Arch Environ Contam Toxicol (2009) 57:561–570 DOI 10.1007/s00244-009-9289-8
Phase I Enzyme Induction in Girardinichthys viviparus, an Endangered Goodeid Fish, Exposed to Water from Native Localities Enriched with Polychlorinated Biphenyls Armando Vega-Lo´pez Æ F. Alejandro Jime´nez-Orozco Æ Luı´s A. Jime´nez-Zamudio Æ Ethel Garcı´a-Latorre Æ M. Lilia Domı´nguez-Lo´pez
Received: 9 October 2008 / Accepted: 20 January 2009 / Published online: 6 February 2009 Ó Springer Science+Business Media, LLC 2009
Abstract The present study examines the induction of mixed-function oxidase (MFO) enzymes, including CYP content CYP1A (EROD) activity and alcohol dehydrogenase activity (ADH), in an endemic Mexican fish species, the black-fin goodeid Girardinichthys viviparus, exposed to the water of two localities, Lake Texcoco (LTX) and Lake Zumpango, and to the same matrices enriched in polychlorinated biphenyls (PCBs) to simulate the potential toxic effects of sublethal increases in these xenobiotics. Fishes of both sexes born in the laboratory were exposed for 1, 2, 4, 8, and 16 days. Water from the two types of localities of the black-fin goodeid contains MFO inducers. Of the two, the most contaminated is LTX water, which also contains PCBs. EROD activity was higher in all treatments with female compared with male fish. This suggests greater metabolic compromise in female fish as a response to damage caused by these xenobiotics. In this species, CYP induction displayed two patterns that were not always concurrent with higher CYP1A activity. In the A. Vega-Lo´pez (&) Laboratorio de Toxicologı´a Ambiental, Escuela Nacional de Ciencias Biolo´gicas, Instituto Polite´cnico Nacional, Avenida Wilfrido Massieu s/n, Unidad Profesional Adolfo Lo´pez Mateos ‘‘Zacatenco’’, Del. Gustavo A. Madero, Me´xico Distrito Federal, CP 07738, Mexico e-mail:
[email protected] F. A. Jime´nez-Orozco Laboratorio de Farmacologı´a Celular y Molecular, Facultad de Medicina, Universidad Nacional Auto´noma de Me´xico, Me´xico Distrito Federal, CP 04510, Mexico L. A. Jime´nez-Zamudio E. Garcı´a-Latorre M. L. Domı´nguez-Lo´pez Laboratorio de Inmunoquı´mica I, Escuela Nacional de Ciencias Biolo´gicas, Instituto Polite´cnico Nacional, Me´xico Distrito Federal, CP 11340, Mexico
enriched matrix system, biotransformation processes were notably altered. Increased ADH may indicate that this enzyme is involved in the biotransformation of PCBs and their metabolites, particularly in male fish, and provides at least a part of reductive power required by the MFO enzymes; however, specific studies are needed to clarify this point.
The Comprehensive Environmental Response, Compensation, and Liability Report ranked polychlorinated biphenyls (PCBs) fifth in a toxicologic study because of the risks they entail (http://www.atsdr.cdc.gov/cercla/07list.html). PCBs and various other xenobiotics are mixed-function oxidase (MFO) phase I agonists. MFO enzymes may be either microsomal, such as cytochrome P450 (CYP), or cytoplasmic, such as alcohol dehydrogenase (ADH) (Gray et al. 1991; Moreno-Grau 2003). Cytochrome P450 1A (CYP1A) is one of the most important isoenzymes induced by arylhydrocarbon receptor agonist, such as PCBs (Whyte et al. 2000; Srogi 2008). However, PCB induction of CYP seems to be dependent on chlorination patterns; it has been considered that non-ortho-substituted chlorinated biphenyls are strong inducers of these cytochromes (Schlezinger et al. 2006), but an important species-dependent variation has been observed. Certain PCB congeners-such as 3,30 ,4,40 tetrachlorobiphenyl (PCB 77) and 2,3,30 ,4,40 ,5-hexachlorobiphenyl (PCB 156)-did not induce CYP1A activity in some fish species (Kuzyk et al. 2005; Yuan et al. 2006). However, under chronic exposure it is possible to observe synergic effects between PCB 77 and other CYP1A inducers (Gunawickrama et al. 2008). It has also been reported that di-ortho-substituted biphenyls do not elicit induction of CYP1A (Machala et al. 1998). Nevertheless, little is known about the role of cytoplasmic MFO
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enzymes, such as ADH, in the biotransformation of PCB metabolites (Moreno-Grau 2003). Alcohol dehydrogenase generates reducing power by way of a redox process, and it is possible that the enzyme has a double function: Nicotinamide adenine dinucleotide could be used as an energy source by different enzymes, or it may be involved in the biotransformation of PCBs and their metabolites through increasing their water solubility by alcohols and aldehyde formation and corresponding acids (Moreno-Grau 2003; Vega-Lo´pez et al. 2007a). Information on the induction of MFO enzymes in species of the Goodeidae family on exposure to PCBs is scarce. A previous study found increases in ADH and ethoxyresorufin-O-deethylase (EROD) activity and total CYP content in the black-fin goodeid Girardinichthys viviparus exposed to PCBs in synthetic water (Vega-Lo´pez et al. 2007a). However, there is still a lack of information on the induction of CYP isoenzymes in this fish species by exposure to water polluted with PCBs. The use of biomarkers in monitoring programs could identify several aspects, including exposure to environmental stressors; monitor spatial and temporal changes on these compounds; and provide early warning indicators on the health status of wild populations (Wu et al. 2005). It has been widely reported that complex matrices, such as contaminated waters, increase CYP1A (EROD) activity (Whyte et al. 2000). Several studies have found that CYP content induction is elicited in some fish species by contaminated water (Maria et al. 2003a, b; Gravato and Santos 2003). However, no studies on complex matrices, such as polluted water, have evaluated ADH activity and its role in biotransformation. All of these biomarkers should be useful in demonstrating a good doseresponse relation to levels of pollutants; exhibiting timeintegrated responses; and showing clear and explicit end points in relation to environmental contamination (Wu et al. 2005). Alterations in the MFO system could also be ecologically relevant. At the beginning of the 16th century, lakes covered approximately 9,600 km2 in the Valley of Mexico. Up to the middle of the 20th century, these contained 15 fish species, 9 of which are endemic; however, these have since ´ lvarez del Villar and Navarro 1957). Habitat disappeared (A disappearance was the principal cause of extinction, but damage by toxic substances may have contributed to the loss of species. The black-fin goodeid G. viviparus (Bustamante 1837) is now the sole surviving endemic species in this area (Vega-Lo´pez et al. 2007b, 2008). It is an endangered species native to Mexico (http://www.semarnat. gob.mx/leyesynormas/Normas%20Oficiales%20Mexicanas %20vigentes/NOM-ECOL-059-2001.pdf) whose natural habitat is now reduced to three localities: the Xochimilco canals, Lake Texcoco (LTX), and Lake Zumpango (LZ). These localities are within the metropolitan area of Me´xico
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City and are fed by domestic and industrial wastewater treated at a secondary level (Vega-Lo´pez et al. 2007b). Wastewater and treated water contain many persistent xenobiotics-such as PCBs, organochlorine pesticides, dioxins, aromatic hydrocarbons, and other toxic substancessome of which are aryl-hydrocarbon receptor agonists (Whyte et al. 2000; van der Oost et al. 2003; Srogi 2008). For the remaining localities of the black-fin goodeid, we detected PCBs in fish from LTX that originated from the discharges of a thermoelectric plant located in the influence zone of this lake (Vega-Lo´pez et al. 2007b). Surprisingly, the black-fin goodeid inhabits only this water body (Alcocer and Willimas 1996). In earlier studies, we reported the estrogenic and antiestrogenic effects elicited by water from LTX and LZ (Vega-Lo´pez et al. 2007b). Furthermore, water from these lakes induces lipid peroxidation and alters the antioxidant response in this fish species (Vega-Lo´pez et al. 2008). In both studies, the response was dependent on the sex of the specimens; higher oxidative stress damage was observed in female than in male fish, whereas male fish were more sensitive to estrogenic substances present in the water. However, the effect of LTX and LZ water on biotransformation processes, such as phase I enzyme induction, in this fish species remains unknown. Given the legal status of the black-fin goodeid, wide environmental monitoring is needed to evaluate the effect of exposure to water from the native localities with the main goal of establishing protective actions for the species. It is thus necessary to use test models that can simulate potential damage elicited by increases in PCBs in the type localities of the black-fin goodeid (Vega-Lo´pez et al. 2007b, 2008). The aim of the present study was to evaluate the short-term effect of exposure to LTX and LZ water, and to the same water enriched with PCBs at sublethal concentrations, on MFO enzymes-including CYP content induction and CYP1A and ADH activity-in both sexes of the black-fin goodeid.
Materials and Methods Specimens and Study Sites Eight-month-old male and female G. viviparus born in the laboratory were used for assays. The parent group of fish was collected after authorization by the Mexican authorities (Permit No./SGPA/DGVD/02750). Multiple 24-h samples were collected from LTX and LZ as previously described (Vega-Lo´pez et al. 2007b). The PCB content in water from the two types of localities was quantified using a Varian 3400 gas chromatograph coupled to a Saturn II mass detector. A Restec capillary column (Supelco; 0.25 mm diameter, 30 m length, 0.5-lm methyl phenyl/
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silicone film) was used for chromatography. The column was first heated to 180°C, then increased by 5°C/min and by 20°C/min to 225°C and held for 10 min. The injector was set to 250°C and the detector to 280°C. The area under the curve was used to determine the relative percentage of each PCB congener in LTX and LZ water as previously reported (Langerbucher 2005). The standard used was a PCB mixture (Inerteen; ITE Circuit Breaker) that was previously characterized (Vega-Lo´pez et al. in press). All PCB residues were neutralized with NaOH to favor chemical decomposition at pH 12 and were then managed by a hazardous waste company. Sublethal Exposure Lethal toxicity tests were conducted for 96 h with water from the native localities. In no case were lethal effects observed with water from LTX and LZ (Vega-Lo´pez et al. 2007b). The lethal toxicity of PCB mixtures has been documented in Vega-Lo´pez et al. (2006). Sublethal exposure experiments were performed as follows. Pure dimethyl sulfoxide (DMSO) dissolved in semihard synthetic water (60 mg CaSO42H2O, 80 mg KCl, 1.2 g MgSO4, and 1.92 g NaHCO3/L deionized water) was used for the control group. Four test groups were set up in addition to the control group: (1) 100% LTX water; (2) 100% LTX water enriched with a nominal concentration of PCB mixture (0.92 mg PCBs/L) dosed at half of LC0 (the estimated concentration at which no deaths occur within 96 h) using DMSO as a vehicle for the PCB mixture; (3) 100% LZ water; and (4) 100% LZ water enriched with a nominal concentration of PCB mixture at half LC0. In no case was the DMSO concentration [ 0.01%. The exposures were performed at 25°C in static medium in glass aquaria with constant low aeration under natural light/dark conditions at a density of 0.8 g fish/L (Vega-Lo´pez et al. 2007a), with 20 specimens of each sex per treatment. The medium was replaced every 4 days, and fish were not fed for that period. After 1, 2, 4, 8, and 16 days, three male fish and three female fish (average size 25.54 ± 2.8 and 39.55 ± 4.6 mm, respectively) from each test group were anesthetized with xylocaine and killed by fast freezing at -70°C. The carcasses were defrosted in ice water to obtain better quality preservation of the tissues. Livers were homogenized (1:10 w/v) at 8,000 rpm in a K T25 basic IKA Labortechnik homogenizer and centrifuged at 9,000 9 g in an Allegra 64R centrifuge (Beckman Coulter) for 20 min. The homogenates were analyzed for specific biomarkers as follows (1) in iced buffer solution (0.25 M sucrose and 50 mM Tris-HCl [pH 7.4]) for CYP content, (2) in buffer A (0.1. mM ethylenediaminetetraacetic acid, 1.0 mM dithiothreitol, and 20% glycerol v/v) chilled on ice for CYP1A activity, and (3) in phosphatebuffered saline [pH 6.5]) for ADH activity. After
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centrifugation all of these fractions were frozen at -70°C for the biomarker assays. In no case were the samples were subjected to [ 1 freezing–unfreezing cycle. Biomarker Assays CYP was evaluated using the procedure of Omara and Sato (Matsubara et al. 1976). The CYP concentration was determined using a molar extinction coefficient of 104 mM-1 cm-1 as the difference in absorbance between the peaks at 450 and 490 nm. The activity of cytochrome P450 1A was determined as EROD using the Parrot et al. (1999) with a 0.1 to 1.0 nM resorufin calibration curve using a Perkin–Elmer LS-55 fluorescence spectrometer (excitation 520 nm and emission 585 nm). ADH activity was determined as reported by Vega-Lo´pez et al. (2007a) using a 30- to 90-nM acetaldehyde curve. Total protein content was determined according to the Bradford (1976) method using bovine serum albumin as standard. All reagents were from Sigma (St. Louis, MO). To eliminate sources of variation, such as sample and fish sizes, the results for CYP1A and ADH activity were normalized by min/mg protein/g tissue/g fish (wet weight [ww]), and CYP content was standardized by mg protein/g tissue/g fish (ww). Statistical Analysis Results were subjected to one-way analysis of variance followed by Dunnett’s multiple comparison test. The minimum probability criterion for significant differences was set at p B 0.05 compared with the control. Differences between water from each native locality and the same water sample enriched with PCBs at a given exposure time were evaluated using Mann-Whitney rank sum test. The minimum probability criterion for significant differences was set at p B 0.05.
Results Evaluation of PCBs in the Water of Native Localities No PCBs were detected in LZ water; however, LTX water contained 9.4 ± 0.3 ng PCBs/L (Table 1). The most abundant PCBs in LTX water were di-ortho-substituted biphenyls (60.41%), followed by mono-ortho-substituted biphenyls (27.20%) and non-ortho-substituted biphenyls (12.39%). 2,20 ,3,30 -Tetrachlorobiphenyl (16.62%), 2,3,6trichlorobiphenyl (5.63%), and 4,40 -dichlorobiphenyl (7.60%) were the must abundant PCBs in each of these classes, respectively. The composition of the PCB mixture
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Table 1 Relative composition of polychlorinated biphenyls from Lake Texcoco water, a remaining locality of the black-fin goodeid G. viviparus Mono-ortho-substituted (%)a
Di-ortho-substituted (%)a
Non-ortho-substituted (%)a
2-chlorobiphenyl (2.63)
2,20 -dichlorobiphenyl (0.50)
3-chlorobiphenyl (0.60)
2,4-dichlorobiphenyl (4.45)
2,20 5-richlorobiphenyl (8.24)
4-chlorobiphenyl (0.74)
20 ,4-ichlorobiphenyl (0.56)
2,20 ,3,30 -tetrachlorobiphenyl (16.62)
4,40 -dichlorobiphenyl (7.62)
2,3,4-trichlorobiphenyl (2.60)
2,20 ,3,40 -tetrachlorobiphenyl (12.42)
3,30 ,5,50 -tetrachlorobiphenyl (3.43)
0
0
0
2,3,4 -trichlorobiphenyl (1.15)
2,2 ,3,5 -tetrachlorobiphenyl (4.54)
2,30 ,4-trichlorobiphenyl (2.92)
2,20 ,5,6-tetrachlorobiphenyl (8.37)
20 ,3,4-trichlorobiphenyl (3.31)
2,20 ,3,30 ,5-pentachlorobiphenyl (3.25)
2,3,6-trichlorobiphenyl (5.63)
2,20 ,3,4,5-pentachlorobiphenyl (4.87)
2,3,40 ,6-tetrachlorobiphenyl (3.36)
2,20 ,3,4,5-pentachlorobiphenyl (1.60)
0
0
2,3 ,5,5 -trichlorobiphenyl (0.59) a
Relative percentage of each chlorinated biphenyl congener
used to enrich the water samples has been reported in Vega-Lo´pez et al. (in press). CYP Content CYP induction was higher in male than in female fish. In male fish exposed to LTX ? PCBs, maximum CYP induction occurred on day 2. In male fish subjected to all other treatments, peak induction occurred on day 4 (Fig. 1a). Female fish treated with water from the two native localities displayed an irregular trend, and only in some cases was there significant induction with respect to control fish (Fig. 1b). In female fish exposed to the PCBenriched matrices, CYP content peaked on day 4. The average response in male fish was higher for treatment with LTX water than that with LZ water (5.44-fold, p B 0.05), whereas the opposite was observed in female fish (0.82fold, p B 0.05).
EROD Activity For all treatments except LTX, EROD induction with respect to the corresponding controls was higher, on average, in female than in male fish. However, the total activity was greater in male than in female fish. Significant increases in EROD activity with respect to controls were observed in male and female fish exposed to LZ water, and this response increased in magnitude with time. In LZ water ? PCBs, EROD activity peaked on day 2 of exposure in male fish (Fig. 2a) and on day 4 in female fish (Fig. 2b), decreasing thereafter. Both sexes of fish treated with LTX water and LTX ? PCBs exhibited an irregular EROD induction trend with time, which was nevertheless significant with respect to controls. Average EROD induction was higher in male fish treated with LTX water than with LZ water (9.83-fold, p B 0.05), whereas the opposite was observed in female fish (0.14-fold, p B 0.05). 3
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Fig. 1 Induction of hepatic CYP content in G. viviparus exposed to its native locality water and to the same water sample enriched in PCBs at sublethal concentrations [half of LC0, which is 0.92 mg PCBs/L]. a Male fish. b Female fish. LZ, Lake Zumpango; LZ ? PCBs, Lake Zumpango enriched sample; LTX, Lake Texcoco;
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* * * a *
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LTX
LTX+PCBs
LTX ? PCBs, Lake Texcoco enriched sample. Error bars represent the standard deviation. Significantly different from controls at * p \ 0.05, ** p \ 0.01 and *** p \ 0.001. Significant differences between the type locality water and the same water sample enriched in PCBs at a set time at bp \ 0.01 and cp \ 0.001
Arch Environ Contam Toxicol (2009) 57:561–570 100
(a)
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Exposure time (d)
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Fig. 2 Induction of hepatic CYP1A (EROD) activity in G. viviparus exposed to its native locality water and to the same water sample spiked with PCBs at sublethal concentrations [half of LC0, which is 0.92 mg PCBs/L]. a Male fish. b Female fish. LZ, Lake Zumpango; LZ ? PCBs, Lake Zumpango enriched sample; LTX, Lake Texcoco;
LTX ? PCBs, Lake Texcoco enriched sample. Error bars represent the standard deviation. Significantly different from controls at * p \ 0.05, ** p \ 0.01 and *** p \ 0.001. Significant differences between the type locality water and the same water sample enriched in PCBs at a set time at bp \ 0.01 and cp \ 0.001
ADH Activity
differences were statistically significant in only a few cases (Fig. 3b). For all treatments for both male and female fish, peak ADH activity preceded or coincided with maximum CYP induction. The response was higher on average in fish of both sexes treated with LTX water than for the LZ treatment (4.23-fold for male fish and 1.05-fold for female fish; p B 0.05). ADH activity induction with respect to the corresponding controls was greater on average in female than in male fish (LZ 1.37-fold and LTX 1.24-fold; p B 0.05).
10000
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* * * ** a * *c *
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* * * c
mM/min/mg protein/g tissue/g fish
mM/min/mg protein/g tissue/g fish
Total ADH activity in male fish was higher than in female fish. In male fish exposed to LZ water and LZ ? PCBs, ADH activity peaked on day 1 and decreased significantly with time. Male fish treated with LTX water and LTX ? PCBs showed a tendency toward increased ADH activity, with peak values occurring on days 4 and 8, respectively, and decreasing thereafter (Fig. 3a). ADH activity was irregular for all treatments for female fish, and
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Fig. 3 Induction of hepatic alcohol dehydrogenase (ADH) activity in G. viviparus exposed to its native locality water and to the same water sample spiked with PCBs at sublethal concentrations [half of LC0, which is 0.92 mg PBCs/L]. a Male fish. b Female fish. LZ, Lake Zumpango; LZ ? PCBs, Lake Zumpango enriched sample; LTX,
LZ
LZ+PCBs
LTX
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Lake Texcoco; LTX ? PCBs, Lake Texcoco enriched sample. Error bars indicate the standard deviation. Significantly different from controls at * p \ 0.05, ** p \ 0.01, *** p \ 0.001. Significant differences between the type locality water and the same water sample enriched in PCBs at a set time at bp \ 0.01 and cp \ 0.001
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Discussion
Fast Induction
CYP Content Induction
A fast-induction response was observed in female fish exposed to enriched matrices and for all treatments with male fish apart from the previously mentioned case. This pattern could be conceived as a ‘‘fast induction–slow recovery’’ response according to Wu et al. (2005). A similar trend has been noted in other species, including the European eel Anguilla anguilla, in which maximum CYP induction occurred after 8 h of exposure to water from a contaminated harbor (Maria et al. 2003a). In this same species exposed to effluent from a wastewater treatment plant, maximum CYP induction occurred at 24 h (Maria et al. 2003b). Exposure to similar conditions elicited significant increases in CYP at 48 h in juvenile sea bass Dicentrarchus labrax (Gravato and Santos 2003). Fast induction in male fish in the present study may be explained by the higher CYP content in this species. In female fish exposed to enriched matrices, high xenobiotic levels may have overcome endogenous regulation by estrogens. Decreased CYP content in both sexes at the end of the observation period is perhaps the result of up-regulation of the xenobiotic response element (Kim et al. 2005). However, induction of CYP isoenzymes by exposure to PCBs could also be dependent on the PCB chlorination patterns of non-ortho-substituted biphenyls, as certain PCBs with two adjacent unsubstituted carbon atoms are considered strong inducers of the CYP system (Van den Berg et al. 1998). If this is true for female black-fin goodeid treated with LTX and LZ water enriched in PCBs, CYP induction could only be elicited by certain PCB congeners, such as 3,30 ,4,40 -tetrachlorobiphenyl and 3,30 ,4,40 ,5-pentachlorobiphenyl (Vega-Lo´pez et al. in press) detected in the PCB mixture used. In addition to 2,3,6-trichlorobiphenyl, 2,2,50 -trichlorobiphenyl and 2,20 ,5,50 -tetrachlorobiphenyl were identified at similar levels in the PCB mixture as in LTX water. This raises some uncertainty, because monoand di-ortho-substituted biphenyls without two adjacent unsubstituted carbon atoms were the most abundant PCB congeners in LTX water and the PCB mixture. Similar findings were noted in female butterfly split-fin goodeid treated with PCBs (Vega-Lo´pez et al. in press). In the present study, the response observed in female black-fin goodeid exposed to PCB-enriched matrices might be explained in terms of synergistic effects of PCB congeners on the CYP system that favor the induction of specific and nonspecific enzymes. This fact remarks the useful of PCBenriched matrices as a powerful tool on the environmental risk-assessment. The observed pattern might be considered as an ‘‘ideal biomarker’’ due to this is able to provide quantitative and time-integrated estimate levels of contaminants in the monitoring environmental (Wu et al. 2005). Few reports are currently available on CYP
A large number of CYP inducers have been reported in various fish species from contaminated waters (Whyte et al. 2000). In the present study, water from the two remaining localities of the black-fin goodeid contained MFO inducers responsible for eliciting CYP synthesis. The CYP superfamily is estimated to include at least 300 proteins assigned to 36 different families of genes (Nelson et al. 1993). Thus, establishing specific links to certain types of xenobiotics that are inducers of this enzyme family and that are present in complex matrices is an extremely complicated task. However, evaluation of CYP as a biomarker of environmental exposure per se is important because of the strong and selective induction of some CYP isoenzymes, which causes a significant increase in CYP content (van der Oost et al. 2003). Using water from LTX and LZ and its corresponding PCB-enriched matrices, we found two patterns of CYP induction with time, i.e., slow induction and fast induction, as follows: Slow or Time-Dependent Induction Slow or time-dependent induction was observed in female fish treated with LTX and LZ water and in male fish exposed to LZ ? PCBs. Wu et al. (2005) reported six hypothetical time-integrated responses of biomarkers; thus, the observed pattern in the current study could be considered a ‘‘slow induction–slow recovery’’ response according to the concept of these investigators. Similar findings were noted in female Ameca splendens, another member of the Goodeidae family, exposed to a PCB mixture in synthetic water (Vega-Lo´pez et al. in press). For female fish, endogenous factors, such as estrogens, inhibit both the activity of CYP isoenzymes (Gray et al. 1991) as well as their transcription (Elskus et al. 1992). Lower levels of these isoenzymes in female black-fin goodeid may explain the slow induction observed. In addition, exogenous factors, such as xenobiotics that compete for estrogen receptors, decrease cytochrome transcription mediated by the aryl-hydrocarbon receptor (AhR) (Sakamoto et al. 2003). The response in male fish exposed to LTX ? PCBs could be a result of these mechanisms or is possibly caused by AhR saturation, as observed in male A. splendens treated with PCBs (Vega-Lo´pez et al. in press). Specific studies to clarify these action mechanisms are necessary. The relevance of this CYP induction pattern confers an advantage to environmental monitoring because the slow recovery may dampen short-term fluctuations (Wu et al. 2005).
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induction by exposure to PCBs in vivo in different goodeid fish species or in male and female fish of this family. There is thus a need for further study of this issue. Induction of EROD Activity In male fish exposed to water from LTX and LZ, maximum EROD activity occurred on days 8 and 16 (5.5- to 36.4fold, p \ 0.01). Similar responses, albeit of lower magnitude, were observed in male rainbow trout exposed to effluent from a treatment plant (Fenet et al. 1998). In male fish of various species exposed to diverse matrices for longer periods, EROD activity was induced up to 79-fold (Livingstone et al. 1993). De novo synthesis of CYP1A in black-fin goodeid female fish may explain the higher EROD induction observed with respect to controls compared with male fish. In female fish exposed to LZ water, the highest increase in EROD activity occurred on day 16 (332.7-fold, p \ 0.001), coinciding with peak CYP content induction. In European eel and rainbow trout female fish exposed in cages for 14 days to a contaminated river, peak EROD activity was 2.2- and 6.6-fold higher, respectively (Fenet et al. 1998). EROD induction in female common carp exposed in cages to effluent from a treatment plant increased from day 4 on, reaching a 25-fold higher level on day 16 (Kosmala et al. 1998). In juvenile Chinook trout (Oncorhynchus tshawytscha) exposed for 28 days to serially diluted treatment plant effluent, EROD was induced approximately 8- to 11-fold (Wilson et al. 2001). The authors found that EROD activity was correlated with para-chloro-dibenzo-dioxin concentrations, a persistent xenobiotic with a long half-life. In contrast, low-persistence EROD inducers such as soluble crude oil fractions elicit a fast response that decreases over time (Pollino and Holdway 2003). Our observations, in agreement with earlier reports by other authors, indicate that EROD activity in female fish is apparently dependent on the persistence of CYP1A inducers; for PCBs this could be explained in terms of the synergistic effects of PCB congeners, as discussed above. Studies must be undertaken to clarify the mechanisms mediating this response. Furthermore, the magnitude of EROD activity in female black-fin goodeid suggests that they are more sensitive than female fish of other species. Compared with our observations with LZ water, LTX water induced irregular EROD activity over time in female black-fin goodeid, although peak activity also occurred on day 16. Wilson et al. (2001) reported that some ortho-chloro-substituted biphenyls decreased EROD activity after an initial period of accelerated induction. The relative bioavailability of mono- and di-ortho-substituted biphenyls (27.20% and 60.41%, respectively) detected in LTX water favored the response observed in female fish exposed to this matrix. In contrast, PCB-enriched matrices
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elicited a fast response (on days 1 and 2) in male fish that was of greater magnitude than the corresponding response in the native matrices. This response was consistent with the pattern of rapid CYP induction. These results show synergistic effects of PCB congeners in black-fin goodeid that induce the CYP system in terms of specific isoenzymes such as CYP1A. These synergistic effects likely occur in this species because of the smaller percentage of nonortho-substituted biphenyls. Our results are in agreement with those of earlier reports in which a rapid EROD response followed by ‘‘stabilization’’ was elicited in fish exposed to high levels of contamination (Kleinow et al. 1987). The observed response could be considered as a ‘‘fast induction-adaptation’’ in agreement with the hypothetical time-integrated responses of biomarkers cited by Wu et al. (2005). This rapid response is the result of early transcription of CYP1A mRNA, which occurs after a few hours (Blanchard et al. 1999). Both the fast response and EROD stabilization may be due to deficiencies in the response of numerous genes regulated by the AhR. Xenobiotics attach to this receptor, allowing its translocation to the nucleus where it dimerizes with the AhR nuclear translocator (ARNT), which then attaches to a specific DNA response element (Goldstone and Stegeman 2006). This complex also activates the transcription of various genes, including CYP1A/1B, by interacting with the xenobiotic response element located in the gene promotor. High concentrations of AhR agonists, such as PCBs, elicit up-regulation of AhR (Hahn 1998). In the present study, lower EROD activity at the end of the exposure period is perhaps due to this up-regulation, which causes deficiencies in formation of the AhR-ARNT complex. EROD induction was over 10-fold higher on average for all treatments for both male and female fish, with the exception of male fish exposed to LZ water and the corresponding PCB-enriched matrix. From the standpoint of environmental monitoring, EROD induction must be 10-fold higher than control values for it to be a reliable early-warning indicator (Wilson et al. 2001). Considering this criterion, we can conclude that the effects elicited by LTX water were stronger than those induced by LZ water. However, it remains unknown if this significant induction is functionally important especially after a long term exposure. PCB enrichment experiments confirmed the risk for black-fin goodeid if PCBs concentration were to increase in its habitat. Nevertheless, it also remains unknown if EROD induction is maintained high or even enhanced after a longer exposure period. Wu et al. (2005) remarks that this sort of responses may merely reflect a faster bioavailability, but it is unable to provided timeintegrated estimations of pollution levels. Also this response could be associated with false-negative or falsepositive results depending on the relative bioavailability
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(Wu et al. 2005). The contamination levels could be underestimated if sampling is carried out after the onset of adaptation in the environmentally exposed populations (Wu et al. 2005). Studies with black-fin goodeid for longer exposure periods are necessary to characterize EROD induction kinetics at subchronic and chronic exposure levels in its natural environment, with the challenge being to develop noninvasive evaluation procedures. Induction of ADH Activity There are a few studies available on the activity of other dehydrogenases in fish exposed to xenobiotics, but they do not document the effect of complex matrices on ADH activity. In the present study, all treatments elicited a rapid ADH response in male fish (6.7- to 55.7-fold with respect to controls). Similarly, in both sexes of Australian crimsonspotted rainbowfish (Melatotaenia fluviatilis), lactate dehydrogenase activity increased 1.33- to 1.48-fold, respectively on day 3 of exposure to a soluble crude oil fraction and to the same fraction with a dispersant (Pollino and Holdway 2003). The authors linked these increases to greater metabolic demand in relation to polyaromatic hydrocarbon detoxification. In male black-fin goodeid exposed to water from LTX and LZ, peak levels of ADH activity preceded or coincided with maximum CYP induction. Cytochrome P450 isoenzymes require (NADPH/NADP?) reductive power, which may be catalyzed by NADPH cytochrome c reductase, to provide the energy for biotransformation (Lemaire et al. 1994). Our findings suggest that ADH is also involved in providing this reductive power, which is of particular importance in male fish, as observed in male butterfly splitfin goodeid treated with PCBs in synthetic water (VegaLo´pez et al. in press). In female black-fin goodeid exposed to water from LTX and LZ, ADH showed an irregular response, which in most cases did not differ significantly, as observed in female black-fin goodeid and butterfly split-fin goodeid, both treated with PCBs in synthetic water (Vega-Lo´pez et al. 2007a, in press). In addition, as CYP isoenzyme activity decreased, so did the ADH response (Figs. 1, 3). An irregular response associated with steroid metabolism has also documented in vitro in female European flounder (Platichthys flesus) hepatocytes for glucose-6-phosphate dehydrogenase (Winzer et al. 2002). In higher organisms, degradation of polyaromatic halogenated compounds due to dehydrogenase activity is among the least understood processes. In microorganisms, short-chain enzymes of the ADH family, the cisbiphenyl-1-2-3-dihydrodiol-2,3-dehydrogenase, have been reported to be involved in aerobic degradation of PCBs to benzoates and 2-hydroxy-penta-2,4-dioenolates (Hu¨lsmeyer et al. 1998). Cytochrome P450 content and EROD activity decreased for LTX ? PCBs treatment toward the end of the observation period, but were concurrent with maximum
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ADH activity. It may therefore be assumed that ADH is also involved in the biodegradation of PCBs into less harmful forms, in agreement with earlier observations of the blackfin goodeid and butterfly split-fin goodeid exposed to PCBs in synthetic water (Vega-Lo´pez et al. 2007a, in press). Based on the present study and a previous report, we propose the existence of two possible mechanisms of ADH involvement in the biotransformation of MFO inducers in the black-fin goodeid. In waters with low levels of contamination, ADH provides the reductive power required by CYP, as in the case of LZ. In waters with high levels of contamination, as in LTX water enriched with PCBs, ADH biotransforms PCBs or their hydroxylated metabolites through increasing its water solubility by alcohols and aldehyde formation and its corresponding acids. Both processes seem to function better in male than in female fish. Many more studies are necessary to clarify these relations and the sex-dependent response. However, it is important to consider that Phase II enzymes play the main role of conjugation and excretion of most of the xenobiotics metabolites (Moreno-Grau 2003).
Conclusion Our findings from this environmental monitoring study and previous studies indicate that evaluation of a single battery of biomarkers is not enough. Biotransformation must be regarded as a set of processes that unleash other responses, including oxidative stress and endocrine disruption. The existence in Lake Texcoco of one remaining wild population of black-fin goodeid indicates that the species has defense mechanisms that are able to compensate for certain increases in contaminant levels, which probably have not yet reached the threshold for damage at the population level. To ensure the survival of the black-fin goodeid, treated wastewater fed into the remaining type localities of this species should be subjected to additional treatment. Acknowledgments We are grateful to the Secretarı´a de Marina, Recursos Naturales y Pesca (SEMARNAT), Direccio´n General de Vida Silvestre (Me´xico). We also thank the staff of the Lago de Texcoco Project, Comisio´n Nacional del Agua, SEMARNAT. This study was financed by the Instituto Polite´cnico Nacional, SIP code 20072167. A. Vega-Lo´pez, E. Garcı´a-Latorre, Luı´s A. Jime´nez-Zamudio and M.L. Domı´nguez-Lo´pez are fellows of Estı´mulos al Desempen˜o en Investigacio´n and Comisio´n y Fomento de Actividades Acade´micas (Instituto Polite´cnico Nacional) and Sistema Nacional de Investigadores (SNI), Me´xico. F.A. Jime´nez-Orozco is a fellow of SNI.
References Alcocer J, Willimas WD (1996) Historical and recent changes in Lake Texcoco, a saline lake in Mexico. Int J Salt Lake Res 5:45–61 ´ lvarez del Villar J, Navarro L (1957) Los peces del Valle de Me´xico. A Secretarı´a de Marina, Direccio´n General de Pesca e Industrias
Arch Environ Contam Toxicol (2009) 57:561–570 Conexas, Comisio´n para el fomento de la Piscicultura Rural. Me´xico, 61 pp Blanchard M, Teil MJ, Carru AM, Ollivon D, Garban B, Chesterikoff A et al (1999) PCB and PAH impacts on cytochrome P-450dependent oxidases in roach (Rutilus rutilus) from the Seine River (France). Arch Environ Contam Toxicol 37:242–250 Bradford 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 Elskus AA, Pruell R, Stegeman JJ (1992) Endogenously-mediated, pretranslational suppression of cytochrome P4501A in PCBcontaminated flounder. Mar Environ Res 34:97–101 Fenet H, Casellas C, Bontoux J (1998) Laboratory and field-caging studies on hepatic enzymatic activities in European eel and rainbow trout. Ecotoxicol Environ Saf 40B:137–143 Goldstone HMH, Stegeman JJ (2006) A revised evolutionary history of the CYP1A subfamily: gene duplication, gene conversion, and positive selection. J Mol Evol 62:708–717 Gravato C, Santos MA (2003) Dicentrarchus labrax biotransformation and genotoxicity responses after exposure to a secondary treated industrial/urban effluent. Ecotoxicol Environ Saf 55: 300–306 Gray ES, Woodin BR, Stegeman JJ (1991) Sex differences in hepatic monooxygenases in winter flounder (Pseudopleuronectes americanus) and scup (Stenotomus chrysops) and regulation of P450 forms by estradiol. J Exp Zool 259:330–342 Gunawickrama SHNP, Aarsæther N, Orbea A, Cajaraville MP, Anders Goksøyr A (2008) PCB77 (3, 30 , 4, 40 -tetrachlorobiphenyl) co-exposure prolongs CYP1A induction, and sustains oxidative stress in B(a)P-exposed turbot, Scophthalmus maximus, in a long-term study. Aquat Toxicol 89:65–74 Hahn ME (1998) The aryl hydrocarbon receptor: a comparative perspective. Comp Biochem Physiol C 121:23–53 Hu¨lsmeyer M, Hecht HJ, Niefind K, Hofer B, Eltis LD, Timmis KN et al (1998) Crystal structure of cis-biphenyl-2, 3-dihydrodiol-2, ˚ resolution. 3-dehydrogenase from a PCB degrader at 2.0-A Protein Sci 7:1286–1293 Kim E-Y, Iwate I, Suda T, Tanabe S, Amano M, Miyazaki N et al (2005) Aryl hydrocarbon receptor (AhR) and AhR nuclear translocator (ARNT) expression in Baikal seal (Pusa sibirica) and association with 2, 3, 7, 8-TCDD toxic equivalents and CYP1 expression levels. Comp Biochem Physiol C 141:281–291 Kleinow KM, Melancon MJ, Lech JJ (1987) Biotransformation and induction: implications for toxicity, bioaccumulation and monitoring of environmental xenobiotics in fish. Environ Health Perspect 71:105–119 Kosmala A, Migeon B, Flamarion P, Garric J (1998) Impact assessment of a water treatment plant effluent using the fish biomarker ethoxyresorufin-O-deethylase: field and on-site experiments. Ecotoxicol Environ Saf 41B:19–28 Kuzyk ZA, Hodson PV, Solomon SM, Reimer KJ (2005) Biological responses to PCB exposure in shorthorn sculpin from Saglek Bay, Labrador. Sci Total Environ 351–352:285–300 Langerbucher F (2005) Interpretation of in vitro–in vivo time profiles in terms of extent, rate and shape. In: Dressman J, Kra¨mer F (eds) Pharmaceutical dissolution testing. Taylor & Francis, Boca Raton, FL, pp 251–279 Lemaire P, Matthews A, Fo¨rlin L, Livingstone DR (1994) Stimulation of oxyradical production of hepatic microsomes of flounder (Platichthys flesus) and perch (Perca fluviatilis) by model and pollutant xenobiotics. Arch Environ Contam Toxicol 26:191–200 Livingstone D, Lemaire P, Matthews A, Peters L, Bucke D, Law R (1993) Pro-oxidant, antioxidant and 7-ethoxyresorufin O-deethylase (EROD) activity responses in liver of dab (Limanda limanda) exposed to sediments contaminated with hydrocarbons and others chemicals. Mar Pollut Bull 26:602–606
569 Machala M, Dra´bek P, Necˇa J, Kola´rˇova J, Svobodova´ Z (1998) Biochemical markers for differentiation of exposures to nonplanar polychlorinated biphenyls, organochlorine pesticides, or 2, 3, 7, 8-tetrachlrodibenzo-p-dioxin in trout liver. Ecotoxicol Environ Saf 41:107–111 Maria VL, Correia AC, Santos MA (2003a) Genotoxic and biochemical responses in caged eel (Anguilla anguilla L.) after short-term exposure to harbour waters. Environ Int 29:923–929 Maria VL, Correia AC, Santos MA (2003b) Genotoxic and hepatic biotransformation responses induced by the overflow of pulp mill and secondary-treated effluents on Anguilla anguilla L. Ecotoxicol Environ Saf 55:126–137 Matsubara T, Koine M, Touchi A, Tochino Y, Sugeno K (1976) Quantitative determination of cytochrome P-450 in rat liver homogenate. Anal Biochem 75:596–603 Moreno-Grau MD (2003) Metabolismo. In: Garcı´a-Brage A (ed) Toxicologı´a ambiental. McGraw-Hill/Interamericana, Madrid, Spain, pp 40–73 Nelson DR, Kamataki T, Waxman DJ, Guengerich FP, Estabrook RW, Feyereisen R et al (1993) The P450 superfamily: update on a new sequences, gene mapping, accession numbers, early trivial names of enzymes, and nomenclature. DNA Cell Biol 12:1–51 Parrot JL, Chong-Kit R, Rokosh R (1999) EROD induction in fish: a tool to measure environmental exposure. In: Rao SS (ed) Impact assessment of hazardous aquatic contaminants, concepts and approaches. Lewis, Boca Raton, FL, pp 99–123 Pollino CA, Holdway D (2003) Hydrocarbon-induced changes to metabolic and detoxification enzymes of Australian crimsonspotted rainbowfish (Melanotaenia fluviatilis). Environ Toxicol 18:21–28 Sakamoto KQ, Nakai K, Aoto T, Yokoyama A, Ushikoshi R, Hirose H et al (2003) Cytochrome P450 induction and gonadal status alteration in common carp (Cyprinus carpio) associated with the discharge of dioxin contaminated effluent to the Hikiji River, Kanagawa Prefecture, Japan. Chemosphere 51:491–500 Schlezinger JJ, Struntz WD, Goldstone JV, Stegeman JJ (2006) Uncoupling of cytochrome P450 1A and stimulation of reactive oxygen species production by co-planar polychlorinated biphenyls congeners. Aquat Toxicol 77:422–432 Srogi K (2008) Levels and congener distributions of PCDDs, PCDFs and dioxin-like PCBs in environmental and human samples: a review. Environ Chem Lett 6:1–28 Van den Berg M, Birnbaum L, Bosveld ATC, Brunstro¨m B, Cook P, Feeley M et al (1998) Toxic equivalency factors (TEFs) for PCBs, PCDDs, PCDFs for humans and wildlife. Environ Health Perspect 106:775–792 Van der Oost R, Beyer J, Vermeulen NPE (2003) Fish bioaccumulation and biomarkers in environmental risk assessment: a review. Environ Toxicol Pharmacol 13:57–149 Vega-Lo´pez A, Martı´nez-Tabche L, Domı´nguez-Lo´pez ML, Garcı´aLatorre E, Ramo´n-Gallegos E, Garcı´a-Gasca A (2006) Vitellogenin induction in the endangered goodeid fish Girardinichthys viviparus: vitellogenin characterization and estrogenic effects of polychlorinated biphenyls. Comp Biochem Physiol C 142(3–4): 356–364 Vega-Lo´pez A, Martı´nez-Tabche L, Galar-Martı´nez M (2007a) Toxic effects of waterborne polychlorinated biphenyls and sex differences in an endangered goodeid fish (Girardinichthys viviparus). Environ Int 33:540–545 Vega-Lo´pez A, Ramo´n-Gallegos E, Galar-Martı´nez M, Jime´nezOrozco FA, Garcı´a-Latorre E, Domı´nguez-Lo´pez ML (2007b) Estrogenic, anti-estrogenic and cytotoxic effects elicited by water from the type localities of the endangered goodeid fish Girardinichthys viviparus. Comp Biochem Physiol C 145:394–403 Vega-Lo´pez A, Jime´nez-Orozco FA, Garcı´a-Latorre E, Domı´nguez Lo´pez ML (2008) Oxidative stress response in an endangered
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570 goodeid fish (Girardinichthys viviparus) by exposure to water from its extant localities. Ecotoxicol Environ Saf 71:94–103 Vega-Lo´pez A, Jime´nez-Orozco FA, Jime´nez-Zamudio LA, Garcı´aLatorre E, Domı´nguez-Lo´pez ML (in press) Prooxidant and antioxidant sex-linked response and its relationship to mixed oxidase function enzymes in liver of Ameca splendens, an endangered goodeid fish exposed to PCBs. Toxicol Environ Chem Whyte JJ, Jung RE, Schmitt CJ, Tillitt DE (2000) Ethoxyresorufin-Odeethylase (EROD) activity in fish as a biomarker of chemical exposure. Crit Rev Toxicol 30:347–570 Wilson JY, Kruzynski GM, Addison RF (2001) Experimental exposure of juvenile Chinook (Oncorhynchus tshawytscha) to
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Arch Environ Contam Toxicol (2009) 57:561–570 bleached kraft mill effluent: hepatic CYP1A induction is correlated with DNA adducts but not with organochlorine residues. Aquat Toxicol 53:49–63 Winzer K, van Noorden CJF, Ko¨ehler A (2002) Glucose-6-phosphate dehydrogenase: the key to sex-related xenobiotic toxicity in hepatocytes of European flounder (Platichthys flesus L.). Aquat Toxicol 56:275–288 Wu RSS, Siu WHL, Shin PKS (2005) Induction, adaptation and recovery of biological responses: implications for environmental monitoring. Mar Pull Bull 51:623–634 Yuan Z, Courtenay S, Chambers C, Wirgin I (2006) Evidence of spatially extensive resistance to PCBs in an anadromous fish of the Hudson River. Environ Health Perspect 114:77–84
