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3.0 FORMS AND TRANSFORMATIONS

This chapter is based primarily on information presented in Moore and Ramamoorthy (1984) and Neff (1979). Where appropriate, other sources of information were also referenced.

3.1 Physico-chemical properties

Polycyclic aromatic hydrocarbons are non-polar, hydrophobic compounds, which do not ionize. As a result, they are only slightly soluble in water (Table 1). In general:

(1) PAH solubility in water decreases as the molecular weight increases.
(2) Alkyl (i.e., CH2- group) substitution of the aromatic ring results in an overall decrease in the PAH solubility, although there are some exceptions to this rule. For example, Benz[a]anthracene is less soluble than either methyl- or ethylbenz[a] anthracene.
(3) Molecules with a linear arrangement tend to be less soluble than angular or perifused1 molecules. For instance, anthracene is less soluble than phenanthrene, and naphthacene is less soluble than chrysene or benz[a]anthracene.

The solubility of PAHs in water is enhanced three- to four-fold by a rise in temperature from 5 to 30 °C. Dissolved and colloidal organic fractions also enhance the solubility of PAHs which are incorporated into micelles (a micelle is composed of an aggregate of surface-active molecules, or surfactants, each possessing a hydrophobic hydrocarbon chain and an ionizable hydrophilic group) (Neff, 1979).

Vapor pressure characteristics determine the persistence of PAHs in the aquatic environment. Two- to 3-ring PAHs are very volatile, while PAHs with 4 or more rings show insignificant volatilizational loss under all environmental conditions (Moore and Ramamoorthy, 1984).

Due to their hydrophobic nature, PAHs entering the aquatic environment exhibit a high affinity for suspended particulates in the water column. As PAHs tend to sorb to these particles, they are eventually settled out of the water column onto the bottom sediments. Thus, the PAH concentrations in water are usually quite low relative to the concentrations in the bottom sediments (Moore and Ramamoorthy, 1984). The sorptive characteristics of PAHs have been exploited in waste treatment processes such as coagulation, flocculation, sedimentation, and filtration with sand or activated carbon.

3.2 Photo-chemical transformations

PAHs are degraded through the process of photooxidation. The photo-induced oxidation of PAHs in the aqueous phase is brought about by singlet oxygen, ozone, HO- radical, and other oxidants. Photooxidation by singlet oxygen appears to be the most dominant process for the breakdown of PAHs and other organics in water (Zafiriou, 1977). Under ozone and light, the half-lives of several PAHs vary between a few minutes to a few hours. The most common products of photolysis are endoperoxides that undergo secondary reactions to yield a variety of products including diones (Figure 3).

PAHs differ in their sensitivity to photooxidation. Nagata and Kondo (1977) studied photodegradation of several PAH compounds in mixed acetone-water or carbon tetrachloride(CCl4)-water solvents. Anthracene, phenanthrene, and benz[a]anthracene were the most sensitive PAHs, whereas chrysene, fluorene, pyrene, and benzo[a]pyrene were relatively resistant to photodegradation. Since the solvents used (e.g., acetone versus CCl4) may exert strong influence on photosensitivity of PAHs, caution must be exercised in predicting the fate of these compounds in natural waters based on laboratory observations.

PAHs attached to particulate matter are more susceptible to photolysis than PAHs in solution. Also, the oxidative pathway for the sorbed PAHs is different from those in solution, and is not intermediated by endoperoxides in yielding quinones or diones (a general name for quinones under IUPAC rules) as the end product. For instance, in the presence of light, anthracene adsorbed to alumina or silica gel is oxidized to anthraquinone. Studies with particulate-associated benzo[a]pyrene show that rates of degradation increase with increasing dissolved oxygen concentration, temperature and light intensity. In a water column, the rate of photodegradation will decrease with depth as a result of (a) decrease in light intensity through absorption and scattering by water and suspended solids, and (b) decrease in temperature and

Figure 3

Photolysis products of 9,10-dimethyl anthracene and benzo[a]pyrene

dissolved oxygen. Photooxidation of PAHs is negligible in bottom sediments (Neff, 1979; Moore and Ramamoorthy, 1984).

Chlorination and ozonation, treatments used for destruction of pathogens in drinking water and sometimes oxidation of organics in industrial wastewater effluents, have been thought to remove PAHs from water. Harrison et al. (1976a, b) studied the influence of chlorination on eight PAHs (i.e., fluoranthene, pyrene, benz[a]anthracene+chrysene, benzo[b+j+k]fluoranthene, benzo[a+e]pyrene, perylene, indeno[1,2,3-cd]pyrene, and benzo[ghi]perylene). The rate of oxidation was dependent on PAH type (e.g., pyrene oxidized most rapidly and fluoranthene most slowly), temperature (the oxidation rate increased slightly from 5 to 20 °C), and pH (PAH concentration in water decreased with decreasing pH). These investigators also evaluated the efficiency of water treatment processes, including chlorination, for removal of these PAHs from drinking water (the concentrations among 8 PAHs ranged from 0.069 to 0.15 µg/L, and the total PAH concentration was 0.628 µg/L). It was found that (a) filtration removed 50% of the total PAH attached to fine particulates in water, while (b) chlorination, at 17 °C and pH = 7.5, removed about 60% of the PAHs remaining in the filtered water. The total PAH concentration in water was 0.13 µg/L after treatment. Ozonation was less efficient than chlorination for the removal of PAHs from water.

Chlorine and ozone react with PAHs to produce quinones and polychlorinated aromatics, some of which may be highly toxic to aquatic organisms (Green and Neff, 1977).

3.3 Biological transformations

3.3.1 Bacteria and fungi

PAHs are subject to biodegradation by microorganisms present in soil, sewage, and water. Microbial metabolism of PAHs may result in either complete or incomplete hydrocarbon degradation, depending upon several environmental (e.g., pH, temperature, dissolved oxygen and redox state) and molecular factors (e.g., PAH type including the number and position of fusion of aromatic rings in the molecule). Lower molecular weight PAHs tend to oxidize completely to form CO2 and H2O while the heavier PAHs will degrade partially to yield various oxygenated metabolites (e.g., various phenolic and acid metabolites, cis-dihydrodiol, etc.).

Lee and Takahashi (1977) studied the degradation of fluorene (initial concentration = 30 µg/L), naphthalene (initial concentration = 50 µg/L), methylnaphthalene (initial concentration = 50 µg/L), and benzo[a]pyrene (initial concentration = 16 µg/L) by marine bacteria isolated from various depths in a controlled ecosystem enclosure in Saanich Inlet, British Columbia. Water in the enclosure was contaminated with No. 2 fuel oil to a concentration of 10-20 µg/L total non-volatile petroleum hydrocarbons. Prior to the addition of oil, naphthalene (10 µg/L/d) and methylnaphthalene (10 µg/L/d) degraded slowly, while fluorene and B[a]P were not metabolized at all in 48 h. Three days after the oil dosing, the metabolism (or degradation) of naphthalene (26 µg/L/d) and methylnaphthalene (250 µg/L/d) increased greatly, but B[a]P degraded at a barely detectable rate (1.0 µg/L/d) while fluorene was still unmetabolized.

Microbial degradation of PAHs is one of the main processes responsible for removing these substances from bottom sediments and the water column. Delaune et al. (1981) noted in their studies that the rate of bacterial metabolism of PAHs in estuarine sediment was significantly lower in acidic, anoxic conditions. Poor water quality or heavy pollution of a water body may increase the residence time of PAHs.

This does not imply that PAHs may reside in the bottom sediments indefinitely. Anaerobic and facultative bacteria present on the sediments are also capable of metabolizing these substances although at a much slower rate than their aerobic counterparts (Delaune et al., 1981). The residence time of a PAH in sediment may thus be longer in anaerobic conditions, but biotransformation will still be occurring. Should the compounds be located deep within the sediment layer, however, degradation may or may not occur depending upon the sediment structure and bioavailability of the PAH.

The degradation of PAHs by fungi is unlike bacterial degradation, but resembles that in mammals as a result of fungi possessing a cytochrome P-450 (a heme protein) enzyme system. For instance, the fungus Cunninghamella elegans degrades naphthalene by the arene oxide-trans-naphthalene dihydrodiol pathway characteristic of mammals (Ferris et al., 1973).

The degradation of PAHs in water, sediment, and soils is shown in Table 3.

TABLE 3
Biodegradation of PAHs in water, sediment, and soils
(From Lee & Ryan, 1983; Heitcamp & Cerniglia, 1987; Sims, 1986; Niimi & Palasso,1986)

Biological Component

PAH

Biodegradation rate

Comment

Water (uncontaminated & from heavily oiled river)+

NA

125 to 320-d half-life (14-d)

10C (22C)

 

2-MNA

390 to 530-d half-life (16-d)

7C (22C)

 

PH

180-d half-life (36-d)

8C (27C)

Sediment (uncontaminated & from heavily oiled river)+

ANTH

95 to 141-d half-life (57-d)

18C

 

B[a]ANTH

1 100-d half-life (16-d)

15C

 

FL

37-d half-life

10C

 

CH

510-d half-life (79-d)

10C

Water & sediment (from pristine contaminated, with petrogenic, ecosystems)+

NA

4.4-wk half-life (2.4-wk)

22C

 

2-MNA

20-wk half-life (14-wk)

22C

 

PH

18-wk half-life (4-wk)

22C

 

PY

not detected (34-wk half-life)

22C

 

B[a]P

not detected (>200-wk half-life)

22C

Soil ( with &without amendment)*

FL

64-d half-life (39-d)

 
 

PH

69-d half-life (23-d)

 
 

ANTH

28-d half-life (17-d)

 
 

FLAN

104-d half-life (29-d)

 
 

PY

73-d half-life (27-d)

 
 

B[a]ANTH

123-d half-life (52-d)

 
 

CH

70-d half-life (42-d)

 
 

B[b]FLAN

85-d half-life (65-d)

 
 

B[k]FLAN

143-d half-life (74-d)

 
 

B[a]P

91-d half-life (69-d)

 
 

B[ghi]PERY

74-d half-life (42-d)

 
 

D[ah]AN

179-d half-life (70-d)

 
 

I[123-cd]PY

57-d half-life (42-d)

 

Rainbow trout (O. mykiss) (mean fish wt. 715-875 g)

FL

7-d half-life

amount fed=3.95 mg

 

PH

9-d half-life

amount fed=3.51 mg

 

ANTH

7-d half-life

amount fed=3.55 mg

 

FLAN

6-d half-life

amount fed=3.28 mg

+ half-lives for contaminated systems are in parenthesis; *half-lives with soil amendments (i.e., manure and lime) are in parenthesis

3.3.2 Animals

In animals, the mixed-function oxygenase(or oxidase) (MFOs) enzyme systems are responsible for the biotransformation of PAHs and other exogenous (e.g., xenobiotics or foreign compounds such as PCBs, pesticides, etc.) as well as endogenous organic substances (e.g., steroids and hormones). The MFO systems are usually associated with the endoplasmic reticulum of microsomal tissues located in the livers of vertebrates and the hepatopancreas of invertebrates; they have also been found in other organs of both groups. Not all invertebrates and vertebrates possess the MFO systems though this may be due to the lack of appropriate technology to detect these enzymes rather than the lack of the system.

The function of the MFOs is primarily to detoxify xenobiotics by converting these lipophilic materials into a more water soluble form, thus expediting their excretion from the organism. Detoxification of PAHs is not a simple process. Before formation of non-toxic and harmless end products by various enzymatic and nonenzymatic reactions, PAHs are converted to arene oxide intermediates followed by formation of derivatives of trans-dihydrodiols, phenols, and quinones. These intermediate products are known to be toxic, carcinogenic, and/or mutagenic.

Biological half-lives for some PAHs in rainbow trout are shown in Table 3.

Aquatic organisms may serve to remove a significant fraction of these compounds from the body of water. Pelagic organisms may take up PAHs directly from the water column or benthic organisms may absorb these substances from contact with both the bottom sediments and the overlying water. Considering the tendency of light molecular weight PAHs to volatilize from the water and of heavier PAHs to settle out with the sediments, it seems logical to assume that pelagic animals are exposed to lower overall PAH concentrations. However, uptake of these compounds tends to occur much more rapidly in the solubilized form. Therefore, in a high concentration, short exposure situation, pelagic organisms may actually be more at risk than their benthic counterparts. The toxicity, carcinogenicity, and mutagenicity of PAHs vary with the molecular weight of the compound, the degree of alkylation, and with the mode of accumulation (water, food or sediment) by the organism ( Neff, 1979; Moore and Ramamoorthy, 1984). Thus, the effects of these compounds upon an aquatic organism are not only highly dependent on the source of PAHs, but also upon the feeding behavior and habitat of the particular species.

3.3.3 Terrestrial plants

Terrestrial plants can take up PAHs through their roots and/or leaves and translocate them to various plant parts (Edwards, 1983). However, relatively little is known about the fate of PAHs within the plants. Dorr (1970) found a decline in B[a]P concentrations in rye plants after 30 days of growth, following a period (20 days) of increasing concentrations due to uptake from nutrient solution and soil containing the PAH. The decline in B[a]P concentration was attributed to degradation or chemical changes in B[a]P within the plants. Using 14C-B[a]P, Durmishidze et al. (1974) demonstrated chemical transformations of B[a]P (mostly to organic acids) within a number of plant species. Durmishidze (1977) reported similar results with both B[a]P and B[a]ANTH. The amount of B[a]P catabolized over a 14-d period varied from 2 to 18% of the B[a]P assimilated and depended upon plant species. The catabolism of anthracene by soybeans was demonstrated by Edwards et al. (1982). More recently, Negishi et al. (1987) demonstrated that a soybean leaves can oxidize B[a]P to its alcohols that are qualitatively similar to those produced by mammalian microsomes and eukaryotic microorganisms.

1 Polycyclic compounds in which two rings have two, and only two, atoms in common are said to be "ortho-fused". Such compounds have n common faces and 2n common atoms. Polycyclic compounds in which one ring contains two, and only two, atoms in common with each of two or more rings of a contiguous series of rings are said to be 'ortho- and peri-fused'. Such compounds have n common faces and fewer than 2n common atoms.

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