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Biochemical biomarkers in Oreochromis niloticus exposed to mixtures of benzo[a]pyrene and diazinon.
ARTICLE IN PRESS Ecotoxicology and Environmental Safety 73 (2010) 858–863
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Biochemical biomarkers in Oreochromis niloticus exposed to mixtures of benzo[a]pyrene and diazinon. Camila Pereira Trı´dico, Aline Cristina Ferreira Rodrigues, Lilian Nogueira, Daniele Caetano da Silva, Altair Benedito Moreira, Eduardo Alves de Almeida n ~ Colombo 2265, CEP 15054-000, Sao ~ Jose´ do Rio Preto, SP, Brazil. ´vao Departamento de Quı´mica e Ciˆencias Ambientais, IBILCE / UNESP, R. Cristo
a r t i c l e in f o
a b s t r a c t
Article history: Received 16 November 2009 Received in revised form 19 January 2010 Accepted 21 January 2010 Available online 9 February 2010
Biochemical biomarkers (the activities of acetylcholinesterase, 7-ethoxyresorufin-O-deetilase, carboxylesterase, catalase, glutathione peroxidase and glutathione S-transferase) were evaluated in Nile tilapia (Oreochromis niloticus) that had been exposed to benzo[a]pyrene (BaP) and the organophosphate pesticide diazinon (DZ), at 0.5 mg/L. The animals were pre-exposed to BaP for three days, and DZ was then added to both non-exposed and pre-exposed groups, being exposed for 2 and 7 additional days. The level of BaP was also measured in the bile. BaP caused the induction of phase I and II enzymes, and DZ caused carboxylesterase inhibition in gills but not in liver. AChE activity was unchanged. No significant modulation was observed in antioxidant enzymes. When in combination with BaP, DZ caused a significant decrease of EROD and GST induction. Levels of BaP in the bile were also increased in fish exposed to BaP combined with DZ, indicating an interference of DZ in responses activated by BaP. & 2010 Elsevier Inc. All rights reserved.
Keywords: Biomarkers Benzo[a]pyrene Organophosphate pesticides Tilapia
1. Introduction The measurement of induction of ethoxyresorufin-O-Deethylase (EROD, indicative of cytochrome P450 1A) in fish has been extensively used as a contamination biomarker to identify the exposure to organic xenobiotics, such as polycyclic aromatic hydrocarbons (PAHs), dioxins and polychlorobyphenyls (Flammarion et al., 1996; Whyte et al., 2000). The inhibition of esterases by organophosphate or carbamate pesticides (OPs and CMs) is also used as exposure and effect biomarkers for these compounds in environmental monitoring (Vioque-Ferna´ndez et al., 2007a,b, 2009). However, data on the effects of complex mixtures of these contaminants in fish biomarkers are scarce. It has been known that phosphorothionate pesticides are poor cholinesterase inhibitors, and must undergo activation to the oxon form by cytochrome P450 isoforms in order to inhibit esterases (Straus et al., 2000). Nevertheless, it has been also demonstrated that metabolites generated during P450-mediated conversion of thionates to the oxon forms are capable of inhibiting cytochrome P450 themselves (Tang et al., 2002). It was proposed that CYP-mediated desulfuration of chlorpyrifos in rats produces chlorpyrifos-oxon (Fukuto, 1990) in the process releasing the sulfur ion, which can then suppress CYP activity through binding onto the heme group (Tang et al., 2002).
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Corresponding author. Fax: + 55 17 3221 2356. E-mail addresses: [email protected], [email protected] (E.A. de Almeida). 0147-6513/$ - see front matter & 2010 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2010.01.016
Suppression of CYP1A by OPs has been also described for fish (Flammarion et al., 1998). Similarly, Wheelock et al. (2005) observed a 30% decrease in CYP1A levels in Chinook salmon (Oncorhynchus tshawytscha) that had been exposed to chlorpyrifos. Apart from desulfuration activity, OPs can also induce CYP isoforms involved in their detoxification (dearylation) (Neal and Halpert, 1982). Therefore, it is evident that other compounds that induce P450s could alter both the toxicity of the parent insecticide as well as the timing of esterase inhibition via the production of the oxon-derivative insecticide. Strauss et al. (2000) demonstrated that P450 isoforms induced by Aloclor 1254 are not able to increase desulfuration or dearylation of the OP chlorpyrifos, but this idea was not tested for other P450 inducers, such as the PAH benzo[a]pyrene (BaP). Depending on the isoform profile of P450s and the substrate specificities of the induced isoform(s), induction can lead to either increases in the production of activated metabolites with enhanced toxicity or increased detoxication with enhanced protection from toxicity (Binder et al., 1984; Stegeman and Hahn, 1994; Straus et al., 2000). Moreover, if cytochrome P450 isoforms are inhibited in the activation process, the inhibition can impair the phase I defense mechanisms involved in detoxification of most contaminants that could be present in combination with OPs in aquatic environments that have been contaminated by other contaminants. Diazinon (DZ) is a moderately persistent organophosphorothionate pesticide largely used in agriculture (Larkin and Tjeerdema, 2000). Its toxicity to animals is due to classical inhibition of cholinesterases, which poses risks to non-target
ARTICLE IN PRESS C. Pereira Trı´dico et al. / Ecotoxicology and Environmental Safety 73 (2010) 858–863
organisms that inhabit natural environments close to agricultural fields. However, there are no data concerning DZs toxic effects when in combination with CYP inducers. It is possible that, in combination with other chemicals, DZ could exert a different toxic response in aquatic organisms, and this remains to be studied. Exposure to PAHs and OPs has been related to oxidative stress generation in aquatic animals due to the production of reactive oxygen species (ROS) during their metabolism in cells (Winston and Di Giulio, 1991; Bagchi et al., 1995; Gultekin et al., 2000, 2001; Valavanidis et al., 2006, van der Oost et al., 2003). In healthy organisms, there is a balance between production and elimination of ROS. When the production of ROS is greater than its removal, an oxidative stress condition can be established. To deal with oxidative stress, the cells possess antioxidant defenses, such as the enzymes catalase (CAT), glutathione peroxidase (GPx) and also GST, a phase II biotransformation enzyme with peroxidase activity (Torres et al., 2002; Almeida et al., 2005, 2007; Limo´nPacheco and Gonsebatt, 2009). It has been shown that BaP exposure can significantly induce phase I and II biotransformation enzymes and oxidative stress. However, the main effect of DZ in organisms is esterase inhibition, even though oxidative stress has also been involved in the toxicology of both compounds. Even so, there are no data on the effect of these contaminants together in aquatic organisms. Thus, the goal of this study was to evaluate various biochemical responses of tilapias after controlled exposure to BaP and DZ alone or in combination. DZ is a thion OP that should be converted to the oxon form by cytochrome P450 to inhibit esterases. Therefore, we were interested to know whether P450 isoforms induced after BaP pre-exposure are able to either increase or decrease esterase inhibition after DZ exposure. We were further interested in knowing whether DZ in combination with BaP is able to decrease the EROD induction caused by BaP, and if there is any other influence of one of these compounds in the oxidative stress responses activated by the other contaminant after exposure to mixtures of both compounds. Levels of BaP in the bile of the tilapias were also monitored during the experiment. Tilapia, a wide-spread fish species in tropical countries, is the most cultivated freshwater fish in Brazil, with increasing occurrence in Brazilian natural environments. Nile tilapia, which is a hardy fish, is an omnivorous feeder and feed by frequent foraging in the whole water column with occasional contact with sediments (Pathiratne et al., 2009). These characteristics make Nile tilapia a suitable indicator species for bio-monitoring of aquatic pollution in tropical environments (Rodrı´guez-Fuentes and Gold-Bouchot, 2004; Gold-Bouchot et al. 2006).
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aquariums and two aquariums that had previously contained BaP. At this point, we had two aquariums containing fish that were not exposed to any compound (control groups), two aquariums treated with DZ (DZ group), two aquariums containing the pre-treatment with BaP (BaP group) and two aquariums that received DZ treatment after BaP exposure (BaP + DZ). After 2 days, one aquarium from each group was collected for testing. The remaining aquariums in each group were left for another 5 days. This left us with exposure periods of 2 days (for the first set of aquariums) and 7 days (for the second set of aquariums). Because we were interested in verifying whether BaP exposure could increase esterase inhibition, we previously tested the doses for DZ in order to establish a dose that did not cause a strong esterase inhibition (data not shown). When testing 2 mg/L of DZ, all tilapias died after 5 days. Strong esterase inhibition was observed after both 2 and 7 exposure days to DZ at 1 mg/L. We then tested 0.5 mg/L and observed only slight and not significant acetylcholinesterase inhibition, and so this concentration was chosen for the present experiment. After 2 days for the first set and after 7 days for the second set, of the addition of DZ to the aquariums, all fish from each group were killed and had their livers, gills, and brains dissected and frozen at 80 1C for biochemical analyses. All experiments were conducted in accordance with national and institutional guidelines for the protection of animal welfare. It is also important to note that, before being discharged into a public sewer, water from the aquaria was filtered through activated charcoal filters in order to retain the toxic waste, thus preventing improper disposal into the environment. 2.2. Sample preparation and enzymatic analyses For the analyses of esterases, tissues were homogenized (1:4, w-vol) in a 0.1 M Tris buffer, pH 8.0, and centrifuged for 30 min at 30,000g. The supernatant fraction was then collected and frozen at 80 1C. The activities of AChE and CbE were measured and then characterized in the supernatant, following the methodology described by Ellman et al. (1961). Acetylthiocholine and phenylthioacetate were used as substrates for the activities of AChE and CbE, respectively, as previously described (Vioque-Ferna´ndez et al., 2007a,b). For the EROD, GST, GPx and CAT analyses, liver and gill tissues were homogenized (1:4, w/v) in a Tris 20 mM buffer (pH 7.4), sucrose 0.5 mM, KCl 0.15 mM and 1 mM protease inhibitor (PMSF). The samples were then centrifuged at 9000g for 20 min at 4 1C. To obtain the microsomal and cytosolic fractions, the supernatant portion was collected and centrifuged at 50,000g for one additional hour at 4 1C. The enzymatic activities of CAT, GPx and GST were analyzed in the cytosolic fraction, while the EROD activity was measured in the liver microsomal fraction only, as indicative of CYP1A. The activity of CAT was quantified at 240 nm by the H2O2 decomposition (Beutler, 1975). GPx activity was assayed by the oxidation of NADPH (linked to GSSG reduction by excess glutathione reductase) at 340 nm, using t-butyl hydroperoxide as substrate (Sies et al. 1979). GST activity was determined by measuring the increase in absorbance at 340 nm, incubating reduced glutathione (GSH) and 1-chloro-2, 4-dinithrobenzene (CDNB) as substrates (Keen et al. 1976). The EROD activity was measured following the method of Burke and Mayer (1974), but with some modifications. The assay mixture contained 1950 mL of potassium phosphate buffer 80 mM (pH7.4), 20 mL of 7- ethoxyresorufin 335 mM, 20 mL of NADPH 20 mM and 10 mL of microsomal liver extract. The reaction was monitored for three minutes at 30 1C. EROD activity (pmol/min/mg of protein) was calculated based on a resorufin standard curve that had been previously prepared. Protein levels were measured following the method of Bradford (1976) and using bovine serum albumin as standard. 2.3. Bile collection and quantification of benzo[a]pyrene
2. Material and methods 2.1. Exposure experiments Sixty Nile tilapia (Oreochromis niloticus) were obtained from the ‘‘Tilapias do Brasil’’ fisheries located in Buritama, Sa~ o Paulo, Brazil, and brought to the laboratory. They were acclimated for one week before the experiments took place. To avoid the interference of gender on biomarker responses, only males were used. The tilapias were divided into ten 50 40 60 cm aquariums, each containing 120 L of water and six animals per aquarium. The aquariums were kept under constant aeration. After the acclimatization period, BaP was added (0.5 mg/L) to five of the ten aquariums for a pre-exposure period of three days. B[a]P dose was defined based on previous works published by Ortiz-Delgado et al. (2005, 2008), in which EROD and P450 1A were highly induced. After the pre-exposure period, the six fish from one of the BaP-exposed aquariums and six fish from one non-exposed aquarium were collected. They were then anesthetized with 2 mL/L of phenoxyethanol, and had their livers, gills and brains collected and frozen at 80 1C. After this process, we had eight remaining aquariums: four without treatment and four pre-exposed to BaP. After the collection of the pre-exposed and non-exposed fish (control and preexposure groups), DZ was added (0.5 mg/L) to two of the remaining untreated
Bile was extracted from the fish with a hypodermic syringe, transferred to an eppendorf microtube, which was then sealed and stored in ice until the analyses. The procedure for BaP determination followed the recommendations described by Ariese et al. (1993). A standard stock solution of BaP (0.15 g/L) in dichloromethane was prepared. From this solution, another standard solution was prepared at 2.00 mg/L in ethanol/water 48% (v/v), for the construction of a standard calibration curve. The bile extracted from the tilapias was diluted in the proportion of 1:1000 with ethanol/water 48% (v/v), and measured by fluorescence in terms of a standard calibration curve of BaP from 1.0 to 20.0 mg/L similar to Ariese et al. (1993). Fluorescence was measured with a Cary Eclipse (Varian) spectrofluorimeter, which consisted of a xenon discharge lamp, two Monk-Gillieson monochromators, a Hamamatsu photomultiplier and quartz cells (1 1 cm). Slits for the excitation and emission monochromators were set at 5 nm, the photomultiplier voltage was adjusted to 600 mV, and the monochromator scan rates were 600 nm/min. The synchronous fluorescence peak of the BaP was located at 323 nm when Dl =25 nm. 2.4. Statistical Analysis For the statistical analysis, Cochrans test of equality error variances was carried out. As p 40.05 was obtained for all data by the Cochrans test, data were
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statistically compared using the Kruskal–Wallis test for non-parametric data followed by a Tukey post-hoc analysis, with the aid of Statistica 7.0 software. Only p o0.05 was accepted as significant for indicating statistical differences between groups with different treatments under the same exposure time. Data are presented as mean + standard deviation.
3. Results Table 1 shows the activities of CbE, GST, GPx, CAT and EROD in the liver of O. niloticus exposed to BaP, DZ and BaP+ DZ for exposure periods of 2 and 7 days. No differences were observed in the activities of CbE, CAT and GPx throughout the experiment for any chemicals tested. However, pre-exposure to BaP for three days caused an increase in GST activity, which did not continue after 2 days, but significantly increased after 7 days of exposure. No such increase was observed in fish exposed for 2 and 7 days to DZ or BaP+ DZ. Moreover, BaP caused a strong increase in EROD activity after the three-day pre-exposure period, and after both 2 and 7 days. When compared to the control group, increased EROD activity was also observed for animals exposed to the mixture of BaP+ DZ for periods of both 2 and 7 days after the pre-exposure period, but this increase was lower than that observed for BaP alone (not significant). Exposure to DZ alone caused no difference in EROD activity when compared to the respective controls. Enzymatic activities measured in the gills are presented in Table 2. In this tissue, CAT activity was the only parameter insensitive to all treatments. DZ was able to decrease CbE activity when alone and when in combination with BaP, but only after 2 days of exposure. No difference in CbE was observed after 7 days of exposure for fish exposed to DZ or BaP+ DZ, nor in fish exposed to BaP alone after 2 and 7 exposure days. GST activity increased in fish exposed to BaP after all exposure times when compared to control values. When in combination with DZ, BaP caused a similar increase in GST activity after 2 days of exposure, but caused no difference (compared to control values) after 7 days. No treatment caused changes in GPx activity compared to the activity in control animals, but those exposed to BaP for 7 days after the
pre-treatment exposure had increased GPx activity compared to those exposed to DZ. The activity of AChE in the brain is presented in Table 3. No differences were observed for all treatments throughout the experiment. In Table 4, BaP levels in the bile of tilapias exposed to the contaminants are shown. BaP levels in bile were significantly increased in the bile of fish exposed to BaP and BaP+ DZ during all exposure periods. Moreover, it seems that the fish exposed to BaP+ DZ presented higher levels of BaP in the bile when compared to animals exposed to BaP alone. This was statistically significant for 7 exposure days. Also, levels of BaP in the bile seemed to decrease during the experiment in fish exposed to BaP and BaP+ DZ. DZ exposure did not cause any variation in BaP levels in the bile when compared to control values.
4. Discussion Pesticides and PAHs are common pollutants in the aquatic environment. In some field studies where oxidative stress parameters, EROD induction, and esterase inhibition can be used to monitor water pollution, the joint presence of pesticides and PAHs may lead to different responses in fish compared to the responses typically observed when the animals are exposed to a single contaminant. Therefore, it is vital to establish the responses that are typically activated in fish species exposed to mixtures of different classes of pollutants under controlled conditions. With this information, scientists can better understand the data that has originated from monitoring studies of aquatic environments contaminated by multi-xenobiotics. In this study, we focused on the effects of BaP plus DZ in O. niloticus that were exposed for 2 and 7 days at sub-lethal concentrations. In general, BaP and DZ alone caused typical responses in tilapias: BaP caused a strong induction of EROD activity in the liver and increased GST activity in the liver and gills, with a stronger effect on the gills. DZ exposure caused the inhibition of CbE but had no effects on AChE activity, since a low dose was
Table 1 Enzymatic activities measured in the liver of O. niloticus exposed to BaP and DZ. Enzymea
Group
CbE
Control BaP DZ BaP +DZ
GST
Control BaP DZ BaP +DZ
Pre-exposure
2 days after pre-exposure
7 days after pre-exposure
1.39 7 0.85 1.75 7 0.66
1.05 7 0.50 1.61 70.85 1.36 70.55 0.85 7 0.06
1.23 70.27 1.21 7 0.20 0.83 7 0.07 0.85 70.37
1.50 7 0.39 2.22 70.48b
1.81 7 70.85 2.30 70.68 2.12 70.54 2.16 70.97
1.13 70.32 1.42 70.31d 1.14 7 0.30 0.87 7 0.06
– –
– –
Control BaP DZ BaP +DZ
0.027 0.01 0.0267 0.01 – –
0.018 7 0.01 0.027 0.01 0.022 7 0.01 0.019 7 0.01
0.037 0.01 0.047 0.01 0.047 0.01 0.037 0.00
CAT
Control BaP DZ BaP +DZ
99.07 7 20.91 89.54 732.15 – –
67.17 720.43 69.80 710.65 75.84 7 15.92 64.24 78.67
72.12 710.70 69.33 7 16.19 76.16 7 14.11 73.44 7 13.65
EROD
Control BaP DZ BaP +DZ
0.86 71.11 21.79 75.03b – –
0.91 7 0.40 44.12 7 13.13b,c 1.01 7 0.90 33.45 721.60b,e
2.64 72.54 23.04 7 17.11b,c 2.26 73.26 10.26 78.76
GPx
a
Activity expressed in mmol/min/mg protein (CbE, GST, GPx and CAT) or pmol/min/mg protein (EROD). Significant difference compared to the respective control. c Significant difference between DZ and BaP at the same exposure time. d Significant difference between BaP and BaP +DZ at the same exposure time. e Significant difference between DZ and BaP +DZ at the same exposure time. Data expressed as mean 7 standard deviation. b
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Table 2 Enzymatic activities measured in the gill of O. niloticus exposed to BaP and DZ. Enzymea
Group
Pre-exposure
2 days after pre-exposure
CbE
Control BaP DZ BaP +DZ
0.087 0.03 0.077 0.03 – –
0.13 70.04 0.117 0.02c,d 0.057 0.01b 0.067 0.01b
0.107 0.05 0.107 0.05 0.077 0.02 0.077 0.00
GST
Control BaP DZ BaP +DZ
0.127 0.02 0.287 0.04b – –
0.12 70.03 0.207 0.03b,c 0.10 70.03e 0.187 0.04b
0.117 0.01 0.16 7 0.05b,c 0.087 0.04 0.127 0.02
GPx
Control BaP DZ BaP +DZ
0.03 70.00 0.037 0.01 – –
0.05 70.01 0.05 70.02 0.05 70.01 0.04 70.01
0.077 0.01 0.107 0.01c 0.077 0.03 0.077 0.01
CAT
Control BaP DZ BaP +DZ
6.317 0.80 6.69 7 1.13 – –
8.04 7 2.59 10.28 7 2.33 9.30 7 3.31 9.76 7 1.57
5.17 71.77 9.68 7 2.90 7.34 73.27 6.90 7 0.79
7 days after pre-exposure
a
Activity expressed in mmol/min/mg protein (CbE, GST, GPx and CAT) or pmol/min/mg protein (EROD). Significant difference compared to the respective control. Significant difference between DZ and BaP at the same exposure time. d Significant difference between BaP and BaP +DZ at the same exposure time. e Significant difference between DZ and BaP + DZ at the same exposure time. Data expressed as mean 7 standard deviation. b c
Table 3 AChE activity in the brain of O. niloticus after xenobiotics exposure. Enzyme
Group
Pre-exposure
2 days after pre-exposure
7 days after pre-exposure
AChEa
Control BaP DZ BaP+ DZ
0.05 7 0.02 0.06 7 0.01 – –
0.0507 0.01 0.053 7 0.01 0.067 0.02 0.067 0.01
0.04 70.01 0.06 70.02 0.06 70.02 0.08 70.01
a
Activity expressed in mmol/min/mg protein. Data expressed as mean7 standard deviation.
Table 4 Levels of BaP in bile of Oreochromis niloticus after xenobiotics exposure.
a
BaP
Group
Pre-exposure
2 days after pre-exposure
7 days after pre-exposure
Control BaP DZ BaP +DZ
0.077 0.02 163.287 78.56b,c – –
0.107 0.02 122.227 48.44b,c 0.10 70.02e 164.687 125.32b
0.057 0.02 78.54 7 28.80b,c,d 0.07 70.02e 110.29 7 0.26b
a
Data expressed in mg/mL of bile. Significant difference compared to the respective control. c Significant difference between DZ and BaP at the same exposure time. d Significant difference between BaP and BaP + DZ at the same exposure time. e Significant difference between DZ and BaP + DZ at the same exposure time. Data expressed as mean 7 standard deviation. b
used. It is possible that AChE activity could be inhibited in other tissues, but this enzyme was evaluated only in brain, and further studies are needed to verify if AChE is inhibited in other tissues by DZ dose used in the present work. As for oxidative stress parameters (GPx and CAT), no effects were observed after BaP or after DZ exposure when compared to control values, even though the tilapias exposed to BaP had higher activity than fish exposed to BaP in combination with DZ, suggesting a possible inhibition effect of DZ on GPx activation by BaP. DZ alone did not cause any effects on EROD activity (compared to control values), which suggests that EROD is not activated by DZ and may not be related to desulfuration or dearylation of this OP. Paolini et al. (1997) have demonstrated that the main P450 isoforms induced by the OP acephate in mice livers were 3A, 1A2
and 2E1, and that this OP suppresses CYPs 2B1 and 1A1. Tang et al. (2002) showed that CYP2B6 has the main in vitro desulfuration activity on the OP chlorpyrifos among all human CYP isoforms tested. They also showed that chlorpyrifos is a suicide substrate for this isoform, which decreased CYP activity during incubation periods. Compared to control values, EROD activity in our study was similar in fishes exposed to DZ, indicating that CYP1A suppression did not occur. However, EROD induction was lower in those fish exposed to BaP and DZ than in those exposed to BaP, though the differences were not statistically significant. It seems that DZ interferes with EROD induction that is promoted by BaP, and this result can have profound implications for monitoring studies that use biomarker analysis. The lack of EROD induction in fish from
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monitoring studies could be misinterpreted as the absence of PAHs in the environment. This study shows, however, that this result could be due to a concomitant presence of OPs and PAHs, with an inhibition effect of these compounds on EROD induction. Flammarion et al. (1998) have presented very similar results with the cyprinid fish gudgeon (Gobio gobio) pre-exposed (intraperitoneal injection) to b-naphtoflavone (BNF, another CYP1A inducer) and the OP methidathion. These results corroborate our study. In that study, methidathion exposure led to a 60–70% decrease of the EROD induction caused by pre-treatment with BNF. A similar profile was also observed for GST activity. BaP was able to cause significant GST induction after the pre-exposure period. After exposure periods of both 2 and 7 days, BaP caused significant induction in the gills, and after the pre-exposure period and 7 days after pre-exposure, BaP caused significant induction in the liver. However, the fish exposed to BaP +DZ presented lower GST induction after 7 days, clearly demonstrating that DZ also interferes with GST induction promoted by BaP. Similarly, Hodge et al. (2000) have shown that DZ and chlorpyrifos are able to cause GST inhibition in an insect (Micromus tasmaniae). Again, the lack of GST induction in fish from field studies could be due to the concomitant presence of PAHs and OPs in the environment, and is not necessarily due to the absence of PAH exposure. EROD induced by BaP had no effects on esterase activities, since no differences were observed in AChE activity throughout the experiments, and considering that the inhibition profile of CbE in gills after two exposure days was identical for fish exposed to DZ or BaP+ DZ. Thus, it can be also supposed that P450 isoforms activated by BaP are not related to desulfuration or dearylation of OPs. Straus et al. (2000) studied the activities of desulfuration and dearylation of the phosphorothionates chlorpyrifos and parathion in channel catfishes exposed to Aroclor 1254, another EROD inducer. Although chlorpyrifos or parathion seems to induce desulfuration and dearylation activities, these enzymes were not altered after Aroclor 1254 exposure, indicating that this compound does not induce the P450 isoforms responsible for desulfuration or dearylation of phosphorothionates, and also that the isoforms responsible for these activities are probably not CYP1A. Although it was not tested for B[a]P exposure, we can assume that B[a]P is also not able to increase the metabolism of the DZ even when EROD activity is increased. Finally, it was observed that the combination of PAH and OP exposure also interferes with levels of BaP excretion in the bile. DZ alone did not cause any effects on BaP excretion, since the levels of this compound in the bile did not change after the animals were exposed to DZ. However, BaP exposure caused a very significant increase in BaP levels in bile. BaP levels were higher after the three days of pre-exposure, and then decreased during the exposure time. Because the fish were exposed to a single dose of BaP, it should be expected that BaP concentration in water diminishes during the exposure period, since it was being metabolized in the fish. A similar trend was observed for fish exposed to BaP + DZ, but in this case, the decrease in BaP concentration in bile was less intense. In our study, it could be supposed that the interference of DZ with EROD and GST induction caused by BaP resulted in a lower BaP metabolization rate. It has been shown that the main product of BaP metabolization through CYP1A and GST involved the formation of hydroxyl, epoxy, diol or quinone BaP derivatives (Miller and Ramos, 2001), and some of these can be conjugated with glutathione by GST. Therefore, there are different BaP metabolites that can be excreted to the bile after phase I and II metabolism. As DZ interferes with phase I and II enzymes, the rate of unmodified BaP excretion could then increase as an alternative pathway to compensate for reduced biotransformation enzymes.
This mechanism could account for the increased levels of BaP in the bile of the fish exposed to BaP+ DZ compared to those exposed to BaP alone after 2 and 7 exposure days. Mdegela et al. (2006) observed an inhibitory effect of the hormone 17b-estradiol on EROD activity induced by BaP in gills of African sharptooth catfish (Clarias gariepinus). They also observed higher levels of BaP in the bile of fish exposed to the hormone in combination with BaP, compared to those fishes exposed to BaP alone, and suggested that the amount of BaP reaching the liver was increased by the hormone through the inhibition of gill filament EROD activities. Relatively lower hepatic EROD activity and accumulation of bile metabolites in fish exposed to BaP alone indicates that much of the BaP was biotransformed in the gills and did not reach the liver. EROD activity was not measured in the gills in the present study, but such suggestions could account for the larger amount of BaP in the bile of fish that were exposed to BaP+ DZ.
5. Conclusions Complex mixtures of OPs and PAHs can account for response mechanisms in tilapias that are different from those observed with a single contaminant alone. Tilapias exposed to BaP alone presented typical responses with respect to phase I and II enzyme, which agreed with numerous other studies found in the literature. DZ alone inhibited CbE in gills but not in liver, and AChE activity in brain was unchanged. In combination, BaP +DZ had different effects on biochemical parameters. In our study, the main finding was that DZ can decrease the main biochemical parameters induced by BaP in fish, along with EROD and GST activities, and can also contribute to increased levels of BaP in the bile. These results have significant implications for monitoring studies of aquatic ecosystems contaminated by both classes of compounds.
Acknowledgements This study had the financial support of the ‘‘Fundac- a~ o de Amparo a Pesquisa do Estado de Sa~ o Paulo’’ (FAPESP, 2006/03873-1). Lilian Nogueira and Camila Pereira Trı´dico are recipients of FAPESP fellowships. The present study was conducted in accordance with national and institutional guidelines for the protection of animal welfare. References Almeida, E.A., Bainy, A.C.D., Dafre, A.L., Gomes, O.F., Medeiros, M.H.G., Di Mascio, P., 2005. Oxidative stress in digestive gland and gill of the brown mussel (Perna perna) exposed to air and re-submersed. J. Exp. Mar. Biol. Ecol. 318, 21–30. Almeida, E.A., Bainy, A.C.D., Loureiro, A.P.M., Martinez, G.R., Miyamoto, S., Onuki, J., Barbosa, L.F., Garcia, C.C.M., Prado, F.M., Ronsein, G.E., 2007. Oxidative stress in Perna perna and other bivalves as indicators of environmental stress in the Brazilian marine environment: antioxidants, lipid peroxidation and DNA damage. Comp. Biochem. Physiol. 145A, 588–600. Ariese, F., Kok, S.J., Verkaik, M., Gooijer, C., Velthorst, N.H., Hofstraat, J.W., 1993. Synchronous fluorescence spectrometry of fish bile: a rapid screening method for the biomonitoring of PAH exposure. Aquat. Toxicol. 26, 273–286. Bagchi, D., Bagchi, M., Hassoun, E.A., Stohs, S.J., 1995. In vitro and in vivo generation of reactive oxygen species, DNA damage and lactate dehydrogenase leakage by selected pesticides. Toxicology 104 (1-3), 129–140. Beutler, E., 1975. Red Cell Metabolism: A Manual of Biochemical Methods.. Grune & Stratton, New York. Binder, R.L., Melancon, M.J., Lech, J.J., 1984. Factors influencing the persistence and metabolism of chemicals in fish. Drug Metab. Rev. 15 (4), 697–724. Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 72, 248–254. Burke, M.D., Mayer, R.T., 1974. Ethoxyresorufin—direct fluorimetric assay of a microsomal o-dealkylation which is preferentially inducible by 3-methylcholanthrene. Drug Metab. Disp. 2, 583–588.
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Contents lists available at ScienceDirect
Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv
Biochemical biomarkers in Oreochromis niloticus exposed to mixtures of benzo[a]pyrene and diazinon. Camila Pereira Trı´dico, Aline Cristina Ferreira Rodrigues, Lilian Nogueira, Daniele Caetano da Silva, Altair Benedito Moreira, Eduardo Alves de Almeida n ~ Colombo 2265, CEP 15054-000, Sao ~ Jose´ do Rio Preto, SP, Brazil. ´vao Departamento de Quı´mica e Ciˆencias Ambientais, IBILCE / UNESP, R. Cristo
a r t i c l e in f o
a b s t r a c t
Article history: Received 16 November 2009 Received in revised form 19 January 2010 Accepted 21 January 2010 Available online 9 February 2010
Biochemical biomarkers (the activities of acetylcholinesterase, 7-ethoxyresorufin-O-deetilase, carboxylesterase, catalase, glutathione peroxidase and glutathione S-transferase) were evaluated in Nile tilapia (Oreochromis niloticus) that had been exposed to benzo[a]pyrene (BaP) and the organophosphate pesticide diazinon (DZ), at 0.5 mg/L. The animals were pre-exposed to BaP for three days, and DZ was then added to both non-exposed and pre-exposed groups, being exposed for 2 and 7 additional days. The level of BaP was also measured in the bile. BaP caused the induction of phase I and II enzymes, and DZ caused carboxylesterase inhibition in gills but not in liver. AChE activity was unchanged. No significant modulation was observed in antioxidant enzymes. When in combination with BaP, DZ caused a significant decrease of EROD and GST induction. Levels of BaP in the bile were also increased in fish exposed to BaP combined with DZ, indicating an interference of DZ in responses activated by BaP. & 2010 Elsevier Inc. All rights reserved.
Keywords: Biomarkers Benzo[a]pyrene Organophosphate pesticides Tilapia
1. Introduction The measurement of induction of ethoxyresorufin-O-Deethylase (EROD, indicative of cytochrome P450 1A) in fish has been extensively used as a contamination biomarker to identify the exposure to organic xenobiotics, such as polycyclic aromatic hydrocarbons (PAHs), dioxins and polychlorobyphenyls (Flammarion et al., 1996; Whyte et al., 2000). The inhibition of esterases by organophosphate or carbamate pesticides (OPs and CMs) is also used as exposure and effect biomarkers for these compounds in environmental monitoring (Vioque-Ferna´ndez et al., 2007a,b, 2009). However, data on the effects of complex mixtures of these contaminants in fish biomarkers are scarce. It has been known that phosphorothionate pesticides are poor cholinesterase inhibitors, and must undergo activation to the oxon form by cytochrome P450 isoforms in order to inhibit esterases (Straus et al., 2000). Nevertheless, it has been also demonstrated that metabolites generated during P450-mediated conversion of thionates to the oxon forms are capable of inhibiting cytochrome P450 themselves (Tang et al., 2002). It was proposed that CYP-mediated desulfuration of chlorpyrifos in rats produces chlorpyrifos-oxon (Fukuto, 1990) in the process releasing the sulfur ion, which can then suppress CYP activity through binding onto the heme group (Tang et al., 2002).
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Corresponding author. Fax: + 55 17 3221 2356. E-mail addresses: [email protected], [email protected] (E.A. de Almeida). 0147-6513/$ - see front matter & 2010 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2010.01.016
Suppression of CYP1A by OPs has been also described for fish (Flammarion et al., 1998). Similarly, Wheelock et al. (2005) observed a 30% decrease in CYP1A levels in Chinook salmon (Oncorhynchus tshawytscha) that had been exposed to chlorpyrifos. Apart from desulfuration activity, OPs can also induce CYP isoforms involved in their detoxification (dearylation) (Neal and Halpert, 1982). Therefore, it is evident that other compounds that induce P450s could alter both the toxicity of the parent insecticide as well as the timing of esterase inhibition via the production of the oxon-derivative insecticide. Strauss et al. (2000) demonstrated that P450 isoforms induced by Aloclor 1254 are not able to increase desulfuration or dearylation of the OP chlorpyrifos, but this idea was not tested for other P450 inducers, such as the PAH benzo[a]pyrene (BaP). Depending on the isoform profile of P450s and the substrate specificities of the induced isoform(s), induction can lead to either increases in the production of activated metabolites with enhanced toxicity or increased detoxication with enhanced protection from toxicity (Binder et al., 1984; Stegeman and Hahn, 1994; Straus et al., 2000). Moreover, if cytochrome P450 isoforms are inhibited in the activation process, the inhibition can impair the phase I defense mechanisms involved in detoxification of most contaminants that could be present in combination with OPs in aquatic environments that have been contaminated by other contaminants. Diazinon (DZ) is a moderately persistent organophosphorothionate pesticide largely used in agriculture (Larkin and Tjeerdema, 2000). Its toxicity to animals is due to classical inhibition of cholinesterases, which poses risks to non-target
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organisms that inhabit natural environments close to agricultural fields. However, there are no data concerning DZs toxic effects when in combination with CYP inducers. It is possible that, in combination with other chemicals, DZ could exert a different toxic response in aquatic organisms, and this remains to be studied. Exposure to PAHs and OPs has been related to oxidative stress generation in aquatic animals due to the production of reactive oxygen species (ROS) during their metabolism in cells (Winston and Di Giulio, 1991; Bagchi et al., 1995; Gultekin et al., 2000, 2001; Valavanidis et al., 2006, van der Oost et al., 2003). In healthy organisms, there is a balance between production and elimination of ROS. When the production of ROS is greater than its removal, an oxidative stress condition can be established. To deal with oxidative stress, the cells possess antioxidant defenses, such as the enzymes catalase (CAT), glutathione peroxidase (GPx) and also GST, a phase II biotransformation enzyme with peroxidase activity (Torres et al., 2002; Almeida et al., 2005, 2007; Limo´nPacheco and Gonsebatt, 2009). It has been shown that BaP exposure can significantly induce phase I and II biotransformation enzymes and oxidative stress. However, the main effect of DZ in organisms is esterase inhibition, even though oxidative stress has also been involved in the toxicology of both compounds. Even so, there are no data on the effect of these contaminants together in aquatic organisms. Thus, the goal of this study was to evaluate various biochemical responses of tilapias after controlled exposure to BaP and DZ alone or in combination. DZ is a thion OP that should be converted to the oxon form by cytochrome P450 to inhibit esterases. Therefore, we were interested to know whether P450 isoforms induced after BaP pre-exposure are able to either increase or decrease esterase inhibition after DZ exposure. We were further interested in knowing whether DZ in combination with BaP is able to decrease the EROD induction caused by BaP, and if there is any other influence of one of these compounds in the oxidative stress responses activated by the other contaminant after exposure to mixtures of both compounds. Levels of BaP in the bile of the tilapias were also monitored during the experiment. Tilapia, a wide-spread fish species in tropical countries, is the most cultivated freshwater fish in Brazil, with increasing occurrence in Brazilian natural environments. Nile tilapia, which is a hardy fish, is an omnivorous feeder and feed by frequent foraging in the whole water column with occasional contact with sediments (Pathiratne et al., 2009). These characteristics make Nile tilapia a suitable indicator species for bio-monitoring of aquatic pollution in tropical environments (Rodrı´guez-Fuentes and Gold-Bouchot, 2004; Gold-Bouchot et al. 2006).
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aquariums and two aquariums that had previously contained BaP. At this point, we had two aquariums containing fish that were not exposed to any compound (control groups), two aquariums treated with DZ (DZ group), two aquariums containing the pre-treatment with BaP (BaP group) and two aquariums that received DZ treatment after BaP exposure (BaP + DZ). After 2 days, one aquarium from each group was collected for testing. The remaining aquariums in each group were left for another 5 days. This left us with exposure periods of 2 days (for the first set of aquariums) and 7 days (for the second set of aquariums). Because we were interested in verifying whether BaP exposure could increase esterase inhibition, we previously tested the doses for DZ in order to establish a dose that did not cause a strong esterase inhibition (data not shown). When testing 2 mg/L of DZ, all tilapias died after 5 days. Strong esterase inhibition was observed after both 2 and 7 exposure days to DZ at 1 mg/L. We then tested 0.5 mg/L and observed only slight and not significant acetylcholinesterase inhibition, and so this concentration was chosen for the present experiment. After 2 days for the first set and after 7 days for the second set, of the addition of DZ to the aquariums, all fish from each group were killed and had their livers, gills, and brains dissected and frozen at 80 1C for biochemical analyses. All experiments were conducted in accordance with national and institutional guidelines for the protection of animal welfare. It is also important to note that, before being discharged into a public sewer, water from the aquaria was filtered through activated charcoal filters in order to retain the toxic waste, thus preventing improper disposal into the environment. 2.2. Sample preparation and enzymatic analyses For the analyses of esterases, tissues were homogenized (1:4, w-vol) in a 0.1 M Tris buffer, pH 8.0, and centrifuged for 30 min at 30,000g. The supernatant fraction was then collected and frozen at 80 1C. The activities of AChE and CbE were measured and then characterized in the supernatant, following the methodology described by Ellman et al. (1961). Acetylthiocholine and phenylthioacetate were used as substrates for the activities of AChE and CbE, respectively, as previously described (Vioque-Ferna´ndez et al., 2007a,b). For the EROD, GST, GPx and CAT analyses, liver and gill tissues were homogenized (1:4, w/v) in a Tris 20 mM buffer (pH 7.4), sucrose 0.5 mM, KCl 0.15 mM and 1 mM protease inhibitor (PMSF). The samples were then centrifuged at 9000g for 20 min at 4 1C. To obtain the microsomal and cytosolic fractions, the supernatant portion was collected and centrifuged at 50,000g for one additional hour at 4 1C. The enzymatic activities of CAT, GPx and GST were analyzed in the cytosolic fraction, while the EROD activity was measured in the liver microsomal fraction only, as indicative of CYP1A. The activity of CAT was quantified at 240 nm by the H2O2 decomposition (Beutler, 1975). GPx activity was assayed by the oxidation of NADPH (linked to GSSG reduction by excess glutathione reductase) at 340 nm, using t-butyl hydroperoxide as substrate (Sies et al. 1979). GST activity was determined by measuring the increase in absorbance at 340 nm, incubating reduced glutathione (GSH) and 1-chloro-2, 4-dinithrobenzene (CDNB) as substrates (Keen et al. 1976). The EROD activity was measured following the method of Burke and Mayer (1974), but with some modifications. The assay mixture contained 1950 mL of potassium phosphate buffer 80 mM (pH7.4), 20 mL of 7- ethoxyresorufin 335 mM, 20 mL of NADPH 20 mM and 10 mL of microsomal liver extract. The reaction was monitored for three minutes at 30 1C. EROD activity (pmol/min/mg of protein) was calculated based on a resorufin standard curve that had been previously prepared. Protein levels were measured following the method of Bradford (1976) and using bovine serum albumin as standard. 2.3. Bile collection and quantification of benzo[a]pyrene
2. Material and methods 2.1. Exposure experiments Sixty Nile tilapia (Oreochromis niloticus) were obtained from the ‘‘Tilapias do Brasil’’ fisheries located in Buritama, Sa~ o Paulo, Brazil, and brought to the laboratory. They were acclimated for one week before the experiments took place. To avoid the interference of gender on biomarker responses, only males were used. The tilapias were divided into ten 50 40 60 cm aquariums, each containing 120 L of water and six animals per aquarium. The aquariums were kept under constant aeration. After the acclimatization period, BaP was added (0.5 mg/L) to five of the ten aquariums for a pre-exposure period of three days. B[a]P dose was defined based on previous works published by Ortiz-Delgado et al. (2005, 2008), in which EROD and P450 1A were highly induced. After the pre-exposure period, the six fish from one of the BaP-exposed aquariums and six fish from one non-exposed aquarium were collected. They were then anesthetized with 2 mL/L of phenoxyethanol, and had their livers, gills and brains collected and frozen at 80 1C. After this process, we had eight remaining aquariums: four without treatment and four pre-exposed to BaP. After the collection of the pre-exposed and non-exposed fish (control and preexposure groups), DZ was added (0.5 mg/L) to two of the remaining untreated
Bile was extracted from the fish with a hypodermic syringe, transferred to an eppendorf microtube, which was then sealed and stored in ice until the analyses. The procedure for BaP determination followed the recommendations described by Ariese et al. (1993). A standard stock solution of BaP (0.15 g/L) in dichloromethane was prepared. From this solution, another standard solution was prepared at 2.00 mg/L in ethanol/water 48% (v/v), for the construction of a standard calibration curve. The bile extracted from the tilapias was diluted in the proportion of 1:1000 with ethanol/water 48% (v/v), and measured by fluorescence in terms of a standard calibration curve of BaP from 1.0 to 20.0 mg/L similar to Ariese et al. (1993). Fluorescence was measured with a Cary Eclipse (Varian) spectrofluorimeter, which consisted of a xenon discharge lamp, two Monk-Gillieson monochromators, a Hamamatsu photomultiplier and quartz cells (1 1 cm). Slits for the excitation and emission monochromators were set at 5 nm, the photomultiplier voltage was adjusted to 600 mV, and the monochromator scan rates were 600 nm/min. The synchronous fluorescence peak of the BaP was located at 323 nm when Dl =25 nm. 2.4. Statistical Analysis For the statistical analysis, Cochrans test of equality error variances was carried out. As p 40.05 was obtained for all data by the Cochrans test, data were
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statistically compared using the Kruskal–Wallis test for non-parametric data followed by a Tukey post-hoc analysis, with the aid of Statistica 7.0 software. Only p o0.05 was accepted as significant for indicating statistical differences between groups with different treatments under the same exposure time. Data are presented as mean + standard deviation.
3. Results Table 1 shows the activities of CbE, GST, GPx, CAT and EROD in the liver of O. niloticus exposed to BaP, DZ and BaP+ DZ for exposure periods of 2 and 7 days. No differences were observed in the activities of CbE, CAT and GPx throughout the experiment for any chemicals tested. However, pre-exposure to BaP for three days caused an increase in GST activity, which did not continue after 2 days, but significantly increased after 7 days of exposure. No such increase was observed in fish exposed for 2 and 7 days to DZ or BaP+ DZ. Moreover, BaP caused a strong increase in EROD activity after the three-day pre-exposure period, and after both 2 and 7 days. When compared to the control group, increased EROD activity was also observed for animals exposed to the mixture of BaP+ DZ for periods of both 2 and 7 days after the pre-exposure period, but this increase was lower than that observed for BaP alone (not significant). Exposure to DZ alone caused no difference in EROD activity when compared to the respective controls. Enzymatic activities measured in the gills are presented in Table 2. In this tissue, CAT activity was the only parameter insensitive to all treatments. DZ was able to decrease CbE activity when alone and when in combination with BaP, but only after 2 days of exposure. No difference in CbE was observed after 7 days of exposure for fish exposed to DZ or BaP+ DZ, nor in fish exposed to BaP alone after 2 and 7 exposure days. GST activity increased in fish exposed to BaP after all exposure times when compared to control values. When in combination with DZ, BaP caused a similar increase in GST activity after 2 days of exposure, but caused no difference (compared to control values) after 7 days. No treatment caused changes in GPx activity compared to the activity in control animals, but those exposed to BaP for 7 days after the
pre-treatment exposure had increased GPx activity compared to those exposed to DZ. The activity of AChE in the brain is presented in Table 3. No differences were observed for all treatments throughout the experiment. In Table 4, BaP levels in the bile of tilapias exposed to the contaminants are shown. BaP levels in bile were significantly increased in the bile of fish exposed to BaP and BaP+ DZ during all exposure periods. Moreover, it seems that the fish exposed to BaP+ DZ presented higher levels of BaP in the bile when compared to animals exposed to BaP alone. This was statistically significant for 7 exposure days. Also, levels of BaP in the bile seemed to decrease during the experiment in fish exposed to BaP and BaP+ DZ. DZ exposure did not cause any variation in BaP levels in the bile when compared to control values.
4. Discussion Pesticides and PAHs are common pollutants in the aquatic environment. In some field studies where oxidative stress parameters, EROD induction, and esterase inhibition can be used to monitor water pollution, the joint presence of pesticides and PAHs may lead to different responses in fish compared to the responses typically observed when the animals are exposed to a single contaminant. Therefore, it is vital to establish the responses that are typically activated in fish species exposed to mixtures of different classes of pollutants under controlled conditions. With this information, scientists can better understand the data that has originated from monitoring studies of aquatic environments contaminated by multi-xenobiotics. In this study, we focused on the effects of BaP plus DZ in O. niloticus that were exposed for 2 and 7 days at sub-lethal concentrations. In general, BaP and DZ alone caused typical responses in tilapias: BaP caused a strong induction of EROD activity in the liver and increased GST activity in the liver and gills, with a stronger effect on the gills. DZ exposure caused the inhibition of CbE but had no effects on AChE activity, since a low dose was
Table 1 Enzymatic activities measured in the liver of O. niloticus exposed to BaP and DZ. Enzymea
Group
CbE
Control BaP DZ BaP +DZ
GST
Control BaP DZ BaP +DZ
Pre-exposure
2 days after pre-exposure
7 days after pre-exposure
1.39 7 0.85 1.75 7 0.66
1.05 7 0.50 1.61 70.85 1.36 70.55 0.85 7 0.06
1.23 70.27 1.21 7 0.20 0.83 7 0.07 0.85 70.37
1.50 7 0.39 2.22 70.48b
1.81 7 70.85 2.30 70.68 2.12 70.54 2.16 70.97
1.13 70.32 1.42 70.31d 1.14 7 0.30 0.87 7 0.06
– –
– –
Control BaP DZ BaP +DZ
0.027 0.01 0.0267 0.01 – –
0.018 7 0.01 0.027 0.01 0.022 7 0.01 0.019 7 0.01
0.037 0.01 0.047 0.01 0.047 0.01 0.037 0.00
CAT
Control BaP DZ BaP +DZ
99.07 7 20.91 89.54 732.15 – –
67.17 720.43 69.80 710.65 75.84 7 15.92 64.24 78.67
72.12 710.70 69.33 7 16.19 76.16 7 14.11 73.44 7 13.65
EROD
Control BaP DZ BaP +DZ
0.86 71.11 21.79 75.03b – –
0.91 7 0.40 44.12 7 13.13b,c 1.01 7 0.90 33.45 721.60b,e
2.64 72.54 23.04 7 17.11b,c 2.26 73.26 10.26 78.76
GPx
a
Activity expressed in mmol/min/mg protein (CbE, GST, GPx and CAT) or pmol/min/mg protein (EROD). Significant difference compared to the respective control. c Significant difference between DZ and BaP at the same exposure time. d Significant difference between BaP and BaP +DZ at the same exposure time. e Significant difference between DZ and BaP +DZ at the same exposure time. Data expressed as mean 7 standard deviation. b
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Table 2 Enzymatic activities measured in the gill of O. niloticus exposed to BaP and DZ. Enzymea
Group
Pre-exposure
2 days after pre-exposure
CbE
Control BaP DZ BaP +DZ
0.087 0.03 0.077 0.03 – –
0.13 70.04 0.117 0.02c,d 0.057 0.01b 0.067 0.01b
0.107 0.05 0.107 0.05 0.077 0.02 0.077 0.00
GST
Control BaP DZ BaP +DZ
0.127 0.02 0.287 0.04b – –
0.12 70.03 0.207 0.03b,c 0.10 70.03e 0.187 0.04b
0.117 0.01 0.16 7 0.05b,c 0.087 0.04 0.127 0.02
GPx
Control BaP DZ BaP +DZ
0.03 70.00 0.037 0.01 – –
0.05 70.01 0.05 70.02 0.05 70.01 0.04 70.01
0.077 0.01 0.107 0.01c 0.077 0.03 0.077 0.01
CAT
Control BaP DZ BaP +DZ
6.317 0.80 6.69 7 1.13 – –
8.04 7 2.59 10.28 7 2.33 9.30 7 3.31 9.76 7 1.57
5.17 71.77 9.68 7 2.90 7.34 73.27 6.90 7 0.79
7 days after pre-exposure
a
Activity expressed in mmol/min/mg protein (CbE, GST, GPx and CAT) or pmol/min/mg protein (EROD). Significant difference compared to the respective control. Significant difference between DZ and BaP at the same exposure time. d Significant difference between BaP and BaP +DZ at the same exposure time. e Significant difference between DZ and BaP + DZ at the same exposure time. Data expressed as mean 7 standard deviation. b c
Table 3 AChE activity in the brain of O. niloticus after xenobiotics exposure. Enzyme
Group
Pre-exposure
2 days after pre-exposure
7 days after pre-exposure
AChEa
Control BaP DZ BaP+ DZ
0.05 7 0.02 0.06 7 0.01 – –
0.0507 0.01 0.053 7 0.01 0.067 0.02 0.067 0.01
0.04 70.01 0.06 70.02 0.06 70.02 0.08 70.01
a
Activity expressed in mmol/min/mg protein. Data expressed as mean7 standard deviation.
Table 4 Levels of BaP in bile of Oreochromis niloticus after xenobiotics exposure.
a
BaP
Group
Pre-exposure
2 days after pre-exposure
7 days after pre-exposure
Control BaP DZ BaP +DZ
0.077 0.02 163.287 78.56b,c – –
0.107 0.02 122.227 48.44b,c 0.10 70.02e 164.687 125.32b
0.057 0.02 78.54 7 28.80b,c,d 0.07 70.02e 110.29 7 0.26b
a
Data expressed in mg/mL of bile. Significant difference compared to the respective control. c Significant difference between DZ and BaP at the same exposure time. d Significant difference between BaP and BaP + DZ at the same exposure time. e Significant difference between DZ and BaP + DZ at the same exposure time. Data expressed as mean 7 standard deviation. b
used. It is possible that AChE activity could be inhibited in other tissues, but this enzyme was evaluated only in brain, and further studies are needed to verify if AChE is inhibited in other tissues by DZ dose used in the present work. As for oxidative stress parameters (GPx and CAT), no effects were observed after BaP or after DZ exposure when compared to control values, even though the tilapias exposed to BaP had higher activity than fish exposed to BaP in combination with DZ, suggesting a possible inhibition effect of DZ on GPx activation by BaP. DZ alone did not cause any effects on EROD activity (compared to control values), which suggests that EROD is not activated by DZ and may not be related to desulfuration or dearylation of this OP. Paolini et al. (1997) have demonstrated that the main P450 isoforms induced by the OP acephate in mice livers were 3A, 1A2
and 2E1, and that this OP suppresses CYPs 2B1 and 1A1. Tang et al. (2002) showed that CYP2B6 has the main in vitro desulfuration activity on the OP chlorpyrifos among all human CYP isoforms tested. They also showed that chlorpyrifos is a suicide substrate for this isoform, which decreased CYP activity during incubation periods. Compared to control values, EROD activity in our study was similar in fishes exposed to DZ, indicating that CYP1A suppression did not occur. However, EROD induction was lower in those fish exposed to BaP and DZ than in those exposed to BaP, though the differences were not statistically significant. It seems that DZ interferes with EROD induction that is promoted by BaP, and this result can have profound implications for monitoring studies that use biomarker analysis. The lack of EROD induction in fish from
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monitoring studies could be misinterpreted as the absence of PAHs in the environment. This study shows, however, that this result could be due to a concomitant presence of OPs and PAHs, with an inhibition effect of these compounds on EROD induction. Flammarion et al. (1998) have presented very similar results with the cyprinid fish gudgeon (Gobio gobio) pre-exposed (intraperitoneal injection) to b-naphtoflavone (BNF, another CYP1A inducer) and the OP methidathion. These results corroborate our study. In that study, methidathion exposure led to a 60–70% decrease of the EROD induction caused by pre-treatment with BNF. A similar profile was also observed for GST activity. BaP was able to cause significant GST induction after the pre-exposure period. After exposure periods of both 2 and 7 days, BaP caused significant induction in the gills, and after the pre-exposure period and 7 days after pre-exposure, BaP caused significant induction in the liver. However, the fish exposed to BaP +DZ presented lower GST induction after 7 days, clearly demonstrating that DZ also interferes with GST induction promoted by BaP. Similarly, Hodge et al. (2000) have shown that DZ and chlorpyrifos are able to cause GST inhibition in an insect (Micromus tasmaniae). Again, the lack of GST induction in fish from field studies could be due to the concomitant presence of PAHs and OPs in the environment, and is not necessarily due to the absence of PAH exposure. EROD induced by BaP had no effects on esterase activities, since no differences were observed in AChE activity throughout the experiments, and considering that the inhibition profile of CbE in gills after two exposure days was identical for fish exposed to DZ or BaP+ DZ. Thus, it can be also supposed that P450 isoforms activated by BaP are not related to desulfuration or dearylation of OPs. Straus et al. (2000) studied the activities of desulfuration and dearylation of the phosphorothionates chlorpyrifos and parathion in channel catfishes exposed to Aroclor 1254, another EROD inducer. Although chlorpyrifos or parathion seems to induce desulfuration and dearylation activities, these enzymes were not altered after Aroclor 1254 exposure, indicating that this compound does not induce the P450 isoforms responsible for desulfuration or dearylation of phosphorothionates, and also that the isoforms responsible for these activities are probably not CYP1A. Although it was not tested for B[a]P exposure, we can assume that B[a]P is also not able to increase the metabolism of the DZ even when EROD activity is increased. Finally, it was observed that the combination of PAH and OP exposure also interferes with levels of BaP excretion in the bile. DZ alone did not cause any effects on BaP excretion, since the levels of this compound in the bile did not change after the animals were exposed to DZ. However, BaP exposure caused a very significant increase in BaP levels in bile. BaP levels were higher after the three days of pre-exposure, and then decreased during the exposure time. Because the fish were exposed to a single dose of BaP, it should be expected that BaP concentration in water diminishes during the exposure period, since it was being metabolized in the fish. A similar trend was observed for fish exposed to BaP + DZ, but in this case, the decrease in BaP concentration in bile was less intense. In our study, it could be supposed that the interference of DZ with EROD and GST induction caused by BaP resulted in a lower BaP metabolization rate. It has been shown that the main product of BaP metabolization through CYP1A and GST involved the formation of hydroxyl, epoxy, diol or quinone BaP derivatives (Miller and Ramos, 2001), and some of these can be conjugated with glutathione by GST. Therefore, there are different BaP metabolites that can be excreted to the bile after phase I and II metabolism. As DZ interferes with phase I and II enzymes, the rate of unmodified BaP excretion could then increase as an alternative pathway to compensate for reduced biotransformation enzymes.
This mechanism could account for the increased levels of BaP in the bile of the fish exposed to BaP+ DZ compared to those exposed to BaP alone after 2 and 7 exposure days. Mdegela et al. (2006) observed an inhibitory effect of the hormone 17b-estradiol on EROD activity induced by BaP in gills of African sharptooth catfish (Clarias gariepinus). They also observed higher levels of BaP in the bile of fish exposed to the hormone in combination with BaP, compared to those fishes exposed to BaP alone, and suggested that the amount of BaP reaching the liver was increased by the hormone through the inhibition of gill filament EROD activities. Relatively lower hepatic EROD activity and accumulation of bile metabolites in fish exposed to BaP alone indicates that much of the BaP was biotransformed in the gills and did not reach the liver. EROD activity was not measured in the gills in the present study, but such suggestions could account for the larger amount of BaP in the bile of fish that were exposed to BaP+ DZ.
5. Conclusions Complex mixtures of OPs and PAHs can account for response mechanisms in tilapias that are different from those observed with a single contaminant alone. Tilapias exposed to BaP alone presented typical responses with respect to phase I and II enzyme, which agreed with numerous other studies found in the literature. DZ alone inhibited CbE in gills but not in liver, and AChE activity in brain was unchanged. In combination, BaP +DZ had different effects on biochemical parameters. In our study, the main finding was that DZ can decrease the main biochemical parameters induced by BaP in fish, along with EROD and GST activities, and can also contribute to increased levels of BaP in the bile. These results have significant implications for monitoring studies of aquatic ecosystems contaminated by both classes of compounds.
Acknowledgements This study had the financial support of the ‘‘Fundac- a~ o de Amparo a Pesquisa do Estado de Sa~ o Paulo’’ (FAPESP, 2006/03873-1). Lilian Nogueira and Camila Pereira Trı´dico are recipients of FAPESP fellowships. The present study was conducted in accordance with national and institutional guidelines for the protection of animal welfare. References Almeida, E.A., Bainy, A.C.D., Dafre, A.L., Gomes, O.F., Medeiros, M.H.G., Di Mascio, P., 2005. Oxidative stress in digestive gland and gill of the brown mussel (Perna perna) exposed to air and re-submersed. J. Exp. Mar. Biol. Ecol. 318, 21–30. Almeida, E.A., Bainy, A.C.D., Loureiro, A.P.M., Martinez, G.R., Miyamoto, S., Onuki, J., Barbosa, L.F., Garcia, C.C.M., Prado, F.M., Ronsein, G.E., 2007. Oxidative stress in Perna perna and other bivalves as indicators of environmental stress in the Brazilian marine environment: antioxidants, lipid peroxidation and DNA damage. Comp. Biochem. Physiol. 145A, 588–600. Ariese, F., Kok, S.J., Verkaik, M., Gooijer, C., Velthorst, N.H., Hofstraat, J.W., 1993. Synchronous fluorescence spectrometry of fish bile: a rapid screening method for the biomonitoring of PAH exposure. Aquat. Toxicol. 26, 273–286. Bagchi, D., Bagchi, M., Hassoun, E.A., Stohs, S.J., 1995. In vitro and in vivo generation of reactive oxygen species, DNA damage and lactate dehydrogenase leakage by selected pesticides. Toxicology 104 (1-3), 129–140. Beutler, E., 1975. Red Cell Metabolism: A Manual of Biochemical Methods.. Grune & Stratton, New York. Binder, R.L., Melancon, M.J., Lech, J.J., 1984. Factors influencing the persistence and metabolism of chemicals in fish. Drug Metab. Rev. 15 (4), 697–724. Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 72, 248–254. Burke, M.D., Mayer, R.T., 1974. Ethoxyresorufin—direct fluorimetric assay of a microsomal o-dealkylation which is preferentially inducible by 3-methylcholanthrene. Drug Metab. Disp. 2, 583–588.
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