ToxSci Advance Access originally published online on May 9, 2006
Toxicological Sciences 2006 92(2):526-536; doi:10.1093/toxsci/kfl011
| ||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||
The Role of the Aryl Hydrocarbon Receptor Pathway in Mediating Synergistic Developmental Toxicity of Polycyclic Aromatic Hydrocarbons to Zebrafish
Nicholas School of the Environment and Integrated Toxicology Program, Duke University, Durham, North Carolina 27708-0328
2 To whom correspondence should be addressed at Nicholas School of the Environment and Integrated Toxicology Program, Duke University, PO Box 90328, Durham, NC 27708-0328. Fax: (919) 668-1799. E-mail: richd{at}duke.edu.
Received February 27, 2006; accepted May 4, 2006
 |
ABSTRACT |
|---|
 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
Planar halogenated aromatic hydrocarbons (pHAHs), such as 2,3,7,8-tetrachlorodibenzo-p-dioxin (dioxin), show strong binding affinity for the aryl hydrocarbon receptor (AHR) and are potent inducers of cytochrome P4501A (CYP1A). It is widely accepted that dioxin toxicity is largely AHR mediated; however, the role of CYP1A activity in causing that toxicity is less clear. Another class of AHR agonists of increasing concern because of their known toxicity and ubiquity in the environment is the polycyclic aromatic hydrocarbons (PAHs). Like dioxin, some PAHs also cause toxicity to early life stages of vertebrates. Symptoms include increased cardiovascular dysfunction, pericardial and yolk sac edemas, subcutaneous hemorrhages, craniofacial deformities, reduced growth, and increased mortality rates. Although developmental effects are comparable between these two types of AHR agonists, the roles of both the AHR and CYP1A activity in PAH toxicity are unknown. As observed in previous studies with killifish (Fundulus heteroclitus), we demonstrate here that coexposure of zebrafish (Danio rerio) embryos to the PAH-type AHR agonist ß-naphthoflavone (BNF) and the CYP1A inhibitor
-naphthoflavone (ANF) significantly enhanced toxicity above that observed for single-compound exposures. In order to elucidate the role of the AHR pathway in mediating synergistic toxicity of PAH mixtures to early life stages, we used a morpholino approach to knock down expression of zebrafish AHR2 and CYP1A proteins during development. We observed that while knock down of AHR2 reduces cardiac toxicity of BNF combined with ANF to zebrafish embryos, CYP1A knockdown markedly enhanced toxicity of BNF alone and BNF + ANF coexposures. These data support earlier chemical inducer/inhibitor studies and also suggest that mechanisms underlying developmental toxicity of PAH-type AHR agonists are different from those of pHAHs. Identifying the pathways involved in PAH toxicity will provide for more robust, mechanistic-based tools for risk assessment of single compounds and complex environmental mixtures. Key Words: AHR; CYP1A; PAH; developmental toxicity; zebrafish; risk assessment.
|
INTRODUCTION |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
Polycyclic aromatic hydrocarbons (PAHs) originate from incomplete combustion such as burning of fossil fuels and invariably occur in the environment as complex mixtures of nonsubstituted, alkyl- and/or N-, S-, and O-ring–substituted PAHs. Although this class of compounds can enter the environment through natural sources, such as ocean oil seeps and forest fires, the majority of PAH deposition can be attributed to anthropogenic activity, including wood treatment and petroleum refining (Douben, 2003
; Latimer and Zheng, 2003), and nonpoint sources such as urban runoff (Van Metre et al., 2000). As such, sediment concentrations of PAHs in urban and industrialized centers can be up to two orders of magnitude higher than those in rural areas (Dabestani and Ivanov, 1999).
Some PAHs and extracts of PAH-contaminated mixtures are toxic to embryonic stages of fish (Billiard et al., 1999; Hawkins et al., 2002; Wassenberg and Di Giulio, 2004a). Collectively, these effects mimic "blue sac disease" characterized for more potent aryl hydrocarbon receptor (AHR) agonists and cytochrome P4501A (CYP1A) inducers such as certain planar halogenated aromatic hydrocarbons (pHAHs), including 2,3,7,8 tetrachlorodibenzo-p-dioxin (dioxin) and coplanar PCBs (e.g., 3,3',4,4',5-pentachlorobiphenyl, "PCB126") (Chen and Cooper, 1999; Elonen et al., 1998; Helder, 1981; Toomey et al., 2001; Walker and Peterson, 1991; Wannemacher et al., 1992). In these studies, the most pronounced effect of chronic, sustained exposure of early life stages of fish species to either pHAHs or PAHs is cardiovascular dysfunction. Recently, dioxin has been shown to adversely affect murine heart development (Thackaberry et al., 2005), consistent with effects previously observed in piscine and avian embryonic models (Ivnitski-Steele et al., 2005).
As evidenced in numerous studies, the general consensus in mammalian models is that dioxin toxicity is largely mediated through binding to the AHR (Fernandez-Salguero et al., 1996), a ligand-activated transcription factor found in vertebrate species from fish to humans (Schmidt and Bradfield, 1996). In general, AHR pathways are similar among mammals and lower vertebrates, including fishes, reptiles, and birds (Hahn, 1998). Activated AHR regulates expression of a number of genes of the AHR battery, including the phase I enzymes CYP1A1, CYP1A2, and CYP1B1 (Nebert et al., 2000; Whitlock, 1999). Two divergent AHRs, AHR1 and AHR2 (Hahn et al., 1997), have been identified and characterized in killifish (Fundulus heteroclitus) (Karchner et al., 1999) and zebrafish (Danio rerio) (Andreasen et al., 2002). In zebrafish, AHR2 binds dioxin with high affinity, is transcriptionally active, and mediates toxic response to dioxin. Conversely, zebrafish AHR1 (AHR1A) lacks the ability to bind pHAHs or PAHs in vitro and is transcriptionally inactive (Andreasen et al., 2002). Since it is nonfunctional in zebrafish, Karchner et al. (2005) suggested ahr1a is a pseudogene in this species. More recently, a novel zebrafish AHR protein belonging to the AHR1 clade, AHR1B, has been discovered (Karchner et al., 2005). Unlike ahr1a, the ahr1b gene encodes a functional, transcriptionally active AHR protein that is expressed early in zebrafish development. In contrast to AHR2A, AHR1B expression is not inducible by dioxin (Karchner et al., 2005). Thus, AHR1B may play an important physiological role during embryogenesis.
The recent advent of antisense morpholino (MO) oligonucleotide gene technology has shown that dioxin-induced toxicity in zebrafish (D. rerio) early life stages is also AHR dependent and corroborates mammalian studies. Specifically, knock down of the AHR2 isoform in zebrafish embryos blocks CYP1A expression in the vascular endothelium and protects against dioxin-induced toxicity, including pericardial edema and craniofacial deformities (Carney et al., 2004; Dong et al., 2004; Teraoka et al., 2003). Several studies have shown that inhibition of CYP1A activity or targeted knock down of CYP1A protein expression also prevents or rescues toxic responses to dioxin or PCB126 in zebrafish embryos, thus suggesting that CYP1A also plays a role in triggering dioxin-mediated developmental toxicity in fish embryos (Cantrell et al., 1996; Dong et al., 2004; Teraoka et al., 2003). However, a recent zebrafish study showed that while MO knock down of AHR2 protected embryos from the developmental toxicity of dioxin, MO knock down of CYP1A had no effect (Carney et al., 2004).
Although the suite of cardiovascular defects are very similar between PAH- and pHAH-type AHR agonists, the roles of the AHR and CYP1A activity in the toxicity of PAH are less clear. In previous studies we have shown that exposure of killifish (F. heteroclitus) embryos to model PAHs that are agonists for the AHR, e.g., ß-naphthoflavone (BNF) and benzo[a]pyrene (BaP), or PAH-contaminated mixtures caused significantly greater-than-additive interactions (in terms of cardiovascular effects) when combined with CYP1A inhibitors (Wassenberg and Di Giulio, 2004a; Wassenberg et al., 2005). This synergy is in marked contrast to responses with dioxin and inhibitors, where CYP1A chemical inhibition or protein knockdown either reduces (Cantrell et al., 1996; Dong et al., 2002) or does not alter toxic response to dioxin during zebrafish development (Carney et al., 2004). We have also shown that killifish embryos cotreated with the pHAH PCB126 and the CYP1A inhibitor -naphthoflavone (ANF) decreased both the frequency and severity of deformities compared with embryos exposed to PCB126 alone (Wassenberg and Di Giulio, 2004a).
The co-occurrence of agonists and inhibitors is typical of environmental PAH mixtures (Wassenberg et al., 2005; Willett et al., 2001). However, current risk assessments for PAHs assume an additive model of toxicity for PAHs (e.g., Barron et al., 2004). Thus, the synergy we have observed has important ramifications for adequately assessing risks posed by exposures to PAHs and also raises compelling questions concerning mechanisms underlying these greater-than-additive toxicities.
In this study, we endeavored to further dissect mechanisms underlying PAH synergy using the genetic tools available in zebrafish. First, we asked whether synergy observed between PAH-type agonists and CYP1A inhibitors could be repeated in zebrafish, or was this phenomenon specific to killifish embryos. Upon confirming that a similar toxic response to inducer/inhibitor mixtures occurred in both species, we then used a morpholino approach in zebrafish to dissect the nature of interaction between PAHs that act as AHR agonists and CYP1A inducers and PAHs that act as inhibitors of CYP1A. Our findings with zebrafish embryos in the current study lead us to conclude that (1) PAH synergy appears to be receptor mediated as AHR2 knockdown reduced cardiovascular toxicity characteristic of BNF, ANF coexposures and (2) CYP1A plays an adaptive or protective role, as blocking translation of this protein greatly enhanced the toxicities of nontoxic doses of BNF alone and of BNF combined with ANF.
|
MATERIALS AND METHODS |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
Animals.
Newly fertilized zebrafish embryos were collected at 20-min intervals in 1x Danieau's water (Nasevicius and Ekker, 2000
) following mating of AB* strain adults maintained at 28°C on a 14-h light and 10-h dark-light cycle. Embryos were maintained in 1x Danieau's water under the same temperature and photoperiod conditions for the duration of experiments.
Morpholinos.
Based on cDNA sequences submitted to GenBank, we used morpholinos (Gene Tools, Philomath, OR) that had been designed previously to block initiation of translation of zebrafish aryl hydrocarbon receptor 2 (zfahr2-MO) mRNA (AF063446
[GenBank]
; Teraoka et al., 2003) and cytochrome P4501A (zfcyp1a-MO) mRNA (AB078927; Carney et al., 2004). Sequences used were zfahr2-MO (5'-TGTACCGATACCCGCCGACATGGTT-3') and zfcyp1a-MO (5'-TGGATACTTTCCAGTTCTCAGCTCT-3'). Gene Tools' standard control morpholino (control-MO) (5'-CCTCTTACCTCAGTTACAATTTATA-3') was used as control-MO in experiments. Morpholinos were diluted to working stocks (control-MO and zfcyp1a-MO to 0.15mM; zfahr2-MO to 0.1mM) in 1x Danieau's solution for injection (estimated injection volume 2–4 nl; Nasevicius and Ekker, 2000). Injection working stocks were stored at 4°C and heated for 5 min at 65°C prior to use. Control- and cyp1a-MOs were tagged with a 3'-end carboxyfluorescein modification (Gene Tools) in order to monitor injection success and validate experimental protocol using epifluorescence. The ahr2-MO was generously provided by Dr Elwood Linney (Duke University) and did not have the 3'-end fluorescein modification.
Embryo microinjection.
Newly fertilized embryos were screened and healthy embryos at the one- to four-cell stage were injected within 2 h postfertilization (hpf) with either zfahr2-MO, zfcyp1a-MO, or control-MO working stocks using a Narishige IM300 Microinjector (Tokyo, Japan). At 24 hpf, dead and damaged embryos were screened and removed. Nondamaged embryos injected with fluoresceinated cyp1a- and control-MOs were assessed for fluorescent intensity at this time to evaluate injection success and even distribution of the morpholino throughout the embryo cell mass. Only embryos exhibiting strong, uniform fluorescence at 24 hpf were subsequently used in dosing experiments.
Dosing experiments.
Chemicals and dosing concentrations were selected based on our earlier observations of synergistic PAH interactions in killifish (Wassenberg and Di Giulio, 2004a). Model PAHs used in this study were BNF (AHR agonist) and ANF (CYP1A inhibitor). Mechanisms of action and chemical properties for compounds used in the current study have been described in detail elsewhere (Wassenberg and Di Giulio, 2004a).
All chemicals were dissolved in DMSO and stored protected from light at – 20°C. Stock solutions were thawed at room temperature and vortexed prior to dosing. Zebrafish embryos were exposed beginning at 24 hpf to waterborne nominal concentrations of BNF alone and in combination with waterborne nominal concentrations of ANF in DMSO (< 0.1% vol/vol). Embryos were left in original dosing solutions for the duration of the experiment (72–96 hpf).
Because of species differences in sensitivity to dioxin-like toxicity, we first tested a range of BNF and ANF concentrations to characterize dose response for embryotoxicity to zebrafish embryos. In these experiments, 30 zebrafish embryos were exposed to graded concentrations of BNF or ANF in glass beakers to 25 ml of egg water containing 21 µg/l 7-ethoxyresorufin (7-ER, in DMSO) until 96 hpf when embryos (n = 20) were scored blind for pericardial edema, a hallmark symptom of blue sac disease. We then tested whether we could repeat synergy between a PAH-type agonist (BNF) with a CYP1A inhibitor (ANF) previously observed in killifish (Wassenberg and Di Giulio, 2004a). Three replicates of 30 pooled embryos were exposed in glass beakers with 25 ml of egg water containing 7-ER to waterborne BNF with or without ANF (100 µg/l). Embryos were subsampled and transferred to clean egg water at 72 hpf for assay of in vivo ethoxyresorufin-O-deethylase (EROD) activity (n = 6–9 per treatment) and again at 96 hpf for deformity assessment (n = 3–9 embryos per treatment), as described below.
In morpholino experiments, replicate groups (n = 2–3) of five injected and noninjected control embryos were exposed in glass scintillation vials in 7.5 ml of egg water containing 21 µg/l 7-ER to waterborne concentrations of BNF with or without ANF (100 µg/l). Larvae were transferred to clean egg water at 96 hpf and individual larvae imaged and scored blind for measurement of in vivo EROD activity and deformities. In order to confirm specificity of both zfahr2- and zfcyp1a-MO phenotypes, treatments were repeated using embryos that had been injected with the standard control-MO. While ideally this control would be included in each experiment, the large numbers of embryos required for binary combinations of chemical treatments in our morpholino experimental design were prohibitive. We are confident with this approach given that the AHR2 and CYP1A morpholinos have been extensively characterized in the zebrafish literature (e.g., Carney et al., 2004; Dong et al., 2004; Teraoka et al., 2003).
In Vivo EROD activity.
We used a slightly modified in ovo EROD method (Nacci et al., 2005) to measure the CYP1A activity of zebrafish larvae. In all experiments, we used in vivo measurement of CYP1A catalytic activity as a surrogate for CYP1A protein induction or inhibition. Briefly, on day 3 or 4 (72 or 96 hpf) of development, hatched zebrafish embryos were anesthetized with MS-222 and immobilized in 3% methylcellulose. The accumulated resorufin fluorescent product of CYP1A metabolism of 7-ER in the gastrointestinal tract was visualized by fluorescent microscopy (50x magnification using a rhodamine red filter set; Axioskop; Zeiss, Thornwood, NY). EROD activity was measured as fluorescent intensity and quantified digitally by IP Lab Software (Scanalytics Inc, Fairfax, VA). In vivo EROD values were expressed as a percentage of fluorescent intensity in control embryos.
Quantification of teratogenesis.
Day 4 larvae (96 hpf) were anesthetized in MS-222 and immobilized in 3% methylcellulose in order to image the lateral left side. Captured digital images were analyzed blind for teratogenic end points previously described in developing embryos exposed to dioxin (Carney et al., 2004). The first, truncation of Meckel's cartilage or lower jaw gap length was measured as the distance from Meckel's cartilage to the anterior edge of the larvae. We also measured total area of the pericardial sac as an index of pericardial sac edema. Captured images were quantified digitally using IP Lab Software (Scanalytics Inc) and expressed as percentage of control values.
Statistical analysis.
Data were analyzed with Statview for Windows (version 5.0.1; SAS Institute, Cary, NC). Dose-response data shown in Figure 1 were analyzed by simple linear regression with Excel (Microsoft 2003). EROD values and deformity measurements (pericardial edema area and lower jaw gap length) were analyzed by factorial analysis of variance (ANOVA). When ANOVA yielded significance (p < 0.05), Fisher's protected least-significant differences was used as a post hoc test. We used significance of the interaction term to test our null hypothesis that mixtures tested would yield additive response.
|
|
RESULTS |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
In initial experiments examining dose-response relationships (0–500 µg/l) for model PAHs, both BNF-alone (p = 0.0076) and ANF-alone (p = 0.016)–treated larvae showed a dose-dependent increase in the occurrence of pericardial edema (Fig. 1). No deformities were observed at doses used in subsequent BNF-alone (1 and 10 µg/l) and ANF-alone (100 µg/l) coexposure experiments.
Consistent with our previous studies in killifish (Wassenberg and Di Giulio, 2004a), coexposure of zebrafish embryos to nominal waterborne concentrations of BNF (1 µg/l) and ANF (100 µg/l) inhibited BNF-induced EROD activity by 72 hpf (p = 0.0003) and increased cardiovascular toxicity at 96 hpf, specifically pericardial edema (p = 0.0002) and craniofacial deformities (p = 0.0006), above those observed for single-compound exposures (Fig. 2).
|
Based on these results we used a morpholino approach to test the hypothesis that targeted knock down of CYP1A protein would mimic the inhibitory effect of ANF on BNF-induced EROD activity and deformities previously observed in zebrafish and killifish embryos (Wassenberg and Di Giulio, 2004a
). In vivo EROD activity was used in all knockdown experiments as an index of morpholino function and as a surrogate measure of CYP1A protein induction. In these experiments we used a higher dose of BNF (10 µg/l) because 1 µg/l with CYP1A knockdown did not show toxicity. However, the goal of this experiment was to see whether the cyp1a-MO would shift BNF toxicity, suggesting that ANF acts to synergize by acting as a CYP1A inhibitor. This dose (10 µg/l) was also effective at inducing EROD activity above controls (significant main effect of BNF, p = 0.0098) but was nonteratogenic to noninjected zebrafish embryos (Fig. 3). As observed in chemical inducer/inhibitor studies, the zfcyp1a-MO reduced induction of CYP1A by BNF (significant effect of morpholino and BNF on EROD activity, p = 0.0198) and enhanced toxicity at 96 hpf in zebrafish morphants (Fig. 3). Compared to noninjected control and BNF-treated embryos, pericardial edema area was significantly greater (significant effect of morpholino and BNF on edema, p = 0.0485) and lower jaw gap length significantly enhanced (significant effect of morpholino and BNF on jaw gap length, p < 0.0001) in BNF-treated cyp1a morphants (Fig. 3). We did not observe any phenotype associated with the cyp1a-MO; injected control morphants also appeared normal.
|
The next sets of experiments were designed to define the roles of the AHR and CYP1A in mediating synergistic interactions with respect to cardiovascular effects between PAH-type AHR agonists and CYP1A inhibitors by selective knock down of these specific genes during zebrafish embryonic development. First, we examined whether knock down of AHR2 protein would block synergistic toxic response to BNF, ANF coexposures in zebrafish embryos. As expected, the zfahr2-MO effectively inhibited downstream BNF-induced EROD activity by approximately 50% at 96 hpf observed for noninjected control embryos exposed to 1 µg/l BNF alone (Fig. 4). Interestingly, lower jaw length appeared to be elongated in noninjected embryos exposed to 100 µg/l ANF alone (Fig. 4). For BNF + ANF coexposure groups, zebrafish ahr2 morphants were significantly less deformed than noninjected controls (Fig. 5). As shown by the significant interaction p values, targeted knock down of AHR2 protein clearly protected zebrafish larvae against the severity of pericardial edema (p = 0.0068) and foreshortened lower jaw (p = 0.0017) caused by BNF + ANF treatment (Fig. 4).
|
|
Finally, we examined what role CYP1A activity played in the synergistic embryotoxic effects observed between a PAH-type CYP1A inducer and inhibitor. The deformities caused by the combination of 1 µg/l BNF + 100 µg/l ANF were so severe that while we observed a visual worsening of deformities in cyp1a-MO–treated embryos, we were not able to quantitatively capture these as we had effectively saturated our measurement indices at those doses. We thus lowered the dose of ANF from 100 to 50 µg/l in order to achieve a dose that when combined with 1 µg/l BNF would put the embryo at deformity control levels. This way we could quantitatively show a worsening of deformities with CYP1A knockdown. This concentration of ANF was sufficient to inhibit BNF-induced EROD activity in zebrafish embryos without causing deformities (Fig. 6). In contrast to results observed with zfahr2-MO, targeted knock down of CYP1A protein significantly exacerbated PAH embryotoxicity (Fig. 7), specifically increased accumulation of pericardial fluid (significant interaction among all factors, p < 0.0001), and retarded growth of the lower jaw (significant interaction among all factors, p < 0.0001; Fig. 6).
|
|
|
DISCUSSION |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
The early life stages of fish species are among the most sensitive vertebrates to dioxin-like cardiovascular effects (Peterson et al., 1993
). For this reason and because of the genetic tools available, we used embryonic stages of zebrafish to elucidate the role of AHR and CYP1A protein in PAH developmental toxicity. To summarize, results from the current study demonstrate a seemingly paradoxical role for the AHR pathway in mediating PAH toxicity during the early life stages of vertebrates. Similar to studies with pHAH-type AHR agonists like dioxin, our data also suggest that PAH-synergistic toxicity is, at least in part, AHR2 dependent. Specifically, AHR2 knockdown in zebrafish embryos significantly reduced toxicity when a PAH-type AHR agonist and CYP1A inducer (BNF) was combined with ANF, a competitive CYP1A inhibitor. However, the opposite effect was observed in zebrafish cyp1a morphants. That is, knock down of CYP1A protein synergized PAH effects in zebrafish, suggesting that CYP1A is an adaptive, protective response and confers protection with defense to the developing vertebrate embryo. This increased PAH teratogenicity is quite different from pHAH studies where CYP1A inhibition or knockdown either reduced toxicity (Cantrell et al., 1996; Dong et al., 2002; Wassenberg and Di Giulio, 2004a,b) or had no effect on dioxin-like symptoms in the developing embryo (Carney et al., 2004). Our data clearly indicate an alternate mechanism by which PAH-type AHR agonists cause developmental toxicity in fish.
Previously, we reported that chemical inhibition of CYP1A activity synergized cardiovascular toxicity of PAH-type AHR agonists (i.e., BaP and BNF) to embryonic stages of killifish (Wassenberg and Di Giulio, 2004a). Similar increases in PAH toxicity were observed when agonists were combined with a variety of CYP1A inhibitors (Wassenberg and Di Giulio, 2004a,b). From these data we concluded that increased toxicity of PAHs was a function of CYP1A inhibition, regardless of chemical properties or mechanism (Wassenberg and Di Giulio, 2004a). Our first objective in this study was to test whether synergistic toxicity observed in killifish could be repeated using the zebrafish developmental model. As expected, coexposure of zebrafish embryos to a nonteratogenic dose of ANF inhibited BNF-induced CYP1A activity in vivo and synergized BNF cardiovascular toxicity.
Relationships among AHR agonism, CYP1A metabolism, and toxicity of PAHs are complicated by rapid oxidation of these compounds by CYP1A enzymes. The few studies that have examined these interactions in fish early life stages have yielded somewhat contradictory results including synergism (this study), additivity (Hawkins et al., 2002), and antagonism (Dong et al., 2002). At odds with our studies, ANF protected zebrafish embryos from circulation failure in the dorsal midbrain following exposure to BNF (Dong et al., 2002). This could be due to differences between the studies with respect to dose and time response. Dong et al. (2002) looked at early-onset (50 hpf) midbrain circulation failure induced by a BNF concentration (272 µg/l) more than 2.5 orders of magnitude higher than the dose used in our studies, which by itself was nontoxic (Wassenberg and Di Giulio, 2004a). In addition, Carney et al. (2004) concluded that the earliest onset for dioxin toxicity in zebrafish is around day 3. Particularly at lower doses of dioxin, and we propose for less potent agonists like BNF, teratogenic end points are modest in zebrafish embryos prior to 72 hpf (Carney et al., 2004).
With respect to ANF, there is evidence to support that its mechanism of action varies by concentration (Aluru et al., 2005; Gasiewicz and Rucci, 1991; Santostefano et al., 1994). Qureshi (2004) observed that at high concentrations (320 µg/l), ANF eliminated symptoms of blue sac disease in rainbow trout larvae caused by exposure to retene and blocked induction of CYP1A protein. This was associated with complete inhibition of phase I metabolism of parent PAH, suggesting that ANF antagonized activation of AHR by retene (Qureshi, 2004). In contrast, lower concentrations of ANF (32–100 µg/l) caused a concentration-dependent increase in CYP1A protein and enhanced mortality and toxicity of retene to larval trout. Overall, lower ANF concentrations were associated with fewer retene metabolites but selective for less polar hydroxylated intermediates (Qureshi, 2004). This is consistent with our observations that inhibition of CYP1A activity is concomitant with BNF and ANF synergies in zebrafish (Wassenberg and Di Giulio, 2004a). Taken together, these data suggest that at lower concentrations, ANF by itself acts to increase CYP1A protein either by acting as a partial AHR agonist and/or by inhibiting CYP1A activity to increase its own half-life to extend the time that it can act as an AHR agonist. However, when combined with a strong AHR agonist such as BNF, its role may change to that of a competitive AHR antagonist or it may inhibit CYP1A and extend the half-life of the other PAH-type AHR agonists (Hawkins et al., 2002). While it is not entirely clear from our studies whether ANF functions as a receptor agonist or CYP1A inhibitor, either mode of action could lead to an increase in CYP1A protein.
Similar to chemical inhibition of CYP1A activity (Wassenberg and Di Giulio, 2004a), we observed that CYP1A knockdown made a nontoxic dose of BNF (10 µg/l) highly teratogenic and that the combined toxicity of BNF (1 µg/l) + ANF (100 µg/l) was also enhanced in cyp1a morphants even though EROD activity was only maximally reduced by half that of BNF-induced embryos. The transient nature of the morpholino approach could account for EROD activity at 96 hpf in CYP1A (and AHR2) knockdowns. Although the same morpholino concentrations were used, it is feasible that incomplete knockdown could be explained by differences in injection volume between this and other studies (e.g., Carney et al., 2004). The fact that we observed such striking effects even without complete knockdown speaks to the biological significance of our results. Collectively, these data show that CYP1A is playing an important role in mediating toxicity of PAHs in early life stages of fish. The fact that CYP1A knockdown mimics ANF effects on BNF also suggests that in our system ANF is acting more as a CYP1A inhibitor than as an AHR agonist or antagonist.
Several studies have clearly shown that pHAH-induced embryotoxicity is receptor mediated in fish (Carney et al., 2004; Prasch et al., 2003; Teraoka et al., 2003). Likewise, our data suggest that synergistic effects of PAH coexposures require functional AHR2. We show that AHR2 knockdown reduces synergistic cardiovascular toxicity in zebrafish embryos cotreated with BNF and ANF, specifically pericardial edema and retarded growth of the lower jaw. Although not directly tested in the current study, we would expect that AHR2 knockdown would block cardiotoxic effects observed in zebrafish exposed to higher embryotoxic doses of BNF alone. Incardona et al. (2004) demonstrated that exposure of zebrafish embryos to the three-ring PAHs abundant in weathered Alaska North Slope (ANS) crude oil induced a specific suite of developmental defects consistent with and secondary to disruption of atrioventricular conduction block. In a subsequent study, the authors speculated that the AHR pathway plays a protective or adaptive role against cardiovascular defects caused by exposure of zebrafish embryos to high concentrations (viz above solubility limits) of low–molecular weight tricyclic PAHs (Incardona et al., 2005). In that study, morpholino knock down of translated zebrafish AHR isoforms and CYP1A protein did not prevent toxicity associated with weathered crude oil and tricyclic PAHs, leading the authors to conclude that cardiotoxicity was independent of AHR activation or CYP1A induction (Incardona et al., 2005). Rather, cardiovascular dysfunction was exacerbated in zebrafish coinjected with ahr1 and ahr2 morpholinos or cyp1a zebrafish morphants exposed to weathered ANS oil. In contrast, we observed that synergistic response caused by coexposure to PAH-type AHR agonists and CYP1A inducers is mediated at least in part by AHR2 since knock down of the receptor protein reduced toxicity. It should be noted that tricyclic PAHs tested in Incardona et al. (2005) are at best weak AHR agonists and CYP1A inducers compared to more potent 4- to 5-ring, PAH-type AHR agonists that have been used as model compounds in our studies (Wassenberg and Di Giulio, 2004a; Wassenberg et al., 2005). Thus, it is likely that we could be looking at different mechanisms of cardiotoxicity that vary with PAH-AHR binding affinity. Similar to Incardona et al. (2005), however, our data do support a protective role of CYP1A activity in PAH toxicity.
Although the exact mechanisms by which PAHs cause developmental toxicity are unclear, it is evident that CYP1A activity can be beneficial in vivo. Thus, part of the difference in the relationships between CYP1A activity and pHAH versus PAH toxicity could be attributed to differential metabolism of these two classes of AHR agonists. Coplanar pHAHs, like dioxin, are recalcitrant to oxidative metabolism because they are poor substrates for the CYP1A enzyme (Brown et al., 2002; White et al., 1997). As such, toxicity of pHAHs is usually attributed to the parent compound, rather than its metabolites (Hankinson, 2005). By comparison, PAHs are excellent substrates for CYP1A metabolism in vertebrates (Brown et al., 2002; Hawkins et al., 2002; Niimi and Palazzo, 1986).
Our data are consistent with the hypothesis that the protective effects of CYP1A1 activity have ensured preservation of this gene in vertebrates for more than 400 million years (Nebert and Dieter, 2000; Uno et al., 2004a), and a review of the mammalian literature supports our conclusion that CYP1A activity is protective with respect to PAH-type AHR agonists. Consistent with knockdown studies in zebrafish, mice lacking a functional cyp1a1 gene are protected against dioxin-induced toxicity at high doses (Uno et al., 2004b). In contrast, cyp1a1 knockout mice are not protected from BaP-induced toxicity (Uno et al., 2004a). Cyp1a1 knockout mice orally dosed with BaP demonstrated significantly higher levels of BaP-DNA adducts than cyp1a1 wild-type mice (Uno et al., 2004a). In the same study, cyp1a1 wild-type mice were also better protected from BaP-induced immunotoxicity, anorexia, bone marrow hypocellularity, and lethality (Uno et al., 2004a). This is in stark contrast to the widely accepted cancer paradigm that CYP1A activity increases PAH toxicity. Important to note, however, is that the evolutionary pressure that may have selected for preservation of CYP1A activity probably did not apply to the development of cancer.
Certainly, caution should be exercised when comparing our studies with fish to the mammalian literature. One key difference is that in contrast to the emphasis on carcinogenic mechanisms of toxicity in mammalian studies of PAHs like BaP, cancer does not appear to be the cause of toxicity of PAHs to early life stages of fish. Rather, the important message to take home from recent mice knockout studies is that functional CYP1A1 appears to be important in detoxification and protection against PAH toxicity.
Other AHR-responsive genes may play a role in embryotoxicity of PAHs. CYP1B1 overlaps in function with CYP1A; however, it is generally thought to have a greater tendency to metabolize substrates to more toxic products than CYP1A. A recent study suggests that CYP1B1 in immune tissues governs metabolic activation and immunotoxicity of BaP when CYP1A1 is genetically ablated (Uno et al., 2006). CYP1B1 has been studied in flatfish (Pleuronectes platessa), carp (Cyprinus carpio), channel catfish (Ictalurus punctatus), and brown bullhead (Ameiurus nebulosus). In flatfish, CYP1B1 expression is restricted to the gills and, unlike CYP1A, is not induced by BNF (Leaver and George, 2000). CYP1B1 is, however, inducible by the AHR agonist 3-methylcholanthrene (3-MC) in carp tissues (El-kady et al., 2004a). With the exception of gill and blood, CYP1B1 is induced by BaP in channel catfish and brown bullhead in most tissues examined (Willett et al., 2003). A second form, CYP1B2, has also been isolated in carp gill and is inducible by 3-MC, although only in this location, and it is not constitutively expressed in any other carp tissues (El-kady et al., 2004b). No studies in zebrafish could be found, although the mRNA has been identified and sequenced (Godard et al., 2005). Two new P450 genes, cyp1c1 and cyp1c2, have recently been isolated from scup (Stenotomus chrysops) (Godard et al., 2005). The zebrafish homologues of these genes have also been identified, and sequence analysis shows that both contain at least one putative dioxin response element in the promoter regions (Godard et al., 2005). Homologues to either cyp1c genes were not identified in mammals, suggesting this is a fish-specific gene (Godard et al., 2005). Although the function of these proteins is currently unknown, sequence similarity suggests a close relationship with CYP1B1. While speculative, these novel AHR-regulated genes could also potentially play a role in PAH-synergistic toxicity observed in our study.
There are a number of questions that arise from the conjecture that PAH developmental toxicity is tied to metabolism. First, does inhibiting CYP1A convert PAHs to a more developmentally toxic form by extending the half-life of metabolically active PAHs so that they resemble dioxin-like compounds which are recalcitrant to CYP-mediated metabolism? Or, does PAH toxicity result from a shift in metabolism when CYP1A activity is blocked or inhibited? We are currently addressing these questions. Additionally, for years, CYP1A activity and induction has been utilized as a biomarker of exposure for both PAHs and pHAHs (Payne et al., 1987). While historically CYP1A activity has been associated with toxicity of pHAHs, increased CYP1A activity does not predict chronic PAH effects to vertebrates (Lee and Anderson, 2005).
Finally, these data create uncertainty as to whether models of risk assessment for PAHs, which assume additivity, are appropriate. Our studies suggest that a toxic equivalent (TEQ) model based on AHR agonism would underestimate the embryotoxic potential of PAH-dominated mixtures. Considering the complexity of environmental mixtures of PAHs, more effective and mechanistic-driven models to predict hazard to vertebrate species would appear to be in order.
|
NOTES |
|---|
1 These authors contributed equally to this work.
|
ACKNOWLEDGMENTS |
|---|
We thank the entire Di Giulio laboratory for their insight and helpful discussions during all phases of this project. We are grateful for the generous access provided to the zebrafish facility and resources of the Linney laboratory (Duke University), to Dr Elwood Linney, Lucia Upchurch, Sue Donerly, Drs Betsey Dobbs-McAuliffe, and Chris Lassiter. We acknowledge Dawoon Jung for excellent technical assistance with zebrafish imaging. This research was supported by the National Institute of Environemntal Health Sciences-supported Duke University Superfund Basic Research Center (P42 ES10356), Integrated Toxicology Program (T32 ES07031), and Environmental Protection Agency Star grant to A.R.T-.L.
|
REFERENCES |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
Aluru, N., Vuori, K., and Vijayan, M. M. (2005). Modulation of Ah receptor and CYP1A1 expression by
-naphthoflavone in rainbow trout hepatocytes. Comp. Biochem. Physiol. C Comp. Toxicol. Pharmacol. 141, 40–49.Andreasen, E. A., Spitsbergen, J. M., Tanguay, R. L., Stegeman, J. J., Heidemann, W., and Peterson, R. E. (2002). Tissue-specific expression of AHR2, ARNT2, and CYP1A in zebrafish embryos and larvae: Effects of developmental stage and 2,3,7,8-tetrachlorodibenzo-p-dioxin exposure. Toxicol. Sci. 68, 403–419.
Barron, M. G., Carls, M. G., Heintz, R., and Rice, S. D. (2004). Evaluation of fish early life-stage toxicity models of chronic embryonic exposures to complex polycyclic aromatic hydrocarbon mixtures. Toxicol. Sci. 78, 60–67.
Billiard, S. M., Querbach, K., and Hodson, P. V. (1999). Toxicity of retene to early life stages of two freshwater fish species. Environ. Toxicol. Chem. 18, 2070–2077.[CrossRef]
Brown, S. B., Fisk, A. T., Brown, M., Villella, M., Muir, D. C. G., Evans, R. E., Lockhart, W. L., Metner, D. A., and Cooley, H. M. (2002). Dietary accumulation and biochemical responses of juvenile rainbow trout (Oncorhynchus mykiss) to 3,3',4,4',5-pentachlorobiphenyl (PCB 126). Aquat. Toxicol. 59, 139–152.[Medline]
Cantrell, S. M., Lutz, L. H., Tillitt, D. E., and Hannink, M. (1996). Embryotoxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD): The embryonic vasculature is a physiological target for TCDD-induced DNA damage and apoptotic cell death in medaka (Orizias latipes). Toxicol. Appl. Pharmacol. 141, 23–34.[ISI][Medline]
Carney, S. A., Peterson, R. E., and Heideman, W. (2004). 2,3,7,8-Tetrachlorodibenzo-p-dioxin activation of the aryl hydrocarbon receptor/aryl hydrocarbon receptor nuclear translocator pathway causes developmental toxicity through a CYP1A-independent mechanism in zebrafish. Mol. Pharmacol. 66, 512–521.
Chen, C. M., and Cooper, K. R. (1999). Developmental toxicity and EROD induction in the Japanese medaka treated with dioxin congeners. Bull. Environ. Contam. Toxicol. 63, 423–429.[Medline]
Dabestani, R. T., and Ivanov, I. N. (1999). A compilation of physical, spectroscopic, and photophysical properties of polycyclic aromatic hydrocarbons. Photochem. Photobiol. 70, 10–34.
Dong, W., Teraoka, H., Tsujimoto, Y., Stegeman, J. J., and Hiraga, T. (2004). Role of aryl hydrocarbon receptor in mesencephalic circulation failure and apoptosis in zebrafish embryos exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. Sci. 77, 109–116.
Dong, W., Teraoka, H., Yamazaki, K., Tsukiyama, S., Imani, S., Imagawa, T., Stegeman, J. J., Peterson, R. E., and Hiraga, T. (2002). 2,3,7,8-Tetrachlorodibenzo-p-dioxin toxicity in the zebrafish embryo: Local circulation failure in the dorsal midbrain is associated with increased apoptosis. Toxicol. Sci. 69, 191–201.
Douben, P. E. T. (2003). Introduction. In PAHs: An Ecotoxicological Perspective (P. E. T. Douben, Ed.), pp. 3–6. John Wiley & Sons, West Sussex, England.
El-kady, M. A., Mitsuo, R., Kaminishi, Y., and Itakura, T. (2004a). cDNA cloning, sequence analysis and expression of 3-methylcholanthrene-inducible cytochrome P450 1B1 in carp (Cyprinus carpio). Environ. Sci. 11, 231–240.[Medline]
El-kady, M. A., Mitsuo, R., Kaminishi, Y., and Itakura, T. (2004b). Isolation of cDNA of novel cytochrome P450 1B gene, CYP1B2, from carp (Cyprinus carpio) and its induced expression in gills. Environ. Sci. 11, 345–354.[Medline]
Elonen, G. E., Spehar, R. L., Holcombe, G. W., Johnson, R. D., Fernandez, J. D., Erickson, R. J., Tietge, J. E., and Cook, P. M. (1998). Comparative toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin to seven freshwater fish species during early life-stage development. Environ. Toxicol. Chem. 17, 472–483.[CrossRef]
Fernandez-Salguero, P. M., Hilbert, D. M., Rudikoff, S., Ward, J. M., and Gonzalez, F. J. (1996). Aryl-hydrocarbon receptor-deficient mice are resistant to 2,3,7,8-tetrachlorodibenzo-p-dioxin-induced toxicity. Toxicol. Appl. Pharmacol. 140, 173–179.[CrossRef][ISI][Medline]
Gasiewicz, T. A., and Rucci, G. (1991). -Naphthoflavone acts as an antagonist of 2,3,7,8-tetrachlorodibenzo-p-dioxin by forming an inactive complex with the Ah receptor. Mol. Pharmacol. 40, 607–612.[Abstract]
Godard, A. J., Goldstone, J. V., Said, M., Dickerson, R. L., Woodin, B. R., and Stegeman, J. J. (2005). The new vertebrate CYP1C family: Cloning of new subfamily members and phylogenetic analysis. Biochem. Biophys. Res. Commun. 331, 1016–1034.[CrossRef][ISI][Medline]
Hahn, M. E. (1998). The aryl hydrocarbon receptor: A comparative perspective. Comp. Biochem. Physiol. C Pharmacol. Toxicol. Endocrinol. 121, 23–53.[CrossRef][Medline]
Hahn, M. E., Karchner, S. I., Shapiro, M. A., and Perera, S. A. (1997). Molecular evolution of two vertebrate aryl hydrocarbon (dioxin) receptors (AHR1 and AHR2) and the PAS family. Proc. Natl. Acad. Sci. U.S.A. 94, 13743–13748.
Hankinson, O. (2005). Role of coactivators in transcriptional activation by the aryl hydrocarbon receptor. Arch. Biochem. Biophys. 433, 379–386.[CrossRef][ISI][Medline]
Hawkins, S. A., Billiard, S. M., Tabash, S. P., Brown, S., and Hodson, P. V. (2002). Altering cytochrome P4501A activity affects polycyclic aromatic hydrocarbon metabolism and toxicity in rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem. 21, 1845–1853.[Medline]
Helder, T. (1981). Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on early life stages of rainbow trout (Salmo gairdneri, Richardson). Toxicology 19, 101–112.[CrossRef][ISI][Medline]
Incardona, J. P., Carls, M. G., Teraoka, H., Sloan, C. A., Collier, T. K., and Scholz, N. L. (2005). Aryl hydrocarbon receptor-independent toxicity of weathered crude oil during fish development. Environ. Health Perspect. 113, 1755–1762.[ISI][Medline]
Incardona, J. P., Collier, T. K., and Scholz, N. L. (2004). Defects in cardiac function precede morphological abnormalities in fish embryos exposed to polycyclic aromatic hydrocarbons. Toxicol. Appl. Pharmacol. 196, 191–205.[CrossRef][ISI][Medline]
Ivnitski-Steele, I. D., Friggens, M., Chavez, M., and Walker, M. K. (2005). 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) inhibition of coronary vasculogenesis is mediated, in part, by reduced responsiveness to endogenous angiogenic stimuli, including vascular endothelial growth factor A (VEGF-A). Birth Defects Res. A Clin. Mol. Teratol. 73, 440–446.[Medline]
Karchner, S. I., Franks, D. G., and Hahn, M. E. (2005). AHR1B, a new functional aryl hydrocarbon receptor in zebrafish: Tandem arrangement of ahr1b and ahr2 genes. Biochem. J. 392, 153–161.[CrossRef][ISI][Medline]
Karchner, S. I., Powell, W. H., and Hahn, M. E. (1999). Identification and functional characterization of two highly divergent aryl hydrocarbon receptors (AHR1 and AHR2) in the teleost Fundulus heteroclitus. J. Biol. Chem. 274, 33814–33824.
Latimer, J. S., and Zheng, J. (2003). The sources, transport, and fate of PAHs in the marine environment. In PAHs: An Ecotoxicological Perspective (P. E. T. Douben, Ed.). John Wiley & Sons, West Sussex, England.
Leaver, M. J., and George, S. G. (2000). A cytochrome P4501B gene from a fish, Pleuronectes platessa. Gene 256, 83–91.[CrossRef][ISI][Medline]
Lee, R. F., and Anderson, J. W. (2005). Significance of cytochrome P450 system responses and levels of bile fluorescent aromatic compounds in marine wildlife following oil spills. Mar. Pollut. Bull. 50, 705–723.[Medline]
Nacci, D., Coiro, L., Wassenberg, D., and Di Giulio, R. (2005). A non-destructive technique to measure cytochrome P4501A enzyme activity in living embryos of the estuarine fish Fundulus heteroclitus. In Techniques in Aquatic Toxicology (G. K. Ostrander, Ed.), Vol. 2, pp. 209–225. CRC Press, Boca Raton, FL.
Nasevicius, A., and Ekker, S. (2000). Effective targeted gene "knockdown" in zebrafish. Nat. Genet. 26, 216–220.[CrossRef][ISI][Medline]
Nebert, D. W., and Dieter, M. Z. (2000). The evolution of drug metabolism. Pharmacology 61, 124–135.[CrossRef][ISI][Medline]
Nebert, D. W., Roe, A. L., Dieter, M. Z., Solis, W. A., Yang, Y., and Dalton, T. P. (2000). Role of the aromatic hydrocarbon receptor and [Ah] gene battery in the oxidative stress response, cell cycle control, and apoptosis. Biochem. Pharmacol. 59, 65–85.[CrossRef][ISI][Medline]
Niimi, A. J., and Palazzo, V. (1986). Biological half-lives of eight polycyclic aromatic hydrocarbons (PAHs) in rainbow trout (Salmo gairdneri). Water Res. 20, 503–507.[CrossRef]
Payne, J. F., Fancey, L. L., Rahimtula, A. D., and Porter, E. L. (1987). Review and perspective on the use of mixed-function oxygenase enzymes in biological monitoring. Comp. Biochem. Physiol. C Pharmacol. Toxicol. Endocrinol. 86, 233–245.
Peterson, R. E., Theobald, H. M., and Kimmel, G. L. (1993). Developmental and reproductive toxicity of dioxins and related compounds: Cross-species comparisons. Crit. Rev. Toxicol. 23, 283–335.[ISI][Medline]
Prasch, A. L., Teraoka, H., Carney, S. A., Dong, W., Hiraga, T., Stegeman, J. J., Heideman, W., and Peterson, R. E. (2003). Aryl hydrocarbon receptor 2 mediates 2,3,7,8-tetrachlorodibenzo-p dioxin developmental toxicity in zebrafish. Toxicol. Sci. 76, 138–150.
Qureshi, P. (2004). Mechanisms of retene toxicity in rainbow trout (Oncorhynchuis mykiss). Undergraduate Thesis, School of Environmental Studies and Department of Biology, Queen's University, Kingston, ON.
Santostefano, S., Liu, H., Wang, X., Chaloupka, K., and Safe, S. (1994). Effect of ligand structure on formation and DNA binding properties of the transformed rat cytosolic aryl hydrocarbon receptor. Chem. Res. Toxicol. 7, 544–550.[Medline]
Schmidt, J. V., and Bradfield, C. A. (1996). Ah receptor signaling pathways. Annu. Rev. Cell Dev. Biol. 12, 55–89.[CrossRef][ISI][Medline]
Teraoka, H., Dong, W., Tsujimoto, Y., Iwasa, H., Endoh, D., Ueno, N., Stegeman, J. J., Peterson, R. E., and Hiraga, T. (2003). Induction of cytochrome P450 1A is required for circulation failure and edema by 2,3,7,8-tetrachlorodibenzo-p-dioxin in zebrafish. Biochem. Biophys. Res. Commun. 304, 223–228.[CrossRef][ISI][Medline]
Thackaberry, E. A., Nunez, B. A., Ivnitski-Steele, I. D., Friggins, M., and Walker, M. K. (2005). Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on murine heart development: Alteration in fetal and postnatal cardiac growth, and postnatal cardiac chronotropy. Toxicol. Sci. 88, 242–249.
Toomey, B. H., Bello, S., Hahn, M. E., Cantrell, S., Wright, P., Tillett, D. E., and Di Giulio, R. T. (2001). 2,3,7,8-Tetrachlorodibenzo-p-dioxin induces apoptotic cell death and cytochrome P4501A expression in developing Fundulus heteroclitus embryos. Aquat. Toxicol. 53, 127–138.[CrossRef][ISI][Medline]
Uno, S., Dalton, T. P., Derkenne, S., Curran, C. P., Miller, M. L., Shertzer, H. G., and Nebert, D. W. (2004a). Oral exposure to benzo[a]pyrene in the mouse: Detoxication by inducible cytochrome P450 is more important than metabolic activation. Mol. Pharmacol. 65, 1225–1237.
Uno, S., Dalton, T. P., Dragin, N., Curran, C. P., Derkenne, S., Miller, M. L., Shertzer, H. G., Gonzalez, F. J., and Nebert, D. W. (2006). Oral benzo[a]pyrene in Cyp1 knockout mouse lines: CYP1A1 important in detoxication, CYP1B1 metabolism required for immune damage independent of total-body burden and clearance rate. Mol. Pharmacol. 69, 1103–1114.
Uno, S., Dalton, T. P., Sinclair, P. R., Gorman, N., Wang, B., Smith, A. G., Miller, M. L., Shertzer, H. G., and Nebert, D. W. (2004b). Cyp1a1(–/–) male mice: Protection against high-dose TCDD-induced lethality and wasting syndrome, and resistance to intrahepatocyte lipid accumulation and uroporphyria. Toxicol. Appl. Pharmacol. 196, 410–421.[CrossRef][ISI][Medline]
Van Metre, P. C., Mahler, B. J., and Furlong, E. T. (2000). Urban sprawl leaves its PAH signature. Environ. Sci. Technol. 34, 4064–4070.[CrossRef]
Walker, M. K., and Peterson, R. E. (1991). Potencies of polychlorinated dibenzo-p-dioxin, dibenzofuran, and biphenyl congeners, relative to 2,3,7,8-tetrachlorodibenzo-p-dioxin, for producing early life stage mortality in rainbow trout (Oncorhynchus mykiss). Aquat. Toxicol. 21, 219–238.[CrossRef]
Wannemacher, R., Rebstock, A., Kulzer, E., Schrenk, D., and Bock, K. W. (1992). Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on reproduction and oogenesis in zebrafish (Branchydanio rerio). Chemosphere 24, 1361–1368.[CrossRef]
Wassenberg, D. M., and Di Giulio, R. T. (2004a). Synergistic embryotoxicity of polycyclic aromatic hydrocarbon aryl hydrocarbon receptor agonists with cytochrome P4501A inhibitors in Fundulus heteroclitus. Environ. Health Perspect. 112, 1658–1664.[ISI][Medline]
Wassenberg, D. M., and Di Giulio, R. T. (2004b). Teratogenesis in Fundulus heteroclitus embryos exposed to a creosote-contaminated sediment extract and CYP1A inhibitors. Mar. Environ. Res. 58, 163–168.[Medline]
Wassenberg, D. M., Nerlinger, A. L., Battle, L. P., and Giulio, R. T. D. (2005). Effects of the polycyclic aromatic hydrocarbon heterocycles, carbazole and dibenzothiophene, on in vivo and in vitro CYP1A activity and polycyclic aromatic hydrocarbon-derived embryonic deformities. Environ. Toxicol. Chem. 24, 2526–2532.[CrossRef][ISI][Medline]
White, R. D., Shea, D., and Stegeman, J. J. (1997). Metabolism of the aryl hydrocarbon receptor agonist 3,3',4,4'- tetrachlorobiphenyl by the marine fish scup (Stenotomus chrysops) in vivo and in vitro. Drug Metab. Dispos. 25, 564–572.
Whitlock, J. P. (1999). Induction of cytochrome P4501A1. Annu. Rev. Pharmacol. Toxicol. 39, 103–125.[CrossRef][ISI][Medline]
Willett, K., Butala, H., Metzger, C., Quiniou, S., and Waldbieser, G. (2003). BaP effects on CYP1B-like mRNA expression and estrogen metabolism in two catfish species. SETAC 24th Annual Meeting Abstract. Austin, TX, p. 89
Willett, K. L., Wassenberg, D., Lienesch, L., Reichert, W., and Di Giulio, R. T. (2001). In vivo and in vitro inhibition of CYP1A-dependent activity in Fundulus heteroclitus</