Chapter 24: Toxic Effects of Solvents and Vapors

The term solvent refers to a class of organic chemicals of variable lipophilicity and volatility. These properties, coupled with small molecular size and lack of charge, make inhalation the major route of exposure and provide for ready absorption across membranes of the lung, gastrointestinal (GI) tract, and skin. In general, the lipophilicity of solvents increases with increasing numbers of carbon and/or halogen atoms, while volatility decreases. Organic solvents are frequently used to dissolve, dilute, or disperse materials that are insoluble in water. As such they are widely employed as degreasers and as constituents of paints, varnishes, lacquers, inks, aerosol spray products, dyes, and adhesives. Other uses are as intermediates in chemical synthesis, and as fuels and fuel additives. Most organic solvents are refined from petroleum. Many such as naphthas and gasoline are complex mixtures, often consisting of hundreds of compounds. Early in the 20th century, there were perhaps a dozen or so known and commonly used solvents. By 1981, this number had climbed to approximately 350 (OSHA, 2006).

Solvents are classified largely according to molecular structure or functional group. Classes of solvents include aliphatic hydrocarbons, many of which are halogenated (ie, halocarbons), aromatic hydrocarbons, alcohols, ethers, esters/acetates, amides/amines, aldehydes, ketones, and complex mixtures that defy classification. The main determinants of a solvent’s inherent toxicity are: (1) its number of carbon atoms; (2) whether it is saturated or has double or triple bonds between adjacent carbon atoms; (3) its configuration (ie, straight chain, branched chain, or cyclic); (4) whether it is halogenated; and (5) the presence of functional groups. Some class-wide generalizations regarding toxicity can be made. For example, the more lipophilic a hydrocarbon, the more potent a central nervous system (CNS) depressant it is; amides/amines tend to be potent sensitizers; aldehydes are particularly irritating; hydrocarbons that are extensively metabolized tend to be more cytotoxic/mutagenic; and many unsaturated, short-chain halocarbons are animal carcinogens. The toxicity of solvents within the same class can vary dramatically. For example, both 1,1,1-trichloroethane (TRI) and 1,1,2-trichloroethylene (TCE) are halocarbons with three chlorine atoms, yet unsaturated TCE is carcinogenic in the rat and mouse, but TRI is not. Similar results have been reported for 2,4- and 2,6-diaminotoluene in rodents, as only the 2,4-isomer is capable of inducing significant hepatocyte proliferation and liver tumors. Slight structural differences in solvent metabolites are also of toxicological consequence. The peripheral neuropathy induced by n-hexane and 2-hexanone is dependent on the production of the γ-diketone metabolite 2,5-hexanedione. Diketones lacking the gamma structure are not neurotoxic. Thus, subtle differences in chemical structure can translate into dramatic differences in toxicity.

Nearly everyone is exposed to solvents during their normal activities. Consider, for example, a person who works in an aircraft factory as a metal degreaser (TCE exposure), drives to the neighborhood bar after work and has a few drinks (ethanol exposure) and cigarettes (benzene and styrene exposure), stops on the way home at a self-service filling station for gasoline (benzene, toluene, 1,3-butadiene, etc, exposure) and the dry cleaner’s for laundry (tetrachloroethylene [PERC] exposure), and after dinner enjoys his hobby of model shipbuilding that requires the use of glue (toluene exposure). While everyone may not identify with the above scenario, detailed surveys of indoor and outdoor air, such as the Environmental Protection Agency’s (EPA’s) Total Exposure Assessment Methodology (TEAM) and National Human Exposure Assessment Survey (NHEXAS) studies, indicate that airborne solvent exposure is unavoidable (Wallace, 1990; Clayton et al., 1999). Drinking water is also a common source of solvent exposure due to discharge of solvents into surface and groundwaters and the presence of disinfection by-products, including the animal carcinogen chloroform (CHCl₃). Trichloroacetic acid (TCA) and dichloroacetic acid (DCA), metabolites of TCE and PERC, are also common drinking water disinfection by-products.

Environmental exposures to solvents in air and groundwater are frequent subjects of toxic tort litigation, despite concentrations that are typically in the low parts per billion (ppb) or parts per trillion (ppt) ranges. Multiple exposure pathways frequently exist (Fig. 24-1). Although not represented in Fig. 24-1, household use of solvent-contaminated water may result in solvent intake from inhalation and dermal absorption as well as ingestion (Weisel and Jo, 1996; Nuckols et al., 2005). In many cases, risk assessment guidelines stipulate that risks be determined for physiologically diverse individuals who are exposed to several solvents by multiple exposure pathways. As an aid to the risk assessment process, the US EPA has derived toxicity factors for many of the most common solvents. These toxicity factors are referred to as reference concentrations (RfCs), reference doses (RfDs), and cancer slope factors (CSFs). Values for a number of these are available from the EPA’s online Integrated Risk Information System (IRIS). Additional sources of exposure guidelines for noncancer end points are found in the Toxicological Profiles of the US Agency for Toxic Substances and Disease Registry (ATSDR). These profiles often contain minimal risk levels (MRLs) that are derived in a similar manner to EPA’s RfCs and RfDs, but are frequently based on different critical studies or derived with different uncertainty factors.

Occupational solvent exposures involve situations ranging from a secretary using typewriter correction fluid to the loading and off-loading of tanker trucks with thousands of gallons of gasoline. The greatest industrial use of solvents is as metal degreasers. This work environment is typically where the highest exposures occur, mainly via inhalation and secondarily via dermal contact. An estimated 10 million people are potentially exposed to organic solvents in the workplace (OSHA, 2006). Many of the most severe exposures to solvents have occurred as a result of their use in confined spaces with inadequate ventilation. While the US Occupational Safety and Health Administration (OSHA) has established legally enforceable permissible exposure limits (PELs) for over 100 solvents, most PELs are outdated. The majority of existing PELs were adopted from a list of threshold limit values (TLVs) previously published by the American Conference of Governmental Industrial Hygienists (ACGIH). Many current TLVs are more stringent than PELs but do not carry the weight of law. Whereas the ACGIH’s TLVs for an eight-hour workday, 40-hour workweek are designed to be protective for a working lifetime, its short-term exposure limits (STELs) and ceiling values are designed to protect against the acute effects of high-level, short-term solvent exposure. If warranted, ACGIH will assign a skin notation to a solvent, indicating that a significant contribution to overall exposure is possible by the dermal route, by either contact with vapors or direct skin contact with the liquid. ACGIH also categorizes chemicals for carcinogenicity in animals and humans. Biological monitoring in the workplace should find increasing use as technologic advances are made, because it often provides a better measure of exposure than classic industrial hygiene monitoring. The ACGIH has published over 50 Biological Exposure Indices (BEIs), on which basis the safety of internal measures of exposure can be judged (ACGIH, 2012).

Most solvent exposures involve a mixture of chemicals, rather than a single compound (see the section “Solvent Mixtures”). Our knowledge of the toxicity of solvent mixtures is rudimentary relative to the toxicology of individual solvents. While the assumption is frequently made that the toxic effects of solvents are additive, the chemicals may also interact synergistically or antagonistically. For example, repetitive alcohol consumption induces certain cytochrome P450s (CYPs), and may therefore enhance the metabolic activation of other solvents to cytotoxic metabolites. Ethanol intake near the time of exposure to such solvents, in contrast, may competitively inhibit their metabolism and be protective. Another well-characterized example of solvent antagonism is the competitive metabolic interaction between benzene and toluene (Medinsky et al., 1994). Coexposure to these chemicals results in diminished benzene metabolism, genotoxicity, and erythropoietic toxicity relative to that which follows benzene exposure alone. It is now recognized that significant data gaps exist in the area of mixtures toxicology, and that these can be significant sources of uncertainty in risk assessments. Physiologically based toxicokinetic (PBTK) models are being developed by different research groups to predict the impact of metabolic induction and inhibition on the kinetics of individual components of specific mixtures (see the section “Physiological Modeling”).

Although some solvents are much less hazardous than others, virtually all can cause adverse effects. Provided that the dose or concentration is sufficient, most have the potential to induce some level of narcosis and cause respiratory and mucous membrane irritation. A number of solvents are animal carcinogens, but only a handful have been classified as known human carcinogens. Herein lies a major challenge for toxicology—determining the human relevance of tumors observed in chronic, high-dose rodent studies. As with other chemicals, whether adverse health effects occur from solvent exposure is dependent on several factors: (1) toxicity/carcinogenicity of the solvent; (2) exposure route; (3) amount or rate of exposure; (4) duration of exposure; (5) individual susceptibility; and (6) interactions with other chemicals. Adverse health effects may occur acutely and be readily discernible, or they may result from chronic exposure and have insidious onset. Numerous epidemiological studies of environmentally and occupationally exposed populations have been conducted for some solvents, but most human risk assessments remain heavily reliant on extrapolation from high-dose animal studies. One must bear in mind that the toxic effects and their underlying mechanisms discussed herein may be operative only in certain animal species or strains and under certain exposure conditions. Care must therefore be taken in generalizing beyond the experimental conditions under which data are collected. While a relatively small number of commercially available solvents is discussed in this chapter, those selected for discussion are thought to best demonstrate principles of solvent toxicology, are of particular commercial importance, and/or are currently garnering significant attention from the toxicological and regulatory communities. A book chapter that examines solvents from an organ systems standpoint, in contrast to discussion of individual solvents, is that of Gerr and Letz (1998).

The CNS depressant effects of acute, high-level exposures and the potential for permanent neurologic damage in chronic solvent abusers are not a matter of debate. It is also clear that chronic, moderate-to-high-level exposure to a few solvents such as n-hexane and carbon disulfide (CS₂) can cause specific degenerative changes in the CNS or the peripheral nervous system (PNS). Far less clear is whether chronic, low-level exposure to virtually any solvent or solvent mixture can produce a pattern of neurologic dysfunction referred to as painters’ syndrome, organic solvent syndrome, psychoorganic syndrome, and chronic solvent encephalopathy (CSE). CSE is characterized by nonspecific symptoms (eg, headache, fatigue, mood disturbances, and sleep disorders) with or without changes in neuropsychological function. There is a reversible form of CSE referred to as neuroasthenic syndrome that consists of symptoms only, and both “mild” and “severe” forms accompanied by objective signs of neuropsychological dysfunction that may or may not be fully reversible. This syndrome was first described in the Scandinavian occupational literature in the late 1970s, in solvent-exposed painters (Axelson et al., 1976; Arlien-Soborg et al., 1979). Since that time, numerous studies from Scandinavia have been published purporting that solvents as a class have chronic neurotoxic properties. These countries, as well as a few others in Europe (ie, Germany, Austria, and Belgium), have passed legislation recognizing CSE as a compensable occupational disability (Triebig and Hallermann, 2001). Scientists outside of Scandinavia, including many in the United States, have generally been less willing to recognize CSE as a legitimate disease state and have published studies to the contrary (Triebig et al., 1988; Bleecker et al., 1991; Spurgeon et al., 1994).

In response to the numerous reports of CSE, two conferences were convened in 1985. The first was held in Copenhagen by the Nordic Council of the World Health Organization (WHO, 1985). The second, in Raleigh, North Carolina, was attended by an international group of scientists from academia, industry, and government (Cranmer, 1986). The categorization scheme that resulted from the Raleigh meeting is presented in Table 24-1. The WHO scheme is similar. Among those who utilize the categorization scheme, it is generally believed that the most severe CSE category, type 3, results from repeated, severe intoxications like those experienced by solvent abusers. CSE types 1 and 2, on the other hand, are thought to be associated with prolonged, low-to-moderate-level exposure common to work environments. A major criticism of the categorization scheme is the lack of consideration of inhaled solvent concentration and exposure duration. While no consensus exists, even most CSE proponents believe that solvent exposure must occur for approximately 10 years before clinical symptoms are manifest. Citing growing acceptance of CSE outside of Scandinavia, some scientists and physicians advocate for refinement of existing diagnostic criteria and a unified categorization scheme (van der Hoek et al., 2000, 2001).

CSE researchers typically rely on self-reported symptoms and a clinical neuropsychological evaluation (Table 24-2), and to a much lesser extent on diagnostic tests such as electroencephalography and computerized brain tomography. It has been argued that the neuropsychological tests are of questionable validity, sensitivity, specificity, and predictive value. It has also been noted that many investigations of CSE are fraught with methodological flaws. For example, CSE investigators frequently fail to measure premorbid function or intellect; employ a reference population; control adequately for the potential confounders of age, alcohol use, other CNS diseases, and other chemicals; corroborate functional deficits with objective evidence of brain disease; and/or examine exposure–response relationships. The importance of doing so is best exemplified by the reanalysis of individuals originally reported in the seminal study by Arlien-Soborg et al. (1979) to have “painters’ syndrome.” When the influences of age, education, and intelligence were considered, the previously reported reduction in neuropsychological test scores disappeared (Gade et al., 1988). Another example is that of Cherry et al. (1985), who demonstrated the importance of matching solvent-exposed and control groups for preexposure intellect before making a diagnosis of CSE. More recently, Albers et al. (2000) reportedly found no objective evidence of toxic encephalopathy among 52 railroad workers with long-term solvent exposure and diagnosis of CSE.

It is evident that resolution of the controversial issue of CSE will come only through the conduct of well-designed and controlled epidemiological studies, especially considering the absence of an appropriate animal model. Brief, but insightful reviews of CSE by Rosenberg (1995) and Schaumburg and Spencer (2000) have been published. These reviews conclude that the current literature, including the “landmark” North American study of 187 paint-manufacturing workers (Bleecker et al., 1991), does not support chronic low-level solvent exposure as a cause of symptomatic CNS or PNS dysfunction. This does not preclude, however, the possibility that such exposure can be associated with subclinical cognitive dysfunction in the form of slight psychomotor and attentional deficit disorders. For the viewpoint of CSE proponents, readers are directed to texts by Arlien-Soborg (1992) and Kilburn (1998).

Inhalants are volatile substances that can be inhaled to induce a psychoactive or mind-altering effect. Their abuse has become a major drug problem worldwide, particularly in disadvantaged populations and among adolescents (Williams and Storck, 2007). The epidemiology of inhalant abuse in the United States is receiving considerable attention (Spiller, 2004; Wu et al., 2004), with almost 13.1% 8th graders claiming to have used inhalants according to a recent National Institute on Drug Abuse (NIDA)–sponsored survey (Johnston et al., 2008). Solvents are among the most popular classes of drugs of abuse, given their presence in a multitude of inexpensive, readily available products that are legal to buy and possess. These products are used for common household and industrial purposes and include paint thinners and removers, dry-cleaning fluids, degreasers, gasoline, glues, typewriter fluid, nail polish remover, felt-tip marker fluids, and aerosols such as fabric protector sprays and spray paints. Solvents are often among the first drugs used by children and adolescents. Early use of inhalants is often a precursor to abuse of multiple illegal substances (Wu et al., 2004). Research suggests that adverse socioeconomic conditions, a history of child abuse, poor grades, and dropping out of school are all factors contributing to inhalant abuse (NIDA, 2012).

Solvent abuse is a unique exposure situation in that participants repeatedly subject themselves to vapor concentrations high enough to produce effects that resemble alcohol intoxication. Solvents can be breathed in through the nose or the mouth by “sniffing” or “snorting” vapors from containers, spraying aerosols directly into the nose or mouth, “bagging” by inhaling vapors from a plastic or paper bag, or “huffing” from a solvent-soaked rag stuffed into the mouth (NIDA, 2012). Although dependent on the pattern of inhalation, blood levels of solvents typically peak a few minutes after inhalation begins, and the abuser can begin to experience effects after a matter of seconds. While intoxication may last only a few minutes, abusers frequently seek to prolong the “high” by inhaling repeatedly over the course of several hours. In extreme circumstances, death may be a consequence of cardiac arrhythmias, asphyxiation, and/or cachexia.

Relatively little is known about the neuropharmacology of abused solvents, although they appear to have much in common with the classical CNS depressant drugs such as ethanol and barbiturates, including the potential for tolerance and dependence (Balster, 1998). Commonly used solvents have been shown to alter the function of a variety of ligand-gated ion channels, including those activated by glycine, gamma-aminobutyric acid (GABA), and N-methyl-d-aspartate (NMDA) (Beckstead et al., 2000; Cruz et al., 2000). Recent evidence links high-level toluene exposure in rats to increased dopaminergic neurotransmission within the mesolimbic reward pathway, an effect thought to underlie the abuse potential of numerous drugs (Reigel and French, 2002; Reigel et al., 2004). Toluene, as well as TCE and TRI, has been shown to enhance serotonin-3 receptor function, which has also been implicated in the reinforcing properties of abused drugs (Lopreato et al., 2003).

Whereas the intoxicating effects achieved through acute solvent abuse are reversible, abuse may be continued for years and result in residual organ damage. For example, chronic abuse of products containing n-hexane and methyl-n-butyl ketone can cause peripheral neuropathies. Blood dyscrasias, liver damage, kidney injury, and hearing impairment are seen in patients who have abused solvents injurious to these organs. It has been known for some time that the brain is not spared residual damage, with long-term neurologic and psychological sequelae (Caldemeyer et al., 1996; Lubman et al., 2008). Rosenberg et al. (2002) reported increased incidences of neurologic and neuropsychological effects in chronic solvent abusers compared with a control group of chronic drug abusers. Solvent abusers did significantly worse on tests of working memory and executive cognitive function, and a much higher percentage of the patients (44.0% vs 25.5%) had structural abnormalities in subcortical regions of the brain (ie, basal ganglia, cerebellum, pons, and thalamus) as visualized by MRI (Fig. 24-2). Solvent abusers also showed moderate to severe diffuse abnormality of the cerebral white matter, consistent with that seen in earlier studies of neuropsychologically impaired toluene abusers (Rosenberg et al., 1988a,b). This condition was termed “white matter dementia,” as myelinated neurons are white in appearance. It was characterized primarily by diffuse cerebral, cerebellar, and brainstem atrophy, and ventricular enlargement (Fig. 24-2). Dementia in toluene abusers has been referred to as “toluene leukoencephalopathy” (Filley et al., 2004). Leukoencephalopathy, also known as multifocal demyelinating disease, involves structural alteration of cerebral white matter, in which the myelin sheaths that cover nerve fibers are destroyed but axons are largely spared. Thus, toluene is among a long list of white matter toxins identified to date.

Widespread use of solvents has resulted in their dissemination throughout the environment (Fig. 24-1). Everyone is exposed daily to solvents, as reflected by their common occurrence in blood of nonoccupationally exposed populations (Churchill et al., 2001; Sexton et al., 2005; Blount et al., 2006). Because solvents as a chemical class are volatile, the preponderance of solvents entering the environment does so by evaporation. The majority of the more volatile organic chemicals (VOCs) volatilize when products containing them (eg, aerosol propellants, paint thinners, cleaners, and soil fumigants) are used as intended. Solvent loss into the atmosphere also occurs during production, processing, storage, and transport activities, resulting in elevated concentrations in air in the proximity of point sources. Winds dilute and disperse solvent vapors across the world. Atmospheric concentrations of most VOCs are usually extremely low (ie, nondetectable to nanograms or a few micrograms per cubic meter of air). Relatively high atmospheric concentrations of certain solvents (eg, 10–520 μg/m³, or 3–163 ppb of benzene) have been measured in urban areas, around petrochemical plants, and in the immediate vicinity of hazardous waste sites. Motor vehicle exhaust is a major contributor to hydrocarbon emissions (Mohamed et al., 2002). Fraser et al. (1998) measured very high levels of benzene, toluene, ethylbenzene, and xylenes in a Los Angeles roadway tunnel.

Solvent contamination of drinking water supplies is a major health concern. Although the majority of a solvent spilled onto the ground evaporates, some may permeate the soil and migrate through it until reaching groundwater or an impermeable material. In years past, the more lipophilic solvents were generally regarded as water insoluble. It is now recognized that all solvents are soluble in water to some extent. Some (eg, alcohols, ketones, glycols, and glycol ethers) are freely water soluble. Maximum solubilities of some common hydrocarbon solvents range from 10 mg/L (ppm) for n-hexane to 24,000 ppm for bromochloromethane. High concentrations of VOCs are sometimes found in the effluent of facilities of rubber producers, chemical companies, petrochemical plants, and paper mills. Concentrations diminish rapidly after VOCs enter bodies of water, primarily due to dilution and evaporation. VOCs in waters rise to the surface or sink to the bottom, according to their density. VOCs on the surface will largely evaporate. VOCs on the bottom must depend on solubilization in the water or mixing by current or wave action to reach the surface. VOCs in groundwater tend to remain trapped until the water reaches the surface, although some are subject to microbial modification. Concentrations in well water are rarely high enough for acute or subacute toxicity to be of concern. The very low levels of some solvents typically found in water have, however, caused a great deal of concern and debate about their toxic and carcinogenic potential (NTP, 1993a).

Potential health effects of solvent contaminants of water have received considerable attention over the past >40 years. A report by Mason and McKay (1975), of an increased incidence of cancer in persons who drank water from the Mississippi River, prompted the EPA to analyze the water supply of New Orleans. The finding of approximately 76 synthetic organic chemicals, many of which were solvents, prompted passage of the Safe Drinking Water Act in 1974. CHCl₃ is the most frequently found VOC in finished drinking water supplies in the United States (ATSDR, 1997a). It and certain other trihalomethanes are formed by reaction of the chlorine added as a disinfectant with natural organic compounds present in the water. Levels of solvents found in drinking water in the United States are typically in the nanogram per liter (ppt) to microgram per liter (ppb) range, although concentrations in the low milligram per liter (ppm) range are found in water from wells situated in solvent plumes from hazardous waste sites and other point discharges. Of the thousands of chemicals found at hazardous waste sites, six of the 10 most commonly present in groundwater are solvents (Pohl et al., 2008).

People are subjected to solvents in environmental media by inhalation, ingestion, and skin contact. Considerable effort was devoted by the EPA, from 1979 to 1985, to assess personal exposure to solvents in different locales of the United States. TRI, PERC, benzene, xylenes, and ethylbenzene were most frequently found in highest concentrations (~1–32 μg/m³) in air (Wallace et al., 1987). Exhaled breath levels of some chemicals (eg, CHCl₃, ethylbenzene) corresponded to indoor air levels. Personal activities (eg, smoking, visiting a dry cleaner or service station) and occupational exposures were thought to be largely responsible for exposures to other VOCs (eg, benzene, toluene, xylenes, and PERC) (Ashley et al., 1994). Subsequent studies, accounting for all pertinent exposure pathways, were conducted by the EPA in the mid-1990s (Gordon et al., 1999). Elliott et al. (2006) more recently reported indoor exposures to VOCs did not have adverse effects on the respiratory function of 953 adults surveyed, with the exception of 1,4-dichlorobenzene, a component of air fresheners, toilet bowl deodorants, and mothballs. It should be recognized that CHCl₃ and other VOCs volatilize to some degree during home water usage, particularly when the water is heated. Thus, a significant proportion of one’s total exposure to VOCs in tap water can occur via inhalation (Weisel and Jo, 1996), although the contribution of dermal exposure is relatively modest (Thrall et al., 2002).

Toxicokinetic (TK) studies are playing an increasingly important role in reducing uncertainties in risk assessments of solvents (Andersen, 2003; Dixit et al., 2003; Clewell et al., 2004). The fundamental goal of TK studies is to delineate the uptake and disposition of chemicals and certain of their metabolites in the body. It is now recognized that toxicity is a dynamic process, in which the degree and duration of injury of a target tissue are dependent on the net effect of toxicodynamic (TD) and TK processes including systemic absorption, tissue deposition, metabolism, interaction with cellular components, elimination, and tissue repair. Estimation of risk of toxicity from TK data is based on the fundamental concept that the intensity of response from an administered dose is dependent on the amount(s) of biologically active chemical moiety(ies) present in a target tissue (ie, the tissue dose). A related concept is that the tissue dose in a given target organ in one species will have the same degree of effect as an equivalent target organ dose in a second species. This concept of tissue dose equivalence appears to be applicable in many cases to solvents (Andersen, 1987), although TD differences are possible. Gaining an understanding of how the processes that govern solvent kinetics vary with dose, route of exposure, species, and even different individuals greatly reduces the number of assumptions that have to be made in assessment of health risks from exposure and effect data.

Volatility and lipophilicity are two of the most important properties of solvents that govern their absorption and deposition in the body. Most solvents are volatile under normal usage conditions, although volatility varies from compound to compound. Lipophilicity can also vary substantially, from quite water soluble (eg, glycols, esters, and alcohols) to quite lipid soluble (eg, halocarbons and aromatic hydrocarbons). Many solvents of particular concern at present are relatively lipid soluble and volatile, hence their designation as VOCs. These compounds have relatively low molecular weight and are uncharged. Thus, they pass freely through membranes from areas of high to low concentration by passive diffusion.

Absorption

The majority of systemic absorption of inhaled VOCs occurs in the alveoli, although limited absorption has been demonstrated to occur in the upper respiratory tract (Stott and McKenna, 1984). Gases in the alveoli are thought to equilibrate almost instantaneously with blood in the pulmonary capillaries. Blood:air partition coefficients (PCs) of VOCs are important determinants of the extent of their uptake. A PC is defined by Gargas et al. (1989) as the ratio of concentration of VOC achieved between two different media at equilibrium. As blood is largely aqueous, the more hydrophilic solvents have relatively high blood:air PCs, which favor extensive uptake. Gargas et al. (1989) determined in vitro PCs for 55 VOCs in F344 rats. Human blood:air values were measured for 36 of the compounds. Because VOCs diffuse from areas of high concentration to those of low concentration, increases in respiratory rate (to maintain a high alveolar concentration) and in cardiac output/pulmonary blood flow (to maintain a large concentration gradient by removing capillary blood containing the VOC) enhance systemic absorption.

Systemic uptake of solvents during ongoing inhalation exposures is dependent on tissue loading and metabolism, in addition to the factors noted above. Percent uptake is initially high, but progressively declines as the chemical accumulates in tissues, and the level of chemical in venous blood returning in the pulmonary circulation increases. A near steady state, or equilibrium, will be reached on inhalation of a fixed concentration of lipophilic solvents. The approach to equilibrium is asymptotic. Despite continued inhalation of lipophilic solvents, concentrations in the blood and tissues (other than fat) generally only increase modestly. Percent uptake remains relatively constant for the duration of exposure, with metabolism and accumulation in adipose tissue largely responsible for the continuing absorption. Hydrophilic solvents take considerably longer to reach steady state, due to the extended time required for equilibration of chemical in inspired air with that in total body water.

Solvents are well absorbed from the GI tract. Peak blood concentrations are observed within a few minutes of dosing fasted subjects, although the presence of fatty food in the GI tract can significantly delay absorption of lipophilic solvents. It is now usually assumed that 100% of an oral dose of most solvents is absorbed systemically. The vehicle or diluent in which a solvent is ingested can affect its absorption and TK. Kim et al. (1990a,b), for example, found that a corn oil vehicle served as a reservoir in the gut to delay GI absorption of carbon tetrachloride (CCl₄) in rats. Although bioavailability of CCl₄ given in corn oil and in an aqueous emulsion was the same, peak blood CCl₄ levels and acute hepatotoxicity were much lower in the corn oil group. Other factors that can influence the oral absorption of certain classes of chemicals have relatively little influence on most lipophilic solvents, due to their rapid passive diffusion in the entire GI tract.

Dermal absorption of solvents can result in both local and systemic effects. Skin contact with vapors and concentrated solutions of solvents is a common occurrence in the workplace. Dermal contact with solvent contaminants of water can also occur in the home and in recreational settings (Weisel and Jo, 1996; Gordon et al., 2006). Lipophilic solvents penetrate the stratum corneum, the skin’s barrier to absorption, by passive diffusion. Important determinants of the rate of dermal absorption of solvents include the chemical concentration, surface area exposed, exposure duration, integrity and thickness of the stratum corneum, and lipophilicity and molecular weight of the solvent (EPA, 1998a). Skin penetration can be quantified in laboratory animals and humans by a variety of in vitro and in vivo techniques (Morgan et al., 1991; Nakai et al., 1999). Although absorption rates measured by these methods may vary numerically, there is often good agreement in their relative ranking of chemicals’ ability to penetrate the skin. Dermal permeability constants are typically two to four times lower for human than for rodent skin (McDougal et al., 1990). The extent of dermal absorption in occupational and environmental exposure settings should be taken into account when conducting risk assessments of solvents.

Transport and Distribution

Chemicals and nutrients absorbed into portal venous blood from the GI tract can be removed from the bloodstream by first-pass, or presystemic elimination. As VOCs are rapidly absorbed and human intestinal epithelium apparently does not contain cytochrome P450 2E1 (CYP2E1) (Paine et al., 2006), the primary CYP isoform that metabolizes low doses of many VOCs, their first-pass GI metabolism should be negligible. Blood in the portal venous circulation passes through the liver before entering the arterial circulation. Solvents are also subject to exhalation by the lungs during their first pass through the pulmonary circulation. Therefore, solvents that are well metabolized and quite volatile are most efficiently eliminated before they reach the arterial blood. The efficiency of the hepatic first-pass elimination is thus dependent on the chemical, as well as the rate at which it arrives in the liver. Pulmonary first-pass elimination, in contrast, is believed to be a zero-order process, as a fixed percentage of the chemical is thought to exit the pulmonary blood at each pass through the pulmonary circulation. Indeed, Lee et al. (1996) demonstrated in rats that pulmonary first-pass elimination of TCE was relatively constant over a range of doses, but that hepatic elimination was inversely related to dose. Thirty years ago Andersen (1981) theorized that the liver was capable of removing “virtually all” of an orally administered VOC, if the amount in the portal blood was not great enough to saturate metabolism. Liu et al. (2009) could not detect TCE (limit of quantitation = 25 pg/ml) in the arterial blood of rats gavaged with 0.1 μg TCE/kg in an aqueous emulsion, but did find TCE at bolus doses ≥1 μg/kg. Interestingly, bioavailability was consistently ~12% in the dosage range of 100 to 5000 μg TCE/kg. These findings indicate that first-pass elimination affords extrahepatic organs substantial, if not complete protection from environmentally encountered levels of well-metabolized VOCs.

Solvents are transported by the arterial blood and taken up by tissues according to their blood flow and mass and tissue:blood PCs (Astrand, 1983). Relatively hydrophilic solvents solubilize in plasma. Nevertheless, as much as 50% of such chemicals may still be carried by erythrocytes (Lam et al., 1990). These researchers found that lipophilic solvents do not bind appreciably to plasma proteins or hemoglobin, but partition into hydrophobic sites in these molecules. Lipophilic solvents also partition into phospholipids, lipoproteins, and cholesterol present in the blood. The brain is an example of a rapidly perfused tissue with a relatively high lipid content. Lipophilic solvents therefore quickly accumulate in the brain after the initiation of exposures (Warren et al., 2000). Inhalation of halothane, TCE, and CHCl₃ can produce CNS effects as profound as surgical anesthesia within as little as one to two minutes. Redistribution to poorly perfused, lipoidal tissues will subsequently occur. Adipose tissue gradually accumulates relatively large amounts of VOCs, and then slowly releases them back into the blood, due to high fat:blood PCs and low blood perfusion rates.

Route of exposure can significantly influence target organ deposition and toxicity of solvents. Much of the pre-1980s toxicology database for solvents comprised results of inhalation studies. Inhalation is the major route of occupational exposure to these chemicals. Since then, much attention has been focused on potential health effects of VOC contaminants of drinking water. Due to the initial paucity of oral data, regulatory agencies sometimes extrapolated directly from inhalation data to predict risks of ingested VOCs. Such a practice is obviously not scientifically valid, when physiological differences in the absorption pathways are taken into account. All the cardiac output passes through the pulmonary circulation versus ~20% for the GI tract. Also, the alveolar surface area for absorption is approximately 20 times that of the entire GI tract in humans. VOCs absorbed in the lungs directly enter the arterial circulation and are transported throughout the body. In contrast, VOCs absorbed from the GI tract largely enter the portal circulation and are subject to first-pass elimination by both the liver and the lungs. The GI epithelium contains certain monooxygenases, notably CYP3A in rats (Mitschke et al., 2008) and CYP3A4 in humans (Ding and Kaminsky, 2003), although these isozymes do not participate to a significant extent in metabolism of most solvents. The liver, however, takes up and metabolizes a portion of many VOCs following ingestion (Sanzgiri et al., 1997). It is not surprising then that extrahepatic organs receive a relatively high dose on inhalation of VOCs.

The pattern of ingestion of solvents can significantly influence their TK and health effects. For convenience, test chemicals are typically given daily to animals as a single bolus by gavage in short- and long-term oral toxicity and carcinogenicity studies. Actual human exposures to solvents in drinking water are quite different, in that people typically ingest water in divided doses. High, daily gavage doses of CHCl₃ and other halocarbons have produced hepatocellular carcinoma in B6C3F1 mice. No evidence of hepatic tumorigenesis was seen, however, when these mice were given the same doses of the chemical in their drinking water (Jorgenson et al., 1985; Klaunig et al., 1986). Female B6C3F1 mice exhibited hepatocellular necrosis and regenerative hyperplasia when gavaged with CHCl₃. Consumptions of the same doses in water were without ill effect (Larson et al., 1994). Similarly, the maximum blood level of CCl₄ was found to be much higher and hepatotoxicity more severe in rats dosed with CCl₄ by gavage, than in those given the same dose over two hours by gastric infusion (Sanzgiri et al., 1995). Oral bolus doses of solvents can cause damage of extrahepatic organs by exceeding the capacity of hepatic and pulmonary first-pass elimination, as well as protection and repair processes of hepatocytes.

Elimination

The rate of systemic elimination of different solvents varies considerably. The two major routes of systemic elimination of VOCs are metabolism and exhalation. The rate and extent of metabolic clearance are dose- and compound-dependent. Exhalation is determined largely by the rate of pulmonary blood flow, the chemical’s blood:air PC, and the alveolar ventilation rate. The more volatile, lipophilic VOCs are exhaled the most readily, because they have the lower blood:air PCs (Gargas et al., 1989). A good “case in point” is the comparison of toluene with acetone. The extent of CNS depression caused by each chemical is dependent on the concentration of the parent compound present in the brain (Bruckner and Peterson, 1981). As can be seen in Fig. 24-3, recovery of rats from toluene anesthesia occurs within minutes of exposure cessation. This can be attributed to redistribution of toluene from the brain to body fat and other tissues, as well as to relatively rapid metabolism and exhalation. In contrast, recovery from acetone narcosis does not occur for at least nine hours postexposure. As acetone is water soluble, limited amounts are deposited in the brain. It is instead distributed in the considerable volume of the blood and other body water. Clearance of acetone is slow due to its large volume of distribution and its relatively slow metabolism and exhalation. Acetone’s water solubility is responsible for its relatively high blood:air PC and retention in the blood. Thus, acetone is available for diffusion into the brain and induction of a modest degree of CNS depression for a prolonged time.

Body fat plays an important role in the elimination of lipophilic solvents. Blood concentrations of such solvents drop very rapidly during the initial elimination phase following cessation of exposure. This so-called redistribution phase is characterized by rapid diffusion of solvents from the blood into tissues. Equilibration of adipose tissue is prolonged due to the small fraction of cardiac output (~3%) supplying fat depots. Body fat increases the volume of distribution and total body burden of lipophilic solvents. De-equilibration from adipose tissue during the terminal elimination phase is prolonged, due to low blood flow and high fat:blood PCs. Sato et al. (1975) demonstrated slower elimination of benzene from female than from male rats and humans, due to the females’ higher body fat content. Rats with excess body fat experienced leukopenia during chronic benzene exposure, whereas their thinner counterparts did not.

Metabolism

Biotransformation plays a key role in modulating the toxicities of many solvents. Organic solvents are poorly soluble in water. Certain cellular enzymes can convert them to relatively water-soluble derivatives, which may be more readily eliminated in the largely aqueous urine and/or bile. Conversion of a bioactive parent compound to a less bioactive or inactive metabolite(s) that is(are) efficiently eliminated is termed metabolic inactivation, or detoxification. Toluene, for example, accumulates in neuronal membranes and inhibits their functions. It is metabolized to hydroxyl and carboxyl metabolites, which are too polar to accumulate or remain in substantial quantities in neuronal membranes. Thus, metabolism serves to detoxify and to accelerate the elimination of toluene. Metabolism of other solvents can produce reactive metabolites that are cytotoxic and/or mutagenic. This phenomenon is known as metabolic activation, or bioactivation. Benzene, for example, is oxidized to a variety of epoxides, quinones, and semiquinones that can produce hematopoietic toxicities and leukemia (Snyder, 2004). Benzene and many other VOCs are converted via multiple metabolic pathways to products of varying toxicity. Some of these competing pathways are considered bioactivation, others detoxification pathways. A variety of factors can influence the prominence of the different pathways and hence alter toxicity outcomes.

The initial step in biotransformation of relatively lipophilic solvents is catalyzed primarily by microsomal CYPs. They are a “superfamily” of hemoproteins that act as terminal oxidases of the mixed-function oxidase (MFO) system. Their mode of action is described in Chap. 6. CYPs can catalyze a number of oxidative reactions, as well as certain reductive reactions (Omiecinski et al., 1999). Fifty seven different human CYP isoforms have been identified, but six (CYPs 2D6, 2C9, 3A4, 1A2, 2C19, and 2E1) account for the majority of xenobiotic metabolism (Rendic, 2002; Guengerich, 2006). CYPs are generally thought to have broad, overlapping substrate specificities, as demonstrated by Kobayashi et al. (2002) in rats. Xenobiotic metabolism under physiological conditions may favor one or two isoforms as the primary catalysts for a given chemical. Different isoforms can predominate at different concentrations of a chemical.

The outcome of exposure to a potentially toxic solvent can depend on the relative abundance of CYP isoforms. Expression of CYPs can differ considerably as a result of genetic polymorphisms and exposure to chemicals that induce or inhibit specific isoforms. A number of factors have been found to be sources of individual variability in CYP induction (Tang et al., 2005). Inheritable gene alterations, such as base changes and deletions, can result not only in functionally deficient enzymes but also in the absence of certain P450s (Ma et al., 2004). As CYPs have different modes of regulatory control, their expression in response to various inducers and inhibitors varies. It should also be recognized that frequently there are interspecies differences in the presence, regulation, and catalytic activities of CYPs (Lewis et al., 1998; Guengerich, 2006).

There has been a focus on identification of CYPs that participate in the metabolism of individual solvents. Nakajima et al. (1992a) found that low concentrations of TCE in rat liver were metabolized primarily by CYP2E1, a constitutive high-affinity, low-capacity isoform. CYP2B1/2, a low-affinity, high-capacity form, predominated under high-dose (substrate) conditions. CYP2E1 was observed to be a major contributor to the metabolism of TCE and benzene, but not toluene. CYP1A1 catalyzed the formation of o-cresol, but not benzyl alcohol from toluene. CYP1A1 oxidized TCE in mice but not in rats. Thus, CYP isoforms exhibit species selectivity, substrate selectivity, and regioselectivity for solvents (Nakajima, 1997).

CYP2E1 is a major catalyst of the oxidation of a variety of low-molecular-weight xenobiotics, including many solvents. It is found predominantly in liver, with lower levels present in kidneys, lungs, brain, testes, and other extrahepatic tissues of rodents and humans. Guengerich et al. (1991) reported that human CYP2E1 is primarily responsible for oxidation of some 16 halogenated and aromatic hydrocarbons, including benzene, styrene, CHCl₃, TCE, and vinyl chloride. These compounds are oxidized to electrophilic metabolites, capable of causing cytotoxicity and/or mutagenicity (Raucy et al., 1993). The isozyme is also responsible for reduction of CCl₄ to free radicals (see the discussion of CCl₄ in this chapter). CYP2E1 activity is associated with the pathogenesis of alcoholic cirrhosis, through formation of highly reactive oxygen radicals and acetaldehyde from ethanol (Caro and Cederbaum, 2004; Gonzalez, 2005; Lu and Cederbaum, 2008). CYP2E1 is inducible by ethanol, acetone, pyridazine, chlorzoxazone, isoniazid, and other of its substrates. Activity of the isoform varies from species to species and from human to human (Snawder and Lipscomb, 2000). Overnight fasting is widely recognized to induce CYP2E1, due in part to release of acetone during lipolysis (Bruckner et al., 2002). Starving rats, however, apparently must engage in coprophagy for induction to occur (Chung et al., 2001). Such differences in CYP2E1 activity/levels may play a major role in susceptibility to the toxicity and carcinogenicity of ethanol and a number of other solvents.

A variety of environmental factors can predispose people to harmful effects of solvents by altering P450s (Lof and Johanson, 1998). Ethanol is an effective CYP2E1 inducer when ingested repeatedly in substantial amounts (see the section “Exogenous Factors”). The person who drinks in such a manner may be subject to potentiation of solvent toxicity, due to increased solvent metabolic activation (Manno et al., 1996). Conversely, ethanol consumed at about the same time as a solvent exposure can be protective by competitively inhibiting metabolic activation of the solvent. Thus, the timing of exposures is important. Folland et al. (1976) reported severe CCl₄-induced hepatorenal injury in workers at an isopropyl alcohol bottling plant. These subjects’ preexposure to the alcohol, which is metabolized to acetone, had markedly induced CYP2E1. Their induced condition resulted in a substantial increase in the metabolic activation and cytotoxicity of CCl₄, on inhalation of normally nontoxic concentrations of the halocarbon.

The metabolic basis of solvent interactions receives considerable attention. The consequence of CYP2E1 induction by one solvent or drug depends on the nature of the second solvent. Kaneko et al. (1994) reported that ethanol pretreatment of rats has a more profound effect on the TK of TRI (poorly metabolized) than of TCE (well metabolized). Similarly, ethanol pretreatment of rats results in a greater increase in metabolism and hepatotoxicity of CCl₄ (relatively poorly metabolized) than of CHCl₃ (well metabolized) (Wang et al., 1997). As capacity-limited (ie, metabolic capacity of the liver) metabolism prevails for poorly metabolized solvents, irrespective of dose, induction of their metabolism could be toxicologically significant in high (occupational) or low (environmental) exposure settings. In contrast, alterations of the bioactivation of low doses of well-metabolized solvents should be of little consequence, as their biotransformation is perfusion (ie, hepatic blood flow) limited. Lipscomb et al. (2003), for example, utilized a PBTK model to calculate that a sixfold interindividual difference in TCE metabolism resulted in <2% change in the total amount of TCE oxidized by persons inhaling only 50 ppm for eight hours or drinking 2 L of water containing 5 ppb TCE. Under such circumstances, interindividual metabolic differences should have little influence on cancer risks posed by trace levels of such solvents.

Physiological Modeling

PBTK models have been developed to provide predictions of chemical concentrations in blood and tissues as a function of time. They are thus used to relate the administered, or external, dose to the blood and/or tissue (internal) dose of parent compound and/or bioactive moiety. If there is sufficient knowledge of the physiology of the test animal/tissue and interactions of the bioactive moiety with cellular components and ensuing responses, physiologically based toxicodynamic (PBTD) models can be developed (Conolly, 2002). Relatively few PBTD models have been published to date for solvents. Tan et al. (2003) linked an existing PBTK model for CHCl₃ with data for hepatocellular death and regenerative hyperplasia to predict liver tumor response in mice. Luke et al. (2010) expanded upon the work of Tan et al. (2003) by developing biologically based dose–response (BBDR) models for CHCl₃ and CCl₄, based on key events of cytotoxicity and proliferation, to forecast liver tumor incidence in female B6C3F1 mice. The BBDR models consisted of a chemical-specific PBPK model for each chemical linked to a PBPD model of cytotoxicity and cellular proliferation, linked in turn to a clonal growth model to predict tumor incidence. Tan et al. (2011) have recently utilized models to evaluate the likelihood of TK and TD interactions of different chemical mixtures.

Development of PBTK models for anesthetic gases was initially undertaken in the 1920s and 1930s. Modeling was used in the design of experiments and dosage regimens for drugs, but the practice was not extended to industrial chemicals until the 1970s and 1980s. A “benchmark” paper by Ramsey and Andersen (1984) described the development of a PBTK model for styrene. By 1995, PBTK models had been developed for ~50 chemical environmental contaminants (Leung and Paustenbach, 1995). Approximately 28 of these 50 chemicals were solvents. Andersen (2003) more recently identified approximately 700 papers that dealt with PBTK models and/or their applications to toxicology and risk assessment. Thus, physiological models are being developed and used with increasing frequency to incorporate dosimetry and mechanistic data into noncancer and cancer risk assessments (Krishnan and Johanson, 2005).

Physiological modeling accounts for species- and dose-dependent shifts from linear to nonlinear kinetics, which impact tissue doses of bioactive moiety or moieties. The models are well suited for species-to-species extrapolations, because human physiological and metabolic parameter values can be inserted into animal models, and simulations of target tissue doses and effects in humans generated. Thus, solvent exposures necessary to produce the same target organ dose in humans, as those found experimentally to cause unacceptable cancer or noncancer risks in test animals, can be determined for some chemicals with reasonable certainty. In the limited number of cases where there may be species differences in tissue sensitivity, PBTD models must be used to forecast toxicologically effective target organ doses.

The influence of interindividual variability in model parameters has been receiving considerable attention (Price et al., 2003). Jonsson and Johanson have performed model fitting in a Bayesian framework using Markov chain Monte Carlo (MCMC) simulation to reduce parameter uncertainty for methylene chloride (MC) (2001) and styrene (2002). A Bayesian analysis of a PBPK model for TCE yielded improved agreement of simulated and empirical TK data (Hack et al., 2006). Covington et al. (2007) utilized this methodology to estimate that only ~2% of low inhaled doses of PERC are metabolized by humans to TCA. Qui et al. (2010) subsequently reported a similar finding when they used the approach in an expanded and more detailed assessment of PERC kinetics and TCA production. The percentage is far lower than has been assumed by regulatory agencies in the absence of data. Hierarchial analyses are proving quite useful in characterization of model parameter uncertainty and variability, and thereby in providing improved predictions of internal dosimetry for risk assessments.

PBTK modeling is finding increasing use in estimating exposure levels that result in concentrations of solvents measured in blood, exhaled breath, or urine (ie, biomarkers) in humans. Bois (2001) pointed out the need to apply PBTK models to large and poorly characterized human populations that have highly variable exposures, activity, physiology, and kinetics. A procedure termed reverse dosimetry has been employed to reconstruct population exposures from limited biomarker data (Clewell et al., 2008). Reverse dosimetry, integrating PBPK modeling with exposure pattern characterization, Monte Carlo analysis, and other statistical tools, has been successfully employed to estimate distributions of exposures consistent with concentrations of TCE (Sohn et al., 2004), CHCl₃ (Tan et al., 2006), and other trihalomethanes (Tan et al., 2007) quantified in human blood or exhaled breath. These exposure distributions can be compared with occupational and environmental exposure guidelines, in order to evaluate potential health risks in different situations.

There is considerable variability in responses of humans to solvents and other xenobiotics. Although information on interindividual differences in the kinetics and toxicity of solvents is usually limited, some significant differences are reported and should be taken into account in risk assessments (Grassman et al., 1998; Hattis et al., 1999). Children within specific age groups appear to be even more variable than adults (Hattis et al., 2003). Current EPA cancer and noncancer risk assessment practices are generally considered to be quite protective of potentially vulnerable subgroups, due to the use of multiple uncertainty factors and conservative default assumptions.

Endogenous Factors

Children

Currently, there is considerable interest in regulatory and scientific issues related to protection of children’s health (Landrigan and Goldman, 2011). The potential sensitivity of infants and children to pesticides and other chemicals was brought to the public’s attention by a National Academy of Sciences’ report (NAS, 1993). Ensuing Presidential Executive Order 13045 in 1997 directed all federal agencies to make protection of children a high priority and to consider risks to children when making policies and setting standards. The resulting Food Quality Protection Act of 1996 dictated that an additional 10-fold safety/uncertainty factor be used in pesticide risk assessments when pediatric toxicity and TK data are unavailable. This is almost always the case for insecticides, solvents, and other chemicals. Application of the 10-fold children’s safety factor to nonpesticides has been considered. The Pediatric Research Equity Act of 2003 requires pharmaceutical companies seeking FDA’s approval of new drugs to assess a compound’s safety and efficacy at different doses in pediatric patients or surrogates.

Very little information is available on the toxic potential of solvents and most other chemicals in infants and children. There are several well-known examples of compounds (eg, lead, thalidomide, chloromycetin, and diazepam) to which fetuses or neonates are particularly susceptible. Results of experiments with chlorpyrifos (CPF) in immature rats raised enough concern about possible neurodevelopmental disorders in children that use of a number of organophosphates was severely restricted, and they have largely been removed from the home-use market (Colborn, 2006). Compilations of maximally tolerated doses (MTDs) of chemotherapeutic drugs are among our most definitive data on age-dependent toxicity in humans. Clinical trials of diverse groups of antineoplastic drugs have revealed that their toxic effects are often similar qualitatively, and usually differ quantitatively by a factor of ≤2 (NAS, 1993). MTDs for anticancer drugs in children are frequently higher than those for adults (Bruckner, 2000). Ginsberg et al. (2002) assembled a database with which TK parameters for adults and children can be compared for 45 drugs. These developmental metabolic profiles are relevant to some solvents and other environmental chemicals.

Increasing numbers of toxicology studies are being carried out in immature laboratory animals. Some of the most comprehensive data sets available are compilations of LD₅₀ values from experiments with immature rodents (Goldenthal, 1971). These and other data reveal that immature rodents are more sensitive to most but not to all chemicals (Hasegawa et al., 2007). Variances are usually no more than two to threefold, but can be as much as 10-fold. The younger and more immature a subject, the more different its response from that of the adult. The newborn mouse and rat more nearly resemble the human fetus during the third trimester than during the human neonate’s initial weeks and months of life. Thus, it is not surprising that newborn rodents would be much more sensitive to many toxicants than adult rodents and human newborns. Growth and maturation are, of course, much more rapid in rodents. A few days’ growth of mice and rats can result in substantial changes in metabolism, disposition, and toxicity (Ginsberg et al., 2004a,b). It is generally accepted that there are windows of vulnerability, or periods during the development of the nervous, audiovisual, immune, reproductive, endocrine, and other organ systems when the intricate, fragile maturational processes may be particularly sensitive to xenobiotics (Dorman et al., 2001; Scheuplein et al., 2002). These windows occur at different ages and last for different periods of time in different species. This makes the choice of an appropriate animal model for children complex and cross-species comparisons more difficult than usual with adult risk assessments (Ginsberg et al., 2004a,b; Morford et al., 2004).

Despite a paucity of data, some logical assumptions can be made about age-dependent changes in the TK and toxicity of solvents. GI absorption of solvents would not be expected to vary with age, as most solvents are absorbed rapidly and completely by passive diffusion from the entire GI tract. Systemic absorption of inhaled solvents may be greater in infants and children. Their cardiac output and respiratory rates are relatively high, although their alveolar surface area is lower than that of adults (Sarangapani et al., 2003). Reduced plasma protein binding in neonates and infants may be of consequence for certain solvent metabolites. Low-binding capacity for TCA, for example, could result in more of the mouse hepatocarcinogen being available for hepatic uptake, as well as for renal clearance. Extracellular water, expressed as percent of body weight, is highest in newborns and diminishes through childhood. The larger volume of distribution for water-soluble solvents results in their slower clearance and longer duration of action. Body fat is high from approximately six months to three years of age, and then steadily decreases until adolescence, when it increases again in females. As lipophilic solvents accumulate in adipose tissue, more body fat would result in higher body burdens, slower clearance, and longer duration of action.

Changes in xenobiotic biotransformation during maturation may have the greatest impact on susceptibility to chemical toxicity and carcinogenicity. Poisonings of premature and term human newborns by benzyl alcohol, hexachlorophene, diazepam, chloramphenicol, and certain other compounds are primarily attributable to deficits in hepatic Phase II metabolic (ie, conjugative) capacity. Maturational studies of human liver reveal that CYP isoforms are expressed asynchronously (Cresteil, 1998). There have been comprehensive efforts to characterize the ontogeny of human hepatic Phase I (Hines and McCarver, 2002; Stevens et al., 2003; Johnsrud et al., 2003) and Phase II (McCarver and Hines, 2002) enzymes from gestation to adolescence. Levels of CYP2E1, the primary catalyst of oxidation of a variety of solvents, are very low in fetal liver, increase steadily for the first year, and remain relatively constant through adolescence (Cresteil, 1998; Johnsrud et al., 2003). Blanco et al. (2000) failed to find an age-related difference in the maximal in vitro activity of CYP2E1 or other CYP isoforms in liver microsomes from donors from infancy to 93 years old. More rapid systemic clearance of many compounds by young children than by adults (Ginsberg et al., 2002) is generally believed to result from the child’s liver’s larger volume and more rapid perfusion rate (Murry et al., 1995; Sarangapani et al., 2002). Glomerular filtration and renal tubular secretion are quite low at birth, but increase substantially during the first six months to one year (Scheuplein et al., 2002). The half-life of creatinine, a substance eliminated entirely by glomerular filtration, is about half the adult value at one year. The net result of multiple age-related TK changes, some with offsetting effects, is difficult to predict.

A number of authorities have advocated the use of PBTK/TD models to collectively account for the unique and changing characteristics of children (Clewell et al., 2004; Daston et al., 2004). A few PBTK models have been published for predicting the internal dosimetry of chemicals in immature rats. Timchalk et al. (2007) developed a BBDR model for chloropyrifos (CPF), a widely used organophosphate insecticide. More pronounced depression of erythrocyte and brain cholinesterase activities, apparently resulting from higher CPF-oxon blood and brain levels, was forecast in preweanling than in adult rats. Rodriguez et al. (2007) used PBTK modeling to forecast the TK of six inhaled VOCs (TCE, PERC, benzene, CHCl₃, MC, and methyl ethyl ketone) in 10-day-old and adult rats. VOC blood levels were forecast to be up to 3.8-fold higher in the preweanlings, due primarily to their lower metabolism and faster respiration. Many of the input parameters were estimated or scaled, however, rather than measured. Empirical data were not provided to assess the accuracy of the simulations. Tornero-Velez et al. (2010) developed a PBTK model that clarified why preweanling rats are more susceptible than adults to the acute toxicity of deltamethrin (DLM), a pyrethroid insecticide. These researchers were among the first to rely primarily on age-specific input parameters and to demonstrate concordance between measured and simulated timecourses of a bioactive moiety (ie, the neurotoxic parent compound). Blood and brain levels of DLM were inversely related to age in 10-, 21-, 40-, and 90-day-old rats. Sensitivity analyses showed V_maxs (Anand et al., 2006a) for CYP- and carboxylesterase (hydrolytic)–mediated clearance pathways significantly influenced both blood and brain dosimetry. Inefficient metabolic detoxification was thus a major contributor to preweanling rats’ susceptibility to DLM, although immaturities in other kinetic processes may also be factors. Age-specific CYP levels and CPF bioactivation and inactivation in liver microsomes and plasma are currently being measured and used to model organophosphate TK and cholinesterase inhibition in humans (Foxenberg et al., 2011; Smith et al., 2011). Determination and utilization of age-specific input parameters will continue to improve the accuracy of immature human PBTK–TD predictions.

Effort is being devoted to development of PBTK models for VOCs in human infants and children. Clewell et al. (2004) constructed what they termed a life-stage model to simulate blood levels of four solvents (isopropanol, vinyl chloride, MC, and PERC) and their primary metabolites in different life stages. This was intended to be a preliminary model, in view of the lack of validated models for the younger groups and the required number of simplifying assumptions in the face of uncertainties about many age-specific parameters. In general, dose metrics for the chemicals across life stages were within a factor of two, except during the early postnatal period when the largest TK differences occurred. Gentry et al. (2003) took a similar approach to forecast internal dosimetry of vinyl chloride, isopropanol, MC, and PERC in human offspring during pregnancy and lactation. In general, blood concentrations of the VOCs were lower in neonates during lactation than in the fetus during gestation. Fetal/neonatal VOC exposures were generally several orders of magnitude lower than maternal exposures. Sarangapani et al. (2002) developed a PBTK model that integrated key age-specific respiratory tract parameters, such that the disposition of four VOCs (styrene, vinyl chloride, isopropanol, and PERC) could be forecast in children and adults. Differences in blood levels of parent compounds in infants and adults were comparable or varied <2-fold during the first year of life. Nong et al. (2006) incorporated measured age-specific liver volumes and CYP2E1 content into a PBPK model for toluene to predict internal doses (blood levels) of the parent compound in adults and in children of different age groups. CYP2E1 content was the key model input parameter for neonates, whereas liver blood flow rate was predominant after a few months of age. Combined interindividual and interage variability in the areas under blood level versus time curves (AUCs) was within a factor of 2, except for the neonates, who had somewhat higher AUCs.

The aforementioned findings suggest that differences in dosimetry of parent solvents across life stages are usually well within the 3.3 TK component of the classical 10-fold intraspecies uncertainty factor. Relative little work has been done, however, to model solvent metabolite dosimetry in different age groups. Potential PD differences should also be taken into account in children's risk assessments.

Elderly

Age-related changes in effects of xenobiotics have primarily been of interest in the field of drug therapy (Bressler and Bahl, 2003; McLean and LeCouteur, 2004). More attention is now being paid to exposures and vulnerabilities of the elderly to other chemicals (Geller and Zenick, 2005). The EPA encourages evaluation of toxicity across the entire lifespan, with a focus on life stages with their own unique sensitivity (Brown et al., 2008). Data are sorely lacking, however, on the relative susceptibility of geriatric populations to solvents and many other industrial chemicals. The aging CNS undergoes changes (eg, neuronal loss, altered neurotransmitter and receptor levels, and reduced adaptability to effects of toxicants) (Ginsberg et al., 2005), which may predispose to more pronounced neurologic effects by solvents. Older people do appear to be more sensitive to the CNS depressant effects of ethanol, due in part to lower gastric alcohol dehydrogenase (ADH) activity (Seitz et al., 1993). Kiesswetter et al. (1997) observed the most pronounced neurobehavorial effects in older workers with the highest single or mixed organic solvent exposures. Memory, attention, visual perception, and motor skills diminish with aging, even in the absence of chemical exposure. Thus, these findings could be interpreted as merely additive (ie, aging + solvents effects), or as exacerbation during a vulnerable life stage. Paradoxically, a subsequent investigation by Kiesswetter et al. (2000), this time of toluene-exposed rotogravure printers, revealed fewer symptoms and better psychometric performance in older workers with higher exposure levels. A limited number of studies of this phenomenon by other investigators are no more definitive.

Investigations of adverse effects of solvents on old animals have been quite limited. Attention has been primarily focused on age-related susceptibility of rodents to liver damage. Susceptibility varies dramatically from one chemical to another. Rikans and Hornbrook (1997) reviewed studies showing increased susceptibility of geriatric F344 rats to allyl alcohol, but there was no apparent age dependency with bromobenzene. Results varied with species, strain, and gender. CCl₄ hepatotoxicity was less severe in 28-month-old female F344 rats than in young adults. Wauthier et al. (2004) saw no changes in CYP2E1 protein or mRNA in 18-month-old male Wistar rats, but did observe decreased CYP2E1 activity. The investigators attributed this phenomenon to oxidative stress on the isoform. More severe CCl₄-induced DNA damage in hepatocytes from older mice was shown to be due to preexisting oxidative DNA damage and a higher incidence of DNA strand breaks in the older animals (Lopez-Diazguerrero et al., 2005). Thus, it appears normal aging processes may be accentuated by chemical stressors.

There are few pertinent geriatric animal or human toxicology studies, but age-specific factors that govern the TK of solvents have received considerable attention. Clewell et al. (2002) and Sarangapani et al. (2003) compiled comprehensive reviews and evaluations of age-specific TK parameters, although data for many age groups are inadequate. Splanchnic blood flow decreases with age more rapidly than cardiac output. This is unlikely to significantly affect the GI absorption of solvents, as they are readily absorbed by passive diffusion. Dermal absorption could be influenced by loss of integrity of the stratum corneum, reduction in skin surface lipids, and atrophy of the dermal capillary network (Roskos et al., 1989). Total lung capacity does not change with age, but vital capacity (mobile volume) diminishes progressively due to loss of lung elasticity and strength of the respiratory muscles. Residual capacity (fixed volume) increases with progressive narrowing and closure of small airways (Ritschel, 1988). Other alterations that may contribute to lower pulmonary absorption include decreases in alveolar surface area, membrane permeability and capillary blood volume, and thicker alveolar cell membranes (Clewell et al., 2002). Nevertheless, inhalation PBTK model predictions of steady-state blood concentrations of isopropanol, styrene, vinyl chloride, and PERC differed very little among 10-, 15-, 25-, 50-, and 75-year-old humans (Sarangapani et al., 2003).

Age influences the distribution of xenobiotics in the body, as well as their metabolism and elimination (Mangoni and Jackson, 2003). With aging, body fat content usually increases substantially at the expense of the lean mass (ie, skeletal muscle) and body water. Thus, polar solvents tend to reach higher blood levels during exposures. Relatively lipid-soluble solvents accumulate in adipose tissue and are released slowly to sites of action, metabolism, and elimination. Ginsberg et al. (2005) assembled clearance and half-life data for 46 drugs in 4500 subjects. Clearance was typically slower and half-life longer after the age of 60 years, particularly in those 80 to 84 years old. Cardiac output diminishes 30% to 40% between the ages of 25 and 65 years, as do renal and hepatic blood flows (McLean and Le Couteur, 2004). This research group and Clewell et al. (2002) concluded that reduced hepatic clearance of flow- and capacity-limited xenobiotics in the elderly was due primarily to reductions in liver blood flow and liver size. McLean and Le Couteur (2004) presented evidence that aging of the sinusoidal endothelium reduced oxygen delivery and drug transfer from blood to hepatocytes. Schmucker (2005) and Herrlinger and Klotz (2001) did not find significant relationships between aging and CYP activities in human liver. Cotreau et al. (2005) did describe some reports of age-related decreases in CYP3A4, an isoform important in metabolism of a number of drugs in humans, but of limited relevance to solvent metabolism.

Recent findings indicate that toxic compounds, both xenobiotics and certain products of intermediary metabolism, are major contributors to macromolecular damage and dysregulation involved in aging (Zimniak, 2008). Important endogenous toxicants include reactive carbonyl compounds and electrophiles formed by peroxidation of polyunsaturated lipids. Enzymes capable of reducing carbonyl groups and glutathione S-transferases (GSTs) are key protection mechanisms. Lee et al. (2008) characterized age-related changes in the expression of a substantial number of xenobiotic-metabolizing enzymes in the liver of aged male F344 and Brown Norway rats. GST expression diminished with age. Most data, however, indicate that Phase II (conjugation) pathways in humans are not altered by aging (Mitchell et al., 2011). Maurya and Rizvi (2010) reported an increase in plasma GST activity and antioxidant potential in a study of 80 healthy male and female human subjects 18 to 85 years old. It was hypothesized that this was in response to increased production of reactive oxygen moieties with advancing age.

Information is available on certain age-related aspects of xenobiotic clearance other than metabolism. Plasma protein binding of drugs generally remains unaltered or diminishes modestly with age (Grandison and Boudinot, 2000). Hepatic and/or renal dysfunction more often account for altered protein binding. Little information apparently exists on effects of aging on most transporters (Kinirons and O’Mahony, 2004), although Lee et al. (2008) did see diminished expression of efflux transporters in the livers of aging male rats. Renal clearance diminishes with advancing age, due to parallel annual decreases of ~0.65% in glomerular filtration and maximal tubular secretory capacity beyond the age of 30 years. The foregoing TK factors may significantly influence the internal dosimetry of polar solvent metabolites, and in turn certain responses of geriatric populations.

PBTK models should be useful for integrating age-related physiological and metabolic changes into a quantitative framework to forecast their net effect on blood and tissue timecourses of parent compounds and metabolites. Very few such “geriatric” models, however, have been published. McMahon et al. (1994) employed a PBTK model to demonstrate that reduced urinary elimination of benzene metabolites by geriatric mice was due to decreased renal blood flow rather than altered formation of metabolites. Corley et al. (2005c) incorporated a number of age-, gender-, and species-specific physiological and biochemical parameters into a preexisting PBTK model for 2-butoxyethanol (BE), in order to predict the kinetics of BE and butoxyacetic acid (BAA), BE’s major metabolite, in male and female mice and rats of different ages. The life-stage PBTK model of Clewell and colleagues was used to simulate human blood concentrations of several VOCs and their major metabolite (or metabolic rate) from infancy to the age of 75 years (Sarangapani et al., 2003). Blood levels of the parent compounds remained relatively constant during adulthood, but metabolite levels or amounts metabolized frequently varied during the pediatric and geriatric years. In a subsequent publication, blood levels of PERC and its major metabolite, TCA, were predicted to progressively rise during adulthood and old age by a PBTK model. Clewell et al., (2004) attributed this phenomenon to relatively low pulmonary and metabolic clearance, coupled with accumulation of the lipophilic solvent in ample amounts of adipose tissue. Much work remains to be done to refine these generic PBTK models and to integrate them with significant geriatric PD changes (Burton et al., 2005).

Gender

Some of the physiological and biochemical differences between men and women have the potential to alter tissue dosimetry and health effects of certain solvents (Vahter et al., 2007). Clewell et al. (2002) have provided one of the most comprehensive reviews of the potential impact of gender differences on chemicals’ TK. There was no convincing evidence of sex differences in GI absorption rate constants. The investigators located few data on gender-dependent dermal or pulmonary absorption of solvents or other chemicals. Sarangapani et al. (2003) published tables of a number of factors (eg, cardiac output, pulmonary ventilation rate, and pulmonary surface area) that dictate pulmonary absorption of VOCs. These rates and values were generally higher for males, but a PBPK model’s predictions of steady-state blood concentrations for inhaled isopropanol, styrene, vinyl chloride, and PERC were largely sex-independent (Sarangapani et al., 2003).

Distribution of water- and lipid-soluble solvents can vary substantially between men and women (Schwartz, 2003). Most men have more lean body mass and larger body size. Women typically have smaller volumes of distribution for polar solvents, but larger volumes of distribution for lipophilic solvents (Gandhi et al., 2004). Lean body mass decreases from 76% to 52%, and body fat increases from 33% to 49% in females between the ages of 25 and 65 to 70 years. Nomiyama and Nomiyama (1974) found that women retain less inhaled acetone and ethyl acetate than similarly exposed men. Levels of toluene and TCE in expired air of females were lower, reflecting greater fat deposition and retention of these lipophilic chemicals. Similarly, women exhibited significantly higher respiratory uptake of 1,3-butadiene, another lipophilic VOC (Lin et al., 2001b). Clewell and Andersen (2004) utilized their life-stage PBTK model to predict blood levels of PERC and TCA, its major metabolite, in men and women over a lifetime of daily ingestion of 1 μg PERC/kg. The women were predicted to attain significantly higher blood PERC and TCA concentrations, ostensibly due to PERC’s greater storage in fat and its relatively low pulmonary and metabolic clearance.

The major sex differences in P450-mediated hepatic metabolism in rats are not seen in humans or most other mammals. Rinn et al. (2004) did see significant differences in gene expression in the liver, kidney, and reproductive tissues of male and female mice. Some of the differentially expressed genes involve drug and steroid metabolism, but the biological significance of these variances is unknown. No marked gender differences in P450-catalyzed oxidation reactions have been identified in humans (Mugford and Kedderis, 1998). Other investigators have described modestly higher levels and activities of certain liver P450s in males, including CYP2E1, CYP1A2, and CYP2D6 (Schwartz, 2003). Many CYP3A substrates do not exhibit sex-dependent differences in clearance, although some researchers report differences for certain substrates (Cotreau et al., 2005). Gender-specific metabolism data for most hydrocarbon solvents are lacking. Men exhibit somewhat lower blood ethanol levels than women on ingestion of equal doses, due in part to more extensive ADH-catalyzed metabolism of the alcohol by the male gastric mucosa (Frezza et al., 1990). UDP-glucuronyl transferase activity appears to be somewhat higher in males. Women seem to have modestly lower glomerular filtration rates, although information on gender-dependent renal tubular secretion and reabsorption processes is lacking.

A gender-specific PBTK model was developed to assess the relative risks of benzene exposures (Brown et al., 1998). Women were found to have a higher blood:air PC, a higher percent of body fat, and higher maximum rate of benzene metabolism than men. They exhibited higher blood benzene levels and a 23% to 26% increase in benzene metabolism, potentially placing them at greater risk than men with equivalent exposures.

Relatively little is known about potential influences of contraceptives, hormone replacement therapy, or pregnancy on the metabolism and disposition of xenobiotics (Gleiter and Gundert-Remy, 1996). The menstrual cycle with its hormone changes may influence the metabolism of some xenobiotics (Fletcher et al., 1994). Watanabe et al. (1997) found that estradiol changes during the estrous cycle of rats affected both UDP-glucuronyl transferase and NADPH-cytochrome c reductase activities. Experiments by Nakajima et al. (1992b) demonstrated that pregnancy of rats resulted in significant reductions in hepatic CYP content, as well as the metabolism of both toluene and TCE. Pregnancy results in a 40% increase in tidal volume, which may increase the absorption of inhaled VOCs. Plasma volume increases ~50% in pregnant women, resulting in a decrease in albumin concentration and plasma protein binding of many drugs (Fletcher et al., 1994). Cardiac output increases ~50%, due to increases in stroke volume and heart rate (Silvaggio and Mattison, 1994). Uterine blood flow, renal plasma flow, and glomerular filtration rise substantially, although no information on hepatic blood flow is apparently available. These TK factors need to be considered when assessing risks and setting occupational exposure limits for these subpopulations.

The effects of pregnancy on peroxisome proliferator–activated receptors (PPARs) are being investigated. PPARs are transcription factors that belong to the nuclear hormone receptor superfamily. They regulate genes involved in cell differentiation, development, and metabolism. The three identified and described PPAR isoforms are peroxisome proliferator–activated receptor alpha (PPARα), PPARβ/δ, and PPARγ. Among the isoforms, PPARγ has the greatest influence on cellular homeostasis and carcinogenicity. However, all three PPAR isoforms play essential roles in physiological change and development in the fetoplacental unit. Abnormalities in PPAR-regulated pathways may be implicated in reproductive and gestational disease (Toth et al., 2007; Borel et al., 2008). Two TCE and PERC metabolites, TCA and DCA, can induce modest PPARα activation in humans. The combined effects of pregnancy and low-level PPAR activation, if any, are unknown.

Gentry et al. (2003) described an approach for use of PBTK modeling to compare maternal and fetal/neonatal blood and tissue dosimetry during pregnancy and lactation. In general, blood levels of isopropanol, vinyl chloride, MC, and PERC were predicted to be lower in the human neonate during lactation than in the fetus during gestation. Simulated vinyl chloride levels, for example, were four orders of magnitude lower in the neonates. Fetal/neonatal isopropanol levels were forecast to be orders of magnitude lower than maternal levels. Corley et al. (2003) reviewed PBTK/TD modeling of the developing embryo, fetus, and neonate.

Genetics

A variety of genetic polymorphisms for biotransformation have been found to occur at different frequencies in different ethnic groups. Polymorphisms for xenobiotic-metabolizing enzymes may affect the quantity and quality of enzymes and the outcomes of exposures to solvents and other chemicals (Wormhoudt et al., 1999). It is important to note that culturally linked environmental factors also contribute to ethnic differences in metabolism and disposition of solvents and other chemicals. It is often difficult to disentangle the influences of genetic traits from those of different lifestyles, socioeconomic status, and geographic settings.

Ethnic differences in regulating the expression of CYP isozymes and other enzymes are associated with some variations in xenobiotic metabolism and susceptibility to certain cancers (Raunio et al., 1995; Weber, 1999). Shimada et al. (1994) found that Caucasians had higher total hepatic CYP levels than Japanese. Caucasians exhibited higher CYP2E1 activities, as reflected by aniline p-hydroxylation and 7-ethoxycoumarin O-deethylation. Individuals with two linked polymorphisms in the transcription regulatory region of the CYP2E1 gene exhibited greater expression of the isoform. Reported frequencies of these rare alleles were 2% in Caucasians, 2% to 5% in African Americans, and 24% to 27% in Japanese (Kato et al., 1992). Kawamoto et al. (1995) found that a CYP2E1 polymorphism in Japanese workers did not affect metabolism of toluene to benzoic acid, but that CYP1A1 and aldehyde dehydrogenase (ALDH2) polymorphisms did. ALDH2 deficiency can result in elevated CNS levels of toluene and benzaldehyde (Wilkins-Haug, 1997) in toluene-exposed individuals. About half of the Japanese population lacks ALDH2, due to a structural point mutation in the ALDH2 gene. Pronounced interethnic differences in rates of ethanol metabolism have been associated with a number of ADH and aldehyde dehydrogenase (ALDH) polymorphisms (Pastino et al., 2000). Results of investigations of relationships between occurrence of defective alleles for CYP2E1 and cancer incidence have been contradictory. Studies with more statistical power of selected ethnic groups are needed. Molecular markers clearly indicate variances in CYP genes, but biologically plausible mechanisms linking specific genotypes with specific outcomes need to be established.

Ethnic differences have been demonstrated for some Phase II biotransformation reactions. GSTs are a family of enzymes, some of which promote the detoxification of electrophilic metabolites by catalyzing their conjugation with reduced glutathione (GSH). In contrast, GST theta–mediated conjugation with GSH is the first step in metabolic activation of MC to formaldehyde, a putative carcinogen (see the section “Methylene Chloride”). Individuals with a null/null genotype for GST theta are thus at a lower risk of developing cancer from MC (El-Masri et al., 1999; Haber et al., 2002). The prevalence of this genotype ranges from 10% in Mexican Americans to 60% to 65% in Chinese and Koreans (Nelson et al., 1995). GST theta also catalyzes dehalogenation of DCA to glyoxylic acid. DCA, a minor product of TCE and PERC oxidation, is a hepatocarcinogen in high doses in mice and rats. It is cleared more slowly from the systemic circulation of persons with a GST theta polymorphism, possibly resulting in increased risk of liver cancer on TCE or PERC exposure (Shroads et al., 2011). GSTs participate in conjugating relatively small percentages of doses of TCE and PERC, but these pathways can lead to production of cytotoxic, mutagenic metabolites in the kidney. Contrary to findings in a more limited study, Wiesenhutter et al. (2007) saw no association between deletion polymorphisms of GST-T1 and GST-M1 and incidence of renal cell cancer (RCC) in 134 RCC cases (20 with TCE exposure) and 401 matched controls. More recently Moore et al. (2010) studied gene variants in 1097 RCC cases and 1476 controls. A significant association was found between TCE exposure and RCC in individuals with at least one intact GST-T1 allele, but not in individuals with two deleted alleles.

Genetic polymorphisms, variable transporter activities, genetic variants of receptors and regulatory proteins, and environmental factors can play roles in individual variability in CYP induction (Tang et al., 2005). Lipscomb et al. (1997) described substantial differences in activity and/or content of CYP2E1 within the US population. Lipscomb (2004) incorporated extremes in CYP2E1-mediated metabolism into a PBTK model for TCE and found there was very little impact on TCA dosimetry. Intersubject differences in hepatic blood flow did have a pronounced effect on TCE biotransformation, as TCE is an extensively metabolized (blood flow limited) compound. Wenker et al. (2001) similarly reported that interindividual genetic polymorphisms in CYP2E1 and several other CYPs did not affect the metabolism of styrene, another flow-limited compound. Kedderis (1997) utilized a PBTK modeling approach to confirm this phenomenon, and found the opposite to be true for poorly metabolized (capacity limited) VOCs such as PERC and CCl₄.

Exogenous Factors

P450 Inducers

Considerable effort has been devoted to investigation of effects of different enzyme inducers on xenobiotic metabolism and toxicity. Preexposure to drugs and other compounds, which induce CYP isoforms that metabolically activate chemicals, can markedly potentiate the chemicals’ toxicity/carcinogenicity (Gut et al., 1993; Ioannides and Lewis, 2004; Dekant, 2009). As described in the subsection “Metabolism” in the section “Toxicokinetics,” CYP2E1 is a high-affinity, low-capacity isoform responsible for activation of a wide variety of low-molecular-weight hydrocarbon and halocarbon solvents to cytotoxins and mutagens (Guengerich et al., 1991). A number of CYP2E1 substrates (eg, ethanol and other alcohols, acetone and other ketones, pyridine and pyridazine, isoniazid, acetaminophen, chlorzoxazone) can act as inducers of the isoform. The minimal hepatocytotoxicity caused by a low dose of CCl₄ alone (Panel B, Fig. 24-4) can be contrasted with the centrilobular necrosis (Panel C, Fig. 24-4) in rats pretreated with 2-butanol before the same CCl₄ dose (Traiger and Bruckner, 1976). CYP2E1 is present in many tissues, as are other CYPs that metabolize solvents. Although levels/activities are highest in liver, induction in extrahepatic tissues can result in increased metabolic activation in situ that may be toxicologically significant. It is worthy of note that many chemicals that induce CYPs also induce detoxifying enzymes (eg, epoxide hydrolase, glucuronyltransferases, and sulfotransferases). Thus, the net outcome of preexposure to an inducing agent on solvent toxicity should be addressed on a case by case basis. A number of naturally occurring organosulfur compounds in allium vegetables have been demonstrated to induce Phase II enzymes in the liver, kidney, intestines, and lungs of Wistar rats (Guyonnet et al., 1999).

It should be recognized that timing of exposures to ≥2 CYP substrates is important. Sato et al. (1991) demonstrated that ethanol competitively inhibited the biotransformation of TCE when the exposures were concurrent. Alternatively, preexposure to the alcohol induced TCE metabolism. As described previously in the subsection “Metabolism” in the section “Toxicokinetics,” increased metabolic capacity in an individual is of little consequence for environmentally encountered levels of blood flow–limited (ie, extensively metabolized) solvents (eg, TCE, vinyl chloride, MC, and CHCl₃). Thus, genetically predisposed or induced individuals should be at no greater cancer risk from oxidative metabolites of low doses of these solvents than the average person.

P450 Inhibitors

An increasing number of drugs and other chemicals, dietary supplements, fruit juices, and vegetable constituents are being identified as inhibitors and/or inducers of CYPs, Phase II enzymes, and efflux transporters (Lin and Lu, 2001; Huang and Lesko, 2004). It should be recalled that mechanisms of P450 inhibition can be categorized as reversible, quasireversible, and irreversible. Reversible inhibition is transient, with the normal function of the CYP continuing after elimination of the inhibitor. Both quasireversible and irreversible inhibitions result from the enzyme’s formation of a reactive metabolite. With quasireversible inhibition, the metabolite and enzyme form a complex that is so stable that the intact enzyme may or may not be available for further metabolism. With mechanism-based or suicide inhibition, the reactive metabolite irreversibly inhibits the enzyme by binding to its active site. Experiments by Lilly et al. (1998) indicate that this is the mechanism by which epoxide metabolites of trans- and cis-1,2-dichloroethylene inhibit CYP2E1. Fisher et al. (2004) subsequently reported that metabolites generated from a single oral dose of CCl₄ as low as 1 mg/kg are suicide inhibitors of CYP2E1 and CYP2B2 in B6C3F1 mice. The net result of this action by CCl₄ was reduced formation of TCA from coadministered PERC. Blobaum (2006) described reversible inactivation of CYP2E and CYP2B isoforms during their mechanism-based inactivation by small tert-butyl terminal alkynes.

Flavanoids in grapefruit juice were one of the first classes of naturally occurring compounds to be documented to produce clinically significant increases in internal levels of therapeutic agents metabolized by CYP3A4 (Chan et al., 1998). Flavanoids in red wine exerted a similar effect on intestinal CYP3A4, but white wine had little influence. Bergamottin, a furanocoumarin in grapefruit and lime juice, was subsequently found to inhibit human CYP3A4 activity (Bailey et al., 2003) as well as intestinal P-glycoprotein efflux transporters (Dresser and Bailey, 2003). St. John’s wort, a top-selling dietary supplement taken for depression, selectively induced CYP3A4 in the small intestine (Huang and Lesko, 2004). Echinacea inhibited human intestinal CYP3A, but induced hepatic CYP3A. Ginsenosides exhibited no or weak inhibition of several human CYPs, but some of the compounds’ major intestinal metabolites inhibited a wide range of CYPs (Liu et al., 2006). Certain components of Ginkgo biloba, frequently used for memory impairment, were potent inhibitors of human recombinant CYP1B1, CYP1A1, and CYP1A2 (Chang et al., 2006). Identification of isoform-selective, nontoxic inhibitors may eventually lead to modulation of human xenobiotic metabolism for therapeutic purposes. There are large interindividual responses to CYP enzyme inducers and inhibitors (Lin and Lu, 2001).

Lifestyle

Exercise can significantly affect the kinetics of xenobiotics, but is often not considered in occupational risk assessments of solvents. It increases two of the major determinants of VOC uptake, alveolar ventilation rate and cardiac output/pulmonary blood flow. Polar solvents with relatively high blood:air PCs (eg, acetone, ethanol, and ethylene glycol [EG]) are very rapidly absorbed into the pulmonary circulation. Alveolar ventilation is rate limiting for these chemicals. In contrast, pulmonary blood flow and metabolism are rate limiting for uptake of the more lipophilic solvents (Johanson and Filser, 1992). Heavy exercise can increase pulmonary uptake of relatively polar solvents as much as fivefold in human subjects (Astrand, 1983). Light exercise doubles uptake of relatively lipid-soluble solvents, but no further increases occur at higher workloads. Blood flow to the liver and kidneys diminishes with exercise, so biotransformation of well-metabolized solvents and urinary elimination of polar metabolites may be diminished (Lof and Johanson, 1998). Dankovic and Bailer (1994) modified the human PBTK model of Reitz et al. (1989) for MC to reflect light work conditions. The modified model predicted that light exercise would result in a twofold increase in hepatic MC deposition and in metabolite formation via CYP- and GSH-dependent pathways. Jonsson et al. (2001) developed a population PBTK model for humans inhaling MC during moderate to heavy exercise. Increased workload was predicted by PBTK modeling to produce an 87% increase in internal exposure to components of a mixture of toluene, ethylbenzene, and xylene (Dennison et al., 2005).

Dietary habits can influence the absorption, metabolism, and toxicity of solvents in several ways. Factors that influence gastric emptying would not be expected to influence systemic absorption of VOCs, as VOCs are very rapidly absorbed by passive diffusion throughout the GI tract. The mere presense of food in the GI lumen, however, can inhibit absorption by reducing contact of the chemical with the GI epithelium. VOCs in the GI tract partition into dietary lipids, largely remaining there until the lipids are emulsified and absorbed. This serves to substantially delay uptake of lipophilic solvents such as CCl₄, resulting in lower, but prolonged blood CCl₄ levels and diminished hepatotoxicity in rats (Kim et al., 1990a). Food intake increases splanchnic blood flow, which favors GI absorption. The elevation in liver blood flow would be expected to enhance the biotransformation of low doses of well-metabolized solvents, but to have relatively little effect on metabolism of relatively high (ie, saturating) doses. Consumption of a high-fat diet results in a selective increase in hepatic CYP2E1 activity and protein levels in male Sprague–Dawley (S-D) rats (Raucy et al., 1991). Elevation of plasma ketone bodies appears to be implicated in this process, because acetone is a potent CYP2E1 inducer. Carbohydrate deficiency has been recognized for many years to enhance the metabolism of a series of VOCs in rats, but not to alter their conjugation (Sato and Nakajima, 1985). Psoralens (found in parsley, parsnip, and celery) reduce CYP1A2 activity, whereas food cooked over charcoal enhances it. Effects of other foods, fruit juices, and food supplements are addressed in the sections “P450 Inducers” and “P450 Inhibitors.”

Fasting for one to three days can significantly alter the hepatotoxicity of medium to high doses of VOCs that undergo metabolic activation. Fasting results in decreased hepatic concentrations of reduced GSH, due to cessation of intake of amino acids required for its synthesis. GSH plays a key role in detoxifying electrophilic metabolites of a number of VOCs, such as 1,1-dichloroethylene (DCE) (Jaeger et al., 1974). Conversely, conjugation of GSH with TCE or PERC can lead to formation of limited quantities of cytotoxic, mutagenic metabolites (see the sections “Trichloroethylene” and “Tetrachloroethylene”). Fasting (~8–24 hours) results in induction of liver CYP2E1 and increase in metabolism of a number of aromatic hydrocarbons and halocarbons (Nakajima et al., 1982). The hepatotoxicity of DCE and CCl₄, which require CYP2E1-catalyzed metabolic activation, is more severe in fasted animals. In contrast, long-term food deprivation (ie, starvation) results in decreased synthesis and activity of CYPs.

Many physiological and biochemical processes that impact solvent TK exhibit circadian, or diurnal, rhythms. A series of investigations has been conducted to delineate molecular control mechanisms of diurnal oscillation in mammals. The phenomenon appears to be controlled by a “master pacemaker” in the suprachiasmatic nuclei (SCN) of the hypothalamus of mice (Teboul et al., 2009). Genetic and genomic analyses of the SCN and liver indicate that a relatively small number of output genes are directly regulated by core oscillator components. Rate-limiting steps in major cellular pathways/processes are being identified as key sites of circadian regulation. Experiments reveal that transcriptional feedback is required for maintenance of circadian rhythmicity (Sato et al., 2006).

Several research groups have discovered circadian rhythms in susceptibility of rats to liver damage by various VOCs, including CCl₄, CHCl₃, DCE, ethanol, and styrene (Labrecque and Belanger, 1991). In each instance, the chemicals were most toxic when given during the initial part of the rodents’ dark/active cycle. Low hepatic GSH levels at this time were an important contributing factor to the increased liver injury by DCE and CHCl₃. Restricted food intake during sleep resulted in lipolysis and formation of acetone, which contributed to the diurnal induction of CYP2E1 and the ensuing increases in hepatic metabolic activation and cytotoxicity of CCl₄ in male rats during the early part of the animals’ active cycle (Bruckner et al., 2002). Skrzypinska-Gawrysiak et al. (2000) also observed that male mice were most susceptible to CCl₄ hepatotoxicity when dosed at a similar time. Clearance rates for a number of well-metabolized compounds, including ethanol (Sturtevant et al., 1978) and nicotine (Gries et al., 1996), are maximal in human subjects during the first part of their active/wake cycle. Hepatic blood flow and acetone levels in expired air of humans peak before breakfast. Thus, humans may also be more susceptible during this time to the toxicity of solvents that undergo metabolic activation.

Ethanol is an effective CYP2E1 inducer when ingested repeatedly in substantial amounts. There are numerous reports of marked potentiation of hepatic or renal damage by ethanol or other alcohols in persons occupationally exposed to potent hepatorenal toxicants, such as CCl₄ (Folland et al., 1976; Manno et al., 1996). CCl₄ is not extensively metabolized, but its free radical metabolites are extremely cytotoxic.

A group of moderate drinkers exposed to TRI vapor at 175 ppm showed a significant increase in metabolism and metabolic clearance of the chemical (Johns et al., 2006). TRI is a relatively nontoxic, poorly metabolized solvent. Kaneko et al. (1994) exposed ethanol-pretreated rats by inhalation of TCE or TRI at 50 to 1000 ppm. TRI metabolism was enhanced at all vapor concentrations, but TCE metabolism was enhanced by ethanol only at the highest vapor concentration (1000 ppm), as its biotransformation is perfusion-limited (limited by hepatic blood flow). Low levels of TCE entering the liver are extensively metabolized, even in nondrinkers who still have substantial CYP2E1 activity. Kedderis (1997) predicted that a 10-fold increase in CYP2E1 activity in humans inhaling TCE at 10 ppm would result in only a 2% increase in TCE metabolism by the liver. Thus, increased bioactivation capacity due to ethanol or other factors should not increase risks of toxicity or cancer from low-level exposures to TCE, DCE, MC, vinyl chloride, benzene, or other extensively metabolized VOCs. In contrast, Kedderis (1997) predicted a 10-fold increase in CYP2E1 activity would result in a 3.8-fold increase in PERC metabolites in the liver of persons inhaling 10 ppm PERC. Increased bioactivation of less extensively metabolized VOCs could thus enhance risks of low-level exposures to such chemicals.

Timing of ethanol consumption and VOC exposure is important. Prior repeated exposure to ethanol is necessary for substantial CYP2E1 synthesis to occur. Concurrent exposure to ethanol and a VOC, however, may often be protective from both well-metabolized and poorly metabolized solvents. If both a VOC and ethanol are metabolized by CYP2E1, the xenobiotics compete for available isozyme. Muller et al. (1975) observed that concurrent intake of ethanol and inhalation of TCE at 50 ppm by human subjects resulted in a marked decrease in urinary excretion of TCE’s major metabolites, TCA and trichloroethanol (TCOH). In this instance, ethanol would afford protection against these oxidative metabolites. Metabolism of ethanol produces an excess of NAD, a cofactor that favors formation of TCOH from chloral hydrate (CH), at the expense of TCA. Reduced formation of TCA would be protective against TCA-induced hepatic tumors. Larson and Bull (1989), however, observed this interaction in rats only with very high doses of TCE and ethanol.

Medications and drugs of abuse that induce or inhibit CYP2E1 and other enzymes involved in the metabolism of VOCs can potentially alter the chemicals’ toxicity or carcinogenicity. Nakajima et al. (1992a) showed that pretreatment of rats with phenobarbital, ethanol, or 3-methylcholanthrene significantly increased TCE oxidation. The same would be expected to occur in humans at high TCE doses. Some drugs (such as cycloheximide, disulfiram, and chloramphenicol) and the aforementioned natural constituents of plants inhibit CYP2E1. Those compounds, in sufficient doses, would be protective against high doses of TCE and other VOCs that are bioactivated by CYP2E1.

Tobacco smoke contains a number of compounds that are strong CYP inducers. Polycyclic hydrocarbons, such as 3-methylcholanthrene, are potent inducing agents. The polycyclic hydrocarbons primarily stimulate synthesis of CYP1A1 and CYP1A2, isozymes that play a modest role in catalyzing the biotransformation of TCE (Nakajima et al., 1992a). Nicotine, however, is a strong CYP2E1 inducer in rats (Micu et al., 2003). Cigarette smoke is known to induce CYP2E1 in both rodents and humans. Yue et al. (2009) recently observed that chronic nicotine treatment also increased voluntary ethanol intake by rats, thereby further enhancing hepatic CYP2E1 and CYP2B1/2 levels.

Solvent Mixtures

Many occupational and environmental exposures to VOCs involve multiple chemicals. That is particularly true of contaminated environmental media, in that widespread use of solvents leads to their volatilization and entry into surface waters and groundwater. VOCs spilled onto the ground evaporate, although some often leaches through porous soils into groundwater and remains trapped there. The groundwater at about 90% of 1608 hazardous waste sites on the US National Priorities List contains VOCs. TCE is the most frequently found of all chemicals, followed by lead, PCE, and benzene (Pohl et al., 2008). The most common four-component VOC mixture is TCE, PCE, TRI, and 1,1-dichloroethane. ATSDR (2004) published a toxicological profile addressing potential health risks posed by this four-component mixture. Many US cities’ drinking water supplies also contain complex mixtures of VOCs. Total concentrations generally range from ppt to ppb (Moran et al., 2007; Rowe et al., 2007). Trace amounts of a variety of VOCs are also present in the blood of many nonoccupationally exposed members of the general population (Churchill et al., 2001; Blount et al., 2006).

Most studies have involved experiments with binary or ternary mixtures. One chemical may have no effect on, potentiate (enhance), or antagonize (inhibit) adverse actions of a second or third chemical (Tan et al., 2011). Knowledge of mechanisms of VOC interactions involves largely the influence of one VOC on the metabolic activation or inactivation of another. Koizumi et al. (1982) published the results of one of the first such studies. They found that coexposure of rats to PERC and TRI resulted in significant suppression of TRI metabolism. Workers exposed to TCE and PERC were found to have lower urinary concentrations of TCE metabolites than workers exposed to TCE alone (Seiji et al., 1989). The interaction resulted from competitive metabolic inhibition, wherein the amounts of the combined chemicals exceeded the metabolic capacity of the study subjects. Such an interaction is protective against cytotoxicity and carcinogenicity, in that the bioactivation of both TCE and PERC is reduced. Conversely, systemic concentrations of parent VOCs would be increased, and this might enhance neurologic effects of sufficiently high doses (Pohl and Scinicariello, 2011).

PBTK modeling has been used by several research groups to predict the metabolic and toxicological consequences of exposure to VOC mixtures. Competitive metabolic inhibition was evident in a PBTK modeling approach to studying TCE and 1,1-DCE (El-Masri et al., 1996) and TCE and vinyl chloride (Barton et al., 1995). Later PBTK modeling efforts predicted interaction thresholds below which competitive metabolic inhibition would not occur. Dobrev et al. (2001), for example, reported that the thresholds for interaction of TCE with PERC and TRI vapor in rats were 25 and 135 ppm, respectively, when the TCE concentration was 50 ppm. Those findings imply that protection from adverse effects caused by oxidative metabolites would occur in occupational settings when vapor concentrations were relatively high. An increase in blood TCE concentrations under these exposure conditions, however, was predicted to result in a disproportionate increase in formation of nephrotoxic GSH conjugation products in humans (Dobrev et al., 2002). Competitive metabolic inhibition, with potentiation or protection from adverse effects of VOCs, would not occur on exposure to environmentally relevant concentrations.

Additivity of toxic effects of chemicals that act by similar mechanisms is typically assumed in the absence of experimental evidence to the contrary. Stacey (1989) studied the joint action of TCE and PERC on the liver and kidneys of rats. Combined administration of near-toxic-threshold doses of the two solvents produced modest hepatorenal toxicity. Jonker et al. (1996) provided evidence that TCE and PERC in combination with two other similarly acting solvents affected kidney weight in rats given subtoxic doses of each chemical by gavage for 32 days. Competitive metabolic inhibition, as described above, would result in less than additive adverse effects when metabolites are the bioactive moieties (Pohl and Scinicariello, 2011). Goldsworthy and Popp (1987) found that the joint effect of TCE and PERC on peroxisome proliferation in the liver and kidneys of mice and rats was less than additive. Competitive metabolic inhibition at relatively high exposure levels of toluene, ethylbenzene, and xylene has been predicted by PBTK modeling to result in higher internal exposures (and CNS depressant effects) than would occur with simple additivity (Dennison et al., 2005).

Few toxicity or carcinogenicity studies of complex chemical mixtures including VOCs have been conducted. The National Toxicology Program (NTP, 1993a) supplied F344 rats and B6C3F1 mice with drinking water containing 24 contaminants for up to 26 weeks. The mixture contained TCE, PERC, MC, TRI, DCE, 1,1-dichloroacetic acid, other solvents, heavy metals, polychlorinated biphenyls, and a phthalate. The total no-observed-adverse-effect level (NOAEL) for histological changes in the liver was 11 ppm in rats. Suppression of immune function occurred in female mice that consumed 756 ppm of the mixture for two weeks or 378 ppm for 13 weeks. A follow-up study in chemically tumor-initiated rats showed that the contaminant mixture did not promote preneoplastic foci in the liver (Benjamin et al., 1999). Wang et al. (2002a) supplied ICR mice with water containing CHCl₃, 1,1-dichloroacetic acid, DCE, TRI, TCE, and PERC for 16 and 18 months. There was a trend of increasing frequency of hepatocellular neoplasms in the male mice and an increasing incidence of mammary adenocarcinomas in the high-dose female mice. The total concentration of VOCs in the drinking water of females was about 1555 ppb. Most of the mixture was TCE (471 ppb) and PERC (606 ppb). These concentrations are far lower than have previously been reported to produce tumors. The results must be regarded as preliminary, in that the study design had a number of limitations, and the results have not been replicated. In addition, male B6C3F1 mice are particularly susceptible to hepatic tumors (Haseman et al., 1998) and mice metabolically activate a substantially greater proportion of VOC doses than do humans.

Diseases

Illness can be a major source of variability in response to solvents. Impaired drug metabolism and clearance are commonly seen in patients with cirrhosis and hepatitis (Welling and Pool, 1996). Reduced metabolism of solvents may result from decrease in hepatic parenchymal mass, diminished enzymatic activity, and/or decreased portal blood flow (Morgan and McLean, 1995). Lower levels of CYP2E1, CYP1A2, and GSH are seen in livers of patients with cirrhosis (Murray, 1992). Reduced P450-catalyzed metabolic activation of certain solvents can be protective, but diminished capacity to conjugate electrophilic metabolites with GSH would have the opposite effect. Plasma levels of albumin and α₁-acid glycoprotein fall in cirrhotic patients. Thus, plasma protein binding of many xenobiotic metabolites decreases and their rate of elimination increases (Morgan and McLean, 1995). Definitive information is lacking, however, on the net effect of common liver diseases on solvents.

Chronic kidney disease is becoming increasingly prevalent in the United States over the past decade (Coresh et al., 2007). Progressive loss of kidney function leads to impaired renal excretion of numerous chemicals and metabolites that may be toxic or pharmacologically active. Kidney disease can affect uptake and efflux transporters and metabolic enzymes in the liver and the GI tract (Nolin et al., 2008). Chronic renal failure can result in downregulation of the expression of some liver CYPs, but have no effect on others (Leblond et al., 2001). Effects on hepatic clearance of therapeutic chemicals can thus vary widely. The mechanisms of these effects on gene expression are unclear, as is their applicability to solvents. The plasma protein binding of many xenobiotics is reduced in patients with compromised renal function, apparently due to retention of substances that compete for protein-binding sites, as well as reduced albumin synthesis. Clearance of highly metabolized xenobiotics thus appears to depend on potentially offsetting influences of altered metabolism, decreased plasma protein binding, and decreased renal excretion (Yuan and Venitz, 2000). Impairment of renal metabolic activation via oxidation or β-lyase-mediated conversion of GSH metabolic intermediates of TCE and PERC (see the sections “Trichloroethylene” and “Tetrachloroethylene” for details) to reactive, mutagenic metabolites would be protective.

Diabetes mellitus is a metabolic disease characterized by hyperglycemia as a result of insulin deficiency (Type I) or insulin resistance (Type II). Type II accounts for 90% of cases. Approximately 21 million people suffer from diabetes in the United States. A prominent effect of chemically induced Type I diabetes in rats is induction of CYP2E1. Elevation of acetone, but apparently not the other ketone bodies, is responsible for the CYP2E1 increase. Hypoinsulinemia may play a role, in that insulin downregulates CYP2E1 mRNA in rat hepatoma cells and primary cultures of rat hepatocytes (Cheng and Morgan, 2001). In contrast, CYP2E1 is not affected in Type II diabetic animals (Sawant et al., 2004). Neither Type I nor Type II diabetes influences CYP2E1 activity in humans (Lucas et al., 1998; Cheng and Morgan, 2001). Nevertheless, animal studies show that Type II diabetes increases susceptibility to hepatotoxicants including allyl alcohol and CCl₄ (Sawant et al., 2004), apparently due to inhibition of tissue repair. Devi and Mehendale (2005, 2006) subsequently demonstrated inhibition of expression of genes involved in cell division and protease inhibitors, as well as enhanced gene expression of proteases in Type I diabetic rats. These events may be important in delayed tissue repair following chemical cytotoxicity in diabetics.

Persons with bacterial infections may be more sensitive to cytotoxic actions of solvents. Endotoxin, which includes a lipopolysaccharide (LPS), is released from the cell wall of gram-negative organisms. The LPS causes the release of inflammatory mediators, which alter cell membranes, intercellular signaling, and gene expression (Roth et al., 1997). These effects may render cells more susceptible to damage by solvents and other chemicals. Exposure of animals to small amounts of endotoxin potentiates liver injury by CCl₄, halothane, allyl alcohol, ethanol, and other solvents (Roth et al., 1997). Endotoxin apparently activates Kupffer cells to release inflammatory mediators and cytotoxic moieties to hepatocytes (Thurman, 1998).

Trichloroethylene

TCE was widely used as a solvent and metal degreasing chemical in the United States from the early 1900s to 1980s (Bakke et al., 2007). Workers were initially subjected routinely to quite high exposures. NIOSH (1989) estimated that 401,000 persons at 23,225 plants in the United States were potentially exposed to TCE. TCE has been identified at over one half of the nearly 1300 hazardous waste sites that make up the EPA’s National Priorities List (ATSDR, 2006a; Pohl et al., 2008). It is released into the atmosphere from vapor degreasing operations; however, direct discharges to surface waters and groundwater from disposal operations have been frequent occurrences. As a result, TCE can be released to indoor air by vapor intrusion through underground walls and floors and by volatilization from the water supply. TCE has recently received a great deal of attention from the scientific and regulatory communities, due in large part to the EPA’s effort to update the chemical’s two-decade-old risk assessment (EPA, 2009).

Moderate to high doses of TCE, as with other halocarbons, are associated with a number of noncancer toxicities (ATSDR, 1997d; Barton and Das, 1996). TCE has been implicated in the development of autoimmune disorders and immune system dysfunction (Blossom and Gilbert, 2006; Cooper et al., 2009; Lan et al., 2010), and has been investigated for its potential as a male reproductive toxicant (Forkert et al., 2003; Xu et al., 2004). The effect of gestational exposure to TCE or its oxidative metabolites on cardiac development is a subject that has invoked considerable debate, as conflicting results have been published (Fisher et al., 2001; Johnson et al., 2003; Watson et al., 2006). The issue of TCE’s potential effect on ocular development has also recently reemerged (Warren et al., 2006). Nevertheless, cancer remains the dominant issue for TCE. TCE and its potential health risks have recently been reviewed by the NAS (2006, 2009), Jollow et al. (2009), and the EPA (2009).

Metabolism

The adverse effects of TCE, other than CNS depression, are generally believed to be associated with TCE’s metabolites. CNS depression is due to the parent compound and a major metabolite, TCOH. Knowledge of TCE’s metabolism in different species and under different exposure conditions is thus a prerequisite to understanding its mechanisms of action and assessing human health risks. A great deal of experimentation at many biological levels has been conducted over the last five decades to these ends.

TCE is rapidly and extensively absorbed into the systemic circulation via the oral and inhalation routes. Dermal absorption is considerably slower and less extensive. The majority of TCE is oxidized in the liver, while a small amount is conjugated in the liver with GSH by GSTs. The oxidative pathway is shown on the left of Fig. 24-5, while the GSH pathway is shown on the right. This diagram is a simplification of a more complex metabolic scheme described in detail by Lash et al. (2000a).

The initial step in the oxidative pathway is catalyzed by microsomal CYPs. CYP2E1, as described previously in the subsection “Metabolism” in the section “Toxicokinetics,” is the primary isozyme responsible for oxidation of low concentrations of TCE (Lipscomb et al., 1997; Ramdhan et al., 2008; Kim and Ghanayem, 2006). CYP-catalyzed oxidation of TCE in rodents and humans, in decreasing order of magnitude, is as follows: mice >> rats > humans (Lash et al., 2000a). Whether or not TCE is initially converted to TCE oxide is controversial. Cai and Guengerich (2001) were able to detect formation of trace amounts of the epoxide by phenobarbital-induced rat liver CYPs, but not by human liver CYPs. The majority of TCE is apparently converted to an oxygenated TCE–CYP intermediate, which rearranges to form chloral, a major metabolic intermediate. Chloral is oxidized to chloral hydrate (CH), a sedative and hypnotic still widely used in medical and dental procedures for infants and children (Buck, 2005; Heistein et al., 2006). CH is both oxidized to TCA and reduced to TCOH. Much TCOH is conjugated with glucuronic acid (GLU) and excreted in the urine. TCOH–GLU that is excreted in the bile is extensively hydrolyzed in the gut, reabsorbed, and oxidized in part to TCA (Stenner et al., 1997). Chiu et al. (2007) observed that concentrations of TCA were significantly lower than TCOH and TCOH–GLU concentrations in the blood of humans who inhaled TCE at 1 ppm for six hours. Modest amounts of DCA apparently are produced from TCA and TCOH in mice, but relatively little DCA is formed in rats or humans. Trace amounts of DCA were detected in one study of TCE-exposed humans (Fisher et al., 1998) but not in other studies (Lash et al., 2000b; Bloemen et al., 2001). Very high doses of both TCA and DCA have been shown to be hepatic carcinogens in mice (Bull, 2000). It is generally accepted that TCA is a nongenotoxic liver carcinogen in B6C3F1 mice, although its ability to cause liver cancer in humans has been discounted by findings in numerous laboratory investigations (Bull, 2000). The possible role of DCA in human liver cancer is even more controversial (Walgren et al., 2005; Caldwell and Keshava, 2006; Klaunig et al., 2007).

The GSH conjugation pathway is quite similar qualitatively, but not quantitatively, in rats and humans. The initial step in this second, minor pathway involves conjugation of TCE with GSH to form S-(1,2-dichlorovinyl)GSH (DCVG). DCVG formation occurs primarily in the liver at a rate about 10 times greater in rats than in humans (Green et al., 1997). Much of the DCVG is excreted via the bile into the intestines and converted to S-(1,2-dichlorovinyl)-l-cysteine (DCVC). That metabolite is reabsorbed and taken up by the liver, where a portion is detoxified by N-acetylation. Bernauer et al. (1996) exposed rats and humans to TCE vapor at up to 160 ppm for six hours. The rats excreted eight times more N-acetyl-DCVC in their urine than did the human volunteers at each exposure level. Some DCVC is taken up by the kidneys and further metabolized by the enzyme β-lyase to S-(1,2-dichlorovinyl)thiol (DCVSH). DCVSH is then converted to highly reactive products, including DCVC sulfoxide (DCVCS), chlorothioketene, and thionoacylchloride (Lash et al., 2000a). Metabolic activation of DCVC to chlorothioketene was shown to occur 11 times more rapidly in rats than in humans (Green et al., 1997). Lash et al. (2001) also demonstrated that cultured rat renal cells were more sensitive to DCVC than human renal cells. Chlorothioketene and similarly unstable congeners are capable of covalently binding to renal cellular proteins and DNA. This results in genotoxicity and cytotoxicity, with ensuing regenerative hyperplasia and potentially renal cell cancer (RCC).

Modes of Carcinogenic Action in Target Tissues

Both metabolic pathways are implicated in the carcinogenicity of TCE: reactive metabolite(s) of the GSH pathway in kidney tumors in rats and oxidative metabolites in liver and lung tumors in mice. That tumor formation in many cases is species-, strain-, sex-, and route of exposure–dependent has provided clues as to TCE’s modes of carcinogenic action. Whereas substantial progress has been made on the mechanistic front, the reader should not infer from the text that follows that all modes of action are known with absolute certainty.

Liver Cancer

It is well established that TCE, when given chronically in very high doses by gavage, can produce an increased incidence of hepatocellular carcinoma in B6C3F1 mice, but not in other strains of mice or in rats. This differential susceptibility can be explained in part by the greater capacity of the mouse to bioactivate relatively large quantities of TCE via the oxidative pathway. The B6C3F1 mouse produces a substantially larger amount of TCA after TCE exposures than do unresponsive strains of mouse, rats, or humans. The susceptibility of the B6C3F1 mouse is also likely related to the high (42.2%) incidence of liver adenoma/carcinoma in male controls (Haseman et al., 1998). This phenomenon may be due to an abnormally high population of spontaneously initiated cells in these animals’ liver. Mice express very low levels of epoxide hydrolase (Lorenz et al., 1984), the enzyme that catalyzes the hydrolytic degradation/detoxification of reactive epoxide metabolites of TCE and other VOCs.

CH induces hepatic tumors in male B6C3F1 mice, but not in F344 rats (Leaky et al., 2003). Female B6C3F1 mice gavaged with up to 100 mg CH/kg per day for 104 weeks showed no increase in liver tumors, but the male mice did exhibit increased incidences of hepatoma and/or hepatocellular carcinoma. CH is rapidly converted to TCA and TCOH in rodents and humans. Merdink et al. (2008) detected only trace amounts of DCA in blood and urine of male human subjects dosed with 500 or 1500 mg of CH. An epidemiology study of the possible association of short-term clinical administration of CH as a sedative–hypnotic and cancer risk in 2290 patients was conducted by Haselkorn et al. (2006). The authors concluded there was no persuasive evidence of a causal relationship between CH and cancer in humans.

TCA and DCA are the most likely candidates for proximate hepatocarcinogens produced by TCE’s oxidative pathway. TCA is a species-specific carcinogen that induces peroxisome proliferation and hepatocellular carcinoma in male and female B6C3F1 mice when administered in very high doses in drinking water or by gavage. It does not produce liver tumors in any strain of rats tested under these conditions (NAS, 2006, 2009). Very high doses of DCA, however, produce hepatic tumors in both B6C3F1 mice and F344 rats. Large, repeated doses of DCA and TCA initially stimulate, and then depress the growth of normal hepatocytes (Bull, 2000). This may confer a growth advantage to initiated cells, and is referred to as negative selection. At high tumorigenic doses, DCA (but not TCA) is thought to stimulate cell replication within liver tumors. If indeed both DCA and TCA contribute to tumorigenesis, findings by Bull et al. (2002) indicate they do so by distinct mechanisms. DCA-promoted liver tumors differed phenotypically from those promoted by TCA. DCA- and TCA-induced tumors also differ as to whether their depression of cytosine methylation is reversible on cessation of treatment (Tao et al., 1998). This is particularly important as DNA hypomethylation, including that of the proto-oncogenes c-jun and c-myc, may be an epigenetic mechanism for the tumorigenicity of DCA and TCA (Tao et al., 2004). It appears that hypomethylation due to TCA and DCA induces DNA replication and prevents the methylation of newly synthesized strands of DNA (Ge et al., 2001). Tao et al. (2000) have reported that DCA and TCA do so by virtue of their depletion of S-adenosylmethionine, which normally supplies the methyl group for the methylation process. It should be recognized that these are not genotoxic mechanisms. The effects disappear as soon as these metabolites are eliminated.

A primary mode of action of TCA and to a smaller extent DCA is activation of PPARα. As peroxisomes contain a variety of oxidative enzymes, PPARα activation produces oxidative stress, which can be manifest as lipid peroxidation, oxidative DNA damage, and transcription factor activation (O’Brien et al., 2005). Stimulation of PPARα can enhance DNA replication, resulting in expansion of some clones of hepatocytes and suppression of apoptosis, so initiated and precancerous cells will be spared. Male wild-type mice dosed orally with TCE exhibit hepatocyte proliferation and changes in expression of genes involved in cell growth (Laughter et al., 2004). PPARα-null mice are refractory to those effects, which are associated with carcinogenesis. Mice expressing human PPARα fail to show increases in markers of cell proliferation and are resistant to liver cancer if treated with PPARα agonists (Morimura et al., 2006; Yang et al., 2008). The concentration of PPARα in human cells is about 10% of that in the livers of rodents (Palmer et al., 1998; Klaunig et al., 2003). Many toxicologists have judged that the mode of action for hepatic carcinogenesis observed in mice after administration of peroxisome proliferation–inducing drugs and other chemicals, such as TCA, makes it unlikely that such chemicals pose a hepatic cancer risk in humans (Cattley et al., 1998; Clewell and Andersen, 2004; Klaunig et al., 2007; Gonzalez and Shah, 2008). It was concluded by an expert review panel that the PPARα mode of action for liver cancer in mice is not relevant to humans (NAS, 2006). However, others have raised questions about the interpretation of PPARα actions and whether it is the only relevant mode of action for such chemicals (Keshava and Caldwell, 2006). This continues to be a subject of active debate (NAS, 2008).

It is important to recognize that stimulation or inhibition of cell growth through PPARα activation ceases when the metabolites are eliminated (Miller et al., 2000). Thus, such alteration of cell signaling is not a genotoxic mechanism of action. Very high concentrations of DCA and CH have a weak genotoxic action in vitro. Bull (2000) and Moore and Harrington-Brock (2000), however, conclude that it is unlikely that those metabolites would cause tumors in any organ through genotoxicity or mutagenicity at exposure concentrations relevant to humans.

Lung Cancer

As discussed in the review paper by Green (2000), inhaled TCE is carcinogenic to the mouse lung but not to that of the rat. Oral TCE is not carcinogenic to the lung, probably due to first-pass hepatic metabolism that limits the amount of TCE reaching the lungs. The primary target of TCE in the mouse lung is the nonciliated Clara cell. Cytotoxicity is characterized by vacuolization and increased replication of these cells in the bronchiolar epithelium. A dose-dependent reduction in the CYP activity in Clara cells is observed as well. This loss of metabolic activation capacity can be thought of as an adaptive response. Clara cells recover morphologically during repeated daily inhalation exposures to TCE.

Chloral is the putative toxicant responsible for pulmonary tumor formation. Clara cells of the mouse efficiently metabolize TCE to chloral. Chloral accumulates, due to its efficient production and low activity of ADH, the enzyme responsible for its reduction to TCOH. The Clara cells’ lack of glucuronosyltransferase, the enzyme that normally catalyzes the formation of TCOH–GLU, has also been implicated in chloral accumulation. Species differences in susceptibility of the lung to TCE are due in part to mouse lung Clara cells having a much higher level of CYP2E1 than those of the rat, and thus a much higher capacity to metabolize TCE to chloral. Also, Clara cells in mice are much more numerous than in rats. Clara cells are rare in human lungs. A critical role for chloral is supported by the findings that its administration to mice, but not TCA or TCOH, causes Clara cell toxicity identical to that of TCE. Chloral does appear to have some genotoxic potential, especially in regard to inducing aneuploidy. However, the fact that tumors are not seen in species where cytotoxicity does not occur strongly implicates cytotoxicity and reparative proliferation in tumor formation. In an effort to test the hypothesis that bronchiolar damage by TCE is associated with bioactivation within Clara cells, Forkert et al. (2006) administered TCE i.p. to CD-1 mice. The result was dose-dependent production of dichloroacetyl lysine adducts in Clara cells (used as an in vivo marker of TCE metabolism) that correlated with bronchiolar damage. The work also suggested that CYP2F2 may play a more important role than CYP2E1 in TCE metabolism and cytotoxicity within the mouse lung (Forkert et al., 2005).

Kidney Cancer

TCE was given in corn oil to F344/N rats and B6C3F1 mice of both sexes by gavage five times weekly at doses up to 1000 mg/kg in rats and 6000 mg/kg in mice in a 13-week study, as well as up to 1000 mg/kg to both species and sexes in a 103-week study (NTP, 1990). A low, but statistically significant increase in renal tumor incidence was observed only in the male rats given TCE at 1000 mg/kg for two years. Two-year gavage studies of TCE, in four additional rat strains, were also conducted (NTP, 1988). In all strains of rats tested, cytomegaly and karyomegaly of tubular cells in the renal corticomedullary region were seen. Frank toxic nephropathy was observed with higher frequency in male rats beginning at 52 weeks of exposure. Renal adenomas or adenocarcinomas were occasionally seen in male rats of different strains after two years of the repetitive, high-dose oral exposure regimen.

Adverse effects of TCE on the kidneys are due largely to metabolites formed via the GSH conjugation pathway (Lash et al., 2000b). As described previously, conjugation of TCE with GSH to form DCVG occurs primarily in the liver. DCVG is secreted into bile and blood. That in the bile is converted in the gut to DCVC, which is reabsorbed into the bloodstream. Humans have a lower capacity than rats to metabolize TCE by the GSH pathway. Lash et al. (1999) were able to detect DCVG in the blood of humans who inhaled TCE at 50 or 100 ppm for four hours, but Bloeman et al. (2001) could not find DCVG or DCVC in the urine of similarly exposed subjects. DCVG in the blood is taken up by the kidneys and metabolized to DCVC by γ-glutamyltransferase and a dipeptidase. Lash et al. (2001) observed the following decreasing order of toxic potency in freshly isolated rat cortical cells: DCVC > DCVG >> TCE. DCVC can be detoxified by acetylation or activated further by two pathways: (1) cleavage by renal cytosolic and mitochondrial β-lyases to dichlorothioketene, which in turn can lose a chloride ion to yield chlorothioketene or tautomerize to form chlorothionacyl chloride (the latter two moieties are very reactive and acylate proteins and DNA); and (2) oxidation by renal CYPs or flavin-containing monooxygenases to DCVCS, a reactive epoxide. Lash et al. (1994) reported that DCVCS was a more potent nephrotoxicant than DCVC in vitro and in vivo in rats. Apoptosis was observed after as little as one hour of incubation of cultured human renal proximal tubular (RPT) cells with DCVC and DCVCS (Lash et al., 2003, 2005). Cellular proliferation, accompanied by increased expression of proteins associated with cellular growth, differentiation, stress and apoptosis, was also an early response to low doses. Necrosis, however, was a late, high-dose phenomenon. Exposure of human RPT cells to DCVC at lower concentrations for 10 days also resulted in expression of genes associated with cell proliferation, apoptosis, and stress (Lash et al., 2005), as well as repair and DCVC metabolism (Lash et al., 2006).

Animal studies typically provide insight into mode of action, but in the case of TCE-induced RCC, human studies have been of significant value. Bruning et al. (1997a) analyzed tumor tissues from 23 RCC patients with occupational histories of long-term, high-level TCE exposure. Tumor cell DNA was isolated and analyzed for somatic mutations of the von Hippel–Lindau (VHL) tumor suppressor gene. Compared with VHL gene mutation rates of 33% to 55% in unexposed RCC patients, all 23 TCE-exposed RCC patients exhibited aberrations of the VHL gene. In a follow-up study, Brauch et al. (1999) sought to determine whether TCE produced a specific mutation of the VHL gene. These investigators analyzed VHL gene sequences in DNA isolated from RCC tissues from patients exposed to high levels of TCE in metal-processing factories. Renal cell tumors of TCE-exposed patients showed somatic VHL mutations in 33 of 44 cases (75%). Of the 33 cases with VHL mutations, a specific mutational hot spot at VHL nucleotide 454 was observed in 13 cases. The nucleotide 454 mutation was not found in any of the 107 RCC patients without TCE exposure or among 97 healthy subjects, 47 of whom had a history of TCE exposure. These data suggest that the VHL gene may be a specific and susceptible target of reactive GSH pathway metabolites, a concept strengthened by a more recent study reporting VHL nucleotide 454 mutation among TCE-exposed but not in nonexposed RCC patients (Brauch et al., 2004). More recently, however, Charbotel et al. (2007) found no associations between the number and type of VHL mutations in TCE-exposed and unexposed RCC patients. The subject awaits further clarification (Chow and Devesa, 2008).

The aforementioned reactive products of DCVC are both cytotoxic and mutagenic. Adverse effects on proximal tubule cells include alkylation of cytosolic and mitochondrial structural and enzymatic proteins, oxidative stress, marked ATP depletion, and perturbations of calcium homeostasis. Tubular necrosis ensues, with subsequent proliferation that can alter gene expression, which may modify cell growth and differentiation. Genes associated with stress, apoptosis, repair, and proliferation were upregulated almost twofold in cultured human renal tubular cells exposed to subtoxic doses of DCVC for 10 days (Lock et al., 2006). Mechanisms of noncarcinogenic and carcinogenic action are discussed in detail by Lash et al. (2000b) and Lock and Reed (2006).

Bruning and Bolt (2000) opine that reactive metabolite(s) of the GSH pathway may have a genotoxic effect on the proximal tubule of the human kidney, but that full development of a malignant tumor requires a promotional effect such as cell proliferation in response to tubular damage. If this is true, RCC secondary to TCE exposure would be a threshold response. The question of whether chronic tubular damage is a prerequisite to renal tumor formation is quite important. Evidence has come from the use of electrophoresis to examine protein excretion patterns in the urine of RCC patients with and without a history of chronic, high-level TCE exposure (Bruning et al., 1996). Protein excretion patterns indicative of tubular damage were identified in all their 17 TCE-exposed cases, but in only about one half of 35 controls. Bruning et al. (1999) subsequently published the results of a larger study supportive of this concept. Approximately 93% of 41 RCC patients with high TCE exposure exhibited elevated urinary α₁-microglobulin excretion versus 46% of 50 RCC patients without a history of TCE exposure. Similar findings were reported in an updated study by Bolt et al. (2004), and in an investigation by Green et al. (2004) of 70 electronics workers who inhaled an average of 32 ppm TCE for four years.

Results of an investigation by Mally et al. (2006) provide additional insight into the TCE renal carcinogenesis threshold premise. A strain of rats (Eker) uniquely susceptible to renal carcinogens was administered 100, 250, 500, and 1000 mg TCE/kg by gavage five days per week for 13 weeks. The Eker rat is a unique animal model for RCC, carrying a germ-line alteration of the Tsc-2 tumor suppressor gene. Results showed a significant increase in cell proliferation in renal tubular cells but no increase in preneoplastic renal lesions or tumor incidence. In vitro studies were conducted on primary Eker rat renal epithelial cells by exposing them to the TCE metabolite DCVC dissolved in water at 10 to 50 μM for eight, 24, and 72 hours. Concentrations of DCVC that reduced rat renal cell survival to 50% also resulted in cell transformation. No carcinogen-specific mutations were identified in the VHL or Tsc-2 tumor suppressor genes in the transformed cells. RCCs in the Eker rat have substantial similarities to human RCC. It is not entirely clear that this or any contemporary experimental animal model adequately mirrors humans with regard to the effects of TCE-induced mutations in the VHL gene, but the authors firmly suggest that TCE-mediated renal carcinogenicity may occur only secondarily to nephrotoxicity and sustained regenerative cell proliferation.

Metabolism of TCE by the GSH pathway is similar qualitatively, but not quantitatively in rats and humans. N-Acetyl-DCVC, the major detoxification product of DCVC, was found in the urine of humans and rats after six hours of inhalation of up to 160 ppm TCE (Bernauer et al., 1996). Cumulative excretion of the N-acetyl derivative was seven- to eight-fold higher in the rats. Bruning et al. (1997b) originally reported an increased likelihood of RCC patients with high TCE exposure having a functional GST isozyme GSTT1 or GSTM1 genotype. This isozyme GSTT1 is thought to be primarily involved in TCE metabolism, while GSTM1 is thought to detoxify epoxides. A reassessment by the investigators of a larger population, however, revealed no such relationships (Wiesenhutter et al., 2007). Moore et al. (2010), recently completed an assessment of an even larger population. There was an increased risk of RCC in workers with at least one intact GSTT1 allele, but not in persons with two deleted alleles (ie, null genotype).

It is also worthy of note that male rats and humans are apparently at greater risk of TCE-induced kidney cancer than their female counterparts. Blood DCVG concentrations were 3.4-fold higher in male than in female volunteers inhaling 50 or 100 ppm TCE (Lash et al., 1999). Male rats display higher GSH conjugation, γ-glutamyl transpeptidase activity, and cysteine conjugate β-lyase activity than female rats. Taken together, results of the cited studies indicate that both male humans and male rats possess GSH conjugation capacity and can produce the critical TCE metabolite DCVC. Renal carcinoma has been observed in male rats and male workers when both have been exposed to very high TCE concentrations for prolonged periods of time. These observations show data congruence, indicating that the conjugation pathway plays a central role in induction of renal carcinoma in males of both species. As described previously, rats have a significantly greater capacity to metabolically activate TCE to DCVC by this pathway than do humans. Rat renal cortical cells, in turn, are more susceptible to injury by DCVC than their human counterparts.

Cancer Epidemiology Studies

There have been many published studies of cancer incidence and mortality in TCE-exposed populations. Most of the epidemiology studies in the United States prior to 2000 involved workers in the aircraft maintenance and manufacturing industries. There were also investigations of Swedish, Finnish, German, and Danish worker cohorts. Results of these assessments have been mixed, ranging from no association to limited evidence. The major studies with some exposure data constituted what the Wartenberg et al. (2000) meta analysis referred to as Tier I studies, which received a greater weighting when making causal inferences than lower tier cohort, case–control, and community-based investigations. Among the Tier I studies, evidence for an excess incidence of cancer was strongest for kidney (rate ratios [RR] = 1.7, 95% confidence interval [CI] = 1.1–2.7), liver (RR = 1.9, CI = 1.0–3.4), and non-Hodgkin lymphoma (NHL) (RR = 1.5, CI = 0.9–2.3) (Wartenberg et al., 2000).

The report of Henschler et al. (1995) was the first in a series of German studies that have provided some of the strongest evidence to date for an association between TCE and RCC. These authors described a cohort of male cardboard factory workers who were exposed to moderate to extremely high concentrations of TCE vapor. By the closing date of the study, five of the 169 exposed workers had been diagnosed with kidney cancer versus none of the 190 controls. This resulted in standardized incidence ratios (SIRs) of 7.97 (CI = 2.59–18.59) and 9.66 (CI = 3.14–22.55), using Danish and German Cancer Registry data for comparison, respectively. German researchers also conducted a hospital-based case–control study with 58 RCC patients and 84 patients from accident wards who served as controls. Of the 58 RCC patients, 19 had histories of occupational TCE exposure of at least two years, compared with only 5 of the controls. After adjustment for potential confounders, an association between RCC and long-term exposure to TCE was reported (odds ratio [OR] = 10.80, CI = 3.36–34.75) (Vamvakas et al., 1998). In an expanded German case–control study of 134 RCC cases and 401 controls, excess risks for those working longest with TCE (OR = 1.8, CI = 1.01–3.20) and those experiencing narcotic symptoms (OR = 3.71, CI = 1.80–7.54) were reported (Bruning et al., 2003). Narcosis was thought to be associated with peak exposures. These workers reported frequent dizziness, requiring them to seek fresh air several times daily. More recent investigations have generally involved surveys of populations with lower exposures. The results of these investigations have been suggestive of an association between TCE and RCC, although some findings were negative or did not reach statistical significance (Alexander et al., 2006; Boice et al., 2006; Chang et al., 2005; Mandel et al., 2006; Raaschou-Nielsen et al., 2003; Wong, 2004). A number of these assessments involved a limited number of subjects. Charbotel et al. (2006) studied 86 RCC patients and 316 matched controls from an area in France with high TCE exposures in local industries. There was a 64% increase in RCC risk with TCE exposure, with the risk doubling in persons with a high cumulative dose, increasing more when peak exposures were also taken into account (OR = 2.74, CI = 1.06–7.07). The OR was still high, but not statistically significant after adjusting for exposure to cutting fluids, because of the study’s lack of power. Chow and Devesa (2008) reviewed recent epidemiological evidence of a rising incidence of RCC in the United States. Cohort studies showed associations with smoking, obesity, diminished physical activity, hypertension, and certain chemical exposures.

Epidemiological evidence of NHL or cancer of the liver, lung, or other organs has been weaker than that for the kidney. There appeared to be a dose-dependent increase in the SIR for NHL with increased duration of TCE exposure in Danish workers (Raaschou-Nielsen et al., 2003). These researchers also reported a SIR of 1.7 for esophageal adenocarcinoma. Seidler et al. (2007) described the association between TCE exposure of >35 ppm per year and malignant lymphoma (OR = 2.1, CI = 1.0–4.8) as being of borderline statistical significance. Lan et al. (2010) recently observed a dose-dependent decline in major types of lymphocytes in TCE-exposed workers in China. Little or no association with NHL incidence was seen by Boice et al. (2006) or Zhao et al. (2005) in aerospace employees. Evidence of lung cancer in persons occupationally exposed to TCE is limited to nonexistent (Boice et al., 2006; Hansen et al., 2001; Zhao et al., 2005; Raaschou-Nielsen et al., 2003). The latter group of scientists’ estimates of SIRs for lung cancer, for example, were 1.4 for men and 1.9 for women. Most estimates of liver cancer risk in TCE-exposed workers have also been low. SIRs of 2.6 (Hansen et al., 2001) and 2.8 (Raaschou-Nielsen et al., 2003), resulting from a small number of cases, were among the highest reported.

Risk Assessment

Attempts to understand the mechanistic underpinnings of TCE’s carcinogenicity in rodent models and their relevance to humans have resulted in a massive body of published data. TCE provides a stellar example of how experimental data from laboratory and epidemiological studies may or may not impact cancer risk assessment in a regulatory context. The EPA began in the 1990s to revise its guidelines for cancer risk assessment. The final EPA Guidelines for Carcinogen Risk Assessment were released in March 2005. TCE was utilized as a pilot chemical for evaluation and implementation of this guidance document. The guidelines emphasize a scientific “weight of evidence” approach that includes characterization of dose–response relationships, modes of action, and metabolic/TK processes. Where adequate data are available to support reversible binding of the carcinogenic moiety to biological molecules as the initiating event, a nonlinear (ie, threshold) risk assessment/approach is to be used. Otherwise, the default assumption of a linear (ie, no-threshold) model/approach is to be used to estimate cancer risk. As described above, there is considerable evidence that TCE’s oxidative metabolites act via nongenotoxic modes of action. PPARα activation and induction of mouse liver tumors were judged to be irrelevant to humans (NAS, 2006). The EPA, in its 2001a draft TCE Risk Assessment, concluded it was difficult to establish with sufficient certainty the important TCE metabolites, the key events they cause, and their relevance to humans. Despite a substantial increase in information in the last decade, EPA scientists (Caldwell and Kesheva, 2006) contend that knowledge of mechanisms and human relevance is still insufficient to depart from the default assumption. This is exemplified by the critical response of EPA scientists (Caldwell et al., 2006) to the attempt of Clewell and Andersen (2004, 2006) to apply a margin of exposure approach (nonlinear dose–response extrapolation) in their TCE risk assessment. In contrast to liver cancer, kidney cancer is widely accepted to be qualitatively similar in rats and humans, although rats form greater quantities of reactive metabolites via the GSH pathway. Despite genotoxic events, kidney tumor formation in humans is generally believed to require promotion resulting from frank cytotoxicity. Caldwell and Kesheva (2006), however, opine that there may be other modes of action of multiple metabolites operative at low doses. This logic has been retained in the EPA’s most recent IRIS Toxicological Review of TCE (2009).

The EPA (2001a) draft TCE Risk Assessment elicited considerable debate, which prompted the EPA and other federal agencies to request a scientific review of the document by a NAS expert panel. Their report was released in 2006. It concluded there was concordance between rat and human studies of renal carcinogenicity. The preponderance of evidence indicated that humans would be much less susceptible than mice to liver and lung carcinogenesis. It was recommended, among other things, that a new meta-analysis of epidemiological data be conducted. Kelsh and his coworkers recently completed meta-analyses of occupational study data. Studies were classified as group I or II, depending on the quality of their design and exposure assessment. The summary relative risk estimate (SRRE) across all such studies was 1.42 (CI = 1.17–1.77) for kidney cancer (Kelsh et al., 2010). The same research group performed meta-analyses for occupational TCE exposures and liver cancer (Alexander et al., 2007) and NHL (Mandel et al., 2006). The highest SSREs calculated for liver cancer and NHL in any study groupings were 1.30 and 1.59, respectively. The authors noted the results of many studies were inconsistent, information on TCE exposure was often quite limited, and recognized confounding factors were not always taken into account. NAS (2009) recently evaluated published occupational epidemiology studies, and concluded there was limited/suggestive evidence of associations between chronic TCE exposure and kidney cancer. Evidence was deemed inadequate/insufficient to determine whether there was an association with NHL or cancer of the liver, lung, or any other organ. In contrast, Scott and Chiu (2006) opined that modest RR elevations (1.5–2.0) typically reported in positive epidemiology studies provided support for the kidney, liver, and lymphatic systems as target organs. The EPA (2009), in its IRIS document released in September 2011, concluded there was convincing evidence of a causal association between TCE exposure and kidney cancer. Epidemiological evidence was said to be less compelling for NHL, and even more limited for liver cancer. Nevertheless, TCE was characterized by the EPA as “carcinogenic in humans by all routes of exposure.” This is a departure from previous national and international classifications.

Tetrachloroethylene

Tetrachloroethylene (perchloroethylene, PERC) is commonly used as a dry cleaner, fabric finisher, degreaser, rug and upholstery cleaner, paint and stain remover, solvent, and chemical intermediate. The highest exposures usually occur in occupational settings via inhalation. Much attention is being focused on adverse health effects that may be experienced by dry cleaners and other persons living in the proximity of such facilities (Garetano and Gochfeld, 2000). Echeverria et al. (1995), for example, reported adverse effects on visuospatial functions in dry cleaners. PERC is frequently detected in the low ppt range in the breath and blood of the general populace (Ashley et al., 1994; Churchill et al., 2001; Blount et al., 2006). Although releases are primarily to the atmosphere, PERC enters surface and groundwaters by accidental and intentional discharges (ATSDR, 1997c). Levels in the ppb range were reported in municipal water in areas of New England, where PERC was used in a process to treat plastic water pipe (Paulu et al., 1999). Pohl et al. (2008) reported that PERC was the third most frequently found chemical contaminant in groundwater at hazardous waste sites in the United States. A dry cleaner adjacent to Camp Lejeune, in North Carolina, was the source of PERC contamination of some wells that supplied drinking water to the marine base. This contamination prompted a recent study of potential health effects in employees and persons stationed at the base (NAS, 2009).

Metabolism

The systemic disposition and metabolism of PERC and TCE are similar in many respects, although PERC is much less extensively metabolized (ATSDR, 1997c; Chiu et al., 2007). Both chemicals are well absorbed from the lungs and GI tract, distributed to tissues according to their blood flow and lipid content, partially exhaled unchanged, and metabolized. PERC, like TCE, is metabolized by CYP-catalyzed oxidation and GSH conjugation. CYP2E1, however, is not thought to play a major role, in that PERC is considered to be oxidized primarily by the CYP2B family in the rat (Hanioka et al., 1995). In humans, CYP2B6 is the primary isoform responsible for PERC metabolism, and there are minor contributions by CYP1A1 and CYP2C8 (White et al., 2001). The initial metabolite is the epoxide PERC oxide. This metabolic intermediate can be biotransformed to several products (Lash and Parker, 2001). The primary one is trichloroacetyl chloride, which reacts with water to form TCA, the predominant PERC metabolite found in the urine of rodents and humans (Birner et al., 1996; Volkel et al., 1998). Some TCA is converted to DCA.

A small proportion of absorbed PERC undergoes conjugation with GSH to form S-(1,2,2-trichlorovinyl)GSH (TCVG). That initial metabolic step is catalyzed by GSTs and occurs primarily in the liver. TCVG is converted to S-(1,2,2-trichlorovinyl)-l-cysteine (TCVC). TCVC, like the DCVC formed from TCE, is both detoxified in the liver by N-acetylation and metabolically activated by β-lyases in the kidneys. 2,2-Dichlorothioketene can decompose to DCA. Hence, DCA is derived from both GSH- and CYP-dependent biotransformation of PERC. PERC is conjugated with GSH more extensively by rats (1%–2%) (Dekant et al., 1986) than is TCE (<0.005%) (Green et al., 1997). The extent of GSH conjugation of PERC increases when the oxidative pathway begins to become saturated at high exposure levels. Metabolic products of GSH conjugates of TCE and PERC are primary contributors to these halocarbons’ nephrotoxicity (Lash et al., 2007).

Modes of Cytotoxicity/Carcinogenicity

PERC-induced hepatic injury is believed to be a consequence of its oxidative metabolism (Lash and Parker, 2001). Although PERC is more poorly metabolized by CYPs than TCE, two additional intermediate metabolites of PERC contribute to its hepatocytotoxicity: the initial oxidation product, PERC oxide, and one of its convertants, trichloroacetyl chloride. TCA is primarily responsible for activation of the nuclear receptor PPARα, which stimulates peroxisomal enzymes and selected CYPs involved in lipid metabolism. This results in peroxisome proliferation, which generates reactive oxygen moieties that can cause lipid peroxidation, cellular injury, and altered expression of cell-signaling proteins (Bull, 2000). Lash et al. (2007) more recently demonstrated that CYP inhibition resulted in reduced injury of hepatocytes isolated from male F344 rats and exposed to PERC. GSH depletion increased cellular injury, apparently because of a shift from the GSH to the oxidative pathway.

The metabolism and mode of nephrotoxicity of PERC and TCE appear to be quite similar, although PERC and its metabolites are somewhat more potent. Renal effects of both halocarbons are due primarily to metabolites formed via the GSH pathway (Lash and Parker, 2001). The sites, enzymes, and products associated with PERC biotransformation are almost identical to those associated with TCE. The primary difference is that TCVG and TCVC are produced from PERC, and DCVG and DCVC are formed from TCE. TCVC can be detoxified by acetylation or cleaved by renal cytosolic and mitochondrial β-lyases to trichlorothioketene, which loses a chloride ion to form dichlorothioketene. The latter is a very reactive moiety that binds to cellular proteins and DNA. TCVC, as noted above, can be enzymatically oxidized to form the very reactive TCVC sulfoxide (TCVCS) (Krause et al., 2003). TCVCS was shown to be more nephrotoxic than TCVC in male rats after i.p. injection (Elfarra and Krause, 2007). TCVC caused more pronounced necrosis of RPT cells in male rats than did DCVC after i.v. injection (Birner et al., 1997). Lash et al. (2002) similarly found that PERC and TCVG were more toxic than TCE and DCVG in vivo to renal cortical cells of F344 rats. Cells from male rats were more sensitive than cells from females to PERC- and TCVG-induced mitochondrial state 3 respiratory inhibition and cytotoxicity. Isolated rat hepatocytes and their mitochondria, however, were unaffected by PERC and TCVC. Elevated GSH levels enhanced TCE-induced and PERC-induced cytotoxicity in suspensions of rat renal cortical cells but not hepatocytes (Lash et al., 2007). Thus, PERC’s GSH metabolites are both sex- and organ-specific.

Hepatorenal Toxicity

PERC, like TCE, has a limited ability to adversely affect the liver or kidneys of rodents. Near-lethal i.p. doses of PERC were required to cause acute liver injury in mice (Klaassen and Plaa, 1966). Buben and O’Flaherty (1985) saw manifestations of modest hepatocellular damage in male mice given 500 to 2000 mg PERC/kg daily for six weeks by corn oil gavage. A lack of dose dependence reflected saturation of metabolic activation in this dosage range. Philip et al. (2007) reported that male mice given 150 to 1000 mg PERC/kg by aqueous gavage initially exhibited a dose-dependent increase in serum alanine aminotransferase (ALT) activity, but levels of the enzyme regressed substantially over a 30-day dosing period. Hayes et al. (1986) found no consistent dose-related changes in any clinical chemistry parameter in male or female rats that consumed up to 1440 mg PERC/kg in their drinking water for 90 days. Liver injury was not seen in B6C3F1 mice or Osborne–Mendel rats gavaged with up to ~1000 mg/kg daily for 78 weeks (NCI, 1977a,b). Dose-dependent karyomegaly was observed in the renal proximal tubule (RPT) epithelium of mice and rats inhaling PERC for 103 weeks (NTP, 1986). This change was most prominent in the male rats. Tinston (1995) reported mild, progressive glomerulonehropathy and increased pleomorphism of renal tubular nuclei in male but not female rats that inhaled 1000 ppm PERC for up to 19 weeks. Green et al. (1990) found increases in protein droplets and cell proliferation in the proximal tubular epithelium of male F344 rats gavaged with 1500 mg PERC/kg daily for up to 42 days. These alterations were accompanied by an increase in α_2u-globulin, a male rat-specific protein (Fig. 24-6). Goldsworthy et al. (1988) reported similar findings in male but not female F344 rats gavaged for 10 days with 1000 mg PERC/kg per day. The accumulation of α_2u-globulin is cytotoxic, causing cellular necrosis and compensatory cellular proliferation in the P2 segment of renal proximal tubules (Borghoff et al., 1990).

Humans should be less susceptible to hepatorenal injury by PERC than rodents, due to lower target organ doses of the parent compound and its bioactive metabolites. Rats achieve a substantially higher internal dose of VOCs than do humans on inhalation exposures. Volkel et al. (1998) subjected rats and people to identical PERC inhalation regimens. Blood TCA concentrations were 3 to 10 times higher in the rats. DCA was not detectable in human urine, but substantial amounts were found in rat urine. A study of the urinary excretion of total trichloro-metabolites by PERC-exposed workers led Ohtsuki et al. (1983) to conclude that the capacity of humans to metabolize PERC was rather low. Lash and Parker (2001) noted that saturation of PERC metabolism occurred at lower doses in humans than in rodents. This indicates that humans have lower capacity to form biologically active metabolites from moderate to high PERC doses. The difference is reflected in the finding of much lower concentrations of protein adducts in the blood of humans than in the blood of rats subjected to equivalent PERC inhalation exposures (Pahler et al., 1999).

PERC has limited ability to damage the liver of humans. Stewart et al. (1977) found no evidence of hepatotoxicity in six male and six female volunteers exposed randomly to PERC at 0, 25, or 100 ppm 5.5 hours per day, five days per week for 11 weeks. Serum ALT activity was not increased in 22 dry cleaners examined by Lauwerys et al. (1983). A research group in Italy studied 141 employees exposed to PCE in small laundries and dry-cleaning shops (Gennari et al., 1992). No worker exhibited clinical signs of hepatic dysfunction or abnormal serum enzyme concentrations, although there did appear to be an increase in one isozyme of γ-glutamyltransferse, which was said to be associated with hepatobiliary impairment. Another investigation of dry cleaners failed to reveal increases in serum enzymes, but did show mild to moderate changes in hepatic parenchyma revealed by ultrasonography (Brodkin et al., 1995). Thus, considerable experience in occupational settings demonstrates that humans may develop mild but reversible liver injury on chronic exposure to high concentrations of PERC (ATSDR, 1997c).

Occupational exposures to PERC vapor have led to several reports of mild renal tubular damage (ATSDR, 1997c). Employees of dry-cleaning shops have been the subjects of a number of studies of potential kidney effects. Increased concentrations of urinary lysosomal or β-glucuronidase activity were described in dry cleaners exposed to 10 ppm (Franchini et al., 1983) and 23 ppm PERC (Vyskocil et al., 1990) for 9 to 14 years. In a more comprehensive study, a number of urinary indices indicative of early glomerular and tubular changes were increased in 50 dry cleaners who inhaled ~15 ppm PERC for 10 years (Mutti et al., 1992). Verplanke et al. (1999) monitored several indices of tubular and glomerular function in Dutch dry-cleaning workers, but found an increase only in retinol-binding protein in their urine. Other investigators have failed to find evidence of renal effects in such populations. A laboratory study of 10 male and 10 female adults, who inhaled PERC at up to 150 ppm for as long as 7.5 hours per day for five days, did not show changes from preexposure baseline BUN levels (Stewart et al., 1981). Hake and Stewart (1977) described a dry cleaner who was found unconscious in a pool of PERC, where he had been for an estimated 12 hours. Laboratory tests revealed hematuria and proteinuria that lasted for 10 and 20 days, respectively. Mild hepatic damage was revealed by transient increases in serum enzymes. On the basis of the foregoing occupational experiences, PERC has shown limited ability to cause diffuse changes along the nephron, although extremely high exposures can lead to pronounced changes in humans.

Cancer Bioassays in Rodents

High, chronic doses of PERC have been demonstrated to produce species-specific and in certain cases strain- and gender-specific tumors in some organs of mice and rats. Male and female B6C3F1 mice, gavaged with very high doses of PERC (stabilized with epichlorohydrin) for up to 78 weeks, exhibited an increase in hepatocellular carcinoma (NCI, 1977a). No increase in tumor incidence was seen in Osborne–Mendel rats. A significant increase in hepatocellular neoplasms (adenomas and carcinomas combined) was seen in B6C3F1 mice inhaling 100 or 200 ppm PERC for two years (NTP, 1986). There was also a low incidence of renal adenomas and carcinomas in male but not female F344 rats exposed for two years to 200 or 400 ppm PERC vapor. The incidence of renal neoplasms in the males was 1 of 49 controls, 3 of 49 exposed at 200 ppm, and 4 of 49 exposed at 400 ppm (NTP, 1986). Even though the changes were not statistically significant, it was noted that these particular tumors are rare in F344/N male rats, so they were believed to have been caused by PERC exposure. There was also a significant increase over controls in mononuclear-cell leukemia (MNL) in the male and female rats, but it was not clearly dose-dependent. The incidence of this form of leukemia can exceed 70% in F344 controls (Caldwell, 1999; Ishmael and Dugard, 2006). MNL in this rat strain apparently arises from large granular lymphocytes. This lukemic origin is very uncommon in humans (Caldwell, 1999). Gliomas were found in two female and four male F344/N rats exposed to PERC at 400 ppm and in one control male (NTP, 1986). The increase in incidence of this tumor was not statistically significant. Thus, the brain tumors were not considered by NTP (1986) to have been induced by exposure to PERC. The overall incidence of Leydig cell (testicular) tumors was 70%, 80%, and 82% in the 0-, 200-, and 400-ppm groups, respectively. Haseman et al. (1998) reported that NTP control F344 rats have an extremely high spontaneous incidence (89.1%) of Leydig cell tumors. Therefore, these tumors are believed to be irrelevant to men (Ishmael and Dugard, 2006). No increases in lung proliferative lesions seen in B6C3F1 mice of either sex after inhalation of PERC at 100 or 200 ppm for 103 weeks, nor were lung neoplasms seen in male or female F344/N rats exposed at 200 or 400 ppm for 103 weeks (NTP, 1986).

Cancer Epidemiology Studies

There have been many epidemiology studies of cancer incidence and mortality in groups of dry cleaners and other persons occupationally exposed to PERC (ATSDR, 1997c; NAS, 2009; EPA, 2008). Some researchers have reported findings of excess incidences of different cancers, while others have not. Frequently, there was not a determination of the degree or duration of PERC exposure, or consideration of major confounding factors (eg, exposure to other solvents/chemicals, smoking, alcohol consumption, socioeconomic status). Nevertheless, there was sufficient information for Weiss (1995) to conclude that cigarette smoking and alcohol consumption could only partially account for an increased rate of esophageal cancer in dry cleaners. In this instance kidney cancer incidence did not appear to be elevated. Blair et al. (2003) conducted and updated a mortality assessment of a cohort of dry cleaners and found an increased risk of death from esophageal cancer (standardized mortality ratio [SMR] = 2.2, 95% CI = 1.5-3.3). No exposure–response pattern was seen. Chang et al. (2005) did not find any cases of esophageal cancer in a Taiwanese cohort of electronics workers, while Lynge et al. (2006) observed a decreased incidence in dry cleaners in Nordic countries. Nevertheless, NAS (2009) concluded there was limited/suggestive evidence of an association between chronic PERC exposure and esophageal cancer.

There has been little evidence of an association between occupational PERC exposure and liver cancer. Increased incidences or mortality were not seen in cohort studies (Blair et al., 2003; Chang et al., 2005) or in a case–control study (Lynge et al., 2006). A RR of just 0.76 (CI = 0.38–1.52) was estimated in the latter investigation of Nordic dry-cleaning workers. A mortality odds ratio (MOR) of 2.57 (CI = 1.21–5.46) was reported by Lee et al. (2003) in another case–control study, but the exposure assessment was weak. The NAS (2009) recently concluded there was insufficient/inadequate evidence to determine whether an association exists between chronic PERC exposure and hepatobiliary cancer.

Epidemiology studies of populations exposed to PERC generally have shown no increase in risk of lung cancer. Blair et al. (2003), however, in an updated mortality analysis of dry cleaners, reported a SMR of 1.5 (CI = 1.2–1.9) for workers with medium to high exposures. This study adequately characterized exposures and included a relatively large number of subjects. Thus, the NAS (2009) concluded there is limited/suggestive evidence of an association between chronic PERC exposure and lung cancer.

Investigations of kidney cancer risk in PERC-exposed populations have yielded mixed results. A number of studies have not revealed an increased risk (Blair et al., 2003; Boice et al., 1999; Lynge et al., 2006; Mundt et al., 2003; Ruder et al., 2001). However, some of the assessments lacked power due to a limited number of study subjects. Although Blair et al. (2003) reported a very low SMR for kidney cancer deaths in dry cleaners with little or no PERC exposures, the SMR was 1.5 for those with medium or high exposure. Increased incidences have also been reported by Mandel et al. (1995), Pesch et al. (2000), and McCredie and Stewart (1993). The NAS (2009) concluded there is limited/suggestive evidence of an association between chronic PERC exposure and kidney cancer in humans.

Risk Assessment

A great deal of research has been conducted to characterize PERC’s dose–response relationships, modes of action, and metabolism/TK in rodents and humans. It is apparent from the previous discussion there is a considerable body of information available from human experience/exposures to PERC, as well as extensive data sets from studies in different species of laboratory animals. Both types of data are important in assessing risks to human health. Knowledge of the relevance of the animal data to humans is essential in order to meaningfully extrapolate from one species to another.

Humans appear to be less susceptible than rodents to the toxic or carcinogenic actions of PERC, as they are to those of TCE. Humans absorb less inhaled PERC and TCE, attain lower target organ doses of the parent compounds, have lower oxidative and GSH conjugation capacity, and inactivate epoxide intermediates more efficiently. On equivalent inhalation exposures of Wistar rats and humans to PERC, the rats excreted substantially larger amounts of TCA, DCA, and acetylated TCVC (Volkel et al., 1998). These observations were consistent with in vitro findings of greater conversion of PERC to TCVC by rat liver cytosol and of 10-fold higher β-lyase-mediated metabolism of TCVC by rat than by human kidney (Green et al., 1990). These data indicate that the human kidney has limited capacity to generate reactive metabolites from PERC by the GSH and β-lyase pathways. Biotransformation of PERC by the GSH conjugation pathway appears to be quite similar qualitatively, but not quantitatively in male rats and humans.

PERC-induced accumulation of α_2μ-globulin in the P2 segment of renal proximal tubules and the accompanying cytotoxic and regenerative changes are unique to the male rat (Borghoff et al., 1990). α_2μ-Globulin is synthesized under androgenic control in the liver of male rats, reabsorbed by the P2 segment, and undergoes hydrolytic digestion. Accumulation of α_2μ-globulin in cytoplasmic droplets elicits tubular cell necrosis and compensatory cell proliferation (Goldsworthy et al., 1988). Sustained cell proliferation can promote clonal expansion of spontaneously or chemically initiated cells in proximal tubules to form preneoplastic and neoplastic lesions (Swenberg, 1993). The increase in cell proliferation is reversible, as is the binding of halocarbon metabolites to α_2μ-globulin. Thus, this mode of action is not genotoxic and is considered by International Agency for Research on Cancer (IARC) and the EPA to be irrelevant to humans. Melnick and Kohn (1999), however, argue that some data are inconsistent with this conclusion, and that alternative mechanisms may exist. Doi et al. (2007) find some contribution of α_2μ-globulin nephropathy to renal tumors produced by several structurally diverse chemicals in rats, but conclude that other critical processes are probably involved.

Several comprehensive reviews of PERC’s toxicity and carcinogenicity have been published. Lash and Parker (2001) compiled an excellent review of the metabolism, toxicity, and modes of toxicity of the chemical in laboratory animals and humans. The NAS (2009) recently published an evaluation of risks of use of TCE- and PERC-contaminated drinking water. It contained an extensive review of the metabolism, toxicity, and carcinogenicity of the two VOCs. The EPA (2008) released a draft document in support of its IRIS risk assessment, entitled Toxicological Review of Tetrachloroethylene (Perchloroethylene). Quite recently, the NAS (2010) published a review of the EPA document. Questions were raised that called into question the soundness and reliability of EPA’s proposed toxicity and cancer risk estimates. Stated weaknesses included a lack of critical analysis of data, preference given to studies (eg, genotoxicity, epidemiological) showing positive results, and lack of an integrated consideration of the weight of evidence. It is anticipated this review will result in EPA’s publication of a revised PERC risk assessment document.

1,1,1-Trichloroethane

TRI (methyl chloroform) is a widely used organic solvent. Its popularity as a metal degreaser, general purpose solvent, spot cleaner, and component of aerosols and a variety of household products increased substantially with the decline in manufacture of other halocarbons found to be high-dose rodent carcinogens and potential human carcinogens (ATSDR, 2006b). Utilization of TRI diminished during the 1990s, however, due to its ozone-depleting properties (Doherty, 2000). The VOC was to be phased out under the Montreal Protocol by 2002, but is still manufactured in the United States and utilized as a precursor for hydrofluorocarbons. TRI is also still present in some cleaning products. It was found in groundwater at 21% of hazardous waste sites surveyed in the United States (Pohl et al., 2008). TRI and other VOCs in groundwater can remain trapped for years and serve as a source of low-level exposure. The highest exposures to TRI occur via inhalation in occupational settings, but many persons encounter the VOC at home by use of commercial products and tap water containing it. Ashley et al. (1994) reported TRI and other VOCs in the blood of 75% of >600 non-occupationally exposed individuals. TRI and TCE were also frequently detected in a subset of 982 adults examined in the NHANES III survey (Churchill et al., 2001), and in 951 members of the general population (Blount et al., 2006).

Toxicokinetics and Metabolism

The TK of TRI has been characterized in rodents and in humans. TRI is rapidly and extensively absorbed from the lungs (Schumann et al., 1982) and GI tract (White et al., 2013). Systemic uptake and blood levels are elevated in humans inhaling TRI and other VOCs when they exercise (Astrand et al., 1973). Dallas et al. (1989) observed that uptake of inhaled TRI by rats diminished from >80% to 50% to 60% during two hours of exposure, due to its systemic accumulation as a result of slow metabolism. Monster et al. (1979) similarly noted in humans that the capacity of the body to absorb TRI vapor was less than for TCE, due to TRI’s relatively poor metabolism. White et al. (2013) reported the bioavailability of comparable oral doses of TRI and TCE in rats to be ~85% and 45%, respectively. TRI, like other VOCs, is distributed throughout the body, with fat achieving the highest concentrations (Schumann et al., 1982). As TRI biotransformation is quite limited, the VOC is cleared primarily by exhalation by rodents (Reitz et al., 1988; Schumann et al., 1982) and humans (Nolan et al., 1984).

A number of PBTK models for TRI in rodents and in humans have been published. Lu et al. (2008) recently evaluated the suitability of each of these for use in the EPA IRIS database. The model of Reitz et al. (1988) proved the most satisfactory, in that it accurately predicted the time-course of TRI in mouse, rat, and human blood for different inhalation scenarios. No empirical data were available for assessing the accuracy of simulations of brain TRI levels. Warren et al. (1998), however, demonstrated a high degree of correlation between brain and blood levels in TRI-exposed mice and rats, as well as reasonable correlation between brain levels and CNS effects of the VOC. TK data from human and rodent studies were used by Lu et al. (2008) to simulate internal dose metrics for appropriate human exposure scenarios for derivation of IRIS reference values.

Toxicity

The primary pharmacological manifestation of acute or chronic inhalation of TRI is CNS depression, ranging in severity from slight headache or dizziness to anesthesia and death (ATSDR, 2006b). The current TLV of 350 ppm was established to prevent decrements in workers’ mental and physical functions. Volunteers who inhaled 450 ppm TRI for four hours showed no effects on psychophysiological functions (Salvini et al., 1971). Two of 11 other subjects inhaling 500 ppm TRI 6.5 to 7.0 hours daily for five days exhibited difficulty balancing on one foot (Stewart et al., 1969). Muttray et al. (2000) found no psychophysiological effects in volunteers breathing 200 ppm TRI for four hours, but did report EEG changes and slight tiredness in their subjects. Warren et al. (2000) observed rapid, parallel increases in blood and brain TRI concentrations in mice inhaling the VOC. Mice breathing high vapor levels initially exhibited increased locomotor activity, followed by decreased activity. Mattsson et al. (1993) saw large evoked potential and EEG changes in F344 rats during inhalation of 2000 ppm TRI, but no evidence of neurotoxicity (eg, residual neurologic functional or morphological changes) after 13 weeks of exposure of ≤2000 ppm six hours daily, five days per week. Most studies of long-term occupational exposures have not revealed residual neurologic effects, although one assessment of workers subject to near-anesthetic vapor levels did reveal deficits in memory and balance (ATSDR, 2006b). It is worthy of note that very high inhaled concentrations of TRI, particularly when accompanied by hypoxia and stress, can sensitize the myocardium to catecholamines, producing cardiac arrhythmias (Reinhardt et al., 1973).

TRI has a very limited cytotoxic potential, ostensibly due to its limited biotransformation to relatively nontoxic metabolites. A near-lethal acute i.p. dose (3350 mg TRI/kg) was required to significantly enhance serum ALT activity in mice (Klaassen and Plaa, 1966). Male rats given a single oral dose of ~2500 mg/kg exhibited a transient increase in serum aspartate aminotransferase activity, but no ALT increase (Tyson et al., 1983). These enzymes are released from damaged or necrotic hepatocytes into the bloodstream. Male rats gavaged five times weekly for as long as 12 days with up to 5 g TRI/kg died from effects of repeated, protracted CNS depression, but exhibited only slight hepatotoxicity (Bruckner et al., 2001). Quast et al. (1988) saw only minimal histological changes in the liver of F344 rats inhaling 1500 ppm TRI six hours daily, five times per week for up to two years. Rats gavaged daily for 21 days with a high dose of TRI showed increased urinary N-acetyl-β-d-glucosaminidase (NAG) activity, but no microscopic evidence of injury indicative of renal toxicity (NTP, 1996). Liver and/or kidney injury are usually absent or quite modest in occupationally exposed populations, even in fatal cases (ATSDR, 2006b). A report by Hodgson et al. (1989) is an exception. They described four TRI-exposed workers who exhibited elevated serum ALT activity and fatty vacuolation of hepatocytes. In addition, Brogen et al. (1986) reported that 10% of a group of metal workers exposed to TRI, TCE, and Freon 113 had elevated NAG activity in their urine.

Potential Carcinogenicity

The NCI (1977b) conducted a 78-week study in which B6C3F1 mice and F344 rats of both sexes received high doses of TRI daily by gavage. There was no increase in cancers attributable to TRI. In a screening study, Maltoni et al. (1986) observed an apparent increase in leukemias in male and female S-D rats gavaged with 500 mg TRI/kg per day for 104 weeks. Statistical analyses were not presented, and the authors stated that definite conclusions could not be drawn from their work due to limitations in the design and number of animals. A two-year inhalation study in B6C3F1 mice and F344 rats of both sexes revealed no evidence of tumorigenicity due to TRI (Quast et al., 1988). Few epidemiological studies of TRI-exposed populations have been conducted. Infante-Rivard et al. (2005) did report a high risk (OR = 7.55, 95% CI = 0.92–61.97) of childhood leukemia in offspring of women exposed to TRI from two years before pregnancy up to birth. Recently, Gold et al. (2011) reported an OR of 1.8 (1.1–2.9) for multiple myeloma in persons exposed to TRI. It should be recognized that workers are commonly exposed to multiple solvents in the workplace. TRI is currently assigned the classification of D (not classifiable as to carcinogenicity in humans) by the EPA (1998b).

Methylene Chloride

MC (dichloromethane) has enjoyed widespread use as a solvent in industrial processes, manufacture of drugs, degreasing agents, aerosol propellants, agriculture, and food preparation. It was commonly used to decaffeinate coffee and tea. Thus, large numbers of people have been exposed occupationally and in the home. The primary route of exposure to this very volatile solvent is inhalation. The preponderance of MC escaping into the environment does so by volatilization (ATSDR, 2000a). The VOC is also frequently found in wastewater discharges and in air and water at hazardous waste sites (Pohl et al., 2008).

Metabolism

The TK of MC has been well characterized in humans and rodents. MC is rapidly absorbed and distributed throughout the body (Angelo et al., 1986). Inhaled MC reached a near steady state in the blood of human subjects with one to two hours of continuous exposure (DiVincenzo and Kaplan, 1981). Less than 5% of the absorbed dose was exhaled unchanged. Approximately 25% to 34% was exhaled as carbon monoxide (CO), the major end metabolite of MC. Exposure of the volunteers to 50, 100, 150, and 200 ppm for 7.5 hours produced peak blood carboxyhemoglobin saturations of 1.9%, 3.4%, 5.3%, and 6.8%, respectively. MC was very rapidly eliminated from the body and did not accumulate over a five-day exposure regimen. As shown in Fig. 24-7, metabolism of MC in humans and rodents is believed to occur via three pathways (Andersen et al., 1987). One entails CYP2E1-catalyzed oxidation to CO via formyl chloride, a reactive intermediate. The second, a GSH-mediated pathway, involves the theta-class GST, GST-T1. Oxidation is a high-affinity, low-capacity pathway that predominates at the relatively low MC concentrations present in occupational and environmental settings. The GST conjugation is a low-affinity, high-capacity pathway operative at the high exposure levels used in cancer bioassays (Green, 1997). With the third and minor pathway, it is postulated that CO₂ is also formed via the oxidative pathway by reaction of formyl chloride with a nucleophile such as GSH (Watanabe and Guengerich, 2006). The abilities of different species to metabolize MC in the liver by the GST pathway are as follows: mouse >> rat > human high conjugators > hamster > human nonconjugators (Reitz et al., 1989; Thier et al., 1998). Interindividual variation in the ability to biotransform MC via GST-T1 is associated with genetic polymorphisms in humans (Haber et al., 2002).

PBTK models for MC have been developed to assess the relative importance of the oxidative and GST pathways in MC’s toxicity and carcinogenicity. A PBTK model by Andersen et al. (1987) was validated by comparing simulations of blood MC time-course data with data from experiments with mice, rats, and humans. Tumor incidences in mice in chronic bioassays by NTP (Mennear et al., 1988) and Serota et al. (1986a,b) were consistent with model predictions of liver and lung doses of GSH metabolites, but not oxidative metabolites. After extensive review, the EPA adopted this model as a reasonable means of extrapolating NTP bioassay results in mice to humans. This was the first use of PBTK modeling by EPA in a cancer risk assessment. Other investigators such as Reitz et al. (1988, 1989) published deterministic models for MC, which also provided internal dosimetry point estimates of GST pathway metabolites. Forecasts of relatively low tissue doses of GST metabolites in humans resulted in part from the need to saturate the oxidative pathway before appreciable GSH metabolites could be formed.

Recently, probabilistic PBTK models for MC have been developed, which allow for inclusion of intraspecies and interspecies variability in model predictions, as well as quantitative assessment of model uncertainty. Values for model input parameters and TK data from rodent studies from a variety of sources were subjected to MCMC analysis, a Bayesian optimization technique. With this approach, the prior input information can be combined to obtain posterior distributions of key model parameters. El-Masri et al. (1999) and Jonsson and Johanson (2001) used Bayesian analysis solely to evaluate the influence of GST-T1 polymorphism on human cancer risk. Marino et al. (2006) used MCMC analysis to develop a probabilistic PBTK model for MC in mice. The resulting dose metrics (mg MC metabolized by GST/L tissue per day) were three- to four-fold higher than contemporary EPA estimates. David et al. (2006) applied this modeling approach to humans. Inclusion of GSH nonconjugators resulted in a unit cancer risk estimation 500-fold lower than the EPA unit risk at that time. The EPA (2010a) recently concluded in an extensive assessment that this was the best available PBTK model, despite some uncertainties.

Modes of Toxicity/Carcinogenicity

MC has a quite limited cytotoxicity potential. Male and female B6C3F1 mice consuming up to ~2000 mg MC/kg per day in their drinking water for 90 days showed no adverse effects (Kirschmann et al., 1986). Similarly dosed male and female F344 rats exhibited mild to moderate hepatocellular lipid vacuolation and elevated serum enzyme activities at daily dosage levels as low as 166 to 209 mg/kg per day (Kirschmann et al., 1986). Hepatic centrilobular vacuolation and focal necrosis occurred in the liver of rats inhaling 500 to 4000 ppm MC six hours per day, five days per week for two years (Burek et al., 1984; Mennear et al., 1988). Manifestations of kidney damage have been rare in laboratory animals, but have occasionally been reported in persons subjected to high vapor levels (ATSDR, 2000a; EPA, 2010a). There is little information on the identity of MC metabolites that adversely affect the liver or kidney. As described previously, the CO formed by oxidation of MC binds to hemoglobin to produce dose-dependent increases in carboxyhemoglobin. Offspring of pregnant rats inhaling low concentrations of CO have been reported to exhibit permanent learning and memory impairment (De Salvia et al., 1995). It is generally accepted that tissue hypoxia can contribute to CNS depressant effects of MC. There are few reports of residual neurologic dysfunction in MC-exposed workers (Lash et al., 1991; ATSDR, 2000a).

There has been a great deal of research to define mechanisms of MC carcinogenicity, in order to more clearly understand the relevance of the murine tumors to humans (Green, 1997). Liver and lung tumors in mice do not seem to be associated with overt cytotoxicity or increased replicative DNA synthesis (Maronpot et al., 1995). Induction of the tumors in mice is generally believed to be due to a reactive intermediate generated via the GST pathway (Andersen et al., 1987). GST-T1 in liver and lung catalyzes conversion of MC to S-(chloromethyl)glutathione (GSCH₂Cl), which apparently breaks down rapidly to GSH and formaldehyde. Both GSCH₂Cl and formaldehyde are reactive with DNA. MC is usually mutagenic in bacterial assays containing GSH/GST activity. MC produces DNA single-strand breaks (SSBs) in vitro in mouse hepatocytes and lung Clara cells, in which GST is localized in the nucleus (Mainwaring et al., 1996). DNA SSBs were induced in mouse hepatocytes by a 60-fold lower MC concentration than in rat hepatocytes. Rat liver does not show preferential nuclear localization of GST-T1. Some human hepatocytes apparently exhibit nuclear localization, others cytoplasmic (Sherratt et al., 2002). Negative results have been seen in a variety of genotoxicity assays with rat or hamster cell lines with little or no GST activity (EPA, 2010a). Positive results for sister chromatid exchanges, chromosomal aberrations, and the micronucleus test have been obtained in experiments with human cell lines and isolated cells. Negative results were seen in unscheduled DNA synthesis, DNA SSBs, and DNA–protein cross-links (DPX). There is limited evidence of formation of GSCH₂Cl DNA adducts in some hybrid in vitro systems, but the adducts’ instability presents considerable technical challenges to their study. They have yet to be isolated in vivo. Formaldehyde produces both DPX and SSBs, indicative of its prominent role in MC’s carcinogenicity (Graves and Green, 1996).

Cancer Bioassays in Rodents

High, chronic exposures to MC have been found to produce species- and gender-specific tumors in some organs of mice and rats. Serota et al. (1986a,b) administered a series of doses of MC to F344 rats and B6C3F1 mice in their drinking water for two years. The male mice showed a trend for an increase in hepatocellular adenomas and carcinomas, but the modest response was not dose-dependent. There was a statistically significant increase in neoplastic nodules or hepatocellular carcinoma in some groups of female F344 rats. Burek et al. (1984) saw a significant increase in salivary gland sarcomas in male S-D rats that inhaled 3500 ppm six hours per day, five days per week for two years. The number of benign mammary tumors per tumor-bearing female S-D rat increased with increasing concentration of exposure, although the number of tumor-bearing rats was not significantly elevated over controls. Similarly exposed hamsters were unaffected. In a two-year follow-up study (Nitschke et al., 1988), female S-D rats inhaling 500 ppm MC also only exhibited an increased number of benign mammary tumors per tumor-bearing rat. No increase in malignant tumors was manifest in male rats. NTP conducted an inhalation study, in which F344/N rats and B6C3F1 mice inhaled up to 4000 ppm MC six hours daily, five times weekly for two years (Mennear et al., 1988). There were weak trends for neoplastic nodules and hepatocellular carcinoma, as well as benign mammary tumors in female rats. Male and female mice inhaling 2000 or 4000 ppm MC exhibited statistically significant, dose-dependent increases in hepatocellular adenoma and carcinoma, as well as bronchoalveolar adenoma and carcinoma. There were similar findings in a follow-up study of female B6C3F1 mice inhaling 2000 ppm MC for two years (Maronpot et al., 1995). The incidences of these hepatic and lung tumors in control B6C3F1 mice are quite high (Haseman et al., 1998).

Cancer Epidemiology Studies

Despite a substantial number of epidemiology studies of MC-exposed workers, evidence of associations between MC and specific tumors is not strong. There have been four cohort mortality studies of employees at facilities where MC was used as a solvent for cellulose acetate. There were no increased risks of cancer mortality for all tissues or for lung or breast. In just one assessment was there an elevated risk (SMR = 2.98 [95% CI = 0.81–7.63]) of death from liver and biliary tract cancer (Lanes et al., 1993). This was apparently the sole report of increased hepatic cancer mortality in an occupational population. No investigations provide evidence of an association between MC and lung or kidney cancer (ATSDR, 2000a; EPA, 2010a). Cantor et al. (1995) conducted a case–control study of 33,509 occupationally exposed women, but found little association between MC exposure probability and breast cancer mortality. Blair et al. (1998), however, reported a RR of 3.0 (1.0–8.8) for breast cancer in 3605 women employed at Hill Air Force Base. Blair et al. (1998) also estimated RRs of 3.0 (0.9–10.0) and 3.4 (0.9–13.2) for NHL and multiple myeloma, respectively, in MC-exposed aircraft maintenance employees of both sexes. Occasionally, excess risks of other cancers have been found in highly exposed groups. Gibbs et al. (1996), for example, computed a statistically significant SMR of 2.08 for prostate cancer death of cellulose acetate workers with a latency period of at least 20 years since their first exposure to 350 to 700 ppm MC. In a study of association between MC and astrocytic brain cancer, Heineman et al. (1994) calculated an OR of 2.4 for males with a high probability of MC exposure and for intense exposure versus unexposed controls. Lastly, Infante-Rivard et al. (2005) reported an increased risk of childhood leukemia (OR = 3.22 [0.88–11.7]) in offspring of mothers with probable or definite occupational MC exposure. Nevertheless, most investigations have revealed weak or no apparent associations between relatively high MC inhalation exposures in industry and cancers (Dell et al., 1999; Starr et al., 2006).

Risk Assessment

A considerable body of scientific information supports the following conclusion: should MC be a carcinogen in humans, it is much less potent than in rodents, notably mice. DPX were detected in hepatocytes that had been isolated from B6C3F1 mice and incubated for two hours with MC. They were not found in hepatocytes of F344 rats, Syrian golden hamsters, or three human subjects (Casanova et al., 1996, 1997). RNA–formaldehyde cross-links, however, were found in hepatocytes of all species. These links were four-, seven-, and 14-fold higher in cells from mice than in cells from rats, humans, and hamsters, respectively. Metabolism of MC via the GSH pathway is an order of magnitude greater in mouse than in rat liver. Metabolic rates in hamster and human liver are even lower (Reitz et al., 1989; Thier et al., 1998; Sherratt et al., 2002). High GST-T1 activity was measured in the nuclei of mouse centrilobular hepatocytes. Mice may be unique in that the extensive metabolic activation of MC to an unstable intermediate occurs in the proximity of the DNA. It would be useful in future PBTK models to include subcompartments for cytosolic and nuclear GST activities (Starr et al., 2006). GST-T1 was also detected in relatively high levels in mouse lung Clara cells and ciliated cells at alveolar/bronchiolar junctions (Mainwaring et al., 1996). Clara cells are present in much lower numbers in rats, and are rare in human lungs. GSCH₂Cl apparently causes SSB in vivo and in vitro in DNA of mouse liver and lung (Graves et al., 1995). No DNA breaks were detected in hamster or human hepatocytes in vitro.

The EPA (2010a,b) has recently concluded that MC is likely to be carcinogenic in humans and appears to act via a mutagenic mode of action. The designation of “likely to be carcinogenic in humans” was based largely on the NTP (Mennear et al., 1988) findings of cancer at two sites (liver and lung) in male and female B6C3F1 mice. More limited findings of certain tumors in MC-exposed rats were considered to be supporting evidence. Epidemiological studies were said to provide some evidence of an association between occupational MC exposure and brain and liver cancer. As described in the previous paragraph, bioactivation of MC is qualitatively, but not quantitatively similar in mice and humans. Due to the similarity, the EPA (2010a) reasons that the apparent mode of action (mutagenicity of GST pathway metabolites) is biologically plausible in humans. The currently recommended inhalation risk value is ~47-fold lower than the previous IRIS value. EPA classifies MC as likely to be carcinogenic in humans.

Carbon Tetrachloride

CCl₄ previously enjoyed widespread use as a solvent, cleaning agent, fire extinguisher, synthetic intermediate, grain fumigant, and human anthelmintic. Its use has steadily declined since the 1970s, due to its hepatorenal toxicity, carcinogenicity, and contribution to atmospheric ozone depletion (ATSDR, 2005). Nevertheless, CCl₄ appears to be ubiquitous in ambient air in the United States, and it is still found in some water wells and waste sites. CCl₄ is a classic hepatotoxin, but kidney injury is often more severe in humans.

The timecourse of CCl₄-induced acute liver injury has been well characterized (ATSDR, 2005). Early signs of hepatocellular injury in rats include dissociation of polysomes and ribosomes from rough endoplasmic reticulum, disarray of smooth endoplasmic reticulum, inhibition of protein synthesis, and triglyceride accumulation. Hypomethylation of RNA is thought to contribute to inhibition of lipoprotein synthesis, thereby playing a role in steatosis (Clawson et al., 1987). Ingested CCl₄ reaches the liver, undergoes metabolic activation, produces lipid peroxidation, covalently binds, and inhibits microsomal ATPase activity within minutes in rats. Single cell necrosis, evident five to six hours postdosing, progresses to maximal centrilobular necrosis within 24 to 48 hours. Most microsomal enzyme activities are significantly depressed (Recknagel et al., 1989). A variety of cytoplasmic enzymes are released from dead and dying hepatocytes into the bloodstream. The activity of these enzymes in serum generally parallels the extent of necrosis in the liver. Cellular regeneration, manifest by increased DNA synthesis and cell cycle progression, is maximal 36 to 48 hours postdosing (Rao et al., 1997).

The metabolism of CCl₄ is required for its conversion to a variety of cytotoxic agents. It is widely recognized that CCl₄ is bioactivated by cytochromes P450 via reductive dehalogenation to the trichloromethyl radical (CCl₃^•), which can react in turn with oxygen to form trichloromethyl peroxy free radicals (CCl₃OO^•). Both unstable radicals bind covalently to a variety of cellular components including enzymatic and structural proteins and polyunsaturated fatty acids in membranes. This results in lipoperoxidation, loss of intracellular and cellular membrane integrity, and leakage of enzymes (Plaa, 2000; Weber et al., 2003). By-products of lipid peroxidation include reactive aldehydes, which can form adducts with proteins and DNA, contributing to cytotoxicity and carcinogenicity, respectively (Manibusan et al., 2007). Liu et al. (1995) have proposed that CCl₄ oxidative stress in the liver enhances nuclear factor kappa B activity, which in turn promotes expression of proinflammatory cytotoxic cytokines. Shi et al. (1998) proposed apoptosis as an additional/alternate mechanism of CCl₄-induced cell death.

Perturbation of intracellular calcium (Ca²⁺) homeostasis appears to be an integral part of CCl₄ cytotoxicity (Stoyanovsky and Cederbaum, 1996). Increased cytosolic Ca²⁺ levels may result from influx of extracellular Ca²⁺ due to plasma membrane damage and from decreased intracellular Ca²⁺ sequestration. Elevation of intracellular Ca²⁺ in hepatocytes can cause activation of phospholipase A₂ and exacerbation of membrane damage (Glende and Recknagel, 1992). Elevated Ca²⁺ may also be involved in alterations in calmodulin and phosphorylase activity, as well as changes in nuclear protein kinase C activity (Omura et al., 1999). High intracellular Ca²⁺ levels activate a number of catabolic enzymes including proteases, endonucleases, and phospholipases, which kill cells via apoptosis or necrosis (Weber et al., 2003). The hydrolytic enzyme calpain mediates progression of acute CCl₄-induced liver injury by leaking from dying hepatocytes and attacking neighboring cells (Limaye et al., 2003). Ca²⁺ may stimulate the release of cytokines and eicosanoids from Kupffer cells. Edwards et al. (1993) demonstrated that destruction of Kupffer cells prior to CCl₄ dosing of rats resulted in significant reductions in neutrophil infiltration and hepatocellular injury. Macrophages are known to release a number of inflammatory mediators, such as tumor necrosis factor alpha (TNF-α), that are cytotoxic (Morio et al., 2001).

Development of cellular resistance and tissue repair are important in limiting CCl₄ hepatotoxicity, and in recovery (Mehendale, 2005). Alterations in transmembrane carrier proteins have been discovered in hepatocytes of CCl₄-treated mice. CCl₄ results in reduced expression of genes associated with extraction of bile acids and organic ions from sinusoidal blood, as well as upregulation of certain detoxification genes (Aleksunes et al., 2005). It also produces differential upregulation of multidrug resistance proteins that are involved in export of oxidative stress products and metabolites. Hepatocellular regeneration has been shown to begin within six hours of a small dose of CCl₄, just as centrilobular necrosis is becoming evident (Lockard et al., 1983). This early phase regeneration (arrested G₂ hepatocytes activated to proceed through mitosis) is followed at ~24 hours by the secondary phase of regeneration (hepatocytes mobilized from G_o/G₁ to proceed through mitosis) (Bell et al., 1988). CCl₄ hepatotoxicity is obviously a complex, multifactorial process that is likely to continue to receive considerable attention.

CCl₄ has frequently been used as a model hepatotoxic compound with which to examine the influence of various factors that alter P450s. CYP2E1 is primarily responsible for catalyzing the bioactivation of low doses of CCl₄ in humans. CYP3A contributes to the metabolism of higher doses (Zangar et al., 2000). The preeminent role of CYP2E1 in animals is clearly demonstrated by the protection afforded to CCl₄-treated rodents by CYP2E1 antibody (Castillo et al., 1992), the CYP2E1 inhibitor 3-amino-1,2,4-triazole (Padron et al., 1996), and the absence of CYP2E1 expression (Wong et al., 1998). As discussed previously in the subsection “P450 Inducers,” a variety of conditions that induce CYP2E1 potentiate CCl₄ hepatotoxicity in test animals and humans. Sufficient doses of CYP2E1 inhibitors, including natural constituents of foods (described in the subsection “P450 Inhibitors”), can inhibit CCl₄ toxicity (Lieber, 1997). Taieb et al. (2005) discovered protein 8, a transcription factor that regulates the expression of genes that protect cells from stress, rapidly triggering CYP2E1 downregulation in CCl₄-dosed mice, thereby minimizing CCl₄ bioactivation.

CCl₄ has been found to be a hepatocarcinogen in rodents, but there is relatively little experimental evidence on whether it is genotoxic or carcinogenic in humans (ATSDR, 2005; EPA, 2010b). There have been extensive studies of its potential genotoxic and mutagenic effects, but the results are largely negative in bacterial and in mammalian systems. Araki et al. (2004) did report that CCl₄ was mutagenic in a strain of E. coli that is particularly sensitive to oxidative damage. As early as the 1940s, the National Cancer Institute conducted a series of chronic bioassays in which very high oral bolus doses of CCl₄ were given to mice and rats. Large increases over controls in incidences of liver tumors were found in male and female mice (ATSDR, 2005). Relatively little was seen in rats, but hamsters were susceptible to CCl₄-induced liver cancer. Recently, Nagano et al. (2007) published the results of a study in which F344 rats and BDF1 mice of both sexes were exposed six hours daily, five times weekly for 104 weeks to 0, 5, 25, or 125 ppm CCl₄ vapor. There was a significant increase in hepatocellular adenomas and carcinomas in the male and female rats at the highest exposure level. Significant increases in these tumors and in adrenal pheochromocytomas were manifest in the 25- and 125-ppm male and female mice. There was a statistically significant, but more modest elevation in hepatocellular adenomas, but not carcinomas, in the 5-ppm female mice. Degenerative and necrotic hepatic changes were seen in livers of all groups of animals with liver tumors except the 5-ppm female mice. There has been limited evidence of associations between occupational CCl₄ exposure and certain cancers described in some epidemiology studies, but the data were not conclusive. Exposures to CCl₄ were poorly characterized and confounded by the concurrent exposures of most subjects to other chemicals in workplaces. There were no reported associations with liver cancer.

The weight of scientific evidence indicates that CCl₄ is more likely an indirect than a direct acting mutagen/carcinogen (EPA, 2010b). Manibusan et al. (2007) concluded that sustained cell death, regeneration, and proliferation enhance the likelihood of unrepaired spontaneous lipid peroxidation and endonuclease-induced mutations that may lead to hepatocarcinogenesis. Jiang et al. (2004) observed changes in expression of genes involved in cell death, proliferation, DNA damage, and fibrogenesis in livers of mice given a high CCl₄ dose daily for four weeks. Four weeks after cessation of this treatment, most gene expression profiles returned to control levels, except fibrogenesis. Bioactivated CCl₄ can apparently exert modest genotoxic effects, such as DNA breakage and related sequelae only under highly cytotoxic conditions. ACGIH (2012) assigned CCl₄ the designation of A2 (suspected human carcinogen), in light of its threshold mode of action and its very weak or absent genotoxicity. Germany placed CCl₄ into its category 4, indicating that genotoxicity plays no, or at most a minor, role in its mode of action (MAK, 2011). The EPA (2010b), however, recently categorized the chemical as “likely to be carcinogenic to humans.” Despite acknowledging the correlation between hepatocellular cytotoxicity, regenerative hyperplasia, and induction of liver tumors in rodents, concern about the reactivity of direct and indirect products of CCl₄ metabolism and about limited knowledge of key events led the agency to conclude that “the mode of action was unknown.” This led to use of a linear model to estimate human cancer risks from oral and inhalation exposures. PBPK models were used to estimate mouse internal doses and human equivalent doses as part of this risk assessment.

Chloroform

The primary use of CHCl₃ (trichloromethane) is in the production of the refrigerant chlorodifluoromethane (Freon 22), but this use is diminishing as chlorine-containing fluorocarbons are phased out under the Montreal Protocol. CHCl₃ was among the first inhalation anesthetics, but it was replaced by safer compounds after about 1940. It is a by-product of drinking water chlorination and has been measured in municipal drinking water supplies in concentrations as high as several hundred ppb, although levels are usually <25 ppb (ATSDR, 1997a). CHCl₃ has also been found in ppb concentrations in swimming pool water and surrounding air (Aggazzotti et al., 1995). Like many other halocarbons, CHCl₃ can invoke CNS symptoms at subanesthetic concentrations similar to those of alcohol intoxication and can sensitize the myocardium to catecholamines, possibly resulting in cardiac arrhythmias.

The reproductive and developmental toxicities of CHCl₃ are rather unremarkable. Schwetz et al. (1974) found that inhalation of 100 to 300 ppm CHCl₃ by pregnant rats caused a high incidence of fetal resorption, retardation of fetal development, and a low incidence of fetal anomalies. Murray et al. (1979) reported that gestational exposure to 100 ppm CHCl₃ resulted in the decreased ability of mice to maintain pregnancy, as well as cleft palate, decreased fetal weight and length, and decreased ossification in pups. Very high CHCl₃ concentrations retarded development and induced diffuse cell death in cultured rat embryos (Brown-Woodman et al., 1998). These studies support CHCl₃ as a weak teratogen, but negative studies employing maternally toxic doses argue against this characterization (Thompson et al., 1974; NTP, 1997a). The EPA’s 2001 IRIS profile for CHCl₃ notes that in reproductive/developmental studies, maternal toxicity and fetal effects occurred at doses higher than those that produced liver toxicity.

Hepatotoxicity serves as the basis for EPA’s benchmark dose–based RfD (EPA, 2001b). Under certain conditions CHCl₃ is hepatotoxic and nephrotoxic. These toxicities are potentiated by aliphatic alcohols, ketones, DCA, and TCA (Davis, 1992). Albeit at low doses, numerous disinfection by-products such as the rodent carcinogens TCA and DCA are routinely consumed with CHCl₃ in finished drinking water. For this reason, mixture studies are particularly relevant. Consider, for example, the studies of Pereira et al. (2001) in which N-methyl-N-nitrosourea-initiated B6C3F1 mice were exposed to DCA with or without CHCl₃. CHCl₃ prevented hypomethylation, which would have resulted in increased mRNA expression of the proto-oncogene c-myc and promotion of liver tumors by DCA. Conversely, CHCl₃ increased DCA-induced DNA hypomethylation and enhanced the DCA promotion of kidney tumors. Thus, concurrent exposure to two rodent carcinogens, CHCl₃ and DCA, resulted in less than additive activity in one organ and synergism in another (Pereira et al., 2001; Tao et al., 2005). This exemplifies the difficulty in assessing the risks posed by solvent mixtures.

Tolerance or adaptation to CHCl₃’s hepatorenal toxicity and carcinogenicity has been observed in some mouse strains after repeated exposure. This phenomenon was first investigated by Pereira and Grothaus (1997), who reported pre-exposure of mice to low doses of CHCl₃ in drinking water–induced resistance to hepatotoxicity and cell proliferation following a higher gavage dose. This was presumed at the time to result from suicidal inhibition of the CYPs responsible for CHCl₃’s activation. More recently, mice exposed daily for 7, 14, and 30 days were found to have a robust regenerative response in target tissues, which prevented the progression of injury (Anand et al., 2005). Blood and tissue levels of CHCl₃ after repeated exposure were substantially lower than those following a single exposure, owing to increased elimination of CHCl₃ via exhalation (Anand et al., 2006b). These same researchers also reported that priming mice with CHCl₃ prior to a lethal dose stimulated compensatory hepatogenic and nephrogenic repair, limiting the progression of injury and resulting in 100% survival. Relative to unprimed mice, there was no difference in hepatic or renal CYP2E1 activity, although GSH and GSH reductase activity were upregulated in the kidney (but not in the liver), with a consequent decrease in renal covalent binding. The area under the blood level versus time curve (AUC) was 40% lower in primed versus unprimed mice, but increased elimination via exhalation was not responsible for the reduction in internal exposure in this particular case (Philip et al., 2006). Taken together, these studies suggest that TK and TD factors contribute to the tolerance observed to CHCl₃ toxicity. Mehendele and colleagues have conducted a series of other studies with CHCl₃, alone or in combination with other hepatotoxicants, to further discern the role of tissue repair in toxicant-induced injury. These studies have provided valuable insight into the importance of both the timing of the tissue repair response and its magnitude as pivotal determinants of the outcome of toxicant-induced injury, emphasizing the need to consider repair processes in predictive toxicology (Anand et al., 2003, 2005; Mehendale, 2005).

The status of CHCl₃ as a rodent carcinogen is indisputable. It causes liver and kidney tumors that are species-, strain-, sex-, and route of exposure–dependent. CHCl₃-induced liver tumors in mice and their dependence on ongoing liver necrosis were reported near the end of World War II (Eschenbrenner and Miller, 1945). These same authors observed that male but not female mice suffered kidney necrosis. This observation was supported by the report of Roe et al. (1979) that CHCl₃ ingested in a toothpaste base resulted in renal tumors in male but not female mice. This sex difference is thought to be attributable to testosterone-mediated differences in renal CYP activity (Smith et al., 1984). The NCI (1976) cancer bioassay demonstrated renal tumors in male rats and an extremely high incidence of liver tumors in both sexes of B6C3F1 mice gavaged with CHCl₃ in corn oil. In 1985, Jorgenson and colleagues reported that daily doses of CHCl₃ in drinking water comparable to those in the NCI gavage assay also produced renal tumors in rats, but failed to cause liver tumors in B6C3F1 mice. This finding provided evidence that the dose rate of CHCl₃ was a determinant of liver tumor formation, supporting the existence of a threshold mechanism. Hard et al. (2000) have reevaluated the kidneys from the Jorgenson et al. (1985) study and have confirmed the presence of chronic renal tubule injury, indicative of renal tumor formation via an epigenetic mechanism. In what may be a landmark study, Larson et al. (1994) compared cytotoxicity and cell proliferation in female B6C3F1 mice given CHCl₃ by gavage in corn oil versus ad libitum ingestion in drinking water. As seen in Fig. 24-8, the hepatocyte labeling index, a measure of the proliferative response, differed between the two exposure regimens at comparable doses. Pereira (1994) reported essentially the same observation. This suggests that ingestion of CHCl₃ in small increments, similar to drinking water patterns of humans, fails to produce a sufficient amount of cytotoxic metabolite(s) per unit time to overwhelm detoxification and other protective mechanisms.

Potentiation of CHCl₃’s toxicity by CYP inducers and protection by GSH and P450 inhibitors suggest that a metabolite, presumably phosgene, is responsible for CHCl₃’s hepatorenal toxicity. Both target organs metabolize CHCl₃ to phosgene. There is evidence that CYP2E1 and CYP2B1/2 metabolically activate CHCl₃. The former isoform is thought to catalyze CHCl₃ metabolism at a lower substrate concentration than the latter (Nakajima et al., 1995). By using an irreversible CYP2E1 inhibitor and CYP2E1 knockout mice, Constan et al. (1999) have demonstrated that metabolism of CHCl₃ by CYP2E1 is required for liver and kidney necrosis and cell proliferation. The electrophilic intermediate generated by CHCl₃’s metabolism (ie, phosgene) is initially detoxified by covalently binding to cytosolic GSH. Once GSH is depleted, phosgene is free to covalently bind to hepatic and renal proteins and lipids. Such binding damages membranes and other intracellular structures, leading to necrosis and subsequent reparative cellular proliferation. Sustained proliferation with repeated exposures promotes tumor formation in rodents by irreversibly “fixing” spontaneously altered DNA and clonally expanding initiated cells. The expression of certain genes, including myc and fos, is altered during regenerative cell proliferation in response to CHCl₃-induced cytotoxicity (Sprankle et al., 1996; Kegelmeyer et al., 1997). While the identity of phosgene’s intracellular targets is largely unknown, Guastadisegni et al. (1999) have reported that phosgene reacts with phosphatidylethanolamine (PE). The adduct formed appears to consist of two PE moieties cross-linked at the amino head groups by the carbonyl moiety of phosgene. CHCl₃-modified PE preferentially accumulates on inner mitochondrial membranes, inducing ultrastructural modifications and inhibiting functions of the organelle. These researchers observed the induction of hepatic apoptosis and necrosis in CHCl₃-treated rats and pointed out that apoptosis may be initiated by the release of regulatory factors normally sequestered in mitochondria, in particular Ca²⁺. Evidence that Ca²⁺ perturbation plays a role in CHCl₃ toxicity comes from a report of Ca²⁺ mobilization in Madin–Darby canine kidney cells using Fura-2 as a Ca²⁺ probe. CHCl₃, albeit in millimolar concentrations, increased the cytosolic Ca²⁺ levels by releasing Ca²⁺ from multiple sites within the cell (Jan et al., 2000).

There is no evidence of covalent binding of CHCl₃ metabolites to nucleic acids. There is binding to nuclear histone, which plays a key role in controlling DNA expression and might be a mechanism of CHCl₃’s carcinogenicity (Diaz and Castro, 1980; Fabrizi et al., 2003). It has been hypothesized that the induction of oxidative stress and depletion of GSH by CHCl₃ may lead to indirect genotoxicity that could contribute to carcinogenicity. This hypothesis is supported by the small dose-dependent increase in M(1)dG adducts (malondialdehyde reacts with DNA to form adducts to deoxyguanosine), DNA strand breakage, and lipid peroxidation in CHCl₃-treated rat hepatocytes in the absence of any increase in DNA oxidation (Beddowes et al., 2003). Such a mechanism would still be threshold dependent, given its reliance on the initial depletion of antioxidants.

The EPA (2001b) classifies CHCl₃ as a probable human carcinogen (group B2), meaning there is sufficient evidence for carcinogenicity in animals and inadequate or no evidence in humans. Experimental evidence and the prevailing opinion that CHCl₃ is nongenotoxic indicates that the relationship between CHCl₃ dose and tumor formation is nonlinear. The EPA has, as is called for in its Guidelines for Carcinogen Risk Assessment (EPA, 2005), considered mode of action in the determination of CHCl₃’s cancer risk and relied on a nonlinear dose–response approach and the use of margin of exposure analysis. In doing so, the Agency concluded that the RfD for noncancer effects, based on the dog study of Heywood et al. (1979), was adequately protective for cancer by the oral route on the basis of cancer and noncancer effects having a common link through cytotoxicity. The wealth of mechanistic data available continues to inform the risk assessment for CHCl₃. For example, Constan et al. (2002) exercised a PBTK dosimetry model to compare hepatic responses in mice and humans with inhaled CHCl₃. They concluded that no safety factor was needed to account for interspecies differences in inhalation cancer risk. Additionally, Tan et al. (2003) have published a PBTK/TD model for CHCl₃ to describe the plausible mechanism linking the hepatic metabolism of CHCl₃ to hepatocellular killing and regenerative proliferation, thereby creating a predictive model that most accurately reflects the current science.

Benzene

Benzene produced commercially in the United States is derived primarily from petroleum. It has been utilized as a general purpose solvent, but it is now used principally in the synthesis of other chemicals (ATSDR, 2006c). The percentage by volume of benzene in gasoline is 1% to 2%. Benzene plays an important role in unleaded gasoline (UG) due to its antiknock properties. Inhalation is the primary route of exposure in industrial and in everyday settings. Benzene is present in the atmosphere both from natural sources such as forest fires and oil seeps and from industrial uses including automobile exhaust and gasoline vapor emission. Cigarette smoke is the major source of benzene in the home (Wallace, 1996). Smokers have benzene body burdens that are six to 10 times greater than those of nonsmokers. Passive smoke can be a significant source of benzene exposure to nonsmokers. Because of the health risks of benzene exposure, the use and environmental release of this chemical has significantly diminished in the last two decades.

There is strong evidence from epidemiological studies that high-level benzene exposures result in an increased risk of acute myelogenous leukemia (AML) in humans (Bergsagel et al., 1999; ATSDR, 2006a). Evidence of increased risks of other cancers in such populations is less compelling. Only AML incidence was significantly elevated in the largest cohort study to date, in which ~75,000 benzene-exposed workers in 12 cities in China were evaluated (Yin et al., 1996). Types of cancer other than AML have been tentatively attributed to benzene. Marginal, nonsignificant increases were seen for lung cancer and chronic myelogenous leukemia. Some investigations of persons exposed to engine exhausts have reported a significant association with multiple myeloma. Bezabeth et al. (1996) and Bergsagel et al. (1999) have concluded, however, that there is no scientific evidence to support a causal relationship between benzene exposure and multiple myeloma or NHL. Increased incidences of malignant lymphomas and a variety of solid tumors were found in male and female B6C3F1 mice dosed orally with high doses of benzene for up to 103 weeks (Huff et al., 1989). Male and female F344 rats in this bioassay exhibited excesses of Zymbal gland, skin, and oral cavity carcinomas. Benzene exposure via inhalation, dermal, and i.p. routes also produced a variety of tumors in rodents. Thus, benzene is clearly an animal and human carcinogen, but major species differences exist.

The most important adverse effect of benzene is hematopoietic toxicity, which precedes leukemia. Chronic, high benzene exposure can lead to bone marrow damage, which may be manifest initially as anemia, leukopenia, thrombocytopenia, or a combination of these. Bone marrow depression appears to be dose-dependent in both laboratory animals and humans. Continued exposure may result in marrow aplasia and pancytopenia, an often fatal outcome. Survivors of aplastic anemia frequently exhibit a preneoplastic state, termed myelodysplasia, which may progress to myelogenous leukemia (Snyder, 2002; Bird et al., 2005). Decreased lymphocyte counts are shown to be the early and consistent finding in animals and humans following benzene exposure (Goldstein, 1988). Hematologic effects in workers with prolonged benzene exposure in the range of one to 30 ppm and above have been well documented (Ward et al., 1996). However, there is conflicting evidence for effects at levels <1 ppm, the permissible exposure level (PEL) recommended by US OSHA. Three large epidemiological studies (two in the United States and one in the Netherlands) have not revealed any difference in blood parameters between exposed (<1 ppm) and nonexposed workers (Collins et al., 1997; Tsai et al., 2004; Swaen et al., 2010). However, Lan et al. (2004) have reported dose-dependent hematotoxicity in approximately 250 Chinese shoe workers who inhaled low (≤1 to ≥10 ppm) levels of benzene. Exposures of 100 of these employees to ≤1 ppm reduced mature white cell and platelet counts, as well as myeloid progenitor cell colony formation. The progenitors were more sensitive to benzene than mature white cells. Polymorphisms in myeloperoxidase (an enzyme that metabolizes benzene to toxic quinones and free radicals) and NAD(P)H:quinone oxidoreductase (an enzyme that protects against these moieties) conferred increased susceptibility to white cell decreases. Turtletaub and Mani (2003) found that formation of DNA and protein adducts in mouse liver and bone marrow was dose-dependent over a range of benzene doses, the lowest of which was 5 μg/kg.

Adverse health outcomes of chronic benzene exposure have been known for a long time, but the mechanisms by which benzene causes effects remain incompletely understood. The toxicity of benzene is dependent on its metabolites. Thus, it is essential to understand the metabolism of benzene in order to address its mechanisms of toxicity. The initial metabolic step (Fig. 24-9), oxidation of benzene to an epoxide (ie, benzene oxide), is catalyzed primarily by hepatic CYP2E1. Valentine et al. (1996) demonstrated that benzene-treated transgenic CYP2E1 knockout mice had relatively low levels of all benzene metabolites in their urine. A five-day, 200-ppm benzene inhalation regimen produced severe genotoxicity and cytotoxicity in wild-type B6C3F1 mice, but no adverse effects in the knockout mice. Benzene oxide, which is in equilibrium with its oxepin form, is further metabolized by three pathways: (1) conjugation with GSH to form a premercapturic acid, which is converted to phenylmercapturic acid; (2) rearrangement nonenzymatically to form phenol; and (3) hydration by epoxide hydrolase to benzene dihydrodiol, which in turn can be oxidized by dihydrodiol dehydrogenase to catechol. If phenol is hydroxylated in the ortho position, more catechol will be formed. Catechol can be converted to o-benzoquinone. If benzene is hydroxylated in the para position, p-hydroquinone is formed. It can be oxidized to p-benzoquinone. The o- and p-benzoquinones are considered to be among the ultimate toxic metabolites of benzene. Another potentially toxic metabolite, muconaldehyde, may arise from ring opening of oxepin. Muconaldehyde undergoes a series of reactions that ultimately lead to t,t-muconic acid, an end metabolite found in the urine (Golding and Watson, 1999; Snyder, 2004).

Livers is the primary metabolic organ, while bone marrow is the target organ for benzene. Transport of primary metabolites and further metabolism in bone marrow is believed by many authorities to play the key role in myelotoxicity. Levels of DNA and protein adducts were significantly higher in bone marrow than in liver of benzene-dosed mice (Turtletaub and Mani, 2003). Bernauer et al. (1999) measured activity and levels of CYP2E1 in liver and bone marrow of five strains of mice of varying sensitivity to benzene. CYP2E1 amounts and activities were considerably greater in liver than in marrow, but no interstrain differences were seen in either tissue. It is likely that non-CYP enzymes (eg, peroxidase, myeloperoxidase, and cyclooxygenase) play an important role in generation of semiquinones and quinones in bone marrow (Ross et al., 1996). Lindstrom et al. (1997) have reported that benzene oxide has an estimated half-life of 7.9 minutes in rat blood, and thus may be able to travel from liver to bone marrow. The quinones are thought to be too reactive to be transported in the bloodstream. Reactivity of muconaldehyde in the presence of GSH suggests that this metabolite would not reach bone marrow; however, metabolites of muconaldehyde, which are less reactive, could be transported to bone marrow and reoxidized. It has been generally accepted that phenolic conjugates are formed in the liver and transported via the blood to the bone marrow, where they are hydrolyzed and oxidized to quinones. Researchers have been unable to reproduce benzene toxicity by giving individual phenolic metabolites to animals, but co-administration of its metabolites, phenol and muconaldehyde with hydroquinone, reproduced the myelotoxic effects of benzene (Witz et al., 1996; Monks et al., 2010). Snyder (2004) concluded that benzene hematopoietic toxicity and leukemogenesis are primarily a function of the bone marrow, a site remote from the liver where substantial benzene metabolism occurs.

Benzene’s metabolic pathways appear to be qualitatively, though not quantitatively, similar in species studied to date (Henderson, 1996). Mice have a greater overall capacity to metabolize benzene than do rats or primates. Mani et al. (1999) reported good correlation between protein–DNA adduct levels in bone marrow and susceptibility to benzene genotoxicity and carcinogenesis in F344 rats and three strains of mice. The B6C3F1 mouse shows the highest adduct levels and is the most sensitive of the animals tested. Powley and Carlson (1999) reported similar findings on measurement of benzene metabolism in mouse, rat, rabbit, and human lung and liver microsomes. A paucity of information is available on the ability of human bone marrow to metabolically activate benzene and/or its metabolites. Rappaport et al. (2009) suggested that humans metabolize benzene to phenol and muconic acid by two enzymes, a high-affinity enzyme which is active primarily below 1 ppm, leading to muconic acid, and a low-affinity enzyme active primarily above 1 ppm, leading to phenol. The authors also suggested that CYP2E1 was active primarily at higher concentrations and CYP2E1 and CYP2A13 could be contributing at lower concentrations.

Factors that alter the metabolism of benzene have the potential to influence the hematopoietic toxicity and carcinogenicity of the VOC. Pretreatment of rats with ethanol, a CYP2E1 inducer, enhanced metabolism of benzene and potentiated its acute and subacute myelotoxicity in rats (Nakajima et al., 1985) and mice (Marrubini et al., 2003). Phenobarbital had a negligible effect. Pretreatment of male B6C3F1 mice with acetone, another CYP2E1 inducer, increased benzene oxidation by about fivefold (Kenyon et al., 1996). Pretreatment with diethyldithiocarbamate, a CYP2E1 inhibitor, completely abolished benzene oxidation. Coexposure of F344 rats to gasoline and benzene resulted in competitive metabolic inhibition (Travis et al., 1992). Metabolism of benzene by hepatic microsomes from 10 human donors was directly proportional to CYP2E1 activity in the samples (Seaton et al., 1994). Thus, it would be anticipated that CYP2E1 polymorphisms may influence susceptibility to benzene toxicity. Male mice have consistently been found to be more sensitive than females to genotoxic effects of benzene. PBTK analysis of data from gas uptake experiments with B6C3F1 mice of both sexes revealed that the optimized maximum rate of benzene metabolism was twice as high in males (Kenyon et al., 1996). It is not known whether there is a sex-dependent difference in benzene metabolism in humans. Sato et al. (1975) saw a more rapid rate of pulmonary elimination of benzene in men than in women following inhalation of 25 ppm for two hours. Benzene was retained longer in the females due to their higher body fat content. Unfortunately, most epidemiology studies of benzene-exposed workers have not provided a gender comparison. Li et al. (1994) did not see a statistically significant difference in cancer mortality between male and female workers exposed to benzene, although risks tended to be somewhat higher for males. Polymorphisms in enzymes involved in benzene metabolism such as CYP2E1, epoxide hydrolase, myeloperoxidase, GSTs, and quinone reductases will also influence the toxicity of benzene.

There are a number of cell populations in the bone marrow that may serve as targets for benzene metabolites. Benzene exposure in vivo results in inhibited growth and development of pluripotential bone marrow stem cells. More mature precursors, such as stromal cells and erythroid and myeloid colony-forming units, are also affected. Trush et al. (1996) point out that hydroquinone-induced inhibition of interleukin (IL)-1 synthesis by stromal macrophages results in altered differentiation of myeloid and lymphoid cells that are normally active in immune responses. These investigators also note that killing of stromal macrophages and fibroblasts could result in such a pronounced reduction of cytokines and growth factors that immature and committed hematopoietic progenitors would die from apoptosis. The erythroid series is more susceptible than the myeloid series to benzene-induced cytotoxicity. Immature myeloid cells can proliferate when the development of erythroid cells is restricted and acquire neoplastic characteristics on dedifferentiation. The end result is AML (Golding and Watson, 1999).

Investigations of benzene toxicity/leukemogenesis have uncovered a variety of potential mechanisms (Golding and Watson, 1999; Snyder, 2002; Bird et al., 2005). As mentioned before, experimental evidence indicates that the complementary actions of benzene and several of its metabolites are required for myelotoxicity. It has been recognized for 20 years that a number of benzene metabolites can bind covalently to GSH, proteins, DNA, and RNA. This can result in disruption of the functional hematopoietic microenvironment by inhibition of enzymes, destruction of certain cell populations, and alteration of the growth of other cell types. Covalent binding of hydroquinones to spindle fiber proteins will inhibit cell replication (Smith, 1996). In vitro studies have established that reactive benzene metabolites bind covalently to DNA of several tissues of different species. Through the use of accelerator mass spectrophotometry, Creek et al. (1997) have demonstrated DNA binding in mice exposed to extremely low ¹⁴C-benzene levels. The binding was dose-dependent over a wide range of doses. DNA adduct levels were so low, however, that this mechanism alone may be insufficient to fully account for leukemogenic effects. The role of the aryl hydrocarbon receptor (AhR) in benzene-induced hematoxicity is emerging. AhR-knockout mice were completely resistant to hematotoxicity of benzene, suggesting a central role of AhR in benzene toxicity (Yoon et al., 2002), but the mechanisms for this receptor are still under investigation. A recent review of literature by Gasiewicz et al. (2010) indicated that AhR regulates the expression of enzymes involved in benzene toxicity and plays a major role in the regulation of hematopoietic stem cells. In order to identify critical data gaps and prioritize research, Meek and Klaunig (2010) proposed five key events in the mode of action for benzene-induced leukemia: (1) benzene metabolism via CYP; (2) interaction of benzene metabolites with target cells in the bone marrow; (3) formation of initiated, mutated target cells; (4) selective proliferation of the mutated cells; and (5) production of leukemia.

It appears likely that oxidative stress contributes to benzene toxicity. As the bone marrow is rich in peroxidase activity, phenolic metabolites of benzene can be activated there to reactive quinone derivatives. Ross et al. (1996) discovered that myeloperoxidase present in murine and human progenitor cells could bioactivate hydroquinone to p-benzoquinone. Electron spin resonance experiments by Hiraku and Kawanishi (1996) revealed the formation of a semiquinone radical in p-benzoquinone-treated HL-60 cells (a human myeloid cell line). This suggested that reactive oxygen moieties (eg, O₂⁻ and H₂O₂) are produced via the formation of the semiquinone radicals. These active oxygen species can cause DNA strand breaks or fragmentation, leading to cell mutation or apoptosis, respectively. Ross et al. (1996), however, pointed out that quinones can also inhibit proteases involved in induction of apoptosis. These authors noted that modulation of apoptosis may lead to aberrant hematopoiesis and neoplastic progression. Badham et al. (2010) recently demonstrated that in utero exposure to benzene in mice increased oxidative stress in fetal tissue. The increase in oxidative stress was also associated with changes in redox-sensitive signaling pathways involved in normal hematopoiesis.

A number of biomarkers of exposure to benzene have been developed and carefully evaluated. Concentrations of the parent compound in exhaled breath parallel blood concentrations. Clayton et al. (1999) reported a high correlation between the extent of smoking-related activities and levels of benzene in exhaled breath of humans. Urinary excretion of a variety of benzene metabolites (ie, phenol, catechol, hydroquinone, 1,2,4-trihydroxybenzene, S-phenylmercapturic acid, and t,t-muconic acid) has been shown to be correlated with benzene exposure in occupational settings. Phenol, catechol, and hydroquinone, however, are neither sensitive nor specific biomarkers, because relatively high levels are found in nonexposed individuals (Medeiros et al., 1997). Similarly, t,t-muconic acid is not specific, because it is a metabolite of sorbic acid, a common food additive. Boogaard and van Sittert (1996) conclude that S-phenylmercapturic acid, an end product of the conjugation of benzene oxide and GSH, is a suitable urinary biomarker for low-level benzene exposure because of its specificity and relatively long half-life. Adducts to hemoglobin and cysteine groups of proteins have been demonstrated in rodents, but not in humans. DNA damage has been detected in benzene-exposed workers (Liu et al., 1996; Andreoli et al., 1997), although such measures have not yet found widespread acceptance as biomarkers of effect.

A number of PBTK models have been developed to aid in predicting risks of myelotoxicity and leukemia posed to humans by benzene. Medinsky et al. (1989) published one of the first such models for benzene. They assumed that benzene metabolism followed Michaelis–Menten kinetics and occurred via benzene oxide. B6C3F1 mice, which are more sensitive than F344 rats to benzene myelotoxicity, were predicted to metabolize two to three times more inhaled benzene. This and other PBTK models do not accurately simulate some laboratory data sets. Modeling benzene metabolism and disposition is difficult because of its inherent complexity and variability. In light of intricate dose–response relationships, Medinsky et al. (1996) emphasized the importance of considering competitive metabolic interactions between benzene and its metabolites, as well as the balance between enzymatic activation and inactivation processes. Cole et al. (2001) employed in vitro metabolic parameters for several pathways to construct a PBTK model that simulated most tissue dosimetry data sets for benzene-exposed mice quite well. Yokley et al. (2006) described a PBTK model that simulated tissue doses of benzene, benzene oxide, phenol, and hydroquinone in humans exposed orally and by inhalation. A MCMC statistical technique was used to assess dissimilarities in population distributions of key model parameters. Variability in metabolic parameters and certain physiological parameters (eg, organ weight) had to be inputed to accurately predict the range of human values.

Benzene is concluded to be a known human carcinogen by regulatory agencies around the world. The IARC classifies benzene as a Group 1 carcinogen (carcinogenic to humans). Benzene is characterized as a known human carcinogen for all routes of exposures based on strong evidence in human and laboratory animals (EPA, 2007). Maximum contaminant levels (MCLs) of 0.005 and 0.01 mg/L have been established for benzene in drinking water (EPA, 2002; WHO, 2004). For the protection of workers, OSHA (2005) has established a TLV of 0.5 ppm as an eight-hour time-weighted average (TWA). ACGIH has also established a BEI for benzene of 25 μg of S-phenylmercapturic acid and 500 μg of t,t-muconic acid/g of creatinine for a urine sample collected at the end of a workshift. Readers are encouraged to review the recent NTP’s 12th Report on Carcinogens and ATSDR’s Toxicological Profile for Benzene (2006c) for other regulatory levels for benzene exposure.

Toluene

Toluene is present in paints, lacquers, thinners, cleaning agents, glues, and many other products. It is also used in the production of other chemicals. Gasoline, which contains 5% to 7% toluene by weight, is the largest source of atmospheric emissions and exposure of the general populace (ATSDR, 2000b). Inhalation is the primary route of exposure, although skin contact occurs frequently. Toluene is a favorite of solvent abusers, who intentionally inhale high concentrations to achieve a euphoric effect (Filley et al., 2004). Large amounts of toluene enter the environment each year by volatilization. Relatively small amounts are released in industrial wastewater. Toluene is frequently found in water, soil, and air at hazardous waste sites (Pohl et al., 2008).

Toluene TK has been thoroughly characterized in humans and laboratory animals. Toluene is well absorbed from the lungs and GI tract. It rapidly accumulates in and can affect the brain, due to that organ’s high rate of blood perfusion and relatively high lipid content. Toluene subsequently is deposited in other tissues according to their lipid content, with adipose tissue attaining the highest levels.

Toluene is well metabolized, but a portion is exhaled unchanged. CYPs catalyze metabolism of toluene primarily to benzyl alcohol and lesser amounts of cresols. Benzyl alcohol is converted by ADH and ALDH to benzoic acid, which is primarily conjugated with glycine and eliminated in the urine as hippuric acid. Nakajima et al. (1992a) reported that CYP2E1 and CYP2C11 are primarily responsible for catalyzing the initial hydroxylation step in rat liver at low and high toluene levels, respectively. CYP2E1 is most active at low doses in humans (Nakajima et al., 1997). Benignus et al. (1998) attempted to use a PBTK model to relate toluene blood concentrations to behavioral effects in rats and humans. It appeared that humans were more sensitive to increases in blood levels, but more rat data and much more human data are needed for model validation.

The CNS is the primary target organ of toluene and other alkylbenzenes. Cardiac, renal, and hepatic toxicities as well as fetal alcohol-like syndrome have occasionally been reported. Increased incidence of spontaneous abortion was linked to toluene exposure at workplaces, but has not been supported by the results of animal testing. Manifestations of acute exposure range from slight dizziness and headache to unconsciousness, respiratory depression, and death. Occupational inhalation exposure guidelines are established to prevent significant decrements in psychomotor functions. Acute encephalopathic effects are rapidly reversible on cessation of exposure (Fig. 24-3), and are not associated with neuroimaging changes (Filley et al., 2004). Subtle neurologic effects have been described in some groups of occupationally exposed individuals. Exposure to ~100 ppm toluene for years may result in subclinical effects, as evidenced by altered brainstem auditory-evoked potentials (Abbate et al., 1993) and changes in visual-evoked potentials (Vrca et al., 1995). Foo et al. (1990) reported a good correlation between toluene exposure and poor scores on neurobehavioral tests of 30 female rotogravure workers. In contrast, severe neurotoxicity is often diagnosed in persons who have abused toluene for a prolonged period. A relatively specific neurobehavioral profile is manifest including inattention, apathy, memory dysfunction, diminished visuospatial skills, frontal lobe dysfunction, and psychiatric symptoms (Filley et al., 2004). Magnetic resonance imaging reveals ventricular enlargement, cerebral atrophy, and white matter hyperintensity, a characteristic profile termed toluene leukoencephalopathy. Such changes represent severe, chronic myelin toxicity. More advanced neuroimaging techniques may be able to reveal mild, subtle changes associated with early alterations in myelin of workers as well as toluene abusers (Caldemeyer et al., 1996; Filley et al., 2004). Rat pups subjected to high doses of toluene on days six to 19 of gestation exhibited a significant reduction in myelination per forebrain cell when they were 21 days old (Gospe and Zhou, 1998).

There is limited information on mechanisms by which toluene and similar solvents produce acute or residual CNS effects (Balster, 1998). The Meyer–Overton theory of partitioning of the lipophilic parent compounds into membrane lipids has been widely accepted for a century. It has been proposed that the presence of solvent molecules in cholesterol-filled interstices between phospholipids and sphingolipids changes membrane fluidity, thereby altering intercellular communication and normal ion movements (Engelke et al., 1996). Such a process is reversible. An alternate hypothesis is that toluene partitions into hydrophobic regions of proteins and interacts with them, thereby altering membrane-bound enzyme activity and/or receptor specificity in a reversible manner (Balster, 1998). Other evidence suggests that toluene and other VOCs may act acutely by enhancing GABA_A receptor function (Mihic et al., 1994), attenuating NMDA receptor-stimulated calcium flux (Cruz et al., 1998), attenuating nicotinic acetylcholine receptors (Bale et al., 2002), enhancing glycine receptors (Bale et al., 2005), and/or activating dopaminergic systems (von Euler, 1994). Using an in vivo microdialysis study and a schedule-controlled operant behavioral task, it was shown that glutamate levels were decreased in mice after acute toluene exposure (Win-Shwe et al., 2009). In addition to neurotransmitter systems, immune and hypothalamic neurohormonal systems also are proposed to play a role in toluene-induced neurotoxicity (Win-Shwe and Fujimaki, 2010). Hester et al. (2011) demonstrated altered expression of genes associated with synaptic transmission and plasticity following acute toluene exposure. The mechanism of chronic toluene neurotoxicity is unknown. Toluene is known to be deposited in brain areas with the highest myelin content. Astrocytosis frequently accompanies myelin disruption (Gotohda et al., 2000; Filley et al., 2004).

As toluene is metabolized by CYPs, ADH, and ALDH, the chemical can interact with other xenobiotics metabolized by these enzymes. Concurrent exposure to solvents metabolized by the same CYP isoforms can result in competitive metabolic inhibition. Inoue et al. (1988) observed that benzene and toluene suppressed one another’s metabolism in humans. Thus, the risk of leukopenia in workers exposed to benzene and toluene should be less than that in workers exposed to benzene alone. Pryor and Rebert (1992) found that toluene greatly reduced manifestations of peripheral neuropathy caused by n-hexane in rats. Although no interaction between toluene and xylenes was seen in humans inhaling low levels of each, simultaneous exposure to higher levels results in mutual metabolic suppression (Tardif et al., 1991). Prior exposure to P450 inducers can result in increased rates of toluene metabolism/elimination and more rapid recovery from toluene-induced CNS depression. Nakajima and Wang (1994) observed that high concentrations of inhaled toluene moderately induced four of the six CYP isoforms that metabolize it, but inhibited the other two in rat liver.

High-level prenatal toluene exposures have produced growth and skeletal retardations in offspring of rodents and humans. The term “fetal solvent syndrome” was used to describe toluene-abusing women’s children who exhibit microcephaly and cranial facial features similar to those with fetal alcohol syndrome (FAS) (Wilkins-Haug, 1997). Exposure to a pattern of binge toluene abuse caused growth retardation, gross morphological anomalies, skeletal abnormalities, and soft tissue anomalies in a dose-related fashion (Bowen et al., 2009). Growth and skeletal retardation were the common findings in the developmental toxicity studies in animals. The reproductive and developmental toxic potential of toluene was evaluated in a two-generation study of S-D rats (Roberts et al., 2003). Both sexes of each generation were exposed to 0, 100, 500, or 2000 ppm toluene vapor six hours per day, seven days per week for 80 days premating and for 15 days of mating. Pregnant animals were also exposed from gestation days one to 20 and lactation days five to 21. No adverse effects were seen at any dose on fertility, reproductive performance, or maternal or pup behavior. Reduced fetal body weight and skeletal anomalies associated with growth retardation were manifest in the offspring of both generations subjected to 2000 ppm, but not to lower vapor levels (Roberts et al., 2003).

Existing studies suggest that toluene is not carcinogenic or mutagenic. IARC classified toluene as not classifiable as a human carcinogen. ACGIH (2012) has established a TLV of 20 ppm as an eight-hour TWA to protect workers from subclinical changes in blue-yellow color vision and the potential for spontaneous abortion. The current ACGIH BEI for toluene is 0.02 mg of toluene/L blood, 0.03 mg toluene/L urine, and 0.3 mg o-cresol/g of creatinine for urine collected at the end of a workshift.

Xylenes and Ethylbenzene

Large numbers of people are exposed to xylenes and ethylbenzene occupationally and environmentally (Wallace et al., 1987). Xylenes (ATSDR, 2006d) and ethylbenzene (ATSDR, 2010a), like benzene and toluene, are major components of gasoline and fuel oil. The primary uses of xylenes and ethylbenzene industrially are as solvents and synthetic intermediates. Most of these aromatics that are released into the environment evaporate into the atmosphere. They may also enter groundwater from oil and gasoline spills, leakage of storage tanks, and migration from waste sites.

The TK and acute toxicity of toluene, xylenes, and other aromatic solvents are quite similar. Xylenes and the others are well absorbed from the lungs and GI tract, distributed to well-perfused and lipophilic tissues such as liver, fat, brain, and skin, exhaled unchanged to some extent, well metabolized by hepatic P450s, and largely excreted as urinary metabolites (Lof and Johanson, 1998). Nielsen and Alarie (1982) state that the potency of benzene and a series of alkylbenzenes, as sensory irritants of the upper respiratory tract of mice, increases with increasing lipophilicity. Acute lethality of hydrocarbons (ie, CNS depressant potency) also varies directly with lipophilicity (Swann et al., 1974). Some of the alkylbenzenes including xylenes and ethylbenzene are shown to be ototoxic in rats. There is limited evidence that chronic occupational exposure to xylenes is associated with residual neurologic effects (ATSDR, 2006d). Limited human data suggest that systemic effects associated with ethylbenzene exposure are respiratory tract and ocular irritation, possible ototoxicity, and hematologic alterations (ATSDR, 2010a). Exposure conditions, including the presence of other solvents, have usually not been characterized in such reports. Uchida et al. (1993), for example, reported increased subjective CNS symptoms in workers exposed for approximately seven years to a mean of 21 ppm xylenes. Neither was the prevalence of symptoms dose-dependent nor were other solvent exposures characterized. This vapor concentration is significantly lower than current occupational exposure standards in the United States, which are established on the basis of acute irritancy or CNS effects. Savolainen et al. (1981) observed improved performance on a series of psychophysiological tests by humans inhaling 200 ppm m-xylene for four hours. An excitatory stage is often initially manifest in subjects, followed by functional inhibition with higher vapor concentrations and/or longer exposure periods. The current PEL in the United States for xylenes and ethylbenzene is 100 ppm.

Xylenes and ethylbenzene appear to have very limited capacity to adversely affect organs other than the CNS. Mild, transient liver and/or kidney toxicity have/has occasionally been reported in humans exposed to high vapor concentrations of xylenes. Little evidence of hepatorenal injury is typically manifest in laboratory animals (ATSDR, 2006c). Xylenes and ethylbenzene increase liver weight and moderately induce liver CYPs in laboratory animals (ATSDR, 2006c, 2010a). The o, p, and m isomers of xylene vary somewhat in their capacity to induce various CYP isoforms in different organs of rats (Backes et al., 1993). Concurrent exposure to an alkylbenzene and another compound metabolized by P450s generally results in their competitive metabolic inhibition. Competitive metabolic interaction between xylene and ethylbenzene following inhalation exposure has been reported (Campbell and Fisher, 2007). Preexposure to an alkylbenzene, conversely, can result in increased metabolism of the other chemical.

The majority of alkylbenzenes do not appear to be genotoxic or carcinogenic. Ethylbenzene and styrene are two exceptions. Kidney injury and an increased incidence of renal adenoma and carcinoma (combined) were found in male F344 rats exposed to 750 ppm ethylbenzene for up to two years (NTP, 1997b). However, no association has been found between the occurrence of tumors and exposure to ethylbenzene in humans. A battery of in vivo and in vitro genetic toxicity assays has shown that ethylbenzene is not genotoxic. IARC classified ethylbenzene as a possible human carcinogen.

Styrene

Styrene is primarily used in the manufacture of polystyrene items and in copolymers with acrylonitrile or 1,3-butadiene to produce synthetic rubber, latex, and reinforced plastics (Gibbs and Mulligan, 1997). Worker exposures in the rubber industry are of greatest concern toxicologically. Styrene is also often detected in the blood of nonoccupationally exposed populations (Churchill et al., 2001; Blount et al., 2006). Sources include tobacco smoke, auto exhaust, and emissions from building materials. Discharges from industry are the major source of environmental pollution (ATSDR, 2010b).

CNS effects and mucous membrane and respiratory irritation are the most commonly reported effects in humans following styrene exposure. Workers exposed to ~25 ppm exhibit signs of mild hepatic injury and cholestasis (Brodkin et al., 2001). Styrene is damaging to the nasal epithelium of rats and mice, and hepatotoxic and pneumotoxic in mice. Species differences in styrene toxicity can be explained by metabolic differences. Styrene is metabolized principally by CYP2E1 in liver and CYP2F2 in lung to styrene-7,8-oxide (SO) (Linhart et al., 2000; Carlson, 2003). SO, the ultimate toxicant, can bind covalently to proteins and nucleic acids. This epoxide is detoxified by the actions of epoxide hydrolase and GST (Carlson, 2010; Chung et al., 2006). SO is detoxified much more efficiently by epoxide hydrolase in rats and humans than in mice. Among the three species, mouse liver and lungs have the highest capacity to form SO from styrene (ATSDR, 2010b).

There is considerable debate as to whether styrene is a human carcinogen. It is carcinogenic to the forestomach of orally dosed mice and rats (Roe, 1994). Results of in vitro and in vivo mutagenicity and genotoxicity studies of styrene have been mixed (ATSDR, 2010b). Chromosomal aberrations, micronuclei, and/or sister chromatid exchange have been reported in employees in some high-exposure occupational settings, but not in others. Increased rates of different cancers have been reported in workers exposed to 1,3-butadiene and styrene in the synthetic rubber industry. The excesses of cancers have largely been attributed to 1,3-butadiene, although styrene may modify the actions of 1,3-butadiene and/or be implicated itself (Matonoski et al., 1997). Occasional findings of small increases in incidences of lymphatic and hematopoietic neoplasms in epidemiological studies of workers exposed primarily to styrene are not very robust (IARC, 2002; ATSDR, 2010b). No increases in any tumors were seen in male or female S-D rats exposed six hours daily, five days weekly for 104 weeks to as high as 1000 ppm styrene vapor (Cruzan et al., 1998). A similar inhalation cancer bioassay, however, revealed increases in pulmonary adenomas and/or carcinomas in male and female CD-1 mice (Cruzan et al., 2001). Gavage studies have yielded variable results (ATSDR, 2010b). In a NCI (1979) study, dose-related increases in the combined incidence of benign and malignant lung tumors were noted in mice following chronic gavage dosing. In an oral study in which pregnant dams were given a single dose of styrene on gestation day 17, and offspring were exposed to high oral doses once weekly for 16 weeks after weaning, a significant increase in lung tumors was observed in these offspring following a 100-week exposure period (Ponomarkov and Tomatis, 1978). The same authors reported no significant tumor incidences in a second study in which a different strain of mice was exposed to much lower doses of styrene on gestation day 17 followed by once a week exposure of offspring for 120 weeks. Lung tumor response observed in mice is most likely due to cytotoxicity caused by the in situ formation of SO, followed by cell proliferation (ATSDR, 2010b). The relevance of mouse lung tumors to humans has been questioned due to species differences in lung metabolism of styrene (Carlson, 2008).

Workers in the reinforced plastics industry are exposed to styrene and SO. Tornero-Velez and Rappaport (2001) used a modified PBTK model to predict the relative contributions of inhaled styrene and SO to systemic levels of SO in humans. Inhaled SO was forecast to present the greater cytogenetic hazard. Sarangapani et al. (2002) created a PBTK model with a multicompartment representation of the respiratory tract that included species-specific quantitative information on physiology, cellular composition, and metabolic capacity. The model-based analysis indicated that humans would be 100-fold less susceptible than mice to pulmonary tumors, based on predictions of SO concentrations in terminal bronchioles. Wenker et al. (2001) attempted to correlate key CYP metabolic capacities of 20 male volunteers with styrene clearance. Lack of correlation was attributed to the low inhaled vapor concentrations and styrene’s blood flow–limited metabolism.

Based on the limited evidence in humans and evidence in mice, NTP (2011) recently classified styrene as “reasonably anticipated to be a human carcinogen,” and IARC classified it as “possibly carcinogenic to humans.” For workers’ protection, OSHA (2006) has established a PEL of 100 ppm, and ACGIH has established a TLV of 20 ppm as an eight-hour TWA (ACGIH, 2012).

Ethanol

Many humans experience greater exposure to ethanol (ethyl alcohol, alcohol) than to any other solvent. Not only is ethyl alcohol used as an additive in gasoline, as a solvent in industry, in many household products, and in pharmaceuticals, but it is also heavily consumed in intoxicating beverages. Frank toxic effects are less important occupationally than injuries resulting from psychomotor impairment. Driving under the influence of alcohol is, of course, the major cause of fatal auto accidents. In many states in the United States, a blood alcohol level of 80 mg/100 mL blood (80 mg%) is prima facie evidence of “driving while intoxicated.” One’s blood alcohol level and the time necessary to achieve it are controlled largely by the rapidity and extent of consumption of the chemical. Ethanol is distributed in body water and to some degree in adipose tissue. Alcohol is eliminated by urinary excretion, exhalation, and metabolism. The blood level in an average adult decreases by ≈ 16 mg% per hour. Thus, a person with a blood alcohol level of 120 mg% would require ≈ eight hours to reach negligible levels.

Ethanol is metabolized to acetaldehyde by three enzymes: the major pathway involves ADH-catalyzed oxidation to acetaldehyde. ADHs are present in various tissues including stomach, lung, eye, and liver, but the highest concentrations are found in the cytoplasm of hepatocytes (Yin et al., 1999). While hepatic ADH largely contributes to ethanol oxidation in humans (Lee et al., 2006), a PBTK model indicates both hepatic and gastric ADH contribute to first-pass metabolism of ethanol in rats, with gastric ADH playing a greater role at lower ethanol doses (Pastino and Conolly, 2000).

The oxidation ADH catalyzes is reversible, but the acetaldehyde that is formed is rapidly oxidized by aldehyde dehydrogenase (ALDH) to acetate. Liver mitochondrial ALDH is the major enzyme responsible for acetaldehyde clearance. However, a recent PBTK model predicted the reversible conversion of acetaldehyde to alcohol by ADH to be an essential step in the systemic clearance of acetaldehyde (Umulis et al., 2005). A second enzyme, catalase, utilizes H₂O₂ supplied by the actions of NADPH oxidase and xanthine oxidase. There is usually little H₂O₂ available in hepatocytes to support the reaction, so it is unlikely that catalase will normally account for more than 10% of ethyl alcohol metabolism. The third enzyme, CYP2E1, is the principal isoform of the hepatic microsomal ethanol oxidizing system.

Ethanol interacts with other solvents that are also metabolized by ADH and CYP2E1. It can be an effective antidote for poisoning by methanol, ethylene glycol (EG), and diethylene glycol (DEG). As ethyl alcohol has a relatively high affinity for ADH, it competitively inhibits the metabolic activation of other alcohols and glycols. Preexposure to a single high dose or multiple doses of ethanol can induce CYP2E1, thereby enhancing the metabolic activation and potentiating the toxicity of a considerable number of other solvents and drugs such as acetaminophen (Lieber, 1997; Klotz and Ammon, 1998). Manno et al. (1996), for example, describe heavy drinkers who developed severe hepatorenal toxicity from CCl₄ exposures, which caused no ill effects in nondrinkers. Other alcohols, such as 2-butanol (Fig. 24-4), can have analogous effects (Traiger and Bruckner, 1976). Chronic heavy drinkers may develop more severe hepatotoxicity associated with acetaminophen overdosing than nonalcoholics, due to increased formation and reduced detoxification of toxic metabolites (Riordan and Williams, 2002; Rumack, 2004). Such interactions are also described in the sections “Metabolism” and “Exogenous Factors.”

A variety of factors that modulate ADH, CYP2E1, and ALDH can influence adverse effects experienced by drinkers. Relatively low ADH and CYP2E1 activities increase systemic alcohol levels and prolong its effects, while low ALDH results in elevated acetaldehyde levels. Long-term ingestion of high alcohol doses can lead to alcohol dependency and cirrhosis, whereas high acetaldehyde levels may result in acute toxicity due to covalent binding to proteins and other macromolecules (Niemela, 1999). Ethanol-metabolizing enzymes exist in multiple molecular forms that are genetically controlled. Genetic polymorphisms of these enzymes contribute to different disease outcomes in different ethnic and racial populations (Crabb et al., 2004; Russo et al., 2004). Asians have higher frequencies of an ADH1*2 variant, which encodes for a rapidly metabolizing form of ADH, preventing alcoholism. White and black populations in the United States predominantly have an ADH1*1 allele (Russo et al., 2004), which encodes for a relatively inactive ADH (Lee et al., 2006), leading to increased risk of alcohol dependency. As for ALDH, approximately 50% of Asians have high levels of the inactive allele, ALDH2*2. Hence, this population experiences acetaldehyde-induced flushing, tachycardia, nausea, vomiting, and hyperventilation on alcohol consumption. Whereas this syndrome offers protection against developing alcoholism, it increases the risk of acetaldehyde-related cancers, including esophageal, stomach, colon, lung, head, and neck tumors (Vasiliou et al., 2004). The proposed mechanism for these cancers is that constant interaction of acetaldehyde with DNA leads to mutations in procarcinogenic or antiapoptotic genes (Israel et al., 2011). While higher frequencies of ADH1*2 and ALDH*2 result in increased internal exposure of acetaldehyde, the incidences of esophageal, stomach, colon, lung, head, and neck tumors in individuals who carry ADH1*2 is markedly less. A possible explanation for this difference is that the increase in acetaldehyde levels in ADH1*2 individuals is brief due to further metabolism of acetaldehyde by ALDH, whereas ALDH2*2 individuals have sustained exposure to acetaldehyde. Hispanics, on the other hand, have very low levels of ADH1*2 and ALDH2*2, predisposing them to alcoholism and their offspring to fetal alcohol syndrome (FAS). Disulfiram, an ALDH inhibitor, is used to treat alcoholism, by enhancing acetaldehyde levels when alcohol is consumed. Whereas CYP2E1 polymorphism is reported in humans (Zintzaras et al., 2006), its role in the pathogenicity of alcoholism is still unclear.

Gender differences in responses to ethanol are well recognized. Women are more sensitive to alcohol, and exhibit higher mortality at lower levels of consumption than men (Sato et al., 2001; Brienza and Stein, 2002). They exhibit somewhat higher blood levels than men following ingestion of equivalent doses of ethanol (Pikaar et al., 1988). This phenomenon appears to be due in part to more extensive ADH-catalyzed metabolism of ethanol by the male gastric mucosa (Frezza et al., 1990; Seitz et al., 1993). Kinetic data indicate that hepatic ADH does not play a role in the sex difference in first-pass elimination of ethanol, although Chrostek et al. (2003) did report higher activities of ADH isoforms that efficiently oxidize ethanol in the liver of male subjects. Sex differences in ethanol metabolism in the remainder of the body appear to be small or nonexistent. A second factor contributing to the higher blood ethanol levels and greater CNS effects in women is their smaller volume of distribution for relatively polar solvents such as alcohols. It is well known that women are more susceptible to alcohol-induced hepatitis and cirrhosis (Thurman, 2000). A postulated mechanism for this sex difference is described below.

Alcohol now is recognized as the leading preventable cause of birth defects and developmental disorders (review in Warren et al., 2011). The severity of birth defects resulting from exposure of the developing embryo or fetus to alcohol is determined by multiple factors, including genetic background, timing and level of alcohol exposure, pattern of drinking, and nutritional status. Peak maternal blood alcohol concentration is apparently the most important determinant of the likelihood of alcohol-related developmental effects. Overconsumption during all three trimesters of pregnancy can result in certain manifestations, dependent on the period of gestation during which ingestion occurs. The most serious adverse consequence of prenatal alcohol exposure is FAS, which has an estimated prevalence that ranges from 0.5 to 7.0 cases per 1000 births in the United States. Diagnostic criteria for FAS include: (1) heavy maternal alcohol consumption during gestation; (2) prenatal and postnatal growth retardation; (3) craniofacial malformations including microcephaly; and (4) mental retardation. However, there are other alcohol-related brain and behavioral abnormalities that are not diagnosed as FAS because of the lack of abnormal facial features. This range of deficits now is referred to as fetal alcohol spectrum disorders (FASD) and is estimated to occur in 1% of births.

Despite an intensive research effort for three decades, the mechanisms underlying FAS remain unclear. Exposure of embryonic tissue to ethanol adversely affects many cellular functions critical to development, including protein and DNA synthesis, uptake of critical nutrients such as glucose and amino acids, and changes in several kinase-mediated signal transduction pathways (Shibley and Pennington, 1997). Alcohol may also harm the fetus indirectly as a result of the mother’s malnutrition. Numerous mechanisms have been suggested as contributing to alcohol-induced fetal damage, although none has been established with certainty (Goodlett and Horn, 2001). Some studies have suggested that oxidative stress on fetal tissues is responsible (Henderson et al., 1999), while others have implicated ethanol’s effects on neurotransmitter-gated ion channels, particularly the NMDA receptor (Costa et al., 2000), and its ability to trigger cell death via necrosis and apoptosis (Ikonomidou et al., 2000; Goodlett and Horn, 2001). Others have reported that ethanol produces a long-lasting reduction in synaptic efficacy (Bellinger et al., 1999) and alters the fetal expression of developmentally important genes such as msx2 and insulin-like growth factors (Singh et al., 1996; Rifas et al., 1997). Recent studies suggest that ethanol-induced NMDA receptor antagonism and GABA receptor agonist activity-mediated apoptosis play critical roles in learning impairment (Ikonomidou et al., 2000; Olney et al., 2002; Toso et al., 2006). Acetaldehyde may also contribute to the development of FAS, as it is shown to accumulate in the fetal brain after prenatal alcohol exposure (Hamby-Mason et al., 1997) and to be cytotoxic to cultured embryonic brain cells (Lee et al., 2005). Other potential mechanisms of FASD include: (1) alterations in the regulation of gene expression (eg, reduced retinoic acid signaling, altered DNA methylation), (2) interference with mitogenic and growth factor responses involved in neural stem cell proliferation, migration, and differentiation; (3) disturbances in molecules that mediate cell–cell interactions (L1, NCAM, loss of trophic support); and (4) derangements of glial proliferation, differentiation, and function (Guerri et al., 2009; Ramsay, 2010).

Genetic variability in metabolic enzymes is another proposed mechanism for FAS. A role for CYP2E1 in the induction of FAS has been hypothesized, given its expression in human cephalic tissues during embryogenesis (Boutelet-Bochan et al., 1997). Results of in vitro experiments indicate that human CYP2E1 is effective in production of reactive oxygen intermediates (eg, hydroxy radicals, superoxide anion, and H₂O₂) and in causation of lipid peroxidation (Dai et al., 1993). Membrane lipids and a variety of enzymes are targets for free radical attack. Albano et al. (1999) have demonstrated a marked increase in covalent binding of hydroxyethyl radicals to hepatic microsomal proteins from rats following chronic ethanol administration. There is not yet direct evidence that hydroxyethyl radicals contribute to lipid peroxidation, but they do readily react with α-tocopherol, GSH, and ascorbic acid, thereby potentially lowering liver antioxidant levels in vivo. Albano et al. (1999) also describe evidence that hydroxyethyl radical–protein adducts in hepatocytes induce immune responses, which may contribute to chronic ethanol hepatotoxicity (ie, alcoholic liver disease [ALD]).

ALD is one of the most prevalent and fatal conditions of alcohol abuse. It is the 12th leading cause of death in the United States. The progression of ALD follows a characteristic pattern marked by appearance of fatty liver, hepatocyte necrosis, inflammation, regeneration nodules, fibrosis, and cirrhosis. Despite extensive research, the mechanisms underlying ALD are still unclear. It is postulated that chronic alcohol consumption increases the gut permeability to lipopolysaccharide (LPS), an endotoxin that is a major constituent of the outer membrane of gram-negative bacteria. LPS binds to receptors in Kupffer cells to produce reactive oxygen species (ROS), which activate nuclear regulatory factor kappa beta (NF-κB), leading to TNF-α and other inflammatory mediators that are cytotoxic to hepatocytes and chemoattractants for neutrophils (Bautista, 2000; Diehl, 2000). These mediators include ILs, prostaglandins, free radicals, and eicosanoids. TNF-α signaling also increases mitochondrial ROS, which trigger apoptotic cell death in the liver. Apoptotic cells in turn result in an increased inflammatory response. Proinflammatory cytokines and oxidative stress stimulate collagen synthesis by hepatic stellate cells, leading to alcoholic fibrosis (Lieber, 2004). Chronic alcohol consumption also inhibits components of innate immunity, such as natural killer (NK) cells. The inhibition of NK cells decreases NK cell-mediated killing of stellate cells, contributing to fibrosis (Miller et al., 2011; Mandrekar and Szabo, 2009). Alcohol-induced oxidative stress is thought to cause liver injury by promoting the development of both humoral and cellular immune responses (Vidali et al., 2008). CYP2E1 can generate ROS during its catalytic cycle. Because CYP2E1 metabolizes ethanol and increases during ethanol exposure, it is suggested to play a significant role in ethanol-induced liver toxicity. Lu et al. (2010) recently reported that wild-type mice and humanized CYP2E1 (knock-in) mice had significant ethanol-induced oxidative stress and fatty liver, but CYP2E1 knockout mice did not have these changes.

Acetaldehyde is capable of upregulating collagen synthesis either directly or indirectly by inducing TGF-β1. Hence, polymorphisms of ALDH can increase the risk of fibrosis (Purohit and Brenner, 2006). Kupffer cells of female rats are more sensitive to endotoxin than those of males. Estrogen increases the sensitivity of Kupffer cells of rats to endotoxin, leading to increased production of TNF-α and death of hepatocytes (Thurman, 2000; Yin et al., 2000; Colantoni et al., 2003). This phenomenon is thought to account in part for the more severe hepatitis and cirrhosis commonly seen in female alcoholics (Thurman, 2000).

Alcohol-induced damage of the liver and other tissues is thought to result in part from nutritional disturbances, as well as toxic effects (Lieber, 2004; DiCecco and Francisco-Ziller, 2006). Lack of money, poor judgment, prolonged inebriation, and appetite loss contribute to poor nutrition and weight loss in alcoholics. A high percentage of calories in the alcoholic’s diet is furnished by alcohol. Malabsorption of thiamine, diminished enterohepatic circulation of folate, degradation of pyridoxal phosphate, and disturbances in the metabolism of vitamins A and D can occur (Mezey, 1985). Prostaglandins released from endotoxin-activated Kupffer cells may be responsible for a hypermetabolic state in the liver. With the increase in oxygen demand, the viability of centrilobular hepatocytes would be most compromised, due to their relatively poor oxygen supply (Thurman, 1998). Metabolism of ethanol via ADH and ALDH results in a shift in the redox state of the cell. The metabolites and the more reduced state can result in hyperlactacidemia, hyperlipidemia, hyperuricemia, and hyperglycemia, leading to increased steatosis and collagen synthesis (Lieber, 2004).

Alcoholism can result in damage of extrahepatic tissues. Alcoholic myopathy is one of the more common consequences, occurring in 50% of alcohol abusers. The condition is characterized by reductions in skeletal muscle mass and strength (Adachi et al., 2003; Preedy et al., 2003). Alcoholic cardiomyopathy is a complex process that occurs in 20% to 25% of alcoholics. It results from decreased synthesis of cardiac contractile proteins, attack of oxygen radicals, increases in endoplasmic reticulum Ca²⁺-ATPase, and an antibody response to acetaldehyde–protein adducts (Richardson et al., 1998; Preedy et al., 2003). Interestingly, light to moderate drinking is reported to protect against atherosclerosis in the carotid artery, a major cause of ischemic stroke (Hillbom, 1999). Alcohol increases high-density lipoprotein (HDL) cholesterol by inducing the constituents, apo I and II, by increasing catabolism of very low-density lipoproteins and chylomicrons, and by delaying HDL catabolism (Kolovou et al., 2006). It is also hypothesized that ethanol metabolism in the vascular wall may inhibit oxidation of low-density lipoproteins (LDL), a requisite for atherogenesis. Phenolic antioxidants in wines may also inhibit LDL oxidation as well as reduce platelet aggregation. Conversely, heavy drinking appears to deplete antioxidants and have the opposite effects (Hillbom, 1999; Kolovou et al., 2006). Recent heavy drinking increases the risk of both hemorrhagic and ischemic strokes. Other organ systems can be adversely affected in alcoholics including the brain and GI tract (Kril et al., 1997; Rajendram and Preedy, 2005).

There is concern about the role of ethyl alcohol in carcinogenesis, due to the frequent consumption of alcoholic beverages by millions of people. IARC (1988) concluded that there was “sufficient evidence” for causation of tumors of the oral cavity, pharynx and larynx, esophagus, and liver of humans. In a 2007 meeting, IARC added breast and colorectal cancers to the list of cancers causally linked with alcohol consumption (Baan et al., 2007). Genotoxic effects by acetaldehyde and elevated estrogen levels are likely mechanisms for breast cancer (Boffeta and Hashibe, 2006). According to the World Health Organization Global Burden of Disease Project, drinking accounts for ~3.2% of all deaths and 3.6% of all cancers (Rehm et al., 2004). The original associations between alcohol and cancers came primarily from epidemiological case–control and cohort studies. One such cohort study of 276,000 American men showed an increase in total cancer risk with increasing ethanol consumption (Boffeta and Garfinkel, 1990). In support of epidemiological studies, a rat study showed that ethanol induces tumors in various organs including oral cavity, tongue, and lips (Soffritti et al., 2002). While mechanisms for alcohol-related carcinogenesis are not fully understood, metabolism plays a major role (Seitz and Stickel, 2007). Ethanol is shown to induce CYP2E1, which is known to generate ROS that damage DNA, lipids, and proteins. Additionally, acetaldehyde, which is carcinogenic to humans, reacts with DNA, forming adducts. Evidence suggests that the risk of alcohol-induced cancer in humans is modulated by genetic polymorphisms in enzymes responsible for alcohol metabolism, folate metabolism, and DNA repair (Boffeta and Hashibe, 2006). Ethanol may also stimulate carcinogenesis by inhibiting DNA methylation and interfering with retinoic acid metabolism. Ethanol and smoking act synergistically to cause oral, pharyngeal, and laryngeal cancers. Age, sex, and ethnicity have also been reported to be factors. It is generally accepted that alcohol induces liver cancer by causing cirrhosis or other chronic liver damage and/or by enhancing the bioactivation of other carcinogens (Table 24-3).

An international symposium (Purohit et al., 2005), sponsored by the National Institute on Alcohol Abuse and Alcoholism in 2004, concluded that chronic ethanol consumption may promote carcinogenesis by: (1) production of acetaldehyde, a weak mutagen and carcinogen; (2) induction of CYP2E1 and its associated oxidative stressors and conversion of procarcinogens to carcinogens; (3) depletion of S-adenosylmethionine (SAM) and, consequently, induction of global DNA hypomethylation; (4) increased production of inhibitory guanine nucleotide regulatory proteins and components of extracellular signal-regulated kinase-mitogen-activated protein kinase signaling; (5) accumulation of iron and associated oxidative stress; (6) inactivation of the tumor suppressor gene BRCA1 and increased estrogen responsiveness (primarily in the breast); and (7) impairment of retinoic acid metabolism. Jeannot et al. (2011) recently reported that ethanol may act by promoting spontaneously initiated cells, especially H-ras mutated cells.

Methanol

Methanol (methyl alcohol, wood alcohol, and CH₃OH) is primarily used as a starting material for the synthesis of chemicals such as formaldehyde, acetic acid, methacrylates, ethylene glycol (EG), and methyl tertiary-butyl ether (MTBE). CH₃OH is found in windshield washer fluid, carburetor cleaners, antifreeze, and copy machine toner, and serves as fuel for Sterno^TM heaters, and model airplanes. It also functions as a denaturant for some ethyl and isopropyl alcohols, rendering them unfit for consumption. It is used to a limited extent as an alternative fuel for fleet vehicles (usually in a mixture of 85% CH₃OH and 15% gasoline) and is being explored as a gasoline additive and hydrogen source for fuel cell vehicles. Exposure of the general population also occurs via the consumption of fruits, fruit juices, vegetables, and alcoholic beverages that contain free CH₃OH or CH₃OH precursors. Indirect exposure occurs via the hydrolysis of the artificial sweetener, aspartame, and subsequent CH₃OH absorption from the gut. Very low-level exposures may occur via ambient air and drinking water. Of all chemicals reported, CH₃OH was ranked No. 1 for fugitive air emissions, No. 3 for point source air emissions, and No. 2 for surface water discharges according to the 2004 Toxic Release Inventory (EPA, 2006a). Nevertheless, persistence and bioaccumulation in the environment are not expected. The consumption of adulterated “bootleg” whiskey is a major cause of CH₃OH poisoning.

Serious CH₃OH toxicity is most commonly associated with ingestion. Left untreated, acute CH₃OH poisoning in humans is characterized by an asymptomatic latent period of 12 to 24 hours followed by formic acidemia, ocular toxicity, coma, and in extreme cases death (Lanigan, 2001). Visual disturbances generally develop between 18 and 48 hours after ingestion and range from mild photophobia and misty or blurred vision to markedly reduced visual acuity and complete blindness (Eells et al., 1996). Although there is considerable variability among individuals in susceptibility to CH₃OH, a frequently cited lethal oral dosage is 1 mL/kg. Blindness and death have been reported with dosages as low as 0.1 mL/kg (ATSDR, 1993). CH₃OH’s target within the eye is the retina, specifically the optic disk and optic nerve. Optic disk edema and hyperemia are seen, along with morphological alterations in the optic nerve head and the intraorbital portion of the optic nerve. Both axons and glial cells exhibit altered morphologies (Kavet and Nauss, 1990). Rods and cones, the photoreceptors of the retina, are also altered functionally and structurally (Seme et al., 1999). Evidence is accumulating that Müller cells, neuroglia that function in the maintenance of retinal structure and in intracellular and intercellular transport, are early targets of CH₃OH (Garner et al., 1995a). There are indications of mitochondrial disruption in Müller and photoreceptor cells, which is consistent with the long-held view that formate inhibits the energy-generating mitochondrial enzyme, cytochrome c oxidase, which is critical for the proper functioning of highly oxidative organs such as the retina. This mechanism might explain, at least in part, CH₃OH’s selective toxicity to photoreceptors and other highly metabolically active cells (Eells et al., 1996). Interestingly, exposure of rats to monochromatic red radiation from light-emitting diode arrays can aid in the recovery of rod and cone function in CH₃OH-treated rats and protect the retina from histopathological changes characteristic of formate toxicity (Eells et al., 2003). This “photobiomodulation” is thought to be mediated by the ability of monochromatic red to near-IR light to improve mitochondrial respiratory chain function (ie, increase cytochrome c oxidase activity), thereby initiating a signaling cascade that promotes cellular proliferation and cytoprotection.

Significant species differences exist in the susceptibility of CH₃OH toxicity. Acute ocular toxicity, metabolic acidosis, CNS depression, and death noted in humans are not seen with rodents. Conversely, rodents can be susceptible to delayed adverse effects of methanol such as developmental toxicity. Therefore, elucidation of the mechanism of CH₃OH’s acute toxicity was hampered for years by the lack of appropriate animal models. Largely based on the work of Gilger and Potts (1955), it became apparent that only nonhuman primates respond to CH₃OH similarly to humans. Humans and nonhuman primates metabolize CH₃OH to formaldehyde (HCOH) by ADH, whereas rodents utilize catalase. In all mammalian species, HCOH is very rapidly converted via formaldehyde dehydrogenase (FLDH) to formate, which is further metabolized to CO₂. The conversion of formate to CO₂ occurs via a two-step, tetrahydrofolate (THF)-dependent pathway. First, formate is converted to 10-formyl-THF by formyl-THF synthetase, after which 10-formyl-THF is oxidized to CO₂ by formyl-THF dehydrogenase (F-THF-DH). Because rodents have much higher hepatic THF levels than primates, formate does not accumulate as it does in humans and monkeys (Medinsky and Dorman, 1995; Martinasevic et al., 1996). Another possible explanation is the lower F-THF-DH activity in primate liver (Johlin et al., 1989). It is further speculated that primates are more susceptible to CH₃OH and formate toxicity due to their relative inability to excrete excess formate via the kidneys (Smith and Taylor, 1982). Irregardless, susceptibility to CH₃OH-induced ocular toxicity is dependent on the relative rate of formate clearance. Dietary and chemical depletion of endogenous folate cofactors in rats has been shown to increase formate accumulation following CH₃OH, resulting in the development of metabolic acidosis and ocular toxicity similar to that observed in humans (Eells et al., 1996, 2000). Sweeting et al. (2011) reported that rabbits, similar to humans and monkeys, metabolize CH₃OH by ADH and exhibit slower CH₃OH clearance and increased formate accumulation, suggesting that rabbits resemble primates and humans more closely than rodents. A simplified scheme of CH₃OH metabolism is presented in Fig. 24-10.

For years there was considerable debate whether HCOH or formate was responsible for CH₃OH’s ocular toxicity. The finding that HCOH does not accumulate following CH₃OH treatment, even in folate-deficient monkeys that are profoundly sensitive to CH₃OH, argues against a role for HCOH. Also, species of differing susceptibility exhibit comparable blood HCOH half-lives. In contrast, formate has been shown to accumulate in the human and monkey following CH₃OH treatment. CH₃OH-derived formate has also been quantified in the vitreous humor, retina, and to a lesser extent the optic nerve (Eells et al., 1996). Moreover, CH₃OH-induced retinal dysfunction, as indicated by diminution of the amplitude of electroretinogram (ERG) a and b waves, is negatively correlated in a linear fashion with blood formate concentrations (Eells et al., 1996). Others have reported similar relationships between blood formate and ERG responses indicative of photoreceptor dysfunction (Seme et al., 1999). Formate has also been shown to induce ocular toxicity in monkeys and formate oxidation-inhibited rats in the absence of metabolic acidosis or a reduction in blood pH (Martin-Amat et al., 1978; Eells et al., 1996). Direct incubation of formate with cultured ocular cells caused ATP depletion and cytotoxicity (Treichel et al., 2003). Thus, formate appears to act as a direct ocular toxin and not indirectly through the induction of an acidotic state, although acidosis may potentiate formate toxicity (since the inhibition of cytochrome oxidase increases as pH decreases and acidosis allows for greater diffusion of formic acid into cells).

The question has been raised whether ocular toxicity is simply a function of circulating formate reaching the visual tract, or whether metabolism in retinal or optic nerve tissues generates toxic metabolites locally. This is a legitimate question considering that metabolism of CH₃OH to HCOH via peroxisomal enzymes (catalase) has been demonstrated in rat retina in vitro (Garner et al., 1995a), and the presence of cytoplasmic formaldehyde dehydrogenase (FLDH) activity has been demonstrated in several regions of the rat and mouse eye, including the retina (Messiha and Price, 1983; McCaffery et al., 1991). By use of a folate-deficient rat model, Garner et al. (1995b) showed that a level of blood formate generated by i.v. infusion of pH-buffered formate did not diminish the ERG b-wave amplitude generated by Müller cells of the retina, as did a comparable blood level of formate derived from CH₃OH. This suggests that the intraretinal metabolism of CH₃OH is necessary for the initiation of retinal toxicity by formate. Not only are the enzymes necessary to produce formate present in the retina, but so too are folate and F-THF-DH. The latter was found to be localized in the mitochondria of Müller cells, prompting the suggestion that F-THF-DH may serve a dual role, one protective of the Müller cell and the other toxic. Protection would come in the form of formate oxidation, and toxicity from the over-consumption or depletion of ATP required for formate metabolism via the folate pathway (Martinasevic et al., 1996).

The aforementioned findings raise concerns about the safety of CH₃OH exposure. Chamber studies of human volunteers exposed to 200 ppm CH₃OH for four or six hours showed no blood formate accumulation above background (Lee et al., 1992; d’Alessandro et al., 1994). Whereas this might be considered evidence that exposure at the 2012 ACGIH TLV and OSHA PEL of 200 ppm poses no risk of ocular toxicity, such an interpretation may not be valid, given that ocular toxicity may be a function of intraretinal CH₃OH metabolism rather than circulating formate levels.

The effects of acute, high CH₃OH exposures are well characterized compared with those of chronic, low-level exposures. Few reproductive/developmental studies and even fewer cancer bioassays have been conducted, due to the fact that rodents are not an ideal animal model for CH₃OH. Nonetheless, Soffritti et al. (2002) conducted a bioassay in which S-D rats received CH₃OH in their drinking water (0, 500, 5000, or 20,000 ppm ad libitum) for 104 weeks. The investigators concluded that CH₃OH was a multipotent carcinogen. However, they referenced two Japanese bioassays that found no evidence of carcinogenicity in B6C3F1 mice or F344 rats exposed 20 hours per day to 10, 100, or 1000 ppm CH₃OH by inhalation for 18 and 24 months, respectively (NEDO, 1987; Katoh, 1989). A recent review of carcinogenicity and genetic toxicity studies concluded that CH₃OH is not likely to be carcinogenic in humans (Cruzan, 2009).

The NTP Center for the Evaluation of Risks to Human Reproduction released a monograph on CH₃OH’s reproductive/developmental toxicity, largely based on an expert panel report (NTP, 2003, 2004b). Among the conclusions was that CH₃OH is a potential developmental toxicant in humans, provided a high enough blood concentration of the parent compound (assumed to be the proximate teratogen) is achieved. As this conclusion was largely based on data in rodents that metabolize CH₃OH much differently than humans, its validity has been called into question (Clary, 2003). As such, the few reproductive/developmental studies in nonhuman primates are particularly informative. These studies reported that maternal exposure of monkeys to inhaled CH₃OH (200, 600, or 1800 ppm, 2.5 hours per day, seven days per week prior to breeding and throughout pregnancy) was not associated with maternal toxicity, reproductive loss, or congenital malformations in offspring, but was associated with a six- to eight-day reduction in the mean length of pregnancy (Burbacher et al., 2004a,b). Studies in rodents and rodent whole embryo cultures have implicated dysregulated cell death in the pathogenesis of CH₃OH-induced malformations, examined GSH status in the developing conceptus as it relates to vulnerability, and explored the basis for the increased sensitivity of the mouse embryo compared with that of the rat (Harris et al., 2003, 2004; Degitz et al., 2004; Hansen et al., 2005). Because the disposition of CH₃OH and formic acid in rabbits was similar to humans, Sweeting et al. (2010) considered rabbits as an appropriate model for assessing the risk of human developmental toxicity. Three strains of mice and New Zealand rabbits were given a single dose of CH₃OH during gestation. Two strains of mice were susceptible to CH₃OH teratogenicity; one strain of mice and the rabbits were resistant. Despite similar CH₃OH and formic acid disposition in three strains of mice, the finding that one strain did not exhibit teratogenic effects suggests mechanisms other than pharmacokinetics could be involved. A role for ROS is suspected. The study also raised concerns about using rodents as a surrogate to predict human developmental toxicity of CH₃OH.

Whereas there is still much to be learned about the mechanisms of CH₃OH toxicity, what is known allows for effective therapies, if they are applied in a timely manner. The American Academy of Clinical Toxicology has published practice guidelines for the treatment of CH₃OH poisoning (Barceloux et al., 2002). In cases of severe CH₃OH poisoning, there is a direct correlation between the formic acid concentration and increased morbidity and mortality. Sodium bicarbonate is usually given i.v. to correct severe acidosis, and case reports suggest that it may enhance renal formate excretion. Metabolic blockade is usually achieved with ethanol or 4-methylpyrazole (fomepizole), both acting as effective competitive inhibitors of ADH. Folate or folinic acid (activated folate) therapy is also indicated to increase the efficiency of formate oxidation. Hemodialysis is generally indicated when acidemia, high CH₃OH concentrations, or visual symptoms are present, although there are conflicting data on whether it appreciably shortens the elimination half-life of formate (Kerns et al., 2002; Hantson et al., 2005). Lastly, the treatment threshold of 20 mg CH₃OH/dL of blood in a nonacidotic patient arriving early for care has been questioned as being overly conservative. This stems from a comprehensive review of worldwide CH₃OH poisonings that identified 126 mg/dL as the lowest early blood CH₃OH level ever clearly associated with acidosis (Kostic and Dart, 2003). CH₃OH exemplifies the benefits of knowing a chemical’s mode of action when treating the poisoned patient. This knowledge also aids in identifying potentially sensitive subpopulations, such as those suffering from dietary folate deficiency. Research suggests, however, that even in a state of folate deficiency, the body probably contains sufficient folate stores to effectively detoxify small doses of CH₃OH-derived formate from exogenous sources (Medinsky et al., 1997).

Ethylene Glycol

EG (1,2-dihydroxyethane) is a constituent of antifreeze, deicers, hydraulic fluids, drying agents, and inks, and is used to make plastics and polyester fibers. Workers may be exposed dermally or by inhalation when solutions containing EG are heated or sprayed. The most important exposure route is ingestion, as EG may be accidentally swallowed, taken deliberately in suicide attempts, or used as a cheap substitute for ethanol. “Antifreeze” poisoning occurs frequently in cats and dogs that find its taste appealing. In 2004, there were 5562 human cases of EG exposure reported by the American Association of Poison Control Centers (23 fatal), nearly 40% of which were treated in a health care facility (Watson et al., 2005). EG enters the environment as a result of disposal of industrial and consumer products containing the chemical. It partitions into surface water and groundwater, but does not persist in any environmental medium and is practically nontoxic to aquatic organisms (Staples et al., 2001).

The TK profile of EG is well characterized. All laboratory mammals and humans metabolize EG similarly. A series of papers has been published describing the TK of EG in S-D rats and CD-1 mice after administration of a single dose by the i.v., oral, or percutaneous routes (Frantz et al., 1996a-c). Absorption from the GI tract of rodents and humans is very rapid and virtually complete, whereas cutaneous and pulmonary absorption is relatively slow and less extensive. Once absorbed, EG is distributed throughout the total body water. As illustrated in Fig. 24-11, EG is metabolized by NAD⁺-dependent ADH to glycolaldehyde and on to glycolic acid (GA). GA is oxidized to glyoxylic acid by GA oxidase and lactic dehydrogenase. Glyoxylic acid may be converted to formate and CO₂, or oxidized by glyoxylic acid oxidase to oxalic acid (OA) (Wiener and Richardson, 1988; Frantz et al., 1996b). The rate-limiting step in the metabolism of EG is the conversion of GA to glyoxylic acid. EG has a half-life in humans of three to 8.6 hours (Leth and Gregersen, 2005). Under conditions of repetitive, low-dose exposure, EG is not expected to bioaccumulate, given its rapid metabolism and elimination. Pregnancy status has no bearing on the TK of EG, GA, or OA, as demonstrated by comparisons of pregnant and nonpregnant S-D rats (Pottenger et al., 2001).

The minimum acute lethal dose of EG in humans is estimated at ~1.4 mL/kg, which equates to 100 mL for a 70-kg adult (LaKind et al., 1999; Hess et al., 2004). Acute poisoning entails three clinical stages after an asymptomatic period, during which EG is metabolized: (1) a period of inebriation, the duration and degree depending on dose; (2) the cardiopulmonary stage 12 to 24 hours after exposure, characterized by tachycardia and tachypnea, which may progress to cardiac failure and pulmonary edema; and (3) the renal toxicity stage 24 to 72 hours postexposure. Metabolic acidosis, due largely to GA accumulation, can develop and become progressively more severe during stages 2 and 3 (Jacobsen et al., 1984; Moreau et al., 1998; Egbert and Abraham, 1999). Hypocalcemia can result from Ca²⁺ chelation by OA to form Ca²⁺ oxalate monohydrate (COM) crystals. Deposition of these crystals in kidney tubules is associated with organ damage and potentially with acute renal failure. Nephrotoxicity appears to be an acute, high-dose phenomenon, as no demonstrable kidney damage has been reported in occupational studies of groups such as airport deicing workers or Finnish auto mechanics (Laitinen et al., 1995; Gerin et al., 1997). COM crystal deposition has been reported in the walls of CNS blood vessels, with associated inflammation, edema, and sometimes neuropathy (Froberg et al., 2006).

Based on high-dose rodent studies, including two lifetime dietary studies, EG has very limited chronic toxicity potential, exhibits no evidence of carcinogenicity, and does not appear to be a reproductive toxicant (DePass et al., 1986; NTP, 1993b). However, it can cause adverse developmental effects such as skeletal and soft tissue malformations and delayed ossification when administered during gestation at high doses (≥500 mg/kg per day) (ATSDR, 1997b). GA appears to be the proximate developmental toxicant, with metabolic acidosis playing an exacerbating, but not obligatory role (Carney et al., 1996, 1999). Pottenger et al. (2001) have demonstrated that at teratogenic doses of EG, blood levels of GA are disproportionately high relative to the EG dose, further supporting GA as the proximate teratogen and implying a role for metabolic saturation. Corley et al. (2005d) have detailed the key events in EG-induced developmental toxicity and suggest the disruption of hox gene expression, fluid imbalance, and dysregulation of cell death in the embryo as mechanistic possibilities. The mechanism of GA’s teratogenic action remains largely unknown, however, as does the relevance of defects in rodents given the absence of reported developmental effects in humans. The NTP Center for the Evaluation of Risks to Human Reproduction has released a monograph on EG’s reproductive/developmental toxicity largely based on an expert panel report (NTP, 2004d,f). The overall conclusion was that EG exposures below the level of metabolic saturation should not result in developmental toxicity in humans. Further, because environmental and occupational exposures to humans are two to three orders of magnitude lower than those expected to result in metabolic saturation, such exposures are of negligible concern. Corley et al. (2005a) have developed a PBTK model of EG and GA and compared internal dose surrogates in rats and humans. Based on the comparisons, they also concluded that occupationally or environmentally exposed humans are unlikely to achieve blood levels of GA that have been associated with developmental toxicity in rats.

The chief concern about EG is renal toxicity after high-dose, acute exposure. Although EG must be metabolized to toxic intermediates to induce kidney damage, the specific metabolite(s) and underlying mechanism(s) responsible remain to be fully elucidated. There is agreement, however, that the process involves a toxicant-induced proximal tubular necrosis leading to loss of renal function. Poldelski et al. (2001) exposed isolated mouse proximal tubule segments and human proximal tubular epithelial cells (HK-2) to GA, glycoaldehyde, glyoxylate, or OA for 15 to 60 minutes, on which basis they concluded glycoaldehyde and glyoxylate are the principal metabolites responsible for EG nephrotoxicity. This is contrary to the prevailing opinion that renal toxicity is due to the terminal metabolite, OA, which precipitates in the kidney in the form of COM crystals (Guo and McMartin, 2005). Corley et al. (2005d) have listed the key mechanistic events in the renal toxicity of EG: (1) metabolism of EG to OA via GA; (2) concentration of OA in tubular urine → precipitation of OA with Ca²⁺ → buildup of COM crystals in renal tubular epithelium → adherence of COM crystals to the plasma membrane of proximal tubular cells → subsequent intracellular uptake of COM crystals by endocytosis; and (3) physical trauma by COM crystals and/or production of free radicals and lipid peroxidation leading to cell necrosis, apoptosis, and renal tubular degeneration. Clearly, the weight of evidence implicates OA is critical in the induction of renal damage, but a possible role for less frequently observed hippuric acid crystals and direct cytotoxicity by other metabolites cannot be ruled out. In vitro assays have indicated that both oxalate and COM crystals are injurious to renal epithelial cells (Thamilselvan and Khan, 1998). Also, renal damage has been observed occasionally following exposure to EG, without documentation of crystals in the kidney.

The intracellular target(s) of COM crystals is(are) not known with certainty, but it is well established that mitochondrial damage is a major mechanism for chemically mediated renal tubular necrosis. As such, COM crystals in the cytoplasm could directly affect mitochondria or be metabolized to release the oxalate ion that could do so. McMartin and Wallace (2005) hypothesized as much and demonstrated that COM produced a dose-dependent decrease in state 3 respiration in isolated rat kidney mitochondria, which they believe may be responsible for induction of mitochondrial permeability transition (MPT). MPT is characterized by an otherwise inpenetrant inner mitochondrial membrane undergoing transformation, whereby it becomes penetrable to solutes or large molecules. As a result of MPT, mitochondria undergo a rapid and progressive osmotic swelling, depolarization of mitochondrial membrane potential, and inhibition of oxidative phosphorylation/ATP synthesis, all of which can lead to either oncotic/ischemic cell death or apoptotic cell death. Additional insight into the pathogenesis of EG-induced renal damage has been provided by Chen et al. (2004), who provided male rats with 0.75% EG in their drinking water for two, four, and eight weeks, after which the kidneys were processed for RNA isolation and microarray analysis using a rat-based chip. Changes in the expression of genes associated with tubular structure and function, oxidative damage, and inflammation were common. Increased expression of mitochondrial uncoupling protein was also observed, providing additional evidence of a mitochondrial target. EG has also been used as a model crystal-forming renal toxicant to determine the basis for COM crystal retention in the kidney. Studies with EG suggest that crystal retention in the kidney may depend on the expression of the proteins hyaluronan, osteopontin, and their mutual cell surface receptor CD44, all of which are upregulated in response to renal injury/inflammation, and under normal circumstances play a role in reestablishment of epithelial barrier integrity and restoration of renal function. However, their upregulation could turn a noncrystal-binding epithelium into a crystal-binding one, thereby setting the stage for crystal retention and renal injury (Asselman et al., 2003; Verhulst et al., 2003).

In cases of severe EG poisoning, early diagnosis and aggressive therapeutic intervention are essential for a favorable clinical outcome. A plasma EG level of 20 mg/dL is considered the threshold for kidney toxicity, if therapeutic strategy is based on EG concentration alone (Hess et al., 2004). However, no signs of renal injury have developed in patients at initial plasma GA concentrations of up to 10.1 mM or 76.7 mg/dL (Brent et al., 1999; Hess et al., 2004). In any case, treatment of EG poisoning involves three primary goals: (1) correction of the patient’s metabolic acidosis; (2) inhibition of EG metabolism to its toxic metabolites; and (3) removal of EG and its toxic metabolites by hemodialysis, if necessary. As Guo and McMartin (2005) demonstrated that the renal cytotoxicity of COM crystals is potentiated by severe acidosis, its correction with bicarbonate should help ameliorate the development of renal toxicity. As with CH₃OH, ethanol and 4-methylpyrazole (fomepizole) are frequently given as antidotes for EG poisoning and can prevent renal injury if given early in the course of intoxication. These ADH inhibitors block EG’s metabolic bioactivation, thus minimizing the formation of toxic metabolites and allowing EG to be eliminated unchanged by the kidneys. There is increasing evidence that i.v. fomepizole may be more efficacious than ethanol as an antidote and obviate the need for hemodialysis in most EG-poisoned patients with normal renal function (Moreau et al., 1998; Brent et al., 1999; Sivilotti et al., 2000; Scalley et al., 2002). Watson (2000) has even stated that the results of fomepizole therapy support the recommendation that ethanol’s use should be limited to settings where fomepizole is not available or contraindicated. The same editorial by Watson also supports the use of hemodialysis for renal insufficiency or metabolic acidosis rather than the traditional criterion of serum EG concentrations >50 mg/dL. Interestingly, Corley and McMartin (2005) have refined a previously published PBTK model for EG and its GA metabolite in rats and humans to include hemodialysis, ethanol, and fomepizole as therapeutic interventions. This enabled the model to describe data from several human case reports of EG poisoning and demonstrated that fomepizole, if administered early enough, can indeed be more effective than ethanol or hemodialysis in inhibiting EG metabolism.

From a regulatory standpoint, EPA has established a RfD of 2 mg/kg per day based on kidney toxicity in the chronic rat feeding study by DePass et al. (1986) (EPA, 2006b). Palmer and Brent (2005) have recently derived a toxicity value of 43.7 mg/kg per day based on a NOAEL for acute renal toxicity in humans using international programme for chemical safety guidelines. Although considerably higher than EPA’s RfD, it is comparable to the benchmark dose 0.5 (BMD_0.5) for EG-induced nephrotoxicity of 49 mg/kg per day derived by Health Canada (2000).

Diethylene Glycol

DEG is similar in physicochemical properties to EG, but has a higher boiling point, viscosity, and specific gravity. It serves as a chemical intermediate in the production of polyester resins and polyurethanes, and as a solvent for shellacs and printing ink. It is hygroscopic, which leads to applications as a drying agent for natural and industrial gases, a humectant for cork and paper, and an additive in cosmetics.

DEG’s use as an excipient in a liquid sulfanilamide preparation resulted in 105 deaths in the United States in 1937. This incident prompted passage of the Food, Drug and Cosmetic Act of 1938 (Wax, 1995). Use of DEG-contaminated propylene glycol (PG) or glycerin in various pharmaceuticals has caused multiple fatalities from renal failure in Nigeria, Bangladesh, India, and Haiti. In the Haitian incident, 109 cases of acute renal failure (with 88 deaths) were identified in children who received a locally manufactured acetaminophen syrup containing DEG-contaminated glycerin (O’Brien et al., 1998). The median lethal dose of DEG was estimated at 1.34 mL/kg. Renal failure was the “hallmark” finding in these cases, but hepatitis, pancreatitis, and severe neurologic manifestations (eg, encephalopathy, optic neuritis with retinal edema, and unilateral facial paralysis) were frequently seen. Alfred et al. (2005) have presented seven cases of DEG poisoning characterized by metabolic acidosis, renal failure, and, in three patients, neurotoxicity.

Compared with EG, toxicological data for DEG in animals are quite limited. Consistent with observations in human poisonings, Fitzhugh and Nelson (1946) reported dose-dependent hepatorenal injury in rats that consumed diets containing DEG. Kraul et al. (1991) reported oliguria or polyuria, proteinuria, and other manifestations indicative of renal tubule injury in rats given a single i.p. dose of DEG. Similar to EG, renal tubular necrosis and COM crystal deposition in tubules have been observed following acute exposure of male rats (Hebert et al., 1978). As for reproductive/developmental effects, high-dose gavage and dietary studies during gestation in mice, rats, and rabbits have largely been negative for embryotoxicity and teratogenicity (Ballantyne and Snellings, 2005). One study in mice employing a continuous breeding protocol has revealed diminished reproductive performance and fertility, as well as limited data on craniofacial malformations in live-born and dead pups at drinking water exposures equivalent to 6.1 g/kg per day (Williams et al., 1990). Citing this lone positive result and the developmental toxicity profile of EG, Ballantyne and Snellings (2005) gavaged CD-1 mice and CD rats with a range of DEG doses (559-11,180 mg/kg per day) on gestation days six to 15. They reported no embryotoxic or teratogenic effects at any dose in either species, except for delayed ossification consistent with reduced fetal body weight in rats.

Like EG, DEG is well absorbed from the GI tract, distributed throughout total body water and organs on the basis of blood flow, and initially metabolized by ADH and subsequently by ALDH (Heilmair et al., 1993). The ether linkage of DEG is not cleaved and no appreciable amounts of EG or EG metabolites are formed from DEG, although small amounts of OA have occasionally been reported (Hebert et al., 1978; Winek et al., 1978). Based on studies in rats and dogs, unchanged DEG recovered in urine constitutes the majority of oral doses, with a single urinary metabolite, (2-hydroxyethoxy)acetic acid, accounting for most of the remainder (Lenk et al., 1989; Wiener and Richardson, 1989; Mathews et al., 1991). Differences in the toxicity/potency of DEG and EG can thus be explained by differences in their TK profiles. DEG’s nephrotoxic moiety(ies) has(have) not been positively identified. As with EG, ADH inhibitors can be effective antidotes for DEG, as demonstrated by the use of fomepizole and hemodialysis to successfully treat a 17-month-old girl who ingested DEG (Brophy et al., 2000).

Propylene Glycol

PG is used as an intermediate in the synthesis of polyester fibers and resins, as a component of automotive antifreeze/coolants, and as a deicing fluid for aircraft. As PG is “generally recognized as safe” by the FDA, it is a constituent of many cosmetics, processed foods, and tobacco products, and serves as a diluent for oral, dermal, and i.v. drug preparations (ATSDR, 1997b). The most important routes of exposure in the general population are ingestion and dermal contact with products containing the compound. The use and disposal of deicing solutions is the major means by which PG is released to the environment. PG has a high mobility in soil and the potential to leach into groundwater, but is neither persistent nor bioaccumulative. Its soil and water half-lives are a few days under aerobic or anaerobic conditions. Workers in industries involved in manufacturing or use of products containing PG may be exposed to concentrations higher than the general population, particularly when these materials are heated or sprayed.

PG has a very low order of acute and chronic toxicity (ATSDR, 1997b; LaKind et al., 1999). No organ system has been identified as a target for acute or chronic injury by PG, and there have been no accounts of human fatalities. Glover and Reed (1996) reported a typical clinical case in which a two-year-old child experienced CNS depression and anion gap acidosis after ingesting a hair gel containing PG. Clinical studies and case reports speak of individuals with reactions to PG-containing drug preparations where preexisting conditions exist. For example, a patient with renal insufficiency secondary to chronic cocaine use developed metabolic acidosis after receiving lorazepam for sedation in which PG was a component of the i.v. formulation (Cawley, 2001). Wilson et al. (2000) reported a case of nearly fatal PG toxicity after i.v. diazepam in high doses for alcohol withdrawal. These same authors report on a case series detailing the risk of PG toxicity from i.v. benzodiazepine therapy (Wilson et al., 2005).

Toxicity studies of PG in laboratory animals can be found in the literature, but findings of adverse effects are rare. Christopher et al. (1990) noted increases in anion gap acidosis, CNS depression, and ataxia in cats ingesting high doses of PG, consistent with progressively elevated plasma lactate levels. PG has not been shown to be mutagenic and was negative for carcinogenicity in a chronic feeding study of male and female rats (Gaunt et al., 1972). As for PG’s reproductive/developmental toxicity, the NTP Center for the Evaluation of Risks to Human Reproduction released a monograph on the subject largely based on an expert panel report (NTP, 2004e,g). The overall conclusion was that there is negligible concern for adverse developmental/reproductive toxicity from PG exposures in humans, because animal studies have shown no such effect even at the highest doses tested.

PG’s TK profile explains its relative lack of toxicity. As with other glycols, PG is readily absorbed from the GI tract and distributed throughout total body water and to organs on the basis of blood flow. Approximately 55% of PG is metabolized by ADH to lactaldehyde, while a significant percentage (~45%) is excreted unchanged by the kidneys (Morshed et al., 1988). PG has a mean serum half-life in humans of two to four hours. Whereas excessive lactic acid from lactaldehyde metabolism is primarily responsible for metabolic acidosis observed in extreme exposures, lactate is a good substrate for gluconeogenesis, an efficient detoxification mechanism (NTP, 2004e,g). This detoxification mechanism typically does not allow lactic acid to accumulate to toxic levels, even under saturable metabolic conditions. The rate-limiting step in PG metabolism is its conversion to lactaldehyde. Saturation of this metabolic step in humans occurs at doses eight- to 10-fold lower than observed in laboratory animals. This is protective, because PG has a lower general toxicity than its metabolites (NTP, 2004e,g). As for EG and DEG, ADH inhibitors may competitively inhibit PG metabolism and thus be beneficial in the PG-poisoned patient. This is exemplified by a case of coingestion of ethanol and PG-containing antifreeze by a 61-year-old man absent of significant acid–base disturbance and minimal lactate elevation (Brooks and Wallace, 2002).

Two structural analogues of PG also have low hazard profiles. Both dipropylene glycol (DPG) and tripropylene glycol (TPG) are widely used in personal care products such as perfumes, facial makeup, stick deodorants, and shaving and skin-care preparations. Their low hazard profiles are predictable, given that DPG is rapidly converted to PG, and TPG is rapidly hydrolyzed to DPG, which is further hydrolyzed to PG. The toxicity of these compounds has been summarized (UNEP, 1994, 2001). A negative DPG cancer bioassay was conducted in F344 rats and B6C3F1 mice receiving the chemical via drinking water (Hooth et al., 2004; NTP, 2004c).

If one alcohol residue of EG (HO–CH₂–CH₂–OH) is replaced by an ether, the resulting compound is a monoalkyl glycol ether such as EG monomethyl ether, also called 2-methoxyethanol (2-ME; CH₃–O–CH₂–CH₂–OH). If both alcohols are replaced by ethers, the result is a dialkyl glycol ether such as EG dimethyl ether (CH₃–O–CH₂–CH₂–O–CH₃). The alkyl group at the end of the ether linkage may be a straight or branched short-chain moiety (eg, methyl, ethyl, n-propyl, isopropyl, or butyl). The butyl moiety results in one of the most widely used glycol ethers, 2-butoxyethanol (2-BE) (CH₃–CH₂–CH₂–CH₂–O–CH₂–CH₂–OH). Acetates of monoalkyl ethers such as 2-methoxyethanol (2-ME) acetate (CH₃–CO–O–CH₂–CH₂–O–CH₃) are also common solvents that undergo rapid ester hydrolysis to their parent glycol ethers in vivo, and thus tend to exhibit the same toxicity profiles as unesterified glycols. Glycol ethers exhibit properties of both alcohols and ethers and are thus soluble in water and most organic solvents. This dual solubility and a favorable evaporation rate make glycol ethers very popular solvents for surface coatings such as varnishes and latex paints. Glycol ethers also find use as solvents in paint thinners and strippers, inks, metal cleaning products, liquid soaps, and household cleaners, and are used as jet fuel anti-icing additives and in semiconductor fabrication. Human exposure occurs mainly via inhalation, but also by dermal absorption.

Although glycol ether metabolism varies with chemical structure, some generalizations are possible (Fig. 24-12). For EG monoalkyl ethers, the major metabolic pathway is oxidation via ADH and ALDH to alkoxyacetic acids. For example, 2-ME and 2-BE are metabolized to methoxyacetic acid (MAA; CH₃–CO–O–CH₂–O–CH₃) and BAA (CH₃–CO–O–CH₂–O–CH₂–CH₂–CH₂–CH₃), respectively. The competing O-dealkylase pathway results in the cleavage of the ether linkage to form EG and an alkyl aldehyde. Medinsky et al. (1990) reported that the relative contribution of the oxidative pathway to metabolism of EG ethers increases with increasing alkyl chain length, whereas that of the O-dealkylase pathway decreases. In contrast, the propylene series of glycol ethers (eg, propylene glycol monomethyl ether [PGME]) is predominantly biotransformed to PG via O-dealkylation. Glycol ethers may also be conjugated with glucuronide or sulfate, but this is thought to occur mainly after saturation of other metabolic pathways.

In vitro and in vivo toxicity studies demonstrate that some glycol ethers and their oxidative metabolites are reproductive, developmental, hematologic, and immunologic toxicants by all exposure routes. A few have also tested positive in rodent cancer bioassays. The metabolism of glycol ethers is considered a prerequisite to their toxicity, as the alkoxyacetic acids are usually regarded as the ultimate toxicants. Their acetaldehyde precursors have also been implicated on occasion. A critical role for metabolism is supported by the differential toxicities of glycol ethers metabolized via the oxidative and O-dealkylase pathways (Miller et al., 1984; Ghanayem et al., 1987). This differential toxicity has resulted in a dramatic shift away from the production and use of certain low-molecular-weight EG ethers and their acetate esters (eg, 2-ME and 2-ethylene glycol monoethyl ether or 2-EE) toward those with more favorable toxicity profiles such as EG, DEG, and triethylene glycol butyl ethers and those in the low-molecular-weight propylene series such as PGME (de Ketttenis, 2005; Spencer, 2005). Like several glycol ethers of the ethylene series, PBTK models have been developed for PGME and its acetate in rats and humans and exercised for risk assessment, including derivation of a RfC and a RfD (Corley et al., 2005b; Kirman et al., 2005; Lemazurier et al., 2005). Structure–activity relationships have been discerned in studies of glycol ethers with various chemical substitution patterns (Hardin et al., 1984; Rawlings et al., 1985; Hardin and Eisenmann, 1987; Ghanayem et al., 1989). The reproductive and developmental toxicities of the monoalkyl glycol ethers, for example, tend to decrease with increasing alkyl chain length, whereas hematotoxicity increases. Thus, structure–activity relationships may not be universally applicable across multiple toxicological end points.

Reproductive Toxicity

Epidemiological studies have reported associations between glycol ether exposure and increased risk for spontaneous abortion, menstrual disturbances, and subfertility among women employed in the semiconductor industry (Schenker et al., 1995; Correa et al., 1996; Chen et al., 2002; Hsieh et al., 2005). These associations appear biologically plausible, given the reproductive toxicity of 2-ME and its active metabolite, MAA, is manifested as ovarian luteal cell hypertrophy and increased progesterone production in the female rat. Furthermore, MAA increased progesterone production in cultured human luteal cells at the same concentration it did in rat luteal cells, implying that it has the potential to alter ovarian luteal function in women (Almekinder et al., 1997; Davis et al., 1997). Reproductive effects, primarily reversible spermatotoxicity, have also been described for men exposed to glycol ethers, in some cases at concentrations well below current OSHA PELs for 2-ME (25 ppm) and 2-EE (200 ppm). For instance, Welch et al. (1988) found that painters exhibited oligospermia and azoospermia following average exposure to 2-EE and 2-ME at 2.7 and 0.8 ppm, respectively. In addition, men exposed to a mean level of 6.6-ppm 2-EE in a foundry had decreased numbers of sperm per ejaculate (Ratcliffe et al., 1989). Lastly, a large case–control study of male fertility clinic patients revealed a highly significant association between a diagnosis of impaired fertility and the detection of ethoxyacetic acid in urine (Veulemans et al., 1993).

Testicular effects in men are supported by experimental observations in animals including seminiferous tubule atrophy, abnormal sperm head morphology, necrotic spermatocytes, decreased sperm motility and count, and infertility (Lamb et al., 1984; Foote et al., 1995; Watanabe et al., 2000). Spermatocytes are among the first cells to be visibly affected following glycol ether exposure and their death involves apoptosis (Chapin et al., 1984; Brinkworth et al., 1995; Ku et al., 1995). Creasy and Foster (1984) noted a consistent order of spermatocyte susceptibility following oral administration of 2-ME or 2-EE to rats: dividing spermatocytes > early pachytene spermatocytes > late pachytene spermatocytes > mid-pachytene spermatocytes. Spermatocytes in the leptotene/zygotene stages of cell division, late-stage spermatids, and spermatogonia can be affected if the dose is increased and exposure prolonged.

As for mechanism of action, studies suggest several possibilities. Beattie et al. (1984) reported that rates of lactate accumulation in cultured rat Sertoli cells were significantly decreased by MAA. Lactate is the preferred metabolic substrate of spermatocytes. Also, 2-EE has been shown to increase oxygen consumption and decrease ATP levels in pachytene spermatocytes in a manner consistent with an uncoupled oxidative state (Oudiz and Zenick, 1986). Mebus et al. (1989) have, in addition, demonstrated that serine, acetate, sarcosine, and glycine attenuated the spermatotoxicity of 2-ME in the rat, suggesting MAA may interfere with the availability of one-carbon units for incorporation into purine and pyrimidine bases necessary for nucleic acid synthesis in pachytene spermatocytes. A mechanistic role for Ca²⁺ has been hypothesized and investigated in a series of studies by Chapin and colleagues. Ghanayem and Chapin (1990) observed that a Ca²⁺ channel blocker afforded protection against 2-ME-induced pachytene spermatocyte cell death. These authors reasoned that 2-ME perturbed Ca²⁺ homeostasis, which is consistent with observations of spermatocyte mitochondrial disruption. Involvement of Ca²⁺ was further suggested by observations that 2-ME activates a Ca²⁺-dependent nuclease, cyclophilin A, found in pachytene spermatocytes and associated with spermatocyte apoptosis (Wine et al., 1997). Whereas an increase in intracellular Ca²⁺ is thought to trigger endonuclease activation, the protection afforded by Ca²⁺ channel blockers against MAA-induced spermatocyte apoptosis is apparently not mediated by preventing a rise in intracellular free Ca²⁺ (Li et al., 1997). Rather, because an intact relationship between Sertoli and germ cells is necessary for the morphological expression of MAA-induced spermatocyte apoptosis, it was reasoned that spermatocyte apoptosis is mediated by Sertoli cell–generated factor(s). This hypothesis proposes that transfer of this factor(s) from Sertoli cells into germ cells (or initiation of spermatocyte apoptosis by 2-ME-damaged Sertoli cells through direct Sertoli cell to germ cell communication) can be inhibited by Ca²⁺ channel blockers through their membrane-stabilizing effects and/or interaction with protein kinase C and/or calmodulin, both of which have demonstrated roles in apoptosis. The inhibition of protein kinase C and calmodulin has been shown to block MAA-induced spermatocyte cell death (Li et al., 1997).

Jindo et al. (2001) advanced the research of their predecessors by using cultured seminiferous tubules of juvenile rats to demonstrate that MAA-induced spermatocyte apoptosis could be blocked with protein kinase inhibitors. Several kinases (eg, Src) increased immediately around dying spermatocytes in the immediate proximity of Sertoli cells. An increase was also noted in the phosphorylation of the endoplasmic reticulum chaperone glucose-regulated protein 94, known also as endoplasmin, which was located inside dying spermatocytes. This work implicates a role for kinase activity in the pathogenesis of MAA-induced spermatocyte apoptosis and suggests the involvement of Sertoli cells. Yet another investigation examined the role of tyrosine kinase pp60 (rat testicular Src), a tyrosine kinase encoded by the Src gene and involved in an array of cell signaling pathways, for its involvement in 2-ME-induced spermatocyte apoptosis. Sertoli cell cytoplasm was observed to be the principal site of Src immunoreactivity in control testis, while 2-ME treatment significantly induced Src expression in dying spermatocytes. In addition, MAA-induced apoptosis was blocked using Src inhibitors, further supporting a role for rat testicular Src in Sertoli–germ cell communication and spermatocyte toxicity of 2-ME (Wang et al., 2000). Furthermore, a suppression subtractive hybridization technique using whole testes from 2-ME-treated mice was employed to create mouse testis cDNA libraries enriched for gene populations either upregulated or downregulated by 2-ME (Wang and Chapin, 2000). A total of 70 clones was screened, and 6 of them were shown to be differentially expressed in the 2-ME lesion, three with increased expression, and three were suppressed. Interestingly, predicted peptide sequences of the six genes revealed several conserved motifs such as phosphorylation sites for protein kinase C and tyrosine kinase. Importantly, these gene changes were apparent at multiple germ cell stages and were localized in multiple germ cell types (Sertoli, interstitial, and peritubular cells). This further suggests the involvement of cell types other than the dying spermatocyte in the pathogenesis of 2-ME-induced spermatocyte death and helps explain the requirement for intact seminiferous tubules for in vitro replication of the pathology observed in vivo.

Developmental Toxicity

Exposure to certain glycol ethers during organogenesis (eg, 2-ME and 2-EE) is toxic to the developing embryo, with effects seen in several animal models including nonhuman primates (Hardin et al., 1986; Scott et al., 1989). Others such as EG butyl, propyl, and monohexyl ethers, and most PG ethers either have not induced fetal malformations or have a lower potential for developmental toxicity (Tyl et al., 1989; Spencer, 2005). Structural anomalies in rodents have included a variety of minor skeletal variations, hydrocephalus, exencephaly, cardiovascular malformations, dilatation of the renal pelvis, craniofacial anomalies, and digit malformations. In the absence of structural defects, electrocardiograms of fetal rats from dams treated with 2-ME during gestation showed persistent, aberrant QRS waves, suggestive of an intraventricular conduction delay (Toraason and Breitenstein, 1988). Neurobehavioral changes and regional brain alterations of several neurotransmitters in offspring of rats treated with 2-ME or 2-EE have been reported (Nelson and Brightwell, 1984).

Little is known about the mechanism by which glycol ethers exert their developmental effects. 2-ME has served as a model toxicant to investigate the disposition of weak acids in the maternal–fetal unit and the hypothesis that weak acids such as MAA exert their effects by altering embryonic pH at critical stages of organogenesis (Nelson et al., 1989; Clarke et al., 1992; O’Flaherty et al., 1995; Terry et al., 1995). Ambroso et al. (1998) have applied confocal laser scanning microscopy, classical histopathology, and in situ immunohistochemistry to demonstrate that 2-ME caused a dose-dependent increase and expansion of apoptosis in gestation day eight mouse embryos that could underlie 2-ME-induced neural tube defects. Such a mechanism has also been hypothesized for malformations induced by several prototypical teratogens such as retinoic acid and ethanol.

Few epidemiological studies have addressed developmental effects of glycol ethers. Saavedra et al. (1997) described facial malformations and varying degrees of mental retardation in 44 offspring of mothers who were exposed occupationally to 2-ME and EG at a factory producing capacitors in Mexico. There are a few reports published from a multicenter case–control study in Europe designed to investigate the role of maternal exposures at work and congenital malformations (Ha et al., 1996; Cordier et al., 1997; Lorente et al., 2000). Preliminary results (Ha et al., 1996) of evaluation of offspring of mothers who were exposed to glycol ethers at work during pregnancy found excesses of oral clefts (OR = 2.0; 95% CI = 1.1– 4.1) and CNS malformations (OR = 1.8; 95% CI = 1.1–3.3). In a study of 984 cases of major congenital malformations, Cordier et al. (1997) reported an overall OR of congenital malformations associated with glycol ether exposure of 1.44 (95% CI = 1.10–1.90), with significant associations for glycol ether exposure with cleft lip, multiple anomalies, and neural tube defects. Lorente et al. (2000) studied 100 mothers of babies with oral clefts and 751 mothers of healthy babies and reported a nonsignificant OR of 1.7 (95% CI = 0.9–3.3) for maternal occupational exposure to glycol ethers and cleft lip, with or without cleft palate. Maldonado et al. (2003) have reviewed the epidemiological evidence and determined that it is insufficient to determine whether occupational exposure to glycol ethers causes human congenital malformations.

Out of concern for the potential of EG monoalkyl ethers as developmental toxicants, several PBTK models have been developed. Hays et al. (2000) developed a PBTK model for 2-ME and MAA in the pregnant rat that was capable of predicting embryonic concentrations. Gargas et al. (2000a,b) applied a PBTK model to estimate inhaled concentrations of 2-EE, its acetate ester, and 2-ME in humans that would result in blood levels equivalent to those observed at the rat NOAELs and LOAELs for developmental effects. Sweeney et al. (2001) applied Monte Carlo simulations to the models of Gargas and coworkers to account for the variability in TK and TD factors among humans and animals and derived occupational exposure limits to protect workers from developmental effects of 2-ME and 2-EE that were one to two orders of magnitude lower than current OSHA PELs. It is worthy of note that 2-ME has been largely removed from commerce due to its teratogenic potency.

Hematotoxicity

Some glycol ethers are hemolytic to red blood cells (RBCs). Typically, the osmotic balance of cells is disrupted, they imbibe water and swell, their ATP concentration decreases, and hemolysis occurs (Ghanayem, 1989). Nyska et al. (1999) reported that subchronic exposure to 2-BE causes disseminated thrombosis and bone infarctions in female, but not male rats, likely due to impedance of blood flow by intravascular hemolysis. It is thought that females might be susceptible, because they are less efficient in eliminating BAA, the hemolytic metabolite of 2-BE, and exhibit higher peak blood BAA levels. Young adult rats are more resistant to the hematologic effects of 2-BE than older rats, an observation attributed to depressed degradation and renal clearance of BAA in the older rats.

Species differ dramatically in their sensitivities to glycol ether–induced RBC deformity and hemolysis. Humans are less susceptible than rodents. This lower susceptibility even applies to RBCs from potentially sensitive subpopulations, such as the elderly and persons with hereditary blood disorders (Udden, 1994; Udden and Patton, 1994). A good example of using PBTK models in human risk assessment has been published by Corley et al. (1994). Based on comparisons of model output with data collected by Udden and colleagues on levels of 2-BE required to affect osmotic fragility of human RBCs, Corley and coworkers concluded that humans are unlikely to achieve hemolytic blood levels of BAA unless very large volumes of 2-BE are intentionally ingested. Udden (2005) has recently reported on the hemolytic effects of diethylene glycol butyl ether (DGBE) and its principal metabolite, butoxyethoxyacetic acid (BEAA), using rat and human RBCs in vitro. BEAA had weak hemolytic activity on rat erythrocytes, which is consistent with the finding of mild hemolysis when DGBE is administered to rats by gavage. However, such effects were absent in human RBCs exposed to DGBE or BEAA, indicating that it is unlikely hemolysis will occur in humans exposed to DGBE. Johnson et al. (2005) recently confirmed DGBE’s low order of hematotoxicity in a 13-week drinking water study in F344 rats that identified a NOAEL of 250 mg/kg per day, with minimal but statistically significant decreases in RBC count, hemoglobin, and hematocrit at 1000 mg/kg per day.

Hoflack et al. (1997) have shown 2-BE capable of inducing apoptosis in a human leukemia cell line and have hypothesized that the hematopoietic toxicity of 2-BE may be the result of its ability to induce apoptotic cell death. However, once inside the cell it is not entirely clear how hemolysis is accomplished, although the RBC membrane has long been the suspected target. Udden and Patton (2005) have utilized BAA to examine the mechanism of glycol ether hemolysis in rat RBCs in vitro. They concluded that the mode of action of BAA is to cause a colloid osmotic lysis of the RBC and speculated the following scenario: BAA causes Na⁺ and Ca²⁺ to enter the cell → Ca²⁺ initially has a protective effect via the Ca²⁺-activated potassium channel, which facilitates the loss of potassium, thereby compensating for the osmotic effect of increased cell Na⁺ → Ca²⁺ subsequently has deleterious effects through activation of proteases and the loss of the normal asymmetric distribution of phospholipids (eg, phosphatidylserine) in the membrane bilayer. These authors noted that preliminary studies in their laboratory have shown the movement of phosphatidylserine from the inner to the outer leaflet of the lipid bilayer of rat RBCs incubated with BAA. This “externalization” of phosphatidylserine is associated with adhesion of RBCs to endothelial cells and the generation of thrombin, which is most interesting given reports of disseminated thrombosis and infarction in 2-BE-treated rats (Nyska et al., 1999; Ghanayem et al., 2001).

Immunotoxicity/Carcinogenicity

Based on changes in thymus and splenic weights/cellularities and a variety of in vitro and in vivo immune function assays, the immune system is a potential target for the oxidative metabolites of some glycol ethers. 2-ME and MAA have been employed almost exclusively in immunotoxicity investigations of glycol ethers. Not only have adult animals proven susceptible, but also 2-ME exposure of pregnant mice induces fetal thymic atrophy/hypocellularity and a reduction in fetal liver prolymphocytes with potential implications for fetal immunity (Holladay et al., 1994). Using B6C3F1 mice and gavage exposure, House et al. (1985) were among the first to report that 2-ME and MAA reduced thymus weight. Kayama et al. (1991) subsequently reported that 2-ME selectively depleted immature thymocytes in mice. Exon et al. (1991) reported not only thymic atrophy in rats exposed to 2-ME in drinking water but also decreased antibody production, decreased splenocyte production of interferon-γ, and a reduction in spleen cellularity. Around the same time, the first in a lengthy series of studies by Smialowicz and colleagues was published (Smialowicz et al., 1991a,b, 1992, 1994; Williams et al., 1995; Kim and Smialowicz, 1997). This series has reported decreased thymus weights, reduced lymphoproliferative responses to mitogens, and reduced IL-2 production in splenocytes of F344 rats exposed to 2-ME by gavage. It has also generated data indicating that not all glycol ethers are immunosuppressive, that mice are relatively insensitive to glycol ether immunosuppression compared with rats, and that rats of various strains show differential sensitivities. The Smialowicz series has further demonstrated that the relative insensitivity of mice is not a function of their more rapid clearance of MAA; that 2-ME is immunotoxic when applied dermally to F344 rats; and that questions remain as to 2-ME’s proximate immunotoxicant, as 2-methoxyacetaldehyde is more immunotoxic than MAA based on the ability to suppress IgM and IgG production by lymphocytes in F344 rats.

As for cancer, only a few chronic bioassays have been conducted with glycol ethers. Two-year inhalation bioassays of 2-BE in F344 rats and B6C3F1 mice revealed some evidence of carcinogenicity in male mice, based on increased incidences of hemangiosarcoma of the liver, as well as some evidence of carcinogenic activity in female mice, based on increased incidences of forestomach squamous cell papilloma or carcinoma (mainly papilloma) (NTP, 2000). In June 2004, an IARC working group evaluated the cancer risk of 2-BE and concluded that it is not likely to be carnicogenic to humans at environmental concentrations at or below the RfD and RfC (Cogliano et al., 2005). Likewise, EPA’s IRIS profile for 2-BE currently indicates that the human carcinogenic potential of 2-BE cannot be determined at this time (EPA, 2012). Since NTP’s bioassays were completed, numerous studies have shed light on 2-BE’s possible modes of action related to liver hemangiosarcomas and forestomach tumors and their implications for risk assessment (Park et al., 2002; Siesky et al., 2002; Poet et al., 2003; Boatman et al., 2004; Klaunig and Kamendulis, 2005; Corthals et al., 2006). As discussed by Gift (2005), these studies suggest the following scenario: 2-BE consumed while grooming is metabolized to irritant metabolites in the forestomach and/or irritant metabolites are formed in the upper respiratory tract and swallowed → chronic irritation → inflammation → hyperplastic effects → forestomach tumors. As for liver hemangiosarcomas, the following is suggested: 2-BE is metabolized to BAA → BAA causes hemolysis of RBCs → hemosiderin (iron) derived from released hemoglobin is taken up by and stored in phagocytic cells (eg, Kupffer cells) of the spleen and liver → oxidative damage and increased synthesis of endothelial DNA are initiated by ROS from excess iron or Kupffer cells, producing cytokines/growth factors that suppress apoptosis and promote cell proliferation → endothelial DNA mutations → potentiation and promotion of hepatic neoplastic cell populations. As further discussed by Gift (2005), the evidence suggests nonlinear modes of action in both cases and questionable human relevance of both tumor types. Several PBTK models have been developed and subsequently refined for 2-BE and BAA to aid in risk assessments (Corley et al., 1994, 2005c; Lee et al., 1998; Franks et al., 2006).

Spencer et al. (2002) have reported a two-year inhalation bioassay of PGME in F344 rats and B6C3F1 mice that did not result in increases in neoplasia in either species except for kidney adenomas in male rats related to α_2u-globulin nephropathy. In contrast, a two-year inhalation bioassay with propylene glycol mono-t-butyl ether (PGMBE) in F344 rats and B6C3F1 mice also resulted in α_2u-globulin nephropathy in male rats, as well as liver tumors in male and female B6C3F1 mice at the highest concentration tested (1200 ppm) (NTP, 2004a; Doi et al., 2004). Dill et al. (2004) have published information on PGMBE TK in rats and mice that demonstrate saturation of PGMBE metabolism/elimination at this tumor-producing concentration. The genotoxicity of some glycol ethers and their metabolites has been evaluated, with most exhibiting a lack of genotoxic potential and others yielding weakly positive responses in certain tests. Therefore, the role of genetic toxicology in the toxicities discussed above cannot be summarily dismissed, but is of unknown significance (Elliot and Ashby, 1997; NTP, 2000; Ballantyne and Vergnes, 2001).

Automotive Gasoline

Automotive gasoline is a complex mixture of hundreds of hydrocarbons predominantly in the C₄ to C₁₂ range. The sheer number of people exposed in the manufacture, distribution, and use of gasoline makes characterization of its acute and chronic toxicities important. Generalizations regarding gasoline toxicity must be made with care, because its composition varies with the crude oil from which it is refined, the refining process, and the use of specific additives. Experiments conducted with fully vaporized gasoline may not be predictive of actual risk, because humans are exposed primarily to the more volatile components in the range of C₄ and C₅. These hydrocarbons are generally regarded as less toxic than their higher-molecular-weight counterparts. Concern about gasoline exposure is fueled in part by the toxicities of certain components, some of which are classified by EPA as known or probable human carcinogens (eg, benzene and 1,3-butadiene). The ACGIH has established a TLV for gasoline of 300 ppm to prevent ocular and upper respiratory tract irritation and a STEL of 500 ppm to avoid acute CNS depression.

Inhalation exposure to gasoline has been measured for service station attendants, self-service customers, truck drivers, distribution workers, and workmen removing leaking underground storage tanks (Kearney and Dunham, 1986; Shamsky and Samimi, 1987). In one survey, short-term exposures of self-service customers averaged about 6 ppm. The TLV is rarely exceeded in occupationally exposed individuals, due in part to the use of vapor scavenging systems. Brief exposures in excess of the STEL have, however, been documented for workers engaged in bulk handling operations (Phillips and Jones, 1978). The most extreme exposures occur to those intentionally sniffing gasoline for its euphoric effects. Several case reports of acute and chronic encephalopathies are testament to the dangers of this habit (Valpey et al., 1978; Fortenberry, 1985). In these cases, the identity of the offending agent(s) is often unclear. Gasoline is one of the most popular and lethal inhalants (Spiller, 2004; Wu et al., 2004), with deaths reported even among Aboriginal people in South Australia (Byard et al., 2003). An all too common occurrence is the ingestion of gasoline during siphoning events. This is typically followed by a burning sensation in the mouth and pharynx, as well as nausea, vomiting, and diarrhea resulting from GI irritation. If aspirated into the lungs, gasoline may produce pulmonary epithelial damage, edema, and pneumonitis. Thus, emetic therapy for gasoline ingestion is usually contraindicated.

Between 1986 and 2004, EPA identified 447,233 releases from underground petroleum storage tanks, many of which threaten groundwater that serves as the primary drinking water source for nearly one half of the US population. Despite the number of releases, few community health studies have been conducted, and those that have are typically driven by concerns over leukemia risk owing to gasoline’s benzene content. Consider, for example, the retrospective cohort study of the “Tranguch Gasoline Spill” in northeastern Pennsylvania (Patel et al., 2004). The standard incidence ratio for leukemia of all types was significantly elevated (4.40; 95% CI = 1.09–10.24), consistent with that reported by the Pennsylvania Department of Health. However, the excess was based on only four cases, two of which had a history of smoking, a potential confounder. In addition, exposure was not well characterized, and only two of the subjects had AML, the leukemia type most strongly associated with benzene. Such a study exemplifies the problem with inferring causation for environmentally exposed populations based on limited data.

Reese and Kimbrough (1993) and Caprino and Togna (1998) have reviewed the acute toxicity of gasoline and its additives. Like some other solvents, gasoline can sensitize the heart to catecholamines, defat the skin on repeated contact, and induce hepatic CYPs and UDP-glucuronyltransferase activities (Poon et al., 1995). The question of whether there is a “fetal gasoline syndrome” has been raised, although case reports are confounded by tetraethyl lead, alcohol abuse, and the possibility that an aberrant gene is distributed within the small Amerindian population where the cases reside (Hunter et al., 1979). There is a paucity of data on the reproductive toxicity of gasoline, but reports of enhanced estrogen metabolism and uterine atrophy among unleaded gasoline (UG)-treated mice suggest that this end point warrants investigation (Standeven et al., 1994a). Although dated, the study of Lykke and Stewart (1978) is of interest, because rats exposed to leaded gasoline at one-third the ACGIH TLV (ie, 100 ppm) for 40 hours per week for six to 12 weeks were observed to have a progressive interstitial fibrosis of the lungs associated with irregular alveolar collapse.

Prior to the identification of α_2u-globulin as the principal accumulating protein in the syndrome referred to as α_2u-globulin nephropathy, Kuna and Ulrich (1984) reported regenerative epithelium and dilated tubules in the kidneys of male rats exposed to 1552 ppm UG for 90 days. At about the same time, a chronic inhalation study revealed not only nephropathy but also increased renal tumors in male rats (MacFarland et al., 1984). Subsequent studies by Halder et al. (1986) and Aranyi et al. (1986) showed that such nephropathy could not be produced by exposure of rats to a mixture of the butane and pentane components of gasoline or the 0°F to 145°F gasoline distillation fraction. These are thought to be more representative of human occupational exposures than wholly vaporized gasoline. In addition, the authors of a gavage screening study of 15 pure hydrocarbons and gasoline fractions concluded that branched aliphatic alkane components were primarily responsible for the nephropathy (Halder et al., 1985).

Investigations of mechanisms of the nephropathy and renal tumors included an assessment of unscheduled (a measure of genotoxicity) and replicative DNA synthesis (a measure of cell proliferation) in rat kidney cells exposed in vitro and in vivo to UG. No unscheduled DNA synthesis occurred, even at a tumorigenic dose, while a five- to eightfold increase in cell proliferation was observed (Loury et al., 1987). In a publication the same year by Olson et al. (1987), UG was reported to result in an increase in hyaline droplets harboring large accumulations of α_2u-globulin within proximal convoluted tubule epithelial cells. It was hypothesized that α_2u-globulin accumulated secondary to a defect in renal lysosomal degradation of the protein (Fig. 24-6). Supportive evidence for this hypothesis came from the demonstration that inhibition of the lysosomal peptidase, cathepsin B, caused a rapid accumulation of phagolysosomes and α_2u-globulin in the kidney similar to that of UG (Olson et al., 1988).

Further progress in elucidating the mechanism of α_2u-globulin nephropathy came from the demonstration that the UG component, 2,2,4-trimethylpentane (TMP), itself an inducer of α_2u-globulin nephropathy, was metabolized to 2,4,4-trimethyl-2-pentanol (TMPOH), which was selectively retained by the kidney of male rats. Subsequently, it was demonstrated that the sex-specific retention of TMPOH in the kidney was due to reversible binding with α_2u-globulin. This binding rendered the protein less digestible by lysosomal enzymes, which accounted for its accumulation (Charbonneau and Swenberg, 1988). This accumulation, in turn, led to cellular degeneration and necrosis, primarily in the P₂ segment of the proximal tubule. In response, regenerative proliferation occurs and promotes formation of renal cell tumors by irreversibly “fixing” spontaneously altered DNA and clonally expanding initiated cells. The promotional effects of gasoline and TMP on atypical cell foci and renal cell tumors have been demonstrated in male rats following initiation with N-ethyl-N-hydroxyethylnitrosamine (Short et al., 1989). NCI–Black–Reiter male rats, the only rat strain not to synthesize α_2u-globulin, are resistant to gasoline- and TMP-induced nephropathy (Dietrich and Swenberg, 1991). Thus, gasoline and TMP have been of great value in elucidating the mechanism of α_2u-globulin nephropathy and shedding light on its implications for renal tumorigenesis. Most toxicologists, and indeed the EPA, have concluded that renal tumors secondary to α_2u-globulin nephropathy are of little relevance, because humans do not synthesize α_2u-globulin.

Chronic inhalation of gasoline at high concentrations has also resulted in increased hepatocellular adenomas and carcinomas in female B6C3F1 mice, possibly due to the promotion of spontaneously initiated cells that occur with unusually high frequency in this mouse strain (MacFarland et al., 1984). This possibility is supported by reports that UG is a CYP inducer, mitogen, and liver tumor promoter in N-nitrosodiethylamine (DEN)-initiated female B6C3F1 mice (Standeven and Goldsworthy, 1993; Standeven et al., 1995; Moser et al., 1996a). CYP induction by UG has been attributed to “heavy UG” (components with boiling points >100°C), whereas mitogenic activity is highly concentrated in UG components boiling from 100°C to 132°C, for which the 2,2,3-, 2,2,4-, and 2,3,4-trimethylpentane isomers appear at least partially responsible (Standeven and Goldsworthy, 1994). It has been hypothesized that the liver tumor-promoting activity of UG is secondary to its estrogen antagonism, given that (1) UG is not a hepatocarcinogen in male mice; (2) estrogen inhibits liver tumor development initiated in mice but potentiates liver tumor promotion by UG; and (3) UG induces hepatic estrogen metabolism (Standeven et al., 1994b). The hypothesis that liver tumor promotion by UG depends on its interaction with estrogen is supported by the demonstration that tumor-promoting activity of UG was greatly attenuated in ovariectomized mice relative to intact mice (Moser et al., 1997). Further, the addition of estrogen to DEN-treated mice substantially reduces the percentage of hepatic foci with decreased levels of TGF-β1 compared with DEN-treated control mice or DEN + UG-treated mice, suggesting a promotional mechanism involving estrogen and the dysregulation of tumor growth factor(s) (Moser et al., 1996c). Whereas much attention has been given to its promotional potential, UG may also damage DNA, as it reportedly induces unscheduled DNA synthesis in hepatocytes from male and female mice treated in vivo and in cultured mouse, rat, and human hepatocytes (IARC, 1989b). The epidemiological evidence for an association between gasoline exposure and cancer in humans is inconclusive. Raabe (1993) has reviewed the carcinogenic potential of gasoline. IARC (1989b) classifies it as possibly carcinogenic to humans (Group 2B) primarily due to its benzene content. A comprehensive review of gasoline toxicity is provided in ATSDR’s (1995) Toxicological Profile for Gasoline.

Vehicle emissions from gasoline combustion are a major contributor to urban air pollution, which is at unhealthy levels in numerous cities. In response, the Clean Air Act Amendments of 1990 require the use of oxygenated gasoline in such areas. Oxygenated gasoline contains additives that add oxygen to gasoline, thereby boosting its octane quality, enhancing combustion, and reducing exhaust emissions. MTBE and ethanol are the two most common oxygenates, although use of the former is being rapidly phased out due to widespread groundwater contamination and health concerns. As a result, the demand for ethanol–gasoline blends is increasing dramatically, raising concerns about how the two components might interact toxicologically. There is a dearth of information on this issue, but a four-week inhalation study of an ethanol–gasoline mixture (6130-ppm ethanol and 500-ppm gasoline) in rats concluded that coexposure showed additive and possibly some synergistic effects on growth, neurochemistry, and histopathology of the adrenal gland and respiratory tract. Effects were described as generally mild and adaptive in nature, and returned to normal after exposure cessation (Chu et al., 2005). The risks and benefits of ethanol as an oxygenate are discussed in detail by Williams et al. (2003), who point out several reasons why ethanol in gasoline can increase groundwater plume lengths and persistence of gasoline constituents in groundwater if ethanol blends are released into the environment. The most obvious concern is that longer or more persistent gasoline plumes could lead to a higher probability of gasoline constituents affecting public water wells. Meanwhile, another fuel additive, methylcyclopentadienyl manganese tricarbonyl (MMT), is receiving attention due to concerns that it could increase manganese inhalation exposures and pose a risk for neurotoxicity. The combustion of MMT gasoline results in the emission of fine Mn particulates mainly as Mn sulfate and Mn phosphate and smaller amounts of oxides. Several studies characterizing vehicular exhaust using MMT gasoline and describing the TK and neurobehavioral toxicity of Mn have been recently published, some of which were mandated by EPA (Normandin et al., 2004; Dorman et al., 2009; Reaney et al., 2006; Tapin et al., 2006). Gasoline engine exhaust has been classified as possibly carcinogenic to humans (Group 2B), based largely on sufficient evidence in animals when condensates and exhaust extracts are tested (IARC, 1989a).

Methyl tertiary-Butyl Ether

MTBE’s high octane rating made it a logical replacement for tetraethyl lead as an octane booster for gasoline, and later as a gasoline oxygenator. As an oxygenator, MTBE makes fuel combustion more complete, thereby reducing pollutant emissions from automobile exhaust. MTBE may be added to gasoline at levels up to 15% by volume in order to comply with the 1990 Amendments to the Clean Air Act. By 1997, it was being used at the rate of 10 million gal per day, with more than one-third of the usage in California (Williams et al., 2000a). While routine, low-level exposure of customers occurs at self-service stations, heightened concern about MTBE has resulted primarily from its contamination of groundwater by leaking underground gasoline tanks. It is highly water soluble, travels faster and farther in water than other gasoline components, and is resistant to degradation.

MTBE is well absorbed following oral, inhalation, and dermal exposure of humans and rats (Dekant et al., 2001; Prah et al., 2004; McGregor, 2006). The majority of absorbed MTBE is exhaled unchanged. Some MTBE is oxidized to tert-butyl alcohol (TBA) and HCOH. Whereas this oxidation is primarily CYP2A6-mediated in humans (Hong et al., 1999), it is largely CYP2B1-mediated in rats (Turini et al., 1998). TBA is relatively water soluble, so it tends to remain in the blood and extracellular fluid, and is slowly exhaled. It is further metabolized, first to 2-methyl-1,2-propanediol and then to 2-hydroxyisobutyrate, the major urinary metabolites of MTBE. In addition, glucuronide and sulfate conjugates of TBA are found in trace amounts in urine. Although HCOH is one of the oxidative metabolites of MTBE, it is undetectable after MTBE exposure in humans and rats, presumably due to its rapid metabolism. PBTK models have been developed to describe the dosimetry of MTBE and TBA following inhalation and oral exposures in rats and humans (Borghoff et al., 1996b; Licata et al., 2001).

Concern about MTBE has led to numerous toxicity studies in humans and rodents, and a number of reviews of MTBE’s toxicity are available (ATSDR, 1996b; EPA, 1997; Borak et al., 1998; McGregor, 2006; WHO, 1998). The review of Borak et al. (1998) focuses on the acute human health effects of MTBE. It concludes, based on 19 reports of inhalation exposure to MTBE alone or in gasoline and 12 reports of parenteral MTBE administration to dissolve cholesterol gallstones, that no significant association exists between MTBE exposure and the acute symptoms commonly attributed to it. These symptoms include headache, eye, nose and throat irritation, cough, nausea, dizziness, and disorientation. The more recent review of McGregor (2006) is supportive of this conclusion.

In 1988, the EPA and industry developed a Testing Consent Order for MTBE under the Toxic Substances Control Act that precipitated investigations of MTBE’s potential two-generation reproductive toxicity, developmental toxicity, in vivo mutagenicity, subchronic inhalation toxicity, oncogenicity, and neurotoxicity. Results of these studies are a major addition to the toxicity literature on MTBE and define several NOAELs (Bevan et al., 1997a,b; Bird et al., 1997; Daughtrey et al., 1997; Lington et al., 1997; McKee et al., 1997). The publication of Bird and colleagues is actually a recapitulation of reports by Chun et al. (1992) and Burleigh-Flayer et al. (1992), both of which are of particular value, as they represent two of only three MTBE cancer bioassays. In the study by Chun et al., male and female F344 rats were exposed to 0, 400, 3000, or 8000 ppm MTBE vapor six hours per day, five days per week for 24 months. In the other inhalation study, Burleigh-Flayer et al. (1992) subjected male and female CD-1 mice to the same exposure regimen for 18 months. The only oral chronic bioassay is that by Belpoggi et al. (1995, 1997), who subjected male and female S-D rats by olive oil gavage to 0, 250, or 1000 mg/kg MTBE four days per week for two years. The results of the three MTBE animal cancer bioassays are presented in Table 24-4. In addition, Cirvello et al. (1995) and NTP (1995) have reported some evidence for the carcinogenicity of TBA in the kidney and thyroid after long-term drinking water exposure.

Taken at face value, one might interpret these cancer bioassay findings as ample evidence of carcinogenicity in animals and suggestive of a cancer risk for humans. The relevance of these findings to humans, however, has been a source of debate among toxicologists. Critics have questioned these studies on the basis of (1) the appropriateness of a combined incidence category for leukemias and lymphomas; (2) the possibility that renal tumors were secondary to male rat-specific α_2u-globulin nephropathy; (3) the possibility that Leydig cell tumors were a function of abnormally low testicular tumor rates in control animals or increased survival time of treated rats; (4) the questionable relevance of testicular tumors in rats to humans, given the species’ differential responses of Leydig cells to proliferative stimuli; (5) the possibility that liver and kidney tumors are the result of high-dose-induced chronic cytotoxicity, cell death, and reparative cell proliferation; (6) the questionable relevance of inhalation bioassays to prediction of drinking water risks; and (7) the use of an oil rather than a water-dosing vehicle, which could unduly influence MTBE’s oral TK. Mennear (1997) and McGregor (2006) have discussed a number of these issues.

As an outgrowth of the uncertainties surrounding MTBE’s human carcinogenicity risk, several mechanistic studies have been published. For example, after only 10 days of MTBE inhalation exposure, a strong positive linear relationship between renal α_2u-globulin concentration and cell proliferation was seen in the male F344 rat (Prescott-Matthews et al., 1997). This study, unlike the chronic bioassay of Chun et al. (1992), definitively identified the accumulating protein as α_2u-globulin. Williams and Borghoff (2001) have shown that TBA interacts with α_2u-globulin, which explains its accumulation in the male rat kidney following MTBE or TBA exposure. MTBE has been shown to be a hepatic mitogen in the female mouse, but not a promoter of tumor formation in DEN-initiated female mouse liver (Moser et al., 1996b). It has been suggested in light of these findings that MTBE may promote the growth of spontaneously initiated cell populations having genetic lesions different from those produced by DEN. Casanova and Heck (1997) have reported a lack of concentration, species, and sex dependence in the formation of HCOH-induced DPX and RNA–HCOH adducts in isolated female CD-1 mouse hepatocytes incubated with MTBE. As the cancer bioassay data suggest that hepatocarcinogenicity varies with all of these factors, these results do not support a role for HCOH in MTBE-induced liver tumor formation. Studies also indicate that MTBE causes endocrine dysregulation in rodents at high doses, suggesting the possibility that MTBE-induced tumor formation is hormonally mediated (Moser et al., 1998; Williams et al., 2000b). Changes in triiodothyronine, luteinizing hormone, testosterone, and estradiol levels have been discussed as possible mechanisms of MTBE-induced Leydig cell cancer (Williams and Borghoff, 2000; Williams et al., 2000b; de Peyster et al., 2003). Although it is generally accepted that MTBE induces certain tumors in animals through nongenotoxic mechanisms, experiments indicate that MTBE is mutagenic in a few in vitro test systems (Williams-Hill et al., 1999; Zhou et al., 2000).

Several mechanistic issues surrounding MTBE’s toxicity, particularly its carcinogenicity, have yet to be resolved. Nonetheless, citing widespread groundwater contamination and health concerns, numerous states have instituted phased-in partial or complete bans of MTBE. This is in keeping with the opinion of EPA’s Blue Ribbon Panel on Oxygenates in Gasoline that, in 1999, agreed that with the use of MTBE and other gasoline additives threats to drinking water supplies should be substantially reduced. To date, independent expert review groups who have assessed MTBE inhalation health risks (eg, “Interagency Assessment of Oxygenated Fuels”) have not concluded that the use of MTBE-oxygenated gasoline poses an imminent threat to public health. EPA’s Office of Water has concluded that available data are not adequate to estimate potential health risks of MTBE at low exposure levels in drinking water, but that the data support the conclusion that MTBE is a potential human carcinogen at high doses. Fortunately, water containing MTBE has an unpleasant taste and odor, which may alert consumers to the fact that their water is contaminated. Based on the results of studies of taste and odor thresholds for humans, an advisory guidance range of 20 to 40 μg/L has been set by the EPA to assure consumer acceptance and provide a large margin of safety from toxicity and carcinogenicity. California has both derived a cancer potency estimate and adopted a Public Health Goal of 13 ppb for MTBE in drinking water. While many advocate the suspension of MTBE’s use in reformulated gasoline, such a decision should weigh the benefits against the risks associated with increased auto emissions of carcinogenic VOCs and the public health impact of increased CO₂ emissions and ozone formation. Two articles that discuss this subject in considerable detail are by Spitzer (1997) and Erdal et al. (1997).

Jet Fuel

Jet A, jet propellant-8 (JP-8), and JP-8+100 are the predominant jet fuels in use today. All are kerosene-like mixtures of hundreds of aliphatic and aromatic hydrocarbons. Jet A is commercial aviation fuel, whereas JP-8 and JP-8+100 are military fuels. JP-8 is a mixture of Jet A plus three additives, whereas JP-8+100 contains JP-8 and an additional additive package. JP-8 is now the recognized battlefield fuel for all NATO forces and is used not only for aircraft but also for ground vehicles and other equipment such as generators, cooking stoves, and tent heaters (NAS, 2003). Owing to slight differences in hydrocarbon composition and additives, JP-8 differs from its predecessor fuels (eg, JP-4 and JP-5) in ways that impart added safety, enhance combat aircraft survivability, simplify battlefield logistics, and promote standardization with commercial jet fuel.

Civilian and military personnel are exposed to jet fuel by inhalation and dermal contact. Exposure can occur to liquid, vapor, or aerosol, each phase having a distinct composition and toxicity profile. Exposure is prevalent in aircraft refueling and maintenance operations and ground crews positioned behind jet aircraft during “cold starts” can become “drenched” in aerosol emissions. Jet fuel can be released into the environment by in-flight jettisoning and spills or leaks to soil or water during use, storage, or transportation. In many cases, the US Department of Defense (DoD) is responsible for the cost of remediating contaminated military sites and contractor facilities, not to mention its responsibility to safeguard the health of military personnel. It is thus in DoD’s interest to accurately characterize the toxicity of jet fuel. Much of the research conducted to date has been funded by the military. The Navy’s Occupational Safety and Health Standards Board has proposed an eight-hour TWA PEL of 350 mg/m³ and a 15-minute STEL of 1800 mg/m³ for jet fuel vapors, but notes that exposure to aerosols that are much more toxic may necessitate reevaluation. The NAS Subcommittee on Jet Propulsion Fuel 8 recommended that the eight-hour TWA be considered interim until further research is completed, and that the STEL be lowered to 1000 mg/m³ to avoid acute CNS effects (NAS, 2003).

The complexity of jet fuel mixtures presents multiple challenges to toxicologists and risk assessors alike. One such challenge is to generate experimental exposures to jet fuel that accurately reflect those occurring in humans. The NAS Subcommittee on JP8 has reviewed the methods used to generate exposure atmospheres in several published studies using aerosol/vapor mixtures and suspects that the JP-8 concentrations may have been underreported, particularly in studies by Witten and colleagues (NAS, 2003). Therefore, it seems prudent at this time to discuss aerosol/vapor exposures in qualitative terms only. Because of concerns surrounding the quantitative accuracy of the exposure atmospheres, the NAS Subcommittee recommended an examination of the methods used for their characterization. To this end, Dietzel et al. (2005) have developed and validated a GC/MS method for JP-8, which was subsequently used to characterize the aerosol and vapor fractions of aerosolized fuel in one of the University of Arizona-based inhalation chambers previously utilized by Witten and colleagues.

Jet A, JP-8, and JP-8+100 have similar toxicity profiles, which suggests their toxicities are largely a function of hydrocarbon content rather than additives. Toxicity data on jet fuels have been well summarized (ATSDR, 1998; NAS, 2003; Ritchie et al., 2003), but there remain several data gaps including genotoxicity. Most toxicity studies of jet fuels have focused on the two main portals of entry (ie, lung and skin) and the immune system. In subchronic and chronic rodent inhalation studies of jet fuel vapor, the chief findings have been renal toxicity and neoplasia in male rats consistent with α_2u-globulin nephropathy (Mattie et al., 1991; Bruner et al., 1993). Whereas the liver, kidneys, and testes in humans are not considered particularly sensitive targets of jet fuels, proteomic analyses of these tissues have been conducted in male rats subchronically exposed to JP-8 vapor (Witzmann et al., 2000a,b, 2003). Similarly, gene expression in the whole brain of rats repeatedly exposed to JP-8 vapor has been examined (Lin et al., 2001a, 2004), driven largely by reports of cognitive and motor deficits and neurochemical changes in jet fuel–exposed workers and experimental animals (Baldwin et al., 2001; Ritchie et al., 2001; Rossi et al., 2001; Bell et al., 2005).

Pulmonary Effects

The pulmonary effects of an aerosol/vapor mixture of JP-8 were initially investigated in rats with nose-only exposures designed to simulate military flightline exposures (Hays et al., 1995; Pfaff et al., 1995, 1996). Functional changes in the form of increases in pulmonary resistance and alveolar permeability were accompanied by a decrease in the concentration of the tachykinin substance P (SP) in bronchoalveolar lavage fluid. Pathological changes were observed in lower pulmonary structures including inflammation of the terminal bronchioles, degeneration of alveolar type II epithelial (AIIE) cells, and disruption of terminal bronchial airway epithelium. Most interestingly, the activity of neutral endopeptidase (NEP), an enzyme responsible for the metabolism of SP in the lung, was increased by exposure and a significant inverse relationship between SP and NEP activity demonstrated. Thus, JP-8 appears to exhibit a rather novel mechanism of lung injury that involves the reduction or depletion of SP due to its enhanced metabolism by NEP.

Because SP participates in the maintenance of airway epithelial cell competency, the effect of JP-8 and n-tetradecane (C₁₄), a primary constituent of JP-8, on epithelial barrier integrity was examined in vitro using paracellular mannitol flux in BEAS-2B human bronchial epithelial cells (Robledo et al., 1999). Noncytotoxic concentrations of JP-8 and C₁₄ produced dose-dependent increases in transepithelial mannitol flux that spontaneously reversed to control values over a 48-hour recovery period. This suggests that JP-8 and C₁₄ compromise the integrity of intercellular tight junctions that may precede and initiate the pathological alterations observed in whole animal studies. Evidence was also collected that SP’s protective effect on the lung is largely mediated through the plasma membrane–bound neurokinin receptor, NK₁, present on airway epithelium from the trachea to the respiratory bronchioles (Robledo and Witten, 1999).

Among the most affected alveolar cells in rodent studies are AIIE cells. In support of the hypothesis that apoptotic cell death is responsible at least partially for JP-8’s cytotoxicity in the lung, Stoica et al. (2001) reported that JP-8 results in morphological and biochemical changes characteristic of apoptosis in the rat AIIE cell line, RLE-6TN. Further, Boulares et al. (2002) have collected data that strongly suggest JP-8 triggers apoptosis in rat lung epithelial cells by inducing the generation of ROS, depleting/reducing intracellular GSH, and markedly decreasing mitochondrial membrane potential, thereby initiating the apoptotic cascade (ie, caspase-3 activation and DNA fragmentation).

A characteristic feature of the lung inflammatory response to JP-8 in rodents is vacuolization of AIIE cells and accumulation of pulmonary alveolar macrophages (PAM). The findings of Wang et al. (2002b) suggest that JP-8 causes proinflammatory cytokine secretion by not only PAM but also AIIE cells. The prolonged production of proinflammatory cytokines, together with the proteases produced by activated macrophages and neutrophils, is capable of producing a sustained immune response with increased risk for lung damage. Moreover, cocultures of AIIE cells and primary PAM indicate that the balance of cytokines released in response to JP-8 could possibly be regulated in vivo by cross-communication between the two cell types. Espinoza et al. (2005) showed that the JP-8-induced expression of proinflammatory cytokine genes in AEII cells was mediated by the activation of PARP-1 (an enzyme coactivator of NF-κB) and NF-κB (a transcription factor that controls the expression of a variety of genes involved in inflammatory responses). The release of cytokines in the lung in response to JP-8 is similar to cytokine release from epidermal cells, which is thought to mediate, at least in part, the dermal and immune toxicities of JP-8 (see below). Further insight into the pulmonary effects of jet fuel has come from examinations of aerosolized JP-8’s effect on gene and protein expression in lung cytosol and lung tissue of rats and mice (Witzmann et al., 1999; Drake et al., 2003; Espinoza et al., 2005).

Immune Effects

The immune system appears to be as susceptible to jet fuel as the lung, if not more so. Detrimental effects on the immune system of mice have been reported for aerosolized Jet A, JP-8, and JP-8+100 (Harris et al., 2001a). Some effects were apparent just one hour after a single inhalation exposure with continued deterioration with each successive exposure (Harris et al., 2002). As first reported by Harris et al. (1997a), mice exposed nose-only to a JP-8 aerosol/vapor mix exhibited decreased spleen and thymus weights and cellularities and an altered number of viable immune cells in lymph nodes, bone marrow, and peripheral blood. Depending on the immune tissue examined, different immune cell subpopulations were lost, including T and B cells and macrophages. In addition, JP-8 affected immune function as demonstrated by a concentration-dependent suppression of T-cell proliferation on stimulation with the mitogen concanavalin A. In a short-term exposure study, Harris et al. (1997b) determined that JP-8-induced immunotoxicity persisted for at least one month after insult. These same authors later expanded the number of immune parameters examined and reported that aerosolized JP-8 exposure of mice nearly completely ablated NK cell function, suppressed the generation of lymphokine-activated killer cell activity, suppressed the generation of cytotoxic T lymphocytes from precursor T cells, and inhibited helper T-cell activity (Harris et al., 2000).

Ullrich (1999) has demonstrated that dermal application of JP-8 to mice can induce immune suppression. Ullrich found IL-10, a cytokine with potent immunosuppressive activity, in the serum of JP-8 dermally treated mice. He interpreted this as suggestive of an immune-suppressive mechanism involving the upregulation of cytokine release. In a follow-up study, Ullrich and Lyons (2000) demonstrated that the immunosuppressive effect of dermally applied JP-8 appears to be specific to cell-mediated immune reactions (ie, T-helper cell-driven cell-mediated immunity), as JP-8 had no effect on antibody production in immunized mice. Further, their study again implicated the release of cytokines from epidermal cells in immunosuppression, particularly prostaglandin E2 (PGE2) and IL-10. Ramos et al. (2004) have subsequently used platelet-activating factor (PAF) receptor antagonists to show that the PAF receptor, a signaling phospholipid which upregulates PGE2 synthesis by keratinocytes, plays a critical role in jet fuel–induced immune suppression. Ramos et al. (2002) have also demonstrated that the dermal application of JP-8 and Jet A suppresses delayed-type hypersensitivity and immunologic memory on rechallenge with a fungal pathogen, suggesting that jet fuel exposure may depress the protective effect of prior vaccination. Similar to that seen for pulmonary toxicity, aerosolized SP can both prevent and reverse some facets of JP-8-induced immunotoxicity, suggesting a key mechanistic role for the neuropeptide (Harris et al., 1997c, 2001b). As reported for AIIE cells, JP-8 also induces apoptosis in primary mouse thymocytes (Stoica et al., 2001). Exposure of mice to JP-8 in utero reportedly has implications for the immune system (Keil et al., 2003).

Dermal Effects

Owing to reports of severe contact dermatitis among military personnel, the dermal toxicity of jet fuel has been the subject of intensive investigation. A recent examination of dermal exposure in 124 US Air Force fuel cell maintenance workers, using a noninvasive tape-strip technique and naphthalene as a surrogate, confirmed that the skin provides a significant exposure route for JP-8 (Chao et al., 2005). Dermal absorption and penetration of JP-8 and its component hydrocarbons have been examined in vitro using dermatomed rat skin and static diffusion cells and in vivo in weanling pigs (McDougal et al., 2000; Singh et al., 2003). Their cutaneous toxicity has been explored in pigs, rats, rabbits, and in vitro using human epidermal keratinocytes (HEK) (Kabbur et al., 2001; Monteiro-Riviere et al., 2001; Chou et al., 2003; Singh and Singh, 2004). Dermal exposure to jet fuel can lead to skin irritation and sensitization and the disruption of skin barrier function. Research implicates cytokine release, oxidative stress, and DNA damage/fragmentation as mechanistic underpinnings (Allen et al., 2000; Rogers et al., 2001; Gallucci et al., 2004). Chronic dermal application of middle distillate fuels such as jet fuel can be weakly carcinogenic, although it has been effectively argued that such tumorigenicity is secondary to chronic irritation (Nessel, 1999; Nessel et al., 1999).

The ultrastructural analysis of pig skin exposed to cotton fabric saturated with jet fuel suggests that the primary effect of exposure is damage of the stratum corneum barrier (Monteiro-Riviere et al., 2004). This same publication reported that IL-8 release from HEK after JP-8 exposure was decreased by SP, which is an agonist for the NK₁ receptor present in keratinocytes and mechanistically linked to IL release. The attenuation of IL release in keratinocytes by SP and the protection SP affords against pulmonary and immunotoxicity suggests that there may be a common mechanistic linkage to these toxicities. However, studies suggest that jet fuel–induced cell death in skin is via necrosis, not apoptosis as observed in some other cell types such as AIIE cells and T lymphocytes (Stoica et al., 2001). This was also demonstrated by Rosenthal et al. (2001), who observed JP-8-induced necrotic rather than apoptotic cell death in mouse skin fibroblasts and HEK in culture or grafted onto nude mice. These authors used immunoblot analysis to determine that necrosis of HEK appeared to be associated with the perturbation of the ratio between antisurvival and prosurvival members of the Bcl-2 family of proteins (ie, the toxic, higher level of JP-8 decreased levels of the prosurvival proteins Bcl-2 and Bcl-x_L, while simultaneously elevating levels of the antisurvival proteins Bad and Bak). This has led to the suggestion that high intrinsic levels of Bcl-2 and Bcl-x_L may prevent apoptotic death of keratinocytes at low concentrations of JP-8, whereas modulation of Bcl-2 family members by high doses may lead to necrotic cell death.

As the skin allows for the selective absorption and penetration of various jet fuel components, one cannot assume that the internal or target tissue dose of chemical is qualitatively or quantitatively the same as that of the external dose (McDougal and Robinson, 2002). This has led to efforts to identify the offending components of jet fuel mixtures. Using static diffusion cells, McDougal et al. (2000) identified 12 components of JP-8 that fully penetrated rat skin and six components, all aliphatic hydrocarbons, which were partially absorbed or retained by the skin. This led these researchers to speculate that the aliphatic components of jet fuel may be the cause of skin irritation. Allen et al. (2001) exposed HEK to micromolar concentrations of four aliphatic hydrocarbon components of jet fuel and found that they all induced IL-8 release at subtoxic doses, further implicating aliphatic components. However, Chou et al. (2002) exposed HEK to 10 aliphatic jet fuel hydrocarbons (C6-C16) and found that the higher cytotoxicity of the shorter chain aliphatics did not correlate with their ability to induce IL-8 release, which peaked at midchain lengths (ie, C9-C13). The toxicological interactions of jet fuel’s aromatic and aliphatic components on HEK cells have also been explored (Yang et al., 2006). The cytotoxicity and IL-8 release from these “mixed” hydrocarbon exposures were not always predictable based on the cytotoxic and IL-8 profiles of individual components. Muhammad et al. (2005) have exposed pigs topically to cotton fabric soaked with one of eight aliphatic hydrocarbons or one of six aromatic hydrocarbons and monitored skin irritation. Based on these data, coupled with data on IL-8 release, in vitro absorption, and cytotoxicity, they postulate that tridecane and tetradecane may be the two most important hydrocarbons responsible for jet fuel–induced skin irritation. Jet fuel exemplifies the difficulty in accurately predicting the dermal risk posed by a complex mixture based on limited knowledge of a few individual components. Progress toward this end has recently been made, as Kim et al. (2006) have published a dermatotoxicokinetic model of the skin that quantitatively characterizes the TK of three aromatic and three aliphatic jet fuel components following application of a single dose of JP-8 to the forearm of human volunteers. As in lung tissue and lung cytosol of rodents, the capacity of JP-8 to alter gene or protein expression in HEK has been thoroughly examined (Espinoza et al., 2004; Witzmann et al., 2005; Chou et al., 2006).

The toxicity of CS₂ was first recognized during the 19th century, when it was widely employed as a solvent to soften rubber. CS₂ is listed as a Clean Air Act Chemical because of reported high emissions and the potential for human exposure. The major uses of CS₂ are in the production of rayon fiber, cellophane, and CCl₄ and as a solubilizer for waxes and oils (ATSDR, 1996a). Historically, exposures were particularly high during the early period of rayon production, and studies of these workers have been very informative as to CS₂’s toxicity. Human exposure is predominantly occupational, although CS₂ has been identified in at least 200 current or former EPA National Priority List hazardous waste sites (ATSDR, 1996a). Most industrial releases are to the atmosphere. The general public may be subjected to low vapor levels as demonstrated by detection of CS₂ in samples of breath and indoor and outdoor air surveyed in and around New York City (Phillips, 1992). Exposure to dithiocarbamate pesticides and drugs (eg, the alcohol aversion drug, disulfiram) can result in indirect exposure to CS₂, as it is a product of their metabolism. However, there is evidence that the metabolic production of CS₂ is not a unifying explanation for the neuropathies frequently associated with dithiocarbamate exposure (Tonkin et al., 2000; Mulkey, 2001).

The relative contributions of parent compound and metabolites to most CS₂-induced toxicities are unknown. Two distinct metabolic pathways for CS₂ exist: (1) the direct interaction of CS₂ with free amine and sulfhydryl groups of amino acids and polypeptides to form dithiocarbamates and trithiocarbonates; and (2) microsomal metabolism of CS₂ to reactive sulfur intermediates capable of covalently binding tissue macromolecules (Graham et al., 1995) (Fig. 24-13). The conjugation of CS₂ with sulfhydryls of cysteine or GSH results in the formation of 2-thiothiazolidine-4-carboxylic acid (TTCA), which is excreted in urine and has been frequently used as a biomarker of CS₂ exposure, especially among viscose rayon workers (Riihimaki et al., 1992; Lee et al., 1995). Several limitations of TTCA as a biomarker have been noted, and covalently cross-linked erythrocyte spectrin and hemoglobin have been discussed as potential alternatives (Valentine et al., 1993, 1998). Nonetheless, the current ACGIH BEI for CS₂ is 0.5 mg of TTCA/g of creatinine for a urine sample collected at the end of a workshift (ACGIH, 2012).

A few comprehensive reviews of CS₂’s toxicity have been published (Beauchamp et al., 1983; ATSDR, 1996a; WHO, 2002; Gelbke et al., 2009). CS₂ is capable of targeting multiple organ systems including the cardiovascular system, CNS and PNS, male and female fertility, and eyes (retinal angiopathy and impairment of color vision). CS₂ toxicity requires frequent and prolonged exposures in occupational settings. The nervous and cardiovascular systems have garnered the most attention. The most common neurotoxic effect is a distal sensorimotor neuropathy that preferentially affects long axons in the PNS and CNS (particularly the ascending and descending tracks of the spinal cord and the visual pathways). Encephalopathy with motor and cognitive impairment has also been reported following chronic, low-level exposure to CS₂ (Graham et al., 1995). Several MRI studies report diffuse white matter lesions in chronically exposed workers similar to that described for “toluene leukoencephalopathy” among solvent abusers (Cho et al., 2002; Ku et al., 2003). For those particularly interested in the nervous system, the classic paper by Richter (1945) detailing his observations of chronic CS₂ poisoning in monkeys is recommended. Rosenberg (1995) described the following clinical syndromes associated with CS₂: (1) acute and chronic encephalopathy (often with prominent psychiatric manifestations); (2) polyneuropathy (both peripheral and cranial); (3) Parkinsonism; and (4) asymptomatic CNS and PNS dysfunction. Pathological changes occur in both the CNS and PNS. CNS pathology consists of neuronal degeneration throughout the cerebral hemispheres, with maximal diffuse involvement in the frontal regions. Cell loss is also noted in the globus pallidus, putamen, and cerebellar cortex, with loss of Purkinje cells. Vascular abnormalities with endothelial proliferation of arterioles may be seen, sometimes associated with focal necrosis or demyelination. PNS changes consist primarily of myelin swelling and fragmentation and large focal axonal swellings, characteristic of distal axonopathy.

Significant contributions to the understanding of CS₂’s neurotoxicity have come from collaborative research by the NIEHS, EPA, and several universities (Harry et al., 1998; Sills et al., 2005). Studies were conducted with F344 rats subchronically exposed to a range of CS₂ concentrations, in order to define the onset and temporal progression of neurotoxicity as manifest by multiple end points. These specific end points included TK changes in blood CS₂ levels and urinary TTCA, covalent cross-linking of blood and spinal cord proteins, alterations in axon/Schwann cell interactions as indicated by nerve growth factor mRNA expression, morphology of distal axonopathy, nerve conduction velocity and action potential, and behavioral assessment using a functional observational battery. This research showed sensitive end points at the cellular level that progressed to alterations in hindlimb and forelimb function, followed by electrophysiological and morphological changes. The TK data provided useful information about internal exposure to CS₂, but their use in the prediction of biological effects was limited (Moorman et al., 1998). The collaborative efforts support the theory that the axonal degeneration that underlies CS₂’s central–peripheral neuropathy results from the reaction of CS₂ (and perhaps carbonyl sulfide [COS]) with protein amino groups to yield initial adducts (dithiocarbamate derivatives). The adducts decompose to an electrophile (isothiocyanate for CS₂ and isocyanate for COS), which in turn reacts with protein nucleophiles on neurofilaments to cause covalent protein cross-linking. Progressive cross-linking of neurofilaments occurs during neurofilament transport along the axon, and covalently cross-linked masses of neurofilaments are thought to occlude axonal transport at the nodes of Ranvier, ultimately resulting in axonal swelling and degeneration (similar to that seen with 2,5-hexanedione). It should be noted that several other mechanisms for the disruption of neurofilament transport that underlie CS₂’s axonopathy have been proposed including impaired energy metabolism, metal ion chelation by CS₂’s dithiocarbamate derivatives, induction of vitamin deficiency, and disruption of cytoskeletal protein association by the increased phosphorylation of neurofilaments (Wilmarth et al., 1993; Graham et al., 1995).

One of the potentially more important outcomes of the collaborative research effort described above was the identification of thiourea cross-linking structures on erythrocyte spectrin and hemoglobin. This cross-linking exhibited a linear dose–response over the range of inhaled CS₂ concentrations examined, was detectable at subneurotoxic exposure levels, preceded axonal structural damage, and was positively correlated with neurofilament cross-linking (ie, spectrin and hemoglobin cross-linking reflects neurofilament cross-linking). These findings suggest the utility of spectrin and hemoglobin cross-linking as sensitive biomarkers of exposure and effect, and potential alternatives or supplements to TTCA, which was found to lack dose proportionality in the same studies (Valentine et al., 1997, 1998; Moorman et al., 1998).

Numerous worker studies support an association between CS₂ exposure and cardiovascular disease and related mortality. Elevated mortality from cardiovascular disease has been reported among viscose rayon workers in Finland, the United Kingdom, the United States, Scandinavia, and Poland (WHO, 2002). Viscose rayon worker studies reporting excess cardiovascular morbidity expand the range to Germany (Drexler et al., 1996) and Japan (Takebayashi et al., 2004). Further support for a CS₂–cardiovascular disease link comes from examination of cardiovascular disease risk factors in workers, and to a lesser extent experimental rodent studies. In workers, CS₂ exposure has been associated with elevations in blood pressure, total and LDL cholesterol, triglycerides, apolipoproteins, and lipid peroxidation in plasma, as well as reductions in HDL and antioxidant status (Vanhoorne et al., 1992; Wronska-Nofer et al., 2002; Luo et al., 2003). Wronska-Nofer (1979) conducted studies in rats supporting a role for CS₂ in the elevation of blood cholesterol, whereas Lewis et al. (1999) found that exposure to as little as 50 ppm CS₂ significantly enhanced the rate of arterial fat deposition in mice placed on a Western style, high-fat diet. Sulsky et al. (2002) have reviewed 37 studies addressing the CS₂–cardiovascular disease association and concluded that epidemiological evidence for an association was “mixed,” with an effect on total and/or LDL cholesterol being the most consistent finding but of limited magnitude and uncertain clinical significance. Tan et al. (2002) conducted a meta-analysis of 11 cohort studies on CS₂’s cardiovascular effects, which showed a small but significant correlation between CS₂ exposure and cardiovascular disease prevalence (pooled RR = 1.56; 95% CI = 1.12–2.1). Taken together, studies suggest that CS₂ has the ability to accelerate atherosclerosis. Further, some have speculated that like neurotoxicity, protein cross-linking may also be involved in CS₂’s promotion of the atherosclerotic process (Lewis et al., 1999).

Price et al. (1997) pointed out that the cardiovascular mortality excesses seen in most published studies were among workers chronically exposed to high concentrations that no longer are observed in the workplace. These authors reviewed historical exposure and mortality data in the viscose rayon industry and estimated that chronic exposures of 15 to 20 ppm would not be associated with an increased risk of mortality due to ischemic heart disease. They were also instrumental in applying the benchmark dose method toward the establishment of EPA’s RfC for CS² of 0.7 mg/m³ (~0.22 ppm), which is based on reduced maximum motor conduction velocity in the peroneal nerves of the NIOSH cohort of viscose rayon workers reported by Johnson et al. (1983) (Price et al., 1996; EPA, 2006c). For the protection of workers, OSHA has established a PEL of 20 ppm as an eight-hour TWA with an acceptable ceiling concentration of 30 ppm. The NIOSH recommended exposure limit (REL) of 1 ppm with a STEL of 10 ppm. The ACGIH TLV was reduced from 10 to 1 ppm in 2006, as an eight-hour TWA based on neurologic end points. As a result, the ACGIH has also reduced CS₂’s BEI from 5 to 0.5 mg TTCA/g creatinine in 2009. Gelbke et al. (2009) reviewed the health effects of CS₂ in viscose industry and concluded that while some uncertainties exist, available data generally support a REL of 10 ppm.