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South American Cyprinodontiform fishes are potential candidates to be used as model species in environmental toxicology. We sought for molecular and biochemical biomarkers of pollution in Poecilia vivipara (Poecilidae) and Jenynsia multidentata (Anablepidae). Partial nucleotide sequences for the cytochrome P450 1A (cyp1A), a classical biomarker of exposure to organic contaminants in fish, were identified in P. vivipara and J. multidentata (~ 650 nucleotides) using degenerated primers and PCR. These sequences shared ~ 90 % identity in the predicted amino acid sequence with the corresponding Cyp1A region of Fundulus heteroclitus. RT-qPCR analysis confirmed that cyp1A transcription was strongly induced in the liver and gills of J. multidentata (~185-fold and ~20-fold, respectively) and P. vivipara (122-fold and 739-fold, respectively), after 24-hrs exposure to 1 μM of the synthetic cyp1A inducer β-naphthoflavone (BNF). After 24 hs of injection with 1 μg.g-1 of the environmental carcinogenic contaminant benzo[a]pyrene (BaP), a decreased total antioxidant capacity against peroxyl radicals was observed both in liver of J. multidentata and gills of P. vivipara. BaP injection in both fishes did not cause changes in lipid peroxides (TBARS) levels, suggesting an absence of an oxidative stress situation caused by BaP injection in this study. The newly identified cyp1As would serve as general biomarkers of exposure to organic contaminant in future studies using P. vivipara and J. multidentata. The results also points out to the important species-specific differences in the biomarker responses in those South American cyprinodontiform fishes, which would suggests distinct resistance/susceptibility to polycyclic aromatic hydrocarbons.
Following the example of the Atlantic killifish Fundulus heteroclitus (Cyprinodontiform, Fundilidae), extensively studied in the North America (Burnett et al., 2007), Cyprinodontiform fishes that inhabit the South American Coast, e.g. Jenynsia multidentata (Anablepidae) (from Argentina to Brazil) and Poecilia vivipara (Poecilidade) (from Argentina to Venezuela) (Froese and Pauly, 2001), have been suggested as “novel” model species in environmental toxicology (Amado et al., 2009; Ame et al., 2009; Cazenave et al., 2008; Mattos et al., 2009). The uses of those model cyprinodontiform fishes in toxicology, relies in the study of the fate of contaminants in the aquatic environment (Burnet et al., 2007, Matson et al., 2008) and also in the understanding of the genetic/evolutionary aspects involved in the adaptation to survive in extreme polluted conditions (Williams and Oleksiak, 2008, Wirgin et al., 2011) that helps to elucidate the fundamental biochemical and molecular mechanisms in toxicology (Hahn et al., 2004).
The understanding of the fate and the toxicity mechanisms caused by organic contaminants requires the study of cytochrome P450s and the biochemical aspects of the oxidative stress. Cytochrome P450 (Cyp) enzymes catalyze oxidative metabolism of thousands of drugs, environmental pollutants, and endogenous compounds. Environmental pollutants includes many substrates of mammalian and fish Cyp1s (e.g., Schober et al., 2006) such as the halogenated aromatic hydrocarbons (HAHs), polycyclic aromatic hydrocarbons (PAH), herbicides, and pesticides (Nebert and Russell, 2002). While metabolism often results in detoxification, the action of Cyp1 enzymes also can generate toxic metabolites, and induced production of reactive oxygen species (ROS), contributing to increased risks of cancer, birth defects, and other toxic effects (Nebert and Karp, 2008).
Mammalian Cyp1A1 and fish Cyp1As could be strongly induced in the mRNA, protein and catalytically activity levels, by environmental contaminants such PAHs, planar polychlorinated biphenyl (PCB), dibenzo-p-dioxin (PCDD), and dibenzofuran (PCDF) congeners, and some natural products, via activation of the aryl hydrocarbon receptor (AHR) (Hahn, 2002). These features have led to the widespread use of Cyp1A as a marker of environmental exposure to AHR agonists in humans and wildlife (e.g., Stegeman, 1986; Fujita et al., 2001; Lambert et al., 2006). Although some substances (e.g. the synthetic β-naphthoflavone, BNF) are potent AHR agonists and Cyp1A inducers without causing major toxicity, other compounds, such as benzo[a]pyrene, BaP and PCB126) are potent AHR agonists, Cyp1A inducers and highly toxic.
It has been reported that both reactive oxygen species (ROS) derived from BaP biotransformation by Cyps (through redox cycling of hydro- and semi-quinone intermediates) and its metabolites as BaP-semiquinone radicals are responsible by oxidative damage to lipids and proteins or to modulate the antioxidant defense system (Kim et al., 2000; Lin et al., 2007). In fact, lipid peroxides are oxidative damage products frequently found after BaP exposure in many animal models (Pan et al., 2006; Alsop et al., 2007).
Our objectives in this study were: (1) to identify Cyp1A sequences for both J. multidentata and P. vivipara; (2) to evaluate the cyp1A transcriptional induction after exposure to BNF, a model Cyp1A inducer (via AHR activation), in the aim to give support for the use of this biomarker of organic contaminant in environmental studies; and (3) to evaluate the oxidative related effects (in terms of total antioxidant capacity and lipid peroxidation) in those fishes injected with BaP, an environmental contaminant that is an Cyp1A inducer and toxic carcinogenic compound. This study will helps to understand, whether or not, inter-specific differences exist between the South American Cyprinodontiform fishes P. vivipara and J. multidentata, and if there is an acquired resistance or increased susceptibility to organic contaminant exposure, as it occurs and have been extensively reported in the N. American F. heteroclitus fish populations.
Male P. vivipara and J. multidentata were caught at Cassino Beach (Rio Grande, Brazil) in March 2011, using minnow traps. Fish were acclimated at 20 °C in 100 L aquarium at UV treated water, adjusted to salinity 15 ppt with dechlorinated water, for one month and were fed twice a day with Alcon BASIC® MEP 200 Complex during the acclimation period. The procedures used in these experiments were approved by the Animal Care and Use Committee (CEUA) at the Universidade Federal do Rio Grande (FURG).
Liver was dissected from one untreated male J. multidentata and one male P. vivipara, randomly selected. Total RNA was isolated using Qiazol reagent (Qiagen). The RNA quality was determined in a standard agarose gel electrophoresis, running 5 μ L of sample plus 2 μ L of loading buffer. The RNA quantity and quality was also determined spectrophotometrically (Nanodrop ND 1000; NanoDrop Technologies, Wilmington, DE). cDNA was synthesized using the High-Capacity cDNA Reverse Transcription Kit (Applied Biosystems), a mix of anchored oligo(dT) primer (MWG Biotech, Inc.) with random hexamer primer (provided in the reverse transcriptase kit) and RNaseOUT Recombinant Ribonuclease Inhibitor (Invitrogen).
Degenerate primers were designed using highly conserved regions of previously known cyp1A sequences from F. heteroclitus (Cyprinodontiform), Japanese medaka Oryzias latipes (Beloniform) and Stickleback Gasterosteus aculeatus (Gasterosteiformes) (Genbank Accession numbers AF026800, NM_001105087 and HQ202281, respectively) and avoiding conserved regions present in other cyp1 subfamilies (e.g. cyp1B1, cypP1C1, cyp1C2 and cyp1D1). The list of primers used is presented in the table 1. PCR reactions were carried out for Cyp1A and β-actin in liver cDNA of P. vivipara and J. multidentata. PCR products were resolved on a 1% agarose gel and then isolated and purified using Illustra GFX PCR DNA and Gel Band PurificationKit (GE Healthcare, Buckinghamshire, UK). Nucleotide sequences were obtained using an ABI Prism 3130×l platform (Applied Biosystems), translated to the predicted amino acid sequences and aligned with other Cyp1 family members using ClustalW (Thompson et al., 1994). Calculation of identities among P. vivipara, J. multidentata and F. heteroclitus nucleotide and amino acid predicted sequences were performed using the Sequence Identity Matrix tool, Bioedit Sequence Alignment Editor Software (Hall, 1999).
Twenty male P. vivipara, with 3-5 cm length and 0.5-2.5 g whole body weight were randomly distributed in two 10 L glass aquaria (10 fish per aquaria). One of those aquaria was further used as control group, and another was used as the BNF exposed group. Fish stayed without food in those aquaria with aerated water, temperature 20 ° C and salinity 15, for 24-h before starting the exposure experiment. After this period, 200 μ L of a 0.05 M BNF stock solution (dissolved in DMSO) was introduced in one of the aquaria to make a final concentration of 1 μM BNF. BNF was purchased from Sigma–Aldrich Co.(MO, USA). An equivalent volume of DMSO was also introduced in the control group aquaria. The BNF concentration used here (1 μM) has been shown to be high enough to induce ethoxyresorufin O-deethylase (EROD) activity in gills of various fish species (Brunstrom et al., 2002, Jonsson et al., 2003). The same experimental procedures described above using P. vivipara, was also done with J. multidentata fish (3-5 cm length, 0.5-2.5 g whole body weight, n=20) in order to setup a control group (n=10 fish) and a BNF exposed group (n=10 fish).
Twenty four hours after the exposure experiment all P. vivipara and J. multidentata fishes (n=40 in total) were killed by cervical transection and liver and gill of individual fish were dissected and immediately placed in RNA later (Ambion). The samples were held 24-h at 4 °C, and then stored at -20 °C according to the RNAlater manufacturer's instructions.
Total RNA was extracted from gills and liver and cDNA was synthesized (n=5 per experimental group), using the methods described in the item 2.2 Gene-specific primers for the P. vivipara and J. multidentata cyp1A were designed with Primer3 (Rozen and Skaletsky, 2000). Specific β-actin primers for J. multidentata and P. vivipara were designed based on GenBank (Acession EF362747) and a recently identified sequence (unpublished data), respectively. Primers were obtained from IDT Integrated DNA Technologies (Primer sequences are shown in Table 1). Real-time PCR (qPCR) was performed in duplicate using GoTaq qPCR Master Mix (Promega, Madison, WI) according to manufacturer's instructions and a 7500 Real-Time PCR System (Applied Biosystems) using the program: 50 °C for 2 min, 95 °C for 2 min and 40 cycles 95 °C for 15s and 60 °C for 30s. Melting curve analysis was performed to all the PCR products at the end of each qPCR run to ensure that a single product was amplified. The EΔct method using β-actin as house-keeping gene was used to evaluate the cyp1A relative expression levels and the fold-induction in response to BNF treatment comparing to the respective control group. Gill's and liver's cDNAs from five fish per experimental group were used in the qPCR analysis.
Ten previously acclimated male J. multidentata and ten male P. vivipara (3-5 cm length, 0.5-2.5 g whole body weight), were individually weighed and intraperitoneally injected with benzo[a]pyrene (BaP) (previously dissolved in DMSO) to make the dose of 1 μg BaP per g of fish (1 mg.kg-1). Previous studies showed that an equivalent injected dose of BaP is high enough to induce ethoxyresorufin O-deethylase (EROD) activity, to alter oxidative stress parameters (Regoli et al., 2009) and to induce cyp1A in the transcriptional level (Bugiak et al., 2009) in fish liver. Ten J. multidentata and ten P. vivipara fishes were also injected with an equivalent volume of DMSO alone (control groups). Twenty four h after the BaP or DMSO injections in the both fish species (four experimental groups in total) the gills and liver were excised from all the forty fishes used in the experiment (n=10 fish per experimental group) and immediately stored at – 80 °C for biochemical analysis.
Total antioxidant competence against peroxyl radicals was determined through reactive oxygen species (ROS) determination in sample tissues treated or not with a peroxyl radical generator. Peroxyl radicals were produced by thermal (35 °C) decomposition of 2, 2′-azobis 2 methylpropionamidine dihydrochloride (ABAP; 4 mM; Aldrich) (Winston et al., 1998). ROS concentration was measured along 30 min with the fluorogenic compound 2′,7′-dichlorofluorescin diacetate (H2DCF-DA) at a final concentration of 40 mM, according to the methodology of Amado et al. (2009).
Results were expressed as relative area difference, using the following expression: Relative area = (FU 30 minABAP – FU 30 minwithout ABAP)/FU 30 minwithout ABAP
According to this expression, low relative area difference means high antioxidant capacity against peroxyl radicals.
Oxidative damage was measured through lipid peroxidation using the TBARS method (Oakes and Van Der Kraak, 2003). This methodology involves the reaction of malondialdehyde (MDA), a degradation product of lipid peroxidation, with thiobarbituric acid (TBA) under conditions of high temperature and acidity, producing a chromogen which was quantified by fluorometry (excitation: 520 nm, emission: 580 nm). Briefly, after homogenization, aliquots were incubated at 95 °C during 30 min with 35 μM of butylated hydroxytoluene (BHT), 8.1% sodium dodecyl sulfate (SDS), 20% acetic acid and 0.8% TBA. After cooling to room temperature it was added n-butanol and centrifuged at 3,000 × g for 10 minutes at 15°C. Tetramethoxypropane (ACROS Organics) was used as standard. Fluorescence was read at room temperature using a plate reader fluorimeter (Victor 2, Perkin Elmer).
All results in the bar graphs were expressed as mean ± standard deviation. For the 1 μM BNF exposure experiment, homogeneity of variance for the cyp1A data was tested through Bartlett's test. Data were logarithmically transformed if the test rejected the assumption of variances homogeneity. Differences between the transcript levels of BNF and control (DMSO), for each fish species and organ, were analyzed using Student's t-Test. For the BaP injection experiment, significant differences in the total antioxidant competence and TBARS were assessed by two-factor analysis of variance (ANOVA), being the factors BaP exposure (Control and 1 mg.kg-1 fish BaP injected) and fish species (P. vivipara and J. multidentata). Post hoc comparisons were performed using the Newman-Keuls test or orthogonal contrasts. In all cases, type I error probability (a) was fixed in 5%. ANOVA assumptions (normality and variance homogeneity) were previously checked (Zar, 1984).
Although poorly studied so far, J. multidentata and P. vivipara, as well as other cyprinodontiform fishes from S. America, possess one or more of the following features that support it use as model species in toxicology: (1) cosmopolitan (see figure 1) and adaptation to live in a wide range of stressing conditions (e.g., pollution, salinity, oxygen and temperature stresses); (2) different from the model species killifish F. heteroclitus, most are ovoviviparous, being interesting alternative models to understand mother embryo toxicological interaction during the embryonic development; (3) a huge diversity of species exists, and possibly a diversity of mechanisms of resistance/sensitivity to chemical stress; (4) some species are rare, thus possibly endangered. In the present study we added important information regarding biomarkers of organic contaminants using J. multidentata and P. vivipara as model species in environmental toxicology.
Using degenerate primers based on teleost Cyp1A, we were able to amplify partial Cyp1A sequences (about 650 bp) from J. multidentata and P. vivipara. Aligning the amino acid predicted sequences from these PCR products with F. heteroclitus Cyp1A, Cyp1B1, Cyp1C1, Cyp1C2 and Cyp1D1 (obtained from GenBank, accession numbers AF026800, FJ786959, DQ133570, FJ786960 and FJ786961, respectively) resulted in the preliminary classification of the J. multidentata and P. vivipara sequences as cyp1As (Figure 2 and Table 2). The partial sequences obtained presents four of the six known Cyp1A substrate recognition sites (SRS2, SRS3, SRS4 and SRS5), and are located between amino acids positions 210 and 420 based on the full-length F. heteroclitus sequence (Figure 2). Comparisons of F. heteroclitus Cyp1A with the other Cyp1A sequences showed a marked identity between the cyprinodontiform Cyp1As in SRSs 2, 4 and 5, and some differences in the SRS3, as it could be expected based on previous studies that compared Cyp1A alignments among fishes before (Goldstone et al., 2009). The partial J. multidentata and P. vivipara Cyp1A sequences display 90 % and 90 % pair-wise identity to Cyp1A of F. heteroclitus, respectively (Table 2), and much lower identity to other F. heteroclitus Cyp1s (Cyp1B1, Cyp1C1, Cyp1C2 and Cyp1D1) that ranged from 38 % to 49 % identity (Table 2). Notably, the Cyp classification based on the amino acid percent identities suggested by Nelson et al. (1996), accepting identities higher than 55 % for the same subfamily, is in agreement with our preliminary annotation of these partial sequences. This preliminary analysis of SRSs and the high identities among the referred sequences (~90-95 %) could suggest that substrate overlapping could exist for Cyp1A in those S. American cyprinodontiform and the N. American F. heteroclitus.
No mortality was observed during the J. multidentata and P. vivipara exposure to BNF or the carrier DMSO (control). The 24-hs exposure to 1 μM of BNF strongly induced the expression of cyp1A in gill and liver of J. multidentata and P. vivipara comparing to the respective control groups (p<0.0001 in all cases) (Figure 3). BNF is a commonly studied AHR agonist and a potent Cyp1A inducer (Jonsson et al., 2007). The present results suggest that the newly identified cyp1As, have potential to be additional sensitive biomarkers of exposure to AHR agonist contaminants in J. multidentata and P. vivipara likely it is well-known in other fishes.
The induction of Cyp enzymes in fish liver was first suggested as an indicator of aquatic contamination in the 1970s (e.g. Payne, 1976). Since then, many studies have shown that Cyp1As in vertebrate liver (often measured by activity assay, e.g. EROD, and protein detection by Western blot) are strongly induced by some organic contaminants that represent risk for human and wildlife (e.g., PAHs, coplanar PCBs, polychlorinated dibenzofurans, and dibenzodioxins) (Bucheli and Fent, 1995). The analysis of Cyp1A transcriptional levels by qPCR has been compared with traditional methods (e.g. EROD) and also has showed to be a sensible and robust biomarker for organic contaminants (Piña et al., 2007). Based on the presented results, we could suggest that the transcript quantification of the newly identified cyp1As by RT-qPCR, could be employed as potential biomarkers for organic contaminant exposure using J. multidentata and P. vivipara S. American cyprinodontiform fishes.
Although strongly induced by BNF, the quantified levels of cyp1A induction in exposed groups vary when different organs/species were considered. The most substantial changes in cyp1A expression in response to BNF were in P. vivipara gills (~739-fold induction) followed by J. multidentata liver (~185-fold), P. vivipara liver (~122-fold) and J. multidentata gill (~20-fold). This result suggests that different patterns for Cyp1A induction could occur when comparing those organs and cyprinodontiform species, after exposure to AHR-agonists, and possibly will reflect differences in the detoxification/bioactivational processes that are linked to Cyp1A activity. It is also possible that the observed differences rely in the characteristic interval of time that each organ/species show the “peak” of induction. It is well documented that in a time-exposure window, few hours can influence strongly the levels of induction at the gene transcription levels for Cyp1A in putterfish organs, after 6-, 12-, 24-, 48- and 96-hs exposure to 1 μM BNF (Kim et al., 2008).
As showed in Figure 4a, total antioxidant capacity against peroxyl radical was decreased (higher relative area difference) in liver of J. multidentata liver and gills of P. vivipara after BaP exposure (p<0.05; Figure 4b). The reduction in antioxidant competence, however, did not result in oxidative stress, since no changes on lipid peroxidation (TBAR) were observed after BaP treatment in tissues of both species (p>0.05; Figure 5). Several evidences indicate that BaP alters the cellular redox state, triggering apoptosis (Soulhaug et al., 2005). Mussels Mytilus galloprovincialis exposed to BaP showed a peak of activity in the antioxidant enzyme catalase after 48 h exposure (Banni et al., 2010). In present study the selected BaP dose and exposure period induced an intermediary response considering the previous evidences that showed an induction of the antioxidant defense (Lin and Yang, 2007; Banni et al., 2010) or the generation of oxidative stress (Costa et al., 2010). Our results indicated that BaP exposure caused loss of antioxidants, with specific differences, since liver was the responsive organ in J. multidentata and gills for P. vivipara 24-hs after injection. However the antioxidants consumption in liver and gills of the two species seemed to cope the pro-oxidant effect of BaP, since no evidences of oxidative damage was observed.
The high identity of Cyp1A sequences observed among P.vivipara, J. multidentata and F. heteroclitus associated with the Cyp1A transcriptional responses observed in fish exposed to BNF and treated with BaP support that proposition to use these South American Cyprinodontiform fishes as model species in environmental toxicology. Interestingly, depending on the species and tissue analyzed, we observed different responses in the oxidative stress associated parameters which suggest distinct resistance/susceptibility to PAHs. Population studies can be subsequently carried out in order to identify possibly polymorphisms in detoxification associated genes that enable the fish to adapt to highly contaminated sites, similarly to what has been described in F. heteroclitus.
This study was supported in part by Brazilian INCT-TA (CNPq 573949/2008-5) and NIH grants to JJS (the Superfund Basic Research Program 5-P42-ES007381 and R01-ES015912). JZ was a Guest Student at the Woods Hole Oceanographic Institution and was supported by a CAPES Ph.D. Fellowship and CNPq Ph.D. Sandwich Fellowship, Brazil. ACDB and JMM are recipients of a CNPQ Productivity Fellowship, Brazil. RSF is recipient of the CNPQ-PIBIC fellowship, Brazil. Study sponsors had no involvement in the studies reported here or in the decision to submit this paper for publication. Thank to J.J. Mattos, M.S. Sopezki, F.S. Guimarães, J.L.R. Scaini, S.I.M. Abril, C. Rodriguez and E.P. Colares, for their assistance in the experimental procedures.