• Petroleum;
  • Combustion;
  • PAH toxicity;
  • Bioavailability


  1. Top of page
  2. Abstract
  8. Acknowledgements

Polycyclic aromatic hydrocarbons (PAHs) are nearly ubiquitous contaminants of freshwater and marine sediments. Sediment PAHs are derived from combustion of organic matter, fossil fuels, and biosynthesis by microbes. Pyrogenic PAHs, particularly those associated with combustion particles (soot), have a low accessibility and bioavailability in sediments. Polycyclic aromatic hydrocarbons associated with petroleum, creosote, or coal tar in sediments may have a moderate accessibility/bioavailability, particularly if the PAHs are part of a nonaqueous phase liquid (NAPL) phase that is in contact with sediment pore water. We present a method for estimating the hazard of complex PAH assemblage in sediments to benthic organisms. Concentrations of all PAHs in sediment pore water are estimated by an equilibrium partitioning model relative to concentrations in bulk sediment. Predicted log Koc values can be used for predicting sediment/water partitioning of petrogenic PAH, but empirically derived log Kd values are needed to predict partitioning of pyrogenic PAH. A hazard quotient (HQ) for each PAH is calculated as the ratio of the estimated concentration in pore water to the chronic toxicity of the PAH determined by a log Kow/toxicity model. Hazard quotients for all PAH in a sample are summed to produce a hazard index (HI), which is a measure of the worst-case estimated hazard of the sediment PAH to benthic organisms. The results of this study show that the integration of HI results with PAH source data provides insights into the causes of sediment toxicity that are useful in an ecological risk assessment.


  1. Top of page
  2. Abstract
  8. Acknowledgements

Polycyclic aromatic hydrocarbons (PAHs) are nearly ubiquitous trace contaminants of freshwater and marine sediments worldwide. They are being recognized with increasing frequency as major contributors to the hazard to aquatic life of contaminated sediments, particularly near areas of intense human activity (Neff 1979, 2002). Polycyclic aromatic hydrocarbons are composed of two or more fused benzene (aromatic) rings (Neff 1979). Aromatic rings are fused when they share two carbon atoms.

Polycyclic aromatic hydrocarbons almost never occur alone in sediments. They usually are present as complex mixtures of hundreds or even thousands of related compounds spanning a wide range of physical/chemical properties and toxicity to aquatic organisms. The composition of PAH assemblages in sediments varies widely depending on the sources of the PAH and the extent of natural degradative processes (called weathering) they have undergone since their release into the environment. A risk assessment for PAH-contaminated sediments requires an estimate of the toxicity to aquatic organisms of the complex PAH assemblage in the sediments. Sediment bioassays alone are often inadequate for identifying the chemicals that pose an ecological risk in sediments. This paper presents an approach to estimating the toxicity of PAH associated with marine and freshwater sediments based on principals of equilibrium partitioning theory (Di Toro and McGrath 2000; Rogers 2002; Hansen et al. 2003). The toxicity of a PAH assemblage in sediments depends on its composition and physical form, both of which depend on the sources of the PAH and on the relative concentrations of different PAH in the sediments. Therefore, an overview of the sources and compositions of the PAH assemblages found in sediments is included.


  1. Top of page
  2. Abstract
  8. Acknowledgements

Polycyclic aromatic hydrocarbons in the environment may be derived from three sources: fossil fuels (petrogenic PAH), burning of organic matter (pyrogenic PAH), and transformation of natural organic precursors in the environment by relatively rapid chemical/biological (diagenic) processes (biogenic PAH) (Neff 1979, 2002). The biogenic PAH assemblages produced naturally (e.g., perylene, retene) are simple and usually do not contribute much to the total mass of PAH in sediments that have received inputs from anthropogenic sources.

Petrogenic PAH

A typical crude oil may contain from 0.2 to more than 7% total PAH. Most of the PAHs in petroleum are low molecular weight hydrocarbons containing two or three fused aromatic rings. Higher molecular weight PAHs, when present, usually are at low concentrations (usually less than 100 mg/kg ppm) (Kerr et al. 1999). Refined petroleum products contain the same PAHs as in the parent crude oil, as well as small amounts of PAHs produced by catalytic cracking and other refining processes (Neff et al. 1994; Stout et al. 2002a). The PAH assemblage in different refined oils varies depending on the distillation temperature range of the product. Gasoline contains mainly low molecular weight aliphatic, olefinic, and monocyclic aromatic hydrocarbons (e.g., benzene and toluene) and 2-ring PAHs (naphthalene and alkylnaphthalenes). Diesel fuels, home heating oils, and engine oils (crankcase oil; middle distillate fuels) may contain aromatic hydrocarbons from benzene through fluoranthene (four aromatic rings).

thumbnail image

Figure Figure 1.. Typical alkyl PAH profiles for naphthalene (N0-N4) and chrysene (C0–C4) in petrogenic (petroleum) and pyrogenic (coal tar) PAH assemblages. From Neff (2002).

Download figure to PowerPoint

Many of the PAHs in petrogenic PAH assemblages contain one or more methyl, ethyl, butyl, or occasionally higher alkyl substituents on one or more of the aromatic carbons (Figure 1). As a general rule, these alkyl PAH are more abundant than the parent compounds in petroleum (Sporstøl et al. 1983; Stout et al. 2002a). Homologues with two to four alkyl carbons usually are more abundant than the less or more highly alkylated homologues (Figure 1).

Pyrogenic PAH

The major source of PAHs containing three or more aromatic rings in the environment is the combustion of organic matter (Neff 1979). During combustion, organic matter is heated to high temperatures, causing it to break up into smaller organic molecules and ultimately into carbon dioxide and water. If combustion is incomplete or the combusted fuel products cool quickly, the small organic chemicals may condense to form new chemicals, including PAH. Polycyclic aromatic hydrocarbons formed during combustion are called pyrogenic PAH, and they often are abundant in the vapor and particulate phases of engine exhaust. Two- and 3-ring PAH are most abundant in the vapor phase; 4- through 6-ringed PAHs often are more abundant in the particulate phase (soot) of engine exhaust or smoke (Neff 2002).

Pyrogenic PAH assemblages are complex, and, unlike the assemblages in petroleum, are dominated by 4-, 5-, and 6-ring PAHs. In pyrogenic PAH assemblages, the dominant compound in each homologous series is the unalkylated parent compound or a homologue with only one or two alkyl substituents (Sporstøl et al. 1983; Stout et al. 2001a). There is an inverse relationship between the temperature of formation and the abundance of alkyl carbons in a pyrogenic PAH assemblage (Neff 1979). The PAH assemblage in coal tar (a product produced by the high-temperature baking of hard coal in a reducing atmosphere to produce coke and manufactured gas) is typical of a high-temperature pyrogenic PAH assemblage (Figure 1). Subsequent distillation of coal (or other) tars alters the composition of the PAH assemblage according to boiling point, sometimes producing pyrogenic mixtures enriched in 2- and 3-ring PAH (e.g., creosote).


  1. Top of page
  2. Abstract
  8. Acknowledgements


Most sediments contain a mixture of PAHs from several petrogenic and pyrogenic sources. It often is desirable or necessary to identify the sources of the PAH in order to identify potentially responsible parties (PRPs) for discharge/spill litigation, to design optimal strategies for source control and remediation, and to aid in characterizing the environmental hazard associated with the sediment contamination. Pyrogenic PAHs in sediments, particularly when they are associated with combustion soot, often are more persistent, less mobile and bioavailable, and less toxic (on a bulk sediment concentration basis) than petrogenic PAHs (Farrington and Westall 1986; Pastorok et al. 1994; Gustafsson et al. 1997). Thus, source identification should be part of any contaminated sediment site assessment.

Historically, the sediment PAHs of primary environmental concern have been the 16 listed on the U.S. Environmental Protection Agency (U.S. EPA) priority pollutant list and similar lists in other countries. However, it has become increasingly evident in recent years that hundreds of PAHs between naphthalene (molecular weight 128.2) and coronene (molecular weight 300.4) are present in environmental matrices that have become contaminated with petrogenic or pyrogenic PAH at concentrations high enough to be of environmental concern (Neff 2002; Barron and Holder 2003). Coronene (log Kow, 6.75, aqueous solubility, 0.14 μg/L) is the highest molecular weight PAH with sufficient environmental mobility to be of potential environmental concern. None of the U.S. EPA priority pollutant PAHs are alkylated; all are unalkylated parent PAHs. However, as discussed above, the most abundant PAH in petrogenic PAH assemblages are alkyl PAHs. Thus, analysis of just 12 to 20 PAHs in sediment samples, as is done for many contaminated site assessments, may be inadequate for providing data needed to identify PAH sources and ecological hazard in the sediments, particularly if they are contaminated primarily with petroleum hydrocarbons.

For source identification, extracts of the nonpolar organic fraction in sediments should be analyzed by a capillary column gas chromatographic method and analyte peaks in the chromatogram identified and quantified by mass spectrometry operated in the selected ion monitoring mode. This analytical method represents a modification of U.S. EPA Method 8270 (Stout et al. 2002a, 2002b). In the modified method, the gas chromatography is operated with a slow oven temperature increase program to optimize separation of target compounds. The mass spectrometer is operated in the selected ion monitoring mode to minimize interferences from nontarget compounds and, when necessary, to improve detection limits for analytes present at low concentrations. The mass spectrometer/selected ion monitoring should target ions of the parent and alkyl PAH analytes of interest. Usually, between 40 and 50 analytes or analyte groups are analyzed, including parent compounds from naphthalene to benzo[ghi]perylene and C1- through C4-alkyl congener groups for naphthalene, fluorene, phenanthrene/anthracene, fluoranthrene/pryrene, and chrysene. Heterocyclic compounds such as dibenzothiophene (a sulfur-containing heterocyclic compound) and their alkyl homologue groups also may be included on the analyte list, particularly when sediments are suspected to contain a substantial contribution of petrogenic PAHs. It is extremely important that the analyses adhere to strict data quality objectives to assure optimal precision, accuracy, specificity, and sensitivity (Boehm et al. 1997). Poor data quality will hamper the ability to accurately identify PAH sources in sediments. Detailed descriptions of these methods can be found elsewhere (Page et al. 1995; Douglas et al. 1996; Stout et al. 2002a, 2002b).

Table Table 1.. Ratios of the PAH isomers phenanthrene to anthracene (PH/AN) and fluoranthene to pyrene (FL/PY) in PAH assemblages from several sources
  1. aAnthracene or fluoranthene concentration was below the detection limit.

Primarily pyrogenic sources
Coke oven emissions1.27–3.570.76–1.31Maher and Aislabe 1992
Iron/steel plant (soot)0.240.62Yang et al. 2002
Iron/steel plant (flue gas)0.061.43Yang et al. 2002
Wood-burning emissions6.411.26Page et al. 1999
Auto exhaust soot (gasoline)1.790.90O'Malley et al. 1996
Diesel engine soot0.061.26Bence et al. 1996
Diesel exhaust particles (n = 22)1.3–780.25–1.38Sjøgren et al. 1996
Highway dust4.71.4Christensen et al. 1999
Urban runoff0.56–1.470.23–1.07Stout et al. 2001a
Creosote0.11–4.011.52–1.70Neff 2002
Coal tar3.111.29Neff 2002
Coke0.241.49S.A. Stout (unpublished data)
Creosote-contaminated sediment in Table 50.341.59Stout et al. 2001a
Urban sediment in Table 50.220.79Stout et al. 2001a
Primarily petrogenic sources
60 crude oils (mean)52.00.25Kerr et al. 1999
Australian crude oil370a0.78Neff et al. 2000
Italian crude oil>232a0.08Neff et al. 1998
Alaska crude oil>262a0.2Bence et al. 1996
Diesel fuel (No. 2 fuel oil)>800a0.38Bence et al. 1996
No. 4 fuel oil11.80.16S.A. Stout (unpublished data)
Bunker C residual fuel oil14.80.14S.A. Stout (unpublished data)
Road paving asphalt20<0.11aKriech et al. 2002
West Virginia coal (2 samples)11.2, 27.90.95, 1.03Neff and Sauer 1993

The differences in ratios of parent to alkyl-substituted PAH congeners can be used to distinguish between petrogenic and various types of pyrogenic PAH assemblages in environmental samples (Bence et al. 1996; Douglas et al. 1996; Zeng and Vista 1997; Stout et al. 2000). If only priority pollutant PAH data are available, ratios of phenanthrene to anthracene (PH/AN) and fluoranthene to pyrene (FL/PY) are useful for differentiating between sediment PAH assemblages containing primarily pyrogenic or petrogenic PAHs (Table 1). Anthracene and fluoranthene are thermodynamically less stable than their isomers, phenanthrene and pyrene, respectively (Baumard et al. 1998). Anthracene and fluoranthene are produced during rapid, high-temperature pyrosynthesis, but are less favored to persist during the slow organic digenesis leading to the generation of fossil fuels. Thus, the PH/AN ratio of pyrogenic PAH assemblages usually is less than 5 and the petrogenic ratio usually is greater than 5 (Table 1). The FL/PY ratio usually approaches or exceeds a value of 1 in pyrogenic assemblages and usually is substantially less than a value of 1 in petrogenic PAH assemblages. Because of the extreme variability in these ratios in PAH assemblages from different sources, and in the absence of additional alkyl PAH and other chemical “fingerprint” data, at a minimum, both ratios should be used to aid in differentiating between petrogenic and pyrogenic PAH in sediments. A plot of PH/AN (y axis) against FL/PY (x axis) for PAH assemblages from different single and mixed sources produces a distribution with a negative slope. Samples containing primarily petrogenic PAHs are clustered in the upper left side of the graph; data points distribute toward the lower right as the fraction of total PAH that is pyrogenic increases.

Several other diagnostic ratios can be used to help distinguish between petrogenic and pyrogenic PAH assemblages in sediments. Because alkyl PAHs are more abundant than the unalkylated parent PAHs in petrogenic PAH assemblages and are less abundant in pyrogenic PAH assemblages, ratios of selected primarily petrogenic alkyl PAHs to selected primarily pyrogenic parent PAHs provide a good indication of PAH source (Sporstøl et al. 1983). The ratios of methylphen-anthrenes to phenanthrene, total methyl-fluoranthenes/pyrenes to fluoranthene/pyrene, and total methylchrysenes to chrysene are the ratios most frequently used (Gustafsson et al. 1997; Zeng and Vista 1997; Pereira et al. 1999; Stout et al. 2001a, 2001b). Although they may be affected by preferential weathering of the parent PAH, these ratios usually are greater than a value of 1 in petrogenic PAH assemblages and less than a value of 1 in pyrogenic assemblages.

Table Table 2.. Fluoranthene plus pyrene to sum of C2- to C4-phenanthrenes (FLPY)/(FLPY + C2- to C4-phenanthrenes [C24PH]) ratios in different types of petrogenic and pyrogenic PAH assemblages. Most data are from the Battelle PAH Forensics database
MaterialNo. of samplesMean ratioRatio range
Petrogenic sources
Crude oils220.0150–0.044
#2 Fuel oil/diesel250.0440.008–0.073
#6 Fuel oil/bunker C430.0500.028–0.143
IBF-380 heavy fuel170.0480.018–0.057
Pyrogenic sources
Coal tar150.9220.838–0.983

The ratio of fluoranthene plus pyrene (FLPY) to the sum of C2- to C4-phenanthrenes (C24PH), expressed as FLPY/(FLPY + C24PH) such that values range from 0 to 1, has been effective in differentiating between petrogenic and pyrogenic PAH assemblages in sediments and biological samples from Prince William Sound, (AK, USA) which was the site of the 1989 Exxon Valdez oil spill (Neff et al. 2004). Petrogenic PAH assemblages nearly always have a FLPY/(FLPY + C24PH) ratio less than 0.1, whereas the ratio in pyrogenic PAH assemblages usually is more than 0.75 (Table 2). Gasoline may have a higher ratio, because it contains only traces of fluoranthene, pyrene, and more highly alkylated phenanthrenes.

More elaborate fingerprinting methods are required for distinguishing among multiple petrogenic sources in sediments (Boehm et al. 1997; Stout et al. 2002b; Neff et al. 2004). Within each alkyl homologue group, alkyl phenanthrenes, dibenzothiophene, and chrysene tend to weather at the same rates and are fairly persistent in contaminated sediments (Douglas et al. 1996; Boehm et al. 1997). The concentrations and relative abundances of different alkyl-PAHs vary widely in crude and refined petroleum from different sources. Thus, alkyl-PAH ratios are useful for identifying PAH assemblages from different petrogenic sources in sediments (Boehm et al. 1997; Burns et al. 1997; Stout et al. 2002a). For example, ratios of total C2-dibenzothiophenes to total C2-phenanthrenes (DT2/PH2) and of the trimethyl homologues (DT3/PH3) were particularly useful for distinguishing among sediment PAH from North Slope crude oil (the oil released in the Exxon Valdez oil spill) and from other petrogenic sources (seep oil, weathered petroleum tar, diesel fuel) in spill path areas of Prince William Sound (Page et al. 1996; Boehm et al. 1997). If the sediment PAH data are graphed in double-ratio plots (e.g., DT2/PH2 vs DT3/PH3), the PAH assemblages from different petrogenic sources cluster separately, often allowing clear differentiation among multiple sources (Brown and Boehm 1993; Boehm et al. 1997).

Higher resolution in the source allocation can be obtained by a comprehensive statistical analysis of the complete PAH profile and diagnostic ratios. One such statistical method is principal component analysis (PCA). The principal component analysis is one of several types of ordination techniques, also known as factor analyses, by which multivariate data sets are explored, reduced, interpreted, and/or studied further (Wold et al. 1987). PCA is used in many types of studies and has been applied to PAH fingerprinting and allocation studies (Boehm et al. 1997; Burns et al. 1997; Naes and Oug 1998; Stout et al. 2001a).

Principal component analysis is an exploratory statistical technique that produces a visual comparison among sediment samples and suspected source materials (e.g., petroleum products from suspected sources, pyrogenic emissions from local point and nonpoint sources, coal tar, and creosote). Figure 2 shows a PCA plot for sediments from an urban waterway in which three sources of PAHs were recognized, namely natural background (arising from preindustrial, natural forest fires); urban runoff; and creosote (from a former tar distillation facility on the waterway) (Stout et al. 2001a, 2002b). Many sediment samples from this urban waterway contained PAHs primarily from one of these three end-members. These “single-source” samples tend to plot as clusters at or near the apices of the trends revealed by the PCA factor score plot. Several other sediments tended to plot in locations intermediate between the three end-members, indicating that they contain a mixture of PAHs from two or three of the sources. Spatial relationships among samples on a PCA score plot can be used to estimate or determine the proportions of each end-member in each sediment sample. Additional calculations involving spatial distributions, concentrations, and volumes of PAH-contaminated sediments from each contaminant source in the study can be used to allocate contributions among the three end-member sources.


Petrogenic PAHs enter freshwater and marine environments from natural oil seeps; erosion of coal, peat, and oil shale deposits; oil and coal spills; discharges of treated and untreated ballast and bilge water from ships; and effluents from oil refineries, oil/water separators on oil production platforms, coal-fired power plants, storm water runoff, and municipal sewage treatment plants (NAS 2002; Neff 2002).

thumbnail image

Figure Figure 2.. Principal component analysis (PCA) factor score plot for sediment PAH data from a contaminated estuary. The PCA identified three dominant PAH sources (natural background, urban runoff, and creosote). Samples falling between the apices contain a mixture of PAH from these sources. From Stout et al. (2001a, 2002b).

Download figure to PowerPoint

A great many domestic and industrial activities, as well as natural events such as forest fires, produce PAH by pyrolysis/pyrosynthesis. Pyrosynthesized PAH may be released to the environment in airborne particles or in the solid or aqueous byproducts of the pyrolysis process. Burning of fossil fuels is an important source of pyrogenic PAHs in the environment. The particulate fractions of exhaust from gasoline and diesel-powered vehicles contain 16 to 2,300 μg/g total 4- through 6-ringed PAHs (Takada et al. 1991; Oda et al. 1998). Nearly all the PAHs derived from vehicular exhaust are deposited within about 50 m of roads (Harrison and Johnston 1985; Hewitt and Rashed 1990). Much of the deposited PAHs, however, find their way to water bodies in surface runoff from land (Hoffman et al. 1984; Sharma et al. 1994).

Several industrial processes, such as coal coking (Lao et al. 1975); carbonization of coal and oil to produce manufactured gas, coal tars, carbon black, and pitch (Villaume 1984; Merrill and Wade 1985; Mueller et al. 1989); catalytic cracking of petroleum feed stocks to produce refined petroleum products (Stout et al. 2001b); manufacture of iron and steel (Yang et al. 2002); and aluminum smelting (Thrane 1987; Näf et al. 1994) produce airborne particulates and solid wastes containing high concentrations of PAH. Coking of coal produces an estimated emission of about 40 mg of benzo[a]pyrene (a carcinogenic PAH) per ton of coke produced (Eisenhut et al. 1990). Polycyclic aromatic hydrocarbon-contaminated wastes and discharges from these industries may reach freshwater and marine environments in wastewater effluents and in deposition of airborne vapor-phase or particulate PAH. For example, an estimated 10 t of PAHs were discharged to the sea from seven Norwegian aluminum smelters in 1992, down from 42 t in 1988 (Knutzen 1995).


  1. Top of page
  2. Abstract
  8. Acknowledgements

Concentration of PAH in sediment pore water

Polycyclic aromatic hydrocarbons in solution in ambient water or pore water of sediments are much more bioavailable and toxic than those adsorbed to particles (particularly combustion soot) (Gustafsson et al. 1997) or associated with a nonaqueous phase liquid (NAPL; e.g., petroleum, creosote, or coal tar) (Pastorok et al. 1994). The dissolved phase of PAHs in sediments can be estimated based on equilibrium partitioning theory (Hansen et al. 2003).

Nonpolar organic chemicals, such as PAHs, have low aqueous solubilities and high affinities for adsorption to sediment and organic particles and absorption (bioconcentration) by living organisms (Neff 2002). Most of the higher molecular weight pyrogenic PAHs entering aquatic environments are sorbed to soot (the particulate fraction of smoke and engine exhaust). Petrogenic lower molecular weight PAHs may enter the water from the vapor phase in rainfall or dry fallout. They quickly adsorb to the organic phase of suspended particles and are deposited with them in sediments (Neff 2002).

Petrogenic PAHs from petroleum and pyrogenic PAHs from creosote and coal tar in sediments may be complexed with the colloidal and particulate organic fraction of sediment or associated with a NAPL, an oil phase, or an oil coating on sediment particles. Because the affinity of hydrocarbons is higher for the oil phase than for the sediment organic matter and sediment porewater phases, partitioning of hydrocarbons into sediment porewater is controlled primarily by the affinity of the hydrocarbons for the NAPL phase (Zemanek et al. 1997). Thus, in estimating the partitioning of PAHs between the NAPL phase (petroleum, coal tar, creosote) and dissolved phase, PAH concentration should be normalized to some measure of total hydrocarbons.

The PAHs in sediments are distributed between the dissolved (porewater) and particulate and NAPL phases of the sediment according to their relative affinities for the three phases. This distribution can be expressed as an organic carbon/water partition coefficient (Koc) or an oil/water partition coefficient (Koil) (Lee et al. 1992a; Neff and Sauer 1995; Di Toro and McGrath 2000; Hansen et al. 2003). Both partition coefficients are similar to the octanol/water partition coefficient (Kow) that is used frequently to model bioconcentration of nonpolar organic compounds from water by aquatic animals (Connell 1993; Neff 2002). The Koc for most nonpolar organic chemicals and colloidal/particulate organic matter in sediments is lower than the Kow (Karickhoff 1981; Di Toro et al. 1991; Neff 2002), whereasthe ow Koil for PAHs in most refined petroleum products and liquid coal tars is about the same as or higher than the Kow and tends to increase with average molecular weight of the NAPL material (Shiu et al. 1990; Lee et al. 1992a, 1992b).

However, high molecular weight, petrogenic PAHs in coal particles and asphalt, pyrogenic PAHs in soot or coal tar, and related viscous liquids often are bound to sediment particles more strongly than predicted by equilibrium partitioning theory. Mitra et al. (1999) reported high, invariant log Kocs for PAHs in sediments from the Elizabeth River (VA, USA), which is heavily contaminated with creosote-contaminated wood particles. Polycyclic aromatic hydrocarbons in sediments of urban estuaries—such as the Tamar River (UK) (Readman et al. 1987); Boston Harbor (MA, USA) (McGroddy and Farrington 1995); and San Francisco Bay (CA, USA) (Maruya et al. 1996)—often are more tightly bound to sediment particles (have higher log Kocs) than predicted. The desorption rate of PAHs from sediments decreases with duration of sediment contamination (Kraaij et al. 2002). Polycyclic aromatic hydrocarbons also are tightly bound to coal (Ghosh et al. 2001). These tightly bound PAHs do not partition effectively into the aqueous phase of porewater.

Table Table 3.. Log Kow' freshwater solubility, and estimated acute and chronic toxicity of PAH frequently found in crude and refined petroleum. Solubility and toxicity values are micrograms per liter (μg/L ppb). Log Kow values and solubilities are from Mackay et al. (1992), Neff and Burns (1996), and Ran et al. (2002)
PAHLog KowFreshwater solubilityAcute toxicityChronic toxicity
  1. aNV = No solubility value could be found.

Benzo[ghi] perylene6.

The use of Kow or Koc tends to overestimate concentrations of dissolved PAHs in porewater of sediments contaminated primarily with pyrogenic PAHs but should give a reasonable upper-limit estimate of dissolved-phase PAHs in porewater of petroleum-, creosote-, or coal tar-contaminated sediments if the oil or other NAPL phase is still liquid and in physical contact with sediment porewater. The NAPL, particularly if it is crude oil or coal tar, may develop a surface “skin” of resins-asphalthenes or other high molecular weight polar compounds, decreasing NAPL/water partitioning (Ghoshal et al. 2004). A NAPL or oil-filled pores also may substantially decrease the permeability of the soil or sediment, decreasing the effective NAPL/water interface and limiting accessibility of the PAH to partitioning into sediment porewater. Empirically determined Kd (particle/water partition coefficients) are best for estimating sediment/water partitioning of pyrogenic PAHs or PAHs from weathered crude oil.

Values for the octanol/water partition coefficient (Kow) have been published for a large number of PAHs (Mackay et al. 1992; Durell et al. 2004). The most accurate current values for log Kow for several PAHs of environmental concern are summarized in Table 3. Koc can be estimated from Kow (Karickhoff et al. 1979), but Kd must be determined empirically on a site-specific basis. The concentration of a PAH in sediment porewater in equilibrium with its concentration in the bulk sediment can be estimated by the simple equation

  • equation image((1))

where Cw is the concentration of the PAH in solution in sediment porewater, Cs is the concentration of the PAH in bulk sediment (measured as concentration per unit mass of sediment organic carbon or concentration per unit mass of total petroleum hydrocarbons [TPH]), and Kx is the sediment organic matter/water partition coefficient (Koc) or sediment particle/water partition coefficient (Kp) for the PAH. The K for pyrogenic PAHs associated primarily with combustion soot requires a variation on Equation 1 to account for the high affinity of soot particles for PAHs (Bucheli and Gustafsson 2000; Cornelissen and Gustafsson 2004).

Concentrations of individual PAHs in bulk sediment, expressed as μg/g dry sediment, should be normalized to the concentration of total extractable (C8+) petroleum hydrocarbons (TPH) in sediments, determined by gas chromatography/flame ionization detection (Sauer and Boehm 1995), if the source of the PAHs in sediments is primarily petroleum. If the PAHs in the sediment are primarily pyrogenic, then concentrations of PAHs should be normalized to sediment total organic carbon if the sediments contain high concentrations (several percent) of particulate organic matter or if a pyrogenic NAPL (e.g., creosote or coal tar) is present. TPH or total extractable organic matter often is the best parameter for normalizing PAH concentrations in urban or industrial sediments, even when the PAHs are primarily from pyrogenic sources because the total extractables analysis quantifies mainly the nonpolar organic fraction in bulk sediment that often is the most important in adsorbing PAHs. This calculation is repeated for all PAHs analyzed in sediment and is the basis for an estimate of the maximum concentration of total PAHs in solution in sediment porewater. Where the estimated concentration of a PAH exceeds its aqueous solubility, the aqueous solubility is used as the water concentration.

Toxicity of dissolved PAH mixtures to aquatic organisms

A search of the U.S. EPA Toxicity Information Retrieval (USEPA 1997) database identified more than 300 values for the acute toxicity (median lethal concentration, LC50) of aromatic hydrocarbons to freshwater and marine invertebrates and fish. The search excluded LC50 concentrations greater than the aqueous solubility of the particular hydrocarbon. Suitable aquatic toxicity data were found for 25 aromatic hydrocarbons, including 14 PAHs (Table 4). Log geometric mean acute toxicity values (in mM/L) for the aromatic hydrocarbons were regressed against log Kow. The regression has a high correlation (r2 = 0.885) and the form

  • equation image((2))

This equation was used to estimate the acute toxicity of each of the PAHs analyzed in sediment (Table 3). Equation 2 is similar to that developed by McCarty et al. (1992) to estimate the toxicity to freshwater fish of a large number of nonpolar organic compounds. Equation 2 considers toxicity data for freshwater and marine invertebrates and fish and applies only to aromatic hydrocarbons.

The chronic toxicity of each PAH was estimated by dividing the acute value by an acute/chronic ratio of 5. An acute/chronic ratio of 5 represents a conservative estimate of the acute/chronic ratio for aromatic hydrocarbons. For example, Suter and Rosen (1988) evaluated the comparative acute and chronic toxicity of several chemicals to marine fish and crustaceans. Acute/chronic ratios for aromatic hydrocarbons calculated from their data are between 2 and 4.

The estimated concentration of each PAH in solution in sediment porewater was divided by its chronic toxicity value to derive a HQ. Hazard quotients for all of the PAHs detected in sediment were summed to produce a HI for total PAHs:

  • equation image((3))
  • equation image((4))

Equations 3 and 4 are based on the reasonable assumptions that the dissolved PAHs are much more bioavailable and toxic than adsorbed PAHs (Neff 2002) and the toxicities of individual PAHs in a mixture in solution are additive (Warne et al. 1989; Di Toro and McGrath 2000; Hansen et al. 2003; Landrum et al. 2003).

Toxicity of PAH assemblages in sediments

Log Kow values for the PAHs most frequently analyzed in freshwater and marine sediments increase with molecular weight from 3.37 for naphthalene to 8.0 for C4-chrysenes (Table 3). Log Koc values estimated by the log Koc/log Kow regression of Karickhoff (1981) are slightly lower than the log Kow values in Table 3, ranging from 2.94 for naphthalene to 6.68 for benzo[ghi]peylene. Values for log Koil, based on the regression of Lee et al. (1992a) for diesel fuel PAHs, are slightly higher than the log Kow values, ranging from 3.81 for naphthalene to 6.72 for benzo[ghi]perylene. The actual value of log Koil varies with the “average molecular weight” of the oil and, therefore, is different for different crude and refined petroleum products and changes with oil weathering (Lee et al. 1992a, 1992b; Shiu et al. 1990). The value of log Koil for a particular PAH decreases as the average molecular weight and density of the bulk oil increases, in agreement with Raoult's Law (Lane and Loehr 1995). Thus, Kow is a reasonable, conservative coefficient to use for estimating the dissolved concentrations of PAHs associated with sediments, most of which contain both weathered petrogenic and pyrogenic PAHs. The use of log Kow may result in approximately 2-fold under- or overestimation of the true concentration of a PAH in solution in equilibrium with the NAPL phase (Shiu et al. 1988).

Table Table 4.. Geometric mean toxicity (LC50 with 48-h or longer exposure) for PAH, based on available data from the AQUIRE database (USEPA 1997)
ChemicalGeomean LC50 (mg/L)ChemicalGeomean LC50 (mg/L)

Polycyclic aromatic hydrocarbon solubility in freshwater decreases with increasing PAH molecular weight (Table 3). The solubility of some alkyl-PAHs is greater than that of the parent PAH, possibly reflecting steric effects of the alkyl carbons. Solubility tends to decrease with increasing seawater salinity and decreasing water temperature (Neff 2002). For example, the solubility of phenanthrene in fresh and salt water at 25°C is 1,080 and 644 μg/L, respectively (Eastcott et al. 1988). The solubility of anthracene in freshwater decreases from 56.5 μg/L at 29.1°C to 15.5 μg/L at 8.9°C (Reza et al. 2002). At all salinities and temperatures, anthracene is much less soluble than its isomer, phenanthrene. These physical properties of PAH affect their bioavailability and toxicity to freshwater and marine organisms.

The measured acute toxicity of aromatic hydrocarbons increases (LC50 decreases) with increasing PAH molecular weight and log Kow (Table 4; Hansen et al. 2003). The estimated acute toxicity values follow the same trend (Table 3). Estimated chronic values for PAHs range from 970 μg/L (ppb) for naphthalene to 0.01 ppb for C4-chrysenes (Table 3). For anthracene and PAHs with molecular weights of 228.3 (chrysene and benz[a] anthracene) or higher, the acutely lethal concentration approaches, or is higher than, the single phase aqueous solubility. Saturated solutions of these highly nonpolar PAHs are not acutely toxic to aquatic organisms.

We selected sediment PAH data from a recent study of the sources of PAHs in sediments of the Wycoff/Eagle Harbor Superfund site in the state of Washington, USA, to demonstrate the method described above for estimating the aquatic toxicity, measured as HI, of sediment-bound PAHs (Table 5). Two sediment samples from Eagle Harbor in Puget Sound were used, one heavily contaminated with creosote and the other contaminated with PAHs from urban runoff and deposition of pyrogenic PAHs from combustion sources. The PAH data for the two sediment samples used in this example are plotted in the PCA plot for the site data (Figure 2) and show that one sample has a clear creosote signature and the other has a clear urban runoff signature.

The creosote-contaminated sediment contained 27,441 μg/g (ppm) of TPH and 17,283 ppm of total PAH. The urban runoff sediment contained 212 ppm of TPH and 25 ppm of total PAH. Total PAH concentrations were much higher than the “high” value reported in the National Status and Trends database of 2.18 ppm of total PAH (24 parent PAH and alkyl homologue groups in sediments) (Daskalakis and O'Connor 1995), indicating that sediments from Eagle Harbor were highly contaminated with PAHs. Daskalakis and O'Connor (1995) identified Eagle Harbor as the location of some of the most heavily contaminated sediments in Puget Sound.

The sediment PAH concentrations were normalized to TPH concentration for calculation of HIs. TPH-normalized PAH concentrations were 629,808 μg PAH/g TPH and 115,655 μg/g in the creosote-contaminated and urban runoff-contaminated sediments, respectively. Estimated concentrations of total PAH in solution in sediment porewater in equilibrium with the two sediments were 17,190 μg/L (ppb) and 10,216 μg/L, respectively (Table 5). However, estimated concentrations of several PAHs in solution in water in equilibrium with the creosote-contaminated sediment were in excess of their single-phase aqueous solubilities (Table 3). Actual concentrations of these PAHs in solution in sediment pore water would not exceed their aqueous solubilities. Therefore, the aqueous solubilities of these PAHs were used as the exposure concentrations.

The estimated concentration of each PAH in solution was divided by its chronic toxicity value (Table 3) to obtain an HQ. The HQs for the PAHs in the creosote-contaminated sediment pore water ranged from 0.1 to 32.1 (C4-phenanthrenes). The sum of HQs (the HI) for this sediment was 250.

The estimated concentration of only benzo[a] pyrene in solution exceeded its water solubility for the urban runoff sediment sample. The difference was small, so no adjustment was necessary. Estimated HQs for PAH in the urban runoff-contaminated sediment porewater ranged from 0.1 to 6.9 (naphthalene), and the estimated HI was 64.

An HI value greater than 1 indicates that the porewater contains in solution a concentration of total PAHs in excess of its estimated chronic toxicity to aquatic animals (Ozretich et al. 2000). Both sediments had HIs substantially greater than 1, suggesting that both sediments would be toxic to the benthic fauna of Eagle Harbor.

Table Table 5.. Hazard quotients (HQ) and hazard indices (HI) for two sediment samples collected from Eagle Harbor, Washington, USA. The weathered creosote-contaminated sediment contained 17,283 μg/g dry wt (ppm) total polycyclic aromatic hydrocarbons (PAH) and 27,441 ppm total petroleum hydrocarbons (TPH). The sediment sample contaminated with urban runoff/fallout PAH contained 25 ppm total PAH and 212 ppm TPH. Concentrations in parentheses are freshwater solubilities. PAH data from Stout et al. (2001a)
 Creosote-contaminated sediment (27,441 μg TPH/g sediment)Sediment contaminated with urban runoff PAH (212 μg TPH/g sediment)
PAHSediment (μg/g TPH)Water (μg/L)HQSediment (μg/g TPH)Water (μg/L)HQ
Anthracene11,280325 (80)(1.3)1,480430.7
Dibenzothiophene86,4582,800 (1,140)(16.3)10,4423404.8
Fluoranthene101,012609 (260)(23.6)6,168373.4
Pyrene63,521420 (130)(10.8)7,808524.3
Benz[a]anthracene14,93418 (15)(7.5)1,99621.3
Chrysene15,69322 (6)(2.7)4,02262.5
Benzo[a]pyrene4,2104 (1)(0.7)1,8672(1)(0.7)
Indeno[1,2,3-cd]pyrene8570.10.7928 0.7
Total PAH629,80817,190 115,65510,216 
HI  250  64

Sediment quality guidelines, based on toxicity to sediment-dwelling marine animals, have been developed for total PAH in marine sediments (Long et al. 1995). The effects range low (ERL) and effects range median (ERM) concentrations for total PAH are 4.022 μg/g dry wt and 44.792 μg/g, respectively. The ERL is the concentration in bulk sediment below which toxicity to benthic organisms is unlikely; the ERM is the concentration above which effects are likely. Concentrations between the ERL and ERM may be toxic and may require additional evaluation. At Eagle Harbor, the creosote-contaminated sediment contained more than 17,000 μg/g of total PAHs, more than 384 times higher than the ERM concentration. The high HI value and substantial exceedence of the ERM value for this sediment indicates that it is likely to be highly toxic to benthic fauna.

Approximately 34% of the HI for the creosote-contaminated sediment at Eagle Harbor was attributable to 4-ring+ (fluoranthene and higher) PAHs, most of which are pyrogenic (Table 5). Other predominantly pyrogenic PAHs also contributed to the HI, including acenaphthene, dibenzofuran, and anthracene. Dibenzothiophenes, which usually are considered primarily petrogenic, make a substantial contribution to the HI. Some of the creosote from the Wycoff wood treatment facility may have been distilled from a high-sulfur petroleum tar, which would contain high concentrations of dibenzothiophenes. These creosote-associated PAHs, particularly the higher molecular weight ones, have much higher log Kocs than predicted (Mitra et al. 1999), indicating a low accessibility and bioavailability. Thus, it is likely that the sediments are much less toxic to benthic animals than predicted by the high HI value and exceedence of the ERM value.

The urban runoff sediment contained 25 ppm of total PAHs, about 55% of the ERM concentration and about 6 times the ERL concentration. The HI of the PAH assemblage in this sediment was 64, indicating a hazard (risk of toxicity) to benthic organisms if the PAH are accessible and bioavailable. This sediment would be toxic to benthic animals if the PAH associated with the sediment particles are accessible. As indicated in Table 5, this sediment sample was enriched in parent PAHs and several 4-ring+ PAHs characteristic of pyrogenic sources. There also is evidence of some petrogenic PAH contributions, particularly the dibenzothiophenes (DBT) that are much more abundant in petrogenic than pyrogenic PAH assemblages (Neff 2002). More than 40% of the HI for this sediment was attributable to 4-ring and higher PAHs (mostly pyrogenic) that tend to sorb to sediment particles much more strongly than predicted (Neff 2002). However, there may be enough alkyl naphthalenes, phenanthrenes, and dibenzothiophenes (mostly petrogenic) in the sediments to elicit effects in some sensitive benthic organisms.

The toxicity of two heavily contaminated Eagle Harbor sediments was evaluated with a sensitive sand dollar embryo test (Meador et al. 1990). The sediment samples contained 33.6 and 37.0 μg/g total PAH. There was 100% mortality of the echinoderm embryos during exposure to both sediments. Ozretich et al. (2000) evaluated the toxicity of 30 creosote-contaminated sediments from nearby Elliott Bay (WA, USA) with two amphipod species. Mean amphipod mortality was less than 10% (sediments were not toxic) in seven sediments containing 12 to 140 μg/g total PAH (34 parent and alkyl PAH groups). There was 100% mortality in eight sediments containing 500 to 25,000 μg/g PAH. Mean amphipod mortality ranged from 13 to 78% in the remaining sediments, containing 19 to 480 μg/g PAH. Thus, the creosote-contaminated sediment used in the present investigation, containing 17,000 μg/g total PAH, probably also would be toxic. The urban runoff sediment containing 25 μg/g PAH probably were either nontoxic or moderately toxic to benthic animals.


  1. Top of page
  2. Abstract
  8. Acknowledgements

The analysis of PAHs in sediments from Eagle Harbor shows how the source and composition of the PAH assemblage in sediment can affect its estimated hazard to aquatic animals. The HI approach to estimating the hazard of a complex mixture, like PAH, in sediments provides more information than the ERL/ERM or sediment toxicity approaches and can be used to aid in interpreting the significance of the hazard estimates and the causes of any risk identified. A combination of HI assessment and sediment toxicity probably would provide the largest amount of useful information upon which to base estimates of ecological risk of PAH-contaminated sediments.


  1. Top of page
  2. Abstract
  8. Acknowledgements

The U.S. Navy, Naval Facilities Engineering Service Center, Port Hueneme, California, USA, and the Marine Spill Response Corp., Washington, DC, supported much of the work upon which this paper is based.


  1. Top of page
  2. Abstract
  8. Acknowledgements
  • Barron MG, Holder E. 2003. Are exposure and ecological risks of PAHs underestimated at petroleum contaminated sites? Hum Ecol Risk Assess 9: 15331545.
  • Baumard P, Budzinski H, Garrigues P. 1998. Polycyclic aromatic hydrocarbons in sediments and mussels of the western Mediterranean Sea. Environ Toxicol Chem 17: 765776.
  • Bence AE, Kvenvolden KA, Kennicutt II MC. 1996. Organic geochemistry applied to environmental assessments of Prince William Sound, Alaska, after the Exxon Valdez oil spill. Org Geochem 24: 742.
  • Boehm PD, Douglas GS, Burns WA, Mankiewicz PJ, Page DS, Bence AE. 1997. Application of petroleum hydrocarbon chemical fingerprinting and allocation techniques after the Exxon Valdez oil spill. Mar Pollut Bull 34: 599613.
  • Brown JS, Boehm PD. 1993. The use of double-ratio plots of polynuclear aromatic hydrocarbon (PAH) alkyl homologues for petroleum source identification. Proceedings of the 1993 International Oil Spill Conference; 1993 Mar 29-Apr 1; Tampa, FL. Washington, DC: American Petroleum Institute. p 799801.
  • Bucheli TD, Gustafsson Ö. 2000. Quantification of the soot-water distribution coefficient of PAHs provides mechanistic basis for enhanced sorption observations. Environ Sci Technol 34: 51445151.
  • Burns WA, Mankiewicz PJ, Bence AE, Page DS, Parker KR. 1997. A principal component and least-squares method for allocating polycyclic aromatic hydrocarbons in sediment to multiple sources. Environ Toxicol Chem 16: 11191131.
  • Christensen ER, Rachdawong P, Karls JF, Van Camp RP. 1999. PAHs in sediments: Unmixing and CMB modeling of sources. J Environ Engin 125: 10221032.
  • Connell DW. 1993. The octanol-water partition coefficient. In: CalowP, editor. Handbook of ecotoxicology. Vol 2. London, UK: Blackwell Scientific. p 311320.
  • Cornelissen G, Gustafsson Ö. 2004. Sorption of phenanthreneto environmental black carbon in sediment with and without organic matter and native sorbates. Environ Sci Technol 38: 148155.
  • Daskalakis KD, O'Connor TP. 1995. Distribution of chemical concentrations in US coastal and estuarine sediment. Mar Environ Res 40: 381398.
  • DiToro DM, McGrath JA. 2000. Technical basis for narcotic chemicals and polycyclic aromatic hydrocarbon criteria. II. Mixtures and sediments. Environ Toxicol Chem 19: 19711982.
  • Di Toro DM, Zarba CS, Hansen DJ, Berry WJ, Swartz RC, Cowan CE, Pavlou SP, Allen HE, Thomas NA, Paquin PR. 1991. Technical basis for establishing sediment quality criteria for nonionic organic chemicals using equilibrium partitioning. Environ Toxicol Chem 10: 15411583.
  • Douglas GS, Bence AE, Prince RC, McMillen SJ, Butler EL. 1996. Environmental stability of selected petroleum hydrocarbon source and weathering ratios. Environ Sci Technol 30: 23322339.
  • Durell G, Røe Utvik T, Johnsen S, Frost T, Neff J. 2004. The use of deployed blue mussels (Mytilus edulis) and semi-permeable membranes (SPMDs) for monitoring produced water originating polycyclic aromatic compounds (PAH) in the North Sea: Results from field surveys and comparison with dispersion model data. Mar Environ Res (forthcoming).
  • Eastcott L, Shiu WY, Mackay D. 1988. Environmentally relevant physical-chemical properties of hydrocarbons: A review of data and development of simple correlations. Oil Chem Pollut 4: 191216.
  • Eisenhut W, Friedrich F, Reinke M. 1990. Coking plant environment in West Germany. Coke Making Intern 1: 7477.
  • Farrington JW, Westall J. 1986. Organic chemical pollutants in the oceans and groundwater: A review of fundamental chemical properties and biogeo-chemistry. In: KullenbergG, editor. The role of the oceans as a waste disposal option. New York (NY), USA: Reidel. p 361425.
  • Ghosh U, Talley JW, Luthy RG. 2001. Particle-scale investigation of PAH desorption kinetics and thermodynamics from sediment. Environ Sci Technol 35: 34683475.
  • Ghoshal S, Pasion C, Alshafie M. 2004. Reduction of benzene and naphthalene mass transfer from crude oils by aging-induced interfacial films. Environ Sci Technol 38: 21022110.
  • Gustafsson Ö, Gschwend PM, Buesseler KO. 1997. Using 234Th disequilibria to estimate the vertical removal rates of polycyclic aromatic hydrocarbons from the surface ocean. Mar Chem 57: 1123.
  • Hansen DJ, Di Toro DM, McGrath JA, Swarz RC, Mount DR, Spehar RL, Burgess RM, Ozretich RJ, Bell HE, Reiley MC, Linton TK. 2003. Procedures for the derivation of equilibrium partitioning sediment benchmarks (ESBs) for the protection of benthic organisms: PAH mixtures. Washington, DC: US Environmental Protection Agency, Office of Research and Development. EPA 600/R-02/013.
  • Harrison RM, Johnston WR. 1985. Deposition fluxes of lead, cadmium, copper, and polynuclear aromatic hydrocarbons (PAH) on the verges of a major highway. Sci Tot Environ 46: 121135.
  • Hewitt CM, Rashed MB. 1990. An integrated budget for selected pollutants for a major rural highway. Sci Tot Environ 93: 375384.
  • Hoffman EJ, Mills GL, Latimer JS, Quinn JG. 1984. Urban runoff as a source of polycyclic aromatic hydrocarbons to coastal waters. Environ Sci Technol 18: 580587.
  • Karickhoff SW. 1981. Semi-empirical estimation of sorption of hydrophobic pollutants on natural sediments and soils. Chemosphere 10: 833846.
  • Karickhoff SW, Brown DS, Scott TA. 1979. Sorption of hydrophobic pollutants on natural sediments. Water Res 13: 241248.
  • Kerr JM, Melton HR, McMillen SJ, Magaw RI, Naughton G. 1999. Polyaromatic hydrocarbon content in crude oils around the world. SPE 52724. Paper presented at the 1999 SPE/EPA Exploration and Production Environmental Conference; Austin, TX. Richardson (TX), USA: Society of Petroleum Engineers. 10 p.
  • Knutzen J. 1995. Effects on marine organisms from polycyclic aromatic hydrocarbons (PAH) and other constituents of waste water from aluminum smelters with examples from Norway. Sci Tot Environ 163: 107122.
  • Kraaij R, Seinen W, Tolls J, Cornellissen G, Belfroid AC. 2002. Direct evidence of sequestration in sediments affecting the bioavailability of hydrophobic organic chemicals to benthic deposit-feeders. Environ Sci Technol 36: 35253529.
  • Kriech AJ, Kurek JT, Osborn LV, Wissel HO, Sweeney BJ. 2002. Determination of polycyclic aromatic compounds in asphalt and in corresponding leachate water. Polycyclic Aromatic Compounds 22: 517535.
  • Landrum PF, Lotufa GR, Gossiaux DC, Gedeon ML, Lee J-H. 2003. Bioaccumulation and critical body residue of PAHs in the amphipod, Diporia spp.: Additional evidence to support toxicity additivity for PAH mixtures. Chemosphere 51: 481489.
  • Lane WF, Loehr RC. 1995. Predicting aqueous concentrations of polynuclear aromatic hydrocarbons in complex mixtures. Water Environ Res 67: 169173.
  • Lao RC, Thomas RS, Monkman JL. 1975. Computerized gas chromatographicmass spectrometric analysis of polycyclic aromatic hydrocarbons in environmental samples. J Chromatogr 112: 681700.
  • Lee LS, Hagwall M, Delfino JJ, Rao PSC. 1992a. Partitioning of polycyclic aromatic hydrocarbons from diesel fuel into water. Environ Sci Technol 26: 21042110.
  • Lee LS, Rao PS, Okuda I. 1992b. Equilibrium partitioning of polycyclic aromatic hydrocarbons from coal tar into water. Environ Sci Technol 26: 21102115.
  • Long EL, MacDonald DD, Smith SL, Calder FD. 1995. Incidence of adverse biological effects within ranges of chemical concentrations in marine and estuarine sediments. Environ Manag 19: 8197.
  • Mackay D, Shiu WY, Ma KC. 1992. Illustrated handbook of physical-chemical properties and environmental fate for organic chemicals. Polynuclear aromatic hydrocarbons, polychlorinated dioxins, and dibenzofurans. Chelsea (MI), USA: Lewis. 597 p.
  • Maher WA, Aislabe J. 1992. Polycyclic aromatic hydrocarbons in nearshore sediments of Australia. Sci Tot Environ 112: 143164.
  • Maruya KA, Risebrough RW, Horne AJ. 1996. Partitioning of polynuclear aromatic hydrocarbons between sediments from San Francisco Bay and their porewaters. Environ Sci Technol 30: 29422947.
  • McCarty LS, Mackay D, Smith AD, Ozburn GW, Dixon DG. 1992. Residue-based interpretation of toxicity and bioconcentration QSARsfrom aquatic bioassays: Neutral narcotic organics. Environ Toxicol Chem 11: 917930.
  • McGroddy SE, Farrington FW. 1995. Sediment porewater partitioning of polycyclic aromatic hydrocarbons in three cores from Boston Harbor, Massachusetts. Environ Sci Technol 29: 15421550.
  • Meador JP, Ross BD, Dinnel PA, Picquelle SJ. 1990. An analysis of the relationship between a sand-dollar embryo elutriate assay and sediment contaminants from stations in the urban embayment of Puget Sound, Washington. Mar Environ Res 30: 251272.
  • Merrill EG, Wade TL. 1985. Carbonized coal products as a source of aromatic hydrocarbons to sediments from a highly industrialized estuary. Environ Sci Technol 19: 597603.
  • Mitra S, Dickhut RM, Kuehl SA, Kimbrough KL. 1999. Polycyclic aromatic hydrocarbon (PAH) source, sediment deposition patterns, and particle geochemistry as factors influencing PAH distribution coefficients in sediments of the Elizabeth River, VA, USA. Mar Chem 66: 113227.
  • Mueller JG, Chapman PJ, Pritchard PH. 1989. Creosote-contaminated sites. Environ Sci Technol 23: 11971201.
  • Naes K, Oug E. 1998. The distribution and environmental relationships of polycyclic aromatic hydrocarbons (PAHs) in sediments from Norwegian smelter-affected fjords. Chemosphere 36: 561576.
  • Näf C, Broman D, Axelman J. 1994. Characterization of the PAH load outside an aluminum smelter on the Swedish Baltic coast. Sci Tot Environ 156: 109118.
  • [NAS] National Academy of Sciences. 2002. Oil in the sea. Inputs, fates and effects. Washington, DC: National Academy Press. 265 p.
  • Neff JM. 1979. Polycyclic aromatic hydrocarbons in the aquatic environment. Sources, fates and biological effects. Barking, Essex, UK: Applied Science. 262 p.
  • Neff JM. 2002. Bioaccumulation in marine organisms. Effects of contaminants from oil well produced water. Amsterdam, The Netherlands: Elsevier. 452 p.
  • Neff JM, Bence AE, Parker KR, Page DS, Brown JS, Boehm PD. 2004. Bioavailability of PAH from buried shoreline oil residues 13 years after the Exxon Valdez oil spill: a multispecies assessment. Environ Toxicol Chem (forthcoming).
  • Neff JM, Burns WA. 1996. Estimation of polycyclic aromatic hydrocarbon concentrations in the water column based on tissue residues in mussels and salmon: An equilibrium partitioning approach. Environ Toxicol Chem 15: 22402253.
  • Neff JM, Langseth DE, Graham EM, Sauer Jr TC, Gnewuch SC. 1994. Transport and fate of non-BTEX petroleum chemicals in soil and groundwater. Washington, DC: American Petroleum Institute. API Publ 4593.
  • Neff JM, Ostazeski S, Gardiner W, Stejskal I. 2000. Effects of weathering on the toxicity of three offshore Australian crude oils and a diesel fuel to marine animals. Environ Toxicol Chem 19: 18091821.
  • Neff JM, Sauer TC. 1993. Assessment of long-term environmental effects and recovery from the 1985 crude oil release to Newton Lake, Illinois. Littleton (CO), USA: Report to Marathon Pipeline Co.
  • Neff JM, Sauer TC. 1995. Reduction in the toxicity of crude oil during weathering on the shore. Washington (DC), USA: Marine Spill Response Corp. Technical Report Series 95–015.
  • Neff JM, Seavy J, McCarthy K. 1998. Changes in the toxicity of oil-contaminated agricultural soils following the Trecate 24 well blowout. Milan, Italy: Report to Agip Oil Co.
  • O'Malley VP, Abrajano TA Jr., Hellou J. 1996. Stable carbon isotopic apportionment of individual polycyclic aromatic hydrocarbons in St. John's Harbour, Newfoundland. Environ Sci Technol 30: 634639.
  • Oda J, Maeda I, Mori T, Yasuhara A, Saito Y. 1998. The relative proportions of polycyclic aromatic hydrocarbons and oxygenated derivatives in accumulated organic particulates as affected by air pollution sources. Environ Technol 19: 961976.
  • Ozretich RJ, Ferraro SP, Lambertson JO, Cole FA. 2000. Test of σ polycyclic aromatic hydrocarbon model at a creosote-contaminated site, Elliott Bay, Washington, USA. Environ Toxicol Chem 19: 23782389.
  • Page DS, Boehm PD, Douglas GS, Bence AE. 1995. Identification of hydrocarbon sources in the benthic sediments of Prince William Sound and the Gulf of Alaska following the Exxon Valdez oil spill. In: WellsPG, ButlerJN, HughesJS, editors. Exxon Valdez oil spill: Fates and effects in Alaskan waters. Philadelphia (PA), USA: American Society of Testing and Materials. STP 1219. p 4183.
  • Page DS, Boehm PD, Douglas GS, Bence AE, Burns WA, Manciewicz PJ. 1996. The natural petroleum hydrocarbon background in subtidal sediments of Prince William Sound, Alaska, USA. Environ Toxicol Chem 15: 12661281.
  • Page DS, Boehm PD, Douglas GS, Bence AE, Burns WA, Manciewicz PJ. 1999. Pyrogenic polycyclic aromatic hydrocarbons in sediments record past human activity: A case study in Prince William Sound, Alaska. Mar Pollut Bull 38: 247260.
  • Pastorok RA, Peek DC, Sampson JR, Jacobson MA. 1994. Ecological risk assessment for river sediments contaminated by creosote. Environ Toxicol Chem 13: 19291941.
  • Pereira WE, Hostettler FD, Luoma SN, van Geen A, Fuller CC, Anima RJ. 1999. Sedimentary record of anthropogenic and biogenic polycyclic aromatic hydrocarbons in San Francisco Bay, California. Mar Chem 64: 99113.
  • Ran Y, He Y, Yang G, Johnson JLH, Yalkowsky SH. 2002. Estimation of aqueous solubility of organic compounds by using the general solubility equation. Chemosphere 48: 487509.
  • Readman JW, Mantoura RFC, Rhead MM. 1987. A record of polycyclic aromatic hydrocarbon (PAH) pollution obtained from accreting sediments of the Tamar Estuary, UK: Evidence for non-equilibrium behaviour of PAH. Sci Tot Environ 66: 7394.
  • Reza J, Trejo A, Vera-Ávila LE. 2002. Determination of the temperature dependence of water solubilities of polycyclic aromatic hydrocarbons by a generator column-on-line solid-phase extraction liquid chromatographic method. Chemosphere 47: 933945.
  • Rogers HR. 2002. Assessment of PAH contamination in estuarine sediments using the equilibrium partitioning-toxic unit approach. Sci Tot Environ 290: 139155.
  • Sauer TC, Boehm PD. 1995. Hydrocarbon chemistry analytical methods for oil spill assessments. Washington, DC: Marine Spill Response Corp. Technical Report Series 95–032. 114 p.
  • Sharma M, Marsalek J, McBean E. 1994. Migration pathways and remediation of urban runoff for PAH control. J Environ Manag 41: 325336.
  • Shiu WY, Bobra M, Bobra AM, Jaijanen A, Suntio L, Mackay D. 1990. The water solubility of crude oils and petroleum products. Oil Chem Pollut 7: 5784.
  • Shiu WY, Maijanen A, Ng ALY, Mackay D. 1988. Preparation of aqueous solutions of sparingly soluble organic substances. II. Multicomponent systems—Hydrocarbon mixtures and petroleum products. Environ Toxicol Chem 7: 125137.
  • Sjøgren M, Li H, Rannug U, Westerholm R. 1996. Multivariate analysis of exhaust emissions from heavy-duty diesel fuels. Environ Sci Technol 30: 3849.
  • Sporstøl S, Gjos N, Lichtenthaler RG, Gustavsen KO, Urdal K, Oreld F, Skel J. 1983. Source identification of aromatic hydrocarbons in sediments using GC/MS. Environ Sci Technol 17: 282286.
  • Stout SA, Magar VS, Uhler RM, Ickes J, Abbott J, Brenner R. 2001a. Characterization of naturally occurring and anthropogenic PAHs in urban sediments: Wycoff/Eagle Harbor Superfund site. Environ Forens 2: 287300.
  • Stout SA, Naples WP, Uhler AD, McCarthy KJ, Roberts LG, Uhler RM. 2000. Use of quantitative biomarker analysis in the differentiation and characterization of spilled oil. SPE 61460. In: SPE International Conference on Health, Safety, and the Environment in Oil and Gas Exploration and Production; 2000 June 26–28; Stavanger, N. Richardson (TX), USA: Society of Petroleum Engineers. 15 p.
  • Stout SA, Uhler AD, Boehm PD. 2002a. Recognition and allocation among multiple sources of PAH in urban sediments. Environ Claims J 13: 141158.
  • Stout SA, Uhler AD, McCarthy KJ. 2001b. A strategy and methodology for defensibly correlating spilled oil to source candidates. Environ Forens 2: 8798.
  • Stout SA, Uhler AD, McCarthy KJ, Emsbo-Mattingly S. 2002b. Chemical fingerprinting of hydrocarbons. In: MurphyBL, MorrisonRD, editors. Introduction to environmental forensics. San Diego (CA), USA: Academic. p 137260.
  • Suter GW Jr, Rosen AE. 1988. Comparative toxicology for risk assessment of marine fishes. Environ Sci Technol 22: 132138.
  • Takada H, Onda T, Harada M, Ogura N. 1991. Distribution and sources of polycyclic aromatic hydrocarbons (PAHs) in street dust from the Tokyo metropolitan area. Sci Tot Environ 107: 4569.
  • Thrane KE. 1987. Deposition of polycyclic aromatic hydrocarbons (PAH) in the surroundings of primary aluminum industry. Water Air Soil Pollut 33: 385393.
  • [USEPA] U.S. Environmental Protection Agency. 1997. Aquatic information and retrieval database. Duluth (MN): USEPA Office of Research and Development.
  • Villaume JF. 1984. Coal tar wastes: Their environmental fate and effects. In: MajumdaySK, MillerEW, editors. Hazardous and toxic waste: Technology, management and health effects. Philadelphia (PA), USA: Pennsylvania Academy of Sciences. p 362375.
  • Warne MStJ, Connell DW, Hawker DW, Schüürmann G. 1989. Prediction of the toxicity of mixtures of shale oil components. Ecotoxicol Environ Saf 18: 121128.
  • Wold S, Esbensen SK, Geladi P. 1987. Principal component analysis. Chemom Intell Lab Syst 2: 3752.
  • Yang H-H, Lai S-O, Hsieh L-T, Sueh H-J, Chi T-W. 2002. Profiles of PAH emission from steel and iron industries. Chemosphere 48: 10611074.
  • Zemanek MG, Pollard SJT, Kenefick SL, Hrudey SE. 1997. Multi-phase partitioning and co-solvent effects for polynuclear aromatic hydrocarbons (PAH) in authentic petroleum- and creosote-contaminated soils. Environ Pollut 98: 239252.
  • Zeng EY, Vista CL. 1997. Organic pollutants in the coastal environment off San Diego, California. 1. Source identification and assessment by compositional indices of polycyclic aromatic hydrocarbons. Environ Toxicol Chem 16: 179188.