- Email: [email protected]
EROD activity in fish as an independent measure of contaminant-induced mortality of invertebrates in sediment bioassays
Marine Environmental Research, Vol. 44, No. I, pp. 4149,
PII:
SOl41-1136(96)00101-g
1997 0 1997 Elsevier !kience Ltd All rights reserved. Printed in Great Britain 0141-1136/97 %17.00+0.00
EROD Activity in Fish as an Independent Measure of Contaminant-Induced Mortality of Invertebrates in Sediment Bioassays Andrew J. Gunther,a Robert B. Spies,” John Stegeman,b Bruce Woodiqb Diane Carney,c James Oakden” & Leslie Hai& “Applied Marine Sciences, 2155 Las Positas Ct. Suite S, Livermore, California 94550, USA bWoods Hole Oceanographic Institution, Woods Hole, Massachusetts 02543, USA CM~~~Landing Marine Laboratory, P.O. Box 450, Moss Landing, California 95039, USA dBodega Marine Laboratory, P.O. Box 247, Bodega Bay, California 94923, USA (Received 22 April 1995; revised version received 1 September 1996; accepted 7 September 1996)
ABSTRACT The survival of amphipods in sediment bioassays was compared to the induction of ethoxyresorufin-0-deethylase (EROD) activity in speckled sanddabs (Citharichthys stigmaeus) to investigate the relative importance of physical and chemical characteristics of sediment in toxicity bioassays. Sediments from San Francisco Bay containing l-30 ppm (dry) of polynuclear aromatic hydrocarbons (PAHs), 3-20 ppb (dry) of polychlorinated biphenyls (PCBs) and variable concentrations of several trace elements were used in separate exposure assays for sanddabs (60 d) and the amphipod Eohaustorius estuarius (10 d). A highly signt&ant correlation (13 = 0.90, p < .OOl) was documented between EROD activity in the sanddabs and Eohaustorius survival for all treatments. The accurate prediction of amphipod survival by the contaminant exposure biomarker in the Jish is consistent with the contention that contamination, rather than physical characteristics of the sediments, is the cause of amphipod mortality. This result supports the use of these bioassays to screen sediments for ecologically significant contaminant concentrations. 0 1997 Elsevier Science Ltd. All rights reserved
INTRODUCTION The identification of contaminated sediments harmful to marine life is a challenging aspect of marine environmental management. The most widely used tools for identification of potentially harmful sediments have been sediment bioassays employing a variety of invertebrate organisms and sediment exposure conditions (Baud0 et al., 1990; Burton, 1992; Burton & Scott, 1992; Dewitt et al., 1989; National Research Council, 1989). While
the endpoints of the tests have been considered as evidence of contaminant 41
effects, and it
42
A. J. Guntheret al.
is clear that contamination can affect survival of benthic organisms, other factors such as the physical characteristics of sediments can also influence the survival of test organisms (Oakden, 1984; Dewitt et al., 1988). This has been particularly true for assays employing amphipods in which exposure to sediment types outside the range of the natural habitat is common. It would be useful to have an independent indication whether mortality in the test population was related to the degree of contamination rather than a result of other characteristics of the sediments. Biochemical, genetic and histological changes due to contaminant exposure could provide such corroborative evidence. Such biomarkers of contaminant exposure and effect are not well developed for invertebrates, however, so routine or widespread use of the biomarker approach with invertebrates must await further studies. Small fishes have occasionally been used in sediment bioassays and in the field (Spies, unpubl.), and a better understanding of contaminant biomarkers exists for fish than for invertebrates (Benson & DiGuilo, 1992). Much of the reported work with fish comes from either field collections in areas of known contamination or from shortterm laboratory exposures to contaminants (Sulaiman & Burke, et al., 1991; Stein et al., 1992; Wirgin et al., 1994). The purpose of this study was to use a biomarker response in a demersal fish as an independent indication of contaminant bioavailability for comparison with the results of a standard amphipod bioassay. Speckled sanddabs (Citharichthys stigmaeus) were exposed for two months and the estuarine amphipod Eohaustorius estuarius for 10 days to contaminated sediments collected along a contamination gradient near an old refinery outfall in San Francisco Bay. The grain size and contaminant concentrations of the sediments varied between samples, but not in the same manner. Ethoxyresorufin-0-deethylase (EROD) activity in the liver of sanddabs was compared to amphipod survival in each sediment. If physical characteristics of the sediment have a predominant influence on amphipod survival, a poor correlation between the EROD activity in fish and amphipod survival for sediments might be expected. If, however, amphipod survival is being controlled predominantly by contamination, then there should be a strong correlation between amphipod survival in a sediment and EROD activity in fish exposed to that same sediment.
METHODS Sediments
Sediments for this experiment were collected from five stations in Castro Cove and San Francisco Bay (Fig. 1) and an uncontaminated reference area (Carr Inlet, WA) during the week 25-29 May 1992. The selected stations were expected to contain a wide range of contaminant concentrations as portions of Castro Cove sediments are known to contain high concentrations of hydrocarbons in the sediments from past discharges of oil refinery effluent. At two of the stations both shallow sediments (top 2 cm) and sediments from ‘full’ cores (to 30 cm) were used, resulting in a total of seven treatments. The percent sand fraction was estimated using sieve analysis (Folk, 1974).
EROD activity inJish
43
44
A. J. Gunther et al.
Organic chemistry
A detailed description of the chemical methods may be found in Risebrough (1994). Briefly, freeze-dried sediments (l&20 g) were mixed with kiln-fired sodium sulfate, soxhlet extracted with methylene chloride, and separated from lipids using florisil-column chromatography. Extract volumes were concentrated to 14 ml, and analyzed by both electron-capture gas chromatography (Varian 3400 GC with 8100 autosampler) and gas chromatography/mass spectroscopy (Varian Saturn II with 8 100 autosampler). DB5 30 m columns (J&W, Scientific, Inc., Folson, CA, USA) were used in both instruments, facilitating confirmation of GC/ECD indentifications by GC/MS. For quality control, internal standards (deuterated phenanthrene, deuterated chrysene and decachlorobiphenyl) were added to the samples prior to extraction to determine recoveries. System blanks (including freeze-dried flourisil), reference sediments (HS-I and HS-6, provided by the National Research Council of Canada), and National Institute of Standards & Technology, NIST Standard Reference Material 1974 (mussel tissue) were also analyzed. CPCB was calculated by summing the concentrations of the following congeners: 005/8, 015, 018, 027124, 028, 029, 031, 040,044,049, 052, 060156, 066195, 070, 074, 087/l 15, 097, 099, 101/90, 105/132, 110/77, 114/131/122, 118, 128, 129/178, 132, 137/176, 138, 141/179, 146, 149, 151/82, 153, 156/171/202, 157/173/201, 158, 170/190, 174, 177, 180, 183, 185, 187, 189, 191, 194, 195/208, 1961203, 199, 205, 206, 207. CPAH was calculated by summing the concentrations of the following compounds: anthracene, benzo[a]anthracene, benzo[a]pyrene, benzo[b]fluoranthene, benzo[e]pyrene, benzo[ghi]perylene, benzo[k]fluoranthene, chrysene, dibenzo[a,h]anthracene, fluoranthene, indeno[l,2,3_cd]pyrene, methylanthracene, phenanthrene, pyrene, all single-substituted methyl-phenanthrenes. Trace metals
A detailed description of the chemical methods may be found in Flegal et al. (1981) (see also Rivera-Duarte & Flegal, 1991). Briefly, homogenized sediment was dried overnight (7O”Q and 1 g of dried sediment was combined with 10 ml aqua regia (3:1 HCl:HN03). The mixture was allowed to stand overnight at room temperature (with a loose cap to let gasses escape), refluxed 4 h (2Oo”Q cooled and then dried on a hot plate. Ten ml 1 N HN03 was added, the sample mixed and allowed to redissolve overnight. The resulting solution was filtered (0.45 pm) and analyzed by graphite furnace atomic absorption spectrometry, flame atomic absorption spectrometry and inductively coupled plasma emission spectrometry. Procedural blanks were analyzed to quantify elemental contamination and extraction efficiency was quantified using Environment Canada standard reference sediments (MESS, BCSS). Fish collection,
incubation and dissection
Speckled sanddabs (Citharichthys stigmaeus) were collected by otter trawl in Tomales Bay (30 km WNW of San Francisco Bay) in the ten days before the start of the experiment, held in a large outdoor aquarium and then transferred to the Bodega Marine Laboratory the day prior to the beginning of the experiment. Since it was possible that the marine laboratory environment itself might have an influence on EROD activity, a small additional sample of fish was collected from the mouth of Tomales Bay during 2-5 June
EROD activity infish
45
1992 as a pre-experimental reference. The available fish ranged in size up to 9 cm Standard Length (SL). Fish were sorted to four size classes (5-6,6-7,7-8 and > 8 cm SL), distributed as evenly as possible to the test aquaria (7 fish in each) and fed a ration of chopped frozen squid at a rate of 5% body weight per day. Uneaten food was routinely removed from the aquaria. The fish collected from Tomales Bay were dissected on 7 June 1992 as pre-experimental controls. The remainder of the fish were sacrificed after two months’ exposure on 30 July to 1 August 1992. (While it is likely, based upon previous studies, that EROD activity was induced within 24 h of exposure to the more contaminated sediments, a 60 d exposure was utilized due to simultaneous histopathological measurements being made on these fish [Spies et al., in prep.].) Each fish was blotted dry, weighed and the standard length determined to the nearest millimeter. The liver was removed, being carefully separated from the gall bladder, and weighed whole. A small cross-section of liver was taken, usually normal to the longest axis, and preserved in 10% neutral formalin for later analysis for histopathological abnormalities (see Spies et al., in prep.). The remaining liver was weighed and placed on ice in phosphate buffered saline (pH 7.2) for several minutes to await further processing for an analysis of EROD activity. Sediment bioassays The toxicity of sediments was evaluated using the estuarine amphipod Eohaustorius estuarius. Studies were conducted at the Moss Landing Marine Laboratory (Moss Landing, CA) using standard methods (ASTM, 1992). All treatments had five laboratory replicates and 20 individuals per replicate. Two negative controls, a reference (Marconi Cove in Tomales Bay, CA, or Carr Inlet, WA) and home sediment (Monterey Bay near Moss Landing, CA), were run with each toxicity test. Acceptability of the test was defined by 290% survivorship in the home sediment. Determination of EROD activity From 0.20 to 0.70 g of liver from each fish was homogenized in phosphate buffered saline within one hour of collection using a hand tissue grinder. The homogenates were centrifuged for 15 min at 13,000 g at 1°C and the supernatant was retained and stored at -75°C until it could be assayed. The EROD activity of the supernatant was evaluated either spectrophotometrically according to the methods of Klotz et al. (1984) or fluorometrically by modifications to the method of Eggens and Galgani (1992) using the Millipore Cytofluor fluorescent plate reader. The hepatic EROD activity is reported as activity per gram of fish for each individual, rather than per gram of sample protein, as samples were too small to prepare a microsomal fraction and thus tests were run on the post-mitochondrial supernatant (PMS). In such cases, normalizing data to liver or body weight can be as or more accurate than normalizing to PMS protein.
RESULTS The survival of amphipods in all sediments except for those from the station furthest away from the refinery (PPP; Fig. 1) was significantly lower than in the reference and the home
*
0.44
95
se = standard error. *Indicates significant difference NA = Not analysed. dw = dry weight. Organic contaminants reported
19
*
NA
NA
NA
1994) trace elements
site.
NA
reported
NA
NA
NA
210 75 163 124 87 55
144
Zn
NA
NA
86 195 29 59 43 26 25
Cu
91 101 39 93 79 48 30
Ni
NA
NA
NA
33 63 21 63 28 25 13
Ph
Induction
NA
EROD
NA
NA
0.47 1.24 0.33 0.65 0.20 0.31 0.59
Cd
NA
NA
0.35 0.42 0.12 0.30 0.27 0.20 0.17
Ag
of Sanddabs
as pg/g dw (data from Flegal et al., 1994).
NA
NA
NA
NA
NA
NA
8.133 1.431 2.493 4.313 2.763 0.469
NA
80 102 53 87 78 59 41
6.98
3.212
Cr
2.26 0.793 3.265 0.947 0.554 0.243
C DDTs
10,397 19.902
C CChlordanes PCB s
1.98 30,461 20.885 0.73 695 9.325 1.3 1,171 12.819 1.12 1,106 14.864 0.81 145 12.74 1 783 3.447
1.42
org. C. C PAH (%)
and Carr Inlet control
94
NA
0 14 9 14 45 7
1
sand (%)
as rig/g dw (data from Risebrough,
between
* *
treatment
NA
Bay
0.16
0.17
* * * *
Bay Monterey
0 70 15 70 85 95
41.7 16.7 24.8 13.6 11.1 8.14
0.89 1.34 0.89 0.89 0.44
0 EVS04 full 14 cc2 full 3 cc4 full 14 PPP top 17 PPP full 19 Carr Inlet top Tomales NA
se 7.12
mean 37.7
se 0.44
% 10
mean 2
survival
EROD Acfivify (amok fish)
TABLE 1 Survival of Amphipod Eohaustorius estuarius in 10 d Bioassays, and Hepatic after 60 d Exposures, for Sediments from San Francisco Bay, CA
Eohaustorius estaurius
of Contaminants,
9.99 6.57 7.42 3.31 2.26 2.58
EVS04 top
Station name
Concentration
a 4 9 2 3 2 g
EROD activity in&h
\
. .
0~~~~~,~~~~,~~~~,,~~~1”“1”“1”“1””,~~~~l~~~~l 0
10
20
30
40
50
60
70
80
90 100
Eohaustorius estuarius survival (percent) Fig. 2. Correlation between hepatic EROD activity in sanddabs (after 60 d exposure) and the survival of amphipod Eohaustorius estuarius (after 10 d exposure) to the same sediment.
(Monterey Bay; Table 1) sediments (ANOVA Fisher multiple range test of arcsin transformed % survival values, p < 0.01). The percentage of sand in gradient test sediments may explain some variability in survival (r2 = 0.36, p > 0.1). Quality control monitoring indicated that ammonia ( < 6 ppm) and pH (7.48.3) were within acceptable bounds and unlikely contributors to toxicity. Rates of hepatic EROD activity were higher in groups of fish exposed to the more contaminated sediments (Table 1) and were highly correlated to EPAH concentrations and CPCB concentrations (?=0.70, p < 0.05). Rates of hepatic EROD activity were highly correlated also with the survival of Eohaustorius in IO-day bioassays (r2 =0.90, p < 0.001) (Fig. 2). All fish held in the laboratory had significantly higher EROD activity (Fisher’s Protected LSD, p < 0.05) than the pre-experimental reference fish from Tomales Bay, indicating some EROD induction due to the experimental regime.
DISCUSSION Factors that co-vary with contamination, such as grain size, can contribute to the death of organisms in sediment bioassays. If such effects are significant but undetected they could lead to ‘false positive’ determinations of unacceptable sediment contamination (Spies, 1989). In this experiment, toxicity from the fine-grain, low-contamination control site of Carr Inlet, WA, was not significantly different from the coarse-grain home sediment control. This result provides further support to the contention that any impacts of the physical characteristics of sediments on amphipod survival in bioassays are not consistent among locations (Chapman et al., 1991).
48
A. J. Gunther et al.
By exposing fish to the same sediments used in the bioassays we introduce an independent basis for interpreting amphipod survival. Grain size or trace metals content of sediments is not known to stimulate hepatic EROD activity in fish. Therefore, the increases in EROD activity in the livers of sanddabs in this study were most likely due to CPAH or XPCB exposure. Assuming that these contaminants are bioavailable both to the fish and the amphipod, the highly significant inverse relationship between hepatic EROD activity in sanddabs and the survival of the amphipod Eohaustorius (Fig. 2) is consistent with an interpretation of PAHIPCB exposure as at least a contributing cause of toxicity in the sediment bioassays. The difference in amphipod survival and EROD activity in the samples from stations ‘PPP top’ and ‘CC4 full’, which have similar concentrations of CPAH or CPCB, suggests that the amphipods and the fish were responding to other contaminants as well. The basis for the use of sediment bioassays to predict the toxicity of contaminated sediments would be strengthened through validation of laboratory results by field research. Unfortunately, our current understanding of processes controlling sediment toxicity in marine ecosystems is inadequate to properly develop corroborative evidence from the field (Luoma & Carter, 1993; Chapman, 1995). In this experiment, the results of the sediment bioassays accurately predict a marker of contaminant exposure in fish, which in turn has been correlated with histopathological abnormahties in the same fish (Spies et al., in prep.). This finding supports the use of these bioassays to screen sediments for ecologically significant concentrations of contaminants, While the sediment bioassay may be the most practical tool marine scientists can currently offer to assist in the management of contaminated sediments, uncertainty remains regarding the relevance of these laboratory tests to the actual toxic effects of contaminants in marine sediments. Further research on the processes contributing to sediment toxicity in specific ecosystems (Luoma & Carter, 1993) should reduce this uncertainty. ACKNOWLEDGEMENTS This study could not have been completed without the cooperation of Jim Clegg, Paul Siri and the staff of the Bodega Marine Laboratory. Michael Carlin and Karen Taberski of the San Francisco Bay Regional Water Quality Control Board provided useful comments on the draft. We also appreciate the efforts of Dr Mark Stephenson and his staff at the Moss Landing Marine Laboratory for providing the sediment samples. Dr Robert Risebrough and the Bodega Bay Institute provided valuable administrative assistance, as did the work of Ms. Barbara Forbes (this paper is dedicated to her memory). Funding for this project was provided by the California Water Resources Control Board. REFERENCES ASTM. (1992). Standard Guide for Conducting IO-day Static Sediment Toxicity Tests with Marine and Estuarine Amphipods (method No. EX367-90). American Society for Testing and Materials, Philadelphia. Baudo, R., Giesy, J. P. & Muntau, H. (eds) (1990). Sediments; Chemistry and Toxicity of In-place Pollutants, Lewis Publishers, Boca Raton, FL. Benson, W. H. & Di Giulio, R. T, (1992). Biomarkers in hazard assessments of contaminated sediments. In Sediment Toxicity Assessment, ed. Burton, G. A., pp. 241-265, Lewis Press, Boca Raton, FL.
EROD activity in jsh
49
Burton, G. A. Jr. (Ed). (1992). Sediment Toxicity Assessment. Lewis Publishers, Boca Raton, FL. Burton, G. A. Jr. & Scott, K. J. (1992). Sediment toxicity evaluations. Environ. Sci. Technol., 26, 2068-2075.
Chapman, P. M., Long, E. R., Swartz, R. C., Dewitt, T. H. & Pastorok, R. (1991). Sediment toxicity tests, sediment chemistry, and benthic ecology do provide new insights into the significance and management of contaminated sediments-a reply to Robert Spies. Environ. Toxicol. Chem., 10, 14. Chapman, P. M. (1995). Do sediment toxicity tests require field validation? Environ. Toxicol. Chem., 14, 1451-1453. Dewitt, T. H., Ditsworth, G. R. & Swartz, R. C. (1988). Effects of natural sediment features on survival of the phoxocephalid amphipod, Rhepoxinius abronius. Mar. Environ. Res., 25, 99-124. Dewitt, T. H., Swartz, R. C. & Lamberson, J. 0. (1989). Measuring the acute toxicity of estuarine sediments. Environ. Toxicol. Chem., 8, 1035-1048. Eggens, M. L. & Galgani, F. (1992). Ethoxyresorufin-0-deethylase (EROD) activity in flatfish: Fast determination with a fluorescence plate reader. Mar. Environ. Res., 33, 213-221. Flegal, A. R., Cutter, L. S. & Martin, J. H. (1981). A study of the chemistry of marine sediments and wastewater sludge. California State Water Resources Control Board, Sacramento, CA. 69 PP. Flegal, A. R., Abu-Saba, K., Rivera-Duarte, I., Crick, J., Ritson, P., Smith, G., Scelfo, G., Owens, K. & Smith, G. (1994). Trace element concentrations of sediments and porewaters. In San Francisco Estuary Pilot Regional Monitoring Program: Sediment Studies. San Francisco Bay Regional Water Quality Control Board, Oakland, CA, pp 3-l - 4-15. Folk, R. L. (1974). Petrology of Sedimentary Rocks. Hemphill Publishing Company. Austin, Texas, 182 pp. Klotz, A. V., Stegeman, J. J. & Walsh, C. (1984). An alternative 7-ethoxyresorufin-0-deethylase activity assay: a continuous visible spectrophotometric method for the measurement of cytochrome P450 monoxygenase activity. Anal. Biochem., 140, 138-145 Luoma, S. N. & Carter, J. L. (1993). Understanding the toxicity of contaminant in sediments: beyond the bioassay-based paradigm. Environ. Toxicol. Chem., 12, 793-796. Oakden, J. M. (1984). Feeding and substrate preference in five species of photocephalid amphipods from central California. J. Crust. Biol., 4, 233-247. National Research Council. (1989). Contaminated Marine Sediments - Assessment and Remediation. Committee on Contaminated Marine Sediments, Marine Board, Commission on Engineering and Technical Systems. National Academy Press, Washington, D.C. Risebrough, R. R. (1994). Organic contaminants in sediments and porewaters. In San Francisco Estuary Pilot Regional Monitoring Program: Sediment Studies. San Francisco Bay Regional Water Quality Control Board, Oakland, CA. pp 41 - 4-52. Rivera-Duarte, I. & Flegal, A. R. (1991). Benthic lead fluxes in San Francisco Bay, California, USA. Geochemica et Cosmochemica Acta, 25, 3307-33 13.
Spies, R. B. (1989). Sediment bioassays, chemical contaminants and benthic ecology: new insights or just muddy water? Marine Environ. Res., 27, 73-75. Spies, R. B., Gunther, A. J., Risebrough, R. R., Stegeman, J. J., Smolowitz, R., Sanders, B. KcHain, L. (in prep.). Effects of long-term exposure to contaminated sediments from San Francisco Bay: Induction of CYPIA, stress proteins and histological change in the sanddab (Citharichthys stigmaeus).
Stein, J. E., Collier, T. K., Reichert, W. L., Casillas, E., Horn, T. & Varanasi, U. (1992). Bioindicators of contaminant exposure and sublethal effects: Studies with benthic fish in Puget Sound, Washington. Environ. Toxicol. Chem., 11, 701-714. Sulaiman, N. S. & Burke, M. D. (1991). Assessment of sublethal pollutant impact on flounders in an industrialized estuary using hepatic biochemical indices. Mar. Ecol. Progr. Ser., 68, 207-212. Wirgin, I., Grunwald, L. C., Courtenay, S., Kreamer, G. L., Reichert, W. L. & Stein, J. E. (1994). A biomarker approach to assessing xenobiotic exposure in Atlantic tomcod from the North American Atlantic coast. Environ. Health Persp., 102, 764770.
PII:
SOl41-1136(96)00101-g
1997 0 1997 Elsevier !kience Ltd All rights reserved. Printed in Great Britain 0141-1136/97 %17.00+0.00
EROD Activity in Fish as an Independent Measure of Contaminant-Induced Mortality of Invertebrates in Sediment Bioassays Andrew J. Gunther,a Robert B. Spies,” John Stegeman,b Bruce Woodiqb Diane Carney,c James Oakden” & Leslie Hai& “Applied Marine Sciences, 2155 Las Positas Ct. Suite S, Livermore, California 94550, USA bWoods Hole Oceanographic Institution, Woods Hole, Massachusetts 02543, USA CM~~~Landing Marine Laboratory, P.O. Box 450, Moss Landing, California 95039, USA dBodega Marine Laboratory, P.O. Box 247, Bodega Bay, California 94923, USA (Received 22 April 1995; revised version received 1 September 1996; accepted 7 September 1996)
ABSTRACT The survival of amphipods in sediment bioassays was compared to the induction of ethoxyresorufin-0-deethylase (EROD) activity in speckled sanddabs (Citharichthys stigmaeus) to investigate the relative importance of physical and chemical characteristics of sediment in toxicity bioassays. Sediments from San Francisco Bay containing l-30 ppm (dry) of polynuclear aromatic hydrocarbons (PAHs), 3-20 ppb (dry) of polychlorinated biphenyls (PCBs) and variable concentrations of several trace elements were used in separate exposure assays for sanddabs (60 d) and the amphipod Eohaustorius estuarius (10 d). A highly signt&ant correlation (13 = 0.90, p < .OOl) was documented between EROD activity in the sanddabs and Eohaustorius survival for all treatments. The accurate prediction of amphipod survival by the contaminant exposure biomarker in the Jish is consistent with the contention that contamination, rather than physical characteristics of the sediments, is the cause of amphipod mortality. This result supports the use of these bioassays to screen sediments for ecologically significant contaminant concentrations. 0 1997 Elsevier Science Ltd. All rights reserved
INTRODUCTION The identification of contaminated sediments harmful to marine life is a challenging aspect of marine environmental management. The most widely used tools for identification of potentially harmful sediments have been sediment bioassays employing a variety of invertebrate organisms and sediment exposure conditions (Baud0 et al., 1990; Burton, 1992; Burton & Scott, 1992; Dewitt et al., 1989; National Research Council, 1989). While
the endpoints of the tests have been considered as evidence of contaminant 41
effects, and it
42
A. J. Guntheret al.
is clear that contamination can affect survival of benthic organisms, other factors such as the physical characteristics of sediments can also influence the survival of test organisms (Oakden, 1984; Dewitt et al., 1988). This has been particularly true for assays employing amphipods in which exposure to sediment types outside the range of the natural habitat is common. It would be useful to have an independent indication whether mortality in the test population was related to the degree of contamination rather than a result of other characteristics of the sediments. Biochemical, genetic and histological changes due to contaminant exposure could provide such corroborative evidence. Such biomarkers of contaminant exposure and effect are not well developed for invertebrates, however, so routine or widespread use of the biomarker approach with invertebrates must await further studies. Small fishes have occasionally been used in sediment bioassays and in the field (Spies, unpubl.), and a better understanding of contaminant biomarkers exists for fish than for invertebrates (Benson & DiGuilo, 1992). Much of the reported work with fish comes from either field collections in areas of known contamination or from shortterm laboratory exposures to contaminants (Sulaiman & Burke, et al., 1991; Stein et al., 1992; Wirgin et al., 1994). The purpose of this study was to use a biomarker response in a demersal fish as an independent indication of contaminant bioavailability for comparison with the results of a standard amphipod bioassay. Speckled sanddabs (Citharichthys stigmaeus) were exposed for two months and the estuarine amphipod Eohaustorius estuarius for 10 days to contaminated sediments collected along a contamination gradient near an old refinery outfall in San Francisco Bay. The grain size and contaminant concentrations of the sediments varied between samples, but not in the same manner. Ethoxyresorufin-0-deethylase (EROD) activity in the liver of sanddabs was compared to amphipod survival in each sediment. If physical characteristics of the sediment have a predominant influence on amphipod survival, a poor correlation between the EROD activity in fish and amphipod survival for sediments might be expected. If, however, amphipod survival is being controlled predominantly by contamination, then there should be a strong correlation between amphipod survival in a sediment and EROD activity in fish exposed to that same sediment.
METHODS Sediments
Sediments for this experiment were collected from five stations in Castro Cove and San Francisco Bay (Fig. 1) and an uncontaminated reference area (Carr Inlet, WA) during the week 25-29 May 1992. The selected stations were expected to contain a wide range of contaminant concentrations as portions of Castro Cove sediments are known to contain high concentrations of hydrocarbons in the sediments from past discharges of oil refinery effluent. At two of the stations both shallow sediments (top 2 cm) and sediments from ‘full’ cores (to 30 cm) were used, resulting in a total of seven treatments. The percent sand fraction was estimated using sieve analysis (Folk, 1974).
EROD activity inJish
43
44
A. J. Gunther et al.
Organic chemistry
A detailed description of the chemical methods may be found in Risebrough (1994). Briefly, freeze-dried sediments (l&20 g) were mixed with kiln-fired sodium sulfate, soxhlet extracted with methylene chloride, and separated from lipids using florisil-column chromatography. Extract volumes were concentrated to 14 ml, and analyzed by both electron-capture gas chromatography (Varian 3400 GC with 8100 autosampler) and gas chromatography/mass spectroscopy (Varian Saturn II with 8 100 autosampler). DB5 30 m columns (J&W, Scientific, Inc., Folson, CA, USA) were used in both instruments, facilitating confirmation of GC/ECD indentifications by GC/MS. For quality control, internal standards (deuterated phenanthrene, deuterated chrysene and decachlorobiphenyl) were added to the samples prior to extraction to determine recoveries. System blanks (including freeze-dried flourisil), reference sediments (HS-I and HS-6, provided by the National Research Council of Canada), and National Institute of Standards & Technology, NIST Standard Reference Material 1974 (mussel tissue) were also analyzed. CPCB was calculated by summing the concentrations of the following congeners: 005/8, 015, 018, 027124, 028, 029, 031, 040,044,049, 052, 060156, 066195, 070, 074, 087/l 15, 097, 099, 101/90, 105/132, 110/77, 114/131/122, 118, 128, 129/178, 132, 137/176, 138, 141/179, 146, 149, 151/82, 153, 156/171/202, 157/173/201, 158, 170/190, 174, 177, 180, 183, 185, 187, 189, 191, 194, 195/208, 1961203, 199, 205, 206, 207. CPAH was calculated by summing the concentrations of the following compounds: anthracene, benzo[a]anthracene, benzo[a]pyrene, benzo[b]fluoranthene, benzo[e]pyrene, benzo[ghi]perylene, benzo[k]fluoranthene, chrysene, dibenzo[a,h]anthracene, fluoranthene, indeno[l,2,3_cd]pyrene, methylanthracene, phenanthrene, pyrene, all single-substituted methyl-phenanthrenes. Trace metals
A detailed description of the chemical methods may be found in Flegal et al. (1981) (see also Rivera-Duarte & Flegal, 1991). Briefly, homogenized sediment was dried overnight (7O”Q and 1 g of dried sediment was combined with 10 ml aqua regia (3:1 HCl:HN03). The mixture was allowed to stand overnight at room temperature (with a loose cap to let gasses escape), refluxed 4 h (2Oo”Q cooled and then dried on a hot plate. Ten ml 1 N HN03 was added, the sample mixed and allowed to redissolve overnight. The resulting solution was filtered (0.45 pm) and analyzed by graphite furnace atomic absorption spectrometry, flame atomic absorption spectrometry and inductively coupled plasma emission spectrometry. Procedural blanks were analyzed to quantify elemental contamination and extraction efficiency was quantified using Environment Canada standard reference sediments (MESS, BCSS). Fish collection,
incubation and dissection
Speckled sanddabs (Citharichthys stigmaeus) were collected by otter trawl in Tomales Bay (30 km WNW of San Francisco Bay) in the ten days before the start of the experiment, held in a large outdoor aquarium and then transferred to the Bodega Marine Laboratory the day prior to the beginning of the experiment. Since it was possible that the marine laboratory environment itself might have an influence on EROD activity, a small additional sample of fish was collected from the mouth of Tomales Bay during 2-5 June
EROD activity infish
45
1992 as a pre-experimental reference. The available fish ranged in size up to 9 cm Standard Length (SL). Fish were sorted to four size classes (5-6,6-7,7-8 and > 8 cm SL), distributed as evenly as possible to the test aquaria (7 fish in each) and fed a ration of chopped frozen squid at a rate of 5% body weight per day. Uneaten food was routinely removed from the aquaria. The fish collected from Tomales Bay were dissected on 7 June 1992 as pre-experimental controls. The remainder of the fish were sacrificed after two months’ exposure on 30 July to 1 August 1992. (While it is likely, based upon previous studies, that EROD activity was induced within 24 h of exposure to the more contaminated sediments, a 60 d exposure was utilized due to simultaneous histopathological measurements being made on these fish [Spies et al., in prep.].) Each fish was blotted dry, weighed and the standard length determined to the nearest millimeter. The liver was removed, being carefully separated from the gall bladder, and weighed whole. A small cross-section of liver was taken, usually normal to the longest axis, and preserved in 10% neutral formalin for later analysis for histopathological abnormalities (see Spies et al., in prep.). The remaining liver was weighed and placed on ice in phosphate buffered saline (pH 7.2) for several minutes to await further processing for an analysis of EROD activity. Sediment bioassays The toxicity of sediments was evaluated using the estuarine amphipod Eohaustorius estuarius. Studies were conducted at the Moss Landing Marine Laboratory (Moss Landing, CA) using standard methods (ASTM, 1992). All treatments had five laboratory replicates and 20 individuals per replicate. Two negative controls, a reference (Marconi Cove in Tomales Bay, CA, or Carr Inlet, WA) and home sediment (Monterey Bay near Moss Landing, CA), were run with each toxicity test. Acceptability of the test was defined by 290% survivorship in the home sediment. Determination of EROD activity From 0.20 to 0.70 g of liver from each fish was homogenized in phosphate buffered saline within one hour of collection using a hand tissue grinder. The homogenates were centrifuged for 15 min at 13,000 g at 1°C and the supernatant was retained and stored at -75°C until it could be assayed. The EROD activity of the supernatant was evaluated either spectrophotometrically according to the methods of Klotz et al. (1984) or fluorometrically by modifications to the method of Eggens and Galgani (1992) using the Millipore Cytofluor fluorescent plate reader. The hepatic EROD activity is reported as activity per gram of fish for each individual, rather than per gram of sample protein, as samples were too small to prepare a microsomal fraction and thus tests were run on the post-mitochondrial supernatant (PMS). In such cases, normalizing data to liver or body weight can be as or more accurate than normalizing to PMS protein.
RESULTS The survival of amphipods in all sediments except for those from the station furthest away from the refinery (PPP; Fig. 1) was significantly lower than in the reference and the home
*
0.44
95
se = standard error. *Indicates significant difference NA = Not analysed. dw = dry weight. Organic contaminants reported
19
*
NA
NA
NA
1994) trace elements
site.
NA
reported
NA
NA
NA
210 75 163 124 87 55
144
Zn
NA
NA
86 195 29 59 43 26 25
Cu
91 101 39 93 79 48 30
Ni
NA
NA
NA
33 63 21 63 28 25 13
Ph
Induction
NA
EROD
NA
NA
0.47 1.24 0.33 0.65 0.20 0.31 0.59
Cd
NA
NA
0.35 0.42 0.12 0.30 0.27 0.20 0.17
Ag
of Sanddabs
as pg/g dw (data from Flegal et al., 1994).
NA
NA
NA
NA
NA
NA
8.133 1.431 2.493 4.313 2.763 0.469
NA
80 102 53 87 78 59 41
6.98
3.212
Cr
2.26 0.793 3.265 0.947 0.554 0.243
C DDTs
10,397 19.902
C CChlordanes PCB s
1.98 30,461 20.885 0.73 695 9.325 1.3 1,171 12.819 1.12 1,106 14.864 0.81 145 12.74 1 783 3.447
1.42
org. C. C PAH (%)
and Carr Inlet control
94
NA
0 14 9 14 45 7
1
sand (%)
as rig/g dw (data from Risebrough,
between
* *
treatment
NA
Bay
0.16
0.17
* * * *
Bay Monterey
0 70 15 70 85 95
41.7 16.7 24.8 13.6 11.1 8.14
0.89 1.34 0.89 0.89 0.44
0 EVS04 full 14 cc2 full 3 cc4 full 14 PPP top 17 PPP full 19 Carr Inlet top Tomales NA
se 7.12
mean 37.7
se 0.44
% 10
mean 2
survival
EROD Acfivify (amok fish)
TABLE 1 Survival of Amphipod Eohaustorius estuarius in 10 d Bioassays, and Hepatic after 60 d Exposures, for Sediments from San Francisco Bay, CA
Eohaustorius estaurius
of Contaminants,
9.99 6.57 7.42 3.31 2.26 2.58
EVS04 top
Station name
Concentration
a 4 9 2 3 2 g
EROD activity in&h
\
. .
0~~~~~,~~~~,~~~~,,~~~1”“1”“1”“1””,~~~~l~~~~l 0
10
20
30
40
50
60
70
80
90 100
Eohaustorius estuarius survival (percent) Fig. 2. Correlation between hepatic EROD activity in sanddabs (after 60 d exposure) and the survival of amphipod Eohaustorius estuarius (after 10 d exposure) to the same sediment.
(Monterey Bay; Table 1) sediments (ANOVA Fisher multiple range test of arcsin transformed % survival values, p < 0.01). The percentage of sand in gradient test sediments may explain some variability in survival (r2 = 0.36, p > 0.1). Quality control monitoring indicated that ammonia ( < 6 ppm) and pH (7.48.3) were within acceptable bounds and unlikely contributors to toxicity. Rates of hepatic EROD activity were higher in groups of fish exposed to the more contaminated sediments (Table 1) and were highly correlated to EPAH concentrations and CPCB concentrations (?=0.70, p < 0.05). Rates of hepatic EROD activity were highly correlated also with the survival of Eohaustorius in IO-day bioassays (r2 =0.90, p < 0.001) (Fig. 2). All fish held in the laboratory had significantly higher EROD activity (Fisher’s Protected LSD, p < 0.05) than the pre-experimental reference fish from Tomales Bay, indicating some EROD induction due to the experimental regime.
DISCUSSION Factors that co-vary with contamination, such as grain size, can contribute to the death of organisms in sediment bioassays. If such effects are significant but undetected they could lead to ‘false positive’ determinations of unacceptable sediment contamination (Spies, 1989). In this experiment, toxicity from the fine-grain, low-contamination control site of Carr Inlet, WA, was not significantly different from the coarse-grain home sediment control. This result provides further support to the contention that any impacts of the physical characteristics of sediments on amphipod survival in bioassays are not consistent among locations (Chapman et al., 1991).
48
A. J. Gunther et al.
By exposing fish to the same sediments used in the bioassays we introduce an independent basis for interpreting amphipod survival. Grain size or trace metals content of sediments is not known to stimulate hepatic EROD activity in fish. Therefore, the increases in EROD activity in the livers of sanddabs in this study were most likely due to CPAH or XPCB exposure. Assuming that these contaminants are bioavailable both to the fish and the amphipod, the highly significant inverse relationship between hepatic EROD activity in sanddabs and the survival of the amphipod Eohaustorius (Fig. 2) is consistent with an interpretation of PAHIPCB exposure as at least a contributing cause of toxicity in the sediment bioassays. The difference in amphipod survival and EROD activity in the samples from stations ‘PPP top’ and ‘CC4 full’, which have similar concentrations of CPAH or CPCB, suggests that the amphipods and the fish were responding to other contaminants as well. The basis for the use of sediment bioassays to predict the toxicity of contaminated sediments would be strengthened through validation of laboratory results by field research. Unfortunately, our current understanding of processes controlling sediment toxicity in marine ecosystems is inadequate to properly develop corroborative evidence from the field (Luoma & Carter, 1993; Chapman, 1995). In this experiment, the results of the sediment bioassays accurately predict a marker of contaminant exposure in fish, which in turn has been correlated with histopathological abnormahties in the same fish (Spies et al., in prep.). This finding supports the use of these bioassays to screen sediments for ecologically significant concentrations of contaminants, While the sediment bioassay may be the most practical tool marine scientists can currently offer to assist in the management of contaminated sediments, uncertainty remains regarding the relevance of these laboratory tests to the actual toxic effects of contaminants in marine sediments. Further research on the processes contributing to sediment toxicity in specific ecosystems (Luoma & Carter, 1993) should reduce this uncertainty. ACKNOWLEDGEMENTS This study could not have been completed without the cooperation of Jim Clegg, Paul Siri and the staff of the Bodega Marine Laboratory. Michael Carlin and Karen Taberski of the San Francisco Bay Regional Water Quality Control Board provided useful comments on the draft. We also appreciate the efforts of Dr Mark Stephenson and his staff at the Moss Landing Marine Laboratory for providing the sediment samples. Dr Robert Risebrough and the Bodega Bay Institute provided valuable administrative assistance, as did the work of Ms. Barbara Forbes (this paper is dedicated to her memory). Funding for this project was provided by the California Water Resources Control Board. REFERENCES ASTM. (1992). Standard Guide for Conducting IO-day Static Sediment Toxicity Tests with Marine and Estuarine Amphipods (method No. EX367-90). American Society for Testing and Materials, Philadelphia. Baudo, R., Giesy, J. P. & Muntau, H. (eds) (1990). Sediments; Chemistry and Toxicity of In-place Pollutants, Lewis Publishers, Boca Raton, FL. Benson, W. H. & Di Giulio, R. T, (1992). Biomarkers in hazard assessments of contaminated sediments. In Sediment Toxicity Assessment, ed. Burton, G. A., pp. 241-265, Lewis Press, Boca Raton, FL.
EROD activity in jsh
49
Burton, G. A. Jr. (Ed). (1992). Sediment Toxicity Assessment. Lewis Publishers, Boca Raton, FL. Burton, G. A. Jr. & Scott, K. J. (1992). Sediment toxicity evaluations. Environ. Sci. Technol., 26, 2068-2075.
Chapman, P. M., Long, E. R., Swartz, R. C., Dewitt, T. H. & Pastorok, R. (1991). Sediment toxicity tests, sediment chemistry, and benthic ecology do provide new insights into the significance and management of contaminated sediments-a reply to Robert Spies. Environ. Toxicol. Chem., 10, 14. Chapman, P. M. (1995). Do sediment toxicity tests require field validation? Environ. Toxicol. Chem., 14, 1451-1453. Dewitt, T. H., Ditsworth, G. R. & Swartz, R. C. (1988). Effects of natural sediment features on survival of the phoxocephalid amphipod, Rhepoxinius abronius. Mar. Environ. Res., 25, 99-124. Dewitt, T. H., Swartz, R. C. & Lamberson, J. 0. (1989). Measuring the acute toxicity of estuarine sediments. Environ. Toxicol. Chem., 8, 1035-1048. Eggens, M. L. & Galgani, F. (1992). Ethoxyresorufin-0-deethylase (EROD) activity in flatfish: Fast determination with a fluorescence plate reader. Mar. Environ. Res., 33, 213-221. Flegal, A. R., Cutter, L. S. & Martin, J. H. (1981). A study of the chemistry of marine sediments and wastewater sludge. California State Water Resources Control Board, Sacramento, CA. 69 PP. Flegal, A. R., Abu-Saba, K., Rivera-Duarte, I., Crick, J., Ritson, P., Smith, G., Scelfo, G., Owens, K. & Smith, G. (1994). Trace element concentrations of sediments and porewaters. In San Francisco Estuary Pilot Regional Monitoring Program: Sediment Studies. San Francisco Bay Regional Water Quality Control Board, Oakland, CA, pp 3-l - 4-15. Folk, R. L. (1974). Petrology of Sedimentary Rocks. Hemphill Publishing Company. Austin, Texas, 182 pp. Klotz, A. V., Stegeman, J. J. & Walsh, C. (1984). An alternative 7-ethoxyresorufin-0-deethylase activity assay: a continuous visible spectrophotometric method for the measurement of cytochrome P450 monoxygenase activity. Anal. Biochem., 140, 138-145 Luoma, S. N. & Carter, J. L. (1993). Understanding the toxicity of contaminant in sediments: beyond the bioassay-based paradigm. Environ. Toxicol. Chem., 12, 793-796. Oakden, J. M. (1984). Feeding and substrate preference in five species of photocephalid amphipods from central California. J. Crust. Biol., 4, 233-247. National Research Council. (1989). Contaminated Marine Sediments - Assessment and Remediation. Committee on Contaminated Marine Sediments, Marine Board, Commission on Engineering and Technical Systems. National Academy Press, Washington, D.C. Risebrough, R. R. (1994). Organic contaminants in sediments and porewaters. In San Francisco Estuary Pilot Regional Monitoring Program: Sediment Studies. San Francisco Bay Regional Water Quality Control Board, Oakland, CA. pp 41 - 4-52. Rivera-Duarte, I. & Flegal, A. R. (1991). Benthic lead fluxes in San Francisco Bay, California, USA. Geochemica et Cosmochemica Acta, 25, 3307-33 13.
Spies, R. B. (1989). Sediment bioassays, chemical contaminants and benthic ecology: new insights or just muddy water? Marine Environ. Res., 27, 73-75. Spies, R. B., Gunther, A. J., Risebrough, R. R., Stegeman, J. J., Smolowitz, R., Sanders, B. KcHain, L. (in prep.). Effects of long-term exposure to contaminated sediments from San Francisco Bay: Induction of CYPIA, stress proteins and histological change in the sanddab (Citharichthys stigmaeus).
Stein, J. E., Collier, T. K., Reichert, W. L., Casillas, E., Horn, T. & Varanasi, U. (1992). Bioindicators of contaminant exposure and sublethal effects: Studies with benthic fish in Puget Sound, Washington. Environ. Toxicol. Chem., 11, 701-714. Sulaiman, N. S. & Burke, M. D. (1991). Assessment of sublethal pollutant impact on flounders in an industrialized estuary using hepatic biochemical indices. Mar. Ecol. Progr. Ser., 68, 207-212. Wirgin, I., Grunwald, L. C., Courtenay, S., Kreamer, G. L., Reichert, W. L. & Stein, J. E. (1994). A biomarker approach to assessing xenobiotic exposure in Atlantic tomcod from the North American Atlantic coast. Environ. Health Persp., 102, 764770.