* Miami Valley Laboratories, The Procter & Gamble Company, P.O. Box 538707, Cincinnati, Ohio 45253;
Pfizer Global Research and Development, Eastern Point Road, Groton, Connecticut 06340; and
National Health Environmental Effects Research Laboratory, Office of Research and Development, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina 27711
Received February 14, 2003; accepted March 31, 2003
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ABSTRACT |
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In parallel with these research efforts that were attempting to define the scope and nature of the endocrine disruptor hypothesis, the U.S. Congress added provisions to the Food Quality Protection Act (FQPA) and the Safe Drinking Water Act of 1996 to require the testing of food-use pesticides and drinking water contaminants, respectively, for estrogenicity and other hormonal activity. These bills were enacted into law, giving the EPA the mandate to implement them. The EPA, with the help of an external advisory committee, the Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC), determined that other hormonal activity should include androgens and compounds that affect thyroid function, and expanded the mandate to include all chemicals under EPAs jurisdiction, potentially including the 70,000 chemicals regulated under the Toxic Substances Control Act (Endocrine Disruptor Screening and Testing Advisory Committee [EDSTAC], 1998). EDSTAC recommended an extensive process of prioritization, screening, and testing of chemicals for endocrine-disrupting activity, including a screening battery that involves a combination of at least eight in vitro and in vivo assays spanning a number of taxa (EDSTAC, 1998
). What started out as a hypothesis has become one of the biggest testing programs conceived in the history of toxicology and the only one that has ever been based on mechanism of action as its premise.
As we pass the 10th anniversary of the emergence of the endocrine disruptor hypothesis, it is useful to look back on the progress that has been made in answering the nine questions posed as data gaps in the EPAs research strategy (EPA, 1998a)not only to see what we have learned, but also to examine whether the questions are still appropriate for the goal, what gaps remain, and what directions should be emphasized in the future.
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Question 1: What Effects Are Occurring in Populations? |
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In humans, the evidence for direct cause-and-effect relationships from environmental exposures remains generally lacking, with the effects of PCBs on neurological development among the strongest possibilities (IPCS, 2002). The diethylstilbestrol (DES) case is clearer, but here the exposure was to pharmacological levels given to prevent miscarriage. DES produced reproductive tract malformations in sons and daughters and vaginal tumors in daughters who were prenatally exposed (for a review, see Mittendorf, 1995
). For perspective, most other agents that have been implicated as endocrine disrupters are much less potent than DES and are present at lower levels. What is at issue, then, is the notion that, by inference, much lower exposures to agents with weaker activity are responsible for widespread effects on development, reproduction, and the production of certain types of cancer. This has been the central question of a number of reviews, including the National Academy of Sciences report, Hormonally Active Agents in the Environment (National Research Council, 1999
). The Academys panel was split on the general hypothesis. The reasons for the lack of consensus are briefly discussed in the Executive Summary of the report; principal among them were the significant limitations and uncertainty in the database on observed effects and their potential etiologies.
There has been considerable retrospective analysis and reanalysis of secular trend data for cancers and birth defects that may have a hormonal etiology (e.g., Jorgensen et al., 2002; Paulozzi et al., 1997
), as well as for sperm counts (e.g., Fisch et al., 1996
; Swan et al., 1997
). Unfortunately, too little of the secular trend data has been critically pursued with etiological studies. There are many alternative explanations for the trends that are at least as plausible as endocrine disrupters in the environment; these include improvements in medical surveillance (e.g., for breast cancer) and changing medical practice (e.g., a possible association between increased use of progesterone in the first trimester to forestall miscarriage and hypospadias [Silver, 2000
]).
Clearly, then, one of the biggest unanswered questions is that of the magnitude of the endocrine disrupter problem. This question can be answered only by epidemiology and ecoepidemiology approaches that fully conform to criteria controlling for bias and that are aimed at determining causation. Such studies will be time-consuming and expensive but are necessary if we are to determine whether the secular trend data for diseases such as breast cancer, testicular cancer (particularly in Denmark), hypospadias (in the U.S.), and others are attributable, in whole or in part, to endocrine disrupters. It is worth noting that some critical analyses have been done, and these will be helpful in addressing the question. In the last decade, a number of studies have evaluated the levels of environmental agents in banked tissue samples and compared them with outcomes such as breast cancer, concluding that there was no association; these studies will provide a useful start for answering this question (Adami et al., 1995; Krieger et al., 1994
). But even here, it is not clear what the critical life stage is that should be targeted for the exposure assessment (Birnbaum and Fenton, 2003
). Continuing follow-up of DES daughters as they enter the stage of life during which breast cancer incidence increases will also be useful in addressing the question of whether prenatal exposure is a prerequisite for increased risk (Hatch et al., 1998
).
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Question 2: What Are the Chemical Classes and Their Potencies? |
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The EDSTAC recommendations for screening were comprehensive in scope: Evaluate all known mechanisms of action for compounds that interfere with estrogen, androgen, and thyroid hormone function and that are representative of a number of vertebrate taxa (EDSTAC, 1998). Known mechanisms are not limited to receptor agonism or antagonism but also include such activities as inhibition of enzymes of hormone synthesis (e.g., via alterations in 5-
-reductase or aromatase). The recommended screening program is ambitious, entailing at least eight different assays that span the gamut from receptor binding assays to more complex whole animal studies in rats, frogs, and fish.
A recent review on the three EDSTAC options discusses the pros/cons of the individual assays based on the data generated since the EDSTAC report was issued (OConnor et al., 2002). The EDSTAC screening program is time- and resource-consuming, with estimates of the cost of running the battery for a single compound in the range of $200$300,000 (EDSTAC, 1998
). Critical decisions remain as to which of the numerous assays will actually be incorporated into the screening battery after factoring in considerations such as the extent of redundancy needed, the efficiency in the use of animals, the sensitivity and specificity of the assays, and need to cover multiple taxa. The generation of screening data for large numbers of compounds will ultimately lead to establishment of triggers for Tier 2 testing, using multigenerational protocols.
A concern with the EDSTAC screening proposal is the fact that prenatal development is not represented, despite the fact that it is generally believed that the earliest life stages are the most sensitive to endocrine disruption. Therefore, it is possible that the screening program could generate a rate of false negatives that would be unacceptable. The main question is whether, indeed, these life stages represent critically unique periods of sensitivity or whether they merely represent quantitatively more sensitive stages. The problem with including prenatal exposure in the screen is that the developmental effects of endocrine disrupters tend to be latent; traditional end points of toxicity (i.e., altered structure or function) may not be detectable until sexual maturity. Because sexual maturity is not reached in lab rodent species until 810 weeks after birth, it has been considered impractical to evaluate the consequences of prenatal treatment in a screening-level assay. Pubertal male and female assays may address this point in part, but most development of the reproductive system has already occurred by the life stage at which dosing is commenced in these protocols. Although traditional end points of toxicity (e.g., structural malformations, diminished reproductive function) may be latent, it is highly likely that they are preceded by immediate and persistent changes in gene expression. It is now well established that the signal transduction pathway for steroid hormones involves changes in gene expression as an integral step; therefore, it should be possible to use gene expression changes in the fetus as a means of detecting endocrine-disrupting potential in the most relevant life stage.
Genomic tools that have become available over the past few years make it feasible to survey tissues for changes in gene expression. Recent work has demonstrated the validity of this approach for endocrine-disrupter screening: Estrogens of varying potencies have been shown to produce a characteristic transcript profile in the fetal reproductive tract of rats transplacentally exposed to these compounds (Naciff et al., 2002). It may be possible to use these transcript profiles to develop an alternative screening assay that is more sensitive than and obviates the need for many of the assays now being considered as part of the EPA battery. Similar activities are also underway for evaluation of effects in wildlife (e.g., Larkin et al., 2002
). Of course, such approaches will need to go through the same standardization and validation process now being followed for the existing proposed screens.
Another aspect of endocrine-disrupter screening that has yet to be addressed is the prioritization of compounds for screening and testing. EDSTAC recommended that EPA extend its screening program beyond the legally mandated testing of food-use pesticides and drinking water contaminants to all the chemicals under its regulatory purview. This includes the inventory of more than 70,000 chemicals regulated under the Toxic Substances Control Acta daunting number of chemicals for which to contemplate any sort of testing. There are some exclusion criteria by which some chemicals will be exempted from testing (most prominently, polymers greater than 1000 m.w.); however, this still leaves many thousands of chemicals in the pipeline for screening. To make good decisions as to which chemicals should be evaluated, tools for prioritization will need to be developed. These will include quantitative structure-activity relationship (QSAR) programs that predict receptor binding and high-throughput in vitro assays that evaluate receptor binding. There have been efforts to develop predictive QSAR programs (e.g., Mekenya et al., 2002; Serafimova et al., 2002
; Shi et al., 2001
; Suzuki et al., 2001
) but the validation and refining of these programs is rate-limited by the available amount of receptor binding data covering a breadth of chemical structures. Indeed, in the recent proposal for identifying the process by which the first 200 chemicals will enter into the EDSTAC screening battery, a number of measures of exposure potential for pesticides was chosen over QSAR because of the lack of validated methodologies (Federal Register, 2002
). High-throughput assays to evaluate receptor-binding activity are technically feasible; in fact, they are now a routine aspect of drug discovery programs in pharmaceutical companies. The challenge for endocrine-disrupter screening will be to adapt these approaches, which are optimized for detecting compounds with high potency, so that they can reliably detect substances with substantially less activity than the prototypical ligands. Even rather simple technical issues, such as compound solubilization, become a substantial obstacle due to the diverse chemical classes that need to be evaluated and will require novel approaches.
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Question 3: What Are the Dose-Response Characteristics in the Low-Dose Region? |
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Additional research will be necessary to resolve this issue of U-shaped dose response curves. Because we dont know which factors are the most important determinants of the development of the morphological end points measured in these studies, it is unlikely that simply conducting more studies using the same end points will do anything to clarify the muddle. Instead, future work will need to concentrate on characterizing the mechanism(s) by which an estrogen (or anti-androgen, for that matter) could produce one developmental effect at a low dose and something entirely different at a high dose. Such work might include an evaluation of whether the underlying gene expression induced by the estrogens is qualitatively different at high versus low doses.
If these nonmonotonic dose-response curves can be verified, it will also be important to determine whether the effects at low doses have any long-term adverse health consequences. Putz and co-workers (2001a,b) reported a nonmonotonic dose-response curve for weights of prostate, testis, and epididymidis in rats neonatally exposed to estradiol, but the low-dose effect was transitory. They surmised that the effect was attributable to acceleration of puberty; organ weights were comparable to controls on maturation. Accelerated puberty would be an important effect; whether this is related to a lasting health consequence (e.g., predisposition to prostatic hyperplasia) is an equally important question. One of the major limitations in assessing a role in prostate cancer is that animal models that are predictive for human prostate cancer are not available.
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Question 4: Do Our Testing Guidelines Adequately Evaluate Potential Endocrine-Mediated Effects? |
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Of perhaps greater consequence for the testing guidelines are the implications of nonmonotonic dose responses, as mentioned earlier, to the practice of risk assessment. Current risk assessment practices involve the application of uncertainty factors to a NOAEL (No Observed Adverse Effect Level) to arrive at an exposure level with minimal risk. However, the NOAEL is derived from study designs that assume that the dose response is monotonic. If there are effects below the presumed NOAEL, these should be considered in setting acceptable exposure levels and would necessitate the redesign of safety studies to include a much wider range of dose levels. This would require increasing the magnitude of toxicity studies, particularly the number of animals used, as well as different statistical approaches to analyze data. At present, the EPA has taken the interim policy position that, pending further research on the relevant mechanisms of toxicity, it would be premature to require testing of low-dose effects in the Endocrine Disruptor Screening Program (see http://www.epa.gov/scipoly/oscpendo/low_dose_statement.pdf).
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Question 5: What Extrapolation Tools Are Needed? |
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Question 6: What Are the Effects of Exposure to Multiple EDCs, and Will a Toxic Equivalency Factor (TEF) Approach Be Feasible? |
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Question 7: How and to What Degree Are Human and Wildlife Populations Exposed to EDCs? |
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The presence and potential hazards of pharmaceuticals in the aquatic environment is a relatively new issue and has recently received increased attention from regulatory and scientific communities, due to a growing number of peer-reviewed papers reporting measurable levels of pharmaceuticals in water (e.g., municipal effluent and surface, ground, and drinking water). For instance, the U. S. Geological Survey detected 82 pharmaceutical and personal care products in surface waters of the U. S. (Kolpin et al., 2002). Even though a drugs affinity to its target molecule is intentionally very high, the measurable levels of pharmaceuticals are generally quite low; therefore, exposure to pharmaceuticals in the water is unlikely to produce a significant acute toxicity hazard in nonmammalian species (Webb, 2001
). Rather, the concern has focused on the potential for chronic, developmental, or reproductive effects to occur in aquatic organisms due to the biological activity of pharmaceuticals. The concern has been raised that these risks are currently not adequately assessed by current test methodologies and data evaluation strategies used for pharmaceuticals. To address this concern, it has been suggested that mammalian pharmacology/toxicology data, clinical efficacious levels in humans, and environmental concentrations can be used to prioritize those pharmaceuticals that may require additional aquatic toxicity evaluation (Huggett et al., 2003
). The premise of this model is that many of the enzyme and receptor targets in mammalian species have significant sequence homology in nonmammalian species (e.g., Andersen et al., 2000
; Menuet et al., 2002
; Todo et al., 2000
). The significance of these findings will continue to be debated until sufficient research is conducted to address the magnitude of this issue. The issue of pharmaceuticals in the environmentillustrates a key need in assessing the magnitude of the endocrine disrupter problem, namely, the prospective monitoring of aquatic species within our waterways to detect population trends has been limited. Expanded routine monitoring is a research need. Perhaps the most useful approach would be to develop a safety assessment study, possibly using fish, that could identify target organ toxicity by utilizing a histopathological examination similar to the screening approaches used in mammalian systems.
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Question 8: What Are the Major Sources and Environmental Fates of EDCs? |
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Question 9: How Can Unreasonable Risks Be Managed? |
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Selected High-Priority Research Needs
Our ability to answer each of the nine questions posed above is limited by the amount of information available. There are some clear research needs that, if addressed, would go a long way toward answering these questions. Here are several that we view as important:
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Conclusion |
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NOTES |
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1 To whom correspondence should be addressed.
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