- Email: [email protected]
Biochemical changes in rockfish, Sebastes schlegeli, exposed to dispersed crude oil
Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223
Contents lists available at ScienceDirect
Comparative Biochemistry and Physiology, Part C j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / c b p c
Biochemical changes in rockfish, Sebastes schlegeli, exposed to dispersed crude oil Jee Hyun Jung, Un Hyuk Yim, Gi Myeong Han, Won Joon Shim ⁎ Oil and POPs Research Group, KORDI, Geoje-shi, 656-834, Republic of Korea
a r t i c l e
i n f o
Article history: Received 3 February 2009 Received in revised form 21 April 2009 Accepted 24 April 2009 Available online 3 May 2009 Keywords: Oil dispersant Rockfish CYP1A Ethoxyresorufin O-de-ethylase Acethylcholinesterase Ovoviviparous
a b s t r a c t This paper describes the response of the ovoviviparous rockfish, Sebastes schlegeli, to hydrocarbons in the water-accommodated fraction (WAF) of crude oil, in the presence or absence of oil dispersants. Concentrations of cytochrome P-450 1A (CYP1A) and levels of its catalytic activity ethoxyresorufin O-de-ethylase (EROD) in rockfish exposed to WAF at concentrations of 0.1% and 1% were significantly increased by the addition of a dispersant, Corexit 9500 after 48 h exposure. After 72 h exposure, the levels of CYP1A and EROD activity were significantly increased in 0.1% and 0.01% chemically enhanced WAF (CEWAF) (Corexit 9500 and Hiclean II dispersant). Bile samples from fish exposed to WAF alone had low concentrations of hydrocarbon metabolites, exemplified by 1-hydroxypyrene. After 72 h exposure, hydrocarbon metabolites in bile from fish exposed to WAF in the presence of either Corexit 9500 or Hiclean II were significantly higher compared with fish exposed to WAF alone or control fish. These experiments confirm that the use of oil dispersants will increase the exposure of ovoviviparous fish to hydrocarbons in oil. © 2009 Elsevier Inc. All rights reserved.
1. Introduction Oil dispersants are potential tools for minimizing the impact of marine oil spills. The dispersants are primarily surfactants that reduce the interfacial tension between the oil and dispersant from micelles that are dispersed into the water column (Canevari, 1978). They do not actually remove the oil. Eventually, it will be broken down by microscopic organisms in form of biodegradation. However, oil dispersant use involves the addition of potentially damaging chemicals to an already stressed environment, which can lead to further problems. Based on a review of the 258 marine incidents that the International Tanker Owners Pollution Federation (ITOPF) were involved with between 1995 and 2005, 46 (18%) involved the use of chemical dispersants at sea. A review of the use of dispersants in different international regions for the 46 cases considered by the ITOPF shows that over half took place in Asian waters (Chapman et al., 2007). On the 7th December 2007, the M/V Hebei Spirit spilled more than 12 million liters of three kinds of crude oil off the west coast of Korea, in a region of high aquaculture and fisheries activity. About 250 kL of oil dispersant were used in the cleanup operations. There is ongoing controversy concerning the impacts of oil dispersant use on marine biota. The observed increased toxicity of dispersed oil has been attributed to dispersed particle sizes (Bobra et al., 1989) and aromatic hydrocarbon concentration (Anderson et al., 1974). Hydrocarbons are accumulated via the gills and mouth and the dispersed fraction would therefore be the
⁎ Corresponding author. Oil and POPs Research Group, Korean Ocean Research and Development Institute, 391 Jangmok-Ri Jangmok-myon, Geoje, Korea. Tel.: +82 55 639 8687; fax: +82 55 639 8689. E-mail address: [email protected] (W.J. Shim). 1532-0456/$ – see front matter © 2009 Elsevier Inc. All rights reserved. doi:10.1016/j.cbpc.2009.04.009
most potentially bioavailable (Thomas and Rice, 1981). Fish can accumulate soluble petroleum hydrocarbons very rapidly (Collier et al., 1995), and these contaminants can concentrate in tissues to factors 10–100 times greater than in water (Ramachandran et al., 2006). Hydrocarbon residues in fish can exert various types of effects. They can alter gonad size, gametogenesis (Nicolas, 1999; Tintos et al., 2007) and brain neurotransmission (Gesto et al., 2006). Some polynuclear aromatic hydrocarbons (PAHs) induce hepatic tumorgenicity in fish (Sonstergard and Leatherland, 1984; McCain et al., 1987; Meyers et al., 2002). Black et al. (1985) have shown that fish from PAHs-polluted waters show a high incidence of liver tumors compared with populations from uncontaminated waters in the same region. Assessment of the biological impacts of oil dispersants on marine ecosystems is usually based on acute toxicity tests (Singer et al., 1993; Cotou et al., 2001). Chemically induced alterations in the biochemistry and physiology of organisms are being used as diagnostic tools in assessing potential biochemical impacts in aquatic systems. However, there are few studies of the impacts of oil and oil dispersants on fish biochemical processes. The induction of cytochrome P-450 1A (CYP1A) enzyme is well established as a “biomarker” to monitor exposure to aromatic hydrocarbons (Payne et al., 1984, 1987). Under actual oil spill conditions in the marine environment, it is unclear how oil and dispersants will combine to act agonistically or antagonistically on fish physiological systems. In this study, the effect of two dispersants on the uptake and metabolism of hydrocarbons and their effects on selected biochemical responses in rockfish were examined. This species was categorized as an ovoviviparous fish, because development of the late embryonic stages is dependent on maternal nutrition as an energy source, in addition to that stored in yolk accumulated in oocytes during vitellogenesis (Boehlert,
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223
1991). Therefore, it is interesting to clarify the impact on ovoviviparous biochemical responses that dispersed oil may alter the hepatodetoxification and neurotransmitter processes. Corexit 9500 is a more recent formulation which provides enhanced penetration and emulsion properties compared with Corexit 9527. In Korea, a common oil dispersant is Hiclean II, which is a known developed product based on a carbonated-aqueous surfactant. Their effects on enhancing hydrocarbon uptake (shown by the presence of PAH metabolites in bile), and on the induction of CYP1A, EROD and inhibition of brain acethylcholinesterase (AChE) are described. 2. Materials and methods 2.1. Corexit 9500 and Hiclean II Corexit 9500 was purchased from Nalco Corporation (Texas, USA); its 96 h LC50 is reported to be 25.2 mg/L for inland silversides (technical note of Nalco). Corexit 9500 is based on aliphatic hydrocarbons, propylene glycol, and some sulfuric acid salt. Hiclean II is a carbonated-aqueous surfactant, which was purchased from Daeil Corporation (Busan, Korea); its 24 h LC50 to rockfish is 2000 mg/L (Technical note of Daeil). 2.2. Water-accommodated fraction (WAF) of crude oil and CEWAF Crude ARM (medium grade, Saudi Arabia) was donated from the laboratory of the Korean Coast Guard. Crude oil and water were mixed at a ratio of 1:49 v/v in sealed containers with minimum head space for 18 h at 18 °C. The vortex was adjusted to no more than a third of the height of the mixture from the oil–water interface (Singer et al., 2000). The mixture was allowed to settle for 1 h for the separation of water and oil phase and the water phase was drained off for the experiment. The dispersants, Corexit 9500 and Hiclean II were added to the surface of the oil–water mixture at the recommended ratio of 1:20 dispersant oil: (technical note of Nalco and Daeil, Gilbert,1996; Ramachandran et al., 2004) and stirred for 1 h. 2.3. Exposure experiments Juvenile rockfish (Sebastes schlegeli), 2 years old,180± 2.0 g (mean± S.D.) weight and 20 ± 2.0 cm length were purchased from a local fish farm in Geoje, Korea. Separate control fish were injected with ethanol or not injected. Fish were acclimated for 1 week in flowing seawater at 16 °C on a 16 h light/8 h dark photoperiod in the laboratory before the beginning of experiments to make sure they were disease free. Water quality variables, including ammonia, nitrate and dissolved oxygen were
Table 1 Experimental summary of exposure groups. Experimental group Treatment CON 0.01 WAF 0.1 WAF 0.01 CEWAFC 0.1 CEWAFC 0.01CEWAFH 0.1CEWAFH DC DH β-NF
Exposure type
100 µL of DMSO (dimethyl sulfoxide) 0.01% WAF (made by crude oil)
Injection Water exposure 0.01% WAF (made by crude oil) Water exposure 0.01% CEWAF (crude oil dispersed by Corexit Water 9500) exposure 0.1% CEWAF (crude oil dispersed by Corexit Water 9500) exposure 0.01% CEWAF (crude oil dispersed by Hiclean II) Water exposure 0.1% CEWAF (crude oil dispersed by Hiclean II) Water exposure 0.1% dispersant (Corexit 9500) Water exposure 0.1% dispersant (Hiclean II) Water exposure 10 µg/kg of β-naphthoflavone (in DMSO) Injection
Note: CON = control; other abbreviations represent experimental chemicals.
219
Table 2 Experimental conditions for rockfish, S. schlegeli. Fish mass (g) 264 ± 26.3
HSI 2.54 ± 0.27
pH 8.4 ± 0.04
Temperature
Dissolved oxygen
(°C)
(% saturation)
14 ± 0.3
97.8 ± 2.1
HSI = hepatosomatic index. Data are means ± standard error of mean (SEM).
monitored daily. Six and seven fish were exposed to series concentrations of WAF (v/v), CEWAFC (WAF treated with Corexit 9500), CEWAFH (WAF treated with Hiclean II) and two 0.1% dispersant exposure groups (Corexit 9500 and Hiclean II) in 60 L water. β-napthoflavone of 10 µg/kg, a known CYP1A inducer, was injected into one group of fish as a positive control (Table 1). Exposure concentrations causing CYP1A induction and lethality were considered. Each experiment was performed continuousflow system which is controlled a light and temperature. They contained a central cylinder where a Teflon-coated main propeller. Exposure concentrations of WAF and CEWAF were 0.1 and 0.01%. The exposure duration was fixed at 48 h and 72 h exposure, which caused comparable CYP1A induction. Immediately after movement ceased, the fish were killed by a blow on the head, weighed, measured, and dissected and various tissues removed for analysis. Tissue samples were stored at − 80 °C until analyzed. 2.4. Fluorescent aromatic hydrocarbon (FAC) metabolites Synchronous scanning of fluorescent bile metabolites, focusing on 1hydroxypyrene (Krahn et al., 1987) was used as a rough indicator of the uptake of PAH by the fish. Frozen gall bladders were disrupted to release the bile, the tissue mass was centrifuged briefly to sediment out unwanted tissue, and a bile sample was diluted 50- or 100-fold in 50% ethanol. Diluted bile samples were scanned with excitation between 300 and 390 nm and emission between 337 and 427 nm (i.e., a wavelength separation of 37 nm: Ariese et al., 1993). 1-hydroxypyrene (Aldrich) was used as a standard at concentrations (in 50% ethanol) in the range of 50–500 nM. 2.5. Enzyme activity Brain tissue samples were thawed on ice and were homogenized and analyzed for AChE activity using a microplate reader adaptation of the Ellman procedure (Ellman et al., 1961) as described by Jung et al. (2006, 2008). Liver samples were homogenized and microsomes were prepared and suspended in 0.1 M phosphate buffer pH 7.6 essentially as described by Addison et al. (1994). Ethoxyresorufin O-de-ethylase (EROD) activity in microsomes was determined using reagent concentrations described by Addison and Payne (1986) in a fluorescence microplate reader with the excitation and emission filter set at 544 nm and 590 nm, respectively. Protein concentrations in the sample were assayed by the bicinconchinic acid (BCA) method using a kit from Pierce Chemicals (Rockford IL); the product was read in a microplate reader at 550 nm, using bovine serum albumin (BSA) as a standard. 2.6. CYP1A protein expression CYP1A was estimated by western blotting using the antibody raised against a 22 amino-acid oligopeptide fragment conserved in several types of vertebrate CYP1A as described by Lin et al. (1997). Liver microsomal proteins were separated by SDS-PAGE and transferred to PVDF nitrocellulose membranes. Nonspecific binding sites on the membrane were blocked with 0.5% goat IgG in TBST (20 mM Tris–HCl pH 8.0, 100 mM NaCl, 0.05% Tween 20) for 2 h at 37 °C. The CYP1A antibody was diluted 1:5000 in TBST and incubated with the membrane for 1 h at room temperature. After washing, the membrane was incubated with a secondary antibody (goat anti-mouse IgG biotin conjugate)
220
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223
for 30 min at room temperature. The blots were developed using an ABC kit (Vector Laboratories, Burlingame, CA) for 30 min. The substrate solution contained 0.02% 3,3′-diaminobenzidine tablets (Sigma) and the development was stopped with distilled water. The reliability of the CYP1A (protein) and EROD measurements was assessed during the BEQUALM intercalibration (Biological Effects Quality Assurance in Monitoring Programs; http://www.bequalm. org).exercise; results were within approx. 1 SD of the mean estimates of nine participating laboratories.
Fig. 2. Hepatodetoxification activity in rockfish exposed to a range of WAF, CEWAF concentrations from crude oils and 10 µg/kg β-napthoflavone after 72 h treatment. (a) Level of CYP1A mRNA expression in rockfish liver based on northern blot analysis of liver samples from individual fish. Relative mRNA levels of rockfish CYP1A were normalized with β-actin. (b) Level of CYP1A (nmol) protein in rockfish liver based on western blot analysis of microsome from individual fish. (c) Level of EROD activity (nmol/ min/mg protein) in rockfish liver microsome from individual fish Data are means ± standard error of mean (SEM) (n = 6–7). Data were subjected to ANOVA followed by Duncan's multiple-range test. Different letters denote statistically significant (P b 0.05).
Fig. 1. Hepatodetoxification activity in rockfish exposed to a range of WAF, CEWAF concentrations from crude oils and 10 µg/kg β-napthoflavone after 48 h treatment. (a) Level of CYP1A mRNA expression in rockfish liver based on northern blot analysis of liver samples from individual fish. Relative mRNA levels of rockfish CYP1A were normalized with β-actin. (b) Level of CYP1A (nmol) protein in rockfish liver based on western blot analysis of microsome from individual fish. (c) Level of EROD activity (nmol/ min/mg protein) in rockfish liver microsome from individual fish. Data are means± standard error of mean (SEM) (n = 6–7). Data were subjected to ANOVA followed by Duncan's multiple-range test. Different letters denote statistically significant (P b 0.05). DC (treated with Corexit 9700), DH (treated with Hiclean II), CEWAFC (crude oil dispersed with Corexit 9700) and CEWAFH (crude oil dispersed with Hiclean II).
2.7. CYP1A mRNA expression cDNA probes were generated by transcription with 5.0 µL of [α− 32P] dATP (Boston, MA NEN) using a random primer labeling kit (Takara, Japan). CYP1A mRNA expression levels were estimated by northern blot analysis as described by Jung et al. (2005, 2006). Templates for each probe were made from cloned CYP1A and β-actin fragments that had been submitted to GenBank in our previous study (accession no. AY 558556; AY 166590).
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223
221
2.8. Statistical analysis All data are expressed as mean± SE. Data were compared by analysis of variance (ANOVA) and significant differences were calculated by Duncan's multiple-range tests. 3. Results 3.1. Test condition for the rockfish 72 h static bioassay with daily renewal Water quality variables (ammonia, nitrate and dissolved oxygen) were monitored daily. These variables did not change over the experimental period and there were no differences in body weight and length for the exposure duration among experimental groups. Fish Hepato Somatic Index (HSI) values are shown in Table 2 and there was no significant difference among experimental groups. 3.2. CYP1A expression and EROD activity in fish exposed to WAF and CEWAF Exposure concentrations causing CYP1A induction and lethality were determined from previous studies (Ramachandran et al., 2004, 2006). A concentration of 0.1% is below the lethal dose. Fig. 1 (a, b), showed that levels of CYP1A mRNA and protein expression in rockfish exposed to CEWAFC (Corexit 9500) increased significantly in response to a concentration of 0.1% (v/v) after 48 h. They were higher than CEWAFH (Hiclean II) and the control group and close to the positive
Fig. 4. Level of bile 1-hydroxypyrene metabolites in rockfish: (a) after 48 h treatment, (b) after 72 h treatment. Data were subjected to ANOVA followed by Duncan's multiplerange test. Asterisks indicate means that were significantly different compared with exposed fish (⁎⁎P b 0.01, ⁎ P b 0.05).
control group level. Fig.1c showed that the EROD activity of fish exposed to CEWAF (Corexit 9500) increased significantly compared with fish exposed to WAF, CEWAFH and the negative control group. After 72 h of exposure, levels of CYP1A mRNA and protein expression were significantly increased at both doses of 0.1% and 0.01% CEWAF (Corexit 9500 and Hiclean II) (Fig. 2a and b). EROD activity in rockfish exposed to all kinds of CEWAF (Corexit 9500 and Hiclean) was also highly increased. Fish in both CEWAF treatments showed up to 15-fold higher EROD activity than fish exposed to water controls. However, we could not detect CYP1A activity in rockfish exposed to 0.1% of either dispersant. 3.3. Brain AChE activity in fish exposed to WAF and CEWAF Brain AChE activity in rockfish exposed to 0.1% and 0.01% CEWAFC was significantly lower than among all experimental groups after 48 h and 72 h treatment. However, there was no significant inhibition in the fish exposed to WAF, or CEWAFH (Fig. 3). 3.4. FAC
Fig. 3. Level of brain AChE activity in rockfish exposed to a range of WAF, CEWAF concentrations from crude oils and 10 µg/kg β-napthoflavone: (a) after 48 h treatment, and (b) after 72 h treatment. Data are means ± standard error of mean (SEM) (n = 6–7). Data were subjected to ANOVA followed by Duncan's multiple-range test. Asterisks indicate means that were significantly different compared with control fish (⁎⁎P b 0.01).
Fig. 4 shows the results of bile analysis for PAH metabolites. Control fish had no fluorescent bile metabolites, and the β-NF treated fish had only background amounts given the detection limit. After 48 h treatment, the WAF-treated groups of fish contained only slightly higher concentrations of fluorescent metabolites than the control group. However, bile metabolites in the 0.1% and 0.01% CEWAFC-treated
222
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223
Fig. 5. Relationship between levels of bile 1-hydroxypyrene metabolites and CYP1A expression in rockfish. Triangle down (nmol/mg/min: EROD activity), circle opened (pixel intensity (%): CYP1A protein) and circle closed (pixel intensity (%): CYP1A gene).
group fish were significantly higher than in fish exposed to WAF and CEWAFH. Fish groups in both CEWAFC and CEWAFH treatments showed up to 5-fold higher concentrations than in control fish. Concentrations of fluorescent metabolites in both CEWAFC treatment groups were significantly higher than CEWAFH after 48 h and 72 h treatment. 3.5. Biomarkers and bile 1-OH-pyrene correlation Hydrocarbon uptake (shown by the presence of PAH metabolites in bile) and level of biochemical responses (CYP1A gene, protein and EROD) showed good correlation (Fig. 5). Indices of CYP1A mRNA were most sensitive in all biochemical responses, but EROD activity didn't respond until bile 1-OH-pyrene reach to the certain level. 4. Discussion PAHs are taken up via gills, skin and mouth; they are transformed to water-soluble polar metabolites and excreted in the bile. This multiresponse study evaluated the biological impacts of PAHs at subcellular and individual levels. Although no water PAHs data exist describing all of the treatment groups, bile fluorescence analysis of fish from each fish group showed evidence of PAH exposure; i.e., there were no clear peaks in control fish bile at the 340–360 (emission) nm regions. Lunnel (1993) reported that when oil is premixed with dispersant, 90% of oil droplets are less than 45 µm, representing 50% of the oil volume, As much as 99% of oil droplets are less than 70 µm, representing 80–90% of the oil volume. WAF-treated groups of fish did not show higher concentrations of fluorescent metabolites after 48 h treatment in the study. However, both CEWAF-treated groups of fish showed much higher concentrations of fluorescent metabolites after 72 h treatment. Metabolic enzyme activities could complement measurements of the induction of the Mixed Function Oxidase (MFO). MFO enzyme activities, as measured by EROD activity, can provide an indication of exposure to petroleum hydrocarbons (Payne et al., 1984; Van Veld, 1990; Goksøyr and Förlin, 1992; Van der weiden et al., 1994; Whyte et al., 2000; Gómez et al., 2006; Damásio et al., 2007). An increased EROD activity was related to higher lethargy in fish. There are some reports that BaP elicit its immunotoxic effects via upregulation of P450 1A in fish. The study suggests that the possible mechanism of BaP-induced immunosupression has indicated that the immunotoxicity of BaP, like its carcinogenicity, requires the production of BaP metabolites by CYP1A-mediated catalysis (Henry et al., 1997; Carlson et al., 2004). For CYP1A mRNA and protein expression, we observed similar positive trends for EROD activity with increasing concentrations of fluorescent metabolites (1-hydroypyrene) in bile. In
our experiments, both dispersants seem to increase the PAHs uptake by fish bioavailability. Little information is available for toxicities of dispersant solution alone. Dispersant alone was not highly toxic to rockfish, however fish were more sensitive to dispersant–oil mixtures than oil or oil dispersant alone as body water content (Ramachandran et al., 2006). These data showed that interactions between bile PAHs metabolites and multi-biomarker response might be affected by the toxicity of dispersant–oil mixtures. In contrast, Adams et al. (1999) showed that dispersed crude oil was less toxic than either oil or dispersant alone, indicating that toxicity data varies for different dispersant and oil types. Results from this study, CYP1A mRNA, protein and EROD activity are products of the same pathway but they have distinguishable responsibility. CYP1A mRNA was the most sensitive predictor of CYP1A evaluation system. But levels of the background in CYP1A gene expression were higher than those of EROD activity. EROD activity didn't respond until bile 1-OH-pyrene reaches to the certain levels of ∼80 intensity. It is not clear whether these variable results reflect sensitiveness differences, differences in threshold value or experimental variation. More studies are needed to understand this problem. AChE activity has been estimated as an enzymatic biomarker in neurotoxicity caused by pesticides. However, it has been recently reported that some environmental chemicals, heavy metals and petroleum-derived products have been reported in marine organisms (Mora et al., 1999; Moreira et al., 2004, Moreira and Guilhermino, 2005). Kang and Fang (1997) have suggested that several PAHs are weaker inhibitors of AChE activity in vitro. CEWAFC-treated groups of fish showed inhibition of brain AChE activity after 48 h and 72 h treatment. The result was similar to the bile fluorescent metabolites induction, which was highest among all of the treatment groups. But the correlation between hydrocarbon uptake and brain AChE activity didn't show good correlation in this experiment (not show the data). When PAHs are released into the marine environment during an oil spill, aquatic organisms are exposed to variable concentrations that can result in abnormal biochemical, physiological and behavioral responses. This study revealed induction of multi-biomarker responses in fish exposed to a dispersant–oil mixture via an increase in PAHs uptake due to their increased bioavailability to fish. Consequently, dispersing the oil would clearly cause an enhanced toxic effect on ovoviviparous fish physiological system. Acknowledgments This work was supported by project no. PM54770 (grants-in-aid from the Ministry of Land, Transport and Maritime Affairs, Korea) and project no. PG47273 (Korea Research Council of Fundamental Science and Technology). We are grateful to Professor Stelvio M. Bandiera, University of British Columbia, Canada, for providing the CYP1A antibody. References Adams, G., Klerks, P., Belanger, S., Dantin, D., 1999. The effect of the oil dispersant Omniclean on the toxicity of fuel oil No. 2 in two bioassays with the sheepshead minnow, Cyprinodon variegates. Chemosphere 39, 2141–2157. Addison, R.F., Payne, J.F., 1986. Assessment of hepatic mixed function oxidase induction in winter flounder (Pseudopleuronectes americanus) as a marine petroleum pollution monitoring technique, with an appendix describing practical field measurements of MFO activity. Can. Tech. Rep. Fish. Aquat. Sci. No. 1505. Addison, R.F., Willis, D.E., Zinck, M.E., 1994. Liver microsomal mono-oxygenase induction in winter flounder (Pseudopleuronectes americanus) from a gradient of sediment PAH concentrations at Sydney Harbour, Nova Scotia. Mar. Environ. Res. 37, 283–296. Anderson, J.W., Neff, J.M., Cox, B.A., Tatem, H.E., Hightower, G.M., 1974. Characteristics of dispersants and water-soluble extracts of crude and refined oils and their toxicity to estuarine crustaceans and fish. Mar. Biol. 27, 75–88. Ariese, F., Steven, S.J., Kok, J., Verkaik, M., Gooijer, C., Velthorst, M.H., Hofstraat, M.W., 1993. Synchronous fluorescence spectrometry of fish bile: a rapid screening method for the biomonitoring of PAH exposure. Aquat. Toxicol. 26, 273–286. Black, J., Fox, H., Black, P., Bock, F., 1985. Carcinogenic effects of river sediment extracts in fish and mice. In: Jolley, R.L. (Ed.), Water Chlorination: Chemistry, Environmental Impact and Health Effects, vol. 5. Lewis Publishers, Chelsea Michigan, pp. 415–427.
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223 Bobra, A.M., Shiu, W.Y., Mackay, D., Goodman, R.H., 1989. Acute toxicity of dispersed fresh and weathered crude oil and dispersants to Daphnia magna. Chemosphere 19, 1199–1222. Boehlert, Y.,1991. Rockfishes of the Genus Sebastes: Their Reproduction and Early Life History. In: Develop. Environ. Bio. Of fishes, vol. II. Klower Academic Publishers, pp. 49–56. Canevari, G.P., 1978. Some observation on the mechanism and chemistry aspects of chemical dispersion. In: McCarthy, L.T.J., Lindblom, G.P., Walter, H.F. (Eds.), Chemical Dispersants for the Control of Oil Spills: Am. Soc. Testing and Materials, Philadelphia, pp. 2–5. Carlson, E.A., Li, Y., Zelikoff, J.T., 2004. Benzo[a]pyrene-induced immunotoxicity in Japanese medaka (Oryzias latipes): relationship between lymphoid CYP1A activity and humoral immune suppression. Toxicol. Appl. Pharmacol. 201, 40–52. Collier, T.K., Anulacion, B.F., Stein, J.E., Goksoyr, A., Varanasi, U., 1995. A field evaluation of cytochrome P450A as a biomarker of contaminant exposure in these species of flatfish. Environ. Toxicol. Chem. 14, 143–152. Cotou, E., Castritsi-Catharious, I., Moraitou-Apostolopoulou, M., 2001. Surfactant-based oil dispersant toxicity to developing nauplii of Artemia: effects on atpase enzymatic system. Chemosphere 42, 959–964. Chapman, H., Purnell, K., Law, R.J., Kirby, M.F., 2007. The use of chemical dispersant to combat oil spills at sea: a review of practice and research needs in Europe. Mar. Pollut. Bull. 54, 827–838. Daeil Corporation. http://daeilchem.net. Damásio, J.B., Barata, C., Munné, A., Ginebreda, A., Guashsch, H., Caixach, J., Porte, C., 2007. Comparing the response of biochemical indicators (biomarker) and biological indices to diagnose the ecological impact of an oil spillage in a Mediterranean river (NE Catalunya, Spain). Chemosphere 66, 1206–1216. Ellman, G.L., Courtney, K.D., Andres Jr., V., Featherstone, R.M., 1961. A new and rapid colorimetric determination of acetylcholinesterase activity. Biochem. Pharmacol. 7, 88–95. Gilbert, T., 1996. Dispersant use/oil/sea temperature suitability matrix and oil chemistry and the impact on the effectiveness of chemical dispersant. Sixth Scientific Support Coordinators Workshop. Launceston, Tasmania, Australian Maritime Safety Authority, Canberra, ACT 2601, Australia. URL://www.amsa.gov.au.me.netplan/disper.htm. Gesto, M., Tintos, A., Soengas, J.L., Miguez, J.M., 2006. Effects of acute and prolonged naphthalene exposure on brain monoaminergic neurotransmitters in rainbow trout (Oncorhynchus mykiss). Comp. Biochem. Physiol. C 144, 173–183. Goksøyr, A., Förlin, L., 1992. The cytochrome P-450 system in fish, aquatic toxicology and environmental monitoring. Aquat. Toxicol. 22, 287–312. Gómez, C.M., Campillo, J.A., Benedicto, J., Fernández, B., Valdés, J., García, I., Sánchez, F., 2006. Monitoring biomarkers in fish (Lepidorhombus boscii and Callionymus lyra) from the northern Iberian shelf after the Prestige oil spill. Mar. Pollut. Bull. 53, 305–314. Henry, E.C., Kent, T.A., Gasiewicz, T.A., 1997. DNA binding and transcriptional enhancement by purified TCDD-Ah receptor complex, Arch. Biochem. Biophys. 339, 305–314. Jung, J.H., Jeon, J.K., Shim, W.J., Oh, J.R., Lee, J.Y., Kim, B.K., Han, C.H., 2005. Molecular cloning of vitellogenin cDNA in rockfish (Sebastes schlegeli) and effects of 2,2′4,4′ 5,5′-hexachlorobiphenyl (PCB153) on its gene. Mar. Pollut. Bull. 794–800. Jung, J.H., Shim, W.J., Addison, R.F., Han, C.H., 2006. Protein and gene expression of VTG in response to 4-nonylphenol in rockfish (Sebastes schlegeli). Comp. Biochem. Physiol. C 143, 162–170. Jung, J.H., Kim, S.J., Lee, T.K., Shim, W.J., Woo, S., Kim, D.J., Han, C.H., 2008. Biomarker responses in caged rockfish (Sebastes schlegeli) from Masan Bay and Haegeumgang, South Korea. Mar. Pollut. Bull. 57, 599–606. Kang, J.J., Fang, H.W., 1997. Polycyclic aromatic hydrocarbons inhibit the activity of acetylcholinesterase purified from electric eel. Biochem. Biophys. Res. Commun. 238, 367–379. Krahn, M.M., Burrows, D.G., MacLeod Jr., W.D., Malins, D.C., 1987. Determination of individual metabolites of aromatic compounds in hydrolyzed bile of English sole (Parophrys vetulus) from polluted sites in Puget Sound, Washington. Arch. Environ. Contam. Toxicol. 16, 511–522. Lin, S., Bullock, P.L., Addison, R.F., Bandiera, S.M., 1997. Detection of cytochrome P450 1A1 in several species using antibody against a synthetic peptide derived from a rainbow trout cytochrome P450 1A1. Environ. Chem. Toxicol. 17, 439–445.
223
Lunnel, T., 1993. Dispersion: oil droplet size measurements at sea. Proceedings of the Sixteenth Arctic and Marine Oil spill Program Technical Seminar, pp. 1023–1056. Meyers, J.N., Nacci, D.E., Di Giulio, R.T., 2002. Cytochrome P450A (CYP1A) in killifish (Fundulus heteroclitus): heritability of altered expression and relationship to survival in contaminated sediments. Toxicol. Sci. 68, 69–81. McCain, M.S., Myers, D.W., Brown, M.M., Krahn, W.T., Roubal, M.H., Schi1ewe, J.T., Landahl, J.T., Chan, S.-L., 1987. Field and laboratory studies of the etiology of liver neoplasms in marine fish from Puget Sound. Environ. Health Perspect. 71, 5–16. Mora, P., Fournier, D., Narbonne, J.F., 1999. Cholinesterase from the marine mussels Mytilus galloprovincialis L. and Mytilus edulus L. from the freshwater bivalve Corbicula fluminea Muller. Comp. Biochem. Physiol., Part C Pharmacol. Toxicol. 122, 353–361. Moreira, S.M., Guilhermino, L., 2005. The use of the Mytilus galloprovincialis acetylcholinesterase and glutathione S-transferases activities as biomarkers of environmental contamination along the North West Portuguese coast. Environ. Monit. Assess. 105, 309–325. Moreira, S.M., Moreira-Santos, M., Ribeiro, R., Guilhermino, L., 2004. The “Coral Bulker” fuel oil spill on the north coast of Portugal: spatial and temporal biomarker responses in Mytilus galloprovincialis. Ecotoxicol 13, 619–630. Nalco Energy Services. http://www.nalco.co.kr. Nicolas, J.M., 1999. Vitellogenesis in fish and the effects of polycyclic aromatic hydrocarbon contaminants. Aquat. Toxicol. 45, 77–90. Payne, J.F., Bauld, C., Dey, A.C., Kiceniuk, J.W., Williams, U., 1984. Selectivity of mixedfunction oxygenase enzyme induction in flounder (Pseudopleuronectes americanus) collected at the site of the Baie Verte, Newfoundland oil spill. Comp. Biochem. Physiol. 79C, 15–19. Payne, J.F., Fancy, L.L., Rahimtula, A.D., Porter, E.L., 1987. Review and perspective on the use of mixed function oxygenase enzymes in biological monitoring. Comp. Biochem. Physiol. 86C, 233–245. Ramachandran, S.D., Hodson, P.V., Khan, C.W., Lee, K., 2004. Oil dispersant increases PAH uptake by fish exposed to crude oil. Ecotoxicol. Environ. Saf. 59, 300–308. Ramachandran, S.D., Michael, J.S., Hodson, P.V., Boudreau, M., Courtenay, S.C., Lee, K., Thomas, K., Jennifer, A.D., 2006. Influence of salinity and fish species on PAH uptake from dispersed crude oil. Mar. Pollut. Bull. 52, 1182–1189. Singer, M.M., George, S., Benner, D., Jacobson, S., Tjeerdema, R.S., Sowby, M.L., 1993. Comparative toxicity of two oil dispersants to the early life stages of two marine species. Environ. Toxicol. Chem. 12, 1855–1863. Singer, M.M., Aurand, D., Bragins, G.E., Clarks, J.R., Coelho, G.M., Sowby, M.L., Tjeerdema, R.S., 2000. Standardization of the preparation and quantitation of water-accommodated fractions of petroleum for toxicity testing. Mar. Pollut. Bull. 40, 1007–1016. Sonstergard, R.A., Leatherland, J.F., 1984. Comparative epidemiology: the use of fishes in assessing carcinogenic contaminants. In: Cairs, V., et al. (Ed.), Contaminant Effects on Fisheries: Adv. Environ. Sci. Technol., vol. 16, pp. 223–232. Thomas, R.E., Rice, S.D., 1981. Excretion of aromatic hydrocarbons and their metabolites by freshwater and seawater Dolly Varden char. In: Vernberg, F.J., Calabrese, A., Thurberg, F., Vernberg, W. (Eds.), Biological Monitoring of Marine Pollutants. InAcademic Press, New York, pp. 425–448. Tintos, A., Gesto, M., Mĺguez, J.M., Soengas, J.L., 2007. Naphthalene treatment alters liver intermediary metabolism and levels of steroid hormone in plasma of rainbow trout (Oncorhynchus mykiss). Ecotoxicol. Environ. Saf. 69, 180–186. Van Veld, P.A., 1990. Absorption and metabolism of dietary xenobiotics by the intestine of fish. Rev. Aquat. Sci. 2, 185–203. Van der weiden, M.E.J., Hanegraaf, F.H.M., Eggens, M.L., Celander, M., Seinen, W., Van derBerg, M.,1994. Temporal induction of cytochrome P450 1A in the mirror carp (Cyprinus carpio) after administration of several polycyclic aromatic hydrocarbons. Environ. Toxicol. Chem. 13, 797–802. Whyte, J.J., Jung, R.E., Schmit, C.J., Tillitt, D.E., 2000. Ethoxyresorufin O-deethylase (EROD) activity in fish as a biomarker of chemical exposure. Crit. Rev. Toxicol. 30, 347–570.
Contents lists available at ScienceDirect
Comparative Biochemistry and Physiology, Part C j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / c b p c
Biochemical changes in rockfish, Sebastes schlegeli, exposed to dispersed crude oil Jee Hyun Jung, Un Hyuk Yim, Gi Myeong Han, Won Joon Shim ⁎ Oil and POPs Research Group, KORDI, Geoje-shi, 656-834, Republic of Korea
a r t i c l e
i n f o
Article history: Received 3 February 2009 Received in revised form 21 April 2009 Accepted 24 April 2009 Available online 3 May 2009 Keywords: Oil dispersant Rockfish CYP1A Ethoxyresorufin O-de-ethylase Acethylcholinesterase Ovoviviparous
a b s t r a c t This paper describes the response of the ovoviviparous rockfish, Sebastes schlegeli, to hydrocarbons in the water-accommodated fraction (WAF) of crude oil, in the presence or absence of oil dispersants. Concentrations of cytochrome P-450 1A (CYP1A) and levels of its catalytic activity ethoxyresorufin O-de-ethylase (EROD) in rockfish exposed to WAF at concentrations of 0.1% and 1% were significantly increased by the addition of a dispersant, Corexit 9500 after 48 h exposure. After 72 h exposure, the levels of CYP1A and EROD activity were significantly increased in 0.1% and 0.01% chemically enhanced WAF (CEWAF) (Corexit 9500 and Hiclean II dispersant). Bile samples from fish exposed to WAF alone had low concentrations of hydrocarbon metabolites, exemplified by 1-hydroxypyrene. After 72 h exposure, hydrocarbon metabolites in bile from fish exposed to WAF in the presence of either Corexit 9500 or Hiclean II were significantly higher compared with fish exposed to WAF alone or control fish. These experiments confirm that the use of oil dispersants will increase the exposure of ovoviviparous fish to hydrocarbons in oil. © 2009 Elsevier Inc. All rights reserved.
1. Introduction Oil dispersants are potential tools for minimizing the impact of marine oil spills. The dispersants are primarily surfactants that reduce the interfacial tension between the oil and dispersant from micelles that are dispersed into the water column (Canevari, 1978). They do not actually remove the oil. Eventually, it will be broken down by microscopic organisms in form of biodegradation. However, oil dispersant use involves the addition of potentially damaging chemicals to an already stressed environment, which can lead to further problems. Based on a review of the 258 marine incidents that the International Tanker Owners Pollution Federation (ITOPF) were involved with between 1995 and 2005, 46 (18%) involved the use of chemical dispersants at sea. A review of the use of dispersants in different international regions for the 46 cases considered by the ITOPF shows that over half took place in Asian waters (Chapman et al., 2007). On the 7th December 2007, the M/V Hebei Spirit spilled more than 12 million liters of three kinds of crude oil off the west coast of Korea, in a region of high aquaculture and fisheries activity. About 250 kL of oil dispersant were used in the cleanup operations. There is ongoing controversy concerning the impacts of oil dispersant use on marine biota. The observed increased toxicity of dispersed oil has been attributed to dispersed particle sizes (Bobra et al., 1989) and aromatic hydrocarbon concentration (Anderson et al., 1974). Hydrocarbons are accumulated via the gills and mouth and the dispersed fraction would therefore be the
⁎ Corresponding author. Oil and POPs Research Group, Korean Ocean Research and Development Institute, 391 Jangmok-Ri Jangmok-myon, Geoje, Korea. Tel.: +82 55 639 8687; fax: +82 55 639 8689. E-mail address: [email protected] (W.J. Shim). 1532-0456/$ – see front matter © 2009 Elsevier Inc. All rights reserved. doi:10.1016/j.cbpc.2009.04.009
most potentially bioavailable (Thomas and Rice, 1981). Fish can accumulate soluble petroleum hydrocarbons very rapidly (Collier et al., 1995), and these contaminants can concentrate in tissues to factors 10–100 times greater than in water (Ramachandran et al., 2006). Hydrocarbon residues in fish can exert various types of effects. They can alter gonad size, gametogenesis (Nicolas, 1999; Tintos et al., 2007) and brain neurotransmission (Gesto et al., 2006). Some polynuclear aromatic hydrocarbons (PAHs) induce hepatic tumorgenicity in fish (Sonstergard and Leatherland, 1984; McCain et al., 1987; Meyers et al., 2002). Black et al. (1985) have shown that fish from PAHs-polluted waters show a high incidence of liver tumors compared with populations from uncontaminated waters in the same region. Assessment of the biological impacts of oil dispersants on marine ecosystems is usually based on acute toxicity tests (Singer et al., 1993; Cotou et al., 2001). Chemically induced alterations in the biochemistry and physiology of organisms are being used as diagnostic tools in assessing potential biochemical impacts in aquatic systems. However, there are few studies of the impacts of oil and oil dispersants on fish biochemical processes. The induction of cytochrome P-450 1A (CYP1A) enzyme is well established as a “biomarker” to monitor exposure to aromatic hydrocarbons (Payne et al., 1984, 1987). Under actual oil spill conditions in the marine environment, it is unclear how oil and dispersants will combine to act agonistically or antagonistically on fish physiological systems. In this study, the effect of two dispersants on the uptake and metabolism of hydrocarbons and their effects on selected biochemical responses in rockfish were examined. This species was categorized as an ovoviviparous fish, because development of the late embryonic stages is dependent on maternal nutrition as an energy source, in addition to that stored in yolk accumulated in oocytes during vitellogenesis (Boehlert,
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223
1991). Therefore, it is interesting to clarify the impact on ovoviviparous biochemical responses that dispersed oil may alter the hepatodetoxification and neurotransmitter processes. Corexit 9500 is a more recent formulation which provides enhanced penetration and emulsion properties compared with Corexit 9527. In Korea, a common oil dispersant is Hiclean II, which is a known developed product based on a carbonated-aqueous surfactant. Their effects on enhancing hydrocarbon uptake (shown by the presence of PAH metabolites in bile), and on the induction of CYP1A, EROD and inhibition of brain acethylcholinesterase (AChE) are described. 2. Materials and methods 2.1. Corexit 9500 and Hiclean II Corexit 9500 was purchased from Nalco Corporation (Texas, USA); its 96 h LC50 is reported to be 25.2 mg/L for inland silversides (technical note of Nalco). Corexit 9500 is based on aliphatic hydrocarbons, propylene glycol, and some sulfuric acid salt. Hiclean II is a carbonated-aqueous surfactant, which was purchased from Daeil Corporation (Busan, Korea); its 24 h LC50 to rockfish is 2000 mg/L (Technical note of Daeil). 2.2. Water-accommodated fraction (WAF) of crude oil and CEWAF Crude ARM (medium grade, Saudi Arabia) was donated from the laboratory of the Korean Coast Guard. Crude oil and water were mixed at a ratio of 1:49 v/v in sealed containers with minimum head space for 18 h at 18 °C. The vortex was adjusted to no more than a third of the height of the mixture from the oil–water interface (Singer et al., 2000). The mixture was allowed to settle for 1 h for the separation of water and oil phase and the water phase was drained off for the experiment. The dispersants, Corexit 9500 and Hiclean II were added to the surface of the oil–water mixture at the recommended ratio of 1:20 dispersant oil: (technical note of Nalco and Daeil, Gilbert,1996; Ramachandran et al., 2004) and stirred for 1 h. 2.3. Exposure experiments Juvenile rockfish (Sebastes schlegeli), 2 years old,180± 2.0 g (mean± S.D.) weight and 20 ± 2.0 cm length were purchased from a local fish farm in Geoje, Korea. Separate control fish were injected with ethanol or not injected. Fish were acclimated for 1 week in flowing seawater at 16 °C on a 16 h light/8 h dark photoperiod in the laboratory before the beginning of experiments to make sure they were disease free. Water quality variables, including ammonia, nitrate and dissolved oxygen were
Table 1 Experimental summary of exposure groups. Experimental group Treatment CON 0.01 WAF 0.1 WAF 0.01 CEWAFC 0.1 CEWAFC 0.01CEWAFH 0.1CEWAFH DC DH β-NF
Exposure type
100 µL of DMSO (dimethyl sulfoxide) 0.01% WAF (made by crude oil)
Injection Water exposure 0.01% WAF (made by crude oil) Water exposure 0.01% CEWAF (crude oil dispersed by Corexit Water 9500) exposure 0.1% CEWAF (crude oil dispersed by Corexit Water 9500) exposure 0.01% CEWAF (crude oil dispersed by Hiclean II) Water exposure 0.1% CEWAF (crude oil dispersed by Hiclean II) Water exposure 0.1% dispersant (Corexit 9500) Water exposure 0.1% dispersant (Hiclean II) Water exposure 10 µg/kg of β-naphthoflavone (in DMSO) Injection
Note: CON = control; other abbreviations represent experimental chemicals.
219
Table 2 Experimental conditions for rockfish, S. schlegeli. Fish mass (g) 264 ± 26.3
HSI 2.54 ± 0.27
pH 8.4 ± 0.04
Temperature
Dissolved oxygen
(°C)
(% saturation)
14 ± 0.3
97.8 ± 2.1
HSI = hepatosomatic index. Data are means ± standard error of mean (SEM).
monitored daily. Six and seven fish were exposed to series concentrations of WAF (v/v), CEWAFC (WAF treated with Corexit 9500), CEWAFH (WAF treated with Hiclean II) and two 0.1% dispersant exposure groups (Corexit 9500 and Hiclean II) in 60 L water. β-napthoflavone of 10 µg/kg, a known CYP1A inducer, was injected into one group of fish as a positive control (Table 1). Exposure concentrations causing CYP1A induction and lethality were considered. Each experiment was performed continuousflow system which is controlled a light and temperature. They contained a central cylinder where a Teflon-coated main propeller. Exposure concentrations of WAF and CEWAF were 0.1 and 0.01%. The exposure duration was fixed at 48 h and 72 h exposure, which caused comparable CYP1A induction. Immediately after movement ceased, the fish were killed by a blow on the head, weighed, measured, and dissected and various tissues removed for analysis. Tissue samples were stored at − 80 °C until analyzed. 2.4. Fluorescent aromatic hydrocarbon (FAC) metabolites Synchronous scanning of fluorescent bile metabolites, focusing on 1hydroxypyrene (Krahn et al., 1987) was used as a rough indicator of the uptake of PAH by the fish. Frozen gall bladders were disrupted to release the bile, the tissue mass was centrifuged briefly to sediment out unwanted tissue, and a bile sample was diluted 50- or 100-fold in 50% ethanol. Diluted bile samples were scanned with excitation between 300 and 390 nm and emission between 337 and 427 nm (i.e., a wavelength separation of 37 nm: Ariese et al., 1993). 1-hydroxypyrene (Aldrich) was used as a standard at concentrations (in 50% ethanol) in the range of 50–500 nM. 2.5. Enzyme activity Brain tissue samples were thawed on ice and were homogenized and analyzed for AChE activity using a microplate reader adaptation of the Ellman procedure (Ellman et al., 1961) as described by Jung et al. (2006, 2008). Liver samples were homogenized and microsomes were prepared and suspended in 0.1 M phosphate buffer pH 7.6 essentially as described by Addison et al. (1994). Ethoxyresorufin O-de-ethylase (EROD) activity in microsomes was determined using reagent concentrations described by Addison and Payne (1986) in a fluorescence microplate reader with the excitation and emission filter set at 544 nm and 590 nm, respectively. Protein concentrations in the sample were assayed by the bicinconchinic acid (BCA) method using a kit from Pierce Chemicals (Rockford IL); the product was read in a microplate reader at 550 nm, using bovine serum albumin (BSA) as a standard. 2.6. CYP1A protein expression CYP1A was estimated by western blotting using the antibody raised against a 22 amino-acid oligopeptide fragment conserved in several types of vertebrate CYP1A as described by Lin et al. (1997). Liver microsomal proteins were separated by SDS-PAGE and transferred to PVDF nitrocellulose membranes. Nonspecific binding sites on the membrane were blocked with 0.5% goat IgG in TBST (20 mM Tris–HCl pH 8.0, 100 mM NaCl, 0.05% Tween 20) for 2 h at 37 °C. The CYP1A antibody was diluted 1:5000 in TBST and incubated with the membrane for 1 h at room temperature. After washing, the membrane was incubated with a secondary antibody (goat anti-mouse IgG biotin conjugate)
220
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223
for 30 min at room temperature. The blots were developed using an ABC kit (Vector Laboratories, Burlingame, CA) for 30 min. The substrate solution contained 0.02% 3,3′-diaminobenzidine tablets (Sigma) and the development was stopped with distilled water. The reliability of the CYP1A (protein) and EROD measurements was assessed during the BEQUALM intercalibration (Biological Effects Quality Assurance in Monitoring Programs; http://www.bequalm. org).exercise; results were within approx. 1 SD of the mean estimates of nine participating laboratories.
Fig. 2. Hepatodetoxification activity in rockfish exposed to a range of WAF, CEWAF concentrations from crude oils and 10 µg/kg β-napthoflavone after 72 h treatment. (a) Level of CYP1A mRNA expression in rockfish liver based on northern blot analysis of liver samples from individual fish. Relative mRNA levels of rockfish CYP1A were normalized with β-actin. (b) Level of CYP1A (nmol) protein in rockfish liver based on western blot analysis of microsome from individual fish. (c) Level of EROD activity (nmol/ min/mg protein) in rockfish liver microsome from individual fish Data are means ± standard error of mean (SEM) (n = 6–7). Data were subjected to ANOVA followed by Duncan's multiple-range test. Different letters denote statistically significant (P b 0.05).
Fig. 1. Hepatodetoxification activity in rockfish exposed to a range of WAF, CEWAF concentrations from crude oils and 10 µg/kg β-napthoflavone after 48 h treatment. (a) Level of CYP1A mRNA expression in rockfish liver based on northern blot analysis of liver samples from individual fish. Relative mRNA levels of rockfish CYP1A were normalized with β-actin. (b) Level of CYP1A (nmol) protein in rockfish liver based on western blot analysis of microsome from individual fish. (c) Level of EROD activity (nmol/ min/mg protein) in rockfish liver microsome from individual fish. Data are means± standard error of mean (SEM) (n = 6–7). Data were subjected to ANOVA followed by Duncan's multiple-range test. Different letters denote statistically significant (P b 0.05). DC (treated with Corexit 9700), DH (treated with Hiclean II), CEWAFC (crude oil dispersed with Corexit 9700) and CEWAFH (crude oil dispersed with Hiclean II).
2.7. CYP1A mRNA expression cDNA probes were generated by transcription with 5.0 µL of [α− 32P] dATP (Boston, MA NEN) using a random primer labeling kit (Takara, Japan). CYP1A mRNA expression levels were estimated by northern blot analysis as described by Jung et al. (2005, 2006). Templates for each probe were made from cloned CYP1A and β-actin fragments that had been submitted to GenBank in our previous study (accession no. AY 558556; AY 166590).
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223
221
2.8. Statistical analysis All data are expressed as mean± SE. Data were compared by analysis of variance (ANOVA) and significant differences were calculated by Duncan's multiple-range tests. 3. Results 3.1. Test condition for the rockfish 72 h static bioassay with daily renewal Water quality variables (ammonia, nitrate and dissolved oxygen) were monitored daily. These variables did not change over the experimental period and there were no differences in body weight and length for the exposure duration among experimental groups. Fish Hepato Somatic Index (HSI) values are shown in Table 2 and there was no significant difference among experimental groups. 3.2. CYP1A expression and EROD activity in fish exposed to WAF and CEWAF Exposure concentrations causing CYP1A induction and lethality were determined from previous studies (Ramachandran et al., 2004, 2006). A concentration of 0.1% is below the lethal dose. Fig. 1 (a, b), showed that levels of CYP1A mRNA and protein expression in rockfish exposed to CEWAFC (Corexit 9500) increased significantly in response to a concentration of 0.1% (v/v) after 48 h. They were higher than CEWAFH (Hiclean II) and the control group and close to the positive
Fig. 4. Level of bile 1-hydroxypyrene metabolites in rockfish: (a) after 48 h treatment, (b) after 72 h treatment. Data were subjected to ANOVA followed by Duncan's multiplerange test. Asterisks indicate means that were significantly different compared with exposed fish (⁎⁎P b 0.01, ⁎ P b 0.05).
control group level. Fig.1c showed that the EROD activity of fish exposed to CEWAF (Corexit 9500) increased significantly compared with fish exposed to WAF, CEWAFH and the negative control group. After 72 h of exposure, levels of CYP1A mRNA and protein expression were significantly increased at both doses of 0.1% and 0.01% CEWAF (Corexit 9500 and Hiclean II) (Fig. 2a and b). EROD activity in rockfish exposed to all kinds of CEWAF (Corexit 9500 and Hiclean) was also highly increased. Fish in both CEWAF treatments showed up to 15-fold higher EROD activity than fish exposed to water controls. However, we could not detect CYP1A activity in rockfish exposed to 0.1% of either dispersant. 3.3. Brain AChE activity in fish exposed to WAF and CEWAF Brain AChE activity in rockfish exposed to 0.1% and 0.01% CEWAFC was significantly lower than among all experimental groups after 48 h and 72 h treatment. However, there was no significant inhibition in the fish exposed to WAF, or CEWAFH (Fig. 3). 3.4. FAC
Fig. 3. Level of brain AChE activity in rockfish exposed to a range of WAF, CEWAF concentrations from crude oils and 10 µg/kg β-napthoflavone: (a) after 48 h treatment, and (b) after 72 h treatment. Data are means ± standard error of mean (SEM) (n = 6–7). Data were subjected to ANOVA followed by Duncan's multiple-range test. Asterisks indicate means that were significantly different compared with control fish (⁎⁎P b 0.01).
Fig. 4 shows the results of bile analysis for PAH metabolites. Control fish had no fluorescent bile metabolites, and the β-NF treated fish had only background amounts given the detection limit. After 48 h treatment, the WAF-treated groups of fish contained only slightly higher concentrations of fluorescent metabolites than the control group. However, bile metabolites in the 0.1% and 0.01% CEWAFC-treated
222
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223
Fig. 5. Relationship between levels of bile 1-hydroxypyrene metabolites and CYP1A expression in rockfish. Triangle down (nmol/mg/min: EROD activity), circle opened (pixel intensity (%): CYP1A protein) and circle closed (pixel intensity (%): CYP1A gene).
group fish were significantly higher than in fish exposed to WAF and CEWAFH. Fish groups in both CEWAFC and CEWAFH treatments showed up to 5-fold higher concentrations than in control fish. Concentrations of fluorescent metabolites in both CEWAFC treatment groups were significantly higher than CEWAFH after 48 h and 72 h treatment. 3.5. Biomarkers and bile 1-OH-pyrene correlation Hydrocarbon uptake (shown by the presence of PAH metabolites in bile) and level of biochemical responses (CYP1A gene, protein and EROD) showed good correlation (Fig. 5). Indices of CYP1A mRNA were most sensitive in all biochemical responses, but EROD activity didn't respond until bile 1-OH-pyrene reach to the certain level. 4. Discussion PAHs are taken up via gills, skin and mouth; they are transformed to water-soluble polar metabolites and excreted in the bile. This multiresponse study evaluated the biological impacts of PAHs at subcellular and individual levels. Although no water PAHs data exist describing all of the treatment groups, bile fluorescence analysis of fish from each fish group showed evidence of PAH exposure; i.e., there were no clear peaks in control fish bile at the 340–360 (emission) nm regions. Lunnel (1993) reported that when oil is premixed with dispersant, 90% of oil droplets are less than 45 µm, representing 50% of the oil volume, As much as 99% of oil droplets are less than 70 µm, representing 80–90% of the oil volume. WAF-treated groups of fish did not show higher concentrations of fluorescent metabolites after 48 h treatment in the study. However, both CEWAF-treated groups of fish showed much higher concentrations of fluorescent metabolites after 72 h treatment. Metabolic enzyme activities could complement measurements of the induction of the Mixed Function Oxidase (MFO). MFO enzyme activities, as measured by EROD activity, can provide an indication of exposure to petroleum hydrocarbons (Payne et al., 1984; Van Veld, 1990; Goksøyr and Förlin, 1992; Van der weiden et al., 1994; Whyte et al., 2000; Gómez et al., 2006; Damásio et al., 2007). An increased EROD activity was related to higher lethargy in fish. There are some reports that BaP elicit its immunotoxic effects via upregulation of P450 1A in fish. The study suggests that the possible mechanism of BaP-induced immunosupression has indicated that the immunotoxicity of BaP, like its carcinogenicity, requires the production of BaP metabolites by CYP1A-mediated catalysis (Henry et al., 1997; Carlson et al., 2004). For CYP1A mRNA and protein expression, we observed similar positive trends for EROD activity with increasing concentrations of fluorescent metabolites (1-hydroypyrene) in bile. In
our experiments, both dispersants seem to increase the PAHs uptake by fish bioavailability. Little information is available for toxicities of dispersant solution alone. Dispersant alone was not highly toxic to rockfish, however fish were more sensitive to dispersant–oil mixtures than oil or oil dispersant alone as body water content (Ramachandran et al., 2006). These data showed that interactions between bile PAHs metabolites and multi-biomarker response might be affected by the toxicity of dispersant–oil mixtures. In contrast, Adams et al. (1999) showed that dispersed crude oil was less toxic than either oil or dispersant alone, indicating that toxicity data varies for different dispersant and oil types. Results from this study, CYP1A mRNA, protein and EROD activity are products of the same pathway but they have distinguishable responsibility. CYP1A mRNA was the most sensitive predictor of CYP1A evaluation system. But levels of the background in CYP1A gene expression were higher than those of EROD activity. EROD activity didn't respond until bile 1-OH-pyrene reaches to the certain levels of ∼80 intensity. It is not clear whether these variable results reflect sensitiveness differences, differences in threshold value or experimental variation. More studies are needed to understand this problem. AChE activity has been estimated as an enzymatic biomarker in neurotoxicity caused by pesticides. However, it has been recently reported that some environmental chemicals, heavy metals and petroleum-derived products have been reported in marine organisms (Mora et al., 1999; Moreira et al., 2004, Moreira and Guilhermino, 2005). Kang and Fang (1997) have suggested that several PAHs are weaker inhibitors of AChE activity in vitro. CEWAFC-treated groups of fish showed inhibition of brain AChE activity after 48 h and 72 h treatment. The result was similar to the bile fluorescent metabolites induction, which was highest among all of the treatment groups. But the correlation between hydrocarbon uptake and brain AChE activity didn't show good correlation in this experiment (not show the data). When PAHs are released into the marine environment during an oil spill, aquatic organisms are exposed to variable concentrations that can result in abnormal biochemical, physiological and behavioral responses. This study revealed induction of multi-biomarker responses in fish exposed to a dispersant–oil mixture via an increase in PAHs uptake due to their increased bioavailability to fish. Consequently, dispersing the oil would clearly cause an enhanced toxic effect on ovoviviparous fish physiological system. Acknowledgments This work was supported by project no. PM54770 (grants-in-aid from the Ministry of Land, Transport and Maritime Affairs, Korea) and project no. PG47273 (Korea Research Council of Fundamental Science and Technology). We are grateful to Professor Stelvio M. Bandiera, University of British Columbia, Canada, for providing the CYP1A antibody. References Adams, G., Klerks, P., Belanger, S., Dantin, D., 1999. The effect of the oil dispersant Omniclean on the toxicity of fuel oil No. 2 in two bioassays with the sheepshead minnow, Cyprinodon variegates. Chemosphere 39, 2141–2157. Addison, R.F., Payne, J.F., 1986. Assessment of hepatic mixed function oxidase induction in winter flounder (Pseudopleuronectes americanus) as a marine petroleum pollution monitoring technique, with an appendix describing practical field measurements of MFO activity. Can. Tech. Rep. Fish. Aquat. Sci. No. 1505. Addison, R.F., Willis, D.E., Zinck, M.E., 1994. Liver microsomal mono-oxygenase induction in winter flounder (Pseudopleuronectes americanus) from a gradient of sediment PAH concentrations at Sydney Harbour, Nova Scotia. Mar. Environ. Res. 37, 283–296. Anderson, J.W., Neff, J.M., Cox, B.A., Tatem, H.E., Hightower, G.M., 1974. Characteristics of dispersants and water-soluble extracts of crude and refined oils and their toxicity to estuarine crustaceans and fish. Mar. Biol. 27, 75–88. Ariese, F., Steven, S.J., Kok, J., Verkaik, M., Gooijer, C., Velthorst, M.H., Hofstraat, M.W., 1993. Synchronous fluorescence spectrometry of fish bile: a rapid screening method for the biomonitoring of PAH exposure. Aquat. Toxicol. 26, 273–286. Black, J., Fox, H., Black, P., Bock, F., 1985. Carcinogenic effects of river sediment extracts in fish and mice. In: Jolley, R.L. (Ed.), Water Chlorination: Chemistry, Environmental Impact and Health Effects, vol. 5. Lewis Publishers, Chelsea Michigan, pp. 415–427.
J.H. Jung et al. / Comparative Biochemistry and Physiology, Part C 150 (2009) 218–223 Bobra, A.M., Shiu, W.Y., Mackay, D., Goodman, R.H., 1989. Acute toxicity of dispersed fresh and weathered crude oil and dispersants to Daphnia magna. Chemosphere 19, 1199–1222. Boehlert, Y.,1991. Rockfishes of the Genus Sebastes: Their Reproduction and Early Life History. In: Develop. Environ. Bio. Of fishes, vol. II. Klower Academic Publishers, pp. 49–56. Canevari, G.P., 1978. Some observation on the mechanism and chemistry aspects of chemical dispersion. In: McCarthy, L.T.J., Lindblom, G.P., Walter, H.F. (Eds.), Chemical Dispersants for the Control of Oil Spills: Am. Soc. Testing and Materials, Philadelphia, pp. 2–5. Carlson, E.A., Li, Y., Zelikoff, J.T., 2004. Benzo[a]pyrene-induced immunotoxicity in Japanese medaka (Oryzias latipes): relationship between lymphoid CYP1A activity and humoral immune suppression. Toxicol. Appl. Pharmacol. 201, 40–52. Collier, T.K., Anulacion, B.F., Stein, J.E., Goksoyr, A., Varanasi, U., 1995. A field evaluation of cytochrome P450A as a biomarker of contaminant exposure in these species of flatfish. Environ. Toxicol. Chem. 14, 143–152. Cotou, E., Castritsi-Catharious, I., Moraitou-Apostolopoulou, M., 2001. Surfactant-based oil dispersant toxicity to developing nauplii of Artemia: effects on atpase enzymatic system. Chemosphere 42, 959–964. Chapman, H., Purnell, K., Law, R.J., Kirby, M.F., 2007. The use of chemical dispersant to combat oil spills at sea: a review of practice and research needs in Europe. Mar. Pollut. Bull. 54, 827–838. Daeil Corporation. http://daeilchem.net. Damásio, J.B., Barata, C., Munné, A., Ginebreda, A., Guashsch, H., Caixach, J., Porte, C., 2007. Comparing the response of biochemical indicators (biomarker) and biological indices to diagnose the ecological impact of an oil spillage in a Mediterranean river (NE Catalunya, Spain). Chemosphere 66, 1206–1216. Ellman, G.L., Courtney, K.D., Andres Jr., V., Featherstone, R.M., 1961. A new and rapid colorimetric determination of acetylcholinesterase activity. Biochem. Pharmacol. 7, 88–95. Gilbert, T., 1996. Dispersant use/oil/sea temperature suitability matrix and oil chemistry and the impact on the effectiveness of chemical dispersant. Sixth Scientific Support Coordinators Workshop. Launceston, Tasmania, Australian Maritime Safety Authority, Canberra, ACT 2601, Australia. URL://www.amsa.gov.au.me.netplan/disper.htm. Gesto, M., Tintos, A., Soengas, J.L., Miguez, J.M., 2006. Effects of acute and prolonged naphthalene exposure on brain monoaminergic neurotransmitters in rainbow trout (Oncorhynchus mykiss). Comp. Biochem. Physiol. C 144, 173–183. Goksøyr, A., Förlin, L., 1992. The cytochrome P-450 system in fish, aquatic toxicology and environmental monitoring. Aquat. Toxicol. 22, 287–312. Gómez, C.M., Campillo, J.A., Benedicto, J., Fernández, B., Valdés, J., García, I., Sánchez, F., 2006. Monitoring biomarkers in fish (Lepidorhombus boscii and Callionymus lyra) from the northern Iberian shelf after the Prestige oil spill. Mar. Pollut. Bull. 53, 305–314. Henry, E.C., Kent, T.A., Gasiewicz, T.A., 1997. DNA binding and transcriptional enhancement by purified TCDD-Ah receptor complex, Arch. Biochem. Biophys. 339, 305–314. Jung, J.H., Jeon, J.K., Shim, W.J., Oh, J.R., Lee, J.Y., Kim, B.K., Han, C.H., 2005. Molecular cloning of vitellogenin cDNA in rockfish (Sebastes schlegeli) and effects of 2,2′4,4′ 5,5′-hexachlorobiphenyl (PCB153) on its gene. Mar. Pollut. Bull. 794–800. Jung, J.H., Shim, W.J., Addison, R.F., Han, C.H., 2006. Protein and gene expression of VTG in response to 4-nonylphenol in rockfish (Sebastes schlegeli). Comp. Biochem. Physiol. C 143, 162–170. Jung, J.H., Kim, S.J., Lee, T.K., Shim, W.J., Woo, S., Kim, D.J., Han, C.H., 2008. Biomarker responses in caged rockfish (Sebastes schlegeli) from Masan Bay and Haegeumgang, South Korea. Mar. Pollut. Bull. 57, 599–606. Kang, J.J., Fang, H.W., 1997. Polycyclic aromatic hydrocarbons inhibit the activity of acetylcholinesterase purified from electric eel. Biochem. Biophys. Res. Commun. 238, 367–379. Krahn, M.M., Burrows, D.G., MacLeod Jr., W.D., Malins, D.C., 1987. Determination of individual metabolites of aromatic compounds in hydrolyzed bile of English sole (Parophrys vetulus) from polluted sites in Puget Sound, Washington. Arch. Environ. Contam. Toxicol. 16, 511–522. Lin, S., Bullock, P.L., Addison, R.F., Bandiera, S.M., 1997. Detection of cytochrome P450 1A1 in several species using antibody against a synthetic peptide derived from a rainbow trout cytochrome P450 1A1. Environ. Chem. Toxicol. 17, 439–445.
223
Lunnel, T., 1993. Dispersion: oil droplet size measurements at sea. Proceedings of the Sixteenth Arctic and Marine Oil spill Program Technical Seminar, pp. 1023–1056. Meyers, J.N., Nacci, D.E., Di Giulio, R.T., 2002. Cytochrome P450A (CYP1A) in killifish (Fundulus heteroclitus): heritability of altered expression and relationship to survival in contaminated sediments. Toxicol. Sci. 68, 69–81. McCain, M.S., Myers, D.W., Brown, M.M., Krahn, W.T., Roubal, M.H., Schi1ewe, J.T., Landahl, J.T., Chan, S.-L., 1987. Field and laboratory studies of the etiology of liver neoplasms in marine fish from Puget Sound. Environ. Health Perspect. 71, 5–16. Mora, P., Fournier, D., Narbonne, J.F., 1999. Cholinesterase from the marine mussels Mytilus galloprovincialis L. and Mytilus edulus L. from the freshwater bivalve Corbicula fluminea Muller. Comp. Biochem. Physiol., Part C Pharmacol. Toxicol. 122, 353–361. Moreira, S.M., Guilhermino, L., 2005. The use of the Mytilus galloprovincialis acetylcholinesterase and glutathione S-transferases activities as biomarkers of environmental contamination along the North West Portuguese coast. Environ. Monit. Assess. 105, 309–325. Moreira, S.M., Moreira-Santos, M., Ribeiro, R., Guilhermino, L., 2004. The “Coral Bulker” fuel oil spill on the north coast of Portugal: spatial and temporal biomarker responses in Mytilus galloprovincialis. Ecotoxicol 13, 619–630. Nalco Energy Services. http://www.nalco.co.kr. Nicolas, J.M., 1999. Vitellogenesis in fish and the effects of polycyclic aromatic hydrocarbon contaminants. Aquat. Toxicol. 45, 77–90. Payne, J.F., Bauld, C., Dey, A.C., Kiceniuk, J.W., Williams, U., 1984. Selectivity of mixedfunction oxygenase enzyme induction in flounder (Pseudopleuronectes americanus) collected at the site of the Baie Verte, Newfoundland oil spill. Comp. Biochem. Physiol. 79C, 15–19. Payne, J.F., Fancy, L.L., Rahimtula, A.D., Porter, E.L., 1987. Review and perspective on the use of mixed function oxygenase enzymes in biological monitoring. Comp. Biochem. Physiol. 86C, 233–245. Ramachandran, S.D., Hodson, P.V., Khan, C.W., Lee, K., 2004. Oil dispersant increases PAH uptake by fish exposed to crude oil. Ecotoxicol. Environ. Saf. 59, 300–308. Ramachandran, S.D., Michael, J.S., Hodson, P.V., Boudreau, M., Courtenay, S.C., Lee, K., Thomas, K., Jennifer, A.D., 2006. Influence of salinity and fish species on PAH uptake from dispersed crude oil. Mar. Pollut. Bull. 52, 1182–1189. Singer, M.M., George, S., Benner, D., Jacobson, S., Tjeerdema, R.S., Sowby, M.L., 1993. Comparative toxicity of two oil dispersants to the early life stages of two marine species. Environ. Toxicol. Chem. 12, 1855–1863. Singer, M.M., Aurand, D., Bragins, G.E., Clarks, J.R., Coelho, G.M., Sowby, M.L., Tjeerdema, R.S., 2000. Standardization of the preparation and quantitation of water-accommodated fractions of petroleum for toxicity testing. Mar. Pollut. Bull. 40, 1007–1016. Sonstergard, R.A., Leatherland, J.F., 1984. Comparative epidemiology: the use of fishes in assessing carcinogenic contaminants. In: Cairs, V., et al. (Ed.), Contaminant Effects on Fisheries: Adv. Environ. Sci. Technol., vol. 16, pp. 223–232. Thomas, R.E., Rice, S.D., 1981. Excretion of aromatic hydrocarbons and their metabolites by freshwater and seawater Dolly Varden char. In: Vernberg, F.J., Calabrese, A., Thurberg, F., Vernberg, W. (Eds.), Biological Monitoring of Marine Pollutants. InAcademic Press, New York, pp. 425–448. Tintos, A., Gesto, M., Mĺguez, J.M., Soengas, J.L., 2007. Naphthalene treatment alters liver intermediary metabolism and levels of steroid hormone in plasma of rainbow trout (Oncorhynchus mykiss). Ecotoxicol. Environ. Saf. 69, 180–186. Van Veld, P.A., 1990. Absorption and metabolism of dietary xenobiotics by the intestine of fish. Rev. Aquat. Sci. 2, 185–203. Van der weiden, M.E.J., Hanegraaf, F.H.M., Eggens, M.L., Celander, M., Seinen, W., Van derBerg, M.,1994. Temporal induction of cytochrome P450 1A in the mirror carp (Cyprinus carpio) after administration of several polycyclic aromatic hydrocarbons. Environ. Toxicol. Chem. 13, 797–802. Whyte, J.J., Jung, R.E., Schmit, C.J., Tillitt, D.E., 2000. Ethoxyresorufin O-deethylase (EROD) activity in fish as a biomarker of chemical exposure. Crit. Rev. Toxicol. 30, 347–570.