Distribution in Aquatic Compartments

39

40 and high organic carbon content (Karickhoff et al., 1979; Maruya et al., 1996; Tolosa et al., 2004).

Fig. 2.12. A conceptual model for the sequestration of hydrophobic organic contaminant by geosorbents.

NAPL= nonaqueous-phase liquids; SOM= sorbent organic matter. Sorption mechanisms include (A) absorption into amorphous or “soft” natural organic matter or NAPL; (B) absorption onto condensed or

“hard” organic polymeric matter or combustion residue (e.g. soot); (C) adsorption onto water-wet organic solvent (e.g. soot); (D) adsorption onto exposed water-wet mineral surfaces (e.g. quartz); (E) adsorption onto microvoids or microporous mineral (e.g. zeolites) with porous surface at water saturation <100%

(adopted from Luthy et al., 1997).

PAHs in surface sediments are, in fact, heterogeneously distributed among the various particle-size fractions (e.g. Maruya et al., 1996; Budzinski et al., 1997; Tolosa et al., 2004; Prahl

& Carpenter, 1983; Simpson et al., 1998; Wang et al 2001; Rockne et al., 2002; Ahrens &

Depree, 2004). PAH enrichment is not found solely in the mud fraction. It has also been observed in the sand fraction, which can also carry significant loads of various PAHs. For example, Rockne et al. (2002) found that PAH levels tend to increase with increasing grain size in sediment samples taken from Piles Creek and Newton Creek in the New York/New Jersey harbor area. Similar trends were reported by many other studies (e.g. Ahrens & Depree, 2004;

Oen et al., 2006; Simpson et al., 1998). Therefore, in this study we chose to investigate the PAH distributions in two general groups of sediment sizes as classified using the Wentworth-Udden scale: sand (coarse) 2 mm – 63μm; and mud (fine) <63μm, which includes the silk and clay fractions.

Sedimentary organic matter (SOM) controls the distribution of PAH in those fractions.

The PAH sorption capacity of SOM depends both on its structure and composition (Grathwohl, 1990; Huang & Weber, 1997; Johnson et al., 2001). Structure and composition vary among

41 sources e.g. peat, fragmented plant materials, black carbon, and kerogen (Grathwohl, 1990). But SOM is essentially the product of diverse geochemical alterations, ranging from biopolymer precursors (e.g. carbohydrate, protein, lipids, lignin, tannin and pigments) to geopolymers (e.g.

fulvic, humic, humin substances and kerogen) through complex diagenesis processes. During diagenesis, SOM experiences compositional changes in polarity and aromatic carbon content which controls its reactivity with hydrophobic organic compounds (Garbarini & Lion, 1986;

Gauthier et al., 1987; Luthy et al., 1997; Chiou et al., 1998). Therefore, a correlation between PAH and SOM is often helpful for pinpointing the significant role of organic matter (e.g.

Karickhoff et al., 1979; Means et al., 1980; Kim et al., 1999; Guinan et al., 2001; Maskaoui et al., 2002; Viguri et al., 2002). Such a correlation applied to sediment size fractions can help elaborate the preferential particle associations of PAHs (e.g. Prahl and Carpenter, 1983).

Coarse and fine sediment fractions contain different types of organic matter which affect PAH distributions. Increasing PAH content for the coarse sediment fraction has been acknowledged, due to the presence of organic particles that have a high affinity for PAH sorption. This is despite the fact that the OM fraction normally represents only a tiny portion of the total sediment mass. Ghosh et al. (2003) petrographically examined carbonaceous particles (coal, coke, charcoal, pitch, cenospheres, and wood). These particles are typically in the size range of 250μm–1mm and comprise only about 5-7% of the total mass, however, they account for 90% of adsorbed PAHs. Oen et al. (2006) emphasized that the presence of decomposed vegetable debris and shiny black particles dramatically increases the PAH content of the sand fraction. Since combustion processes leave behind soot and black carbon materials as residue, it is entirely possible that PAHs have already been adsorbed onto those particles before they ever reach the sediment. Therefore, examining the PAH content of the various sediment size fractions can provide important information on the mode of PAH transportation into the aquatic environment. On the other hand, humic substances are organic matter specifically associated with the fine fraction. This is due to the fact that the sorption of humic substances (normally in the form of dissolved organic matter, DOM) onto the fine fraction is typically a direct function of uptake onto its exceedingly large surface area. However, PAHs may be reluctant to associate with fine fraction-OM in the presence of combustion-derived OM (e.g. Prahl & Carpenter, 1983). Furthermore, different sources of SOM in the fractions have also been recognized. Evans et al. (1990) illuminated two different types of organic materials which were responsible for bimodal distribution of PAHs in coarse (>250μm – 2mm) and fine (<63 μm) fractions. First, fragmented plant materials are assumed to be responsible for high levels of organic matter (OM) in the coarse fraction. This is in addition to combustion-associated particles such as coal, soot or black carbon. Second, condensed organic matter (humic substances) is mostly associated with the fine fractions. Therefore, PAH enrichment in the sand fraction might be an indication of a strong anthropogenic source for PAHs. An investigation of these PAH-sized fraction

42 associations would provide us with significant additional information about PAH delivery modes into aquatic environments, whether as a result of combustion-derived particle association or from the sorption equilibrium on the surface of fine sediments.

2.3.2. Suspended Particulate Matter and Water

Suspended particulate matter (SPM) is the main carrier by which most terrestrially-derived materials - including anthropogenic pollutants - are transferred from land to aquatic environments. It also provides a fundamental link between chemical constituents and the water column, bed sediments and food chain (Turner & Milward, 2002; Suzumura et al., 2004). The term SPM refers to all particulate matter with different natures and origins, but is operationally defined as those materials retained by a filter with a specific pore size. Therefore, the definition of SPM can operationally vary between studies. However, a maximum pore size of 0.7 μm (GF/F) is often employed for differentiating between the particulate and dissolved phases from river, estuary and coastal waters for pollution and geochemistry studies (e.g. Zhou et al., 1998;

Luo et al., 2006; Suzumura et al., 2004; Boldrin et al., 2005; Gebhardt et al., 2004).

PAHs move from terrestrial environments into the oceans via river, estuary, and coastal pathways. The concentrations of PAHs and SPM have been shown to be positively correlated (e.g. Fernandes et al. 1997). This means that PAH loads entering the ocean are closely linked to the overall SPM load emerging from riverine, estuarine and coastal drainage basins. Any change in the hydrologic cycle (e.g. increased precipitation or extreme floods caused by climate change) can therefore modify and possibly intensify the temporal load of PAHs. The implications of flood events for PAH loading have been studied. For example, Witt and Siegel (2000) observed a significant flux (two orders of magnitude higher than normal) of PAHs into the Baltic Sea, stemming from municipal and industrial areas in the Oder River basin as a consequence of a 1997 flood event. Sicre et al. (2008) calculated that 90% of the annual load of particulate PAHs flowing from the Rhône River (France) into Mediterranean Sea took place during flood episodes in 1994. These facts are extremely important for tropical rivers, particularly in Sumatra where overall precipitation is high, and floods are occurring more frequently due to climate changes. Thus, the transport of pollutants via Sumatran waterways into the ocean should be expected to be significant.

Interactions between hydrophobic organic pollutants and SPM depend on the composition of SPM in the water column. This is due to the fact that different types of particles embody various binding capacities for specific, hydrophobic pollutants. In general, SPM in riverine, estuarine and coastal waters represents a composite of lithogenous, hydrogenous, biogenic and anthropogenic particles. Lithogenous particles are inorganic materials derived from the weathering of rocks and other substances in the Earth's crust, which are composed mainly of quartz and other aluminosilicate minerals. Hydrogenous matter is generated in-situ by chemical

43 processes, resulting in such materials as humic substances, carbonates and both iron and manganese oxides. They occur either as coatings or discrete phases. Biogenic particles include those stemming from microorganisms, plankton and the decaying remains of macroorganisms and terrestrial plant debris. Bio-particles can also refer to those derived from proteins, carbohydrates, lipids and pigments. Anthropogenic particles consist mainly of combustion by-products such as dust and fly ash, but also include other widely varied synthetic materials such as plastic, tar, solvents and surfactants (Turner & Millward, 2002). However, SPM particles are often divided into inorganic and organic particles based on their chemical properties. Particulate inorganic material (PIM) includes lithogenous matter (minerals and insoluble salts), whereas particulate organic materials (POM) are composed of a rather broad mixture of hydrogenous, biogenic and anthropogenic particles. PIM and POM are straightforward measures for the overall SPM composition in a sample. The latter also acts as an effective sorbent for hydrophobic organic pollutants.

The molecular composition of PAHs in particulate matter is somewhat different from that of those found in sediments. Luo et al. (2006) concluded that SPM samples taken from the Pearl River estuary contained large amounts of both 2- and 3-ring PAHs. On the other hand, these sediments were also characterized by high levels of 5- and 6-ring PAHs. Similar findings have been reported for other riverine and estuarine systems (e.g. Witt, 1995; Shi et al., 2005). These patterns were driven by the fact that the SPM was continually receiving fresh PAH inputs, either from the atmospheric deposition of combustion by-products or from direct oil spills, which are predominantly characterized by low molecular weight compounds (see 2.1.2). However, this is not always the case in every situation. PAH profiles in SPM samples represent the local conditions where the transfer of high molecular weight molecules into the aquatic environment occurs. This can also be mainly derived from land-water interactions. For instance, Witt &

Siegel (2000) observed that the distribution of PAHs resulting from Oder River floods were characteristic of high-combustion profiles, which are predominated by HMW compounds. On the other hand, the concentration of LMW PAHs in the water column and sediments was subject to great variation due to differing degradation processes (photonic and microbial). Therefore, examination of a specific PAH profile in SPM yields clues to the fate of PAHs in a particular environmental setting.

2.3.3. Water Solution as dissolved PAHs

In addition to SPM association, PAH compounds can alternately remain in solution as truly dissolved substances or also be bound by dissolved organic matter. Differentiation between those two aqueous fractions is important for particular purposes like bioavailability or toxicity studies (Hawthorne et al., 2007). However, for the general assessment of the distribution of PAH in natural waters, separation of those two aqueous fractions is not

44 substantial. This is due to the fact that the fate of truly dissolved PAHs is appreciably controlled by the existence of dissolved organic matter (DOM) as a geosorbent. Schlautman & Morgan (1993) observed in an experiment that PAH-DOM binding could be completed within a timeframe of only 3 minutes. As a consequence, the free fraction in natural waters is unstable and is readily associated with DOM. Thus, it is quite challenging in terms of analytical techniques to effectively separate these two aqueous fractions. And this is despite the availability of various analytical techniques such as fluorescence quenching, purging or sparging techniques, solid-phase microextracion (SPME), equilibrium dialysis, solubility enhancement, ultrafiltration, size exclusion chromatography, and liquid-liquid extraction (Burkhard, 2000 and references therein). This study simply considers these two fractions of PAHs as the "dissolved phase", which operationally defines dissolved PAHs as those passing through the GF/F. The most popular extraction techniques for determining dissolved PAHs are solid phase extraction (SPE) and stir bar microextraction (SBME) (Falcon et al., 2004;

Fernández-Gonzáles et al., 2007; Poerschmann et al., 1997).

The composition of unsubstituted PAHs in the dissolved phase depends on the solubility and hydrophobicity (Kow) of a given compound. The former decreases as molecular weight increases and the latter increases with increasing molecular weight (see Appendix 1). With regard to a compound’s solubility, we can expect a predominance of low molecular weight PAHs (2 and 3 rings) as compared to HMW compounds in the absence of DOM, salinity and pH effects. The opposite also holds true. Furthermore, angular compounds are more soluble than corresponding linear isomers, such as phenanthrene (1.18 mg/L) compared to anthracene (0.08 mg/L) or fluoranthene (0.26 mg/L) compared to pyrene (0.14 mg/L). With regards to Kow, we might assume that the presence of DOM should increase the concentration of HMW PAHs in the dissolved phase. However, interactions between PAHs and DOM prove themselves to be DOM source-dependent (Liu & Amy, 1993). Therefore, examining the concentration and composition of materials in the dissolved phase can reveal the relevance of particular environmental settings on the fate of particular PAHs.

2.4. The fate of PAHs in the water: a partitioning concept and the role of