LEVELS OF POLYCYCLIC AROMATIC HYDROCARBON IN FRESHWATER FISH DRIED UNDER DIFFERENT DRYING REGIMES

ABSTRACT

Preservation of fish by drying over different types of heat regimes have been known. However, there has not been a comprehensive comparison in terms of the possible contamination associated with these drying regimes. This work was set to evaluate the levels of PAHs that are likely to accumulate in the bodies of freshwater fishes dried under heat from charcoal, sun (sun drying), electric oven and polythene augmented drying regimes (burning of used cellophane materials). The levels of sixteen PAHs were determined in fish samples harvested from Otuocha River in Anambra State, Nigeria. The fish samples were dried, pulverized and subjected to soxhlet extraction using n-hexane at 600c for 8hrs. The water content of the eluents were further removed with florisil clean-up before Gas chromatographic – mass spectrometric analysis. Results obtained showed that sun-dried fish had PAHs concentration to be 35.7+ 0.2µg/g; oven dried gave 47.7+ 0.2µg/g and charcoal dried 79.53+ 0.2µg/g, while drying with firewood resulted in 188.1+ 0.2µg/g. Charcoal drying augmented with polythene resulted into PAHs level of 166.2+ 0.1µg/g while fish dried under heat generated from burning firewood and polythene material resulted into PAHs concentration of 696.3+0.2µg/g. Preliminary analysis of the freshwater samples and the undried fish samples (control) revealed that the fresh water contained total PAHs level of 2.86+ 0.1µg/ml, while the fresh fish 4.97+ 0.2µg/g. The concentration of PAHs in all the dried fish under different drying agents were significantly higher than the control. The result is more worrisome in that even the fishes dried under the sun have PAHs significantly higher than that of the control (p<0.05). It is apparent that the increase in PAHs must have come from the environmental PAHs (exposure) under which the fishes were dried (under sun). For the other drying regimes, in which the levels of PAHs were significantly higher than that of sun-dried, it can be concluded that the excessive PAHs in the body of the dried fish were from the “burning” or drying agents. More significantly are the observed very high increase in PAHs when drying was augmented with polythene, an agent known to be a high source of PAHs when incinerated. Consumers of dried fish should therefore beware of the dried fish they purchase from the local market. CHAPTER ONE

1.1 INTRODUCTION

Polycyclic aromatic hydrocarbons (PAHs) are a group of organic compounds consisting of two or more fused benzene rings (linear, cluster or angular arrangement), or compounds made up of carbon and hydrogen atoms grouped into rings containing five or six carbon atoms. They are called “PAH derivatives” when an alkyl or other radical is introduced to the ring, and heterocyclic aromatic compounds (HACs) when one carbon atom in a ring is replaced by a nitrogen, oxygen or sulphur atoms. PAHs originate mainly from anthropogenic processes particularly from incomplete combustion of organic fuels. PAHs are distributed widely in the atmosphere. Natural processes, such as volcanic eruptions and forest fires, also contribute to an ambient existence of PAHs (Suchanova et al., 2008). PAHs can be present in both particulate and gaseous phases, depending on their volatility. Low molecular weight PAHs (LMW PAHs) that have two or three aromatic rings (molecular weight from 152 to 178g/mol) are emitted in the gaseous phase, while high molecular weight PAHs (HMW PAHs), molecular weight ranging from 228 to 278g/mol, with five or more rings, are emitted in the particulate phase, (ATSDR, 1995) . In the atmosphere, PAHs can undergo photo-degradation and react with other pollutants, such as sulfur dioxide, nitrogen oxides, and ozone. Due to widespread sources and persistent characteristics, PAHs disperse through atmospheric transport and exist almost everywhere. There are hundreds of PAH compounds in the environment but in practice PAH analysis is restricted to the determination of six (6) to sixteen (16) compounds. Human beings are exposed to PAH mixtures in gaseous or particulate phases in ambient air. Long term exposure to high concentration of PAHs is associated with adverse health problems. Since some PAHs are considered carcinogens, inhalation of PAHs in particulates is a potentially serious health risk linked to lung cancer (Philips, 1999).

1.2. Physical and Chemical Characteristics of PAHs.

PAHs are a group of several hundred individual organic compounds which contain two or more aromatic rings and generally occur as complex mixtures rather than single compounds. PAHs are classified by their melting and boiling points, vapour pressure, and water solubility, depending on their structure. Pure PAHs are usually coloured, crystalline solids at ambient temperature. The physical properties of PAHs vary with their molecular weight and structure (Table1). Except for naphthalene, they have very low to low water solubilities, and low to moderately high vapour pressures. Their octanol-water partition coefficients (Kow) are relatively high, indicating a relatively high potential for adsorption to suspended particles in the air and in water, and for bioconcentration in organisms (Sloof et al., 1989). Table 1 shows physical and chemical characteristics of few selected PAHs from the sixteen (16) priority PAHs, listed by the US EPA. (see appendix). Most PAHs, especially as molecular weight increases, are soluble in non-polar organic solvents and are barely soluble in water (ATSDR, 1995).

Most PAHs are persistent organic pollutants (POPs) in the environment. Many of them are chemically inert. However, PAHs can be photochemically decomposed under strong ultraviolet light or sunlight, and thus some PAHs can be lost during atmospheric sampling. Also, PAHs can react with ozone, hydroxyl radicals, nitrogen and sulfur oxides, and nitric and sulfuric acids which affect the environmental fate or conditions of PAHs (Dennis et al., 1984; Simko, 1991).

PAHs possess very characteristic UV absorbance spectra. Each ring structure has a unique UV spectrum, thus each isomer has a different UV absorbance spectrum. This is especially useful in the identification of PAHs. Most PAHs are also fluorescent, emitting characteristic wavelengths of light when they are excited (when the molecules absorb light). Generally, PAHs only weakly absorb light of infrared wavelengths between 7 and 14µm, the wavelength usually absorbed by chemical involved in global warming (Ramanathan, 1985).

Polycyclic aromatic hydrocarbons are present in the environment as complex mixtures that are difficult to characterize and measure. They are generally analyzed using gas chromatography coupled with mass spectrometry (GC-MS) or by using high pressure liquid chromatography (HPLC) with ultraviolet (UV) and fluorescence detectors (Slooff et al., 1989)

Source and Emission of PAHs

PAHs are mainly derived from anthropogenic activities related to pyrolysis and incomplete combustion of organic matter. Sources of PAHs affect their characterization and distribution, as well as their toxicity. Major sources of PAH emissions may be divided into four classes: stationary sources (including domestic and industrial sources), mobile emission, agriculture activities, and natural sources (Wania et al, 1996).

1.3. Stationary Sources

Some PAHs are emitted from point sources and this is hardly shifted (moved) for a long period of time. Stationary sources are further subdivided into two main sources: domestic and industrial.

1.3.1. Domestic Sources1.3.2. Industrial Sources

Sources of PAHs include emission from industrial activities, such as primary aluminum and coke production, petrochemical industries, rubber tire and cement manufacturing, bitumen and asphalt industries, wood preservation, commercial heat and power generation, and waste incineration (Fabbri and Vassura , 2006).

1.3.3. Mobile Sources

Mobile sources are major causes of PAHs emissions in urban areas. PAHs are mainly emitted from exhaust fumes of vehicles, including automobile, railways, ships, aircrafts, and other motor vehicles. PAHs emissions from mobile sources are associated with use of diesel, coal, gasoline, oils, and lubricant oil. Exhaust emissions of PAHs from motor vehicles are formed by three mechanisms: (1) synthesis from smaller molecules and aromatic compounds in fuel; (2) storage in engine deposits and in fuel; (3) pyrolysis of lubricants (Baek et al., 1991). One of the major influences on the production of PAHs from gasoline automobiles is the air-to-fuel ratio. It has been reported that the amount of PAHs in engine exhaust decreases with leaner mixtures (Ravindra et al., 2006b). A main contribution to PAH concentrations in road dust as well as urban areas is vehicle exhaust. Abrantes et al., (2009) reported that the total emissions and toxicities of PAHs released from light-duty vehicles using ethanol fuels are less than those using gasohol. Low molecular weight PAHs are the dominant PAHs emitted from light duty vehicles and helicopter engines.

1.3.4 Agricultural Sources

Open burning of bush wood, straw, moorland heather, and stubble are agricultural sources of PAHs. All of those activities involve burning organic materials under sub optimum combustion conditions. Thus it is expected that a significant amount of PAHs are produced from the open burning of biomass. PAH concentrations released from wood combustion depend on wood type, kiln type, and combustion temperature. Between 80 – 90% of PAHs emitted from biomass burning are low molecular weight PAHs, including naphthalene acenaphthylene, phenanthrene, fluoranthene and pyrene. Lu et al., (2009) reported that PAHs emitted from the open burning of rice and bean straw are influenced by combustion parameters. Total emissions of 16 PAHs from the burning of rice and bean straw varied from 9.29 to 23.6µg/g and from 3.13 to 49.9µg/g respectively. PAH emissions increased with increasing temperature from 200 to 7000c.

Maximum emissions of PAHs were observed at 40% O2 content in supplied air. However, emission of PAHs released from the open burning of rice straw negatively correlate with the moisture content in the straw (Lu et al., 2009).

1.3.5. Natural Sources

Accidental burning of forests, woodland, and moorland due to lightning strikes are natural sources of PAHs. Furthermore, volcanic eruptions and decaying organic matter are also important natural sources, contributing to the levels of PAHs in the atmosphere. The degree of PAH production depends on meteorological conditions such as wind, temperature, humidity, and fuel characteristics and type; such as moisture content, green wood, and seasonal wood (Wild and Jones, 1995).

1.3.6 Uses of PAHs

PAHs are not synthesized chemically for industrial purposes. Rather than industrial sources, the major source of PAH is the incomplete combustion of organic material such as coal, oil, and wood. However, there are a few commercial uses for many PAHs. They are mostly used as intermediaries in pharmaceuticals, agricultural products, photographic products, thermosetting plastics, lubricating materials, and other chemical industries. Acenaphthene, Anthracene, Fluoranthene, Fluorene, Phenanthrene and Pyrene are used in the manufacture of dyes, plastics, pigments, pharmaceutical and agrochemicals such as pesticides, wood preservatives resins and so on.

Other PAHs may be contained in asphalt used for the construction of roads, as well as roofing tar. Precise PAHs, specific refined products, are used also in the field of electronics, functional plastics, and liquid crystals. (Katarina, 2011).

1.4 Routes of Exposure for PAHs

PAH exposure through air, water, soil, and food sources occurs on a regular basis. The routes of exposure include ingestion, inhalation, and dermal contact in both occupational and non-occupational settings. Some exposure may involve more than one route simultaneously, affecting the total absorbed dose (such as dermal and inhalation exposure from contaminated air). All non-workplace source of exposure such as diet, smoking, and burning of coal and wood should be taken into consideration (ATSDR, 1995).

1.4.1 Air

PAHs concentrations in air can vary from less than 5 to 200,000 (ng/m3) (Cherng et al., 1996; Georgiadis and Kyrtopoulos, 1999). Although environmental air levels are lower than those associated with specific occupational exposure, they are of public health concern when spread over large urban populations (Zmirou et al., 2000).

The background levels of the Agency for Toxic Substances and Disease Registry’s toxicological priority for PAHs in ambient air have been reported to be 0.02 – 1.2 ng/m3 in rural areas and 0.15 – 19.3 ng/m3 in urban areas (ATSDR, 1995).

Cigarette smoking and environmental tobacco are other sources of air exposure. Smoking one cigarette can yield an intake of 20-40ng of benzo (a) pyrene (Philips, 1996; O’Neill et al., 1997). Smoking one pack of unfiltered cigarette per day yields 0.7µg/day benzo (a) pyrene exposure. Smoking a pack of filtered cigarette per day yields 0.4 µg/day (Sullivan and Krieger 2001).

Environmental tobacco smoke contains a variety of PAHs, such as benzo (a) pyrene, and more than 40 known or suspected human carcinogens. Side-stream smoke (smoke emitted from a burning cigarette between puffs) contains PAHs and other cytotoxic substances in quantities much higher than those found in mainstream smoke (exhaled smoke of smoker) (Jinot and Bayard, 1996; Nelson, 2001).

1.4.2. Water

PAHs can leach from soil into groundwater. Water contamination also occurs from industrial effluents and accidental spills during oil shipment at sea. Concentrations of benzo (a) pyrene in drinking water are generally lower than those in untreated water and about 100 fold lower than the US Environmental Protection Agency’s (EPA) drinking water standard. (EPA’s maximum contaminant level (MCL) for benzo (a) pyrene in drinking water is 0.2 parts per billion {ppb}(US EPA, 1995).

1.4.3 Soil

Soil contains measurable amounts of PAHs primarily from airborne fallout. Documented level of PAHs in soil near oil refineries have been as high as 200,000 micrograms per kilogram (µg/kg) of dried soil. Levels in soil samples obtained near cities and areas with heavy traffic were typically less than 2,000 µg/kg (IARC, 1973).

1.4.4 Food Stuffs

In non-occupational settings, up to 70% of PAH exposure for non-smoking person can be associated with diet (Skupinska et al., 2004). PAH concentrations in foodstuffs vary. Charring meat or barbecuing food over a charcoal, wood, or other type of fire greatly increase the concentration of PAHs. For example, the PAH level for charring meat can be as high as 10-20 µg/kg (Philips, 1999). Charbroiled and smoked meats and fish contain more PAHs than do uncooked products, with up to 2.0 µg/kg of benzo (a) pyrene detected in smoked fish. Tea, roasted peanuts, coffee, refined vegetable oil, cereals, spinach, and many other foodstuffs contain PAHs. Some crops such as wheat, rye and lentils, may synthesize PAHs or absorb them via water, air, or soil (Grimmer, 1968; Shabad and Cohan 1972; IARC, 1973).

1.4.5 Other Sources of Exposure

PAHs are found in prescription and non-prescription coal tar products used to treat dermatologic disorders such as psoriasis and dandruff (Van Schooten, 1996). PAHs and their metabolites are excreted in breastmilk, and they readily cross the placenta.

Antracene laxative use has been associated with melanosis of the colon and rectum (Badiali et al., 1985).

1.5 Individuals at Risk of Exposure

Workers in industries or trades using or producing coal or coal products are at highest risk for PAHs exposure. Those workers include, but are not limited to Aluminum workers, Asphalt workers, Carbon black workers, Chimney sweeps, Coal-gas workers, Fishermen (coal tar on nets), Graphite electrode workers, Machinists, Mechanics (auto and diesel engine), Printers, Road (pavement) workers, Roofers, Steel foundry workers, Tire and rubber manufacturing workers, and Workers exposed to creosote, such as Carpenters, Farmers, railroad workers, Tunnel construction workers, and Utility workers.

Exposure is almost always to mixtures that pose a challenge in developing conclusions (Samet, 1995). Fetuses may be at risk for PAH exposure. PAH and its metabolites have been shown to cross the placenta in various animal studies (ATSDR, 1995). Because PAH are excreted in breast milk, nursing infants of exposed mothers can be easily exposed.

1.6 Standard and Regulations of PAHs Exposure.

The United States Government Agencies have established standards that are relevant to PAHs exposure in the workplace and the environment. There is a standard relating to PAHs in the workplace, and also a standard for PAHs in drinking water.

Occupational safety and health administrations (OSHA) have not established a substance-specific standard for occupational exposure to PAHs. Exposures are regulated under OSHA’s Air contaminants standard for substances termed coal tar pitch volatiles (CTPVs) and coke oven emission. Employees exposed to CTPVs in the coke oven industry are covered by the coke oven emissions standard.

The OSHA coke oven emission standard required employers to control employee exposure to coke oven emissions by the use of engineering controls and work practices.

Whenever the engineering and work practices control that can be instituted are not sufficient to reduce employee exposure to or below the permissible exposure limit (PEL), the employer shall nonetheless use them to reduce exposure to the lowest level achievable by these controls and shall supplement them by the use of respiratory protection. The OSHA standards also include elements of medical surveillance for workers exposed to coke oven emissions (ATSDR, 1995).

Air

The OSHA PEL for PAHs in the workplace is 0.2 milligram/cubic meter (mg/m3). The OSHA – mandated PAH workroom air standard is an 8-hour time-weighted average (TWA) permissible exposure limit (PEL) of 0.2 mg/m3, measured as the benzene-soluble fraction of coal tar pitch volatiles. The OSHA standard for coke oven emissions is 0.15 mg/m3. The National Institute for Occupational Safety and Health (NIOSH) has recommended that the workplace exposure limit for PAHs be set at the lowest detectable concentration which was 0.1 mg/m3 for coal tar pitch volatile agents at the time of the recommendation (ATSDR, 1995).

•TLV: threshold limit value.

•TWA (time – weighted average), concentration for a normal 8-hour workday and a 40-hour workweek to which nearly all workers may be repeatedly exposed.

•REL (recommended exposure limit): recommended airborne exposure limit for coal pitch volatiles (cyclohexane – extractable fraction) averaged over a 10 – hour work shift.

•PEL (permissible exposure limit): the legal airborne permissible exposure limit (PEL) for coal tar pitch volatiles (Benzene soluble fraction) averaged over an 8 – hour work shift.

•MCL: maximum contaminant level. (ATSDR, 1995).

Water

The maximum contaminant level goal for benzo (a) pyrene in drinking water is 0.2 parts per billions (ppb). In 1980, EPA developed ambient water quality criteria to protect human health from the carcinogenic effects of PAH exposure. The recommendation was a goal of zero (non-detectable level for carcinogenic PAHs in ambient water). EPA, as a regulatory agency, sets a maximum contaminant level (MCL) for benzo (a) pyrene, the most carcinogenic PAH at 0.2ppb. EPA also sets MCLs for five other carcinogenic PAHs (see table 2) (ATSDR, 1995).

Food

The U.S. Food and Drug Administration has not established standard governing the PAH content of foodstuffs but the Food and Agricultural Organization (FAO) and World Health Organization (WHO) have set a maximum permissible level for total polycyclic aromatic hydrocarbons and benzo (a) pyrene in certain foods. Recently the maximum permissible level of health hazard dietary intake of the PAHs in cooked and processed food are not defined accurately and varies from one country to another. Janoszka et al., (2004) reported that the health hazard level of the PAHs daily ingested in diet was found to be 3.7µg/kg in Great Britain, 5.17µg/kg in Germany, 1.2 µg/kg in New Zealand and 3 µg/kg in Italy. Generally it is known that the maximum permissible level (MPLs) of total PAHs and BaP are 10 and 1µg/kg wet cooked or processed meat and fishery products respectively as reported by FAO/WHO and Story How and Sikorski (2005). The above and the Health hazard level of 5.7µg/day as reported by Janoszka et al., (2004) are the accepted reference standards even in Nigeria.

1.7 Metabolism of PAHs

Once PAHs enter the body they are metabolized in a number of organs (including liver, kidney, lungs), excreted in bile, urine or breast milk and stored to a limited degree in adipose tissue. The principal routes of exposure are: inhalation, ingestion, and dermal contact. The lipophilicity of PAHs enables them to readily penetrate cellular membranes (Yu, 2005). Subsequently metabolism renders them more water-soluble making them easier for the body to remove. However, PAHs can also be converted to more toxic or carcinogenic metabolites.

Phase I metabolism of PAHs

There are three main pathways for activation of PAHs: the formation of PAH radical cation in a metabolic oxidation process involving cytochrome P450 peroxidase, the formation of PAH-o-quinones by dihydrodiol dehydrogenase-catalysed oxidation and finally the creation of dihydrodiol epoxides, catalysed by cytochome P450 (CYP) enzymes (Guengerich, 2000). The most common mechanism of metabolic activation of PAHs, such as Benzo (a) pyrene (B(a)P), is via the formation of bay-region dihydrodiol epoxides eg. Benzo (a)pyrene-7, 8-dihydrodiol-9,10-epoxide (BPDE), via CYP450 and epoxide hydrolase (EH) as seen in figure 1 below.

The most important enzymes in the metabolism of PAHs are CYPs IA1, IA2, IB1 and 3A4. CYP IAI is highly inducible by PAHs such as B(a)P and some polyhalogenated hydrocarbons. Recombinant human CYP IAI metabolizes compounds such as B(a)P, 2-acetylamino-fluorene and 7,8-diol, 7-12-dimethylbenz (a) anthracene (Kim et al., 1998). CYPIA2 and CYPIB2 are also inducible by the exposure to PAHs. These enzymes share the same mechanism with which PAH molecules interact with the aryl hydrocarbon receptor (AHR). The AHR is present in the cytoplasm as a complex with other proteins such as heat shock protein 90 (HSP 90), p23 and AhR-interacting protein. After forming a complex with PAHs, the Hsp90 is released and the AhR-PAH complex translocates to the nucleus as seen in Figure 2.

Here, it creates a heterodimer with a ARNT (Ah Receptor Nuclear Translocator) and afterwards binds to DNA via the xenobiotic response element (XRE) situated in the promoter region of CYPIA and CYPIB genes (Shimada et al., 2002).

Other phase I enzymes related to PAHs metabolism are the aldo-keto reductases. These enzymes oxidize polycyclic aromatic (PAH) trans-dihydrodiols to reactive and redox-active O-quinones in vitro (Quinn and Penning, 2006). Specifically, AKRIAI, and members of the AKRIC dihydrodiol/hydroxysteroid dehydrogenase subfamily, AKRICI-AKRICA are involved in metabolic activation of PAH trans-dihydrodiol. Production of O-quinone metabolites by these enzymes has been shown in vitro and in cell lines to amplify ROS and oxidative damage to DNA bases to form the highly mutagenic lesion 8-oxo-deoxyguanosine (8-oxo-Guo) and render damaged and carcinogenic DNA (Quinn et al., 2008).

Phase II metabolism of PAHs

Phase II metabolism includes conjugation of metabolites from phase I with small molecules catalysed by specific or glutathione S-transferases (GSTs). SULTs have been shown to activate some metabolites of PAHs such as 7, 12-dimethylbenz(a)anthracene and its methyl-hydroxylated derivatives, in different tissues (Chou et al., 1998). Polymorphisms of SULTIAI have been associated with PAH-DNA adduct levels (Tong et al., 2003). Like sulfation, glucuronidation produces polar conjugates that are readily excreted. Oxygenated benzo (a) pyrene derivatives are common substrates of UDP-glucuronyl-transferase (Bansal et al., 1981), the resulting metabolites, I-hydroxypyrene glucuronide, and the parental I-hydroxypyrene are used as biomarkers of PAH exposure (Strickland et al., 1994). Finally, GSTs are also involved in conjugation of PAH derivatives. Glutathione conjugates are further metabolized to mercapturic acids in the kidney and are excreted in the urine. On the other hand, polymorphisms of phase II metabolism are associated with carcinogenesis and with DNA damage. For instance, there is an important association between GSTMI gene polymorphism and the DNA adduct levels (Binkova et al., 2007). The classification of some PAHs by some agencies and their carcinogenic tendencies as shown in table 3.

1.7.1. Fate of PAHs in Soil and Groundwater Environment

Low molecular weight (LMW) PAHs (two or three rings) are relatively volatile, soluble and more degradable than are the higher molecular weight compounds. High molecular weights (HMW) (four or more rings) sorb strongly to soils and sediments and are resistant to microbial degradation (Sikkema et al., 1995).

Because of the very low water solubility and high Kow values, they will tend to be sorbed to the organic matter in the soil instead of being solubilized in the infiltrating water and through this be transported downwards to the groundwater reservoirs. The sorption process is therefore counteractive to efficient biodegradation since it will decrease bioavailability (Zhang et al., 1998). Bacterial strains that are able to degrade aromatic hydrocarbons have been repeatedly isolated mainly from soil. These are usually gram negative bacteria (especially germs Pseudomonas). It has been claimed that a slow sorption following the initial rapid and reversible sorption lead to a chemical fraction that is very resistant to desorption. This phenomenon is called aging, and the existence of such a desorption – resistant residues may increase with time as the compound stay in the soil (Hatzinger and Alexander, 1995). PAHs have also been shown to be partitioned or incorporated more or less reversibly into the humic substances of the soil after partial degradation and thereby be even more immobilized in the soil (Kastner et al., 1999; Ressler et al., 1999). They also show very low aerobic degradability depending on the environmental conditions and the available concentration. Only two-and three-ringed components have been shown to be degraded under anaerobic conditions with nitrate or sulphate as the terminal electron acceptor (Mihelic and Luthy, 1988; Coates et al., 1996). Low concentrations of bacteria have a strong influence on the biodegradation of such hydrophobic compounds, and some studies have indicated that the process stops below a certain threshold concentration (Alexander, 1985). The low mobility and persistence means that PAHs can stay in the soil for decades, and even at sites with contamination dating at least fifty (50) years back with 4- or 5- ringed PAHs found near the soil surface.

1.7.2. Fate of PAHs in Air and their Ecotoxicological consequences

PAHs are usually released into the air or they evaporate into the air when they are released to soil or water. PAHs often adsorb to dust particles in the atmosphere, where they undergo photo oxidation in the presence of sunlight, especially when they are adsorbed to particles. This oxidation process can break down the chemicals over a period of days to weeks. Since PAHs are generally insoluble in water, they are generally found adsorbed in particulates and precipitated in the bottom of lakes and rivers or solubilized in any oily matter which may contaminate water, sediments and soil. Mixed microbial populations in sediments/water systems may degrade some PAHs over a period of weeks to months. The toxicity of PAHs to aquatic organisms is affected by metabolism and photo-oxidation, and they are generally more toxic in the presence of ultraviolet light. PAHs have moderate to high acute toxicity to aquatic life and birds. PAHs in soil are unlikely to exert toxic effects on terrestrial invertebrates, except when the soil is highly contaminated. Adverse effects on these organisms include tumors, adverse effects on reproduction, development, and immunity. Mammals can absorb PAHs by various routes e.g. inhalation, dermal contact, and ingestion (ATSDR, 1995).

Plants can absorb PAHs from soils through their roots and translocate them to other plant parts. Uptake rates are generally governed by concentration, water solubility, and their physicochemical state as well as soil type. PAH-induced phytotoxic effects are rare, however the database on this is still limited. Certain plants contain substances that can protect against PAH effects, whereas others can synthesize PAHs that act as growth hormones. PAHs are moderately persistent in the environment and can bioaccumulate. The concentration of PAHs found in fish and shellfish are expected to be much higher than in the environment from which they were taken. Bioaccumulation has also been shown in terrestrial invertebrates, however PAH metabolism is sufficient to prevent biomagnifications (Katarina, 2011).

1.8 Human Health Effects

1.8.1 Acute or Short-term Health Effects

The effect on human health will depend mainly on the length and route of exposure, the amount or concentration of PAHs one is exposed to, and of course the innate toxicity of the PAHs (IPCS, 1998). A variety of other factors can also affect health impacts including subjective factors such as pre-existing health status and age. The ability of PAHs to induce short-term health effects in humans is not clear. Occupational exposure to high levels of pollutant mixtures containing PAHs has resulted in symptoms such as eye irritation, nausea, vomiting, diarrhea and confusion (IPCS, 1998). However, it is not known which component of the mixture were responsible for these effects and other compounds commonly found with PAHs may be the cause of these symptoms. Mixtures of PAHs are also known to cause skin irritation and inflammation. Anthracene, benzo (a) pyrene and naphthalene are direct skin irritants while anthracene and benzo (a) pyrene are reported to be skin sensitizers (cause an allergic skin response in animals and human) (Rom, 1998). Some PAHs have low acute toxicity, other more acutely toxic agents probably cause the acute symptoms attributed to PAHs. Hydrogen sulfide in roofing tars and sulfur dioxide in foundries are examples of concomitant, acutely toxic contaminants. Naphthalene, the most abundant constituent of coal tar, is a skin irritant, and its vapors may cause headache, nausea, vomiting, diaphoresis (Rom, 1998).

1.8.2 Chronic or Long-term Health Effect

Health effects from chronic or long-term exposure to PAHs may include decreased immune function, cataracts, kidney and liver damage (e.g. jaundice), and breathing problems, asthma – like symptoms, and lung function abnormalities, whereas repeated contact with skin may induce redness and skin inflammation (IPCS, 1998). Naphthalene, a specific PAH, can cause the breakdown of red blood cells if inhaled or ingested in large amounts.

Many PAHs are only slightly mutagenic or even non-mutagenic in vitro. However, their metabolites or derivatives can be potent mutagens (Gupta et al., 1991). Reported health effects associated with chronic exposure to coal tar and its by-products (e.g. PAHs) are:

• Skin: erythema, burns, and warts on sun-exposed areas with progression to cancer. The toxic effects of coal tar are enhanced by exposure to ultraviolet light.

• Eyes: irritation and photosensitivity

• Respiratory system: cough, bronchitis, and bronchogenic cancer.

• Gastrointestinal system: leukoplakia, buccal-pharyngeal cancer and cancer of the lip.

• Hematopoietic system: leukemia (inconclusive) and lymphoma.

• Genitourinary system: hematuria and kidney and bladder cancers (Rom, 1998).

1.8.3 Carcinogenicity

The carcinogenicity of certain PAHs is well established in laboratory animals. Both the International Agency for Research on Cancer (IARC, 1987) and US EPA (1994) classified a number of PAHs as carcinogenic to animals and some PAH-rich mixtures as carcinogenic to humans. The EPA has classified seven PAH compounds, as probable human carcinogens these include, Benz (a) anthracene, Benzo (a) pyrene, Benzo (b) fluoranthene, Benzo (k) fluoranthene, Chrysene, Dibenz (a, h) anthracene and Ideno (1,2,3-cd) pyrene.

Researchers have reported increased incidences of skin, lung, bladder, liver and stomach cancers, as well as injection-site sarcomas, in animals (Blanton 1986, 1988). Animal studies show that certain PAHs also can affect the hematopoietic and immune systems and can produce reproductive, neurologic, and developmental effects (Dasgupta and Lahiri, 1992; Zhao, 1990). It is difficult to ascribe observed health effects in epidemiological studies to specific PAHs because most exposures are to PAH mixtures. Increased incidences of lung, skin, and bladder cancer are associated with occupational exposure to PAHs. Epidemiologic reports of PAH-exposed workers have noted increased incidences of skin, lung, bladder, and gastrointestinal cancer. These reports however provide only qualitative evidence of the carcinogenic potential of PAHs in humans because of the presence of multiple PAH compounds and other suspected carcinogens. Some of these reports also indicate the lack of quantitative monitoring data (Hammond, et al., 1976; Lloyd, 1971).

1.8.4 Teratogenicity

Embryotoxic effects of PAHs have been described in experimental animals exposed to PAHs such as benzo (a) anthracene, benzo (a) pyrene, and naphthalene. The laboratory studies conducted on mice have demonstrated that ingestion of high levels of benzo (a) pyrene during pregnancy resulted in birth defects and decreased body weight in the offspring. It is not known whether those effects can occur in humans. However, the centre for children’s environmental health reports studies that demonstrate that exposure to PAH pollution during pregnancy is related to adverse birth outcomes including low birth weight, premature delivery, and heart malformations.

High prenatal exposure to PAH is also associated with lower 1Q at age three, increased behaviourial problems at ages six and eight, and childhood asthma. Cord blood of exposed babies shows DNA damage that has been linked to cancer. (IARC, 2010).

1.8.5 Genotoxicity

Genotoxic effects for some PAHs have been demonstrated both in rodents and in vitro tests using mammalian (including human) cell lines. Most of the PAHs are not genotoxic by themselves and they need to be metabolized to the diol epoxides which react with DNA, thus inducing genotoxic damage. Genotoxicity plays important role in carcinogenicity process and may be in some forms of developmental toxicity as well (IARC, 2010).

1.8.6 Immunotoxicity

PAHs have also been reported to suppress immune reaction in rodents. The precise mechanisms of PAH-induced immunotoxicity are still not clear; however, it appears that immuno supression may be involved in the mechanisms by which PAHs induce cancer (IARC, 2010).

1.8.7 Effect of PAHs on Pathogenic Change

A key factor in PAH toxicity is the formation of reactive metabolites. Not all the PAHs are of the same toxicity because of differences in structure that affect metabolism.

Another factor to consider is the biologic effective dose, or the amount of toxics that actually reaches the cells or target sites where interaction and adverse effects can occur. Because of solid state, high molecular weight and hydrophobicity PAHs are very toxic to whole cells. CYPIAI, the primary cytochrome P-450 isoenzyme that biologically activates benzo (a) pyrene, may be induced by other substances (Kemena et al., 1988; Robinson et al., 1975).

The mechanism of PAH-induced carcinogenesis is believed to be via the binding of PAH metabolites to deoxyribonucleic acid (DNA). Some parent PAHs are weak carcinogens that require metabolism to become more potent carcinogens. Diol epoxides – PAH intermediate metabolites – are mutagenic and affect normal cell replication when they react with DNA to form adducts. A theory to explain the variability in the potency of different diol epoxides, “the bay theory”, predicts that an epoxide will be highly reactive and mutagenic if it is in the “bay” region of the PAH molecule (Jerina, 1976 and 1980; Weis, 1998). The bay region is as indicated in Figure 3 below using the structure of Benzo(a) pyrene, Chrysene and Dibenz(a,h) anthracene

REFERENCES

Abrantes, R., Assunção, J.V., Pesquero, C.R., Bruns, R.E. and Nobrega, R.B. (2009) emission of polycyclic aromatic hydrocarbons from gasohol and ethanol vehicles. Atmospheric Environment, 43:648-654.

Agency for Toxic Substances and Disease Registry (ATSDR) (1995). Toxicological Profile for Polyaromatic hydrocarbons-Update US Department of Health and Human Services, Atlanta, GA: 150-155.

Agersted, M.J. and Slog, K. (2005). Review: Genotoxicity of heat processed foods. Mutation Research, 574:156-172

Alexander, M. (1985). Biodegradation of Organic Chemicals. Environmental Science Technology, 18:106-111.

Al-Jedah, J.H., Ali, M.Z. and Robinson, R.K. (1999). The nutritional importance to local communities of fish caught off the coast of Quatar. Nutritional Food Science, 6:288-294.

Ambrosone, C.B., Freudenhein, J.L. Graham, S., Marshall, J.R., Vena, J.E. and Brasure, J.R. (1995). Cytochrome p450/A and glutathione S. transferase (MI) genetic polymorphisms and Post Menopausal Breast Cancer Risk. Cancer Research 55 (16): 3483-3485.

Annual Book of ASTM standards (2005). Extraction of solid Waste sample for chemical analysis using soxhlet extraction, environmental assessment, hazardous substances and oil spill responses, Practice for D5369, 11:196-201.

Badiali, D., Marcheggiano, A., Pallone, F., Paoluzi, P., Bausano, G. and Tannoni, C. (1985). melanosis of the rectum in patients with chronic constipation. Dis Colon Rectum, 28(4): 241-245.

Baek, S., Goldstone, M., Kirk, P., Lester, J. and Perry, R. (1991). phase distribution and particle size dependency of polycyclic aromatic hydrocarbons in the urban atmosphere. Chemosphere, 22:503-520.

Bansal, S.K., Zaleski, J. and Gessner, T. (1981). Glucuronidation of oxygenated benzo(a)pyrene derivation by UDP-glucuronyltransferase of nuclear envelop. Biochemical and Biophysical Research Communications, 98(1):131-139.

Request Complete Work

Get Complete Work Now!