Ah receptor ligands

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9.3.5. Ah Receptor Ligands

A wide variety of structurally diverse anthropogenic and natural chemicals are capable of interacting with the AhR. These chemicals also have a broad range of potencies at inducing dioxin-like effects in experimental systems. One of the major differences between the anthropogenic chemicals included in the TEF methodology and the natural AhR ligands is their pharmacokinetics. The anthropogenic chemicals included in the TEF methodology are persistent and bioaccumulate in wildlife and humans. In contrast, most if not all of the natural AhR ligands are rapidly metabolized and eliminated from biological systems. The following section will examine the differences between the chemicals included in the TEF methodology and remaining AhR ligands not included in this approach.[1] Industrial/Synthetic AhR Ligands

The synthetic compounds that bind to AhR include a number of different classes of chemicals, most notably the PCDDs, PCDFs, and PCBs. Other synthetic AhR ligands include industrial chemicals (polybrominated biphenyls, polychlorinated napthalenes, chlorinated paraffins, etc.), pesticides (hexachlorobenzene), and contaminants (polybrominated dioxins, dibenzofurans, and napthalenes) associated with various manufacturing, production, combustion, and waste disposal processes. In addition, pyrolysis of organic material can produce a number of non-halogenated polycyclic aromatic hydrocarbons (PAHs) with moderate to high affinity for AhR (Poland and Knudson, 1982; Nebert, 1989; Chaloupka et al., 1993).

Not all of the anthropogenic sources of dioxin-like compounds are included in the TEF methodology. Many of these chemicals, such as hexachlorobenzene and the brominated diphenyl ethers, are only weakly dioxin-like and have significant toxicological effects that are not mediated by the AhR. For these chemicals, it is not clear that adding them to the TEF methodology would decrease the uncertainty in the risk assessment process. For other classes of chemicals, such as the chlorinated napthalenes, environmental concentrations and human exposures are largely uncertain.

The PAHs are one class of anthropogenic chemicals not included in the TEF scheme despite evidence for AhR binding. The PAHs are not included in the TEF methodology because of their short half-lives and relatively weak AhR activity. In addition, the role of the Ah receptor in the toxicity of the PAHs is uncertain. For example, both benzo[a]pyrene and chrysene induce CYP1A1 activity through an AhR-mediated mechanism (Silkworth et al., 1995). However, while the Ah receptor also plays a role in the immune suppressive effects of benzo[a]pyrene it does not appear to be involved in the immune suppression induced by chrysene (Silkworth et al., 1995). Furthermore, PAHs are DNA reactive and mutagenic and these mechanisms play a large role in both the carcinogenicity and immunotoxicity of the PAHs (Ross and Nesnow, 1999). In contrast, TCDD and other dioxin-like compounds are not DNA reactive. While there are PAHs that bind to the AhR, the role of AhR or other competing pathways in the toxicity of these compounds has not been clearly defined.

Brominated dioxins, dibenzofurans, biphenyls, and napthalenes also induce dioxin-like effects in experimental animals (Miller and Birnbaum, 1986; Zacherewski et al., 1988; Birnbaum et al., 1991; Hornung et al., 1996; DeVito et al., 1997; Weber and Greim, 1997). The brominated dioxins and dibenzofurans may be more or less potent than their chlorinated orthologues, depending on the congener (Birnbaum et al., 1991; DeVito et al., 1997). The sources of the brominated dioxin-like compounds are not well characterized. Some of the chemicals, such as the brominated biphenyls and their contaminants the brominated naphthalenes, were synthesized and sold as commercial flame retardants. The manufacture and use of polybrominated biphenyls has been prohibited. Brominated dibenzofurans are produced as byproducts of synthesis and pyrolysis of some brominated flame retardants. There is some evidence of human exposure to brominated dioxins and dibenzofurans from extruder operators (Ott and Zober, 1996). Polybrominated, polychlorinated, and mixed bromo- and chloro- dioxins and dibenzofurans have been found in soot from textile processing plants (Sedlak et al., 1998). Although these chemicals have been found in humans, these studies are limited to a small population and exposure to the general population remains undetermined. Future examinations of the TEF methodology should include a more detailed discussion of the brominated dioxins and dibenzofurans. Naturally Occurring AhR Ligands

The evolutionary conservation of AhR and its biological function following activation by dioxin-like compounds have led to the hypothesis that there must be an endogenous or physiological ligand(s) for this receptor. Presently, the endogenous ligand remains undetermined. However, efforts to discover the natural ligand have led to the discovery of a number of naturally occurring AhR ligands. A number of naturally occurring chemicals present in the diet are capable of binding to AhR and inducing some dioxin-like effects in experimental animals (Bradfield and Bjeldanes, 1984; 1987) and humans (Michnovicz and Bradlow, 1991; Sinha et al., 1994). The question of how the interaction of these chemicals relates to the toxicity of those chemicals designated as dioxin-like has become the subject of much debate.

One class of naturally occurring chemicals that activate the AhR is the indole derivatives. Indole derivatives, naturally present in a variety of cruciferous vegetables, are capable of modulating the carcinogenicity of PAHs (Wattenberg and Loub, 1978). Indole-3-carbinol (I-3-C) and 3,3'-diindolylmethane (DIM) are major secondary metabolites found in cruciferous vegetables and induce both phase I and II metabolic enzymes (CYP1A-dependent glutathione and glucuronyl transferases, oxidoreductases) in experimental animals (Bradfield and Bjeldanes, 1984, 1987), human cell lines (Bjeldanes et al., 1991; Kleman et al., 1994), and humans (Michnovich and Bradlow, 1990, 1991). Although both compounds induce CYP450 enzymes under AhR transcriptional control, they exhibit relatively low binding affinity for the Ah receptor (Gillner et al., 1985). Further investigation revealed that I-3-C is relatively unstable in the acidic environment of the digestive tract and readily forms DIM. In turn, DIM can participate in acid condensation reactions to form indolocarbazoles (ICZs) (Chen et al., 1995). ICZs are also produced by bacterial metabolism of the common dietary amino acid tryptophan. ICZs, in particular indolo[3,2b]carbazole, exhibit high binding affinity for the rodent AhR, approximately equipotent to 2,3,7,8-tetrachlorodibenzofuran, and can induce CYP1A1 activity in cultured cells (Bjeldanes et al., 1991; Gillner et al., 1993; Chen et al., 1995). ICZ and a methylated derivative, 5,11-dimethylindolo[3,2b]carbazole (MICZ), are also capable of binding to and activating the AhR in human hepatoma cells (HepG2) (Kleman et al., 1994). With considerably lower efficacy, I-3-C and DIM can partially displace TCDD from the AhR from human breast cancer cells (T47D) (Chen et al., 1996). These results would suggest that this group of compounds may represent a class of physiologically active AhR ligands derived from natural sources, which could either mimic dioxin-like compounds in their action or act as competitors for AhR binding.

In addition to the plant-derived indoles, experimental animals consuming thermally treated meat protein as well as humans fed cooked meat can exhibit induced CYP1A2 activity (Degawa et al., 1989). High-temperature cooking (250"C, 22 minutes) of ground beef resulted in the formation of a number of heterocyclic aromatic amines (HAAs) in part-per-billion levels, which were thought to be responsible for the observed CYP1A2 induction in human volunteers (Sinha et al., 1994). Mechanistic analysis of one particular HAA, 2-amino-3,8-dimethylimidazo[4,5-f]quinoxaline (MeIQx), has shown that it is capable of both interacting with the AhR and inducing CYP1A1/A2 activity in rats (Kleman and Gustafsson, 1996). These data should be viewed cautiously because recent data indicate that CYP1A2 can be induced through non-AhR mechanisms (Ryu et al., 1996). Because there are multiple pathways to induce CYP1A2, the increase in CYP1A2 activity following exposure to complex mixtures, such as cooked meat, does not necessarily indicate the presence of dioxin-like compounds.

Other diet-derived chemicals that can interact with the AhR include oxidized essential amino acids. UV-oxidized tryptophan is capable of inducing CYP1A1 activity in mouse hepatoma cells through an AhR-dependent mechanism (Sindhu et al., 1996). Rats exposed to UV-oxidized tryptophan in vivo also exhibited induction of hepatic and pulmonary CYP1A1 activity. Both in vitro and in vivo enzyme induction were transient, with the oxidized tryptophan possibly being metabolized by CYP1A1 (Sindhu et al., 1996). Tryptanthrins, biosynthetic compounds produced from the metabolism of tryptophan and anthranilic acid by yeast commonly found in food, are agonists for the rat AhR (Schrenk et al., 1997). Various tryptanthrins also induce CYP1A1-related enzyme activity in mouse hepatoma cells with the approximate efficacy of indolo[3,2b]carbazole.

Recent studies have demonstrated that physiological chemicals can bind to the AhR. Bilirubin was recently found to transform the AhR from mouse hepatoma cells into its DNA-binding state, resulting in CYP1A1 induction. Hemin and biliverdin can also be metabolically converted to bilirubin, resulting in AhR-dependent gene activation (Sinal and Bend, 1997). Despite these results, there is no clear evidence that these are the physiological ligands for the

AhR, nor is there evidence that these compounds can modulate the activity of dioxin-like compounds or lead to dioxin-like toxic effects in humans or animals. Industrial vs. Natural AhR Ligands

There are a number of structurally diverse chemicals that bind to the Ah receptor. Some of these chemicals are industrial chemicals produced intentionally (PCBs, PBBs, etc.). Others are by-products of industrial processes (PCDDs and PCDFs). There are also a number of "natural" AhR ligands that are either plant derived (i.e. I-3-C) or are synthesized endogenously, such as bilirubin. It has been postulated that the natural ligands could be the major contributors to the daily dose of TEQs, because of their higher estimated intakes (Safe, 1995). The natural ligands tend to have short half-lives and do not accumulate. The PCDDs/PCDFs and PCBs included in the TEF methodology clearly bioaccumulate. If contributions to the total TEQ are estimated on steady-state body burdens of these chemicals instead of daily intake, then TCDD and other PCDDs/PCDFs and PCBs contribute more than 90% of the total TEQ compared to the natural ligands (DeVito and Birnbaum, 1996). The difference in the results of these analyses demonstrates our uncertainty of the relative potencies, exposures and dose metrics used in the comparisons of the synthetic dioxins vs. the natural AhR ligands.

When a comparison is attempted between the perceived relative risk from natural vs. anthropogenic AhR agonists, a number of factors should be taken into consideration. The potency of AhR ligands depends on several factors, including AhR binding affinity and pharmacokinetic properties. When estimating the relative potency of a chemical compared to TCDD, the larger the difference in pharmacokinetic properties, the greater the effect that study design has on the relative potency. Initial studies comparing the potency of indolo[3,2b]carbazole to TCDD demonstrate the importance of the pharmacokinetic differences between these chemicals. In Hepa-1 cells exposed for 4 hours, the relative potency for induction of CYP1A1 mRNA of indolo[3,2b]carbazole compared to TCDD is 0.1 (Chen et al., 1995). If the relative potency is determined after 24 hours of exposure, the potency of indolo[3,2b]carbazole drops 1,000-fold to 0.0001 (Chen et al., 1995). In addition, the dioxin-like effects of low doses of indolo[3,2b]carbazole in Hepa-1 cells are transient. Similar transient effects of other dietary-derived AhR ligands have also been reported (Xu and Bresnick, 1990; Berghard et al., 1992; Riddick et al., 1994). These data demonstrate that the relative potencies of these chemicals compared to TCDD are dependent upon the pharmacokinetic properties of the chemicals and the experimental design used in the comparisons. In addition, these data also demonstrate that for rapidly metabolizable AhR ligands, the effects are transitory and not persistent like TCDD. It appears that the transient nature of the effect is due to the transient concentrations of these chemicals in these experimental systems. These data also demonstrate our uncertainty of the relative potency of the dietary-derived AhR ligands.

The chemicals included in the TEF scheme are those that not only bind to AhR but also bioaccumulate and have long biological half-lives in humans, typically on the order of years. In contrast, the pharmacokinetics of the endogenous or natural group are not well studied, but these chemicals tend to be short-lived, with half-lives on the order of minutes to hours. Although both PAHs and the halogenated aromatics bind to AhR and induce cytochrome P450-related enzyme activities, only the latter group produces the additional dioxin-like spectrum of toxicological responses. These toxicities are thought to be due to the persistent exposures attributable to the long half-lives of these chemicals (Riddick et al., 1994).

One caution when comparing the relative exposures to dietary AhR ligands and the anthropogenic AhR ligands is that few in vivo studies have examined the relative potency of the dietary or natural AhR ligands for toxic responses. Using the criteria of the WHO workgroup for PCDDs, PCDFs, and PCBs results in only two in vivo studies of I-3-C which compared the REP to TCDD (Wilker et al., 1996; Bjeldanes et al., 1991). In an in vivo study of the developmental effects of I-3-C, in utero exposure of rats to I-3-C resulted in a number of reproduction-related abnormalities in male offspring, only some of which resemble those induced by TCDD (Wilker et al., 1996). Because of the different spectrum of effects of I-3-C compared to TCDD in these developmental studies, it is likely that mechanisms other than AhR activation are involved in these effects. I-3-C and some of its acid catalyzed oligomerization products alter androgen and estrogen metabolism (Wilson et al., 1999; Telang et al., 1997), which may contribute to the developmental effects of I-3-C. While a number of in vitro studies have demonstrated dioxin-like enzyme induction of the indole derivatives, in order to have REP values that adequately describe the in vivo potency of these chemicals, future in vivo studies examining toxic responses are required. Limitations in Comparing the Quantitative Interactions between Industrial/Synthetic and Natural AhR Ligands

Although there are limited data on the in vivo biochemical and toxicological effects of these ligands, the effects of mixtures of anthropogenic and natural AhR ligands is altogether lacking. There is one study examining the interactions of I-3-C and DIM on the effects of TCDD in cell culture systems. However, it is uncertain how to extrapolate these in vitro concentrations to present human in vivo exposures. The limited data available do not adequately address the interactions between these chemicals. Future in vivo studies are required in order to better understand the potential interactions between these classes of AhR ligands.

Another limitation in comparing the natural AhR ligands to the dioxins is the multiple effects induced by the natural AhR ligands. In vivo and in vitro studies of I-3-C indicate that it induces a number of biochemical alterations that are not mediated through the AhR (Broadbent and Broadbent, 1998). The activation of these additional pathways creates difficulties in making direct comparisons with TCDD and related chemicals. Similarly, the PAHs also have non-AhR-mediated biochemical and toxicological effects that also complicate direct comparisons with TCDD and related dioxins. For example, co-exposure to TCDD and PAHs have demonstrated both synergistic and antagonistic interactions in mice depending upon the endpoint examined (Silkworth et al., 1993).

Presently, there are several limitations in our understanding of the importance of naturally occurring dioxin-like compounds vs. the dioxin-like compounds included in the TEF methodology. First is the limited data available on the dioxin-like toxicities of the natural ligands. In addition, there is a lack of data on the interactions between these classes of chemicals. Few if any mixtures of natural AhR ligands and PCDDs or PCDFs examining a toxic response have been published. Many of the natural AhR ligands have multiple mechanisms of action that presently cannot be accounted for in the TEF methodology. For example, I-3-C has anticarcinogenic properties in tumor promotion studies, and these effects may or may not be mediated through AhR mechanisms (Manson et al., 1998). The lack of data and the role of non-AhR mechanisms in the biological effects of these chemicals prohibit a definitive conclusion on the role of natural vs. anthropogenic dioxins in human health risk assessment. Though it is important to address these issues, the available data do not lend themselves to an appropriate quantitative assessment of these issues.

One of the most significant differences between the industrial Ah receptor ligands (i.e. dioxins) and the natural Ah receptor ligands is the persistence of the dioxins in biological systems. Because of their long half-lives, dioxins provide a persistent activation of the Ah receptor. In contrast, the natural ligands are rapidly metabolized and the activation of the Ah receptor is short-lived. Determining the relative potency of the natural ligands compared to TCDD is not necessarily a trivial matter. The relative potency of these chemicals is due to their ability to bind and activate the Ah receptor and the persistence of this signal. Most of the studies examining the relative potency of the natural ligands are based on in vitro or short-term in vivo studies. The estimates of the relative potencies of these chemicals is greatly exaggerated in these short-term assays because of the bioaccumulative nature of TCDD. Studies comparing the relative potency of TCDD to TCDF have demonstrated that due to the differences in the half-lives of TCDF and TCDD, short-term studies overestimate the relative potency of TCDF compared to the relative potency observed in longer-term studies (DeVito and Birnbaum, 1995). The relative potencies of the natural ligands would best be estimated following long term exposures. These data are unavailable and thus the estimates of the relative potencies of these chemicals is unreliable.

Although Safe has suggested that exposure to natural AhR ligands is 100 times that of TCDD and other dioxin-like compounds (Safe, 1995), the impact of the natural AhR ligands remains uncertain. Epidemiological studies suggest that human exposures to TCDD and related chemicals are associated with adverse effects, such as developmental impacts and cancer. In many of these studies, the exposed populations have approximatley 100 times more TCDD exposure than background populations (see Part II, Chapter 7). If the exposure to natural AhR ligands is included in these comparisons, then the exposed populations should have approximatley double the total TEQ exposures than the background population. It seems unlikely that epidemiological studies could discriminate between such exposures. These data suggest that the estimates of the contribution of the natural AhR ligands to the total TEQ exposure are overestimated. In addition, regardless of the background human exposure to "natural" AhR ligands, the margin of exposure to TCDD and related chemicals between the background population and populations where effects are observed remains a concern.

See also


  1. U.S.EPA (2003): Toxic Equivalency Factors (TEF) for Dioxin and Related Compounds. In: Exposure and Human Health Reassessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and Related Compounds. Part II: Health Assessment for 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and Related Compounds. Chapter 9. NAS Review Draft NCEA-I-0836. December 2003. www.epa.gov/ncea.
    DISCLAIMER This document is a draft. It has not been formally released by the U.S. Environmental Protection Agency and should not at this stage be construed to represent Agency policy. Mention of trade names or commercial products does not constitute endorsement or recommendation for use.