Download Bioavailability and Risk Assessment of Orally Ingested Polycyclic

ORIGINAL RESEARCH
Bioavailability and Risk Assessment of Orally Ingested
Polycyclic Aromatic Hydrocarbons
Aramandla Ramesh,1 Stormy A. Walker,1 Darryl B. Hood,1,2 Maria D. Guillén,3
Klaus Schneider,4 and Eric H. Weyand5
1
Department of Pharmacology and 2 Center in Molecular and Behavioral Neuroscience,
Meharry Medical College, Nashville, Tennessee, USA
3
Tecnologia de Alimentos, Facultad de Farmacia, Universidad del Pais Vasco, Vitoria, Spain
4
FoBig, Forschungs und Beratunginstitut Gefahrstoffe GmbH, Freiburg, Germany
5
Maple City Research, Inc., Hornell, New York, USA
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.
Keywords
Benzo(a)pyrene, Bioavailability, Dietary Exposure,
Metabolism, PAHs
Received 26 July 2004; accepted 27 July 2004.
The work presented in this review was supported in part by NIH
grants ES012168 to AR, GM08037 to AR and DBH, and NS41071,
ES00287, and RR03032 to DBH. Stormy Walker was supported
by a graduate research assistantship through the NIH-RISE grant
2R25GM59994.
Address correspondence to Dr. Aramandla Ramesh, Department of
Pharmacology, Meharry Medical College, 1005 D.B. Todd Boulevard,
Nashville, TN 37208, USA. E-mail: [email protected]
International Journal of Toxicology, 23:301–333, 2004
c American College of Toxicology
Copyright ISSN: 1091-5818 print / 1092-874X online
DOI: 10.1080/10915810490517063
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, carbonelectrode, 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
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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. Catacondensed 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 SN 2 reaction or, (after forming a carbonium ion) in an SN 1 reaction (Guillén, Sopelana, and Partearroyo
1997). Oxidative metabolism of diols such as B[a]P 7,8-diol can
FIGURE 1
Different types of PAH structures (Guillén and Sopelana 2003; copyright permission obtained from CAB International).
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
FIGURE 2
Regions related to biological activity in PAHs (Guillén and
Sopelana 2003; copyright permission obtained from CAB
International).
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
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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
FIGURE 3
Pathways of benzo(a)pyrene metabolism and possible effects (Guillén and Sopelana 2003; copyright permission obtained from
CAB International).
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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
1 7,12-Dimethylbenz[a]anthracene (DMBA) is not an environmental contaminant. It is a potent synthetic PAH compound (methyl-substituted PAH) used
extensively as a model in carcinogenicity, toxicity, and bioavailability studies.
(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 flavincontaining 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
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
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
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the logarithm of a PAH compound’s octanol/water partition coefficient [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 casecontrol 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 detoxification 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
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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, 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, 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 cellu-
lar 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-methylcholanthene 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 concentrationdependent 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 Cγ 1
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 reuptake of Ca2+ into the endoplasmic reticulum (ER) by the sarcoendoplasmic 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
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
307
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-molecularweight 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 H2 O2 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 cHa-ras in cultured rat through Ahr signaling and oxidative stress
(Zhao and Ramos 1998).
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.
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
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
308
A. RAMESH ET AL.
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. Broadleaved 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.
309
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
TABLE 1
PAH concentrations (µg kg−1 dry weight in grain, fluor, bran, and coffee; µg kg−1 wet weight in vegetables and fruits;
µg kg−1 in oils) in foods of plant origin∗
Compound
Acenaphthylene
Phenanthrene
Anthracene
1-Methylphenanthrene
2-Methylphenanthrene
2-Methylanthracene
4,5-Methylphenanthrene
9-Methylanthracene
Fluoranthene
Pyrene
1-Methylpyrene
Benzo[ghi]fluoranthene
Cyclopenta[c,d]pyrene
Benz[a]anthracene
Chrysene
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo[e]pyrene
Benzo[a]pyrene
Perylene
Indeno[1,2,3-cd]pyrene
Dibenz[a,c]anthracene
Dibenz[a,h]anthracene
Benzo[ghi]perylene
Coronene
Wheat Wheat
Olive Sunflower Soybean Groundnut Coconut
Vegetables Fruits grain flour Bran Coffee oil
oil
oil
oil
oil
—
2.2
0.1
0.0
0.0
—
—
n.d.a
1.3
0.3
—
—
—
—
—
—
—
0.2
0.1
0.0
—
n.d.
n.d.
—
—
0.14
0.36
0.02
—
—
—
—
—
0.037
0.034
—
—
—
0.014
0.025
0.016
0.014
—
0.014
—
0.014
—
0.014
0.014
—
—
—
—
—
—
—
—
—
0.6
0.4
—
0.0
0.1
0.2
0.8b
0.6c
—
0.3
0.3
—
0.3
—
—
0.3
0.1
—
—
—
—
—
—
—
—
0.2
0.5
—
—
—
0.1
0.1
0.0
0.1
0.2
0.1
n.d.
0.1
—
—
0.1
—
—
—
—
—
—
—
—
—
0.7
0.1
—
—
—
0.7
0.8
0.3
0.5
0.4
0.4
—
1.1
—
—
0.5
—
—
—
—
—
—
—
—
—
8.0
8.1
—
—
—
0.3
1.8b
1.9c
—
0.8
0.9
0.3
0.5
0.1e
—
0.6
—
—
15
0.9
—
—
—
—
—
4.2
5.0
—
—
—
0.2
0.5
0.1
0.1
—
0.1
—
0.2
—
—
0.0
—
4.4
2.3
0.0
—
—
—
—
—
6.7
5.0
—
—
—
3.1
1.7b
2.2
2.0d
4.1
1.5
0.6
1.3
0.0e
—
1.7
0.3
0.9
2.2
2.1
—
—
—
—
—
8.9
2.6
—
—
—
22
17
25
28d
25
28
10
23
4.7e
—
17
2.1
—
—
—
—
—
—
—
—
17
7.2
—
—
—
79
63b
85
99d
88
106
36
81
13e
—
66
7.4
—
2.8
0.3
—
—
0.6
1.5
—
18
20
3.6
—
—
1.3
4.1b
0.7
—
0.4
0.2
<0.1
<0.1
—
—
<0.1
—
∗
Fruits (Falcó et al. 2003); Wheat grain (Jones et al. 1989); Wheat fluor (Dennis et al. 1991); Bran (Dennis et al. 1991); Coffee (Klein, Speer,
and Schmidt 1993); Vegetables (Tateno, Nagumo, and Suenaga 1990); Olive oil (Moret and Conti 2000); Sunflower oil (Kolarovic and Traitler
1982); Soybean oil (Kolarovic and Traitler 1982); Groundnut oil (Kolarovic and Traitler 1982); Coconut oil (Larsson et al. 1987).
a
n.d. = not detected.
b
Concentration of chrysene + triphenylene.
c
Concentration of benzo[b]fluoranthene + benzo[ j]fluoranthene + benzo[k]fluoranthene.
d
Concentration of benzo[ j]fluoranthene + benzo[k]fluoranthene.
e
Concentration of dibenz[a,c]anthracene + dibenzo[a,h]anthracene.
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
310
A. RAMESH ET AL.
TABLE 2
PAH concentrations (µg kg−1 dry weight) in oyster, mussel, and fresh and smoked fish samples∗
Oyster Catfish Mussel Fish1
Compound
Methylnaphthalene
Dimethylnaphthalene
Acenaphthylene
Phenanthrene
Anthracene
Fluoranthene
Pyrene
Benz[a]anthracene
Chrysene
Triphenylene
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo[ j]fluoranthene
Benzo[e]pyrene
Benzo[a]pyrene
Perylene
Indeno[1,2,3-cd]pyrene
Dibenz[a,c]anthracene
Dibenz[a,c]anthracene
Benzo[ghi]perylene
Anthanthrene
—
—
—
76
30
680
410
230
260
—
53
161
—
—
31
—
—
—
—
41
—
6.0
—
100.0
—
270.0
—
2700.0 10
—
1.1
1800.0
5.7
1500.0
4.7
22
1.8
—
5.4
—
—
—
2.8d
—
—
—
—
14
2.2
7.0
0.7
8.0
1.0
—
1.2
—
0.1d
—
—
—
1.8
—
—
—
—
—
19
0.8
26
35
0.5
2.1c
—
1.7d
—
—
0.8
1.3
0.0
0.0
0.0d
—
0.0
—
Fish2
Pork
Beef
Lamb
Eggs Cow milk sausage3 meat4 Frankfurters5 sausage6
—
—
—
—
—
—
62
19
—
30
21
1.2
120
5.5
4.5 4.5
290
5.6
—
—
30
3.5
1.0 4.5
—
—
—
—
48.0 7.5
—
—
—
8.7
8.0 —
—
—
200
1.2
—
—
—
—
—
3.0
0.5
3.4
36
2.4
8.6
—
3.1
n.da
—
n.da
1.5
—
n.d.
—
4.7
0.0
—
—
—
1.3
—
1.4
1.8
2.0
0.1
0.51
—
0.1
0.1
—
—
n.d.
—
n.d.
—
n.d.
n.d.
—
—
—
—
—
—
11
1.3
0.5
25
—
1.2
0.6
—
—
1.4
—
—
—
1.5
0.0
—
—
—
—
168
35
120
127
45
44b
—
30
42c
—
22
54
7.9
41
—
3.5d
36
15
—
—
14
18
1.4
3.8
4.4
0.44
1.5
—
0.34
0.48
—
—
0.32
—
0.30
n.d.
—
0.45
—
∗
Oyster (Sanders 1995); catfish (Vassilaros et al. 1982); mussel (Baumard et al. 1998); fish (1 Baumard et al. 1998; 2 Akpan, Lodovici, and
Dolara 1994); eggs (Husain et al. 1997; data in µg kg−1 wet weight); cow milk (Husain et al. 1997; data in µg kg−1 wet weight); pork sausage
(3 Mottier, Parisod, and Turesky 2000); beef meat (4 Lodovici et al. 1995); frankfurters (5 Larsson et al. 1983); lamb sausage (6 Mottier, Parisod,
and Turesky 2000).
1
Fresh fish; 2 smoked fish; 3 raw beef meat; 4 barbecued beef meat, 5 grilled Frankfurters.
a
n.d. = not detected.
b
Concentration of chrysene + triphenylene.
c
Concentration of benzo[b]fluoranthene + benzo[ j]fluoranthene + benzo[k]fluoranthene.
d
Concentration of benzo[ j]fluoranthene + benzo[k]fluoranthene.
e
Concentration of dibenz[a,c]anthracene +dibenzo[a,h]anthracene.
TABLE 3
Dietary intake of PAHs in various countries
Country
U.S.A.
Intake
(µg/person/day)
United Kingdom
Germany
0.16–1.6
0.04–0.06∗
0.12–2.8∗
3.7
0.02–0.04
Austria
Italy
Spain
Greece
The Netherlands
Sweden
28
3.0
6.3–8.4
1.6–4.5
5–17
0.08
∗
Reference
Santodonato, Howard, and Basu (1981)
Kazerouni et al. (2001)
Hattemer-Frey and Travis (1991)
Dennis et al. (1991)
State Committee for Polution Control (1992),
cited in IPCS (1998)
Pfannhauser (1991)
Lodovici et al. (1995)
Falco et al. (2003)
Voutsa and Samara (1998)
de Vos et al. (1990)
Larsson (1986), cited in Beckman Sundh,
Thuvander, and Anderson (1998)
Values reported were for benzo[a]pyrene concentrations only.
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
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 expo-
311
sure 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 1OHP, 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 [14 C]phenanthrene and [14 C]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, phenanthrene, 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
312
A. RAMESH ET AL.
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 isoforms 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 extrahepatic tissues albeit to a minor extent. The expression and ac-
tivity 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
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
313
sophageal 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.
FIGURE 4
Schematic representation of fate of ingested PAHs.
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 oe-
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 firstpass metabolism of ingested and absorbed PAHs in view of its
positioning in the anatomical system, the expression of drugmetabolizing 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.
314
A. RAMESH ET AL.
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.
315
As above
Male F344 rats
Male F344 rats
Male Wistar rats (Y)
Male Wistar rats (Y)
Male Syrian golden
hamsters
Male Wistar rats
Female CD rats
Female CD rats
Female CD rats
Male Wistar rats
Methylcholanthrene (MC)
B[a]P
B[a]P
B[a]P
B[a]P
B[a]P
Male F-344 rats
Male SD rats
Male Dunkin-Hartley
guinea pigs
Male SD rats
As above
As above
Female BALB/c mice
Male Wistar rats
As above
Rats
Rats
Rats
Rats
Male Wistar rats
As above
As above
Wistar rats
Female SD rats
As above
As above
Rats
2-nitrofluorine
B[a]P
B[a]P
1,8-dinitropyrene
Pyrene
As above
B[a]P
B[a]P
B[a]P
B[a]P
Fluoranthene
Pyrene
B[a]A
B[a]P
Phenanthrene
As above
As above
B[a]P, pyrene, chrysene,
and phenanthrene
1-nitropyrene
Naphthalene
B[a]P
Chrysene
B[a]A
Chrysene
Triphenylene
B[a]P
Female merino sheep
Animal used
B[a]P
PAH
0.20
4–15
NA
NA
NA
NA
20
20
20
1 mmol
0.07
As above
As above
NA
0.7–4.4 nmol
0.002
0.25
25
20
2 mg/rat
4 mCi/animal
0.22
71.0–143
71.0–143
71.0–143
10, 20, 50 µmol
1000
0.16–5.5
0.002
1000
10–20 µci (labeled)
+ 1 mg (unlabeled)
0–20 µci (labeled)
+ 1 mg (unlabeled)
0.37–3.7
Dose (mg/kg)
33% DMSO/66% corn oil
20% Emulphor/80% saline
Aged (1–30 days) clay soil
Aged (180 days) clay soil
Aged (1–30 days) sandy soil
Aged (180 days) sandy soil
Tween 80/isotonic saline (1:5)
As above
As above
Olive oil
5% gum acacia w/o soil
5% gum acacia + 0.5 g sandy soil
5% gum acacia + 0.5 g clay soil
Soil-spiked
Intraduoedenal infusion
Gavage
Gavage
Gavage
Gavage
Gavage
Commercial diet
33% DMSO/66% corn oil
10% emulsifer/90% olive oil
10% emulsifer/90% olive oil
10% emulsifer/90% olive oil
Gavage
Synthetic diet
Corn oil + commercial diet
Char-broiled hamburger
Synthetic diet
5 ml of toluene containing BaP
+ 100 g of Lucerene chaff
5 ml of toluene containing MC
+ 100 g of Lucerene chaff
Peanut oil
Vehicle administered
47
87
44–53
34
64–64
51
62
42
102
28
73
76
78
90
30
>80
19
13
20
76
57+
86.9
94
75
75
90
99.6
96.7
89
88.7
91.2
64–70
54–67
Value (%)
TABLE 4
Summary of absorption/bioavailability data for orally administered PAHs‡
(Continued on next page)
Möller, Rafter, and Gustafsson
(1987)
Foth, Kahl, and Kahl (1988)
Yamazaki and Kakiuchi (1989)
Shah, Rowland, and Combes
(1990)
Jacob et al. (1989)
Withey, Law, and Endrenyi (1991)
Goon et al. (1991)∗
Goon et al. (1991)∗
Goon et al. (1991)∗
Goon et al. (1991)∗
Lipniak and Brandys (1993)∗
Lipniak and Brandys (1993)∗
Lipniak and Brandys (1993)∗
Van de Wiel et al. (1993)
Kadry et al. (1995)
As above
As above
Stroo et al. (2000)
Grimmer et al. (1988)
Bartosek et al. (1984)
Bartosek et al. (1984)
Bartosek et al. (1984)
Jongeneelen, Leijdekkers, and
Henderson (1984)
Dutcher et al. (1985)
Bakke et al. (1985)
Bowes and Renwick (1986)
Hecht, Grabowski, and Groth
(1979)
As above
Rabache, Billaud, and Adrian
(1985)
As above
Mirvish et al. (1981)
West and Horton (1976)
West and Horton (1976)
Reference
316
0.07
Male Lewis rats
As above
As above
As above
As above
As above
Male SD rats
As above
Female B6C3F1 mice
Male F-344 rats
Female CD rats
Female B6C3F1 mice
Female DBA/2J mice
Female C57BL/6J mice
Female DBA/2J mice
Lactating alpine goats
As above
As above
Castrated pigs
Human
Humans
Humans
As above
As above
As above
As above
As above
B[a]P
Pyrene
Pyrene
B[a]P
B[a]P
B[a]P
B[a]P
Pyrene
Pyrene
B[a]P
Pyrene
Phenanthrene
B[a]P
B[a]P
Pyrene
Pyrene
Dose (mg/kg)
Olive oil ingested as a bolus
Prepared meal
99.5
5% gum acacia + 0.5 g
contaminated soil
As above
As above
Sunflower oil
As above
As above
20% emulphor: 80% isotonic
glucose solution
20% emulphor: 80% isotonic
glucose solution
Powdered rodent diet
containing manufactured gas
plant tar
Peanut oil
Diet containing coal tar
Basal gel diet containing coal
tar
As above
As above
As above
Vegetable oil
Vegetable oil
Vegetable oil
Milk (1000 ml) spiked with
labled B[a]P
Char-broiled hamburger
97.3
90
98.8
86
99.8
99.8
5.5
63
38
33,.
40
65–80
95
9–75
86–94
99.7
99.6
99.8
99.7
99.6
>90
90
Value (%)
Lamp back
Vehicle administered
Hecht, Grabowski, and Groth
(1979)
Viau et al. (1995)#
Viau et al. (2002)
As above
As above
As above
Grova et al. (2002)
Grova et al. (2002)
Grova et al. (2002)
Laurent et al. (2002)
Ramesh et al. (2001)∗
Weyand et al. (2002)
Nemeth and Weyand (2002)
Koganti et al. (1998)
Bouchard and Viau (1998)
As above
As above
As above
As above
As above
Bouchard and Viau (1997)
van Schooten et al. (1997)
Stroo et al. (2000)
Reference
‡ Studies that were conducted after 1975 and that reported the absorption or bioavailability data were included in this table. Where the administered concentration/dose of a PAH
is given per animal basis, either the animal body weight provided by the authors or the mean body weight of that strain of rodent from published literature was used to compute the
dose levels. Where necessary, data were not given in literature, the urinary and fecal excretion data of unchanged compound (represents as % of administered dose) was subtracted
from the initial dose to obtain a nearest estimate of absorption/bioavailability data. Several bioavailability studies in animal models and human volunteers did not provide the
kinetics data, but used DNA damage or one particular metabolite level in target tissues as end points. These studies were cited in the text appropriately. NA = data not available;
∗
where available, bioavailability data were used in lieu of absorption data.
+
Assuming that 57% of the administered dose was absorbed after 35% and 8% of the dose was excreted through feces and urine respectively.
#
Estimated on the basis of urinary excretion of 1-hydroxypyrene.
◦
On mg/mouse basis.
¶ On µg/mouse basis.
Total recovery of labeled compound was used for calculation of absorption data; parent compound and metabolite concentrations were not differentiated.
.
Portal absorption values.
0.007
0.0007–0.003
0.0001
140◦
42.3◦
52.4◦
2.5 × 106 Bq
2.5 × 106 Bq
2.5 × 106 Bq
50 µci
100
0.76–157
110◦
1.5, 5, 15, 50 and 100
µmol
0.12–430¶
0.018
0.007
0.07
0.034
0.0084
2, 6, 20 and 60
NA
Rats
Animal used
BaP, pyrene, chrysene,
and phenanthrene
Pyrene
PAH
TABLE 4
Summary of absorption/bioavailability data for orally administered PAHs‡ (Continued)
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
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
317
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 biotransformation 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).
318
A. RAMESH ET AL.
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 2000, 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 organochlorine 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
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
care centers that are old (Heudorf and Angerer 2001), located
in low-income areas (Wilson, Chuang, and Lyu 1999, 2000,
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 [14 C]phenanthrene
or [14 C]B[a]P in milk, the time-course plasma concentrations
319
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 triacylglycerols 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
320
A. RAMESH ET AL.
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).
AAF =
fraction absorbed from the environmental exposure
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 matrixspecific 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;
321
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
ingestion of soil, drinking water, homegrown produce, homegrown 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:
Ba =
tissue concentrations (mg/kg)
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
TABLE 5
Metabolism factors reported for some PAHs (Hofelt et al.
2001)
Compound
Animal model
Metabolism factor
Benz[a]anthracene
Benzo[a]pyrene
Pyrene
Rat
Mouse
Rat
0.001
0.004
0.003
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
322
A. RAMESH ET AL.
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 (Livingstone 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
TABLE 6
Incidence and tumor multiplicity of lung and forestomach tumour in female B6C3F1 mice fed coal tar mixtures (CTM1 and
CTM2) in a long-term feeding study (Culp et al. 1998)
CTM
concentration
in food (%)
0
CTM1
0.01
0.03
0.1
0.3
0.6
1
CTM2
0.03
0.1
0.3
a
B[a]P
concentration
in food (ppm)
B[a]P daily
dose/animal (mg day−1 )
B[a]P daily dose
(mg kg−1 day−1 )
Incidence
in forestomach (%)
Incidence
in lung (%)
TBA
(%)a
0
0
0
0/47 (0)
2/47 (4)
5/48 (10)
0.22
0.22
2.2
6.6
13.4
22
0.8
2.4
8
23.7
36.2
63
0.027
0.08
0.27
0.79
1.45
3.15
2/47 (4)
6/45 (13)
3/47 (6)
14/46 (30)
15/45 (33)
6/41 (15)
3/48 (6)
4/48 (8)
4/48 (8)
27/47 (57)
25/47 (53)
21/45 (47)
12/48 (25)
14/48 (29)
12/48 (25)
40/48 (83)
42/48 (88)
43/48 (90)
1.1
3.7
11.1
4
13.2
36.3
0.13
0.44
1.21
3/47 (6)
2/47 (4)
13/44 (30)
4/48 (8)
10/48 (21)
23/47 (49)
17/48 (35)
23/48 (48)
44/48 (92)
Numbers of tumor-bearing animals, calculated using individual animal data for tumors of the liver, lung, forestomach, and small intestine,
hemangiosarcomas, histiocytics sarcomas, and sarcomas of the mesentery, forestomach, skin, and kidney.
CTM1: coal tar mixture 1; CTM2: coal tar mixture 2.
323
Mice B6C3F1
Mice B6C3F1
Mice B6C3F1
Mice B6C3F1
Mice A/J
Mice A/J
Mice A/J
Mice A/J
∗
Coal tar mixture 1
Coal tar mixture 2
Coal tar mixture 1
Coal tar mixture 2
Manufactured gas plant residue
Manufactured gas plant residue
Coal tar paint
Coal tar paint
Type of PAH mixture
263% (20 PAHs)
250% (20 PAHs)
263% (20 PAHs)
250% (20 PAHs)
244% (20 PAHs)b
244% (20 PAHs)b
n.a.
n.a.
carcinogenic potency of a PAH mixture
carcinogenic potency of B(a)P as a single substance
Forestomach
Forestomach
Lung
Lung
Forestomach
Lung
Forestomach
Lung
Potency ratio =
Oral
Oral
Oral
Oral
Oral
Oral
Gavage
Gavage
Species stain Application Tumor site
1.2
0.7
>15a
>27a
1
>33c
1
>13d
Culp et al. (1998)
Culp et al. (1998)
Culp et al. (1998)
Culp et al. (1998)
Weyand et al. (1995)
Weyand et al. (1995)
Robinson et al. (1987)
Robinson et al. (1987)
Reference
b
No lung tumors after application of B[a]P alone.
Calculated with data from Goldstein et al. (1998) for the composition of the mixture.
c
Dose-response data do not allow for modeling: lung tumor incidence after exposure to 3 ppm B[a]P in manufactured gas plant residue is higher than after exposure to
98 ppm B[a]P alone.
d
Dose-response data do not allow for modeling: lung tumor incidence after exposure to 20 µg B[a]P (three times per week) in coal tar paint is higher than after exposure to
250 µg B[a]P alone (three times per week).
a
Complete carcinogenesis
Complete carcinogenesis
Complete carcinogenesis
Complete carcinogenesis
Lung adenoma assay
Lung adenoma assay
Lung adenoma assay
Lung adenoma assay
Bioassay
Sum of BAP
equivalents (number Potency
of PAHs analyzed)
ratio
TABLE 7
Carcinogenicity studies with BaP and PAH mixtures reanalyzed: Study details, sum of BAP equivalents (with relative potency values according to the US
EPA; Brown and Mittelman 1993), and calculated potency ratios∗ (Schneider et al. 2002)
324
A. RAMESH ET AL.
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 PAHrich 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
FIGURE 5
Comparison of potency ratios (carcinogenic potency of PAH
versus B[a]P in the same bioassay) with predictions based on
relative potency values (B[a]P equivalency factors). FS1 and
FS2 refer to forestomach tumors developed in B6C3F1 mice
that were administered with coal tar mixtures 1 and 2,
respectively.
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.
FIGURE 6
Potency ratios (potency of mixture divided by potency of
B[a]P) for oral exposure route and tumor locations. CTM1 and
CTM2 refer to coal tar mixtures 1 and 2, respectively.
BIOAVAILABILITY AND RISK ASSESSMENT OF PAHs
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 neuroendocrine 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.
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