 10-6 0-25 0-53
Ni 26 4 45 58 28 GB1-8 28-9 4-5 15-7 12-2 0-29 0-52
Pb 25 12 46 59 99 GB1-9 83-3 3-8 4-5 39-8 0-44 0-61
Zn 55 6 116 123 44 GB1-10 45-7 2-7 6-0 21-5 0-55 0-73
_________________________________________________________________ GB1-11 12-3 0-3 2-5 6-0 0-32 0-76
Concentrations are total in pg g . Concentration* concentration normalized GB1-13 147-3 4-0 2-7 71-6 0-76 0-79
to the average fraction of sediment (0-45) in the less than 63 pm grain-size GB1-14 199-7 0-0 0-0 99-9 0-77 0-77
range. Reactive-metal fraction = 1 N HCI extractable-metal + pyrite-metal. GB1-16 24-8 2-1 8-4 11-4 0-35 0-50
GB1-17 18-4 1-9 10-5 8-3 0-29 0-35
All sediments were observed to be anoxic with the active sulphide min- GB1-18 7-9 0-4 4-7 3-8 0-16 0-24
eral formation occurring. Data on the concentrations of AVS and TRS. GB1-19 5-2 0-1 1-8 2-6 0-12 0-09
and th degree of pyritization (DOP) of iron using both citrate dithionite GB1-20 23-0 1-4 6-1 10-8 0-23 0-50
and 1 N HCI extration techniques to remove reactive-Fe are presented in GB1-21 17-2 2-3 13-6 7-4 0-10 0-24
Table 5. DOP is delined as (berner, 1970): GB1-22 41-4 7-1 17-2 17-2 0-34 0-38
Pyrite Fe GB1-23 7-0 0-6 8-7 3-2 0-17 0-28
DOP= GB1-24 8-5 0-3 4-0 4-1 0-27 0-21
Pyrite Fe + Reactive Fe GB1-25 28-8 0-4 1-5 14-2 0-21 0-27
where pyrite-Fe is assumed equal to 0-5 (TRS AVS), based on the 1:2 GB1-26 64-6 4-8 7-4 29-9 0-47 0-62
stoichiometry of FE:S in pyrite. The AVS concentration averages 3-6 GB1-27 21-2 1-2 5-8 10-0 0-14 0-17
and ranges from 0-1 to 16-3 TRS concentrations GB1-28 43-3 0-4 0-9 21-5 0-28 0-28
average 58-6 and range from 5-2 to 314-2 AVS GB1-29 49-7 9-8 19-8 20-0 0-17 0-16
generally represents a small fraction of TRS averaging 7-6% and never GB1-30 23-6 0-8 3-3 11-4 0-27 0-12
exceeding 24 of TRS. Neither TRS nor the fractin of TRS as AVS are GB1-31 88-7 1-9 2-2 43-4 0-66 0-48
well correlated with the <62 pm fraction of the sediment (r2=0-25 and GB1-32 58-9 9-8 16-7 24-5 0-42 0-28
0-08, respectively)> GB1-33 95-8 5-8 6-1 45-0 0-62 0-52
The DOP values determined by the citrate Jithionite and 1 N HCI GB1-34 175-2 3-9 2-2 85-6 0-50 0-41
methods are in good general agreement averaging, respectively. 0-35 and GB1-35 65-5 10-4 15-8 27-9 0-48 0-41
0-42. This is consistent with the findings of previous studies (e.g. Huerta- GB1-36 56-9 7-7 13-5 24-6 0-29 0-22
Diaz & Morse. 1990). Because of difficulties in applying the citrate GB1-37 314-2 4-0 1-3 155-1 0-62 0-58
dithionite method when determining some trace metals of interest (e.g. GB1-38 38-8 4-0 10-3 17-4 0-61 0-54
Zn), results using the 1 N HCI extraction method for charactyerizing the GB1-39 62-6 1-6 2-6 30-5 0-37 0-32
non-pyritized reactive fraction were used. The DOP (1 n HCI) of GB1-40 38-5 0-9 2-4 18-8 0-22 0-18
GB1-41 68-8 16-3 23-7 26-2 0-30 0-24

Average 58-6 3-6 7-6 27-5 0-35 0-42
SD 60-1 3-6 6-2 29-8 0-18 0-19

(Column 1)

20 J.W. Morse, R.J. Presley, R.J. Taylor, G. Benoit, P. Santschi

sediments is highly variable, ranging from 0.09 to 0.79. The upper range
of values is indective that Fe may be approaching being the limiting
factor for pyrite formation in some of the sediments.
In areas where sediment grain size varies substantially, such as Galve-
ston Bay, total metal concentrations (Table 4) by themselves often are
not particularly informative. This is because the metals are often domi-
nantly associated with the fine-grained material and, consequently, varia-
tions in metal concentrations given in Table 4 may largely reflect grain
size differences (e.g. Trefry & Presley, 1976). The correlation coefficients
(f) for the total reactive fraction of metals studied in Galveston Bay
with the <62 pm grain size fractions are: Fe = 0.08. Mn = 0.52, Ni =
0.66, Cu = 0.57, Zn = 0.73, Cd = 0.16, Pb = 0.47, Cr = 0.69, Mo = 0.18,
As = 0.00, and Hg = 0.08. These relationships indicate that most metals,
with the exceptions of Cd, Mo, As, and Hg are dominantly associated with
the fine-grained fraction of the sediment. The metals not well associated
with this fraction may be incorporated into the sediment by processes
other than simple sedimentation (e.g. diffusion of dissolved arsenate and
molybdate into the sediment followed by reduction and precipitation) or
be associated with fractions such as large carbonate particles.
It is common in the study of trace metal geochemistry to attempt to
'normalize' concentrations by ratioing the trace metals to some other
more abundant chemical components such as Fe or Al. Because Al was
not determined in this study, the correlations (r2) of total trace metal
concentrations with total Fe concentrations were calculated. The results
are: Mn = 0.60, Ni = 0.79, Cu = 0.56, Zn = 0.88, Cd = 0.42, Pb = 0.46,
Cr = 0.90, Mo = 0.42, As = 0.01, and Hg = 0.06. Mn, Ni, Zn, Cd, Cr,
Mo all exhibit significantly better correlations with Fe than with the fine-
grained fraction, whereas Cu and Pb exhibit about the same correlation
by both approaches. Arsenic and Hg were the only metals exhibiting a
poor correlation with both grain size and Fe.
The primary concern in this study is not with total metal concentrations,
but rather with the total reactive fraction which is operationally defined here
as reactive metal (l M HCI extractable) + pyrite-metal. The citrate dithionite
extractable fraction was not used since the reagent is substantially con-
taminated with several of the metals of interest and in some cases
this fraction is not well analyzed by graphite furnace atomic absorption
spectrometry (see Huerta-Diaz & Morse, 1990, for discussion). The con-
centration ratios of the total reactive fraction of trace metals relative to
total reactive-Fe (Me*) are presented in Table 6. Normalization to total
reactive-Fe concentrations was done in order to observe if anomalously
high total reactive trace metal concentrations are present in any of
the sediments. Such anomalies may be indicative of contamination.

(Column 2)

Trace metal chemistry of Galveston Bay

TABLE M

Ratios (Me*) of Total Reactive-Metal Concentration Total Reactive-Iron Concentrations
---------------------------------------------------------------------------------------------------------------------------------------

Sample Mn* Ni* Cn* Zn* Cd* Pb* Cr* Mo* As* Hg*

---------------------------------------------------------------------------------------------------------------------------------------

Galveston Bay
GBI-1 52.0 1.03 1.40 6.50 0.110 1.42 1.91 1.11 2.14 0.0024
GBI-2 45.3 0.80 1.63 7.17 0.222 3.74 3.73 3.16 12.48 0.0371
GBI-3 86.2 0.86 1.31 4.57 0.228 2.11 2.16 2.55 8.76 0.0019
GBI-4 60.6 0.87 1.59 6.75 0.238 2.88 2.82 3.96 16.08 0.0089
GBI-5 75.0 1.42 1.10 3.38 0.235 1.58 2.29 1.76 3.57 0.0023
GBI-6 42.6 1.13 1.62 5.36 0.381 2.49 2.62 2.05 7.97 0.0018
GBI-7 97.6 2.91 1.77 17.44 1.067 4.81 4.69 4.5 40.28 0.0026
GBI-8 56.1 2.27 1.82 7.12 0.577 3.96 7.25 4.24 7.91 0.0026
GBI-9 58.1 1.68 1.52 3.49 0.143 1.96 2.41 1.62 3.95 0.0418
GBI-10 47.1 2.13 2.39 5.76 0.287 4.04 6.24 2.57 7.40 0.0035
GBI-11 87.5 2.43 1.96 6.49 0.240 8.88 9.13 8.78 33.30 0.0061
GBI-13 47.4 2.73 1.09 3.15 0.382 1.24 1.27 0.96 1.24 0.0022
GBI-14 27.0 1.08 0.39 2.29 0.064 0.67 1.06 0.91 1.30 0.0009
GBI-16 182.1 18.28 8.64 18.64 3.330 8.36 5.87 6.21 26.11 0.0331
GBI-17 126.0 5.69 1.69 6.82 0.981 3.36 2.28 5.07 7.32 0.0032
GBI-18 126.2 6.40 1.45 6.70 0.868 5.69 4.56 8.08 145.34 0.0063
GBI-19 47.3 0.80 0.17 6.75 0.000 2.27 2.22 3.94 12.30 0.0018
GBI-20 226.0 3.26 4.42 3.72 0.461 2.37 1.38 2.21 5.15 0.0031
GBI-21 183.3 6.04 1.67 4.75 0.963 3.44 3.12 3.48 5.73 0.0067
GBI-22 69.7 1.17 0.66 4.11 0.039 1.28 1.32 1.48 3.12 0.0111
GBI-23 84.6 6.37 4.45 5.40 0.889 4.42 2.99 4.98 6.63 0.0062
GBI-24 57.9 1.01 0.93 5.32 0.105 1.75 1.96 2.01 4.11 0.0020
GBI-25 46.7 1.01 4.52 5.44 0.042 4.61 2.54 1.04 3.47 0.0068
GBI-26 45.9 1.47 3.91 7.33 0.131 1.53 5.65 1.18 7.20 0.0035
GBI-27 96.9 3.97 2.21 7.47 0.569 2.60 4.80 1.61 3.71 0.0069
GBI-28 31.5 0.89 1.00 5.38 0.059 1.08 2.20 0.63 1.36 0.0021
GBI-29 49.2 1.21 1.15 4.17 0.061 0.90 2.82 0.74 2.84 0.0017
GBI-30 45.0 1.02 0.33 3.85 0.38 1.01 1.89 1.17 4.32 0.0037
GBI-31 26.6 0.85 0.56 2.61 0.019 0.57 1.29 0.74 2.42 0.0023
GBI-32 39.1 0.91 0.83 3.07 0.028 0.74 1.57 0.52 2.12 0.0027
GBI-33 29.4 0.97 0.78 2.89 0.027 0.65 1.71 0.58 2.69 0.0056
GBI-34 31.5 0.96 0.96 3.05 0.000 0.80 1.88 0.92 2.23 0.0028
GBI-35 29.6 0.84 0.89 5.67 0.000 0.79 1.83 0.97 1.95 0.0106
GBI-36 96.9 1.02 1.07 3.64 0.072 0.84 2.04 0.78 1.65 0.0032
GBI-37 24.9 0.99 1.07 2.02 0.089 0.79 1.98 0.91 2.36 0.0023
GBI-38 24.8 1.05 5.97 6.05 0.094 1.83 2.05 3.78 7.97 0.0049
GBI-39 50.0 0.89 0.73 3.40 0.094 0.82 1.66 0.95 2.09 0.0022
GBI-40 55.0 1.02 0.52 3.85 0.065 1.16 2.01 1.12 3.67 0.0027
GBI-41 40.0 0.91 1.48 4.87 0.097 0.92 2.11 0.51 1.53 0.0020

Average 67.9 2.32 1.48 5.47 0.341 2.34 3.73 2.40 6.56 0.0065
46.3 3.09 1.05 3.05 0.580 1.98 1.81 2.07 6.73 0.0094

Baflin Bay
86 BB2 102.9 2.97 1.93 3.59 0.128 1.32 2.62 - - -
88 BB1 112.0 1.85 1.68 3.53 0.102 5.71 1.87 0.04 0.41 0.00244
88 BB2 94.1 1.61 1.64 3.36 0.120 1.83 1.24 0.21 0.83 0.00461

Average 102.7 2.14 1.75 3.48 0.12 2.85 1.93 0.13 0.63 0.00362

GB/BB 0.66 1.08 0.85 1.57 2.93 0.82 1.93 18.76 10.45 1.80
---------------------------------------------------------------------------------------------------------------------------------------
Values exceeding twice the average are in italics. Baflin Bay sediment data from Huerta-Diaz
& Morse (1992). GB/BB = Galveston Bay results divided by Baflin Bay results.
Ratios are molar x 1000.

(Column 1)

22 J.W. Morse, R.J. Presely, R.J. Taylor, G. Benoit, P. Santschi

TABLE 6B
Samples in Which the Value Exceeds Two Times the Standard Deviation of the Mean.
Numerical Values Represent (Sample Average)/Standard Deviation
------------------------------------------------------------------------------------------------------------------------------
Sample Mn* Ni* Cu* Zn* Cd* Pb* Cr* Mo* As* Hg*
------------------------------------------------------------------------------------------------------------------------------

GBI-2 -- -- -- -- -- -- -- -- -- 3.25
GBI-7 -- -- -- 3.93 -- -- -- -- -- --
GBI-9 -- -- -- -- -- -- -- -- -- 3.76
GBI-11 -- -- -- -- -- 3.30 2.98 3.08 3.97 --
GBI-16 2.47 5.17 2.06 3.34 5.15 3.04 1.18 -- 2.91 2.83
GBI-18 -- -- -- -- -- -- -- 2.74 -- --
GBI-20 3.41 -- -- -- -- -- -- -- --
GBI-21 2.49 -- -- -- -- -- -- -- -- --
GBI-26 -- -- 2.31 -- -- -- -- -- -- --
GBI-38 -- -- 4.27 -- -- -- -- -- -- --
------------------------------------------------------------------------------------------------------------------------------

However, if contaminant sources introduce large quantities of iron, this
procedure might not be appropriate as an indicator of excess metals.
Observed anomalics and the relationship of metal concentrations in
Galveston Bay to other sediments will be dealt with in the Discussion
section.
Huerta-Diaz & Morse (1990) introduced the concept of degree of trace
metal pyritization (DTMP) which is equivalent to DOP for iron. By
comparing DTMP with DOP it is possible to relate the pyritization of a
given trace metal to that of Fe, which is the dominant metal that is pyri-
tized. The metals can be arbitrarily divided into three major groups:
these generally exhibiting considerably less, about the same, and greater
pyritization than Fe. Results are shown graphically in Fig. 4. Mn, Zn, Ni,
and Pb generally fall into the first category, not being as extensively pyri-
tized as Fe. Cu, Cr, and Mo exhibit a large scatter, but most samples fall
into the second category, being pyritized about the same as Fe. Arsenic
and Hg are often much more extensively pyritized than Fe. The concen-
trations of Cd were often close to or below detection limits and such a
relationship is not reliable for this metal.

DISCUSSION

To our knowledge these are the first reliable, systematic measurements of
trace metal concentrations other than As and Sb in the water column of
Galveston Bay. Data bases of the Texas Water Commission (TWC),
Army Corps of Engineers, and US Geological Survey contain many

(Column 2)

Trace metal chemistry of Galveston Bay 23

(3 Graphs, ABC)

Fig. 4. The relationship between trace metal (DTMP) and iron (DOP) pyritization,
(A) Metals generally undergoing less pyritization than Fe (white square = Mn; black
square = Sn; white triangle = Pb; black triangle = Ni). (B) Metals generally having
a similar degree of pyritization to that of Fe (+ = Cu; x = Mo; black diamond
= Cr). (C) Metals generally undergoing greater pyritization than Fe (white circle =
As; black circle = Hg). Dashed lined is for 1 to 1 ratio of DTMP to DOP.

24 J.W. Morse. R.J. Presley, R.J. Taylor, G. Benoit, P. Santsclu
Trace metal chemistry of Galveston Bay 25

years of trace metal data from surface water, but these measturemtns In Fig. 3 the concentrations of metal on suspended particles vary over
are invalid due to sample contamination and/or insufficient analytical a range that brackets the average values in surface sediments. This is ex-
sensitivity. Our values are typically 100 to 1000 times lover than the ear- actly what would be expected if metals on the suspended particles were
lier analyses. The difference most probably reflects our more careful mea- derived directly from resuspended bottom sediments. The range in con-
surements rather than any real change in metal concentrations over time. centrations in the water column can be explained by the range in concen-
Galveston Bay is very shallow, so SPM levels, and their trace metal tration on bottom sediments, but this is probably modulated by the
burdens, should be dominated by resuspension and settling of bottom preference of metal for fine sediments and differntial resuspension and
sediments. Because of this close coupling between the water and sediment settling of different sediment size classes. In particular, under conditions
colums, it seems likely that particulate metals in the water column will of low wind stress and turbulence, SPM concentrations are lower and are
reflect levels in surficial bottom sediments. Supporting this expectation. probably composed largely of finer sediments, which are enriched in
particulate metal concentrations broadly mirror SPM levels, with Zn and trace metals. COnversely, stronger winds resuspend a greater number of
Pb showing mid-salinity maxima in August, and corresponding minima larger particles, which are comparatively depleted in metals. Figure 3
in Octor. Figure 5 shows the good corrclation between particulate Zn shows that trace metals concentrations on suspended particles do, in fact,
and Pb concentrations with SPM concentrations for October samples. decrease with increasing SPM concentration for all four metals. This is
The linear corrclation is remarkable. since it implies a nearly constant consistent with previous observations by DUinker (1983) who associated
concentration of these two metals in bottom sediments across the entire this inverse correlation to the fact that high levels of SPM include a sub-
estuary. The slopes of the two lines give average concentrations (Zn 4I stantial fraction of heavier, larger-sized particles and aggregates, origi-
pg kg . Pb--12 pg kg'). The other data show weaker correlations, nating from resuspension of surface sediments. Larger particles are often
perhaps because metal concentrations in surficial bottom sediments mainly composed of quartz-grains which have a lower trace metal con-
ordinarily exhibit a wide range. rather than a uniform value. centration thus diluting the pool of line suspended particles enriched in
organic carbon and metal oxides, the most efficient carrier phases for
many trace metals. Alternatively, this behavior can also be produced by
the particle-concentration effect, which has as a cause the presence of
trace metal-containing colloidal particles in the filter-passing fraction
(Honeyman & Santschi, 1989: Baskaran & Santschi, 1992: Baskaran
et al., 1992). One way to test this hypothesis would be to measure
particulate trace metal levels as a function of suspended particle size,
including colloidal particles.
Dissolved trace metals do not show a simple pattern relative to salinity
or SPM concentration (not shown). Based on our research unsing
naturally occurring radionuclides (Baskaran & Santschi, 1992; Baskaran
et al., 1992), we believe that trace metals are scavenged and released very
rapidly in Galveston Bay waters. This means that dissolved trace metal
levels are the result of rapid particle uptake (adsorption and colloid
aggregation) and release from particles (desorption and colloid disaggre-
gation). as well as rapid cycling of particles derived by resuspension of
bottom sediments through the water column. Thus, dissolved trace metal
levels are controlled mainly by the combination of three processes: (1) re
suspension of bottom sediments to yield particulate and colloidal metals,
(2) steady-state equilibrium partitioning of metals between solid and
liquid phases, (3) steady-state partitioning between particles of different
sizes including those in the colloidal size range. As a result, dissolved

Fig. 5. Filter-retained Pb and Zn concentration in the water column as a function of
SPM. The good corrclations support the hypothesis that particulate metals in the water
column are derived from resuspended bottom sediments. The linear relationship implies
that the suspended sediments have nearly a constant metal conccutration (pg g') at all
locations sampled.

26 J. W. Morse, R. J. Presley, R. J. Taylor, G. Benoit, P. Santschi
trace metal levels should depend on bothe the quantity and size spectrum of suspended particles, and on their partitioning characteristics. Further research is needed to elusidate the details of such a relationship.
although the behaviour of the dissolved metals was complex, certain simple trends are worth noting. In general, metal concentrations were lowest near the fulf salinity end member. this is to be expected, since rivers act as a source of metals to the ocean. Also, at all salinities, dissolved Pb was much lower than particulate Pb, while the opposite was true for Cu. Zinc and Ag showed approximately equal partitioning between dissolved and particulate fractions. This trend matches our expectations ince Pb is highly particle reactive (e. g. Balistrieri & Murray, 1984: Santschi el al., 1984), while Cu tends to form soluble organic com-
8000
1986
1987
6000
1988
1989
1990
4000
2000
0
GBR GBHR GBOB GBSC GBID GBYC
Zinc (ppm)
10
8
6
4
2
0
GBCR GBHR GBOB GBSC GBID GBYC
Cadmium (ppm)
Fig. 6. annual mean and standard deviation of metal concentrations in oysters collected at six sites in Galveston Bay from 1986 90. Solid horizontal line represents mean of 874 oyster samples collected alon US Gulf of Mexico coast from 1986 to 1990. (A) An. (B) CJ.
Trace metal chemisty of Galveston Bay 27
plexes (e. g. Sunda & Hanson, 1987). dissolved Cu showed the simplest pattern against salinity, decreasing from the fresh to the saline end member via a mid-salinity maximum. this pattern would be consistent with release of dissolved Cu from particles or sediments at mid-salinities. Another possibility is addition of dissolved Cu in water originating from the San Jacinto River or Clear Lake area.
Trace metal concentrations in various estuaries exhibit a wide range, reflecting large local differences in inputs and removal processes. For that reason, it is difficult to compare Galveston Bay to other estuaries, except to say that concentrations are in the same range as measurements elsewbere. Since, unfortunately, these are the first reliable trace metal data for this estuary it is impossible to draw conclusions about historical trends. Because metals in Galveston Bay waters are dominated by exchange with sediments, it seems reasonable that water column metal concentrations are as variable as sediment metal concentrations (see below).
Discussion of metals in Galveston Bay oysters averaged over all sites and all years obviously cannot show possible geographic and temporal trends within the Bay. In the case of Zn, for example, it can be seen (Fig. 6) that three of the six Galveston Bay sites had oysters with near Gulf average Zn, with relatively little year to year variations. The other three sites, two of which were not sampled in the first two years of the program, had much higher Zn concentrations. Two of the sites with high Zn concentratiosn, Ship Channel and Yacht Club, are in northwestern Galveston Bay near industrial waste water inputs and boat basins where Za contamination might be expected.
The other site with high Zn concentrations was in Offatts Bayou on Galveston Island and is surrounded by residential development and private boat moorings. This site was moved a few hundred metres between year three and year four, and the Zn concentration in oysters was lower by about 50%. This shows the extremely local influence on Zn content of oysters. Local control can be seen even more dramatically in the variations between stations at a given site for a given year (data not given here). Local control on Zn, and on other metals is seen not only in Galveston Bay but also throughout the Gulf of Mexico. For example, oysters from one site in Tampa BAy averaged 200-300 mgg Zn over the first five years of NS&T while a nearby site averaged 6000-8000 mgg (GERG, 1990). On a temporal scale, particularly large changes in Zn were found in Sabine Lake, Texas, another industrialized site. It seems likely that the big changes in Zn from time to time and place to place are caused by human activities, but the exact activity responsible for such a pattern has not been identified. Alternatively, it could be caused by

28 J.W. Morse, R.J. Presley, R.J. Taylor, G. Benoit, P. Santschi Trace metal chemistry of Galveston Bay 29

natural variations in concentrations and chemical form of sedimentary Fe, as suggested by Luoma & Bryan (1978).
Cadmium, Pb, Ag and Hg are often added to the environment in industrial areas by human activities in amounts rivalling those added by natural processes (e.g. Bowen, 1979; Fergusson, 1990) but there is little evidence of anthropogenic inputs of these metals in the Galveston Bay oyster data, except possibly for Pb. Confederate Reef and especially Hanna Reef are in open areas of the Bay, well away from industrial activity, yet oysters from these reefs are similar in Cd, Ag and Hg content to those from reefs along the highly industrialized northwestern shore of the Bay and are only slightly lower in Pb. The Cd pattern (Fig. 6) is thus representative of the distribution of these four toxic materials. Note that the highest Hg value found was at Hanna Reef, apparently the most pristine site sampled. Furthermore, the most anomalous Ag value was found at pristine Confederate Reef where all three stations sampled in 1990 were extremely enriched in Ag. We have no explanation for these results.
Attempting to assess whether sediments in a region such as Galveston Bay are contaminated with respect to a given metal is difficult. Much of this difficulty arises from the heterogenous nature of the sediments. For many of the metals under consideration the variation in grain-size distribution can easily lead to variations of a factor of two or more in absolute metal concentrations (Table 4). Consequently, simply giving average total concentrations can be quite misleading, if it is not normalized to grain size (Table 4).
Secondly, in order for contamination to cause a major change in average concentrations, over the entire estuarine system, large amounts of the metal from anthropogenic sources would have to be added (e.g. for Cu which is of intermediate concentration for the trace metals studied, about 60 mg m year of Cu, equivalent to 80 tons year for all of Galveston Bay, would have to be added to the sediments to double their average Cu concentration. This extimate assumes 13 mgg Cu interface sediments,and a sediment delivery rate to the Bay of 6 X 10 tons year (GURC. 1965). Thus, unless truly massive contamination with a given metal occurs it is generally not possible to detect the enrichment against the background of natural variability using regional averages.
However, while it is difficult to asses whether or not an entire area may be contaminated with a given metal, it is frequently possible to identify local sub-areas that have anomalously high concentrations of a given metal relative to the region under study as a whole. Such areas haveoften been observed to be close to the source of anthropogenic metal inputs. It must be kept in mind aht formation of such 'pockets of contamination' is strongly dependent on the degree of dispersion of the introduced metal, and a variety of other processes such as biological uptake and sedimentation patterns.
As discussed in the Results section, in this sutdy the search for sediments with anomalous metal concentrations was undertaken by ratioingtotal reactive-metal concentrations to total reactive-Fe is designated Me*. Average values of this parameter in Galveston Bay are compared to equivalent values in Baflin Bay, Texas (Huerta-Diaz & Morse, 1992) in Table 6. Baflin Bay was chosen for comparison because it is remote from major population centres and similar data exist for sediments from this Bay. (It should be noted that Baflin Bay is generally hypersaline whereas Galveston Bay has a salinity less than that of seawater.) Mn* is less inGalveston Bay, and Ni*, Cu*, and Pb* are similar in both bays. Zn*, CD*, Cr*, and Hg* ar at least 1-5 times higher in Galveston Bay. Mo* and As* are at least an order of magnitude higher in Galveston Bay than Baflin bay, but total As and other metals in the sediments of Galveston Bay are similar to concentrations in other Texas and Louisiana bays (GERG,1990).
It is not possible to unambiguously ascertain whether the trace metals that are higher in Galveston Bay are so as a result of anthropogenic inputs or differing natural sources and processes in the two bays. For example, Mo and As, which are highly elevated in Galveston Bay sediments relative to Falin Bay, exist in the water column dominantly as molybdate and arsenate, and As has been observed to be at close to normal concentrations in the open water column (Tripp, 1988). In the sediments they are extensively reduced along with sulphate and incorporated into iron sulphide minerals. The salinity in Baflin Bay is typically about four times higher than in jGalveston Bay. Consequently, even if Mo and As were in similar concentrations in the water column, their ratio relative to sulphate would be about four times higher in Galveston Bay. This difference in ratios could then well be reflected in their incorporation into sediments.
Three approaches have been made to try to identify sites in Galveston jBay which have 'anomalous' metal concentrations using Me* values. The first two are given in Table 6A, where samples having twice the average value are given in italics, and Table 6B where samples having a difference of over two standard deviations from the average value are given. The third approach is non-statistical and consists of observing Me* values in histograms (Fig. 7). The choice of values above which Me* is considered 'anomalous' is arbitrary. For some metals, such as Zn, it is relatively obvious, for others, such as Cr it is difficult. The histograms in Fig. 7 are

30 J W Morse, R. J. Presir"I., ft. J. TaYlor, G. Benoic 1@ Sanfvchi Trace tPietal chentistry of Galveston Bay 31

is
Nickel* Mercury*
V) 12 15 (X 1000)
ILL 12
E
15 25 E 9
13
Zinc* Cadmium* !n 'B
01 12 20
E E
10 -3 3 7)
z z
T .12MMI 1 :@ -m-lawk
2 10 0 0 02 2.4 46 60 8,10 10
do W
KI
E

0 24 40 60 8.10 10 O@O 2 0 2 OA 0 4 0.6 0.6 O's 0.0.1 1 15 15
Molybdenum* Chromium*
12 12

9 E 9
Manganese*
10 Copper*
0 12 6
T 0

3
9
6 6 01 0 -- B14
Ol 12 ZI 3,4 45 56 6 01 12 23 34 45 56 .6

E
3 E Fig. 7. conlit
01 Oil
'j
'V JP AV
arranged in increasing order of difficulty in making such a judgement.
The values chosen arc Zn* > 10, Cd* > 0.8, Mn* > 120, Cu* > 3.5,
15 12 As* > 12, Pb* > 3. Ni* > 3, I-lg* > 0-01, Mo* > 6, and Cr* > 6.
M 12 Arsenic* 10 Lead* Almost half (17 out of 39) sites exhibited an anomalous value for at
2 0 least one metal by a( least one of the above criteria (Table 6B). Site GBI-
CL W
E LL 8
2 16 off Eagle Point had anomalous value,.; for all metals. The number of
sites having anomalous values for each metal is Pb = 9, Ni 7, Cd and
0
As = 6, Mn and Mo = 5. Cr and fig = 4, Cu = 3. and Zn 2. Exccp-
4
E tionally higher values (delined its -4 times higher the standard deviation
E
z above the average) were encountered for Ni* and Cd* a( site Gill-16,
2L -38, Zn* at site 6131-7, and As* at site G111-1 1. While the
0 2 2.4 4.6 G-a 6 tO 10A2 .12 a0-05 0.5-1 1-1.5 15 222.5 2.53 13 Cu* at Gill I
high Cu* (Clear Lake site) may possibly be explained by leaching of Cu
froin anti-fouling paints on the many boats moored in this small area, the
possible reasons for the other high values are not obvious. It is interest-
ing to note that few of' the exceptionally high values occurred near Texas
L

City or directly in the Houston Ship Channel, where higher contaminant
Fig. 7. tlistograms of Mle* (= reac(ive-mcial to reaciive-iron concentralion ratio) values metal concen t rations might be expected, but that about 6(YY,. of the
for different metals in Galivesion flay sedimenis, dredge spoil sites had anomalous metal coriccritrations.

32 J.W. Morse, R.J. Presley, R.J. Taylor, G.Benoit, P. Santschi Trace metal Chemistry of Galveston Bay 33
SUMMARY AND CONCLUSIONS
100

0 Zn Mn Mi Pb Cd Cr Cu Mo As Hg

Fig 8. Histogram of percentage of samples for different metals having greater than 20% of the total reactive fraction pyritized in the top 10 cm of sediment.

A major objective of thsi investigation of trace metal levels in sediments was to investigate if extensive pyritization of reactive trace metal takes place near the sediment water interface (approximately top 10cm In the Results section data were presented indicating that this does occur for several of the metals of interest. Figure 8 is a histogram giving the percentage of each metal that was significantly (here defined as >20% pytitized. Metals in which less than 15% of the samples fell in this catagory include Zn, Mn, Ni, and Pb. As previously discussed, the data for Cd are uncertain due to its low concentration. Over 75% of the Cr, Cu, Mo, As and Hg samples were significantly pyritized. It should be noted that previous studies (e.g. Huerta-Diaz & Morse, 1992) indicate that pyritization of metals dominantly occurs in the upper 10 cm of sediment in fine-grained sediments associated with coastal environments. Consequently, these observations are of likely general validity for Galveston Bay. It is also interesting to note that the most extensively pyritize metals, As and Mo, both have total reactive concentrations normalize to Fe over an order of magnitude higher than in Baflin Bay sediments.
Based on extensive investigations into concentrations and chemical forms of selected trace metals in water, sediments and biota (e.g. oysters), we come to the following conclusions:
(1)Trace metal concentrations in the open water column of Galveston Bay are similar to those in apparently more pristine bays nad estuaries. Cu, Zn, Pb, and Ag concentrations in the water column of Galveston Bay are low, and are mostly regulated by sediment dynamics (wind and tide generated sediment suspension and settling), leading to significant association with SPM and correlations of their particulate concentrations in the water with suspended matter concentrations. Concentrations of these metals in suspended particles >0-4mm, diameter resemble those inthe sediments and are higher at low particle concentrations, indicating enrichment in the finer, slower settling fraction.
(2)Except in Zn, trace metal conceantrations in oysters from Galveston Bay are similar to those in oysters from pristine areas elsewhere and do not reflect the relative differences inproximity to population and industrialization centres of the different sampling sites in the Bay.
(3) Average trace metal concentrations in the sedimnets are similar to those from other estuaries. However, due to the large range of concentrations observed for many trace metals, meaningful comparisons require normalization to grain size and reactive-Fe. In Galveston Bay sediments, total reactive Zn, Cd, Cr, and Hg are at least 1-5 times higher, and As and Mo over an order of magnitude higher when normalized to reactive-Fe, than in sediments from Baflin jBay (Huerta-Diaz & Morse, 1992). Since trace metals in the water column closely reflect those in sediments of this estuary, it is likely that the same metals may have been historically elevated in teh water column as well. Forty four percent of the individual sites in jGalveston Bay exhibit and 'anomalous; concentration with respect to at least one of the metals studied. About half of these sites were directly associated with dredge spoils. It is not possible to unequivacally determine if this is the result of contamination from anthropogenic sources, however, it is probable that these elevated concentrations of metals in the sediment reflect past conditions during which anthropogenic metal inputs were higher.
(4) A major fraction of reactive Cr, Cu, Mo, As, and Hg are immoblized by incorporation into authigenic pyrite in the top 10 cm of sediment. These metals may be transformed via pyrite oxidation (Morse, 1991) to more bioavailable species if the sediments are resuspended in teh oxic water column by storms or activities such as dredging and bottom trawling.

34 J. W. Morse, R. J. Presley, R. J. Taylor, G. Benoit, P. Santschi

ACKNOWLEDGEMENTS

The authors were supported in this research by the NOAA Sea Grant Program (J.W.M.), NOAA Status and Trents Program (B.J.P.
and R.J.T.), the Texas Chemical Council and the Texas Institute of Oceanography (P.S. and G.B.).
REFERENCES
Amer. Public Health Assoc. (1985). Standard Methods for the Analysis of Water and Wastewater, 16th edn. American Public
Health Assoc., American Water Works Assoc., Water Pollution Control Federation, American Public Health Assoc., Washington,
DC, 1134 pp. Armstrong, N.E. (1982). Responses of Texas estuaries to freshwater inflows. In Estuarine Comparisons, ed.
V. S. Kennedy, Academic Press, New York, pp. 103-20. Balistrieri, L. S. & Murray, J. W. (1984). Marine scavenging: Trace
metal adsorption by interfacial sediment from MANOP site H. Geochim. Cosmochim. Acta, 48, 921-9. Baskaran, M. & Santschi,
P. H. (1992). The role of particles and colloids in the transport of radionuclides in coastal environments of Texas. Mar.
Chem., (in press). Baskaran, M., Santschi, P. H., Benoit, G. & Honeyman, B. D. (1992). Scavenging of thorium isotopes by
colloids in seawater of the Gulf of Mexico. Geochim. Cosmochim. Acta (in press). Benoit, G., Oktay, S., Cantu, A., Hood,
M. O., Coleman, C. H., Corapeioglu, O. & Santschi, P. H. (1992). Partitioning of Cu, Pb, Ag, Zn, Fe, Al and Mn between
filter-retained particles, colloids and solution in 6 Texas estuaries. Mar. Chem. (submitted). Berner, R. A. (1970)
Sedimentary pyrite formation. Amer. J. Sci., 268, 1 23. Bevington, P. R. (1969). Date Reduction and Error Analysis for
the Physical Sciences. McGraw-Hill Book Co., New York, 336 pp.

36 J.W. MORSE, R.J. PRESLEY. R.J TAYLOR, G. BENOIT, P. SANTSCHI

Martin.J.II., Knauer. G.A. & Gordan.R.M.(1983). Silver distributions and
fluxes in north-east Pacific waters. Naure,305,306-9.
Morse.J.W.(1991). Sedimentary pyrite oxidation kinetics in seawater. Geochim.
Cosmochim. Acta. 55.3665-8.
NOAA (1985). The National Status and Trends Program for Marine Enviromen-
tal Quality. NOAA Technical Memorandum NOS OMA 45. NOAA Office
of Oceanography and Marine Assessments. Ocean Assessments Division.
Rockville, MD, 13pp.
NOAA (1989a). Galveston Bay: Issues, Resources Status, and Management.
NOAA Estuary of the Month Seminar Series No. 13. NOAA EStuarine
Programs Office. Washington, DC, 114pp.
NOAA (1989b). National Status and Trends Program for Marine Enviromental
Quality. A summary of data on tissue contaminations from the first three
years (1986 1988) of the Mussel Watch Project. NOAA Technical Memo-
randum NOS OMA 49. NOAA office of Oceanography and Marine
Assessments. Ocean Assessments Division. Rockville, MD. 22pp. and
appendix.
Patterson, C.C.& Settle, D.M. (1976). The reduction in order of magnitude
errors in lead analysis by biological materials in natural waters by evaluat-
ing and controlling the extent in source of industrial lead contaminiation
introduced during sample collection and analysis. National Bureau of
Standard special Publication. 422(2). 321-51.
Presley, B.J., Taylor. R.J. & Boothe. P.N. (1990). Trace metals in Gulf of
Mexico oysters. Sci. Total Environ., 97/98. 551-93.
Santschi, P. II., Nixon, S. & Pilson. M. (1984). Accumulation of sediments,
trace metals (Pb. Cu) and hydrocarbons in Narragansett Bay. Rhode
Island. Estuary Coastal Shelf Sci., 19.427 50.
Schaule, B.K. & Patterson, C.C. (1983). Perturbations of the natural lead
depth profile in the Sargasso Sca by industrial lead. In Trace Metals in Sca
Water, cd. C.S. Wong, E. Boyle, K.W. Bruland, J.D. Burton & E.D.
Goldberg. Plenum Press, New York, pp. 487-503.
Shao. E.Y. & Winefordner, J.D. (1989). Determiniation of lead in coal by elec-
trothermal atomization-Zeeman atomic absorption spectrophotometry.
Microchem.J., 39. 229-34.
Shiller. A.M. & Boyle, E.A. (1985). Dissolved zinc in rivers. Nature, 317,
49 52.
Sturgeon, R.E., Berman, S.S., Willie, S.N. & Desaulniers. J.A.H. (1981). Pre-
concentration of trace elements from seawater with sifica-immobilized
8-hydroxyquinoline. Anal. Chem., 53,2337-40.
Sugimura, Y. & Suzuki, Y. (1988). A high-temperature catalytic oxidation
method for the determination of non-volatile dissolved organic carbon in
seawater by direct injection of a liquid sample. Mar. Chem., 24, 105 31.
Sunda, W.G. & Hanson, A.K. (1987). Measurement of free cupric ion concen-
tration in seawater by a ligand competition technique involving copper
sorption onto C1b SEP-PAK cartridges. Limnol. Oceanogr.,32,537-51.
Texas Water Commission (1987). Intensive Shady of the Houston Ship Channel.
(1S 87-09) Texas Water Commission, Austin, Texas. 89 pp.
Trefry. J.H. & Presley, B.J. (1976). Heavy metal transport from the Mississippi
River to the Gulf of Mexico. In Marine Pollutant Transfer, cd. 11. L.

Trace metal chemistry of Galveston Bay 37
Windom & R.A. Duce. D.C. Heath and Company, Lexington, MA,
pp.39 76.
Tripp, A.R. (1988). Geochemistry of arsenic and antimony in Galveston Bay,
Texas. MS Thesis, Texas A&M Univsersity, 75 pp.
Valenta. P., Duursma, E.K., Merk, A.G.A., Rutzel, H & Nurnberg. H.W.
(1986). Distribution of Cd, Pb, and Cu between the dissolved and particu-
late phase in the Eastern Scheldt and Western Scheldt estuary. Sci. Total
Environ., 53.41 76.
Windom, II. L., Wallace. G., Smith. R.G., Dudek, N., Macda, M., Dulmage.
R. & Storti, F. (1983). Behavior of copper in southeastern United States
estuaries. Mar. CHem., 12. 183 93.
Windom, H.L., Smith, R.G. & Macda,M. (1985). The geochemistry of lead in
rivers, estuaries and the continental shelf of the southeastern United States.,
Mar. Chem., 17.43-56.
Zhabina. N.N. & Volkov.I.I. (1978). A method of determiniation of various
sulfur compounds in sea sediments and rocks. In Environmental Biogeo-
chemistry and Geomicrobiology, 3: Methods, Metals and Assessment, cd. W.
E. Krumbein. Ann Arbor Sci. Publ., Ann Arbor, pp. 735-46.

Reprint 8

Mercury Bioaccumulation by Shrimp
(Penaeus aztecus) Transplanted to
Lavaca Bay, Texas

Sally J. Palmer and Bobby J. Presley

1-81

Marine Pollution Bulletin

DouAbul.A.A.Z. & Al-Saad. H.T. (1985). Seasonal variations of oil
residues in water of SHatt Al-Arab River. Iraq. Water,Air, Soil, Pollut.
24, 237-246.
El-Wakeel. S.K. & Riley. J.P. (1957). The determination of organic
carbon in marine mud.J. conseil int. pour lexplor. mer. 12. 180-183.
Goutx. M. & Saliot. A. (1980). Relationship between dissolved and
particulate fatty acids and hydroacarbons, chlorophyll a and
zooplanton biomass in Villefranche Bay, Mediterranean Sea. Mar.
Chem. 8. 299-318.
Jeng. W.L. (1981). ALiphatic hydrocarbons in river and estuarine
sediment of western Taiwan. Acta Oceanographic Taiwanica 12, 16-
27.
Metcalfe, L.D. & Schmitz. A.A. (1961). The rapid preparation of fatty
acid esters for gas chromatographic analysis. Anal. Chem. 33. 363-
364.
Saliot.A., Tissier. M.J. & Boussuge.C. (1980). Organic geochemistry
of the deep ocean, sediment interface and interstisial water. In
Advance Organic Chemistry (A.G. Douglas & J.B. Maxwell. eds).
pp. 333-341. Pergamon. Oxford.
Simoneit. B.R.T. (1977). The black sea a sink for terrigenous lipids.
Deep Sea Res. 24. 813-830.
Simoneit. B.R.T.(1978). The organic chemistry of marine sediment. In
Chemical Oceanography (J.P. Riley & R. Chester. eds). pp. 234-311.
Academic Press. New York.
Tynni. R. (1983). Geochemical survey of Finland. Report of Investiga-
tion. No. 60.
Volkman. J.K., Johns. R.B., Gillan, F.T. & Perry. G. J. (1980)
Microbial lipids of an interridal segment-1. Fatty acids and Hydro-
carbons. Geochim. Cosmochim. Acta 44. 1133-1143.
Welte. D.H. & Ebhardt. G. (1968). Distribution of long chain
n-paraffins and fatty acids in sediment from the Arabian Gulf.
Geochim. Cosmochim. Acta 32. 465-466.

Marine Pollution Bullene., Volume 26. No. 10.pp. 562-566.1993. (4)25-326x '93 s6.00-0.00
Printed in Great Britain. O 1993 Pergamon Press Ltd.

Mercury Bioaccumulation by Shrimp
(Penaeus aztecus) Transplanted to
Lavaca Bay, Texas

SALLY JO PALMER and BOBBY JOE PRESLEY
Department of Oceanography, Texas A&M University, College Station. TX 77843. USA

A field study was conducted to determine mercury
accumulation rates by brown shrimp. Penaeus aztecus,
transferred to a mercury contaminated estuary, Lavaca
Bay, Texas. Mercury levels in the caged shrimp rose
from an average baseline value of 347 163 ppb to
1170 107 ppb in 36 days, resulting in an average rate
of mercury uptake of 22 ppb per day. Our results show
that shrimp rapidly accumulate Hg when confined to a
contaminated area, even though the natural population
of shrimp in Lavaca Bay is not conataminated

As much as 29.9 kg day -1 of mercury was released into
Lavaca Bay, Texas from 1966 to 1970 by waste water
from a chlor-alkali plant (Fig. 1). Since 1970 the Texas
Department of Health (TDH) has issued periodic
health warnings and bay closures due to elevated (>0.5
pm wet wt) Hg levels in Lavaca Bay organisms, only to
reopen the bay to fishing when the Hg levels decreased.
Portions of the bay closed to commercial and sport
fishing of finfish and crabs in 1988 have not been
reopened as of this writing. Unlike natural population
of oysters (Crassostrea virginica) and blue crabs
(Callenectes sapidus) in Lavaca Bay no high (>0.5 ppm
wet wt) mercury levels in shrimp have been reported
by the TDH (Trebatoski & Gooris. 1990) or other
researchers who worked in the area (Blanton &
Blanton. 1972: Palmer. 1992). Therefore all of Lavaca
Bay remains open to shrimping.

Numerous laboratory metal accumullation studies
have been conducted using a variety of invertebrates
over the years (e.g. King & Davis. 1987: Riisgard &
Famme. 1986: Zanders & Rojas. 1992). In the work
reported here instead of a laboratory study, field caging
experiments were used to determine the uptake rate of
mercury by shrimp confined to a contaminated area of
Lavaca Bay. To our knowledge. this is the first time that
transplanted shrimp have been used to determine
mercury accumulation rates in the field. Although it is
impossible to control variables such as temperature,
salinity, food supply and turbidity in a field study, the
authors felt that a field caging study woult better reflect
natural conditions than a laboratory study using Lavaca
Bay sediment of mercury contaminated food.

Materials and Methods

In July 1991 similarly sized (3.2 0.31 cm rostral
length: 3.16 0.31 g wet wt) adult brown shrimp
(penaeus aztecus) were collected from Matagorda Bay
in a relatively uncontaminated area about 10 km from

564 1-83

Fig. 2 Mercury concentrations is shrimp from Matagorda Bay
confined to Lavaca and keller Bays. Each symbol represents an
individual shrimp sacrificed on that day. Lines are best fit
through all data from each bay.

which feed on shrimp. were a prime interest in this
study, therefore, the shrimp were analysed whole. Once
thawed, the shrimp were rinsed, weighed, freeze dried,
and digested according to a modification of USEPA
method 245.1 (USEPA. 1990) All samples were
analysed in replicate for total mercury using cold
vapour atomic absorption spectrophotometry (Hatch &
Ott. 1968).

Data quality control
Included with each set of samples analysed were
blanks and a dogfish muscle reference material.
DORM-1. certified for Hg by the National Research
Council of Canada. Analyses of DORM-1 run with
each set of shrimp samples were within the certified
value for Hg 95% of the time.

Statistical analysis
To determine relation ships between total Hg. caging
sites and time, statistical analyses using SAS Institute
Inc. software (SAS Institute Inc. 1985) were performed.
The general linear model (GLM) was used to test for
significant differences in Hg levels between Lavaca and
Keller Byas and day of the caging experiment. The
Least Square Means test. LSMEANS. was used to
verify changes in shrimp Hg levels over the duration of
the experiment.

Results

Slightly, Hg contaminated Matagorda Bay shrimp
caged in the highly contaminated portion of Lavaca Bay
readily accumulated additional Hg While shrimp caged
in uncontaminated Keller Bay did not sighificantly
change in Hg concentrations during the 36 day
experiment (Fig 2) Average Hg concentrations in
shrimp caged in lavaca Bay climbed from 347+163
ppb dry wt (n=9) on day 0 to 1170+107 ppb (n=3)
on day 36 Using the shrimp baseline and final mercury
concentrations, the average daily rate of Hg uptake was
22 ppb over the36 day experiment.
The LSMEANS test showed that Hg levels in shrimp
caged in Lavaca Bay on day 22 were significantly
higher (p<0.05) than baseline concentrations. The
GLM procedure indicated a significant difference in Hg
concentrations between Lavaca and Keller Bays at
every sampling period after day 0 at the p<0.05 level.

Fig. 1 Map of Lavaca Bay area showing the Matagorda Bay shrimp
collection site and the caging sites in Lavaca and Keller Bays.

the most heavily Hg contaminated part of Lavaca Bay.
These were transferred to the caging experiment sites in
Lavaca Bay and the control site, Keller Bay (Fig. 1).
The cages, 20X20X20 cm plastic storage crates, were
lined with 3.3 mm plastic mesh and held tow shrimp
each. To facilitate handling, groups of eight crates were
attached to 0.5X1.0 m plastic grates. Individual cages
were spaced approximately 7 cm apart on the grates to
minimize restriction of water flow around them. At each
site the three grates holding the cages were tied
together, weighted and pushed at least 1 cm into the
sediment. Therefore the caged shrimp could derive
food from organic detritus in the bottom sediment
(Britton & Morton 1989) as well as plankton and
demersal fauna that entered the cage through the mesh
(Gleason & Wellington. 1988).
The caged shrimp were sampled six times over a
period of 36 days. At each sampling as many as nine
individuals were collected from each of the two
locations. Immediately after collection the animals were
placed in plastic bags and frozen until analysed. The
cages that remained in the field after each sampling
episode were shifted slightly within the site in hopes of
renewing the food supply, but it is nevertheless possible
that the shrimp suffered food depravation. Although the
average dry weight of the whole shrimp decreased over
the experimental period from 1.25+0.39 g (n=9) to
0.97+0.09 g (n=6) in Lavaca Bay and 0.71+0.11 g
(n=7) in Keller Bay, the shrimp were vigorous and
appeared to be healthy when sampled.
Food chain relationships, especially potential routes
of Hg transfer to large commercially important finfish
565

1-84

Marine Pollution Bulletin

Discussion and Conclusions showed that average Hg levels in abdominal tissue were
2.25 times greater than in head and exoskeleton. There-
The accumulation of mercury by shrimp confined to fore it is likely that the edible muscle tissue of the
a contaminated area of Lavaca Bay shows that shrimp shrimp caged in Lavaca Bay became more con-
can become contaminated with Hg if forced to remain taminated during the 36 day exposure period than did
in a contaminated area for three weeks or more. The the whole organism.
fact that the natural population of shrimp collected in
the contaminated area are not contaminated implies The observation that the natural population of
shrimp caught from Lavaca Bay are not contaminated
that they spend less than three weeks at a time in this with mercury suggests that shrimp spend much of their
area. Additionally, the caging experiment in Keller Bay
time and obtain much of their food in non-contamina-
suggests that slightly contaminated shrimp are slow to ted areas of the bay or coastal ocean. The relative
depurate Hg. This contrasts with results from a similar
importance of water. sediment. and food in the accumu-
experiment where contaminated oysters rapidly lation of Hg by shrimp is still poorly understood and
depurated Hg when placed in Keller Bay (Palmer et al., could not be resolved in this study because all three
1993). The slow depuration of Hg by shrimp and the
media are known to be enriched in Hg at the caging site
low Hg in the natural population of shrimp in con- near the old chlor-alkali plant (Palmer. 1992). taminated Lavaca Bay implies that the shrimp do not
move into and out of the contaminated area on a time Blanton. W. G. & Blanton. C. J. (1972). Final report: a study of the
cycle that would result in their spending more than a concentrations of mercury in tissues of selected animals from Lavaca
Bay, Texas. The Texas Water Quality board. Austin. Texas
total of three weeks in the contaminated area during Britton. J.C. & Morton. B (1989). Shore Ecology of the gulf of mexico.
their lifetime. University of Texas Press.
The difference in Hg loss rates between oysters and Gleason. D.F. & Wellington, G. T (1988). Food resources of postlarval
brown shrimp (Penaeus aztecus) in a Texas salt marsh. Mar. Biol. 97.
shrimp may be due to differences in Hg speciation 329-337.
within the organisms, but we have no data to document Hatch. W. R. & Ott. W.L. (1968). determination of sub-microgram
this, Riisgard & Famme (1986) for example found the quantities of mercury by atomic absorption spectrophotometry.
Anal.Chem. 40.2085-2087.
retention efficiency. defined as the amount of accumu- King. D.G. & Davies. J.M. (1987). Laboratory and field studies of the
lated mercury divided by the amount of ingested accumulation of inorganic mercury by the mussel Mytilus edulis (L.)
mercury, in shrimp. Crangon crangon. to be 4% for Mar. Polha.Bull. 18,40-45.
Palmer. S J.(1992). Mercury bioaccumulation in Lavaca Bay. Texas.
inorganic and 75% for organic mercury during their 28 Thesis. College Station. Texas.
day experiment. It is well known that methyl mercury Palmer. S.J.,Presley. B. J. Taylor. R. J. & Powell. E.N. (1993).Field
is more efficiently accumulated and retained than studies using the oyster Crassostrea virginica to determine mercury
accumulation and depuration rates. Bull.Environ. Contamin.
inorganic mercury (Riisgard et al.. 1985). Shrimp. Toxicol (in press).
unlike oysters consume sediment dwelling organisms. Riisgard. H.U.& Famme. P, (1986). Accumulation of inorganic and
These may contain a higher proportion of methyl organic mercury in shrimp. Crangon crangon. Mar. Pollut. Bull. 17.
mercury than plankton and organic detritus found in Riisgard.H.U. Kjorbe.T., Mohlenberg. F., Drabaek I.& Pheiffer
the water column. even though our data of total Hg Madsen. P.(1985). Accumulation. elimination and chemical
shows these two food sources to be similarly con- speciation of mercury in the bivalves Mytilus edulis and Macoma
taminated (Palmer. 1991). balthica Biol. 86.55-62.
SAS Institute Inc. (1985) SAS' User's Guide: Statistics. Version 5
Since an aliquot of a whole homogenized shrimp was Edition. SAS Institute Inc.
used for analysis in this study. concentrations in Muscle Trebaloski.B. & Gooris. J. (1990). Texas Water Commission Natural
Resource Damage Assessment: Pre-assessment Screening
tissue cannot be obtained directly from this data.
Document of Lavaca Bay 2453. (Corpus Christi. Texas).
However. in a separate study (Palmer. 1992), 18 USEPA (1990). Contract Laboratory Program Statement of Work for
Matagorda Bay shrimp collected with those used in the Inorganics Analysis document Number ILMO1.0.
Zanders. P I. & Rojas. W. E. (1982). Cadmium accumulation. LC, and
caging accumulation studv were dissected into muscle 0xygen. consumption in the tropical marine amphipod Elasmopus
rapax. Mar. Biol. 113.409-413.
(abdominal tissue). exoskelton. and head. The results

566
1-85

Reprint 9

Pol nuclear Aromatic Hydrocarbon
Contaminants in Oysters from the Gulf of
Mexico (1986-1990)

Thomas J. Jackson, Terry L. Wade,
Thomas J. McDonald, Dan L. Wilkinson
and James M. Brooks

1-87

Environmental Pollution 83 (1994) 291-298

POLYNUCLEAR AROMATIC HYDROCARBON
CONTAMINANTS IN OYSTERS FROM THE
GULF OF MEXICO (1986-1990)

Thomas J. Jackson, Terry L. Wade, Thomas J. McDonald, Dan L. Wilkinson
& James M. Brooks
Geochemical and Environmental Research Group, College of Geosciences and Maritime Studies,
Texas A & M University College Station. Texas 77845. USA

(Received I July 1992: accepted 25 September 1992)

Abstract equilibrium concentration for trace organic contami-
Polynuclear aromatic hYdrocarbon (PAH) contaminant nants such as PAHs within approximately one month
concentrations in 870 composite oYster samples from (Sericano & Wade. unpublished data).
coastal and estuarine areas of the Gulf of Mexico ana- To assess the spatial and temporal variation of con-
lyzed as part of National Oceanographic and Atmo- taminant levels of coastal and estuarine environments.
spheric Administrations(NOAA's) National Status and the National Oceanic and Atmospheric Administration
Trends (NS& T) Mussel Watch Program exhibit a log- (NOAA) instituted the National Status and Trends
normal distribution. There are two major populations in (NS&T) Mussel Watch Program under its Program for
the data. The cumulative freuency function was used to Marine Environmental Quality (O'Connor. 1990). The
deconvolute the data distribution' into two probabilitY sample sites were selected to characterize the overall
densitY functions and calculate summarY statistics for concentration of contaminants in coastal and estuarine
each population. The first population consists of sites ecosystems away from known point-sources of contam-
with lower PAH concentration probably due to back- ination.
ground contamination i.e. stormwater runoff, atmo- The focus of this paper is to examine the distribution
spheric deposition;. The second population are sites with of the PAH contaminant concentrations in oysters
higher concentrations of PAHs associated with local collected from the Gulf of Mexico as part of NOAA's
point sources of PAH input ,i.e. small oil spills, etc. NS&T Mussel Watch Program. and determine the
The temporal pattern for the mean concentration of the environmental factors controlling, the concentration of
populations from the Gulf of Mexico is consistent with PAHs.
large-scale climatic factors such as the El Nino cycles
which affect the precipitation regime.
METHODS
INTRODUCTION Sample collection
Oysters (Crassostrea virginica) were collected from
Oysters and other bivalve molluscs have been used for three stations at each site durine the winter of each
monitoring contaminants in the environment (Farring- year (1986-1990). The number of sites per year varied
ton et al 1983). Oysters are sentinel organisms which from 48 to 68. In some years not all sites had three
concentrate contaminants from the marine environ- stations due to the low abundance of oysters at a specific
ment. yet do not readily metabolize contaminants such site (Table 1). Sample sites give coverage of the Gulf of
as polynuclear aromatic hydrocarbons (PAHs) (Far- Mexico coastal and estuarine areas from southern-most
rinton & Quinn, 1973). PAHs enter the near-coastal Texas to southern-most Florida (Fig. 1). Individual
environment through a number of mechanisms (e.g. stations at each site are generally from 100 to 1000 m
runoff. discharge of industrial waste or sewage. natural apart. An analysis at each station represents a corn-
or industrial combustion processes. natural oil seep- posite of twenty individual oysters. Each year. the field
ages. and spills of petroleum or petroleum products). sampling returned to as many sites as possible. In some
The contaminants found in oysters reflect the current instances it was necessary to relocate or abandon an
contaminant burden of an ecosystem. The concentra- Table 1. National Status and Trends Oysters Gulf of Mexico
tion of a contaminant in an oyster is the difference Sampling Program-Summary of sampling
between uptake and excretion of that contaminant.
Galveston Bay oysters transplanted from a 'high' level 1986 1987 1988 1989 1990
site to a 'low' level site. and vice versa. come to a new
Year I II III IV V
Number of sites 49 48 65 62 68
Environ. Pollut. 00269-7491'94'506.00 C 1993 Elsevier Science Number of samples 142 144 195 186 203
Publishers Ltd. England. Printed in Great Britain

291

1-89

MISSISSIPPI ALABAMA GE

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PAH contaminants in oysters from the Gulf of Mexico

established oyster site due to lack of suitable seed Gas chromatography - mass spectrometry (GCC-MS)
bivalves (Wilkinson et al., 1991). The locations and PAHs were separated and quantified by GC-MS
designator for the oyster sites are found in Wilkinson et (HP5980-GC interfaced to a HP5970-MSD). The sam-
al. (1991). Sericano et al. (1990) and Wade et al. ples were injected in the splitless mode on to a 30 m
X0.25 mm (0.32 m film thickness) DB-5 fused silica
Tissue extraction capillary column (J&W Scientific Inc.) at an initial tem-
The tissue extraction process used was adapted from a perature of 60 C and temperature programmed at
method developed by MacLeod et al. (1985). Approxi- 12 C/min to 300 C and held at the final temperature
mately 15 g of wet tissue were used for the PAH for 6 min. The mass spectral data were acquired using
analysis. After the addition of internal standards (surro- selected ions for each of the PAH analytes. The
gates) and 50 g of anhydrous NA2SO4 the tissue was GC-MS was calibrated and linearity determined by
extracted three times with dichloromethane using a injection of a standard containing all analytes at five
tissuemizer. A 20 ml sample was removed from the total concentrations ranging from 0 01 to 1
solvent volume and concentrated to one ml for Sample componenet concentrations were calculated
percentage determination. The 280 ml of remaining from the average response factor for each analyte
solvent was concentrated to approximately 20 ml in a Analyte identifications were based on correct retention
flat-bottomed flask equipped with a three-ball time of the quantitation ion (molecular ion) for the
column condenser. The tissue extract was then trans- specific analyte and confirmed by the ratio of quantita-
ferred to a Kuderna-Danish tube heated in a water tion ion to confirmation ion.
(60 C) to concentration the extract to a final volume of Calibration check samples were run with each set of
2 ml. During concentration. the dichloromethane was samples (beginning, middle, and end). with no more
exchanged from hexane. than 6 h between calibration checks. The calibration
The tissue extracts were fractionated by aluminatisilica check must maintain an average response factor within
(80-100 mesh open column chromatography. The 10% for all analytes with no one analyte greater than
silica gel was activated at 170 C for 12 h and partially +25% of the known concentration. A laboratory refer-
deactivated with 3% distilled water (ww). Twenty ence sample (oil spiked solution) was also analyzed
grams of silica gel were slurry-packed in dichloro- with each set of samples to confirm GC-MS system
methane over 10 g of alumina. Alumina was activated performance and calibration.
at 400 C for 4 h and partially deactivated with
distilled water (ww). The dichloromethane was replaced RESULTS AND DISCUSSION
with pentane by elution. The extract was then applied
to the top of the column. The extract was sequentially Oyster site variations
eluted from the column with 50 ml of pentane (aliphane During the first five years of this study a total of 870
fraction) and 200 ml of 1:1 pentane: dichloromethane composited oyster samples have been analyzed for
(aromatic fraction). The aromatic fraction was further PAHs. The PAH (total NS&T PAHs) is the sum of the
purified by HPLC to remove the lipids. The lipids were eighteen aromatic hydrocarbon analytes. as measured in
removed by size exclusion using dichloromethane as Year 1. with concentrations greater than 20 ngg dry wt
an isocratic mobile phase (7 ml min) and two 225 (Table 2): this was the reporting limit for Year 1 data
250 mm Phenogel 100 columns (Krahn et al., 1955). (Wade et al., 1981). The median PAH concentration at
The purified aromatic fraction was collected from a site is used as a measure of the best indicator of the
15 min prior to the elution of 4.4 - dibromofluoro- concentration. The median is a more stable (or resistant)
biphenyl to 2 min. after the elution of perylene. The
retention times of the two marker peaks were checked Table 2. National Status and Trends oysters polynuclear
prior to the beginning and at the end of a set of 10 aromatic hydrocarbon analytes
samples. The purified aromatic fraction was concen- Aromatic hydrocarbons
trated to 1 ml using a Kuderna-Danish tube heated in
a water bath at 60 C.
Quality assurance for each set of ten samples in- Low molecular weight High molecular weight
cluded a procedural blank. matrix spike. duplicate. and Biphenyl Fluoranthene
tissue standard reference material (NIST-SRM 1974) Naphthalene Pyrene
which were carried through the entire analytical scheme. 1-methylnaphthalene Benz(a)anthracene
Internal standards (surrogates) were added to the sample 2-methylnaphthalene Chrysene
prior to extraction and were used for quantitation. The 2.6-dimethylnaphthalene Indeo[1,2,3-cd]pyrene
surrogates were d-naphthalene. d-acenaphthene. 1.6.7-trimethylnaphthalene Benzo(a)pyrene
d10-phenanthrene d -chrysene. and d-perylene. Surro- Acenaphthene Benzo(e)pyrene
gates were added at a concentration similar to that Acenaphthylene Perylene
expected for the analytes of interest. To monitor the Fluorene Dibenz(a,h)anthracene
recovery of the surrogates. chromatography internal Phenanthrene Benzo(g,h,f)perylene
standards d10-fluorene and d12-benzo(a)pyrene were Anthracene
added just prior to GC-MS analysis. 1-methylphenanthrene
Analytes not used in tPAH summation.

1-91

294 T J Jackson et al.

Table 3. Total N'S&T PAK concentration in oysters

"'0, Site Median concenir---ori of tPAH Bay group No. Site Median concentration of tPAH BaN gr1up
code median crkje! median
V IV 111 11 1 (ng:g) V IV III if I (ng;g)
1990 1989 198@ @4@'7 1996 1990 1989 1988 1987 1986
Texas (11grg) (ng gi kng SP g@ (ng,g) Louisiana-cont (ng g) (ng g) (ng,g) (ngrg) Ing,g)
I LMSB 22 20 30 20 25 65 MRTP 212 310 1410 - - 391 t 582
52 LMPI - - 31,80 - - 30 � 58 64 \IRPL 403 '30 695 - -
78 LMAC 120 - - - - 31 BSSI 185 71 484 68 177 181 t 134
53 CCBH 1530 - 1 600 - - 30 BSBG 45 202 213 118 265
2 CC\B 161 264 59S 45 565 � 725 32 LBMP 20 84 89 26 20 39+-59
3 CCIC 137 430 848 1 140 62 LBNO - - 81 - -
54 ABHI - - 1870 -
4 ABLR 20 20 20 1 20 Mississippi
10 10 33 MSPC 103 300 175 114 99
5 CBCR 88 - - 22 20-- 1 4 MSBB 1 110' 893 1500 4 110 1 600 3,22 t 6_54
(3 \A BN R 20 20 20 10 21 35 MSPB 59 306 776 300 246
7 SAPP 26 - - ; 1 45 Alabama
8 S.4,\lp - - - .19 93 25 3 6 M BCP 20 90 288 1 -3, 7 1
9 ESSP 20 - - 20
66 MBHI 761 554 1110 295 -40
10 ESBD 11 70 11 - 79 MBDR 1 520 - -
12 MBGP - 20 86 20 Florida
I 1 96 348 - ;9 90 45 48
67 PBPH 168 369 842
56 MBCB 20 - ;6 - -
3,7 PBIB - ZI 204 250 406 197 198
13 MBTP 20 20 @6 10 20 8@ PBSP 130 - - - -
MBDI - -
73 CBJB 1 680 8590 - - -
14 MBEM 201 200 78 138 � 119 -
39 CBSP 225 4@ 703 543 418 429� 1 140
7? BRCL 761 60 38 CBSR 69 21 24;40 2470 209
57 BRFS 95 5 1 670 682 - 792 � 792 74 PCLO 98 129 - - -
18 GBCR 370 1 170 52@ 1 070 68 PCMP 1 210 26()0 4 750 - -I goo � 1 ;90
5 8 GBOB 593 54" - 40 S. -% V @ B1 150 2 090 1990 1 970 11 800
16 GBTD 44 20 i2 149 259 � 606 14
41 APDB 20 2 800 20 20 57 � 530
t; GBYC 247. 1 207 -@s 1 0110 - 740 t)
I - 42 APCP 269 1110 109
.9 GBSC 1290 . @O 3100
17 GBHR 20 119 -'4 10 75 AESP -3 74 - - -
Louisiana 69 SRN\'P - - 119 - -
19 SLBB 108 154 169 26 247 1 @4 72 43, CKBP 20 74 1-4 69 22 46 103
20 CLSJ 180 1-28 10-1 '76 220 218 /6 TBN? 269 194
60 CLLC 404 726 20 47 TB%IK 101 1@0 -?0 49
44 TBPB 20 2!7 2 86 68 9@;
21 JHJH 88 72 20 S4 43 44-- 50 70 TBOT 112 ',;7 __12 - - 126 t 165
22 VBSP 189 1 79 79 108 TBKA 252 S,4 - - -
'N ABOB 20 18 192 -'2 '2 42 45 TBHB - - 1 460
2 5 CLCL 20 54 20 0 20 46 TBCB 20 6; 94 12 20
10 48 CBBI 20 31 41 20 ;1 _+ 180
26 TBLB 49 306 20 40+- 162
_27 TBLF 101 50 8-' -15 71 CBFM 69 ;16 27-1 - -
61 BBTB - - - 49 \BNB 8_1 20", 2 4; 11 108 1-1-8 7_1 � 129
50 RBHC 20 @7 6-7 10 4-7
28 96' � 1 020
BBSD 963 5 480 44 -7
29 BB%IB 1080 1 ' 80 1 46C) 41) 822 1 ENTU 47 68 2 20 112 68 � 125

estimator of the tvpIcal value than, the mean for data MBLR. MBCB. MBTP & MBDI) and Aransas bavs
which ma,, contain outliers (Hem.-I. 1990). (ABLR. CBCR & MBAR) which exhibit low median
in con-
The data in Table 3 presents th, spatial and temporal concentrations of tPAH and small variability '
variation for the median tPAH concentration in the centration. The highest median tPAH concentration for
coastal and estuarine areas of the Gulf of Mexico. The a bay group in Texas is the Brazos River (BRCL &
sites are separated into Bay groups (Wilson et al.. 1992) BRF'S). which carries the runoff from agriculture and
for data comparison. The vaniability for each Bay wastewater discharge from industrial point-sources
group is the standard deviation @_;; computed from the (NOAA. 1985). For the entire coastal and estuarine
interquartile range OQR) for th.-. five years of data area of the Gulf of Mexico (Table 3). the highest
(Hensel. 1990). In Texas. Co-,pus Ch'risti (CCBH. median tPAH concentration for a bay group is near
CCNB, CCIC & ABHl) and Gaixeston bays (GBCR. Panama Cirv. Florida (PCLO. PCMP & SAWB),
GBOB. GBTD. GBYC. GBSC & GBHR) are near which is close to a paper mill (NOAA. 1985. Wilkinson
industrial and population cent.-7s and exhibit high ei al.. 1991 @. I
median concentrations of tPAH and laree variability in There are fifteen sites (LMSB. ABLR. CBCR.
concentration compared to Matagorda (ESBD. MBGP. MBAR. SAPP. ESSP. ESBD. MBGP. MBCB. MBTP.

1-92

PAH contaminants in oysters from the Gulf of Mexico 295

NS&T PAH Data - Years I to V NS&T PAH Data - Years I to V
500

400
300

200

100
CLCL-1 CLCL-2 CLCL-3
Site and Station Median of Site NS&T

Fig. 2. Total NS&T PAH concentration distribution during Fig. 4. Frequency distribution of the median total NS&T
the first five years for all three stations: Caillou Lake in PAH (tPAH) concentration in the Gulf of Mexico during the
Louisiana (Site 25-CLCL). first five years of the program.

CLCL. LBMP. TBCB. CBBI & RBHC) with low Cumulative frequency model
concentration of tPAH (< 100 ng/g) and little variation Bar graphs (Wade et al. 1990) or crossplots (Wade &
in the observed values (Fig. 2). There are also six sites Sericano. 1989) of data comparing one year's data with
(GBSC, BBMB, MSBB. CBJB. PCMP & SAWB). of another have been used to display the general trend for
the seventy-eight different sites. where high concentra- tPAH data (Wade & Sericano. 1989: Wade et al. 1990:
tions of tPAH (>1000 ng/g) are observed. Four sites Wade et al. 1991). These data presentations easily
(CCIC PBPH. PBIB & PCMP) exhibited a decrease in visualize the variation in concentration for a particular
the tPAH each year during the first five years of this site. In this report the cumulative frequency function is
study. Many sites exhibited a cyclic variation with time. used to examine the heterogeneous distribution of PAHs
At Choctawatchee Bay off Santa Rosa (CBSR. Fig. 3). in Gulf of Mexico oysters (Mackay & Paterson. 1984).
the order of magnitude increase in concentration of This approach has the advantage of examining the Gulf
tPAH in Years II and III is probably due to relocation of Mexico as a single environmental system. determining
of the collection site to an area containing wood pilings. the percentage of sites exposed to a particular threshold
which if treated with creosote. are a source of PAHs. concentration. and providing information for environ-
The decrease in Years IV and V probably reflects relo- mental evalualion.
cation of the collection stations to an oyster reef away The distribution of the PAH data in Table 3 is best
from wood pilings. Due to prolonged freshwater condi- described by a lognormal distribution i.e. the distribu-
tons in San Antonio Bay during 1988 and 1989 (Years tion of data is skewed to low concentrations and has a
III IV). the oyster reefs experienced a die-off resulting in fraction which extends to high concentrations (Fig. 4).
no oysters being taken from SAPP. SAMP and ESSP. O'Connor (1990) used the lognormal distribution.
typical of environmental data. to define high concentra-
NS&T PAH Data - Years I to V tions as those whose logarithmic value is more than the
mean plus one standard deviation of the logarithms for
all conctntrations. The tPAH data in Fig. 4 is further
4500 skewed in that analytes with concentrations less than
4000 20 ng g are not included in the sum of eighteen 2-5
3500 ring aromatic hydrocarbon analytes in Table 2. i.e. the
3000
data has been censored. For Years I-III, on1y censored
2500 data was available. whereas for Years IV and V both
2000 censored and uncensored data was available. A regres-
sion analysis of the censored (tPAH) data versus
1500 uncensored data for the sum of all analytes (T-PAH) in
1000 Table 2 from Years IV and V yields the best fit line as
y = 153-0 + 0.9834 x (r2 = 0.99289): where y = uncen-
50O sored data. and x = censored data. Using the best fit
0 line from the Year IV and V data. the censored data
CBSR-1 CBSR-2 CBSR-3
site and station for the cumulative frequency data was corrected to be
Fig. 3. Total NS&T PAH concentration distribution during the same as the uncensored cumulative frequency data.
the first five years for all three stations: Choctawatchee Bay Distribution functions are useful measures of environ-
off Santa Rosa (Site 38-CBSR). mental quality data in that changes with time can be

1-93

296 T J. Jackson et al.

Total NS&T PAHs (ppb) Total NS&T PAHs (ppb)
Fig. 5. Plot of the cumulative frequency distribution Fig. 6. Plot of the cumulative frequency distribution for Year
V total NS&T PAH (PAH) concentration. compared to the V NS&T PAH (PAH) concentration. compared to the
Gaussian curve and its cumulative frequency distribution gen- Gaussian curves and their cumulative frequency distributions
erated from a lognormal model with a mean of 250 ppb and generated from a two population lognormal model with a
standard deviation of 218. mean of 214 ppb for Population I and a mean of 1205 ppb
ascertained without being influenced by outliers. For for Population 2.
tile cumulative distribution plot. the data is sorted from computed. but did not compare as well with the actual
the lowest value to the highest. similar to rank trans- data for Year V.
formation (Conover & Iman. 1981). Each observation The implication of the two populations in the data is
is I n fraction of the data set. where n is the number of that there are two primary mechanisms accounting for
samples in the data set. The sum of the fraction of the the distribution of T-PAH concentration in the Year V
samples less than the concentration is plotted against data. The sites with lower concentration PAHs are prob-
the concentration., From this plot the median can be ably due to low level background inputs from storm-
determined. since it is defined as the 50th percentile. water runoff. atmospheric deposition and sewage
The interquartile range (IQR) is used a measure of effluents. etc. (NOAA. 1985). The sites with higher con-
variability. The lQR is the 75th percentile minus the centration PAHs are probably due to local point-sources
25th percentile and equals 1:35 times the standard of PAH contamination (i.e. small spills). From the log
deviation for a normal distribution (Hensel. 1990). normal cumulative frequency function two probability
To begin the examination of the distribution of the density functions were derived. the relative proportion of
PAH concentration data, the logarithm of the sum of the two populations were estimated to be 0.9 for popula-
all PAH analytes (T-PAH) for Year V data was plotted tion one and 0.25 for population two. Comparison of
as a cumulative frequency distribution. The 50th the cumulative frequency distribution derived from the
percentile was 250 ppb and the standard deviation as sum of the two probability density functions. in the
determined from the IRQ was 218. The log of the data above proportions. with the actual data for the cumula-
versus fraction of the samples was plotted and com- tive frequency distribution (Fig.7) indicates a good
pared with a lognormal distribution (fig. 5). The shape correlation.
of the cumulative frequency curve (i.e. the positive
deviation from the lognormal model for the T-PAH
data suggests two overlapping lognormal distributions.
Making the assumption that there is a 25 overlap for
the two distributions. the mean and standard deviation

were computed for each data set. or population (Table
4). Tile cumulative frequency distribution from the two
population model data compare well with the actual
T-PAH data (Fig. 6). Other increments of overlap were

Fig. 7. comparison of the cumulative frequency distributions
for the actual Year V total NS&T PAH (PAH) concentra-

tion data and the cumulative frequency distribution generated
from the two population model.

PAH contaminants in oysters from the Gulf of Mexico 297

Table 5. Two population lognormal distribution model. Corrected tPAH data-ng/g dry weight

Year Median Population I Population 2
1 mean (log) std (log) mean (log) std (log)
11 197(2.294 5) 108 (0.229 8) 1 075 (3.031 4) 714 (0.277 2)
111 186 (2.269 5) 87 (0.196 7) 1 150 (3.059 9) 1 100 (0.381 1)
IV 259 (2.413 3) 216 (0.343 5) 1 910 (3.280 8) 1 190 (0.261 8)
V 269 (2.429 8) 174 (0 250 0) 1 350 (3.131 6) 1 190 (0.303 9) 212 (2 326 3) 131 (0 263 9)
1 170 (3.068 9) 637 (0.243 5)

Since historical NS&T data (Table 3) is censored tration. while Year III had 80%. Year IV had 83%.
data ( Wade et al lass: Wade & Sericano. 1989: Wade and Year V had 87%. Alternatively. the cumulative
et al.. 1990 the acmulative frequency distribution of frequency data can be used to calculate the percentage
this censored (pah data was corrected using the best- of sites exposed to a concentration in excess of a partic-
fit-line from the or Years IV and V. Data below ular threshold.
the reporting limit were extrapolated (Hensel. 1990: The cumulative frequency distribution was used in
Mackay & Paterson 1984). The summary statistics for this study as an environmental evaluation tool to
the corrected date using the two population model for examine the heterogeneous distribution of total PAH
Years I-V data (TABLE 5) were calculated using the data contaminants in Gulf of Mexico oysters from coastal
from 0-80 for the original cumulative frequencv distri- and estuarine areas collected during the winters of
bution for popultion 1 and from 77.5-100% for the 1986-1990. The PAH concentrations exhibits a log-
original cumulative frequency distribution for popula- normal distribution with two major populations in the
tion 2 (Table 6). data for each year. The two populations were decon-
The summary statisitcs for the first five vears of voluted into probability density functions and sum-
measuring PAH conaminants in the Gulf of Mexico mary statistics for each population were calculated.
for NOAA's NS&T Mussel Watch Program (Table 5) The lower PAH concentrations are probably related to related to
show variation in the means for both populations. indi- chronic inputs. Many of these low PAH concentration
cating temporal change in the total Gulf of Mexico sites show little variability from year to year. support-
data and with the highest values found in Years III and ing the contention that the PAH contamination is on a
IV. The higher mean concentrations of PAHs in Years continual basis. The higher concentration PAHs are
III and IV and the lower abundance in Years 1. 11 and probably associated with local point-sources of PAH
V is a pattern which is probably related to large-scale contamination or spills. Most of the NO concentration
climatic factors such as the El Nino cycles (Philander. sites (>1000 ng g dry tissue) show largee variability
1989) which affects the precipitation regime (Wilson et from year to year supporting the contention that PAH
al.. 1992). Examination of the PAH data for individual contamination for these sites is on an episodic basis. In
sites. as discussed above does not show this pattern. addition 20% of Gulf of Mexico sites Year III were
The cumulative frequency data for Years I-V gives exposed to a PAH threshold concentration of greater
the percentage of sites whose PAH concentration is less than 1000 ng/g of dry oyster tissue. Whereas. in Years I
than a particular concentration (Table 6). As an exam- and 11 only II of the Gulf of Mexico sites had
ple. using 1000 ppb as an arbitrary concentration. 89% concentrations greater than 1000 ng g of total NS&T
of the sites for Years I and 11 are less than this concen- PAHs. The changes in the mean concentration of the
two populations between years display a cyclic patte
which is probably due to large scale climatic factors
Table 6. NS&T concentration distribution data (cumulative
frequency). Corrected tPAH data-n/g dry weight such as the El Nino cycles which affects the precipita-
tion regime (Wilson et at..1992). The cyclic pattern
1990 1989 1988 1987 1986 was obtained by examining the Gulf of Mexico as a
Year V Year IV Year III Year 11 year I single heterogenous system. since the PAH concentra-
tion data for individual sites does not clearly show this
10% 110 171 110 110 110 pattern.
20% 140 200 153 140 140 226 206 162 169 140
30% 164 249 259 186 197
40% 212 352 345 208 229
50% 270 435 445 258 286 ACKNOWLEDGEMENTS
60% 3188 539 832 370 378
70% 397 510 1030 480 557 Funding for this research was supported by the
80% 597 869 2090 1300 1180 National Oceanic and Atmospheric Administration.
90% 1020 1440 contract number 50-DGNC-5-00262 (National Status
and Trends Mussel Watch Program),through the Texas
A & M Reseach Foundation. Texas A& M University.

298 T J. Jackson et aL

REFERENCES ment and Tissues. A Special NOAA 20th Anniversary
Report. 34 pp.
Conover. W.J. & Iman. R. L.(1981). Rank transformations Philander. G. (1989). El Nino and La Nina. Amer scientist.
as a bridge between parametric and nonparametric statis- 77,451-9.
tics. The Amer Statistician, 35. 124-9. Sericano, J. L., Wade. T. L.. Atlas, E. L. & Brooks. J. M.
Farrington, J. W. & Quinn. J. G. (1973). Petroleum hydro- (1990). Historical perspective on the environmental
carbons in Narragansett Bay. 1. Survey of hydrocarbons in bioavailability of DDT and its derivatives to Gulf of
sediments and clams (Mercenaria mercenaria). Estuar. and Mexico oysters. Environ. Sci. Technol., 77. 154-8.
Coast. Afar. Sci.. 1. 71-9. Wade, T. L.. Atlas, E. L., Brooks. J. M., Kennicutt 11. M. C.,
Farrington. J.V.. Goldberg, E. D., Risebrough, R. W., Fox, R. G., Sericano. J. L.. Garcia-Romero. B. & Defireitas,
Martin. J. H. & Bowen. V. T. (1983). US Mussel Watch D. A. (1988). NOAA Gulf of Mexico Status and Trends
1976-1978: An overview of the trace metal, DDE. PCB. Program: Trace organic contaminant distribution in sedi-
hydrocarbon and artificial radionuclide data. Environ. Sci. ments and ovsters. Estuaries. 11, 171-9.
Technol.17. 490-6. Wade. T. L. & Sericano. J. L. (1989). Trends in organic
Hensel. D. R. (1990). Less than obvious. Statistical treatment contaminant distribution in oysters from the Gulf of
of the data below the detection limit. Environ, Sci. Tech- Mexico. Oceans '89 Proceedings. Marine Technology
nol. 24. 1766-74. Society. IEEE Publication Number 89CH2780-5. pp. 585-9.
Krahn. M. M. Moore. L. K.. Bogar. R. G. Wigren. C. A.. Wade. T. L.. Sericano. J. L.. Garcia-Romero. B.. Brooks.
Chan. S-L. & Brown. D. W. (1988). High-performance J. M. & Presley. B. J. (1990). Gulf Coast NOAA National
liquid chromatography method for isolating organic con- Status & Trends Mussel Watch: The first four years. Proc.
taminants from tissue and sediment extracts. J. chromatogy.. mar. Tech. Soc.. 1990. 274-80.
437. 161-75. Wade, T. L., Brooks. J. M.. Kennicutt 11. M. C.. Denoux.
Mackay. D. & Paterson. S. (1984). Spatial concentration G. J. & Jackson. T. J. (1991). Oysters as bimonitors of
distributions environ. Sci. Technol., 18.207A-1 4A. oil in the ocean. Proceedings of the Annual Offshore
MacLeod. W. D.. Brown. D. W.. Friedman, A.J. Burrows, Technology Conference, OTC 6529. pp.
D. G.. Maynes. 0.. Pearce. R. W.. Wigren. C. A. & Bogar. Wilkinson, D. L.. Brooks. J. M. & Fay.R. R. (1991). NOAA
R. W. (1985). Standard analytical procedures of the NOAA Status and Trends: Mussel Wtch Program-Field
National Analytical Facility 1985-1986. Extractable Toxic Sampling and Logistics Report-Year VI. GERG Technical
Organic Compounds. 2nd Ed. US Department of Commerce. Report 91-046. US department of commerce national
NOAA/NMFS NOAA Tech. Memo NMFS F/NWC-92. oceanic & atmospheric administration. ocean assessment
Division.
Strategic Assessment: Data Atlas. United States Depart- Wilson. E. A.. Powell. E. N.. Wade. T. L.. Taylor. R. J..
ment of Commerce. National Oceanic and Atmospheric Presley. B. J. & Brooks, J. M. (1992)Spatial and temporal
Administration. pp. 4.0-5.32. distributions of bodv burden and disease in the Gulf of
O'Connor. T. P. (1990). Coastal Environmental Quality in Mexico oyster populations: The role of local and large-
the United States. 1990. Chemical Contamination in Sedi- scale climatic controls. Helgol. Measurments. (in press).

1-96

Reprint 10

Butyltin Concentrations in Oysters from the
Gulf of Mexico During 1989-1991

Bernardo Garcia-Romero, Terry L. Wade,
Gregory G. Salata and James M. Brooks

1-97

Environmental Pollution 81 (1993) 103-4 11

BUTYLTIN CONCENTRATIONS IN OYSTERS FROM
THE GULF OF MEXICO FROM 1989 TO 1991

Bernardo Garcia-Romero, Terry L. Wade,* Gregory G. Salata & James M. Brooks
Di@parlment of Oceanoyraphy, Texas A & M University, Geochemical and Environmental Research Group,
833 Graham Road, College Station, Texas 77845, USA

(Received 4 March 1992; accepted 2 June 1992)

Abstract 1991; Waite et al., 1991). In the USA, however, contin-
Oyster samples from 53 Gu?f qf Mexico coastal sites uous monitoring is needed in order to provide informa-
were collected and analyzed for butyltins during 1989, tion on the long-term response of butyltin concentrations
1990, and 1991. The geomeiric-mean tributylin concen- in the marine environment to these regulations.
trations were 85, 30, and 43 ng Sng for 1989, 1990, and Oysters are excellent sentinels of TBT contamina-
1991, respectively. 77te tributyltin concentrations are best tion. Bivalves have been used in uptake and depuration
represented by a log-normal disribution. A decline in the studies (Laughlin et al., 1986; Langston & Burt, 1991;
butyltin concentrations at sites with relatively low Sericano ei al. (in press); Alzieu ei al., 1991; Ritsema
butyltin concentrations for 1989 compared with 1990 and et al... 1991; Salazar & Salazar, 1991) and to determine
1991 was observed, and, at relarively high butyltin con- temporal and spatial variations of butyltin concentra-
centrations (>400 ng Sn1g), there was hardly any differ- tions (Short & Sharp, 1989; Wade ey al., 1988; Page &
ence between 1989 and 1991, but lower concentrations Widdows, 1991). These studies indicated that oysters
were present in 1990. Continued monitoring is needed in integrate bioavailable TBT with equilibration rates in
order to determine if butyltin contamination of the the order of weeks. This indicates that continuous and
coastal marine environment is decreasing in response to carefully planned sampling should be carried out in
use limitations. order to deterrnine trends in the variation of TBT
concentration in the environment.
LNITRODUCTION Tributyltin and its degradation products were deter-
mined in oysters from 53 sites in the Gulf of Mexico
The presence of tributyltin and its degradation products from 1989 to 1991. The over-all butyltin concentrations
in the environment continues to be of environmental showed a decline from 1989 to 1990 (Wade et a].,
concern. Tributyltin (TBT) ann-fouling paints are a 1991a.b). If this decline resulted from the implementa-
solution to the costly problem of fouling organisms tion of the limitations on the use of TBT in the USA by
that attach to the bottom of the hulls of boats and ships the Organotin Anti-Fouling Paint Control Act of 1988
(Huggett et aL, 1992). Although an effective anti-fouling (OAPCA). a continuous decline would be expected. The
agent, tributyltin, was found to a5ect non-target organ- results are now available for 1991. This report compares
isms adversely (Bushong et al.. 1987; Hall & Pinkney, three years of data for the Gulf of Mexico to determine
1985; Minchin et aL. 1987; Short & Thrower, 1986; if there is a trend in butyltin concentrations.
Thain, 1986, Thompson et al., 19S5; Alzieu, 1991). For
example, commercially valuable species were adversely
affected in France (Alzieu, 19911. The presence of TBT N1ETHODS
and its degradation products., dibutyltin (DBT) and
monobutyltin (MBT), in samples removed from input Oyster (Crassostrea virginica) samples were collected
sources (Wade et al., 1988; 1991b) suggests that envi- at 73 different sites along the Gulf of Mexico coast in
the winters of 1989, 1990, and 1991. Table I shows
ronmental half-lives in the marine environment may be the geographic location of the sites sampled and the
longer than reported values (Lee et al., 1987; Olson & symbols used to identify each site. Although knoAm
Brinckman, 1986; Seligman et aL. 1986ah, 1988). After point sources of TBT such as marinas or dry docks
the use of TBT-based pain@s was limited in countries were avoided, some locations are closer to such TBT
such as France, England, and the USA, the concentra- sources. A complete description of field sampling and
tion of organotins in water and oysters was shown to logistics has been reported (GERG, 1991).
decline (Short & Sharp, 1989, Wade et al., 1991b; The same sampling and analytical procedures were
Alzieu, 1991; Page & Widdows. 1991; Valkirs et aL, used for all oyster samples reported. A detailed descrip-
. To whom correspondence should be addressed. tion of these procedures has been previously reported
Environ. Pollut. 0269-7491/93/SO6.00 C 1993 Elsevier Science (Wade et al., 1988; Wade & Garcia-Romero. 1989).
Publishers Lid, En,land, Printed in G-mat Britain 103 Brieft, oyster tissues were homogenized, weighed,
1-99

Table 1. Sampling locations and site designators

Designation Site Location Latitude Longitude
(dog) (-in) (deg) i.min)

TEXAS
LMSB Sou,.h Rav Lower Laguna Madre 26 02-58 97 10-49
LMAC' Arro%,- 61orado Laguna Madre 26 16-80 97 1730
CCBR* Boa-. Harbor Corpus Christi 27 50-00 97 2@-00
CCNIV Nuo.---@i Bay Corpus Christi 27 51-70 97 21-00
CCIC Ingimie Cove Corpus Christi 27 50-30 97 14-25
ABLR Long Reef Aransas Bay 28 03-30 % 57-50
CDCRO Copan, Reef Copano Bay 28 08-20 97 07-58
MBAR A)T-- Reef Mesquite Bay 28 10-30 % 49-70
SAPP` Panthz.- Pt. Reef San Antonio Bay 28 13-20 % 43-00
SAMPO Moscuito Point San Antonio Bay 28 19-00 % 42-20
ESSPO South" Pass Reef Espiritu Santo Bay 28 17-83 % 37-50
ESBDP Bill Dz%s Reef Espiritu Santo Bay 28 25-00 % 27-00
MBGPO G&U-- p'per Pt. Matagorda Bay 28 35-00 % -14-00
MBLR Lava: River Mouth Matagorda Bay 28 39-30 96 35-00
MBCB4 Caranzabua Day Matagorda Bay 28 40-00 96 23-20
MBTP Tres Pz:acios Bay Matagorda Bay 28 39-00 96 15-50
MBEM Eas: Matagord' Matagorda Bay 28 42-30 95 511-00
BRCL* Ccd@- Lakes Brazos River 28 51-50 95 27-90
BRFS Frernor, River Brazos River 28 55,00 95 20 50
GBCR Con3::1Z'-ratc Reef Galveston Bay 29 15-75 94 50-50
GBOB Offa-:@ Bayou Galveston Bay 29 16-70 94 50-70
GBTD Todd% Dump Galveston Bav 29 30-10 94 54-00
GBYC Yach: Club Galveston Ba@ 29 37-00 94. 59-50
GBSC4 Ship Channel Galveston Bay 29 42-50 94 59-50
GBHR Hanrz Reef Galveston Bay 29 29-50 94 42 50
SLBB Blue B---zk Point Sabine Lake 29 48-00 94 .14-42
LOUISIANA
CLS] St. Jo@- Island Calcasieu Lake 29 50-00 93 32-00
CLLC Lake Charles Calcasieu Lake 30 03-50 93 1750
JHJH Joseph Harbor Bayou Joseph Harbor Bayou 29 37-75 92 45-75
VBSP Southu-st Pass Vermillion Bay 29 34-70 92 04-00
ABOB Oysic.- Bayou Atchafalaya Bay 29 13-00 91 08-00
CLCL Ciillo-- Lake Caillou L@k 29 15-25 90 55 50
TBLB Lake Ba-re Terrebonne Bay 29 15-00 90 16-00
TBLF Lake Ftlicity Terrebonne Bay 29 16-00 90 2450
BBSD Bavo-.; S,.. Denis Barataria Bav 29 24-10 89 ;9180
BBMB M;ddiz Bank Barataria Bay 29 17-20 89 56-60
MRT? Tile. P--5s Mississippi River 29 08-69 89 25-67
MRPV Pass a Loutre Mississippi River 29 04-30 89 04 60
BSSI Sabie as'and Breton Sound 29 24-70 89 28-70
BSBG Ba, G:-derne Breton Sound .19 35-87 89 38 50
LBMP Milhe-.:7eux Point Lake Borgne 29 52-30 89 4070
LPGOO Gull a-let Lake Ponchartrain 30 02-20 89 01-00
MISSISSIPPI
MSPC Pass C@%-nstian Mississippi Sound 30 19-75 89 1958
MSBB Bilm i.". Mississippi Sound 30 23-38 88 15-42
MSPR Pasmzo@iia Bav Mississippi Sound 30 21-05 88 17-00
ALABAMA
MBCP Ceda- Point Reef Mobile BaN 30 19-40 88 07 , .30
MBHJ Har6,-r Island Mobile Ba% 30 33-59 88 02-80
MBDRO Dog R--%-er Mobile BaN 30 35-50 88 0272
FLORIDA
PBPH Public Harbor Pensacola Bay 30 M-80 87 11-50
PRIBO Indian Bayou Pensacola BaY 30 30-83 87 04-00
PBS1319 Sabint Point Pensacola Ba% 30 20-80 87 OS-10
CBJB Joes BZ%ou Choctawhatc6e Bay 30 24-70 86 29-55
CBSP Shirk Point Choctawhatchee Bay 30 28-95 86 2S-60
CBSR Off Saw-c ia Rosa' Choctawhatch" Bay 30 23 50 86 10-60
PCLO Little 0,-.ster Ba)- Panama City 30 15-00 85 40-87
PCMPO Munmnaj Pier Panama City 30 08-20 85 37-50
SAWB Watson' Bavou St. Andrew Bav 30 0-850 85 5 8
APDB Drv Bar Apalachicola liay 29 41-50 85 05-00
APCP Cai Point Bar Apalachicola Bay 29 43-00 84 52-50
AESP Spring Creek Apalachee Bay 30 30-30 94 19 38
CKBP Black Point Cedar Key 29 10-25 83 03-00
TBNP Navar:z Park Tampa Bay 27 48-30 82 4528
TBMK Mullet Key Bayou Tampa Bay 27 37-17 82 43-62
TBPB Papys Bay-pu Tampa Bay 27 50-72 82 36-75
TBOT Old Tampa Bay Tampa Bay 28 01-48 92 37-95
TBKA@ K. Airport Tampa Bay 27 54-46 82 27-29
TRCB Cockroach Bay Tampa Bay 28 40-55 82 30-56
CBB1 Bird Island Charlotte Harbor 26 31@00 82 02-60
CBFM1 Fort %levers Charlotte Harbor 26 38-64 81 52-48
NBNB Nap)-- Ba% Naples Bay 26 00-00 81 32-00
RBHC' Hend-son Crock Rookerv gav 26 01-83 81 43-75
EVFU Faka 1:nion Bay Everglades 25 54-27 81 30-60
BHKF'
Sites that were not sampled coniccutively from 1989 to 1991. 1-100

Butyltin concentrations in oystersfrom the Gulf of Mexico 105

spiked will a urro,ale standard, extracted will 0*1% 14,5 � 5,11 for DBT; and 11*1 1*5 for MBT, Method-
tropolone in methylene chloride, hexylated, purified by detection limit (MDL) on average was 5 ng Sn/g for
using Si/Al columns, and analyzed by gas chromatog- TBT and DBT and 10 ng Sn/g for MBT.
raphy with a tin-selective-flame photometric detector.
Quality control consisted in using duplicate samples, RESULT'S AND DISCUSSION
procedural blanks, and spike blanks. Quadruplicate
analysis of one sample yielded the following means Annual variation of butyltins at individual sites
and standard deviations 395 14.5 ng Sn/g for TBT; Oyster butyltin concentrations determined in 1989,

Tributyltin (ng Sn/g)
1-6-19-89 13 1990 1991

low -

600-

6DO M
C!
400. L6

0-
Co 0. to Cr
S

< cl)
_j

1000,

00.
_j
:00

400 _j

2W
URI Rift
1B R FS
1@0u. RTC. eV5 2rL0

IWO,

6W X
LU

400

2DO
0 Wo
U Cc IL Q- IL 0 CL X
8 5? 2 g . j i.

SITE
Fig. 1. Geographical distribution of tributyltin concentrations in oysters (Crassostrea virginica) from tl@e Gulf of Mexico coast.
Asterisks indicate those sites which were not sampled in consecutive years.

1-101

0 M.

0
1< 0

S; 0 9 8
0
LMSB
I CLSJ Pop"
.0 0 Imc
CLLC
CCAfl
0 0
l3 M
0) 0 TCNII ClUll
;@ L:11: ccic vosp CBSP
53 ;; r- c r,
V C3 R = A13LR ABW MR
9. . 29."
10 V),7 0 -4 -, a -cl3cR CLCL PCLO
p - m C, m maAn 7BLO
:: cl,
to C3 le 'SAPP
Q. 8 0 co tr Er T13LF SAW13
t, >( C) '$AMP APDO
q M W 0 to
0 813SD
APCP
ts BOMB
QQ 0 TSBD AESP
0 -0 MBGP MTP CKOP
MBLR TBNP

OQ tj m P) assf TBMK
(v (m A t3 ts
00 t3 kA M m (a. 04 WTP
V, 5-S, BM TBPO
z m co MEM 7BDT
91 8 2' L13MP
%0 go
- T13KA
tr U 0 enFs IPGO LA TOM
. .......... .......... . ............. ..... . .
MSPC
W cr 0 qQ GMA Cool
tl r- CD
CL 6130B MOO CBrM
00 0 p) W GOTO NBNB
:3 ts " MMSPO ms
an, - 0" 0
t3 Gaye
q 5 0 M13CP
0 W
ts =Ml EVFU
R w E; M8111
GOIAR
'MoDn :3
0 SLOB AL
\0 - I I
0. co 00 0
0
TEXAS LA-MS-AL

>
CL to

P) CL 00
M 0 \0
ti a. s
\0 0
-tI

Mo
80 C.
o
z;;
0
-.z
no
o o
0,

o r) to
@4 0. 0 0
0
Od :j 0 0 0
0-1 0 LmS8
cr o Z tog. @ M . I
tj a 9 IMAC CLSJ PBPH
Co 0 CLLC 'Pain
FJ L, 0 Iccall
0 C)* M
2, tr 0 'CCND 22"
(V cr -.1
C: ccoc
tr al - cr vesp CBJB
1< 0 CBSP
j ABLR ABOR
CL CRSR
0 or TFRICR
tz CLCL
q PCLO
,a o ::r o- b MBAR
UQ P, :@ TOM
'SAPP
C) T13LF SAVVB
Er ts' 8 'SAMP
ft 0 :e t3 APOO
:r A R - 'ESSP ()BSD
0 to APCP
@t 7, -ES81) BBMB
cob 0 tj
W) w M AESP
0 -MBGP MRTP
:J MBLR CKBP
TBNp w
H o 'MI3CB
w 10 :3. 0 Cx assl law
El morp
tz o MBFM OSBO TOP13
,j Lamp
o E-r TOOT
R. TIFIKA
o 0 C, Q BRFS
0
9= LA TWA
kO no E@ CISCR MSPC
0 g C891
0 -1 w
00 M 0 < GOOB
0 tj H MSBB CBFM
0. G13TO
MSPB N13NB
0 GOYC M s
115 ........ . . ....... . . .. ....................................
j wer-P
t, .013sc
rb :3 to - M
EVFU
GE"n M13MI
gi 'R , 'fVfKr
sl nil DII
f% AL
to ,
0 TEXAS
LA-MS-AL
0 0 r*
a- M Lr to oo
go., co:. 0",o
-q 0,0 IP
00 t3
CL

C3
....................................

108 B. Garcia-Romero, T L. Wade, G. G. Salata, J. M. Brooks

Of a dynamic equilibrium between uptake, metabolism, in 1989, 1990, and 1991, respectively. Generally, sites
and depuration. with high TBT concentrations had high MBT concen-
The TBT concentrations determined for each site trations. MBT was detected in 21, four, and nineteen of
during 1989, 1990, and 1991 are shown in Fig. 1. Sites the 53 sites during 1989, 1990, and 1991, respectively.
are shown in geographical order from Texas to During all three years, MBT was detected at only three
Florida. Tributyltin concentrations ranged from <5 ng sites in Florida (CBJB, TBKA and CBFM) and at one
Sn/g to 1450 (TBKA), 770 (BBMB), and 1160 ng Sn/g site in Texas (CCIC). The fact that MBT was found in
(BBMB) in 1989, 1990, and 1991, respectively. TBT lower concentrations than DBT and DBT was found in
concentrations increase monotonically at some sites lower concentrations than TBT is consistent with the
from 1989 to 1991, whereas, at other sites, concentra- fact that TBT is the major constituent of anti-fouling
tions decreased monotonically. For example, oyster paints, while DBT and MBT are environmental-break-
TBT concentrations increased from 1989 to 1991 at down products of TBT. This may indicate that only a
CLLC, BBMB, and GBTD (Fig. 1). Decreasing TBT limited degradation of TBT has occurred or that the
concentrations from 1989 to 1991 were observed for more water-soluble DBT and MBT were assimilated by
oysters from PBPH, SAWB, TBCB, MBLR, and MBEM the oysters at a slower rate than TBT.
(Fig. 1). Concentrations of TBT were the same at
TBOT and GBCR during all three years. In general, Annual variation of butyltins in the Gulf of Mexico
higher concentrations of TBT were determined in A graphic representation of the TBT data for the 53
Florida sites than in Texas, Louisiana, Mississippi. or sites sampled in 1989, 1990, and 1991 is shown in
Alabama sites. TBT was below the detection limit Fig. 4. The graph is a plot of 1989 concentrations
at one of 53 sites in 1989 and at ten and eleven against those of 1990 and 1991. The x and Y scales are
sites during 1990 and 1991, respectively. Although the identical. If no change occurs in the TBT concentration
concentrations were low, butyltins were detected in at a site, those data will be plotted on the center line.
oysters from every site sampledin at least one sampling Sites that fall below the line show a decrease, whereas
year. points that rise above the line show an increase com-
Dibutyltin concentrations determined in oysters during pared with 1989. Two other lines also appear in Fig. 4.
1989, 1990, and 1991 are shown in Fig. 2. Dibutyltin These are the fines that form the boundary of sites with
concentrations ranged from <5 ng Sn/g to 380 (TBKA)@ a factor of two increase (top line) or decrease (bottom
160 (TBKA), and 200 ng Sn/g (TBKA), in 1989, 1990, line). Only six sites for 1990 and eight for 1991 of the
and 1991, respectively. Sites sampled in Florida had the 53 sites plotted for each year are above the center line.
highest DBT concentrations. With the exception of five Hence, over 85% of the TBT concentrations in 1990
sites (CBJB. TBKA. CBFM, BBMB, and BRFS), the and 1991 were less than the concentration measured in
annual variation of DBT concentrations did not mimic oysters, at that site in 1989. There were 30 sites (57%) in
the annual variation of TBT concentrations. Ship and 1990 and 20 sites (38%) in 1991 that had decreases, of
boating activities have been cited as potential factors more than a factor of two. There was only one site that
that may affect DBT fluctuations (Short and Sharp, had an increase of TBT concentration of more than a
1989; Uhler et aL, 1989). Furthermore, the commercial factor of two.
usage of DBT as a stabilizer for plastics, including In order to detect temporal trends, the butyltin
PVC pipes, may be another important source of input oyster concentrations for the entire Gulf of Mexico
to the marine environment and may result in DBT from 1989 to 1991 are compared. Annual variations of
fluctuations that do not mimic TBT fluctuations (Fent butyltins for the entire Gulf of Mexico are not readily
et aL, 1991: Maguire, 1991). At this point, it is not apparent in Figs. 1, 2, or 3, where only annual concen-
possible to estimate the influence of the factors dis- trations at'individual sites are compared. Comparisons
cussed above on the.DBT concentrations present in the of arithmetic mean, geometric mean.. and medians
oysters. Monotonic increases or decreases of DBT were (Table 2) for butyltin concentrations determined during
observed at specific sites during the three-year period. 1989, 1990, and 1991 are based only on the 53 sites that
For example, Middle Bank (BBMB, Figs I and 2) not were sampled in all three years. All these parameters
only showed. increasing concentrations of TBT during were calculated by assigning 5 ng Sn/g to all those
the three-year sampling period but also showed a samples with concentrations below the limit of detec-
steady increase in DBT in the sample period. DBT was tion. The percentage of samples below the detection
detected in 39, 38, and 33 out of the 53 sites sampled in limit is listed in Table 2. The median and geometric
each of the three years. In many instances, DBT was means are similar in all cases, whereas the arithmetic
not detected in any of the sampling years. mean is always higher. The median or the geometric
Regional MBT concentrations are shown in Fig, 3. mean appears to be the better estimator of the central
Since the MBT concentrations are low, annual varia- tendency of the data. On the basis of the median or the
tions in MBT concentrations for each site are large. geometric mean, there was a decrease in TBT oyster
The precision of MBT determination is also not as concentrations when 1989 is compared with 1990 or
good as that of TBT and DBT (Wade et aL, 1988). 1991.
Monobut-0tin concentrations ranged from <5 ng Sn/ A complete view of butyltin' concentrations for the
9
to 145 (NBNB), 25 (CCIC), and 42 ng Sn/g (TBKA). whole Gulf of Mexico for a given year can be achieved

1-104

Butyllin concentrations in oystersfrom the Gutf of Mexico 109

10000-.

GULY COAST OYSTERS

1000".

100

10
M 0 91: Wom

1 10 100 low 10"

TBT (ng Sn/g) 1989

Fig. 4. Tributyltin concentrations determined in 1989 plotted against the tributyltin concentrations determined in 1990 and 1991.
Points falling along the oenter line have equal concentrations, colateral lines indicate a factor of two greater or less than the con-
centrations determined in 1989.

by using either cumulative-percentage-distribution or (Fig. 5) are Gaussian with some degree of skewness.
probability-distribution curves (Mackay & Paterson. DBT had a log-normal distribution only in 1989 and
1984; O'Connor & Ehler, 1990; Jackson et at, in the 1990, whereas MBT does not follow a log-normal dis-
press). Although both types of curve may describe a tribution for any of the years. The geometric mean
distribution of butyltin concentration for each year, concentrations are indicated by solid lines for 1989,
probability-distribution curves were chosen because dotted lines for 1990, and dashed lines for 1991 (Fig. 5)
they are more easily compared, Use of this type of and are also reported in Table 2a, The TBT, DBT, and
curve assumes that the log of the concentration pro- MBT concentrations for � I standard deviation from
duces a normal distribution. Log-normal distributions the geometric mean are listed in Table 2b. Probability-
have already been reported for environmental data distribution curves of TBT in oysters from the Gulf
obtained in the NOAA National Status and Trends of Mexico provide information about annual variations
Mussel Watch Program (O'Connor & Ehler, 1990; at low, medium, and high ranges of concentration.
Jackson er at, in the press). Although the standard deviations quantify the spread
TBT log (distribution) curves are shown in Fig. 5 for of a data set, they provide no information about how
1989, 1990, and 1991. These curves were obtained by low or high concentrations changed with time. The
using the following equation (Milton & Arnold, 1986): TBT concentration decreased from 1989 to 1990 at all
f(x) = js[4(27r)])-' exp - 10.5[(x - A)lsf) (1) concentrations, whereas it decreased from 1989 to 1991
at low and medium concentrations but was similar for
where f(x) is the distribution probability of the log the two years at high concentrations (Fig. 5). This de-
(butyltin concentration), s is the standard deviation, x crease may be the result of the TBT regulation of 1988
is the log of the butyltin concentration, and X is the and/or development and use of lower-rciease-rate TBT
geometrical mean. Each f(x) was then divided by the paint formulations. Initial TBT regulations probably
sum of the f(x) values shown by eqn (2): resulted in a marked reduction in private boat owners
painting their own vessels. The facts that newer TBT-
r(x)i = fwily. RX)i (2)
containing pamts are rated to be good for up to five
in such a way that years and that TBT was not banned but its use limited
probably lead to decreased TBT inputs. This may have
Nx)i (3) resulted in the observed decreases in TBT concentra-
TBT-concentration curves from 1989, 1990, and 1991 tions in 1990 and 1991. The decrease observed at high
1991

1-105

110 B. Garcia-Romero, T L. Wade, G. G. Salata. J. M. Brooks

0.03- -0-- 1989
--(>,- 1990
A : : I
'dr 1991

AA
0.02-

0.01-

0.00 H
1 10 100 1000 10wo
TBT (ng Sn/g)
Fig. 5. Log-normal distribution of tributyltin concentrations determined in oysters in 1989, 1990, and 1991.

concentrations from 1989 to 1990 but not in 1991 may chanees in TBT-based paint formulations, but the
be due to the naturally higher variation of TBT con- effects are not as apparent at sites with high TBT
centrations near input areas (Seligman et aL, 1988). concentrations. Distribution curves for DBT and
TBT lower-concentration ranges may therefore have MBT concentrations did not follow a log-normal
decreased as a consequence of TBT regulations or distribution but also showed annual variations. This
Table 2a. Arithmetic and geometric means and medians may be due to the high percentage of values below
(ng Sat), the MDL (Table 2).
TBT DBT MBT CONCLUSION
1989 Oysters are valuable biomonitors for butyltins. The
Arithmetic mean 176 32 13 percentage of TBT present with respect to the total
Geometric mean 85 14 8
Median 77(2%) 12(26%) 5(60-/.) butyltins oscillated around 85% during the three years
1990 sampled. There was a decrease in the butyltin concen-
Arithmetic mean 96 17 6 tration from 1989 to 1990 or 1991. However, at high
Geometric mean 30 8 6 concentrations. there was little difference between 1989
Median 24(17%) 5 (72%) 5(90-1.)
1991 and 1991. Environmental response to the TBT regula-
Arithmetic mean 150 25 8 tion in 1988 is not yet apparent. The decline between
Geometric mean 43 13 6 1989 and 1990 or 1991 may have resulted from pre-
Median 42(17%) 8(40-/.) 5(66%) vious chanLyes in anti-fouling paint formulation with
Numbers in parenthesis indicate percentage of samples lower TBT-release rates or the suspension of painting
below MDL. activities by individual boat owners after 1988. Because
the newer TBT paints were formulated to last five years
Table 2b. Geometric mean �1 standard deTiation of the log or more, there are many boats still in use that were
(butyltin concentrations) (ng Sn/g) painted with TBT-containing paints before the ban.
TBT DBT MBT Consequently. continuous monitoring is necessary to
determine trends in butyltin contamination of the
1989 marine environment.
Plus 29i 44 18
Minus 25 5 3
1990 ACKINOWLEDGENIENT
Plus 141 21 8
Minus 6 3 4 This research was supported by the National Oceanic
1991 and Atmospheric Administration Grant No. 50-DGNC-
Plus 233 37 10 5-00262 (National Status and Trends Mussel Watch
Minus 8 4 4 Program).

1-106

Buyltin concentrations in oyster from the Gulf of Mexico

REFERENCES ings of the Oceans '86 Organotin Symposium, Vol. 4, pp.
1196-201.
Alziu, C. (1991). Environmental problems caused by TBT in Page, D. S. & Widdows, J. (1991). Temporal and spatial vari-
France: assessment, regulation, prospects. Mar. Environ. ation in levels of alkyltins in mussel tissues: a toxicological
Res., 32, 7-17. interpretation of field data. Mar. Environ. Res.. 32, 113-29.
Alzieu, C. Michel, P., Tolosa, L, Bacci, E., Mee, L. D. & Ritscma, R., Laane, R. W. P. M. & Donard, 0. F.X. (1991).
Readman, J. W. (1991). Organotin compounds in the Butyltins in marine waters of the Netherlands in 1988 and
Mediterranean: A continuing cause for concern. Mar. 1989: concentrations and effects. Mar. Environ. Res., 32,
ron. Res., 32, 261-70. 243-60.
Bushong, S. J., Hall, W. S., Johnson, W. E. & Hall, L. W., Salazar, M. H. & Salazar, S. M, (1991). Assessing site-specific
Jr. (1987). Toxicity of tributyltin to selected Chesapeake effects of TBT contamination with mussel growth rates.
Bay biota. In Proceedings of the Oceans '87 International Mar. Environ. Res.. 32, 131-50.
Organotin Symposium, Vol. 4, pp. 1499-503. Sericano, J. L., Wade, T. L., Garcia-Romero, B. & Brooks, J.
Clearly, J. J.(1991) Organotin in the marine surface micro. (Submitted) Environmental Accumulation and Depuration
layer and subsurface waters of Southwest England: relation of Tributyl tin by the American Oyster, Crassostrea
to toxicity thresholds and the UK environmental quality virginica. Mar. Biol., submitted for publication.
standard. Mar. Environ. Res., 32, 213-22. Seligman, P. F., Valkirs, A. 0. & Lee. R. F. (1986). Degra-
Fent, K.. Hunn. J., Renggli, D. & Siegrist. H. (1991). Fate of dation of tributyltin in San Diego Bay, California. waters.
tributyltin in sewage sludge treatment. Mar. Environ. Res., Environ. Sci.Technol., 20. 1229-35. '
32,223-31. Seligman, P. F., Valkirs. A. 0. & Lee, R. F. (1986a). Degra-
GERG (1991). NOAA Status and Trends, Mussel Watch dation, of tributyltin in marine and estuarine waters. In
Program. Field Sampling and logistics. Year VI. The Proceedings of the Oceans '86 Organoiin Symposium. Vol.
Geochemical and Environmental Research Group, Texas 4, pp. 1189-95.
A&M Research Foundation. Technical Report 91-046. US Seligman, P. F., Valkirs. A. 0., Stang, P. M. & Lee. R. F.
Department of Commerce, National Oceanic and Atmo- (1988). Evidence for rapid degradation of tributyltin in a
spheric Administration. Ocean Assessment Division. ma
Hall, L. W., Jr. & Pinkney, A. E. (1985). Acute and sublethal rine. Mar. Pollut. Bull., 19, 531-4.
effects of organotin compounds on aquatic biota: An inter- Short, J. W. & Thrower, F. P.(1986). Tri-n-butyltin caused
pretative literature evaluation, CRC Critical Reviews in mortality of chinook salmon, Oncorhynchus ishawyicha. on
Toxicology. 14, 159-209. transfer to TBT-treated marine net pen. In Proceedings of
Huggett. R.J., Unger, M. A., Seligman, P.F. & Valkirs. A. the Oceans '86 Organotin Symposium. Vol. 4, pp. 1202-5.
0. (1992). The marine biocide tributyltin. Assessing and Short, J. W. & Sharp, J. L. (1989). Tributyltin in bay mussels
managing the environmental risks. Environ. Sci. Technol.. (Mytilus edulic) of the Pacific Coast of the United States.
26. 232-7. Environ. Sci. Technol.. 23. 740-3.
Jackson. T. J.. Wade, T. L., McDonald, T. J., Wilkinson, D. Thain, J. E. (1986). Toxicity of TBT to bivalves: Effects on
L & Brooks. J.M. (Submitted) Polyaromatic hydrocarbon reproduction, growth and survival. In Proceedings of the
contaminants in National Status and Trends oysters from Oceans '86 Organozin Symposium, Vol. 4. pp. 1306-13.
the Gulf of Mexico (1986-1990). Oil Chem. Poll., sub- Thompson, J. A. J.. Sheffer. M. G.. Pierce. R. C.. Chau. Y.
mitted for publication. K., Cooney. J. J.. Cullen. W. R. & Maguire, R. J. (1985).
Langston, W. J. & Burt, G. R. (1991). Bioavailability and Organotin Compounds in the Aquatic Environment:
effects of sediment-bound TBT in deposit-feeding clams. Scientific Criteria for Assessing their Effects on Environmen-
Scrobicularia plana. Mar. Environ. Res.. 32. 61-77. tal Quality. National Research Council of Canada (NRCC
Laughlin. R. B.. French, W. & Guard. H. E. (1986). Accu- Associate Committee on Scientific Criteria for Environ-
mulation of bis(tributyltin) oxide by the marine mussel mental Quality).
mytilus edulis. Environ. Sci. Technol., 20. 884-90. Uhler, A. D..Coogan, T. H.. Davis, K. S., Durell. G. S..
Lee., R. F. (1985). Metabolism of tributyltin oxide by crabs, Steinhauer, W. G.. Freitas. S. Y. & Boehm. P. D. (1989).
oysters and fish. Mar. Environ. Res.. 17, 145-8. Findings of tributyltin in bivalves from selected U.S.
Lee, R. F.. Valkirs, A. 0. & Seligman, P. F. (1987). Fate of coastal waters. Environ. Toxicol. Chem.. 8. 971-9.
tributyltin in estuarine waters. In Proceedings of the Oceans Valkirs, A.0.. Davidson, B.. Kear. L. L.. Fransham. R. L..
'87 International Organotin Symposium, Vol. 4, pp. Grovhoug, J. G. & Seligman. P. F. (1991). Long-term
1411-15. monitoring of tributyltin in San Diego Bay California.
Lee, R. F. (1991). Metabolism of tributyltin by marine Mar. Environ. Res., 32, 151-67.
animals and possible linkages to effects. Mar. Environ. Wade, T. L. & Garcia-Romero. B. (1989). Status and trends
Res.. 32. 29-35. of tributyltin contamination of oysters and sediments from
Mackay. D. & Paterson, S. (1984). Spatial concentration dis- the Gulf of Mexico. In Proceedings of the Oceans '89
tributions. Environ. Sci. Technol.. 18. 207A-14A. Organotin Symposium, Vol. 2, pp. 550-3.
Maguire. R. J. (1991). Aquatic environmental aspects of non- Wade, T. L., Garcia-Romero, B. & Brooks. J. M. (1988).
pesicidal organotin compounds. Water Pollut. Res. J. Tributyltin contamination in bivalves from US coastal
Canada, 26, 243-360. estuaries, Environ. Sci. Technol.. 22, 1488-92.
Milton. J. S. & Arnold, (1986). Probabilit and Statistics Wade, T.L., Garcia-Romero, B. & Brooks. J. M. (1991).
in the Engineering and Computing Sciences. McGraw-Hill, Bioavailability of Butyltins. In Organic Geochemistry:
New York& Toronto. Advances and Applications in the Natural Environment,
Minchin, D., Duggan, C. B.,& King, W. (1987). Possible ed. D. A. C. Manning. Manchester University Press.
effects of organotins on scallop recruitment. Mar. Pollut. Manchester, U K, pp. 571-3.
Bull., 18, 604-8. Wade, T. L., Garcia-Romero, B. & Brooks, J. M. (1991b).
O'Connor. T. P. & Ehler, C. N. (1990). Results for the Oysters as biomonitors of butyltins in the Gulf of Mexico.
NOAA National Status and Trends Program on distribu- Mar. Environ. Res 32, 233-42.
tion and effects of chemical contamination in the coastal Waite, M. E., Waldock, M. J.. Thain. J. E., Smith, D. J. &
and estuarine United States. Environ. Monitor. Assessment, Milton. S. M. (1991). Reductions in TBT concentrations in
16, 1-17. UK estuaries following legislation in 1986 and 1987. Mar.
Olson, G. J. & Brinckman, F. E. (1986), Biodegradation of Environ.rs.. 32.89-111
tributyltin by Chesapeake Bay microorganisms. Proceed-

1-107

Reprint I I

The American Oyster (Crassostrea virginica)
as a Bioindicator of Trace Organic
Contamination

Jos6 L. Sericano

1-109

THE AMERICAN OYSTER (CRASSOSTREA VIRGINICA) AS A BIOINDICATOR

OF TRACE ORGANIC CONTAMINATION

A Dissertation

JOSE LUIS SERICANO

Submitted to the Office of Graduate Studies of
Texas A&M University
in partial fulfillment of the requirement for the degree of

DOCTOR OF PHILOSOPHY

May 103

Major Subject: Oceanography

THE AMERICAN OYSTER (CRASSOSTREA VIRGINICA) AS A BIOINDICATOR

OF TRACE ORGANIC CONTAMINATION

A Dissertation

JOSE LUIS SERICANO

Ap o as to style and content by:
I - - X, /,r,@ IL
james'M. Bi6oks Teriy L. Wade
hair of Committee) (Co-Chair of Committee)

15obby J. PreiWey
(W@ber) (Member)

Ste*n H. Safe Samily M. Ray
((Member) (Member)

Gilbert T. Rowe
(Head of Department)

May 1993

ABSTRACT

The American Oyster (Crassostrea virginica) as a Bioindicator of

Trace Organic Contamination. (May 1993)

Josd Luis Sericano,

Qufmico, Universidad Nacional del Sur, Argentina;

Lic. en Bioqufmica, Universidad Nacional del Sur, Argentina;

Lic. en Qui'mica, Universidad Nacional del Sur, Argentina;

M.S., Texas A&M University, U.S.A.

Co-Chairs of Advisory Comn-dttee: Dr. James M. Brooks
Dr. Terry L. Wade

This study was designed to examine the uptake and depuration of trace organic

contaminants, e.g. polynuclear aromatic hydrocarbons (PAHs), polychlorinated

biphenyls (PCBs), including planar congeners, and butyltin species, by oysters

(Crassostrea virginica) through transplantation experiments in Galveston Bay, Texas.

PAHs, low molecular weight PCBs and tributyltin (TBT) were rapidly

bioaccumulated by transplanted oysters and apparent equilibrium concentrations were

reached after 20 to 30 days of r@xposure. In contrast, high molecular weight PCBs did

not reached an equilibrium plateau at the end of the seven-week exposure period. When

oysters'were back-transplanted to their former location, PAHs, low molecular weight

PCB congeners and TBT were depurated at similar rates while the high molecular,
weight PCBs were depurated at considerably slower rates. The original background

concentrations were not reached after the 50-day depuration period. Chronically

contaminated Ship Channel oysters, simultaneously transplanted to the uncontaminated

area, showed lower clearance rates than those encountered for originally

uncontaminated bivalves.

Oysters exposed in the laboratory to PCBs and PAHs, preferentially bioaccumulated

four to six chlorine-substituted PCBs and four- and five-ring PAHs. Oysters exposed

simultaneously to PAHs plus PCBs depurated PAHs at a faster Tate than oysters that

were exposed solely to PAHs. The half-lives for individual PAHs encountered in the

first group of oysters were similar to those found in the field.

The present study presents evidence to substantiate the theory that bioconcentration

and clearance of different PCB congeners by oysters appear to be more affected by

molecular size than by hydrophobicity. The influence of the chlorine substitution

patterns in the bioaccumulation of PCBs by oysters is particularly evident in the case of

the highly toxic planar congeners.

Indigenous oysters can be valuable bioindicators of environmental contamination by

trace organic compounds paly if their limitations are fully understood. Within these

limitations, transplanted oysters can be succesfully used to monitor environmental

contamination by PAHs and TBTs in areas lacking indigenous bivalves if deployed in-

situ for a period of time of at least 30 days; for PCBs, however, a much longer time

period, Le over 6 months, may be required.

ACKNOWLEDGEMENTS

I would like to express my sincere gratitude to the co-chairmen of my Graduate

Committee Dr. James M. Brooks and Dr. Terry L. Wade for lending support and advice

during the preparation of this dissertation. I am also grateful to Dr. Eric N. Powell, Dr.

Bobby J. Presley, Dr. Stephen H. Safe and Dr. Sammy M. Ray for their assistance and
helpful suggestions as members of my Graduate Committee. I thank Dr. Don E.

Albrecht who served as the Graduate College Representative.

My deepest gratitude is expressed to Dr. Elliot L. Atlas who got me started in the

analysis of polychlorinated biphenyl by high-resolution gas chromatography. I also

acknowledge Dr. Roger R. Fay for allowing me to use the boats everytime that I needed

to do the samplings and sincerely appreciate Ken McCormick's generous help during

these activities. I thank Mrs. Amani M. El Husseini for helping me to develop the

carbon:silica column chromatographic technique used during this study. I am truly

indebted to Dr. Thomas J. McDonald and Mr. Bernardo Garcfa-Romero for the

analyses of polynuclear aromatic hydrocarbons and butyltins, respectively. Thanks also

goes to Dr. Thomas J. Jackson for helpful suggestions while preparing this manuscript.

Thanks to everyone at GERG for their friendship.

Finally, I would like to thank my wife, Ndlida, for her love, understanding and

constant support throughout my college career and apologize to my son, Mauro, and

daughter, Gisella, for the many weekends lost to this project.

TABLE OF CONTENTS

Page
ABSTRACT ................................................................................ iii

ACKNOWLEDGEMENTS ............................................................... v

TABLE OF CONTENTS .................................................................. vi

LIST OF TABLES ......................................................................... xi

LIST OF FIGURES ........................................................................ X111

CHAPTER

I INTRODUCTION ............................................................ I
Statement of Purpose .................................................... I
Research Objectives ...................................................... 5
Galveston Bay System ................................................... 7

11 BIOAVAILABILITY OF PAHs TO THE AMERICAN OYSTER
(CRASSOSTREA VIRGIMCA): AFIELD STUDY ................... 10

Introduction ............................................................... 10

PAHs: A Review ........................................................ I I
Background Information ............................................ I I
Distribution and Occurrence in Galveston Bay ................... 14
Bivalve Uptake and Depuration Studies ........................... 20
Uptake and Depuration of PAHs ....................................... 22
Experimental Design, Sample Collection and Methods ......... 22
Extraction and fractionation of PAHs ......................... 24
Instrumental analysis ............................................ 28
Ancillary parameters ............................................ 28
Statistical analysis ............................................... 29
Uptake of PAHs by Transplanted Oysters ........................ 29

vii

CHAPTER Page
Depuration of PAHs by Newly and Chronically
Contaminated Oysters ............................................... 39
Concluding Remarks ..................................................... 45

III UPTAKE, RETENTION AND RELEASE OF PCBs BY THE
AMERICAN OYSTER (CRASSOSTREA VIRGINICA ............... 48

Introduction ............................................................... 48

PCBs: A Review ......................................................... 49
Background Information ............................................ 49
Distribution and Occurrence in Galveston Bay ................... 53
Bivalve Uptake and Depuration Studies ........................... 56
Uptake and Depuration of PCBs ....................................... 58
Experimental Design, Sample Collection and Methods ......... 58
Extraction and sample fractionation of PCBs ................ 58
Instrumental analysis ............................................ 58
Ancillary parameters ............................................ 60
Statistical analysis ............................................... 60
Uptake of PCBs by Transplanted Oysters ........................ 60
Depuration of PCBs by Newly and Chronically
Contaminated Oysters ................................................ 70
Concluding Remarks ..................................................... 76

IV UPTAKE AND DEPURATION OF PLANAR PCB CONGENERS
BY THE AMERICAN OYSTER (CRASSOSTREA VIRGMCA):
A SPECIAL"CASE OF PCBs ............................................... 80

Introduction ............................................................... 80
Planar PCBs: A Review ................................................. 81
Background Information ............................................ 81
Distribution and Occurrence in Galveston Bay ................... 83
Bivalve Uptake and Depuration Studies ........................... 84
Uptake and Depuration of Planar PCBs ............................... 85

Viii

CHAPTER Page
Experimental Design, Sample Collection and Methods ......... 85
Extraction and initial sample fi-actionation .................... 85
Isolation of planar PCB congeners ............................ 85
Instrumental analysis ............................................ 86
Planar PCB Congener Analysis .................................... 86
Uptake of Planar PCBs by Transplanted Oysters ................ 90
Depuration of Planar PCBs by Newly Contaminated Oysters.. 95
Concluding Remarks ..................................................... 96

V UPTAKE AND DEPURATION OF TRIBUTYLTIN BY THE
AMERICAN OYSTER (CRASSOSTREA VIRGINICA ............... 97

Introduction ............................................................... 97

TBT: A Review .......................................................... 97
Background Inforniation ............................................ 97
Distribution and Occurrence in Galveston Bay ................... 99
Bivalve Uptake and Depuration Studies ........................... 99
Uptake and Depuration of TBT ......................................... 101
Experimental Design, Sample Collection and Methods ......... 101
Extraction and sample fractionation ........................... 101
Instrumental analysis ............................................ 102
Ancillary parameters ............................................. 102
Statistical analysis ............................................... 102
Uptake of TBT by Transplanted Oysters .......................... 102
Deepuration of TBT by Newly and Chronically
Contaminated Oysters ............................................... 105
Concluding Remarks ..................................................... 106

VI MECHANISM OF THE UPTAKE AND RELEASE OF TRACE

ORGANIC CONTAMINANTS BY THE AMERICAN OYSTER
(CRASSOSTREA VIRGINICA ........................................... 107

Introduction ............................................................... 107

CHAPTER Page
Mechanisms of Bioconcentration ....................................... 108

Kinetics ............................................................... 108
Polynuclear aromatic hydrocarbons ........................... Ill
Polychlorinated biphenyls ...................................... 116
Tributyltin ........................................................ 120
The Octanol-to-Water Partition Coefficient ....................... 122
The Two-Compartment Model Approach ......................... 125
Depuration versus Degradation ..................................... 127
Concluding Remarks ..................................................... 128

VII SIMULTANEOUS UPTAKE AND DEPURATION OF PAHs AND
PCBs BY THE AMERICAN OYSTER (CHASSOSTRE
YIRGINICA) .................................................................. 130

Introduction ........................ 130
Simultaneous Exposure to PAHs and PCBs: A Laboratory
Study .................................................................. 131
Aquarium Exposure ................................................. 131
Extraction, fractionation and instrumental analyses of
PAHs and PCBs ................................................. 135
Polynuclear Aromatic Hydrocarbons .............................. 135
Polychlorinated Biphenyls; .......................................... 141
PAH-PCB Interactions .............................................. 148
'Concluding Remarks ..................................................... 151

VIII NOAAS NATIONAL STATUS AND MENDS "MUSSEL
WATCH" PROGRAM ....................................................... 153

Introduction ............................................................... 153
Polynuclear Aromatic Hydrocarbons .................................. 155
Polychlorinated Biphenyls .............................................. 157
7,PCB Congeners/Total PCB Relationship ....................... 161
Planar PCB Congeners ............................................... 168

CHAPTER Page
Butyltin Species .......................................................... 177

IX SUMMARY AND PROSPECFIVES ...................................... 182

REFERENCES ............................................................................. 187

APPENDIX .................................................................................. 209

VITA ......................................................................................... 242

LIST OF TABLES

TABLE Page

1 Hydrocarbon Concentrations in Samples from the Galveston Bay Area.
Except Where Indicated, Concentrations in Organisms Are Expressed in
ng g- I on a Wet-Weight Basis. Concentrations in Sediment and Water
Samples Are Expressed in ng g-I, on a Dry-Weight Basis, and in ng 1-1,
Respectively. Ranges in Parenthesis ........................................... 15

2 Biological Half-Lives (Days) of PAHs in Hanna Reef and Ship Channel
Crassostrea virginica Oysters ..................................................... 46

3 Polychlorinated Biphenyl Concentrations in Samples from the Galveston.
Bay Area. Except Where Indicated, Concentrations in Organisms Are
Expressed in ng g- I on a Wet-Weight Basis. Concentrations in Sediment
and Water Samples Are Expressed in ng g- 1, on a Dry-Weight Basis, and
in ng 1-1, Respectively ............................................................ 54

4 Biological Half-Lives (Days) of PCBs in Hanna Reef and Ship Channel
Crassostrea virginica Oysters .................................................... 77

5 Recoveries of Four Planar PCB Congeners from Spiked Aroclor 1254
and Dolphin Blubber Samples Using Activated Carbon:Silica Columns ... 91

6 TBT, DBT and MBT Concentrations (in ng Sn g-.1 on a Dry-Weight
Basis) in Oyster and Sediment Samples from the Galveston Bay Area ..... 100

7 Estimated Days to 90% Uptake Equilibrium (t90%), Bioconcentration
Factors (BCF), Depuration Rates (kd) and Biological Half-Lives (BHL)
for PAHs and PCB Congeners 'in Hanna Reef and Ship Channel Oysters
During the Field Studies .......................................................... 112

xii

TABLE Page

8 Characteristics of the Relationships Between log Kb and log Kow for
Bioconcentration of Trace Organic Contaminants in Different Organisms . 126

9 Biological Half-Lives of PAHs in Crassostrea virginica Oysters Exposed
in the Laboratory to Particle-Associated PAHs; Alone (Aquarium C) and
PAHs + PCBs (Aquarium D) .................................................... 140

.10 Biological Half-Lives of PCBs in Crassostrea virginica Oysters Exposed
in the Laboratory to Particle-Associated PCBs Alone (Aquarium B) and
PCBs + PAHs (Aquarium D) .................................................... 149

11 PCB Congeners/Total PCB Relationships in Gulf of Mexico Oyster
Samples ............................................................................. 164

12 Planar and Total PCB Concentrations in Oysters (Crassostrea virginica)
from Galveston and Tampa Bays ................................................ 170

13 2,3,7,8-TCDD Equivalents (pg g-1) Corresponding to Non-Ortho
Substituted PCB in Oysters (Crassostrea virginica) from Galveston and
Tampa Bays ........................................................................ 172

14 Selected Mono- and Di-Ortho Substituted PCB and Total PCB Average
Concentrations (ng g-1) in Oysters (Crassostrea virginica) from
Galveston and Tampa Bays ...................................................... 174

15 Average 2,3,7,8-TCDD Equivalents (pg g-1) Corresponding to Selected
Mono- and Di-ortho Substituted PCBs in Oysters (Crassostrea virginica)
from Galveston and Tampa Bays ................................................ 175

xiii

LIST OF FIGURES

FIGURE Page

I Galveston Bay, Texas ............................................................ 8

2 Structures and common names of selected aromatic hydrocarbons
discussed in the text ............................................................... 12

3 Galveston Bay transplantation sites ............................................. 23

4 Exposure (a) and depuration (b) sites, respectively. Approximately 250
adult oysters (c) were transplanted in net bags (d) ............................ 25

5 Trace organic analytical scheme ................................................. 26

6 Concentrations of polynuclear aromatic hydrocarbons, grouped by
number of rings, in transplanted Hanna Reef and indigenous Ship
Channel oysters during the 48-day exposure period near the Houston
Ship Channel. Ship Channel oysters were not sampled on day 7 .......... 31

7 Concentrations of selected polynuclear aromatic hydrocarbons in tissues
of Hanna Reef and Ship Channel oysters during exposure to the Ship
Channel area contaminant levels and following transplant to the Hanna
Reef. area ............................................................................ 32

8 Concentrations of individual polynuclear aromatic hydrocarbons in
tissues of Hanna Reef and Ship Channel oysters at the end of the 48-day
.exposure period .................................................................... 36

9 Concentrations of polynuclear aromatic hydrocarbons, grouped by
number of rings and individually, in Ship Channel sediment sarnples ...... 37

xiv

FIGURE Page

10 Concentrations of polynuclear aromatic hydrocarbons, grouped by
number of rings and individually, in Ship Channel seawater samples ...... 38

11 Concentrations of polynuclear aromatic hydrocarbons, grouped by
number of rings, in back-transplanted Hanna Reef and transplanted Ship
Channel oysters during the 50-day depuration period in the Hanna Reef
area .................................................................................. 40

12 Concentrations of individual polynuclear aromatic hydrocarbons in
tissues of Hanna Reef and Ship Channel oysters at the end of the 50-day
depuration period .................................................................. 41

13 Comparison of the concentrations of polynuclear aromatic hydrocarbons,
grouped by number of rings and individually, in tissues of Hanna Reef
oysters before exposure to the Ship Channel contaminant levels and after
depuration at the Hanna Reef site ................................................ 42

14 Concentrations of polynuclear aromatic hydrocarbons, grouped by
number of rings and individually, in Hanna Reef sediment samples ........ 44

15 General formula of polychlorinated biphenyls and examples of major
congeners commonly found in environmental samples ....................... 50

16 Examples of high-resolution gas chrornatograms of Hanna Reef oysters
transplanted to the Ship Channel area during different stages of the 48-
day exposure period. PCB congeners are numbered according to
Bal1schmitter & Zell, 1980 ........................................................ 61

17 Concentrations of polychlorinated biphenyl congeners, grouped by level
of chlorination, in transplanted Hanna Reef and indigenous Ship Channel
oysters during the 48-day exposure period near the Houston Ship
Channel. Ship Channel Oysters were not sampled on day 7 ................ 62

FIGURE Page

18 Concentrations of selected polychlorinated biphenyl congeners in tissues
of Hanna Reef and Ship Channel oysters during exposure to the Ship
Channel area contaminant levels and following transplant to the Hanna
Reef area ............................................................................. 64

19 Concentrations of selected polychlorinated biphenyl congeners in tissues
of Hanna Reef and Ship Channel oysters at the end of the 48-day
exposure period .................................................................... 67

20 Percent differences in concentrations of selected polychlorinated
biphenyls between Hanna Reef and Ship Channel oyster tissues at the end
of the 48-day exposure period. Positive values indicate congeners with
greater accumulation in Ship Channel oysters ................................. 69

21 Concentrations of polychlorinated biphenyls, grouped by level of
chlorination, in Ship Channel sediment and seawater samples .............. 71

22 Concentrations of polychlorinated biphenyl congeners, grouped by level
of chlorination, in back-transplanted Hanna Reef and transplanted Ship
Channel oysters during the 50-day depuration period in the Hanna Reef
area .................................................................................. 73

23 Concentrations of selected polychlorinated biphenyl congeners in tissues
of Hanna Reef and Ship Channel oysters at the end of the 50-day
depuration period ... .............................................................. 74

24 Comparison of the concentrations of selected polychlorinated biphenyl
congeners measured in tissues of Hanna Reef oysters before-exposure to
the Ship Channel contaminant levels and after deputation at the Hanna
Reef site ............................................................................ 75

xvi

FIGURE Page

25 Concentrations of polychlorinated biphenyls, grouped by level of
chlorination, in Hanna Reef sediment samples ................................ 78

26 General formula of polychlorinated biphenyls. Three of the most toxic
planar PCB congeners, i.e. PCB 77, 126 and 169, are shown together
with the compounds they mimic in toxic effects ............................... 82

27 High-resolution gas chromatographic analyses of (a) Aroclor 1254 spiked
with planar PCB congeners (i.e., 77, 81, 126 and 169), (b) PCB
congeners recovered in Fraction 1, and (c) planar PCB congeners eluted
in Fraction 2. PCB congeners 103 and 198 are external standards. PCB
congeners are numbered according to Ballschrrdtter & Zell, 1980 .......... 88

28 High-resolution gas chromatographic analyses of (a) dolphin blubber
extract spiked with planar PCB congeners 77, 81, 126 and 169, (b)
chlorinated hydrocarbons recovered in Fraction 1, and (c) planar PCB
congeners eluted in Fraction 2. PCB congeners are numbered according
to Ballschmitter & Zell, 1980 .................................................... 89

29 Example of high-resolution gas chromatograms obtained from an extract
of indigenous Ship Channel oysters. PCB congeners are numbered
according to Ballschmitter & Zell, 1980 ........................................ 92

30 Concentrations of planar polychlorinated biphenyl congeners 77 and 126
in tissues of Hanna Reef oysters during the uptake and depuration phases
of the transplantation experiments at Galveston Bay .......................... 94

31 Concentrations of tributyltin in tissues of Hanna Reef and Ship Channel
Oysters during exposure to the Ship Channel area contaminant levels and
following transplant to the Hanna Reef area ................................... 104

xvii

or-

FIGURE Page

32 Depuration constant (kd) and biological lialf-lives (BHL) of planar PCB
congeners compared to ranges of values calculated for non-plartar PCBs.. 117

33 Bioconcentration factors of six selected PCB congeners in relation to their
liphophilicity and size ............................................................. 119

34 Relationship between chlorine-substitution patterns in PCBs and &@e_;r
depuration half-lives. See text for explanation ................................ 121

35 Bioconcentration factors of polynuclear aromatic hydro@arbons and
poIrchlorinated biphenyls calculated for transplanted Hanna Reef and
indigenous Ship Channel oysters during the exposure period versus log
octanol to water partition coe-fficients (I(o,,,) ................................... 123

36 General laboratory set-up 'a), details of an aquarium and water
recirculation system (b), dosing and feeding system (c), and oysters used
during the experiment (d) ......................................................... 133

37 Concentrations of selected polynuclear aromatic hydrocarbons in tissues
of oysters during exposure to particle-associated PAHs alone (Aquariurn
C) and PAHs + PCBs (Aquarium D) and following transplant to
contaminant-free aquariums. Aquarium A was used as control ............. 136

38 Concentrations of individual polynuclear aromatic hydrocarbons in
tissues of laborutory expo sed oysters after the 30-day exposure period to
particle-associated PAHs (Aquarium C) and PAHs + PCBs (Aquarium
D) ................................................................. * .................... 138

39 Concentrations of individual polynuclear aromatic hydrocarbons ill
tissues of oysters previously exposed in the laboratory to particle-
associated PAHs (Aquarium C) and PAHs + PCBs (Aquarium D) after
the 30-day depuration period in contaminant-free aquarium,,; ................ 139

xviii

FIGUREE Page

40 Concentrations of selected polychlorinated biphenyls in tissues of oysters
during exposure to particle-associated PCBs alone (Aquarium B) and
PCBs +* PAHs (Aqualium D) and following transplant to contaminant-
free aquariums. Aquarium A was used as control ............................ 142

41 Comparison of the distribution of PCB congeners, grouped by level of
chlorination, encountered in laboratory exposed oysters at the end of the
30-day exposure period with that of the exposure mixture (i.e. 1:1:1:1
Aroclor 1242, 1248, 1254 and 1260). Distribution of PCB congeners it]
individual Aroclor mixtures are also shown .................................... 14.."t

42 Selective accumulation or depletion of PCB congeners in laboratory
exposed oysters relative to the exposure Aroclor mixture .................... 145

43 Concentrations of selected PCB congeners in tissues of laboratory
exposed oysters after the 30-day exposure period to particle-associated
PCBs (Aquarium B) and PC13s + PAIIs (Aquarium D) ...................... 146

44 Concentrations of selected PCB congeners in tissues of oysters
previously exposed in the laboratory to particle-associated PCBs
(Aquarium B) and PCBs + PAHs (Aquarium D) after the 30-day
- depuration period in contaminant-free aquariums .............................. 147

45 NOAA' s National Status and Trends sampling locations in the Gulf of
Mexico .............................................................................. 154

46 Average distributions of PCB congeners grouped by level of
chlorination, in Gulf of Mexico sediments and oysters ....................... 158

47 Average d.istributions of selected PCB congeners in Gulf of Mexico
sedimens and oysters ............................................................... i60

xix

FIGURE Page

48 Relationships between the sum of 18 selected PCB congeners and the
total PCB load encountered in Gulf of Mexico oysters for the first year of
NOAA's National Status and Trends Program. See text for discussion ... 163

49 Three different examples of the bias introduced in the report of total PCB
concentrations by using the regression equation (see text) compared to the
total PCB load calculated as the sum of all measurable individual,
congeners ........................................................................... 167

50 NOAA's National Status and Trends sampling locations in Galveston and
Tampa Bays ........................................................................ 169

51 Toxic equivalents corresponding to three planar PCBs and selected mono-
and di-ortho chlorine-substituted congeners in oyster samples collected
from six different locations in Galveston and Tampa Bays .................. 176

52 Contribution of planar and selected mono-ortho chlorine substituted PCB
congeners to the total toxicity in oysters ........................................ 178

53 Total butiltin concentrations at selected sites in the Gulf of Mexico
sampled between 1986 and 1992 as part of NOAA's National Status and
Trends Program ................................................................... 180

CHAPTERI

INTRODUCTION

STATEMENT OF PURPOSE

Many toxic organic compounds of both synthetic and natural origin, such as

polychlorinated biphenyls (PCBs), polynuclear aromatic hydrocarbons (PAHs) and

butyltin compounds, e.g. tributyltin (TBT), can be present at high levels, i.e. ppm, in the

coastal marine environment (e.g. Kerkhoff er al., 1982; Malins er al., 1984, 1987; Wade

et al., 1988a) and may not only affect the productivity of marine organisms but may

ultimately be hazardous to human health.

PCBs and PAHs enter the marine environment from several sources including

precipitation, land runoff, atmospheric fallout, industrial and municipal was te discharge

and accidental spills (e.g. Hoffman et al., 1984; Prahl et al., 1984). PAHs are also

known to enter the marine environment from natural oil seepage (Venkatesan & Kaplan,

1982; Anderson et al., 1983; Kvenvolden & Harbough, 1983; Venkatesan et al., 1983).
The major source of tributyltin (TBT), the most toxic of the butyltin species (Davis &

Smith, 1980), to the marine environment is the use of antifouling paints containing this

compound.

7bis dissertation follows the format of the Marine Environmental Research journal.

Because of their low aqueous solubilities (e.g. Mackay et al., 1980; Whitehouse,

1984), PCBs, PAHs and butyltin compounds are rapidly adsorbed onto particulate matter,

which can result in their deposition with estuarine and coastal sediments (Herbes, 1977;

Pavlou & Dexter, 1979; Means et al., 1980; Langston et al., 1987). Sediments may serve

as a storage compartment for long-term release of contaminants by biogeochemical

processes (Sodergren & Larsson, 1982; Prahl & Carpenter, 1983; Coates & Elzerman,

1986; Unger et al., 19 87); however, the extent of residue accumulation in sediments is

largely determined by the chemical nature of the compounds and the sediment

characteristics, e.g. texture and organic matter content (Choi & Chen, 1976; Karickhoff et

al., 1979; Chiou et al., 1983; MacIntyre & Smith, 1984). Natural organic materials can

enhance the partition of hydrophobic compounds into the bottom sediments and pore
water (Brownawell & Farrington, 1985, 1986; Brownawell, 1986; Chin & Gschwend,

1992). This partition process influences the availability of PCBs, PAHs and butyltin

compounds to the overlying seawater and, in turn, to the aquatic organisms where these

compounds can accumulate by passive adsorption directly from water or by partitioning

into food. Both routes have been shown to contribute significantly to levels found in

fishes (Rubistein et al., 1984; Malins et al., 1987; Oliver & Niimi, 1983; Opperhuizen &

Schrap, 1988) and benthic organisms (Clement et al., 1980; Stekoll et al., 1980; Laughlin

er al., 1986; Salazar, 1986; Oliver, 1987).

A considerable body of knowledge exists on the dynamics of PCBs and PAHs uptake

and depuration in marine species (e.g., Stegeman & Teal, 1973; Lee, 1977; Clement et

al., 1980; Riley et al., 1981; Opperhuizen et al., 1985; Jovanovich & Marion, 1987;

Pruell et al., 1986, 1987; Tanabe et al., 1987a). Most of the previous studies, however,

described the steady-state bioconcentration factors of PCBs or PAHs by a variety of

marine organisms while only a few of them discussed the dynamics by which the final

levels are achieved. Until recently, investigations on PCBs and PAHs focused mainly on

the commercial Aroclor mixtures or whole petroleum (e.g. Stegeman & Teal, 1973;

Courtney & Denton, 1976; Shaw & Connell, 1982; Stickle et al., 1984) rather than on the

individual PCB congeners or specific PAHs (e.g. Duinker et al., 1983; Frank et al., 1986;

Pruell et al., 1986, 1987; Jovanovich & Marion, 1987; Tanabe et al., 1987a). Now, the

importance of considering individual compounds is well recognized in view of differences

in both toxicity (e.g. Safe, 1984, 1985, 1990; Tanabe et al., 1987b, 1987c) and

physicochemical properties controlling their assimilation by organisms (e.g. Shaw &

Connell, 1980, 1984; Opperhuizen er al., 1985, 1988). This is particularly important in

the case of planar PCB congeners (McFarland & Clarke, 1989). In a recent review on

PCBs, Oliver et al. (1989) described the current information on the behavior of individual

PCB congeners as limited; however, research on environmental pathways and hazard

assessments of specific congeners can be determined with the currently available analytical

procedures for PCBs. Similarly, Smith et al. (1990) stated that our understanding of the

hazards by PCBs to various animal species and ecosystems remains inadequate.

Moreover, little is known about the effects that a group of xenobiotic compounds, for

example PCBs, will have on the uptake and/or depuration dynamics of other

contaminants, for example PAHs. Interaction between PCBs and PAHs during uptake

has been reported to occur in fish and oysters (Stein et al., 1984; Collier et al., 1985;

Fortner & Sick, 1985).

Comparatively, the knowledge of the dynamics of butyltin compounds uptake and

depuration by different marine organisms is limited. Although contamination of the

coastal environment by TBT has been investigated since early 1980s, it was not until the

late 1980s that this compound was considered to be a real threat to the quality of coastal

waters.

Monitoring of PCBs, PAHs and butyltin compounds, at trace levels, in the aquatic

environment using various organisms is well-establi shed. Bivalves are generally

preferred for this purpose because of their wide geographic distribution, sedentary form

of life, ability to bioconcentrate organic and inorganic contaminants, comparatively low

enzyme activity for metabolizing xenobiotics, ability to survive under extreme pollution

conditions and commercial value (Goldberg et al., 1978; Bums & Smith, 1981;

Farrington et al., 1983). The use of bivalves as bioindicators has grown rapidly over the

last decade and the "Mussel Watch" concept is now being used by many national and

international programs (National Academy of Sciences, 1980; Farrington et al., 1983;

Risebrough et al., 1983; May-tin, 1985; Wade et al., 1988a, 1988b; Sericano et al., 1990a,

1990b, Tripp et al., 1992).

Laboratory experiments have been carried out in order to have a better understanding

of the uptake and depuration processes taking place in the environment; however,

extrapolations from laboratory tests to natural environmental conditions are not always

possible. For example, results of laboratory-based studies of PCB and PAH kinetics in

bivalves and field data revealed various inconsistencies, including which PCB isomers or

PAHs are preferentially accumulated and/or released by different bivalves and their half-

lives (e.g. Boehm & Quinn, 1976; Fucik & Neff, 1977; Jackim & Lake, 1978; Langston,

1978; Lee et al., 1978; Bjorseth er al., 1979; Calambokidis et al., 1979; Obana et al.,

1983; Pruell er al., 1986, 1987; Weigelt, 1986; Jovanovich & Marion, 1987; Tanabe et

al., 1987a; Fox, 1988; Wade et al., 1988c, Tanacredi & Cardenas, 1991). Although the

causes of such disagreements are not clear, the uptake of PCBs and PAHs from solution

in laboratory experiments may be different from real situations since the routes of

contaminant uptake may differ. Methods using contaminants adsorbed onto particles, e.g.

clay, might produce more realistic results in uptake/depuration studies since they closely

simulate the manner by which filter-feeding bivalves are likely to be exposed to organic

xenobiotics in the coastal marine environment. Also, the effects of using solubilizing

agents and exposure concentrations much higher than those measured in the field are

difficult to evaluate and extrapolate to real situations. In the case of experiments using

naturally contaminated sediments, synergistic or antagonistic effects between different
organic contaminants are likely to influence the uptake kinetics of these compounds.

Furthermore, certain techniques, such as breaking open the bivalve shell to permit

continuous contact with the contaminated medium (see, for example, Fortner & Sick,

1985), obviously produce conditions that are not normally encountered in the

environment.

Finally, marine organisms in the environment are exposed to complex contaminant

mixtures rather than to individual compounds. Therefore, detailed information on uptake

and deputation kinetics of xenobiotics for organisms exposed to contaminant mixtures

must include both field and laboratory studies to assess the effects of anthropogenic

chemicals on marine biota.

RESEARCH OBJECTIVES

In view of the preceding discussion, this study was designed to:
1. Evaluate the uptake of selected PCB congeners, PAHs and butyltin species in

transplanted American oysters, Crassostrea virginica, under field conditions in Galveston

Bay, Texas. Transplanted organisms from a clean environment, Hanna Reef, are

compared to native oysters from a chronically contaminated area near the Houston Ship

Channel where relatively high concentrations of PCBs, PAHs and organotin compounds

are known to exist. Body burdens of both oyster populations are compared to water and

sediment concentrations.

2. Evaluate the depuration of selected PCB congeners, PAHs and butyltin species in

newly and chronically contaminated American oysters, Crassostrea virginica, under field
conditions in Galveston Bay, Texas. As a continuation of the uptake experiment

mentioned above, both originally clean and chronically contaminated oysters were

transplanted from the area near the Houston Ship Channel to the Hanna Reef area and

their depuration kinetics were compared.

3. Evaluate the potential for highly toxic coplanar PCBs to bioaccumulate in oysters

under field conditions. Depuration rates of these PCB congeners by newly and

chronically contaminated individuals are compared.

4. Assess the usefulness of transplanted oysters in biomonitoring studies involving

these trace organic contaminants.

5. Compare, under laboratory conditions, accumulation and depuration dynamics of

selected individual PCB congeners and PAHs by the American oyster, Crassostrea

virginica, when simultaneously exposed to particle - associated PCBs, PAHs and PCBs

plus PAHs, at environmentally realistic levels.

6. Use the experimental results to better understand the PCB, PAH and butyltin data

in oyster samples collected along the northern Gulf of Mexico coast during the NOAA's

National Status and Trends (NS&T) "Mussel Watch" Program.

The American oyster, Crassostrea virginica, was proposed as the organism of interest

for this study due to its wide distribution in the U.S.A. coastal areas, its importance as an

economic resource, and its suitability as a sentinel organism for monitoring coastal

pollution (Goldberg et al., 1978; Bums & Smith, 1981; Farrington et al., 1983; Wade et

al., 1988a). PCBs, PAHs and tributyltin species were selected for study because of their

toxicity and ubiquitous distributions in the marine environment.

GALVESTON BAY SYSTEM

It is not the purpose of this section to present a thorough description of the Gal veston

Bay system. The intention, instead, is to briefly describe the system in order to provide a

basic background for this study. Most of the following paragraphs are summarized from

Stanley's work (1989).

The Galveston Bay system (Fig. 1) includes the Galveston, Trinity, East and West
Bays, with a total area of nearly 1,430 km2. Water depth through the area is very

shallow. Average depths range from <1 m, in East and West Bays, to 2-4 m, in the lower

Galveston Bay. Upper Galveston and Trinity Bays average about 1.6 m in depth. The

maximum depths (up to 12 m) are found in the dredged channels, e.g. Houston Ship

Channel.

Main freshwater inflows to the Galveston Bay system include those from the San

Jacinto and Trinity River drainage areas. The Trinity River basin is the largest with a
drainage area of approximately 46,540 km2 and supplies about half of the total freshwater
input to Galveston Bay. The San Jacinto River basin has a much smaller drainage area
(10,230 km2). While the San Jacinto River is generally the main source of freshwater to

the lower Houston Ship Channel, the principal source of inflow during dry periods is

wastewater discharges. Smaller coastal drainage areas also contribute freshwater to the
bay system. Total coastal inputs represent a drainage area of about 2,000 km2. The
combined annual' freshwater inflow to the system averages 11.6 km3. In addition to these
overland runoffs, there is an average input of about 1.9 km3 each year fromprecipitation

directly onto the bays.

Mean salinity in the Galveston Bay system is around 177oo, but is highly variable in

time and space. Trinity Bay has generally the lower salinity, mainly because of the Trinity

River's outflow. Salinity along the western pan of Galveston Bay is typically higher than

95- 94@30'

TEXA S

EAS

29*30'

.. .. . ...

GALVESTON

Site 1: Hanna Reef

WES Site 2: Ship Channel

Fig. 1. Galveston Bay, Texas.

that on the eastern section. This is due to the Trinity River discharge from the east and to

the barrier formed along the Houston Ship Channel. This channel is the primary path for

salinity intrusion into upper Galveston Bay. Mean water temperature for the entire

Galveston Bay system averages about 220C, although the water temperature follows very

closely the seasonal changes in air temperature.

The Galveston Bay system constitutes one of the largest and most economically

important estuaries along, the Texas Gulf coast. For many years, this area has been the

recipient of various environmental injuries because of an aggressively growing urban and

industrial development. Houston, Deer Park, Baytown, Texas City and Galveston,

surrounding Galveston Bay to the north and west, are some of the most ' heavily

industrialized areas in Texas. Hundreds of industrial plants, including petrochemical

complexes and refineries, bordering the Galveston Bay estuarine system are likely to

introduce significant amounts of organic pollutants into the Bay. Early ecological studies

showed the damaged suffered by different areas in Galveston Bay. Hohn (1959) and

Chamber & Sparks (1959) reported significant decreases in diatom species diversity and

number of invertebrates and fish in the upper Houston Ship Channel. Fish species

diversity indices were also used to assess the health of Galveston Bay (Betchtel &

Coperland, 1970). A change in species diversity from sciaenids to anchovy was related to

the influx of pollutants into the Bay. In general, these ecological studies suggested that

the waters of Galveston Bay contained pollutants in sublethal amounts, which caused

stress to organisms resulting in significant changes in the estuarine community structure.

CHAPTER H

BIOAVAnABILITY OF PAHs TO THE AMERICAN OYSTER (CRASSOSTREA

VIRGINICA): A FEELD STUDY

INTRODUCTION

Ile ability of marine invertebrates to incorporate polynucleararomatic hydrocarbons

(PAHs) from polluted aquatic environment has been documented by different authors

(e.g. Mix, 1984; Pruell er al., 1986, 1987; McElroy et al., 1989). In this chapter, the

uptake and release of PAHs, under field conditions, by two groups of American oysters

(Crassostrea virginica) with different pollution histories, are reported and compared.

Oysters from Hanna Reef, a relatively uncontaminated area in Galveston Bay, were

transplanted to a site near the Houston Ship Channel, a highly polluted area, to assess the

accumulation of PAHs over a period of seven weeks. Concentrations in transplanted

oysters were compared to the levels encountered in indigenous Ship Channel oysters.

After the uptake period, the remaining Hanna Reef oysters were back-transplanted to their

original geographic location to monitor the depuration of the bioaccumulated organic

contaminants. At the same time, indigenous Ship Channel organisms were transplanted to

the Hanna Reef area to compare depuration rates of PAHs between both groups of

oysters, i.e. newly and chronically contaminated oysters.

PAHs: A REVIEW

Background Information

The literature regarding the analytical chemistry and occurrences in environmental

samples of polynuclear aromatic hydrocarbons (PAHs) has been adequately reviewed in

several recent articles and books (e.g. National Academy of Sciences). Therefore, only a

general discussion of the most important aspects of these trace organic contaminants

related to this study is presented here.

Polynuclear aromatic hydrocarbons (Fig. 2) are one of several classes of organic

pollutants that are released into the environment, due in large part to human activities, and

are widely distributed in soils, waters, sediments and organisms throughout the world

(e.g. National Academy of Sciences, 1975, 1985; Neff, 1979; Giesy et al., 1983). PAHs

are composed of carbon and hydrogen atoms arranged in the form of two or more

aromatic (benzene) rings that are either fused (e.g. naphthalene) or linked (e.g. biphenyl)

with occasional incorporation of cyclopentene or cyclohexene rings (e.g. indeno[1,2,3-

c,d]pyrene). These hydrocarbons range in molecular weight (MW) from naphthalene

(C I ()H 12, MW 128.16) to coronene (C24H 12, MW 300.26).

Until the late 1970s, it was generally considered that PAHs were formed only during

high-temperature (e.g. 700'Q pyrolysis of organic materials. The discovery in fossil

fuels of complex mixtures of PAHs spanning a wide molecular weight range has led to the

conclusion that, given sufficient time (e.g. millions of years), pyrolysis of organic

materials at temperature as low as 100- 1 50'C can lead to production of PAHs (Blumer,

1976). In addition, there is some experimental evidence that a wide variety of organic

molecules containing fused-rings polyaromatic systems are synthesized by bacteria, fungi

and plants; although this contributes little to the global PAH burden in the environment

(Neff, 1979).

PAHs

2-Rings:

Naphthalene 2-Methyl Naphthalene

3-Rinas:

Anthracene Phenanthrene

4-Rings:

Senz(a)anthracene Chrysene

5-Rings:

Senzo(a)pyrene Dibenz(a,h)anlhracene

6-Rings:

lndeno(1,2,3-c,d)pyrene

Fig. 2. Structures and common names of selected aromatic hydrocarbons discussed in
the text.

Although petroleum is not the only source of hydrocarbons to an ecosystem, most of

the evaluations of environmental concentrations of hydrocarbons are based on the analysis

of total or selected individual compounds that are indicative of petroleum pollution. Major

inputs of petroleum hydrocarbons to the coastal marine environment include drilling

operations and petroleum production, transportation activities, coastal and/or riverine

inputs, combustion of fossil fuels and atmospheric fallout. Cycloalkanes, branched

alkanes, n-alkanes and low molecular weight aromatic compounds are the predominant

hydrocarbons present in petroleum.

In addition to petroleum sources, aromatic hydrocarbons, particularly high molecular

weight PAHs, are introduced into the environment from different sources, e.g. pyrolysis

of organic materials, municipal incinerators, natural fires and coal production and

burning. Because of the persistence and lipophilic nature of PAHs, it is not surprising

that they have been frequently detected in biota, sediment and water samples from a wide

variety of polluted and unpolluted habitats. In general, the presence of petroleum

hydrocarbons in earlier studies has been inferred from the distribution of normal alkanes

and the presence or absence of an unresolved complex mixture (UCM) in the aliphatic

fractions. Since most of these studies were conducted before the introduction of capillary

columns, identifications of individual aromatic compounds were not confirmed by gas

chromatography/mass spectrometry (GC/MS).

It is estimated that more than 230,000 metric tons of PAHs reach the aquatic

environment each year by a variety of routes and accumulate in estuaries and coastal

marine areas (Giesy el al., 1983). Particularly important sources of PAHs are the

discharges of domestic and industrial wastes and runoff from land. For example, urban
runoff entering Narragansett Bay account for 7 1 % of the total inputs of PAHs (Hoffman

ei al., 1984), whereas riverine contribution of PAHs to coastal sediments off Washington

State was reported to be >30% of the total sediment load (Prahl et al., 1984). Generally,

PAHs are detected in part per million (ppm) in organisms and sediments and in part per

billion (ppb) in water samples.

Toxicity of the different PAH compounds differs. Unsubstituted two- or three-ring

PAHs such as, for example, anthracenes, fluorenes, naphthalenes and phenanthrenes,

exhibit significant acute toxicity and other adverse effects on organisms but are

noncarcinogenic. In contrast, four- to seven-ring PAHs such as, for example,

benzo(a)pyrene, are significantly less toxic but are demonstrably carcinogenic, mutagenic

or teratogenic to a wide variety of animals including mammals (Kennish, 1992).

Polynuclear aromatic hydrocarbons of environmental concern are those compounds

having relatively high volatility and/or solubility.

Distribution and Occurrence in Galveston Bay

A number of studies have been conducted in the Galveston Bay area to establish

baseline concentrations of petroleum hydrocarbons in organisms; however, reports of

individual aromatic concentrations or distributions are limited (Table 1). Most of these

studies were conducted with organisms, particularly bivalves.

Oysters collected from several polluted and unpolluted locations in Galveston Bay in
November 1969 and January 1971 had total PAHs that ranged from 11.0 to 237 ng g- I

(Fazio, 1971). The highest PAHs in oyster tissues from contaminated sites were
fluoranthene (7.8 ng g-1), pyrene (6.5 ng g-1), benzo(b)fluoranthene (2.2 ng g-1), and
benzo(e)pyrene (2.1 ng g-1). Benzo(a)pyrene was below detection in samples from both

contaminated and uncontaminated stations.

Much higher concentrations were reported for oyster samples from a heavily polluted

area, Morgan's Point Reef, near the entrance of the Houston Ship Channel (Ehrhardt,

1972). Concentrations of aromatic hydrocarbons, mainly mono-, di-, and tricyclic
aromatics, were higher than those of alkanes (134,000 and 102,000 ng g- 1, respectively).

TABLE 1
Hydrocarbon Concentrations in Samples from the Galveston Bay Area. Except Where Indicated, Concentrations in Organisms Are Expressed in
ng g- I on a Wet-Weight Basis. Concentrations in Sediment and Water Samples Are Expressed in ng g-1, on a Dry-Weight Basis, and in ng 1-1,
Respectively. Ranges in Parenthesis.

Location Sample Total HCs Total Aromatic HCs Individual PAHs Reference

Galveston Bay oysters (11-237) fluoTanthene= 7.8 Fazio, 1971
pyrenc= 6.5
benzo(b)fluoranthenc= 2.2
benzo(e)pyrene= 2.1
benzo(a)pyrene= n.d.
Houston Ship Channel oysters 236,000 134,000 Ehrhardt, 1972
Morgan's Point Reef oysters 160,000 Anderson, 1975
Halfway Reef oysters 26,000 Anderson, 1975
East Bay oysters <2,000 Anderson, 1975
West Bay oyster <2,000 Anderson, 1975
Galveston Bay oysters pyrene= 1,010 Farrington et al, 1980
fluoranthenc= 940
Morgan's Point Reef oysters benzo(a)pyrene= 0. 12 Murray et al., 1980
YachtClub oysters 615 pyrene= 212 (55-481) Fox, 19880)
(319-1,020) fluoranthene= 112 (55-219)
chrysene= 97 (<20-146)
Todd's Dump oysters 134 pyrene= 31 (<20-63) Fox, 19880)
(94.7-183) fluoranthene= 12 (<20-57)
chrysene= <20 (<20-36)

TABLE I
(continued)

Location Sample Total HCs Total Aromatic HCs Individual PAHs Reference

Confederate Reef oysters 610 pyrene= 146 (40-293) Fox, 1988(l)
(259-1,120) fluoranthene= 210 (55-404)
chrysene= 61 (28-87)
Hanna Reef oysters ill pyrcne= <20 (<20-25) Fox, 19880)
(21.3-228) fluoranthcnc= <20 (<20-37)
chryscnc= <20
Morgan's Point Reef oysters 5,783 pyrcnc= 2,170 (669-3.9 10) Scricano(l)
(2,270-10,120) nuoranthene= 738 (317-1,120) (unpublished data)
chrysene= 632 (260-1,090)
San Luis Pass Fish, crab, benzo(a)pyrene= <0.01 Murray et al., 198 1 a
shrimp
Houston Ship Channel Cormorants naphthalene= (20-40) King et al., 1987(2)
fluoranthene= (n.d.-70)
pyrene-- (20-240)
benzo(a)pyrene= (40-110)
chrysene= 130
benzo(g,hj)perylene= 590
benzoWfluoranthene= 40
1,2,4,5-dibenzoanthracene= 20

TABLE I
(continued)

Lmation Sample Total HCs Total Aromatic HCs Individual PAHs Reference

Trinity Bay sediments 96,100 34,200 dimethyl naphthalenes= 8,000 Armstrong et al., 1979
tfimethyinaphthalenes= 10,000
C4- naphthalenes= 9,000
dimcthylbiphenyls= 800
San Luis Pass sediments bcnzo(a)pyrenc= 2.2 (0.01-6.0) Murray et al., 198 1 a

Trinity Bay water 10,500 10,500 benzene= 1,500 Armstrong et al., 1979
toluene= 3,200
C2- benzene= 3, 100
C3- benzene= 800
dimethyinaphthalenes= 700
San Luis Pass water benzo(a)PyTcne= n.d. Murray et aL, 198 1 a

n.d.= not detected; (1) ng g- I on a dry-weight basis; (2) geometric mean

Anderson (1975) reported similar concentrations of total hydrocarbons in oysters collected
at the same general location (160,000 ng g-1). At Halfway Reef, a few miles farther

away from the entrance of the Houston Ship Channel toward the center of Galveston Bay,
26,000 ng g-1, wet weight, of total hydrocarbons were detected while oyster samples
collected in the East and West Bays had less than 2,000 ng g-I of total hydrocarbons in

their tissues. Benzo(a)pyrene in oysters collected during May 1979 near Morgan's Point
Reef ranged from 0.07 to 0. 14 ng g- I with a mean of 0. 12 ng g- I (Murray et al., 1980).

In 1980, Farrington et al. published the hydrocarbon concentrations measured in

bivalves collected from 90 to 100 stations around the U.S. coastline during the EPA

"Mussel Watch" Program (1976-1978). Oysters collected in the Galveston Bay area
during 1977-1978 had concentrations of 940 and 1,010 ng g- I for fluoranthene and

pyrene, respectively.

. Fox (1988), in a study designed to examine the spatial and temporal variations in

concentrations of selected organic contaminants in Galveston Bay, reported the PAHs

concentrations in oysters from three stations at four sites sampled during 1986. Total

PAHs were higher in samples from sites located closer to urban areas. Oysters collected
in the proximity of the Houston Yacht Club (615 ng g- 1, range = 319-1,020 ng g- 1) and
Confederate Reef (610 ng g-1, range = 259-1,120 ng g-I), near the city of Galveston,

had annual average concentrations higher than samples collected in the Todd's Dump area
(134 ng g-1, range = 94.7-183 ng g-I), located in the middle of Galveston Bay, and
Hanna Reef (I'll ng g-I, range = 21.3-228 ng g-I), in the East Bay. Pyrene,

fluoranthene, chrysene, phenanthrene and 1-methyl phenanthrene were the most

frequently detected analytes. Although temporal variations of individual PAHs in oysters

from the Galveston Bay area did not present an easily recognizable trend during this

study, it seemed evident that total PAHs in samples from the most polluted sites, i.e.

Houston Yacht Club and Confederate Reef, were lower during the summer. This

observation is confirmed by data produced during a six month study with oysters from the

upper part of Galveston Bay. Oysters collected monthly near the entrance to the Houston

Ship Channel were analyzed for a number of organic contaminants between December
1988 and June 1989. Ile maximum total PAHs measured in February (10,100 ng g-l'
range = 9,680-10,600 ng g- 1) decreased to 2,270 ng g- I (range = 1,840-2,7 10 ng g- 1) in

May. Pyrene, fluoranthene, chrysene, benz(a)pyrene and benzo(e)pyrene were the most

abundant PAHs detected during that study (Sericano, unpublished data). Temporal

variations of trace organic contaminants in bivalves were also reported for DDT (Butler,

1973) and PCBs (Farrington et al., 1993). Other marine organisms collected at San Luis

Pass, located in West Galveston Bay at the west end of the Galveston Island, were

analyzed for benzo(a)pyrene (Murray et al., 198 1 a). In all cases, concentrations were
below the detection limit (<O.O I ng g- 1).

In 1987, King et al. reported the concentrations of selected PAHs in double-crested

cormorants, a fish-eating bird near the top of an aquatic food web, wintering in the

Houston Ship Channel. This cormorant is rarely found in the area during summer

months. Naphthalene and fluoranthene were the only PAHs present in individuals

collected at the beginning of the study. After the three-month winter period, eight

aromatic hydrocarbons were detected in bird carcasses (Table 1).

Reports of PAHs in sediment and water samples from the Galveston Bay area are

limited. In 1979, Armstrong et al. reported the results of a study conducted from April

1974 to December 1975 to examine the effects of brine effluents from a producing

platform in Trinity Bay on the surrounding benthic communities. Total petroleum
hydrocarbons measured in sediments collected near the platform were 96,100 ng g-l.
Approximately one third of this total (i.e. 34,200 ng 9-1) were aromatic hydrocarbons,

mainly dimethyl-, trimethyl-, and tetramethylnaphthalenes. Bottom water samples

collected at the same site contained mostly monoaromatic compounds, e.g. toluene,

benzene, and C2-benzene, in the 200-3,200 ng 1-1 range. Total PAHs in water was
10,500 ng 1-1. Sedimentary PAHs decreased with distance from the platform to near
background levels (2,000-6,000 ng g-1). There was a definite inverse correlation

between sedimentary PAHs and the number of benthic species and individuals present.

The Bay bottom was almost completely devoid of organisms within 15 m of the effluent

outfall. Stations located 455 m from the platform were unaffected. Sediment samples

collected in the San Luis Pass area had an average benzo(a)pyrene concentration of 2.2 ng
9_1 (range = 0.01-6.0 ng g-1; Murray et al., 1981a). Benzo(a)pyrene was not detected in

water samples from that area.

Bivalve Uptake and Depuration Studies

A considerable number of reports on the uptake and deputation of petroleum

hydrocarbons by bivalves have been published over the last two decades. In general,

bivalves can be exposed to petroleum hydrocarbons in the laboratory by any one or a

combination of several methods including water-soluble fractions (Neff & Anderson,

1975; Neff et al., 1976; Lee et al., 1978; Nunes & Benville, 1979; Jovanovich & Marion,

1987; Axiak et al., 1988), water dispersion s/sol ution s (Boehm & Quinn, 1973; Stegeman

& Teal, 1973; Stainken, 1975; Wong, 1976; Fossato & Canzonier, 1976; Stainken, 1977;

Fucik & Neff, 1979; Riley er al., 198 1; Tanacredi & Cardenas, 199 1), contaminated food

(Roesijadi er al.,-1978; Fortner & Sick, 1985) and contaminated sediments (Palmork &

Solbakken, 198 1; Obana et al., 1983; Pruell et al., 1986, 1987). Similarly, experimental

field studies include exposures to water soluble fractions (Wolfe et al., 1981), water

dispersion$/solutions (Fucik et al., 1977) and contaminated sediments (Roesijadi er al.,

1978). Alternativelyj uptake and/or deputation studies can be performed in the field by

transplanting uncontaminated bivalves to contaminated areas (e.g. Sericano et al., in

press) or relocating chronically contaminated bivalves into pristine environments (e.g.

Pittinger et al., 1985) or in tanks in the laboratory (e.g. Boehm & Quinn, 1977).

Bivalves are generally reported to preferentially bioaccumulate four-, five- and six-ring

PAHs when exposed, in the laboratory, to naturally contaminated sediments (e.g. Pruell

et al., 1986, 1987) or in the environment (e.g. Bjorseth et al., 1979), with little, if any,

uptake of two- and three-ring PAHs. However, oysters from the Gulf of Mexico were

reported to preferentially uptake two- and three-ring PAHs when compared to four-, five-

and six-ring PAHs (Wade et al., 1988c).

There is some disagreement in the published literature regarding the accumulation of

individual PAHs and their half-lives in different bivalves. For example, the order

chrysene > benzo(b)fluoranthene > fluoranthene > benzo(a)pyrene > benzo(a)-anthracene

encountered in mussels (Pruell et al., 1986) do not agree with the accumulation order

pyrene > benzo(e)pyrene > benzo(b)fluoranthene > benzo(a)anthracene reported for clams

(Obana et al., 1983). Both bivalves were exposed in the laboratory to contaminated

sediments. Similarly, the estimated half-lives for fluoranthene and benzo(a)anthracene

reported for mussels (30 and 18 days, respectively; Pruell et al., 1986) disagree with the

half-lives encountered in oyster (5 and 9 days, respectively; Lee et al., 1978)

Most previous studies have indicated significant but incomplete depurations of

aromatic hydrocarbons by different bivalves (e.g. Fossato & Canzonier, 1976; Pruell et

al., 1986, 1987; Sericano et al., in press). However, some studies reported a complete

depuration of different PAHs to levels below detection limits after relatively short periods

of time, i.e. less than a week. Wormell (1979), for example, reported that depuration
studies with chronically contaminated oysters showed no preferential retention"Of

saturated or aromatic hydrocarbons. Depuration was rapid and nearly complete with a

biological half-life of 4.4 days for total accumulated hydrocarbons. The report suggests,

however, that seasonally related conditions might be a significant factor in the ability of

oysters to clean themselves since individuals depurated in December and January retained

a significant fraction of the bioaccumulated hydrocarbons.

Pittinger et al. (1985) indicated that contaminated oysters (Crassostrea virginica),

transplanted to a nonimpacted site depurated PAHs to undetectable levels within four days

of relocation. Other studies did not detect any depuration. Tanacredi & Cardenas

(1991), for example, reported that laboratory exposed clams (Mercenaria mercenaria) did

not show evidences of decreasing trends in accumulated PAHs after a 45-day depuration

period. Similar results were reported by Boehm & Quinn (1976). In that study,

chronically contaminated clams (Mercenaria mercenaria) failed,to release the accumulated

PAHs when relocated to clean seawater over a four-month period. Both reports,

however, seem to be in disagreement with the vast majority of previous investigations

involving bivalves.

In spite of the abundant information, it is clear from the preceding discussion that

there are many contradictions in the published literature. The reasons for these

disagreements are difficult to explain. It is possible, however, that the extremely high oil

or individual analyte concentrations used in some studies, the presence of stressed animals

and the use of different experimental designs or analytical techniques could be responsible

for the observed discrepancies.

UPTAKE AND DEPURA71ON OF PAHs

Experimental Design, Sample Collection and Methods

In December 1988, approximately 250 oysters were collected by dredge at Hanna
Reef, a relatively pristine area in Galveston Bay (Fig. 3). Within 24 hr., these oysters
were transplanted live, in net bags, to a site near the Houston Ship Channel, an area

where oysters have high PAH concentrations (Wade et al., 1988a). Photographs of both

94@30'

T E X A S

EAST

29*30,

40 GALVESTON
0

A
Site 1: Hanna Reef

W 2: c?hip Channel
Site

Fig. 3. Galveston Bay transplantation sites.

sites, nets and oysters are shown in Fig. 4. Thereafter, oysters were sampled in groups

of 20 individuals during the 3rd, 7th, 17th, 30th and 49th days after transplantation.

During the uptake period, native oysters were collected from the Ship Channel area to

compare their concentrations of these trace organic contaminants with those encountered

in transplanted Hanna Reef oysters.

The remaining transplanted oysters, i.e. approximately 150 individuals, were re-

located to the Hanna Reef area and sampled in groups of 20 individuals during the 3rd,

6th, 18th, 30th, and 50th days after transplantation. The transplant experiment to the

Hanna Reef area was duplicated with approximately 150 native oysters from the Houston

Ship Channel area in order to compare depuration dynamics in both populations. In the

following sections, Ship Channel, Hanna Reef, transplanted Hanna Reef-to-Ship

Channel, transplanted Ship Channel-to-Hanna Reef and relocated Hanna Reef-to-Ship

Channel-back to-Hanna Reef oysters are refered as SC, HR, HRSC, SCHR and

HRSCHR oysters, respectively. Sediment and water samples were collected during

oyster sampling days for PAH analyses.

Extraction andfi-actionation of PAHs

The analytical procedure used was based on a method developed by MacLeod et al.

(1985) with a few modifications that proved to be equivalent or superior to the original

techniques. The analytical scheme is summarized in Fig. 5 Precleaning of all glassware

involved extensive washing with Micro cleaning solution, rinsing with distilled water and

combustion at 400*C for 4 hrs. All solvents were glass-distilled nanograde purity, e.g.

Burdick & Jackson. Solvent purity was checked, after 300-fold concentration, by gas

chromatography/mass spectrometry (GC/MS). Each set of samples (8-10) was

accompanied by a complete system blank and spiked blank or reference material that were

carried through the entire analytical procedure. Before extraction, PAH internal standards

Alf, Or

-,FAA,.
A,, ilbm

Fig. 4. Exposure (a) and depuration (b) sites. respectively. Approximately 250 adult oysters (c) w
transplanted in net bags (d).

ST-D I M I, NT
TISSUE i
WARM TO ROOM HOMOGENIZE
TEMPERATURE DRY .911 1
W IGHT ADDITJONOF
INTERNAL STANDARVS
HOMOGLNUE

EXTJLACTTO%i
ADDITION OF
M.FERNAL STANDA RDS t. C5 em 100 mi
2. CH, OMICK2 CS 000-1
J. v@ Cs I 1 100 ml K j
COLD EXTRACTION PA RTITION Wate,
cm, q 1 3 N 100 mi WITH S..CL SOLUTION pho"
Organic phole
CENTRI]FUGE CONCE.NTRATE

CON RATE
CENT
i COPPER TREATMENT

COLUNIN,
CHROMATOGR.Aplly@ U, =Rol NAPH Y

CONCENTRATED EXTRA--)
I

ALUMINA / SILICA GFL
CHROMAT'DGILAPSY

ALIPHATIC PAH. @6 OTHER POLAR
HYDROCARBONS PEST. & PCB's ORGANICS

GC/FID; GCI%4S SEPIIADEX ccrlf); GC/MS

TOT. PAH', DERI%IZATION

Siructure Coortr", ation

GC-'FID; CC/.-.fS

GCIFID; GCISIS CCIECD; CCIMS T

--@o p- R REDUCTION/RNRLYSIS
L T
@
UOM MZE

A
R
AL 11ANIA..11

FLORISIL
HROMATOGRAPHY

Fig. 5. Trace organic analytical scheme.

(d8-naphthalene, djo-acenaphthene, djo-phenathrene, d12-chrysene and d12-perylene)

were added to all samples, blank and spiked blank. These standards were added at a

concentration level similar to that expected for the sample components of interest.

Approximately 15 g of wet tissue were used for the analysis of PAHs in oysters.

After the addition of 50 g of anhydrous Na2SO4, the tissue was extracted three times with

100 ml of methylene chloride using a "Tissurnizer" homogenizer. The organic phase was

concentrated to 10-15 ml in a flat-bottom flask equipped with a three ball Snyder

condenser. Kudema-Danish tubes were heated in a water bath at 60'C, to concentrate the

extracts to a final volume of 2 ml in hexane.

Approximately 50 g of sediment (wet weight) were used for analysis. The sediments

were sequentially extracted on a roller table with 100 n-d of methanol (I h), 100 ml of 1: 1

methanol:methylene chloride (I h) and three portions of 100 ml of methylene chloride (16,

3 and I h, respectively). The combined extracts were partitioned into two phases by

addition of acidic NaCl solution (10%, pH=2). The combined extracts were concentrated

to 1-2 ml as previously described for oyster extracts.

Fifteen to 17 1 of seawater samples were acidified with HCl (pH=2) and extracted, for

15 min, with 500 ml of methylene chloride. The extraction was repeated three times. The

organic phase was partitioned against an acidic NaCl solution (pH=2). The extract was
then dried with anhydrous Na2SO4 and concentrated to 1-2 ml as previously described

for oyster extracts.

Tissue, sediment and water extracts were fractionated by alumina: silica (80-100

mesh) column chromatography. The silica gel was activated at 170'C for 12 h and

partially deactivated with 5% distilled water. Twenty grams of silica gel were slurry

packed in methylene chloride over 10 g of alumina. Alumina was activated at 400"C for 4

h and partially deactivated with 1 % distilled water. Activated copper was added to the top

of the column for sediment samples to remove any residue of elemental sulfur. The

methylene chloride was replaced with pentane and the extract was applied to the surface of

the column. The column was sequentially eluted with 50 ml of pentane (f 1), 200 ml of

1:1 methylene chloride:pentane (U) and, for sediments, 50 ml of methanol (6). The f2

fraction, which contains the polynuclear aromatic and chlorinated hydrocarbons, was

concentrated as previously described. The f2 fraction from oyster samples was further

purified by Sephadex LH-20 column (25-100 mesh) to remove lipids (Ramos &

Prohaska, 1981). The column was eluted with 140 ml of a cyclohexane:methanol:

methylene chloride (6:4:3) mixture. The first 40 ml were discarded and the next 100 nil

fraction was concentrated to a final volume of 0.5-1 ml, in hexane, for gas

chromatographic/mass spectrometry analysis.

1nstrwwnW analysis

PAHs were quantitatively analyzed by GC/MS in a selected ion mode (SIM) utilizing

the molecular ions of the components of interest. A 30 m DB-5 capillary column (0.31

mm i.d., 0.052 mm film thikness) was temperature programmed from 40 to 3000C at
10'C min-I and hold at 300'C for 10 min. The GC/MS was calibrated by injections of

standard solutions at three different concentrations. Sample analytes were quantified from
a first degree calibration curve with an R2 value equal to or greater than 0.99. Analyte

identity was confirmed by their molecular weights and retention times of authentic

standards.

Ancillwyparairwers

Grain size analysis was performed by the procedure of Folk (1974). Briefly,
reffigerated samples were homogenized, treated with 30% H202 to oxidize organic matter

and washed with distilled water to remove soluble salts. Sodium hexametaphosphate was

added to deflocculate each sample before they were wet-sieved though a 62.5 micron (4.0

phi) sieve to separate gravel and sand from the silt and clay fraction. ne total gravel and

sand fraction was then oven dried at 40*C and weighed. The silt-clay fraction was

analyzed for particle size distribution by the pipette (settling rate) method.

Extractable lipids in oysters were determined on an aliquot of the sample extracts.

Twenty nil of the combined oyster extracts were withdrawn from the total volume and

evaporated to dryness under N2 gas. The residue was redisolved in I ml of methylene

chloride, 0. 1 ml was evaporated on a paper pad and the residual materials weighed using a

Cahn 29 electrobalance.

Statistical analysis

During the uptake period, one-way analyses of variance (ANOVA) were performed on

the analyte concentrations to evaluate the bioconcentration by transplanted oysters relative

to indigenous individuals. Slopes of the least-square linear regressions of the logarithm

of the concentrations (log Q versus time (t) for the depuration period were tested for

statistical significance.

Uptake of PAHs by Transplanted Oysters

Average PAH concentrations measured in SC and HRSC oyster, sediment and water

samples, during the first part of this experiment at the Ship Channel site, are reported in

Tables A-2 and A-3 (Appendix).

The concentrations in Ship Channel oysters represent the time-integrated amounts of

trace organic contaminants accumulated from solution, particles and/or food minus any

metabolism and/or depuration of these comp ounds. The concentrations of most of these

PAHs in SC oysters did not change significantly during the first phase of the experiment.
Total aromatic hydrocarbon concentrations in SC oysters averaged 3,800�590 ng g-l
(range 3,200 to 4,400 ng g-1) over the seven-week uptake period. The fluctuations

observed in the concentrations of the lower molecular weight PAHs with time were,

however, comparatively greater than the variability encountered in the concentrations of

the higher molecular weight analytes. Naphthalene, 2,6-dimethyl-naphthalene and

phenanthrene, for example, had an overall average of 13�8.3, 25�16 and 54�37 ng g-

during this period with coefficient of variations of 64, 64 and 69%, respectively. These

coefficients of variations were larger than those observed for pyrene (1,500�290 ng g-
19%), benzo(e)pyrene (270�65 ng g- 1; 24%) and perylene (I 30�22 ng g- 1; 17 %) during

the same period of time.

Concentrations of PAHs in HRSC oysters increased dramatically during the seven-

week exposure period. Concentrations of total PAHs in transplanted HRSC oysters
increased from an initial total concentration of 290 to 4,400 ng g- I during this period.

Four- and five-ring PAHs were rapidly accumulated by HRSC oysters; comparatively,
two-, three- and six-ring PAHs were detected in low concentrations in both transplanted

HRSC and indigenous SC oysters (Fig. 6). One month after the experiment started, no

statistically significant differences were observed in the distributions of PAHs, by ring

number, between HRSC and SC oysters.

Generally, the PAHs measured in HRSC oysters had similar concentrations to those

found in SC oysters in less than 20 days (Fig. 7). Four- and five-ring PAHs had

increased in the HRSC oysters while the two- and three-ring PAHs concentrations either

did not change (e.g. naphthalene, 2-methylnaphthalene) or increased only slightly (e.g.

2,3,5-trimethylnaphthalene, I-methylphenanthrene) during the seven-week exposure

period. A decrease was observed in the concentrations of some lower molecular weight

PAHs in transplanted oysters during the first week of exposure. These decreases in

concentrations were not observed in indigenous SC oysters. However, since this initial

decrease in the concentrations of low molecular weight hydrocarbons was not observed

for the higher molecular weight PAHs, it is unlikely that it was a consequence of stressed

4000 Ship Channel Site 0 day 4000, Ship Channel Site 17 days
M Banns ReerOysLers 0 Bass& ReelOysters
0 Ship Channel Oysters It M Ship Channel Oysters
P..
3000, 3000'

at
20M 2000

6
Iwo low

Mniiiiiiiiii viiiii=
2 3 4 5 6 2 3 4 5 6
Number of Rings Number of Rings
Ship Channel Site 3 days 4000- Ship Channel Site 30 days
4000 0 Hanna Reeroysters
0 Hanna Reer Oysters
0 Ship Channel Oysters
0 Ship Channel Oy.slers
3000-
6 3000'

-fAt
2000 2000
C

low 5 1000
W Cc
U 0
A moog-
2 3 4 6 2 3 4 5 6
Number or Rings Number or Rings
4000' Ship Channel Site 7 days 4000- Ship Channel Site 48 days
M Hanna Reef Oysters 0 Hanna Reef Oysters
0 Ship Channel Oysters 1: Ship Channel Oysters
3000. V

2000- 2000
r

V
6
r 1000 low
Cc

Number of Rings Number of Rings

Fig. 6. Concentrations of polynuclear aromatic hydrocarbons, grouped by number of
rings, in transplanted Hanna Recf and indigenous Ship Channel oysters during the 48-
day exposure period near the Houston Ship Channel. Ship Channel oysters were not
sampled on day 7.
L Ion

Naplithalene 2-Methy1naphthalene

Hanna Reef Oystm
J J
It Ship chound Oysters Ship Channel Oystas

to"

too- too

U U

6 1$ 20 30 40 so 60 70 so 90 too 0 to 20 30 40 so 60 70 go 98 too

Time (days) Time (days)

2,3,5-TrImethyInaphthaIene I-Methylphenanthrene
Hanna Reef Oysters 10000. Hanna Reer Oysters
V Ship Channel Oysters Ship Channel Oystffs

10" 100,
V V

too fee =--46
C - ----------
0

C
0
U I U
0 10 20 30 40 50 60 76 Be 90 foe 0 10 20 30 40 50 60 70 Be 90 100
11me (days) Time (days)

Fig. 7. Concentrations of selected polynuclear aromatic hydrocarbons in tissues of Hanna Reef and Ship Channel oysters
during exposure to the Ship Channel area contaminant levels and following transplant to the Hanna Reef area.

Pyrene Denz(a)anthracene
HannaltedOysters Hanna Red Oynters
Ship Channel Oysten Ship Channel Oyden

c at
IGO

It
c 10

0 10 20 36 40 so 60 70 as 90 too 0 to 20 30 40 50 60 70 so 90 too
71 me (days) Time (days)
100" Chrysene Hanna ReefOysters too". Benzo(b)nuoranthene flannalledOyssers
Ship Channel Oyders Ship Channel Oysters

to".

100
C

9 10 20 30 40 50 69 70 80 90 too 0 10 20 30 40 50 60 70 so 90 too
Time (days) Time (days)

Fig. 7. (Continued)

Denzo(&)pyrene ftenzo(e)pyrene
Hanna Red 0yaterv Hanna Red Oysten
It Ship Channel Oysters It Ship Channel Oysters
If" - 0111111
V

be am

If
c

8
U U
I
0 It 20 30 40 50 60 79 so 90 100 0 10 20 30 40 50 60 70 so 90 100
Time (days) Time (days)
Perylene Benzo(g,h,I)perylene
Hanna Rftf0y.,Urs Hanna Reef Oysters
---w- Ship Channel Oysters Ship Channel Oysters

Joe@-

114111- too

it, c is

C
0
U U
I
0 10 20 30 40 50 66 70 so 90 100 0 10 20 30 40 50 60 70 so 96 100
Time (days) Time (days)

Fig. 7. (Continued)

animals. By the, end of the exposure period, the concentrations of all the individual PAH

measured in transplanted oysters were not statistically differentiable from those

encountered in indigenous oysters (Fig. 8). The observed rapid bioconcentration of

PAHs is similar to the uptake curves reported in different studies involving bivalves either

in laboratory experiments (Nunes & Benville, 1979; Pruell et al., 1986; Tanacredi &

Cardenas, 1991) or in transplantion studies (Pittinger et al., 1985). The PAHs

accumulated to the highest concentrations in SC and HRSC oysters were: pyrene >

fluoranthene > chrysene > benzo(e)pyrene > benzo(b)anthracene. Three PAHs, pyrene,

fluoranthene and chrysene, accounted for about 60% of the total PAH load measured in

these oysters. Other uptake studies, using a variety of organisms exposed to different

sources of PAHs, produced a different order for the concentration of four- and five-ring

PAHs. For example, the relative abundances reported by Pruell et al. (1986) for mussels

(chrysene > benzo(b)nuoranthene > fluoranthene > benzo(a)pyrene > benzo(a)anthracene)

or by Obana et al. (1983) for clams (pyrene > benzo(e)pyrene > benzo(b)fluoranthene >

benzo(a)anthracene) are diferent from that found in this study for oysters. These

discrepancies might reflect the different PAH compositions in the sources or a different

uptake ability of the organisms.

The average concentrations of PAHs, by number of rings and individually, in Ship

Channel sediment and water samples are shown in Fig. 9 and 10, respectively. Water and

sediment samples were collected each time the oysters.were collected. Sediment samples

had higher relative concentrations of four- and five-ring PAHs when compared to

seawater samples. The relative abundances of PAHs in sediments were pyrene >

benzo(b)fluoranthene > benzo(e)pyrene > chrysene > benzo(a)pyrene > fluoranthene >

benzo(k)fluoranthene (Fig. 9). Two-ring PAHs and most of the three-ring PAHs were

detected at low concentrations in sediments. In contrast, two-ring PAHs, i.e.

naphthalenes, were the predominant PAHs in the seawater samples (Fig. 10). The

Naphthalene
2-Methylnaphthalene
1-Methyinaphthalene
Biphenyl
2,6-Dimethylnaphthalene
Acenaphthylene
Acenaphthene
2.3,5-Trimethylnaphthalene
Fluorene

Phenanthrene

Anthracene
I-Methylphenanthrene
Fluoranthene
Pyrene
Benz(a)anthracene
Chrysene
Benzo(b)nuoranthene
Benzo(k)fluoranthene
Benzo(e)pyrene
Benzo(a)pyrene
Perylene Ship Channel Oysters
Indeno[1,2,3-c,dlpyrene
Dibenz(a,h)anthracene Hanna Reer Oysters
Benzo(g,hJ)perylene r 7,

0 500 1000 1500 2000

Concentration (ng/g, dry wt.)

Fig. 8. Concentrations of individual polynuclear aromatic hydrocarbons in tissues of
Hanna Reef and Ship Channel oysters at the end of the 48-day exposure period.

0 200 400 600 800 1000

Naphthalene
2-Methylnaphthalene
I-Methyinaphthalene
Biphenyl
2,6-Dimethyinaphthalene
Acenaphthylene
Acenaphthene
2,3,5-Trimethylnaphthalene
Fluorene

Phenanthrene

Anthracene
I-Methylphenanthrene
Fluoranthene
Pyrene
Benz(a)anthracene
Chrysene
Benzo(b)fluoranthene
Benzo(k)fluoranthene
Benzo(e)pyrene
Benzo(a)pyrene
Perylene
Indeno[1,2,3-c,d]pyrene
Dibenz(a,h)anthracene
Benzo(g,h,i)perylene

0 so 150 200 250 300

Concentration (ng/g, dry wt.)

Fig. 9. Concentrations of polynuclear aromatic hydrocarbons, grouped by number of
rings and individually, in Ship Channel sediment samples.

- ------ - ------- ---
2

4
Lw

z
z 0 5 10 15 20

Naphthalene
2-Meth)-inaphthalene
1-Methyinaphthalene
Biphenyl
2,6-Ditnethyinaphtholene
Acenaphthylene
Acenaphthene
2,3,5-Trimethylnaphthalene
Fluorene

Phtnanthrene

Anthracene
1-Methylphenanthrene
Fluoranthene
Pyrene
Benz(a)anthracene
Chrysene
Benzo(b)fluoranthene
Benzo(k)fluoranthene
Benzo(e)pyrene
Benzo(a)pyrene
Perylene
Indeno[1,2,3-cdjpyrene
Dibenz(a,h)anthracene
Benzo(g,hj)perylene

0 1 2 3 4 5 6

Concentration (ng/1)

Fig. 10. Concentrations of polynuclear aromatic hydrocarbons, grouped by number
of rings and individually, in Ship Channel seawater samples.

observed decreasing PAH concentrations in seawater with increasing molecular weight is

consistent with published solubility data (e.g. Whitehouse, 1984). The distribution of

PAHs by ring numbers in oyster tissues showed lower concentrations of five-ring PAHs

relative to the surrounding sediments. Compared to the seawater PAH distribution, oyster

tissues were depleted in the more soluble two- and three-ring aromatic hydrocarbons. The

PAH distribution for oysters is intermediate when compared to sediments and seawater.

Depuration of PAHs by Newly and Chronically Contaminated Oysters

Average concentrations of selected PABs measured in HRSCHR and SCHR oyster

and sediment samples from the Hanna Reef area are shown in Table A-4 and A-5

(Appendix). Sediment concentrations were normalized by dividing by the percentages of

silt and clay in the samples to decrease the variability observed among the samples. After

relocation to the Hanna Reef area, HRSCHR and SCHR oysters showed statistically

significant depuration of accumulated PAHs. At the end of the depuration period, the total

PAH concentrations in HRSC oysters were about 40% higher than the final

concentrations in HRSCHR individuals in the same period of time. Total PAH
concentrations decreased from 4,400 to 360 ng g-1, in HRSCHR oysters, and from
4,400 to 500 ng g-1, in SCHR oysters. In both cases, most of the decreases in

concentrations were due to the depuration of four- and five-ring PAHs (Fig. 11). The

diffences in the final concentrations of some individual three-, four- and five-ring

hydrocarbons between SCHR and HRSCHR oyster populations at the end of the 50-day

depuration period are evident (Fig. 12). The largest percent differences were observed for

fluoranthene (38%), pyrene (70%), chrysene (80%) and benzo(e)pyrene (I 10%).

Although HRSCHR and SCHR oysters significantly depurated most of the

bioaccumulated individual PAHs, their concentrations did not decrease to the levels

encountered for HR oysters at the begining of this study. For example, Fig. 13 shows

Hanna Reef Site - 0 day Hanna Reer Site IS days
4000" 4000'
a Hopes Reer Oysters 0 Hansa RedOysteri
Ship Channel Oysters 0 Ship Channel Oysters

3000'
In

2M - 2000

low 1000

Man- 0.
2 3 4 S 6 2 3 4 5 6
Number of Rings Number of Rings
Hanna Reef Site - 3 days 4000- Hanna Reer Site 30 days
a Boons Reef Oysters
U Hansa Reef Oysters
6 Ship Channel Oysters it M Ship Channel Oysters
3000'
39W V

2000- 5 2000-.
C

1000-
1000,

2 3 4 S 6 2 3 4 S 6
Number of Rings Number or Rings
Hanna Reef Site - 6 days Hanna Reef Site 50 days
a Fianna RecrOysters 4000., M Hansa Reer Oysters
0 Ship Channel Oysters It 0 Ship Channel Oysters

A. 30M 3w
V V

at
2M. 2000

0
low"

L) Ld

2 a 4 3 6 2 3 4 5 6
Number of Rings Number of Rings

Fig. 11. Concentrations of polynuclear aromatic hydrocarbons, grouped by number
of Tings, in back-transplanted Hanna Reef and transplanted Ship Channel oysters during
the 50-day depuration period in the Hanna Reef area.

Naphthalene
2-Methylnaphthalene
I-Methyinaphthalene
Biphenyl
2,6-Dimethylnaphthalene
Acenaphthylene
Acenaphthene
2,3,5-Trimethylnaphthalene
Fluorene

Phenanthrene

Anthracene
I-Methylphenanthrene
Fluoranthene

Pyrene
Benz(a)anthracene
Chrysent
Benzo(b)fluoranthene
Benzo(k)fluoranthene
Benzo(e)pyrene
Benzo(a)pyrene
Perylene Ship Channel Oysters
Indeno[1,2,3-c,d]pyrene Hanna Reef Oysters
Dibenz(a,h)anthracene
Benzo(g,h,I)perylene

0 50 100 150 200

Concentration (ng/g, dry wt.)

Fig. 12. Concentrations of individual polynuclear aromatic hydrocarbons in tissues of
Hanna Reef and Ship Channel oysters at the end of the 50-day depuration period.

42)

4
E . . . . . . .......... ...
6
z . . . . . . . . . . . . . . . .
0 so 1@0 1;0 2@0 2;0

Naphthalene
2-Methyinaphthalene
1-Methylnaphthaleruie
Bipbenyl
2,6-Dimetbylnaphthalene
Acenaphthylene
Acenaphthene
W,S-Trimethylnaphthalene
Fluorene

Phenanthrene

Anthracene
I-Methylphenanthrene
Fluoranthene

Pyrene
Benz(a)anthracene
Chrysene
Benzo(b)fluoranthene
Benzo(k)fluoranthene
Benzo(e)pyrene
Benzo(a)pyrene
Perylene Before Transplantation
Indeno[I,2,3-c,dlpyrene
Dibenz(a,h)anthracene After Depuration
Benzo(g,h,i)perylene

0 20 40 60 80 100 120 140

Concentration (ng/9, dry wt.)

Fig. 13. Comparison of the concentrations of polynuclear aromatic hydrocarbons,
grouped by number of rings and individually, in tissues of Hanna Reef oysters before
exposure to the Ship Channel contaminant levels and after depuration at the Hanna Reef
si te.

the concentrations of PAHs, according to the number of rings and individually, in Hanna

Reef oysters before transplantation to the contaminated site and after depuration for 50

days at the Hanna Reef site. The original distribution of PAHs in Hanna Reef oysters

showed a predominance of the more volatile and soluble compounds, i.e. two- and three-

ring PAHs. When these oysters were exposed to higher PAH concentrations at the Ship

Channel area, they bioconcentrated four- and five-ring PAHs. This resulted in higher

concentrations for total PAHs as well as in a different distribution of PAHs when the

oysters were back-transplanted to the Hanna Reef location. These different distributions

are probably the consequences of two different sources of PAHs. In the first case, the

PAH distribution reflects the remote location of the Hanna Reef site and indicates

atmospheric inputs and water transport of the more soluble PAHs as their most probable

sources. A slight increase in the concentration of naphthalenes with time in SCHR and

HRSCHR oysters during the depuration part of this study was observed. When oysters

are exposed to the significantly higher PAH concentrations, over a wider molecular

weight range, present in the Ship Channel area, they readily bioconcentrate the higher

molecular weight PAHs. Because of the low water solubility of the predominant

compounds encountered in Ship Channel oysters, it is probable that the main route of

uptake is through food ingestion. In contrast, most of the uptake in the Hanna Reef area

probably occurs through gills. Unfortunately, the concentrations of PAHs in Hanna Reef

water samples were below the detection limit; therefore, no firm conclusion can be drawn.

Sediment samples collected from the Hanna Reef area had a distribution of PAHs, by

zing numbers, similar to that encountered near the Houston Ship Channel; however, the

concentrations of individual PAHs were about an order of magnitude lower (Fig. 14).

With the exception of the high concentration of perylene, a compound with natural as wen

as combustion sources, in Hanna Reef sediments, the differences in concentrations

6
z
0 so 1 00 ISO 200

Naphthalene
2-Nlethylnaphthalene
1-Methyinaphthalene
Biphenyl
2,6-Dimethylnaphthalene
Acenaphthylene
Acenaphthene
2,3,5-Trimethyinaphthalene
Fluorene

Phenanthrene ......

Anthracene
I-Methylphenanthrene
Fluoranthene
Pyrene
Benz(a)anthracene
Chrysene
Benzo(b)fluoranthene
Benzo(k)fluoranthene
Denzo(e)pyrene
Benzo(a)pyrene
Perylene . ......
1ndeno[I,2,3-c,djpyrene
Dibenz(a,h)anthracene
Benzo(g,hJ)perylene

0 20 40 60 so 100

Concentration (ng/g, dry wt.)

Fig. 14. Concentrations of polynuclear aromatic hydrocarbons, grouped by number
of rings and individually, in Hanna Reef sediment samples.

between the lower and higher molecular weight PAHs in these samples is less marked

than in the case of the Ship Channel sediment samples.

Clearance rates of aromatic compounds by both groups of oysters were approximately

exponential. This is indicated in Fig. 7 where the concentrations of selected PAHs,

during the depuration phase of this study, plotted on semi-log plots, approximate straight

lines. Original Hanna Reef oysters depurated PAHs at a faster rate than Ship Channel

oysters. Differences in the slopes of the depuration curves are reflected in the lower PAH

half-lives for SCHR oysters. Calculations to estimate the half-lives and related kinetic

parameters for the different trace organic pollutants by Crassostrea virginica oysters will

be discussed in more details in Chapter VI. The depuration half-lives for PAHs ranged

from 9 to 24 days and from 10 to 24 days in originally uncontaminated Hanna Reef and

chronically exposed Ship Channel oysters, respectively. Most of PAH half-lives were

between 10 and 13 days and 13 and 16 days for HRSCHR and SCHR oysters,

respectively. These values are within the ranges of previously reported PAH half-lives

(Table 2). A few studies report complete depuration of PAHs by bivalves after they are

relocated to a clean environment over short periods of time (less than a week). However,

the reported minimum detection limits for PAHs in those studies were generally high. For

example, Pittinger et al. (1985) reported minimun detection limits for PAHs ranging from
93 to 222 ng g-l. If those minimum detection limits are applied to this study, most of the

measured PAHs, in both groups of oysters, would be below detectable levels after one

week of relocation to the Hanna Reef area.

CONCLUDING REMARKS

PAHs were rapi dly accumulated by uncontaminated oysters to final concentrations that

were statistically undistinguished from the concentrations encountered in indigenous Ship

TABLE 2
Biological Half-Lives (Days) of PAHs in Hanna Reef and Ship Channel Crassostrea virginica
Oysters.

Analyte Hanna Ship Dun & Stich Lee et al. Pruell et al.
Reef Channel (1976) (1978) (1986)
Oysters Oysters Mussels Oysters Mussels

2,3,5-Trimethyinaphthalene 24 22 - -
Anthracene 24 42 - 3
I-Methylphenanduene 23 24 - - -
Huoranthene 26 3 '1 - 5 30
Pyrene 10 12 - - -
Benz(a)andiracene 13 15 - 9 18
Chrysene 12 16 - - 14
BenzoNfluoranthene - - 17
Benzo(k)fluoranthene - - 12
Benzo(e)pyrene 12 16 - - 14
Benzo(a)pyrcne 9 10 16 18 15
Perylene 11 13 - - -
Indeno[ 1.2,3-c.d]pyrene 10 11 - - 16
Dibenz(ab)anthracene 16 14 - -
Benzo(g,h,i)perylene 11 12 - - 15

Channel oysters within 20 to 25 days after transplantation. The PAHs accumulated to the

highest concentrations in SC and transplanted HRSC oysters were: pyrene > fluoranthene

> chrysene > benzo(e)pyrene > benzo(a)anthracene. Although there are some

discrepancies when comparing the order of uptake of these PAHs, encountered in the

present study, with previously published works using different bivalves, these

discrepancies might reflect the different PAH compositions in the sources or a different

uptake ability of the organisms. Ile final distributions of individual PAHs in transplanted

and indigenous oysters during the uptake phase of this experiment were intermediate

between the profiles encountered in sediment and seawater samples from the Ship

Channel area.

When transplanted to the relatively uncontaminated Hanna Reef area, both groups of

oyster depurated the bioaccumulated PAHs. Calculated depuration rates were higher for

the originally uncontaminated oysters. Most of individual PAH deputation half-lives were

between 10 and 13 days and 13 and 16 days for HRSCHR and SCHR oysters,

respectively. The depuration of individual PAHs by HRSCHR oysters was, however,

not complete and the concentrations encountered at the end of the depuration period were

higher than the levels measured before their exposure to the Ship Channel concentrations.

Comparing the distribution profiles of PAHs encountered in HRSCHR oysters at the end

of the depuration period with the distribution they had before the transplant experiment,

i.e. HR oysters, seems to indicate that the sources of PAHs to Hanna Reef and Ship

Channel are different. While the original distribution of PAHs in Hanna Reef oysters

showed a predominance of the more volatile and soluble compounds, i.e. two- and three-

ring PAHs, the distribution of PAHs after the deputation phase of this experiment shows

predominance of four- and five-ring PAHs. It can be speculated that petroleum

background and water transport are the most probable sources for the predominance of the

lower molecular weight PAHs to the Hanna Reef area.

CHAPTER III

UPTAKE, RETENTION AND RELEASE OF PCBs BY THE AMERICAN OYSTER

(CRASSQSTREA VIRGINICA

INTRODUCTION

Polychlorinated biphenyls (PCBs) are of particular concern in pollution studies

because of their widespread occurrence, environmental persistence and bioaccumulation

properties. For these reasons, these compounds have been included as analytes of interest

in many national and international programs (see, for example, Farrington et al., 1980;

Sericano et al., 1990a). In most of these monitoring programs, bivalves were preferTed

as sentinel organisms.

Despite the overwhelming popularity that the "Mussel Watch" concept has obtained

since its introduction in the 1970s, both monitoring data and, particularly, laboratory-

generated data on PCB kinetics in bivalves show discrepancies. For example, there are

disagreements over which PCB congeners are preferentially accumulated by bivalves and

the length of time bivalves need to reach an equilibrium with environmental concentrations

or the time needed for different PCB congeners to be depurated, if they are depurated,

when the environmental concentration is reduced. Such inconsistencies decrease the

usefulness of the Mussel Watch concept in environmental studies.

This chapter reports the uptake and release of PCBs by two groups of American

oysters (Crassostrea virginica) with different pollution histories. Oysters from Hanna

Reef, a relatively uncontaminated area in Galveston Bay, were transplanted to a site near

the Houston Ship Channel, a highly polluted area, to assess the accumulation of PCBs

over a period of seven weeks. . Concentrations in transplanted oysters were compared to

the levels encountered in indigenous Ship Channel oysters. After the uptake period, the

remaining Hanna Reef oysters were back-transplanted to their original geographic location

to monitor the depuration of the bioaccumulated organic contaminants. At the same time,

indigenous Ship Channel oysters were transplanted to the Hanna Reef area in order to

compare depuration rates of PCBs between the two groups of oysters, i.e. newly and

chronically contaminated.

PCBs: A REVEEW

Background Information

PCBs are the subject of several monographs and books (e.g. Safe, 1984; Erickson,

1986; Safe et al., 1987; Tanabe & Tatsukawa, 1986; Voogt & Brinkman, 1989; Lang,

1992). Polychlorinated biphenyls (PCBs), systematically called 1,1'-biphenyl, chloro

derivatives, is the generic name of many isomers and congeners with I

(monochlorobiphenyls) to 10 (decachlorobiphenyl) chlorine atoms substituted on both

biphenyl rings (Fig. 15). The synthesis of PCBs was first described by Schmidt &

Schultz (1881); however, commercial production in the U.S.A. did not begin until 1929.

The rings in the biphenyl molecule are joined by a single carbon-carbon bond allowing

free rotation of both rings. The presence of one or more chlorine in ortho positions (2,

2', 6 and/or 6') results in an inter-ring angle of up to 90' (McKinney et al., 1983).

Although there are 209 possible PCB congeners, the catalytic electrophilic substitution of

chlorines is favored at the ortho and para positions on the biphenyl molecule. Thus,

several congeners have been found to be absent (or present at levels below 0.05% total

PCBS

C12H I O-nC In
Cl (n-itoio)

NomgUgLaLurQ:

3 2
4G 2- 3- 4'
5 6 Q65 -

FAMM22 Of M-WQ-r f4g-0192-n em Ln E n vi ro n m e n1a I Ba m4gU:
C1 C1 C1 C1 C1 C1 C1
C4@ C4 C1
a C1
PCB #52 PCB #101 PCB #105
C1 C1 C1 C1 C1 C1 C1 C1 C1
0-0 C1 cl@@C, C,
C, C1
PCB#110 PCB #138 C1 PCB #180 C1
Fig. 1S. General formula of polychlorinated biphenyls and exLmples of major
congeners commonly found in environmental samples.

concentration) from technical PCB mixtures (Schulz et al., 1989). Unique properties,

including thermal stability and resistance to oxidation, resulted in the use of PCBs as

adhesives, heat transfer fluids, wax extenders, hydraulic fluids, lubricants, flame

retardants and as dielectric fluids in transformers and capacitors.

Different PCB formulations were graded and marketed according to their chlorine

content. Monsanto Chemical Corporation produced, for example, Aroclor 1221, 1232,

1254 and 1260, which contained 21, 32, 54 and 60 percent of chlorine by weight,

respectively. Many comparable commercial PCB formulations have been produced by

different chemical companies in several countries including Kanegafuchi Chemical Co. in

Japan (Kaneclor), Prodelec in France (Phenoclor), Bayer in West Germany (Clophen),

Deutchen Solvay Werken in East Germany (Orophene), Caffaro in Italy (Fenclor) and

Soval in the U.S.S.R. (Sovol and Sovtol) (Onuska & Comba, 1980). It has been
reported that between 1930 and 1975 the U.S.A. production of PCB:s was 570xlO3 tons

(U.S. Environmental Studies Board, 1979) whereas the total worldwide production of
PCBs through 1980 was estimated to be 11 OOx 103 tons (Erickson, 1986). In 1977, the

major U.S. producer, Monsanto Chemical Corporation, ceased manufacturing PCBs

partly due to their widespread detection in the environment. A recent estimation,

however, indicates that more than two-thirds of the cumulative world PCB production

may still be in use mainly in older transformers and capacitors (Tanabe, 1985).

Dr. Soren Jensen, a Swedish environmental chemist, first reported the presence of

several unknown peaks that interfered with quantitative determinations of DDT in

environmental samples (Jensen, 1966); those peaks were soon identified as a complex

mixture of PCBs by GC and GC/MS (Widmark, 1967). The earliest analyses of PCBs,

e.g. before 1980, were performed with packed columns. The results of these studies

have provided valuable information on hot spots and general trends in concentration

distributions; however, because of the low-resolution chromatograms most of the

information regarding individual congeners, which are important when determining

source identification, sink, toxicity and biological uptake or depuration, was limited. The

importance of considering individual PCB congeners in view of their differences in both

toxicity and physico-chemical properties is well recognized (see, for example, Shaw &

Connell, 1984; Opperhuizen et al., 1985, 1988; Tanabe er al., 1987b, 1987c). Although

a complete separation of all 209 PCB congeners with a single gas chromatographic run

has not been achieved yet, the introduction of the capillary column greatly improved the

separation of individual congeners.

The synthesis and chromatographic properties of all 209 PCB congeners have been

reported (Mullin et al., 1984). Certain PCB congeners are considered to be the most toxic

because they can attain a planar stucture similar to the highly toxic dibenzo-p-dioxins and

dibenzofurans (McKinney et al., 1976, 1985; Hansen, 1987; McFarland & Clarke,

1989). Because of their environmental significance, these PCB congeners will be

discused separately in Chapter IV.

PCBs are ubiquitous contaminants of the global environment. The physicochemical

proper-ties of these components vary widely depending on the number and position of

chlorine atoms in the biphenyl rings. In general, vapor pressure, water solubility and

biodegradability decrease with increasing number of chlorine atoms, whereas lipophilicity

and adsorption capacity show a reverse trend (Tanabe et al., 1984). Large variations of

PCB compositions are found in different environmental compartments resulting from this

wide range of properties. PCBs have been found in air, water, soil and sediment samples

throughout the world (e.g. Atlas & Giam, 198 1; Tanabe et al., 1983a). Nearly all marine

plant and animal specimens, fish, mammals, birds (especially fish-eating birds), bird eggs

and humans have measurable PCB concentrations (Tanabe et al., 1983b, 1986, 1987c).

In general, PCBs are detected in parts per billion (ppb) in organism, soil and

sediment samples and in parts per trillion (ppt) in water samples; however, concentration

levels vary over a large range from highly polluted to pristine.

Distribution and Occurrence in Galveston Bay

A variety of organochlorine residues have been determined in organisms, e.g.

bivalves and various species of fishes and birds, sediment and water samples, from the

Galveston Bay area. PCB congeners were one of the most commonly found compounds

in Galveston Bay samples (Table 3).

The ubiquity of PCBs in Galveston Bay was demonstrated by Fox (1988). Oyster

samples were collected from four different sites. PCBs were detected in every sample

analyzed during the study. Concentrations measured in oysters collected near the

Houston Yacht Club were higher than the levels found in samples from Hanna Reef,

Todd's Dump and Confederate Reef areas.

A number of different species of fish were also analyzed for PCBs. Concentrations

ranged over two orders of magnitude. Finfishes such as mullet, croaker and Florida

pompano, collected in the vicinity of a power plant (Houston Lighting and Power

Company) near the upper Trinity Bay, contained PCB concentrations in the range of 50-
500 ng g- I (Strawn er al., 1977). Lower concentrations were reported in juvenile
croakers (9-43 ng g-1; Stahl, 1980). Similar ranges to those published by Strawn et al.

(1977) were reported for tidewater silverside, sheepshead minnow and striped mullet

(King, 1989a, 1989b). These fishes are the food source of some birds such as black

skimmer and olivaceous cormorant.

A few waterbirds, e.g. olivaceous and double-crested cormorants, laughing gulls and

black skimmers, nesting in the upper Galveston Bay were also analyzed for chlorinated

hydrocarbons (King & Krynitsky, 1986; King et al., 1987). Average concentrations and

concentration ranges encountered in these birds were similar. PCB average

TABLE 3
Polychlorinated Biphenyl Concentrations in Samples from the Galveston Bay Area. Except Where Indicated, Concentrations in Organisms Are
Expressed in ng g-1 on a Wet-Weight Basis. Concentrations in Sediment and Water Samples Are Expressed in ng g-I, on a Dry-Weight
Basis, and in ng 1-1, Respectively.

Location Sample PCBs Range Reference

Yacht Club oysters 966 120-4,025 Fox, 19880)
Todd's Dump oysters 155 47.4-283 Fox, 19880)
Confederate Reef oysters 131 94.7-171 Fox, 19880)
Hanna Reef oysters 59.4 32.5-107 Fox, 19890)
Trinity Bay fish 50-160 Strawn et al., 1977
fish 50-500 Strawn el al., 1977
fish 60-150 Strawn et al., 1977
Galveston Bay fish 9-43 Stahl, 1980
crab 18-42
Galveston Bay fish 310 70-540 King, 1989a(2)
Galveston Bay fish 350 100-620 King, 1989b(2)
Galveston Bay cormorants 6,990 2,600-24,000 King & Krynitsky, 1986(2)
gulls 4,210 1,500-11,000
skimmers 3,880 800-11,000
Galveston Bay cormorants 1,580 1,100-3,300 King et al., 1987(2)

TABLE 3
(continued)

Location Sample PCBs Range Reference

Houston Ship Channel sediments 3,250 Salch & Lm 1976
Texas City Channel sediments 2,860 Salch & Lee, 1976
Galveston Bay sediments 15-68 Stahl, 1980
San Luis Pass sediments 0.52 0.25-0.78 Murray et al., 1981 a
Morgan's Point sediments 1.5 <0. 14-3.3 Murray et al., 198 1 b
Trinity 'Bay sediments 1.2 <0. 14-7.1 Murray et al., 198 1 b
Texas City Channcl sediments 2.8 <0. 14-5.6 Murray et al., 19 8 1 b

Galveston Bay water 2-15 Stahl, 1980
Morgan's Point water 1.1 <0.0 1 -4,6 Murray et al., 198 1 b
Trinity Bay water 1.8 <0.0 1 -4.1 Murray et al., 198 1 b
Texas City Channel water 18 <0.0 1 -70 Murray et al., 198 lb

n.d.= not detected; (1) ng g- 1 on a dry-weight basis; (2) geometric mean

concentrations ranged from 1,580 to 6,990 ng g-1, respectively. These concentrations are

one order of magnitude higher than concentrations listed on Table 3 for Galveston Bay

fish samples. Since these fish-eating birds are at the top of an aquatic food chain, a

bioaccumulation of organic contaminants is seen. With overall average PCB
concentrations of 4,170 ng g-1 in waterbirds and 330 ng g-1 in fishes, a biaccumulation

factor of 13 can be calculated for PCB; residues in Galveston Bay waterbirds.

Reports of PCB s concentrations in sediment and water samples from the Galveston

Bay area are limited. In general, PCB concentrations in sediments were in the <0.14 to
7.1 ng g- I range (Murray et al., 1981, 1981 b). Stahl (1980) reported a slightly higher
concentration range for PCBs in sediments (15-68 ng g- 1). These concentrations are two

to three orders of magnitude lower than those previously reported in dredged sediments

from the Houston Ship and Texas City Channels (Saleh & Lee, 1976). The samples

collected during that study corresponded to sediments disturbed by the construction of

underwater pipelines; therefore, they might represent sediments deposited before the

restrictions of the use of PCBs in the U.S.A. in the 1970s. Water samples collected at

different locations in Galveston Bay had PCB concentrations ranging from <0.01 to 70
ng 1-1 (Saleh & Lee, 1976). The higher PCB concentrations were measured near Texas

City.

Bivalve Uptake and Depuration Studies

There are published works reporting uptake and depuration of PCBs by a variety of

organisms; however, the number of studies involving bivalves are limited. The methods

generally used to expose bivalves to PCB congeners in the laboratory (e.g. Lowe et al.,

1972; Vreeland, 1974; Courtney & Denton, 1976; Pruell et al., 1986, 1987) are similar to

those mentioned for petroleum hydrocarbon exposures (Chapter II). Transplanting

bivalves from an uncontaminated area to contaminated areas or vice versa has also been

done in uptake and depuration studies (e.g. Calambokidis et al., 1979; Tanabe et al.,

1987a; Kannan et al., 1989; Sericano et al., in press).

In a laboratory study with blue mussels (Mytilus edulis) exposed to contaminated

sediments, Pruell et al. (1986) reported an equilibration time of about 20 days for four

PCB congeners although this bivalve failed to accumulate the highly chlorinated

congeners present in the exposure sediments after a 40-day exposure period. Sin-dlar time

scales were reported for the uptake of the lower molecular weight PCB congeners, i.e.

those congeners having 2, 3 or 4 chlorines in the molecule, by transplanted green-lipped

mussels (Perna viridis) in contaminated Hong Kong waters (Tanabe et al., 1987a).

Lower equilibration rates, i.e. more time, for high er-chlorinated PCB congeners are

reported. Vreeland (1974) suggested that, even after an equilibrium with the PCB

congener concentrations is attained, the total amount of PCB per oyster increases as the

oyster grows. Langston (1978) observed some differences in the depuration rates of

selected PCB congeners by bivalves (Cerastodernia edule and Macoma balthica). Di-, ni-

and tetrachlorobiphenyls, with half-lives ranging from 5 to 21 days, were depurated faster

than hexachlorobiphenyls and some pentachloro-biphenyls. Most of the

hexachlorobiphenyls did not show any decrease after 21 days. Pruell et al. (1986)

reported that about 50% of the total PCBs accumulated by exposed blue mussels (Mytilus

edulis) were lost after 40 days in clean seawater with half-lives for tri- to

hexachlorobiphenyls ranging from 16.3 to 45.6 days. Contrasting with this study,

Courtney & Denton (1976) reported that clams exposed to a PCB mixture, Aroclor 1254,

in the laboratory did not depurate the accumulated PCBs during a three-month period in

control seawater.

UPTAKE AND DEPURATION OF PCBs

Experimental Design, Sample Collection and Methods

The experimental design and sample collection used for the study of PCBs were the

same as those discussed for PAHs (Chapter II).

Ejaraction and saVlefiractionation of PCBs

As described in the previous Chapter, the analytical procedure used in the extraction,

fractionation and cleanup of PCBs in oyster, sediment and water samples, which is done

concurrently with the extraction, fractionation and cleanup of PAHs (Fig. 5, Chapter II),

was based on a method developed by MacLeod et al. (1985) with a few modifications that

proved to be equivalent or superior to the original technique. The important steps of this

method have been previously discussed for PAHs. The only differences for PCB

analysis are:

a. PCB congener quantitations were done using 4,4 dibromoocta-fluorobipbenyl

(DBOFB) and PCBs congeners 103 and 198 as internal standards. As previously

discussed, these standards were added at concentrations similar to those expected in the

samples for the compounds of interest.

b. After the final extract concentration to I ml, and before the addition of the GC

internal standard for GC-ECD analysis, a 250 ul fraction was reserved for further planar

PCB analyses.

Instrurwntal analysis
PCB s were analyzed by fused-silica capillary column GC-ECD (Ni63) using either a

Varian 3500 GC or a Hewlett Packard 5880A GC in the splitless mode. Capillary

columns, 30 meters long x 0.25 mm i.d. with 0.25 mm DB-5 film thickness, were

temperature-programmed from 100 to 1401C at 5'C min- 1, from 140 to 250'C at 1.5*C
min- I and from 250 to 3001C at I O'C min-1 with I min hold time at the beginning of the

program and before each program rate change. A hold time of 5 min was used at the final

temperature. Total run time was 94.33 min. Injector and detector temperatures were set

at 275 and 325)C, respectively. Helium was used as carrier gas at a flow velocity of 30.0
cm sec-1 at 1000C. Nitrogen or argon/methane (95:5) were used as make-up gases at a
flow rate of 20 ml min-1. The volume injected was 2 gl. The numbering of PCB

isomers, after Ballschmiter & Zell (1980), is as follows: numbers 1-3 represent mono-, 4-

15 di-, 16-39 tri-, 40-81 tetra-, 82-127 penta-, 128-169 hexa-, 170-193 hepta-, 194-2Q5

octa-, 206-208 nonachlorobiphenyls and 209 decachloro-biphenyl. A group of PCB

congeners (i.e. 8, 18, 28, 44, 52, 66, 101, 105, 110, 118, 128, 138, 153, 170, 180,

187, 195, 206 and 209) were quantitated against a set of authentic standards, which were

injected at four different known concentrations to calibrate the instrument and to

compensate for non-linear response of the electron capture detector. The remaining PCB

congeners were quantitated by comparison to a single reference congener of the same

degree of chlorination injected at four different known concentrations. The reference PCB

congeners used for quantitation were 8, 28, 52, 101, 138, 170, 195, 206 and 209 for di-,

tri-, tetra-, penta-, hexa-, hepta-, octa-, nonachlorobiphenyls and decachloro-biphenyl ,

respectively. Tetrachloro-m-xylene (TCMX) was used as the GC internal standard to

calculate the recoveries of the internal standards. The detection limits for organochlorines

and individual PCB isomers, calculated on the basis of 15 g (wet weight) tissue and 50 g

(wet weight) sediment sample sizes with 0.2% by volume of the extract injected into the
GC-ECD, were 0.25 and 0.02 ng g- I dry weight for oysters and sediments, respectively.

AncilLarypararneters

Methodologies for the sediment grain-size analysis and extractable lipids percentage

were discussed in the materials and methods section of Chapter II.

SwistictV wialysis

The statistical analyses performed on the PCB data were previously discussed in the

materials and methods section of Chapter Il.

Uptake of PCBs by Transplanted Oysters

In the following sections, Ship Channel, Hanna Reef, transplanted Hanna Reef-to-

Ship Channel, transplanted Ship Channel-to-Hanna Reef and relocated Hanna Reef-to-

Ship Channel-back to-Hanna Reef oysters are refered as SC, HR, HRSC, SCHR and

HRSCHR oysters, respectively.

Average concentrations of the predominant PCB congeners found during the first part

of this experiment in SC and HRSC oyster, sediment and water samples are reported in

Tables A-6 and A-7 (Appendix). Total PCB concentrations in indigenous Ship Channel

oysters were fairly constant during the seven-week uptake period with values fluctuating
between 960 and 1,500 ng g-1. In contrast, concentrations of total PCBs in transplanted
HRSC oysters increased from 30 ng g- I to 830 ng g- I after the 48-days exposure period

to the Ship Channel conditions. Typical PCB chromatograms of extracts obtained from

transplanted HRSC oysters during the uptake phase of this study are shown in Fig. 16.

Pentachlorobiphenyls accumulated to the highest concentrations in HRSC and native

SC oysters (Fig. 17). In comparison, practically no octa-, nona- or decachloro-biphenyls

were de tected in either oyster group. Not all the PCB homologs measured in transplanted

oysters reached the concentration encountered in indigenous individuals by the end of the

first phase of this experiment. While there were not statistically significant differences

HRSC-3

UL HRSC-7

HRSC-17

HRSC-30
ux't@

HRSC-48

Fig. 16. Examples of high-resolution gas chromatograms of Hanna Reef oysters
transplanted to the Ship Channel area during different stages of the 48-day exposure
period. PCB congeners are numbered according to Balischmitter & Zell, 1980.

Ship Channel Site 0 day Ship Channel Site 17 days
100 low
"o 0 Henna Reer Oysters 0 Ban" Red Oysters
N Ship Channel Oysters 0 Ship channel Oysters
we W

V 700 700

SM Soo

MW 40
me Z' 3W
C

C r
Soo 100

0 0
1 2 3 4 S 6 7 8 9 10 1 2 3 4 5 6 7 8 9 10

Level of Chlorination Level of Chlorination
1000 Ship Channel Site - 3 days 1000 Ship Channel Site - 30 days
E Henna Reer Oysters - M Hones Reer oysters
-r 9W 900
0 Ship Channel Oysters It 0 SbIpChaosel Oysters
Soo so
700 700
GN at 600
Sail Soo
0 400 400
300 300
200
W
A
100 100
U 0 Idah Ak 0 AFAM X.T.In
1 2 3 4 5 6 7 3 9 10 1 2 3 4 5 6 7 a 9 10
Level or Chlorination Level of Chlorination

Ship Channel Site 7 days Ship Channel Site 48 days
1000 a Hanna RetrOysters 1000 a Hansa Reeroysters
900
0 ship Channel oysters It 2 Sbipchaeftel oysters
$00

V 700 700
t 600 600
SOD
0 400 Cc 400
Z
300 300

200 200

100 100

0 0
1 2 3 4 5 6 7 8 9 10 1 2 3 4 5 6 7 8 9 10

Level or Chlorination Level or Chlorination

Fig. 17. Concentrations of polychlorinated biphenyl congeners, grouped by level of
chlorination, in transplanted Hanna Reef and indigenous Ship Channel oysters during
the 48-day exposure period near the Houston Ship Channel. Ship Channel Oysters
were not sampled on day 7.
L L

(alpha = 0.05) in the total tri- and tetrachlorobiphenyl concentrations measured in HRSC

and SC oysters, significant differences were observed in the total concentrations of penta-

and hexachlorobiphenyls. Concentrations of these two homolog groups in transplanted

HRSC oysters, at the end of the uptake period were about 30 and 50% lower than the total

concentrations measured in indigenous SC oysters, respectively.

Uptake and depuration curves observed for different PCB congeners are shown in

Fig. 18. Some of these represent coeluting congeners, for example PCBs 101 and 90 or

PCBs 110 and 77. However, the first congeners listed, i.e. 101 and 110, would be

expected to be highly dominant over the others. For example, in one of the most common

PCB mixtures, Aroclor 1254, the contribution of the PCB congener 101 to the total

101/90 peak is close to 90%; similarly, PCB congener 110 contributes almost 100% of

the total 110177 peak (Schulz et al., 1989). Therefore, it is assumed that the uptake and

depuration curves represent the first PCB congener; although all the co-eluting congeners

are indicated. When comparing the concentrations of individual PCBs measured in

transplanted and indigenous oysters after about one month of exposure, the concentrations

of the lower-chlorinated congeners, i.e. tri- and tetra- chlorinated biphenyls, in HRSC

oysters were similar to the levels encountered in indigenous individuals

Although increasing trends in the concentrations of all the predominant PCB

congeners were observed in HRSC oyster tissues, the concentrations of the higher-

chlorinated biphenyls, i.e. penta-, hexa- and heptachlorobiphenyls, did not always

reached full equilibrium with the levels measured in SC oysters. This results in qualitative

as well as quantitative differences between the PCB profiles in HRSC and SC oyster

samples at the end of the exposure period. The total PCB concentration at the end of the
exposure period in HRSC oysters (830 ng g-1) was about 25% lower than the levels
encountered in SC oysters (1,100 ng g-1). This is clearly shown in Fig. 19 where the

concentrations of selected PCB congeners measured at the end of the seven-week uptake

2,4,4' Trichlorobiphenyl (IUPAC No 28) 2,21,4,51 Tegrachlorobiphenyl (FUPAC No 49)

628

C
owl Hanna Red oysters Hanna Reef Oysters
Ship Channel Oysters
Ship Channel Oysters
U

a Ill 24 30 40 56 60 79 Of 90 too 0 Ill 20 30 46 50 60 70 so 90 too

Time (days) Time (days)

2,21,5,51 Tetrachlorobiphenyl (IUPAC No 52) 2,31,41,5 Tetrachlorobiphenyl (IUPAC No 70)

foo.
V V

flanna Reef Oysters Hanna Reef Oysters
Ship Channel Oysters Ship Channel Oysters
U

0 Is 20 30 40 58 60 79 so 90 too 0 to 20 30 40 50 60 1@ as to too

Time (days) Time (days)

Fig. 18. Concentrations of selected polychlorinated biphenyl congeners in tissues of Hanna Reef and Ship Channel oysters
during exposure to the Ship Channel area contaminant levels and following transplant to the Hanna Reef area.

Montilla

2,4,4',S Tttrachlorobiphenyl (TUPAC No 74) 2,2',4,5,5' & 2,2',394',S Pentachlorobiphenyis
100 1000 (IUPAC No 101 & 90)

too

.2
I;

Ilanna Reef Oysters Hanna Reer Oysters
c
Ship Channel Oysters 0 : , Ship Channel Oysters
U U

0 10 20 30 40 so 60 70 80 90 too 0 to 20 30 40 so 60 ?o so 90 1 0

Time (days) Time (days)
2,3,3',4',6 Pentachlorobiphenyl (IUPAC No 110) & 2,3',4,4'S Pentachlorobiphenyl (IUPAC No 118)
1000. 3,3'4,4' Teirachlorobiphenyl (IUPAC N.) 77) 1000.

fool

cc w
-S 10
C
.2

Hanna Reef Oysters Hanna Reer Oysters
Ship Channel Oysters Ship Channel Oysters
U

a 10 20 30 40 so 60 70 8 0 90 too 0 10 20 30 40 so 60 70 so 90 100

Tit me (days) Time (days)

Fig. 18. (Continued)

ON
f-A

2,2',3,4,4',S' & 2,3,3',4,5,6 Hexachlorobiphenyis 2,20,3,4',S',6 Hetachlorobiphenyl ("AC No 149) &
(JUPAC No 138 & 160) 2',3,4,4',S Pentachlorobiphfnyl (IUPAC No 123)

100. 100
V

c 10, C

ILI

IlannakeefOyst Ilanna Reef Oysters
er3
0
Ship Channel Oysters Ship Channel Oysters
U U

0 10 20 30 40 50 60 70 90 90 100 0 10 20 30 40 50 60 70 so 90 100

Time (days) Time (days)
2,2',4,4',5,5' & 2,2',3,3',4,6 Ilexachlorobiphenyis 2,2',3,4,4',5,6 Heptachloroblphenyl (JUPAC No 180)
1000. (IUPAC No 153 & 132) 1000.

L. 100, 100,
V

10 10

Ilanna ReefOyster3 Ilanna Red Oysters
Ship Channel Oysters Ship Channel Oysters
U U
.I IN P I @ - I i @ @ , - .
0 10 20 30 40 50 60 70 so 90 100 0 10 20 30 40 SO 60 70 so to M

Time (days) Time (days)

Fig. 18. (Continued)
I Red Oyst

llarin.Reef@0

Ship Channel Oysters
2,2 "8'
2,
3::: (26)
Hanna Reef Oysters
2,4,4' '2:1
2,2',3,4 (4
2,2',,3,5' (44)
2,2',4,1' (4
(5 9)
2,2',5,5' 2)
2,3',4'5 (70)
2,4,4',5 (74)
2,2',3,3',6 (84)
2,2',3,4,5' (87)
2,2',3,4',6 (91)
2,2',3,5,5' (92)
2,2',3,,S',6 (95)
2,2',3',4,5 (97)
2,2',4,4',5 (99)
2,2',4,5,5'(101) ..........
2,3,3',4,4' (105)
2,3,3',4',6 (110)
2,3',4,4',5 (118)
2,2',3,4,4',5' (138)
2,2',3,4',5',6 (149)
2,2',4,4',5,5' (153) . ..... ....
2.2',3,4,4',5,5' (180)

0 20 40 60 80 100 120

Concentration (ng/g, dry wt.)

Fig. 19. Concentrations of selected polychlorinated biphenyl congeners in tissues of
Hanna Reef and Ship Channel oysters at the end of the 48-day exposure period.

period are presented. In general, the higher the number of chlorines substituted in the

biphenyl molecule, the larger the difference between the concentrations encountered in

HRSC and SC oysters. Typically, these differences, in percentages, ranged from -20%

to 20% for the lower molecular weight congeners, and from 40% to 60% for the higher

molecular weight PCBs (Fig. 20). A negative percentage indicates a higher concentration

in HR oysters compared to SC oysters, i.e. PCB congeners 28, 41, 44 and 91. The

predominant PCB congeners in HRSC individuals were l53(6)/l32(6Ml0(5)/77(4),

95(5),52(4), 101(5)/90(5), 118(5) and 70(4) compared to 153(6)/132(6), 110(5)/77(4),

101(5)/90(5), 95(5), 118(5), 52(4) and 138(6)/160(6) in SC oysters (the "/" indicates co-

eluting congeners; the numbers given in parentheses indicate the level of chlorination).

Combined, these congeners accounted for more than 40% of the total PCB load.

This study confirms previously published reports which indicate that the less

lipophilic congeners reach equilibrium concentrations, during both uptake and depurution,

at faster rates than the more lipophilic compounds (Ellegehausen et al., 1980; Bruggeman

et al., 1981; Tanabe et al., 1987a), For example, there is an initial enrichment of the

lighter PCB fraction, e.g. congeners 95 and 52, in HRSC oysters. The concentrations of

these congeners, bowever, reached constant concentrations while the concentrations of the

more lipophilic congeners, e.g. 10 1 and 118, continued increasing. If the oysters were

allowed enough time, the final PCB distribution in HRSC oysters would probably have

approximated the distribution observed in SC oysters. Despite the differences observed in

equilibration rates of the various congeners, the composition of PCB homologs, in both

oyster populations, were largely dominated by penta-, tetra- and hexachloro-biphenyls

and had low concentrations of octa-, nona- and decachloro-biphenyls.

The dominant PCB congeners and homolog distribution encountered in newly and

chronically contaminated oysters at the end of the uptake period of this study are similar to

those reported for benthic invertebrates (Macoma balthica and Arenicola marina) and

difference

2,2',5 (18)
0
2,3',5 (26)
2,4,4' (28)
PO ................. .. . .............................. ..
CL 2,2',3,4 (4l)-
2,2',3,5' (44)-
rA = N
CL
2,2',4,5' (49)-
2,2',5,5' (52)
2,3',4'5 (70)-
2,4,4',S (74)-
2,2',3,3',6 (84)-
............................. ...............

2,2',3,4,5' (87)
En
2,2',3,4',6 (9l)-
Un 2,2',3,5,5' (92)-
0
2,2,3,5 6 (95)
2,2',3',4,5 (97)
2,2',4,4',5 (99)-
2,2',4,5,5' (101) . . .... . .... .................. .... . ................ ....... - - - - - -------
2,3,3' 4,4' (105)-
2,3,3':4',6 (110)-
2,3',4,4',5 (118)-
0.a 0 1
2,21,3,4,41,5, (138) . ............ ..... ..... ......R................................

Cr
2,2',3,4',5',6 (149)
2,21,4,41,5,51 (153)
Co
j...........
2,2',3,4,4',5,5' (180)]

sediments from the Dutch Wadden Sea where 10 1, 118, 138, 149, 153, 180 and 187, and

15, 18, 28, 118, 138, 153, and 187 were the dominant PCB congeners, respectively

(Duinker et al., 1983). Dominant PCB congeners in benthic polychaetes (Nephrys spp.)

from the southern North Sea were 118, 138, 149, 153 and 180, while in sediments the

highest concentrations corresponded to congeners 15, 18, 118, 138, and 153 (Boon et al.,

1985). Recently, Niimi & Oliver (1989) reported that the 10 most common congeners

detected in trout and salmon from Lake Ontario were 101, 84, 118, 110, 87/97, 153, 138,

149 and 180.

PCB congeners in Ship Channel sediments were dominated by pentachloro-biphenyls

and, to a lesser extent, by hexa- and tetrachloro-biphenyls (Fig. 21). Combined, these

three homologs represented more than 90% of the total sedimentary PCB load. Dominant

PCB congeners in sediments were 110(5)/77(4), 138(6)/160(6), 101(5)/90(5), 153(6)/

132(6) and 52(4). Each of these compounds accounted for more than 5% of the total

PCB load in the average sediment sample. This sedimentary PCB distribution is similar

to the distribution profiles encountered in HRSC and SC oysters.

Comparatively, PCB concentrations measured in water samples were significantly

lower (Fig. 21). The homolog PCB group with six chlorines represents the largest

portion of total PCBs in water, mainly because of the relatively high concentrations of

PCB 138 and 153.

Depuration of PCBs by Newly and Chronically Contaminated Oysters

When relocated to the Hanna Reef area, Hanna Reef and Ship Channel oysters

showed statistically significant depuration of total PCBs. Tables A-8 and A-9 (Appendix)

list the average concentrations of predominant PCB congeners in HRSCHR and SCHR

oysters. Also listed are the average concentrations encountered in Hanna Reef sediments.
Total PCB concentration decreased from 830 to 380 ng g- I and from 1,100 to 730 ng g- I

Ship Channel Sediments
20-

10-

2 3 4 5 6 9 10

Level of Chlorination

Ship Channel Seawater
3.0

2.5

2.0

C
ro-1 1.51
1.0

0.5,

0.0
2 3 4 5 6 7 8 9 1*0

Level of Chlorination

Fig. 21. Concentrations of polychlorinated biphenyls, grouped by level of
chlorination, in Ship Channel sediment and seawater samples.
I . ii,

in HRSCHR and SCHR oysters, respectively, after seven weeks at the Hanna Reef

location.

The concentrations of PCBs, grouped by level of chlorination, in both oyster

populations at different stages during the 50 days depuration period are shown in Fig. 22.

Different PCB congeners were depurated at different rates by SCHR and HRSCHR

oysters. Also, a marked decrease in the depuration efficiencies of the bioaccumulated

homologs with increasing number of substituted chlorines was observed in both groups

of oysters. For example, three-, four-, five- and six-chlorine substituted homologs

decreased 80, 70, 47, 20%, in HRSCHR oysters, and 73, 50, 24, 17%, in SCHR

individuals, respectively. This differential depuration of the accumulated PCBs can be

observed in Fig. 23 where the concentrations of selected PCB congeners in HRSCHR

and SCHR oysters at the end of the depuration period are shown. This retention of the

highly lipophilic congeners was more evident in chronically contaminated oysters.

Because of the incomplete depuration, the total PCB concentration in Hanna Reef

oysters, after 50 days, remained one order of magnitude higher than the original levels
(380 ng g- I versus 30 ng g- 1). The concentrations of homologs and selected PCB

congeners in Hanna Reef oysters before the transplantation to the polluted Ship Channel

site and 50 days after their relocation to the Hanna Reef area are shown in Fig. 24. The

distribution of PCBs in originally uncontaminated, i.e. HR, oysters shows a relafive

predominance of five- > four- > six-chlorine substituted homologs whereas the

predominant homologs in HRSCHR oysters were those having five, six and four

chlorines.

Depuration of PCBs by HRSCHR and HRSC oysters were approximately exponenfial

(Fig. 18). The clearance rates for high molecular weight PCBs were significantly slower

in both oyster populations. Transplanted HRSCHR oysters depurated most of the

recently incorporated PCB congeners at a faster rate than SCHR oysters. Detailed

Hanna Reef Site - 0 day Hanna Reef Site - 19 days
1000 0 Boom* Reer oysters 1000, M Hanna Reef Oysters
9w 900.
0 Ship Channel Oysters w Ship Channel Oysters
11100 No.

10 700 V 700-

6w 600

so Soo

400

300 "o
r
200 200
w w
C
too INTall U0 Joe
0 0
1 2 3 4 5 6 7 9 9 10 1 2 3 4 3 6 7 8 9 to

Level of Chlorination Level of Chlorination
Hanna Reer Site - 3 days 1000 Hanna Reef Site 30 days
Hanna Reef Oysters 1:1 900 � Hanna Reeroysters
Ship Channel Oysters i E Ship Channel Oysters
so
700 V 700
6oo Gi 600
at at
Soo Soo
400 400
300 30
C 200
200 w
100 0 100
AIIIIIIIIIIII - U 0
1 2 3 4 5 6 7 9 9 10 1 2 3 4 5 6 7 3 9 to
Level or Chlorination Level of Chlorination

Hanna Reef Site 6 days Hanna Reer Site 50 days
1000 0 Hanna ReefOysters 1000 0 Hamm ReerOysters
900 900
0 Ship Channel Oysters It 0 Ship Channel Oysters
s.
700 700

60 6"

Soo Soo.
r
0 400 400-
at 300 300-

200 ow 200-
C C
0 100 0 A-11
U U 100
0 0
1 2 3 4 3 6 7 8 9 10 1 2 3 4 5 6 7 8 9 10

Level of Chlorination Level of Chlorination

Fig. 22. Concentrations of polychlorinated biphenyl congeners, grouped by level of
chlorination, in back-transplanted Hanna Reef and transplanted Ship Channel oysters
during the 50-day depuration period in the Hanna Reef area.
LA
I
L.i L 4,

Ship Channel Oy sters
2,2',5 (18)
2,3',5 (26) Hanna Reef Oysters
2,4,4'(28)
2,2',3,4 (41)
2,2*,3,5' (44)
2,2',4,5' (49)
2,2',5,5' (52)
2,,3',4'5 (70)
2,4,4',S (74)
2,2',3,3',6 (94)
2,2',j3,4,5' (87)
2,2',3,4',6 (91)
2,2',3,5,5' (92)
2,2',3,,5',6 (95)
2,2',3',4,,S (97)
2,2',4,4',5 (")
2,2',4,5,5' (101)
2,3,3',4,4' (105)
2,3,3',4',6 (110)
2,3',4,4',5 (118)
2,2',3,4,4',S' (138)
2,2',3,4 5',6 (149)
2,2',4,4',5,5' (153)
2,2',3,4,4',S,S' (180)

0 20 40 60 80 100 120

Concentration (ng/g, dry wt.)

Fig. 23. Concentrations of selected polychlorinated biphenyl congeners in tissues of
Hanna Reef and Ship Channel oysters at the end of the 50-day depuration period.

Before Transplantation
2,2',@5 (18)
2,3',5 (26) After Depuration
2,4,4'(28)
2,2',3,4 (41)
2,2',3,5' (44)
2,2',4,5' (49)
2,2',5,5' (52)
2,3',4'5 (70)
2,4,4',5 (74)
2,2',"',6 (84
2,2',3,4,5' (87)
2,2',3,4',6 (91)
2,2',3,5,5' (92)
2,2',3,5',6
(95
2,2',3',4,5 (97)1
2,2',4,4',5 (99)
2,2',4,5,5' (10 1)
2,3,3',4,4* (105)
2,3,3',4',6 (110)
2,3',4,4',5 (119)
2,2',3,4,4',5'(138)
2,2',3,4',5',6 (149)
2,2',4,4',5,5' (153)
RIM,,

2,2',3,4,4',5,5' (180)

Concentration (ng/g, dry wt.)

Fig. 24. Comparison of the concentrations of selected polychlorinai-ld biphenyl
congeners measured in tissues of Hanna Reef oysters before exposure to the Ship
Channel contaminant levels and after depuration at the Hanna Reef site.

discussion of the PCB biological half-lives and related kinetic parameters are presented in

Chapter VI. The estimated half-lives of selected PCB congeners in Hanna Reef and Ship

Channel oysters are listed in Table 4 for comparison purposes. Calculated PCB biological

half-lives ranged from 14 to 200 days in Hanna Reef oysters and from 18 to 595 days in

Ship Channel oysters. Similarly to previous studies, the biological half-lives of PCB

congeners increased with the number of chlorine atoms in the biphenyl rings. With the

exception of the values reported by Tanabe et al. (1987a), the estimated half-lives for

different PCB congeners during this study were comparable to most of the values

previously reported for a number of different organisms. In Tanabe's study, most of the

PCB congeners were depurated with extremely short half-lives, i.e. less than 10 days.

The average homolog concentrations in Hanna reef sediments (Fig. 25) were one

order of magnitude lower than the levels encountered in the Ship Channel area.

Comparing sediment samples from the Ship Channel area to the Hanna Reef location, it is

possible to observe some differences in the relative contribution of the different homologs

to the total sedimentary PCB load. While the average PCB distribution in Ship Channel

sediments is largely don-@dnated by pentachlorobiphenyls, sediment samples from the

Hanna Reef area show a slight predominance of hexachlorobiphenyls.

CONCLUDING REMARKS

Low molecular weight PCB congeners, i.e. those substituted with two, three and four

chlorines, were rapidly accumulated by transplanted oysters to final concentrations that

were not statistically differentiable from the concentrations encountered in indigenous

oysters. In most cases, these concentrations were reached in 30 days. Comparatively,

the bioaccumulation of higher molecular weight PCB congeners was much slower. As a

consequence of this slower uptake rate, the high molecular weight PCB congeners did not

TABLE 4
Biological Half-Lives (Days) of PCBs in Hama Reef and Ship Channel Crassostrea vir&ka Oysters.

Congener Hanna Ship Pruell et al Tanabe et al Bruggeman et al.
Reef Channel (1986) (1987a) (1981)
oysters oysters mussels mussels fish

2.2',5 (18) 14 19 6 14
2,3',5 (26) 22 22 - -
2,4,4' (28) 17 34 16 7 -
2,2',3,3' (40) 14 18 - 4 -
2.2'.3,4 (41) 23 55 - 5 -
2,2',3.5' (44) 27 45 - 6 -
2,2',4,5' (49) 39 61 - 5 -
2,2',5,5' (52) 27 45 - 6 46
2,3',4'.5 (70) 30 58 - 6 69
2,4,4'.5 (74) 30 47 - 7 -
2,2',3.3',6 (84) 37 80 - 6
2,2',3,4.5'/2.3,4,4',6 (87/115) 55 132 - 5 -
2,2',3,4',6 (91) 25 50 - 5 -
2,2',3,5,5' (92) 31 63 6 -
2,2',3',5',6 (95) 45 95 5 -
2.2',4,4',5 (99) 49 91 - 6 -
2,2',4,5,5'/2,2'.3,4',5 (101/90) 55 116 28 6 -
2,3,3',4,4' (105) 63 120 - 6 -
2,3,3',4',5 (107) 30 46 - - -
2,3.3',4',6/3.3',4,4' (1 lOn7) 45 103 - 6 -
2,3',4,4',5 (118) 73 299 - 7 -
2,21,3,3,,4,4' (128) 76 229 37 7 -
2,2',3,4,4',5/2,3,3',4,5,6 (138/160) 200 595 - 8 -
2.2'.3.4',5,5' (146) 111 239 - -
2,2',3,4',5',6/2',3,4,4'.5 (149/123) 130 439 - 7 -
2,2',4,4',5,5/2,2',3,3',4,6' (153/132) 51 102 46 9 -
2,2',3,3',4',5,6 (177) 52 145 - 11 -
2,2'.3.3'.5,5',6 (178) 52 91 - 8
2.2',3,4.4',5.5' (180) 50 142 - 7
2,2',3,4',5,5',6 (187) 70 258 - 10

2.5 Hanna Reef Sediments

2.0-

US
C

I
C 1.0
Ix
0-5 T
C

0.0
2 3 4 7 10

Level of Chlorination

Fig. 25. Concentrations of polychlorinated biphenyls, grouped by level of
chlorination, in Hanna Reef sediment samples.

attain equilibrium concentration by the end of the exposure period in this study and

statistically significant differences were evident between SC and HRSC oysters. In

general, the higher the number of chlorines substituted in the biphenyl molecule, the larger

the difference between the concentrations found in HRSC and SC oysters. In spite of

their lower uptake rates, pentachlorobiphenyls were the PCBs accumulated to the highest

concentrations in HRSC and SC oysters. In comparison, practically no congeners having

eight, nine or ten chlorines were accumulated by either oyster group. At the end of the

seven-week exposure period, the final distribution profiles of PCB homologs and

individual congeners in both transplanted (HRSC) and indigenous (SC) oysters were

similar to the profile encountered in sediment samples collected in the Ship Channel area.

When transplanted to the Hanna Reef location, both groups of oysters (i.e. HRSCHR

and SCHR) depurated the low molecular weight congeners at faster rates than the

clearance rates observed for the heavier PCBs. However, individual tetra- and

pentachlorobiphenyl congeners were depurated at a faster rate by HRSCHR than by

SCHR oysters. The concentration at the end of the 50-day depuration period measured in

FIRSCHR oysters was about one order of magnitude higher than the original level. In

both groups of oysters, the depuration efficiency decreased with the increasing number of

substituted chlorines in the biphenyl rings. This observed decrease in the clearance

efficiency is reflected in the estimated half-lives. In general, the less lipophilic congeners

reach equilibrium concentrations, during both uptake and depuration, at faster rates than

the most liphophilic PCB congeners.

CHAPTER IV

UPTAKE AND DEPURATION OF PLANAR PCB CONGENERS BY THE

AMERICAN OYSTER (CRASSOSTREA VIRGIMCA):A SPECIAL CASE OF PCBs

IN7MODUCnON

One of the objectives of this study was to evaluate the bioaccumulation of the highly

toxic planar PCB congeners, i.e. PCBs 77, 126 and 169, by bivalves under

environmental conditions. Most of the effort, however, was dedicated to the development

of a reliable technique for the isolation of non-ortho substituted tetra-, penta- and

hexachlorobiphenyl congeners that could be coupled to the existing cleanup procedures in

the laboratory.

This chapter serves two purposes. First, a new method for the isolation of the three

most toxic planar PCB congeners is presented and evaluated. As compared to previously

published methods for planar PCB analysis, this methodology saves both time and

materials, i.e. solvents, and eliminated the use of benzene, a highly carcinogenic solvent,

that requires extreme care in handling by the analyst. Second, the uptake and depuration

of these planar PCB congeners by transplanted oysters in Galveston Bay are determined

and discussed.

PLANAR PCBs: A REVEEW

Background Information

Of the 209 possible PCB congeners, only 20 have non-ortho chlorine substitutions in

the biphenyl rings. Some of these congeners can attain planarity, which makes them

sterically similar to the highly toxic dibenzo-p-dioxins and dibenzofurans (McKinney et

al., 1976, 1985; Hansen, 1987; McFarland & Clarke, 1989). Particularly important

within this group are the PCBs with no ortho, two para and at least two meta chlorines.

For example, congeners 3,3',4,4' tetrachlorobiphenyl (1UPAC No 77), 3,3',4,4',5

pentachlorobiphenyl (IUPAC No 126), and 3,3',4,4',5,5' hexachlorobiphenyl (1UPAC

No 169), shown in Fig. 26, are very potent mimics of the 2,3,7,8 tetrachlorodibenzo-p-

dioxin (TCDD) and 2,3,7,8 tetrachlorodibenzofuran (TCDF) both in P-450 induction and

toxic effects, e.g. body weight loss, dermal disorders, liver damage, thymic atrophy,

reproductive toxicity and immunotoxicity (Goldstein & Safe, 1989; Poland & Knutson,

1982; Safe, 1984, 1986, 1990; Tanabe, 1988). These planar PCBs are the most potent

pure 3-methylcholanthrene-type (3-MC-type) inducer congeners. Some studies have

indicated that not only non-ortho chlorine substituted PCBs but also some mono- and di-

ortho analogs of planar PCBs possess similar toxic potential (e.g. Robertson et al., 1984;

Safe, 1985; Bryan et al., 1987; Hansen, 1987; Olafsson et al., 1987; Tanabe et al.,

1987c; McFarland & Clarke, 1989).

In a recent review, Safe (1990) discussed the environmental and mechanistic

considerations behind the development of the Toxic Equivalent Factor (TEF) concept for

different PCBs. Safe proposed provisional TEF values of 0.01, 0.1 and 0.05 for planar

congeners 77, 126 and 169, respectively. Recently, the validation and limitations of these

factors have been reported (Safe, 1992).

PCBS
General Form 1a: C12HIO-ncin
@@Cl en = Ito 10)

Nomenclature:

3 2 2' 3'
4Q-Q 4'

Planar Conaeners:

C1 C1 C1 C1 C1 C1
C"(D --- OC, CIIL@@C' cllc:@@C'
C1 C1 C1
PCB #77 PCB #126 PCB *169

Related _TD-x:i.Q Compounds:
Cla 0'-'OC1
C, 0 C1 C 0 C1
2,3,7,&Tetrachlorodibenzo-p.dioxin 2,3,7,8-Tetrach1orodibefvofuran

Fig. 26. General formula of polychlorinated biphenyls. Three of the most toxic
planar PCB congeners, i.e. PCB 77, 126 and 169, are shown together with the
compounds they mimic in toxic effects.

Although these planar PCB congeners represent a small portion of the total technical

PCB mixtures (Duinker & Hillebrand, 1983; Kannan et al., 1987; Schulz et al., 1989),

monitoring these compounds is needed because of their high toxicity. However,

quantitation of individual non-ortho substituted PCB congeners is very difficult because

of their extremely low concentrations. Routine high-resolution capillary gas

chromatography analyses fails to separate some of these planar PCBs from other ortho-

PCB congeners, although this separation can now be achieved with more expensive and

complicated techniques such as multidimensional gas chromatography (Duinker et al.,

1988a).

During the last decade, a wide variety of different methodologies have been reported

for the separation of individual PCBs, according to the number of chlorines in the ortho

positions, using different adsorbents, such as. florisil and activated carbon. In general,

the existing methods for the separation of planar PCBs from other congeners use an

extremely large volume of eluant per sample, i.e. over 1000 ml (e.g. Huckins et al., 1980;

Stalling et al., 1980), involve a carcinogenic solvent, i.e. benzene, (e.g. Tanabe et al.,

1987; Hong & Bush, 1990;"Kuehl et al., 1991, or are extremely complicated for routine

analysis (e.g. Smith et al., 1984, Patterson Jr. et al., 1989).

Distribution and Occurrence in Galveston Bay

Although PCB congeners have been widely reported in Galveston Bay samples (Table

3) and have been one of the most commonly found chlorinated compounds in oyster

samples from Galveston Bay (Sericano et al, 1990a), the occurrence of planar PCB

congeners in this area have not, until recently, been reported (Sericano et al., 1992).

This study, which is discussed in greater details in Chapter VIII, reports the

occurrence of three highly toxic PCB congeners (PCBs 77, 126 and 169) in oysters

(Crassostrea virginica) fi-om different locations in Galveston Bay using a newly developed

carbon chromatographic method (Sericano et al., 1991). The highest concentrations of

planar PCB congeners in Galveston Bay were reported in samples collected near the area

where the Houston Ship Channel enters the upper Galveston Bay (2,000, 2,200 and 790
pg 9_1 for congeners 77, 126 and 169, respectively) and decreased seaward. The second

highest concentrations were encountered in samples from near the city of Galveston (500,
400 and 93 pg g- I for congeners 77, 126 and 169, respectively). The lowest

concentrations were measured in samples collected near Hanna Reef in East Bay (89, 110
and 89 pg g-I for congeners 77, 126 and 169, respectively). The general distribution of

planar PCB congeners in Galveston Bay clearly correlates high concentrations with highly

populated areas. The same correlation between urban centers and concentrations was

observed in Tampa Bay (Sericano er al., 1992).

Bivalve Uptake and Depuration Studies

The number of studies reporting the uptake and depuration of PCBs by different

bivalves is limited. Even more limited is the number of studies reporting the uptake,

persistency and release of highly toxic planar PCB congeners. In one of the first reports

regarding the bioconcentration of planar PCB congeners in lower aquatic organisms, e.g.

Green-lipped mussels (Perna viridis Linnaeus) and possible transfer through food chain to

higher animals, it was concluded that these compounds are highly bioaccumulated by

lower organisms and, because of their persistence, they may reach higher consumers,

including humans, in quantities of toxicological concern (Kannan et al., 1989).

UPTAKE AND DEPURATION OF PLANAR PCBs

Experimental Design, Sample Collection and Methods

The experimental design and sample collection used for the study of planar PCBs

were the same as those discussed for PAHs (Chapter 11).

Extraction and initial sarVIefractionation

The extraction, initial fractionation and cleanup of planar PCBs were done

simultaneously with the rest of the ortho-substituted PCBs. After the final extract

concentration to I ml, and before the addition of the GC internal standard for GC-ECD

analysis, a 250 gl fraction was reserved for the analysis of planar PCB congeners.

Before proceeding to the next step, PCB 81 was added to the extracts as an internal

standard.

Isolation ofplanar PCB congeners

The methodology to analyze planar PCBs in transplanted oyster tissues is published

elsewhere (Sericano et al., 1991). Glass chromatographic columns (10 mm i.d.) were

packed in methylene chloride. Two g of the adsorbent, a 1:20 mixture of activated AX-21

charcoal (Super-A activated carbon) and LPS-2 silica gel (Low-pressure silica gel, particle
size 37-53 gm, 450 m2g-1), were packed between two layers of anhydrous sodium

sulfate. The adsorbent mixture was carefully checked for interfering compounds by

running blanks with the solvent mixtures used to elute the column. Oyster tissue extracts

were sequentially eluted from the column with 50 ml of 1:4 methylene chloride and

cyclohexane, 30 ml of 9:1 methylene chloride and toluene, and 40 ml of toluene. The
flow rate through the column was 1.5 to 2.0 ml min- 1. The first two solvent mixtures

were collected as one fraction (f 1) and contained the bulk of PCB congeners. The second

fraction 02), containing the ortho unsubstituted PCB congeners with four, five and six

chlorines in meta and para positions, was concentrated to a final volume of 0. 1 ml, in

hexane, for GC-ECD analysis.

Instrumental analysis

Planar PCB congeners were analyzed by fused-silica capillary column GC-ECD
(Ni63) using a Hewlett Packard 5880A GC in splitless mode. Capillary columns, 30 m

long x 0.25 mm i.d. with 0.25 gm DB-5 film thickness, were tempera ture-programmed
from 100to 150OCatlOOC min-l and from 150to270OCat6'Cmin-I with I min hold

time at the beginning of the program and before the program rate change. A hold time of

3 min was used at the final temperature. Total run time was 30 min. Injector and detector

temperatures were set at 275 and 325'C, respectively. Helium was used as the carrier gas
at a flow velocity of 30.0 cm sec-1 at 100'C. Nitrogen or argon/ methane (95:5) were
used as the make-up gas at a flow rate of 20 ml min-1. The volume injected was 2 gl.

Planar PCBs were quantitated against a set of authentic standards that were injected at four

different known concentrations to calibrate the instrument and to compensate for a non-

linear response of the electron capture detector. Tetrachloro-m-xylene (TCMX) was used

as the GC internal standard to estimate the recoveries of the internal standards. The

detection limits for organochlorines and individual PCB congeners, calculated on the basis

of 15 g (wet weight) oyster tissue sample size with 0.2% by volume of the extract injected
into the GC-ECD, was 0.05 ng g- I dry weight.

Planar PCB Congener Analysis

Activated carbon has been previously used to separate chlorinated compounds based

on the degree of chlorination as well as molecular planarity (e.g. Jensen & Sundstr6m,

1974; Stalling er al., 1980). In the case of PCBs, for example, the planar structure is

related to the number of chlorines in the ortho positions. Based on the high surface area

of activated carbon and its selective adsorptive capacity of planar structures, this adsorbent

can be successfully used to isolate planar PCB congeners having four or more chlorines in

meta and para positions. Although PCB congeners with a decreasing number of ortho

substituted chlorines were differentially retained in the column (Stalling et al., 1980), all

the PCBs with at least one ortho chlorine were eluted by the first two solvent mixtures and

collected in one fraction. Thd mixture of I part of AX-21 activated carbon and 20 parts of

LPS-2 silica gel was relatively easy to pack and use.

The efficiency of the column was initially checked with a mixture of PCBs, Aroclor
1254 (5,000 ng ml-1), spiked with the four planar PCB congeners. These analytes were

added in triplicate to the Aroclor mixture at three different concentrations (20, 50 and 100
ng m.1-1). Fig. 27 shows the chrornatograms of spiked Aroclor 1254 (a), PCB congeners

recovered in the first fraction, i.e. 50 ml of 1:4 methylene chloride and cyclohexane

followed by 30 ml of 9:1 methylene chloride and toluene (b), and planar PCBs eluted in

the second fraction, i.e. 40 ml of toluene (c). Recoveries of planar PCB congeners are

reported in Table 5. Recoveries for the three highly toxic planar PCB congeners were

above 90%, whereas that for PCB 81 was slightly lower. Recoveries in the first fraction

of PCB congeners having one to four chlorines in the ortho-ortho' positions were, in all

cases, near 100%.

To investigate the efficiency of the column with environmental samples with high lipid

concentrations, dolphin blubber extracts were spiked with the same four planar PCB
congeners at a concentration of 50 ng g- I each. Total PCB concentration in the dolphin
blubber was 3,700 ng g-l. Fig. 28 shows the chromatograms of the spiked dolphin

blubber sample (a).as well as the ortho- and non-orthochlorine substituted PCB congeners

recovered in the first and second fractions, b and c respectively. Also, these planar PCB

congeners were isolated from other organochlorine compounds present in the blubber

-tg! 49 S7.
44

00 41/64
103 103 103

74 70
X 0
0- = a- 91
(IQ
@l (D r. 60 92
E3, = 84
e- CD =1
99

83
97
w CL 0 al --
= E4 w : lz _F;
Cl) CL 0 --1 -1 - 77
f) 0

00
0 i49
134
N)
C) C6
146
(IQ =s los
(1) 141/179
:3
ED

129 178 129/126
00 4n -126 1178
co CD 187
cr V) 183
0 7 128

174
tj 177
156/171
>
172
w N '= ISO
0 tz QQ
0 co 0
0 o :3 11 1
0 Cb _, -169 169
= 7 1,j 170/190
m fm tA 198 198
-198
< lul
0
C) a
w CL @4
9
@@103 @41/64
,4
gl 4 6; 7@i@49
'0!
141/179
F
9/
187
IS3 2
67 1 I'l'
174
77

103
Oxychlordan

V
oil Gamma-Chlordone
0

0
W 0 ;5wimmi-i D i e I d r i n

a, CL
149

-p,DD
0, @@@o-p-DDT P

0 00
OV co P-W DDT
V
129/126
0 187
103
S 128

0 (7,

0
Mirex
2L 169 @169
170/190
0 0
:3 198

00 0 =r
o
LGm7m.
V@- ,
@
77
:@149

W
206

extract, i.e. chlordane-related compounds and DDT and its metabolites DDD and DDE.

Spiked dolphin blubber samples also had excellent recoveries for these PCB congeners,

comparable to those calculated for the spiked Aroclor mixture (Table 5). Recoveries for

all the chlorinated hydrocarbons originally present in the dolphin blubber sample were

close to 100%. Only a negligible concentration of p-p'DDE, detected in the original
sample at a very high concentration (1,450 ng g-1) was present in the final fraction (Fig.

28c).

Overall, this method yielded higher or similar recoveries for PCB congeners 77, 126

and 169 than those reported by Kamops et al. (1979), Huckins et al. (1980), Smith et al.

(1984), Tanabe et al. (1987) and Hong & Bush (1990) at comparable concentrations

using either florisil or carbon chromatography on pure standard solutions or spiked

samples.

Uptake of Planar PCBs by Transplanted Oysters

Ship Channel, Hanna Reef, transplanted Hanna Reef-to-Ship Channel, transplanted

Ship Channel- to-Han na Reef and relocated Hanna Reef-to-Ship Channel-back to-Hanna

Reef oysters are refered as SC, HR, HRSC, SCHR and HRSCHR oysters, respectively,

in this and the following sections.

Concentrations of planar congeners in transplanted HRSC oysters were encountered at
very low concentrations, e.g. parts per trillion (pg g- 1) to parts per billion (ng g- 1). The

lowest concentrations corresponded to congener 3,3',4,4',5,5' (169), which was present

at concentrations near or below the detection limits. Fig. 29 illustrates both the

applicability of this technique to real environmentally contaminated samples and the
difficulties involved in the analysis of planar PCB's because of their extremely low

concentrations. The chromatograms correspond to an extract of indigenous SC oysters.

Fig. 29a shows the ortho substituted PCBs eluted in the first fraction with the two

TABLES
Recoveries of Four Planar PCB Congeners from Spiked Aroclor 1254 and Dolphin
Blubber Samples Using Activated Carbon:Silica Columns.

PCB congeners Aroclor 1254 Blubber Average
Level I Level H Level HI
20 ng ml- I 50ngml-l lOOngml-I 50 ng g-I

3,4,4',5 (81) 85�1.8 82�8.6 79�4.0 82�5.4 70�2.3
3,3',4,4'(77) 100�4.8 94�9.0 90�4.1 94�7.2 87�4.0
3,3',4,4',5 (126) 90.�0.6 96�5.5 94+' 2. 1 93�3.9 91�2.8
3,3',4,4',5,5'(169) 94�1.0 96�3.4 97�1.1 96�2.5 97�2.3

Cot-
t-

Fig. 29. Example of high-resolution gas chromatograms obtained from an extract of
indigenous Ship Channel oysters. PCB congeners are numbered according to
]Ballschn-@tter & Zell, 1980.

different solvent mixtures. Fig. 29b shows, for comparison, the planar congener

fraction, i.e. second fraction, at the same magnification as the chromatogram

corresponding to the first fraction and Fig. 29c shows the second fraction magnified 20

times.

Both, 3,3',4,4' tetraCB (77) and 3,3',4,4',5 pentaCB (126) exhibit fairly well-

defined uptake and depuration curves. Fig. 30 shows the concentrations of PCB

congeners 77 and 126 versus time during the uptake and depuration pe riods. The

concentrations of these two planar PCB congeners in HRSC oysters increased over the

seven week exposure period. PCB congener 77 reached a concentration similar to that

encountered in indigenous SC oysters within a month. The uptake of congener 126 was

slower and only approximated the concentration of SC oysters by the end of the exposure

period. Contrasting with planar PCBs 77 and 126, it was not possible to observe a clear

trend in the concentration of congener 169 versus time. This is mainly because of its

extremely low concentration.

The final concentrations of the accumulated congeners decreased as the number of

chlorines substituted in the biphenyl rings increased. This trend was also reported in

transplanted green-lipped mussels (Perna viridis Linnaeus) during an exposure experiment

in Hong-Kong waters (Kannan et al., 1989). Kannan et al. (1987) reported the

concentrations of these planar congeners in different commercial PCB mixtures. In

general, congener 77 is 1-2 and 3-5 orders of magnitude higher than congeners 126 and

169, respectively. Comparing these relative concentrations with those observed in

transplanted oyster samples, it appears that congeners 126 and 169 in oyster tissues were

enriched with respect to congener 77. The same observation was made by Kannan et al.
(1989). This is not. surprising since the log Kow (octanol-to-water coefficient) increases

with the number of chlorines substituted in the biphenyl rings (6.36, 6.89 and 7.42 for
congeners 77, 126 and 169, respectively; Hawker & Connell, 1988). In general,

3,3'4,4' Tetrachlorobiphenyl (IUPAC No 77)
20000-. Hanna Reef Oysters
Ship Channel Oysters

3LOOO

100 .. .............
0 10 20 30 40 50 60 70 go 90 100

Time (days)

3,3',4,4',5 Pentachlorobiphenvi (IUPAC No 126)
1000-.
Hanna Reef Oysters
Ship Channel Oysters

loo-,

10
0 10 20 30 40 so 60 70 so 90 100

Time (days)

Fig. 30. Concentrations of planar polychlorinated biphenyl congeners 77 and 126 in
tissues of Hanna Reef oysters during the uptake and depuration phases of the
transplantation experiments at Galveston Bay,

concentrations for congener 77 in oyster tissue were 3-5 and 10-12 times higher than

those measured for congeners 126 and 169, respectively.

Depuration of Planar PCBs by Newly Contaminated Oysters

When transplanted to the Hanna Reef area, exposed oysters slowly depurated the

concentrated planar congeners. These PCBs were still present at high concentrations,

relative to original HR oysters, by the end of the 50-days depuration period. Kannan et

al. (1989) also observed that the concentrations of these planar PCB congeners in

transplanted green-lipped mussels (Perna viridis Linnaeus), at the end of the exposure

period (32 days), were substantially higher than those found in native individuals.

Depuration of congener 77 was comparatively faster than the clearance rate observed

for congener 126. Calculation of half-lives and related kinetic parameters of these trace

organic pollutants will be discussed in greater details in Chapter VI. For comparison

purposes, the estimated biological half-lives of these toxic PCBs were 88 and 107 days

for congeners 77 and 126, respectively. These estimated values were significantly higher

than those reported by Kannan et al. (1989) for mussels (9 and 13 days, respectively).

However, it must be noted that, as previously discussed in Chapter III, all the biological

half-lives reported for different PCB congeners in that transplantation study (i.e. Tanabe

et al., 1987a; Kannan et al., 1989) were significantly lower than the estimated half-lives

during this study and previous reports involving bivalves as well as many other

organisms (Table 4). The estimated biological half-lives for tetra- and pentachloro

substituted PCB congeners during this study were in the 14 to 39 and 25 to 73 days

ranges, respectively (Chapter VI, Table VII). It is clear that, compared to other ortho-

substituted congeners with the same number of chlorine per molucule, planar PCBs are

removed more slowly from the lipid pool of oysters. The same observation was reported

for transplanted green-lipped mussels (Perna viridis Linnaeus) in Hong-Kong (Kannan er

aL, 1989).

CONCLUDING REMARKS

A simple, sensitive, precise and specific method for the isolation of planar PCB

congeners, with four or more chlorines in non-ortho positions, from other PCBs in

environmental samples was developed for this study. This method, which can easily be

coupled to existing cleanup procedures in most environmental laboratories currently

involved in the high-resolution gas chromatographic analysis of PCBs, yields acceptable

recoveries of these PCB congeners. Compared to other methods, this methodology saves

both time and materials, i.e. solvents, and is safer for the analyst and the environment.

Two of the most toxic planar PCB congeners, i.e. congeners 77 and 126, were

biococentrated by transplanted oysters during the seven-week exposure period. Congener

77 attained an equilibrium concentration with the indigenous oysters in a shorter period of

time than congener 126. Because of the low concentrations, it was not possible to

observe a clear trend in the uptake of PCB congener 169.

When newly contaminated oysters were transplanted back to the Hanna Reef area,

they depurated both 77 and 126 planar PCB congeners; however, the estimated depuration

half-lives were significantly longer than those corTesponding to non-planar PCBs with the
I
same number of chlorines substituted in the biphenyl molecule. Also, the final

concentrations of these planar PCB congeners in HRSCHR oysters at the end of the

depuration phase of this experiment remained relatively high. Because of their toxicity

and per sistency, these planar PCB congeners are of importance in environmental studies.

These congeners are bioconcentrated and retained by bivalves and constitute a potential

health hazard for higher consumers, including human beings.

01-

CHAPTER V

UPTAKE AND DEPURATION OF TRIBUTYLTIN BY THE AMERICAN OYSTER

(CRASSOSTREA VIRGINICA

INMODUCTION

Tributyltin is the active component in antifouling paints. However, this compound

has been shown to be highly toxic to a wide variety of aquatic organisms rather than being

specific to the target individuals. This observation has generated a growing interest in the

bioaccumulation of TBT by marine organisms. In this chapter, the bioconcentration of

TBT and its depuration by transplanted and chronically contaminated oysters are

discussed.

TBT: A REVEEW

Background Information

Evans & Karpel (1985) defined organotin compounds as compounds in which at least

one direct tin-carbon bond exists. Most of the organotin compounds have fin in the IV+

oxidation state giving four series of organotin compounds: mono-, di-, tri- and

tetraorg anotins. Properties of these organotin classes are different. While

monoorganotins have low toxicity, for example, the triorganotin compounds have biocidal
properties. Diorganotins are used as stabilizers in the plastic industry.

These organotin compounds, which were introduced commercially in U.S.A. in the

1940s (Evans & Karpel, 1985), found use as stabilizers of polyvinylchloride (PVC),

industrial catalyzer in the synthesis of polyurethane foams, epoxy resins, plastic materials,

wood preservative and biocide. Within the scope of the last application, butyltins are the

organotin compounds most widely used. Butyltins, in the form of tributyltin (TBT)-

based paints, are highly effective as antifouling agents. With a useful life between 5 and 7

years (Champ & Pugh, 1987) and an effectiveness 10 to 100 times greater than copper-

based paints (Anderson & Dalley, 1986; Ludgate, 1987), the use of paints containing

TBT presents important economic benefits. For example, it has been reported that a six-

month accumulation of fouling organisms, e.g. barnacles, seaweeds and tubeworms, on

ship bottoms increases up to 40% the non-nal fuel consumption (Hall & Pinkney, 1985).

Associated with the economical benefits, there were environmental risks. Because of the

slow mode of action of TBT, standard, short duration tests failed to indicated its toxicity

(Laughlin & Linden, 1987).

Contamination of the coastal marine environment by tributyltin has been investigated

since the early 1980s when French workers discovered that TBT caused malformations

and reduced growth in the Pacific oyster, Crassostrea gigas (Alzieu et al., 1980, 1982).

Similar effects have been reported in England (Anonymous, 1980; Abel et al., 1986) and

the United States of America (Stephenson et al., 1986; Salazar et al.. 1987, Salazar and

Salazar, 1987, 1988; Valkirs et al., 1987a). As a result, the use of antifouling paints

containing TBT on vessels under 25 m has been banned in France (1982), England

(1987), and the United States of America (1988) (Anonymous, 1980; Knipe, 1989; U.S.

Environmental Protection Agency, 1987). The increased concern about the adverse

effects of TBT to non-target organisms led to the decision in the U.S. to include the

analysis of butyltin compounds as part of the National Oceanic and Atmospheric

Administration's National Status and Trends "Mussel Watch" (NOAA's NS&T) Program

(e.g. Wade et al., 1988a).

Distribution and Occurrence in Galveston Bay

Because TBT was not considered to be an environmental threat until the late 1980s,

studies directed at understanding the occurrence and fate of this contaminant in Galveston

Bay are recent and very limited (Table 6). Wade et al. (1988) reported the results of a

study designed to understand its temporal and spatial variations in bivalves collected from

four sites in Galveston Bay. This study indicated that the TBT concentrations were higher

in samples from sites located closer to known sources of inputs, i.e. the Galveston Bay

Yacht Club. A decrease in TBT concentrations is reported toward the outer part of the

Bay. Similarly, the highest TBT concentrations in Galveston Bay sediment samples were

reported near the Galveston Bay Yacht Club (Wade et al., 1990).

Bivalve Uptake and Depuration Studies

Compared to PAHs and PCBs, the number of studies of uptake and depuration of

TBT by bivalves is limited. Although contamination of the coastal environment by TBT,

for example, has been investigated since early 1980s, it was not until the late 1980s that

this compound was considered to be a real threat to the quality of coastal waters.

Controlled flow-through experiments have shown that mussels accumulate increasing
amounts of TJ3T over a 60 day period, reaching a steady state concentration thereafter

(Salaza er al., 1987). Reported half-life for TBT in field studies with diffesent bivalves

were relatively short (Mytilus edulis, 14 days, Laughlin et al., 1986; Crassostrea gigas

and Ostrea edulis, 10 days, Waldock et al., 1983). Depuration rate constants calculated

from laboratory data were found to be much lower than those obtained in environmental

studies (Laughlin et al., 1986). For example, the longer half-life (40 days) recently

TABLE6
TBT, DDT and MBT Concentrations (in ng Sn g- on a Dry-Weight Basis) in Oyster and Sediment Samples from the Galveston Bay Area.

Location Sample TBT DBT MBT Total OTs Reference

Yacht Club oysters 660 70 10 740 Wade et al., 198 8a
Todd's Dump oysters 380 30 10 420 Wade el al., 1988a
Confederate Reef oysters 380 30 10 420 Wade el al., 1988a
Hanna Reef oysters 110 10 <5 120 Wade et al., 1988a

Yacht Club sediments 13 <5 <5 13 Wade et al., 1990
Todd's Dump sediment% 7 <5 <5 7 Wade et al., 1990
Confederate Reef sediments 6 <5 <5 6 Wade et al., 1990
Hanna Reef sediments 11 <5 <5 11 Wade et al., 1990

101

reported for mussels (Mytilus edulis) in laboratory depuration studies (Zuolian & Jensen,
1989) might reflect the effects of bivalve manipulation.

UPTAKE AND DEPURATION OF TBT

Experimental Design, Sample Collection and Methods

The experimental design and sample collection used for the study of TBT were the

same as those discussed for PAHs (Chapter 11).

Exraction and sarVIefractionation

The analytical procedure used during this study is a modification of previously

reported methods (Maguire, 1984; Unger et al., 1986; Matthias et al., 1986) and is

discussed in detail elsewhere (Wade et al., 1988a). Approximately 15 g (wet weight) of

tissue sample was weighed into a 250 ml centrifuge tube. Anhydrous sodium sulfate (40

g), 0.2% tropolone in methylene chloride (100 ml) and tripropyltin chloride as an internal

standard were added. The sample was extracted for 3 min with a Tekmar Tissuemizer,

centrifuged, and the supernatant was decanted into a 500 ml flat-bottom flask. The

extraction was repeated two more times with 0.2% tropolone in methylene chloride (100

ml). The combined extracts were concentrated in a water bath (600C) and the methylene

chloride was replaced by hexane. The sample was then purged with nitrogen and hexyl-
magnesium bromide (2 ml, 2 M Grignard reagent) was added. The hexylation reaction

was carried out for 6 h at 50'C. HCl (5 ml, 6 M) was then added to neutralize the excess

Grignard reagent. The sample was shaken vigorously and the organic and aqueous

phases were allow to separate. Two more extractions were perforned using a mixture of

pentane:methylene chloride (2:1, 125 ml). The organic phase was dried with anhydrous

sodium sulfate and concentrated to 2 ml in hexane. The hexylated organotin compounds

102

were isolated on a column containing combusted alumina (4000C, 17 g) and silica (170'C,

13.5 g). The column was eluted with pentane (60 n-d). The sample was then concentrated

to 500 gl. Samples were spiked with tetrapropyltin before analysis to determine recovery

of the internal standard for the whole analytical procedure.

Instrwnental analysis

Butyltin species were analyzed by gas chromatography on a Hewlett-Packard 5790

gas chromatograph (GQ equipped with a capillary column (DB-5, 30 m x 0.25 mm i.d. x

0.25 pm coating thickness) and a flame photometric detector (FPD). The GC temperature

was programmed from 60'C to a final temperature of 300'C, at a rate of 12*Qmin, with a

final 10 min hold time. Injector and detector temperatures were 300 and 250'C,

respectively. Helium was used as the carrier gas. The response of the FPD was selective

for Sn using a 610 nm filter. The limit of detection of TBT and breakdown products,
dibutyltin (DBT) and monobutyltin (MBT), was 5 ng Sn g- I dry weight.

Ancillary pararwrers

Methodologies for the sediment grain-size analysis and extractable lipids percentage

were discussed in the materials and methods section of Chapter 11.

Statistical analysis

The statistical analyses performed on the TBT data were previously discussed in the

materials and methods section of Chapter 11.

Uptake of TBT by Transplanted Oysters

In this and the following sections, Ship Channel, Hanna Reef, transplanted Hanna

Reef-to-Ship Channel, transplanted Ship Channel-to-Hanna Reef and relocated Hanna

103

Reef-to-Ship Channel-back to-Hanna Reef oysters are refered as SC, HR, HRSC, SCHR

and HRSCHR oysters, respectively.

Average concentrations of TBT encountered in oyster and sediment samples, are

reported in Table A-10 (Appendix). TBT concentrations in SC oysters during the uptake
phase of this experiment were very stable (mean= 340�39 ng Sn g-1, relative standard
deviation = I I%, range = 330-420 ng Sn g- 1) suggesting that bioavailable TBT to the

oysters was,relatively constant. Therefore, it is assumed that the accumulation rate in

HRSC transplanted oyster was not affected by changes in the concentration of

bioavailable TBT.

Approximately a 10-fold increase in TBT concentrations was observed by the end of

the exposure period in HRSC oysters (Fig. 31). By the end of the seven-week exposure

period, the concentration of TBT in HRSC oysters was similar to the level found in SC

oysters. This increase is similar to previously reported uptake data in exposed mussels
after about 50 days (Laughlin, Jr., et al., 1986; Zuolian & Jensen, 1989). Controlled

flow-through experiments have shown that mussels reach a steady state concentration of

TBT after a 60-day exposure period (Salazar et al., 1987). A steady state concentration

was not attained in this study; however, the continued increasing concentrations of TBT

measured in transplanted oysters by the end on the seven-week uptake period seem to

indicate that, given enough time, a true equilibrium concentration comparable to the levels

measured in native oysters would have been reached.
DBT, the major breakdown product of TBT (Maguire, 1984; Seligman el al., 1986),

did not show any accumulation during the exposure period and was only detected at low

concentrations in both groups of oysters. This might suggest that DBT, a more polar and

soluble 'compound than TBT, may be quickly depurated from the oyster tissues. DBT
concentrations ranged from 22 to 34 ng Sn g-l and <5 to 22 ng Sn g-I in SC and HRSC

104

1000 Tributyltin

100.

Hanna Reef Oysters
Ship Channel Oysters
10 ............ I
a 10 20 30 40 so 60 7 0 80 90 100
Time (Days)

Fig. 31. Concentrations of tributyltin in tissues of Hanna Reef and Ship Channel
Oysters during exposure to the Ship Channel area contaminant levels and following
transplant to the Hanna Reef are.

105

oysters, respectively, whi.le MBT concentrations were, in all cases, below the 5 ng Sn g-I

Umit of detection.

Sediment samples collected from the Ship Channel and Hanna Reef locations during

this study had TBT, DBT, and MBT concentrations below the detection limits.

Depuration of TBT by Newly and Chronically Contaminated Oysters

Both oyster populations showed statistically significant depuration of TBT after back-

transplantation to the Hanna Reef area. The final TBT concentration encountered in

HRSCHR individuals at the end of the 50- day depuration period was over 100% higher

than the levels measured in the same group of oysters before the transplantation

experiment to the Ship Channel area.

The calculated half-life for TBT in the originally uncontaminated Hanna Reef oysters

(21 days) was higher than the values reported for mussels (Mytilus edulis, 14 days,

Laughlin et al., 1986) and the Pacific (Crassostrea gigas) and European (Ostrea edulis)

oysters (10 days, Waldock er al., 1983). Depuration rate constants calculated from

laboratory data were found to be much lower compared to environmental studies

(Laughlin et al., 1986). For example, the longer half-life (40 days) recently reported for

mussels (Mytilus edulis) in laboratory depuration studies (Zuolian & Jensen, 1989) might

reflect the effects of bivalve manipulation. Comparatively, the TBT half-life in chronically

exposed oysters, i.e. SCHR oyster, was 27 days.

As discussed in previous chapters, similar differences in the depuration rates, i.e.

half-lives, were observed for other trace or-anic contaminants between newly and

chronically contaminated oysters. In the specific case of TBT, a possible explanation of

these different depuration rates could be the existence of ligands within the oyster body as

suggested by Laughlin (1990). These ligands, which do not seem to be induced by TBT

exposure, might be produced slowly by chronically exposed bivalves. Tissue molecules

106

with groups containing sulfur, oxygen or nitrogen are mentioned as the obvious ligand

candidates. Ilerefore, the existence of these ligands in one group of oysters, but not in

the other, might explain the difference observed in depuration rates.

CONCLUDING REMARKS

Although a steady state concentration was not reached, transplanted oysters rapidly

accumulated TBT to practically reach an equilibrium with the concentrations encountered

in indigenous oysters at the end of the 48-day exposure period. DBT, a more polar

compound than TBT, has only been detected at low concentrations in the oyster tissues.

When relocated to the Hanna Reef area, both oyster populations significantly

depurated TBT; however, the observed depuration rates were different. HRSCHR

oysters depurated at a rate about 30% faster than the clearance rate observed in SCHR

oysters. 17his is reflected in the estimated TBT half-lives for HRSCHR and SCHR

oysters (21 and 27 days, respectively). The same observation was made when comparing

half-lives for PAHs and PCBs in HRSCHR and SCHR oysters.

107

CHAPTER VI

MECHANISM OF THE UPTAKE AND RELEASE OF TRACE ORGANIC

CONTAMINANTS BY THE AMERICAN OYSTER (CRASSOSTREA VIRGINICA)

WIRODUCTION

The relationship between a pollutant concentration in organisms and their aquatic

habitat was first explained as a simple partition process across external membrane surfaces

(Hamelink et al., 197 1). Since then, the dynamic equilibrium between uptake from and

depuration to water, together with the balance between ingested and excreted matter, has

been widely used to explain bioaccumulation data. After the introduction of the n-octan6l-

water partition coefficient (Kow) to assess the pot ential of different organic compounds to

be bioaccumulated under equilibrium conditions, several studies have reported a

correlation between the concentration factors of organic contaminants in tissues and the

logarithms of their Kow coefficients (Geyer et al., 1982; Mackay, 1982; Pruell et al.,

1986). The Kow coefficient has been found to be very useful in predicting the

environmental partitioning of some lipophilic compounds.

In this chapter, the kinetics involved in the uptake and release of selected trace organic

contaminants (PAHs, PCBs, including planar congeners, and TBT), as well as their

concentration factors by American oyster (Crassostrea virginica) during transplant

experiments, are reported. In the following sections, Ship Channel, Hanna Reef,

transplanted Hanna Reef-to-Ship Channel, transplanted Ship Channel-to-Hanna Reef and

109

relocated Hanna Reef-to-Ship Channel-back to-Hanna Reef oysters are refered as SC,

HR, HRSC, SCHR and HRSCHR oysters, respectively.

WCEAMSM OF BIOCONCENTRATION

Kinetics

Bioconcentration is defined as the balance between uptake and depuration processes,

which may proceed by first order kinetics characterized by the rate constants ku and kd,

respectively (Shaw & Connell, 1984). The bioconcentration factor (BCF) is defined as

the proportionality constant relating the concentration of a chemical in water to the

concentration in an aquatic organism at steady state equilibrium.

Ile following characteristics describing the kinetics of bioconcentration using a single

compartment model was adapted from Connell (1990). The one-compartment approach is

the mathematical expression of the hydrophobicity model, which considers

bioconcentration as the partitioning of a chemical between the exposure media and the

lipidic pools of an organisms, and vice versa, with no physical barriers (Barron, 1990).

The general first-order equation that describes the uptake and depuration of lipophilic

compounds, such as PAHs and PCBs, is expressed by

dCddt = ku Cw - kd Ct (1)

where CL is the concentration in the organism at time = t and Cw is the concentration in

water. Since the net amount of an analyte in the water represents a large reservoir

compared to the relatively lower amount taken up by organisms, Cw can be regarded as

constant. By integration and rearrangement

Ct = (ku/kd )Cw(l - e-kdt) + Ce-kdt (2)

109

where Cto is the initial concentration in the organism. This equation predicts that Ct will

increase in concentration with time and with a declining rate of increase. Thus, at time

infinity, t.

Ct.= ku/kd CW (3)

and

CtjCw = kw'kd = Kb (4)

where Kb is the bioconcentration factor (BCF). Similarly, the BCF can be calculated

from equation (1) when uptake and depuration are in equilibrium, i.e. at t = infinity

dCVdt = 0 = ku Cw - kd Ct. (5)

The theoretical time period to reach equilibrium occurs when e-kdt is zero, which is
when t is infinity. However, effective equilibrium can be considered to be reached at-teq

when Ct is 0.99 of the concentration value at infinity. Thus, from equation (2)

Cteq = (ku/kd) Cw (I - e-kdL--q) (6)

CLeq = 0.99 (kulkd) Cw (7)

From equations (6) and (7)

0.99 = I - e-kdteq (8)

and

teq = 4.605/kd (9)

110

Because of the lenghts of time required for transplanted oysters to reach a concentration

equal to 99% the equilibrium concentration, a more time-realistic approach would be to

consider the time to attain 90% of the equilibrium concentration for organic contaminants.

Equation (9) is then modified to

tg()%= 2.303/kd (10)

If exposure to the compound is terminated by transfer to uncontaminated water or,

more realistically, to a site were environmental concentrations are negligible, then C,, can

be regarded as zero and

dCVdt = -kd Ct (11)

Ibis indicates that during exposure, both uptake and depuration were operating, but now,

in very low concentration or uncontaminated water, uptake can be neglected. By

integration

Ct = Cto e-kd t (12)

log Ct = log Cto - kd t/2.303 (13)

where Cto is @ow the initial concentration at time zero for the depuration period. This

equation shows that as t increases Ct declines, but the rate of decline decreases with
increasing time. Also, since Cto and kd are constants, log C, is linearly related to time.
When half of the initial compound has been cleared, then Ct= Ctd2 and the half life, tj/2,

is represented by

t1/2 = log 2 (2.303/kd) = 0.693/kd (14)

The kinetic parameters obtained for PAHs, PCBs, planar PCBs and T]3T are given in

Table 7. Concentration factors for PAHs and PCBs were calculated comparing the

concentrations measured in HRSC and SC oyster tissues at the end of the seven-week

exposure period and the average concentrations encountered in water samples (Tables A-

2, A-3, A-6 and A-7, Appendix).

Polynuclear aromadc hydrocarbons

As previously discussed in Chapter II, transplanted HRSC oysters bioconcentrated

most of the PAHs to concentrations that were not significantly different from the

concentrations encountered in indigenous SC oysters at the end of the uptake period.

Bioconcentration factors in both groups of oysters increased with the number of aromatic

rings for PAHs having two-, three- and four-rings per molecule and decreases thereafter.

The maximum concentration factors in both group of oysters were for pyrene, chrysene

and benzo(a)anthracene. The lowest values were for compounds with molecular weights

less than or greater than pyrene.

Depuration constant values for PAHs can be divided in two groups in both oyster

populations. The first one represents the two- and three-ring PAHs, which ranged from
0.0268 to 0.0297 days- I in HR oysters and from 0.0166 to 0.320 days-1 in SC oysters.

The second group includes the remaining PAHs with kd values ranging from 0.0439 to
0.0787 days- I in HR oysters and from 0.0430 to 0.0708 days- I in SC oysters. In the

first case, the low kd values result in longer biological half-lives and longer time to reach a

concentration within 10% the concentration at equilibrium. PAHs in the second group

had considerably shorter half-lives and reached within 10% of the equilibrium
concentration in a shorter period of time. Estimated PAH half-lives ranged from 9 days

TABLE 7
Estimated Days to 90% Uptake Equilibrium (tgo%), Bioconcentration Factors (BCF), Depuration Rates (kd) and Biological Hal f-Lives (BHL)
for PAHs and PCB Congeners in Hanna Reef and Ship Channel Oysters During the Field Studies.

Hanna Reef Oysters Ship Channel Oysters
Analytc t90% BCFa kd BHL R2b BCF kd BHL R2
(days) (days- 1) (days) (days) (days-1) (days)

PAHs

2,3,5-Trimcthynaphthalene 80 16,0W 0.0287 24 0.74 23,000 0.0320 22 0.83
Anthracene 78 29,000 0.0295 24 0.67 29,000 0.0166 42 0.68
I-Mcthylphenanthrenc 78 43,000 0.0297 23 0.97 60,000 0.0283 24 0.96
Fluoranthene 86 310,000 0.0268 26 0.90 310,000 0.0215 32 0.69
Pyrene 35 890,000 0.0663 10 0.95 880,000 0.0557 12 0.98
Benz(a)anthracene 44 450,000 0.0525 13 0.96 490,000 0.0453 15 0.99
Chrysene 41 490,000 0.0565 12 0.99 530,000 0.0439 16 0.99
Benzo(b)fluoranthene 38 290,000 0.0601 12 0.95 270,000 0.0488 14 0.98
Benzo(k)fluoranthene 34 340,000 0.0674 10 0.96 330,000 0.0561 12 0.98
Benzo(e)pyrene 38 300,000 0.0602 12 0.97 310,000 0.0430 16 0.98
Benzo(a)pyrene 29 200,000 0.0787 9 0.98 210,000 0.0708 .10 0.99
Perylene 35 140,000 0.0649 11 0.94 140,000 0.0532 13 0.99
Indeno[ 1,2,3-c,dlpyrene 35 44,000 0.0665 10 0.96 42,000 0.0647 I'l 0.93
Dibenzo(a,h)anthracenc 52 27,000 0.0439 16 0.93 24,000 0.0506 14 0.90
Benzo(g,b,i)perylene 38 120,000 0.0610 11 0.96 110,000 0.0574 12 0.98

TABLE 7
(confinued)

Hanna Reef Oysters Ship Channel Oysters
Analyte tgo% BCFa kd BHL R2b BCF kd BHL R2
(days) (days-1) (days) (days) (days-1) (days)

PcBs

2,2',5 (18) 48 110,000 0.0480 14 0.82 110,000 0.0362 19 0.93
2,3',5 (26) 73 - 0.0314 22 0.88 - 0.0327 22 0.99
2,4,4' (2 8) 55 - 0.0419 17 0.96 - 0.0205 34 0.93
2,2',3,3' (40) 47 - 0.0488 14 0.87 - 0.0383 18 0.97
2,2',3.4 (41) 77 - 0.0299 23 0.83 - 0.0127 55 0.92
2,2',3,5' (44) 91 97,000 0.0253 27 0.83 83,000 0.0153 45 0.94
2,2',4,5' (49) 131 210,000 0.0176 39 0.84 220,000 0.0114 61 0.94
2,2',5,5' (52) 88 100,000 0.0261 27 0.80 100,000 0.0155 45 0.91
2.3',4',5 (70) 100 610,000 0.0235 30 0.88 740,000 0.0120 58 0.96
2,4,4',5 (74) 100 200,000 0.0231 30 0.87 220,000 0.0147 47 0.95
2,2',3,3',6 (84) 123 0.0187 37 0.84 - 0.0087 80 0.79
2,2',3,4,5'/2.3,4,4',6 (87/115) 181 0.0127 55 0.73 - 0.0038 132 0.45
2,2',3,4',6 (91) 84 370,000 0.0275 25 0.78 330,000 0.0140 50 0.89
2,2',3,5,5' (92) 99 - 0.0233 31 0.89 - 0.0108 63 0.93

TABLE 7
(continued)

Hanna Reef Oysters Ship Channel Oysters
Analyte t90% BCF1 kd BHL R2b BCF kd BUL R2
(days) (days-1) (days) (days) (days- 1) (days)

2.2',3',5',6 (95) 149 - 0.0155 45 0.81 - 0.0073 95 0.79
2,2'.4,4',5 (99) 165 210,000 0.0140 49 0.88 330,000 0.0076 91 0.79
2,2',4,5,5'/2,2',3,4',5 (101/90) 184 160,000 0.0125 55 0.86 260,000 0.0060 116 0.78
2.3.3',4,4' (105) 211 120,000 0.0109 63 0.76 140,000 0.0058 120 0.76
2,3,3',4',5 (107) 100 - 0.0231 30 0.82 - 0.0151 46 0.75
2.3,3',4',6/3,3',4,4' MOM) 149 250,000 0.0155 45 0.74 330,000 0.0067 103 0.67
2.3',4,4',5 (118) 242 300,000 0.0095 73 0.79 480,000 0.0023 299 0.19
2,2',3,3',4,4' (128) 253 - 0.0091 76 0.75 - 0.0030 229 0.42
2,2',3,4,4*,5'/2,3,3',4,5,6 (138/160) 658 46,000 0.0035 200 0.32 79,000 0.0012 595 0.11
2,2',3,4',5,5' (146) 371 0.0062 111 0.60 - 0.0029 239 0.27
2,2'.3,4',5',6/2',3,4,4',5 (149/123) 435 74,000 0.0053 130 0.46 150,000 0.0016 439 0.24
2,2',4,4',5,5'/2,2',3,3'.4,6' (153/132) 169 8-7,000 0.0136 51 0.71 140,000 0.0068 102 0.90
2,2',3.3',4',5,6 (177) 171 50,000 0.0135 52 0.83 65,000 0.0048 145 0.54
2,2',3,3',5,5',6 (178) 172 - 0.0134 52 0.62 - 0.0076 91 0.83
2,T,3,4.4',5,5' (180) 167 12,000 0.0138 50 0.94 14,000 0.0049 142 0.63
2,2'.3.4',5,5',6 (187) 233 61,000 0.0099 70 0.65 82,000 0.0027 258 0.56

TABLE 7
(continued)

Hanna Reef Oysters Ship Channel Oysters
Analyte t90% BCFa kd BHL R2b BCF kd BHL R2
(days) (days- 1) (days) (days) (days-1) (days)

Coplanar PCBs

3,3',4,4' (77) 291 - 0.W79 88 0.85
3,3',4,4',5 (126) 360 - 0.0064 107 0.50 - - -

Butyltin species

Tributyltin 68 72,000c 0.0251 27 0.95 78,000c 0.0202 34 0.96

a Bioconcentration factor concentration in oyster tissue at the end of the uptake period / conccntration in water.
b R2 = square of the correlation coefficient for the regression equation to obtain kd.
C = estimated BCF (see text).

116

for benzo(a)pyrene to 26 days for fluoranthene and from 10 days for benzo(a)pyrene to

42 days for anthracene in HRSCHR and SCHR oysters, respectively. Most of the values

were, however, between 10 and 13 days for HRSCHR oysters and between 13 and 16

days for SCHR oysters. In general, originally uncontaminated oysters depurated faster

d= chronically exposed individuals.

Polychlorinated biphenyls

Bioconcentration factors for PCB congeners show approximately the same general

behavior discussed for PAHs. Concentration factors are higher for tetra- and

pentachlorobiphenyls congeners and lower for tri-, hexa- and heptachlorobiphenyls.

Octa-, nona- and decachlorobiphenyls were detected at very low concentrations.

Ile decreasing values of kd and the increasing values of t90% and half-lives with the

higher degree of chlorination of the biphenyl molecule reflect the more rapid uptake and

release of the lower chlorinated biphenyls. Previous reports indicate that the less

lipophilic congeners reach an equilibrium concentration, either during uptake or

depuration, at a faster rate than those compounds that are more lipophilic (e.g.

Ellegehausen et al., 1980; Bruggernan et al., 198 1; Tanabe et al., 1987a).

Biological half-lives for PCB congeners in HRSCHR and SCHR oysters ranged from

14 to 200 days and from 19 to 595 days for congeners 2,2',5-trichlorobiphenyl (18) and

2,2',3,4,4',5'-hexachlorobiphenyI (138), respectively. Planar PCBs showed a slower
clearance rate'than other PCBs within the same homolog groups (Fig. 32). These slower

depuration rates are reflected in longer half-lives and time periods to reach 90% of

equilibrium concentrations. As in the case of PAHs, most of the bioconcentrated PCB

congeners were eliminated faster by originally clean oysters than by- chronically

contaminanted bivalves.

117

PCB 77
4

PCB 126
5

-8
0.00 0.01 0.02 0.03 0.04 OAS
kd (1/day)

PCB 77
4
U
0
.C PCB 126

0 so 100 150 200
BHL (day)

Fig. 32. Depuration constant (kd) and biological half-lives (BHL) of planar PCB
congeners compared to ranges of values calculated for non-planar PCBs.

118

Several reports have suggested that bioaccumulation of PCBs by different organisms

might be influenced by physicochernical factors (Jan & Josipovic, 1978; Tulp &

Hutzinger, 197 8; Matsuo, 1980; Shaw & Connell, 1980, 1982, 1984; Opperhuizen et al.,

1985; Samuelian & O'Connor, 1985). Several parameters have been suggested that may

be suitable to measure the effect of these factors on the kinetics of bioaccumulation, e.g.

molar volume, parachor, steric effect coefficients.

Competitive partition between aqueous and nonpolar phases, e.g. lipids, as well as

stereochemistry appear to be significant factors influencing the uptake of these

compounds. As discussed earlier, maximum PCB uptake by organisms is observed for

congeners having four to six chlorine atoms. Low chlorinated congeners have higher

water solubilities and, as a consequence, lower lipophilicity. In contrast, isomers in the

higher homolog groups have unfavorable steric configurations (Shaw & Connell, 1984).

Opperhuizen et al. (1985) reported that the BCFs of polychlorinated naphthalenes and

biphenyls depend on molecular sizie, e.g. molecular volume and cross-sectional area that

are directly related to chlorine substitution patterns, rather than hydrophobicity.

As an example of the antagonistic effects that lipophilicity and size of the different

congeners have on their accumulation, the bioconcentration factors of six related PCB
congeners are compared in Fig. 33. Log Ko.. and total surface area (TSA X 10-20 m2)

values are also indicated (Hawker & Connell, 1988). These congeners have a common

2,4,5-chlorine distribution in one ring while one to four chlorines are sequentially

substituted on the second ring. It is clear that the more favorable lipophilicity/size

conditions for bioaccamulation are present in PCB 99. Congener 180 is the most

lipophilic of the six PCBs shown and also has the largest total surface area. On the other

hand, the smaller size of congener 74 for membrane transport into the tissue is countered

by its higher water solubility.

119

CI CI

CI
500- tc, 11 CI CI CI
CI
400 CI CI CI
CI CI
CI CI C III
X, 3W CI CI CI
CI CII CI CI CI
200 CA CI CI CI
100 CI CI
CI
CI

0 -----JEML--
PCB Congener 74 99 118 138 153 180

Log K,,, 6.20 6.39 6.74 6.83 6.92 7.36
TSAx10-20 (M2) 246 252 262 265 267 280

Fig. 33. Bioconcentration factors of six selected PCB congeners in relation to their
liphophilicity and size.
Nk

120

Not only the bioconcentration of PCB congeners seems to be affected by the different

chlorine substitution in the biphenyl rings, but their depuration may be affected by these

factors. For example, Fig. 34 shows PCB congeners at different levels of chlorination

that have a fixed substitution pattern for all chlorines but one. The extra chlorine, in bold,

is sequentially substituted in para, meta and ortho positions giving three different

substition patterns. Also indicated in the figure are the estimated half-lives for these

congeners in chronically contaminated Ship Channel oysters. The experimental data

shows that there is a decrease in the estimated biological half-lives when the extra chlorine

is substituted in thepara > ortho > meta positions.

Tributyltin

Unfortunately, the lack of data on concentrations of TBT in seawater samples from the

Ship Channel area do not permit the calculation of a bioconcentration factor. However,

when the limit of detection of the analytical method for seawater is used to calculate the

bioconcentration factor, the minimum value can be estimated. The detection limit for TBT
in seawater is 5 ng Sn L-1; this gives a bioconcentration factor for transplanted and

indigenous oysters, on a dry weight basis, on the order of 72,000 and 78,000 at the end

of the exposure period, respectively. On wet weight basis, these estimated values convert

to 9,000 and 9,750, respectively, which compare well with previously published

concentration factors in mussels (up to 5,500; Laughlin et al., 1986) and oysters (up to

6,000; Waldock et al., 1983).

Depuration constants for TBT in newly and chronically contaminated oysters were
0.0251 and 0.0202 days-1, respectively. These values compare well with those

encountered for the lower molecular weight PAHs and PCBs. Similarly to those organic

contaminants, the estimated half-life for TBT in originally uncontaminated HR oysters (27

121

cl C1 cl C1
(D-QCI &_O
C1 PCB 49 C1 PCB 52 C1
BHL = 61 Days BHL = 45 Days

C1 C1 C1 C1 C1 C1 C1 C1 C1
a_OCI &_O 0-0
C1 C1 C1 C1 C1
PCB 87 PCB 95 PCB 92
BHL = 132 Days BHL = 95 Days BHL = 63 Days

C1 C1 C! C1 C1 C1 C1 C1 C1
CIO-OCI CIC)_O CIO-0
cl CA
PCB 105 PCB 110 PCB 107
BHL = 120 Days BHL = 103 Days BHL = 46 Days

C1 C1 C1 C1 C1 C1 C1 C1 C1
C1 O_Q cl C1 C@_Q C1 C)_@
C1 cl . C1 C1 cl
PCB 138 PCB 149 PCB 146
BHL = 595 Days BPL = 439 Days BHL = 239 Days

Fig. 34. Relationship between chlorine-substitution patterns in PCBs and their
depuration half-lives. See text for explanation.

122

days) was lower than the corresponding value for chronically exposed individuals (34

days).

The Octanol-to-Water Partition Coefficient

The octanol to water partition coefficien (Kow) is defined as:

Kow = CC/CW (15)

where CO and Cw are the concentrations of the analyte in n-octanol and water,

respectively. Although many organic solvents have been used for this purpose, n-octanol

is considered to be the best surrogate for organism lipids. Since the introduction of the

Kow partition coefficient in early 1960s (Hansch & Fujita, 1964), it has been used in

numerous studies to explain the concentration of different organic compounds in

biological tissues. The observed bioconcentration factor, or more commonly its log

value, is related to log Kow by the following equation

log Kb = a Kow + b (16)

If n-octanol behaves as a perfect surrogate for organism lipids, the constant (a) should be

equal to 1. A deviation from unity indicates how much octanol differs from biological

lipids.

Fig. 35 shows the log of the calculated bioconcentration factor of individual PAHs or

selected PCB congeners in both oysters populations plotted against the log of their
octanol-to-water coefficients, respectively. Values -of the log of the PAH and PCB

octanol-to-water partition coefficients are from Isnard and Lambert (1989) and Patil

(1991), respectively. The plots of the log of calculated concentration factors versus log
Kow have bell shaped curves. In general, an increase of the bioconcentration factors is

123

Polynuclear Aromatic Hydrocarbons
1000000.

U
1000007

10000-

:
0
0 0 Hanna Reef Oysters
0 0 Ship Channel Oysters
1000 .................... ......
3.0 4.0 S.0 6.0 7.0 8.0

Log Kow

1000000. Polychlorinated Biphenyls

0
16.
q a 0 0
0
0 0
1000DO 0
0

a
15

0 0 Hanna Reef Oysters
to 0 Ship Channel Oysters
10000 1 - - - - r- .-r- . --r - r @. - I . - . -I
S.0 5.5 6.0 6:5 7.0 7.5

Log Kow

Fig. 35. Bioconcentration factors of polynuclear aromatic hydrocarbons and
polychlorinated biphenyls calculated for transplanted Hanna Reef and indigenous
Ship Channel oysters during the exposure period Versus log octanol to water partition
coefficients (Kow).

124

observed until log Kow reaches about 6, then there is a decline for the more lipophilic

compounds with high Kow. Dobbs & Williams (1983) indicated that compounds with

Kow greater than 6 exhibit a decrease in their lipid solubility. These compounds, often

referred as "superlipophilic" were redefined by Connell (1990) as "superhydrophobic."

A departure from the predictive relationship (16) has been observed in other studies

(e.g. Oliver, 1984, 1987). Hawker & Connell (1985) have indicated that the little

attention devoted to the time required by highly lipophilic chemicals to reach equilibrium

might result in underestimated kd values which cause the observed change of the slope in

the plot. They estimated that the necessary time for "superhydrophobic" analytes to attain

an equilibrium concentration in exposed organisms range from a minimum 0.5 years, for

log Ko.. = 6, up to 12 years, for log K0,, = 8. That study, however, was done by

exposing uncontaminated organisms in the laboratory to different organic compounds.

Besides the fact that extrapolations of laboratory produced data to real world situations

are not always possible, and in most cases inaccurate, this extrapolation is further

complicated if the uptake of these xenobiotics by an uncontaminated adult organisms is

compared to the uptake by organisms that are developing in a chronically contaminated

area. It seems obvious that tissues and lipid pools being formed by juvenile organisms in

chronically contaminated environments will have a better chance to truly incorporate

xenobiotic compounds than have tissues and lipid pools already formed in uncontaminated

adult individuals later exposed to the same xenobiotic compounds.

In the present study, Ship Channel oysters have been chronically exposed to the high

xenobiotic concentrations present in the Ship Channel area since the earliest stages of their

lives. With average lengths between 7 and 9 cm, the oysters used in this study were adult

organisms at the time of sampling and the analyte concentrations in their bodies can be

assumed to represent the equilibrium concentration at infinity (t.). Because of the
similarities in the shapes of the curves obtained for SC and HRSC oysters, it is clear that

125

less than two months are needed for newly exposed oysters to reach equilibrium

concentrations that are, in general, comparable to those encountered in chronically

exposed individuals.

Table 8 compares the relationships between log Kd and log Kow (16), obtained for this

study with previuosly reported works. Since the linear relationship between log Kb and

log Kow does not exist for log Kow values higher than 6, values for constants a and b are

commonly calculated considering only log Kow values up to 6. Connel & Hawker (1988)

have suggested that octanol is not a good surrogate for "superhydrophobic" compounds

(i.e. those with log Kow > 6) and, therefore, the concentration of these compounds by

organisms can not be accurately predicted from their log Kow values.

The Two-Compartment Model Approach

In depuration studies, it is also important to understand the effects that partitioning

among different body compartments might have on the elimination of accumulated organic

contaminants. The observed differences in depuration rates and, consequently, in the

half-life estimations between HRSCHR and SCHR oyster populations seems to indicate

that at least a two compartment model rather than the single compartment approach would

be more accurate in describing the depuration kinetics in exposed oysters. The two

compartment model, with the first compartment representing a peripheral system and the

second a central system, was described by Moriarty (1975). Briefly, the initial rapid

exchange across external membranes is followed by a slower but more persistent

accumulation in fatty tissues through the circulatory system. Chronic, or long term,

exposure would result in accumulation of organic xenobiotics in deeper deposits of lipids

stored as energy reserves. In this study, for example, xenobiotic chemicals may be better

partitioned between both compartments in chronically contaminated SC oysters than in

newly exposed HRSC oysters. If the accumulation of organic xenobiotics in bivalves is

126

TABLE8
Characteristics of the Relationships Between log Kb and log Kow for
Bioconcentration of Trace Organic Contaminants in Different Organisms.

Analytes n a b Organism Reference

PAHs 22 0.78 -0.35 fish SchUdrmann & Klein (1988)
PAHs 30 0.95 -1.06 fish Connell & SchUUrmann(1988)
PAHs 6 0.97 -1.40 mussels Pruell et al. (1986)
PAHs 13 1.17 -0.88 oysters (SC) This study
PAHs 13 1.15 -0.77 oysters (HR) This study

Crude oil 14 0.49 1.03 oysters Ogata et al. (1984)

PCBs 4 0.59 1.73 mussels Pruell et al. (1986)
PCBS 7 0.69 1.22 oysters (SQ This study
PCBs 7 0.65 1.46 oysters (HR) This study

Pesticides 8 0.70 -0.26 alga Ellegehausen et al. (1980)
Pesticides 8 0.83 -1.71 fish Ellegehausen et al. (1980)

Organics 16 0.86 -0.81 mussels Geyer er al. (1982)

127

the result of a simple partition between tissues and seawater, then the elimination of the

source of contamination should reverse the process. Within the scope of the two

compartment model, tissues with low lipid contents, i.e. muscle and mantle, are generally

reported to have relatively lower tissue burdens than those with higher lipid levels, i.e.

gonads and gills (e.g. Laughlin et al., 1986). Clearly, with two different compartments

capable of accumulation of organic xenobiotics, kd values in exposed oysters can no

longer represent the simple partition between oyster and ambient seawater but a more

complicated and longer process involving different equilibrium constants among the

compartments and with the environment.

Depuration versus Degradation

Biotransformation of organic compounds into more polar and, therefore, more soluble

metabolites decreases the equilibrium level of the accumulated chemicals by increasing the

rate of depuration. Thus, the observed kd constants for PAHs and TBT might be a

combination of at least two constants, the physical partition rate between the oyster tissues

and ambient seawater and the biological breakdown rates to more polar compounds. This

will result in an increase of the clearance rate beyond that due solely to the physical

process described in the preceding sections.

Although early reports indicated that bivalves could not metabolize petroleum

hydrocarbons (e.g. Lee er al., 1972; Payne & Penrose, 1975; Vandermeulen & Penrose,

1978), later studies have shown that bivalves are able to metabolize PAHs (e.g.

Anderson, 1978a, 1978b, 1979; Payne & May, 1979; Stegeman, 1980). However, the

metabolic rates are relatively lower than those observed in other marine animals. For

example, Anderson (1978a) detected comparatively low concentrations of aryl

hydrocarbon hydroxylase, an enzyme inducible by exposure to benzo(a)-pyrene, in the

digestive gland of oysters. Similarly, metabolism of TBT has been shown by Lee (1985,

128

1986). Lee (1985) reported that after three days exposure 10% of the radioactivity of the
applied [14C]TBT was found in the digestive gland of oysters in the form of DBT and

other more polar metabolites. Other biodegradation studies have shown that DBT, a more

polar compound than TBT, was the major breakdown product of TBT, whereas MBT

was detected at very low concentrations (Maguire, 1984; Seligman et al., 1986). In the

case of PCBs, there is no report that indicates that bivalves are able to metabolize these

compounds.

CONCLUDING REMARKS

Bioconcentration factors (BCF) for PAHs and PCBs increased with the increasing

octanol-to-water partition coefficient up to a value of Kow around 6 and decreased

thereafter. Maximum BCF were observed for four-ring PAHs (e.g. pyrene, chrysene and

benzo(a)anthracene) and for PCB congeners having four and five substituted chlorines.

In general, depuration of PAHs was faster than the clearance rates observed for PCB

congeners. Bioconcentration and release of different PCB congeners seemed to be more

affected by physicochernical factors such as molecular size and chlorine substitution

patterns than by hydrophobicity, Thus, the magnitude of the accumulation of

hydrophobic organic compounds by oysters is not exclusively determined by the contents

of lipids in the organisms as previously speculated.
The most commonly reported distribution profile of PCB congeners in different

organisms is similar to that encountered in the commercial PCB mixture Aroclor 1254 or

in mixtures of Aroclors; 1254 and 1248 or 1260. This might point to these mixtures as the

most probable sources of the PCB congeners in different organisms. However, as a

consequence of physicochemical factors that discriminate against the uptake of different

congeners, the organisms may present a distribution of PCB congeners similar to that

129

found in Aroclor 1254 or in mixtures of Aroclors 1254 and 1248 or 1260 even if the

organisms are exposed to a more complete suite of PCB congeners in the environment.

Ile influence of these physicochernical factors in bioaccumulation is particularly evident

in the case of planar PCB congeners. Compared to other PCBs within the same level of

chlorination, planar congeners take a longer time to equilibrate into the lipid pools of the

oysters (t90%). Similarly, the time needed for depuration of these from the oyster tissues

was longer and this is paralleled by their significantly longer biological half-lives.

In general, TBT showed a biological half-life slightly longer than those observed for

most PAHs, comparable to the values calculated for low molecular weight PCB congeners

and significantly shorter than the half-lives estimated for most high molecular weight PCB

congeners.

PAHs, PCBs and TBT were, in most cases, depurated faster by newly contaminated

oysters than by chronically exposed individuals. A two-compartment model seems to be

more appropriate to explain this phenomenon than a single one.

130

CHAPTER VII

SIMULTANEOUS UPTAKE AND DEPURAnON OF PAHs AND PCBs BY THE

AMERICAN OYSTER (CRASSOSTREA VIRGIMCA)

IN'17RODUCTION

PCBs and PAHs are known to be highly toxic to marine organisms (Hargis er al.,

1984; Malins et al., 1984, 1987) and interactions between the two groups of contaminants

are known to occur. PCBs, for example, can affect the toxicity of PAHs. Hawkes

(1979) observed a more severe intestinal sloughing in marine chinook salmon when

simultaneously exposed to PCBs and petroleum incorporated in food relative to separate

exposures at similar concentrations. PCB exposure induces in vitro hepatic metabolism

and DNA binding of benzo(a)pyrene in rainbow trout (Egaas & Varanasi, 1982).

Not only is the toxicity of these xenobiotics affected by PAH-PCB interactions but

also their biological fate. ne availability of PCBs to benthic organisms, for example, is

reduced by the presence of oil in the substrate (Meier & Rediske, 1984). A prior
exposure of Coho salmon to Aroclor 1254 substantially altered the biological disposition
of [14C]-labeled dimethylnaphthalene in the fish by increasing the levels of
dimethy1naphthalene metabolites (Collier et al., 1985). Bioaccumulation of [14C]-
naphthalene by oysters decreased with the simultaneous exposure to [14C]-labeled PCBs
,and 3H-benzo(a)p yrene (Fortner & Sick, 1985). The same study indicated that
accumulation of a (14C]-labeled PCB mixture by oysters was not always antagonistically

131

affected by simultaneous exposure to the three contaminants. Tissue accumulation of
[3H]-benzo(a)pyrene was not significantly affected. It has also been shown that
simultaneous exposure of English sole to sediment- associated [3H]-benzo(a)pyrene (BaP)
and [14C]-PCBs significantly increased the concentrations of BaP-derived radioactivity

and decreased the concentrations of PCB-derived radioactivity in some tissues (Stein et

al., 1984). Several other studies have demonstrated PCB induction of many of the

enzymes responsible for the metabolism of PAHs in different aquatic species (Moore er

al., 1980; Spies et al., 1982; Anderson, 1985; Livingston, 1985).

In this chapter the uptake and depuradon of selected PAHs and PCBs by the American

oyster (Crassostrea virginica), exposed in the laboratory to particle-associated PAHs,

PCBs and PAHs plus PCBs, are discused.

SIMULTANEOUS EXPOSURE TO PAHs AND PCBs: A LABORATORY STUDY

Aquarium Exposure

The aquarium exposures were conducted simultaneously with the transplant

experiments in Galveston Bay (Chapters 11 and III). Oysters collected by dredge from the

Hanna Reef area were transfered as soon as possible to 40 1 glass aquariums and adapted

to laboratory conditions for 7 days prior to the experiments. One hundred and twenty

oysters were exposed in three aquariums, forty organisms per aquarium, to particles

containing PCBs, PAHs and both PCBs and PAHs. Kaolin (Al2H2Si2O8.H20), which

was found to have benefical effects on oystes growth (Langdon & Siegfried, 1984), was

used as the adsorbant for PCBs and PAHs in this study. A fourth aquarium was used as

a control. Uptake experiments were perfon-ned at one dosing level. The aquarium set-ups

were as follows*.

132

Aquarium A. Oysters were exposed to uncontaminated suspended particles at a nominal
concentration of 10 mg 1-1.
Aquarium B. Oysters exposed to particle-associated PCBs. Nominal concentrations of
suspended particles and total exposure PCBs were 10 mg 1-1 and 10 Ag g-1,
respectively. Nominal aquarium total PCB concentration was 0.10 gg 1-1

(approximately 0. 1 ppb).
Aquarium C. Oysters exposed to particle-associated PAHs. Nominal concentrations of
suspended particles and total exposure PAHs were 10 mg 1-1 and 240 Ag g-I

respectively. Nominal aquarium total PAH concentration was 2.4 gg

(approximately 2.4 ppb).

Aquarium D. Oysters exposed to particle-associated PCBs and PAHs. Nominal

concentrations of suspended particles and total exposure PCBs and PAHs were 10 mg
1-1, lOgg g-I and 240 gg g-1, respectively. Nominal aquarium total PCB and PAIH
concentrations were 0. 10 gg 1- 1 (0. 1 ppb) and 2.4 gg V 1 (2.4 ppb), respectively.

Ile general experimental set up is shown in Fig. 36a. Each aquarium was designated

to use recirculated seawater simulating a flow-through system (Fig. 36b). Briefly,

seawater was recirculated by gravity through an activated charcoal-glass wool-

polyurethane foam filter and air pumped back into the aquarium. Activated charcoal and

polyurethane foam plugs act as solid adsorbant retaining dissolved PCBs and PAHs as

well as gases and metabolic products from the test organisms (Gesser et al., 1971; Basu

& Saxena, 1978; Afghan et al., 1984). Activated charcoal and foam plugs were changed

daily. Foam plugs were washed with water, extracted three times with acetone and air
dried prior to their, use in the filtering systems. PCBs, PAHs and PCBs plus PAHs,
adsorbed onto particles at environmental realistic concentrations, were pumped with a

peristaltic pump (Fig. 36c) and mixed with the seawater entering the aquariums after

A-

'#4

Air

Fig. 36. General laboratory sct-up (a), details of an aquarium and water recirculation systent (1)), (1
and feeding system (c), and oysters used during the experiment (d).

134

aeration to prevent losses of the most volatile contaminants during this process. A
0'
different line out of the peristaltic pump was used to feed the oysters. The bivalves were

continuously fed with a mixture of two algae, Thalassiosira fluviatilis and Isochrysis r-

galbana, raised on an f/2 algae food mixture. Temperature, pH, salinity, suspended

particles and recirculation flow for each aquarium were monitored daily. Uptake studies

lasted one month. Groups of five oysters, water and suspended particle samples were

collected from each aquarium during the 3rd, 7th, 15th, and 30th days after the

experiments started. A total of 20 oysters per aquarium were sampled during uptake

experiments. Fig. 36d shows the sizes of the laboratory exposed oysters.

For depuration studies, groups of five oysters were sampled from each aquarium

during the 3rd, 7th, 15th, and 30th days after the contaminant inputs were discontinued

and the organisms were transferred to clean seawater. A total of 20 oysters per aquarium

were sampled during the depuration period. Water samples from each aquarium were also

collected.

Temperature, pH, salinity and recirculation flow for each aquarium were checked
daily. With an average flow of approximately 5 1 h- I and a volume of water in the

aquariums of 40 1, 95% of the water in the aquariums was filtered every 24 h (Spague,

1969). When necessary, the salinity of the aquarium was adjusted with HPLC water to

the starting value of 18%o. Particle concentrations in the aquariums were in the range of
6.4 to I I mg 1- 1, a concentration range commonly reported for coastal marine

environments (see, for example, Cadee, 1982; Colijn, 1982). Mortality of the exposed

oysters was minimal throughout the experiment (one oyster in Aquarium D). Control

oysters showed little change in analyte concentrations during the 60-day exposure and

depurati on experiments. The reported concentrations correspond to five pooled oysters.

135

Extraction,fractionation and instrumental analyses of PAHs and PCBs

The extraction and fractionation, as well as instrumental analyses of PAHs and PCBs

were discussed in Chapters II and III, respectively.

Polynuclear Aromatic Hydrocarbons

PAH concentrations measured in oysters collected from Aquariums A (control), C

(PAHs), and D (PAHs plus PCBs) during the uptake and depuration experiments are

plotted in Fig. 37. In general, exposed oysters rapidly accumulated four- and five- and

some three-ring compounds. In this molecular range, some PAHs reached an apparent

steady state concentration 10 days after the start of the experiments (e.g., I-

methylphenanthrene and pyrene). Most of the analytes, however, had not reached a

concentration plateau after 30 days (e.g., benz(a)anthracene, chrysene, benzo(e)pyrene

and perylene). Two- and most of the three-ring PAHs were detected at low

concentrations in both groups of oysters.

The PAHs accumulated in highest concentration were the same in organisms exposed
to PAHs alone or simultaneously to PAHs plus PCBs (Fig. 38). However,

concentrations of individual PAHs in oysters exposed solely to PAHs were, at the end of

the 30-day exposure period, lower than the concentrations encountered in oysters exposed

to the mixture PAHs plus PCBs. The PAHs accumulated to the highest concentration

after 30 days in oysters exposed only to PAHs were: ben zo(b)fluoranthene,

benzo(e)pyrene, benz(a)anthracene, chrysene and indeno-[1,2,3-c,d]pyrene whereas the

accumulation order found in second group of oysters was: benzo(b)fluoranthene,
benz(a)anthracene, benzo(e)pyrene, chrysene and indeno[1,2,3-c,d]pyrene. Most of

these PAHs were also preferentially accumulated by HRSC and SC oysters under field
conditions. Therefore, the laboratory uptake confirms the environmental findings. When

exposed to a wide molecular weight range of PAHs, i.e. two- to six-ring compounds,

1000. Benz(a)anthracene 1000. Chrysene

fool 100,

Aquarium A
Aquarium A at
to- Aquarium C 10 Aquarium C
Aquarium D
Aquarium D

U U
A - 4
0 10 20 30 40 50 60 0 to 20 30 40 so 60

Time (days) Time (days)

Denzo(e)pyrene Peryleve
1000 loon
Aquarium A
Aquarium C
100 100 Aquarium D
V
Aquarium A a
Aquarium C
10 to,
Aquarium S

0 to 20 30 40 50 60 0 to 20 30 40 so 60

Time (days) Time (days)
Fig. 37. Concentrations of selected polynuclear aromatic hydrocarbons in tissues of oysters during exposure to particle-
associated PAHs alone (Aquarium Q and PAHs + PCBs (Aquarium D) and following transplant to contaminant-free
aquariums. Aquarium A was tised as control.

137

Aquarium C
Naphthalene
2-Methylnaphthalene Aquarium D
I-Methylnaphthalene
Biphenyl
2,6-Dirnethy1naphthalene
Acenaphthylene
Acenaphthene
2,3,5-Trimethy1naplithalene
Fluorene

Phenanthrene

Anthracene
1-Methylphenanthre e

Fluoranthene
Pyrene
Benz(a)antliracene
Chrysene
Benzo(b)(luoranthene . . . . . . . . .
Benzo(k)fluoranthene ..........
Benzo(e)pyrene
Benzo(a)pyrene
Perylene
Indenoj1,2,3-c,djpyrene
Dibenz(a,h)anthracene
Benzo(g,hJ)perylene

0 100 200 300 400 500 600 700

Concentration (ng/g, dry wt.)

Fig. 38. Concentrations of individual polynuclear aromatic hydrocarbons in tissues of
laboratory exposed oysters after the 30-day exposure period to parficle-associated PAHs
(Aquarium Q and PAIIs + PCBs (Aquarium D).

138

oyster preferentially bioconcentrate those analytes having four and five rings. This

preferential uptake is unrelated to the presence or absence of PCBs.

When the input of contaminants was stopped, both oyster groups showed statistically

significant depuration of most of the PAHs accumulated during the first phase of these

experiments. However, as previously discussed for environmentally contaminated

oysters, they were unable to reach the low concentrations encountered for some of these

analytes before the exposure. Fig. 39 compares the final concentrations of PAHs at the

end of the 30-day depuration period in oysters exposed to particle-associated PAHs and

PAHs plus PCBs. At the end of the 30-day depuration period, the total PAR loads in

both groups of exposed oysters were dominated by heavier molecular weight PAHs, i.e.

four- and five-ring compounds.

Oysters that had been exposed simultaneously to PAHs and PCBs depurated PAHs at

a faster rate than oysters exposed only to PAHs. Calculated half-lives for both groups of

oysters are shown in Table 9. In PAH exposed oysters, the estimated half-lives ranged

from 9 (fluoranthene and pyrene) to 25 (ben zo (b)fl uoranthene/benzo(k)fl uoranthene)

days. Comparatively, PAHs plus PCBs exposed oysters yielded PAH half-lives ranging

from 6 (pyrene) to 15 (benzo(e)pyrene) days. Most of the values were, however, in the

range 15 to 17 days and 8 to 10 days, for the first and second groups of oysters,

respectively.

In general, the estimated half-lives are in good agreement with previously published
values and with the calculated clearance rates for HRSCHR and SCHR oysters presented

in Chapters II and VI; however, some differences exist. First, it is evident that, except for

2,3,5-trimethylnaphthalene, anthracene, 1-methylphenanthrene and fluoranthene, the half-

lives calculated for oysters exposed simultaneously to a mixture of PAHs and PCBs

compare better to the environmental half-life values estimated for HRSCHR oysters than

the values obtained from the oysters exposed only to PAHs. This is consistent with the

139

Aquarium C
Naphthalene
2-Methyinaphthalene Aquarium D
I-Methylnaphtha)ene
Biphenyl
2,6.Dimethylnaphthalene
Acenaphthylene
Acenaphthene
2,3,5.Trimeth3,lnaphthalene
Fluorene

Phenanthrene

Anthracene
1-Methylphenanthrene
Fluoranthene

Pyrene
Benz(a)anthracene
Chrysene
Benzo(b)nuoranthene
Benzo(k)nuoranthene
Benzo(e)pyrene
Betizo(a)pyrene
Perylene
Indenol 1,2,3-c,dlpyrene
Dibenz(a,h)anthracene
Benzo(g,h,i)perylene

100 200 300

Concentration (ng/g, dry wt.)

Fig. 39. Concentrations of individual polynuclear aromatic hydrocarbons in tissues of
oysters previously exposed in the laboratory to particle-associated PAHs (Aquarium Q
and PAHs + PCBs (Aquarium D) after the 30-day depuration period in contaminant-free
aquariums.

140

TABLE9
Biological Half-Lives of PAHs in Crassostrea virginica Oysters Exposed in the
Laboratory to Particle-Associated PAHs Alone (Aquarium Q and PAHs + PCBs
(Aquarium D).

Analyte Biological Half-Lives (R2)a
Aquarium C Aquarium D

2,3,5-Trimethylnaphthalene 16(0.75) 10(0.91)
Anthracene 16(0.68) 8(0.89)
I-Methylphenanthrene 16(0.72) 7(0.88)
Fluoranthene 9(0.78) 7(0.85)
Pyrene, 9(0.75) 6(0.89)
Benz(a)anthracene 16(0.81) 9(0.99)
Chrysene, 22(0.86) 12(0.98)
Benzo(b)fluoranthene/
Benzo(k)fluoranthene 25(0.94) 13(0.95)
Benzo(e)pyrene 21(0.93) 15(0.98)
Benzo(a)pyrene 12(0.87) 9(0.78)
Perylene 15(0.96) 10(0.89)
Indenof 1,2,3-c,d]pyrene 10(1.00) 8(0.78)
Dibenz(a,h)anthracene 11(0.97) 10(0.69)
Benzo(g,h,i)perylene 16(0.96) 11(0.92)

a R2 = square of the correlation coefficient for the regression equation.

161

When combined, these congeners accounted for 43.9 and 36.7% of the total PCBs in

oysters and sediments, respectively.

XPCB Congeners/Total PCB Relationship

Several methods have been used to quantitate PCBs in environmental samples. In the

past, for example, PCB concentrations have been expressed as the equivalent Aroclor

mixtures (e.g. Bopp er al., 198 1; Pugsley et al., 1985; Brownawell & Farrington, 1986)

or as their similar foreign technical formulations, e.g. Clophen (Eder et al., 1981) or

Phenochlor (Elder er al., 1979). An accurate determination of total PCBs in

environmental samples would have to be carried out with the use of each individual

congener as reference material (Duinker et al., 1980). With the introduction of capilary

columns and the availability of almost every individual PCB congener as a standard,

several reserchers have attempted to report total PCB concentrations as the sum of all the
measurable individual congeners. However, some of these' congeners are not alway s

separated from other congeners on a single GC capillary column (Duinker et al., 1988a).

Duinker et al. (1988b) suggested a number of different congeners, in addition to those

recommended by ICES, that could be accurately analyzed in environmental samples.

These congeners cover all the levels of chlorination and satisfy the condition of good GC

separation on an SE-54 or similar capillary column. A variant of this approach was

initially used in reporting the total PCB concentrations in oyster and sediment samples

collected from the Gulf of Mexico as pan of the NS&T "Mussel Watch" Program.

During 1986 and 1987, 18 different congeners (i.e. PCB 8, 18, 28, 44, 52, 66, 101,

105, 118, 128, 138, 153, 170, 180, 187, 195, 206 and 209) were supplied by NIST,

formerly NBS, for making quantitation standards. These congeners, which are some of

the major congeners found in commercial Aroclor mixtures, are among those commonly

reported in environmental samples. Nine of these congeners (i.e. PCB 8, 28, 52, 101,

162

153, 170, 195, 206 and 209), specified by NOAA, were used as reference congeners,

representative of a given degree of chlorination from C12 to Clio, to determine other

congener concentrations at each level of chlorination. Results were reported as the sum of

congeners within each level of chlorination and total PCB as the sum of these amounts.

One obvious problem with this method of quantitation is the different relative response

factor for each congener. Thus, a congener that does not have a standard to be directly

compared to might be underestimated or overestimated because of the difference between

its relative response factor and that of the corresponding representative congener.

Discussions among the different participating laboratories in this program directed to

improve the PCB reporting led to the adoption of an equation that relates the sum of the 18

individual congener concentrations in the samples with the total PCB loads. Fig. 48

shows, for example, the correlation encountered in oyster samples collected in the Gulf of

Mexico during 1986. Therefore, starting in 1988, total PCB concentrations in oyster and

sediment samples from the Gulf of Mexico, Atlantic and Pacific coasts, including Hawaii,

were estimated and reported using this new approach.

During the first year of the NS&T program total, PCBs ranged from 10 to 4,020 ng
9_1 (Sericano et al., 1990a). Nearly 95% of the samples had a total PCB load below 500
ng g-l. Therefore, the correlation between the sums of the 18 individual PCB congeners

and the total PCB concentrations in oyster samples is likely to be affected by a small

percentage of samples collected from heavily contaminated sites. Table I I shows

correlations for the same data set when ranges are chosen to eliminate the bias introduced

by the highly contaminated samples. For example, just eliminating the highest PCB

concentration measured in a sample collected near the Yacht Club, in Galveston Bay,

reduces the near-perfect correlation of 0.99 to 0.97. Considering samples with total PCB
concentrations of 500 ng g- I or lower (i.e. 95% of the samples) decreases the correlation

to 0.92 (Fig. 48b). Further reductions of the data set to maximum concentrations lower

A
Q 4000, 300.

400-
3000

me
20"'.
2"'.
It"
I @7,2574 JAMx RA2 0."3 Y = 12.777 I-7771z RA2 0.922
= 144 0 = 137
0 so@ Is'6 2;00' 2SOO IO'o 200 300
Total 1: congen:rSsOO(ng1g) Total IS congeners (ng/g)

200- C

a a Joe a 60-

40 *0

Y= 11.198 + 1.78S3z RA2 0.902
Y = 18-192 + 1.29531 RA2 0.491
a= = 72

0 20 44 60 100 0 Is 20 30 40
Total IS congeners (ng/g) Total 18 congeners (ng/g)

Fig. 48. Relationships between the sum of 18 selected PCB congeners and the total PCB load encountered in Gulf of
Mexico oysters for the first year of NOAA's National Status and Trends Program. See text for discussion.

164

TABLE 11
PCB Congeners/Total PCB Relationships in Gulf of Mexico Oyster Samples.

n Total PCBa Fraccion Regression equation R2
ngg-I %

1986
144 all data 100 YPCB=7.25+1.89ICongb 0.99
143 :52000 99 Y-PC13=6.52+1.901Cong 0.97
140 :51000 97 Y-PCB=12.29+1.79Y-Cong 0.95
137 :5500 95 YPCB=l2.78+l.78YCong 0.92
108 :5157 75 YPCB=l 1. 1 1+1.71YCong 0.80
66 :586 50 I:PCB =I 8.19+1.29Y-Cong 0.49

1987c
149 all datad 99 Y-PCB--0.81+2.30ICong 0.96
140 :5300 95 Y-PCB=13.8+1.89Y-Cong 0.81
129 -<-200 87 Y-PCB=l 2.5+1.837-Cong 0.75

aupper limit of the data range corresponding to the total PCBs calculated from
level of chlorination; bsurn of 18 individual PCB congeners; CBrooks et al. (1988);
dtwo outliers eliminated from regression analysis.

165

than or equal to 157 ng g- 1 (75% of the samples) or to 86 ng g- 1 (50% of the samples)

will reduce the correlation coefficient to 0.80 and 0.52, respectively (Fig. 48c and d,

respectively). Similar correlations were reported for oyster and sediment samples

collected during 1987 (Brooks et al., 1988; Table 11). Joiris & Overloop (1991) showed

the correlations between the sum of nine of the most "classical" congeners (i.e. PCB

congeners 28, 52, 101, 118, 138, 153, 170, 180 and 194) and total PCBs expressed as

Aroclor 1254 as well as correlations between individual congeners and total PCBs in

particulate matter (mainly phytoplankton) and netplankton (mainly zooplankton with some

phytoplankton) samples collected in the Indian sector of the Southem Ocean. Although
no correlation factors are given in the report, coefficients of regressions (R2) between the

sum of the nine congeners and total PCB concentrations, estimated ftom new plots made

from their figures, were about 0.85 and 0.56 for particulate matter and netplankton

samples, respectively. Total PCB concentrations in particulate matter and netplankton
samples ranged from about 200 to 2900 n g g- I and from 70 to 5 10 ng g- I on a dry

weight basis, respectively. This data analysis indicates that the correlation between the

sums of individual PCB congeners and the total PCB concentrations in environmental

samples appears to be dependent on the total PCB load.

Although it has been shown that after exposure to a wide range of molecular weight

PCBs oysters will preferentially uptake four-, five-, and six-chlorine substituted

congeners (see Chapter VII), there might be differences in the residual PCB profiles

among oyster samples collected from different'geographical areas as a result of a variety
of local sources. The 1congeners/total PCB ratios calculated from the data reported by

Schulz et al. (1989) encountered in Aroclor mixtures 1016, 1242, 1254 and 1260 were

2.61, 2.86, 2.63 and 2.55, respectively. These ratios can be modified in the environment

as a consequence of the differential physico-chemical and biological properties of

individual congeners controlling their water transport, bioaccumulation, etc. Different

166

residual PCB compositions in oysters will obviously produce different results. For

example, total PCB concentrations, calculated as the sum of all measurable PCB

congeners, in oyster samples collected near the Houston Ship Channel area in Galveston

Bay over a seven-week period (see Chapter III for more details) yielded concentrations

that constantly were between 30 to 35% higher than the concentrations estimated with the

above correlation (Fig. 49). Another way to illustrate this assertion is to consider the

PCB, profiles encountered in uncontaminated Hanna Reef oysters when transplanted to the

upper Galv eston Bay area near the Houston Ship Channel (Fig. 17, Chapter III). During

this experiment, low molecular weight PCBs were bioaccumulated at a faster rate than

congen ers with higher level of chlorination. By the end of the 48-day exposure period,

the amount of total PCB estimated by the equation was up to 60% lower than the total

PCB load measured as the sum of all individual congeners (Fig. 49).

Although the transplantation experiment can be compared to an extreme case of a rapid

environmental PCB, contamination, similar disequilibrium between PCB concentrations in

oysters and environmental leves might also be a consequence of natural processes related

to the bivalves themselves such as spawning. It has been suggested that high variability

in xenobiodc concentrations in bivalves from a given location might be more related to the

stage of the reproductive cycle and its associated biochemical modifications than to

environmental changes (Jovanovich & Marion, 1987). Most organisms have a marked

increase in their lipid contents during gametogenesis, which is followed by a drastic loss

of lipidic material with the gametes at spawning (Phillips, 1986). Since most
hydrophobic trace organic c6ntaminants will tend to concentrate in lipid-rich tissues, such

as eggs, it is evident that their concentration will vary with the sexual cycle. Release of

hydrocarbons and pesticides during spawning has been reported for Mytilus edulis and

Crassostrea virginica, respectively (Lowe et al., 197 1; Fossato & Canzonier, 1976).

167

2000- Indigenous
Ship Channel Oysters 0 Regression Equation
a Sum of Individual Congeners

1500-.

100

7 17 30 49
Sampling Day
12oo - Transplanted s 0 Regression Equa tion
Hanna Reef Oyster
U Sum ofIndividual Congeners
1000'

4",
U
G:
U 2"

3 7 17 30 48
Length of Exposure (days)

$00. Laboratory Exposed
Hanna Reef Oysters 0 Regression Equation
0 Sum of Individual Congeners
400,

3 7 is 30 48
Lenght of Exposure (days)

Fig. 49. Three'different examples of the bias introduced in the report of total PCB
concentrations by using the regression equation (see text) compared to the total PCB
load calculated as the sum of all measurable individual congeners.

168

It may be concluded that even though there is a reasonable correlation between the

sums of 18 individual PCB congeners and the total PCB concentrations in oyster samples

and it might provide an estimation of the total PCB load, the preceeding discussion

indicates that it must be applied with caution when reporting and interpreting

environmental data. The greatest disadvantage of this procedure is that much of the

information is lost when complete congener characterization of PCB residues in

environmental samples is not reported. This is emphasized by the fact that PCB

composition changes drastically as they move from one environmental compartment to

another.

Planar PCB Congeners

PCB congeners have been widely reported in oyster samples collected as part of this

program in the Gulf of Mexico (Seiicano et al, 1990a); however, the occurence of toxic

planar PCB congeners, i.e. 77, 126 and 169, have not until recently been reported

(Sericano et al., 1992).

The concentrations of planar congeners, as well as the concentrations of selected

predominant mono- and di-ortho substituted congeners and total PCBs in oyster samples

from sites in Galveston and Tampa Bays (Fig. 50), collected during winter 1990-1991

(year 4 of the NS&T program), are summarized in Table 12. In Galveston Bay, the

highest concentration of these planar PCBs was found in samples collected near the area

where the Houston Ship Channel enters the upper Galveston Bay (GBSC) and decreases

seaward. Ile second highest total concentration was encountered in samples from a site

near the, city of Galveston (GBOB). The general distribution of planar congener

concentrations in Galveston Bay clearly shows high values near population centers. The

same correlation between urban centers and concentrations of planar PCBs can be

tEXAS
A - TAMPA BAY OLD
A B - TAMPA BAY KMGHT
C - TAMPA BAY PAPYS,
D - TAMPA BAY NARVAJ
E - TAMPA BAY COCKR
F - TAMPA BAY AwLLET

).A
Dunedin. -
qD D
s- LE@
Clea- let.,
0 !-:Tamp

-VESTON C
G@d
F
GULF SL P@etersburg
D
fs-t OF
@b
PL Pmew@
e E
4@b
MEXICO

A - GALVESTON BAY SHIP CHAN74EL (GBSC)
B - GALVESTON BAY YACHT CLUB (GBYC)
C - GALVESTON BAY TODDs DUMP (GBTD)
D - GALVESTON RAY HANNA REEF (GBHR)
E - GALVESTON RAY CONFEDERATE REEF (GBCR) Anna
F - GALVESTON DAY OFFATS BAYOU (GBOB) Maria Island

Fig. 50. NOAA's National Status and Trends sampling locations in Galveston and Tampa Bay

170

TABLE 12
Planar and Total PCB Concentrations in Oysters (Crassostrea virginica) from
Galveston and Tampa Bays.

Sample Concentration of Planar PCBs Total PCBs
77 126 169
p9 9-1 p9 9-1 p9 9-1 ng g-I

Galveston Bay

GBSC 2,000 2,200 790 1,100�120
GBYC 330 210 190 210�14
GBID 140 120 54 110�18
GBHR 89 110 89 50�7.0
GBCR 100 94 51 77�9.6
GBOB 500 400 93 160�44

Tampa Bay

TBOT 170 320 280 55�8.5
TBKA 1,500 330 84 580�230
TBPB 85 100 51 75�27
TBNP 260 140 150 120�31
TBCB 200 290 100 49�20
TBMK ND ND ND 38�14

ND = not detected

171

observed in Tampa Bay. The highest concentrations were measur in samples collected

near Tampa (TBKA).

As expected from the small contributions of these planar congeners to the total

commercial PCB mixtures (Kannan et al., 1987; Schulz et al., 1989), these congeners

were detected at much lower concentrations than other mono- and di-ortho substituted

PCB congeners. In commercial PCB mixtures, the concentration of congener 77 is one to

two and three to five orders of magnitude higher than concentrations of congeners 126

and 169, respectively (Kannan et al., 1987). Therefore, it appears that congeners 126 and

169 are enriched with respect to congener 77 in oyster samples from Galveston and

Tampa Bays. This is not surprising since the log Kow (octanol-to-water coefficient)

increases with the number of chlorines substituted in the biphenyl rings (6.36, 6.89 and

7.42 for congeners 77, 126 and 169, respectively; Hawker & Connell, 1988). On

average, the sum of these three highly toxic congeners ranged from 0.26 to 0.62% and

from 0.31 to 1.40% of the total PCB load in Galveston and Tampa Bays, respectively.

In a review, Safe (1990) discussed the environmental and mechanis tic considerations

behind the development of the Toxic Equivalent Factor (TEF) concept. Safe proposed

provisional TEF values of 0.01, 0.1 and 0.05 for planar congeners 77, 126 and 169,

respectively. Recently, the validation and limitations of these factors have been reported
(Safe, 1992). Calculated 2,3,7,8-TCDD equivalents, in pg g- 1, in oyster tissues collected

from Galveston and Tampa Bay, as well as their averages, are listed in Table 13. In

Tampa and Galveston Bays, the total 2,3,7,8-TCDD equivalents ranged from 14 to 52 pg
9- 1 and from 13 to 280 pg g-1, respectively. The data show that, except for the sample

collected near the Houston Ship Channel, oysters from Tampa and Galveston Bays are

similar in terms of total toxicity. Oysters collected near the Houston Ship Channel

(GBSQ in Galveston Bay were clearly the most toxic. Thisarea is closed to commercial

172

TABLE 13
2,3,7,8-TCDD Equivalents (pg g- 1) Corresponding to Non-Ortho Substituted PCB
in Oysters (Crassostrea virginica) from Galveston and Tampa Bays.

Sample Congener TOW
77 126 169

Galveston Bay

GBSC 20 220 40 280
GBYC 3.3 21 9.5 34
GBTD 1.4 12 2.7 16
GBHR 0.9 11 4.5 16
GBCR 1.0 9.4 2.6 13
GBOB 5.0 40 4.7 50

Tampa Bay

TBOT 1.7 32 14 48
TBKA 15 33 4.2 52
TBPB 0.9 10 2.6 14

TBNP 2.6 14 7.5 24
TBCB 2.0 29 5.0 36

TBNM - - - -

173

or sport oystering due to bacteria concentrations; therefore, the high PCB levels are not a

human health thmat.

As discussed earlier, congeners 77, 126 and 169 are present at trace concentrations in

commercial PCB mixtures and at very low concentrations in environmental samples;

however, their mono-ortho derivatives (e.g. congeners 105, 118, 156 and 189) may be

more important in terms of both TCDD-like activity and occurrence (Safe, 1984). Certain

di-ortho derivatives of the m,m'pp' sustitution pattern (e.g. congeners 128, 138, 153

and 170) are significant components of PCB residues (Duinker et al., 1988a; Schulz et

al., 1989; Schwartz et al., submitted). Congeners 128, 138 and 170 have reduced

TCDD-like activity compared to their parent planar congeners whereas PCB 153 lacks of

TCDD-like responses (Hansen, 1987). Safe (1990) proposed provisional TEF values of

0.001 and 0.00002 for mono- and di-ortho chlorine substituted PCB congeners,

respectively.

The concentrations of PCBs 105, 118, 128 and 138 as well as total PCBs in oyster

samples from sites in Galveston and Tampa Bays are summarized in Table 14. These

congeners are derivatives of planar PCB 77. Individually, the concentrations of these

mono- and di-ortho congeners were, as expected, one to two orders of magnitude higher

than planar PCB concentrations. In order to assess the environmental significance of

these congeners in terms of TCDD-like effects in oyster samples from Galveston and

Tampa Bay, the calculated 2,3,7,8-TCDD equivalents (Table 15) are compared to those

corresponding to planar congeners. In spite of the relatively lower toxic effect of

congeners 105 and 118 compared to planar PCBs, these congeners might have a

significant toxic impact in the environment. Most of the relative toxicity in oyster,

however, are due to the presence of planar PCBs (53.8 to 94.3%; Fig. 51). Contribution

of congeners 105 plus 118 to the total 2,3,7,8-TCDD equivalents was as high as 45.4%;

in contrast, the contribution of di-ortho congeners is negligible (<1.0%). The lesser

174

TABLE 14
Selected Mono- and Di-Ortho Substituted PCB and Total PCB Average
Concentrations (ng g-1) in Oysters (Crassostrea virginica) from Galveston and
Tampa Bays.

Sample Congener Total PCBs
105 118 128 138

Galveston Bay

GBSC 39�4.1 48�5-8 4.4�0.6 50�6.7 1,100�120
GBYC 4.1�1.7 9-0�0.3 1.5�0.2 13�3.2 210�14
GBTD 1.3�0.2 5.2�1.0 0.6�0.2 5.7�1.1 110�18
GBHR 0.6�0.5 1.2�0.3 0.6�0.2 4.3�0.8 50�7.0
GBCR 0.7-+0.6 2.8�0.2 0.7�0.3 5.0-+1.4 77�9.6
GBOB 3.2�1.8 10�2.7 1.0�03 8.7�3.4 160�44

Tampa Bay

TBOT 0.4�0.2 2.4�1.6 0.2�0.2 4.0-+0.8 55�8.5
TBKA 7.6�3.7 36�15 2.0�1.0 30�13 580�230
TBPB, 0.4�0.1 3.0-+0.7 0.3�0.2 6.1�2.6 75�27
TBNP 1.3�0.2 7.3�1.8 0.6�0.2 8.9-+3.1 120�31
TBCB 0.4�0.2 3.0�1.1 0.2�0.2 2.8�1.2 49:L20
TBMK 0.3�0.2 1.6�0.3 0.2-+0.1 3.3�2.0 38�14

175

TABLE 15
Average 2,3,7,8-TCDD Equivalents (pg g-1) Corresponding to Selected Mono-
and Di-Ortho Substituted PCBs in Oysters (Crassostrea virginica) from Galveston
and Tampa Bays.

Sample Congener TOW Totala
105 118 128 138

Galveston Bay

GBSC 39 48 0.1 1.0 89 379
GBYC 4.1 9.0 <0. 1 0.3 13 47
GBTD 1.3 5.2 0.1 0.1 6.7 23
GBHR 0.6 1.2 <0. 1 0.1 1.9 18
GBCR 0.7 2.8 <0. 1 0.1 3.6 17
GBOB 3.2 10 <0. 1 0.2 14 64

Tampa Bay

TBOT 0.4 2.4 <0. 1 0.1 2.9 51
TBKA 7.6 37 <0. 1 0.6 45 97
TBPB 0.4 3.0 <0. 1 0.1 3.6 18
TBNP 1.3 7.3 <0. 1 0.2 8.8 33
TBCB 0.4 3.0 <0. 1 0.1 3.5 40
TBMK 0.3 1.6 <0. 1 0.1 2.0 2.0

aIncludes PCB congeners 77, 126, 169, 105, 118, 128 and 138.

176

0 PCBs 77,126,169
300-
12 PCBs 105, 118
250' 0 PCBs 128,138

200

150

100

50
el@ N
el@

0
GDSC GErYC GErM GBHR GBCR C33013 IMOT TBKA TBPB TBNP TBCB 7BMK
Galveston Bay Tampa Day

Fig. 51. Toxic equivalents corresponding to three planar PCBs and selected mono-
and di-ortho chlorine-substituted congeners in oyster samples collected from six
different locations in Galveston and Tampa Bays.
L d- L,_d

177

toxicity of the di-ortho congeners is a consequence of their much lower TCDD-like

activity rather than lower concentrations. As shown in Fig. 52, most of the toxicity

corresponding to planar PCBs is contributed by congener 126 while that corresponding to

mono-ortho derivative s is due to congener 118.

Although none of the other PCB congeners considered to be inducers of hepatic aryl

hydrocarbon hydroxylase (AHH) activity, i.e. congeners 123, 114, 158, 166, 167, 156,

157, 170 and 189), have been quantitated in Galveston and Tampa Bay oyster samples,

the concentration of most of them in commercial Aroclor mixtures are very low (Schulz et

al., 1989). With the exemption of congeners 123, 158 and 170, which range from the

minimum reporting level (<0.05%) to 0.81, 1.55 and 3.91%, respectively, in different

Aroclor mixtures, the contributions of the rest of the individual AHH-active congeners are

below 0.30%. Therefore, the contribution of these mono- and di-ortho AHH-active

PCBs to the total toxicity of environmental samples is expected to be negligible. For

example, congeners 77, 126, 169, 105 and 118 accounted for nearly 99% of the total

toxicity, calculated as the sum of the toxic equivalents of each individual AHH-active

congener, encountered in eggs (Smith et al., 1990). Thus, it can be speculated that the

total toxic equivalents reported for oyster samples collected in Galveston and Tampa Bays

(Table 15) would not increase by more than 10% of the total TEF values if all the AHH

active PCBs had been analyzed.

BUTYLTIN SPECIES

The decision to include butyltin compounds as pan of NOAA's NS&T Program was a

consequence of the increasing concern about the adverse effects of TBT to non-target

organisms. Thus, butyltin compounds have been monitored in Gulf of Mexico oyster

samples since 1986.

Percentage

0 .........

Cr
C
M.
0=
0
0

E@
Co

C
Cr 13

179

The use of TBT in antifouling paints in the U.S., on vessels under 25 m, was banned

in 1988 (U.S. EPA., 1987). In that year, the reported average concentration of TBT in
bivalves for U.S. coastal sites was 366 ng Sn g- I (Wade et al., 1988b). In general, the
body burden of the butyltin species was TBT>DBT>MBT. With a half-life of 34 days-1

in chronically contaminated oyster (see Chapter V), it would have taken about 240 days

(0.6 years) for the average concentrations encountered in Gulf of Mexico oysters of TBT
to be below the present detection limit (5 ng Sn g- 1) if all the inputs were stopped at that

time. Obviously, this is not a realistic estimation because the use of TBT was not

completely banned and because there might be a number of boats in use that had been

painted just before the ban. Also, TBT present in sediments, with a reported half-life of

more than 20 weeks, may be a long term source of TBT to the environment (Harris &

Cleary, 1987; Johnson et al., 1987; Maguire, 1986; Stang & Seligman, 1987; Unger et

al., 1987; Valkirs et al., 1986, 19 87 b).

Some changes can be seen, however, at some areas that have been followed since the

beginning of the Status and Trends Program. As mentioned before, it coincides with the

ban on the use of TBT-containing paints in U.S. waters. For example, Naples Bay,

Florida, has a very heavy recreational boating activity. At this site, a decreasing trend in

the total concentration of butyltins has been observed since 1988 (Fig. 53). A similar

decrease has been detected at Biloxi Bay, Mississippi. Under the actual input/degradation

conditions, it seems that a decrease of 50% in environmental TBT concentrations in these

and other areas similar to Naples and Biloxi Bays takes about 2-3 years. This is about an

order of magnitude larger than the time needed for oysters to depurate when transplanted

to a clean environment. At this rate, and assuming no environmental redistribution of

TBT, its concentrat ions in oysters from sites like Naples and Biloxi Bays should be below
the present detection limits (i.e. 5 ng Sn g-1) in 8 to 12 years. The much slower

decreases at sites with extensive recreational boating suggests that in spite of the

200' . Naples Day (FL) a TOT 2000 Biloxi Day (FL) a TOT
0 DOT a DOT
III MOT 0 MOT
IS" -
to to

Hn 1000

C
0 soll Soo
U U

1986 1987 1999 1989 1990 1991 1992 1986 1987 1988 1989 1990 1991 1992
Year Year

20" 2ooo-
Galveston Bay (TX) a TOT Galveston Bay (TX) III TOT
Todd's Dump
Yacht Club a DOT 0 DOT
a MDT
0 MOT rA 1500-
V) Moo-
to to

10"

C
500-
500
U
0 M-A M.- 0
1986 1997 1999 1999 1990 1991 1992 1996 1997 1988 1989 1990 1991 1992
Year Year
Fig. 53. Total butiltin concentrations at selected sites in the Gulf of Mexico sampled between 1986 and 1992 as part of
NOAA's National Status and Trends Program.
M

181

restrictions applied in 1988 to the use of paints containing TBT a decreasing but still

significant amount of TBT is being introduced into the coastal marine environment. These

inputs may be from boats painted before 1988, TBT in sediments and/or TBT usage on

larger vessels. Other areas with important recreational boating activities but also heavy

maritime usages, like Galveston Bay, Texas, did not show any decrease and, even with

the actual restrictions to the TBT usage, no decreases in the near future may be found.

182

CHAPTERIX

SUMMARY AND PROSPEC77VES

Polynuclear aromatic hydrocarbons (PAHs), low molecular weight polychlorinated

biphenyls (PCBs), i.e. di-, tri- and tetrachlorobiphenyls, and tributyltin (TBT) were

rapidly bioaccumulated by oysters under environmental conditions. Apparent steady state

concentrations for these analytes were reached after 20 to 30 days of exposure. In

contrast, high molecular weight PCBs did not reached an equilibrium plateau at the end of

the seven week exposure period to relatively high PCB concentrations. However, the still

increasing concentrations encountered for these PCBs by the end of the exposure period

suggest that, given enough time, the equilibrium concentrations would eventually be

reached. When back-transplanted to their former location near Hanna Reef, originally

uncontaminated oysters depurated PAHs, low molecular weight PCB congeners and TBT

at similar rates while the heavier molecular weight PCB congeners were depurated at

considerably slower rates. In neither case, however, the original background

concentrations were reached after the 50-day depuration period.

Chronically contaminated Ship Channel oysters were also transplanted to the Hanna

Reef area during the second phase of the field experiment in Galveston Bay to allow for a

direct comparison with newly contaminated Hanna Reef individuals. In general, the

observed clearance, rates in Ship Channel oysters were slower than those exhibited by for

Hanna Reef bivalves. The differences might be explained as a consequence of different

distributions of PAH, PCB and TBT in the various body compartments in chronically

183

exposed oysters compared to recently contaminated individuals or a more effective

clearance response by originally uncontaminated oysters. A combination of both of these

processes should not be disregarded.

The present study presents evidence to substantiate the theory that the rates of uptake

and depuration of PCB congeners by the oyster Crassostrea virginica decreases as the

number of substituted chlorines in the two phenyl rings increases. However, in spite of

their lower uptake rates compared to low molecular weight congeners, the

pentachlorobiphenyls were the congeners bioaccumulated to the highest concentrations. It

was also observed that although heavier molecular weight congeners, i.e. heptachlorinated

biphenyls or higher, are more liphophilic, they have less favorable steric configurations,

which antagonistically affected their bioaccumulation and latter depuration by oysters.

Thus, bioconcentration and clearance of different PCB congeners appear to be more

affected by molecular size, e.g. molecular volume and cross-sectional area, which are

directly related to the number of chlorines substituted in the two phenyl rings and their

substitution patterns, rather than by hydro phobicity.

The influence of the chlorine substitution patterns in the bioaccumulation of PCBs by

oysters is particularly evident in the case of the highly toxic planar congeners, i.e. PCBs

77 and 126. Compared to other PCBs within the same level of chlorination, these planar

congeners take a longer time to equilibrate into and out of the organism's lipid pools.

Because of this, the importance of lipid content in oysters in determining potential

environmental hydrophobic organic accumulation might not be as significant as usually

speculated. Furthermore, the tendency for larger organic molecules to be less

concentrated in the lipidic pools of the organisms as a consequence of unfavorable steric

configurations suggests that these large molecules may also partition less easily into the

cells. Because of this low diffusivity among the different compartments in the organism,

it may take longer for the larger molecules to reach toxic concentrations.

184

The identification of the source of PCB congeners can also be confounded by the

differential PCB congeners uptake by oysters. Oysters exposed in the laboratory to a

wide molecular range of PCB congeners (1: 1: 1: 1 mixture of Aroclor 1242, 1248, 1254

and 1260), preferentially bioaccumulated congeners with four, five and six chlorines per

molecule resulting in a PCB profile similar to the distribution of homologs that would be

encountered in an approximately 2:1 mixture of commercial Aroclors 1248 and 1254.

Similar distribution of homologs has been observed in transplanted Hanna Reef oysters

during the field study near the Houston Ship Channel in Galveston Bay. Comparatively,

the profile of PCB homologs in indigenous Ship Channel oysters, exposed longer to the

the local levels, had a distribution profile with a slighly larger contribution of Aroclor

1254 (approximately 1: 1 Aroclor 1248 and 1254). Although it can be speculated that the

profile distributions encountered in chronically contaminated and newly exposed oysters

are the result of exposure to Aroclors 1248 and 1254 sources, it seems very probable that

the observed profiles are a consequence of the congener uptake discrimination from a

more complex mixture. However, it could also be that, even with a more complex

mixture of different Aroclors, there was a natural fractionation of the low, middle and

high molecular weight congeners. It is well known that a PCB mixture can not be

considered as a simple chemical contaminant but as a theoretical mixture of 209 congeners

with distinctive physico-chemical properties that can be environmentally fi-actionated. The

loss of the lowest and the highest molecular weight PCB congeners from a more complex

Aroclor mixture by evaporation/dissolution and adsorption/deposition, respectively, after

input can result in a profile distribution similar to that of Aroclor 1254 or a mixture of

Aroclors 1254 with 1248 or 1260. Therefore, it seems that, independently of the

composition of the. original PCB mixture, the environmental fractionation together with

preferential uptake will indicate Aroclor 1254 as the most probable contaminant source.

Incidentally, Aroclor 1254, which is one of the most commonly reported PCB mixtures as

185

the source in environmental pollution studies, is not the one that was produced in the

largest quantities. The most popular blend in the U.S. was Aroclor 1242, which

comprised over 50% of the total domestic production between 1957 and 1970 (Cairns et

aL, 198 6).

Similarly to what was observed for PCB congeners, oysters exposed in the laboratory

to a wide molecular range of PAHs showed the preferential uptake of four- and five-ring

PAHs. This observation was confirmed by the results obtained from the field
experiments in Galveston Bay. If oysters preferentially bioaccumulate combustion-

derived PAHs, i.e. four- and five-ring compounds, compared to petroleum-derived

PAHs, i.e. two- and three-ring compounds, then how accurate do they represent the

contamination at a site where most of the PAHs are petroleum-derived? This particular

area requires further attention.

Other areas requiring more investigation are the effect that simultaneous exposure to

PCBs and PAHs have on the concentrations encountered in environmentally contaminated

oysters and how well these concentrations correlate with local environmental levels.

During this study, it has been shown that oysters exposed in the laboratory to a mixture of

PCBs and PAHs, depurated PAHs at a faster rate when the contaminants input was

stopped than oysters that were not simultaneously exposed to PCBs. The half-lives for

individual PAHs encountered in oysters exposed in the laboratory to a mixture of PCBs

and PAHs compared more closely to those found during the field experiment in Galveston

Bay.

It can be concluded that indigenous oysters can be valuable bioindicators of

environmental contamination by trace organic compounds, particularly PAHs, PCBs and

TBT, gnly if their limitations are fully understood. Within these limitations, transplanted

oysters can be succesfully used to monitor environmental contamination by PAHs and

TBTs in areas lacking indigenous bivalves if deployed in-situ for a period of time of at

186 1
1
least 30 days; for PCBs, however, much longer time period, Le over 6 months, may be
required. I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I

187

REFERENCES

Abel, R. King, N. J., Vosser, J. L., & Wilkinson, T. G. (1986). The control of
organotin use in antifouling paint. The UK's basis for action. Proceedings of the
Oceans86 Conference, Washington DC 23-25 Sept. 1986, pp. 1314-23.

Afghan, B. K., Wilkinson, R. J., Chow, A., Findley, T. W., Gesser, H. D. &
Srikameswaran, K. 1. (1984). A comparative study of the concentration of
polynuclear aromatic hydrocarbons by open cell polyurethane foams. Water Res.,
18,9.

Alzieu, C., Thiabaud, Y., Heral, M. & Boutier, B. (1980). Evalution des risques dus
1'emploi des peintures antisalissures dans les zones conchylicoles. Rev. Trav. Inst.
Pech. Marit., 44, 30 1.

Alzieu C., Heral, M., Thibaud, Y., Dardignac, M. & Fenillet, M. (1982). Influence des
peintures antisalissures a base d'organostanniques sur la calcification de la coquille de
1'huitre Crassostrea gigas. Rev. Trav. Inst. Pech. Marit., 45, 101.

Anderson, C. D. & Dalley, R. (1986). Use of organotin in antifouling paints.
Proceedings of the Oceans'86 Conference, Washington DC 23-25 Sept. 1986, pp.
1108-13.

Anderson, R.D. (1975). Petroleum hydrocarbons and oyster resources of Galveston
Bay, Texas. In: Conference on Prevention and Control of Oil Pollution, American
Petroleum Institute, Washington DC, pp. 541-8

Anderson, R. K., Scalan, R. S., Parker, P. L. & Berhens, E. D. (1983). Seep oil and
gas in the Gulf of Mexico slope sediments. Science, 222, 619.

Anderson, R.S. (1978a). Benzo(a)pyrene metabolism in the American oyster,
Crassostrea virginica. Ecological Research Series EPA-600/3-78-009. U.S.
Environmental Protection Agency, Washington DC.

Anderson, R.S. (1978b). Developing an invertebrate model for chemical carcinogenesis:
metabolic activation of carcinogens. Comp. Pathobiol., 4, 11.

Anderson, R.S. (1979). Immune competence in marine invertebrates. Symposi .um on
Pollutant Responses in Marine Animals. Abstracts. October 21-23, 1979, Texas
A&M University, College Station, TX

Anderson, R.S. (1985). Metabolism of a model environmental carcinogen by bivalve
molluscs. Mar. Environ. Res., 17, 137.

188

Anonymous (1980). Organotin in antifouling paints. Environmental considerations.
Department of the Environment. Pollution paper NO 25, London.

Armstrong, H.W., Fucik, K., Anderson, J.W. & Neff, J.M. (1979). Effects of oilfield
brine effluent on sediments and benthic organisms in Trinity Bay, Texas. Mar.
Environ. Res., 2, 55.

Atlas, E. & Giam, C. S. (1981). Global transport of organic pollutants: Ambient
concentrations in the remote marine atmosphere. Science, 211, 163.

Axiak, V., George, JJ. & Moore, M.N. (1988). Petroleum hydrocarbons in the marine
bivalve Venus verrucosa: accumulation and cellular responses. Mar. Biol., 97, 225.

Ballschmiter K. & Zell, M. (1980). Analysis of polychlorinated biphenyls (PCB) by
glass capillary gas chromatography. Composition of technical Aroclor and Clophen-
PCB mixtures. Z Analyt. Chem., 302, 20.

Barron, M.G. (1990). Bioconcentration. Will water-bome organic chemicals accumulate
in aquatic animals? Environ. ScL Technol., 24, 1612.

Basu, D. K. & Saxena, J. (1978). Monitoring of polynuclear aromatic hydrocarbons in
water. 11- Extraction and recovery of six representative compounds with polyurethane
foams. Environ. ScL Technol., 12, 791.

Betchtel, T.J. & Coperland, B.J. (1970). Fish species diversity indices as indicators of
pollution in Galveston Bay, Texas. Contrib. Mar. Sci., 15, 103.

Bjorseth, A., Knutzen, J. & Skei, J. (1979). Determination of polycyclic aromatic
hydrocarbons in sediments and mussels from Sauda@jord, W, Norway, by glass
capillary gas chromatography. Sci. Total Environ., 13, 7 1.

Blumer, M. (1976). Polycyclic aromatic compounds in nature. Sci. Am., 234, 34.

Boehm, P. D. & Quinn, J.G. (1973). The persistence of chronically accumulated
hydrocarbons in the hard shell clam Mercenaria mercenaria. Mar. Biol., ", 227.

Boon, J. P., Van Zantvoort, M. B., Govaert, M. J. M. A. & Duinker, J. C. (1985).
Organochlorines in benthic polychaetes (Nephtys spp.) and sediments from the
southern North Sea. Identification of individual PCB components. Neth. J. Sea
Res., 19, 93.

Bopp, R.P., Simpson, H.J., Olsen, C.R. & Kostyk, N. (1981). Polychlorinated
biphenyls in sediments of the tidal Hudson River, New York. Environ. Sci.
Technol., 15, 210.

Brooks, J.M., Wade, T.L., Atlas, E.L., Kennicutt II, M.C., Presley, D.J. Fay, R.R.,
Powell, E.N. & Wolff, G. (1987). Analysis of Bivalves and Sediments for Organic
Chemicals and Trace Elements. First annual report for the NOOA's National Status
and Trends Program. Contract 50-DGNC-5-00262, 618 pp.

189

Brooks, J.M., Wade, T.L., Atlas, E.L., Kennicutt 11, M.C., Presley, B.J. Fay, R.R.,
Powell, E.N. & Wolff, G. (1988). Analysis of Bivalves and Sediments for Organic
Chemicals and Trace Elements. Second annual report for the NOOA's National Status
and Trends Program. Contract 50-DGNC-5-00262, 644 pp.

Brooks, J.M., Wade, T.L., Atlas, E.L., Kennicutt II, M.C., Presley, B.J. Fay, R.R.,
PoweU, E.N. & Wolff, G. (1989). Analysis of Bivalves and Sediments for Organic
Chemicals and Trace Elements. Third annual report for the NOOA's National Status
and Trends Program. Contract 50-DGNC-5-00262, 692 pp.

Brownawell, B. J. (1986). The Role of Colloidal Organic Matter in the Marine
Geochemistry of PCBs (Diss: Ph. D.). Tech. Report, WHOI-86-19. Woods Hole
Oceanogr. Inst., MA, 318 pp.

Brownawell, B. J. & Farrington, J. W. (1985). Partition of PCB's in marine sediments.
In: Marine and Estuarine Geochemistry. Sigleo, A. C. & Hattori, A. (Eds.). Lewis
Publisher, Chelsea, MI, pp. 97-120.

Brownawell, B. J. & Farrington, J. W. (1986). Biogeochernistry of PCBs in interstitial
waters of a coastal marine sediment. Geochem. Cosmochim. Acta, 50, 157.

Bruggeman, W.A., Martron, L.B.J.M., Kooiman, D. & Hutzinger, 0. (1981).
Accumulation and elimination kinetics of di-, tri- and tetrachlorobiphenyls by goldfish
after dietary and aqueous exposure. Chemosphere, 10, 811.

Bryan, A.M., Stone, W.B. & 01afsson, P.G. (1987). Disposition of toxic PCB
congeners in snapping turtle eggs expressed as toxic equivalents of TCDD. Bull.
Environ. Contam. Toxicol., 39, 79 1.

Bums, K. A. & Smith, J. L. (1981). Biological monitoring of ambient water quality: the
case for using bivalves as sentinel organisms for monitoring petroleum pollution in
coastal waters. Estuar. Coastal Shelf Sci., 13, 433.

Butler, P.A. (1973). Organochlorine residues in estuarine mollusks, 1965-72. National
Pesticide Monitoring Program. Pestic. Monit. J., 6, 238.

Cadee, G. C. (1982). Tidal and seasonal variation in particulate and dissolved organic
carbon in the western Dutch Wadden Sea and Marsdiep tidal inlet. Neth. J. Sea Res.,
15, 228.

Caims, T., Doose, G.M., Froberg, J.E., Jacobson, R.A. & Siegmund, E. G. (1986).
Analytical chemistry of PCBs. In: PCBs and the Environment, Vol. 1, Waid, J.S.
(Ed.), CRC Press, Boca Raton, FL, pp. 1-45.

Calambokidis, J., Mowrer, J., Beug, M.W. & Herman, S.G. (1979). Selective retention
of polychlorinated biphenyl components in the mussel, Mytilus edulis. Arch.
Environ. Contam. Toxicol., 8, 299.

Chamber, G.V. & Sparks, A.K. (1959). An ecological survey of Houston ship channel
and adjacent bays. Contrib. Mar. Sci., 10,213.

190

Champ, M. A. & Pugh, W. L. (1987). Tributyltin antifouling paints : Introduction and
overview. Proceedings of the Oceans86 Conference, Washington DC 23-25 Sept.
1986, pp 1296-1308.

Chin, Y.P. & Gschwend, P.M. (1992). Partitioning of polycyclic aromatic hydrocarbons
to marine porewater organic colloids. Environ. Sci. TechnoL, 26, 1621.

Chiou, C. T., Porter, P. E. & Schmedding, D. W. (1983). Partition equilibria of
nonionic organic compounds between soil organic matter and water. Environ. Sci.
Technol., 17, 227.

Choi, W. W. & Chen, K. Y. (1976). Association of chlorinated hydrocarbons with fine
particles and humic substances in nearshore surficial sediments. Environ. Sci.
Technol., 10, 782.

Clement, L. E., Stekoll, M. S. & Shaw, D. G. (1980). Accumulation, fractionation and
release of oil by the intertidal clam Macoma balthica. Mar. BioL, 57,41.

Coates, J. T. & Elzerman, A. W. (1986). Desorption kinetics for selected PCB
congeners from river sediments. J. Contam. HydroL, 1, 191.

Colijn, F. (1982). Light adsorption in the waters of the Ems-Dollard estuary and its
consequences for the growth of phytoplankton and microphytobentos. Neth. J. Sea
Res., 15, 196.

Collier, T. K., Gruger Jr., E. H. & Varanasi, U. (1985). Effect of Aroclor 1254 on the
biological fate of 2-6 dimethylnaphthalene in Coho Salmon (Oncorhynchus kisutch).
BuIL Environ. Contam. ToxicoL, 34, 114.

Connell , D.W. (1990). Bioconcentration of lipophilic and hydrophobic compounds by
aquatic organisms. In: Bioaccumulation of Xenobiotic Compounds. Connell, D.W.
(Ed.), CRC Press, B oca Raton, FL, pp. 97-144.

Connell, D.W. & Hawker, D.W. (1988). Use of polynomial expressions to describe the
bioconcentration of hydrophobic chemicals by fish. EcotoxicoL Environ. Saf., 16,
242.

Connell, D.W. & Schihirmann, G. (1988). Evaluation of the various molecular
parameters as predictors of bioconcentration in fish. EcotoxicoL Environ. Saf, 15,
324.

Courtney, W. A. M. & Denton, G. R. W. (1976). Persistence of polychlorinated
biphenyls in the hard-clam (Mercenaria mercenaria) and the effect upon the
distribution of these pollutants in the estuarine environment. Environ. PolL, 10, 55.

Davis, A. G. & Smith, P. J. (1980). Recent advances in organotin chemistry. Adv.
Inorg. Chem. Radiochem., 23, 1.

Dobbs, A.J. & Williams, N. (1983). Fat solubility - A property of environmental
relevance? Chemosphere, 12, 97.

191

Duinker, J.C. & Hillebrand, M.T.J. (1983). Characterization of PCB components in
Clophen formulations by capillary GC-MS and GC-ECD techniques. Environ. Sci
Technol., 17, 449.

Duinker, J.C. & Boon, J.P. (1985). PCB congeners in the marine environmnet - A
review. In: Proceedings of the Fourth European Symposium on Organic
Micropollutants in the Aquatic Environment. Vienna, Austria, October 22-24, 1985,
pp. 187-205.

Duinker, J. C., Hillebrand, M. T. J. & Boon, J. P. (1983). Organochlorines in benthic
invertebrates and sediments from the Dutch Wadden Sea. Identification of individual
PCB components. Nether. J. Sea Res., 17, 19.

Duinker, J.C., Schulz, D.E. & Petrick, G. (1988a). Multidimentional gas
chromatography with electron capture detection for the determination of toxic
congeners in polychlorinated biphenyl mixtures. Anal. Chem., 60, 478.

Duinker, J.C., Schultz, D.E. & Petrick, G. (1988b). Selection of chlorinated biphenyl
congeners for analysis in environmental samples. Mar. Poll. Bull., 19, 19.

Duinker, J.C., Hillebrand, M.T.J., Palmork, K.H. & Wilhelmsen, S. (1980). An
evaluation of existing methods for quantitation of polychlorinated biphenyls in
environmental samples and sugestions for an improved method based on measurement
of individual components. Bull. Environ. Contaim Toxicol., 25, 956.

Dunn, B.P. & Stich, H.F. (1976). Monitoring procedures for chemical carcinogens in
coastal waters. Can. J. Fish. Aquat. Sci., 33, 2040.

Eder, G., Ernst, W., Goerke, H., Duinker, J.C. & Hillebrand, T.J. (1981).
Organochlorine residues analysed in invertebrates of the Dutch Wadden Sea by two
methods. Neth. J. Sea Res.,15, 78.

Egaas, E. & Varanasi, U. (1982). Effects of polychlorinated biphenyls and
environmental temperature on in vitro formation of benzo(a)pyrene metabolites by
liver of trout (Sabno gairdneri). Biochem. Pharmacol., 31, 561.

Ehrhardt, M. (1972). Petroleum hydrocarbons in oysters from Galveston Bay. Environ.
Poll., 3, 257.
Elder, D.L., Fowler, S.W. & Polikarpov, G.G. (1979). Remobilization of sediment-
associated PCBs by the worm Nereis diversicolor. Bull. Environ. Contam. Toxicol.,
21,448.

Ellegehausen, H., Guth, J.A. & Esser, H.O. (1980). Factors determining the
bioaccumulation potential of pesticides in the individual compartments of aquatic food
chains. Ecotoxicol. Environ. Saf., 4, 134.

Erickson, M. D. (1986). Analytical Chemistry of PCBs. Butterworth Publishers,
Stoneham, MA, 509 pp.

192

Evans, C.J. & Karpel, S. (1985). Organotin Compounds in Modern Technology.
Journal of Organometallic Chemistry Library 16. Elsevier Science Publisher, New
York, 279 pp.

Farrington, J. W., Goldberg, E. D., Risebrough, R. W., Martin, J. H. & Bowen, V. T.
(1983). U.S. "Mussel Watch" 1976-1978: An overview of the trace-metal, DDE,
PCB, hydrocarbon, and artificial radionuclide data. Environ. Sci. Technol., 17, 490.

Farrington, J.W., Albaiges, J., Burns. K.A., Dunn, B.P., Eaton, P., Laseter, J.L.,
Parker, P.L. & Wise, S. (1980). Fossil fuels. In: The International Mussel 141a.-ch,
Report of a workshop sponsored by the Environmental Studies Board Commision on
Natural Resources, National Research Council, pp. 7-77.

Farrington, J. W., Risebrough, R. W., Parker, P. L., Davis, A. C., DeLappe, B. W.,
Winters, J. K., Boatwright, D. & Freq, N. M. (1982). Hydrocarbons,
Polychlorinated Biphenyls, and DDE in Mussels and Oysters from the U. S. Coast,
1976-1978. Tech. Report, WHOI-82-42. Woods Hole Oceanographic Institution,
MA, 106 pp.

Fazio, T. (1971). Analysis of oyster samples for polycyclic hydrocarbons. In:
Proceedings of the 7thlVaiional Shelffish Sanitation Workshop. FDA, Washington
DC, pp. 238-43.

Folk, R.L. (1974). Petrology of Sedimentary Rocks. Hemphill's, Austin, TX. 160 pp.

Forlin, L. (1980). Effects of Clophen A50, 3-methylcholanthrene, pregnenolone- 16
alpha-carbonitrile, and phenobarbital on the hepatic microsomal cytochrome P-450-
dependent monooxygenase system in rainbow trout, Sah-no gairdncri, of different age
and sex. Toxicol. Appl. Pharmacol., 54,420.

Fortner, A. R. & Sick@ L. V. kl985;. Simultaneoui accumulation of naphthalene. a PCB
mixture, and benzo(a)pyrene by the oyster Crassostrea virginica. Bull. Enviror.
Contam. Toxicol., 34, 256.

Fossato, U.V. & Canzonier, W.J. (1976). Hydrocarbon uptake -and loss by the mussel
Mytilus edulis. Mar. Biol., 36, 243.

Fox, R. G. (1988). Spatial and Temporal Variations of Polynuclear Aromatic
Hydrocarbons, Pesticides, and Po4ychlorinated Biphenyls in Crassostrea virginica and
Sediments from Galveston Bay, Texas. M. S. Thesis. Texas A&M University,
College Station, TX, 92 pp.

Frank, A. P., Landrum, P. F. & Eadie, B. J. (1986). Polycyclic aromatic hydrocarbon
rates of uptake, depuration, and biotransformation by lake Michigan Stylodrilus
heringianus. Chemosphere, 15, 317.

Fucik, K.,M. & Neff, J. Tv!. (.1977). Effects of temperature and salinity on naphthalene
uptake in the temperate clam Rangia cuneata and the b oreal clam Protothaca staminca.
In: Fate and Effects of Petroleum Hydrocarbons in Marine Organisms and
Ecosystems. Wolfe, D. A. (Ed.), Pergarnon Press, NY, pp. 305-12.

193

Fucik, K.W., Armstrong H.W. & Neff, J.M. (1977). Uptake of naphthalenes of the
clam, Rangia cuneata, in the vicinity of an oil separator platform in Trinity Bay,
Texas. In: Proceedings of the 1977 Oil Spill Conference (Prevention, Behavior,
Control, Cleanup.). Ameiican Petroleum Institute, Washington DC, pp. 637.

Gesser, H. D., Cbow, A., Davis, F. C., Uthe, J. F. & Reinke, J. (1971). The extraction
and recovery of polychlorinated biphenyls (PCB) using porus polyurethane foam.
Anal. Letters, 4, 883.

Geyer, H., Sheehan, P., Kotzias, D., Freitag, D. & Korte, F. (191-112). Prediction of
ecotoxi.col.ogical behavior of chemicals: Relationship between physico-chernical
properties and bioaccumulation of organic chemicals in the mussel Mytilus edulis.
Chemosphere, 11, 1121.

Giesy, J. P., Bartell, S. M., Landrum, P. F., Leversee., G. J. & Bo-xling, J. W. (1983).
Fates and Biological Effects of Polycyclic Aromatic Hydrocarbons in Aqua.ric
Systems. U.S. Environmental Protection Agency, EPA-600/S3-83-053, 5 pp.

Goldberg, E. D., Bowen., V. T., Farrington, J. W., Harvey, G., Martin, J. H., Parker,
P. L., Risebrough, R. W., Robertson, W., Schneider, E. & Gamble, E. (1978). The
mussel watch. Environ. Conserv., 5, 101.

Goldstein, J. A. & Safe, S. (1989). Mechanism of action and structure-activity
relationships for the chlorinated di benzo-p -dioxins and related compounds. in:
Halogenated Biphenyls, Terphenyls, Naphthalenes, Dibenzodioxins and Related
Products. Kimbrough & Jensen (Eds.). Elsevier Science Publishers, New York, up.
239-93.

Hall, L. W. & Pinkney, A. E. (1985). Acute and sublethal effects of organotin
compounds on aquatic biota: An interpretative literature evaluadon. CRC Critical
Reviews in Toxicol., 14, 159.

Hamelink, J.L., Waybrant, R.C. & Ball, R.C. (1971). A proposal: exchange equilibria
control the degree chlorinated hydrocarbons are biologically magnified in lentic
environments. Trans. Am. Fish. Soc., 100, 207.

Hansch, C. & Fugita, T. (1964). p-c;-n analysis. A method for the correlation of
biological activity and chernical structure. J. Am. Chern. Soc., 86, 1616.

Hansen, L.G. (1987). Environmental toxicology of polychlorinated biphenyls. In:
Environmental Toxin Series 1, Polychlorinated Biphenyls (PCBs): Mammalian and
Environmental Toxicology. Safe, S. & Hutzirger, 0. (Eds.), Springer-Verlag,
Berlin, Germany, 152 pp.

Hargis Jr., W. J., Roberts Jr., M. H. & Zwerner, D. E. (1984). Effects of contaminated
sediments and sediment-ex posed effluent water on ar estuarine fish: Acute toxicity.
Mar. Environ. R es., 14, 337.

Harris, J.R.W. & Cleary, J.J. (11087). Particle-water partitioning and organotin dispersal
in an estuary. In: Proceedings of the 0ceans87 Conference, Halifax, Nova Sc'.)tia,
Canada, 2.8 Sept.-I Oct. 1987, pp. 1370-4.

194

Hawker, D.W. & Connell, D.W. (1985). Relationships between partition coefficient,
uptake rate constant, clearance rate constant, and time to equilibrium for
bioaccumulation. Chemosphere, 14, 1205.

Hawker, D.W. & Connell, D.W. (1988). Octanol-water partition coefficients of
polychlorinated biphenyl congeners. Environ. Sci. Technol., 22, 382.

Hawkes, J.W. (1979). Morphological effects of petroleum and chlorobiphenyls on fish
tissue [Abstract only]. In: Animals as Monitors of Environmental Pollutants.
Symposium of Pathobiology of Environmental Pollutants: Animals Models and
Wildlife as Monitors. National Academy od Sciences, Washington DC, pp. 381-2.

Herbes, S. E. (1977). Partitioning of polycyclic aromatic hydrocarbons between
dissolved particulate phases in natural waters. Water Res., 11, 493.

Hill, D.W., Hejtmancik, E. & Camp, B.J. (1976). Induction of hepatic microsomal
enzymes by Aroclor 1254 in Ictaluruspunctatus (channel catfish). Bull. Environ.
Contam. Toxicol., 16, 495.

Hoffman, E. J., Mills, G. L., Latimer, J. S. & Quinn, J. G. (1984). Urban runoff as a
source of polycyclic aromatic hydrocarbons to coastal waters. Environ. Sci.
Technol., 18, 580.

Hohn, M.H. (1959). The use of diatom population as a measure of water quality in
selected areas of Galveston and Chocolate Bays, Texas. Contrib. Mar. Sci., 6,206.

Hong, C-S & Bush, B. (1990). Determination of mono- and non-ortho coplanar PCBs in
fish. Chemosphere, 21, 173.

Huckins, J.N., Stalling, D.L. & Petty, J.D. (1980). Carbon-foam chromatographic
separation of non-o-o'-chlorine substituted PCBs from Aroclor mixtures. J Assoc.
Off. Anal. Chem., 63, 750.

Isnard, P. & Lambert, S. (1989). Aqueous solubility and n-octano/water partition
coefficient correlations. Chemosphere, 18, 1837.

Jackim, E. & Lake, C. (1978). Polynuclear aromatic hydrocarbons in estuarine and
nearshore environments. In: Estuarine Interactions. Wiley, M.L. (Ed.), Academic
Press, New York, 415 pp.

James, M.O. & Little, P.J. (1981). Polyhalogenated biphenyls and phenobarbital:
evaluation as inducers of drug metabolizing enzymes in the sbeepshead, Archosargus
probatocephalus. Chem-Biol. Interact., 36, 229.

Jan, J. & Josipovic, D. (1978). Polychlorinated terphenyls in hens - The behaviour of
ortho, meta and para isomers. Chemosphere, 11, 863.

Jensen, S. (1966). Report of a new chemical hazard. New Scientist, 32, 612.

Jensen, S. & Sundstr6m, G. (1974). Structures and levels of most chlorobiphenyls in
two technical PCB products and in human adipose tissues. Ambio, 3, 70.

195

Johnson, W.E., Hall Jr., L.W., Bushong, S.J. & Hall, W.S. (1987) Organotin
concentrations in centrifuged versus uncentrifuged water column samples and in
sediment pore waters of Northern Chesapeake Bay tributary. In: Proceedings of the
Oceans'87 Conference, Halifax, Nova Scotia, Canada, 28 Sept.-I Oct. 1987, pp.
1364-9.

Joiris, C.R. & Overloop, W. (1991). PCBs and organochlorine pesticides in
phytoplankton and zooplankton in the Indian sector of the Southern Ocean. Antarctic
Science, 3, 371.

Jovanovich, M.C. & Marion, K.R. (1987). Seasonal variation in uptake and depuration
of anthracene by the brackish water clam Rangia cuneata. Mar. Biol., 95, 395.

Kamops, L.R., Trotter, W.J., Young, S.J., Smith, A.C., Roach, J.A.G., & Page, S.W.
(1979). Separation and quantitation of 3,3,4,4'-tetrachlorobiphenyl and
3,3',4,4',5,5'-hexachlorobiphenyls in Aroclors using Florisil column
chromatography and gas-liquid chromatography. Bull. Environ. Contain. Toxicol.,
23, 51.

Kannan, N., Tanabe, S., Wakimoto, T. & Tatsukawa, R. (1987). Coplanar PCBs in
Aroclor and Kanechlor mixtures. J. Assoc. Off. Anal. Chein., 70, 451,

Kannan, N., Tanabe, S., Tatsukawa, R. & Phillips, D.J.H. (1989). Persistency of
highly toxic coplanar PCBs in aquatic ecosystems: uptake and release kinetics of
coplanar PCBs in green-lipped mussels (Perna viridis Linnaeus). Environ. Pollut.,
56, 65.

Karickhoff, S. W., Brown, D. S. & Scott, T. A. (1979). Sorption of Hydrophobic
pollutants on natural sediments. Water Res., 13, 241-.

Kennish, M.J. (1992). Polynuclear aromatic hydrocarbons. In: Ecology of Estuaries:
Anthropogenic Effects. CRC Press Inc., Boca Raton, FL, pp. 133-81.

Kerkhoff, M. A. T., de Vries, A., Wegman, R. C. C. & Hofstee, S. W. M. (1982).
Analysis of PCB's in sediments by glass capillary gas chromatography.
Chemosphere, 11, 165.

King, K.A. (1989a). Foods habits and organochlorine contaminants in the diet of black
skimmers, Galveston Bay, Texas, USA. Colonial Waterbirds, 12,109.

King, K.A. (1989b). Food habits and organochlorine contaminants in the diet of
olivaceous cormorants in Galveston Bay, Texas. Southwestern Nat., 34, 338.

King, K.A. & Krynitsky, A.J. (1986). Population trends, reproductive success and
organochlorine chemical contaminants in waterbirds nesting in Galveston Bay, Texas.
Arch. Environ. Contain. Toxicol., 15, 367.

King, K.A., Stafford, C.J., Cain, B.W., Mueller, A.J. & Hall H.D. (1987). Industrial,
agricultural and petroleum contaminants in cormorants wintering near the Houston
Ship Channel, Texas, USA. Colonial Waterbirds, 10, 93.

196

Knipe, T. (1989). Toxics and the sea: The lesson of the dancing cats. Calypso Log,
August 1989, pp. 14-16.

Kuehl, D.W., Butterworth, B.C., Libal, J. & Marquis, P. (1991). An isotope dilution
high resolution gas chromatographic - high resolution mass spectrometric method for
the determination of coplanar polychlorinated biphenyls: application to fish and marine
mammals. Chemosphere, 22, 849.

Kvenvolden, K. A. & Harbough, J. W. (1983). Reassessment of the rates at which oil
from natural sources enters the marine environment. Mar. Environ. Res., 10, 223.

Lang, V. (1992). Polychlorinated biphenyls in the environment. J. Chromatogr. 595, 1.

Langdon, C.J. & Siegfried, C.A. (1984). Progress in the development of artificial diets
for bivalve filter-feeders. Aquaculture, 39, 135.

Langston, W. J. (1978). Persistence of polychlorinated biphenyls in marine bivalves.
Mar. Biol., 46, 35.

Langston W. J., Burt, G. R. & Mingjiang, Z. (1987). Tin and organotin in water,
sediments, and benthic organisms of Poole Harbor. Mar. Poll. Bull., 18, 634.

Laughlin Jr., R. B. (1986). Bioaccumulation of tributyltin: the link between environment
and organisms. Proceedings of the Oceans86 Conference, Washington DC 23-25
Sept. 1986, pp. 1206-9.

Laughlin Jr., R.B. (1990). Mechanisms of TBT bioaccumulation by marine bivalve
molluscs. In: Proceedings of the 3rd International Organotin Symposium, Monaco,
17-20 April 1990, pp. 77-8.

Laughlin Jr., R.B. & Linden, 0. (1987). Tributyltin - Contemporary environmental
issues. Ambio 16, 252.

Laughlin Jr., R.B., French, W. & Guard, H. E. (1986). Accumulation of
bis(tributyltin)oxide by the marine mussel Mytilus edulis. Environ. Sci. Technol.,
20, 884.

Lee, R. F. (1977). Accumulation and turnover of petroleum hydrocarbons in marine
organisms. In: Fate and Effects of Petroleum Hydrocarbons in Marine Ecosystems
and Organisms. D. A. Wolfe (Ed.). Pergamon Press, New York, pp. 60-70.

Lee, R.F. (1985). Metabolism of tributyltin oxide by crabs, oysters and fish. Mar.
Environ. Res., 17, 145.

Lee, R.F. (1986). Metabolism of bis(tributyltin)oxide by estuarine animals. In:
Proceedings of the Oceans86 Conference, Washington DC 23-25 Sept. 1986, pp.
1182-8.

Lee, R.F., Sauerheber, R. & Dobbs, G.H. (1972). Petroleum hydrocarbons: uptake and
discharge by the marine mussel Mytilus edulis. Science, 177, 344.

197

Lee, R. F., Gardner, W. S., Anderson, J. W., Blaylock, J. W. & Barnell-Clarke, J.
(1978). Fate of polycyclic aromatic hydrocarbons in controlled ecosystem enclosures.
Environ. Sci. Technol., 12, 832.

Livingston, D. R. (1985). Responses of the detoxication/toxication enzyme systems of
molluscs to organic pollutants and xenobiotics. Mar. Poll. Bull., 16, 158.

Lowe, J.I., Wilson, P.D., Rick, A.J. & Wilson Jr., A.J. (1971). Chronic exposure of
oysters to DDT, toxaphene and parathion. Proc. Nat. Shellf. Assoc., 61, 7 1.

Lowe, J.I., Parrish, P.R., Patrick, J.M. & Forrester, J. (1972). Effects of the
polychlorinated biphenyl Aroclor 1254 on the American oyster, Crassostrea virginica.
Mar. Biol., 17, 209.

Ludgate, J. W. (1987). The economic and technical impact of TBT legislation on the
USA marine industry. Proceedings of the Oceans87 Conference, Halifax, Nova
Scotia, Canada, 28 Sept.-1 Oct. 1987, pp. 1309-13.

MacIntyre, W. G. & Smith, C. L. (1984). Comment on partition equilibria of nonionic
organic compounds between soil organic matter and water. Environ. Sci. Technol.,
18, 295.

Mackay, D. (1982). Correlation of bioconcentration factors. Environ. Sci. Technol.,
16, 274.

Mackay, D., Mascarenhas, R. & Shiu, W. Y. (1980). Aqueous solubility of
polychlorinated biphenyls. Chenzosphere, 9, 257.

MacLeod, W. D., Brown, D. W., Friedman, A. J., Burrows, D. G., Maynes, 0.,
Pearce, R. W., Wigren, C. A. & Bogar, R. G. (1985). Standard Analytical
Procedures of the NOAA National Analytical Facility, 1985-1986. Extractable Toxic
Organic Components. Second edition, U.S. Department of Commerce,
NOAA/NMFS. NOAA Tech. Memo. NMFS F/NWC-92.

Maguire, R. J. (1984). Butyltin compounds and organic tin in sediments in Ontario.
Environ, Sci. Technol., 18, 291.

Maguire, R.J. (1986). Review of the occurrence, persistence and degradation of
tributyltin in fresh water ecosystems in Canada. In: Proceedings of the Oceans'86
Conference, Washington DC 23-25 Sept. 1986, pp. 1252-5.

Malins, D. C., Mc Cain, B. B., Brown, D. W., Myers, M. S., Krahn, M. M. & Chan,
S-L. (1987). Toxic chemicals, including aromatic and chlorinated hydrocarbons and
their derivatives, and liver lesions in White Croaker (Genyonemus lineatus) from the
vicinity of Los Angeles. Environ. Sci Technol., 21, 765.

Malins, D. C., Mc Cain, B. B., Brown, D. W., Chan, S-L., Myers, M. S., Landahl, J.,
T., Prohaska, P. G., Friedman, A. J., Rhodes, L. D., Burrows, D. G., Gronlund,
W. D. & Hodgins, H. 0. (1984). Chemical pollutants in sediments and diseases of
bottom-dwelling fish in Puget Sound, Washington. Environ. Sci. TechnoL, 18, 705.

198

Martin, M. (1985). State Mussel Watch: Toxic survillance in California. Mar. Poll.
Bull., 16, 140.

Matsuo, M. (1980). A thermodynamic interpretation of bioaccumulation of Aroclor 1254
(PCB) in fish. Chemosphere, 9, 67 1.

Matthias, C. L., Bellania, J. M., Olson, G. J. & Brinckman, F. E. (1986).
Comprehensive method for determination of aquatic butyltin and butylmethyltin
species at ultratrace levels using simultaneous hydridization/extraction with gas
chromatography-flame photometric detection. Environ. Sci. Technol., 20, 609.

Mazurek, M.A. & Simoneit, B.R.T. (1984). Characterization of biogenic and petroleum-
derived organic matter in aerosols over remote, rural and urban areas. In:
Identification and Analysis of Organic Pollutants in Air. Keith, L.H. (Ed.), Ann
Arbor Science, Butterworth Publishers, Boston, MA, pp. 353-70.

McElroy, A.E., Farrington, J.W. & Teal, J.M. (1989). Bioavalaibility of polycyclic
aromatic hydrocarbons in the aquatic environment. In: Metabolism of Polycyclic
Aromatic Hydrocarbons in the Aquatic Environment. Varanasi, U. (Ed.), CRC Press
Inc., Boca Raton, FL, pp. 1-39.

McFarland, V.A. & Clarke, J.U. (1989). Environmental occurrence, abundance, and
potential toxicity of polychlorinated biphenyl congeners: Considerations for a
congener-specific analysis. Environm. Health Perspectives, 81, 225.

McKinney, J.D., Gottschalk, K.E. & Pedersen, L. (1983). The polarizability of planar
aromatic systems. An aplication to polychlorinated biphenyls (PCBs). J. Mol.
Struct., 105, 427.

McKinney, J.D., Chae, K., McConnel, E.E. & Birnbaum, L.S. (1985). Structure-
induction versus structure- toxicity relationships for polychlorinated biphenyls and
related aromatic hydrocarbons. Environ. Health Perspect., 60, 57.

McKinney, J., Chae, K., Gupta, B.N., Moore, J.A. & Goldstein, J.A. (1976).
Toxicological assessment of hexachlorobiphenyl isomers and 2,3,7,8-
tetrachlorodibenzofuran in chicks. I. Relationship of chemical parameters. Toxicol.
Appl. Pharmacol., 36, 65.

Means, J. C., Wood, S. G., Hassett, J. J. & Banwart, W. L. (1980). Sorption of
polynuclear aromatic hydrocarbons by sediments and soils. Environ. Sci. Technol.,
14, 1524.

Meier, P. G. & Rediske, R. R. (1984). Oil and PCB interactions on the uptake and
excretion in Midges. Bull. Environ. Contam. Toxicol., 33, 225.

Mix, M.C. (1984). Polycyclic aromatic hydrocarbons in the aquatic environment:
occurrence and biological monitoring. In: Reviews in Environmental Toxicology, Vol
50, Hodgson, E. (Ed.). Elsevier, Amsterdam, 51 pp.

199

Moore, M. N., Livingston, D. R., Donkin, P. Bayne, B. L., Widdous, J. & Lowe, D.
M. (1980). Mixed function oxygenase and xenobiotic, detoxication/toxication
systems in bivalve molluscs. Helgol. Wiss. Meeresunters, 33, 278.

Moriarty, F. (1975). Exposure and residues. In: Organochlorihe Insecticides. Persistent
organic pollutants, Moriarty, F. (Ed.) Academic Press, London.

Muller, M.D., Renberg, L. & Rippen, G. (1989). Tributyltin in the environment -
sources, fate and determination. An assessment of present status and research needs.
Chemosphere, 18, 2015.

Mullin, M.D., Pochini, C.M., McCrindle, S., Romkes, M., Safe, S. H. & Safe, L.M.
(1984). High-resolution PCB analysis: synthesis and chromatographic properties of
all 209 PCB congeners. Environ. Sci. Technol., 18, 468.

Murray, H., Lee, E. & Giam, C.S. (1981a). Analysis of marine sediments, water and
biota for selected organic pollutants. Chemosphere, 10, 1327.

Murray, H., Lee, E. & Giam, C.S. (1981b). Phthalic acid esters, total DDTs and
polychlorinated biphenyls in marine samples from Galveston Bay, Texas. Bull.
Environ. Contain. Toxicol.,26,769.

Murray, H., Neff, G.S., Hrung, Y. & Giam, C.S. (1980). Determination of
benzo(a)pyrene, hexachlorobenzene and pentachlorophenol in oysters from Galveston
Bay, Texas. Bull. Environ. Contain. Toxicol., 25, 663.

National Academy of Sciences (1975). Petroleum in the Marine Environment. U.S.
National Academy of Sciences, Washington DC, 107 pp.

National Academy of Sciences. (1980). The International Mussel Watch. U.S. National
Academy of Sciences, Washington DC.

National Academy of Sciences (1985). Oil in the Sea. Inputs, Fates, and Effects. U.S.
National Academy of Sciences, Washington DC, 601 pp.

Neely, W.B., Branson, D.R. & Blau, G.E. (1974). Partition coefficient to measure
bioconcentration potential of organic chemicals in fish. Environ. Sci. Tech. 13,
1113.

Neff, J. M. (1979). Polycyclic Aromatic 1@ydrocarbons in the Aquatic Environment:
Sources, Fates, and Biological Effects. Applied Science Publishers, London, 262 pp.

Neff, J.M. & Anderson, J.W. (1975). Accumulation, release, and distribution of
benzo(a)pyrene-14C in the clam Rangia cuneata. In: Proceedings of the 1975
Conference on the Prevention and Control of Oil Pollution. American Petroleum
Institute, Washington DC, pp. 469-7 1.

Neff, J.M., Cox, B.A., Dixit, D. & Anderson, J.W. (1976). Accumulation and release
of petroleum -derived aromatic hydrocarbons by four species of marine animals. Mar.
Biol., 38, 279.

200

Niimi, A. J. & Oliver, B. G. (1989). Distribution of polychlorinated biphenyl congeners
and other halocarbons in whole fish and muscle among Lake Ontario salmonids.
Environ. Sci. Technol., 23, 83.

Nunes, P. & Benville, Jr., P.E. (1979). Uptake and depuration of petroleum
hydrocarbons in the Manila clam, Tapes semidecussata Reeve. BulL Environ.
Contam. Toxicol., 21, 719.

Obana, H. Hori, S., Nakamura, A. & Kashimoto T. (1983). Uptake and release of
polynuclear aromatic hydrocarbons by short-necked clams (Tapes japonica). Water
Res., 17, 1183.

Ogata, M., Fujisawa, K., Ogino, Y. & Mano, E. (1984). Partition coefficients as
measure of bioconcentration potential of crude oil compounds in fish and shellfish.
Bull. En viron. Con tam. Toxicol., 33, 5 6 1.

Olafsson, P.G., Bryan, A.M. & Stone W. (1987). PCB congener-specific analysis: a
critical evaluation of toxic levels in biota. Chemosphere, 16, 2585.

Oliver, B. G. (1984). Uptake of chlorinated organics from anthropogenically
contaminated sediments by Oligochaete worms. Can. J. Fish. Aquat. Sci., 41, 878.

Oliver, B. G. (1987). Biouptake of chlorinated hydrocarbons from laboratory-spiked and
field sediments by Oligochaete worms. Environ. Sci. Technol., 21,785.

Oliver, B. G. & Niimi, A. J. (1983). Bioconcentration of chlorobenzenes from water by
Rainbow Trout: Correlations with partition coefficients and environmental residues.
Environ. Sci. Technol., 17, 287.

Oliver, B. G., Baxter, R. M. & Lee, H-B. (1989). Polychlorinated biphenyls. In:
Analysis of Trace Organics in the Aquatic Environment. Afghan, B. K. & Chau, A.
S. Y. (Eds.). CRC Pres, Inc, Boca Raton, FL, pp. 31-68.

Onuska, F. 1. & Comba, M. (1980). Identification and quantitative analysis of
polychlorinated biphenyls on WCOT glass capillary columns. In: Hydrocarbons and
Halogenated Hydrocarbons in the Aquatic Environment, Afghan, B. K. & Mackay,
D. (Eds.). Plenum Press, New York, pp. 285-302.

Opperhuizen, A. & Schrap, S. M. (1988). Uptake efficiencies of two
polychlorobiphenyls in fish after dietary exposure to five different concentrations.
Chemosphere, 17, 253.

Opperhuizen, A., Gobas, F. A. P. C., Van der Steen, J. M. D. & Hutzinger 0. (1988).
Aqueous solubility of polychlorinated biphenyls related to molecular structure.
Environ. Sci. Technol., 22, 638.

Opperhuizen, A., Van der Velde, E. W., Gobas, F. A. P. C., Liem, A. K. D., Van der
Steen, J. M. D. & Hutzinger, 0. (1985). Relationship between bioconcentration in
fish and steric factors of hydrophobic chemicals. Chemosphere, 14, 187 1.

201

Palmork,K.H. & Solbakken, J.E. (1981). Distribution and elimination of [9-14C]
phenanthrene in the horse mussel (Modiola modiolus). Bull. Environ. Contain.
Toxicol. 26: 196.

Patil, G.S. (1991). Correlation of aqueous solubility and octanol-water partition
coefficient based on molecular structure. Chemosphere, 22, 723.

Patterson, Jr., D.G., Lapeza, Jr., C.R., Barnhart, E.R., Groce, D.F. & Burse, V.W.
(1989). Gas chromatographic/mass spectrometric analysis of human serum for non-
ortho (coplanar) and ortho substituted polychlorinated biphenyls using isotope-
dilution mass spectrometry. Chemosphere, 19, 127.

Pavlou, S. P. & Dexter, R. N. (1979). Distribution of polychlorinated biphenyls (PCB)
in estuarine ecosystems. Testing the concept of equilibrium partitioning in the marine
environment. Environ. Sci. Technol., 13, 65.

Payne, J.F. & Penrose, W.R. (1975). Petroleum induction of benzo(a)pyrene
hydroxylase in marine organisms. Can. Fed. Biol. Soc., 18, 56.

Payne, J.F. & May, N. (1979). Further studies on the effect of petroleum hydrocarbons
on mixed-function oxidases in marine organisms. In: Pesticide and Xenobiotic
Metabolism in Aquatic Organisms, Khan, M.A.Q., Lech, J.J. & Menn, J.J. (Eds.),
American Chemical Society, Washington DC, pp. 339-47.

Phillips, D.J.H. (1986). Use of organisms to quantify PCBs in marine and estuarine
environments. In: PCBs and the Environment. Vol. 11, Waid, J.S. (Ed), CRC
Press, Inc, Boca Raton, FL, pp. 127.

Pittinger, C.A., Buikema, Jr., A.L.Hornor, S.G. & Young, R.W. (1985). Variation in
tissue burdens of polycyclic aromatic hydrocarbons in indigenous and relocated
oysters. Environ. Toxicol. Chem., 4, 379.

Poland, A. & Knutson, J.C. (1982). 2,3,7,8 tetrachlorodibenzo-p-dioxin and related
halogenated aromatic hydrocarbons: examination of the toxicity. Ann. Rev.
Pharmacol. Toxicol., 22, 517.

Prahl, F. G. & Carpenter, R. (1983). Polycyclic aromatic hydrocarbon (PAH)-phase
associations in Washington coastal sediment. Geochim. Cosmochim. Acta, 47,
1013.

Prahl, F. G., Crecelius, E. & Carpenter, R. (1984). Polycyclic aromatic hydrocarbons in
Washington coastal sediments: an evaluation of atmospheric and riverine routes of
introduction. Environ. Sci. Technol., 18, 687.

Pruell, R. J., Lake J. L., Davis, W. R. & Quinn, J. G. (1986). Uptake and depuration
of organic contaminants by the blue mussels (4vtilus edulis) exposed to
environmentally contaminated sediments. Mar. Biol., 91, 497.

Pruell, R. J., Quinn, J. G., Lake, J. L. & Davis, W. R. (1987). Availability of PCB's
and PAH's to Mytilus edulis from artificially resuspended sediments. In: Oceanic
Processes in Marine Pollution. Vol. 1. Biological Processes and Wastes in the

202

Ocean. Capuzzo, J. M. & Kes ter, D. R. (Eds.), Kriegel Publ. Co., Malabar, FL,
265 pp.

Pugsley, C.W., Hebert, P.D.N., Wood, G.W., Brotea, G. & Obal, T.W. (1985).
Distribution of contaminants in clams and sediments from the Huron-Eire corridor. I-
PCBs and octachlorostyrene. J. Great Lakes Res., 11, 275.

Ramos, L. & Prohaska, P. G. (1981). Sephadex LH-20 chromatography of extracts of
marine sediments and biological samples for the isolation of polynuclear aromatic
hydroc arbons. J. Chromatogn, 211, 284.

Riley, R. T., Mix, M. C., Schaffer, R. L. & Bunting, D. L. (1981). Uptake and
accumulation of naphthalene by the oyster Ostrea edulis in a flow-through system.
Mar. Biol., 61, 267.

Risebrough, R. W., DeLappe, B. W., Walker II, W., Simoneit, B. T., Grimalt, J.,
Albaiges, J. & Regueiro, J. A. G. (1983). Application of the Mussel Watch concept
in studies of the distribution of hydrocarbons in the coastal zone of the Ebro Delta.
Mar. Poll. Bull., 14, 181.

Robertson, L.W., Parkinson, A., Bandiera, S., Lambert, I., Merril, J. & Safe, S.
(1984). PCBs and PBBs: biology and toxic effects on C57BU6J and DBA/2J inbred
mice. Toxicol., 31, 191.

Roesijadi, G., Anderson, J.W. & Blaylock, J.W. (1978). Uptake of hydrocarbons from
marine sediments contaminated with Prudhoe Bay crude oil: influence of feeding type
of test species and availability of polycyclic aromatic hydrocarbons. J. Fish. Res.
Board Can., 35, 608.

Rubistein, N. I., Gilliam, W. T. & Gregory, N. R. (1984). Dietary accumulation of
PCB's from a contaminated sediment source by a dernersal fish (Leiostomus
xanthurus). Aquatic Taxicol., 5, 331.

Safe, S. (1984). Polychlorinated biphenyls (PCBs) and polybrominated biphenyls
(PBBs): biochemistry, toxicology and mechanisms of action. CRC Crit. Rev.
Toxicol., 26, 319.

Safe, S. (1986). Comparative toxicology and mechanisms of action of polychlorinated
dibenzo-p-dioxins and dibenzofurans. Ann. Rev. Pharmacol. Toxicol., 26, 371.

Safe, S. (1990) . Polychlorinated biphenyls (PCBs), dibenzo-p-dioxins (PCDDs),
dibenzofurans (PCDFs), and related compounds: environmental and mechanistic
considerations which support the development of toxic equivalency factors (TEFs).
Crit. Rev. Toxicol., 21, 5 1.

Safe, S. (1992). Development, validation and limitations of toxic equivalency factors.
Chemosphere, 25, 61.
Safe, S., Safe, L. & Mullin, M. (1987). Polychlorinated biphenyls: Environmental
occurrence and analysis. In: Polychlorinated Biphenyls (PCBs): Maminalian and
Environmental Toxicology. Safe, S. (Ed.), Springer-Verlag, Berlin, pp. 1-13.

203

Salazar, M. H. (1986). Environmental significance and interpretation of organotin
bioassays. Proceedings of the Oceans86 Conference, Washington DC 23-25 Sept.
1986, pp. 1240-5.

Salazar, M. H. & Salazar, S M. (1987). TBT effects on juvenile mussel growth.
Proceedings of the Oceans87 Conference, Halifax, Nova Scotia, Canada, 28 Sept.-1
Oct. 1987, pp. 1504-10.

Salazar, M. H. & Salazar, S M.. (1988). Tributyltin and mussel growth in San Diego
Bay. Proceedings of the Oceans88 Conference, Baltimore, MD, 31 Oct.-2 Nov.
1988, pp. 1188-95.

Salazar, S. M., Davidson, B. M., Salazar, M. H., Stang, P. M. & Meyer-Schulte, K.
(1987). Field assessment of a new site-specific bioassay system. Proceedings of the
Oceans'87 Conference, Halifax, Nova Scotia, Canada, 28 Sept.-I Oct. 1987, pp.
1461-70.

Saleh, F.Y. & Lee, G.F. (1976). Analytical Methodologyfor Chlorinated Hydrocarbon
Pesticides and Polychlorinated Biphenyls in Water, Elutriate and Sediments Using
EC-GC. Center for Environmental Studies, University of Texas at Dallas, Paper #12.

Samuelian, J. & O'Connor, J. M. (1985). Structure-activity relationships and
accumulation of PCB congeners in estuarine fishes: A field study (Abstract only).
Estuaries, 8, p. 83A.

Schmidt, H. & Schultz, G. (1881). Uber benzidin (alpha-di-amidodiphenyl). Ann.
Chem. Liebigs., 207, 320.

Schulz, D.E., Petrick, G. & Duinker, J.C. (1989). Complete characterization of
polychlorinated biphenyl congeners in commercial Aroclor and Clorphen mixtures by
multidimentional gas chromatography-electron capture detection. Environ. Sci.
Technol., 23, 852.

SchijOrmann, G. & Klein, W. (1988). Advances in bioconcentration prediction.
Chemosphere, 17, 1551.

Schwartz, T.R., Tillitt, D.E., Feltz, K.P. & Peterman, P.H. Determination of mono- and
non-o,o'-chlorine substituted polychlorinated biphenyls in Aroclors and
environmental samples. Chemosphere (submitted).

Seligman, P.F., Grovhoug, J.G. & Ritcher, K.E. (1986). Measurement of butyltins in
San Diego Bay, CA: A monitoring strategy. In: Proceedings of the Oceans86
Conference, Washington DC 23-25 Sept. 1986, pp. 1289.

Sericano, J.L., El-Husseini, A.M. & Wade, T.L. (1991). Isolation of planar
polychlorinated,biphenyls by carbon column chromatography. Cheinosphere, 23,
1541.

Sericano, J. L., Wade, T.L. & Brooks, J.M. The usefulness of transplanted oysters in
biomonitoring studies. In: Proceedings of The Coastal Society 12th International
Conference, San Antonio, TX (in press).

204

Sericano, J. L., Atlas, E. L., Wade, T. L. & Brooks, J. M. (1990a). NOAA's Status
and Trends Mussel Watch Program: Chlorinated pesticides and PCBs in oysters
(Crassostrea virginica) and sediments from the Gulf of Mexico, 1986-1987. Mar.
Environ. Res., 29, 161.

Sericano, J. L., Wade, T. L., Atlas, E. L. & Brooks, J.M. (1990b). Historical
perspective on the environmental bioavailability of DDT and its derivatives to Gulf of
Mexico oysters. Environ. Sci. Technol., 24, 1541.

Sericano, J.L., Wade, T.L., El-Husseini, A.M. & Brooks, J.M. (1992) Environmental
significance of the uptake and depuration of planar PCB congeners by the American
oyster (Crassostrea virginica). Mar. Poll. Bull., 24, 537.

Shaw, G. R. & Connell, D. W. (1980). Relationships between steric factors and
bioaccumulation of polychlorinated biphenyls (PCBs) by the sea mullet (Mugil
cephalus Linnaeus). Chemosphere,9,731.

Shaw, G. R. & Connell, D. W. (1982). Factors influencing concentrations of
polychlorinated biphenyls in organisms from an estuarine ecosystem. Aust. J. Mar.
Freshwater Res., 33, 1057.
Shaw, G. R. & Connell, D. W. (1984). Physicochernical properties controli@g
polychlorinated biphenyl (PCB) concentrations in aquatic organisms. Environ. Scl.
Technol., 8, 18,

Smith, L.M., Schwartz, T.R. & Feltz, K. (1990). Determination and occurrence of
AHH-active polychlorinated biphenyls, 2,3,7,8-tetrachloro-p-dioxin and 2,3,7,8
tetrachlorodibenzofuran in Lake Michigan sediment and biota. The question of their
relative toxicological significance. Chenwsphere, 21, 1063.

Smith, L.M, Stalling, D.L. & Johnson, J.L. (1984). Determination of part-per-trillion
levels of polychlorinated dibenzofurans and dioxins in environmental samples. Anal.
Chem., 56, 1830.

Sodergren, A. & Larsson, P. (1982). Transport of PCB's in aquatic laboratory model
ecosystems from sediments to the atmosphere via the surface microlayer. Ambio, 11,
41.

Spague, J.B. (1969). Measurement of pollutant toxicity to fish. Part I. Water Research,
3, 793.

Spies, R. B., Felton, J. S. & Dillard, L. (1982). Hepatic mixed function oxidases in
California flat fishes are increased in contaminated environments or by oil and PCB
ingestion. Mar. Biol., 70, 117.

Stahl, R.G. (1980). Polychlorinated Biphenyls in Water, Sediments and Selected
Organisms of Galveston Bay) Texas. Environmental Levels and Bioaccumulation.
M. S. Thesis, Texas A&M Univeristy, College Station, TX, 151 pp.

Stainken, D.M. (1975). Preliminary observations on the mode of accumulation of #2 fuel
oil by the soft shell clam, Mya arenaria. In: Proceedings of the 1975 Conference on

205

Prevention and Control of Oil Pollution, American Petroleum Institute, Washington
DC, pp. 463-8.

Stainken, D.M. (1977). The accumulation and depuration of No 2 fuel oil by the soft
shell clam, Mya arenaria. In: Fate and Effects of Petroleum Hydrocarbons in Marine
Organisms and Ecosystems, Wolfe, D.A. (Ed.), Pergamon Press, New York, 313
pp-

Stalling, D.L., Huckins, J.N. & Petty, J.D. (1980). Presence and potential significance
of o-o-unsubstituted PCB isomers and trace Aroclor 1248 and 1254 impurities. In:
Hydrocarbons and Halogenated Hydrocarbons in the Aquatic Environment Afghan,
B.K. and Mackay, D. (Eds.). Plenum Press, New York, pp. 131-9.

Stang, P.M. & Seligman, P.F. (1987). In situ adsorption and desorption of butyltin
compounds from Pearl Harbor, Hawaii, sediments. In: Proceedings of the Oceans87
Conference, Halifax, Nova Scotia, Canada, 28 Sept.-I Oct. 1987, pp. 1386-91.

Stegeman, J. (1980). Mixed-function oxygenase studies in monitoring for effects of
organic pollution. V Reun. Cons. Int. Explor. Mer, 179, 33.

Stegeman, J. J. & Teal, J. M. (1973). Accumulation, release and retention of petroleum
hydrocarbons by the oyster Crassostrea virginica. Mar. Biol., 22, 37.

Stein, J. E., Hom. T. & Varanasi, U. (1984). Simultaneous exposure of English Sole
(Parophrys vetulus) to sediment-associated xenobiotics: Part 1- Uptake and depuration
of 14C-polychlorinated biphenyls and 3H-benzo(a)pyrene. Mar. Environ. Res., 13,
97.

Steinhauer, M. S. & Boehm, P.D. (1992). The composition and distribution of saturated
and aromatic hydrocarbons in nearshore sediments, river sediments, and coastal peat
of the Alaskan Beaufort Sea: Implications for detecting anthropogenic hydrocarbon
inputs. Mar. Environ. Res., 33, 223.

Stekoll, M. S., Clement, L. E. & Shaw, D. G. (1980). Sublethal effects of chronic oil
exposure on the intertidal clam Macoma balthica. Mar. Biol., 57, 51.

Stephenson, M. D., Smith, D. R., Goetzl, J., Ichikawa, G. & Martin, M. (1986).
Growth abnormalities in mussels and oysters from areas with high levels of tributyltin
in San Diego Bay. Proceedings of the Oceans86 Conference, Washington DC, 23-
25 Sept. 1986, pp. 1246-51.

Stickle, W. B., Rice, S. D. & Moles, A. (1984). Bioenergetics and survival of the snail
nais lima during long-term oil exposure. Mar. Biol., 80, 281.

Strawn, K., Aldrich, D.V., Wilson, W.B., Wisepape, L.M., Jones, F.V., Gibbard, G.,
Branch, M., Fredieu, B., St. Clair, L.A. & Strong, C. (1977). The effects on
selected organisms of water passing through the Cedar Bayou generating station.
Texas Agricultural Experiment Station, Project 1869, 144 pp.

Tanabe, S. (1985). Distribution, behavior and fate of PCB's in the marine environment.
J. Oceanog. Soc. Japan, 41, 358.

206

Tanabe, S. (1988). PCB problems in the future: foresight from current knowledge.
Environ. Poll., 50, 5.

Tanabe, S. & Tatsukawa, R. (1986). Distribution, behavior and load of PCBs in the
oceans. In: PCBs and the Environment, Vol. 1, Waid, J.S. (Ed.), CRC Press, Boca
Raton, FL, pp. 143-61.

Tanabe, S. Hidaka, H. & Tatsukawa, R. (1983a). PCB's and chlorinated hydrocarbon
pesticides in antarctic atmosphere and hydrosphere. Chemosphere 12, 277.

Tanabe, S., Mori, T. & Tatsukawa, R. (1983b). Global pollution of marine mammals by
PCBs, DDTs and HCHs (BHCs). Chemosphere, 12, 1269.

Tanabe, S., Tanaka, H. & Tatsukawa, R. (1984). Polychlorinated biphenyls, DDT and
hex achlorocyclohexane isomers in the western North Pacific ecosystems. Arch.
Environ. Contain. Toxicol., 13, 7 3 1.

Tanabe, S., Subramanian, A., Hidaka, H. & Tatsukawa, R. (1986). Transfer rates and
pattern of PCB isomers and congeners and p-p'DDE from mother to egg in Adelic
Penguin (Pygoscelis adeliae). Chemosphere, 15, 343.

Tanabe, S., Tatsukawa, R. & Phillips, D. J. H. (1987a). Mussels as bioindicators of
PCB pollution: A case study on uptake and release of PCB isomers and congeners in
green-lipped mussels (Perna viridis) in Hong Kong waters. Environ. Poll., 47, 4 1.

Tanabe, S., Kannan, N., Wakimoto, T. & Tatsukawa, R. (1987b). Method for the
detemination of three toxic non-ortho chlorine substituted coplanar PCB's in
environmental samples at par-per- trillion levels. Int. J. Environ. Anal. Chem., 29,
199-

Tanabe, S., Kannan, N., Subramanian, A., Watanabe, S. & Tatsukawa, R. (1987c).
Highly toxic coplanar PCB's: occurrence, source, persistency and toxic implications
to wildlife and humans. Environ. Pollut., 47, 147.

Tanacredi, J.T & Cardenas, R.R. (1991). Biodepuration of polynuclear aromatic
hydrocarbons from a bivalve mollusc, Mercenaria mercenaria L. Environ. Sci
Technol., 25, 1453.

Tripp, B.W., Farrington, J.W., Goldberg, E.D. & Sericano, J. (1992). International
Mussel Watch: the initial implementation phase. Mar. Poll. Bull, 24, 371.

Tulp, M.T.M. & Hutzinger, 0. (1978). Some thoughts on aqueous solubilities and
partition coefficients of PCB and the mathematical correlation between
bioaccumulation and phisicochemical properties. Chemosphere, 10, 849.

Unger, M. A., Maclntyre, W. G., Greaves, J. & Huggett, R. J. (1986) GC
determination of butyltins in natur"al waters by flame photometric detection of hexyl
derivatives with mass spectrometric confirmation. Chemosphere, 15, 461.

207

Unger, M. A., Mac Intyre, W. G. & Huggett, R. J. (1987). Equilibrium sorption of
tributyltin chloride by Chesapeake Bay sediments. Proceedings of the Oceans'87
Conference, Halifax, Nova Scotia, Canada, 28 Sept.-1 Oct. 1987, pp. 1381-5.

U.S. Environmental Protection Agency (1987). Tributyltin technical support document -
Position document 2/3. Office of Pesticides and Toxic Substances, Washington DC.

U. S. Environmental Studies Board (1979). PCB transport throughtout the environment.
In: Polychlorinated biphenyls. National Academy of Sciences, Washington) DC, pp.
11-76.

Valkirs, A.O., Seligman & P.F., Lee, R.F. (198b). Butyltin partitioning in marine
waters and sediments. In: Proceedings of the Oceans'86 Conference, Washington
DC, 23-25 Sept. 1986, pp. 1165-70.

Valkirs, A. 0., Davidson, B. H., Seligman, P.F. "1987a). Sublethal growth effects and
mortality to marine bivalves from long-term exposure to tributyltin. Chemosphere,
16, 201.

Valkirs, A.O., Stallard, M.O. & Seligman, P.F. (1987b). Butyltin partitioning in marine
waters. In: Proceedings of the Oceans'87 Conference, Halifax, Nova Scotia, Canada,
28 Sept.-1 Oct. 1987, pp. 1375-80.

Vandermeulen, J.H. & Penrose, W.R. (197/8). Absence of aryl hydrocarbon
hydroxylase (AHH) in three marine bivalves. J. Fish. Res. Board Can., 35, 633.

Venkatesan, M. 1. & Kaplan, 1. R. (1982). DiStribution and transport of hydrocarbons in
surface sediments of the Alaskan outer continental shelf. Geochim. Cosmochim.
Acta, 46, 2135.

Venkatesan, M. I., Kaplan, 1. R. & Ruth, E. (1983). Hydrocarbon geochemistry in the
surface sediments of the Alaskar outer continental shelf: Part I C15+ hydrocarbons.
Am. Assoc. Geolog. Bull., 67, 831.

Voogt, P de & Brinkman, U.A.Th (1989). Production, properties and Usage of
polychlorinated biplhenyls. In: Halogenated Biphenyls, Terphenyls, Naphthalenes,
Dibenzodioxins and Related Products. Kimbrough & Jensen (Eds.). Elsevier
Science Publishers, New York, pp. 1-45.

Vreeland, V. (1974). Uptake of chlorobiphenyls by oysters. Environ. Pollut., 6, 135.

Wade, T.L. & Garcia-Romero, B. (1989). Status and Trends of tributyltin contamination
of oysters and sediments from the Gulf of Mexico. In: Proceedings of the Oceans'89
Conference, Seattle WA 18-21 Sept. 1989, pp, 550-3.

Wade, T. L., Garcia-Romero, B. & Brooks, J. M. (1988a). Tributyltin contamination of
bivalves from U.S. coastal estuaries. Eviron. Sci. Technol., 22, 1488.

Wade, T.L., Garcia-Romero, B. & Brooks, J.M. (1988b). Tributyltin analyses in
association with NOAA's National Status and Trends Mussel Watch Program. In:

208

Proceedings of the Oceans88 Conference, Baltimore, MD, 31 Oct.-2 Nov. 1988, pp.
1198-201.

Wade, T.L., Garcia-Romero, B. & Brooks, J.M. (1990). Butyltins in sediments and
bivalves from U.S. coastal areas. Chemosphere, 20, 647.

Wade, T. L., Atlas, E. L, Brooks, J. M., Kennicutt 11, M. C., Fox, R. G., Sericano, J.
L., Garcia, B. & DeFreitas, D. (1988c). NOAA Gulf of Mexico Status and Trends
Program: Trace organic contaminant distribution in sediments and oysters. Estuaries,
11, 171.

Wakeham, S.G., Schaffner, C. & Giger, W. (1980). Polycyclic aromatic hydrocarbons
in Recent lake sediments- I. Compounds having anthropogenic origins. Geochim.
Cosmochim Acta, 44, 403.

Waldock, M.J., Thain, J.E. & Miller, D. (1983). The accumulation and depuration of
bis(tributyltin) oxide in oysters: A comparison between the Pacific oyster
(Crassostrea gigas) and the European flat oyster (Ostrea edulis). Int. Counc. Explor.
Sea, CM 1983[E:52.

Weigelt, V. (1986). Kapillargaschromatographische PCB-musteranalyse mariner spezies
- Betrachtungen zwischen anreicherung und chlorsubstitution ausgewahlter PCB-
komponenten in marinen organismen aus der Deutschen Bucht (in german).
Chemosphere, 15, 289.

Whitehouse, B. G. (1984). The effects of temperature and salinity on the aqueous
solubility of polynuclear aromatic hydrocarbons. Mar. Chem., 14, 319.

Widmark, G. (1967). Possible Interference by chlorinated biphenyls. J. Assoc. Off.
Anal. Chem., 50, 1069.

Windsor Jr., J.G. & Hites, R.A. (1979). Polycyclic aromatic hydrocarbons in Gulf of
Maine sediments and Nova Scotia soils. Geochim. Cosmochim Acta, 43, 27.

Wolfe, N.A., Clark, R.C., Foster, C.A., Hawkes, J.W. & MacLeod, W.D. (1981).
Hydrocarbon accumulation and histopathology in bivalve molluscs transplanted to the
Baie de Morlaix and the Rade de Brest. In: Amoco Cadiz: Fates and Effects of the Oil
Spill. CNEXO, Paris, pp 599-616.

Wong, W.C. (1976). Uptake and retention of Kuwait crude oil and its effect on oxygen
uptake by the soft-shell clam, Mya arenaria. J. Fish. Res. Board Can., 33, 2774.

Wormell, R.L. (1979). Petroleum Hydrocarbons Accumulation Patterns in Crassostrea
virginica: Analyses and Interpretations. Ph.D. Dissertation, Rutgers University, New
Brunswick, NJ, 189 pp.

Zuolian, C. & Jensen, A. (1989). Accumulation of organic and inorganic tin in blue
mussel, Mytilus edulis, under natural conditions. Mar. Poll. Bull., 20, 281.

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TABLE A-1
Biological Ancillary Parameters in Transplanted and Indigenous Oysters.

Sample Days after Shell length Wet weight Lipids
u=splants (cm) (g) M

Hanna Reef-to-Ship Channel (HRSC)
HRSC 3 7.6+0.9 7.6+2.4 10+2.2
HRSC 7 7.5+0.8 7.9+3.1 10+3.4
HRSC 17 7.8+1.2 10+3.7 9.7+1.7
HRSC 30 7.3+0.5 8.3+1.9 11+4.0
HRSC 48 8.1+1.2 11+3.7 11+2.8

Hanna Reef-Ship Channel-Hanna Reef (HRSCHR)
HRSCHR 51 7.2+0.5 7.9+1.3 13+2.3
HRSCHR 54 7.9+1.1 10+3.8 11+0.6
HRSCHR 66 7.6+1.2 9.3+3.5 12+2.4
1HRSCHR 78 7.4+0.8 9.6+3.6 11+0.8
HRSCHR 98 7.7+1.0 13+3.9 12+2.5

Ship Channel (SC)
Sc 3 7.4+1.3 7.3+1.6 14+3.6
Sc 7 Sample was not collected
SC 17 10+1.1 17+4.4 14+0.6
Sc 30 8.9+1.1 11+2.3 15+1.3
Sc 48 9.3+1.1 11+3.0 15+0.3

Ship Channel-to-Hanna Reef (SCHR)
SCHR 51 10+1.5 16+3.4 13+1.0
SCHR 54 8.7+1.1 13+3.0 13+1.4
SCHR 66 8.7+1.5 12+6.0 12+0.9
SCHR 78 8.4+1.2 11+4.1 12+1.3
SCHR 98 7.3+1.7 14+1.6 13+0.8

TABLE A-2
Average PAH Concentrations I S.D.) in Seawater, Sediment and Indigenous Ship Channel (SC) oyster Samples During the Uptake Phase
of the Experiment at the Ship Channel Site and Estimated Bioconcentration Factors (BCF).

Uptake phase (days)
Analyte Water Sediment 0 3 7 17 30 48 BCFa
ng 1-1 ng g-I ng g-I

Naphthalene 3.8�1.0 10�23 - 25�2.9 - 12+0.4 7.8�0-8 7.2�1.8 1,900
2-Methyinaphthalenc 2.5�0.7 15�4.6 - 25�4.5 - 20�6.7 19�2.9 9.1�0.6 3,600
I-Methyinaphthalene 1.9�0.6 7.8�1.4 - 13�3.1 - 13�4.0 9.1�2.1 3.8�0.8 2,0M
Biphenyl 1.2�0.3 3.6�0.2 - 11�2.1 - 6.2�1.1 7.1�0.4 3.6�0.2 3,000
2,6-Dimethylnaphthaicnc 1.6�0.5 12�4.8 - 22�2.2 - 17�0.6 49�3.3 13�3.5 8,100
Acenaphthylene 0.7�0.1 23�7.2 14�3.3 - 7.9�0.3 6.8�0.2 6.9�0.2 9,900
Acenaphthene 0.6�0.1 9.5�3.0 21�2.6 - 13�1.0 31�1.3 5.8�0.4 9,700
2,3.5-Trimethyinaphthalene 2.4�0.6 32�5.3 53�9.1 - 83�14 120�6.3 56�6.2 23,000
Fluorene 1.0�0.2 15�7.2 23�1.8 - 22�2.2 34�4.1 15�1.7 15,000
Phenanthrene 2.2�0.7 51�5.8 45�3.6 - 39�2.7 110�12 23�2.9 10,000
Andvacene 1.0�0.3 54�11 31�9.4 - 29�3.8 46�1.4 30�3.5 30,000
I-Methylphenanthrene 1.5�0.6 37�15 63�12 - 110�2.8 88�13 89�11 59,000
Fluoranthene 1.6�1.2 140�58 - 390�45 - 560�95 790�83 490�51 310,000
Pyrene 2.1�1.6 190�38 - 1,200�130 - 1,400�270 1,400�59 1,900�80 900.000
Bcnz(a)anthracene 0.8�0.1 110�42 - 130�25 - 160�18 220�13 280�13 350,000
Chrysene 0.9�0.3 150�38 - 400�51 - 310�26 380�25 490�21 540,000
Benzo(b)fluoranthene 0.8�0.2 170�41 170�30 - 130�14 170�14 210�24 260,000

TABLE A-2
(confinued)

Uptake phase (days)
Analyte Water Sediment 0 3 7 17 30 48 BCFa
ng 1-1 ng g- I ng g-I

Benzo(k)fluoranthene 0.3�0.1 140:1:38 - 52�15 48�2.8 66�7.2 82�14 270,000
Benzo(e)pyrene 1.1�0.3 160�30 - 310�49 - 200�24 240�20 340�5.8 310.000
Benzo(a)pyrene 0.8�0.1 142�45 - 57�16 - 73�7.1 97�11 160�6.3 200,000
Perylene 1.1�0.3 110�18 - 110�25 - 110�11 140116 160�4.1 150,000
Indenot 1.2.3-c.d]pyrene 0.5�0.1 86�15 16�3.4 11�1.4 16�2.3 22�0.4 44,000
Dibenzo(a,b)anthracene 0.6�0.1 37�4.3 - 16�3.2 - 9.9�0.6 10�0.7 15�3.2 25,000
Benzo(g,h,i)peryiene 0.6�0.1 65�20 - 54�8.5 - 33�2.9 45D.7 6810.8 110,000

Total 2-Rings 13�3.1 80�7.4 - 150�18 - 150�23 210�14 93�7.5 7,200
Total 3-Rings 7.1�1.9 190�24 - 200�31 - 220�12 310�26 170�6.5 24,000
Total 4-Rings 5.2�2.1 580�170 - 2,100�250 - 2,400�390 2,800�68 3,100�120 600,000
Total 5-Rings 4.6�0.8 800�170 - 720�140 - 570�57 710�67 970�42 210,000
Total 6-Rings 1.1�0.2 150�26 - 70�11 - 44�4.2 61�6.0 90�1.2 82,000

Total PAHs 32�7.0 1,800�380 - 3,200�420 - 3.400�470 4,100�170 4.400�146 140,000

a Bioconcentration factor concentrafion in transplanted oyster fissue at the end of the uptake pcriod/concentration in water.

TABLE A-3
Averap PAH Concentrations (i I S.D.) in Seawater, Sediment and Hanna Reef-to-Ship Channel (HRSC) Transplanted oyster Samples
During the Uptake Phase of the Experiment at the Ship Channel Site and Estimated Bioconcentration Factors (BCF).

Uptake phase (days)
Analyte Water Sediment 0 3 7 17 30 48 BCFa
ng 1-1 ng g- I ng g-I

Naphthalene 3.8�1.0 10�2.7 22�7.5 12�1.9 12�3.0 14�3.6 8.3�2.2 7.5�1.1 2,0M
2-Methyinaphthalene 2.5�0.7 15�4.6 16�7.8 11�2.5 6.7�1.5 14�1.9 12�3.8 7.5�1.2 3,000
1 -Methylnaphthalene 1.9�0.6 7.8�1.4 9.0�3.6 5.7�0.9 3.9�1.0 9.9�1.6 5.6�1.7 3.9�0.6 2,100
Biphenyl 1.2�0.3 3.6�0.2 8.3�3.2 7.2�1.8 5.6�1.6 6.8�2.2 6.8�1.6 4.3�0.5 3,600
2,6-Dimethylnaphthalcne 1.6�0.5 12�4.8 15�3.2 9.8�0.9 6.8-f 1. 1 12�1.3 39�6.5 10�1.5 6,300
Acenaphthylcne 0.7�0.1 23�7.2 7.3�2.1 7.4�1.5 6.5�1.8 6.7�1.5 6.1�1.2 6.5�1.7 93M
Acenaphthene 0.6�0.1 9.5�3.0 4.0�1.7 9.2�0.8 9.5�1.1 9.1�1.6 25�6.6 7.6�1.3 13,000
2,3,5-Trimethylnaphthalene 2.4�0.6 32�5.3 26�5.9 22�4.9 14�3.3 50�11 90�20 38�11 16,000
Fluorene 1.0�0.2 15�7.2 9.3�1.5 11�1.2 11�1.4 16�2.3 26�5.9 16�3.1 16,000
Phenanthrene 2.2�0.7 51�5.8 14�4.2 20�1.8 21�1.6 28�5.4 87�19 48�21 22,000
Anthracene 1.0�0.3 54�11 16�1.6 11�1.6 12�3.1 20�4.1 36�11 30�6.7 30,000
1 -Methylphenanthmne 1.5�0.6 37115 62�31 23�4.2 20�5.2 65�16 96�19 64�27 43,000
Fluoranthene 1.6�1.2 140�58 11�2.5 110�18 130�19 410�45 620�110 490�76 310,000
Pyrene 2.1�1.6 190�38 16�3.5 300�44 340�50 1,000�100 1,300�190 1,900�190 900,000
Benz(Oanthracene 0. 8�0. 1 110�42 8.3�5.0 22�3.3 30�4.9 140�17 180�26 260�27 330,000
Chrysene 0.9�0.3 150�38 7.7�3.5 71�9.5 110�16 250�20 320�38 450�49 500,000
Ben7o(b)fluoranthene 0.8�0.2 170�41 3.0�1.0 28�4.9 39�3.7 100�2.7 140�14 220�20 280,000

TABLE A-3
(wntinued)

Uptake phase (days) 30 48 BCFa
Analyle water Sediment 0 3 7 -1 17
ng 1-1 ng 9-1 ng g
lucranthene, 0.3�0.1 140�38 3.3-+0.6 1 I:t 1. 1 14jo.9 41�4.3 50�9-1 86tl 1 290,000
Bcn7.0(k)j 160DO 4.0:t 1 .0 4913.2 76-19.1 1500.2 210jJ4 330-t24 300,000
Henzo(Opyrene 1.1�0.3 64�6.6 74�9-1 1 SOD i 190,WO
Benzo(a)pyrene 0.9�0.1 142-145 6.1-15.0 M3.0 16�0-9 10019.6 1 OOV 3 150127 140'OM
Perytcnc 1.1�0.3 110:H8 3.00-1 M2.9 27�2.4 1210.5 14:0-1 23-13.3 46,000
Indenot 1.2.3-c.dlpyrene 0.50. 1 g6:t , 5 j():j3.5 7.9f2.0 7.1+10-9 17�2.7 28,OM
Dibenzo(a.b)anthraccne 0.60.1 3714.3 3.0�2.6 6.7�1.9 5.6�0.5 9.7�1.6 11 �3.5
Benzo(g,b,i)perylcne 0.6�0.1 65:t2O 7.00.0 15�3.9 20�2.4 31�2.1 43�3.9 73�9.9 120,000
13�3.1 80�7.4 97�16 69�10 4913.8 110�15 160J:33 7)�12 5,500
TOW 2-Rings 7.1�1.9 190�24 110�27 8 1 -t9- 5 80�14 150123 280�62 .170�43 24,OW
Total 3-Rings 5.2�2.1 580�170 43�12 5OU69 610�97 1,8()O:t 180 2,400�340 3,100�290 600,000
Total 4-Rings 8OO�i7O 30�9.0 1301'16 190�17 470�1 g 590�46 95()�110 210,000
Total 5-Rings 4.6�0.9 .9 44�2.4 58�43 96�13 97,OM
ToW 6-Rings 1.1�0-2 15ft26 17�7.5 23�5.8 27+2
Total PAH9 32�7.0 i,soo-1380 -290�4o 810�83 950�110 2,600t220 3,500�390 4,40W30 138,000
a Vloconcentration factor concentration in transplanted oyster tiswe at the end of the Uptake Pcri(WcOncenvation in water- tj

ago

TABLE A-4
Average PAH Concentrations (� 1 S.D.) in Sediment and Ship Channel-to-Hanna Reef (SCHR) Transplanted Oyster Samples During the
Deputation Phase of the Experiment at the Hanna Reef Site and Estimated Biological Half-Lives (BHL).

Deputation phase (days)
Analyte Sediment 0 3 6 18 30 50 BHL(R2)a
ngg-I ng g- I

Naphthalene 3.9�0.3 7.2�1.8 8.5t].5 14�6.4 9.1�0.6 9.0�1.2 12�2.8
2-Methylnaphthalene 5.6�0.3 9.1�0.6 11�1.9 9.7�1.2 6.1�1.3 6.5�1.5 11�2.4
I-Mclhyinaphthalenc 4.1�0.4 3.8�0.8 5.5�0,8 6.0�1.4 3.9�0.7 5.3�1.6 7.8�1.0
Biphenyl 4.3�0.9 3.6�0.2 6.0�1.1 5.9�0.7 5.6�1.3 5.7�0.6 7.1�1.5
2,6-Dimcthyinaphthalenc 6.1�2.0 13�3.5 17�4.9 10�1.2 3.8�1.0 7.9�1.0 6.2�2.4
Acenaphthylene 4.8�0.4 6.9�0.2 5.1�2.6 3.1�3.3 3.7�2.7 3.8�2.0 3.1�1.2
Acenaphthene 2.9�0.4 5.8�0.4 4.2�0.9 3.1�0.9 2.2�0.7 14�2.1 3.3�1.0
2,3,5-Trimethyinaphthalene 5.6�0.9 56�6.2 44�10 23�6.8 17�5.4 16�2.8 9.1�2.6 22(0.83)
Fluorene 4.9�0.8 15�1.7 14�1.4 11�2.5 5.7�1.0 15�2.1 7.5�2.8
Phenanthrene 12�3.4 23�2.9 24�3.9 18�1.1 15�2.9 46�4.6 29�8.8
Anthracene 7.6�1.0 30�3.5 19�7.0 16�4.2 13�3.6 16�4.4 9.5�2.8 42(0.68)
I-Methylphenanthrene 4.8�2.4 89�11 66�16 54�17 47�15 36�6.5 18�5.4 24(0.96)
Fluoranthene 29�8.4 490�51 350�89 250�45 260�110 320�18 110�29 32(0.69)
Pyrene 29�7.7 1,900�80 1,300�290 870�230 500�140 300�37 95�28 12(0.98)
Benz(a)anthracene 18�4.2 280�13 200�48 170�44 110�32 59�9.4 26�7.2 15(0.99)
Chrysene 15�2.8 490�21 380�53 340�48 230�52 110�18 54�10 16(0.99)
Benzo(b)fluoranthenc 17�3.0 210�24 200�34 200�9.5 140116 59�11 18�5.6

TABLE A-4
(continued)

Depuration phase (days)
Anw@ Sediment 0 3 6 18 30 50 BtIL(R2)a
ngg-I ng g-l

Benzo(k)fluoranthene, 131l.7 82�14 71�12 72�8.8 35�2.4 17�3.9 5.5�2.7
Benzo(e)pyrene 17�2.8 340�5.8 310�16 300�6.9 210�20 89�21 44�8.7 16(0.98)
Benzo(a)pyrene 15:1. 1. 2 160�6.3 110�22 79�17 39�8.7 14�2.2 4.4�1.3 10(0.99)
Perylene 74�6.3 IW4.1 130118 120�23 71�18 28�6.6 11�3.8 13(0.99)
Indenol 1,2,3-c,d]pyrene 15�3.8 22�0.4 23�7.2 22�1.2 14�6.2 2.3�1.3 1.2�0.2 11 (0.93)
Dibenzo(a,h)anffiracene 4.9�0.9 15�3.2 17�7.5 18�4.3 15�6.1 3.3�0.8 1.7�0.6 14(0.90)
Benzo(g,h,i)perylene 14�3.5 68�0.8 70�18 73�6.1 37�7.2 13�0.4 4.8�1.7 12(0.98)

Total 2-Rings 30�3.8 93�7.5 92�18 68�4.9 45�3.8 50�4.4 53�13
Total 3-Rings 37�3.6 170�6.5 130:01 110�26 87�23 130�16 70�22
Total 4-Rings 90�22 3,100�120 2.200�450 1,600�360 1.100�310 780�76 290�74
Total 5-Rings 140�9.4 970�42 850�48 790�56 510�43 210�40 85�23
Total 6-Rings 30�6.9 90�1.2 93�25 95�7.2 50�13 15�1.0 6.0�1.9

Total PAHs 330�32 4,400�150 3.400�440 2,700�430 1,800�330 1200�120 500�130

a R2 = square of correlation coefficient for regression equation.

TABLE A-5
Average PAH Concentrations I S.D.) in Sediment and Hanna Roet-Ship Cbannel-Hanna Reef (14RSCHR) Transplanted Oyster Samples
During the Depuration Phase of the Experiment at the Hanna Reef Site and Esdmated Biological Half-Lives (BHL).

Depuration phase (days)
Analyte Sediment 0 3 6 18 30 50 BtIL(R2)a
ngg-I ng g-l

Naphthalene 3.9�0.3 7.5�1.1 7.5�0.5 9.3:LI.4 6.6�0.8 8.3�1.1 13�5.6
2-Methylnaphthalene 5.6�0.3 7.5�1.2 12�1.9 11�3.1 5.3�0.8 6.0�0.3 11�0.6
I-Methyinaphihalene 4. 110. 4 3.9�0.6 6.0�0.7 5.6�2.) 2.8�0.4 4.0�0.6 6.811.2
Biphenyl 4.3�0.9 4.3�0.5 5.0�0.3 4.7�1.0 4.8�1.0 6.4�1.1 7.7�1.9
2,6-Dimethyinaphthalene 6.1�2.0 10�1.5 19�0.8 7.3�0.6 3.6�1.1 7.9�1.8 6.4�2.0
Acenaphthylcne 4.8�0.4 6.5�1.7 6.3�1.7 3.9�1.2 3.2�1.2 3.5�1.5 3.6�1.4
Accnaphlbene 2.9�0.4 7.6�1.3 5.0�0.4 2.7�0.4 2.0�0.3 13�3.0 3.4�0.7
2,3,5-Trimethyinaphthalene 5.6�0.9 38�11 41�9.2 15�4.2 16�6.5 12�0.9 8.2�0.6 24(0.74)
Fluorene 4.9�0.8 16�3.1 15�1.4 9.0�1.7 5.1�0.5 13�2.0 8.1�1.9
Phenanthrene 12�3.4 48�21 33�2.8 15�2.2 12�1.6 "�l 1 24�2.8
Andiracene 7.6�1.0 30�6.7 26�4.8 15�2.8 17�4.9 11�2.2 5.8�2.4 24(0.90)
1-Methylphenandurne 4.8�2.4 64�27 46�8.4 47�4.5 29�10 26�4.0 12�2.4 23(0.97)
Fluoranthene 29�8.4 490�76 350�7.4 160�42 190�35 220�43 80�20 26(0.67)
Pyrene 29�7.7 1,900�190 1.400�86 570�170 350�100 170�41 56�10 10(0.95)
Benz(a)anthracene 18�4.2 260�27 250�22 150�47 70�25 39�5.4 20�1.7 13(0.96)
Chrysene 15�2.8 450�49 420�15 280�69 150�35 58�11 30�4.7 12(0.99)
Benzo(b)fluoranthene 17�3.0 220�20 200�9.5 210�16 76�6.3 25�8.7 15�4.2 12(0.95)

TABLE A-5
(confinued)

Dcpuration phase (days)
Analyte Sediment 0 3 6 18 30 50 BHL(R2)a
ng 9-1 ng g- I

Benzo(k)fluoranthene 13�1.7 86�11 76:L8.1 78�5.4 19�4.2 6.8�2.0 4.4�1.6 10(0.96)
Benzo(e)pyrene 17�2.8 330�24 310�21 280�19 110�26 39�12 21�43 12(0.97)
Benzo(a)pyrene MA.2 150�31 130�25 81�19 32�3.8 8.6�2.7 3.5�1.9 9(0.99)
Perylene 74�6.3 150�27 140117 100�19 28�8.1 13�2.1 7.5�1.1 11 (0.94)
Indenol 1,2.3-cdipyrene 15�3.8 23�3.3 18�3.1 16�2.6 8.4�1.8 1.8�0.9 1.0�1.1 10(0.96)
Dibenzo(a,h)anthracene 4.9�0.9 17�2.7 15�1.9 13�2.8 7.9�2.8 2.3�1.2 2.3�1.7 16(0.93)
Benzo(g,h,i)peryiene 14�3.5 73�9.9 61�1.2 51�3.3 17�3.6 7.7�1.1 4.1�1.6 11(0.96)

Total 2-Rings 30�3.8 71�12 90�12 53�4.8 39�6.1 45�0.4 52�7.4
Total 3-Rings 37�3.6 170�43 130�15 92�63 68�16 110�9.4 57�6.6
Total 4-Rings 90�22 3.100�290 2,400�96 1,200�320 750�180 490�97 190�20
Total 5-Rings 140�9.4 950�110 870�73 750�22 270�17 95�24 54�13
Total 6-Rings 30�6.9 93�13 79�4.0 67�33 26�4.2 9.5�1.4 5.1�2.4

Total PAHs 330�32 4,400�330 3,600�170 2,100�320 1.200�190 750�120 360�18

a R2 = square of correlation coefficient for regression equation.

TABLE A-6
Average PCB Congener Concentrations (� I S.D.) in Seawater. Sediment and Ship Channel (SC) Indigenous Oyster Samples During the
Uptake Phase of the Experiment at the Ship Channel Site and Estimated Bioconcentration Factors (BCF).

Uptake phase (days)
Analyte Water Sediment 0 3 7 17 30 48 BCFa
ng 1-1 ng g-I ng g-l

18 0.05�0.03 3.0�0.7 4.7�0.3 6.8�2.1 6.1�0.6 120,000
15/17 0.03�0.02 0.16�0.11 3.1�0.9 4.2�0.3 5.6�1.8 4.8�0.5 160,000
26 35�5.3 34�2.0 30�5.9 36�8.0
50/31 0.15�0.07 2.5�0.4 3.4�0.2 4.3�1.4 5.2�0.8
28 6.5�1.1 7.1�0.6 6.7�1.4 8.0�1.4
52 0.50�0.29 1.58�0.17 62�7.0 55�6.1 47�11 52�4.0 100,000
49 0.14�0.09 0.65�0.17 36�2.2 31�4.5 29�3.8 31�3.2 220,000
47/48/75 25�1.1 20�0.9 16�5.7 19�2.0
44 0.20�0.22 0.56�0.19 27�1.1 21�1.6 19�3.9 20�3.2 100,000
37/42/59 0.09�0.09 0.21�0.06 26�0.8 28�3.0 21�3.6 33�5.3 370,0M
41 0.44�0.09 21�4.6 17�4.7 15�2.3 16�2.8
40 0.14�0.02 9.7�0.4 10�0.3 10�1.2 13�0.5
74 0.07�0.03 0.23�0.07 17�0.9 13�3.2 12�0.9 15�0.9 210,000
70 0.05�0.03 0.87�0.41 46�2.5 34�0.9 32�0.4 38�2.0 760,000
95 95�9.0 71�0.8 56�12 63�8.2
91 0.07�0.04 0.40�0.14 27�2.6 20�2.7 15�4.9 22�1.0 310,000
60/56 10�0.8 8.2�1.1 8.9�1.5 9.5�1.4
92 0.45�0.21 31�7.2 32�5.8 18�5.7 26�6.6

TABLE A-6
(continued)

Uptake phase (days)
Analyte Water Sediment 0 3 7 17 30 48 BCFa
ng 1-1 ng g-I ng 9-1

84 0.71�0.23 46�0.8 35�4.9 27�7.9 29�7.3
101)90 0.27�0.12 2.19�0.65 110�13 85�2.2 67�8.1 71�8.5 260,000
99 0.14�0.08 1.44�0.09 62�5.5 52�1.1 41�4.9 45�4.6 320.000
97 0.68�0.26 33�3.0 26�1.4 20�2.1 20�3.7
87/11 1.33�0.40 53�4.9 40�1.4 30�3.6 32�3.1
1 10fl7 0.24�0.08 3.84�1.00 130�12 110�6.8 76�15 80�12 330,0(X)
82 0.30�0.06 14�1.3 9.2�0.8 7.7�1.0 6.3�2.3
151 0.08�0.05 0.22�0.04 16�1.1 13�0.8 9.1�2.1 10�2.2 130.000
135 0.10�0.02 0.16�0.02 14�5.0 10�1.8 7.4�1.9 8.7�2.1 87,000
107 15�1.9 13�2.8 12�0.2 15�3.1
149/123 0.19�0.12 0.90�0.23 44�4.5 36�2.4 31�1.9 36�3.5 190,000
118 0.12�0.04 1.35�0.37 88�12 68�5.6 51�8.3 56�6.3 470,000
146 0.13�0.05 13�3.4 10�0.9 8.0�1.1 7.8�2.1
1531132 0.59�0.32 2.15�0.54 170�11 110�19 93�10 100�10 170,000
105 0.12�0.06 0.67�0.16 30�7.8 23�5.2 15�3.5 17�6.1 140,000
141/179 0.12�0.11 0.34�0.03 7.3�1.7 4.5�0.5 3.7�1.0 5.2�0.9 43,000
138/160 0.59�0.17 2.52�0.12 77�12 60�5.5 44�7.3 47�10 80,000
187 0.14�0.09 0.11�0.04 18�6.9 15�2.4 13�1.3 15�2.5 110,000

TABLE A-6
(continued)

Uptake phase (days)
Analyte Water Sediment 0 3 7 17 30 48 BCFa
ng 1-1 ngg-I ng g-l

128 0.0810.06 0.42�0.09 10�0.6 7.7�1.3 6.3�0.1 5.9�1.0
177 0-05�0.04 0.18�0.06 5.3�0.2 4.3�0.9 3.7�0.2 4.6�0.1 92,000*
180 0.33�0.18 0.40�0.06 6.5�0.6 5.5:EO.9 4.1�0.5 4.6�0.2 14,000

Total dichlorobiphenyls 1.5
Total trichlorobiphenyis 0.13�0.08 0.31�0.17 59 64 64 71 540,000
Total tetrachlorobiphenyls 1.00�0.56 6.16�1.58 270 240 210 250 250,000
Total pentachlorobiphenyls 0.99�0.38 13.5�3.49 740 580 440 490 490,000
Total hexachlorobiphenyls 1.64�0.81 6.98�1.13 360 250 210 230 140,000
Total heptachlorobiphenyls 0.78�0.50 1.14�0.12 57 46 40 52 67,000
Total octachlorobiphenyis 0.09�0.05 0.25�0.03 1.6 1.3 0.5 1.9 21,000
Total nonachlorobiphenyis 1.0 0.7 0.5 0.8
Total decachlorobiphenyl 0.2 0.6 0.4 0.4

Total PCBs 4.62�2.15 28.4�6.41 1500 1200 960 1100 240,000

a Bioconcentration factor = concentration in transplanted oyster tissue at the end of the uptake period/concentration in water.

TABLE A-I
Average PCB Congener Concentrations (� I S.D.) in Seawater, Sediment and Hanna Reef-to-Ship Channel (HRSQ Transplanted Oyster
Sarnple@ During the Uptake Phase of the Experiment at the Ship Channel Site and Estimated Bioconccntration Factors (13CF).

Uptake phase (days)
Analyte Water Sediment 0 3 7 17 30 48 BCFa
ng 1-1 ng g- ng g-I

18 0.05�0.03 1.6�0.4 2.2�0.3 3.7�1.0 4.4�0.7 6.0�1.0 120,000
15/17 0.03�0.02 0.16�0.11 1.2�0.5 11�2.7 2.9�0.8 3.0�1.2 4.5�1.3 150,000
26 6.5�1.2 11�2.8 19�2.5 23�2.1 30�9.3
50/31 0.15�0.07 1.0�0.3 3.8�2.1 2.7�0.9 5.2�2.8 9.8�4.6
28 2.4�0.2 3.8�1.1 5.6�0.7 7.9�1.3 9.9�1.4
52 0.50�0.29 1.58�0.17 17�2.9 28�2.1 32�2.3 37�3.7 51�5.4 100,000
49 0.14�0.09 0.65�0.17 6.6�0.7 13�1.6 16�1.0 22�1.5 291:1.5 210,000
47/48175 5.6�1.6 9.6�1.1 11�1.3 14�1.3 18:L2.2
44 0.20�0.22 0.56�0.19 6.9�0.7 14�2.0 14�1.3 17�1.9 22�2.3 110,000
37/42/ 0.09�0.09 0.21�0.06 12:LO.7 17�3.3 19�1.9 22�1.6 42�8.3 470,000
41 0.44�0.09 4.3�0.8 8.0�0.1 8.9�3.3 12�2.1 20�1.4
40 0.14�0.02 1.4�0.5 4.6�1.4 6.7�0.7 9.4�0.9 14�2.5
74 0.07�0.03 0.23�0.07 2.9�0.5 5.2�0.5 7.3�0.9 9.4�2.1 13�2.7 190,000
70 0.05�0.03 0.87�0.41 4.7�1.2 9.7�1.9 15�1.6 23�3.5 31�2.8 620,000
95 11�1.2 22�2.1 29�1.2 41�5.5 55�5.3
91 0.07�0.04 0.40�0.14 4.2�0.5 9.1�1.1 11�0.9 14�2.0 24�8.6 340.000
60/56 2.2�0.5 3.5�0.6 5.4�0.8 7.3�1.4 10�1.9
92 0.45�0.21 4.8�1.3 6.5�1.1 11�2.5 15�1.0 22�8.0

TABLE A-7
(continued)

Uptake phase (days)
Analyte Water Sediment 0 3 7 17 30 48 BCFa
ng 1-1 ng g-I ng g-I

84 0.71�0.23 5.9�0.5 13�2.2 17�1.5 24�3.8 29�0.9
101/90 0.27�0.12 2.19�0.65 11�1.6 20�4.1 27�1.1 37�4.3 43�5.2 160,000
99 0. 14:LO.09 1.44�0.09 8.7�1.0 15�1.6 18�3.3 27�2.9 29�1.3 210.000
97 0.68�0.26 3.9�1.6 6.8�1.1 8.9�1.7 11�1.8 14�3.2
87/115 1.33�0.40 4.7�0.6 8.6�1.3 13�0.6 16�2.0 20�1.4
1 IOM 0.24�0.08 3.84�1.00 13�1.8 28�4.2 37�1.7 50�5.2 61�4.2 250,000
82 0.30�0.06 1.8�0.5 3.2�0.4 3.3�0.7 4.4�0.1 5.9�1.0
151 0-08�0-05 0.22�0.04 1.7�0.5 2.6�0.9 4.0�0.8 4.9�0.6 6.0�1.0 75,000
135 0.10�0.02 0.16�0.02 2.6�1.4 2.2�0.7 3.6�0.9 4.0�0.6 5.0�0.5 50,000
107 1.3�0.3 2.4�0.7 3.6�0.6 5.3�1.1 6.6�1.0
149/123 0.19�0.12 0.90�0.23 4.7�0.4 7.8�1.1 11�0.4 14�1.5 18�1.6 95,000
118 0.12�0.04 1.35�0.37 7.6�1.4 14�1.6 21�1.4 29�2.0 34�2.5 280,000
146 0.13�0.05 3.4�0.5 4.2�1.9 3.6�0.6 4.0�0.6 4.1�1.1
153/132 0.59�0.32 2.15�0.54 16�4.9 18�6.3 28�4.8 38�7.5 64�12 110,000
105 0.12�0.06 0.67�0.16 7.6�2.7 11�1.1 13�2.0 13�1.4 14�3.0 120,000
141/179 0.12�0.11 0.34�0.03 1.7�0.6 1.6�0.5 2.3�1.0 3.3�0.7 4.7�0.5 39,000
138/160 0.59�0.17 2.52�0.12 12�2.0 17�2.7 20�0.7 27�5.2 27�4.7 46,000
187 0.14�0.09 0.11�0.04 4.3�0.4 3.4�1.3 5.9�1.7 6.4�1.1 11�2.2 78,000

TABLE A-7
(continued)

Uptake phase (days)
Analyte Water Sediment 0 3 7 17 30 48 BCFa
ng 1-1 ng g- I ng g-l

128 0.42�0.09 1.7�0.3 2.5�0.3 2.9�0.4 3.8:10.6 3.9�0.6
177 0.05�0.04 0.18�0.06 1.0�0.2 1.3�0.5 1.6�0.8 1.7�0.1 3.5�1.6 70,000
180 0.33.+0.18 0.40�0.06 1.5�0.3 2.4�0.5 2.3�03 3.4�0.7 3.8� f.7 12,000

Total dichlorobiplicnyls 0.1�0.1 0.3�0.6 0.1�0.2
Total trichlorobiphcnyis 0.130.08 0.310.17 16�2.5 31�5.5 46�6.5 50�6.2 66�9.3 500,000
Total tetrachlorobiphenyis 1.000.56 6.161.58 58�13 110�9.2 130�7.6 170�19 240�24 240,000
Total pentachlorobiphenyls 0.990.38 13.53.49 84�7.6 160�17 210�9.2 290�15 360�26 360,000
Total hexachlorobiphenyis 1.640.81 6.981.13 44�1.9 58�12 77�6.5 100�17 120�42 74,000
Total heptachlorobiphenyis 0.780.50 1.140.12 17�3.8 20�7.6 27�10 31�9.9 45�25 58,000
Total octachlorobiphenyls 0.090.05 0.250.03 0.9�0.9 1.5�1.2 1.1�0.5 1.1�0.9 1.3�0.9 14,000
Total nonschlorobiphenyls 0.3�0.1 0.4�0.1 0.6�0.2 0.5�0.2 1.0�0.3
Total decachlorobiphenyl 0.3�0.2 0.2�0.1 0.3�0.2 0.3�0.2 0.9�1.0

Total PCBs 4.622.15 28.46.41 220�20 380�49 500�25 650�58 830�110 180,000

a Bioconcentration factor = concentration in transplanted oyster tissue at the end of the uptake pcriod/concentration in water.

TABLE A-9
Average PCB Congener Concentrations I S.D.) in Sediment and Hanna Reef-Ship Channel-Hanna Reef (HRSCHR) Transplanted
Oyster Sm, nples During the Depuration Phase of the Experiment at the Hanna Reef Site and Estimated Biological Half-Lives (BHL).

Depuration phase (days)
Analyte Sediment 0 3 6 18 30 50 BHL(R2)a
ng g-l ng g-l

18 6.0�1.0 6.5�1.9 3.2�1.3 1.3�0.3 0.7�0.3 0.7�0.2 14(0.82)
15/17 4.5�1.3 4.0�1.2 2.3�0.2 1.8�1.0 0.5�0.3 0.4�0.1
26 30�9.3 33�10 16�1.0 15�4.6 13�8.5 5.5�2.5 22(0.88)
50/31 9.8�4.6 4.9�0.9 3.4�1.0 2.1�1.1 3.1�1.3 3.4�1.2
28 10�1.4 9.3�1.2 7.2�1.7 3.4�0.7 2.3�0.3 1.3�0.3 17(0.96)
52 51�5.4 45�3.4 33�8.4 22�3.2 13�1.9 15�3.4 27(0.80)
49 29�1.5 30�2.8 23�4.6 18�3.5 13�1.1 13�2.8 39(0.84)
47/48fl5 18�2.2 18�1.8 14�2.2 9.4�2.1 7.1�1.7 6.6�1.8
44 22�2.3 22�2.2 16�3.2 9.7�2.3 7.2�1.0 6.7�2.2 27(0.83)
37/42159 42�8.3 25�8.1 13�4.3 5.5�1.9 5.1�0.2 5.4�1.9
41 20�1.4 20�4.2 13�4.7 7.3�1.9 5.3�1.4 5.1�0.7 23(0.83)
40 14�2.5 13�0.8 9.0�2.5 2.3�1.5 2.5�2.1 1.3�0.8 14(0.87)
74 13�2.7 14�2.2 10�3.3 6.9�1.4 5.0�0.8 4.7�1.1 30(0.87)
70 31�2.8 34�3.9 27�4.3 18�2.7 12�0.7 11�3.3 30(0.88)
95 55�5.3 49�4.4 45�6.9 31�5.2 25�0.4 26�5.8 45(0.81)
91 24�8.6 18�1.4 13�4.7 7.6�1.7 5.8�0.1 5.9�1.6 25(0.78)
60/56 10�1.9 11�0.8 8.3�2.0 5.2�0.9 3.7�1.0 3.3�1.0
92 22�8.0 22�2.4 17�2.0 11�2.1 9.4�1.2 7.8�2.5 31(0.89)

TABLE A-8
(confinued)

Depuration phase (days)
Analyte Sediment 0 3 6 18 30 50 BtIL(R2)a
ng g-1 ng g-1

84 29�0.9 24�7.3 2015.2 14�3.0 11�0.8 11�3.0 37(0.84)
101/90 43�5.2 47�7.6 43�8.0 34�2.3 26�5.6 26�3.0 55(0.86)
99 29�1.3 32�5.4 28�5.9 21�2.5 17�4.3 16�1.5 49(0.88)
97 14�3.2 16�2.2 14�3.6 10�1.3 9.4�0.8 10�2.2
87/115 20�1.4 21�2.6 18�3.9 12�1.8 11�1.4 12�2.0 55(0.73)
1 IOM 61�4.2 61�6.9 50�11 32�8.4 30�0.9 31�8.3 45(0.74)
82 5.9�1.0 6.5�1.2 5.1�2.1 6.2�2.4 4.0�0.1 4.6�0.8
151 6.0�1.0 7.5�0.6 6.8�2.3 5.0�1.0 5.0�1.0 4.9�0.5
135 5.0�0.5 5.7�0.7 4.9�1.2 3.3�0.7 3.8�0.4 3.7�0.5
107 6.6�1.0 8.3�3.1 5.7�1.8 3.5�1.1 2.7�0.8 2.6�0.5 30(0.82)
149/123 18�1.7 20�2.9 19�3.1 14�1.6 14�1.0 16�1.7 130(0.46)
118 34�2.5 35�4.5 32�7.9 25�2.5 32�18 23�1.5 73(0.79)
146 4.1�1.1 5.0�1.3 5.2�1.1 4.3�0.9 4.2�1.7 3.4�0.5 111(0.60)
153/132 64�12 59�16 48�13 37�9.4 33�7.6 34�3.5 51(0.71)
105 14�3.0 14�4.4 14�5.2 9.6�1.2 8.8�1.0 9.0�1.9 63(0.76)
141/179 4.7�0.5 4.1�1.3 3.0�0.9 2.1�0.7 2.1�0.4 2.0�0.3
138/160 28�4.7 30�3.8 33�6.0 25�4.2 24�3.8 26�1.9 200(0.32)
187 11�2.2 11�3.5 8.1�1.6 6.7�2.2 6.5�1.8 6.6�0.6 70(0.65)

TABLE A-8
(continued)

Depuration phase (days)
Analyte Sediment 0 3 6 18 30 50 Bt]L(R2)a
ngg-I ng g-l

128 3.9�0.6 4.10.9 4.1�1.0 2.9�0.5 2.7�0.2 2.8�0.5 76(0.75)
177 3.5�1.6 3.8�0.8 3.0�1.4 1.9�0.5 1.1�0.9 1.9�0.1 52(0.83)
180 3.8�1.7 3.9�0.6 3.7�0.5 2.7�0.-5 2.9�0.6 1.9�0.4 50(0.94)

Total dichlorobiphenyls 0.1�0.2
Total trichlowbiphenyls 66�9.3 73�13 29�10 22�10 20�9.4 13�2.9
Total teirachlorobiphenyis 240�24 230�17 170�38 100�17 75�5.0 74�17
Total pentachlorobiphenyls 360�26 360�28 310�76 220�25 200�32 190�30
Total hexachlorobiphenyls 120�42 140�25 130�19 97�19 91�14 95�4.7
Total heptachlowbiphenyls 45�25 41�13 32�9.7 19�8.1 16�2.8 16�1.6
Total octachlorobiphenyls 1.3�0.9 1.0�0.2 2.5�1.7 1.2�1.5 1.7�2.7 0.4�0.5
Total nonachlorobiphenyls 1.0�0.3 0.7�0.2 0.9�1.2 0.4�0.5 0.2�0.2 0.4�0.1
Total decachlorobiphenyl 0.9�1.0 0.5�0.4 0.5�0.9 0.1�0.1 0.2�0.1

Total PCBs 830�110 850�82 670�120 470�76 400�59 380�50

a R2 = square of correlation coefficient for regression equation.

TABLE A-9
Average PCB Congener Concentrations (� I S.D.) in Sediment and Ship Channel -to-Hanna Reef (SCHR) Transplanted Oyster Samples
During the Depuration Phase of the Experiment at the Hanna Reef Site and Estimated Biological Half-Lives (BHL).

Depuration phase (days)
Analyte Sediment 0 3 6 18 30 50 BHL(R2r
ng g-I ng 9-1

18 6.1�0.6 6.5�2.0 3.8�1.0 2.3�0.6 1.7�1.0 1.1�0.2 19(0.93)
15/17 4.8�0.5 4.8�0.7 4.0�0.6 2.4�1.6 1.2�0.4 0.9�0.2
26 36�8.0 37�4.6 28�11 23�10 14�10 7.1�1.5 22(0.99)
50/31 5.2�0.8 7.9�1.3 3.20.4 6.7�3.1 7.6�1.7 5.5�1.1
28 8.0�1.4 6.5�0.9 8.5�0.6 5.3�0.6 4.2�1.0 2.9fO.3 34(0.93)
52 52�4.0 46�7.0 50�6.1 32�1.6 30�9.7 24�3.5 45(0.91)
49 31�3.2 29�4.6 32�5.4 27�2.4 24�7.6 17�1.0 61(0.94)
47/48M 19�2.0 17�1.9 19�1.8 14�1.4 13�3.9 12�2.0
44 20�3.2 20�2.2 22�2.5 15�0.9 14�5.0 9.7�0.8 45(0.94)
37/42/59 33�5.3 23�6.4 18�3.0 9.5�0.9 8.1�2.8 6.6�0.5
41 16�2.8 16:0.2 16�3.8 11�1.2 11�2.9 9.2�1.5 55(0.92)
40 13�0.5 11�1.6 10�2.3 5.7�1.4 3.2�0.9 2.1�0.2 18(0.97)
74 15�0.9 14�1.3 15�0.9 11�0.7 8.5�2.2 7.4�0.7 47(0.95)
70 38�2.0 33�4.6 36�3.2 28�2.9 26�7.6 20�2.4 58(0.96)
95 63�8.2 59�8.4 70�8.9 54�4.8 55�18 44�0.3 95(0.79)
91 22�1.0 18�2.7 20�2.4 13�1.7 13�3.9 11�0.8 50(0.89)
60/56 9.5�1.4 10�1.4 11�1.4 7.4�0.5 7.0�2.6 5.1�0.3
92 26�6.6 26�4.8 25�5.5 20�2.1 20�5.9 15�03 63(0.93)

TABLE A-9
(continued)

Depuration phase (days)
Analyte Sediment 0 3 6 18 30 50 Bt]L(R2)a
ng g-I ng g-I

84 29�7.3 31�3.6 33�2.7 23�3.1 23�8.1 21�1.4 80(0.79)
101/90 71�8.5 69�7.4 80�7.1 64�3.8 62�12 54�3.0 91(0.79)
99 45�4.6 44�4.7 52�0.6 39�4.1 3913.3 32�2.2
97 20�3.7 20�2.3 25�1.9 21�2.3 21�6.4 21�2.7
87/115 32�3.1 28�3.9 34�3.5 27�2.6 27�7.6 26�2.6 132(0.45)
1 IOM 80�12 76�9.5 88�10 67�8.2 58�8.8 62�4.7 103(0.67)
82 6.3�2.3 8.5�0.6 12�1.1 9.8�1.7 7.6�2.1 8.2�2.0
151 10�2.2 10�1.0 13�1.5 11�0.6 11�1.7 10�1.3
135 8.7�2.1 7.3�1.1 8.9�1.7 7.5�0.7 7.4�1.4 7.1�1.0
107 15�3.1 11�4.1 10�1.1 7.3�1.0 6.6�0.7 6.3�1.4 46(0.75)
149/123 36�3.5 31�3.5 36�5.1 32�2.7 33�6.9 31�2.0 439(0.24)
118 56�6.3 55�5.8 68�7.4 65�18 55�5.8 52�5.3 299(0.19)
146 7.8�2.1 8.0�1.3 10�1-9 8.2�1.2 7.9�1.9 7.3�0.9 239(0.27)
153/132 100�10 95�24 98�27 80�10 77�10 72�5.5 102(0.90)
105 17�6.1 17�4.7 23�2.1 20�2.8 19�6.4 17�2.5 120(0.76)
141/179 5.2�0.9 5.8�1.1 5.4�1.2 4.7�0.9 4.6�0.5 4.6�0.9
138/160 47�10 48�4.0 56�6.9 49�4.1 48�9.7 47�1.8 595(0.11)
187 15�2.5 14�1.3 13�3.2 12�2.6 13�1.8 13�2.1 258(0.56)

TABLE A-9
(continued)

Depuration phase (days)
Analyte Sedimcnt 0 3 6 18 30 50 BIIL(R2)a
ng 9-1 ng g-l

128 5.9�1.0 5.9�0.4 6.9�0.8 1.2�0.2 5.5�2.0 6.4�1.3 229(0.42)
177 4.6�0.1 3.6�0.3 4.4�0.8 3.6�0.2 3.7�0.6 3.4�0.4 145(0.54)
180 4.6�0.2 5.2�0.8 5.1�0.7 4.4�1.4 3.8�0.5 4.1�0.9 142(0.63)

Total dichlorobiphenyis 0.0 0.6�1.1 0.0 0.4�0.7 0.3�0.3 0.0
Total trichlorobiphenyis 71�12 67�10 53�5.1 40�12 31�12 19�3.6
Total tetrachlorobiphenyls 250+71 220�29 230�26 160�9.4 150�43 120�11
Total pentachlorobiphenyis 490�64 470�42 550�38 430�41 390�59 370�30
Total hexachlorobiphenyis 230�31 220�32 240�45 200�16 200�30 190�16
Total heptachlorobiphenyls 52�5.9 47�1.5 46�6.3 33�4.4 31�6.3 27�3.0
Total octachlorobiphenyls 1.9�0.2 1.3�0.9 2.8�2.0 0.9�0.9 0.2�0.3 0.3�0.3
Total nonachlorobiphenyls 0.8�0.3 1.4�1.0 1.6�0.7 0.7�0.4 0.5�0.2 0.6�0.2
Total decachlombiphenyl 0.4�0.3 0.9�0.7 1.2�0.3 0.3�0.3 0.1�0.1 0.1�0.1

Total PCBs 1100�130 1000�110 1100�110 870�64 800�140 730�65

a R2 = square of correlation coefficient for regression equation.

231

TABLE A-10
TBT, DBT and MBT Concentrations in Indigenous Ship Channel and Transplanted Hanna
Reef Oysters.

Sample Days after TBT DBT MEBT
transplant (ng Sn g- 1)

Hanna Reef-to-Ship Channel (HRSC)
HRSC 0 40 13 9

HRSC 3 68 <5 <5
HRSC 7 130 10 <5
HRSC 17 210 <5 <5

HRSC 30 230 6 <5

HRSC 48 360 22 <5

Hanna Reef-Ship Channel-Hanna Reef (HRSCHR)
HRSCHR 51 330 <5 <5
HRSCHR 54 290 21 <5
HRSCHR 66 180 <5 <5

HRSCHR 78 130 6 <5

FIRSCHR 98 110 <5 <5

Ship Channel (SC)
SC 3 350 24 <5
SC 7 Sample was not collected
SC 17 310 22 <5
SC 30 320 32 <5
SC 48 390 34 <5

Ship Ci,annel-to-Hanna Reef (SCHR)
SCHR 51 320 31 <5
SCHR 54 340 62 <5
SCHR 66 240 24 <5
SCHR .78 220 16 <5
SCHR 98 130 10 <5

TABLE A-11
PCB Congener Concentrations in Oysters During Exposure in the Laboratory to a 1: 1: 1: 1 Mixture of Aroclors 1242, 1248, 1254 and 1260
and Following Deputation in Contarninant-Free Aquariurns.

Uptake phase (days) Depuration phase (days)
Analyte 0 3 7 15 30 3 7 15 30 BHL (112)a
ngg-1 ng g-I

18 1.36 1.93 3.58 6.46 7.78 6.42 5.77 2.50 28(0.87)
15/17 0.21 0.54 0.72 1.55 1.61 1.44 1.43 0.66
50/31 1.02 1.70 3.03 5.54 6.04 4.44 5.34 2.80
28 1.60 2.93 3.90 8.78 10.6 8.49 9.04 3.90 33(0.78)
52 2.16 3.22 5.56 12.5 12.9 11.4 10.3 5.17 28(0.96)
49 1.38 2.42 3.39 8.63 9.03 7.72 7.91 4.04 38(0.86)
47/48fl5 0.90 1.23 2.85 7.93
44 1.26 2.03 3.33 8.27 9.10 7.66 7.56 3.65 34(0.87)
37/42/59 0.87 2.05 3.07 5.64
41 1.28 2.52 4.46 15.1 14.7 15.4 12.5 4.74 25(0.88)
40 0.48 0.62 1.09 2.72 3.26 2.59 2.45 1.05 28(0.86)
74 1.40 1.96 3.82 10.3 10.3 9.29 9.10 6.18 57(0.93)
70 3.17 4.60 8.29 20.4 21.4 19.5 18.8 112.6 58(0.90)
91 0.18 0.21 0.51 1.31 1.43 1.51 1.15 1.01 83(0.73)
60/56 0.97 1.68 2.71 6.93 7.51 7.56 6.06 4.40
92 0.17 0.16 0.43 1.35 1.53 1.87 1.11 1.18 99(0.31)

TABLE A-11

Uptake phase (days) Depuration phase (days)
Analyte 0 3 7 15 30 3 7 15 30 BHL (R2)a
n99-I ng g-I

84 0.45 0.62 0.97 2.55 2.81 3.04 2.25 1.85 66(0.75)
101/190 1.92 2.29 4.41 11.2 11.4 11.5 - 9.10 8.55 90(0.83)
99 1.46 1.42 2.08 5.15 4.61 4.26 4.01 3.73 101 (0.83)
97 0.70 0.87 1.86 4.31 4.19 3.90 3.86 3.08
87/115 1.40 1.74 3.19 8.74 9.04 7.39 7.18 5.21
110M 1.92 2.79 4.85 16.0 13.7 12.6 13.6 10.2 80(0.78)
82 1.38 1.33 1.84 2.61 2.44 2.65 1.86 2.25
151 0.41 0.45 1.05 3.00 2.59 2.68 2.38 1.96
135 0.34 0.40 0.73 2.27 1.76 1.78 1.54 1.60
149/123 1.45 1.88 3.14 7.49 7.11 7.20 7.90 7.05
118 2.21 1.86 4.35 10.1 9.74 11.5 10.1 7.71 103(0.58)
146 0.41 0.50 0.78 1.39 1.55 1.34 1.58
153/132 2.61 3.23 5.55 17.0 13.5 11.9 11.9 11.1 90(0.59)
105 0.42 0.46 1.56 3.45 3.51 4.02 2.81 3.38
141/179 0.47 0.76 0.74 1.82 1.89 1.51 1.33 1.42
138/160 2.90 4.08 7.24 16.2 14.6 13.6 11.0 10.1 63(0.86)

TABLE A-11
(continued)

Uptake phase (days) Depuration phase (days)
Analyte 0 3 7 15 30 3 7 15 30 BHL (R2)a
ng g-I ng g-l

187 1.05 1.48 2.20 5.75 4.59 4.20 4.37 4.34
128 0.30 0.45 0.85 2.01 1.45 1.24 1.05 0.98 48(0.72)
177 0.35 0.50 0.87 2.36 1.81 1.77 1.62 1.80
180 0.37 0.40 0.35 0.40 0.18 0.10 0.16 0.08

Total dichlorobiphenyls
Total trichlorobiphenyls 4.93 8.21 12.7 24.8 29.7 24.9 24.2 11.7
Total tetrachlorobiphenyls 14.9 24.4 41.6 105 95.5 87.7 82.5 47.6
Total pentachlorobiphenyls 13.1 14.8 27.6 67.6 64.6 64.7 61.0 49.4
Total hexachlorobiphenyls 9.05 12.0 20.5 51.9 44.8 41.6 36.8 34.6
Total heptachlorobiphenyls 2.15 2.82 3.91 9.74 7.46 6.58 6.62 6.62
Total octachlorobiphenyls 0.09 0.12 0.18 0.20 0.32 0.24 0.15 0.35
Total nonachlombiphenyis
Total decachlombiphenyl

Total PCBs 44.1 62.3 106 259 242 226 211 150

a R2 = square of correlation coefficient for mgression equation.

TABLE A-12
PCB Congener CDncentrations in Oysteis During Exposure in the Laboratory to a Mixture, of PCBs Plus PAHs and Following Depuration in
Contaminant-Free Aquariums.

Uptake phase (days) Depuration phase (days)
Analyte 0 3 7 15 30 3 7 15 30 BHL (112)a
ng g-I ng g-l

18 2.25 2.10 6.79 10.5 45(0.82)
15/17 0.16 0.52 1.23 2.16 1.90 2.35 1.78 1.21
50/31 0.66 1.53 3.49 7.35 6.32 7.75 5.53 4.22
28 1.27 2.15 5.57 13.5 11.3 12.0 10.2 7.41 52(0.94)
52 2.44 2.88 7.13 17.2 16.9 15.8 14.6 11.9 78(1.00)
49 1.05 2.14 4.87 12.7 10.3 10.1 9.57 8.59 93(0.77)
44 1.02 1.80 4.65 11.6 10.4 10.6 9.04 6.96 59(0.98)
41 3.18 2.96 7.95 23.0 21.8 18.6 15.6 11.9 44(0.98)
40 0.64 0.55 1.34 3.73 3.70 3.44 2.85 2.17 51 (0.99)
74 2.12 2.65 5.99 16.1 13.5 13.8 12.2 9.50 61 (0.94)
70 2.14 4.47 11.4 29.1 26.1 25.6 22.9 19.1 74(0.97)
91 0.34 0.43 0.77 1.75 1.70 1.65 1.51 1.52 143(0.77)
60/56 0.83 1.72 4.06 10.4 11.3 9.70 8.66 7.21
92 0.20 0.45 0.97 2.25 2.01 1.98 2.01 1.67 119(0.84)
84 0.37 0.59 1.49 3.78 3.76 3.36 2.54 72(0.96)
101/90 1.30 2.60 6.37 14.4 16.5 13.2 13.7 12.5 155(0.50)

TABLE A-12
(continued)

Uptake phase (days) Depuration phase (days)
Analyte 0 3 7 15 30 3 7 15 30 BHL (R2)a
ngg-I ng g-1

99 0.48 1.00 2.18 6.57 5.43 4.80 4.90 4.74 120(0.50)
97 0.33 0.76 2.12 5.88 4.51 4.27 4.37 3.94
87/115 0.57 1.47 3.88 11.2 9.61 8.77 7.51 7.22 72(0.79)
1 10M 0.61 2.35 6.82 20.7 17.7 17.1 15.7 15.6 125(0.65)
82 1.46 1.60 2.05 3.13 2.32 3.87 2.57 2.01
151 0.29 0.44 1.34 3.71 2.94 2.50 2.71 2.44
135 0.30 0.42 0.90 2.42 1.92 1.76 1.61 1.61
149/123 0.86 1.76 4.08 9.96 9.76 7.90 6.89 7.72
118 1.31 1.95 5.33 11.8 10.4 9.35 10.7 9.79 272(0.23)
146 0.33 0.47 0.99 1.21 1.70 1.65 1.75 1.45
153/132 1.41 3.08 7.69 20.7 13.5 12.9 12.7 13.1 103(0.30)
105 0.30 0.84 2.12 4.26 5.60 3.77 3.81 4.24
141/179 0.17 0.35 0.73 1.77 1.59 1.34 1.43 1.22
138/160 1.96 3.52 7.93 16.5 13.9 11.6 11.5 11.0 86(0.62)
187 0.73 1.17 2.57 6.16 4.50 4.23 4.29 3.88 89(0.53)
128 0.21 0.35 0.82 1.82 1.25 1.10 0.95 0.82 44(0.76)

TABLE A-12
(continued)

Uptake phase (days) Depuration phase (days)
Analyte 0 3 7 15 30 3 7 15 30 BHL (112)a
ngg-I I ng g-I

177 0.18 0.43 0.95 2.40 1.82 1.49 1.54 1.56 95(0.38)
180 0.33 0.24 0.51 0.37 0.19 0.09 0.11 0.18

Total dichlorobiphenyls
Total trichlorobiphenyls 5.27 7.58 19.4 37.4 32.5 38.3 28.6 21.0
Total tetrachlorobiphenyls 14.4 20.9 51.7 135 125 118 106 85.2
Total pentachlorobiphenyls 7.44 14.8 37.4 89.4 83.8 74.5 73.7 68.6
Total hexachlorobiphenyls 5.77 10.7 25.1 58.8 47.7 41.9 40,8 40.4
Total hCptachlorobiphenyls 1.35 2.04 4.40 9.70 7.13 6.19 6.27 5.89
Total octachlorobiphenyls 0.09 0.09 0.14 0.12 0.16 0.27 0.11
Total nonachlorobiphenyls
Total decachlorobiphenyl

Total PCBs 34.3 56.1 138 330 296 279 255 221

a R2 = square of correlation coefficient for regression equation.

TABLE A-13
Pblynuclear Aromatic Hydrocarbon Concentrations in Oysters During Exposure in the Laboratory to a Mixture of Selected PAHs and
Following Depuration in Contaminant-Free Aquariums.

Uptake phase (days) Depuration phase (days)
Analyte 0 3 7 15 30 3 7 15 30 BHL (112)a
ngg-I ng g-I

NaphthWene 22 26 13 9.4 10 9.8 16 8.6 8.6
2-Methyinaphthalene 16 31 21 9.0 8.7 5.9 9.5 5.2 6.8
I-Methyinaphthalene 9.0 27 16 6.2 7.7 5.3 9.7 5.0 5.0
Biphenyl 8.3 64 56 37 16 10 11 7.3 6.4
2,6-Dimethyinaphibalene 15 57 50 35 13 6.9 12 7.2 7.4
Acenaphthylene 7.3 30 27 24 22 12 13 9.1 5.7
Acenaphthene 4.0 48 48 48 65 9.1 8.8 8.3 5.3
2,3,5-Trimethyinaphthalene 26 69 73 61 51 21 23 15 11 16(0.75)
Fluorene 9.3 54 57 48 44 9.1 12 7.9 7.0
Phenanthrene 14 69 64 46 49 25 31 28 22
Anthracene 16 33 42 50 70 24 29 19 13 16(0.68)
I-Methylphenanthrene 62 119 115 92 86 37 60 22 19 16(0.72)
Fluoranthene 11 84 91 76 144 44 68 15 11 9(0.78)
Pyrene 16 86 82 88 83 31 47 8.7 8.4 9(0.75)
Bm(a)anthracene 8.3 85 67 154 305 279 228 94 93 16(0.81)
Chryscne 7.7 52 55 112 286 268 276 144 124 22(0.86)
Bmzo(b)fluoranthene/
Benzo(k)fluoranthene 3.3 183 153 300 7" 705 742 570 330 25(0.94)

TABLE A-13
(continued)

Uptake phase (days) Dcpuradon phase (days)
Analpe 0 3 7 15 30 3 7 15 30 BHL (112)a
ng g- I ng g-l

Benzo(e)pyrene 4.0 104 92 154 430 436 446 262 174 21(0.93)
Benzo(a)pyrene 6.3 29 16 35 97 61 82 25 17 12(0.87)
Perylene 3.0 8.9 12 22 47 42 45 27 13 15(0.96)
Indeno[ 1,2,3-c,dlpytene 10 73 47 102 200 171 137 71 28 10(1.00)
Dibenzo(a,h)anthracene 3.0 19 9.6 26 45 36 24 13 6.5 11(0.97)
Benzo(g,h,i)peryiene 7.0 36 27 65 144 161 133 94 43 16(0.96)

Total 2-Rings 97 106 44
Total 3-Rings 113 337 72
Total 4-Rings 43 818 237
Total 5-Rings 23 1370 541
Total 6-Rings 17 344 71

Total PAHs 293 2970 965

a R2 square of correlation coefficient for regression equation.

TABLE A-14
Polynucl.ear Aromatic Hydrocarbon Concentrations in Oysters During Exposure in the Laboratory to a Mixture of Selected PAHs and
Following Depuration in Contaminant-Free Aquariums.

Uptake phase (days) Depuration phase (days)
Analyte* 0 3 7 15 30 3 7 15 30 BHL (R2)a
ng g-l ng g-I

Naphthalene 22 21 10 9.3 7.9 12 16 17 13
2-Methyinaphthalene 16 36 16 10 9.2 9.9 16 14 9.1
1 -Methyinaphthalene 9.0 29 12 7.7 5.5 5.8 9.2 9.5 6.8
Biphenyl 8.3 94 49 40 19 9.8 7.9 6.7 5.0
2,6-Dimethylnaphthalene 15 81 39 28 27 12 12 13 5.8
Accnaphthyleric 7.3 50 27 36 37 18 11 12 5.1
Accnaphthenc 4.0 74 54 87 49 8.3 7.4 4.8 3.9
2,3,5-Trimethylnaphthalenc 26 99 68 56 78 34 28 14 6.9 100.91)
Fluorene 9.3 90 43 62 51 11 12 8.6 5.5
Phenanthrene 14 113 77 56 60 38 37 24 21
Anthracene 16 49 64 81 118 55 45 14 8.9 8(0.89)
I-Methylphenanthrene 62 125 123 114 144 78 73 12 8.2 7(0.88)
Fluoranthene 11 84 118 162 194 79 77 11 7.8 7(0.85)
Pyrene 16 81 120 138 150 74 63 9.4 5.4 6(0.89)
Benz(a)anthracene 33 85 169 519 342 238 127 46 9(0.99)
Chrysene 20 59 120 357 304 198 155 62 12(0.98)
Benzo(b)fluoranthene/
Benzo(k)fluoranthene 58 144 317 859 662 521 474 148 13(0.95)

TABLE A-14
(continued)

Uptake phase (days) Dcpuration phase (days)
Analyte 0 3 7 15 30 3 7 15 30 BHL (112)a
ng 9-1 ng g-l

Benzo(e)pyrene 31 88 170 491 452 303 227 120 15(0.98)
Bcnzo(a)pyrene 12 27 57 202 60 37 48 12 9(0.78)
Perylene 5.2 13 21 65 40 18 20 5.8 10(0.89)
Indeno[ 1,2,3-c,dlpyrcnc 14 42 90 329 113 47 81 14 8(0.78)
Dibenzo(a.h)anthracene 5.6 16 26 72 24 9.4 17 4.5 10(0.69)
Benzo(g,h,i)peryiene 7.4 21 47 136 85 51 50 17 11 (0.92)

Total 2-Rings 97 146 45
Total 3-Rings 113 459 53
Total 4-Rings 43 1,220 120
Total 5-Rings, 23 1,690 290
Total 6-Rings 17 466 31

Total PAHs 293 3,980 540

a R2 = square of correlation coefficient for regression equation.

242

VITA

Josd Luis Sericano was born in Puerto Belgrano, Buenos Aires, Repdblica

Argentina, on October 10, 1953. He is the son of Vicente Luis and Margarita Sericano.

He attended public schools and graduated from Colegio Nacional Punta Alta, Punta

Alta, Rep6blica Argentina, in December 197 1. He attended Universidad Nacional del

Sur, Bahia Blanca, Repdblica Argentina, and graduated as Quimico, Lic. en Bioquft:nica

and Lic. en Qufmica in August 1975, December 1976 and August 1977, respectively.

He enrolled in the Graduate College at Texas A&M University in August 1983 and

received a M.S. in Oceanography in May, 1986. In February 1981, he married Ndlida

Maria Cavallfn and their family consists of one son, Mauro Luis, born in January 1982

and one daughter, Gisella Maria, born in June 1987.

His permanent mailing address is Murature 68, (8109) Punta Alta, Buenos Aires,

Repdblica Argentina.

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