POPs in Marine and Freshwater Environments

Chapter 8

POPs in Marine and Freshwater Environments Richard J. Wenning and Linda Martello ENVIRON, Emeryville, CA, USA

8.1 INTRODUCTION Persistent organic pollutants (POPs) are organic compounds of natural or anthropogenic origin that resist photolytic, chemical, and biological degradation. Examples include polychlorinated biphenyls (PCBs), dioxins and furans, many pesticides and certain metals, particularly mercury. Although many countries have banned or severely restricted the production and use of POPs in recent decades, these substances are pervasive and can be found in remote environments around the world [1,2]. It is well established that POPs have the ability to migrate long distances from their original source. Inputs of POPs from the atmosphere and surface waters, and rereleases from sediments and removal pathways such as volatilization and sedimentation may explain, in part, why countries that banned the use of certain POPs are experiencing less dramatic declines in environmental concentrations nearly a decade later [3]. Research conducted as part of the mandate under the Stockholm Convention to identify POPs of potential global concern has identified nearly 40 substances in three categories (pesticides, industrial chemicals, and by-products) for elimination (Annex A), restriction (Annex B), or work to reduce unintentional production (Annex C).1 While many of these substances have been the subject of extensive research and are well-represented in the literature, such as PCBs, dioxins, and pesticides such as Dichloro-diphenyl-trichloroethane (DDT), there are other classes of substances that have not received the same level of scrutiny or are only now being recognized as possible POPs. Five classes of substances were selected for discussion in this chapter because while not as frequently found in the literature they are now either formally recognized as POPs or as emerging contaminants with comparable

1. http://chm.pops.int/Convention/ThePOPs/ListingofPOPs/tabid/2509/Default.aspx. Environmental Forensics for Persistent Organic Pollutants. © 2014 Elsevier B.V. All rights reserved.

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physical and chemical properties to POPs. These chemicals warrant concern because of their broad global distribution and mounting evidence of potentially harmful effects to aquatic life. The five substances to be discussed in this chapter include certain polyfluorinated compounds (specifically perfluorooctane sulfonic acid and its salts and perfluorooctane sulfonyl fluoride, which are listed under Annex B), certain brominated flame retardant (BFR) compounds (specifically the tetra-, penta-, hexa- and heptabromodiphenyl ethers, which are listed under Annex A), certain pharmaceutical compounds and ingredients in personal care products, certain types of nanomaterials (NMs), and certain polyaromatic hydrocarbons (PAHs). POPs are typically hydrophobic (water-hating) and lipophilic (fat-loving) chemicals. In aquatic environments and soils they partition strongly to organic matter and avoid the aqueous phase. They also partition into the lipids of organisms rather than entering the aqueous environment of cells and become stored in fatty tissue. This results in the persistence of these chemicals in biota since metabolism is slow and POPs accumulate in food chains. POPs also tend to enter the gas phase under typical ambient temperatures. They can therefore volatilize from soils, vegetation and water into the atmosphere and, because of their resistance to breakdown reactions in air, they can travel long distances before being re-deposited. In both the marine and freshwater environments, several processes influence the fate and transport of POPs. It is well established that POPs distribute between particles, colloids, and the water phase. Atmospheric deposition across the air-sea interface is the main input route for POPs to the marine environment and the processes contributing to the air-sea exchange of POPs are diffusive vapor exchange, aerosol-vapor partitioning, precipitation scavenging of vapors and particle-sorbed chemicals, and dry particle deposition [5,6]. A second important route is POPs binding to settling particles in industrial or municipal effluents deposited to bottom sediment, where burial with sediment particles constitutes a long-term, and possibly permanent, removal process (Figure 8.1; [8]). When POPs sorb to particles and colloids, the freely dissolved concentrations decrease and, thus, also the bioavailability to aquatic organisms [9]. On the other hand, the sorption of POPs to colloids can enhance environmental transport due to the higher mobility of small colloids compared to larger particles. For instance, the groundwater transport of polychlorinated dibenzopdioxins and dibenzofurans (PCDD/Fs) from contaminated soil was found to be mediated by colloid-facilitated transport processes [10]. Colloids can also facilitate the transport of POPs over the thin stagnant aqueous films (i.e., diffusive boundary layers) that are present at the interface between water and sediment or on passive samplers [11,12]. The extent of sorption to particles and colloids and the distribution between different phases in the aquatic environment has a large impact on POP migration in the environment and fate processes, such as sedimentation, bioavailability, and degradation [13].

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Gas–particle partition Air–soil or air–vegetation exchange

Gas Dry deposition Wet deposition Photodegradation Air–water exchange Hydraulic transport Bioaccumulation

Dissolved phase

Sedimentation FIGURE 8.1 Main environmental processes during long-range atmospheric transport of POPs. Reproduced with permission from Fernandez and Grimalt [7].

The purpose of this chapter is to highlight the challenges associated with conducting environmental forensic work to better understand source inputs, environmental behavior, fate and transport, and analytical methods for measuring the five classes of substances listed above in the aquatic environment. The challenges are similar for both freshwater and marine environments. A brief summary of the current understanding of environmental levels is included that environmental forensic specialists may find useful for investigating sources and distribution in the environment.

8.2 POLYFLUOROALKYL COMPOUNDS The occurrence and fate of fluorinated compounds in the aquatic environment is recognized as an important emerging contaminant issue. Knowledge about this large and complex family of chemicals and their worldwide distribution, environmental fate, and transport pathways has advanced substantially since the 1990s, concurrent with improving analytical methods. Several reviews and extensive compilations of the science literature on the most common substances, such as perfluorooctane sulfonate (PFOS) and perfluoroalkyl sulfonic and carboxylic acids and their anions and salts, have been published recently [14 28]. Clarification of chemical terminology is the first challenge to understanding scientific research on sources, environmental levels, fate, and ecological and human health effects. This paper follows recommendations by Buck et al. [28] regarding the use of the term PFASs (singular PFAS) as an acronym for “perfluoroalkyl and polyfluoroalkyl substances” and the term PFCs

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(singular PFC) exclusively for “perfluorocarbons.” The acronym PFC has been used in official Kyoto Protocol documents since its adoption in 1997 to specifically designate perfluorocarbons [29], one of the families of greenhouse gases regulated by this important multilateral international agreement. PFCs are not addressed in this chapter.

8.2.1 Sources to the Environment Since 1950, PFASs and surfactants and polymers made with the aid of PFASs have been widely used in numerous industrial and commercial applications [30]. The carbon-fluorine bond is extremely strong and stable [31]. The chemical and thermal stability of a perfluoroalkyl moiety, in addition to its hydrophobic and lipophobic nature, lead to highly useful and enduring properties in surfactants and polymers into which the perfluoroalkyl moiety is incorporated [30,31]. Polymer applications include textile stain and soil repellents and grease-proof, food-contact paper [32]. Surfactant applications that take advantage of the aqueous surface tension-lowering properties include processing aids for fluoropolymer manufacture, coatings, and aqueous filmforming foams used to extinguish fires involving highly flammable liquids [30,31,33]. Numerous additional applications have been described [30].

8.2.2 Environmental Fate PFASs have been detected in the environment, wildlife, and humans as a consequence of their widespread use. The global regulatory community is specifically interested in “long-chain” perfluoroalkyl sulfonic acids (CnF2n11SO3H, n $ 6, PFSAs) and perfluoroalkyl carboxylic acids (CnF2n11COOH, n $ 7, PFCAs) and their corresponding anions [34,35], which have been shown to be more bioaccumulative than their short-chain analogs [36 39]. The PFSAs include perfluorohexane sulfonic acid (PFHxS), PFOS, other higher homologs, and their salts and precursors. The PFCAs include perfluorooctanoic acid (PFOA, sometimes called C8), other higher homologs, and their salts and precursors. Some PFCA precursors include chemicals known commercially as fluorotelomers. PFOS and PFOA are the two “long-chain” perfluoroalkyl acids most often reported and discussed in the scientific literature.

8.2.3 Environmental Levels PFASs are persistent and bioaccumulative and have been detected in various environmental matrices, including freshwater [40] and marine waters [41 43]. Forty PFASs from different classes are commonly detected in the aqueous environment at concentrations ranging between picogram and nanogram per liter (pg/L and ng/L, respectively) levels for individual compounds [44].

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A global comparison of PFAS levels from around the world is given in Figure 8.2. Generally, PFOA and PFOS are the dominant compounds in the aquatic environment. Mo¨ller et al. [45] report concentrations of PFASs in the North Sea range from 0.4 to 11.6 ng/L and identified the River Rhine and the River Scheldt as possible major sources to the North Sea and into the German Bight. In general, environmental concentrations of PFOA are higher than PFOS. For example, in a large study of PFOS and PFOA in surface water, sediment, and fish in Japan’s rivers, the concentration of PFOA was generally significantly higher than PFOS and other perfluorinated substances [46]. Investigation of PFAS levels in surface water, suspended particulate matter, and sediment in Tokyo Bay, Japan conducted by Ahrens et al. [25] suggests that the distribution of PFASs depends on physicochemical characteristics; short-chain PFCAs (C , 7) were exclusively detected in the dissolved phase, while longer chain (C . 7) substances appear to bind more strongly to

Antarctic, coastal area (a,b) Mid to South Pacific Ocean (b) Indian Ocean (b) Central to Eastem Pacific Ocean (b) Atlantic Ocean (a, c, d,e) Germany, coastal area (f,g) Korea, coastal area (h) China, Pearl River Delta (h) Hong Kong, coastal area (h) China, Dalina coastal area (i)

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PFOA concentration (pg/L) Antarctic, coastal area (a,b) Mid to South Pacific Ocean (b) Indian Ocean (b) Central to Eastem Pacific Ocean (c) Atlantic Ocean (a, c, d,e) Germany, coastal area (f,g) Korea, coastal area (h) China, Pearl River Delta (h) Hong Kong, coastal area (h) China, Dalina coastal area (i) Japan, coastal area (j)

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PFOS concentration (pg/L)

FIGURE 8.2 Concentrations (minimum, maximum, median (circles)) of PFOA and PFOS in seawater in the open ocean and coastal area in picogram per liter. Reproduced with permission from the Royal Society of Chemistry [44].

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particles. PFOA and PFOS (C=8) have higher water solubility [47], lower bioaccumulation potential [38], and lower sorption potential to sediment.

8.2.4 Analysis Methods Several comprehensive studies describe sample pretreatment, extraction, cleanup strategies, and instrument measurement of PFASs [18,48,49]. Sampling strategy, the use of certain sample and analysis materials, sample contamination, and matrix effects are vexing challenges for environmental sampling and testing [25]. Comparing results to published data of PFC levels in the environment must be done cautiously, because of the many different sampling methods, pretreatments, and instruments. Furthermore, different matrices (i.e., precipitation, groundwater, seawater, river water, lake water, and wastewater) with different characteristics (e.g., pH and organic carbon) have been shown to greatly influence test results. The most frequently used instrument for the measurement of PFCs is high-performance liquid chromatogram coupled with a tandem mass spectrometry operated in a negative electrospray mode (HPLC ESI-MS/MS) or high-resolution time-of-flight (TOF) MS [18,50]. The most common extraction method for aqueous samples is solid-phase extraction (SPE), which has been optimized by Taniyasu and coworkers to determine a wide range of PFASs, including short- and long-chain PFCs [51,52]. Alternatively, liquid liquid extraction (LLE) has been used without prior filtration, but this method is limited to the longer chain PFASs (C . 8) [53]. The sampling method, sampling period, sampling container, and sampling depth can have a significant influence on the results. For example, the vertical profile of PFASs in seawater has been shown to change by a factor of 24 109 in the surface microlayer compared to the corresponding subsurface water layer (.30 centimeter (cm) depth) [54]. Sample blank contamination has been attributable to adsorption on glass or polypropylene plastic sample containers [51,55]. For volatile neutral PFASs (e.g., fluorotelomer alcohols (FTOHs), perfluoroalkyl sulfonamides (FASAs), fluorotelomer acids (FTAs)), evaporation or degradation to ionic PFASs (e.g., PFCAs, PFSAs) is possible [56]. If filtration is necessary for water samples with high suspended particulate matter content, PFCs have been shown to adsorb to the filtration equipment and to filter material (e.g., glass fiber filter (GFF) or syringe nylon membrane filter); the filtration equipment itself may be a source of sample contamination [57,58]. The effect of matrix-induced signal suppressions during the instrumental analysis has been observed for some compounds depending on the extraction volume and the sample type [48,59]. A method based on combustion ion chromatography has been developed for the determination of total fluorine (TF), followed by fractionation of the samples to determine inorganic

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fluorine and organofluorine separately [60]. The method has been shown to detect TF in low microgram per liter levels in aqueous matrices by reducing high background levels. This method may help identify unknown PFASs in water. Recently, the quantification of branched fluorine isomers has become more important because of the possible relationship to relative toxicity among the perfluorinated compounds [61,62]. The fluorine isomer pattern may also be useful to identify the dominant source from either historical releases like electrochemical fluorination (ECF) fluorochemicals or current releases like telomer-derived products [63].

8.2.5 Summary While the primary sources of PFASs are understood, there are continuing efforts to better understand historical sources [64], fate and transport processes, and the numerous different precursors in the environment [64 66]. An important research topic, directly related to environmental fate and transport, is the question of how and how fast PFOS and PFOA in particular, as well as their homologs and precursors, are transported away from their emission sources over long distances in air and/or water [64,67 74].

8.3 PHARMACEUTICALS AND PERSONAL CARE PRODUCTS Pharmaceuticals and personal care products (PPCPs) are a diverse collection of thousands of chemical substances, including prescription and over-thecounter therapeutic drugs, veterinary drugs, fragrances, sunscreens, detergents, and cosmetics. Among this category of compounds, some are capable of disrupting the endocrine system of animals, including fish, wildlife, and humans; these substances are termed endocrine disrupting chemicals (EDCs). PPCPs known or suspected to have EDC properties are widely considered to be an emerging class of contaminants and considered generally to behave similarly to POPs because many PPCP substances resist degradation in the environment and bioaccumulate in the food chain [75]. PPCPs, however, are not formally listed as POPs under the Stockholm Convention and there is an active debate in regard to whether PPCPs fall into the category of POPs. Nonetheless, PPCPs have become emerging contaminants of concern because of their potential to affect drinking water supplies and the uncertain consequences of chronic low-level exposures to wildlife.

8.3.1 Sources to the Environment Municipal wastewater, attributed to the widespread use of PPCPs both in the home and in health care and personal care facilities, is the primary pathway

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by which chemicals in prescription and over-the-counter products find their way into the aquatic environment [76]. According to Boxall et al. [75], regulatory environmental risk assessment approaches for PPCPs consider releases to surface waters from wastewater treatment systems, aquaculture facilities, and runoff from fields, as well as releases to soils during biosolid and manure application, emissions from manufacturing sites, disposal of unused medicines to landfills, runoff of veterinary medicines from hard surfaces in farmyards, irrigation with wastewater, and the disposal of carcasses of treated animals. The release of pharmaceuticals from manufacturing facilities is heavily regulated and is not a major contributor to the environment.

8.3.2 Environment Fate Figure 8.3 illustrates possible sources and pathways for the occurrence of PPCP residues in the aquatic environment. Few studies trace the fate of PPCPs in wastewater treatment or in biosolids, and much of what is known is based on extrapolation from human and laboratory animal metabolic studies or insights gleaned from mass balance studies of general pharmaceutical usage in a population [78] or fate studies involving widely recognized substances, such as ibuprofen, 17α-ethinylestradiol, diatrizoate, and cyclophosphamide

Medicinal products for animal use

Medicinal products for human use

Excretion (hospital effluents) usually

Excretion (private households)

Waste disposal (unused medicine)

Excretion

Manure

Domestic waste

Municipal waste water

Sewage farms Sewage treatment plants (STPs)

Sewage sludges

Waste disposal site Soil

Surface water Groundwater Aqua cultures Pharmaceutical production plants

Drinking water

FIGURE 8.3 Possible sources and pathways for the occurrence of pharmaceutical residues in the aquatic environment. Reproduced from Th. Heberer, K. Reddersen and A. Mechlinski (2002) From municipal sewage to drinking water: fate and removal of pharmaceutical residues in the aquatic environment in urban areas. Water Science & Technology Vol 46 No 3 pp 81 88, with permission from the copyright holders, IWA Publishing [77].

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[79]. Daily and seasonal changes in PPCP levels associated with sewage discharge to freshwater are generally recognized [80], but the significance for source identification and environmental fate is largely unknown. Several excellent reviews of PPCPs in the aquatic environment reveal the need for research to better understand abiotic levels and effects on both humans and wildlife [81 87].

8.3.3 Environment Levels With advances in analytical technology, scientists have been able to detect trace levels of PPCP residues in the aquatic environment with several reports appearing in the literature over the past decade. In general, PPCP residue concentrations detected in drinking water are typically measured in the nanogram per liter range and typically at levels more than 1000-fold lower than minimum therapeutic doses [76].

8.3.4 Analysis Methods Xia et al. [79] summarize analytical methods for wastewater and biosolids, the two predominant pathways by which PPCPs enter the aquatic environment. SPE isolation and high-performance liquid chromatography electrospray ionization mass spectrometry (HPLC ESI-MS) analytical procedures are often used for the routine determination of the presence and concentration in surface water [88]. In a 2010 study of PPCPs in reservoirs, the New York City Department of Environmental Protection (DEP) relied principally upon two methods for PPCP analyses: Montgomery Watson (MWH) Method EDC2SR and Underwriters Laboratory (UL) Method 220. MWH Method EDC2SCR (a peer-reviewed isotope dilution based SPE LC MS/MS method using a sensitive API4000 instrument) was used to analyze and quantify 21 compounds, including most of the compounds detected frequently in a similar 2009 study. UL Method 220 was used to detect a broader range of 44 PPCPs than provided by MWH [89]. Recently, Lausier [90] used synchronous-scan fluorescence spectroscopy (SFS) to detect caffeine, 17α-ethynylestradiol, and triclosan in three freshwater lakes in Maine. Detection and quantification of a mixture of compounds in environmental samples using SFS is a novel and a relatively inexpensive method that could provide an approximate estimate of compound levels in water. The benefit of this method is that individual compounds can be detected rapidly in a mixture without prior separation [90].

8.3.5 Summary There is a general lack of environmental data on PPCPs in the aquatic environment in regard to sources, pathways, environmental fate, and potential

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long-term ecotoxicological effects. The literature suggests that different PPCPs have widely varying impacts and modes of action. Prioritizing those PPCPs most likely to cause adverse impacts to aquatic and terrestrial communities should be a focus of future research.

8.4 ENGINEERED NANOMATERIALS NMs have a less troublesome terminology than PFASs, but encompass an equally diverse and large family of substances. The definition of NMs, as adopted by the British Standards Institution [91], the American Society for Testing Materials [92], and the Scientific Committee on Emerging and Newly Identified Health Risks [93], is materials with one dimension ,100 nm. Scientists typically refer to engineered or manufactured NMs (hereafter referred to as ENMs) to distinguish industrial and commercial applications from sea salt, volcanic dust, particulate from forest fires, and other naturally occurring particles. Within this group of materials, nanoparticles (NPs) are defined as materials with at least two dimensions between 1 and 100 nm [92]. Environmental release of ENMs, and particularly NPs, into the aquatic environment poses new environmental problems. While there are quite a few studies focusing on the toxicologic effects of ENMs and NPs, research is lacking regarding potential exposures of aquatic organisms to these materials. Global trends suggest rapid increase in production and use [94]. However, measurements of environmental concentrations of ENMs are almost completely absent in the scientific literature, in large part due to the analytical challenges. Among the earliest studies, Benn and Westerhoff [95] and Kaegi et al. [96] reported detection in water, wastewater, and biosolids of two common applications, nanosilver (n-Ag) and nano-titanium oxide (n-TiO2), used in fabrics and exterior paints. Murr et al. [97] reported the presence of carbon nanotubes (CNTs) and fullerene nanocrystal in a 10,000-year-old ice core melt sample, suggesting a natural presence for some microparticles. There are generally five classes of ENMs—CNTs, ENMs containing metal and metal oxide, semiconductor nanocrystals, zero-valent metal ENMs, and dendrimers. By far the largest consumer product applications involve the use of CNTs and metal oxide ENMs, particularly n-Ag. Ionic silver is highly reactive, is readily adsorbed by both macroparticles and colloidal particles such as iron oxyhydroxides or natural organic matter in natural waters, and ranges in size from ,1 kilodalton (kDa) to .0.45 micrometer (μm) [98]. Nanosilver particles are one of the most commonly used ENMs owing to their strong antimicrobial activity [99].

8.4.1 Sources to the Environment ENMs enter the aquatic environment through controlled and uncontrolled atmospheric emissions and solid or liquid waste streams from several

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industrial sectors involving semiconductors; memory and storage technologies; display, optical, and photonic technologies; energy; biotechnology; paints and pigments; pharmaceuticals; textiles; and health care and personal care product manufacturers. Estimates of NP production are in the range of 500 tons/year and 50,000 tons/year for silver and TiO2, respectively [100]. Wastewater and biosolids are increasingly becoming a primary pathway to the aquatic environment as the use of ENMs in consumer products increases. Diffuse releases are widespread due to their rapidly expanding presence in a wide range of common consumer products, such as house paints, sunscreen lotions, cosmetics, and the wear and erosion of treated clothing and furniture fabrics from general use [101]. At present, researchers are generally challenged by the detection of microparticles in surface waters and determination if these substances are naturally occurring or ENM. As instrumentation capabilities improve, improvements are anticipated in identification and quantification of ENM in water samples with the potential to determine the source. Aside from incidental releases, nanotechnology is increasingly used in environmental remediation and several environmental treatment technologies [102]. Nanoremediation has emerged as a new discipline for application of reactive ENMs to facilitate transformation and detoxification of pollutants. Purposeful releases to the aquatic environment include the use of ENMs for remediation of contaminated soils and groundwater [103] and water and wastewater treatment [104,105]. The use of ENMs in environmental remediation will inevitably lead to the release of NPs to the environment and ecosystems.

8.4.2 Environmental Fate The environmental fate of ENM is reviewed elsewhere [106 109]. NPs released to the environment have the potential to contaminate soil, migrate into surface and groundwater, and interact with biota. Particles in solid wastes, wastewater effluents, direct discharges, or accidental spillages can be transported to aquatic systems by wind or rainwater runoff. With increasing control of fugitive releases arising within the manufacturing process, the biggest risks for environmental release come from spillages associated with the transportation of manufactured NPs from production facilities to other manufacturing sites, intentional releases for environmental applications, and diffuse releases from consumer products. Figure 8.4 illustrates the pathways and the behavior of ENMs in the environment. Comparatively little work has been done on ecological systems and few relevant data are available regarding the fate and behavior of manufactured NMs in the aquatic environment. Released ENMs may have a greater or lesser environmental impact than the starting materials, depending on the transformation reactions and the material. Little is known about the environmental behavior and the effects of released and transformed ENM [110]. ENM tends

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(a)

Non point source

AIR Wildlife and humans

UV degradation Deposition

Food chain Volatilization dust Sorption to organic matter

WATER

Food chain

Spillage

Source

Filter feeders Biodegradation

Sediment Chemical degradation

Benthic organisms Leaching

Soil

Aquifer

(b)

Aggregates Environmental transformation

Nanomaterial Aggregates

Free nanoparticles

Environmental transformation

Functionalized nanoparticles Nanoparticles composites Environmental transformation

Functionalized nanoparticles

Environment

FIGURE 8.4 Pathways illustrating the behavior of ENMs in the environment. (a) Most important pathways involving engineered NPs in the environment. (b) Modification of NPs in the environment. Reproduced with permission from Anal. Bioanal. Chem. 393 (2009) 81 95 [108].

to form aggregates that can be trapped or eliminated through sedimentation. Aggregates or adsorbed ENM are less mobile, but can undergo uptake by filter feeders and sediment-dwelling animals. For these reasons biomagnification in the food chain is possible, but too date few data are available to confirm it is occurring. Although the likelihood for ENMs to enter surface waters, and consequently drinking water sources, is high, few studies have investigated their

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fate in drinking water treatment processes. These studies have found that ENMs can be removed by conventional treatment, but that the removal efficiencies are highly dependent on the specific nature of the ENM and water characteristics [111,112]. Factors such as pH, organic matter content, and salt composition influence the size, aggregation, dissolution, and stability of ENMs in water. To date, few systematic studies have investigated how changes in abiotic factors such as pH, ionic strength, or the presence of organic ligands in the water influence environmental fate. The high ionic strength of seawater compared to freshwater will encourage ENM aggregation. Experimental evidence from colloid chemistry in saline conditions suggests that even small increases in salinity above that of freshwater (about 2.5%) could dramatically decrease colloid concentrations by aggregation and precipitation processes [113]. The behavior of ENM in even very dilute seawater is, therefore, likely to be different from behavior in freshwater. Given this uncertain environmental fate profile, the risks from ENM are unclear.

8.4.3 Environmental Levels Studies based on real measurements of NMs in the environment are currently almost nonexistent, despite the growing use of NMs in commercial products. Therefore, simulations that predict how NMs may be released into the environment, and in what quantities, can be helpful in estimating risks and providing guidance for legislation on the use and disposal of NMs. Gottschalk et al. (2009) [178] calculated probable environmental concentrations (PECs) based on a probabilistic material flow analysis from a life-cycle perspective of ENM containing products. Nano-TiO2, nano-ZnO, nano-Ag, CNTs, and fullerenes were modeled for the United States, Europe, and Switzerland. The environmental concentrations were calculated as probabilistic density functions and were compared to data from ecotoxicological studies. The simulated modes (most frequent values) range from 0.003 ng/L (fullerenes) to 21 ng/L (nano-TiO2) for surface waters and from 4 ng/L (fullerenes) to 4 μg/L (nano-TiO2) for sewage treatment effluents (Figure 8.5). For Europe and the United States, the annual increase of ENMs on sludge-treated soil ranges from 1 ng/kg for fullerenes to 89 μg/kg for nano-TiO2. The results of this study indicate that risks to aquatic organisms may currently emanate from nano-Ag, nano-TiO2, and nano-ZnO in sewage treatment effluents for all considered regions and for nano-Ag in surface waters. For the other environmental compartments for which ecotoxicological data were available, no risks to organisms are presently expected. The results agree well with the limited data that are available from studies providing actual measurements of environmental concentrations. However, researchers stress the need to eliminate uncertainties in the available environmental models by improving current knowledge about the volumes of NMs

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FIGURE 8.5 Predicted NM concentrations (US) in sediment for nano-TiO2 (mg/kg) and for nano-ZnO, nano-Ag, CNTs, and fullerene (μg/kg) for the period 2001 2012. Reprinted with permission from Gottschalk et al. 2009. Modeled Environmental Concentrations of Engineered Nanomaterials (TiO2, ZnO, Ag, CNT, Fullerenes) for different regions. Environmental Science & Technology, 43 (24), pp 9216 9222. Copyright American Chemical Society.

used in different products and the specific types of NMs studied. For instance, nanotubes and fullerenes vary in their properties according to their specific forms. Regional databases containing region-specific product lifecycle information could be helpful to future forensics work.

8.4.4 Analysis Methods As yet, no peer-reviewed literature is available on concentrations (or speciation) of ENM in natural waters (marine or freshwater) or sediments. Further, considerable analytical problems require resolution before these measurements can be performed reliably for routine monitoring, regulatory purposes, or research [114]. Speciation (the physicochemical form or distribution of forms) analysis may require refinement of current research methodologies used for concentration analysis, which include inductively coupled plasma/ mass spectrometry (ICP MS) or atomic absorption spectrometry. Hassellov et al. [115] describes the possibilities for identification and characterization of ENM using analytical tools such as dynamic light scattering (DLS), transmission electron microscopy (TEM), and scanning electron microscopy (SEM). These techniques have successfully been applied to the identification of fullerenes and CNTs in sediments [97,116]. Although quantification cannot be achieved with these methods alone, these methods could

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help understand surface characteristics and reactivity that, in turn, could contribute to the development of selective and sensitive analytical methods useful to investigate biological and aquatic matrices. Difficulties measuring ENMs in the aquatic environment are related to their ability to form colloidal phases and aggregates, their different adsorption and absorption characteristics, and the variety of shapes and sizes that make quantification challenging [117]. Measuring trace levels against a high background of natural colloids further complicates reliable detection in the aquatic environment. For the specific (and likely least amenable to analysis) case of iron oxide ENMs, there is a large background of naturally occurring iron either in the dissolved phase (generally at low pH and reducing conditions) or in the solid phase [118,119]. Into this complex milieu, iron oxide or zero-valent iron ENMs may be discharged, and clearly distinguishing between the natural and manufactured materials may be extremely complex. For aquatic systems, isotopic labeling may be essential to perform any meaningful analyses. For other and, arguably, more important ENMs, such as n-Ag or cerium oxide, understanding background conditions will be less important because of their absence in natural systems.

8.4.5 Summary At present, few analytical tools are reliable for detecting nanoscale particles. Since the ultimate sink for ENMs may be sediment, standardized testing protocols for ENMs are needed. Similar to other emerging POPs, the analytical challenges can be overcome by accurately understanding the potential for environmental releases and characterizing environmental behavior, fate, and bioavailability.

8.5 FLAME RETARDANT COMPOUNDS Flame retardants are chemicals used in thermoplastics, thermosets, textiles, and coatings to inhibit flammability or resist the spread of fire. There are more than 175 different types of flame retardants, which are generally divided into four classes that include the halogenated organic (usually brominated or chlorinated), phosphorus-containing, nitrogen-containing, and inorganic flame retardants. BFRs are the most significant class of substances and include five major types—brominated bisphenols, diphenyl ethers, cyclododecanes, phenols, and phthalic acid derivatives. The five major BFRs are tetrabromobisphenol A (TBBPA), hexabromocyclododecane (HBCD), and three commercial mixtures of polybrominated diphenyl ethers (PBDEs), which are known as decabromodiphenyl ether (deca-BDE), octabromodiphenyl ether (octa-BDE), and pentabromodiphenyl ether (penta-BDE). Until recently, TBBPA, PBDEs, and HBCD were the most commonly used BFRs [120]. Penta-BDE and octa-BDE are being

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phased out or banned in North America and Europe. Deca-BDE is used restrictively, and production, importation, and sales will be discontinued in the United States for all uses by the end of 2013. New BFRs are increasingly used as replacements and some of these new BFRs are already found in the environment, including the Arctic, confirming their potential for long-range atmospheric transport [121].

8.5.1 Sources to the Environment HBCD, TBBPA, and PBDEs are used as additive or reactive ingredients in polymers and other materials used in a wide variety of consumer- and commercial manufactured products that must meet fire safety standards such as carpets, computers, clothing and bedding, electrical equipment, furniture upholstery, insulating foams, televisions, and other household appliances. Among the BFRs, the family of 209 different PBDE congeners poses a significant challenge to environmental forensics work because of their widespread use worldwide [122,179].

8.5.2 Environmental Fate Industrial facilities that produce BFRs, as well as manufacturing facilities that incorporate BFRs into consumer products, release these chemicals during polymer formulation, processing, or manufacturing practices. Disintegration of foam products, volatilization (especially under conditions of high temperature), and leaching from products during laundering or use results in the release of BFRs from products in homes and businesses. Disposal of products, including combustion and recycling of waste products, as well as leaching from landfills, is the final route of entry for BFRs into the environment [123,124]. BFR levels in the environment are highest near industrial sources such as facilities involved in the production of flame retardants, manufacturing facilities that incorporate BFRs into their products, and in electronics recycling facilities [123,124]. In general, BFRs are highly lipophilic (fat soluble) rather than water soluble. BFRs also have a high affinity for binding to particles, which is reflected in low measurements in water and higher measurements in sediment, sewage sludge, and particulate samples like dust particles [125]. Transportation as particle-bound contaminants on airborne dust may explain the wide distribution of BFRs to remote areas. Scientists have found PBDEs and HBCD in air samples collected from remote areas like the Arctic and in marine mammals from the deep seas, which indicate long-range transport of BFRs [126,127]. Aquatic sediment provides a sink for PBDEs due to the fact that the water solubility and vapor pressure of these chemicals are very low and they adsorb quickly onto solid particles like sediment [128]. Half-lives of PBDEs in sediment are short compared to those of other POPs such as dioxin-like

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compounds [129,130]. Ahn et al. [129] showed that in sediment, photodegradation is the main degradation process, but it occurs only on the surface and in environments where the light reaches the sediment. In most cases, once adsorbed onto sediment, PBDEs are only slowly degraded and can accumulate over time. Several studies have found that deca-BDE breaks down to lower brominated congeners (nona- to hexa-BDEs) in sand, sediment, and soils in laboratory conditions of both artificial and natural sunlight [125,131]. The breakdown of deca-BDE occurs much more quickly in UV light (half-life,30 min) compared to natural sunlight where the estimated half-life may be as high as 53 h on sediment and 150 200 h on soil [131]. Depending on the season, photolytic degradation of TBBPA has a half-life of 7 81 days in water. HBCD has a half-life of 3 days in air and 2 25 days in water [125].

8.5.3 Environmental Levels In 2009, US National Oceanic and Atmospheric Administration (NOAA) reported PBDEs throughout the US coastal zone. New York’s Hudson Raritan Estuary had the highest overall concentrations of PBDEs, both in sediments and shellfish. Other coastal locations with the highest PBDE measurements included Anaheim Bay, CA, the Southern California Bight, Puget Sound, the central and eastern Gulf of Mexico near the Tampa St. Petersburg, Florida coast, and Lake Michigan waters near Chicago and Gary, IN [132]. Møskeland [133] found concentrations of deca-BDE and HBCD are increasing in biota and environmental samples collected and measured over several years, whereas levels of penta-BDE and octa-BDE are decreasing and TBBPA detected rarely [133]. Law et al. [134] reported similar trends in the European environment, including a survey of sediment and fish from UK rivers and estuaries. Atmospheric transportation is a major pathway for PBDEs into the marine environment. PBDEs have been found to concentrate in the Arctic and bioaccumulate in wildlife and humans. High levels have been reported in fish, crabs, Arctic-ringed seals, and other marine mammals in northern Canada [135], and also in fish and mussels from Greenland [136]. Ikonmou et al. [135] reported PBDE concentrations (Σ13 congeners) between 350 and 2300 μg/kg lipid weight in porpoise blubber from Northern Canada and between 22 and 340 μg/kg lipid weight in sole [135]. In Greenland, Christensen et al. [136] reported the highest PBDE concentrations (Σ4 congeners) in mussels were 0.11 μg/kg wet weight and in marine fish liver as high as 12.0 μg/kg wet weight. In the Danube River delta in Romania, Webster et al. [137] reported PBDEs in zooplankton (1 to 7.2 μg/kg dry weight) and in several species of fish (,0.1 to 14.3 μg/kg lipid weight), with penta-BDE being the main congener. PBDEs in sediments from the Niagara River in western New York are indicative of general trends observed elsewhere. Samara et al. [138] reported

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PBDE concentrations (Σ9 congeners) ranged from 0.72 to 148 μg/kg dry weight with the highest concentrations found close to wastewater treatment plants. Penta-BDE is typically the dominant PBDE congener.

8.5.4 Analysis Methods PBDEs are analyzed in environmental and biological samples by methods similar to those used for PCBs [139 141]. Covaci et al. [142] reviewed the determination of BFRs with emphasis on PDBEs in environmental and human samples. PBDE samples could be collected by grab sampling and are typically separated from the biological and environmental media by extraction with organic solvents. Liquid solid extraction (e.g., Soxhlet apparatus) remains a widely used technique for solid samples. Typical solvents are hexane, toluene, hexane/acetone mixtures, or dichloromethane. New extraction techniques such as accelerated solvent extraction (ASE) or microwave-assisted extraction (MAE) are currently used by a number of laboratories. Supercritical fluid extraction (SFE) with solid-phase trapping and extraction with pressurized hot water (PHWE) has been used for the extraction of BFRs from sediment. Liquid-liquid extraction (LLE) has been applied for river and seawater samples using hexane/acetone mixtures. SPE has been used for the analysis of acidic and neutral BFRs from human plasma [142]. Gas chromatography mass spectrometry (GC MS) with capillary columns (i.e., congener specific) is the primary analytical technique now used for PBDEs [143].

8.6 POLYCYCLIC AROMATIC HYDROCARBONS Considerable research on the characterization of PAHs for source identification has been conducted and several excellent reviews are available [144,145]. PAHs belong to the group of organic compounds consisting of 2 to13 aromatic rings. PAHs are weakly volatile, dissolve in water with solubility decreasing with an increase in the number of aromatic rings, and are chemically inactive but bond to particulate matter. When adsorbed at the surface of dust, PAHs are highly thermo- and photosensitive. Photooxidation is one of the most important ways of removing PAHs from the atmosphere[146]. A potential source of PAHs to the environment that is receiving increased attention are PAHs generated from oil sands. It has been suggested that oil sands development sites such as those located in the Alberta, Canada region generate unique PAH signatures that can be useful to determine the likely origin of the PAHs. Oil sands as a source of PAHs are discussed further below.

8.6.1 Sources to the Environment PAHs are ubiquitous in the environment, generated during incomplete combustion of materials containing carbon and hydrogen, which includes coal

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fuel, crude oil, wood, gas, and organic materials, as well as combustion of polypropylene and polystyrene, communal and industrial waste, and used tires [147]. Releases to air include those from natural sources [148], such as volcanoes, forest fires, or industrial sources such as from stack emissions and combustion. Releases to water, soil, and sediment include industrial and wastewater treatment plant discharges, precipitation of industrial and natural dust particles, leaks from containers and pipelines, urban runoff [149], and from seepage through and erosion of PAH-containing materials [150]. Synthetic sources provide a much greater release volume than natural sources; the largest single source is the burning of wood in homes (contributing approximately 52% of the total US atmospheric PAH load [151]). Automobile and truck emissions are also major sources of PAHs. Environmental tobacco smoke, unvented radiant and convective kerosene space heaters, and gas cooking and heating appliances may be significant sources of PAHs in indoor air. Hazardous waste sites can be concentrated sources of PAHs on a local scale. Examples of such sites are abandoned wood treatment plants (sources of creosote e.g., Brenner et. al. 2002 [152]) and former manufactured gas sites (sources of coal tar). Figure 8.6 shows the Residential Wood Stoves and Fireplaces Creosote Railway Ties Creosote Utility Poles Vehicle Exhaust Nonroad Internal Combustion Power Generation Gasoline Distribution Tire Fires Commercial Heating Fuel Oil Leaks Open Burning of Household Wastes Residential Heating Fuel Cement Production Coal Tar Sealants Tire Wear Industrial Fuel Combustion Creosote Marine Pilings Incineration Locomotives Refineries Personal Watercraft Vessels Used Motor Oil Disposal Steel Production Cigarette Smoke Port-related Activity Airplanes Oil Spills Pulp and Paper Production

35.7% 30.5% 12.8% 9.6% 3.4% 1.3% 1.3% 0.9% 0.9% 0.5% 0.5% 0.5% 0.4% 0.4% 0.3% 0.3% 0.2% 0.1% 0.1% 0.1% 0.1% 0.1% 0.04% 0.02% 0.02% 0.01% 0.01% 0.01% 0.003% 0%

5%

10%

15%

20%

25%

30%

35%

40%

Percent of Total Release

FIGURE 8.6 Total PAH Release Within the NY/NJ Watershed. Data used for the generation of the figure was obtained from Valle et al. (2007) [180].

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contribution of different sources of PAH within the Hudson River watershed showing a consistent trend with U.S. PAH loads with residential wood stoves and fireplaces contributing the greatest share of PAHs into the environment. Different types of combustion yield different types of PAHs. Those produced from coal burning are different from those produced by motor-fuel combustion, which differ from those produced by forest fires. Some PAHs occur within crude oil, arising from chemical conversion of natural product molecules. They can be summarized into two distinct groups: G

G

Petrogenic hydrocarbon compounds associated with petroleum (petro 5 petroleum) Pyrogenic hydrocarbon compounds associated with the combustion of petroleum, wood, coal etc. including creosote, coal tar (pyro 5 fire/burn)

Total US-reported releases and disposals of PAHs were nearly 2 million pounds in 2005; total releases are likely to be greater than this estimate because not all sources of PAH releases are required to report [153]. US EPA’s 2002 National Emissions Inventory (NEI) lists aggregated emissions for 15 individual PAH compounds (15-PAH). While the dominant PAH emission source reported by US EPA in the 2002 NEI was residential wood heating and open burning due to forest and wildfires, comparisons with the National Toxics Inventory Third Report to Congress for 1990 1993 show consumer products usage as the major source for 16-PAHs. However, this category is almost completely attributed to naphthalene, which is no longer included in the aggregated list of 15-PAH for the NEI [151]. This reflects changes due to regulations, changes in industry, and changes in knowledge and information on the part of the organizations submitting data that are included in these national emission inventories. The Alberta oil sands serve as an interesting case study for the evaluation of PAHs in a regional environment. Recent investigations into the effects of the oil sands on the environment have shown that concentrations of PAHs are increasing in water bodies in the vicinity of bitumen (oil sands) production. There was little monitoring of the air and water in the Alberta region before bitumen production started and there is a polarized debate about what is considered “natural” occurrence of petroleum deposits in lakes and rivers versus PAHs contributed by bitumen production specifically. In a 2010 study, University of Alberta scientists discovered deformed fish in Lake Athabasca downstream from oil sand deposits [154]. The study caused a public outcry and eventually led to a federal provincial environmental monitoring plan for the Alberta oil sands region. The study indicated that there is “little doubt of the unprecedented increases of PAHs” in northeastern Alberta’s lakes, and warns of “striking contaminant increases consistent with the prevailing winds blowing across local upgrading facilities and surfacemining areas.” The study warns of the unknown long-term ecological effects of PAHs in freshwater lakes [154].

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In January 2013, scientists from Queen’s University published a report analyzing lake sediments in the Athabasca region over the past 50 years. They found that levels of PAHs had increased as much as 23-fold since bitumen extraction began in the 1960s [155]. Levels of carcinogenic, mutagenic, and teratogenic PAHs (particularly C1-C4-alkylated PAHs) were substantially higher than guidelines for lake sedimentation set by the Canadian Council of Ministers of the Environment in 1999. Kurek et al. [155] reported a temporal shift in PAH ratios indicative of a shift from primarily wood combustion to petrogenic sources that coincide with greater oil sands development. PAH (congener) ratios proved to be a powerful forensic tool in understanding these sources of PAHs.

8.6.2 Environmental Fate Global transport of PAHs generally involves releases to the atmosphere and removal by wet and dry deposition onto soil, water, and vegetation. In surface water, PAHs volatilize, photolyze, oxidize, biodegrade, bind to suspended particles or sediments, or accumulate in (lowest trophic level) aquatic organisms. In the atmosphere heavier PAHs (containing more than four rings) are adsorbed on dust particles, whereas the lighter ones, which are not adsorbed, remain in the gas phase [156]. In the particle phase, PAHs can be carried by wind and may remain in the atmosphere until they are removed with precipitation. PAHs are accumulated in soil and can also be absorbed by plants. PAHs can penetrate into water with precipitation or with refuse water. Due to their weak solubility, PAH concentrations in water are low (approximately 100 ng/L), instead they accumulate in sediments and aquatic organisms [157]. Approximately 89% of PAHs are accumulated in soil, 10% in sediments, and 0.5% in air and water [158]. Soil contamination by PAHs occurs from airborne dust, sludge that is used in agriculture as a fertilizer, compost and other organic fertilizers, refuse water and water that flows from asphalt roads, fuel and grease used in agriculture, and accidental contamination by oil derivatives [159]. PAHs in soil can also enter groundwater and be transported within an aquifer. Key sources of PAHs in surface waters include deposition of airborne PAHs directly onto water bodies, municipal wastewater discharge, urban storm water runoff, runoff from coal storage areas, effluents from wood treatment plants and other industries, oil spills, and petroleum pressing [160]. Studies have identified industrial effluents, road runoff, and oil spills as the major contributors in specific bodies of water [161,162]. Because of their low solubility and high affinity for organic carbon, PAHs in aquatic systems are primarily found sorbed to particles that either have settled to the bottom or are suspended in the water column. PAHs partition preferentially to sediments because of their hydrophobicity (log Kow: 3.37-7) and are deposited for a long time with half-lives of 0.2 5 years in

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sediment and soil [163,164]. It has been estimated that two-thirds of PAHs in aquatic systems are associated with particles and only about one-third are present in dissolved form [165]. In an estuary, volatilization and adsorption to suspended sediments with subsequent deposition are the primary removal processes for medium and high molecular weight PAHs, whereas volatilization and biodegradation are the major removal processes for low molecular weight compounds [166]. In an enclosed marine ecosystem study, 1% of the original amount of radiolabeled benz[a]anthracene added to the system remained in the water column after 30 days; losses were attributed to adsorption to settling particles and to a lesser extent to photodegradation [167]. Higher levels of total high molecular weight (HMW) PAHs are generally correlated with proximity to urban areas, indicating that levels are likely due to human activity. Whereas control of point sources can lead to localized decreased PAH concentrations in surface sediment over time [152], Longterm trends of PAH concentrations in lake sediments confirm recent increases related to urban pollution. For example, 38 lakes were studied representing a diverse group of geographic regions and ecoregions, and categorized by land use: densely urban, light urban, and reference (1.5% urban land use). Between 1970 and 2001, concentrations of total PAH in sediment increased at 42%, decreased at 5%, and showed no trend at 53% of the lakes. None of the reference lakes showed a trend in total PAH concentrations in this period. To evaluate the potential impact to aquatic biota, the researchers compared the mean concentrations in the sediment in the decades from 1965 to 1975 and the 1990s to a consensus-derived probable effect concentration (PEC). For PAHs, the frequency of exceedances of the PEC approximately doubled in the 1990s compared to the decade from 1965 to 1975 with the highest frequency of exceedances occurring in densely urban lakes. An analysis of trends for individual PAH compounds in these 38 lakes found that most of the lakes had increasing trends of compounds with higher molecular weight than those with lower molecular weight. The higher molecular weight compounds are more typical of combustion by-products. In an earlier analysis for 10 urban lakes (a subset of the 38), the researchers found that increases in PAH concentration followed closely with increases in automobile use, even in urban areas where there was a relatively minor increase in the degree of urbanization over the same time period. The authors noted that there are several sources of vehicle-related PAHs in addition to exhaust, including asphalt wear, tire wear, and leaks and spills of engine oil [168].

8.6.3 Environmental Levels PAHs are ubiquitous environmental contaminants. Although they can be formed naturally (e.g. forest fires), their predominant source is anthropogenic emissions, and the highest concentrations of PAH are generally found around urban centers. Concentrations of PAHs in the aquatic environment are

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generally highest in sediment, intermediate in biota and lowest in the water column [169]. Significant concentrations of PAH can be found in some major estuaries. However, PAH concentrations at offshore sites were generally low or undetectable. For sediments, while PAH concentrations are generally low or undetectable at most intermediate and offshore sites, further work should be concentrated on fine sediments and depositional areas. The main sources of PAHs in water bodies are atmospheric particulate matter deposition, runoff of polluted ground sources and pollution of river and lakes by industrial effluents, municipal wastewater discharge, and oil spills. Since PAHs have low solubility and tend to adsorb to particulate matter, they are usually found in low concentrations in water bodies. Some PAH concentrations that have been measured in water include: marine waters with levels of non-detected to 11 μg/L, wastewater in North American and European municipalities with levels of ,1 to 625 μg/L and urban runoff in the U.S. with levels of ,0.05 to 560 μg/L [170]. Concentrations of PAHs in sediment can range from μg/kg to g/kg levels depending on the proximity of the area to PAH sources such as industries, municipalities, and on water currents. In North America, total PAH concentrations in marine sediments usually range from 2.17 170,000 ng/g sediment [170]. Sediment core studies have shown an increase in PAH concentrations in the past 100 150 years with concentrations peaking in 1950 [171]. PAH profiles in sediments are usually dominated by the more hydrophobic 4-, 5-, and 6-ring compounds. In a study on PAH concentrations in urban samples, Fatoki et al. 2009 [172] reported that environment river water samples had PAH concentrations ranging between 0.1 53.5 μg/L and between 22.8 9,870 ug/kg in river sediments. PAH levels in runoff sediments ranged from 72.5 34,000 μg/kg [172]. PAHs behave differently than most POPs in that they do not biomagnify in aquatic food chains, In fact, top predators which often include various fish, birds, and marine mammals, generally contain lower tissue residues of PAHs than do animals such as snails and bivalve mollusks that occupy lower trophic levels. PAH levels in fish are usually low because this group rapidly metabolizes PAHs [173]; furthermore, higher molecular weight PAHs, which include the largest class of chemical carcinogens, do not seem to accumulate in fish [174]. Raw fish from unpolluted waters usually do not contain detectable amounts of PAHs, but smoked or cooked fish contain varying levels. The concentration of benzo(a)pyrene in skin of cooked fish was much higher than in other tissues, suggesting that skin may serve as a barrier to the migration of PAHs in body tissues [175]. The NOAA National Status and Trends (NS&T) Mussel Watch Program has monitored concentrations of trace chemicals in shellfish in the coastal United States since 1986. The sites were selected to be representative of large areas rather than smaller scale areas that would be influenced directly

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by particular local sources of contaminants. Since mollusks concentrate chemicals from surrounding waters in their tissues, they provide an integrated measurement of contamination over time. For PAH compounds, the Mussel Watch Program groups low molecular weight (LMW) PAH (two- and three ring compounds) separately from HMW PAH (four- and five-ring compounds), because the LMW compounds are relatively more concentrated in oil than in combustion products [176]. Data from the Mussel Watch Program from 1986 to 1996 showed no trends for PAHs. By contrast, a new national scale analysis of Mussel Watch Program data showed that the median concentration in mollusks decreased between 1986 and 2002 for both total LMW and total HMW PAHs [177].

8.6.4 Analysis Methods PAHs are commonly analyzed by gas chromatography coupled with either mass spectrometry (GC-MS) or flame ionization detector (GC-FID), or by liquid chromatography coupled with mass spectrometry (LC-MS) or UVdiode array detection (LC-UV-DAD). GC-FID is generally considered a simpler and more direct method for determination of PAHs compared to other methods and use of GC-FID results in high reproducibility, sensitivity, and resolution [172]. GC-FID is valuable for fingerprinting PAH sources as this method can G G G

identify the source of contamination track the timeframe for contamination differentiate between pollutants from oil and petrol and biogenic sources such as plant material

This method provides detailed information that is not available from other methods, including aging, weathering, alkane distribution, biomarkers, PAH distribution and unknown chemical contaminants. GC-FID analyses can be used to clarify the responsibility for the contamination or to determine the most appropriate remediation techniques.

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