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Evaluation of impact of PAH on a tropical fish, Oreochromis mossambicus using multiple biomarkers
Chemosphere 53 (2003) 835–841 www.elsevier.com/locate/chemosphere
Evaluation of impact of PAH on a tropical fish, Oreochromis mossambicus using multiple biomarkers M.S. Shailaja *, Classy DÕSilva National Institute of Oceanography, Dona Paula, Goa 403004, India Received 3 February 2003; received in revised form 26 June 2003; accepted 2 July 2003
Abstract Treatment of the widely occurring tropical cichlid, Oreochromis mossambicus with a pure polycyclic aromatic hydrocarbon (PAH), phenanthrene, induced a concentration-dependant formation of the enzyme, 7-ethoxyresorufin-Odeethylase (EROD). Concomittant increase (65–669%) in the activity of sorbitol dehydrogenase in the serum (SSDH) occurring with EROD induction denoted liver cell damage, which was more severe in fish exposed to lower concentrations (0.4–4 lg g1 ) of the chemical. In O. mossambicus exposed to 25% refinery effluent liver damage associated with cell death was indicated by the twin analyses of SSDH and liver somatic index. Cell injury appeared to have occurred at low PAH concentrations due to inadequately induced phase II-related detoxification of metabolites. This was indicated by the nearly 33% higher activity of hepatic glutathione S-transferase (GST) in fish exposed to higher PAH concentrations as compared to low-exposure animals. Tilapia such as O. mossambicus were found to be eminently suited for biomonitoring in tropical coastal waters. A combination of EROD and serum sorbitol dehydrogenase activity measurements serves as an excellent tool for biomonitoring sublethal effects of PAH pollution in fish. 2003 Elsevier Ltd. All rights reserved. Keywords: Tilapia; Polycyclic aromatic hydrocarbon (PAH); Sublethal effects; 7-Ethoxyresorufin-O-deethylase (EROD); Glutathione S-transferase (GST); Serum sorbitol dehydrogenase
1. Introduction Aqueous environments around the world, estuaries and coastal waters in particular, receive large inputs of anthropogenic pollutants through industrial and urban discharges, atmospheric deposition and terrestrial drainage. With the steady growth of industrial activities in the developing countries, environmental degradation has become a potential problem. Industrial and municipal wastes contain numerous chemicals, both organic [e.g., polycyclic aromatic hydrocarbons (PAHs), pesticides, polychlorinated biphenyls (PCBs), etc.] and inor-
*
Corresponding author. Tel.: +91-832-45-6700; fax: +91832-245-6702. E-mail address: [email protected] (M.S. Shailaja).
ganic (viz., heavy metals such as Hg, Cd, Pb and Cu), capable of causing deleterious effects in aquatic biota. PAHs are potentially carcinogenic, ubiquitous contaminants of much ecotoxicological concern in the coastal region. They induce the formation of mixed function oxygenase (MFO), specifically the isozyme, cytochrome P4501A1 (CYP1A1) in the fish liver which has been employed as a biomarker of exposure to hazardous organic pollutants in monitoring studies (Spies et al., 1982; Payne et al., 1987; Stegeman and Lech, 1991; Collier et al., 1992a; Goksøyr and F€ orlin, 1992; Livingstone, 1993). While the MFO system is essential for the biotransformation of PAHs, its induction could produce damaging side effects through the formation of intermediates that are highly reactive, mutagenic and carcinogenic (Gelboin, 1980; Buhler and Williams, 1988; Stegeman and Lech, 1991). Indeed, a strong linkage has
0045-6535/$ - see front matter 2003 Elsevier Ltd. All rights reserved. doi:10.1016/S0045-6535(03)00667-2
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been shown between the prevalence of several toxicopathic hepatic lesions and CYP1A induction in english sole (Myers et al., 1998). To ensure the health of the coastal oceans around them developing nations in the tropical regions in particular, need to adopt viable monitoring programmes. However, deficiencies in information often limit their ability to do so. For instance, data on biomarker responses to organic contaminants in tropical fish are few (Stegeman et al., 1990, 1997; Gadagbui and Goksoyr, 1996; Pathiratne and George, 1996, 1998; Al-Arabi and Goksoyr, 1998; Bainy et al., 1999; Shailaja and Rodrigues, 2001, 2003) whereas such information is absolutely essential for embarking on PAH monitoring with a suitable species of fish. This paper describes our studies with Oreochromis mossambicus, a widely distributed tropical euryhaline fish exposed in the laboratory to a pure PAH compound (phenanthrene) and a refinery effluent containing a mixture of PAHs with the objective of assessing the suitability of O. mossambicus for biomonitoring PAH impact using a suite of biomarkers. The enzyme biomarkers assayed included the phase I enzyme, CYP1A1 represented by 7-ethoxyresorufin-O-deethylase (EROD), phase II enzyme, glutathione S-transferase (GST) and serum sorbitol dehydrogenase, an indicator of liver cell damage (Gerlach, 1983). Liver somatic index (LSI) was used as a physiological biomarker while biliary fixed wavelength fluorescence was determined as an indicator of PAH metabolites.
2. Materials and methods Juvenile O. mossambicus (weight 49–67 g) were obtained from a local estuarine fish farm. These fish having a high tolerance towards salinity changes were maintained in fresh water for seven days before being administered a single intra-abdominal injection of phenanthrene (Sigma, USA). Due to the small size of the experimental fish, the volume injected into each individual fish in the low and high exposure groups, respectively, consisted of 100 ll of 0.2 or 2 mg ml1 solution of phenanthrene in sunflower oil. The exact dose received by the fish was in the range of 0.4–32 lg g1 , based on the weight of individual fish. Control fish received equivalent amounts (100 ll) of the carrier alone. The experimental fish were maintained in 30 l glass aquaria (up to n ¼ 6, per tank) with continuous aeration and the water being changed every 24 h. Commercial feed was provided ad libitum during the period of experiment. The animals were sacrificed three days after exposure. A second experiment consisted of exposing O. mossambicus (66–71 g body weight; n ¼ 6) to a secondarytreated oil refinery effluent (25% dilution) for two weeks.
The undiluted effluent had a total PAH content of 478 ± 24 lg l1 . The test medium was renewed every alternate day, all other conditions remaining the same as described earlier. Untreated fish (n ¼ 6) formed the control. At the conclusion of each experiment, liver samples were obtained by immediate dissection and homogenates prepared from fresh livers for EROD assay according to the procedure of Gunther et al. (1997). EROD activity was determined in the postmitochondrial supernatant (PMS) with NADPH as the electron donor and resorufin (85 nM) as internal standard (Stagg and Addison, 1995). The activity is reported in terms of nmol resorufin formed min1 mg1 protein. GST activity in the PMS of liver extract was evaluated as the formation of a conjugate between glutathione and 1-chloro-2,4-dinitrobenzene (CDNB) at 340 nm (extinction coefficient ¼ 9.6 cm1 mM1 ) in 50 mM potassium phosphate buffer, pH 6.5 with 1 mM EDTA (Habig et al., 1974). The activity of the enzyme is reported as the number of micromoles of conjugate formed per min mg1 protein. Sorbitol dehydrogenase (SSDH) activity was measured as the rate of reduction of D -fructose by NADH (Gerlach, 1983) in composite serum samples obtained from 2 to 3 fishes by cardiac puncture. One unit of SSDH activity refers to a change in absorbance of 0.001 min1 at 366 nm in 3 ml of the reaction mixture. Protein was estimated in the PMS by the method of Lowry et al. (1951) using bovine serum albumin as the standard. PAH concentrations in the liver were estimated fluorimetrically (Shimadzu spectrofluorophotometer RF-1501) after a 4-h saponification of the tissue using alkaline methanol (Law and Whinnett, 1992) and subsequent hexane extraction. Fluorescent aromatic compounds (FACs) in the bile were determined in fish exposed to refinery effluent by fixed wavelength fluorescence measurements at specified excitation/emission wavelength pairs for selected PAH metabolites (Krahn et al., 1987, 1993). LSI was calculated as (weight of liver/total body weight) · 100. Where statistical evaluation was required, correlation and F-tests were performed using the Analysis Tool Kit of Excel 2000 (Microsoft). p < 0:05 was considered as statistically significant.
3. Results and discussion 3.1. Phenanthrene concentrations and EROD activity in the liver Hepatic concentrations of phenanthrene in O. mossambicus exposed to high (6–32 lg g1 ) and low
M.S. Shailaja, C. DÕSilva / Chemosphere 53 (2003) 835–841
µM Phenanthrene g-1 fish
25
Control Low High
20
* *
15 10 5 0 Phenanthrene exposure
Fig. 1. Hepatic concentrations of phenanthrene (mean ± SD) in O. mossambicus treated with low (0.4–4 lg g1 , n ¼ 10) and high (6–32 lg g1 , n ¼ 12) concentrations of phenanthrene. Control fish (n ¼ 5) were treated only with sunflower oil. Values significantly different from the control (p < 0:05) are marked with asterisk.
(0.4–4 lg g1 ) levels of the compound are presented in Fig. 1. Compared to control fish, both the treatment groups had significantly higher phenanthrene levels in the liver (p < 0:003 and <0.0004, respectively, F test). However, between test groups the values were not significantly different (p > 0:05). Further, a significant inverse relationship was observed between phenanthrene levels in the liver and exposure concentrations (r ¼ 0:585, p < 0:02) suggesting a concentration-dependant increase in the induced metabolism of PAH. The large (35–38%) deviations from the mean observed in the hepatic concentrations of phenanthrene in the two test groups (Fig. 1) were likely due to biotransformation of PAH. Mean hepatic EROD activities in the control, lowexposure and high-exposure groups of fish were, respectively, 0.721 ± 0.045, 5.42 ± 2.08 and 17.31 ± 5.45 nmol resorufin min1 mg1 protein (Table 1). In general,
in the two groups of exposed fish, based on the lowest and highest values, EROD activity (range 1.36–28.49 nmol resorufin min1 mg1 protein) was 2–40 times higher than the control value. Also, the induced activity correlated well (r ¼ 0:799, p < 0:001) with phenanthrene exposure levels more so when the concentrations were >18 ppm (Fig. 2). However, the relationship between phenanthrene concentration in the liver and enzyme activity was not significant (p > 0:05). Induction of CYP1A in fish by small molecular size PAHs (having 64 fused aromatic rings) such as chrysene, phenanthrene, pyrene and naphthalene is irregular and species dependant. For example, phenanthrene did not induce CYP1A biotransformation activity in a rainbow trout liver cell line (Bols et al., 1999) but did so in cod and other bony fishes such as flounder (Goksøyr et al., 1986) and scup (Stegeman et al., 1998). Similarly, pyrene and chrysene did not evoke EROD activity in the rainbow trout (Bols et al., 1999) but again, both of them did so in the carp (van der Weiden et al., 1994). It is thus evident that the response to PAHs in particular, to
EROD activity (nmol resorufin min-1 mg-1 protein)
30
837
30
R = 0.799 20
10
0 0.1
1
10
100
-1 Phenanthrene exposure (µg g fish)
Fig. 2. Exposure dependence of EROD induction in O. mossambicus injected with phenanthrene (n ¼ 21). EROD activity values in control fish are given in Table 1.
Table 1 Ethoxyresorufin-O-deethylase (EROD), glutathione S-transferase (GST) and serum sorbitol dehydrogenase (SSDH) activities in O. mossambicus exposed to phenanthrene Phenanthrene exposure level
ERODa
GSTb
SSDHc
Control
0.72 ± 0.04 (n ¼ 4) 5.42 ± 2.08* (n ¼ 9) 17.31 ± 5.45* (n ¼ 12)
1.932 ± 0.302 (n ¼ 4) 2.255 ± 0.566 (n ¼ 9) 2.994 ± 0.332* (n ¼ 12)
0.67 ± 0.28 (n ¼ 5) 30.09 ± 17.2* (n ¼ 10) 8.67 ± 5.99* (n ¼ 11)
Low (0.4–4 lg g1 ) High (6–32 lg g1 )
Values are mean ± SD and marked with an asterisk when significantly different from the control value (p < 0:05). a nmol resorufin min1 mg1 protein. b mol product min1 mg1 protein. c Units ml1 serum.
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compounds having <4 fused aromatic rings, is not uniform among different fish species nor within the same species. This could be due to intrinsic variations in the affinity of the aryl-hydrocarbon receptor (AhR) of different fish species to small molecular size PAH compounds. Further, large variations in the induction of EROD activity between tropical and temperate fishes (Oreochromis niloticus and rainbow trout, respectively) have also been reported (Pathiratne and George, 1996). 3.2. Hepatic phase II enzyme activity The metabolites formed by phase I biotransformation activity are conjugated via phase II enzymes (e.g., GST) before excretion. In O. mossambicus exposed to 6– 32 lg g1 phenanthrene GST activity (2.994 ± 0.332 lmol min1 mg1 protein, Table 1) was 32.8% higher (p < 0:05) than in the fish subjected to low (<4 lg g1 ) phenanthrene exposure (2.255 ± 0.566 mol min1 mg1 protein). 3.3. Sorbitol dehydrogenase activity in serum Serum sorbitol dehydrogenase activity is a sensitive biochemical indicator of chemically induced liver damage in fish (Dixon et al., 1987). SSDH activity (mean value 20.21 ± 14.31 units min1 ml1 serum) was substantially increased on exposure of O. mossambicus to phenanthrene (0.4–32 lg g1 ) as compared to control fish (0.67 ± 0.28 units min1 ml1 serum), indicating liver cell injury. Also, SSDH activity showed a very distinct inverse correlation (p < 0:01) with PAH exposure levels (Fig. 3). Significantly, the damage was greater in the fish exposed to lower concentrations than in those treated with >18
3.4. Exposure of O. mossambicus to refinery effluent
SSDH activity (units ml-1 serum)
100 80
R = - 0.797
60 40 20 0 0.1
lg g1 phenanthrene as observed from the nearly 3 times higher SSDH values in the former (Table 1). As the liver cell injury apparently was greater in the fish exposed to low concentrations of the toxicant despite the insignificant difference in the hepatic concentrations of phenanthrene in the two test groups (Fig. 1) it is presumed that the metabolic products of biotransformation rather than the parent compound, phenanthrene were responsible for the observed liver damage. The activity of hepatic GST which plays a major role in detoxifying xenobiotics was nearly 33% higher in the fish exposed to higher PAH concentrations as compared to the low-exposure animals. Therefore, possibly, a mismatch in the kinetics of induction of phase I (bioactivation) and phase II (conjugation/ detoxification) systems in O. mossambicus exposed to low PAH concentrations resulting in an inadequate conjugation potential leads to accumulation of highly reactive oxygenated metabolites in the liver and subsequently, tissue damage. This is somewhat similar to the case of english sole, a flatfish species showing a higher predisposition to contaminant-associated hepatic neoplasms versus starry flounder. The activity of the PAH-activation MFO, aryl hydrocarbon hydroxylase in english sole was reported to be higher just as GST was lower than in starry flounder (Collier et al., 1992b). Likewise, Willett et al. (2000) have hypothesized that the observed difference in the sensitivity to PAH-mediated liver cancer in channel catfish (less sensitive) and brown bullhead (more sensitive) is due to differences in their hepatic phase II activities. We are in the process of testing other PAHs including small molecular size compounds such as anthracene and chrysene for similar effects.
1 10 Phenanthrene exposure
100
(µg g-1 fish)
Fig. 3. Relationship between sorbitol dehydrogenase activity in serum of O. mossambicus (n ¼ 21) and phenanthrene exposure concentration. Data points represent SSDH values measured in composite (n ¼ 2 or 3) serum samples expressed against corresponding pooled exposure concentrations. Values of SSDH in control fish are given in Table 1.
Changes in biochemical (EROD and SSDH) and physiological (LSI) parameters measured in O. mossambicus exposed to 1:4 diluted refinery effluent are shown in Fig. 4. Compared to control, EROD was enhanced by 375% (significance 99.9%) in fish exposed to the effluent having a total PAH content of approximately 120 ppb. Serum SDH which was almost nil in control fish measured 73 units in the exposed fish. LSI, on the other hand, was reduced by nearly 27% which was significant (p < 0:05). LSI values are generally elevated in vertebrates experiencing induction of hepatic microsomal P-450 for detoxification of organic compounds (Huuskonen and Lindstr€ om-Seppa, 1995). Hence, the observed reduction in LSI when considered together with increased SSDH activity would indicate hepatocellular injury associated with cell death. It is suspected that the toxic impact of the effluent stemmed essentially from benzo(a)pyrene. The metabolites present in the bile were examined by fixed fluores-
M.S. Shailaja, C. DÕSilva / Chemosphere 53 (2003) 835–841 8 EROD activity SSDH activity (x 10-1) Liver Somatic Index (%)
EROD
6
LSI
SSDH
*
4
* 2
0 Control
839
ticularly valid since reports indicate lowering of CYP1A1 activities when contamination is high (Peters et al., 1994; Celander et al., 1996) while some PAHs (e.g., fluoranthene) could inhibit CYP1A dependant activity (Willett et al., 2001). CYP1A-dependant bioassays may therefore lead to an underestimation of PAH exposures. Besides, lowered EROD induction is also associated with degenerating, preneoplastic and neoplastic livers which could be misleading if employed as the only biomarker and can only be interpreted in relation to the state of health of the liver (K€ ohler and Pluta, 1995).
Exposed
Fig. 4. Impact of exposure of O. mossambicus to 25% refinery effluent. EROD: 7-ethoxyresorufin-O-deethylase (nmol resorufin min1 mg1 protein); SSDH: serum sorbitol dehydrogenase (units min1 ml1 serum); LSI: liver somatic index (%). n ¼ 6 for EROD and LSI; one composite sample (n ¼ 3) for SSDH. Values shown with asterisk were significantly different (p < 0:05) from the control value.
cence measurements at wavelength pairs specific for naphthalene (290/335 nm), pyrene (341/383 nm), benzo(a)pyrene (380/430 nm) and phenanthrene (260/380 nm). Significant differences compared to the values in control fish were seen only in the BaP-type metabolites (Table 2). This study is the first in tropical fish linking liver damage to PAH biotransformation. Our observation that relatively low concentrations of PAH, including non-carcinogenic ones such as phenanthrene, could lead to sublethal hepatic cell injury in fish is important because the liver is an essential organ having many vital physiological functions including that of production of yolk proteins during oocyte development. Secondly, the apparent sublethal toxicity of low levels of innocuous PAHs would also mean a potential enlargement of the sphere of influence of discharge points. Our findings strongly support the collaborative use of enzyme biomarkers such as CYP1A- and SSDH-activities for biomonitoring PAH impact in fish. This is par-
3.5. Tilapia––a tropical fish for PAH effect monitoring The clear relationship seen between EROD activity and exposure to PAH in O. mossambicus qualifies them as suitable organisms for pollution monitoring especially considering their wide geographical distribution, diverse food habits and their ability to grow in both fresh and marine waters. Besides, the induction of CYP1A activity in tilapia is not influenced by sexual maturation (Pathiratne and George, 1998), as is the case with several temperate species used in biomarker studies (Lindstr€ omSeppa and Stegeman, 1995). Tilapia have also been found to be well suited for performing field caging experiments (Gadagbui and Goksoyr, 1996). However, CYP1A induction is a non-specific biomarker (it can be induced by PCBs and PCDDs, besides PAHs) but could be made relevant to PAHs by studying it in combination with biliary FACs specific for PAH metabolites (Beyer et al., 1996).
4. Conclusions The MFO system concerned with biotransformation of PAHs is highly active in the tilapia, O. mossambicus. Thus, tilapia make good sentinel organisms for PAH pollution monitoring in tropical waters. Exposure to low concentrations of PAH (<4 lg g1 ) leads to sublethal hepatic toxicity in fish as shown by the increased activity
Table 2 Fixed fluorescence evaluation of biliary PAH metabolites of O. mossambicus exposed to 25% refinery effluent Bile metabolites B(a)P-type Naphthalene-type Pyrene-type Phenolphthalene-type
Fluorescence intensitya Control fish
Effluent-exposed fish
380 ± 30 240 ± 16 26 ± 4 314 ± 21
670 ± 18* 240 ± 22 27 ± 4 354 ± 15
Values are mean ± SD; n ¼ 6 for control and effluent exposed fish. [Excitation/emission k used for B(a)P-, naphthalene-, pyrene- and phenanthrene-type metabolites were 380/430 nm; 290/335 nm; 341/383 nm and 260/380 nm, respectively.] *Value significantly different from the control value (p < 0:05). a Arbitrary units.
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of serum sorbitol dehydrogenase, an indicator of liver damage. The use of multiple enzyme biomarkers such as CYP1A- and SSDH-activities in combination with bile FACs is especially recommended for biomonitoring PAH impact in fish, considering the ease of analyses of the parameters. Acknowledgements We thank Nancy Rodrigues, Angelo Rodrigues and Christine Fernandes for their excellent technical assistance. This is National Institute of Oceanography contribution No. 3838. References Al-Arabi, S.A.M., Goksoyr, A., 1998. Species characteristics of hepatic biotransformation enzymes in two tropical fish of Bangladesh, one riverine catfish (Rita rita) and one marine mudskipper (Apocryptes bato). Marine Environmental Research 46, 121–122. Bainy, A.C.D., Woodin, B.R., Stegeman, J.J., 1999. Elevated levels of multiple cytochrome P450 forms in tilapia from Billings Reservoir––Sao Paolo, Brazil. Aquatic Toxicology 44, 289–305. Beyer, J., Sandvik, M., Hylland, K., Fjeld, E., Egaas, E., Aas, E., Skare, J.U., Goksøyr, A., 1996. Contaminant accumulation and biomarker responses in flounder (Platichthys flesus L.) and Atlantic cod (Gadus morhua L.) exposed by caging to polluted sediments in Soerfjorden, Norway. Aquatic Toxicology 36, 75–98. Bols, N.C., Schirmer, K., Joyce, E.M., Dixon, D.G., Greenberg, B.M., Whyte, J.J., 1999. Ability of polycyclic aromatic hydrocarbons to induce 7-ethoxyresorufin-O-deethylase activity in a trout liver cell line. Ecotoxicology & Environmental Safety 44, 118–128. Buhler, D.R., Williams, D.E., 1988. The role of biotransformation in the toxicity of chemicals. Aquatic Toxicology 11, 19–28. Celander, M., Stegeman, J.J., Forlin, L., 1996. CYP1A1-, CYP2B- and CYP3A-like proteins in rainbow trout (Oncorhynchus mykiss) liver: CYP1A1-specific down-regulation after prolonged exposure to PCB. Marine Environmental Research 42, 283–286. Collier, T.K., Connor, S.D., Eberhart, B.L., Anulacion, B.F., Goksoyr, A., Varanasi, U., 1992a. Using cytochrome P450 to monitor the aquatic environment: initial results from regional and national surveys. Marine Environmental Research 34, 195–199. Collier, T.K., Singh, S.V., Awasthi, Y.C., Varanasi, U., 1992b. Hepatic xenobiotic metabolizing enzymes in two species of benthic fish showing different prevalences of contaminant–– associated liver neoplasms. Toxicolology & Applied Pharmacology 113, 319–324. Dixon, D.G., Hodson, P.V., Kaiser, K.L.E., 1987. Serum sorbitol dehydrogenase activity as an indicator of chemically induced liver damage in rainbow trout. Environmental Toxicology & Chemistry 6, 685–696.
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Stagg, R.M., Addison, R.F., 1995. An inter-laboratory comparison of measurements of 7-ethoxyresorufin O-de-ethylase activity in dab (Limanda limanda) liver. Marine Environmental Research 40, 93–108. Stegeman, J.J., Lech, J.J., 1991. Cytochrome P450 monooxygenase systems in aquatic species: carcinogen metabolism and biomarkers for carcinogen and pollutant exposure. Environmental Health Perspectives 90, 101–109. Stegeman, J.J., Renton, K.W., Woodin, B.R., Zhang, Y.-S., Addison, R.F., 1990. Experimental and environmental induction of cytochrome P450E in fish from Bermuda waters. Journal of Experimental Marine Biology & Ecology 138, 49–67. Stegeman, J.J., Woodin, B.R., Singh, H., Oleksiak, M.F., Celander, M., 1997. Cytochrome P450 (CYP) in tropical fishes: catalytic activities, expression of multiple CYP proteins and high levels of microsomal P450 in liver of fishes from Bermuda. Comparative Biochemistry & Physiology C 116, 61–75. Stegeman, J.J., Soballa, V., Morrison, H.G., Jacob, J., Doehmer, J., 1998. Catalytic properties of scup cytochrome P4501A expressed in Chinese hamster V79 cells. Marine Environmental Research 46, 395. van der Weiden, M.E., Hanegraaf, F.H.M., Eggens, M.L., Celander, M., Seinen, W., van den Berg, M., 1994. Temporal induction of cytochrome P4501A in the mirror carp (Cyprinus carpio) after administration of several polycyclic aromatic hydrocarbons. Environmental Toxicology & Chemistry 13, 797–802. Willett, K., Rogers, J., Di Guilio, R., 2000. Differential phase I and phase II enzyme activities in brown bullhead and channel catfish exposed to naphthoflavone. Marine Environmental Research 50, 67–68. Willett, K.L., Wassenberg, D., Lienesch, L., Reichert, W., Di Guilio, R.T., 2001. In vivo and in vitro inhibition of CYP1A-dependent activity in Fundulus heteroclitus by polynuclear aromatic hydrocarbon fluoranthene. Toxicology & Applied Pharmacology 177, 264–271.
Evaluation of impact of PAH on a tropical fish, Oreochromis mossambicus using multiple biomarkers M.S. Shailaja *, Classy DÕSilva National Institute of Oceanography, Dona Paula, Goa 403004, India Received 3 February 2003; received in revised form 26 June 2003; accepted 2 July 2003
Abstract Treatment of the widely occurring tropical cichlid, Oreochromis mossambicus with a pure polycyclic aromatic hydrocarbon (PAH), phenanthrene, induced a concentration-dependant formation of the enzyme, 7-ethoxyresorufin-Odeethylase (EROD). Concomittant increase (65–669%) in the activity of sorbitol dehydrogenase in the serum (SSDH) occurring with EROD induction denoted liver cell damage, which was more severe in fish exposed to lower concentrations (0.4–4 lg g1 ) of the chemical. In O. mossambicus exposed to 25% refinery effluent liver damage associated with cell death was indicated by the twin analyses of SSDH and liver somatic index. Cell injury appeared to have occurred at low PAH concentrations due to inadequately induced phase II-related detoxification of metabolites. This was indicated by the nearly 33% higher activity of hepatic glutathione S-transferase (GST) in fish exposed to higher PAH concentrations as compared to low-exposure animals. Tilapia such as O. mossambicus were found to be eminently suited for biomonitoring in tropical coastal waters. A combination of EROD and serum sorbitol dehydrogenase activity measurements serves as an excellent tool for biomonitoring sublethal effects of PAH pollution in fish. 2003 Elsevier Ltd. All rights reserved. Keywords: Tilapia; Polycyclic aromatic hydrocarbon (PAH); Sublethal effects; 7-Ethoxyresorufin-O-deethylase (EROD); Glutathione S-transferase (GST); Serum sorbitol dehydrogenase
1. Introduction Aqueous environments around the world, estuaries and coastal waters in particular, receive large inputs of anthropogenic pollutants through industrial and urban discharges, atmospheric deposition and terrestrial drainage. With the steady growth of industrial activities in the developing countries, environmental degradation has become a potential problem. Industrial and municipal wastes contain numerous chemicals, both organic [e.g., polycyclic aromatic hydrocarbons (PAHs), pesticides, polychlorinated biphenyls (PCBs), etc.] and inor-
*
Corresponding author. Tel.: +91-832-45-6700; fax: +91832-245-6702. E-mail address: [email protected] (M.S. Shailaja).
ganic (viz., heavy metals such as Hg, Cd, Pb and Cu), capable of causing deleterious effects in aquatic biota. PAHs are potentially carcinogenic, ubiquitous contaminants of much ecotoxicological concern in the coastal region. They induce the formation of mixed function oxygenase (MFO), specifically the isozyme, cytochrome P4501A1 (CYP1A1) in the fish liver which has been employed as a biomarker of exposure to hazardous organic pollutants in monitoring studies (Spies et al., 1982; Payne et al., 1987; Stegeman and Lech, 1991; Collier et al., 1992a; Goksøyr and F€ orlin, 1992; Livingstone, 1993). While the MFO system is essential for the biotransformation of PAHs, its induction could produce damaging side effects through the formation of intermediates that are highly reactive, mutagenic and carcinogenic (Gelboin, 1980; Buhler and Williams, 1988; Stegeman and Lech, 1991). Indeed, a strong linkage has
0045-6535/$ - see front matter 2003 Elsevier Ltd. All rights reserved. doi:10.1016/S0045-6535(03)00667-2
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been shown between the prevalence of several toxicopathic hepatic lesions and CYP1A induction in english sole (Myers et al., 1998). To ensure the health of the coastal oceans around them developing nations in the tropical regions in particular, need to adopt viable monitoring programmes. However, deficiencies in information often limit their ability to do so. For instance, data on biomarker responses to organic contaminants in tropical fish are few (Stegeman et al., 1990, 1997; Gadagbui and Goksoyr, 1996; Pathiratne and George, 1996, 1998; Al-Arabi and Goksoyr, 1998; Bainy et al., 1999; Shailaja and Rodrigues, 2001, 2003) whereas such information is absolutely essential for embarking on PAH monitoring with a suitable species of fish. This paper describes our studies with Oreochromis mossambicus, a widely distributed tropical euryhaline fish exposed in the laboratory to a pure PAH compound (phenanthrene) and a refinery effluent containing a mixture of PAHs with the objective of assessing the suitability of O. mossambicus for biomonitoring PAH impact using a suite of biomarkers. The enzyme biomarkers assayed included the phase I enzyme, CYP1A1 represented by 7-ethoxyresorufin-O-deethylase (EROD), phase II enzyme, glutathione S-transferase (GST) and serum sorbitol dehydrogenase, an indicator of liver cell damage (Gerlach, 1983). Liver somatic index (LSI) was used as a physiological biomarker while biliary fixed wavelength fluorescence was determined as an indicator of PAH metabolites.
2. Materials and methods Juvenile O. mossambicus (weight 49–67 g) were obtained from a local estuarine fish farm. These fish having a high tolerance towards salinity changes were maintained in fresh water for seven days before being administered a single intra-abdominal injection of phenanthrene (Sigma, USA). Due to the small size of the experimental fish, the volume injected into each individual fish in the low and high exposure groups, respectively, consisted of 100 ll of 0.2 or 2 mg ml1 solution of phenanthrene in sunflower oil. The exact dose received by the fish was in the range of 0.4–32 lg g1 , based on the weight of individual fish. Control fish received equivalent amounts (100 ll) of the carrier alone. The experimental fish were maintained in 30 l glass aquaria (up to n ¼ 6, per tank) with continuous aeration and the water being changed every 24 h. Commercial feed was provided ad libitum during the period of experiment. The animals were sacrificed three days after exposure. A second experiment consisted of exposing O. mossambicus (66–71 g body weight; n ¼ 6) to a secondarytreated oil refinery effluent (25% dilution) for two weeks.
The undiluted effluent had a total PAH content of 478 ± 24 lg l1 . The test medium was renewed every alternate day, all other conditions remaining the same as described earlier. Untreated fish (n ¼ 6) formed the control. At the conclusion of each experiment, liver samples were obtained by immediate dissection and homogenates prepared from fresh livers for EROD assay according to the procedure of Gunther et al. (1997). EROD activity was determined in the postmitochondrial supernatant (PMS) with NADPH as the electron donor and resorufin (85 nM) as internal standard (Stagg and Addison, 1995). The activity is reported in terms of nmol resorufin formed min1 mg1 protein. GST activity in the PMS of liver extract was evaluated as the formation of a conjugate between glutathione and 1-chloro-2,4-dinitrobenzene (CDNB) at 340 nm (extinction coefficient ¼ 9.6 cm1 mM1 ) in 50 mM potassium phosphate buffer, pH 6.5 with 1 mM EDTA (Habig et al., 1974). The activity of the enzyme is reported as the number of micromoles of conjugate formed per min mg1 protein. Sorbitol dehydrogenase (SSDH) activity was measured as the rate of reduction of D -fructose by NADH (Gerlach, 1983) in composite serum samples obtained from 2 to 3 fishes by cardiac puncture. One unit of SSDH activity refers to a change in absorbance of 0.001 min1 at 366 nm in 3 ml of the reaction mixture. Protein was estimated in the PMS by the method of Lowry et al. (1951) using bovine serum albumin as the standard. PAH concentrations in the liver were estimated fluorimetrically (Shimadzu spectrofluorophotometer RF-1501) after a 4-h saponification of the tissue using alkaline methanol (Law and Whinnett, 1992) and subsequent hexane extraction. Fluorescent aromatic compounds (FACs) in the bile were determined in fish exposed to refinery effluent by fixed wavelength fluorescence measurements at specified excitation/emission wavelength pairs for selected PAH metabolites (Krahn et al., 1987, 1993). LSI was calculated as (weight of liver/total body weight) · 100. Where statistical evaluation was required, correlation and F-tests were performed using the Analysis Tool Kit of Excel 2000 (Microsoft). p < 0:05 was considered as statistically significant.
3. Results and discussion 3.1. Phenanthrene concentrations and EROD activity in the liver Hepatic concentrations of phenanthrene in O. mossambicus exposed to high (6–32 lg g1 ) and low
M.S. Shailaja, C. DÕSilva / Chemosphere 53 (2003) 835–841
µM Phenanthrene g-1 fish
25
Control Low High
20
* *
15 10 5 0 Phenanthrene exposure
Fig. 1. Hepatic concentrations of phenanthrene (mean ± SD) in O. mossambicus treated with low (0.4–4 lg g1 , n ¼ 10) and high (6–32 lg g1 , n ¼ 12) concentrations of phenanthrene. Control fish (n ¼ 5) were treated only with sunflower oil. Values significantly different from the control (p < 0:05) are marked with asterisk.
(0.4–4 lg g1 ) levels of the compound are presented in Fig. 1. Compared to control fish, both the treatment groups had significantly higher phenanthrene levels in the liver (p < 0:003 and <0.0004, respectively, F test). However, between test groups the values were not significantly different (p > 0:05). Further, a significant inverse relationship was observed between phenanthrene levels in the liver and exposure concentrations (r ¼ 0:585, p < 0:02) suggesting a concentration-dependant increase in the induced metabolism of PAH. The large (35–38%) deviations from the mean observed in the hepatic concentrations of phenanthrene in the two test groups (Fig. 1) were likely due to biotransformation of PAH. Mean hepatic EROD activities in the control, lowexposure and high-exposure groups of fish were, respectively, 0.721 ± 0.045, 5.42 ± 2.08 and 17.31 ± 5.45 nmol resorufin min1 mg1 protein (Table 1). In general,
in the two groups of exposed fish, based on the lowest and highest values, EROD activity (range 1.36–28.49 nmol resorufin min1 mg1 protein) was 2–40 times higher than the control value. Also, the induced activity correlated well (r ¼ 0:799, p < 0:001) with phenanthrene exposure levels more so when the concentrations were >18 ppm (Fig. 2). However, the relationship between phenanthrene concentration in the liver and enzyme activity was not significant (p > 0:05). Induction of CYP1A in fish by small molecular size PAHs (having 64 fused aromatic rings) such as chrysene, phenanthrene, pyrene and naphthalene is irregular and species dependant. For example, phenanthrene did not induce CYP1A biotransformation activity in a rainbow trout liver cell line (Bols et al., 1999) but did so in cod and other bony fishes such as flounder (Goksøyr et al., 1986) and scup (Stegeman et al., 1998). Similarly, pyrene and chrysene did not evoke EROD activity in the rainbow trout (Bols et al., 1999) but again, both of them did so in the carp (van der Weiden et al., 1994). It is thus evident that the response to PAHs in particular, to
EROD activity (nmol resorufin min-1 mg-1 protein)
30
837
30
R = 0.799 20
10
0 0.1
1
10
100
-1 Phenanthrene exposure (µg g fish)
Fig. 2. Exposure dependence of EROD induction in O. mossambicus injected with phenanthrene (n ¼ 21). EROD activity values in control fish are given in Table 1.
Table 1 Ethoxyresorufin-O-deethylase (EROD), glutathione S-transferase (GST) and serum sorbitol dehydrogenase (SSDH) activities in O. mossambicus exposed to phenanthrene Phenanthrene exposure level
ERODa
GSTb
SSDHc
Control
0.72 ± 0.04 (n ¼ 4) 5.42 ± 2.08* (n ¼ 9) 17.31 ± 5.45* (n ¼ 12)
1.932 ± 0.302 (n ¼ 4) 2.255 ± 0.566 (n ¼ 9) 2.994 ± 0.332* (n ¼ 12)
0.67 ± 0.28 (n ¼ 5) 30.09 ± 17.2* (n ¼ 10) 8.67 ± 5.99* (n ¼ 11)
Low (0.4–4 lg g1 ) High (6–32 lg g1 )
Values are mean ± SD and marked with an asterisk when significantly different from the control value (p < 0:05). a nmol resorufin min1 mg1 protein. b mol product min1 mg1 protein. c Units ml1 serum.
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compounds having <4 fused aromatic rings, is not uniform among different fish species nor within the same species. This could be due to intrinsic variations in the affinity of the aryl-hydrocarbon receptor (AhR) of different fish species to small molecular size PAH compounds. Further, large variations in the induction of EROD activity between tropical and temperate fishes (Oreochromis niloticus and rainbow trout, respectively) have also been reported (Pathiratne and George, 1996). 3.2. Hepatic phase II enzyme activity The metabolites formed by phase I biotransformation activity are conjugated via phase II enzymes (e.g., GST) before excretion. In O. mossambicus exposed to 6– 32 lg g1 phenanthrene GST activity (2.994 ± 0.332 lmol min1 mg1 protein, Table 1) was 32.8% higher (p < 0:05) than in the fish subjected to low (<4 lg g1 ) phenanthrene exposure (2.255 ± 0.566 mol min1 mg1 protein). 3.3. Sorbitol dehydrogenase activity in serum Serum sorbitol dehydrogenase activity is a sensitive biochemical indicator of chemically induced liver damage in fish (Dixon et al., 1987). SSDH activity (mean value 20.21 ± 14.31 units min1 ml1 serum) was substantially increased on exposure of O. mossambicus to phenanthrene (0.4–32 lg g1 ) as compared to control fish (0.67 ± 0.28 units min1 ml1 serum), indicating liver cell injury. Also, SSDH activity showed a very distinct inverse correlation (p < 0:01) with PAH exposure levels (Fig. 3). Significantly, the damage was greater in the fish exposed to lower concentrations than in those treated with >18
3.4. Exposure of O. mossambicus to refinery effluent
SSDH activity (units ml-1 serum)
100 80
R = - 0.797
60 40 20 0 0.1
lg g1 phenanthrene as observed from the nearly 3 times higher SSDH values in the former (Table 1). As the liver cell injury apparently was greater in the fish exposed to low concentrations of the toxicant despite the insignificant difference in the hepatic concentrations of phenanthrene in the two test groups (Fig. 1) it is presumed that the metabolic products of biotransformation rather than the parent compound, phenanthrene were responsible for the observed liver damage. The activity of hepatic GST which plays a major role in detoxifying xenobiotics was nearly 33% higher in the fish exposed to higher PAH concentrations as compared to the low-exposure animals. Therefore, possibly, a mismatch in the kinetics of induction of phase I (bioactivation) and phase II (conjugation/ detoxification) systems in O. mossambicus exposed to low PAH concentrations resulting in an inadequate conjugation potential leads to accumulation of highly reactive oxygenated metabolites in the liver and subsequently, tissue damage. This is somewhat similar to the case of english sole, a flatfish species showing a higher predisposition to contaminant-associated hepatic neoplasms versus starry flounder. The activity of the PAH-activation MFO, aryl hydrocarbon hydroxylase in english sole was reported to be higher just as GST was lower than in starry flounder (Collier et al., 1992b). Likewise, Willett et al. (2000) have hypothesized that the observed difference in the sensitivity to PAH-mediated liver cancer in channel catfish (less sensitive) and brown bullhead (more sensitive) is due to differences in their hepatic phase II activities. We are in the process of testing other PAHs including small molecular size compounds such as anthracene and chrysene for similar effects.
1 10 Phenanthrene exposure
100
(µg g-1 fish)
Fig. 3. Relationship between sorbitol dehydrogenase activity in serum of O. mossambicus (n ¼ 21) and phenanthrene exposure concentration. Data points represent SSDH values measured in composite (n ¼ 2 or 3) serum samples expressed against corresponding pooled exposure concentrations. Values of SSDH in control fish are given in Table 1.
Changes in biochemical (EROD and SSDH) and physiological (LSI) parameters measured in O. mossambicus exposed to 1:4 diluted refinery effluent are shown in Fig. 4. Compared to control, EROD was enhanced by 375% (significance 99.9%) in fish exposed to the effluent having a total PAH content of approximately 120 ppb. Serum SDH which was almost nil in control fish measured 73 units in the exposed fish. LSI, on the other hand, was reduced by nearly 27% which was significant (p < 0:05). LSI values are generally elevated in vertebrates experiencing induction of hepatic microsomal P-450 for detoxification of organic compounds (Huuskonen and Lindstr€ om-Seppa, 1995). Hence, the observed reduction in LSI when considered together with increased SSDH activity would indicate hepatocellular injury associated with cell death. It is suspected that the toxic impact of the effluent stemmed essentially from benzo(a)pyrene. The metabolites present in the bile were examined by fixed fluores-
M.S. Shailaja, C. DÕSilva / Chemosphere 53 (2003) 835–841 8 EROD activity SSDH activity (x 10-1) Liver Somatic Index (%)
EROD
6
LSI
SSDH
*
4
* 2
0 Control
839
ticularly valid since reports indicate lowering of CYP1A1 activities when contamination is high (Peters et al., 1994; Celander et al., 1996) while some PAHs (e.g., fluoranthene) could inhibit CYP1A dependant activity (Willett et al., 2001). CYP1A-dependant bioassays may therefore lead to an underestimation of PAH exposures. Besides, lowered EROD induction is also associated with degenerating, preneoplastic and neoplastic livers which could be misleading if employed as the only biomarker and can only be interpreted in relation to the state of health of the liver (K€ ohler and Pluta, 1995).
Exposed
Fig. 4. Impact of exposure of O. mossambicus to 25% refinery effluent. EROD: 7-ethoxyresorufin-O-deethylase (nmol resorufin min1 mg1 protein); SSDH: serum sorbitol dehydrogenase (units min1 ml1 serum); LSI: liver somatic index (%). n ¼ 6 for EROD and LSI; one composite sample (n ¼ 3) for SSDH. Values shown with asterisk were significantly different (p < 0:05) from the control value.
cence measurements at wavelength pairs specific for naphthalene (290/335 nm), pyrene (341/383 nm), benzo(a)pyrene (380/430 nm) and phenanthrene (260/380 nm). Significant differences compared to the values in control fish were seen only in the BaP-type metabolites (Table 2). This study is the first in tropical fish linking liver damage to PAH biotransformation. Our observation that relatively low concentrations of PAH, including non-carcinogenic ones such as phenanthrene, could lead to sublethal hepatic cell injury in fish is important because the liver is an essential organ having many vital physiological functions including that of production of yolk proteins during oocyte development. Secondly, the apparent sublethal toxicity of low levels of innocuous PAHs would also mean a potential enlargement of the sphere of influence of discharge points. Our findings strongly support the collaborative use of enzyme biomarkers such as CYP1A- and SSDH-activities for biomonitoring PAH impact in fish. This is par-
3.5. Tilapia––a tropical fish for PAH effect monitoring The clear relationship seen between EROD activity and exposure to PAH in O. mossambicus qualifies them as suitable organisms for pollution monitoring especially considering their wide geographical distribution, diverse food habits and their ability to grow in both fresh and marine waters. Besides, the induction of CYP1A activity in tilapia is not influenced by sexual maturation (Pathiratne and George, 1998), as is the case with several temperate species used in biomarker studies (Lindstr€ omSeppa and Stegeman, 1995). Tilapia have also been found to be well suited for performing field caging experiments (Gadagbui and Goksoyr, 1996). However, CYP1A induction is a non-specific biomarker (it can be induced by PCBs and PCDDs, besides PAHs) but could be made relevant to PAHs by studying it in combination with biliary FACs specific for PAH metabolites (Beyer et al., 1996).
4. Conclusions The MFO system concerned with biotransformation of PAHs is highly active in the tilapia, O. mossambicus. Thus, tilapia make good sentinel organisms for PAH pollution monitoring in tropical waters. Exposure to low concentrations of PAH (<4 lg g1 ) leads to sublethal hepatic toxicity in fish as shown by the increased activity
Table 2 Fixed fluorescence evaluation of biliary PAH metabolites of O. mossambicus exposed to 25% refinery effluent Bile metabolites B(a)P-type Naphthalene-type Pyrene-type Phenolphthalene-type
Fluorescence intensitya Control fish
Effluent-exposed fish
380 ± 30 240 ± 16 26 ± 4 314 ± 21
670 ± 18* 240 ± 22 27 ± 4 354 ± 15
Values are mean ± SD; n ¼ 6 for control and effluent exposed fish. [Excitation/emission k used for B(a)P-, naphthalene-, pyrene- and phenanthrene-type metabolites were 380/430 nm; 290/335 nm; 341/383 nm and 260/380 nm, respectively.] *Value significantly different from the control value (p < 0:05). a Arbitrary units.
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of serum sorbitol dehydrogenase, an indicator of liver damage. The use of multiple enzyme biomarkers such as CYP1A- and SSDH-activities in combination with bile FACs is especially recommended for biomonitoring PAH impact in fish, considering the ease of analyses of the parameters. Acknowledgements We thank Nancy Rodrigues, Angelo Rodrigues and Christine Fernandes for their excellent technical assistance. This is National Institute of Oceanography contribution No. 3838. References Al-Arabi, S.A.M., Goksoyr, A., 1998. Species characteristics of hepatic biotransformation enzymes in two tropical fish of Bangladesh, one riverine catfish (Rita rita) and one marine mudskipper (Apocryptes bato). Marine Environmental Research 46, 121–122. Bainy, A.C.D., Woodin, B.R., Stegeman, J.J., 1999. Elevated levels of multiple cytochrome P450 forms in tilapia from Billings Reservoir––Sao Paolo, Brazil. Aquatic Toxicology 44, 289–305. Beyer, J., Sandvik, M., Hylland, K., Fjeld, E., Egaas, E., Aas, E., Skare, J.U., Goksøyr, A., 1996. Contaminant accumulation and biomarker responses in flounder (Platichthys flesus L.) and Atlantic cod (Gadus morhua L.) exposed by caging to polluted sediments in Soerfjorden, Norway. Aquatic Toxicology 36, 75–98. Bols, N.C., Schirmer, K., Joyce, E.M., Dixon, D.G., Greenberg, B.M., Whyte, J.J., 1999. Ability of polycyclic aromatic hydrocarbons to induce 7-ethoxyresorufin-O-deethylase activity in a trout liver cell line. Ecotoxicology & Environmental Safety 44, 118–128. Buhler, D.R., Williams, D.E., 1988. The role of biotransformation in the toxicity of chemicals. Aquatic Toxicology 11, 19–28. Celander, M., Stegeman, J.J., Forlin, L., 1996. CYP1A1-, CYP2B- and CYP3A-like proteins in rainbow trout (Oncorhynchus mykiss) liver: CYP1A1-specific down-regulation after prolonged exposure to PCB. Marine Environmental Research 42, 283–286. Collier, T.K., Connor, S.D., Eberhart, B.L., Anulacion, B.F., Goksoyr, A., Varanasi, U., 1992a. Using cytochrome P450 to monitor the aquatic environment: initial results from regional and national surveys. Marine Environmental Research 34, 195–199. Collier, T.K., Singh, S.V., Awasthi, Y.C., Varanasi, U., 1992b. Hepatic xenobiotic metabolizing enzymes in two species of benthic fish showing different prevalences of contaminant–– associated liver neoplasms. Toxicolology & Applied Pharmacology 113, 319–324. Dixon, D.G., Hodson, P.V., Kaiser, K.L.E., 1987. Serum sorbitol dehydrogenase activity as an indicator of chemically induced liver damage in rainbow trout. Environmental Toxicology & Chemistry 6, 685–696.
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