Bioavailability and Risk Assessment of Orally Ingested Polycyclic Aromatic Hydrocarbons

Abstract

Polycyclic aromatic hydrocarbons (PAHs) are a family of toxicants that are ubiquitous in the environment. These contaminants generate considerable interest, because some of them are highly carcinogenic in laboratory animals and have been implicated in breast, lung, and colon cancers in humans. These chemicals commonly enter the human body through inhalation of cigarette smoke or consumption of contaminated food. Of these two pathways, dietary intake of PAHs constitutes a major source of exposure in humans. Although many reviews and books on PAHs have been published, factors affecting the accumulation of PAHs in the diet, their absorption following ingestion, and strategies to assess risk from exposure to these hydrocarbons following ingestion have received much less attention. This review, therefore, focuses on concentrations of PAHs in widely consumed dietary ingredients along with gastrointestinal absorption rates in humans. Metabolism and bioavailability of PAHs in animal models and the processes, which influence the disposition of these chemicals, are discussed. The utilitarian value of structure and metabolism in predicting PAH toxicity and carcinogenesis is also emphasized. Finally, based on intake, disposition, and tumorigenesis data, the exposure risk to PAHs from diet, and contaminated soil is presented. This information is expected to provide a framework for refinements in risk assessment of PAHs from a multimedia exposure perspective.

Polycyclic aromatic hydrocarbons (PAHs) are a family of ubiquitous environmental contaminants that consist of more than 100 chemicals. These chemicals are released into the environment during volcanic eruptions, forest fires, burning of coal, wood, municipal refuse, expulsion of fumes from manufacturing industries such as coke, aluminum, graphite-electrode, carbon-electrode, and petroleum, and life style/domestic activities such as smoking, incense, candle, mosquito coil burning, and cooking. These contaminants accumulate to toxic levels in the body within a short period of time (IARC 1983; ATSDR 1995; Baird and Ralston 1997; IPCS 1998).

Some PAHs have been implicated as causative agents of lung (Smith et al. 2000), breast (Li et al. 2002), esophageal (Ward et al. 1997; Roth et al. 1998, 2001), pancreatic (Z’graggen et al. 2001), gastric (Ward et al. 1997), colorectal (Sinha et al. 1999; Wiese, Thompson, and Kadlubar 2001), bladder (Boffetta, Jourenkova, and Gustavsson 1997), skin (Boffetta, Jourenkova, and Gustavsson 1997; Elmets et al. 2001), prostate (Kizu et al. 2003) and cervical (Wu et al. 2004) cancers in humans and animal models.

Aside from carcinogenicity, PAHs have also been reported to cause hemato- (Romero et al. 1997; Knuckles, Inyang, and Ramesh 2001, 2004), cardio- (Miller and Ramos 2001), renal (Parrish et al. 2002), neuro- (Saunders, Ramesh, and Shockley 2002; Wormley, Ramesh, and Hood 2004), immuno- (Mounho, Davila, and Burchiel 1997), reproductive (Sram et al. 1999; Inyang et al. 2003), and developmental (Perera et al. 1998; Hood et al. 2000; Archibong et al. 2002; Wu et al. 2003) toxicities in humans and laboratory animals.

In view of the pluripotential of PAHs to cause adverse health effects to humans, studies on the bioavailability of these contaminants from various entry routes and the assessment of risk from exposure are important not only from the standpoint of scientific research, but also from the standpoint of policy making and formulating regulations to reduce the exposure to these chemicals. This review focuses on the dynamics of PAHs consumed through the diet and methods to assess the extent of risk posed by exposure to these common environmental toxicants.

STRUCTURE, BIOTRANSFORMATION, TOXICITY, AND CARCINOGENESIS OF PAHs

Structure

Figure 1 shows the structural diversity of PAHs. Polycyclic aromatic hydrocarbons have been categorized into two groups, the peri- and cata-condensed. Peri-condensed PAHs can be defined as those whose lines connect the ring centers, and form cycles. Peri-condensed PAHs can be subdivided into two classes: alternants, which are formed exclusively by six-membered rings, and nonalternants that include some five-membered rings. Cata-condensed PAHs can be defined as those systems whose lines do not form cycles, and can be classified as branched or not branched; the former are thermodynamically more stable and chemically less reactive than their nonbranched counterparts of the same size. Cata-condensed PAHs are always alternant systems.

The structure of PAHs (some regions and carbon atom positions) determine their biological activity. The structure of benz-[a]anthracene, a representative PAH molecule detailing the various regions is depicted in Figure 2. The ‘K’ region is defined as the external corner of a phenanthrene moiety; the ‘L’ region consists of a pair of opposed open anthracenic point atoms; the ‘bay’ region is defined as an open inner corner of a phenanthrene moiety; the distal bay region also known as the ‘M’ region; and the peri position, that corresponds to the carbon atom opposite the bay region and adjacent to the angular ring.

Biotransformation

PAHs undergo metabolic transformation in the organism, resulting in polar products that are destined for excretion or in reactive metabolites that can form covalent adducts with DNA. Given that chemical:DNA adduct formation is considered the initiation event in the three-stage model of chemical carcinogenesis, understanding the formation of reactive PAH metabolites is critical to understanding the carcinogenicity of PAHs. As an example of the metabolic transformations of PAHs, the metabolic pathway of benzo[a]pyrene, B[a]P (including both activation and detoxification routes) is shown in Figure 3. The PAH biotransformation process begins with a cytochrome P450 (CYP)-mediated epoxidation of the molecule (Figure 3). The initial epoxidation is catalyzed by an enzyme complex called mixed-function oxidase (MFO), which is located in the endoplasmic reticulum. The second step involves hydroxylation with the formation of diols. This is catalyzed by epoxide hydrolase (EH), which is closely linked to the MFO enzyme complex. The enzyme complex including the hydrolase is often referred to as an aryl-hydrocarbon hydroxylase (AHH). The diols formed can be converted further into dihydrodiol epoxides. From a chemical and biological point of view, the dihydrodiol epoxides (especially those formed in the bay region such as benzo[a]pyrene diol epoxide, BPDE) are very reactive to nucleophilic attack by nucleophilic sites in DNA, either directly in an S_N2 reaction or, (after forming a carbo-nium ion) in an S_N1 reaction (Guillén, Sopelana, and Partearroyo 1997). Oxidative metabolism of diols such as B[a]P 7,8-diol can also be catalyzed by prostaglandin H synthase (Marnett, Reed, and Dennison 1978; Eling and Curtis 1992), a myeloperoxidase system (Mallet, Mosebrook, and Trush 1991), lipoxygenases (Hughes et al. 1989), or cyclooxygenase-2 (Wiese, Thompson, and Kadlubar 2001). These reactions are of importance in situations in which there are relatively low levels of CYP or when chronic irritation and/or inflammation occurs (Kensler et al. 1987; Ji and Marnett 1992). Nevertheless, the intermediate diols can also undergo a detoxification process by conjugating with glucuronic acid or glutathione, leading to conjugated metabolites, which can be excreted by renal or biliary routes. The steps leading to the formation of trans-dihydrodiols (diols) is called phase I metabolism, and further reactions of the phase I metabolites are referred to as phase II metabolism. Metabolites of PAHs with two and three rings are excreted preferentially in the urine, whereas higher-molecular-weight metabolites are excreted in the feces.

Liver is the major organ for PAH metabolism. However, other organs may play a greater role depending on the site of PAH entry. In the case of ingestion, gut micro flora and intestinal cytochrome P450 enzymes can contribute to PAH metabolism. Several cytochrome P450 enzymes involved in the metabolism of PAH compounds are given below.

CYP1A1: This enzyme is capable of metabolizing a wide spectrum of PAHs. The constitutive expression of this enzyme is low in tissues (Guengerich and Shimada 1991). The induction of CYP1A1 is controlled by the Ah (aryl-hydrocarbon) receptor, a transcription factor activated by ligands such as PAHs. This indicates that individual PAH compounds or mixtures can regulate their own metabolism by inducing CYP1A1. After induction, the level of CYP1A1 expression is high in extrahepatic tissues (Moorthy 2000; Wu et al. 2003), but the level of expression is low in human liver, compared to rodents.

CYP1A2: The capacity of this enzyme to metabolize PAHs is less than that of CYP1A1. Nevertheless, this enzyme has been reported to metabolize B[a]P to 3-hydroxy B[a]P and B[a]P 7,8-dihydrodiol to its epoxides in rodents (Shou, Wells, and Elkind 1994) and humans (Bauer et al. 1995). CYP1A2 was reported to be induced by B[a]P, benzo[b]-fluoranthene, benzo[k]fluoranthene, benzo[a]anthracene, and dibenzo[a,h]anthracene (Vakharia et al. 2001).

CYP1B1: This enzyme has been reported to act on a number of PAHs (Shen, Wells, and Elkind 1994) and expressed in rodent (Savas et al. 1994; Bhattacharya et al. 1995; Eltom, Zhang, and Jefcoate 1999; Moorthy et al. 2002), goat (Eltom and Schwark 1999), and human (Sutter et al. 1994: Larsen et al. 1998; Shimada 1999) tissues.

CYP2B: This enzyme has been reported to be involved in the metabolism of 7,12-dimethylbenz[a]anthracene (DMBA1; Morrison, Burnett, and Craft 1991), and 6-aminochrysene (Yamazaki et al. 1993). A recombinant version of this enzyme was reported to metabolize B[a]P in human cell lines (Shou, Wells, and Elkind 1994). The constitutive levels of this enzyme are low in human liver, but when induced can metabolize PAHs considerably (Hall et al. 1989).

CYP2C: This enzyme is abundant in human liver and several members of this family are capable of metabolizing PAHs in human liver (Yun, Shimada, and Guengerich 1992), and other tissues (Fisslthaler et al. 1999; Riddick et al. 2003).

CYP3A: This enzyme is also abundant in human liver and can metabolize B[a]P (Shimada et al. 1989; Yun, Shimada, and Guengerich 1992), and 6-nitrochrysene (Chae et al. 1993) extensively.

The intestinal epithelium contains all the enzymes, which have been identified as being involved in activation and detoxification of PAHs, although these activities are generally much lower than in the liver (Benford and Bridges 1985). The low levels of inducible CYP isozymes in the intestinal tract could influence the occasional development of tumors in the small and large intestines as a consequence of the ingestion of PAH-containing food (Stavric and Klassen 1994). Colon tissues have been reported to express CYP enzymes though at a low level (reviewed in Ding and Kaminsky 2003). Also, colon tissues have been shown to express prostaglandin H synthase (COX; Kargman et al. 1995). In this context, it is interesting to note that B[a]P and its metabolites has been reported not only to induce the expression of COX-2, but also function as substrates for COX-2 (Kelley et al. 1997). Human COX-1 and COX-2 have been reported to activate B[a]P 7,8-diol to intracellular electrophiles (Wiese, Thompson, and Kadlubar 2001). Thus, the activation of B[a]P/its metabolites by COX isozymes is relevant to colorectal carcinogenesis as this region of the gastrointestinal (GI) tract receive direct exposure to PAHs through diet.

Polycyclic aromatic hydrocarbons also induce the flavin-containing monooxygenases (FMOs), another superfamily of metabolizing enzymes that oxidize numerous nucleophilic compounds and drugs (Hines et al. 1994). Because FMO resides in the same organs and requires the same cofactors as the cytochrome P450 monooxygenases (NADPH and oxygen; Ziegler 1993), PAHs are expected to be good substrates for FMO. Chung et al. (1997) reported induction of FMO1 (an isoform of MFO) in rat liver by 3-methylcholanthrene. In a recent study, distinct cell-, tissue-, sex-, and developmental stage–specific patterns of MFO expression in mice was reported (Janmohamed et al. 2004). The differential expression of the MFO suggests a specific role in chemical defense. Because the study of Janmohamed et al. (2004) is not aimed at looking the metabolism of xenobiotics by MFOs, the toxicological significance of different FMO isoforms is unclear. The ability to bioactivate PAHs by MFO to mutagenic metabolites is not established in mammals. Hence it is not yet known whether organs capable of such activation are at an increased or decreased risk to PAH-induced genotoxic responses such as cancer.

Notwithstanding the organ-specific variation in metabolizing enzymes, the resulting biological activity of ingested PAHs is determined notably by their degree of absorption and overall metabolism as well as the presence of compounds that can act as inducers, promoters, or inhibitors of the PAH metabolism by acting on enzymatic factors. Thus, drugs, certain vegetables, other environmental pollutants such as polychlorinated biphenyls, and gastric hormones (Benford and Bridges 1985) induce the activity of intestinal enzymes that metabolize PAHs into ultimate carcinogens.

The BPDE reacts with DNA and form adducts. The (+)- and (−)-anti-BPDE reacts with DNA. The trans addition of the exocyclic amino group of guanine (N ²-dG) to the C10 position of the PAH gives rise to major adducts. Additionally, the cis addition at N ²-dG and from cis and trans addition at N ⁷-dG and the exocyclic amino group of adenine (N ⁶-dA) gives rise to minor adducts (Cheng et al. 1989). The diol epoxides of all bay region PAHs do not always specifically bind to guanine as BPDE. For example, benzo[c]phenanthrene diol epoxides bind equally to adenine and guanine (Dipple et al. 1987). The biological significance of minor adducts is not fully established. The formation of DNA adducts or premutagenic lesions leads either to an altered function of the gene product or to disturbance of the normal regulation of the expression of that product. The processing of a lesion by polymerases and repair enzymes may repair the lesion or fix it as a permanent alteration of the DNA sequence. The altered DNA sequence leads to alterations in heritable information or mutation (Josephy 1997; Hanawalt, Fort, and Lloyd 2003). If the mutation occurs in an active region of the genome, i.e., if it acts on a proto-oncogene (abnormal activation or overexpression) or a tumor suppressor gene (inactivation), this alteration has the potential to initiate carcinogenesis. Among PAHs, B[a]P has been widely studied with regard to the mechanism of carcinogenesis. B[a]P was reported to induce G → T transversions in specific codons (the 12th) of the ras family of proto-oncogenes, which may convert the gene into an active oncogene (Boelsterli 2003). These DNA adducts have been used as biomarkers of exposure to PAHs in susceptible populations (Beach and Gupta 1992; Talaska, Roh, and Getek 1992; Brandt and Watson 2003; Farmer et al. 2003).

Though B[a]P has been widely studied as a prototypical representative of PAHs with regard to its metabolism and toxicity, exceptions to the bay-region diol epoxide activation pathway with some PAHs does occur. For example, some PAHs such as fluoranthene, benzo[j]fluoranthene, and benzo[ghi]perylene that lack a bay region are metabolically activated to vicinal diol epoxides, with the epoxide function in a hindered position (Phillips and Grover 1994).

Mechanisms of Carcinogenesis

Harvey (1996) summarized four mechanisms for carcinogenesis of PAHs: the diol epoxide mechanism, the radical-cation mechanism, the quinone mechanism, and the benzylic oxidation mechanism. The diol epoxide mechanism involves metabolic activation by CYP enzymes to reactive epoxide and diol epoxide intermediates that interact with DNA, leading to mutations and ultimately to cancer (Harvey 1996). This mechanism is widely accepted as an important pathway. The radical-cation mechanism involves one electron oxidation to generate radical-cation intermediates that may attack DNA resulting in depurination (Cavalieri and Rogan 1995). The quinone mechanism involves enzymatic dehydrogenation of dihydrodiol metabolites to yield quinone intermediates that may combine directly with DNA or enter into a redox cycle with O₂ to generate reactive oxygen species, such as hydroxyl radicals and superoxide anion, capable of attacking DNA. Formation of small amounts of quinone metabolites may result in generation of high ratios of reactive oxygen species (Flowers-Geary, Harvey, and Penning 1993; Penning et al. 1996).

Carcinogenicity Predictions

In recent years, investigations have used the structural features of PAHs to predict their toxicity and carcinogenicity. Klopman, Tu, and Fan (1999) used META, a metabolic transformation software system (Klopman and Tu 1997), to predict the metabolism and understanding the carcinogenic potential of PAHs. The system is based on the postulation that for an enzyme catalyzed site-specific metabolism to occur, two criteria have to be met: sufficient reactivity and ready accessibility. The authors used a quantum index (QI) to express the reactivity of a site (atom or bond) in a target molecule, and a graph index (GI) to express the structural features of a molecule respectively. Another parameter, the lipophilicity (expressed as the logarithm of a PAH compound’s octanol/water partition co-efficient [Kow]; log P) affects the reactivity and accessibility. Because one of the most important steps in the mechanism of oxidation of PAHs by CYP is the approach of the PAH to the lipophilic region of CYP, log P was incorporated in the modeling. Using these three parameters, the P value (an integer) was calculated to determine the reactivity of a PAH for transformation reactions. The program was designed in such a way that if several metabolic pathways exist for a single molecule, then the program chooses the reaction with the highest reactivity (lowest P value) as the major biotransformation pathway. These authors tested this approach for 42 PAHs and compared it with the experimental data. META predicted epoxidation sites with good accuracy more so than the hydroxylation sites. These variations could be minimized by employing advanced quantum mechanical calculations to predict the QI and GI accurately (Klopman, Tu, and Fan 1999). Because META is partly based on using the Kow, its usefulness for carcinogenicity prediction of PAHs with differing hydroxylation properties is not yet validated.

Recent work from He and Jurs (2003) provides information on predicting the genotoxicity of PAHs using the topological, geometrical, electronic, and polar surface area properties. Compounds were tested for their ability to cause DNA damage using SOS Chromotest in the presence or absence of S9 rat liver homogenate. This assay is an early monitor to measure the induction of a lacZ reporter gene of Escherichia coli in response to DNA damage (Hofnung and Quillardet 1988). Parameters such as nearest neighbor, linear discriminant analysis, and probabilistic neural network were used to develop a consensus model that correctly predicted the genotoxicity of 81.2% of the 277 PAH compounds tested.

Another recent study focused on modeling the metabolic pathways (Conti et al. 2003). As mentioned elsewhere, different enzymes are involved in the metabolism of PAHs. In this study, two methods have been proposed to model the complex sequence of biochemical reactions, one based on Bayesian model averaging, and the other based on pharmacokinetic modeling. These two approaches were tested utilizing data from a case-control study of colorectal polyps in relation to consumption of well-cooked red meat, a rich source of PAHs. Both of these approaches focused on conversion of PAHs to DNA-adducting metabolites by distinct biochemical routes with several enzymes involved in the pathway such as CYP1A1, epoxide hydrolase, and glutathione S-transferase. This information is incorporated into various categories such as exposure quantities, genotypes, metabolic phenotypes, rates of metabolic activation, and detox-ification in the model. After fitting the data to these models, the authors (Conti et al. 2003) found that the Bayes model averaging approach, which is less parameterized, is more promising than the pharmacokinetic approach, which is highly model dependent. There is scope for improvement of these approaches. For example, additional biological information such as enzyme kinetics could be incorporated in the Bayes-model approach. Similarly, averaging of parameters over several alternative model specifications help avoid the uncertainty in pharmacokinetic modeling.

Because PAHs are widely distributed environmental and dietary contaminants, many of which have not been investigated with regard to their potential to cause toxicity and cancer, development and refinement of algorithms like META will help in categorizing the chemicals on the basis of their biotransformation (epoxidation or hydroxylation) properties.

Mechanisms of Toxicity

PAHs were repoted to cause a wide range of toxicities encompassing different organ systems. PAHs are cytotoxic at high doses, producing lymphoid atrophy, whereas at low doses, cause immunotoxicity without cytotoxicity (Davila, Mounho, and Burchiel 1997). Exposure of lymphoid cells to PAHs may suppress B-cell response to both T cell–dependent and T cell–independent antigens and B-cell mitogens. PAHs also may affect T cell–mediated immune function and macrophage function. Furthermore, they affect innate immune responses such as phagocytosis and interferon production (reviewed in Davila, Mounho, and Burchiel 1997).

Studies have revealed that a dose of 50 to 100 mg/kg body weight is required to elicit significant immunosuppression or cytotoxicity following oral or subcutaneous exposure to B[a]P and DMBA (Davila, Mounho, and Burchiel 1997). In this context it should be mentioned that acute and subchronic oral exposures of rats to B[a]P and fluoranthene (FLA) affected white blood cell counts in a dose-dependent manner, with a maximum decrease of 45% compared to the controls (Knuckles, Inyang, and Ramesh 2001, Knuckles, Inyang, and Ramesh 2004). The doses used in these studies were high (5, 50, 100 mg/kg body weight/day for B[a]P and 150, 750, 1500 mg/kg body weight/day for FLA). Similarly, De Jong et al. (1999) observed immunotoxicity of B[a]P (at doses 3, 10, 30, and 90 mg/kg bodyweight/day) in a subacute oral toxicity study . The reduction in white blood cell (WBC) counts reported in the aboveare consistent with immunosuppression associated with PAH exposure.

Other organs to which PAHs are toxic include kidneys and brain. Tubular casts (molds of distal nephron lumen) in the kidneys were reported to be caused by subchronic oral exposure to B[a]P and FLA (Knuckles, Inyang, and Ramesh 2001, Knuckles, Inyang, and Ramesh 2004). A wide range of neurobehavioral toxicities caused by acute oral exposure to B[a]P and FLA were reported by Saunders, Ramesh, and Shockley (2002).

The toxicities caused by PAHs are due to interference with cellular signal transduction pathways within the cell (Romero et al. 1997). Ah receptor (Ahr) signaling plays an important role in regulation and induction of several drug metabolizing enzymes such as CYP1A1, CYP1A2, CYP1B1, glutathione S-transferase, UDP-glucoronyltransferase, quinone oxidoreductase, etc. (Hankinson 1995; Rowlands and Gustafasson 1997; Ramesh et al. 2000; Nebert et al. 2004). These enzymes process toxicants to reactive metabolites that interact with cellular macromolecules contributing to toxicity or carcinogenesis. Using Ahr (+/+) and Ahr (−/−) mice, Shimada et al. (2002) showed induction of CYP1A1, CYP1A2, and CYP1B1 in Ahr (+/+) mice. Among the PAHs, B[a]P, 7,12-DMBA, dibenz[a,1]-pyrene, 3-methylcholanthene, 1,2,5,6-dibenzanthracene, benzo-[b]-fluoranthene, and benzo[a]anthracene strongly induced CYP1A1 and CYP1B1. The PAH compounds 6-aminochrysene, chrysene, benzo[e]pyrene, and 1-nitropyrene weakly induced CYP1A1 and CYP1B1. Anthracene, pyrene, and fluoranthene were weak inducers. The induction of CYP1A2 was less than CYP1A1 and CYP1B1. The studies of Shimada et al. (2002) support the view that PAH toxicity occurs by Ahr-dependent mechanisms.

Dertinger et al. (2000) conducted experiments to see whether B[a]P toxicity occurs through Ahr-independent mechanisms. The authors focused on blocking the signal transduction occurring through Ahr using a synthetic flavone derivative. They challenged wild-type and Ahr null allele mice with B[a]P and without a flavone cotreatment. In mice that received isoflavone and B[a]P, the genotoxicity of B[a]P was significantly altered. These studies also reported the inhibition of basal and induced CYP1A1/2 activities by isoflavone. The protection of Ahr null allele mice from B[a]P toxicity by isoflavone treatment suggest the involvement of Ahr-independent mechanisms.

Guigal et al. (2000, 2001) reported induction of CYP1A1 by transcriptional activation of serum, independent of Ahr pathway. The authors found an increase in CYP1A1 mRNA levels in serum when CaCo-2 cells were treated with fetal bovine serum. The increase was similar to that observed after 3-methylchol-anthene induction. Furthermore, experiments done using heteronuclear RNA on CaCo-2 cells revealed that CYP1A1 mRNA induction by serum was due to transcriptional activation.

Delescluse et al. (2000) outlined several signaling pathways related to CYP1A1 induction by various chemicals. The relevance of some of these pathways to PAHs is not clear and hence they are not mentioned in detail here. Protein kinase C (PKC) activity is required for Ahr-mediated signal transduction. Hence PKC inhibitors may inhibit the transcription of CYP1A1 (Delescluse et al. 2000). Benzo[a]pyrene was reported to inhibit PKC activity in vascular smooth muscle cells in a concentration-dependent manner. A comparable response was seen with 3-OH B[a]P, indicating that oxidative metabolites of the parent PAH compounds are capable of modulating the PKC activity (Ou et al. 1995). Similarly, tyrosine kinase activation may influence CYP1A1 induction (Delescluse et al. 2000). PAHs were reported to activate src family-associated protein tyrosine kinases (PTKs) Lck and Fyn in T cells that initiate phospholipase Cγ1 activation, the production of 1,4,5-inositol triphosphate, and the release of Ca²⁺ from intracellular storage pools in the endoplasmic reticulum (Archuleta et al. 1993). PAHs also inhibit the reup-take of Ca²⁺ into the endoplasmic reticulum (ER) by the sarco-endoplasmic reticulum Ca²⁺ ATPases (Krieger et al. 1995). These findings indicate that alterations in Ca²⁺ homeostasis may be responsible for the immunotoxic effects of PAHs. Recent studies (Page et al. 2004) have shown that DMBA-induced bone marrow toxicity is dependent on tumor necrosis factor alpha receptor (TNFR) and double-stranded RNA–dependent protein kinase (PKR). Further, Page et al. (2004) have demonstrated that PKR is activated by TNFR-mediated signaling. The changes in intracellular calcium flux due to PAH exposure may have partly contributed to the upregulation of PKR (Page et al. 2004).

Many PAHs were reported to cause epigenetic toxicity (Upham, Weis, and Trosko 1998). Epigenetic chemicals alter the genetic phenotype of a cell is by interacting with a finite number of intracellular biochemical pathways that turn genes on and off. These intracellular pathways involve intercellular communication through gap junctions. Gap junctions are channels formed by connexin proteins that permit regulatory molecules and ions (glutathione, cAMP, Ca⁺², inositol triphosphate, etc.) to pass directly between adjacent cells. Gap-junctional intercellular communication (GJIC) has been linked to the regulation of development, cellular proliferation, differentiation, and apoptosis (Upham et al. 1997). GJIC function is regulated by the extracellular receptor kinases (ERK), a class of mitogen-activated protein kinases (MAPK; Rummel et al. 1999). Rummel et al. (1999) reported that PAHs containing bay or bay-like regions (methylated and chlorinated isomers of anthracene) initially inhibited GJIC followed by MAPK activation. These findings suggest that altered GJIC affect MAPK activation, instead of MAPK regulating GJIC. Bláha et al. (2002) reported that low-molecular-weight PAHs (such as fluorine, phenanthrene, fluoranthene, and their methylated derivatives) were strong inhibitors of GJIC. On the other hand, high molecular weight PAHs (such as B[a]P, dibenzo[a,e]pyrene, and dibenzo[a,h]pyrene were found to be weak inhibitors. The studies of Rummel et al. (1999), and Bláha et al. (2002) collectively indicate that PAH structure and lipophilicity may play a role in exerting their toxicity. Inhibition of GJIC may be an important mode of action for low molecular weight PAHs whereas for high molecular weight PAHs, other mechanisms could be involved that needs a detailed study.

Metabolism of PAHs generates reactive oxygen species that are capable of causing cell injury. Studies have shown that the B[a]P metabolites 3-OH B[a]P, 3,6-quinone and H₂O₂ activate transcription of c-Ha-ras oncogene in vascular smooth muscle cells (Miller and Ramos 2001). Benzo[a]pyrene was also reported to inhibit the expression of c-fos, c-jun, c-myc, and c-Ha-ras in cultured rat through Ahr signaling and oxidative stress (Zhao and Ramos 1998).

Toxicity Predictions

Most of the toxicity data for environmental chemicals are available for individual components. Hence risks are calculated for individual compounds. Because PAHs occur in the environment as complex mixtures of varying composition, there is a need to develop reliable estimates of toxicity for these chemicals. Developing such estimates require using toxicity data derived from experiments with the mixture of interest. PAHs are handicapped by the lack of mixture-specific toxicity data and thus necessitate the use of approximations to predict toxicity (Reeves et al. 2001). To overcome these difficulties, toxic equivalency factors (TEFs) were developed by Nisbet and LaGoy (1992) for PAHs. These factors serve as a means of ranking the toxicity of PAHs relative to B[a]P. However, difficulties in implementing the TEFs for PAHs include lack of knowledge about PAH toxicity, and scarcity of information about PAH interactions (U.S. EPA 1993). Additional deficiencies in adopting TEF approach are brought into light by the mice tumor and particulate matter mutagenicity studies. Studies by Weyand et al. (1995) have shown significant differences in numbers and sites of tumors between B[a]P- and coal tar–dosed B6C3F1 mice (discussed in detail in the section on risk assessment). Similarly, atmospheric PAH pollution surveys conducted by Matsumoto et al. (1998) revealed no correlation between mutagenicity of air particulates and B[a]P content in air samples.

To provide an approximation of toxicity of PAH mixtures, Reeves et al. (2001) conducted a series of rapid in vitro and in vivo bioassays. The authors isolated PAHs from a sample of coal tar and separated via normal-phase high-performance liquid chromatography (HPLC) into five fractions. Each fraction was tested in the Salmonella/microsome assay, the chick embryo screening test, and the gap junction intercellular communication assay and their ability to induce cytochrome P450 enzymes in hepatic cells. However, Reeves et al. (2001) observed a lack of agreement between assay-predicted potencies and chemical analysis predicted potencies. The mixture used in these studies was a complex one with more than 70% of the PAH fractions could not be identified by gas chromatography–mass spectrometry (GC/MS) analysis. Hence interaction among various components in the mixture and their effect on cell-to-cell communication may have influenced the outcome of this study (Reeves et al. 2001). Similar studies with mixtures of known PAH composition may provide a reasonably accurate outcome.

The TEF approach is based on the premise that the Ahr mediates the toxic action for PAHs. Exceptions to this approach were mentioned in the section on ‘mechanisms of toxicity’ wherein evidence was presented that PAHs exert their toxic action through an Ahr-independent pathway. Other limitations to TEF approach may involve developing consistent TEF values for a range of species/animal models. Therefore, mechanistic studies using PAH mixtures are warranted to provide a sound scientific basis for prediction of toxicity.

PAHs IN FOODS AND DIETARY EXPOSURE OF HUMANS

Food ingestion is the major route of exposure compared to inhalation for a large section of general population exposed to PAHs (Butler et al. 1993; Van Rooij et al. 1994). Studies conducted on human exposure to B[a]P revealed that the range and magnitude of dietary exposures (2 to 500 ng/day) were larger than for inhalation (10 to 50 ng/day; Lioy et al. 1988). Diet makes a substantial contribution (more than 70% in nonsmokers) to PAH intake other than occupational PAH exposure (Beckman Sundh, Thuvander, and Andersson 1998; Phillips 1999). For a nonsmoking ‘reference male’ between 19 and 50 years (on a total body basis), a mean total PAH intake of 3.12 mg/day was estimated of which dietary intake contributed 96.2%, air 1.6%, water 0.2%, and soil 0.4% (Menzie, Potoki, and Santodonato 1992). Among PAHs, FLA and B[a]P are the two compounds detected in high levels in food with FLA exceeding the levels of B[a]P (Larsson et al. 1983).

PAH contamination of food arises from two sources, environment and food-processing technique. Unprocessed food consists of vegetables, fruits, grains, vegetable oils, dairy products, and seafood. For plants (leafy vegetables and tubers), uptake through atmosphere and soil are prime sources of contamination. Broad-leaved vegetables such as lettuce have a larger surface area that is ideal for deposition of airborne particles containing PAHs (Wickstrom et al. 1986). The accumulation of PAHs in foods of animal origin, especially livestock, is due to consumption of contaminated pastures and vegetation (Crepineau et al. 2003). PAHs in fish and shellfish are a result of contamination of fresh and coastal waters. Processed food (through smoking; Yakibu, Martins, and Takahashi 1993) and cooked food (charcoal cooked; Knize et al. 1999) also contribute substantially to the intake of PAHs. The type of cooking, cooking temperature, time, amounts of fat, and oil influence the formation of PAHs (Vainiotalo and Matveinen 1993; Perez, Lopez de Carain, and Bello 2002). Drying techniques used for cereal preservation such as combustion gas heating and smoking increase the PAH concentrations (Klein, Speer, and Schmidt 1993). Grilled vegetables contain high PAH concentrations more so than raw vegetables (Tateno, Nagumo, and Suenaga 1990). Similarly, the variations in refining processes contribute to the differences in PAH concentrations in oils of plant origin (Guillén and Sopelana 2003).

Table 1 lists the concentrations of PAHs found in products of plant origin. Vegetables are considerably contaminated (Zhong and Wang 2002; Tao et al. 2004) if they are grown in soil in close proximity to highways (Camargo and Toledo 2003). Petersen et al. (2002) have experimentally demonstrated the uptake of PAHs by fruit and vegetables grown in contaminated soils. PAH levels are low in cereals and beans. Contamination of these food products is due to aerial deposition (Jones et al. 1989), which explains the occurrence of PAHs in higher concentrations in bran than flour (Dennis et al. 1991). Processing techniques used for cereal and bean preservation (Klein, Speer, and Schmidt 1993) in different countries also may contribute to their PAH levels. The residue levels of PAHs in nuts, roots, and tubers are low (Dennis et al. 1991). Oils from different sources such as virgin, refined olive oil, sunflower, soybean, maize, coconut, rapeseed, cotton, groundnut, grape seed, rice, palm, and palm kernel oils harbor considerable levels of PAHs. Contamination from PAHs during processing of seeds may exceed the air-borne contamination. Nonetheless, the high PAH concentrations are a matter of concern because ingredients such as vegetable oils and fats are common in a variety of manufactured or cooked foods.

Table 2 gives the concentrations of PAHs found in products of animal origin. Products such as milk, butter, cheese, and eggs have low levels of PAHs. Additionally cooking procedures (Chen and Chen 2003) and heat sources used for cooking (Oanh, Nghiem, and Phiu 2002) contribute to PAH contamination in foods. The levels of certain PAHs are high in barbecued or grilled beef, lamb, and pork (Larsson et al. 1983; Lodovici et al. 1995). Consumption of seafoods also contributes to considerable intake of PAHs. The residue levels of PAHs in aquatic organisms depend on contamination of their habitat and ability of these organisms to metabolize the contaminants. Sediment feeding shellfish and finfish accumulate a great amount of PAHs (Sanders 1995; Vassilaros et al. 1982; Baumard et al. 1998). Food-chain bioaccumulation of PAHs is significant for organisms at lower levels (i.e., crustaceans, mollusks, and fishes), but not for high trophic level consumers, such as humans, probably due to higher capacity of metabolism.

There are several published reports on contamination of a wide variety of foodstuffs by PAHs. It is beyond the scope of this review to categorize them. Interested readers may refer to the comprehensive reports/reviews on this subtopic by Phillips (1999), Kazerouni et al. (2001), ECSCF (2002), Guillén and Sopelana (2003).

Two studies significantly contributed to our knowledge on dietary intake of PAHs. The first comprehensive study of Kazerouni et al. (2001) reported B[a]P concentrations in more than 200 food items procured in the Washington DC area. Because B[a]P is a known carcinogen, it has been used as a surrogate by the authors to gain an understanding of the contribution of diet for the intake of carcinogenic PAHs. This database (Kazerouni et al. 2001) was linked to a Food Frequency Questionnaire (FFQ). The parts of composite food samples for each item in FFQ were determined from the National Health and Nutrition Examination Survey. The FFQ consisted of details on the frequency of consumption and portion sizes (small, medium, and large) for different food items by subjects from a disease-free population. The authors used responses from this questionnaire to estimate daily consumption (in grams) of different food items using the frequency and portion size. Benzo[a]pyrene concentrations were derived by multiplying weight of food items by their B[a]P content. The B[a]P content was totaled for all the items in the diet to estimate intake for individual subjects. About 31% of the subjects had a daily B[a]P intake from 40 to 60 ng. The bread/cereal/grain, meat, vegetables, and fruit groups contributed highly to B[a]P intake. On the contrary, fat, sweet, and dairy food groups contributed lower amounts to daily B[a]P intake.

The second study was from Catalonia, Spain, covering 108 food samples (Falcó et al. 2003). The design of this study is different from Kazerouni et al. (2001). In this study concentration in foodstuffs for 16 individual PAHs along with their daily intake in a population comprising different age groups were reported. Most subjects who participated in this study had a daily B[a]P intake of 20 ng. The total amount of PAH intake from various food categories ranged from 100 to 150 ng. Cereals, meat, and meat products contributed almost half of the total amount of PAH intake. This study found notable differences in dietary intake of PAHs for different age groups, which were attributed to their dietary habits. When the intake values were adjusted for body weight, the highest intake was for children (0.307 μg/kg/day) followed by adolescents (0.150 μg/kg/day), and senior citizens (0.102 μg/kg/day).

The dietary intake of PAHs in various countries as reported in literature is summarized in Table 3. The methodology used for extraction of PAHs from diverse matrices of food, clean up, separation, identification, and quantification of PAHs in food samples (Chen 1997; Moret and Conte 2000) decisively influences the results obtained. This aspect has to be taken into consideration when results from various studies are reconciled.

The above approaches were adopted for measuring dietary intake (combining residue data with food consumption data) of PAHs and are similar to that adopted for pesticides in food implemented by the U.S. Environmental Protection Agency (USEPA) as per the mandate of the U.S. Food Quality Protection Act of 1996 (USEPA 1993). The limitation of this approach is that it provides little information about source-to-intake relationships for dietary exposures to atmospherically emitted semivolatile organic chemicals such as PAHs. The physicochemical properties of these combustion byproducts enable them to partition from environmental media to the food chain. These characteristics were utilized by Lobscheid, Maddalena, and McKone (2004) to estimate the contribution of locally grown produce in assessing cumulative exposure to PAHs. Using a CalTOX multimedia mass-balance model (McKone, Maddalena, and Bennett 2004), these authors (Lobscheid, Maddalena, and McKone 2004) determined the variation in population-based intake fraction for PAHs among food commodities and different exposure pathways. This approach takes the food pathway into account when considering the dietary exposure of PAHs for populations that are remote from the pollutant source, given the pollutants can migrate to agricultural regions and subsequently to agricultural commodities distributed to that population. For this study Lobscheid, Maddalena, and McKone (2004) used two PAHs, B[a]P and FLA, an established carcinogen, and a nonclassifiable carcinogen (USEPA 2001) respectively. The results indicated that fruits, vegetables, and grains contributed largely to the total intake of airborne PAHs. These findings are in broad agreement with those of Kazerouni et al. (2001) in that grain and raw/leafy vegetables had relatively high B[a]P levels.

The contribution of contaminated drinking water (Rodrigues, Lacerda, and Lancas 2002) to PAH intake is relatively low compared to food. These chemicals are carried via long-range atmospheric transport to contaminate the open-air drinking water supplies. Runoff from road traffic (Ishimaru, Inouye, and Morioka 1990), oil fields (Literathy, Quinn, and Al-Rashed 2003), and former industrial sites (Boyce and Gary 2003) contaminates groundwater. Additionally, the use of tar-coated water pipes was reported to release PAHs into drinking water (Maier, Maier, and Lloyd 2002).

Controlled feeding studies have shown a strong association between consumption of PAH-rich foodstuffs and urinary excretion of their metabolites in humans (Buckley and Lioy 1992; Sithisarankul et al. 1997). Published evidence indicates a strong relationship between diet and toxicity. A relationship between intake of PAH-rich foods and cancer incidences were shown for stomach and esophagus (Ward et al. 1997), and colon and rectum (Sinha et al. 1999) in humans. Grilled and barbecued meats were reported to contain high levels of B[a]P when compared to pan-fried or boiled foods (Knize et al. 1999). They contributed to 21% of mean daily intake of B[a]P (Kazerouni et al. 2001). Pyrolysis of fats that occur during grilling of meat creates smoke that deposits relatively large amounts of PAHs onto the surface of the meat (Lijinsky 1991). Epidemiological studies revealed a positive association between consumption of red meat cooked by deep-fried method and risk of breast cancer (Dai et al. 2001). Orally administered B[a]P and 7,12-DMBA were reported to induce nuclear anomalies in the mouse digestive tract (Reddy et al. 1991; Brooks et al. 1999). Similarly, DMBA administered through a high-fat diet induced ductal pancreatic cancer in rodent models (Z’graggen et al. 2001). All these studies are consistent with a potential role for dietary PAHs in gastric/colorectal carcinogenesis.

The 1-hydroxypyrene (1-OHP; a metabolite of pyrene) concentrations have been used as a biological marker of global exposure to environmental PAHs (Bouchard et al. 2001; Sithisarankul et al. 1997; Scherer et al. 2000). Controlled studies using human volunteers have shown a strong association between urinary excretion of 1-OHP and consumption of PAH contaminated diet (Buckley and Lioy 1992; Kang et al. 1995; Viau et al. 1995). On the other hand, such an association was not evident in some studies (Roggi et al. 1997; Scherer et al. 2000; Viau et al. 2002). This could be attributed to interindividual differences in the bioavailability of ingested pyrene, extent of pyrene transformation to 1-OHP, and polymorphism of drug-metabolizing enzymes (Viau et al. 2002). Pyrene and fluoranthene together account for half of the total PAH levels measured in diet (Phillips 1999). Also, a significant correlation exists between pyrene and B[a]P in diet (Viau et al. 2002), suggesting that pyrene could be used as a surrogate to estimate dietary exposure to PAHs (Viau et al. 2002).

BIOLOGICAL FATE OF PAHs TRANSFERRED THROUGH FOOD

Transfer Pathways of PAHs Through Diet

A limited number of studies have been conducted on the transfer pathways of PAHs from food to animals. Laurent et al. (2001) used [¹⁴C]phenanthrene and [¹⁴C]B[a]P to study the transfer of these chemicals to pigs through milk. Milk spiked with these chemicals was fed to the animals. Portal and arterial blood samples were collected to study the kinetics of PAH transfer. Peak absorption in the blood occurred 6 h after ingestion of B[a]P and 5 h for phenanthrene. The absence of a time shift in absorption for these compounds indicate that PAHs and milk fat were absorbed during the same time period. Absorption rates were high for phenanthrene (95%) compared to B[a]P (33%), reflecting a high water solubility and low lipophilicity of the former compound (Laurent et al. 2002). This research group extended these studies in lactating goats (Grova et al. 2002). B[a]P, phenan-threne, and pyrene were given to goats through oral gavage and the kinetics of elimination was studied. Relative to milk, excretion through urine and feces were high for the three chemicals studied. Although 75% of pyrene and phenanthrene each were assimilated, the value was 12% for B[a]P. The authors attributed the absorption characteristics to the molecular size of B[a]P (5 benzene rings), compared to phenanthrene (3) and pyrene (4) rings. West and Horton (1976) reported similar transfer kinetics of PAHs from diet to milk in lactating ewes (sheep), rats, and rabbits. Though studies like these use labeled chemicals and yield interesting information, the lack of details on the PAH dose administered (mg/kg basis) makes it difficult to normalize the data for dose and compare these values to others for risk assessment purposes.

The types of food ingested (oil, fat, etc.), composition of gastrointestinal fluids, and transport processes across intestine influence the bioavailability of PAHs. The first two aspects are discussed elsewhere in this review. Cellular transport processes influence the PAH compound mobilized from diet or soil across the intestine. Part of the ingested PAH compound is metabolized in the small intestine (Zhang et al. 1997b). The ATP-binding cassette (ABC) transport proteins in the apical membrane of small intestine mediate the transport of B[a]P metabolites back into the intestinal lumen. This helps in preventing resorption of metabolites by decreasing the oral bioavailability of B[a]P, thus rendering a beneficial effect against potential carcinogenic and mutagenic metabolites. Human Caco-2 cells were used as a model to understand the interplay between metabolism and redirected transport into the intestinal lumen (Buesen et al. 2002, 2003). These studies indicated that Caco-2 cells reduced the resorption of B[a]P through metabolism and the redirected transportation of metabolites back into the intestinal lumen. Thus, these processes serve as an effective intestinal barrier preventing buildup of a reactive metabolites in the target organs as a result of absorption.

Hepatic Metabolism of PAHs

It is a well-established fact that liver plays a predominant role in the metabolism of PAHs. However, it is impossible to define the origin of metabolites as PAH administration results in accumulation of metabolites in target tissues as a result of transport from the liver and local metabolism as well. To assess the role of liver in the mechanism of toxicity/carcinogenesis, Wall et al. (1991) conducted experiments using orthotopic liver transplantation. Labeled B[a]P was infused into the portal vein of rats, and livers were perfused and either transplanted to another rat or sham operated and left in situ (non-transplant group). After 4 h, lungs, kidneys, spleen, colon, adrenals, heart, and brain tissues were collected and polar metabolites and DNA adducts were measured. Concentrations of B[a]P was higher in livers of nontransplant group compared to the transplant ones. Concentrations of polar metabolites were nearly identical in peripheral tissues from both groups. There were no differences in DNA binding between the transplant and nontransplant groups. If metabolism in target tissue exceeds that of liver, one would expect to observe more DNA adducts in the nontransplant group than the transplant group. These findings indicates that liver is the major reservoir for PAHs, capable of extracting circulating B[a]P from blood (Wiersma and Roth 1983) and facilitate the delivery of metabolites to target tissues. Although this argument may hold good for PAH exposures of a short duration, what would happen in long term exposure conditions is not yet known. Because various isoforms of CYP are reported in liver and other tissues, experiments are needed to inhibit the iso-forms expressed predominantly in liver and follow the course of metabolite formation. Employing some inhibitor compounds to hinder the uptake of circulating metabolites by target tissues may provide some new information on the contribution of liver to PAH toxicity/carcinogenesis.

Extrahepatic Metabolism of PAHs

In addition to liver, PAH metabolism also occurs in extra-hepatic tissues albeit to a minor extent. The expression and activity of several CYP enzymes that are known metabolizers of PAHs in extrahepatic tissues of humans and other mammals have been extensively studied and reviewed (Guengerich and Shimada 1991; Whitlock 1999; Ding and Kaminsky 2003; Shimada et al. 2003; Uno et al. 2004). Following oral administration, PAHs and/or their metabolites have been detected in gastrointestinal tracts (Ding and Kaminsky 2003), lung (Ramesh et al. 2001), brain (Saunders, Ramesh, and Shockley 2002), and kidney (Roos 2002).

The distribution of administered PAH dose to various target organs and the ability of these organs to process the chemical determine the extent of damage caused by the parent compound. Helleberg et al. (2001) measured the interorgan distribution of BPDE dose in mice (deficient in excision repair) fed with 13 mg B[a]P/kg body weight. Because BPDE is the major mutagenic metabolite of B[a]P, reacts with deoxyguanosine-N ² and forms adducts, which are indicators of premutagenic events, these authors measured the dGuo-N ²-BPDE adducts in different organs. The organs studied were stomach, small intestine, colon, spleen, lung, and liver. Adduct levels were highest in liver and lung, probably due to induced CYP1A activities. On the other hand, the adduct concentrations were approximately the same in rest of the tissues and two- to threefold lower when compared to liver. The low adduct levels nevertheless indicate a decreased bioactivation of B[a]P in target tissues. This pattern of low adduct levels may change in a chronic exposure conditions as tumors have been diagnosed in extra hepatic tissues (Chen and Chu 1991; Culp et al. 1998). Furthermore, PAHs such as B[a]P were known to generate metabolites other than the diol epoxides that are reactive towards DNA (Penning et al. 1999). The extent to which some of these metabolites such as redox active o-quinones are produced in extrahepatic tissues is not clear. Hence there is a need to conduct experiments on the contribution of extrahepatic tissues to B[a]P metabolism. This could be done in an appropriate rodent model by blocking the blood supply to the liver and bypassing the portal vein. The concentrations of PAH parent compound/metabolites including diol epoxides and reactive quinones could be measured in plasma.

Esophageal Metabolism of PAHs

In addition to PAH intake through diet, PAH intake through inhalation also enters the GI tract, albeit to a minor extent. Particle-bound PAHs are cleared rapidly into the GI tract via the mucociliary system (Sun et al. 1984; Bevan and Ruggio 1991). A schematic representation of PAH intake, biotransformation, and excretion is shown in Figure 4.

The initial site of metabolism after oral ingestion of xenobiotics is esophageal epithelium. Tumors of the epithelium of esophagus may arise from chronic exposure to dietary-borne and/or tobacco smoke–borne toxicants such as PAHs. The tongue and esophagus were found to be the major targets of tumor formation in mice that were fed with B[a]P for a 2-year period. Fifty percent of mice employed in this study developed tumors in tongue and esophagus (Culp et al. 1998). Using a MutaMouse model, von Pressentin, Kosinska, and Guttenplan (1999) reported that the oral tissues and esophagus have a significant capacity for metabolic activation of B[a]P and 7,12-DMBA. Their findings also suggested DNA damage in the aforementioned sites can be converted to mutations. Studies have associated high levels of environmental PAH exposure with a high incidence of esophageal and colorectal cancers in humans. Inhalation of PAH-laden fumes released from combustion of soft coal used in unvented stoves and consumption of food tainted with PAHs were presumed to be associated with high rates of oesophageal cancer in Linxian, China (Roth et al. 1998, 2001). A recent report (van Gijssel et al. 2004) demonstrated the presence of PAH-DNA adducts from archived esophageal endoscopic biopsy samples obtained from Linxian. Furthermore, the UDP-glucuronosyltransferase genes (genes expressing detoxifying enzymes for these chemicals) showed a polymorphism in humans afflicted with orolaryngeal carcinomas (Elahi et al. 2003). G → T transversions of the tumor suppressor gene T53 has been observed in lower esophagus and cardia of humans (Breton et al. 2003). A link between B[a]P exposure and T53 mutations at the above mentioned substitutions has already been established in lung cancer cases (Denissenko et al. 1998). Hence it is likely that the T53 mutations in cancers of esophagus and cardia are likely to have arisen from PAHs contained in tobacco smoke and food. These findings attribute a role for PAHs in the causation of upper digestive tract cancers.

Intestinal Metabolism of PAHs

The intestines are the functional interface through which ingested PAHs are absorbed into the body (Mirvish et al. 1981; Ramesh et al. 2001), facilitated by bile (Laher et al. 1983). The gastrointestinal uptake was estimated to be at least 30% (Barrowman et al. 1989). Investigations on the role of bile in the bioavailability of PAHs in the rat intestine revealed that increasing ring number and solubility plays a role in absorption in the presence of bile (Rahman, Barrowman, and Rahimtula 1986). The values for absorption without bile were 92%, 97%, 71%, 43%, and 23% for 2,6-dimethylnaphthalene, phenanthrene, anthracene, 7,12-DMBA, and B[a]P, respectively. These results indicate that absorption of the four- and five-membered ring compounds (DMBA and B[a]P) was strongly dependent on the presence of bile in the intestinal lumen. Studies of Cavret et al. (2003) provide further evidence that physicochemical properties of PAHs play a key role in intestinal permeability and bioavailability. The absorption of phenanthrene and B[a]P respectively were 9.5 and 5.2% after a 6-h exposure in Caco-2 cells, and 86.1% and 30.5%, respectively, 24 h following ingestion in pigs. Consistent with these findings was a report from an in vitro PAH solubilization study (Roos et al. 1996) that showed a decrease in extraction efficiency with an increase in molecular weight of the PAHs. A correlation observed between increased proportion of total recovered metabolites in bile and increased size of PAHs support the notion that PAHs with poor water solubilities are significantly dependent on bile salts for their absorption (Rahman, Barrowman, and Rahimtula 1986). Prior to systemic uptake, small intestine contributes to the first-pass metabolism of ingested and absorbed PAHs in view of its positioning in the anatomical system, the expression of drug-metabolizing enzymes and the large surface area available in this organ (Kaminsky and Zhang 2003). CYP1A1 (Zhang et al. 1996; Spatznegger, Horsmans, and Verbeck 2000; Sterling and Cutroneo 2002; Roos 2002) and UDP-glucuronyltransferases (Dubey and Singh 1988) that are involved in the metabolism of PAHs have been isolated from the small intestine.

Ingestion of PAHs causes a very rapid and profound autoinduction of intestinal CYP1A1 that occurs as a consequence of the kinetics of intestinal absorption. For example, orally administered B[a]P produced strong induction of CYP1A1 in the villi of proximal intestine with a distal decline in mice. The larger number of mutations in proximal than distal intestine support this observation (Brooks et al. 1999). Zhang et al. (1996, 1997a) reported that the induced levels of CYP1A1 and the levels of Ahr decrease from proximal to distal part of the small intestine. Similarly, Roos (2002) observed a significant CYP1A1 induction in duodenum after oral intake of PAH-contaminated soils in rat. This study (Roos 2002) confirmed the earlier report of Zhang et al. (1997a) that CYP1A1 is more rapidly responsive to induction in the small intestine than it is in the liver. However, Zhang et al. (1997a) found that small intestinal CYP1A1 was maintained for shorter periods than in liver due to the short-life span of enterocytes (Iatropolous 1986). Fat consumed through diet may change the fatty acid composition of the membrane phospholipids that in turn alter the configuration of the CYP1A1. The changes in induction or kinetic properties of these enzymes are likely to affect the distribution of PAH parent compound and their metabolites to other target sites in the body.

After initial metabolism, most of the detoxified portion of PAH compounds are excreted into the bile as metabolites, then subsequently eliminated through the feces. A smaller proportion is excreted through urine. Table 4 summarizes the absorption and excretion data for several orally administered PAHs. Because most of the PAH metabolites excreted by the liver (products of phase II biotransformation) into the bile are water soluble, they are more likely to be eliminated (Chipman et al. 1981; Grimmer et al. 1988). However, enzymatic hydrolysis of the glucuronide and sulfate conjugates by intestinal microflora release the less polar compounds that may be reabsorbed into the portal circulation. The process of excretion of PAH compounds into the intestine via the bile, reabsorption, and return to the liver by the portal circulation is termed enterohepatic recirculation and has been demonstrated to occur for DMBA (Laher et al. 1983), B[a]P (Bowes and Renwick 1986; Bevan and Weyand 1988; van Schooten et al. 1997), anthracene (van Schooten et al. 1997), pyrene (van Schooten et al. 1997; Viau et al. 2002), and 1-nitropyrene (Medinsky et al. 1985). The glucuronic acid conjugate of B[a]P 4,5-diol and glutathione conjugate of B[a]P 4,5-epoxide in rodents were reported to undergo extensive enterohepatic circulation (Elmhirst et al. 1985). Likewise, the dihydrodiol glucuronide and glutathione conjugate of naphthalene were reported to undergo enterohepatic circulation (Bakke et al. 1985). Enterohepatic circulation extends the residence time of PAHs in the body. Continuous enterohepatic recycling may lead to long half-lives of reactive PAH metabolites that were reported to be formed in colon and cause genetic damage (Autrup et al. 1978; Reddy et al. 1991). These findings suggest a role for PAHs in colorectal carcinogenesis. Though enterohepatic recycling of PAH has been studied in animal models, it has not been confirmed in human studies (de Kok and van Maanen 2000). The reflux of bile into the pancreatic duct of people occupationally exposed to PAHs may subject them to an increased risk of pancreatic cancer (Wang et al. 1998). However, using laparoscopic cholecystectomy, de Kok et al. (2000) determined the biliary concentrations of 1-hydroxypyrene and 3-hydroxy B[a]P, two established biomarkers of PAH activation, in a population of smokers and nonsmokers. The nondetectable concentrations of PAHs or their metabolites in bile imply that enterohepatic circulation of PAH and/or metabolites and subsequent pancreatic reflux is unlikely in humans. Thus, there is a clear need for research to support the involvement of enterohepatic cycling of PAHs in colorectal carcinogenesis.

For carcinogenic PAHs ingested through a lipid-rich diet, the intestine is the first sorting station where digestion, dispersion, and membrane and cytosolic transport occur before the lipids are packaged for delivery to the general circulation (Laher et al. 1984). Two pathways are involved in transport of absorbed PAH from intestine to the rest of the body. Soon after absorption, hydrophilic compounds are carried into systemic circulation through lymphatic transport, whereas hydrophobic compounds are carried into the portal circulation reaching liver where they are metabolized. Using closed-loop sections of rat intestine, Bock, Clausbruch, and Winne (1979) found that >90% of dietary B[a]P was recovered in portal blood as metabolites. To study the relative importance of lymphatic and portal routes, rats were given doses of 10 μg, 10 mg, and 20 mg of radiolabeled DMBA in olive oil by intraduodenal infusion (Laher and Barrowman 1983). Biliary and mesenteric lymphatic catheters were installed to allow collection of excreted metabolites and the proportion of the compound transported by lymph after absorption. Exogenous bile was infused into the duodenum to replace bile diverted by the biliary cannula. With complete intestinal lymphatic diversion, a significant amount of radiolabel can be detected in bile before any radiolabel could be detected in lymph, suggesting portal venous transport. In 24 h, the combined recoveries of radiolabel in bile and lymph were 23%, 13%, and 20% for the 10-μg, 10-mg, and 20-mg dose of DMBA, respectively, and the proportions recovered in bile were 82%, 75%, and 77%, respectively. The above findings suggest that portal venous transport of absorbed DMBA, probably as polar metabolites, is of much greater importance than lymphatic transport. Recent studies by Laurent et al. (2002) and Cavret et al. (2003) also demonstrated that transfer of labeled B[a]P and phenanthrene in pigs occur through absorption of portal system.

In a similar study using radiolabeled B[a]P, the same research team (Laher et al. 1984) followed the appearance of isotope in bile and lymph. The hydrocarbon was administered in two different amounts (0.4 μmol in 50 μmol and 500 μmol triolein) via intraduodenal infusion (Laher et al. 1984). Total radiolabel recovered in a 24-h period was 20% and 17% of the administered dose of B[a]P for the small and large amounts of carrier lipid, respectively. Almost 80% of the total radiolabel recovered was in the form of biliary metabolites despite complete intestinal lymph diversion.

Though most of the ingested PAH is metabolized and excreted, in some instances, unmetabolized parent compound is directly passaged into the lumen of the GI tract and eliminated through feces. Studies have reported unchanged B[a]P (Foth, Kahl, and Kahl 1988; Ramesh et al. 2001), chrysene (Grimmer et al. 1988), and pyrene (Jacob et al. 1989) in feces of rats orally exposed to this chemical. Oral administration of soil contaminated with PAHs resulted in an increased amount of unchanged B[a]P, pyrene (van Schooten et al. 1997), and phenanthrene (Kadry et al. 1995) in feces. The passage of unchanged PAH in feces depends on the bioavailability of administered dose. The high doses of B[a]P administered to animal models (Ramesh et al. 2001) may have caused a poor extraction of B[a]P by the GI fluids and a capacity limited absorption and biotransformation eventually contributing to unmetabolized B[a]P in the feces.

Metabolism by Intestinal Microflora

Using colon microflora from the simulator of the Human Intestinal Microbial Ecosystem (Molly, Woestyne, and Verstraete 1994), Van de Wiele et al. (2004) demonstrated the formation of hydroxylated metabolites of PAHs. Microbial production of 1-hydroxypyrene and 7-hydroxybenzo[a]pyrene was observed in colon suspensions incubated with PAHs. The relevance of gut microbial metabolism of PAHs to the overall absorption and toxicity of these compounds remain to be explored in detail. Studies have demonstrated that glucuronic acid conjugates of biliary PAH metabolites can be deconjugated by some intestinal microflora to potentially reactive species that are reabsorbed, entering the portal circulation (Ball et al. 1991). Intragastric administration of 2-nitrofluorene in germ-free rats showed no hemoglobin (Hb) adducts compared with rats equipped with a bacterial flora derived from human or rat feces that showed Hb adducts (Scheepers et al. 1994). Furthermore, pure and mixed cultures of intestinal microflora from humans and rodents reduced 1-nitropyrene to 1-aminopyrene, a biologically active isomer (Cerniglia and Somerville 1995). In germ-free rats, B[a]P metabolism was reported to proceed through a novel ring opening process, leading to the formation of 7-oxo-benz[d]-anthracene-3,4-dicarboxylic acid, a genotoxic metabolite (Yang et al. 2000). On the basis of these findings, it is reasonable to assume that intestinal microflora have a role in modulating toxicity/carcinogenesis of B[a]P.

Several factors are important for the metabolic activation of PAHs by gut microflora in an in vivo situation. These are metabolic interactions between enzymes of the host gut mucosa and microflora, the composition of mixed populations of microflora, the geometric structure of the PAH compound ingested, and the type of diet (WHO 2003).

There is no experimental evidence that the reactive metabolites generated by gut microflora contribute to toxicity or carcinogenicity of the tissues through which they pass. On the other hand, recent studies (Lo et al. 2004) using Ames test have demonstrated that several bifidobacteria and lactic acid bacteria possesses antimutagenic activities against B[a]P. When probiotic foods such as whole milk, semiskimmed milk, and skimmed milk were preincubated with B[a]P and Bifidobacterium lactis, the antimutagenic activities increased to more than 99%.

Relative Contribution of Liver and Intestine to PAH Metabolism

Studies have revealed the small intestine is of less relative importance when compared to liver in the first-pass metabolism of PAHs. For example, in humans, the greater weight of liver (around 1.5 kg) relative to that of small intestine (around 0.7 kg), which when the CYP concentrations and microsomal protein contents are taken into account, attributes a greater metabolic capacity for liver (Lin, Chiba, and Baillie 1999; Doherty and Charman 2002). A recent review by Ding and Kaminsky (2003) indicates that intestinal CYP metabolism serve as a barrier (through detoxification) to the systemic uptake of toxic chemicals.

To investigate the relative contribution of liver and intestine towards toxicity of orally ingested PAHs, a multicompartment perfusion system (biohybrid simulator) was developed to mimic the PAH absorption process across the small intestine and bio-transformation in the small intestine and liver (Sakai et al. 2003; Choi et al. 2004). This system consists of three interconnecting physiologically relevant compartments: the top compartment houses the Caco-2 cell membranes, the middle compartment houses the methylcholanthrene induced Hep G2 cells, and the bottom compartment was designed to house other tissues. B[a]P was chosen as a model PAH to examine the metabolic contribution of Caco-2 and Hep G2 cells to toxicity under individual and co-culture experimental conditions. An enhanced CYP1A1/2 activity was seen in both cell lines. When B[a]P was introduced to the apical side of the Caco-2 cell layer, CYP activities were enhanced. This resulted in the production of high concentrations of B[a]P metabolites in the apical side of Caco-2 compared to basolateral, liver, and other tissue compartments presumably due to permeation of B[a]P across the Caco-2 cells. The toxicity (measured by the cell viability) of B[a]P to Hep G2 cells, was more than that of Caco-2 cells in pure cultures, despite the low production of B[a]P 7,8-diol (a precursor of BPDE) in Hep G2 cell compared to Caco-2 cell pure cultures. This study provides further evidence that the intestinal pathway of B[a]P biotransformation, if not as substantial as the hepatic pathway, should not be overlooked.

The differential expression of genes encoding for PAH metabolism in intestine and liver also play an important role in the bioavailability of ingested PAHs. Lampen et al. (1998) and Lindell, Lang, and Lennernas (2003) reported that CYP1A1 is expressed strongly in the rat small intestine as opposed to liver. On the other hand, CYP1A2 is expressed highly in liver (Ding and Kaminsky 2003).

BIOAVAILABILITY OF PAHs

Bioavailability studies are warranted to address the concern whether sufficient doses of PAHs are absorbed into the blood for distribution to other tissues and thereby eliciting toxic effects. From the ecotoxicology standpoint, bioavailability can be broadly defined as the portion of toxicant in the ambient environment that is available for uptake by an animal/organism and the ensuing biological actions (Rand 1995). This approach involves feeding groups of laboratory animals with a known amount of toxicant over a finite period of time and measuring the fraction of total dose retained in the animal after allowing time for clearance of food from the digestive tract. Bioavailability is calculated by mass balance under the assumption that all toxicant measured in the organism is assimilated (McCloskey, Schultz, and Newman 1998).

From a classical pharmacology perspective, bioavailability of a toxicant can be defined as the fraction of administered dose reaching the systemic circulation of the animal (Gibaldi 1991). In pharmacokinetic studies, oral bioavailability is calculated from the ratio of areas under the blood concentration-time curves (AUCs) of orally and intravenously administered doses of a toxicant.

The relative merit of each approach to measure bioavailability is not discussed here as it is beyond the scope of this review. Interested readers may refer to various white papers, reports, and books (Hrudey, Chen, and Rousseaux 1996; NEPI 2000; NRC 2003) for a detailed review. However, for environmental chemicals including PAHs, both approaches have been used to address bioavailability issues.

Estimating bioavailability from one matrix may not be sufficient enough to generalize for other matrices. Usually in oral bioavailability studies, the widely used media are food, water, suspensions, etc. In vivo toxicity studies using lab rats suggest that oral bioavailability of soil-borne toxic chemicals might be less than that found for these chemicals in food (Hrudey, Chen, and Rousseaux 1996). As a result of this matrix-related difference in bioavailability, there is a possibility of erroneously over estimating the risks for toxicants in soil. In this context, it should be noted that soil ingestion is a major route of exposure to PAHs in children (Wilson, Chuang, and Lyu 1999 Wilson, Chuang, and Lyu 2000 Wilson, Chuang, and Lyu 2001; Wilson et al. 2003).

In Vitro Models

To account for the effect of matrix in risk assessment, researchers have developed in vitro models that simulate human physiological conditions. These models have advantages in that (i) they reduce the need for laboratory animals and suggest alternatives; (ii) various parameters that contribute to variability in bioavailability can be investigated in a systematic manner; (iii) these models are less expensive and reproducible. This approach has successfully been used for contaminants like organo-chlorine chemicals (Oomen et al. 2002) and metals (Hainel, Buckley, and Lioy 1998).

The bioavailability of orally ingested PAHs depends on mobilization of these compounds from the surrounding matrix under physiological conditions within the GI tract. To elucidate the factors controlling the mobilization or gastrointestinal solubilization of PAHs, studies were conducted using in vitro models (Hack and Selenka 1996; Holman, Mao, and Goth-Goldstein 1997; Holman et al. 2002) that simulate the conditions within the GI tract. One of the models (Hack and Selenka 1996) tests the fate of ingested PAH in two steps in a successive manner, the first one representing the stomach and the second one representing the intestine. Compared to the gastric juice, intestinal juice extracted PAHs effectively due to the action of bile to form micelles with fatty acids. When lyophilized milk powder (a source of protein and lipids) was added to gastric juice, extraction efficiency was more (at least a twofold increase). This model used an autotitration unit (pH meter, dosage pumps, and control unit) and a temperature-controlled water bath with shaker to simulate the conditions in the gut.

Another model (Holman et al. 2002) explored the concept of using diminished bioavailability of weathered petroleum residues (petroleum residues in highly weathered soils collected from diesel- and crude oil–contaminated sites) as a function of solubilization of these compounds in soil. The authors initially used a mammalian (human) digestive tract model to study the bioavailability of PAHs (Holman, Mao, and Goth-Goldstein 1997) and then extended these studies to petroleum hydrocarbons. They studied the fate of PAH compounds in fasted and fat digestion states using a synthetic upper small intestinal digestive fluid that included mixed bile salts and intestinal lipids. Fasted state covers mixed bile salts and fat digestion state includes mixed bile salts as well as intestinal lipids. The GI solubility of hydrocarbons increased from the fasting phase to fat digestion phase. This model used a patented stepwise solubilization system that involves absorption of bile salts to the soil surface, reacting with hydrocarbons to form micelles and desorption of hydrocarbons from the soil and diffusing into lumen’s fluid. Subsequently the micelles penetrate the unstirred layer, hydrocarbons adsorb to the microvilli of enterocytes, diffuse across the cells, and enter the blood and lymph circulation.

The sources of bile used seem to have no bearing on the extractability of PAHs in in vitro digestion models. Studies were conducted recently to see if bile of animal origin will contribute to an increased bioaccessibility of soil-borne PAHs when compared to purified bile salts in in vitro models (Oomen et al. 2004). The differences in bioaceessibility were less than 10% when four soils spiked with B[a]P were extracted individually with ox, pig, and chicken bile.

Though both models (Hack and Selenka 1996; Holman et al. 2002) provide valuable information, they do not account for the interactions between chemicals of interest and digestive enzymes. Although petroleum hydrocarbons are not expected to have significant interactions with digestive enzymes, the converse may be true for other chemicals. For example, toddlers and children of other age groups from residential dwellings and day care centers that are old (Heudorf and Angerer 2001), located in low-income areas (Wilson, Chuang, and Lyu 1999 Wilson, Chuang, and Lyu 2000 Wilson, Chuang, and Lyu 2001; Wilson et al. 2003; Chuang et al. 1999), on superfund sites (Calabrese et al. 1997), former coal mine tailings (van Wijnen et al. 1996), tar ponds (Lambert and Lane 1994), and polluted urban areas (Vyskocil et al. 2000; Fiala et al. 2001) get exposed to many soil-bound contaminants, including PAHs through ingestion of contaminated soil, household dust, and food. To explain the fate of ingested PAHs in such scenarios, in vivo studies are warranted using animal models to evaluate the integrity of in vitro models.

In Vivo Models

The formation of reactive metabolites and carcinogen-DNA adducts are critical steps in carcinogenesis and are considered to be important biomarkers during the initiation stage for carcinogenesis (Dipple 1995). The cumulative levels of CYP, urinary PAH metabolites, and adducts are expected to be proportional to the dose. This approach has successfully been utilized to study the rodent bioavailability of soil-bound PAHs from coal tar contaminated soils (Weyand et al. 1991; Koganti et al. 1998; Bordelon et al. 2000) and from manufactured gas plant residue (Weyand et al. 1994). Furthermore, dietary exposure to PAHs resulted in enhanced DNA adduct levels in humans (Rothman et al. 1993; van Maanen et al. 1994). Therefore, an understanding of the influence of diet on DNA adduct formation in target organs in animals is important in discovering which organ systems are more vulnerable to damage from continuous intake of PAHs when diets are contaminated.

Role of Dietary Fat in PAH Bioavailability

In higher animals, lipophilic toxic environmental contaminants are taken up via coabsorption with dietary lipids across the wall of the small intestine (Dulfer, Groten, and Govers 1996). In sheep, gastrically instilled B[a]P was absorbed via uptake of chylomicrons and carried to the vascular circulation via lymph flow (Busbee, Norman, and Ziprin 1990). Studies conducted in vitro and in vivo provided evidence that fatty foods facilitate transfer of B[a]P from food particles and enhance the transfer of B[a]P to the intestinal wall (Stavric and Klassen 1994). The question of whether increased fat intake translates into increased bioavailability of ingested toxic chemicals is unresolved. Increased lipid feeding resulted in a similar bioavailability and lymphatic transport (Laher et al. 1984) in rats administered with B[a]P intraduodenally. Bowes and Renwick (1986) observed no change in the inducible levels of B[a]P hydroxylase and DNA binding in the intestine of guinea pigs that were fed normal or high-fat diets. On the contrary, Clinton and Visek (1989) reported efficient absorption and bioavailability of DMBA in rats fed high-fat diets. A direct relationship between PAH absorption and fat absorption was demonstrated by the studies of Laurent et al. (2001). When pigs were administered [¹⁴C]phenanthrene or [¹⁴C]B[a]P in milk, the time-course plasma concentrations of these chemicals showed a peak 5 to 6 h post administration, corresponding with the period of maximum fat absorption.

O’Neill et al. (1990a, 1990b) conducted studies in rodents to see whether the intake of B[a]P through dietary fat similar to that of humans would affect B[a]P metabolite formation. Mice were fed diets containing the principal ingredients within the normal human intake range. Increased dietary fat led to an increased production of B[a]P metabolites. Zaleski et al. (1991) have shown that orally administered B[a]P is sequestered in lipid droplets. An inverse relationship was observed between the levels of tria-cylglycerols and B[a]P metabolism in hepatocytes isolated from rats that were maintained on high-fat and food-restricted diets. Decreases in levels of glucuronide and sulfate conjugates in stomach, lung, and kidney of rats maintained on food-restricted diet compared to high-fat diet were reported (Kwei et al. 1991). These findings indicate that dietary modulation influenced carcinogen metabolism by altering the levels of hydrolases involved in the metabolism of conjugates. To date, few studies have been attempted to investigate the effect of dietary fat on metabolic activation versus detoxification processes.

Studies have shown that it is not the volume (Laher et al. 1984) but the fatty acid composition (Laher, Chernenko, and Barrowman 1983) of lipid administered with PAH that influences bioavailability. Yoo, Norman, and Busbee (1984) showed that the triglyceride content of plasma lipoproteins was positively correlated with PAH intake. Long-chain triglycerides were reported to promote more efficient absorption of 7,12-DMBA compared to medium-chain triglycerides (Laher, Chernenko, and Barrowman 1983). Gower and Willis (1986) found that the rate of B[a]P metabolism in the intestine depended on the quantity of dietary fat as reflected by the positive correlation between amount of metabolites produced by the intestine and the type of diet. Contrary to the notion that increased intake of only animal fat contributes to an increase in cancer incidence, studies in rodents have revealed that plant derived oils such as corn oil, safflower oil, and sunflower oil enhance cancer development (Carroll 1991; Fay et al. 1997). These oils are rich in unsaturated fatty acids. On the other hand, olive oil, which is also rich in unsaturated fatty acids, has no effect on cancer development (Carroll 1991; Fay et al. 1997). Is fat intake itself important for carcinogenesis? Is the amount of intake of dietary carcinogen also important for carcinogenesis? From a mechanistic standpoint, is there any interplay between these factors? These aspects are far from understood and require further study.

Studies of Gower and Willis (1987) revealed that an increase in polyunsaturated fat content in the diet will greatly elevate the conversion of B[a]P 7,8-dihydrodiol to its ultimate carcinogenic metabolite B[a]P 7,8-dihydrodiol, 9,10-epoxide and also a greater DNA binding. Thus, the type of lipids available in plasma and their levels may play a role in delivering absorbed PAHs to target organs. Busbee, Norman, and Ziprin (1990) have shown that high-density lipoproteins facilitate B[a]P uptake into hepatocytes whereas low-density lipoproteins inhibit the uptake. Therefore it is conceivable that differences in intake of PAHs from the dietary lipid matrix and transport of this chemical to the liver may determine the differences not only in organ specific metabolism but also in organ specific DNA adduct formation.

RELEVANCE OF ORALLY INGESTED PAHs IN RISK ASSESSMENT

Apart from the scientific standpoint, studies on bioavailability are also important from the regulatory or public health perspective. PAHs are the principal contaminants in hazardous waste superfund sites, 600 of which are on the U.S. National Priority List and targeted for federal clean up. This necessitates using appropriate animal models with necessary correction factors for PAH risk assessment in humans. Because humans and rodents have analogous patterns of PAH metabolism (Selkirk 1985), studies on biomarkers of exposure such as the concentrations of parent compound and/or reactive metabolites and adducts in various tissues help to establish the link between exposure events, the resulting toxic effects and extrapolate the findings to humans.

Absorption Adjustment Factors

The administered dose of a toxic chemical is often not completely absorbed in animals and humans due to differences in media and routes of exposure. Hence, it is not appropriate to directly apply a dose-response value from the laboratory studies to the human exposure dose. This necessitates introducing a correction factor in the calculation of risk to account for differences between absorption in the study from which the dose-response value or toxicity criterion was derived and absorption likely to occur upon human exposure to a toxic chemical. In other words, these correction factors help avoid over- or underestimation of human health risk from exposure to toxicants (Magee, Anderson, and Burmaster 1996).

This correction factor is defined as the absorption adjustment factor (AAF). The AAF is used to adjust the human exposure dose so that it is expressed in the same terms as the doses used in the dose-response study performed with laboratory animals. Thus, the AAF is the ratio between the estimated absorption factor for the specific matrix and route of exposure, and the reported or estimated absorption factor for the laboratory study from which the dose-response value was derived (Magee, Anderson, and Burmaster 1996).

AAFs can be derived from data within a single or multiple experiments if an appropriate measure of absorption is compared between different routes of administration and/or sample matrices.

In the absence of data from a single experiment that quantitates the absorption from the similar route and matrix, the AAF is derived using the following equation (Magee, Anderson, and Burmaster 1996).

$$\text{AAF}=\frac{\text{fraction absorbed from the environmental exposure}}{\text{fraction absorbed in the dose - response study}}$$

Absorption adjustment factors can be less than 1.0 or greater than 1.0. If the absorption from the site-specific exposure is the same as absorption in the laboratory study, then the AAF is 1.0. An AAF of 1.0 indicates that absorption is known or estimated to be the same as that in the dose-response study. The use of an AAF permits the risk assessor to make necessary adjustments for differences in bioavailability between laboratory vehicles and environmental matrices. The Guidelines for Exposure Assessment prepared by the Environmental Protection Agency (USEPA 1993) discusses the appropriateness of using properly documented absorption/bioavailability factors in risk assessment process.

Using these guidelines, Magee, Anderson, and Burmaster (1996) derived the AAFs for PAHs by taking into account matrix-specific bioavailability and knowledge of PAH pharmacokinetics obtained from the studies of Goon et al. (1991), Rozett et al. (1996), and Weyand et al. (1996). The above-mentioned research groups used soils of various particle sizes and chemical composition to administer PAHs or manufactured gas plant residue containing PAHs to rats and mice. Hence each data point for absorption values were given equal weightage by Magee et al. in the derivation of AAF. The data were subjected to curve fitting and simulation exercises to obtain a distribution of AAF values that were used to make probabilistic risk assessments. The mean oral-soil AAFs derived from the probabilistic risk assessments was 0.31, and the 50th percentile oral-soil AAF was 0.27. The mean value of oral-soil AAF derived from the above-mentioned three studies was 0.29, which is in agreement with the probabilistic risk assessment values. Hence a value of 0.29 was used as a point estimate of the oral-soil AAF for deterministic risk assessment purpose.

As mentioned earlier, the AAF approach is based on the premise that PAHs in soil pose a potential risk to human health. Because AAFs are derived from soil-borne PAHs, they are heavily dependent on soil characteristics. Although AAFs are useful for dermal exposure scenarios, they may have limited use for adult humans from an oral exposure standpoint. On the other hand, the AAF values will be of help to risk assessors in estimating the risk for children from soil-borne PAH exposure.

Intermedium Transfer Coefficients (Biotransfer Factors)

The National Hazardous Waste Reduction and Combustion Strategy (National Strategy) policy established by the USEPA requires conducting multipathway risk assessments for all hazardous waste combustion facilities in accordance with USEPA guidance provided in the Human Health Risk Assessment Protocol for Hazardous Waste Combustion Facilities (HHRAP, USEPA 1998). The HHRAP protocol uses site-specific land use and activity pattern information to determine plausible exposure scenarios and pathways. The recommended set of exposure scenarios are defined as receptors (i.e., resident adult, resident child, farmer adult, farmer child, fisher adult, and fisher child) and each receptor is assumed to be exposed via multiple applicable exposure pathways (i.e., inhalation of vapors and particles; ingestion of soil, drinking water, homegrown produce, home-grown beef, milk from homegrown cows, homegrown chickens, eggs from homegrown chickens, homegrown pork, locally caught fish, and breast milk). The HHRAP modeling protocol uses stack emission rates and air dispersion modeling to predict concentrations of contaminants in air, soil, and water. Additionally, these medium concentrations are used to predict concentrations of toxicants in other environmental matrices such as plants and animals. The intermedium transfer coefficients (ITCs), which describe the partitioning between various environmental compartments (McKone and Hammond 2000), were used to accurately predict the concentrations of contaminants in various media.

According to Travis and Arms (1988), the ITC or biotransfer factor (Ba) can be calculated by the following equation:

$$\text{Ba}=\frac{\text{tissue concentrations (mg/kg)}}{\text{daily intake of contaminants of concern (mg/day)}}$$

The ITCs that are currently used in HHRAP to describe uptake into the food chain are based on equilibrium-partitioning theory. This approach is based on the assumption that a descriptor of lipophilicity such as the octanol: water partition coefficient (K _ow) could be used for individual contaminants (most of the lipophilic and persistent ones) to predict their concentrations in animals. Because this assumption is based on thermodynamics, loss of the contaminant through mechanisms such as metabolism cannot be accounted for.

METABOLISM FACTORS

To test the validity of the ITC approach, contaminant accumulation in beef was used as an example. The concentrations of contaminants in beef were calculated as the product of the daily intake of contaminated medium (e.g., soil, fodder) times the ITC for that contaminant. In addition, the equation contains a “metabolism factor” (MF). The MF in other words is an “elimination rate factor” as it is used to modify the calculated tissue concentration of contaminants that are rapidly metabolized and eliminated and hence does not bioaccumulate in proportion to their K _ow values. However, for compounds like PAHs that are extensively metabolized and eliminated, this approach may erroneously overestimate the tissue concentrations (as PAHs move up through the food chain), and the resulting risk as well. To overcome these limitations, Hofelt et al. (2001) derived MFs for use in multipathway risk assessment approach for PAHs. Benz[a]anthracene (B[a]A) was chosen as the representative PAH chemical to calculate the MF as it had a complete set of absorption, distribution, metabolism, and excretion (ADME) data. These authors used the ADME data to compute the toxicokinetic parameters and the resultant data were used to estimate the MF. Further, to derive the MFs, an uncertainty factor of 10 was applied to account for interspecies differences in metabolism. Using the MF, the final predicted concentrations for PAHs in diverse matrices such as milk, chicken, eggs, and pork were calculated. The MF data for B[a]A were in agreement with calculated MF data for other PAHs such as B[a]P, and pyrene (Table 5). This shows a similarity in metabolic pathways for all PAHs, thus validating the PAH-specific metabolism factor. However, the revised MF approach of Hofelt et al. (2001) is very useful to compute the body burden of dietary PAHs from a risk assessment perspective, but will not be of much relevance to toxicity.

Bioconcentration Factors

Risk assessment of consumption of PAH contaminated aquatic biota has largely been based on federal ambient water quality criteria for PAHs, issued more than 20 years ago. The vast amount of literature that has been accumulated since then on the uptake, accumulation, and metabolism of these chemicals in aquatic animals warrant a revisiting of this topic. Because PAHs have limited environmental mobility, the transport and biological fate of these chemicals in aquatic environment is as important as their terrestrial counterparts.

Bioconcentration factor (BCF) is an important criterion in considering accumulation of toxicants in aquatic biota. This factor is defined as the ratio of a chemical concentration in tissue to the concentration of that chemical in water. It could be normalized on lipid basis also. Some limitations in this approach are using the BCFs derived from studies that were conducted in closed systems, for a limited period of time, that do not necessarily approximate the environmental conditions under evaluation. Furthermore, steady-state conditions, necessary to measure uptake and elimination rates, may underestimate the uptake in aquatic systems. To avoid assessing risk from these confounding factors, the concept of bioavailability has been proposed. PAHs that are more hydrophobic tend to be partitioned more onto organic carbon in the water column, making them less bioavailable. Data on PAH concentrations in the surrounding water alone tend to overestimate the amount available for uptake. Other modifying factors are habitat specific. For example, areas in the aquatic milieu that are heavily colonized would influence the levels. The ability of many aquatic organisms to metabolize PAHs will reduce the bioaccumulation/persistence of these chemicals in tissues. Consequently, a decrease in bioconcentration would occur. The BCF values available from literature are based on whole-body concentrations only. Thus these are cumulative values, regardless of the extent of accumulation of these chemicals by various tissues. Moreover, the differential or tissue-specific distribution of PAH compounds in aquatic animals need to be taken into consideration. Different sentinel organisms differ in their CYP biotransformation of PAHs. For example, PAHs tend to partition into organs with high biotransformation activity such as liver in fish or hepatopancreas in crabs, lobsters, and shrimp (Living-stone 1998), resulting in concentrations that are 10 to 50 times greater than those measured in fish muscle tissue. Because muscle tissues are the ones that are most consumed by humans, the tissue-specific distribution should be taken into account when human PAH intake via fish ingestion is estimated.

In the light of the above-mentioned confounding factors in consideration of the risks from exposure to PAHs in the aquatic environment, Boyce and Gary (2003) presented a new approach. This approach is based on developing an “alternative risk-based target concentration” for PAHs in aquatic animals assuming human consumption of these. These values were calculated assuming that PAH contaminants from a former creosote-handling facility are leached into the surface waters through groundwater transport and taken up by biota. In support of their work, the authors used the Model Toxics Control Act (MTCA) of Washington State, toxicity equivalency factors or potency equivalency factors or relative potency estimates (estimates developed from studies using B[a]P and at least another PAH compound (USEPA 1993). After making appropriate adjustments based on all these algorithms, the alternative risk-based concentrations developed for PAHs were greater by a factor of 30 than the default concentrations calculated using the original assumptions. Thus, these revised values do not raise any concern for human health in terms of consumption of aquatic biota from lakes and other water bodies adjacent to the former creosote-handling facilities. The scenario might, however, change in the face of increasing discharge of pollutants into aqueous media as a result of dredging and industrial runoffs.

Carcinogenic Potency Ratios

Because certain PAHs are potent carcinogens, risk assessment from a cancer standpoint needs consideration. For inhalation exposure of PAHs, cancer potency estimates have been derived from epidemiological data (Boström et al. 2002). On the other hand, cancer risk assessment for oral exposure is handicapped by inadequate data. As oral uptake of contaminated food and soil is of special concern to general population and people who live near hazardous waste sites, respectively, cancer risk estimates for oral uptake of PAHs is warranted.

Most if not all exposures of humans and animals to PAHs involve complex organic mixtures. Metabolism and bioactivation of individual PAH compounds are influenced by the presence of other PAHs (Warshawsky 1999; Goldstein 2001). Ironically, most studies have been done with pure compounds. B[a]P an important component of PAH mixtures, has often been used as a surrogate compound for risk estimates for PAH exposure. These estimates were based on animal studies, which were either incomplete or insufficient. Additionally, adequate long-term studies have not been available to make meaningful cancer risk estimates. Toxicity equivalency factors (Collins et al. 1998) have been proposed for various PAHs relative to B[a]P. This approach is based on assumption of additive risks for individual PAHs in a mixture. However, the carcinogenic risk of PAH mixtures is highly dependent on the exposure pathway. This necessitated a revisiting of this topic by Schneider et al. (2002) who made a reassessment of oral cancer potency of PAHs using recent oral carcinogenicity studies with B[a]P and coal tar mixtures, as well as some older studies for a critical reappraisal.

As a first step, Schneider et al. (2002) selected carcinogenicity studies with oral exposure that allow a direct comparison of the carcinogenic potency of pure B[a]P and coal tars (that contain high amounts of PAH mixtures). Two studies provided reliable dose-response data to assess the carcinogenic risk of PAH mixtures. Culp et al. (1998) applied two different coal tar mixtures (CTM; CTM1 or CTM2) or B[a]P in food to B6C3F1 mice over a lifetime. Dose-response assessment was carried out using B[a]P uptake as a surrogate for PAH uptake. The amount of B[a]P fed as a component of CTM was given in Culp et al. (1998) for most dose groups. With B[a]P alone (up to 100 ppm), they observed a higher incidence of fore stomach tumors compared with controls, whereas with both of the PAH-rich coal tar mixtures (up to 10000 and 3000 ppm, respectively) the tumor incidence was increased in a dose-dependent way for various locations, most prominently in lung, fore stomach, small intestine and for various types of sarcomas (Table 6). Weyand et al. (1995) used A/J mice for a similar feeding study with B[a]P and a PAH-rich manufactured gas plant residue (MGP). The A/J mice were prone to develop lung tumors after exposure towards various carcinogens and, indeed, the authors found increased numbers of tumors in lung after exposure with B[a]P (16 and 98 ppm) and MGP (1000 or 2500 ppm) in food for 260 days. Moreover, with B[a]P (but not with MGP) the forestomach tumor incidence was increased.

The individual PAHs concentrations in these mixtures were used for calculating B[a]P equivalents (Schneider et al. 2002). By using relative potency values according to the USEPA (Brown and Mittelman 1993), the sum of B[a]P equivalents was calculated as a percentage of B[a]P. Furthermore, potency ratios (carcinogenic potency of a PAH mixture divided by the carcinogenic potency of B[a]P as a single substance) were determined for these studies and the results are given in Table 7. The calculated potency ratios were then compared with predictions based on relative potency values (B[a]P equivalents). The results showed that potency estimates predicted by relative potency values were poorly correlated with the potency of PAH observed in bioassays (Figure 5). Thus the calculation of potency by B[a]P equivalents will underestimate the real potency for most PAH mixtures.

The authors (Schneider et al. 2002) also found a direct relationship between potency ratios and tumor locations for oral exposure route. B[a]P is responsible for forestomach tumors in mice as indicated by potency ratios of about unity. On the other hand, B[a]P’s contribution to lung carcinogenesis is small (Figure 6). These data strongly suggest that for oral exposures, the potency ratio between pure B[a]P and the PAH mixture is dependent on the target organ. To describe risk for humans after oral intake of PAH mixtures, Schneider et al. (2002) derived a cancer slope factor using data from a coal tar mixture feeding study (Culp et al. 1998), making necessary adjustments for body weight and caloric demand. A slope factor of 11.5 was obtained for humans, which translates into human excess risk per oral lifetime exposure with 1 mg B[a]P kg⁻¹ day⁻¹ in a PAH mixture.

On the whole, these findings indicate that the contribution of B[a]P to the carcinogenic potency of various PAH mixtures from industrial sources are relatively constant. Gaylor et al. (2000) expressed risk in terms of coal tar concentration in the diet and proposed its use in assessing manufactured gas plant waste sites. Because coal tar is not a defined entity, their approach may not accurately reflect risk from contaminated soil exposure. On the contrary, the oral slope factor derived by Schneider et al. (2002) is recommended for assessing health hazards by oral exposure due to PAH contaminations at hazardous waste sites.

CONCLUSION

The above account articulates the importance of orally ingested PAHs in risk assessment processes. As PAHs have been implicated as causative agents of breast, lung, and colon cancers and have been associated with neuro-, reproductive, and developmental toxicities, the processes governing the disposition of these chemicals in the body and their subsequent metabolic fate assume a greater importance. Towards this end, there is a growing need for studies that involve physiologically based pharmacokinetic (PBPK) and pharmacodynamic (PBDK) models. These models are ideal for integrating in vitro and in vivo pharmacokinetic, mechanistic, and toxicological data of PAHs. As PAHs are suspected neurotoxicants, ingestion by children will have profound implications on the development of neu-roendocrine system. The magnitude of PAH insult to developing brain regions depends on the developmental stage and duration of exposure. Hence, there is an additional need for construction of PBPK models to study the disposition of orally ingested PAHs during pregnancy. These models will help understand the distribution, metabolism, and elimination of chemicals in both maternal and fetal systems.

Dietary habits will play a greater role in PAH intake. In this context, information on disposition of PAHs in different types of dietary fat, protein and carbohydrates is lacking. Humans are seldom exposed to individual PAH compounds, but mostly as complex mixtures. Hence, the acute and subchronic toxicity of PAH mixtures to laboratory animals at environmentally relevant levels would help provide an integrated picture on the relationship among exposure levels (doses), disposition in the body, and toxicity/carcinogenesis, and would be useful for risk assessment.