[advances in molecular toxicology] volume 7 || genotoxicity of polycyclic aromatic hydrocarbon...

CHAPTER THREE

Genotoxicity of PolycyclicAromatic HydrocarbonMetabolites: Radical Cationsand KetonesSushmita Sen, Jeffrey M. Field1Department of Pharmacology and Center of Excellence in Environmental Toxicology, Perelman Schoolof Medicine, University of Pennsylvania, Philadelphia, Pennsylvania, USA1Corresponding author: e-mail address: [email protected]

Contents

1.

AdvaISSNhttp:/

Introduction

nces in Molecular Toxicology, Volume 7 # 2013 Elsevier B.V.1872-0854 All rights reserved./dx.doi.org/10.1016/B978-0-444-62645-5.00003-1

84
2. Exposure to PAHs 88 3. Exposure Risks of PAH 89 4. DNA Damage by PAH 90 5. PAH Metabolism to DNA-Reactive Metabolites 90 6. Diol-Epoxide Pathway 93 7. The Radical Cation Pathway 93 8. Depurinating PAH Adducts and AP Sites 97 9. Mutagenesis Studies to Address PAH Radical Cations 101
10.
Evidence for AP Site Formation Due to Radical Cations 102 11. Mutagenicity Studies with B[a]P Radical Cation Products 106 12. The Quinone Pathway 109 13. Exposure to OPAHs 111 14. Biological Relevance of OPAH 113 15. Conclusion 115 References 116
Abstract

Polycyclic aromatic hydrocarbons (PAHs) are widespread environmental pollutants, sowidespread that it is impossible for anyone to avoid exposure to them. They are theproducts of partial combustion, and exposure comes from the fossil fuels that weuse to drive our cars, cook our food, warm our home, and fuel our industry. Other expo-sure comes from tobacco smoke and oil spills. PAHs are responsible for more cancers,primarily lung cancers, than any other carcinogen. The most studied PAHs areunsubstituted multi-ring structures. These compounds are not DNA reactive, and mustbe modified to become carcinogenic. Metabolic pathways modify PAH during thedetoxification process. Three pathways have been extensively documented: the

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http://dx.doi.org/10.1016/B978-0-444-62645-5.00003-1

84 Sushmita Sen and Jeffrey M. Field

diol-epoxide pathway, the radical cation pathway, and the o-quinone pathway. SomePAHs are generated during combustion in oxygenated forms, usually as quinones. Qui-nones also form as PAHs are exposed to sunlight, for example, in an oil spill. These met-abolic products damage DNA through distinct mechanisms: diol epoxides form bulkyadducts with DNA, radical cations form depurinating adducts and quinones undergofutile redox cycling to generate reactive oxygen. Here, we discuss the evidence for thesepathways in PAH carcinogenesis.

1. INTRODUCTION

Polycyclic aromatic hydrocarbons (PAHs) are widespread environmen-

tal contaminants as a consequence of the ubiquitous use of fossil fuels. Their

major sources of origin are hydrocarbons and partial combustion products.

The hydrocarbon sources of PAH are coal tar, crude oil, creosote, and roofing

tar. The combustion sources are incomplete combustion of gas, coal, gasoline

and diesel fuel, organicwaste, charbroiledmeat, and tobacco. PAHs can persist

in the environment for years and can work their way through the food chain,

so ingestion of contaminated plants and animals presents as a health risk beyond

direct environmental exposure [1]. Themajor toxicity of PAH is to cause can-

cer through mutagenesis. PAHs, primarily through tobacco smoke, probably

cause more cancers than any other carcinogen. The toxic and mutagenic

effects of these compounds have been reviewed [2–9].

The carcinogenicity of PAHs was recognized as early as 1921 [10]. By

1933, benzo[a]pyrene (B[a]P) was isolated from coal tar and its high carci-

nogenic activity was reported [11]. Eventually, several other PAHs were

synthesized to use as models to understand carcinogenesis [12,13].Metabolic

activation of PAH was first noted by Miller et al. when B[a]P applied to

mouse skin was found to be covalently bound to proteins [14]. Miller pro-

posed the principle of carcinogenesis by chemical carcinogens in 1970 that

states that chemical carcinogens bind covalently to cellular macromolecules,

and the ultimate electrophilic species react with nucleophilic groups of

DNA, RNA, and proteins to initiate and propagate carcinogenesis

[15,16]. The International Agency for Research on Cancer (IARC) has clas-

sified some of these compounds such as B[a]P and dibenz[a,h]anthracene as

carcinogenic (Group 1) or likely carcinogenic (Group 2A) to humans. Based

on several decades of evidence, the United States Environmental Protection

Agency (USEPA) has designated 28 PAH compounds as priority pollutants

(Table 3.1). TheUSEPA has established safe exposure levels for many PAHs,

Table 3.1 USEPA priority list PAHs

Acenaphthene 7,H-Dibenzo[c,g]carbzole

Acenaphthylene 7,12-Dimethylbenz[a]

anthracene

Phenanthrene 3-Methylcholanthrene

Fluorene 5-Methylchrysene

Anthracene Benzo[j]fluoranthene

Benz[a]anthracene Dibenz[a,j]acridine

Chrysene Dibenz[a,h]acridine

Pyrene Benzo[r,s,t]pentaphene

1-Nitropyrene Dibenzo[a,e]pyrene

Fluoranthene Dibenzo[a,l]pyrene

Benzo[b]fluoranthene Dibenzo[a,e]fluoranthene

Benzo[k]fluoranthene Benzo[g,h,i]perylene

Benzo[a]pyrene Dibenz[a,h]anthracene

Indeno[1,2,3-c,d]

pyrene

Dibenzo[a,h]pyrene

85Genotoxicity of Polycyclic Aromatic Hydrocarbon Metabolites

Table 3.2 Screening levels for 1-year average exposure to PAHs commonly found incrude oilPAH Exposure limits (ng/m3)

Dibenz[a,h]anthracene �580

Benzo[a]pyrene �640

Benzo[b]fluoranthene �6400

Benzo[k]fluoranthene �6400

Indeno[1,2,3-c,d]pyrene �6400

Benz[a]anthracene �6400

Naphthalene �30,000

Chrysene �64,000


based on their health effects, which take into account levels that might pose

an increased risk of cancer. Table 3.2 shows the amounts of each PAH above

which the risk of adverse health effects could increase.

USEPA also monitors and screens for PAH levels during accidental oil

spills. PAHs are among the most toxic compounds contained in crude

oil. On April 20, 2010, an explosion on the Deepwater Horizon drilling

platform at the Macondo oil well in the Gulf of Mexico, operated by BP

Oil, caused the largest offshore oil spill in US history. The explosion resulted

in the release of an estimated 5.3�108 kg of oil (4.9 million barrels) and

1.7�108 kg of gas from the Macondo well. To evaluate the exposure

and subsequent health effects of the Gulf oil spill, USEPA scientists are com-

paring PAH in air samples to health-based screening concentrations (also

called “screening levels”) in exposed populations. PAHs are also being mon-

itored in fish and crustaceans in the Gulf coast. At the peak of the disaster,

36% of the federal fishing areas were closed; reduced fish catches are being

reported in affected areas as of this writing (May 2013).

In an oil spill, PAHs are found not only in the crude oil but also in the

smoke plume formed when oil is burned as a part of the cleanup procedures

to contain the spill. Furthermore, the crude oil in a spill changes with time.

Some of the most volatile compounds evaporate soon after the spill and

other compounds evaporate slowly over weeks to months. With time

though, many of the compounds can travel over a distance. The oil that

reaches the shore is commonly known as the “weathered oil.” In weathered

oil, the lighter PAHs have evaporated, while the remaining PAH can

become oxygenated so that the oxygenated PAH (OPAH) eventually makes


up a significant fraction of the total PAH. The PAHs measured by USEPA

when sampling contaminated sites, such as the Deepwater Horizon oil spill,

are given in Table 3.1. However, OPAHs are not routinely monitored and

they are not studied as frequently.

OPAHs are PAHs with one or more carbonyl oxygen(s) attached to the

aromatic ring structure. They can be classified as a ketone, quinone, or anhy-

dride. Figure 3.1 shows structures and names of some of the representative

OPAHs. OPAHs are abundant in the environment and are generated from

the same sources as PAHs, because they are also products of incomplete

combustion [17–20]. OPAHs are also formed by oxidation of PAHs as they

weather in the environment. PAHs can be oxidized to corresponding

OPAHs through chemical oxidation, photooxidation, or biological trans-

formation [21,22]. Chemically, environmental PAHs can react with oxi-

dants such as singlet oxygen, peroxides, peroxy radicals, and hydroxy

Figure 3.1 Representative environmental oxygenated polycyclic aromatic hydrocarbons(OPAHs).


radicals to generate OPAHs [23]. Photooxidation of PAHs involves absorp-

tion of ultraviolet (UV) light and reaction of the activated PAHwith ground

state oxygen to ultimately produce OPAH [22,24]. During biodegradation,

PAHs can be converted into OPAH either by enzymatic reactions carried

out by microorganisms in the environment or by P450-mediated metabolic

pathways in cells [23,25]. The carbonyl groups in the OPAHs make them

more polar andmore water soluble compared to the parent PAH. This prop-

erty of OPAHs increases their tendency and capability to spread to the sur-

roundings through soil, water, and moisture in the air. The role of OPAHs

in oxidative stress and DNA damage will be discussed in the later sections of

this chapter.

2. EXPOSURE TO PAHs

PAHs have several routes of exposure, including breathing polluted air,

contaminated food, direct skin contact with PAH-contaminated soils, heavy

oils, but the greatest risk factor is exposure to tobacco smoke. Tobacco smoke

is the major cause of lung cancer in the world and estimates of its health toll

range as high as 6 million deaths per year [26–29]. Over 500 PAHs have been

identified in tobacco smoke and are considered to be a major carcinogen in

tobacco smoke [29,30]. The generation of mouse skin tumors after exposure

to cigarette smoke is suggested to be mediated by PAHs [7,31].

Chinese women living in XuanWei County are mostly nonsmokers, but

have the highest lung cancer rates in China [32,33]. Their cancer rates are so

high because they use smoky coal for cooking and can inhale 10 times as

much PAH as a 20-cigarette-per-day active smoker, from air concentrations

of PAHs approaching levels experienced by workers on the topside of coke

ovens. Several studies conclude that the excess lung cancer in this region is

caused by PAHs derived from smoky coal exposure [34,35].

Air pollution is another source of PAH, primarily from combustion pro-

cesses such as burning coal or oil for heating, emissions from power plants

and other industrial sources, burning of garbage, volcanic emissions, and for-

est fires. In urban areas, PAH exposure is predominantly attributed to diesel

exhaust from automobiles and operation of machines run by oil. Urban par-

ticulate phase PAH concentrations have been measured by filter sampling

studies [36–38]. The PAH concentrations are generally in the range

0.1–30 ng/m3, but these concentrations vary greatly between sampling sites

and over time at the same sites.


As mentioned in the earlier sections, oil spills such as the BP oil spill can

lead to exposure through contaminated air, water, and soil [39]. Several

aquatic life forms have been affected by the oil spill and can enter the food

chain increasing the risk of exposure. Oil spill remediation processes such as

burning the oil can generate PAH products such as OPAHs that can persist

in the environment over a long period of time, so legacy accumulations pre-

sent a risk.

Someworkers are highly exposed to PAHs, including those in aluminum

production, coal gasification, coke production, iron and steel production, tar

distillation, road paving, roofing, and shale oil extraction.

3. EXPOSURE RISKS OF PAH

The major risk of PAH exposure, primarily through exposure to

tobacco smoke, is lung cancer. But numerous other respiratory diseases

are likely to be caused by PAH exposure, including asthma, chronic obstruc-

tive pulmonary disease, and chronic bronchitis. Lung cancer develops

through a progression of pathological mutations in the respiratory epithe-

lium. While tobacco exposure is the primary risk factor, many studies sug-

gest that an increase in lung cancer risk is due to exposure to urban air

pollutants. For research purposes, particulate matter (PM) is divided up into

inhalable particles (PM10): fine particles (PM2.5) and ultrafine particles

(PM0.1). Lung cancer risk in particular is believed to be due to exposure

to PM of size 10 or 2.5 [40,41]. A major source of PM is diesel exhaust,

which was recently declared a lung cancer carcinogen [42].

Numerous in vitro and in vivo studies of the PAHs in the PM of polluted

air report lung toxicity, lung inflammation, genotoxicity, and rodent carci-

nogenicity [43–54]. There are data from cell free systems and tissue culture

experiments that show these PM (both from diesel exhaust and urban air)

cause oxidative DNA damage, mainly single-strand breaks (SSBs) and

8-oxo-dG adducts [55]. In vivo experiments in rodents have demonstrated

that diesel exhaust particles induce oxidative DNA damage in lung tissues,

even at low doses [50,56–59].

The role of PAHs in tumor initiation has been investigated in carcinoge-

nicity experiments, in metabolic studies, and by the identification and quan-

tification of DNA adducts. There is a range in the biological activity of

different PAH and their metabolic products, so the mutagenicity of a sample

depends greatly on the PAH present. Among all the PAHs, the toxicity and

mutagenicity of benzo[a]pyrene (B[a]P) are the most widely studied [60,61].


The IARC has determined B[a]P to be carcinogenic to humans. B[a]P cau-

sed significant mutations in human cells at concentrations of 14.9 ng/mL

while pyrene showed no mutagenic activity at 100,000 ng/mL. In 1979,

an estimated 5000 tons of B[a]P was released globally. B[a]P will be discussed

mostly in this chapter as a representative PAH.

4. DNA DAMAGE BY PAH

To establish mechanisms of DNA damage by suspect carcinogens the

type of DNA damage observed in populations must be correlated with the

chemistry of the carcinogens. In some cases, the relationship of PAH–DNA

adducts correlate with oncogenic mutations [62,63].

PAH exposure causes two types of DNA damage: bulky adducts and oxi-

dative damage. Bulky adducts are formed by the covalent attachment of

PAH to DNA bases. Oxidative damage is caused by the generation of reac-

tive oxygen species (ROS) from the futile redox cycling of PAH metabo-

lites. The oxidative damage leads to the formation of 8-oxo-dG, the most

mutagenic oxidative lesion [64]. A third type of adduct is the formation

of apurinic sites (AP sites), which arise when unstable bulky adducts are lost

from DNA. While bulky adducts and oxidative damage have been exten-

sively documented, AP sites are observed less frequently.

PM, both from diesel exhaust and urban air, cause oxidative DNA dam-

age, mainly SSBs and stimulate 8-oxo-dG adducts (reviewed in Ref. [55]).

In rodents, diesel exhaust particles induce oxidative DNA damage in lung

tissues, even at low doses [50,56–59]. PM also cause bulky adducts

[52,65]. A relatively new PAH, 3-nitrobenzanthrone (3-NBA), is highly

mutagenic and can form bulky adducts as well as redox cycle [66–68].

Tobacco exposure also causes oxidative damage and reduces the levels of

antioxidants in blood [69].

5. PAH METABOLISM TO DNA-REACTIVEMETABOLITES

Metabolic activation of chemical carcinogens to more reactive com-

pounds is usually required for mutagenesis and subsequent carcinogenesis

since the parent PAHs are generally inert chemicals that do not damage

DNA. Damage is caused by PAHmetabolites created as PAHs are converted


from hydrophobic compounds into relatively hydrophilic compounds dur-

ing detoxification. Detoxification begins with oxidation by P450 to produce

hydroxylated compounds that can be either eliminated by conjugation with

phase III enzymes or further activated to form highly electrophilic

species [25]. The electrophilic species are capable of interacting with nucleic

acids to form aDNA adduct, whichmay be unstable, or it may be incorrectly

repaired leading to a mutation later. Damage can also be caused by initiating

a futile redox cycle that generates ROS.

A number of criteria must be met to establish that specific metabolic

pathways are carcinogenic. (1) Biomarker metabolites must be observed

in exposed populations; (2) DNA adducts of specific metabolites must be

detectable in the exposed populations; (3) suspect metabolites must

be carcinogenic; (4) metabolites must be mutagenic (or tumor promoters);

(5) preventing metabolites from forming must reduce mutagenesis;

(6) metabolites must leave a “signature” in DNA similar to the one seen

in exposed populations. There are multiple reactive metabolites created

by the breakdown of PAH through different enzymatic pathways. Although

none of these metabolites meet all of the criteria in each experimental

system, there are some that have come close to meeting the criteria.

PAHs are metabolically activated via three major pathways in cells,

namely: (1) the diol-epoxide pathway, (2) the radical cation pathway, and

(3) the aldo–keto reductase (AKR) (or o-quinone) pathway. Figure 3.2 illus-

trates the three metabolic pathways in the representative PAH B[a]P. In the

diol-epoxide pathway, activation by P450 monooxygenases catalyzes the

formation of highly mutagenic bay-region diol epoxides, while in the radical

cation pathway, one-electron oxidation by P450 peroxidases leads to short-

lived and reactive PAH radical cations. In the o-quinone pathway, activation

of intermediate dihydrodiols by AKRs leads to formation of redox-active

o-quinones [70].

Establishing the relative contribution of these pathways to carcinogenesis

has been the topic of major research efforts because of their relevance to can-

cer, especially lung cancer. If one pathway is found less mutagenic, then it

may be possible to intervene pharmacologically to direct metabolites to that

pathway, which may reduce the risk for cancer from PAH exposure. Each of

the three metabolic pathways has been documented in cells, and in most

cases treated animals. A quantitative analysis of all of the pathways was car-

ried out by Lu et al., who found that all three pathways are metabolizing

benzo[a]pyrene (B[a]P) simultaneously in lung cells [71].

Figure 3.2 PAH metabolic pathways in representative benzo[a]pyrene.


6. DIOL-EPOXIDE PATHWAY

The most extensively documented intermediate is the anti-

diol-epoxide of benzo[a]pyrene ((�)-anti-BPDE) [72]. It is formed by the

combined action of P4501A1/P4501B1 and epoxide hydrolase. Initially,

B[a]P is converted to the (�) trans-7,8-dihydroxy-7,8-dihydrobenzo[a]

pyrene (B[a]P-7,8-diol), which is subsequently converted to (þ)-anti-7,8-

dihydroxy-9R,10R epoxy-7,8,9,10-tetrahydrobenzo[a]pyrene, (þ)-anti-

BPDE (Fig. 3.2). (�)-anti-BPDE is mutagenic in a number of paradigms

and can form stable DNA adducts. If these adducts are unrepaired, trans-

lesional DNA synthesis by error-prone bypass polymerases can introduce

an A opposite the bulky adduct, which upon replication, leads to a G to

T transversion.

Some of the strongest data documenting (�)-anti-BPDE metabolites as

carcinogens were obtained using the technique of ligation-mediated PCR

(LMPCR). In LMPCR, adduct-specific nucleases are used to cleave

DNA at the sites of adducts; cleavage is followed by ligation-assisted

PCR to amplify the DNA. The codons where the original adducts formed

can then be mapped by running the products on high-resolution gels. In

analyzing p53, one of the most common genes mutated in lung cancer,

Pfeifer and colleagues found that (�)-anti-BPDE formed adducts at pre-

ferred codons. Remarkably, the preferred codons, 157, 158, 245, and

273, corresponded to cancer “hot spots” mutated on p53 in patients with

lung cancer [73]. Although this made a compelling story that (�)-anti-

BPDE is the major carcinogen in cigarette smoke, the conclusion has been

challenged for a number of reasons. Firstly, many nonsmoke accessible can-

cers such as colon cancer also have the same hotspots. Secondly, other sus-

pect carcinogens, includingROS form adducts at these sites.While there are

some challenges to the paradigm that (�)-anti-BPDE forms adducts at spe-

cific sites, the case for this metabolite as an ultimate carcinogen is the stron-

gest for any PAH metabolite [74].

7. THE RADICAL CATION PATHWAY

Miller’s hypothesis of chemical carcinogenesis mediated by covalent

binding of PAH to cellular macromolecules led Wilk, Benz, and Rochlitz

to postulate that radical cations arising from one-electron oxidations of

PAH were electrophilic species capable of reacting with cellular

Figure 3.3 Mechanism of radical cation formation of representative PAH benzo[a]pyrene. Modified from Ref. [85].


nucleophiles such as DNA and proteins [75]. Others soon developed the

mechanism further [76–80]. In the radical cation pathway, cytochrome

P450 peroxidases, H2O2-dependent peroxidases, and prostaglandin

H synthase [81–84] catalyze one-electron oxidations of PAHs to generate

highly reactive PAH radical cations (Fig. 3.3).

Radical cations can also be generated in vitro by chemical, electrochemical,

and enzymatic reactions. PAH radical cationswere generated chemically using

chemical oxidants such as ferric iron III iodine in pyridine [79] and manganic

acetate in acetic acid [86]. The radical cations generated chemically were

trapped with nucleophiles to generate adducts. When B[a]P was reacted with

2 equivalents (eq) of manganese acetate, the predominant product was

6-acetoxy-B[a]P (via 6-hydroxy-B[a]P) along with minor B[a]P diones.

Figure 3.4 illustrates the radical cation pathway and end products for the rep-

resentative PAH B[a]P. For the most potent environmental carcinogen,

dibenzo[a,l]pyrene (DB[a,l]P), only 10-acetoxy-DB[a,l]P was obtained

[87–90].

Several studies showed that PAH radical cations could form covalent

adducts with DNA in vitro. An initial concern with the radical cation theory

was the absence of the monooxygenase enzyme system outside the endo-

plasmic reticulum, but the monooxygenases were later found in nuclear

membranes [91,92]. Cavalieri et al. showed that PAH covalently bound

to DNA in the presence of rat liver nuclei, which they hypothesized

occurred via one-electron oxidation [93]. Subsequently, using tritiated B

[a]P, they found one-electron oxidation in rat liver microsomes, rat liver

nuclei, and mouse skin [94].

The chemical and structural properties of PAHs determine the likeli-

hood of formation and stability of PAH radical cations. The two important

features of PAH for radical cation formation are the ionization potential (IP)

and the charge localization of the resulting radical cation. Formation of PAH

radical cations requires a low IP. The proposed cutoff IP above which one-

electron oxidation cannot occur is 7.35 eV. This cutoff IP value was derived

Figure 3.4 Benzo[a]pyrene radical cation pathway.


by testing DNA-binding capacities of series of PAHs with IP ranging from

6.68 to 8.19 eV in the presence of horseradish peroxidase (HRP) or pros-

taglandin H synthase [95]. Above the cutoff IP, activation by one-electron

oxidation is unlikely due to the difficulty in removal of an electron from a

more stable carbon–hydrogen bond, leading to a less stable radical cation.

Activation of PAH via one-electron oxidation is not only dependent on

the IP value of the PAH but also the charge localization capacity of the PAH

radical cation. Due to the fused nature of PAH benzene rings, delocalization

of pi electrons found in benzene is reduced. This results in positions of

unequal electron distribution, therefore, not every carbon in a PAH is able

to sustain a high electron density. The extent of unequal charge distribution

depends on the overall structural symmetry of the PAH and typically meso-

anthracenic positions can hold high electron density in a neutral PAH. In

brief, when one electron is removed to form the radical cation, the position

with the highest electron density in the neutral molecule usually is the posi-

tion where the positive charge of the radical cation can be found. For exam-

ple, in B[a]P, C-6 is the position with the highest electron density in the

neutral molecule and therefore is the only position where a radical cation

is formed. B[a]P also has a relatively low IP (7.23 eV) and therefore is readily

metabolized into a radical cation. Unlike B[a]P, the charge of the radical

cation for DB[a,l]P is mainly localized at the C-10 position.

Lehner et al. found that there is generally a poor correlation between a

carcinogen’s ability to form radical cations and its mutagenicity. For exam-

ple, anthracene and pyrene form radical cations easily, but are poor carcin-

ogens [96]. Competingmetabolic pathways in different tissues, in part, could

explain this lack of correlation. IP should distinguish pathways, as PAHwith

IPs above 7.35 eV should be activated primarily by the diol-epoxide path-

way and should be carcinogenic in mouse skin that is rich in P450 mono-

oxygenases. PAHs with IPs of 7.35 or below can be activated by both

mechanisms and should be carcinogenic in both mouse skin and rat mam-

mary gland (mostly rich in P450 peroxidases). In carcinogenicity studies,

PAHs that have high IPs, such as 5-methylchrysnene (7.73 eV) and

dibenz[a,h]anthracence (7.61 eV), were found to be potent carcinogens in

mouse skin and inactive in rat mammary glands, whereas 7-methylbenz[a]

anthracence (7.37 eV) with an IP value closer to the cutoff value is a potent

carcinogen in mouse skin and also weakly active in rat mammary gland

[97,98]. The results of the carcinogenicity studies corresponding with the

electron potential of some selected PAHs are listed in Table 3.3.

Figure 3.5 shows the relative carcinogenicity of the three widely studied

Table 3.3 Carcinogenicity of selected PAHs in mouse skin (repeated topicallyapplication) and rat mammary gland (injection) and their corresponding ionizationpotentials

PAHIonizationpotential (eV)

Carcinogenicity

Mouse skinRat mammarygland

1,2,3,4-Tetrahydro-7,12-

dimethylbenzanthracene

6.94 Moderate Extreme

3-Methylcholanthrene 7.12 High High

7,12-Dimethylbenzanthracene 7.22 Extreme Extreme

Benzo[a]pyrene 7.23 High Moderate

Dibenz[a,l]pyrene 7.27 Extreme Extreme

Benzo[a]pyrene-7,8-dihydrodiol 7.49 High Poor

Benzanthracene 7.54 Poor Poor

Figure 3.5 Relative carcinogenicity based on ionization potentials and structuresof PAHs.


potent carcinogens with respect to their IPs and chemical structure. It can be

concluded from the above data that PAHs with high IPs are not efficient

in generating radical cations and also may not be able to regioselectively

localize the charge in their radical cation forms. Therefore, activation of

the diol-epoxide pathway rather than the radical cation pathway to generate

carcinogenicity seems more logical for these PAHs.

8. DEPURINATING PAH ADDUCTS AND AP SITES

Radical cations are electrophilic in nature, capable of interacting with

nucleophilic centers of macromolecules such as DNA and proteins to form

covalent adducts. There are two types of DNA adducts: stable DNA adducts,

which remain incorporated in the DNA unless repaired, and unstable

Table 3.4 Formation of PAH depurinating adducts in vitro and in vivoRadical cation DNAadducts

Amount formed with ratliver microsomes (in vitro) (%)

Amount formed inmouse skin (in vivo) (%)

B[a]P-6-C8Gua 12 34

B[a]P-6-N7Gua 10 10

B[a]P-6-N7Ade 58 30

BPDE-10-N7Ade 0.5 <0.05

BPDE-10-N2dG 15 22

7-MBA-12-CH2-N7Gua 17 20

7-MBA-12-CH2-N7Ade 82 79

DB[a,l]P-10-C8Gua 6 Not tested

DB[a,l]P-10-N7Gua 2 Not tested

DB[a,l]P-10-N7Ade 14 Not tested

DB[a,l]P-10-N3Ade 28 Not tested

(þ/�)-anti-DB[a,l]PDE-

14-N7Gua

3 Not tested

(þ/�)-syn-DB[a,l]PDE-

14-N7Ade

31 Not tested


depurinating DNA adducts, which are spontaneously released fromDNA by

cleavage of the glycosidic bond. Cavalieri et al. provided evidence that PAH

radical cations bind to N7-, N3-, or C8-positions of guanine and adenine to

form DNA adducts both in vivo and in vitro as illustrated in Table 3.4 [99].

These DNA adducts are unstable because of the presence of a labile

N-glycosidic bond, resulting in spontaneous depurination to yield AP sites

within the sugar–phosphodiester backbone.

Themechanism of formation of an AP site via radical cations involves the

following steps: (1) addition of an electrophile on the N-7 of Gua in the

DNA gives rise to a positive charge on the N atom, which makes the highly

electronegative N atom highly unstable; (2) the adjacent C–N bond cleaves

most likely by participation of one pair of nonbonding electrons of the oxy-

gen in the deoxyribose (sugar) ring to form an intermediate oxocarbenium

ion intermediate and a neutral Gua-electrophile adduct; (3) the

oxocarbenium ion adds a water molecule, giving deoxyribose as the end


product. The AP sites can be mutagenic when DNA polymerases replicate

the DNA. The overall mechanism is depicted in Fig. 3.6.

According to Cavalieri et al., formation and inefficient repair of these AP

sites are responsible for the carcinogenic and mutagenic effects of PAHs such

as B[a]P, DMBA, and DB[a,l]P both in vitro and in mouse skin. The G to

T and A to T transversions could occur by DNA misreplication across

AP sites with adenine being inserted most frequently opposite to the lesion.

During the next round of DNA replication, the most likely base to be

inserted opposite to the AP site is A [101]. When the coding strand of

the DNA is then replicated, a T is inserted opposite to the new A. This

results in the G to T transversion. When an adenine is lost by depurintion,

the preferential insertion of A in the opposite DNA strand leads to an A to

T transversion at the AP site.

PAH–DNA adducts formed in vitro by PAH radical cations have been

identified and quantified using a combination of HPLC and fluorescence

line-narrowing spectroscopy. Out of the four DNA bases, adenine and gua-

nine are most susceptible to electrophilic attack compared to thymine and

cytosine. As mentioned earlier, covalent binding to adenine is usually at

N-3 and N-7 positions whereas N-7 and C-8 positions of guanine are

Figure 3.6 Mechanism of AP site formation. Modified from Ref. [100].


the most reactive. These adducts are unstable and eventually depurinate,

which generates AP sites.

B[a]P-6-N7Gua was the first depurinating adduct identified in vitro using

HPLC and mass spectrometry after activation of B[a]P by P450 and HRP in

the presence of DNA [102]. Eventually, profiles of all stable and unstable

adducts formed after activation of B[a]P by P450 and in mouse skin were

documented [103–105]. According to Cavalieri et al., activation of B[a]P

in mouse skin formed both stable (29%) and depurinating (71%) adducts.

The depurinating adducts that were formed as illustrated in Fig. 3.7 are

B[a]P-6-C8-Gua (34%), B[a]P-6-N7-Gua (10%), B[a]P-6-N7-Ade

(22%), BPDE-10-N7-Gua (2%), and BPDE-N7-Ade (3%). Only 23% of

the total adducts were the ultimate carcinogen diol-epoxide BPDE

Figure 3.7 Depurinating PAH–DNA adducts. Modified from Ref. [85].


(BPDE-10-N2-dG) and the remaining 6% were unidentified stable adducts.

They also found that B[a]P-7,8-dihydrodiol (precursor of BPDE) forms

more stable adducts (63%) in mouse skin than depurinating adducts (37%).

Another PAH, the highly potent carcinogen DB[a,l]P, was tested using

activated rat liver microsomes (containing mostly P450s). Analogously to

B[a]P, the major adducts formed were depurinating adducts (85%).

However, the depurinating adducts were formed almost equally by radical

cations and diol epoxides. Unlike B[a]P, a majority of the depurinating

adducts formed by DB[a,l]P were adenine adducts and not guanine adducts

(Fig. 3.7). Based on the results showing formation of depurinating adducts as

major PAH–DNA adducts, AP site formation was postulated to be the most

important mutagenic DNA lesion for carcinogenesis.

9. MUTAGENESIS STUDIES TO ADDRESSPAH RADICAL CATIONS

PAHcarcinogenicity via the radical cationpathwaywas studied inmouse

skin and rat mammary gland [89,90,106–108]. The authors used rat mammary

glands because, as mentioned earlier, the P450 peroxidases that catalyze the

one-electron oxidation of PAH radical cations are abundant in mammary

gland, whereas P450 monooxygenases are not. On the other hand, the P450

monooxygenases responsible for diol epoxides are found at higher concentra-

tions inmouse skin. In tumorigenicity studies of B[a]P, B[a]P-7,8-dihydrodiol

and anti-BPDE,B[a]P andB[a]P-7,8-dihydrodiol (the enantiomer that leads to

the highly carcinogenic anti-BPDE) were found to exhibit similar potency in

mouse skin, while anti-BPDE was found to be less potent. These results indi-

cated that the tumorigenicity of B[a]P and its metabolites had no relationship

with amounts of stable adducts formed since the less potent anti-BPDE formed

the highest amounts of stable adducts. The authors thus concluded that the diol

epoxides were not the exclusive pathway of activation for B[a]P and that the

radical cation from B[a]P played a more important role in tumor initiation.

In another approachbyCavalieri et al., a fluoro substituentwasused toblock

the bay region formation of the diol epoxides in B[a]P, 7,12-dimethylbenz[a]

anthracene (DMBA), and 3-methylcholanthrene (MC). The idea being if the

diol-epoxide pathway is the only mechanism of activation, the fluoro-

substituted derivative should be inactive. When 7-fluoro-B[a]P, 8-fluoro-

B[a]P, and 9-fluoro-B[a]P were tested in mouse skin and rat mammary gland,

they were found to be moderately to weakly active; since they retained some

activity, this suggested that othermechanisms besides diol epoxides can activate


these compounds [107]. An analysis of the metabolites of 8-fluoro-B[a]P and

9-fluoro-B[a]P did not yield 7,8-dihydrodiol and the 9,10-dihydrodiol,

suggesting fluoro-substitution effectively inhibited BPDE formation.

In addition, other fluoro-PAHs such as fluoro-DMBA and fluoro-MC

tested for carcinogenicity in rat mammary glands suggested that diol epoxides

did not play a role in the activation of these compounds. 1-fluoro-DMBA,

2-fluoro-DMBA, and 4-fluoro-DMBA, inwhich formation of the diol epox-

ide is blocked by the fluoro substituent, were carcinogenic [109]. Similarly,

bay-region fluorinated derivatives of methylcholanthrene (8-fluoro-MC and

10-fluoro-MC) were carcinogenic in the rat mammary gland [107]. Because

they foundmutagenesis with compounds that were unable to from diol epox-

ides, they concluded that radical cations are ultimate carcinogens.

10. EVIDENCE FOR AP SITE FORMATIONDUE TO RADICAL CATIONS

An important prediction of the radical cation pathway is that PAHwill

stimulate the production of AP sites. To test for AP sites, Melendez-Colon

et al. searched for AP sites in PAH-treated cells. They developed a procedure

to detect and measure the relative amounts of stable and unstable adducts

formed by DB[a,l]P and DB[a,l]PDE with DNA. Stable adducts were mea-

sured by 33P postlabeling and quantified by HPLC. To measure AP sites,

they were first converted to strand breaks, and examination of the integrity

of the dihydrofolate reductase gene correlated to the amount of AP sites

formed. DNA from Chinese hamster ovary B11 cells were treated with

DB[a,l]PDE and the human mammary carcinoma cell line MCF-7 was

treated with DB[a,l]P. However, they were unable to detect any increases

in AP sites in DB[a,l]P-treated cells. No significant amount of depurinating

DNA adducts was found in human mammary carcinoma cells, MCF-7,

treated with DB[a,l]P. Instead, the corresponding diol-epoxide adduct

was the major adduct [110]. These results suggest that the carcinogenicity

of DB[a,l]P is mostly due to the diol-epoxide pathway and not the radical

cation pathway.

In a follow-up study by Melendez-Colon, the proportions of stable

DNA adducts and AP sites formed by carcinogenic PAHsDB[a,l]P, DMBA,

and B[a]P were investigated in mouse epidermis [111]. After topical appli-

cation of the PAHs on the skin of female SENCAR mice, epidermal DNA

was isolated and stable DNA adducts were measured by 33P postlabeling and

HPLC. AP sites were measured with an aldehyde-reactive probe assay. After


4 and 24 h of exposure, DB[a,l]P formed significantly higher amounts of sta-

ble DNA adducts than DMBA and B[a]P. In contrast, the amount of AP sites

was found highest in DNA exposed toDMBA and least in the DNA exposed

to DB[a,l]P. Exposure of mice to the carcinogenic PAHs B[a]P, DMBA, and

DB[a,l]P resulted in a dose-dependent formation of high levels of stable diol

epoxide–DNA adducts but only low levels of AP sites within epidermal

DNA. On the basis of the results from the above two studies, the authors

argued that mutagenesis is caused by stable diol epoxide–DNA adducts

rather than the formation of AP sites. AP sites were also not found in a study

by Baird’s group in HL-60 cells, which are rich in CYP peroxidases and

devoid of other P450s [112].

Two separate studies conducted by labs other than the Baird lab led to

similar conclusions that one-electron oxidation is not a major route of PAH

activation, in contrast to what was proposed by the Cavalieri lab. In the first

study, the formation of specific 7H-dibenzo[c,g]carbazole (DBC)–DNA

adducts formed by one-electron oxidation was analyzed in the mouse liver

at 4 h following a single intraperitoneal dose of 40 mg/kg of radio-labeled

DBC [113]. In addition, the resulting AP sites were compared with the level

of DBC–DNA adducts detected in the target tissues liver and lung. Interest-

ingly, the authors of the study reported that only one of the three expected

depurinating adducts (DBC-5-N7Gua) was detected. There was no clear

dose–response for the formation of AP sites; however, a significant increase

over control levels was observed at 4 and 40 mg/kg dose groups at 4 and 12 h

postdosing in the mouse liver, respectively. However, in mouse lung

there was no significant difference between control and treated groups.

On the other hand, there was a clear and significant dose and time response

of stable DBC–DNA adducts detected by 32P-postlabeling. Quantitation of

the DNA adducts showed that only 0.4% of the total adducts was the

depurinating adduct DBC-5-N7Gua, whereas the total amount of stable

adducts detected was 99.6%. The data from this study clearly indicate that

one-electron oxidation of DBC in mouse liver is only a minor pathway.

A 6–18-fold excess of stable DNA adducts compared with AP sites were

observed following activation of DBC with HRP in vitro. The major muta-

tion observed following DBC administration in mouse lung was an A to

T transversion and not a G to T transversion as expected.

Results from another study by Leavitt et al., in 2008, also supported the

role of stable PAH–DNA adducts as the major pathway contributing

to PAH-induced mutagenesis in mouse lung in vivo [114]. In this study,

the authors tested the effect of two dosing regimens on mutagenicity of


DB[a,l]P and B[a]P in vivo using the Big Blue transgenic mouse system. The

Big Blue assay detects mutations produced in the lacI target gene, which is

stably integrated in the mouse genome. A single highly mutagenic dose of

each PAH was compared with a fractionated delivery of the same total dose

administered over 5 days. The authors of the study were hoping to induce

higher AP site production in the animals receiving single high dose com-

pared to the animals receiving the fractionated doses. G to

T transversions were the most common mutations observed both with

DB[a,l]P and B[a]P, whereas G to A transitions were the most common

mutations in the controls. The mutations induced by DB[a,l]P and B[a]P

were found to correlate with the stable DNA adduct formation and not with

the formation of depurinating adducts, thus supporting a stronger role of

stable DNA adducts of DB[a,l]P and B[a]P in mouse lung carcinogenesis.

It must be noted that detection of AP sites in vivo may be challenging since

the repair of AP sites is assumed to be a rapid process. If the AP sites are

formed and repaired at similar rates, then mutagenesis by this mechanism

of DNA damage may not be measurable in vivo.

In a contrasting study published in 2001 by Casale et al., urine samples of

seven cigarette smokers, seven women exposed to coal smoke, and thirteen

controls were analyzed using HPLC and tandem mass spectrometry [115].

Depurinating B[a]P DNA adducts were detected in the urine from three of

the seven cigarette smokers and three of the seven women exposed to coal

smoke, and no adducts were detected in the urine from the 13 controls. The

depurinating adduct B[a]P-6-N7Gua was found to be present at �20–300

times the concentration of B[a]P-6-N7Ade in the urine of the women

exposed to coal smoke but was not detected in the urine of the cigarette

smokers. The authors claimed that this difference could be due to difference

in B[a]P exposures experienced by the two groups. They further rec-

ommended use of depurinated B[a]P adducts as potential biomarkers for

acute exposure of B[a]P and PAH-associated cancer risk.

Previously, our lab had demonstrated mutagenicity of three PAH

o-quinones, B[a]P-7,8-dione, BA-3,4-dione, and DMBA-3,4-dione, and

compared them with the ultimate carcinogen anti-BPDE in a yeast reporter

gene assay based on p53 transcriptional activity [116–118]. Yeast offers some

advantages to studyingmutagenesis over bacterial systems.The primary advan-

tage is that the assay system utilizes the major activity of p53, its transcriptional

function, to isolate biologically relevant mutations. In the yeast mutagenesis

assay, a plasmid expressingp53 is exposed to amutagen in vitro.The treatedplas-

mid is then transformed into thehost yeast strainYIG397, an ade2 strain,which


has multiple p53-binding sites engineered into a promoter-driving expression

of ADE2.When wild-type p53 is expressed, it drives the expression of ADE2,

which will turn the host yeast strain from red to white. Mutant p53 colonies

remain red.The ratioof red towhite colonies canbeused tocomparemutagens

and the p53plasmidcanbe isolated and sequenced.ThePAH o-quinones tested

gave a dose-dependent increase in mutation frequency in the range of

0.160–0.375 mM quinone, provided redox-cycling conditions were used.

A dose-dependent mutation frequency was also observed with anti-BPDE

but at micromolar concentrations. The dominant mutation observed with

the quinones was G to T transversion and with anti-BPDE was G to

C transversion [118]. In the yeast system, the locations of the mutant codons,

or spectrum of themutations, are random. If, however, selection for dominant

mutations is performed, the spectrum then closelymatches the one seen in lung

cancers [63,74].

One of the difficulties in addressing radical cation mutagenesis is that they

are unstable and cannot be tested directly. Moreover, in most cells, multiple

metabolic pathways are active at the same time.To isolate andquantify the con-

tribution of radical cations, we developed a system to study them in isolation of

the other pathways using the yeast p53 mutagenesis system described above.

Because radical cations are short-lived, they were generated in situ by reacting

B[a]P with CuOOH and HRP. Formation of the B[a]P radical cation was

monitored by the generation of the stable downstream products, B[a]P-1,6-

dione andB[a]P-3,6-dionewithHPLCandmass spectrometry.Quantification

using standard curves indicated that typically about 2 mM of B[a]P-1,6-dione

was generated in a reaction. Assuming that about 40% of the radical cation is

converted into B[a]P 1,6-dione (the rest is converted to other products such

as B[a]P-3,6-dione and B[a]P-6,12-dione), we calculated the radical cation

concentration in a typical reaction to be 3–5 mM (�8% conversion of B[a]P).

In themutagenesis assayswith the radical cations, adose-dependent increase

inmutagenicity from250 nMto10 mMB[a]Pwas observedwithno significant

increase seen with further escalation to 50 mM B[a]P. However, mutagenesis

was 200-fold less than with the AKR pathway derived B[a]P-7,8-dione

(BPQ). At concentrations as high as 50 mM, very low mutation frequencies

(0.2–0.5% red colonies) were observed with the B[a]P radical cations. For

comparison, 250 nM BPQ yielded mutation frequencies of 1.5–2%, under

redox-cycling conditions.Mutant p53plasmids,whichyield redcolonies,were

recovered from the yeast to study the pattern and spectrum of mutations. The

mutation pattern observed was G to C (36%)>G to T (29%)>G to A (18%).

The frequency and location of codons mutated by the B[a]P radical cations


were essentially random and not enriched for cancer hotspots. These data sug-

gest that radical cations are poorly mutagenic [119].

Most studies agree that PAHs can generate radical cations, as inferred from

their signaturemetabolites.However, there is less evidence that they aremuta-

genic.Because radical cation adducts areunstable andcause depurination, some

investigatorsmeasuredAP sites, but so farhavenotbeen able to finda significant

increase in them upon PAH exposure [110–114]. One of the reasons radical

cations may not be mutagenic is because they do not reach DNA. The

DNA is in the nucleus and radical cations are predominantly generated in

the endoplasmic reticulum. Because the radical cations are short-lived, they

may not survive long enough to reach the nucleus.

11. MUTAGENICITY STUDIES WITH B[a]P RADICALCATION PRODUCTS

As discussed above, the one-electron oxidation of B[a]P produces an

unstable, highly reactive radical cation at the C-6 position. The short-lived

radical cation is eventually converted into the phenol metabolite

6-hydroxy-benzo[a]pyrene (6-OH-B[a]P). 6-OH-B[a]P autooxidizes to

form three B[a]P quinones, namely, B[a]P-1,6-dione, B[a]P-3,6-dione,

and B[a]P-6,12-dione (Fig. 3.3). This pathway of oxidation of B[a]P pro-

duces additional quinones that are different from the ortho-quinone (B[a]

P-7,8-dione) generated by the AKR pathway. The phenol metabolite

3-hydroxy-benzo[a]pyrene (3-OH-B[a]P) can also be converted to B[a]

P-3,6-dione. For the purpose of this chapter, mutagenicity studies of only

6-OH-B[a]P, B[a]P-3,6-dione, B[a]P-1,6-dione, and B[a]P-6,12-diones

will be discussed in this section.

Several studies have looked at the mutagenic strength of various B[a]P

metabolites in bacterial and mammalian cells. The Ames assay can measure

the mutagenicity of compounds that are directly mutagenic and those

requiring metabolic activation. Mutagenic activity of PAH metabolites in

the Ames test system using Salmonella typhimurium TA98 both in the pres-

ence of S9 activation system and in the absence of liver microsomes has been

reported [120,121]. To measure direct mutagens in a mammalian system,

V79 hamster lung cells, which contains no activating enzymes, are grown

in the presence of PAH and microsomal suspensions or other cell prepara-

tions that are capable of activating the PAH [122]. The genetic markers used

in these mammalian cell assays are predominantly 8-azaguanine resistance,

thioguanine resistance, and ouabain sensitivity.


A study by Kodama et al. reported the participation of 6-oxy-B[a]P rad-

ical, which is an intermediate formed during autooxidation of 6-OH-B[a]P

to quinones, inDNAbinding [123]. Indirect evidence basedon loss of tritium

label in a later publication by Rogan et al. [94] suggests that B[a]P maybe

bound at the 1-, 3-, and 6-positions indicating a role of 6-OH-B[a]P and

its subsequent quinones. 6-OH-B[a]P was also found to cause transforma-

tions in cell culture [124]. In a study by Wislocki et al., mutagenicity and

cytotoxicity of 29 B[a]P derivatives were tested for mutagenic activity with-

out activation in S. typhimurium strains TA98, TA100, and TA1538 and in

Chinese hamster V79 cells [121]. 6-OH-B[a]P was found to be moderately

mutagenic in both strain TA98 of S. typhimurium and in V79 cells.

Studies with 6-OH-B[a]P injected subcutaneously into rats found it to

be carcinogenic but similar studies in mice did not find 6-OH-B[a]P to be

active [125]. 6-OH-B[a]P was also not carcinogenic in mouse skin [126].

Another study, by Slaga et al., looking at the skin tumor-initiating activities

of B[a]P phenols in mice found that 6-OH-B[a]P had less than 5% of the

tumor-initiating activity of parent B[a]P when the data were expressed as

papillomas per mouse [127].

The results from the above studies indicate that 6-OH-B[a]P is not

sufficiently carcinogenic or tumorigenic and has a much lower mutagenic

potency compared to parent B[a]P and B[a]P metabolites such as BPDE.

The mutagenesis studies in the bacterial system do not correlate with the

carcinogenic potential of 6-OH-B[a]P. Therefore, the ability of 6-OH-

B[a]P to be mutagenic but not carcinogenic may support the hypothesis that

the radical cation pathway of B[a]P is a minor pathway of chemical carcino-

genesis compared to the diol epoxide and AKR pathway.

The ultimate products of the radical cation pathway in B[a]P are the oxy-

genated B[a]P diones that are capable of redox cycling thereby generating

ROS. All the three diones (1,6-, 3,6-, and 6,12-B[a]P-diones) generated in

the radical cation pathway can undergo reversible, univalent oxidation–

reduction cycles involving the correspondingB[a]P catechols and intermediate

semiquinone radical anions (Fig. 3.8).The1,6-,3,6-, and6,12-B[a]Pdiones are

readily reduced by NADH and GSH to their corresponding catechols, which

rapidly autooxidize back into the diones when exposed to air. Chesis et al. car-

ried out amutagenicity study inSalmonella and found thatB[a]P-1,6-dione and

B[a]P-3,6-dionewere highlymutagenic, whereas B[a]P-6,12-dionewas a rel-

ativelyweakmutagen [128].A study fromWislocki et al. came todifferent con-

clusions finding that (in Chinese Hamster V79 cells and Salmonella TA98,

TA100, and TA1538 tester strains) 1,6-, 3,6-, and 6,12-B[a]P diones were

Figure 3.8 Quinone pathway of representative PAH benzo[a]pyrene (B[a]P).


either inactive or weakly mutagenic compared to B[a]P dihydrodiols. How-

ever, Wislocki et al. did not perform experiments under redox-cycling condi-

tions. Thus, the differences in mutagenicity of 1,6 and 3,6 B[a]P diones

between the two studies can be attributed to the requirement for redox cycling

by quinones for the mutagenicity as noted in study by Chesis et al. [128].

In biological studies conducted by Lesko et al., 1,6-, 3,6-, and 6,12-B[a]P

diones were cytotoxic at low concentrations in cultured hamster cells and

induced strand scissions when incubated with T7 DNA [129]. The same

study found the cytotoxic effect could be substantially reduced by depletion

of oxygen from the growth medium and the atmosphere in which the cells

are cultured. This observation supports the observation that ROS genera-

tion might be responsible for the cytotoxicity seen in these cells.

In a study to address the role of B[a]P-diones associatedwithmixtures of air

pollutants andPMin toxicity, theSV-40-transformedhumanbronchialepithe-

lial cell lineBEAS-2Bwas treatedwith1,6-, 3,6-, and6,12-B[a]P-diones [130].

At nanomolar concentrations, all three diones inhibited epithelial cell prolifer-

ation and were cytotoxic. The cell growth inhibition induced by the diones

could be partially reversed by co-incubation with N-acetyl-L-cysteine and


ascorbate and formation of ROS in the presence of the diones was observed.

The authors suggested that ROS generation and macromolecule alkylation

are possible mechanisms of cytotoxicity of these diones.

In a more recent study, Burchiel et al. used mRNA expression arrays and

qRT-PCR to determine potential pathways and mechanisms by which 1,6-

and 3,6-B[a]P diones induce strong proliferative and EGF receptor activating

activity inhumanmammaryepithelial cells [131].They found that1,6- and3,6-

B[a]P dione activate genes through transcription factors that bind to dioxin

response elements (DRE) (also called xenobiotic response elements) and anti-

oxidant response elements (ARE). B[a]P-3,6-dione had higher activity in

upregulating DRE compared to B[a]P-1,6-dione, whereas opposite was true

for the activation ofARE.Both theB[a]P diones inducedoxidative stress genes

such asHMOX1,GCLC,GCLM, and SLC7A11; phase II enzyme genes such

as NQO1, NQO2, and ALDH3A1; PAH metabolizing genes such as

CYP1B1, EPHX1, and AKR1C1; and certain EGF receptor-associated genes

such as EGFR, IER3, ING1, SQSTM1, and TRIM16. Other PAH

o-quinones also induce oxidative stress genes [132,133].

As discussed in the previous section, we found that radical cations syn-

thesized in vitro were very poor mutagens. However, when tested under

redox conditions, as seen in other studies, the quinone products of radical

cations, B[a]P-1,6-dione and B[a]P-3,6-dione were more mutagenic than

the B[a]P radical cation, although significantly less mutagenic than B[a]P-

7,8-dione. We concluded that B[a]P radical cations are weakly mutagenic

compared to ROS generating BPQ and that mutations caused due to

smoking in p53 are less likely to be attributed to B[a]P radical cations com-

pared to BPDE and BPQ [119].

Evidence from several studies over the years has supported the hypothesis

that tumorigenesis by several oxygenated metabolites of PAH can result

from the generation of ROS [63,64,134–136]. The conversion of B[a]P into

redox-active B[a]P diones that are distinct from the quinones generated by

the AKR pathway may account for the carcinogenicity of the radical cation

pathway of B[a]P.

12. THE QUINONE PATHWAY

In the late 1980s, Smithgall et al. found the first evidence that PAH

dihydrodiols such as B[a]P-7,8-dihydrodiol can be transformed into cate-

chols by dihydrodiol dehydrogenase [137]. The catechols derived via this

pathway are unstable and undergo two-step autooxidation to initially form


a semiquinone anion-radical plus hydrogen peroxide which rapidly converts

into the quinone and superoxide anion (Fig. 3.8) [135]. There are two routes

by which PAH o-quinones can damage the DNA: they can enter into a futile

redox cycle using NADPH to produce ROS, which causes oxidative dam-

age to DNA. ROS causes oxidative damage through the formation of

8-oxo-dG, which leads to G to T transversions [117,135,138]. Park et al.,

in 2008, reported the activation of the quinone pathway in human lung ade-

nocarcinoma (A549) cells suggesting its role in human lung carcinogenesis

[136]. Because the PAH o-quinones are Michael acceptors they have the

ability to form stable adducts (e.g., BP-7,8-dione-N2-dGuo and BP-7,8-

dione-N6-dAdo) and depurinating adducts (e.g., BP-7,8-dione-N7-Gua

and BP-7,8-dione-N7-Ade) which may have mutagenic potential

[139,140]. As discussed above, PAH o-quinones are highly mutagenic,

but only under redox-cycling conditions, so it is unlikely that the direct

adducts contribute much to carcinogenesis.

Several studies have implicated a role of ROS in lung carcinogenesis

[141,142]. This may be supported by the finding that one allele of hOGG1,

a mammalian base excision repair enzyme that excises 8-oxo-dG by acting as

an AP-glycosylase/lyase, is deleted in some human lung cancers [143]. The

hOGG1gene is located on the distal end of the short arm of chromosome

(3p25/26) and nearly 100% of cells derived from small cell lung cancers

(which represents, one-third of all lung cancers) are characterized by a loss

of heterozygosity at chromosome 3. This suggests that during the develop-

ment of lung cancer individuals lose the ability to excise the most common

DNA lesion associated with ROS exposure and this could increase the

mutational load [144]. Because of the association of smoking with lung can-

cer, these observations raise the issue of whether there is a known pathway of

activation of tobacco carcinogens that increases the production of ROS.

In humans, at least five members of the AKR superfamily have

dihydrodiol dehydrogenase activity (AKR1C1–AKR1C4 and AKR1A1).

Of these, AKR1C1 is inducible by planar aromatics and ROS [145], and

differential display shows that it is highly expressed in non-small cell lung

carcinoma [146]. This high expression has been correlated to a poor prog-

nostic outcome. One interpretation of these data is that AKRs are over-

expressed as a result of PAH exposure and are involved in their metabolic

activation in situ. If PAH o-quinones produced by AKRs cause DNA lesions

via ROS these lesions should be observed in vivo. Indeed, rat hepatocytes

treated with B[a]P-7,8-diol show strand scission, which is attenuated with

AKR inhibitors, [147] and lung cells treated with B[a]P-7,8-diol show


ROS accumulation and increased levels of 8-oxo-dG [136], demonstrating

that the ROS lesions are AKR dependent. In conclusion, PAH o-quinones

are formed in human lung by the inducible AKR1C isoforms following

exposure to the parent hydrocarbon. This results in ROS formation and oxi-

dative damage to DNA creating the mutagenic lesion 8-oxo-dG, which in

turn causes the signature G to T transversions in p53.

13. EXPOSURE TO OPAHs

The presence of OPAHs in the environment is widespread with

significant levels found in diesel and gasoline exhaust [148–150], emission

of gases from combustion processes [151], fly ash [152], urban aerosols

[18,19,150,151,153] sediments [154,155], river and coastal waters [156], sew-

age sludge [157], industrial waste [158], and soil [159–163]. The concentration

of atmospheric OPAHs might be higher because PAHs in the atmosphere are

more susceptible to photochemical transformation thanPAHs in the soil [155].

Atmospheric sampling studies of OPAH have found concentrations in the

approximate range 0.1–10 ng/m3 for a number of OPAH [164,165].

In seminal work done by Lundstedt et al., OPAHs in soils from seven

different PAH-contaminated sites were quantified [159]. The seven sites

chosen by the authors were a gasworks site, a coke production site, four

wood preservation sites in Sweden, and a Superfund site in the United States

(Laramie, WY). The results of the study found the levels of OPAHs com-

parable to the PAHs levels in all the study sites. In six of the seven study sites,

concentrations of 17 OPAHs varied between 10% and 30% of the total con-

centrations of the PAHs that are considered harmful by the USEPA

(Table 3.1). In soil from the seventh site, the concentration of OPAHs

was 66% of the PAH concentration. It is worthwhile to note that all the

OPAHs studied in the above study were not quantified using authentic ref-

erence compounds; some had to be quantified against 4H-cyclopenta[def]

phenanthrene, which was one of the major OPAHs peaks in the samples.

Nevertheless, the authors found levels of OPAHs too high to be ignored

in any of the samples. In some cases, concentrations of OPAH such as

9-fluorenone were higher than the concentrations of the parent PAH

(fluorene) from which they had most likely originated.

In another study by the Lundstedt group, the concentrations of OPAHs

in the soil samples from the gasworks site were analyzed prior to remedial

treatments of the soil [159,160]. The results were similar to the aforemen-

tioned study although the soils sampled for remedial treatments contained


somewhat lower concentrations of both PAHs and OPAHs. In addition to

the above studies to quantify OPAH concentrations, two other studies have

reported identification of OPAHs in PAH-contaminated soil prior to bio-

degradation experiments. The first study by Brooks et al. identified

9-fluorenone, anthracene-9,10-dione, 7H-benzanthracen-7-one, and 4H-

cyclopenta[def ]phenanthrene in a creosote-contaminated soil from a Super-

fund site in Minnesota [161]. In the second study, Saponaro et al. found

anthracene-9,10-dione and two isomers of benzanthracenone in soil from

a former gasworks site in Italy [166]. The only shortcoming of the last

two studies is that mass spectroscopy was used to identify the OPAHs in

the study without any other analytical techniques.

Studies were also performed to quantify the formation of OPAHs after a

soil remediation process. The Lundstedt group also demonstrated OPAH

accumulation during 3 different remedial processes: a bioslurry treatment,

treatment with wood-rotting fungi and treatment with a combination of

ethanol and Fenton’s reagent [160,167,168]. The same soil samples from

the gasworks site were used in the bioslurry and the chemical (ethanol

and Fenton’s reagent) studies. During the bioslurry treatment, overall

PAH removal was poor but OPAHs were found to accumulate in the pro-

cess at significantly higher concentrations at the end of the treatment than

before the soil was treated. In the chemical degradation process, PAHs were

depleted faster, accompanied by the accumulation of OPAHs. Anthracene-

9,10-dione, 1-methylanthracene-9,10-dione, 2-methylanthracene-9,10-

dione, benzo[a]anthracene-7,12-dione, and 4-hydroxy-9-fluorenone were

found at higher concentrations in the soil samples after the treatment than

before the treatment.

Studies with treatment of wood-rotting fungi to an artificially contam-

inated soil, also found accumulation of OPAHs as the PAHs were degraded.

In a study by Andersson et al., accumulation of anthracene-9,10-dione was

predominant after an artificially PAH-contaminated soil was treated with

white-rot fungi [169]. In another study, 9 fluorenone, anthracene-9,10-

dione, 2-methylanthracene-9,10-dione, and benz[a]anthracene-7,12-dione

were found to accumulate during the biodegradation of PAHs in an artifi-

cially contaminated soil [170]. In an interesting study conducted by Eriksson

et al., OPAHs were quantified before and after biodegradation at the gas-

works site mentioned earlier [162]. Before biodegradation, the relative pro-

portions of PAHs and OPAHs were similar. After biodegradation, OPAHs

identified in the gasworks site soil were found to accumulate, as the soil was

biologically degraded under various conditions.


In an older study by Durant and coworkers, extracts of an urban aerosol

sample from Washington, DC were analyzed by human cell bioassay-

directed chemical analysis. The nonpolar organic fraction containing

PAH and the semipolar fraction containing OPAH were both human cell

mutagens. Further work to test individual compounds for human cell muta-

genicity found a number of PAH and OPAH to be human cell

mutagens [171].

The Macondo oil spill is showing evidence that there may be composi-

tional changes in the PAH as the oil weathers [172,173]. A study by Aeppli

et al. looked at the fate of the oil by analyzing samples at different stages if

weathering during an 18-month period following the spill [39]. These sam-

ples were collected from the impacted beaches and included the oil from the

Macondo well, sea surface oil slicks, oil soaked sands (“sand patties”), and

oil-covered rocks (“rock scrapings”). Thin layer chromatography–flame

ionization detector and gas chromatography–flame ionization detector were

used to characterize the components of the samples. The authors measured

PAHs for a subset of 10 samples and the total concentration of 30 PAHs

decreased from oil at the Macondo well to oil from surface slicks to extract-

able material from sand patties and rock scrapings, mostly because the smaller

PAH, which are water soluble, evaporated or dissolved. However, the

amount of oxygenated oil residues increased. The authors hypothesized that

biodegradation and photooxidation were responsible for accumulation of

oxygenated compounds that after a period of time dominate the solvent-

extractable material from oiled samples. Although OPAHs were not specif-

ically addressed, it is likely that they were formed as the oil weathered.

14. BIOLOGICAL RELEVANCE OF OPAH

The genotoxic and mutagenic effects of OPAHs are generally under-

rated compared to PAHs. OPAHs are products of combustion, especially

diesel fuel. Many PAHs can degrade both chemically and biologically into

OPAH. Even bioremediation of contaminated water and soil can lead to the

accumulation of OPAH as dead end products of metabolism [159,168].

OPAHs also persist in the environment and cause adverse effects in humans.

OPAHs are acutely toxic to the marine bacterium Vibrio fischeri, the

aquatic invertebrate Daphnia magna, various microalgae and the aquatic

and terrestrial plants Lemna ibba and Brassica napus [24,174–176]. Some

OPAHs are mutagenic in bacterial and human cell mutation assays

[165,177–179]. In addition, OPAHs are compounds enriched in the


semipolar fractions of atmospheric particulate extracts; fractions that are

highly mutagenic in bacterial and human cells [165]. In addition to the gen-

otoxic effects of PAHs in the PM, particulate air pollution also induces oxi-

dative damage to DNA [55]. The modified base 8-oxo-dG, one of the

mutagenic lesions, is elevated in lymphocytes of people exposed to urban

air pollution [180,181].

OPAHs are cytotoxic, induce oxidative stress and disrupt the endocrine

system [182–186]. Mutagenic activity of OPAHs has also been observed in

mammalian cells [171]. In a case–control study nested in a large prospective

study in Europe (EPIC), cases of newly diagnosed lung cancer (N¼115),

upper respiratory cancers (pharynx, larynx) (N¼82), bladder cancer

(N¼124), leukemia (N¼166), and COPD or emphysema deaths

(N¼77) were compared between ex-smokers and never-smokers [187].

The authors found that the patterns of adduct formation in healthy non-

smokers who developed lung cancer was stronger in comparison to smokers

and the observed adduct profile has been previously linked to exposure to

environmental air pollution. Interestingly, the authors also found an associ-

ation of adduct levels with the concentration of ozone in the nonsmokers,

suggesting a possible role of OPAHs formed during photochemical reac-

tions. These observations are in agreement with a previous study in

Florence, where a significant relationship was found between cumulative

ozone expose and bulky DNA adduct formation in nonsmokers [188].

Ozone is a marker of photochemical smog that contains complex mixtures

of combustion products of fossil fuel and their reaction products with UV

radiation in sunlight. B[a]P, in the presence of UV radiation, can signifi-

cantly increase the expression levels of tumor suppressor gene P53 and

the production of G to T transversions, a signature mutation caused by

oxidative damage [189,190].

Even though mammalian metabolic processes may enhance the mutage-

nicity of OPAHs, exogenous metabolic activation of OPAHs may or may

not be required for the manifestation of mutagenic response and can directly

react with DNA and other biomolecules [17,22,151,174,178,179,191]. Sev-

eral in vitro studies in bacteria have been done to study the mutagenic effects

of outdoor air pollution and have been reviewed [43]. The mutagenesis

studies suggest that mixtures of environmental PAH constitute an important

source of genotoxicity. For instance, a study by Lemieux et al. found that

mixtures of OPAHs contributed significantly to the total mutagenicity of

PAH-contaminated soil samples in the Ames test [192]. In the study, the

authors combined chemical separation methods with biological tests for


toxicity assessment to narrow down the most mutagenic component of the

contaminated soil. Fractionation of the samples based on the polarity of the

components and subsequent toxicity testing of the fraction with Salmonella

revealed that fractions enriched with OPAHs and nitro-PAHs contributed

to the total genotoxicity of the semipolar fraction that was extracted. Other

studies that have investigated toxicity of complex environmental pollutant

mixtures using chemical fractionation also concluded that OPAHs predom-

inate in the most toxic fractions [148,154,193,194].

In an interesting study published by Wei et al., in 2010, the burden of

oxidative stress in humans exposed to particulate PAHs and PAH-quinones

was examined [195]. The authors recruited two nonsmoking security guards

who worked at a heavy traffic spot on a university campus. Each subject

wore a personal air sampler for 24 h per day to estimate exposures to

24 PAHs and anthraquinone. Analysis of urine samples from the subjects

found a threefold increase of 8-oxo-dG after an 8 h work shift compared

to the samples taken during the prework shift. All the 24 PAHs and anthra-

quinone levels were positively and significantly associated with the postwork

urinary 8-oxo-dG.

The mechanisms governing the cytotoxicity and genotoxicity of

OPAHs are complex and have not been clearly elucidated yet. It has been

posited that OPAHs, such as the quinones, may be converted to electro-

philic intermediates that may form adducts with cellular DNA and proteins.

This may ultimately lead to genotoxicity mediated by DNA adduct forma-

tion and possibly also cytotoxicity via depletion of cellular glutathione

levels [64]. In addition, quinones can undergo futile enzymatic and non-

enzymatic redox reactions generating ROS in cells. The ROS produced

by the quinones can cause severe oxidative damage to the cellular compo-

nents by reacting with macromolecules such as DNA, proteins, and lipids.

The toxic effects of OPAHs can also be caused by activation of a series of

signaling pathways and cellular events in the cell due to ROS

generation [64].

15. CONCLUSION

PAHs are the most toxic group of environmental toxins. Most

research has focused on the unsubstituted parent compounds. However,

OPAH poses a risk that has not been adequately evaluated. OPAHs are

derived from PAH combustion, PAH decomposition and also PAH metab-

olism by either the AKR pathway or the radical cation pathway. Although


there is mixed evidence that the radical cations themselves are genotoxic,

because their signature AP sites are not reproducibly observed, their metab-

olites can redox cycle, like other OPAH, to generate ROS. Due to unprec-

edented levels of sampling, the Macondo oil spill presents an opportunity to

quantify the formation of OPAH through weathering and to monitor their

persistence in the environment.

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[advances in molecular toxicology] volume 7 || genotoxicity of polycyclic aromatic hydrocarbon...

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