Science CornerScience CornerMajor Risk Assessment Issues
Assessing health risks is a scientific process that can be described as analyzing the relationships among exposure, dose, and response. This paper will focus on these relationships. Approaches to be discussed include uses of methods, measurements, and models, best done in an highly integrated fashion. Exposure assessment, which describes pathways by which chemical or physical insults move from a source to the exterior of the body, is not included.
One of the first questions to be asked when discussing "dose" is what the term means. Is it dose to the organism, or dose to the tissue? The level of refinement may be even more specific with target dose being the dose to a cell or even a specific macromolecule within the cell. Methods must be developed and measurements made that will allow predictive models to be formulated to estimate the appropriate dose metric. Lack of agreement between model predictions and experimental measures leads to development of improved methods, better measurements, and models with not only more validity but more predictive power as well.
Once the appropriate dose metric has been defined, the question must be addressed whether "dose" to a target tissue (cell, macromolecule) means a response will occur. For example, does binding of a ligand to its cognate receptor necessarily result in signal transduction? Does formation of a DNA adduct mean a mutation has occurred? The answer to both of these questions is no. Interaction of a dose with a target allows the potential for a response. Additional steps are always necessary to actually bring about a biologic effect. The inherent sensitivity of the tissue plus the dose determines whether a response will occur. For example, if a given cell type does not have a necessary factor, a response cannot occur. If a chemical causes an anti-mitogenic signal that says "stop dividing," but the cell is already terminally differentiated, it cannot respond to the signal.
Risk assessment involves two basic types of extrapolation: dose extrapolations and response extrapolations. Most experimental studies are usually conducted at high dose for short times by the oral route, when in reality, concern for environmental health usually involves low-dose exposure over long periods of time and often by inhalation or dermal exposure. Recent applications of techniques of physiologically based pharmacokinetic modeling have led to major advances in the area of dose extrapolations. This approach incorporates realistic physiologic and biochemical parameters into quantitative descriptions of what the body does to a chemical: absorption, distribution, metabolism, and elimination.
By contrast, extrapolation of animal data to human response involves the development of biologically based, dose-response models. These models incorporate mechanistic information and understanding of inherent tissue sensitivity and involve quantitative descriptions of what the chemical in question does tothe body. Mechanistic studies may be conducted in vivo or in vitro and may involve animals or humans. The classic parallelogram approach is often applied. However, a three-dimensional array involving high-to-low dose, interspecies, and in vivo-in vitro studies is probably a better representation of the many relationships that need to be understood to incorporate mechanistic data into risk assessment processes.
Incorporation of New Scientific Information on Risk Assessment
The remainder of this paper will focus on how the incorporation of new scientific information is affecting the process and findings of risk assessment. Some major chemical classes found at hazardous waste sites include volatile organic hydrocarbons, heavy metals, and dioxins/PCBs. Examples of recent investigations into the pharmacokinetics and mechanisms of a representative chemical from each of these classes will be discussed.
Benzene
Benzene, the simplest aromatic hydrocarbon, is a product of both natural and industrial sources. Produced by plants, animals, and forest fires, its major sources in the environment are from commercial uses of coal and oil products. Benzene is heavily used as an industrial solvent and chemical intermediate and in plastics, pesticides, and detergents. People are exposed occupationally in the rubber, oil, chemical, gasoline, and shoe industries. Consumer products resulting in exposure include glues, adhesives, cleaners, paint strippers, art supplies, cigarette smoke, and gasoline.
Environmental exposures result from gasoline, vehicle exhaust, cigarette smoke, chemical spills, storage-tank leaks (contaminating groundwater and air), toxic waste sites, and food products. Benzene is both an acute and chronic animal and human toxicant and carcinogen (ATSDR 1989). Acute exposure in humans results in effects on the central nervous system, such as drowsiness, dizziness, and headaches. Long-term human exposure primarily affects the bone marrow, leading to immune suppression, aplastic anemia, and leukemia. The first reports of benzene's carcinogenicity in humans appeared in the 1920s, followed by additional epidemiologic studies resulting in the International Agency for Research on Cancer (IARC 1987a) and the U.S. Environmental Protection Agency (EPA 1984a) classifying benzene as a human carcinogen in the 1980s. Recent studies have suggested that non-leukemic malignancies may also occur in exposed persons (Yin et al. 1989). In experimental animals as in humans, acute exposure to benzene also results in effects on the central nervous system. Longer-term exposure results in bone-marrow suppression and chromosome abnormalities. Chronic treatment, following both inhalation and oral exposures, results in tumors in both sexes of rats and mice at multiple sites (NTP 1986, Maltoni et al. 1989). However, leukemia is rarely seen in experimental animals following chronic exposure to benzene.
Benzene undergoes biotransformation in the liver and in extrahepatic tissues of the body to metabolites that are more water soluble and thus more readily eliminated. However, this process also results in production of reactive intermediates. Benzene is initially oxidized via the cytochrome P450 mixed-function oxidase system to benzene oxide, which can then be either conjugated with glutathione leading to production of mercapturic acids or spontaneously rearranged to phenol. Phenol can be either conjugated with sulfate or glucuronic acid, leading to excretion, or further oxidized to hydroquinone or catechol, both of which can be conjugated and excreted. Catechol can be oxidized to produce trihydroxybenzene, while hydroquinone can undergo autoxidation and redox cycling, leading to transient production of semiquinone and benzoquinone, both highly reactive products. An additional pathway of benzene metabolism, which is not as well understood and may or may not involve benzene oxide, leads to production of muconaldehyde, a highly reactive dialdehyde, which is further oxidized to muconic acid and readily eliminated. Benzene metabolism can, in fact, be considered to consist of multiple pathways, some of which lead to more toxic products, others to detoxification (Henderson et al. 1992). Toxification pathways lead to ring breakage and benzoquinone, with muconic acid and hydroquinone, respectively, as markers. Detoxification pathwayslead to mercapturic acid products and phenyl conjugates, with prephenyl and phenyl mercapturic acids and phenyl conjugates as the markers, respectively.
Metabolism is required for benzene toxicity (Cooper and Snyder 1988). Mice are more sensitive to the toxic and carcinogenic effects of benzene than are rats. Can metabolic differences explain, at least in part, differences in species response to benzene? A series of studies by Sabourin and coworkers (1987, 1988, 1989) demonstrated that mice produce more of the "toxic" metabolites (hydroquinone, muconaldehyde) than do rats. From analysis of metabolites present in human urine (Inoue et al. 1988, 1989) and in-vitro studies (Brodfuehrer et al. 1989), humans appear to metabolize benzene more like the mouse. However, not only do species differences exist in favored metabolic pathways, but toxic pathways are saturated at relatively low-exposure concentrations in both rats and mice. That is, toxic pathways are high-affinity, low-capacity pathways, whereas pathways that produce detoxified metabolites have high capacity but low affinity. These enzymatic differences are also observed in monkeys and chimpanzees (Sabourin et al. 1992). Because most animal studies are conducted at high doses--doses at which detoxification pathways would be favored-- linear extrapolation of responses could underestimate the toxicity of benzene following low-dose exposures. Detailed studies on effects of route, species, dose, and rate of exposure on distribution and metabolic fate of benzene demonstrated that inhalation exposures resulted in greater tissue concentrations than did oral treatments (Henderson et al. 1992). Tissue concentrations following identical exposures were also lower in rats than in mice. Prior exposure to benzene had little effect on subsequent metabolism (Sabourin et al. 1990). Thus, although benzene oxidation is catalyzed primarily by cytochrome P4502E1 (Koop et al. 1989), which is highly inducible by some of its substrates such as acetone and ethanol, inhalation exposure to benzene fails to induce this isozyme.
Several physiologically based pharmacokinetic (PBPK) models have been developed to describe the behavior of benzene following oral and inhalation exposures (Medinsky et al. 1989, Travis et al. 1990, Bois et al. 1991). These models are flow limited, with ventilation rate and cardiac output being critical parameters. The lipophilicity of benzene is input by use of a high fat-to-blood partition coefficient. The rate of metabolism is also a critical parameter in describing benzene's behavior. At high levels of exposure, much of the benzene may be exhaled unchanged because metabolism is saturated. All these models adequately describe the kinetics of benzene in a given species following a single exposure route and successfully predict behaviors at different doses. Use of models to predict effects via a second route of exposure or in another species have promise. In addition, Medinsky et al. (1989) have modeled the behavior of major benzene metabolites that are markers for toxic and detoxification pathways. Travis et al. (1990) has a model with the advantage of including bone marrow as a specific physiological compartment. Recent studies suggest that in-situ metabolism in this tissue may be critical to its toxicity (Subrahmanyan et al. 1990). Although these PBPK models appear to allow dose and route extrapolations, inherent differences in species response still need to be explained.
Arsenic
In contrast to volatile organic compounds such as benzene, metals can neither be created nor destroyed. Industrial processes can result in their environmental release, as can increased acidification of rain water and other environmental processes. Arsenic is widely distributed in the environment, being present in measurable levels in certain foods and drinking water, resulting in ingestion. Although the maximum level allowed in drinking water is currently 50ęg/L ( EPA 1962), these levels are readily exceeded in areas of the United States, and in other parts of the world, even by a factor of 10 (Stohrer, 1991). The presence of arsenic in water and soil also leads to dermal exposure.
Occupational exposure to arsenic often involves inhalation. Arsenic is both an animal and human toxicant causing skin, gastrointestinal, and liver problems, as well as peripheral neuropathy. It is considered a human carcinogen by both EPA (1984b) and IARC (1980, 1987b). Occupational exposure is associated with lung cancer. Skin cancer has been associated with both oral and dermal exposures. In addition, recent studies have linked exposure to arsenic via drinking water with cancer of the urinary bladder, kidney, and liver (Cuzick et al. 1992, Smith et al. 1992). Although demonstrating arsenic carcinogenicity in standard animal studies has been difficult, arsenic has been shown to cause tumors in hamsters (Pershagen and Bjorklund 1985) and has been shown to be a "progressor" in in-vitro experiments (Lee et al. 1988).
Arsenic exists in several chemical forms in the environment (Cullen and Reimer 1989). Arsenic may be present as arsine gas. Other inorganic forms include the fully oxidized form of arsenic with a +5 valency, arsenate (AsV), and the partially reduced +3 state, arsenite (AsIII). Arsenic can also exist in organic forms involving methylation of either AsV or AsIII. Mono-, di-, and tri-methyl arsenate have all been detected in urine (McKinney 1992). Methylated arsenic species have been postulated based on the obligate requirement of glutathione in metabolism of arsenic species (Thompson 1993). Recent studies have demonstrated the existence of glutathion(GSH)-arsenate and GSH-arsenite species (Delnomdedieu et al. 1993, 1994; Scott et al. 1993). Recent advances in methodology now enable complete separation of major inorganic and organic arsenic forms for analysis (Yarley et al. 1993). Great care must be taken to insure that spontaneous oxidation does not occur during sample preparation and analysis.
In contrast to mercury for which methylated forms are the most toxic, organic arsenic species appear to be less toxic than the inorganic species (Squibb and Fowler 1983, Aposhian 1989). Depletion of GSH, an obligate cofactor in the methylation of arsenic, enhances the toxicity of arsenic (Buchet and Lauwerys 1987). Many studies have indicated that arsenite is more toxic than arsenate (Smith et al. 1992). Using the mouse lymphoma assay, sodium arsenite was shown to be approximately 10 times more cytotoxic than sodium arsenate (Harrington-Brock et al. 1993). However, on a weight basis, arsenite was approximately 3000 times more cytotoxic than methylarsonic acid, which was about 3 times more toxic than dimethylarsinic acid. Similar relationships hold for induction of gene mutations and chromosome aberrations. These recent data support the conclusion that methylation results in a decrease in arsenic's cytotoxicity, mutagenicity, and clastogenicity. However, dimethylarsinic acid is a developmental toxicant (Rogers et al. 1981), albeit at doses much higher than those reported to cause other forms of toxicity by inorganic arsenic species.
In general, however, methylation of arsenic appears to be a detoxification process if more reactive AS(III) intermediates are not allowed to accumulate. Recent studies in mice (Hughes et al. 1994) have indicated that metabolism of arsenic is dosed dependent. At high exposure concentrations, methylation becomes saturated resulting in a buildup of arsenic in the tissues. Levels of both monomethylarsonic acid (MMA) and arsenite also increase in urine, while levels of dimethylarsinic acid declined. Increased elimination of arsenite and MMA suggests that, at high doses of arsenate, the more reactive arsenic(III) intermediates may be allowed to accumulate and thus potentially increase the toxicity of arsenic. Although little evidence exists of toxicity in humans following exposure to low arsenic concentrations (Mass 1992), humans do excrete a relatively greater amount of MMA. Does this suggest that humans are more susceptible to arsenic toxicity? What levels of arsenite are eliminated in human urine? Thompson (1993) has proposed that amounts of both MMA and arsenite in human urine could be important biomarkers for high-dose exposure and adverse health effects.
Dioxins/PCBs
The polyhalogenated dibenzo-p-dioxins (PHDDs), dibenzofurans (PHDFs), and biphenyls are widespread persistent environmental contaminants. The most toxic of these is the PHDD isomer, 2,3,7,8-tetrachlorodibenzo-p-dioxin, TCDD or "dioxin." TCDD has a wide spectrum of effects, many of which are tissue-, species-, and developmental-stage specific (Birnbaum, In press). It is best considered as an environmental hormone because of the broad variety of its effects.
At the highest doses, dioxin causes delayed lethality, with the time to death being species specific. Because the "high" dose for dioxin is extremely low compared with doses for most other chemicals, dioxin is often considered the most toxic synthetic compound. At doses just below those that cause lethality, dioxin causes a severe wasting syndrome in which animals may lose as much as half their body weight. Slightly lower doses result in atrophy of lymphoid tissues and of the testis. Liver toxicity occurs in many species as a correlate of dioxin toxicity. This involves liver enlargement, fatty deposits, and even necrosis. Hyperplasia may be observed in epithelial tissues lining the bile duct, urinary bladder, and gastrointestinal tract. Metaplasia occurs in the meibomian glands of the eyelids and ceruminous glands of the ear canal resulting in waxy deposits. Dermal effects result in chloracne, a severe form of cystic acne that has often been called the hallmark of dioxin toxicity. Dioxin also causes birth defects, immunosuppression, and tumors. Some of the best-studied biologic responses to dioxin involve biochemical alterations. Dioxin results in induction of many enzymes leading to alterations in metabolism. Effects on multiple growth factors and their receptors leads to altered differentiation and proliferation. Perturbation of multiple hormone systems disturbs the normal homeostatic controls of an organism.
Some key issues in dioxin risk assessment involve understanding how dioxin causes its biologic effects. What is the mechanism of action of TCDD? Are humans a sensitive species to the biologic effects of dioxin? Do other chemicals act like TCDD? A specific cellular protein, known as the Ah receptor, binds dioxin and is essential for all its effects (Birnbaum 1994). However, although the Ah receptor is necessary, it is NOT sufficient to bring about responses to dioxin. Binding of dioxin is just the first step in a cascade of events that vary for different responses. The activated Ah-receptor/dioxin complex interacts with other proteins and regulatory factors to alter gene expression. Although functionally similar to members of the steroid receptor superfamily of nuclear receptors, the Ah receptor is a basic helix/loop/helix protein (Burbach et al. 1992, Ema et al. 1992). Clearly, however, interaction with the receptor is necessary but not sufficient for dioxin's effects. Although the dioxin/Ah receptor complex functions as a transcriptional enhancer for CYP1A1, the Ah receptor may have additional effects on signal transduction, such as direct activation of tyrosine kinases. Phosphorylation in response to dioxin has been demonstrated to be a rapid and sensitive response in human and mouse cells and tissues (Bombick et al. 1988, Clark et al. 1992, DeVito et al. 1994, Ma et al. 1992).
The species specificity of dioxin is most evident in the varying oral doses that result in lethality. Although guinea pigs die following an exposure to approximately 1 µg TCDD/kg, hamsters require approximately 5000 times as much. However, most animals examined have oral LD50s in the range of 100mg/kg. In fact, for any biologic response, a given species may be an outlier. Overall, however, most species exhibit similar responses following comparable treatments. What about humans? The ability to induce drug-metabolizing enzymes is similar in humans and experimental animals. In fact, the body burdens associated with induction of cytochrome CYP1A1 in humans may be even lower than those required in animals (Lucier 1991). The body concentrations of dioxins and related chemicals associated with chloracne are similar in humans, hairless mice, rabbit ears, and monkeys (Ryan et al. 1990). Using cells in culture, suppression of the immune response in lymphocytes occurs at similar concentrations of TCDD with cells from mice, monkeys, and humans (Neubert et al. 1992, Wood et al. 1992). Organ cultures of embryonic palatal shelves from rats and humans fail to fuse when exposed to the same concentration of TCDD (Abbott and Birnbaum 1991). And dioxin appears to cause cancer in both animals and humans. At least 18 positive cancer studies have been conducted in both sexes of rats, mice, and hamsters. Tumors are produced at multiple sites in all species (Huff 1992). In addition, dioxin induces high incidence of tumors in fish (Johnson et al. 1992) with short latency and at multiple sites. Three recent epidemiologic studies (Fingerhut et al. 1991, Manz et al. 1991, Zober et al. 1990), all with the advantage of measuring serum dioxin levels to validate exposure matrices, have indicated that elevated risk from all cancers is observed in highly exposed occupational cohorts. Commonalities observed in mechanisms of dioxin effects between animals and humans support response information indicating similar sensitivities. Thus, the weight of evidence is compatible with the conclusion that exposure to high levels of dioxin may be associated with increased incidence of cancer in both animals and humans.
Examination of dose-response relationships suggest that cancer may not be the most sensitive endpoint resulting from exposure to dioxin. Because all responses to dioxin require interaction with the Ah receptor, ligand binding occurs at the lowest exposure concentration. Biochemical responses such as induction of CYP1A1 and CYP1A2 can be detected at very low levels using molecular techniques such as quantitative PCR (Heuvel et al. 1992). At slightly higher doses, induction of enzymatic activity can be detected. These doses are similar to those at which immunotoxic and developmental effects have also been noted. Chloracne is a relatively high-dose response, and body burdens associated with detectable increase in tumors are still higher.
As mentioned, 2,3,7,8-TCDD is but one member of a family of chemicals that causes the same spectrum of biologic responses via its interaction with the Ah receptor. Approximate stereoisomers include the laterally halogenated dibenzo-p-dioxins, furans, biphenyls, naphthalenes, and azo- and azoxybenzenes, among others. The structural similarity, common mechanism of action, and common responses have led to development of toxic equivalency factors (TEFs) (EPA 1989). The toxic equivalency approach is a relative- potency scheme in which 2,3,7,8-TCDD is assigned a value of 1.0, and the potency of related compounds is expressed as some fraction of the TCDD potency. The toxic equivalency of a mixture can be expressed as a sum of the relative toxicities of each component. This approach has been widely used to predict and to describe the toxicity of complex environmental mixtures of PCDDs and PCDFs. However, use of TEFs for PCBs is more complex because only a small subset of PCBs are dioxin-like in their activity (Barnes et al. 1991). Non-dioxin-like PCBs have inherent toxicity of their own. Some appear to have neurotoxic properties, while metabolites of others are reproductive toxicants. Several non-dioxin-like PCBs are potent tumor promoters. Even for dioxin-like PCBs, the appropriate TEFs have demonstrated a wide range of values, depending on the response and system used (Safe 1990, Walker and Peterson 1991, DeVito et al. 1993).
Although differences in relative potency of dioxin-like PCBs are in part due to pharmacokinetic differences, such as differential rates of metabolism of various congeners, recent studies have provided increasing understanding of factors controlling the dosimetry of dioxin. In fact, several recent PBPK models have been developed to describe the behavior of TCDD and related compounds (Andersen et al. 1993, Kedderis et al. 1993, Kohn et al. 1993). In these models, disposition of dioxin is described as a diffusion-limited process, where tissue distribution is governed by its lipophilicity, metabolism, and induction of a hepatic-binding protein. These models have also linked kinetic behavior with the induction of a biochemical response, such as the increase in CYP1A1 and CYP1A2.
Estimates of human serum concentrations suggest that the average TCDD level, on a lipid-adjusted basis, is approximately 6 parts per trillion (ppt) (Needham et al. 1991). Inclusion of other laterally substituted PCDDs and PCDFs raises the average toxic equivalency to approximately 30 ppt. If recent estimates of TEFs for dioxin-like PCBs are included, dioxin equivalency might be in the range of 50 ppt. Where does this body burden come from? The majority of general population exposure to dioxin and related compounds comes from food, especially meat, dairy products, and fish. Persons in the United States, Canada, and western Europe ingest an estimated 1-3 pg PCDD/PCDF equivalents/kg body weight/day (Furst et al. 1991). Inclusion of dioxin-like PCBs in this estimate would raise the level of daily exposure to approximately 5 pg/kg/day. Of course, occupationally exposed workers, subsistence fishermen, and nursing infants would receive higher exposures. Whether general population exposures or those of sensitive subpopulations are at the level where adverse effects would occur is unclear. However, some evidence exists that this background level is not far below that where effects have already been detected in the human population (Egeland 1992, Sweeney et al. 1992, Wolfe et al. 1992).
Conclusion
Scientific risk assessment requires integration of exposure information, hazard identification, and dose/response data. Incorporating new scientific data decreases uncertainties in the assessment. PBPK modeling of benzene has demonstrated nonlinearities in production of toxic metabolites that could lead to underestimation of risk. Disposition studies of arsenic indicate the existence of nonlinearities in the methylation of arsenic, generally a detoxification process, which could result in overestimation of risk. Better understanding of the mechanism of action of dioxin and related compounds is leading to improved characterization of potential risk from this ubiquitous contaminant. The better the science, the better the risk assessment. Integrated approaches that involve both experiments and modeling will work in an iterative fashion to advance knowledge most quickly and efficiently. Incorporating data from both in-vivo and in-vitro studies allows mechanistic information to greatly improve our ability to extrapolate, not only from one experimental situation to another, but ultimately from animals to humans.
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