* Nicholas School of the Environment and Earth Sciences and Integrated Toxicology Program, Duke University, Durham, North Carolina, 277080328; and
United States Environmental Protection Agency, National Health and Environmental Effects Research Laboratory, Atlantic Ecology Division, Narragansett, Rhode Island 02882
Received December 14, 2001; accepted February 15, 2002
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ABSTRACT |
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Key Words: cytochrome P4501A; polycyclic aromatic hydrocarbons; creosote; adaptation; acclimation; nongenetic inheritance; killifish; mummichog; Fundulus heteroclitus.
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INTRODUCTION |
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Hahn (1998a) reviewed the transcriptional regulation of CYP1A in fish through a transcription factor known as the aryl hydrocarbon receptor (AhR). The transcriptional regulation of CYP1A in fish and mammals appears to be similar. CYP1A has been studied extensively due to its importance in the metabolism and toxicity (especially carcinogenicity) of certain xenobiotics (Hankinson, 1995). Briefly, CYP1A protein expression and activity, although constitutively present at significant levels, are markedly upregulated in response to the binding of AhR agonists; potent agonists include certain PAHs as well as certain halogenated aromatic hydrocarbons (HAHs) including coplanar PCBs (polychlorinated biphenyls), dioxins, and furans (subsequently collectively referred to as coplanar HAHs). Binding of these agonists to the cytosolic AhR leads to translocation of the AhR to the nucleus, dimerization of the AhR with ARNT ("aryl hydrocarbon receptor nuclear translocator," a misnomer), and subsequent binding to an enhancer region usually termed the XRE (xenobiotic response element) or DRE (dioxin response element). CYP1A is one of several (possibly many) genes possessing at least one 5` DRE. In addition, there is evidence in mammals for regulation of CYP1A by other mechanisms by some chemicals (e.g., Hoffer et al., 1996
; Ledirac et al., 1997
). CYP1A, a monooxygenase, plays an important role in the Phase-I metabolism of many xenobiotic and endogenous chemicals, including PAHs. Thus, many PAHs are both inducers and substrates for CYP1A. In contrast, coplanar HAHs, while often good inducers, are frequently poor substrates for CYP1A.
It may be that for organisms inhabiting environments highly contaminated with coplanar HAHs, persistently upregulated CYP1A activity could be detrimental, in part due to the fact that many of these compounds are such poor CYP1A substrates. For example oxidative stress, generated during the ineffective interaction of CYP1A with coplanar HAHs, may represent an important part of this toxicity (Alsharif et al., 1994; Cantrell et al., 1996
; Schlezinger et al., 1999
; Shertzer et al., 1998
). Similarly, while PAHs are generally good substrates for CYP1A, there is evidence that some CYP1A-produced PAH metabolites (e.g., quinones) may produce reactive oxygen species (ROS) and oxidative stress and/or other toxicities (Di Giulio et al., 1995
; Kim and Lee, 1997
; Livingstone et al., 1990
). Such evidence for toxicity associated with CYP1A metabolism of CYP1A inducers suggests that a refractory CYP1A phenotype could be advantageous to organisms inhabiting environments highly contaminated with either coplanar HAHs or PAHs (Van Veld et al., 1992
; Van Veld and Westbrook, 1995
).
A refractory CYP1A phenotype could be caused by alterations in factors upstream or downstream of CYP1A transcription (Hahn 1998b). Furthermore, those alterations could be the result of genetic changes in the entire population of fish (adaptation), or physiologically mediated alterations in gene expression patterns (acclimation). If low CYP1A inducibility is in fact advantageous in environments highly contaminated with AhR agonists, then the increased toxicity exerted by those chemicals on the individuals with the highest CYP1A responsiveness could drive selection at the population level for individuals with less responsive CYP1A. On the other hand, there are physiological mechanisms that would allow for expression of a refractory phenotype, and examples of the operation of some such mechanisms in AhR-mediated pathways have been found in both piscine and mammalian models (e.g., Celander and Forlin, 1995
; Davarinos and Pollenz, 1999
; Melancon and Lech, 1983
; Mimura et al., 1999
; Schlezinger et al. 1999
). If sufficient physiological plasticity were present to allow for acclimation without toxicity, then selection would not be expected to play a major role.
In the case of either genetically or physiologically based alterations, it seems probable that the organisms would be affected in ways other than just CYP1A inducibility. For example, it has been shown in mammals that binding of the AhR to the DRE regulates the transcription of several (and probably many) other important genes including glutathione S-transferase Ya, NADPH quinone reductase, a UDP glucuronosyltransferase, an aldehyde dehydrogenase, and others (Hankinson, 1995; Nebert et al., 2000
). Furthermore, the AhR pathway plays important roles in development, immune function, and normal liver function (Birnbaum, 1995
; Fernandez-Salguero et al., 1995
; Kerkvliet, 1995
; Schmidt et al., 1996
), and may crosstalk with other signaling pathways such as those involved in cell cycling, oxidative stress responses, and hypoxia responses (Chan et al., 1999
; Nebert et al., 2000
). Thus, an alteration in this pathway seems likely to alter responses other than CYP1A, possibly resulting in reduced fitness for the organisms affected.
Distinguishing whether the CYP1A refractory phenotype is genetically or physiologically based, then, is not just interesting from the perspective of basic biology and toxicology, but potentially important in understanding the long-term population-level impacts of chronic contamination by AhR agonists. In attempting to make such distinctions, it is sometimes assumed that if a phenotype is heritable to F1 offspring in the absence of the stressor, it is genetically based, and if not, then it is physiologically based. As discussed below, however, heritability should be demonstrated at least until the F2 generation, in order to support the hypothesis of genetic alteration. Thus, although the term heritable, when used in biology, is often assumed to imply a genetic basis, its meaning in this paper is restricted to "passed on to offspring." The heritability of the CYP1A refractory phenotype has been investigated in the case of 3 of the adapted populations of fish cited earlier. In the case of the New Bedford Harbor killifish, the refractory phenotype was heritable for 2 generations in laboratory-raised offspring of fish from the contaminated site (Nacci et al., 1999). The refractory phenotype was also heritable to 1- and 8-day-old larvae bred from feral Newark Bay killifish (Elskus et al., 1999
). On the other hand, in the case of the Hudson River Atlantic tomcod, the refractory phenotype, while persistent during several months of laboratory depuration in feral fish, was not heritable to the 1st generation of laboratory-raised offspring (Roy et al., 2001
).
Another important question is whether reduced CYP1A inducibility actually confers a fitness advantage on killifish exposed to AhR agonists in general, and the mixture of contaminants present at the Elizabeth River site in particular. This question has previously been addressed only, to our knowledge, by Nacci and Coiro (manuscript in preparation), who found that low in ovo EROD (ethoxyresorufin-O-deethylase; a measure of CYP1A activity) activity in killifish embryos correlated with decreased susceptibility to the acute toxicity of PCB-126 (3,3`,4,4`,5-pentachlorobiphenyl, a prototypical coplanar HAH), but increased susceptibility to the acute toxicity of benzo[a]pyrene.
The primary hypotheses addressed by this study were: (1) the altered CYP1A refractory phenotype is heritable for multiple generations (i.e., is genetically-based) in the Elizabeth River killifish, and (2) the altered phenotype is adaptive in the context of the Elizabeth River (i.e., lower CYP1A responsiveness results in greater fitness in the presence of Elizabeth River sediment contaminants). In order to test these hypotheses, feral killifish from the Elizabeth River and their laboratory-raised offspring were tested for CYP1A responsiveness, and the results were compared to the CYP1A responsiveness of killifish (feral and laboratory-raised) from 2 reference sites (King's Creek, Virginia, and Beaufort, North Carolina). Multiple developmental stages (embryonic, larval, and adult) of each generation were also tested, in order to further characterize the persistence of the CYP1A refractory phenotype within single generations. The results from this study suggest that the CYP1A refractory phenotype is neither genetically based (although it is somewhat heritable to 1st-generation offspring), nor associated with enhanced fitness in Elizabeth River killifish.
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MATERIALS AND METHODS |
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Chemicals, sediments, and sediment pore water.
All chemicals were purchased from Sigma Chemical Company (St. Louis, MO) except PCB-126, which was purchased from Chem Service (West Chester, PA), and acetone, Tris-HCl, sucrose, and dimethylsulfoxide (DMSO), which were purchased from Mallincrodt Baker, Inc. (Phillipsburg, NJ).
Sediments were collected at the Elizabeth River and King's Creek (reference site), homogenized, sealed, and stored at 4°C for up to 6 months. "Sediment pore water" was obtained via centrifugation (2500g, 20 min, 4°C) of contaminated-site sediments; the supernatant was sealed and stored until use (up to 1 week) at 4°C. This sediment pore water (supernatant) was then diluted to varying degrees with clean artificial seawater (Instant Ocean) for exposures. Sediment pore water was used rather than simple sediment exposures in order to ensure even dosing of different individuals and because of the difficulty of observing very small larvae in sediments without excessive disturbance of the larvae (which are hard to see in beakers containing sediments).
Dosing and Exposures
ß-naphthoflavone (BNF) injection.
For the experiment shown in Figure 1, BNF was administered intraperitoneally in corn oil carrier; controls were injected with corn oil. Injection volumes were 5 µl/g wet weight; the dose of BNF used to induce was 50 mg/kg, which was determined in previous range-finding experiments (not shown) to lead to maximal EROD induction by day 2 in reference site killifish; in the experiment shown, fish were sacrificed on day 3.
Waterborne and sediment exposures.
Adults exposed to Elizabeth River and King's Creek sediments were exposed in divided tanks (i.e., all dosage groups exposed to a given sediment were in the same tank) with a 20:1 (v:v) ratio of artificial salt water to sediment. Larvae and embryos were exposed to dilutions of Elizabeth River sediment pore water, obtained as described above (dilutions specified in figure legends). 3-Methylcholanthrene (3-MC) and BNF were delivered waterborne in an acetone carrier for embryo and larval exposures; ethoxyresorufin was delivered waterborne in a DMSO carrier for embryo exposures. In all cases, controls were treated with the appropriate carriers. In addition, early in ovo EROD experiments (including the experiment presented in Fig. 3) were conducted with clean artificial saltwater controls. However, since no significant difference was seen between that and carrier controls, carrier controls only were used in all subsequent experiments.
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Preparation of microsomes.
Tissues (individual adult livers, pools of 6 larvae, or pools of 10 embryos) were rinsed in ice-cold KCl (0.15 M, pH 7.4) and homogenized in 4 volumes of ice-cold buffer (0.1 M TrisHCl, 0.25 M sucrose, 1 mM EDTA, 1 mM phenylmethylsulfonylfluoride (pmsf), pH 7.4). The homogenate was centrifuged at 4°C at 10,000 x g for 20 min; the resulting supernatant (excluding the fatty layer) was centrifuged at 4°C at 105,000 x g for 1 h. After aspirating the supernatant, the resulting pellet was resuspended (buffer identical to the homogenization buffer except for the inclusion of 20% glycerol) and either analyzed immediately or snap-frozen in liquid nitrogen. Samples were aliquoted and stored at -80°C and never freeze-thawed more than once before analysis. Protein analysis was performed on a Perkin-Elmer HTS7000 plate reader using the BioRad Protein Assay dye reagent concentrate (Hercules, CA) and bovine serum albumin (BSA) as a standard.
Microsomal EROD measurements.
EROD measurements on microsomes were carried out by a modification of the method of Kennedy et al (1993). Briefly, samples (roughly 50 µg microsomal protein for controls, 5 µg for induced samples from adults; 75 and 50 µg for larval control and induced sample, respectively), blanks, and resorufin standards were loaded in triplicate on a 96-well plate, cofactor buffer was added (0.1 mM NADPH, 0.12 mM NADH, 5 mM MgSO4 in 0.1 M HEPES, pH 8.0), and the reaction was initiated by addition of ethoxyresorufin to a final concentration of 1.25 mM. Blank wells contained all reagents except microsomes. Fluorescence (535 nm excitation, 595 nm emission) was read on a Perkin-Elmer HTS7000 plate reader at 2830°C for 6, 11, or 14 min, depending on the activity level of the microsomes; in all cases, reactions were linear over the time course read. Standards and blanks contained 50 µg BSA. Activities were calculated based on a 7-point standard curve as pmol of resorufin produced per min per mg protein (150 pmol for controls, 10500 for induced).
In ovo EROD.
These activities were measured as described by Nacci et al (1998) in the case of the experiment presented in Figure 3, or by a modification of that technique (Figs. 4 and 5
). The original method involves quantifying, via fluorescence microscopy, the resorufin accumulated in an embryo's bladder after waterborne exposure to ethoxyresorufin (21 µg/l in all cases) and, in some cases, an inducer (chemicals and concentrations varied, and are described in the results section or in figure legends). The embryo and bladder are imaged microscopically, and the fluorescence is quantified with a photometer after excitation of the resorufin, using a rhodamine-red filter set. The initial in ovo EROD measurements were performed at the laboratory of Dr. Diane Nacci, according to these previously described methods, and subsequent experiments were carried out at Duke University using fluorescence microscopy coupled with software-based image analysis for quantification. Extensive initial methods development and testing with the image analysis system (IP Lab software, Scanalytics, Inc., Fairfax, VA) showed that under standardized image capture and analysis conditions, software-based analysis of the bladder fluorescence, like photometric analysis, was sensitive and reproducible. Photometric analysis has the advantage of allowing quantification of a physical unit (i.e., millivolts). Image analysis can only provide relative fluorescent intensities, based on pixel values; however, it has the advantage that images can be saved and retrieved for later inspection. In these experiments, waterborne exposures were begun 48 h postfertilization, embryos were transferred to clean water 7 days postfertilization, and fluorescence was read on day 7, 8, or 9 postfertilization.
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Statistical analysis.
Data analysis was performed with Statview© for Windows (Version 5.0.1, SAS Institute Inc., Cary, NC). Most data were analyzed after log-transformation by ANOVA and Fisher's Protected Least-Significant Differences (FPLSD) post hoc analysis was performed where appropriate. Simple regression analysis was performed for the data presented in Figures 1 and 2, and Spearman's rank order method was used to test for correlations between EROD inducibility and resistance to sediment pore water toxicity. In all cases where induced and control EROD responses were compared in the same analysis, analyses were performed on log-transformed data because of heterogeneous variance. Results are graphed as nonlog-transformed data for clarity, however. Details for each particular experiment are included in the appropriate figure legend or section in Materials and Methods. In all cases, initial analysis included all possible factors; pertinent results are presented in figure legends or presented and discussed in the results and discussion sections. In all figures, error bars represent one standard error of the mean.
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RESULTS |
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Figure 2 is a Western blot of CYP1A protein from the microsomal samples analyzed for activity level in Figure 1
. Protein levels were good predictors for activity (r2 = 0.86, p < 0.0001) in a regression analysis comparing protein levels and EROD activity; in particular, the population differences in activity were also observed in protein levels. Protein levels were also determined for all samples from the first adult sediment exposure, subsets of the samples from the second sediment exposure, and subsets of in ovo EROD studies (not shown); in all cases, protein levels correlated reasonably well with EROD activity (r2
0.60).
EROD and CYP1A levels in ovo. EROD activity in Elizabeth River F1 embryos was not detectably induced upon exposure to 268 ng/l 3-MC, while the EROD activity in Elizabeth River F2 embryos was nearly as high as in the King's Creek F1 and F2 embryos (Fig. 3) (FPLSD: p < 0.0001 for all pair wise comparisons except KCF1 vs. KCF2, p = 0.0004; ERF2 vs. KCF1, p = 0.0836).
In a second in ovo experiment, the EROD activity in Elizabeth River F1 embryos again was not induced, now upon exposure to 100 ng/l BNF (Fig. 4). The EROD activity in Elizabeth River F3 embryos and embryos from reference sites (King's Creek F2s and Beaufort F1s) was induced normally, and the EROD activity in hybrid crosses (offspring of feral, laboratory-held Elizabeth River females and King's Creek F1 males are indicated as "EK," and offspring of King's Creek females and Elizabeth River males labeled "KE") was induced in an intermediate fashion (Fig. 4
). The EROD activity of the two reference site embryos was similar (p = 0.0587, FPLSD), and the EROD activity in Elizabeth River F3 embryos was not distinguishable from that of either reference site (p = 0.3802 vs. King's Creek, p = 0.2942 vs. Beaufort, FPLSD). The EROD activity in Elizabeth River F1 embryos was not induced detectably, and the basal EROD activities were far below those of the reference site and Elizabeth River F3 embryos (p < 0.0001 for all comparisons, FPLSD). Hybrid offspring were significantly different from all other groups (p < 0.0001 for all comparisons except vs. each other, p = 0.0028), and in general exhibited an intermediate response between each set of half-siblings, with the offspring of Elizabeth River males and King's Creek females showing a slightly greater degree of activity than the hybrid offspring of Elizabeth River females and King's Creek males.
When exposed to a range of concentrations of BNF (Fig. 5), the EROD activity of Elizabeth River F1s was not induced detectably, even at the highest nominal concentration of 1 mg/l (p = 0.3745 for effects of dose on ERF1 embryos only). King's Creek F2 embryos showed a maximal EROD response at 10 mg/l BNF, with decreasing activity thereafter. This decrease in EROD activity correlated with an increase in deformities (not seen in Elizabeth River offspring), as described below. The EROD activity in the hybrids was induced in a manner intermediate between the EROD activity in the 2 nonhybrid groups, and in this case no difference was observed between the EROD activities of offspring of Elizabeth River males and King's Creek females and offspring of Elizabeth River females and King's Creek males (p = 0.6882, group effect, p = 0.2375, group by treatment interaction).
EROD activity in microsomes from whole-body larval homogenates.
Both population and generational differences were observed in larval EROD activity. CYP1A activity in Elizabeth River F1 larvae was significantly lower than in either King's Creek F1 (p = 0.0006, FPLSD) or Elizabeth River F2 larvae (p = 0.0386, FPLSD) (Fig. 6) after exposure to BNF. EROD activity in Elizabeth River F2 larvae was not significantly different from those in King's Creek F1s (p = 0.0691, FPLSD). Qualitatively identical (same rank order and presence/absence of statistical significance) results were observed when a similar experiment was performed with exposure to Elizabeth River sediments rather than BNF (results not shown).
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Relationship between in ovo EROD activity and survival in Elizabeth River sediment pore water.
When larvae that had been characterized before hatch, regarding EROD inducibility (presented in Fig. 3) were exposed to a 10:1 dilution of Elizabeth River sediment pore water, no relationship was observed between EROD inducibility and survival time (Spearman rank order correlation coefficient = -0.008, p = 0.9334), although visual inspection indicated that none of the embryos with the highest EROD activities were among the longest survivors (Fig. 9
). The same lack of relationship was obtained when survival time was correlated to absolute EROD level, rather than inducibility (not shown). It is important to note that in this experiment, no mortality was observed in Elizabeth River F1 larvae; therefore, this group could not be included in the correlation analysis between EROD induction and survival time. All Elizabeth River F2 and King's Creek embryos were eventually killed by the exposure.
Subsequently, larvae hatched from the embryos used in the experiment presented in Figure 5 were exposed to a 1:1 dilution of Elizabeth River sediment pore water, which was expected to be lethal to at least some of even the most resistant larvae, based on previous data (Meyer and Di Giulio, manuscript submitted). Only larvae hatched from embryos exposed to 1 µg/l BNF in the in ovo EROD experiment (Fig. 5
) were used. In this case, all larvae from all groups died by the end of the experiment, and a significant negative association (Spearman rank order correlation coefficient = -0.659, p < 0.0001) existed between in ovo EROD activity and survival time for larvae.
There was no mortality among the control larvae kept in clean artificial seawater in the experiments shown in Figures 9 and 10.
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DISCUSSION |
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CYP1A of adult, feral Elizabeth River killifish, as previously reported (Carey, 1998; Van Veld and Westbrook, 1995
), was induced poorly compared to reference site killifish (Fig. 1
). The strong correlation between EROD activity and CYP1A protein levels across populations (r2 = 0.86, p < 0.0001 for the data shown in Figures 1 and 2
, and unpublished data on other samples including embryo homogenates) suggests that most of the population-level differences observed in EROD activity could be attributed to altered induction of CYP1A protein synthesis, rather than altered enzyme function.
The refractory CYP1A phenotype observed in the feral Elizabeth River killifish was also observed in early developmental stages of their F1 offspring, which as embryos did not display detectable EROD increases following exposure to any concentration of any compound tested (Figs. 35). The same phenomenon was observed at the levels of both CYP1A protein and activity upon exposure to PCB-126 up to 1 µg/l (Meyer and Di Giulio, in press
), although Elizabeth River sediment pore water exposure led to a small degree of induction (Meyer and Di Giulio, unpublished data). The fact that the EROD activity of feral Elizabeth River killifish was inducible (although minimally), while their offspring showed no detectable EROD induction as embryos, is currently unexplained mechanistically. One possibility is that there are inducers present in the Elizabeth River sediments which singly or in combination are more powerful even than very high concentrations of prototypical inducers such as BNF and PCB-126; this possibility is supported by the observation that the only EROD induction we detected in Elizabeth River F1 embryos occurred after exposure to Elizabeth River sediment pore water.
Further evidence for the heritability of the refractory phenotype is offered by the intermediate CYP1A response observed in the hybrid crosses (Figs. 4 and 5). The results with the hybrids also indicate that the refractory CYP1A phenotype can be inherited from both male and female Elizabeth River parents.
Interestingly, however, the highly refractory CYP1A phenotype observed in Elizabeth River F1 embryos was less pronounced at the larval stage (Fig. 6), where EROD induction in the F1 larvae, while poor, did occur. Finally, the EROD activity of mature adult Elizabeth River F1 and F2 offspring was induced relatively well, although the response was still statistically distinct from that of the reference site killifish (Figs. 7 and 8
). The normal maximum lifespan for killifish in the wild is probably 45 years (Kneib and Stiven, 1978
), and sexual maturity is reached at 912 months under our laboratory conditions. Thus, despite the fact that the refractory phenotype was completely heritable at the embryonic stage, the refractory nature of the CYP1A response slowly diminished over the lifespan of the F1 offspring, suggesting a nongenetic, persistent, but not permanent pattern of inheritance.
A significant sex effect on EROD activity (males on average had higher EROD activity than females) was present in both of the experiments carried out with reproductively active fish (Figs. 1 and 7). This effect is consistent with previously published results (Elskus et al., 1992
; Forlin and Hansson, 1982
). However, our results do not definitively indicate whether the nature of the refractory CYP1A phenotype in Elizabeth River killifish is modulated by sex; the refractory phenotype was consistent in both sexes in the feral fish (Fig. 1
), but was sex-dependent (observed only in the males) in reproductively active laboratory-raised F2 adults (Fig. 7
).
At first glance, the changes in CYP1A phenotype observed in different developmental stages and generations described above might seem to reflect a developmental stage-specific or tissue-specific adaptation to the Elizabeth River sediments. A developmental stage-specific pattern of adaptation theoretically seems plausible given the relatively increased vulnerability of early life stages. Similarly, a tissue-specific adaptation also appears theoretically reasonable. For this study, EROD activities and CYP1A protein levels were measured in different experiments on liver microsomes, in whole-body larval homogenates, and in ovo (presumably incorporating various tissues). It is well known that CYP1A is present in many tissues in fish (e.g., Bello, 2001), and it is conceivable that part of the change observed in the Elizabeth River killifish is tissue-specific; this has indeed been demonstrated to be the case for feral Elizabeth River killifish (Van Veld and Westbrook, 1995
).
However, the trend in CYP1A responsiveness through multiple generations of laboratory-raised Elizabeth River killifish supports the hypothesis of a heritable, nongenetic effect. The refractory phenotype diminishes from one generation to the next, when analyzed in the same tissue, whether embryos (Figs. 3 and 4), larvae (Fig. 6
), or adult livers (Figs. 1, 7, and 8
). The F3 generation, in the one experiment presented here, was not distinguishable from reference site embryos (Fig. 4
; it should be noted that Elizabeth River F3 generation embryos were statistically, although minimally, distinct from reference site embryos in a separate experiment reported in Meyer and Di Giulio, in press). Even the F2 generation, while always exhibiting a lower CYP1A induction response than reference site fish, was not always statistically distinguishable from the reference site killifish (Figs. 3 and 6
). The fact that CYP1A in F2 and F3 generation embryos and F1 and F2 generation adults was induced almost normally indicates that this is not a developmental stage- or tissue-specific adaptation. In summary, we have observed a trend towards a less refractory CYP1A response that is evident both in successive generations of Elizabeth River killifish raised in the laboratory, and in time or developmental stages within the F1 generation.
A possible genetically based explanation for the loss of the refractory CYP1A phenotype during the lifetime of the Elizabeth River F1 offspring (and in subsequent generations) is selection in the laboratory for the F1 individuals that are the most "normal" (reference site type) in terms of the CYP1A phenotype. In fact, the Elizabeth River F1 generation is characterized by poorer than normal survivorship during development (Meyer and Di Giulio, manuscript submitted). However, this explanation requires that there be significant variability in the CYP1A phenotype in the early stage F1 offspring (i.e., a process of laboratory selection cannot select for the most "normal" responders if there is no significant variability in the response among individuals). As previously discussed, we were unable to detect CYP1A induction in any of the Elizabeth River F1 embryos (the only life stage for which we have a noninvasive assay). Thus, we have not been able to provide any evidence to support the hypothesis of laboratory selection for a more "normal" CYP1A phenotype.
What sorts of mechanisms could lead to such a heritable but impermanent and nongenetic (i.e., unrelated to changes in the actual DNA sequence) effect? This question can be divided into 2 parts: first, what mechanisms might permit a physiologically based downregulation of CYP1A expression and activity, and second, how could these mechanisms be maintained in the absence of the stressor and be transmitted to offspring, without postulating population-level genetic change? The first question has received considerable attention, and relatively long-term repression of inducibility following chronic (months-long) exposure to AhR agonists (PCBs) has been observed in fish (Celander and Forlin, 1995; Celander et al., 1996
). Short-term inactivation of CYP1A via substrate inhibition (Goeptar 1995
; Melancon and Lech, 1983
; White et al., 1997
) and oxidative damage (Schlezinger et al. 1999
) in response to AhR agonists have been documented, but appear to be unlikely candidates for long-term inactivation, especially in the context of PAHs (typically metabolized relatively quickly, compared to HAHs). Alterations in expression patterns of genes related to CYP1A expression have also been observed in response to exposure to AhR agonists, as recently discussed by Bello et al. (2001) and Hahn (1998b). For example, the AhR has been shown to be downregulated proteolytically after binding by AhR agonists in mammalian models (Davarinos and Pollenz, 1999
), and the AhR repressor has been shown to be upregulated by AhR agonists in both mammals (Mimura et al., 1999
) and killifish (Karchner et al., 2002
). Two forms of AhR, ARNT, and the AhR repressor have been identified in killifish (Karchner et al., 1999
, 2002
; Powell et al., 1999
), and possible alterations in the AhR pathway in HAH-resistant killifish from New Bedford Harbor are currently being studied (Bello0 et al., 2001
; Hahn 1998b
; Karchner et al., 2002
; Powell et al., 2000
). Oxidative stress has been shown to downregulate CYP1A1 expression in mammals (Morel and Barouki, 1998
; Nebert et al., 2000
), and significant crosstalk exists between oxidant-responsive signaling pathways and the AhR pathway (Nebert 2000
). We have observed that Elizabeth River sediments elicit biochemical responses indicative of oxidative stress (Meyer and Di Giulio, in preparation), suggesting that this may be another way in which Elizabeth River sediments influence CYP1A expression in killifish.
Thus, there is evidence for multiple nongenetic mechanisms that downregulate CYP1A expression in the context of AhR agonists, both in mammals and in fish, although there is currently no evidence regarding the operations of such mechanisms in the Elizabeth River killifish. How, in the absence of population-level genetic change, could physiologically based alterations in CYP1A expression be maintained in the absence of Elizabeth River sediments, and passed on to F1 generation offspring?
This question has recently been addressed in a similar context by other researchers (Roy et al., 2001). One possibility is a maternal effect of some sort. Maternal loading of PAHs can be significant in killifish and other fish (Hall and Oris, 1991
; Monteverdi and Di Giulio, 2000
), and it is known that mRNA for genes related to CYP1A activity is maternally loaded in zebra fish (Wang et al., 1998
). Maternally loaded mRNA and proteins guide the early formation of developing organisms prior to activation of the zygotic genome (Gilbert 1997
), so that environmentally induced alterations in the gene expression pattern of females could well be transmitted to their offspring. However, if a maternal effect of any sort were present, it would be expected to manifest as a differential CYP1A response in Elizabeth River hybrids (i.e., offspring of Elizabeth River females would be more like the Elizabeth River parents than would offspring of Elizabeth River males), and we did not observe a difference. Thus, it would seem that gene imprinting (McLachlan et al., 2001
) or another relatively persistent, physiologically based mechanism that we have not thought of are better hypotheses.
The second major hypothesis these experiments addressed was that reduced EROD activity would be adaptive in the context of Elizabeth River sediments. This hypothesis is not supported by our data; no relationship was observed when fish demonstrating relatively normal (i.e., reference type; in this experiment, King's Creek F1 and F2 as well as Elizabeth River F2 offspring) EROD and sediment tolerance responses were used (Fig. 9). While some correlation was observed when Elizabeth River F1 purebred and hybrid offspring were included (Fig. 10
), the interpretation of that relationship is complicated by the fact that there was very little variability in EROD response in the Elizabeth River F1 embryos. Furthermore, it is clear from observation of the data that nearly all of the relationship observed can be attributed to the fact that Elizabeth River F1 embryos both showed very little EROD induction and were very tolerant of exposure to Elizabeth River sediment pore water, the hybrid crosses were intermediate for each phenotype, and King's Creek purebred offspring showed relatively strong induction of CYP1A and were intolerant of exposure to the sediment pore water. This correlation could be confounded by other (nonCYP1A related) mechanisms of resistance; i.e., the resistant phenotype could be dependent on a trait completely unrelated to CYP1A, yet still correlate with CYP1A, simply because the CYP1A response is so altered in the Elizabeth River F1 embryos. The lack of a correlation between EROD response and survival time in King's Creek F1 and F2 and Elizabeth River F2 fish (Fig. 9
), all of which showed significant variability in both EROD and survival responses, suggests that EROD activity has little relationship to relatively short-term larval survival upon exposure to Elizabeth River pore water. An alternative explanation is that this relationship exists only in the case of very low EROD activity (like that observed in the Elizabeth River F1s), but is lost above a relatively low threshold of EROD activity (i.e., in the case of EROD activity in the range of that observed in King's Creek F1 and F2 and Elizabeth River F2 fish). It is also possible that exposure to sediments rather than sediment pore water would produce a different result, or that a relationship between EROD activity and toxicity might exist for other endpoints not measured in this study (e.g., teratogenesis, chronic toxicity, carcinogenesis). For example, Shimizu et al. (2000) recently showed that lack of CYP1A inducibility in AhR-deficient mice correlated with resistance to benzo[a]pyrene-mediated tumorigenesis.
This result (lack of relationship between EROD activity and survival time) is in contrast to that reported with killifish embryos exposed to either PCB-126 or benzo[a]pyrene, as discussed in the introduction (Nacci and Coiro, manuscript in preparation). This difference is currently unexplained, but may be related to the different contaminants used in the 2 experiments. For example, it may be that high CYP1A activity, while potentially harmful in some contexts involving high levels of certain PAHs (e.g., Kim and Lee, 1997), does not lead to the production of sufficient reactive metabolites in the context of Elizabeth River sediment pore water contaminants to drive acute toxicity in killifish.
Elizabeth River sediment pore water and BNF both produced deformities in King's Creek embryos, but not in Elizabeth River embryos. Similarly, Elizabeth River embryos were resistant to the induction of deformities upon exposure to PCB-126 (Meyer and Di Giulio, in press). The deformities induced by BNF, Elizabeth River sediment pore water, and PCB-126 are not identical, although heart and tail deformities were consistent. The heart and tail deformities, in particular, seem reminiscent of those described for "blue sac" syndrome (Henry et al., 1997
; Walker et al., 1991
), and similar to those observed in killifish exposed to high levels of coplanar HAHs (Prince and Cooper, 1995b
).
There are several important implications of these results. One, and perhaps the most important, is to add weight to the growing body of literature suggesting that chronic contamination is leading to long-term, heritable (genetic or otherwise) effects on entire populations of organisms; this is a type of contaminant effect which has not been well researched in the past. It is possible that such changes may lead to decreased fitness for the populations involved in other contexts ("fitness costs"); this possibility has been discussed previously in the context of environmental pollution (Coustau et al., 2000; Fox, 1995
; LeBlanc, 1994
; Meyer and Di Giulio, manuscript submitted; Shaw, 1999
; Weis et al., 1999
).
A second implication is that, as suggested previously (e.g. Klerks and Levington, 1989; Theodorakis and Shugart, 1999
), it is not sufficient to look at just one generation of offspring or hybrid crosses in order to distinguish between physiologically based acclimation and genetically based adaptation; nongenetically based acclimations can be quite persistent and even heritable (i.e., capable of being passed on to offspring). From the perspective of basic developmental and evolutionary biology, it is becoming evident that even within the relatively limited number of fish populations studied to date that inhabit AhR agonist-contaminated environments, there is a range of responses in terms of modes of adaptation.
Finally, the marked upregulation of CYP1A mRNA, protein, and activity levels in response to exposure to AhR agonists has led to the frequent use of CYP1A as a biomarker in environmental contexts (Di Giulio et al., 1995; Goksøyr and Husoy, 1998
; Livingstone and Goldfarb, 1998
; Stegeman 1992
). The results of this study support the observation by other researchers (e.g., Elskus et al., 1999
; Hahn 1998b
; Roy et al., 2001
) that the ability of fish or populations of fish to respond to the presence of high and persistent levels of AhR agonists by presenting a phenotype of refractory CYP1A induction may compromise the use of CYP1A as a biomarker under those circumstances.
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ACKNOWLEDGMENTS |
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