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Applying Mode-of-Action and Pharmacokinetic Considerations in Contemporary Cancer Risk Assessments: An Example With Trichloroethylene

Posted on: Sunday, 31 October 2004, 03:00 CST

ABSTRACT: The guidelines for carcinogen risk assessment recently proposed by the U.S. Environmental Protection Agency (U.S. EPA) provide an increased opportunity for the consideration of pharmacokinetic and mechanistic data in the risk assessment process. However, the greater flexibility of the new guidelines can also make their actual implementation for a particular chemical highly problematic. To illuminate the process of performing a cancer risk assessment under the new guidelines, the rationale for a state-of- the-science risk assessment for trichloroethylene (TCE) is presented. For TCE, there is evidence of increased cell proliferation due to receptor interaction or cytotoxicity in every instance in which tumors are observed, and most tumors represent an increase in the incidence of a commonly observed, species-specific lesion. A physiologically based pharmacokinetic (PBPK) model was applied to estimate target tissue doses for the three principal animal tumors associated with TCE exposure: liver, lung, and kidney. The lowest points of departure (lower bound estimates of the exposure associated with 10% tumor incidence) for lifetime human exposure to TCE were obtained for mouse liver tumors, assuming a mode of action primarily involving the mitogenicity of the metabolite trichloroacetic acid (TCA). The associated linear unit risk estimates for mouse liver tumors are 1.5 10^sup -6^ for lifetime exposure to 1 g TCE per cubic meter in air and 0.4 10^sup - 6^ for lifetime exposure to 1 g TCE per liter in drinking water. However, these risk estimates ignore the evidence that the human is likely to be much less responsive than the mouse to the carcinogenic effects of TCA in the liver and that the carcinogenic effects of TCE are unlikely to occur at low environmental exposures. Based on consideration of the most plausible carcinogenic modes of action of TCE, a margin-of-exposure (MOE) approach would appear to be more appropriate. Applying an MOE of 1000, environmental exposures below 66 g TCE per cubic meter in air and 265 g TCE per liter in drinking water are considered unlikely to present a carcinogenic hazard to human health.

KEYWORDS: Cancer risk assessment, Mode-of-action, Pharmacokinetics, Trichloroethylene

I. INTRODUCTION

Assessing the potential risk associated with human exposure to carcinogenic environmental contaminants represents an uncomfortable admixture of scientific evaluation and political policy, with the potential for enormous impact on both the public health and the economic well-being of the nation. A difficult challenge facing cancer risk assessors today is to realistically consider the implications of the chemical's mechanism(s) of carcinogenicity in developing a risk assessment approach for a particular carcinogenic effect. It has become increasingly difficult to justify the use of the same linear risk assessment approach both with chemicals that act through a purely radiomimetic, genotoxic mechanism, and with chemicals for which carcinogenicity is mediated by an inherently nonlinear process, such as increased cell proliferation secondary to cytotoxicity or receptor interaction. Mechanismdependent risk assessment approaches are the only alternative for maintaining the credibility of cancer potency estimates in the face of increasing sophistication in the understanding of the mechanisms of carcinogenicity.

The guidelines for carcinogen risk assessment recently proposed by the U.S. Environmental Protection Agency (U.S. EPA)1 would appear to provide the flexibility necessary to move forward in this area.2 Under the new guidelines, multiple options are available for performing a carcinogenic dose-response assessment, ranging from the use of a biologically based dose-response model to the use of linear or margin-of-exposure (MOE) default approaches. The selection of the dose-response approach to be used with a particular chemical is determined on the basis of the information available on the chemical, which considers both pharmacokinetic and mechanistic data. Significantly, the new guidelines depart from the usual definition of a default used in the past. Under the new guidelines, a default is defined as the "no-information" option, the use of which must be justified by the agency on the basis of the lack of sufficient information on a specific chemical to support a more preferred, chemical-specific approach. This definition stands in contrast to past practice, in which the default position was treated as the preferred approach, and justification was required for departing from it on the basis of chemical-specific information.

Risk assessments for chemical carcinogens must necessarily be iterative in nature. It is in the nature of scientific inquiry that understanding develops slowly, as experimental information accumulates and theories can be tested and refined. Risk assessments, however, cannot be postponed indefinitely until an adequate understanding of the carcinogenicity of a particular chemical has been achieved. Therefore, it is necessary to attempt to perform the most scientifically defensible assessment possible, given the information available at that time, and to be ready to revise the estimate, periodically, whenever important new information is developed. In the last few years there has been a significant improvement in the level of understanding regarding chemical carcinogenesis in general and the mechanisms of carcinogenicity of trichloroethylene (TCE) in particular. The purpose of the study reported here was to attempt to perform a state- ofthe-science risk assessment for TCE in the spirit of the new U.S. EPA cancer guidelines, using to as great an extent as possible the information currently available on pharmacokinetics and carcinogenic mode of action. An attempt will be made to provide a fairly thorough discussion of the available data and the decision criteria associated with each step in the risk assessment process for TCE, in the hope of illuminating some of the issues and considerations that must be addressed in the implementation of the new U.S. EPA cancer guidelines for any chemical.

A. Consideration of Pharmacokinetics and Metabolism

As is described later, the primary target tissues for the carcinogenicity of TCE identified in animal studies are the liver, lung, and kidney. For each of these target tissues there is evidence that the carcinogenicity of TCE may be associated with one or more of its metabolites: trichloroacetic acid (TCA) and dichloroacetic acid (DCA) in the liver,3,4 chloral (CHL) in the lung,5 and 1,2- dichlorovinylcysteine (1,2-DCVC) in the kidney.6 Thus a comprehensive cancer risk assessment for TCE should be based on an analysis of tissue dosimetry for all three target tissues, including a description of the kinetics of the metabolites imputed to play a role in the carcinogenic activity.

A powerful technique for performing tissue dosimetry is physiologically based pharmacokinetic (PBPK) modeling. Briefly, PBPK modeling attempts to describe the relationship between external measures of applied dose (e.g., amount administered or concentration in food, water, or air) and internal measures of biologically effective dose (e.g., amount metabolized or concentration in the tissue displaying the toxic response), using as realistic a description of mammalian physiology and biochemistry as is necessary and feasible.7-10 Thus nonlinear biochemical processes can be incorporated into the model and the behavior of the animal-chemical system can be predicted for different routes of exposure over a wide range of exposure conditions.

The ability of PBPK modeling to support crossspecies dosimetry is of particular importance for risk assessment. The physiological and biochemical parameters in the model can be changed from those for the test species to those that are appropriate for humans to provide a biologically meaningful animal-to-human extrapolation. The ultimate aim of using PBPK modeling in risk assessment is to provide a measure of dose that better represents the "biologically effective dose"-that is, the dose that causally relates to the toxic outcome. The improved dose metric can then be used in place of traditional dose metrics (such as total amount absorbed) in the appropriate dose- response model to provide a more accurate extrapolation to human exposure conditions. Advantages of applying PBPK modeling in risk assessment have been discussed for both cancer11-15 and noncancer endpoints.16-18 In addition, the use of PBPK modeling has been recommended to improve route-to-route extrapolation19,20 and the estimation of risk for chemical mixtures.21

Simple pharmacokinetic dosimetry approaches have previously been used by regulatory agencies in cancer risk assessment; for example, the U.S. EPA made use of estimates of total metabolized dose in its cancer risk assessments for TCE.22,23 However, the first case where an agency used a full PBPK approach in a published chemical risk assessment was in the U.S. EPA's latest revision of its inhalation risk assessment for methylene chloride.24 In 1989, after a detailed multiagency evaluation of the available PBPK information and a review by the U.S. EPA Scientific Advisory Board, the U.S. EPA revised the inhalation unit risk and risk-specific air concentrations for methylene chloride i\n its Integrated Risk Information System (IRIS) database,25 citing the PBPK model of Andersen et al.26 The resulting risk estimates were lower than those obtained by the default approach by nearly a factor of 10. Application of the PBPK model for methylene chloride in a cancer risk assessment for occupational exposure has also been described,27- 29 and a PBPK model was used by the Occupational Safety and Health Administration (OSHA) in their rulemaking for methylene chloride.30 More recently, the U.S. EPA has used a PBPK model for vinyl chloride31 in its risk assessment for this chemical.32

Cancer risk assessments using PBPK models have also been proposed for many other chemicals, including TCE.33-36 However, methylene chloride and vinyl chloride represent the only cases to date in which cancer risk assessments conducted by federal regulatory agency have used a PBPK model for estimating cancer risks. Part of the reason for the slow progress of incorporating PBPK modeling in cancer risk assessment is the concern of regulatory agency risk assessors about uncertainties in its implementation.30 These concerns are not without basis; the potential impact of uncertainty in PBPK-based risk assessment has been a subject of some controversy in the literature.27-29,37-42 Nevertheless, the new U.S. EPA cancer guidelines strongly encourage the use of pharmacokinetic and tissuedose information both in the qualitative hazard identification step and in the quantitative dose-response evaluation for a chemical.

B. Consideration of Carcinogenic Mode of Action

Clearly, pharmacokinetic information describing the relationship between the environmental exposure to a chemical and the target- tissue exposure to a primary carcinogenic metabolite can be an important factor in conducting a chemical-specific risk assessment. However, the incorporation of pharmacokinetic information alone may not provide a more accurate assessment of human risk if modeof- action information is ignored. This problem can be illustrated by a comparison of inhalation risk estimates for liver cancer from TCE with similar estimates for the known human carcinogen, vinyl chloride. In the case of vinyl chloride, there is strong epidemiological evidence from several occupational cohorts demonstrating that vinyl chloride is a potent human liver carcinogen. Indeed, it was possible to use the results of these epidemiological studies to estimate the cancer potency of vinyl chloride directly, resulting in risk estimates that were very close to those derived from animal bioassays.31 In the case of TCE, on the other hand, although there have been a large number of well- conducted studies of chronic occupational exposure to TCE at concentrations similar to or even higher than the exposures to vinyl chloride, none of these studies has provided unequivocal evidence of carcinogenicity, as will be discussed later.

Despite this strong contrast in the human occupational evidence regarding carcinogenic risk, the comparison of pharmacokinetically based risk estimates for TCE and vinyl chloride provided in Table 1 demonstrates the implausible result of very similar risk estimates for these two chemicals when low-dose linear approaches are used with animal bioasseay data. As already mentioned, the most recently published U.S. EPA risk assessments for TCE22,23 make use of pharmacokinetic information. Specifically, estimates of total metabolism from pharmacokinetic studies in rodents and humans, scaled by body surface area, were used as the basis for the potency calculations. The resulting U.S. EPA estimates, based on rodent bioassays, of the human cancer risk from lifetime exposure to 1 g/ m^sup 3^ of TCE in air are on the order of 1.31.7 per million.22,23 A more recent pharmacokinetic risk assessment using a PBPK model arrived at a similar estimate using model-predicted area under the concentration-time curve (AUC) in the blood for the metabolite TCA as the basis for the potency calculations.35 The current U.S. EPA risk assessment for vinyl chloride also considers pharmacokinetic information, and the potency calculations are based on metabolism estimates obtained with a PBPK model.31 The risks of vinyl chloride exposure predicted with this model from rodent inhalation bioassays are shown in Table 1 for comparison with the pharmacokinetically based risk estimates for TCE. As can be seen in Table 1, the pharmacokinetically based risk estimates for vinyl chloride are very similar to those calculated for TCE. However, this similarity of the pharmacokinetically based low-dose linear risk estimates for liver cancer from these two chemicals stands in stark contrast to the marked difference in human evidence for liver carcinogenicity from occupational exposures. Thus it would appear that pharmacokinetics alone is inadequate to provide a reasonable comparison of the human risk for liver cancer from these two chemicals.

TABLE 1

Comparison of Human Liver Cancer Risk Estimates for Inhalation of TCE and Vinyl Chloride Using Animal Bioassay Data and Pharmacokinetic Dose Metrics(a)

Just as the pharmacokinetics and metabolism of a chemical must be considered in order to obtain a realistic measure of internal tissue exposure to the active form of the chemical, the pharmacodynamics and mechanism of action of the chemicalinduced carcinogenic process must also be considered in order to obtain a realistic measure of the expected response to that exposure. A concept that has proven useful in support of these considerations is the "mode of action," a term coined by the U.S. EPA in its new guidelines for carcinogen risk assessment. In the U.S. EPA's Draft Final Guidelines for Carcinogen Risk Assessment,1 the term "mode of action" is defined as:

sequence of key events and processes, starting with interaction of an agent with a cell, proceeding through operational and anatomical changes, and resulting in cancer formation. A "key event" is an empirically observable precursor step that is itself a necessary element of the mode of action or is a marker for such an element. Mode of action is contrasted with "mechanism of action," which implies a more detailed understanding and description of events, often at the molecular level, than is meant by mode of action. The toxicokinetic processes that lead to formation or distribution of the active agent to the target tissue are considered in estimating dose but are not part of the mode of action as the term is used here. There are many examples of possible modes of carcinogenic action, such as mutagenicity, mitogenesis, inhibition of cell death, cytotoxicity with reparative cell proliferation, and immune suppression.

The draft guidelines provide a discussion of the desired elements of a mode of action and a description of the kinds of data that can inform its development, as well as a conceptual framework for modeof- action evaluation that has also been adopted by IPCS.43 Indeed, one of the more striking features of the proposed U.S. EPA cancer guidelines is the emphasis they place on the consideration, throughout the risk assessment, of a chemical's carcinogenic mode of action.

While mode of action has occasionally been considered in risk assessments in the past, either to help in the determination of whether a particular carcinogenic effect seen in animals was relevant to humans or to support the use of a threshold approach for estimating safe human exposures, such cases served as exceptions to a standard approach that was applied across chemicals regardless of differences in mode of action. In 1991, the U.S. EPA's Risk Assessment Forum recommended that male rat renal tubule tumors arising as a result of a process involving accumulation of α^sub 2^-microglobulin should not contribute to the qualitative weight of evidence that a chemical poses a human carcinogenic hazard, and should not be included in dose-response calculations for the estimation of human risk.44 The qualitative implications of mode- of-action information have also been discussed for other carcinogenic processes such as rodent forestomach tumors (for which there is no corresponding organ in the human), in the case of butylated hydroxyanisole,45 and bladder tumors resulting from irritation by crystalline deposits, in the case of saccharin.46 An example of an agency mode-of-action evaluation leading to a conclusion that an animal tumor endpoint was not relevant to the human health assessment can be found in the inhalation cancer risk assessment for 1,1-dichloroethylene.47 Mode-of-action information has also been used by the U.S. EPA's Risk Assessment Forum to modify the quantitative portion of the cancer risk assessment: specifically, to justify the application of a threshold dose- response paradigm rather than the customary U.S. EPA assumption of low-dose linearity in the case of thyroid follicular cell carcinogenesis.48 The quantitative use of chemical-specific pharmacokinetic and mode-ofaction information to provide a basis for departing from the default linear dose-response approach has also been suggested for a number of chemicals where carcinogenicity was believed to be secondary to toxicity, including ethylene dichloride,49 ethyl acrylate,50 and chloroform.51 The U.S. EPA's recent risk assessment for chloroform52 represents the first case of a cancer risk assessment that departs from the standard low-dose linear dose-response paradigm on the basis of chemical-specific modeof-action information.

FIGURE 1. Diagram of the factors that must be considered in a cancer risk assessment for trichloroethylene (TCE). Information on the pharmacokinetics of TCE and its active metabolites, trichloroacetic acid (TCA), dichloroacetic acid (DCA), chloral (CHL), and dichlorovinylcysteine (DCVC), is necessary to support target tissue dosimetry. Information on the carcinogenic mode of action in each target tissue is then necessary to determine the appropriate measure of tissue \dose, as well as to suggest the nature of the relationship (linear, nonlinear, threshold) between tissue dose and tumor response.

C. Risk Assessment Approach

Figure 1 illustrates the process involved in performing a risk assessment for TCE that considers both pharmacokinetics and mode of action. The process, of course, begins on the right side of the diagram, with the hazard identification studies that provide evidence of the carcinogenicity of TCE and identify the tumors and target tissues of concern. Mechanistic information specific to each of the tumors of concern must then be evaluated to develop a hypothesis for the mode of action in each target tissue. The role of the mode of action is to provide the basis for linking target tissue chemical exposure with the biological or biochemical effects in the target tissue that lead to the observed cancer response. In parallel, the information shown on the left side of the diagram must also be gathered: Data on the pharmacokinetics and metabolism of TCE and its metabolites, including each of its putative carcinogenic metabolites (CHL, TCA, DCA, and 1,2DCVC). The actual risk assessment then proceeds from left to right in Figure 1. The pharmacokinetic and metabolism data are used to build a pharmacokinetic description, such as a PBPK model, in order to provide predictions of the concentration profiles for TCE and its active metabolites in each of the target tissues for cancer, whether associated with exposure to TCE in the animal bioassays or in potential human exposure scenarios. The specific mode of action associated with the production of a particular tumor provides the basis for identifying the dose metric to be used in these calculations. The mode of action plays a fundamental role in driving expectations regarding both the dose-response for tumor incidence within a species and the nature of crossspecies scaling. These expectations, in turn, drive decisions concerning the most appropriate risk assessment approach and the assumptions to be made where chemical-specific data are lacking.

II. EVIDENCE FOR THE CARCINOGENICITY OF TRICHLOROETHYLENE

The first step in the cancer risk assessment for a chemical is the hazard identification, in which the evidence for carcinogenicity, and its relevance to potential human environmental exposures, is evaluated. Under the new U.S. EPA guidelines, not only chronic epidemiological and animal bioassay data are evaluated, but also genotoxicity, pharmacokinetic, and mechanistic data. The possibility that the chemical may only be carcinogenic under certain conditions of exposure (route, dose, etc.) can also be considered. Only tumor data for TCE will be examined in this section; pharmacokinetic, genotoxicity, and mechanistic data will be examined in subsequent sections.

Although a large number of studies have demonstrated tumors in animals following exposure to TCE,53 the relevance of these animal results to the question of the human carcinogenicity of TCE has frequently been questioned.54-56 The American Conference of Government Industrial Hygienists now classifies TCE into Group A5, not suspected as a human carcinogen, based on a well-conducted, negative epidemiological study performed in an aircraft maintenance facility at Hill Air Force Base.57,58 The International Agency for Research on Cancer (IARC), on the other hand, assigned TCE to Group 2A, probably carcinogenic to humans, based on its assessment of sufficient data in animals and limited data in humans.59 The human evidence considered significant by IARC was the consistency of an association of TCE exposure with slightly increased incidences of liver/biliarytract tumors and non-Hodgkins lymphoma in studies of three cohorts in the United States,57 Sweden,60 and Finland,61 despite the fact that all three studies were characterized as negative by the original investigators because the increases were not statistically significant. Since that time, several additional well-conducted studies of occupational exposure to TCE in the United States62-64 and Denmark65,66 have been reported; these studies were also characterized by the authors as failing to support an association between TCE exposure and increased cancer mortality.

In contrast to these essentially negative results, a study in what was formerly East Germany67 found that workers repeatedly exposed to very high airborne concentrations of TCE (estimated to have sometimes been on the order of 1000 ppm) showed a highly significant incidence (7/169) and mortality (2/169) from kidney cancer as compared to controls (0/190 and 0/190). Moreover, the tumors seen in these workers were reported to be similar, both in locality and in histology, to the tumors observed in rats exposed to TCE. However, this study has been criticized because it was based on a previously recognized cluster of kidney cancer cases.68 Nevertheless, two subsequent case control studies of renalcell cancer patients have found an association with a history of occupational TCE exposure.69,70 Further evidence for the association of TCE with the occurrence of these kidney lesions is the observation of detoxification products of a potentially reactive and mutagenic TCE metabolite in the urine of human workers exposed to TCE by inhalation. Because the development of the renal-cell carcinoma observed in these subjects has been associated with mutations in the von Hippel-Lindau (VHL) tumor suppressor gene, two independent studies were performed to investigate whether TCE exposure could result in specific mutations of the VHL gene. Unfortunately, the results of these two studies are contradictory, with one study in Germany71 finding a unique mutation pattern, involving multiple mutations and loss of hererozygosity, associated with TCE exposure, while the other study in Switzerland72 revealed no unique mutation pattern. Despite the equivocal nature of the human evidence, the correspondence between rodents and humans for these rare kidney tumors, similar to the case of liver angiosarcoma from vinyl chloride, justifies increased concern for kidney lesions in a revised cancer risk assessment for TCE.

The animal cancer bioassays that have been conducted for TCE are summarized in Table 2. By far the most common carcinogenic outcomes associated with TCE exposure in animal studies are increased liver and lung tumors in several strains and both sexes of mice.73-78 Statistically increased tumor outcomes observed in only a single study include malignant lymphoma in HAN:NMRI mice exposed by inhalation,79 renal tubular-cell adenoma and carcinoma in male F344 rats exposed by oral gavage,74 and benign testicular (Leydig-cell) tumors in Sprague-Dawley rats exposed by inhalation.77,78 Of these less commonly found outcomes, the kidney tumors have raised the greatest concern since they are not observed in control animals. Bioassay results for metabolites of TCE are also shown in Table 2. Hepatocellular carcinoma is induced by several of the oxidative metabolites of TCE: CHL, TCA, and DCA.3,80-84

With regard to conducting a quantitative risk assessment for the carcinogenicity of TCE, none of the human studies just described provide the necessary combination of a statistically significant association and an adequate characterization of exposure. Therefore, the animal bioassays provide the only reliable basis for a quantitative risk estimate. The U.S. EPA's published quantitative cancer risk estimates for TCE have been based on animal bioassays, specifically liver and lung tumors in mice. In 1983, the U.S. EPA calculated unit risks for TCE of 4.1 10^sup -6^ (g/m^sup 3^)^sup - 1^ for inhalation and 0.54 10^sup -6^ (g/L)^sup -1^ for drinking water using data on the incidence of liver tumors in male B6C3F1 mice given TCE in an oil vehicle by gavage73,85; the linearized multistage model86 was used with a calculation of absorbed TCE dose scaled by body surface area to obtain these estimates.87 In 1985, lower unit risks of 1.3 10^sup -6^ (g/m^sup 3^)^sup -1^ for inhalation and 0.32 10^sup -6^ (g/L)^sup -1^ for drinking water were recalculated on the basis of the same oral gavage bioassays using the results of pharmacokinetic studies54,88,89 to calculate total metabolized dose in both animals and humans, rather than absorbed dose, although the body surface area adjustment was still applied to obtain the human equivalent dose.22 In 1987, the U.S. EPA calculated a new inhalation unitriskof 1.7 10^sup -6^ (g/m^sup 3^)^sup -1^ based on the incidence of mouse lung and liver tumors in inhalation bioassays,76-78 again using a calculation of metabolized dose and the body surface area adjustment.23

III. PHARMACOKINETICS AND METABOLISM

For ease of presentation, information on the pharmacokinetics and metabolism of TCE and its metabolites is described before discussing mode-ofaction considerations. However, it should be noted that the two areas must actually be considered together. Pharmacokinetic and metabolic data provide insights into the possible carcinogenic modes of action and the potential roles of parent chemical, metabolites, and the metabolic processes themselves. Mechanistic data, on the other hand, may identify minor metabolites that are crucial to the risk assessment, which might otherwise have been overlooked due to their low production relative to other metabolites. The following discussion summarizes the experimental evidence for the pharmacokinetics and metabolism of TCE and its major metabolites, TCA and trichloroethanol (TCOH), as well as the minor metabolites, DCA, CHL, and 1,2-DCVC, that have been suggested to play an important role in the carcinogenicity of TCE.

TCE is a volatile, lipophilic chemical that distributes readily throughout all tissues, including the brain, but partitions preferentially into fat tissue. In contrast, its major metabolite TCA is a watersoluble chemical that preferentially distributes into the plasma and rich\ly perfused organs, and is found only in relatively lower concentrations in the muscle and fat. The properties of TCOH are somewhat intermediate between the other two compounds.90 Clearance of TCE occurs both by exhalation and by metabolism.91 A schematic of the metabolic pathways for TCE is shown in Figure 2, with the major oxidative pathway, which takes place primarily (but not exclusively) in the liver, shown to the right of the diagram and the minor glutathione-dependent pathway, which involves several locations including the liver and kidney, shown to the left.

A. Oxidative Metabolism

TCE is much more extensively metabolized in the mouse than in the rat, whether TCE is administered orally54 or by inhalation.92 The primary route of metabolism for TCE, shown on the right side of the diagram in Figure 2, is oxidation via the microsomal mixed function oxidase (MFO) system, now referred to as cytochrome P-450, or CYP.93- 99 A minor pathway for TCE metabolism, involving conjugation with glutathione (GSH) by glutathione transferase (GST), has also been observed100; this pathway, which is shown on the left side of the diagram, is described in the next section.

TABLE 2

Summary of Bioassay Results for Trichloroethylene and Its Metabolites

TABLE 2

Summary of Bioassay Results for Trichloroethylene and Its Metabolites

TABLE 2

Summary of Bioassay Results for Trichloroethylene and Its Metabolites

FIGURE 2. Metabolism of TCE. Abbreviations not given in text: (right pathway) CDH, chloral dehydrogenase (aldehyde oxidase); EHR, enterohepatic recirculation; FA, formic acid; GA, glyoxylic acid; OA, oxalic acid; TCE-O-P450, oxygenated TCE-cytochrome P-450 transition state complex; TCOG, TCOH glucuronide; UGT, UDP glucuronosyl transferase; (left pathway) BL, cysteine conjugate - lyase; CGDP, cysteinyl-glycine dipeptidase; DCVG, dichlorovinyl glutathione; DCVSH, dichlorovinyl mercaptan; GGTP, y-glutamyl transpeptidase; NADCVC, N-acetyldichlorovinylcysteine; NAT, A/- acetyltransferase.

The principal oxidative metabolite formed in vitro is CHL (in its hydrated form, chloral hydrate),93,94,97 which is subsequently reduced to TCOH in the cytosol or oxidized to TCA in either the cytosol or mitochondria.93 CHL is not stable in vivo, and circulating concentrations are relatively low compared to its breakdown products, TCA and TCOH.90 Within a few hours of the administration of 50 mg/kg chloral hydrate to a child, the rapid initial clearance of CHL was essentially complete, with subsequent blood concentrations paralleling the time course for TCOH but approximately an order of magnitude lower, suggesting a continuing production of CHL from TCOH.101

The principal circulating metabolite of TCE in the blood is TCA, which accumulates in the body due to protein binding102 and slow excretion,103 while TCOH is readily excreted.104105 TCA appears to be derived both directly from CHL and indirectly from TCOH.104-'06

Based on both in vitro and in vivo studies, the metabolism of TCE has been suggested to consist of oxidation of TCE to CHL by the MFO system, followed by either oxidation of CHL to TCA by an aldehyde oxidase, also known as chloral dehydrogenase (CDH), or reduction to TCOH by alcohol dehydrogenase (ADH) with subsequent glucuronidation. Oxidation of TCOH to TCA by the MFO system has also been proposed.106

Four different isozymes of cytochrome P-450 have been found to play a role in the oxidative metabolism of TCE in rodents, IAl/2, 2Bl/2, 2C11 /6, and 2El ; CYP 2E1 appears to have the highest affinity for TCE, although the other isozymes can become important at higher TCE concentrations.97 In humans, CYP 2E1 was found to account for more than 60% of TCE metabolism in vitro, with smaller contributions from IAl, 1A2, and 3A4; variation in CYP 2E1 activity across tissues from 23 humans was less than 10-fold.107 Sex-, pregnancy-, and age-related differences in metabolism can also result from normal variations in CYP 2El content.98 Moreover, increased metabolic capacity can result from the induction CYP 1A1/ 2 (e.g., by aromatics), 2B1 (e.g., by phenobarbital), or 2E1 (e.g., by ethanol).6,96 However, due to the high affinity of CYP 2El metabolism, the clearance of TCE is typically flow-limited at low TCE concentrations; therefore, variations in enzyme activity may not result in significant changes in TCE metabolism for human exposure conditions.108

Inhalation exposures of human volunteers to TCE concentrations from 27 to 201 ppm showed no evidence of metabolic saturation or of a change in the proportion of TCA to TCOH.109 Saturation of TCE metabolism has been observed in mice, rats, and dogs.110,111 The relative proportion of the major metabolites does not appear to be a strong function of dose; however, repeated dosing does appear to increase the production of TCA at the expense of TCOH,112 and the relative production of CO2 increases with increasing dose in mice.6 The production of TCA in humans appears to be highly variable, and generally somewhat higher than in other animals. For example, in one study the production of TCA from chloral hydrate in different individuals varied from 5 to 47%.104

Human in vivo studies with TCE98 have identified the major urinary metabolites to be TCOH (50% of the administered TCE dose), primarily as the glucuronide, and TCA (19%); monochloroacetic acid (MCA) was also identified as a minor metabolite (4%) in these studies. Another minor metabolite, yV-(hydroxyacetyl)-aminoethanol, has also been identified in human (and rodent) urine following TCE exposure, and TCE-derived oxalic acid has been detected in the rodent."3 A study of TCE metabolism in nonhuman primates"4 found that TCA was partially excreted as the glucuronide, particularly at longer times after dosing; the authors suggest that since the detection of TCA glucuronide had not been reported previously, TCA excretion may have been underreported in earlier studies (such as the human study cited earlier). The glucuronidation of TCA is supported by the observation that TCA is excreted in the bile of rats and mice."2 Urinary excretion represents the major route of elimination of the metabolites, with fecal excretion, in the form of TCOH glucuronide, accounting for less than 5% of the total."4 The low fecal excretion is apparently associated to some extent with enterohepatic recirculation of TCOH (i.e., biliary excretion of the glucuronide, followed by hydrolysis and reabsorption of TCOH), which has also been suggested to occur in rats.115

DCA has been reported to be a minor urinary metabolite of TCE (less than 1 %) in both rats and mice,6,112,113,116 while MCA has been reported to be present at less than 0.1%."2 DCA has also been reported in the urine of mice dosed with TCA.117 Both MCA and DCA were detected in the urine of an acute intoxication patient who had ingested approximately 70 ml of TCE,118 and DCA was reported to reach maximum concentrations on the order of 0.1 M in subjects exposed to 100 ppm TCE for 4 h.119 A recent in vitro study with mouse and rat liver tissues concluded that unlike most other chlorinated compounds, which are metabolized primarily by the microsomal enzymes of the mixedfunction oxygenase (MFO) system, DCA degradation appears to occur primarily in the cytosol in a process that requires GSH.120 The enzyme responsible for this process has been identified as glutathioine transferase zeta, and the principal product isglyoxylic acid.121 The kinetics of DCA have been extensively studied in the human due to its clinical use.122"125 The peak concentration and AUC of DCA in the plasma after intravenous administration are linear up to approximately 20 mg/kg, but above 20 to 30 mg/kg some individuals display evidence of saturation of metabolism.123 The apparent volume of distribution and half-life for DCA are 0.3 L/kg (range: 0.09 to 0.60) and 1.05 h (range: 0.25 to 1.87), respectively.124 Significantly, the clearance in humans appears to be much more rapid than would be expected from allometric scaling of animal data; in a comparative study,122 the half-lives in rats, dogs, and humans were 2.1-4.4 h, 17.1-24.6 h, and 0.330.6 h, respectively.

It has recently been reported that under some conditions DCA can be artificially produced from TCA in blood samples (but not serum, plasma, or urine samples) as an artifact of the analysis. Specifically, if acidification of the blood sample is performed rapidly after collection of the sample, before the iron in hemoglobin has been oxidized by exposure to air, DCA can be generated from TCA present in the sample.126 Subsequent to this discovery, a study127 identified additional sources of artifactual formation of DCA from TCA during analysis. Using analytical methods that avoided artifactual DCA formation, these investigators were unable to identify DCA in blood (with a 1.9 μ? detection limit) after either intravenous dosing with 100 mg/kg TCE or oral dosing with 1000 mg/kg TCE. Thus this artifact may have severely compromised the data discussed in this section on DCA blood concentrations following administration of TCA or TCE.110,128,129 The artifactual formation of DCA from TCA ex vivo calls into serious question the evidence, discussed in the next paragraph, that DCA can be produced in vivo from TCA, but does not necessarily invalidate the conclusion that DCA may be produced in vivo from the metabolism of TCE.

Studies in rats have suggested that DCA may be produced from TCA.128 It has also been reported that DCA was produced in a roughly linear fashion from perchloroethylene,130'131 at levels consistent with production from TCA, the principle metabolite of perchloroethylene. However, an analysis of data on the dose- response and elimination kinetics of DCA formed from TCE led to the conclusion that another source of DCA was required in addition to TCA, but that the data were inconsistent with the second source being the initial oxidation step129 ; inste\ad, the production of DCA from TCOH was hypothesized. A metabolic study of TCA and DCA in rats and mice128 found that DCA was more rapidly metabolized than TCA: More than 50% of the administered TCA from an oral dose was excreted unchanged in the urine, as compared to only 2% of administered DCA. Plasma concentrationtime curves for TCA were similar in mice and rats, while those for DCA were greater in rats than in mice, both when DCA was administered and when it was derived from TCA.128

There is evidence that exposure to high concentrations of DCA inhibits its own metabolism by glutathione transferase zeta.132 In studies with human volunteers, 124,125 the excretion half-life for DCA increased 2- to 6-fold following repeated intravenous doses of DCA on the order of 10 to 50 mg/kg. Inhibition was only slowly reversible, taking from 1 wk to greater than 3 mo to resolve. In the rat and mouse, exposure to 0.2 or 2 g/L of DCA in drinking water for 14 d caused a significant increase in the blood concentration-time profiles of a subsequent dose of DCA.133134 The half-maximal inhibition concentration in mouse, rat, and human are similar, on the order of 40 M.132 Therefore, while DCA selfinhibition can be observed at high doses of DCA, it is not of concern for human environmental exposures to DCA or its chemical progenitors.

B. Metabolism in the Lung

As with most chemicals, the preponderance of the metabolic clearance of TCE appears to take place in the liver. It has been demonstrated, however, in studies with an isolated ventilated perfused lung preparation,135 that the male F344 rat lung also possesses a limited oxidative metabolic capability for TCE. While the affinity (K^sub m^) for the lung metabolism observed in that study was similar to the affinity observed in the liver, the capacity (V^sub max^) of the lung metabolism was less than 1% of the capacity of the liver. These results suggest that lung metabolism is not an important contributor to total in vivo metabolism in the rat, and that the rat lung does not possess a significant first-pass (presystemic) clearance capability for inhaled TCE. However, these results do not rule out the possibility that metabolism in the lung could produce sufficient local exposure to metabolites to produce toxicity and/or carcinogenicity.

C. Conjugative Metabolism

In mouse, rat, and human, a small proportion of TCE appears to be metabolized by enzymatic conjugation with GSH, principally by GST in the liver,136,137 followed by further metabolism in the kidney to the cysteine conjugate, 1,2-DCVC.138 The GST metabolic pathway is shown on the left side of Figure 2. Delivery of 1,2-DCVC to the kidney may also be mediated by enterohepatic recirculation, in which glutathione conjugate excreted in the bile is converted by gut bacteria to the cysteine conjugate, which is then reabsorbed.139 The GSH conjugate has been identified both in vitro, with rat liver microsomes, and in the bile of rats given 2.2 g/kg TCE in corn oil.100 The TCE glutathione conjugate has also been observed in the blood of human subjects exposed to TCE at 100 ppm for 4 h.140 The cysteine conjugate has also been identified in the urine of animals dosed with TCE.141

The bioactivation of 1,2-DCVC to a reactive and mutagenic thioacetylating intermediate is performed by cysteine conjugate β-lyase in the kidney.142 Although similar β-lyase activity has been measured in the kidney and liver, the two enzymes are distinct.143 Detoxification and clearance of 1,2-DCVC takes place by urinary excretion of the N-acetyl derivative.100,144 In a study with perchloroethylene,145 it was determined that the excretion of the N-acetyl derivative was dose-related (a higher fraction of N-acetyl derivative was excreted at doses where the oxidative pathway was saturated), and was significantly greater in the rat than in the mouse. However, measurements of acid-labile protein adducts associated with 1,2-DCVC suggest that the activation of 1,2-DCVC in the kidney may be as much as 12-fold greater in mice than in rats, and that the kidney tissue exposure to 1,2- DCVCderived reactive species from oral dosing with TCE may be twice as great in the mouse as in the rat.146,147

The activity of β-lyase has been measured in the liver and kidney of both humans and rats. One research group has reported a specific activity In hu= man kidney on the order of 2.0 to 3.6 nmol/ min/mg eytosol,148,149 as compared to 6,45 to 7,6 nmol/ min/mg eytosol in the rat,141 Another research group, however, has reported a maximum velocity (V^sub max^) of only 0,8 nmol/min/mg eytosol in the human, with an affinity (K^sub m^) of 0,29 mM, compared to V^sub max^ = 7.5 nmol/min/mg cytosol and ATn, = 1.6 mM in rat kidney cytosol in the same study,130 Data for perehloroethylene on the relative activity of liver ey= tosolic glutathione transferase and kidney oytosoliq cysleine conjugate /Hyase suggest that the human activity of both enzymes Is roughly 10=fold lower than the rut.145 On the other hand, the specific activity of N-acetyltransferase in kidney cytosol appears to be very similar across species; 0,41 nmol/ min/m cytosol in the human, as compared to 0,35-0.61 in the rat and 0,94 in the mouse.151

The fact that yv=acetyl=DCVC has been identi= fled in the urine of humans exposed to TCE, both OQQupationully151 and in controlled exposures,152 indicates that exposure of the kidney to DCVC does occur in the human. In the occupational study,151 the concentrations of total W-acetyl-DCYC (both the 1,2 and 2,2 isomers) in the workers' urine was about one third ofthat measured in rats dosed orally with 50 nig/kg TCE, The ratio of total A'-aeetyl-DCVC to TCA in the workers' urine ranged from 0,03 to 0,3, while in rats it ranged from 0.025 to 0.045, and in mice from 0,014 to 0,065, However, more re= cent data,1M obtained in controlled inhalation stud= les with both rats and human subjects, suggest that relative urinary excretion of total N-aeetyl-DCVC metabolites Is actually somewhat lower in humans than in rats,

IV. MODE OF ACTION

The discussion in this section focuses on the three animal target tissues of greatest concern: liver, lung, and kidney, In each tissue, evidence for the possible mode (or modes) of action of TCl is ex= plored in order to lay the basis for (1) the identification of elements that must be included ill the phai'= macokinetic model, (2) the selection of the most appropriate dose metrics for exposure of that tis= sue, and (3) the determination of the approach to be used for the dose=response assessment for that tissue, In particular, the selection of the mode of action plays a major role in determining the ba= sis for eross=speeies extrapolation and the method to be used for low-dose extrapolation (linear or MOE),

In general, much mere information has been collected relevant to the liver carcinogenicity of TCE than for the lung and kidney tumors, but uncertainties remain in all three eases, As mentioned earlier, risk assessments must be conducted in the face of such uncertainty, Science continually progresses, and risk assessors Can only hope to provide an interim best estimate based on a snapshot of the state of the science at that point in time, It will always be true that the use of ongoing research when it becomes available could improve and perhaps radically change that risk assessment, To the extent possible, the risk assessor must therefore attempt to estimate the potential impact of existing uncertainties on the current assessment.

A. Liver

As understanding of carcinogenic proceises in the rodent liver and of the effects of TCE has evolved, a number of possible mechanisms for the liver carcinogenicity of TCE have been suggested, These alternative mechanisms can be loosely characterized as involving genotoxicity, cytotoxieity, and promotion. More recently, as evidence became available that several of the metabolites of TCE also induce liver tumors (CHL, TCA, and DCA), attention has focused on the possible role of these metabolites in the induction of liver cancer by TCE. In particular, it has been suggested that the production of TCA and DCA from TCE under bioassay exposure conditions is adequate to account for the incidence of liver tumors in the TCE bioassays.4,110

1. Genotoxicity

At one time, covalent binding to DNA was considered to be a common feature of all chemical carCinogenicity, and the mutagenicity, and hence the aarcinogenicity, of chlorinated elhylenes was expeeted to correlate with the reactivity of their epoxlde intermediates, which was taken as a predictor of the potential for DNA adduct formation,153 In the ease of TCE, it was believed that rapid eonverof the presumed epoxide intermediate to chloral was consistent with a low mutagenic and carelnogenic activity,153 Investigators have repeatedly emphasized the lack of mutagenic or DNA-binding evidence for a direct genotoxic effect of TCE in the liver. 88,154 Nevertheless, in an in vivo study, a low rate of covalent binding to DNA was taken as an indication that TCE is possibly a weak initiator, and the production of adducts was imputed to the GST pathway.155

There is a similar sparsity of direct evidence for genotoxic potential for the major circulating metabolites of TCE. CHL has been found to be genotoxic in a number of in vitro studies,156 but is present only at relatively low concentrations, on the order of 1 mg/ L or less, after TCE doses associated with liver tumors in bioassays.54 The generation of free radicals from CYP2E1 oxidation of CHL, TCA, and TCOH in vitro has been shown to produce lipid peroxidation to carcinogenic species156; the relative production of free radicals from the three compounds, CHL, TCA, and TCOH, is consistent with the source of the free radicals being the oxidation of TCA in every case. However, ras mutational frequencies in TCA- induced tumors are essentially identical to those in spontaneous tumors, suggesti\ng that TCA treatment does not lead to genotoxicity in vivo.157 Evidence of possible genotoxicity in the form of single- strand breaks in DNA,158,159 and increased DNA synthesis as evidenced by thymidine incorporation,160 has been observed for TCA and DCA. However, other investigators have been unable to reproduce the reported DNA strand break effects,161'162 and the thymidine incorporation data are susceptible to a variety of explanations, including mitogenicity, cytotoxicity,91 and the potential for direct effects of DCA and TCA on the thymidine pool.160 No increases in oxidative DNA damage were seen with TCA or DCA at carcinogenic drinking water concentrations in mice, whereas brominated analogues administered at the same concentrations produced measurable, dose- related increases in 8-hydroxydeoxyguanosine.163

Two studies of in vivo ras mutation spectral data for DCA157,164 reported a shift in the frequency with which specific ras mutations were detected in DCAinduced tumors compared to spontaneous tumors. These data were interpreted as evidence of either initiation activity or selective promotion of specific cell subpopulations. There has also been a report of evidence for the mutagenicity of DCA,165 although a more recent in vitro mutagenicity study on DCA was entirely negative,166 and an in vivo genotoxicity study167 suggested that DCA might be an extremely weak inducer of genetic damage in mice at a drinking water concentration of 3.5 g/L, but not at lower concentrations. Since DCA has been shown to produce tumors in mice at a drinking water concentration of 0.5 g/L,81 it is unlikely genotoxicity is the primary mechanism of DCA carcinogenicity.

Overall, the evidence relating to the genotoxicity of TCE and its metabolites is inadequate to support a primarily genotoxic mechanism for the observed liver carcinogenicity of TCE, but it does raise some concerns relevant to the question of residual low-dose risks below the threshold for other potential modes of action.

2. Cytotoxicity

As an explanation for the observed carcinogenicity of TCE in the mouse liver, but not in the rat liver, an epigenetic mechanism was proposed88 in which the higher rates of metabolism observed in the mouse would produce recurrent toxicity, and a consequent reparative hyperplasia, resulting from the binding of an unspecified reactive intermediate with cellular macromolecules; consistent with this mechanism, toxicity was observed at high doses in mice but not rats, and there was little evidence of DNA binding. A subsequent pharmacokinetic study168 demonstrated that the relationship between the acute hepatotoxicity of TCE and the total production of urinary metabolites was linear, and it was suggested that this result was consistent with the hypothesis that the toxicity was produced by a reactive metabolite. Citing the proposed cytotoxic mechanism for the carcinogenicity of TCE,88 it was recommended that total metabolism be used as a pharmacokinetic dose measure for the carcinogenicity of TCE in place of the administered dose. Largely in response to this study, the U.S. EPA recalculated the carcinogenic potency of TCE, based on the NCI73 and NTP85 gavage studies, using pharmacokinetic data on total metabolism of TCE in the mouse and human22; the pharmacokinetically derived potencies were almost identical to the potencies calculated previously using administered dose.87 Total metabolism was also used as the basis for the later derivation of a revised inhalation carcinogenic potency of TCE based on inhalation bioassay data.23

A cytotoxic mechanism has recently been proposed as the basis for the liver carcinogenicity of TCE.169 Recurrent toxicity, leading to reparative hyperplasia, has also been suggested to be a likely contributing factor to the carcinogenicity of other halogenated hydrocarbons,170,171 particularly in the case of chloroform, where toxicity was believed to result from binding of reactive metabolites to critical cellular macromolecules.17,51,172 However, studies on other chemicals that are similar to TCE do not support the possibility of a carcinogenic mechanism based solely on binding to cellular macromolecules. For example, the binding in mice from two noncarcinogens, 1,1-dichloroethane and 1,1,1trichloroethane, was 2- to 18-fold greater than from two closely related carcinogens, 1,2- dichloroethane and 1,1,2-trichloroethane.173

There is a considerable body of evidence supporting the production of unidentified reactive species in the initial metabolism of TCE,95,174-177 similar to that observed with ethanol.178 Moreover, repeated exposure to TCE has been associated with cytotoxicity in mice, but not rats, when the TCE was administered in a corn oil gavage,88,179 although not when an aqueous vehicle was used.179 However, there has been no evidence demonstrating a link between the liver toxicity of TCE and a sustained, generalized reparative hyperplasia under bioassay conditions that could plausibly lead to increased tumor incidence, such as has been shown with chloroform.I80'~83 Indeed, there is molecular biological evidence that chloroform and TCE have quite different modes of action. An evaluation of in vivo ras mutation frequency data for several chemicals184 concluded that the presence of H-ra.c proto-oncogene activation in only a small fraction (21%) of chloroform-induced hepatocellular neoplasms, as opposed to its activation in more than 60% of control animal tumors, was supportive of a nongenotoxic mode of action. In contrast, H-ra.v activation in TCE- and DCA-induced tumors were on the order of 60%, similar to spontaneous tumors.

Glycogen accumulation, vacuolization, focal necrosis, and reparative hyperplasia have been reported at drinking-water bioassay concentrations with DCA, but not TCA, in mice.3,83,160 However, a recent study with rats demonstrated similar tumor potency for DCA in the rat compared to the mouse, but with no evidence of necrosis or increased cell proliferation,185 effectively ruling out a cytotoxic mode of action for this chemical. Thus while cytotoxicity may contribute to the incidence of liver tumors in some TCE bioassays, particularly those performed by gavage in corn oil, it appears highly unlikely that cytotoxicity and reparative hyperplasia are the principal mode of action underlying the production of liver tumors from TCE.

3. Promotion

TCE administered by gavage at 1000 mg/kg daily for 7 wk acted as a weak promoter of rat liver altered foci in one initiation/ promotion study using diethylnitrosamine as the initiator.186 The TCE metabolites TCA and DCA have also been shown to promote foci and tumors initiated by W-methylN-nitrosourea.187 A promotional mechanism has sometimes been proposed as an explanation for the fact that the liver carcinogenicity of TCE has been observed in mice but not rats, based on the observation that a much greater peroxisomal proliferation response is observed in mice, as compared to rats, exposed to TCE.54-154,188-190 With regard to the major circulating metabolites of TCE, peroxisome proliferation is also produced by TCA and, with less potency, by DCA, with the mouse again being the more sensitive species.191 In fact, based on structure-activity considerations, it is likely that the metabolite TCA is responsible for the peroxisomal activity of TCE: Most of the known peroxisome proliferators are acids or are metabolized to acids.192,193 This conclusion is supported by in vitro studies showing that TCA- but not TCE-activated cloned peroxisome proliferators activated receptor alpha (PPAR-α).194

A number of peroxisome proliferators, including DCA,195 have been used successfully in humans for the treatment of hyperlipidemia, but their use has been criticized due to the evidence that they induce liver cancer in rodents.196 In general, the carcinogenic peroxisomal proliferators have not been found to be mutagenic or directly genotoxic,55 and it is widely believed that they act through one or more epigenetic mechanisms: (a) oxidative stress, leading to oxidative damage of DNA,196 (b) generalized cell proliferation,197 (c) focal cell proliferation,198 and (d) interaction with steroid hormones.199 That these effects are secondary to activation of PPAR- a has been demonstrated by the lack of liver carcinogenicity and related early events in PPAR-a knockout mice following exposures to a peroxisome proliferator.200,201

It has been observed that some peroxisomal proliferators are much more potent carcinogens than others, even though they produce similar levels of peroxisomal proliferation and enzyme induction.202 The key difference affecting potency appears to lie in the ability to induce sustained cell proliferation in a subpopulation of cells.197,198,203 A rationale for the appearance of focal, as opposed to general, hyperplasia from exposure to peroxisomal proliferators '98 is that normal cells apparently cease to respond to the chemical's mitogenic signal after a short period of time, but that some altered cells (i.e., the preneoplastic cells) are able to maintain a sustained proliferative response leading eventually to tumor formation. A study of proto-oncogene activation in tumors produced by TCE, perchloroethylene, and DCA164 concluded that exposure to these chemicals provides a selective growth advantage to spontaneously occurring mutations. Interestingly, studies in PPAR- α knockout mice demonstrated a lack of altered cell cycling for treatment with TCE or TCA, whereas DCA induced similar alterations in both the control and knockout mice.204

An attractive hypothesis developed to explain the promotional effects of phenobarbital,205,206 but that is consistent with the observed effects of many peroxisome proliferators, suggests that the mitogenic signal from a chemical leads to an initial burst of organ- wide proliferation that in turn engenders the production of an opposing cytostatic signal (e.g., TGF-β) from the s\urrounding epithelial tissues and downregulation of the cellular receptors for growth factors. The prolonged maintenance of this dynamically unstable situation by chronic exposure to the mitogen provides a selective pressure that favors the development of a subpopulation of altered cells that have lost the ability to respond to the cytostatic signal. These altered cells have therefore escaped from the inhibition and can respond to the mitogenic signal, greatly increasing their rate of growth as compared to normal cells, thus enhancing the probability of cell division errors leading to cell transformation. A suppression escape mechanism has been proposed for the liver carcinogenicity of dioxin.207,208 This mechanism is consistent with the observed promotion of basophilic foci by peroxisome proliferators, including TCA.83

Although there is still considerable uncertainty regarding the mechanism of carcinogenicity of peroxisomal proliferators,209 the conclusion often reached by those who have evaluated this mechanism has been that the liver tumorigenicity is essentially irrelevant to humans.192,193 This conclusion is based primarily on the observation that humans (and nonhuman primates) appear to be much less sensitive to hepatic peroxisomal proliferation than rodents.210"213 However, attempting a quantitative comparison between the rodent and the human on this basis is problematic, since it assumes that the cross- species sensitivity for the peroxisomal response can be assumed to apply to the carcinogenic response as well. The adequacy of this assumption depends on the mechanism of tumorigenicity and its relationship to the peroxisomal response. If, for example, the tumorigenicity were thought to be produced by oxidative stress secondary to increased peroxisomal oxidation, it might seem reasonable that the much lower human peroxisomal response implies a much lower relative cancer risk compared to the rodent. However, a recent study has demonstrated that oxidative damage from several haloacetates was independent of peroxisome proliferation.163

Similarly, if mitogenesis is the underlying mechanism of carcinogenicity, the relationship of the tumor responses across species cannot be based on the peroxisomal response unless it can be demonstrated that the dose response for the two effects, peroxisomal and mitogenic, is directly related. In fact, there does appear to be some interrelationship between the peroxisomal proliferation and mitogenicity produced by these chemicals. Avidity of binding to a cytoplasmic transport protein, liver fatty acid binding protein (L- FABP), has been shown to correlate with the potency of peroxisome proliferators, and its function is required for the production of the mitogenic response to these chemicals in vitro. This protein functions as an intracellular carrier of fatty acids in hepatocytes, and binds such endogenous substrates as linoleic acid, arachidonic acid, and prostaglandin EI , as well as a number of growth- modulatory compounds.214 However, the relationship between the dose responses for peroxisome proliferation and mitogenesis varies from strain to strain in the mouse: In the BALB/c mouse the threshold for the induction of peroxisomal oxidation and mitogenesis occurs at similar exposure levels, while in the C57B1/6N mouse mitogenesis occurs at exposure levels at least one order of magnitude lower than the threshold for increased peroxisomal oxidation.215 This lack of correspondence clearly argues against the use of peroxisomal proliferation as a surrogate for the variation of the mitogenic response across species.

In the case of DCA, the onset of neoplasia in the rat from chronic treatment occurs at concentrations (0.5 g/L) well below the onset of increased peroxisomal activity (1.6 g/L), indicating that the peroxisomal response is not a requirement for the tumorigenicity of this chemical.185 However, DCA has been shown to alter hepatic glucocorticoid receptor binding activity, and the dose response for the receptor interaction correlated with that for DCA liver tumorigenicity.216 Glucocorticoids exert a significant effect on cellular metabolism, differentiated function, and proliferation. DCA also stimulates glycogen accumulation in normal hepatocytes through an insulin-dependent mechanism,217 but DCA-induced tumors possess significantly different insulin receptor levels than marginal tissues and do not accumulate glycogen.218 In general, data on the effects of DCA exposure are consistent with a mitogenic mode of action in which a subpopulation of cells is promoted by selective pressure. Although transient increases in organwide cell proliferation have been reported,83 repeated exposure of mice to DCA in drinking water eventually results in generalized inhibition of both mitosis219 and apoptosis,220 while continued treatment results in the formation of focal hyperplastic lesions.3,221,222 It has been suggested that the tumors produced by DCA arise from these areas of focal hyperplasia.222 Indeed, it has been demonstrated that the hyperplastic nodules produced by DCA include clusters of cells expressing the same tumor markers as the adenomas and hepatocarcinomas, supporting their preneoplastic role.220,221 Significantly, no such clusters were observed outside of the hyperplastic nodules, and very few of the commonly observed altered hepatic foci were found, suggesting that the hyperplastic nodules are the only significant preneoplastic lesion in DCAinduced hepatocarcinogenesis.221,222

Marked differences have been noted in the phenotype and cell replicative behavior of tumors induced by DCA and TCA.223,224 The altered cells induced by DCA exposure, which are predominantly

Source: Critical Reviews in Toxicology

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