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Molecular markers and their use in environmental organic geochemistry
Geochemical Investigationsin Earth and Space Science: A Tribute to Isaac R. Kaplan 9 The GeochemicalSociety,PublicationNo. 9, 2004 Editors: R.J. Hill, J. Leventhal,Z. Aizenshtat, M.J. Baedecker, G. Claypool, R. Eganhouse,M. Goldhaberand K. Peters
Molecular markers and their use in environmental organic geochemistry ROBERT P. EGANHOUSE Water Resources Discipline, US Geological Survey, 12201 Sunrise Valley Drive, MS 432, Reston, VA 20192-0002, USA Abstract--Molecular markers are organic substances that carry information about sources of organic matter or contamination. The source/marker relation can be used to indicate the presence of a given source material (qualitative), or, under appropriate conditions, to estimate the amount of a source material (quantitative source apportionment) in the environment. Assemblages of markers can also be used as process probes. In this instance, systematic differences and/or similarities in the physical-chemical properties of markers are coupled with compositional changes in marker composition to infer the operation of natural processes. This paper provides an overview of what molecular markers are, what types of markers are present in the environment, the requirements for the use of markers, and some common applications. To illustrate how molecular markers can answer specific environmental questions, three case studies are presented. The first case study examines the impact of municipal waste on a large urban harbor (Boston Harbor). Linear alkylbenzenes (unreacted residues of linear alkylbenzenesulfonate surfactants) and coprostanol (a fecal indicator) provide information on the sources and likely transport pathways of municipal wastes in a complex hydrologic system. The marker data are also used to estimate the proportion of sewage-derived polychlorinated biphenyls (PCBs) in polluted harbor sediments. The second case study concerns a portion of the continental shelf off southern California (Palos Verdes) where discharge of municipal wastewaters has led to extensive contamination of sediments and biota. Long-chain alkylbenzenes (surfactant residues), PCBs and the pesticide, DDT (dichlorodiphenyltrichloroethane), are used to develop sedimentation rate estimates for several time periods by molecular stratigraphy. These results, when combined with other information, allow conclusions to be drawn about the most likely transport pathway of sediments at the study site and to predict the fate of historically deposited contaminants. Finally, an investigation of a crude-oil spill in Bemidji, MN illustrates how monoaromatic hydrocarbons can be exploited as process probes, providing insights into the relative importance of different attenuation processes in a contaminated aquifer. The results show that natural attenuation of the monoaromatic hydrocarbons is occurring at this site and is dominated, not by physical and/or chemical processes, but by biodegradation.
INTRODUCTION ORGANIC CHEMICALS generated as a result of human activity enter the environment in vast quantities (on the order of hundreds of tons/yr). These inputs span a large range of temporal and spatial scales, and the media to which the chemicals are introduced (air, water, soil) vary. Some chemicals are applied intentionally over large areas (e.g. pesticides) or were used in a wide range of commercial products (e.g. polychlorinated biphenyls (PCBs)). Others are undesirable byproducts of industrial activity (e.g. combustion-derived polycyclic aromatic hydrocarbons (PAHs), polychlorinated dibenzodioxins (PCDDs)) or are man-made chemicals that emanate from a limited number of point sources (surfactants, pharmaceuticals). Many synthetic organic chemicals are individual compounds (e.g. lindane, atrazine), but complex mixtures of congeneric substances are not u n c o m m o n (e.g. PCBs, toxaphene). For these reasons, identification of the sources, transport pathways and fates of anthropogenic organic contaminants represents a significant challenge. Molecular markers, compounds whose structures or isotopic compositions are related to specific sources, offer a means of extracting such information directly from the environment. 143
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The mobility of a chemical depends on its tendency to partition among environmental compartments (atmosphere, hydrosphere, lithosphere, biosphere). Environmental partitioning, in turn, is determined by the physical-chemical properties of a substance and where and how it is released. GOUIN et al. (2000) have shown that the behavior of different chemicals can, to a first approximation, be predicted based on where they plot on a two-dimensional property space diagram (log air-water partition coefficient versus log octanol-water partition coefficient). Compounds, such as the low-molecular-weight n-alkanes (C1-8), catechol, and DDT, have physical-chemical properties that strongly favor partitioning into air, water, and soil/sediment phases, respectively. Other chemical classes or mixtures, such as the chlorobenzenes and PCBs, have a wide range of properties that favor partitioning among multiple phases. These are multimedia contaminants, and, if sufficiently persistent, they can be transported long distances and become globally distributed (WANIAand MACKAY,1996). This paper presents some basic background information about what molecular markers are, requirements for their use, and the types of applications that have been developed over the last 40 yr. Three case studies are offered as a means of illustrating some of the ways in which molecular markers can be used to answer specific environmental questions. These questions probably could not have been readily addressed using conventional environmental chemistry approaches. The objective of this paper is to stimulate interest in the molecular marker approach and to foster its use in environmental organic geochemistry. The interested reader is directed to EGANHOUSE (1997) for a more extensive discussion of molecular markers, their uses and limitations. BACKGROUND
Molecular markers as information carriers Molecular markers are carriers of information. This information is contained in their isotopic composition (613C, A14C, 615N, 6D, 637C1, etc.), chemical structure, or the particular mixture of molecules composing an assemblage. If the isotopic composition, structure, or molecular distributions of an individual or an assemblage of chemicals are unique to a given source, these substances can be used as source indicators. Molecular markers can also be used to investigate the operation of natural processes (i.e. process probes). The most common applications involve assemblages of compounds whose properties differ systematically (i.e. homologsman homologous series comprises compounds that differ by the chain length of a single alkyl substituent) or whose properties are similar or identical (i.e. isomers). Because chemicals partition among or are degraded within environmental media on a structure-specific basis, changes in marker composition can be used to infer the relative importance of different processes (e.g. volatilization, sorption, biodegradation, etc.).
Types of molecular markers As discussed in EGANHOUSE (1997), molecular markers can be classified as belonging to one of the three groups. Contemporary biogenic markers are synthesized by living organisms and can be found in the source organisms themselves (e.g. abietane) or, with little or no alteration, in the contemporary environment (e.g. dehydroabietic acid). Fossil biomarkers, good examples of which are the ubiquitous hopanoids, are biomolecules (e.g. bacteriohopanepolyols) that have been transformed through diagenetic and/or catagenetic processes. Even so, they largely retain the stereospecific carbon skeleton of the original biological precursor. The third group comprises anthropogenic markers. The presence of these compounds in the environment signals the influence of human activity or the input of wastes (TAKADA and EGANHOUSE, 1998). Anthropogenic markers can be broken into subgroups that include: (1) non-toxic synthetic chemicals (e.g. linear alkylbenzenesulfonate surfactants (LAS) used in commercial detergents), (2) non-toxic biogenic chemicals (e.g. coprostanol, urobilin), (3) toxic synthetic chemicals (e.g. PCBs, toxaphene), and (4) by-product chemicals (e.g. PCDDs or PAHs). The latter arise
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incidentally from human activities such as combustion of organic matter, synthesis of chemicals, or other industrial processes.
Requirements There are three principal requirements for the ideal molecular marker: (1) source-specificity, (2) persistence or predictable behavior, and (3) massive production/usage (TAKADA and EGANHOUSE, 1998). Ideally, a molecular marker should derive from a single source. However, this condition is rarely met. The Boston Harbor case study describes the use of waste-specific molecular markers (coprostanol, linear alkylbenzenes (LABs)) that serve as particle tracers in a complex urban harbor. In this instance, the markers arise exclusively from municipal wastes. Thus, the requirement of source-specificity is satisfied. Alternatively, when there is a massive local input of what ordinarily are non-source-specific contaminants, such compounds can be used as molecular markers of that local input. This situation is typified by the Palos Verdes Shelf, CA and the Bemidji, MN case studies where DDT (§ and volatile monoaromatic hydrocarbons (MAHs), respectively, serve as markers of local inputs. A molecular marker should be persistent or its behavior should be predictable. Ideally, one would desire conservative behavior, but given the diversity of catabolic enzymes, this is a rather demanding requirement for organic substances. In practice, information on the persistence and behavior of markers is often limited. Few studies have been undertaken to systematically determine the susceptibility of any of the commonly used markers to photolysis, chemical reaction, or biodegradation under an appropriate range of environmental conditions (EGANHOUSE, 1997). Even fundamental physical-chemical properties (e.g. vapor pressure, aqueous solubility, octanol-water partition coefficient), which are essential for multimedia fate modeling, are rarely known with confidence, and the database, at best, is spotty. This highlights an area where more reliable data are clearly needed. It is important to add that knowledge of the environmental behavior of a marker (e.g. LABs) vis-gt-vis a non-sourcespecific contaminant (e.g. PCBs) is essential for quantitative source apportionment of the latter (HEDGES and PRAHL, 1993; TAKADA et al., 1997). Finally, a molecular marker should be produced on a massive scale and/or its use should be widespread. This requirement effectively defines the range of applicability of a marker insofar as a substance must be readily detectable in a variety of environments for it to yield useful results. Markers, like other organic contaminants, are subject to the same range of spatial and temporal input functions and they are discharged to various environmental media.
Applications Table 1 summarizes various ways in which molecular markers have been applied in environmental studies during the last 40 yr. Source indicators-As source indicators, molecular markers are applied either qualitatively (source identification) or quantitatively (source apportionment). The requirements for quantitative source apportionment are considerably more stringent and the number of examples in the literature is fewer than for source identification (e.g. SCHAUER et al., 1996; EGANHOUSE and SHERBLOM,2001; TAKADA et al., 1997). Specific biogenic markers have been used for chemotaxonomic purposes (KATES, 1997) and to characterize the composition of benthic and pelagic communities present in aquatic ecosystems (FINDLAYand WATLING,1997). Many fossil biomarkers have served as indicators of organic matter provenance (PETERSand MOLDOWAN, 1993) or for paleoclimate reconstruction (BRASSELe t al., 1986). Biomarkers are also used frequently to identify sources of fossil fuel contamination in the contemporary environment (KAPLANet al., 1997; VOLKMAN et al., 1997). Finally, there are numerous studies in which molecular markers associated with municipal wastes, urban runoff, or combustion of fossil fuels have been used to infer the effect of various point and non-point sources of contamination (see references in TAKADA and EGANHOUSE, 1998).
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R . P . Eganhouse Table 1. Applications of molecular markers
Source indicators Source identification (qualitative) Contemporary biogenic markers: chemotaxonomy, biogenic inputs Petroleum source rock identification, archaeological geochemistry and paleoclimate reconstruction, forensic geochemistry Impacts of major pollutant sources (e.g. municipal wastewaters, urban runoff, combustion of fossil fuels) a Source apportionment (quantitative) Microbial community composition Oil spills Emissions to the atmosphere Non-specific contaminants a Process probes Transport Pathways a,b,c Processes Dilution Resuspension Burial b Mixing b Fate Phase transfer processes Volatilization Vaporization Dissolution c Sorption c Degradation processes c Bioaccumulation Superscripted applications indicate how case studies in this paper reflect the use of molecular markers as source indicators and process probes: aBoston Harbor case study; bpalos Verdes Shelf, CA case study; CBemidji, MN oil-spill site case study.
Process probes-There are two general ways in which molecular markers are applied as process probes. Both rely upon the ability to relate a prospective marker to a specific source and, in some cases, a known input function. Markers can yield information on the pathways along which particles move (hydrophobic markers) or the direction of ground- and surface-water flow (hydrophilic markers). In addition, molecular markers provide unique approaches with which to understand processes of dilution, resuspension, burial, and mixing. Case studies presented below describe some aspects of these applications (see footnote to Table 1 for a key to these applications). In terms of the environmental fate of contaminants, three general process types can be distinguished: phase transfer, degradation, and bioaccumulation. All of these process types depend upon knowledge of the physical-chemical properties of the markers, transformation rates and pathways, measurements of the compositions of mixtures of compounds, and an understanding of the physics of the system under investigation. CASE STUDIES
Boston Harbor Boston Harbor is a shallow, glacially carved, tidally dominated estuary. Flushing occurs primarily through two channels, President Roads and Nantasket Roads, which connect the harbor with Massachusetts Bay (Fig. 1). For nearly 125 yr (BOTHNER et al., 1998), the wastes of metropolitan Boston and surrounding communities were discharged directly into the harbor with minimal or no treatment. Treated effluent and sludge were released from two plants (Deer Island and Nut Island) through a system of outfalls. When the capacities of the treatment plants were exceeded during wet weather, untreated sewage + runoff was discharged through a relief system consisting of 108 combined sewer overflows (CSOs). As recently as 1990, almost
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FIG. 1. Map of Boston Harbor showing locations of municipal wastewater and sludge outfalls, major channels to Massachusetts Bay and combined sewer overflows (CSOs). Inset depicts study area adjacent to the Fox Point CSO including stations where surficial sediment samples were collected (modified from EGANHOUSE and SHERBLOM,2001).
nothing was known about the effects these CSOs were having on water quality in the harbor. For this reason, a study was undertaken to determine the sources of organic contaminants to one of the largest CSOs (at Fox Point; see inset, Fig. 1) and to assess the fate of these materials after discharge to Boston Harbor (EGANHOUSE and SHERBLOM, 2001). In this instance, molecular markers not only served as qualitative source indicators, but also revealed information on transport pathways and allowed for quantitative source apportionment of a class of contaminants of concern, the PCBs. Samples of Fox Point CSO effluent and harbor receiving water along with sediments near the CSO outfall (Fig. 1) were collected during varying weather conditions. These samples, as well as sludges obtained from Deer Island and Nut Island treatment plants, were analyzed for two molecular markers of municipal wastes, the LABs (surfactant chemicals) and coprostanol (a fecal indicator). PCBs were also measured. LABs and coprostanol were found in the effluent during all sampling periods, indicating that even when weather conditions were dry, sewage entered the CSO system, presumably by leakage or through illegal connections. Moreover, concentrations of the LABs and PCBs in the CSO effluent were highly c o r r e l a t e d ( r 2 - - 0 . 8 7 ) suggesting that both of these substances were coming from the same source (namely sewage). Compositional features of the LABs, notably the I/E ratio (the ratio of internal to external isomers; see TAKADA and ISHIWATARI, 1989 for definition), which increases with biodegradation, indicated that sewage entering the CSO had not undergone appreciable alteration. Apparently the CSO was carrying untreated sewage even during dry weather. The chemical composition of harbor receiving waters was found to resemble that of the CSO effluent under all weather conditions, but more so during periods of rainfall (EGANHOUSE and SHERBLOM, 2001). Thus, the CSO discharge was having a measurable effect on the local aquatic environment. By contrast, sediments deposited near the CSO outfall differed in chemical composition from the CSO effluent and harbor receiving waters. WALLACE et al. (1991) compared mass-emission rates of suspended solids, particulate organic carbon and trace metals from the Fox Point CSO with the rates of accumulation of these materials in nearby sediments (based on 21~ dating) and showed that less than 4% of the sedimentation near the CSO could be accounted for by the discharge of effluent solids. Thus, the provenance of the sediments, originally expected to be mainly the CSO, appeared to be non-local. Molecular markers were used to elucidate the non-local sources of sediment contamination. Data on the concentration of coprostanol and the ELAB/coprostanol concentration ratios of CSO effluent (E4), receiving water (W4), and sediment ($7) samples, as well as Deer Island and
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FIG. 2. (a) Concentrations of coprostanol and total linear alkylbenzene/coprostanol (~LAB/cop) concentration ratios of CSO effluent particles (E4), municipal sludges, receiving water particles (W4), and surficial sediments collected near the Fox Point CSO outfall ($7) and (b) scatter plot of total linear alkylbenzene (~LAB) and coprostanol concentrations in surficial sediments from the study site (modified from EGANHOUSE and SHERBLOM, 2001). Note: E4 and W4 refer to CSO effluent and receiving water samples, respectively, collected during high flow conditions as described in EGANHOUSEand SHERBLOM (2001). TEO, total extractable organic matter. Sediment station ($7) location is shown in Fig. 1.
Nut Island sludges, are shown in Fig. 2a. Whereas coprostanol concentrations of sludges from the two treatment plants are similar, the ~LAB/coprostanol ratios differ by a factor of 10. This difference is explained by the fact that at the time of this study LABs in Deer Island waste effluent (and sludge) came principally from residues normally present in linear alkylbenzenesulfonate surfactant-based detergents (EGANHOUSE et al., 1983). By comparison, Nut Island treatment plant received additional inputs from an industrial LAB sulfonating facility. The paired CSO effluent and receiving water samples (E4 and W4), both have low ~LAB/ coprostanol concentration ratios that are nearer the ratio found for Deer Island sludge than Nut Island sludge. This result is consistent with the fact that the Fox Point CSO serves a portion of the same drainage area as the Deer Island treatment plant. As noted above, the Deer Island plant did not receive inputs from the LAB sulfonating facility. Sediments at station $7, which have much lower coprostanol concentrations (because of dilution during sedimentation), are characterized by a ELAB/coprostanol concentration ratio that is virtually identical to that of the Nut Island sludge. This is illustrated in Fig. 2b, where the slope of the least-squares line (~LAB versus coprostanol) for all sediment samples is approximately 0.2. Assuming the ~LAB/coprostanol marker ratio is conserved, this indicates that the sediment contamination at station $7 is not primarily the result of wastes released through the CSO (or the Deer Island outfall). Rather, it most likely reflects transport of wastewater and/or sludge particles from the Nut Island treatment plant outfalls (see Fig. 1 for locations of Nut Island wastewater and sludge outfall termini). Possible transport pathways from these outfalls are shown in Fig. 3. As discussed in EGANHOUSE and SHERBLOM (2001), the proposed pathways are consistent with available information on tidal currents in Boston Harbor. Transport of municipal wastes, including sludge, out of the harbor and into Massachusetts Bay was an inefficient process (SIGNELL and BUTMAN, 1992; STOLZENBACH et al., 1993), and a large quantity of this material apparently found its way to quiet embayments along the inner margins of Boston Harbor where rapid sedimentation is favored (KNEBEL et al., 1991). Although the number of sediment samples in this study was limited, significant correlations between the concentrations of the S'PCBs and the molecular markers were found (Fig. 4). The similarity of the y-intercepts of the regression lines indicates that the background concentration of non-sewage derived PCBs is about 0.1-0.2/zg/g. Thus, as much as 60-80% of the PCBs in the heavily contaminated sediments near the Fox Point CSO are derived from sewage (EGANHOUSE and SHERBLOM, 2001).
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FIG. 3. Map of Boston Harbor showing probable transport routes (arrows) of effluent particles from wastewater and sludge outfalls to study site.
Palos Verdes Shelf CA Sediments deposited on the narrow continental shelf off Palos Verdes, CA are heavily contaminated by a variety of inorganic and organic substances (see references in EGANI-IOtJSE et al., 2000). This contamination is the direct result of discharge, through a submarine outfall system, of treated municipal wastewaters (Fig. 5). Among the contaminants of greatest concern in the sediments are DDT (and its metabolites) and PCBs. For example, it has been estimated that 100- 250 tons of total DDT presently reside in the shelf and upper slope sediments off Palos Verdes (LEE et al., 2002; MURRAY et al., 2002; MACGREGOR, 1976; MCDERMOTT et al., 1974). The DDT originated from chemical manufacturing wastes that were introduced through sewer lines to the Los Angeles County Sanitation Districts' (LACSD) treatment plant in Carson, CA and subsequently released through the outfall system. In the early 1990s, a major multidisciplinary research program was initiated to determine the quantity of contaminants in these sediments and to predict the long-term fate of DDT and the PCBs. Central to these investigations was a predictive modeling effort (SHERWOOD et al., 2002)
FIG. 4. Scatter plots of total polychlorinated biphenyl (~PCB) versus (a) coprostanol and (b) ~LAB concentrations in surficial sedimentsfrom study site. Lines and statisticsfor linear regression analysis are shown (modifiedfrom ECANHOUSEand SHERBLOM,2001).
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FIG. 5. Map of the Palos Verdes Shelf showing locations of municipal wastewater outfall system operated by the Los Angeles County Sanitation Districts (LACSD), the Portuguese Bend Landslide (PBL), and sediment-coring stations (3C, 522) (modified from EGANHOtJSEet al., 2000).
that relied on establishing sedimentation rates at key locations on the shelf. It was recognized that numerous sources (e.g. effluent particles, cliff erosion, landslide, biogenic detritus, fiver inputs, etc.) contributed sediment to the shelf and slope. During the period when DDT inputs were occurring (1947-1971), however, wastewater solids emanating from the outfall system were believed to represent the largest source. Correspondence between the vertical distribution of total organic carbon (TOC) at station 3C (approximately 6 - 8 km downcurrent from the outfall system) and the mass emissions of suspended solids from the outfall system during the period 1946-1981 is illustrated in Fig. 6. Following World War II and up until 1971, the monotonic increase in emissions of suspended solids from the LACSD paralleled the population trend in Los Angeles. Thereafter, solids emissions declined in response to improved source control and advances in waste treatment (STULL et al., 1996). The vertical concentration profile of TOC in the 3C (1981) core records the historical trend in effluent solids emissions and indicates that, for this period, the outfalls dominated sedimentation of organic carbon on the shelf. The decline in emissions of suspended solids from the outfalls after 1971 became a matter of concern because of the potential for remobilization of heavily contaminated sediments that had been laid down in earlier years. Because of difficulties in applying conventional radioisotopic methods of geochronology (e.g. 2~~ at the site, a study was undertaken to estimate sedimentation and mass accumulation rates by molecular stratigraphy. Selected persistent organic contaminants (i.e. DDT + metabolites, PCBs) and waste-specific molecular markers (the long-chain alkylbenzenes; EfaNnot;sz and PONTOLILLO, 2000) were used to reconstruct the history of waste emissions. Molecular stratigraphy involves correlating known or predicted rates of input of a chemical to a sedimentary system with the vertical distribution of that same chemical in a sediment column (VALETTE-SILVER, 1993). Dates are assigned to inflection points in vertical concentration profiles, and average sedimentation rates (cm/yr) can then be estimated for the corresponding
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FIG. 6. (a) Vertical concentration profile of total organic carbon (TOC) in sediment core collected from station 3C in 1981 and (b) historical emissions (mta -- metric tons yr -~) of suspended solids from the LACSD wastewater outfall system, 1946-1981 (modified from EGANI-IOtJSSand PONTOLILLO,2000). depth intervals. Chemical analyses of a gravity core collected by the L A C S D in 1981 at station 3C were c o m p a r e d directly with those made on a box core collected by the U S G S in 1992 at station 522, slightly inshore of station 3C (Fig. 5; EGANI-IOVSE and PONTOLILLO, 2000). The objective was to examine the variation in sedimentation (cm/yr) and mass accumulation (g/cm 2 yr) rates over time and to assess the relative importance of wastewater emissions vis-gt-vis other sediment sources such as the Portuguese Bend Landslide (PBL; see Fig. 5). The vertical concentration profiles of three molecular markers in the 3C (1981) and 522 (1992) cores are shown in Fig. 7. p,p~-DDE (1,1-dichloro-2,2-bis ( p - c h l o r o p h e n y l ) ethylene) is the major persistent D D T metabolite in shelf sediments ( 6 0 - 7 0 % of all D D T compounds; EGANHOUSE et al., 2000), 6-C12 (6-phenyldodecane) is the most persistent and abundant
FIG. 7. Vertical concentration profiles of molecular markers (6-C12 (6-phenyldodecane), TAB3 (a tetrapropylene-based alkylbenzene) and p,p/-DDE) in sediment cores collected from the Palos Verdes Shelf in (a) 1981 (station 3C) and (b) 1992 (station 522) and age-dates assigned based on historical discharge/usage patterns (modified from EGANHOUSEand PONTOLILLO,2000). Location of sediment coring stations is shown in Fig. 5.
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R.P. Eganhouse Table 2. Estimatedsedimentation (cm/yr) and mass accumulation (g/cm2-yr)rates for cores from stations 3C (1981) and 522 (1992) based on molecular stratigraphy Station/date
1955-1965
1965-1971
1971-1981
3C/1981 cm/yr g/cm2-yr
0.8 0.6
1.3 0.7
0.7 0.3
522/1992 cm/yr g/cmZ-yr
0.9 1.1
1.3 1.4
0.8 0.9
1981-1992
2.1 2.2
constituent of the LABs (see TAKADA and EGANHOUSE, 1998), and TAB3 is a highly persistent member of the tetrapropylene-based alkylbenzenes (TABs: see TAKADA and EGANHOUSE, 1998), a class of surfactant markers related to the LABs. Dates have been assigned according to known and/or estimated historical discharge/usage patterns for these compounds as discussed in EGANHOUSE and PONTOLILLO (2000). Based on the date assignments, average sedimentation (cm/yr) and mass accumulation (g/cm 2 yr) rates were computed by dividing the differences in 2 depth (cm) and mass accumulation ( g / c m ) by the time interval between collection of the two cores (11.2 yr). These results are given in Table 2. The estimated sedimentation and mass accumulation rates for three time periods (1955-1965, 1965-1971, 1971-1981) at station 3C are reasonably consistent with the historical trends in solids emissions from the outfall system (cf Table 2, Fig. 6). The sedimentation rates for station 522 during the same time periods are similar to those for station 3C. However, mass accumulation rates are systematically higher at station 522 than at station 3C, and there is a marked increase in both sedimentation rate and mass accumulation rate in the 1981-1992 period based on molecular stratigraphy (station 522 only). The latter observation is significant because emissions of wastewater effluent solids from the LACSD outfall system during the post-1981 period continued to decline (see Fig. 5 in EGANHOUSE and PONTOLILLO, 2000). Thus, the increase in sedimentation for the period 1981-1992 near station 522 (and presumably 3C) cannot be ascribed to an overriding influence of the outfall system. As discussed by EGANHOUSE and PONTOLILLO (2000), the cause of the increased sedimentation in the 1980s and early 1990s was most likely mobilization of coarser sediment from the Portuguese Bend Landslide, which resulted from large winter storms in 1982-1983 and 1988. This hypothesis is supported by the observation of sedimentological and mineralogical indicators of PBL sediment in the vicinity of stations 3C/522 (DRAKE, 1994; WONG, 2002; REYNOLDS, 1987). Prior to this work, it was presumed that transport of PBL debris primarily would be to the southeast via littoral currents (Fig. 8; KAYEN et al., 2002). Offshore transport and subsequent entrainment by poleward-moving currents could be envisioned as a means of bringing PBL debris to the vicinity of stations 3C/522. However, a more plausible explanation for the observed increase in sedimentation near stations 3C/522 in the 1980s and early 1990s and the systematically higher sediment accumulation rates at station 522 than station 3C for all time periods is that PBL debris was mobilized and transported offshore near to its source. Offshore transport would be favored in the vicinity of the PBL because poleward moving subthermocline currents are deflected to the west at Long Point owing to coastal bathymetry (Fig. 8; DRAKE, 1994). Under these circumstances, sediments deposited at station 522 would be expected to ,receive PBL inputs earlier and to a greater extent than those at station 3C. This case study provides an example of how data from molecular markers combined with sediment data can be used to elucidate sedimentation processes and transport pathways that might not have been identified using other techniques.
Bemidji, MN oil-spill site In 1979, a high-pressure pipeline carrying a light paraffinic crude oil ruptured near the town of Bemidji, MN. The nearly instantaneous release of approximately 1670 m 3 of oil resulted in its
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FIG. 8. Map of the Palos Verdes Shelf showing locations of wastewater outfall system, Portuguese Bend Landslide (PBL), sediment coring stations (3C, 522) and possible transport routes (arrows) of PBL debris.
accumulation within topographic lows and spraying of oil to the southwest of the pipeline (Fig. 9). After cleanup operations were completed, approximately 410 m 3 of oil remained unaccounted for. A portion percolated through the unsaturated zone and formed an irregularly distributed body of residual oil that came to rest on the water table. Soluble constituents of the oil entered ground water (forming a contaminant plume), and degradation of these substances resulted in production of a complex mixture of organic transformation products, as well as alteration of the geochemical conditions in the aquifer (EGANHOUSE et al., 1993a; BAEDECKER et al., 1993; BENNETT et al., 1993). Within a few years after the spill, degradation of the oil led to development of a lens of highly contaminated anoxic ground water beneath and 7 0 - 8 0 m downgradient from the oil body. Although a complex mixture of volatile aliphatic, aromatic, and alicyclic hydrocarbons was found in ground water near the residual oil body, roughly 85% of the volatile organic compounds was composed of MAHs (EGANHOUSE et al., 1993a). A number of these compounds are known or suspected carcinogens. Therefore, a study was undertaken to understand processes controlling the transport and fate of the MAHs within the anoxic plume. Because the source of these aromatic hydrocarbons was known with certainty, they were considered molecular markers of the crude oil and were used as process probes. Paired oil and ground-water samples were collected along the axis of the oil body in the general direction of ground-water flow; ground-water samples were also collected upgradient and downgradient from the oil body (Fig. 9). These samples were analyzed for benzene and all alkylbenzenes having 1 - 4 alkyl carbon substituents (Cl_4-benzenes; 36 compounds in all) using purge-and-trap high-resolution gas chromatography/mass spectrometry (EGANHOUSE et al., 1993b). Analysis of the paired oil and ground-water samples revealed that oil from the trailing edge of the oil body was depleted in MAHs compared with oil from the leading edge of the oil body (see EGANHOUSE et al., 1996). The extent of depletion of each MAH was found to be directly related to its aqueous solubility. A sequential batch equilibration simulation showed that most of the depletion could be explained by water washing of the oil as ground water passed beneath and through the residual oil body. At the same time, ground-water samples collected
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FIG. 9. Map of the Bemidji, MN crude oil-spill site showing approximate location of the residual oil body, the oil spray zone, and ground-water sampling wells along the general direction of ground-water flow. Main sampling transect is depicted as the bold line in general direction of ground-water flow (modified from EGANHOUSEet al., 1996).
along the axis of the oil body in the direction of ground-water flow (Fig. 9) showed increasing concentrations of MAHs. Together, these observations indicated that MAHs were dissolving from the oil into the ground water and that this process was leading to minor, but measurable, changes in composition of the residual oil. In order to assess the causes of observed downgradient changes in MAH composition, it was necessary to determine if ground water had reached saturation with respect to MAHs prior to being transported away from the oil body. Oil-water equilibration experiments were carried out using several oil samples collected from the field site (EGANHOUSEet al., 1996). The laboratory results were then compared to theoretical calculations based on the aqueous solubilities of individual MAHs. Good agreement was found between the laboratory results and theoretical calculations. This set the stage for computing the degree of saturation of MAHs in ground water within the anoxic plume downgradient from the oil. The 21 MAHs for which data were available fell into two groups. Compounds that clearly were undersaturated ( - 30 to - 90%; e.g. toluene, xylenes) tended to be less persistent within the anoxic plume downgradient from the oil. Compounds that were near saturation (_+ 30%; e.g. benzene, ethylbenzene) were more persistent within the anoxic plume downgradient from the oil. These findings are summarized in Fig. 10, where the minimum apparent disappearance rate (in units of % change in concentration/meter) within the anoxic plume is plotted versus the % difference between measured and equilibrium concentrations. Additional data on the temporal variability of MAH concentrations in ground water immediately downgradient from the oil showed the same type of compound grouping (EGANHOUSE et al., 1996). The more persistent
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MAHs (e.g. benzene, ethylbenzene) showed little temporal variation (< about 20%), whereas the less persistent MAHs (e.g. toluene, xylenes) showed greater temporal variability (20-139%). These findings suggest that compounds that are persistent under the conditions prevailing within the anoxic plume approach or attain saturation and show little variation in concentration with time because dissolution rates exceed removal rates. On the other hand, compounds that are subject to rapid removal under the prevailing environmental conditions are undersaturated because removal rates exceed dissolution rates. To gain some understanding of the removal processes responsible for attenuation of MAHs at the Bemidji site, a number of homologous series (e.g. the n-alkylbenzenes) and isomer groups were measured. The physical-chemical properties of homologous series differ systematically. For example, the log octanol-water partition coefficient (log Kow), which is a predictor of the tendency of a compound to sorb to particulate matter, increases from a value of 2.16 for benzene to 3.71 for n-propylbenzene (see Fig. lla; data from SHERBLOM and EGANHOUSE, 1988). If sorption was the principle mechanism responsible for the decreasing concentration of MAHs with distance downgradient from the oil, it is expected that benzene would be transported farthest and n-propylbenzene least. As shown in Fig. 1 l a, this clearly is not the case at the Bemidji oil-spill site. Here, the distributions of benzene + n-C]_3-benzenes are plotted against distance. The rapid attenuation of toluene and, by comparison, the persistence of npropylbenzene indicate that sorption does not control the transport and fate of these compounds. Because chemical degradation is not likely to be important, biological degradation is the most plausible explanation for attenuation of these MAHs within the anoxic plume. By the same reasoning, isomers having similar physical-chemical properties would be expected to behave the same if physical processes were dominant. However, three C3-benzene isomers having near identical log Kow values (SHERBLOM and EGANHOUSE, 1988) show strikingly different distributions (Fig. l lb). This result, again, indicates that structure-specific biological degradation, not sorption, exerts a dominant control on the transport and fate of these compounds. Although ancillary supporting evidence for biodegradation of MAHs at this site exists (BAEDECKERet al., 1993; COZZARELLIet al., 1994), this case study is a good illustration of how molecular markers can serve as process probes.
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CONCLUSIONS Molecular markers can be valuable tools for environmental investigations. With appropriate selection of marker compounds that are related to specific sources of organic matter and/or contamination, many types of information can be obtained. This information includes a better understanding of sources in complex multi-input systems as well as insights into the transport and fate of markers and associated contaminants. Unfortunately, the lack of data on the behavior of these compounds in the environment limits their utility for quantitative source apportionment. Studies are needed to ascertain the susceptibility of molecular markers to photolytic, chemical, and biological degradation processes. In addition, there is an ongoing need for reliable physicalchemical property data and systematic studies of the behavior of marker compounds vis-gt-vis non-source-specific organic contaminants of concern. These data are essential for successful application of modeling approaches to predicting the transport and fate of hazardous materials in air, land, and sea. Hydrophilic markers increasingly are being discovered as analytical limits are overcome (e.g. pharmaceuticals by electrospray high-performance liquid chromatography/mass spectrometry/mass spectrometry). Such developments and more widespread adoption of the molecular marker approach should advance our understanding of Man's impact on the environment. A c k n o w l e d g e m e n t s - - T h e author wishes to thank Dr Ian Kaplan, to w h o m this v o l u m e is dedicated, for providing the excellent scientific e n v i r o n m e n t at the University of California, Los Angeles, where the author c o m p l e t e d his doctoral and post-doctoral studies. The author also would like to a c k n o w l e d g e former students and research associates who have contributed to the data and interpretations discussed in this paper: P. Sherblom, C. Phinney (Boston Harbor), J. Pontolillo, T. Leiker (Palos Verdes, CA), M. J. Baedecker, I. Cozzarelli, T. Dorsey, C. Phinney, S. Westcott (Bemidji, MN). Finally, the author appreciates the helpful c o m m e n t s and suggestions of M. J. Baedecker, I. Cozzarelli, and Angel Martin who r e v i e w e d an early draft of this paper.
REFERENCES BAEDECKER M. J., COZZARELLII. M., EGANHOUSER. P., SIEGEL D. I. and BENNETT P. C. (1993) Crude oil in a shallow sand and gravel aquifer: III. Biogeochemical reactions and mass balance modeling in anoxic groundwater. Appl. Geochem. 8, 569-586. BENNETT P. C., SIEGEL D. I., BAEDECKER M. J. and HULT M. F. (1993) Crude oil in a shallow sand and gravel aquifer: I. Hydrogeology and inorganic geochemistry. Appl. Geochem. 8, 529-549. BOTHNER M. H., BUCHHOLTZTEN BRINK M. and MANHEIM F. T. (1998) Metal concentrations in surface sediments of Boston Harbor-changes with time. Mar. Environ. Res. 45(2), 127-155. BRASSEL S. C., EGLINTON G., MARLOWE I. T., PFLAUMANNU. and SARNTHEIN M. (1986) Molecular stratigraphy: a new tool for climatic assessment. Nature 320(6058), 129-133. COZZARELLI I. M., BAEDECKERM. J., EGANHOUSER. P. and GOERLITZ D. F. (1994) The geochemical evolution of lowmolecular-weight organic acids derived from the degradation of petroleum contaminants in groundwater. Geochim. Cosmochim. Acta 58, 863-877. DRAKE D. E. (1994) The natural recovery of contaminated effluent-affected sediment on the Palos Verdes Margin: background information and results of the USGS natural recovery research. Southern California Damage Assessment Witness Reports, 131 pp. EGANHOUSE R. P. (1997) Molecular markers and environmental organic geochemistry: an overview. In Molecular Markers in Environmental Geochemistry (ed. R. P. EGANHOUSE), pp. 1--20. American Chemical Society, Washington, DC. EGANHOUSE R. P. and PONTOLILLO J. (2000) Depositional history of organic contaminants on the Palos Verdes Shelf, California. Mar. Chem. 70, 317-338. EGANHOUSE R. P. and SHERBLOM P. M. (2001) Anthropogenic organic contaminants in the effluent of a combined sewer overflow: impact on Boston Harbor. Mar. Environ. Res. 51, 51-74. EGANHOUSE R. P., BLUMFIELD D. L. and KAPLAN I. R. (1983) Long-chain alkylbenzenes as molecular tracers of domestic wastes in the marine environment. Environ. Sci. Technol. 17, 523-530. EGANHOUSE R. P., BAEDECKER M. J., COZZARELLII. M., AIKEN G. R., THORN K. A. and DORSEY T. F. (1993a) Crude oil in a shallow sand and gravel aquifer: II. Organic geochemistry. Appl. Geochem. 8, 551-567. EGANHOUSE R. P., DORSEY T. F., PHINNEY C. S. and WESTCOTT A. M. (1993b) Determination of C6-C~0 aromatic hydrocarbons in water by purge-and-trap capillary gas chromatography. J. Chromatogr. 628, 81-92. EGANHOUSE R. P., DORSEY T. F., PHINNEY C. P. and WESTCOTT A. W. (1996) Processes affecting the fate of monoaromatic hydrocarbons in an aquifer contaminated by crude oil. Environ. Sci. Technol. 30, 3304-3312. EGANHOUSE R. P., PONTOLILLOJ. and LEAKERT. J. (2000) Diagenetic fate of organic contaminants on the Palos Verdes Shelf, California. Mar. Chem. 70, 289-315.
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FINDLAY R. H. and WATLING L. (1997) Seasonal variation in sedimentary microbial community structure as a backdrop for the detection of anthropogenic stress. In Molecular Markers in Environmental Geochemistry (ed. R. P. EGANHOUSE), pp. 49--64. American Chemical Society, Washington, DC. GOUIN T., MACKAY D., WEBSTER E. and WANIA F. (2000) Screening chemicals for persistence in the environment. Environ. Sci. Technol. 34(5), 881-884. HEDGES J. I. and PRAHL F. a. (1993) Early diagenesis: consequences for applications of molecular biomarkers. In Organic Geochemistry: Principles and Applications (eds. M. H. ENGEL and S. A. MACKO). Plenum Press, New York. KAPLAN I. R., GALPERIN Y., Lu S.-T. and LEE R.-P. (1997) Forensic environmental geochemistry: differentiation of fuel-types, their sources and release time. Org. Geochem. 27(5/6), 289-317. KATES M. (1997) Diether and tetraether phospholipids and glycolipids as molecular markers for archaebacteria (Archaea). In Molecular Markers in Environmental Geochemistry (ed. R. P. EGANHOUSE), pp. 35--48. American Chemical Society, Washington, DC. KAYEN R. E., LEE H. J. and HEIN J. R. (2002) Influence of the Portuguese Bend landslide on the character of the effluentaffected sediment deposit, Palos Verdes margin, southern California. Cont. ShelfRes. 22, 911-922. KNEBEL H. J., RENDIGS R. R. and BOTHNER M. H. (1991) Modem sedimentary environments in Boston Harbor, Massachusetts. J. Sedim. Petrol. 61, 791-804. LEE H. J., SHERWOOD C. R., DRAKE D. E., EDWARDS B. D., WONG F. and HAMER M. (2002) Spatial and temporal distribution of contaminated effluent-affected sediment, Palos Verdes margin, southern California. Cont. ShelfRes. 22, 859-880. MACGREGOR J. S. (1976) DDT and its metabolites in the sediments off southern California. Fish. Bull. 74, 27-35. MCDERMOTT D. J., HEESEN T. C. and YOUNG D. R. (1974) DDT in bottom sediments around five southern California outfall systems. Southern California Coastal Water Research Project, 54 pp. MURRAY C. J., LEE H. J. and HAMPTON M. A. (2002) Geostatistical mapping of effluent-affected sediment distribution on the Palos Verdes shelf. Cont. Shelf Res. 22, 881-897. PETERS K. E. and MOLDOWAN J. M. (1993) The Biomarker Guide. Prentice-Hall, Englewood Cliffs, NJ. REYNOLDS S. (1987) Sediment dynamics on the Palos Verdes Shelf. Report to the Los Angeles County Sanitation Districts, 59 pp. SCHAUER J. J., ROGGE W. F., HILDEMANN L. M., MAZUREK M. A., CASS G. R. and SIMONEIT B. R. T. (1996) Source apportionment of airborne particulate matter using organic compounds as tracers. Atmos. Environ. 30(22), 3837-3855. SHERBLOM P. M. and EGANHOUSE R. P. (1988) Correlations between octanol-water partition coefficients and reversedphase high-performance liquid chromatography capacity factors. Chlorobiphenyls and alkylbenzenes. J. Chromatogr. 454, 37-50. SHERWOOD C. R., DRAKE D. E., WIBERG P. L. and WHEATCROFT R. A. (2002) Prediction of the fate of p,p~-DDE in sediment on the Palos Verdes shelf, California, USA. Cont. ShelfRes. 22, 1025-1058. SmNELL R. P. and BUTMAN B. (1992) Modeling tidal exchange and dispersion in Boston Harbor. J. Geophys. Res. 97, 15591-15606. STOLZENBACH K. D., ADAMS E. E., LADD C. C., MADSEN O. S. and WALLACEG. (1993) Boston Harbor study of sources and transport of harbor sediment contamination, Part I: Transport of contaminated sediments in Boston Harbor. MWRA Technical Report, No. 93-12. STULL J. K., SWIFT D. J. P. and NIEDORODA A. W. (1996) Contaminant dispersal on the Palos Verdes continental margin: I. Sediments and biota near a major California wastewater discharge. Sci. Total Environ. 179, 73-90. TAKADA H. and EGANHOUSER. P. (1998) Molecular markers of anthropogenic waste. In Encyclopedia of Environmental Analysis and Remediation (ed. R. A. MYERS), pp. 2883-2940. Wiley, New York. TAKADA H. and ISHIWATARI R. (1989) Biodegradation experiments of linear alkylbenzenes (LABs): isomeric composition of C12 LABs as an indicator of the degree of LAB degradation in the aquatic environment. Environ. Sci. Technol. 24(1), 8 6 - 91. TAKADA H., SATOH F., BOTHNER M. H., TRIPP B. W., JOHNSON C. G. and FARRINGTONJ. W. (1997) Anthropogenic molecular markers: tools to identify the sources and transport pathways of pollutants. In Molecular Markers in Environmental Geochemistry (ed. R. P. EGANHOUSE), Vol. Symposium Series 671, pp. 178-195. American Chemical Society, Washington, DC. VALETTE-SILVERN. J. (1993) The use of sediment cores to reconstruct historical trends in contamination of estuarine and coastal sediments. Estuaries 16, 577-588. VOLKMAN J. K., REVILE A. T. and MURRAY A. P. (1997) Applications of biomarkers for identifying sources of natural and pollutant hydrocarbons in aquatic environments. In Molecular Markers in Environmental Geochemistry (ed. R. P. EGANHOUSE), pp. 110--132. American Chemical Society, Washington, DC. WALLACE a. T., KRAHFORSTC. F., PITTS L. C., SHINE J. P., STUDER M. M. and BOLLINGERC. R. (1991) Assessment of the chemical composition of the Fox Point CSO effluent and associated subtidal and intertidal environments: analysis of CSO effluents, receiving water and surface sediments for trace metals prior to CSO modification. Massachusetts Department of Environmental Protection. WANIA F. and MACKAY D. (1996) Tracking the distribution of persistent organic pollutants. Environ. Sci. Technol. 30(9), 390A-396A. WONG F. L. (2002) Heavy mineral provinces of the Palos Verdes margin, southern California. Cont. Shelf Res. 22, 899-910.
Molecular markers and their use in environmental organic geochemistry ROBERT P. EGANHOUSE Water Resources Discipline, US Geological Survey, 12201 Sunrise Valley Drive, MS 432, Reston, VA 20192-0002, USA Abstract--Molecular markers are organic substances that carry information about sources of organic matter or contamination. The source/marker relation can be used to indicate the presence of a given source material (qualitative), or, under appropriate conditions, to estimate the amount of a source material (quantitative source apportionment) in the environment. Assemblages of markers can also be used as process probes. In this instance, systematic differences and/or similarities in the physical-chemical properties of markers are coupled with compositional changes in marker composition to infer the operation of natural processes. This paper provides an overview of what molecular markers are, what types of markers are present in the environment, the requirements for the use of markers, and some common applications. To illustrate how molecular markers can answer specific environmental questions, three case studies are presented. The first case study examines the impact of municipal waste on a large urban harbor (Boston Harbor). Linear alkylbenzenes (unreacted residues of linear alkylbenzenesulfonate surfactants) and coprostanol (a fecal indicator) provide information on the sources and likely transport pathways of municipal wastes in a complex hydrologic system. The marker data are also used to estimate the proportion of sewage-derived polychlorinated biphenyls (PCBs) in polluted harbor sediments. The second case study concerns a portion of the continental shelf off southern California (Palos Verdes) where discharge of municipal wastewaters has led to extensive contamination of sediments and biota. Long-chain alkylbenzenes (surfactant residues), PCBs and the pesticide, DDT (dichlorodiphenyltrichloroethane), are used to develop sedimentation rate estimates for several time periods by molecular stratigraphy. These results, when combined with other information, allow conclusions to be drawn about the most likely transport pathway of sediments at the study site and to predict the fate of historically deposited contaminants. Finally, an investigation of a crude-oil spill in Bemidji, MN illustrates how monoaromatic hydrocarbons can be exploited as process probes, providing insights into the relative importance of different attenuation processes in a contaminated aquifer. The results show that natural attenuation of the monoaromatic hydrocarbons is occurring at this site and is dominated, not by physical and/or chemical processes, but by biodegradation.
INTRODUCTION ORGANIC CHEMICALS generated as a result of human activity enter the environment in vast quantities (on the order of hundreds of tons/yr). These inputs span a large range of temporal and spatial scales, and the media to which the chemicals are introduced (air, water, soil) vary. Some chemicals are applied intentionally over large areas (e.g. pesticides) or were used in a wide range of commercial products (e.g. polychlorinated biphenyls (PCBs)). Others are undesirable byproducts of industrial activity (e.g. combustion-derived polycyclic aromatic hydrocarbons (PAHs), polychlorinated dibenzodioxins (PCDDs)) or are man-made chemicals that emanate from a limited number of point sources (surfactants, pharmaceuticals). Many synthetic organic chemicals are individual compounds (e.g. lindane, atrazine), but complex mixtures of congeneric substances are not u n c o m m o n (e.g. PCBs, toxaphene). For these reasons, identification of the sources, transport pathways and fates of anthropogenic organic contaminants represents a significant challenge. Molecular markers, compounds whose structures or isotopic compositions are related to specific sources, offer a means of extracting such information directly from the environment. 143
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The mobility of a chemical depends on its tendency to partition among environmental compartments (atmosphere, hydrosphere, lithosphere, biosphere). Environmental partitioning, in turn, is determined by the physical-chemical properties of a substance and where and how it is released. GOUIN et al. (2000) have shown that the behavior of different chemicals can, to a first approximation, be predicted based on where they plot on a two-dimensional property space diagram (log air-water partition coefficient versus log octanol-water partition coefficient). Compounds, such as the low-molecular-weight n-alkanes (C1-8), catechol, and DDT, have physical-chemical properties that strongly favor partitioning into air, water, and soil/sediment phases, respectively. Other chemical classes or mixtures, such as the chlorobenzenes and PCBs, have a wide range of properties that favor partitioning among multiple phases. These are multimedia contaminants, and, if sufficiently persistent, they can be transported long distances and become globally distributed (WANIAand MACKAY,1996). This paper presents some basic background information about what molecular markers are, requirements for their use, and the types of applications that have been developed over the last 40 yr. Three case studies are offered as a means of illustrating some of the ways in which molecular markers can be used to answer specific environmental questions. These questions probably could not have been readily addressed using conventional environmental chemistry approaches. The objective of this paper is to stimulate interest in the molecular marker approach and to foster its use in environmental organic geochemistry. The interested reader is directed to EGANHOUSE (1997) for a more extensive discussion of molecular markers, their uses and limitations. BACKGROUND
Molecular markers as information carriers Molecular markers are carriers of information. This information is contained in their isotopic composition (613C, A14C, 615N, 6D, 637C1, etc.), chemical structure, or the particular mixture of molecules composing an assemblage. If the isotopic composition, structure, or molecular distributions of an individual or an assemblage of chemicals are unique to a given source, these substances can be used as source indicators. Molecular markers can also be used to investigate the operation of natural processes (i.e. process probes). The most common applications involve assemblages of compounds whose properties differ systematically (i.e. homologsman homologous series comprises compounds that differ by the chain length of a single alkyl substituent) or whose properties are similar or identical (i.e. isomers). Because chemicals partition among or are degraded within environmental media on a structure-specific basis, changes in marker composition can be used to infer the relative importance of different processes (e.g. volatilization, sorption, biodegradation, etc.).
Types of molecular markers As discussed in EGANHOUSE (1997), molecular markers can be classified as belonging to one of the three groups. Contemporary biogenic markers are synthesized by living organisms and can be found in the source organisms themselves (e.g. abietane) or, with little or no alteration, in the contemporary environment (e.g. dehydroabietic acid). Fossil biomarkers, good examples of which are the ubiquitous hopanoids, are biomolecules (e.g. bacteriohopanepolyols) that have been transformed through diagenetic and/or catagenetic processes. Even so, they largely retain the stereospecific carbon skeleton of the original biological precursor. The third group comprises anthropogenic markers. The presence of these compounds in the environment signals the influence of human activity or the input of wastes (TAKADA and EGANHOUSE, 1998). Anthropogenic markers can be broken into subgroups that include: (1) non-toxic synthetic chemicals (e.g. linear alkylbenzenesulfonate surfactants (LAS) used in commercial detergents), (2) non-toxic biogenic chemicals (e.g. coprostanol, urobilin), (3) toxic synthetic chemicals (e.g. PCBs, toxaphene), and (4) by-product chemicals (e.g. PCDDs or PAHs). The latter arise
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incidentally from human activities such as combustion of organic matter, synthesis of chemicals, or other industrial processes.
Requirements There are three principal requirements for the ideal molecular marker: (1) source-specificity, (2) persistence or predictable behavior, and (3) massive production/usage (TAKADA and EGANHOUSE, 1998). Ideally, a molecular marker should derive from a single source. However, this condition is rarely met. The Boston Harbor case study describes the use of waste-specific molecular markers (coprostanol, linear alkylbenzenes (LABs)) that serve as particle tracers in a complex urban harbor. In this instance, the markers arise exclusively from municipal wastes. Thus, the requirement of source-specificity is satisfied. Alternatively, when there is a massive local input of what ordinarily are non-source-specific contaminants, such compounds can be used as molecular markers of that local input. This situation is typified by the Palos Verdes Shelf, CA and the Bemidji, MN case studies where DDT (§ and volatile monoaromatic hydrocarbons (MAHs), respectively, serve as markers of local inputs. A molecular marker should be persistent or its behavior should be predictable. Ideally, one would desire conservative behavior, but given the diversity of catabolic enzymes, this is a rather demanding requirement for organic substances. In practice, information on the persistence and behavior of markers is often limited. Few studies have been undertaken to systematically determine the susceptibility of any of the commonly used markers to photolysis, chemical reaction, or biodegradation under an appropriate range of environmental conditions (EGANHOUSE, 1997). Even fundamental physical-chemical properties (e.g. vapor pressure, aqueous solubility, octanol-water partition coefficient), which are essential for multimedia fate modeling, are rarely known with confidence, and the database, at best, is spotty. This highlights an area where more reliable data are clearly needed. It is important to add that knowledge of the environmental behavior of a marker (e.g. LABs) vis-gt-vis a non-sourcespecific contaminant (e.g. PCBs) is essential for quantitative source apportionment of the latter (HEDGES and PRAHL, 1993; TAKADA et al., 1997). Finally, a molecular marker should be produced on a massive scale and/or its use should be widespread. This requirement effectively defines the range of applicability of a marker insofar as a substance must be readily detectable in a variety of environments for it to yield useful results. Markers, like other organic contaminants, are subject to the same range of spatial and temporal input functions and they are discharged to various environmental media.
Applications Table 1 summarizes various ways in which molecular markers have been applied in environmental studies during the last 40 yr. Source indicators-As source indicators, molecular markers are applied either qualitatively (source identification) or quantitatively (source apportionment). The requirements for quantitative source apportionment are considerably more stringent and the number of examples in the literature is fewer than for source identification (e.g. SCHAUER et al., 1996; EGANHOUSE and SHERBLOM,2001; TAKADA et al., 1997). Specific biogenic markers have been used for chemotaxonomic purposes (KATES, 1997) and to characterize the composition of benthic and pelagic communities present in aquatic ecosystems (FINDLAYand WATLING,1997). Many fossil biomarkers have served as indicators of organic matter provenance (PETERSand MOLDOWAN, 1993) or for paleoclimate reconstruction (BRASSELe t al., 1986). Biomarkers are also used frequently to identify sources of fossil fuel contamination in the contemporary environment (KAPLANet al., 1997; VOLKMAN et al., 1997). Finally, there are numerous studies in which molecular markers associated with municipal wastes, urban runoff, or combustion of fossil fuels have been used to infer the effect of various point and non-point sources of contamination (see references in TAKADA and EGANHOUSE, 1998).
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Source indicators Source identification (qualitative) Contemporary biogenic markers: chemotaxonomy, biogenic inputs Petroleum source rock identification, archaeological geochemistry and paleoclimate reconstruction, forensic geochemistry Impacts of major pollutant sources (e.g. municipal wastewaters, urban runoff, combustion of fossil fuels) a Source apportionment (quantitative) Microbial community composition Oil spills Emissions to the atmosphere Non-specific contaminants a Process probes Transport Pathways a,b,c Processes Dilution Resuspension Burial b Mixing b Fate Phase transfer processes Volatilization Vaporization Dissolution c Sorption c Degradation processes c Bioaccumulation Superscripted applications indicate how case studies in this paper reflect the use of molecular markers as source indicators and process probes: aBoston Harbor case study; bpalos Verdes Shelf, CA case study; CBemidji, MN oil-spill site case study.
Process probes-There are two general ways in which molecular markers are applied as process probes. Both rely upon the ability to relate a prospective marker to a specific source and, in some cases, a known input function. Markers can yield information on the pathways along which particles move (hydrophobic markers) or the direction of ground- and surface-water flow (hydrophilic markers). In addition, molecular markers provide unique approaches with which to understand processes of dilution, resuspension, burial, and mixing. Case studies presented below describe some aspects of these applications (see footnote to Table 1 for a key to these applications). In terms of the environmental fate of contaminants, three general process types can be distinguished: phase transfer, degradation, and bioaccumulation. All of these process types depend upon knowledge of the physical-chemical properties of the markers, transformation rates and pathways, measurements of the compositions of mixtures of compounds, and an understanding of the physics of the system under investigation. CASE STUDIES
Boston Harbor Boston Harbor is a shallow, glacially carved, tidally dominated estuary. Flushing occurs primarily through two channels, President Roads and Nantasket Roads, which connect the harbor with Massachusetts Bay (Fig. 1). For nearly 125 yr (BOTHNER et al., 1998), the wastes of metropolitan Boston and surrounding communities were discharged directly into the harbor with minimal or no treatment. Treated effluent and sludge were released from two plants (Deer Island and Nut Island) through a system of outfalls. When the capacities of the treatment plants were exceeded during wet weather, untreated sewage + runoff was discharged through a relief system consisting of 108 combined sewer overflows (CSOs). As recently as 1990, almost
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FIG. 1. Map of Boston Harbor showing locations of municipal wastewater and sludge outfalls, major channels to Massachusetts Bay and combined sewer overflows (CSOs). Inset depicts study area adjacent to the Fox Point CSO including stations where surficial sediment samples were collected (modified from EGANHOUSE and SHERBLOM,2001).
nothing was known about the effects these CSOs were having on water quality in the harbor. For this reason, a study was undertaken to determine the sources of organic contaminants to one of the largest CSOs (at Fox Point; see inset, Fig. 1) and to assess the fate of these materials after discharge to Boston Harbor (EGANHOUSE and SHERBLOM, 2001). In this instance, molecular markers not only served as qualitative source indicators, but also revealed information on transport pathways and allowed for quantitative source apportionment of a class of contaminants of concern, the PCBs. Samples of Fox Point CSO effluent and harbor receiving water along with sediments near the CSO outfall (Fig. 1) were collected during varying weather conditions. These samples, as well as sludges obtained from Deer Island and Nut Island treatment plants, were analyzed for two molecular markers of municipal wastes, the LABs (surfactant chemicals) and coprostanol (a fecal indicator). PCBs were also measured. LABs and coprostanol were found in the effluent during all sampling periods, indicating that even when weather conditions were dry, sewage entered the CSO system, presumably by leakage or through illegal connections. Moreover, concentrations of the LABs and PCBs in the CSO effluent were highly c o r r e l a t e d ( r 2 - - 0 . 8 7 ) suggesting that both of these substances were coming from the same source (namely sewage). Compositional features of the LABs, notably the I/E ratio (the ratio of internal to external isomers; see TAKADA and ISHIWATARI, 1989 for definition), which increases with biodegradation, indicated that sewage entering the CSO had not undergone appreciable alteration. Apparently the CSO was carrying untreated sewage even during dry weather. The chemical composition of harbor receiving waters was found to resemble that of the CSO effluent under all weather conditions, but more so during periods of rainfall (EGANHOUSE and SHERBLOM, 2001). Thus, the CSO discharge was having a measurable effect on the local aquatic environment. By contrast, sediments deposited near the CSO outfall differed in chemical composition from the CSO effluent and harbor receiving waters. WALLACE et al. (1991) compared mass-emission rates of suspended solids, particulate organic carbon and trace metals from the Fox Point CSO with the rates of accumulation of these materials in nearby sediments (based on 21~ dating) and showed that less than 4% of the sedimentation near the CSO could be accounted for by the discharge of effluent solids. Thus, the provenance of the sediments, originally expected to be mainly the CSO, appeared to be non-local. Molecular markers were used to elucidate the non-local sources of sediment contamination. Data on the concentration of coprostanol and the ELAB/coprostanol concentration ratios of CSO effluent (E4), receiving water (W4), and sediment ($7) samples, as well as Deer Island and
148
R.P. Eganhouse
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FIG. 2. (a) Concentrations of coprostanol and total linear alkylbenzene/coprostanol (~LAB/cop) concentration ratios of CSO effluent particles (E4), municipal sludges, receiving water particles (W4), and surficial sediments collected near the Fox Point CSO outfall ($7) and (b) scatter plot of total linear alkylbenzene (~LAB) and coprostanol concentrations in surficial sediments from the study site (modified from EGANHOUSE and SHERBLOM, 2001). Note: E4 and W4 refer to CSO effluent and receiving water samples, respectively, collected during high flow conditions as described in EGANHOUSEand SHERBLOM (2001). TEO, total extractable organic matter. Sediment station ($7) location is shown in Fig. 1.
Nut Island sludges, are shown in Fig. 2a. Whereas coprostanol concentrations of sludges from the two treatment plants are similar, the ~LAB/coprostanol ratios differ by a factor of 10. This difference is explained by the fact that at the time of this study LABs in Deer Island waste effluent (and sludge) came principally from residues normally present in linear alkylbenzenesulfonate surfactant-based detergents (EGANHOUSE et al., 1983). By comparison, Nut Island treatment plant received additional inputs from an industrial LAB sulfonating facility. The paired CSO effluent and receiving water samples (E4 and W4), both have low ~LAB/ coprostanol concentration ratios that are nearer the ratio found for Deer Island sludge than Nut Island sludge. This result is consistent with the fact that the Fox Point CSO serves a portion of the same drainage area as the Deer Island treatment plant. As noted above, the Deer Island plant did not receive inputs from the LAB sulfonating facility. Sediments at station $7, which have much lower coprostanol concentrations (because of dilution during sedimentation), are characterized by a ELAB/coprostanol concentration ratio that is virtually identical to that of the Nut Island sludge. This is illustrated in Fig. 2b, where the slope of the least-squares line (~LAB versus coprostanol) for all sediment samples is approximately 0.2. Assuming the ~LAB/coprostanol marker ratio is conserved, this indicates that the sediment contamination at station $7 is not primarily the result of wastes released through the CSO (or the Deer Island outfall). Rather, it most likely reflects transport of wastewater and/or sludge particles from the Nut Island treatment plant outfalls (see Fig. 1 for locations of Nut Island wastewater and sludge outfall termini). Possible transport pathways from these outfalls are shown in Fig. 3. As discussed in EGANHOUSE and SHERBLOM (2001), the proposed pathways are consistent with available information on tidal currents in Boston Harbor. Transport of municipal wastes, including sludge, out of the harbor and into Massachusetts Bay was an inefficient process (SIGNELL and BUTMAN, 1992; STOLZENBACH et al., 1993), and a large quantity of this material apparently found its way to quiet embayments along the inner margins of Boston Harbor where rapid sedimentation is favored (KNEBEL et al., 1991). Although the number of sediment samples in this study was limited, significant correlations between the concentrations of the S'PCBs and the molecular markers were found (Fig. 4). The similarity of the y-intercepts of the regression lines indicates that the background concentration of non-sewage derived PCBs is about 0.1-0.2/zg/g. Thus, as much as 60-80% of the PCBs in the heavily contaminated sediments near the Fox Point CSO are derived from sewage (EGANHOUSE and SHERBLOM, 2001).
Molecular markers and their use in environmental organic geochemistry
149
FIG. 3. Map of Boston Harbor showing probable transport routes (arrows) of effluent particles from wastewater and sludge outfalls to study site.
Palos Verdes Shelf CA Sediments deposited on the narrow continental shelf off Palos Verdes, CA are heavily contaminated by a variety of inorganic and organic substances (see references in EGANI-IOtJSE et al., 2000). This contamination is the direct result of discharge, through a submarine outfall system, of treated municipal wastewaters (Fig. 5). Among the contaminants of greatest concern in the sediments are DDT (and its metabolites) and PCBs. For example, it has been estimated that 100- 250 tons of total DDT presently reside in the shelf and upper slope sediments off Palos Verdes (LEE et al., 2002; MURRAY et al., 2002; MACGREGOR, 1976; MCDERMOTT et al., 1974). The DDT originated from chemical manufacturing wastes that were introduced through sewer lines to the Los Angeles County Sanitation Districts' (LACSD) treatment plant in Carson, CA and subsequently released through the outfall system. In the early 1990s, a major multidisciplinary research program was initiated to determine the quantity of contaminants in these sediments and to predict the long-term fate of DDT and the PCBs. Central to these investigations was a predictive modeling effort (SHERWOOD et al., 2002)
FIG. 4. Scatter plots of total polychlorinated biphenyl (~PCB) versus (a) coprostanol and (b) ~LAB concentrations in surficial sedimentsfrom study site. Lines and statisticsfor linear regression analysis are shown (modifiedfrom ECANHOUSEand SHERBLOM,2001).
150
R.P. Eganhouse
FIG. 5. Map of the Palos Verdes Shelf showing locations of municipal wastewater outfall system operated by the Los Angeles County Sanitation Districts (LACSD), the Portuguese Bend Landslide (PBL), and sediment-coring stations (3C, 522) (modified from EGANHOtJSEet al., 2000).
that relied on establishing sedimentation rates at key locations on the shelf. It was recognized that numerous sources (e.g. effluent particles, cliff erosion, landslide, biogenic detritus, fiver inputs, etc.) contributed sediment to the shelf and slope. During the period when DDT inputs were occurring (1947-1971), however, wastewater solids emanating from the outfall system were believed to represent the largest source. Correspondence between the vertical distribution of total organic carbon (TOC) at station 3C (approximately 6 - 8 km downcurrent from the outfall system) and the mass emissions of suspended solids from the outfall system during the period 1946-1981 is illustrated in Fig. 6. Following World War II and up until 1971, the monotonic increase in emissions of suspended solids from the LACSD paralleled the population trend in Los Angeles. Thereafter, solids emissions declined in response to improved source control and advances in waste treatment (STULL et al., 1996). The vertical concentration profile of TOC in the 3C (1981) core records the historical trend in effluent solids emissions and indicates that, for this period, the outfalls dominated sedimentation of organic carbon on the shelf. The decline in emissions of suspended solids from the outfalls after 1971 became a matter of concern because of the potential for remobilization of heavily contaminated sediments that had been laid down in earlier years. Because of difficulties in applying conventional radioisotopic methods of geochronology (e.g. 2~~ at the site, a study was undertaken to estimate sedimentation and mass accumulation rates by molecular stratigraphy. Selected persistent organic contaminants (i.e. DDT + metabolites, PCBs) and waste-specific molecular markers (the long-chain alkylbenzenes; EfaNnot;sz and PONTOLILLO, 2000) were used to reconstruct the history of waste emissions. Molecular stratigraphy involves correlating known or predicted rates of input of a chemical to a sedimentary system with the vertical distribution of that same chemical in a sediment column (VALETTE-SILVER, 1993). Dates are assigned to inflection points in vertical concentration profiles, and average sedimentation rates (cm/yr) can then be estimated for the corresponding
Molecular markers and their use in environmental organic geochemistry
151
FIG. 6. (a) Vertical concentration profile of total organic carbon (TOC) in sediment core collected from station 3C in 1981 and (b) historical emissions (mta -- metric tons yr -~) of suspended solids from the LACSD wastewater outfall system, 1946-1981 (modified from EGANI-IOtJSSand PONTOLILLO,2000). depth intervals. Chemical analyses of a gravity core collected by the L A C S D in 1981 at station 3C were c o m p a r e d directly with those made on a box core collected by the U S G S in 1992 at station 522, slightly inshore of station 3C (Fig. 5; EGANI-IOVSE and PONTOLILLO, 2000). The objective was to examine the variation in sedimentation (cm/yr) and mass accumulation (g/cm 2 yr) rates over time and to assess the relative importance of wastewater emissions vis-gt-vis other sediment sources such as the Portuguese Bend Landslide (PBL; see Fig. 5). The vertical concentration profiles of three molecular markers in the 3C (1981) and 522 (1992) cores are shown in Fig. 7. p,p~-DDE (1,1-dichloro-2,2-bis ( p - c h l o r o p h e n y l ) ethylene) is the major persistent D D T metabolite in shelf sediments ( 6 0 - 7 0 % of all D D T compounds; EGANHOUSE et al., 2000), 6-C12 (6-phenyldodecane) is the most persistent and abundant
FIG. 7. Vertical concentration profiles of molecular markers (6-C12 (6-phenyldodecane), TAB3 (a tetrapropylene-based alkylbenzene) and p,p/-DDE) in sediment cores collected from the Palos Verdes Shelf in (a) 1981 (station 3C) and (b) 1992 (station 522) and age-dates assigned based on historical discharge/usage patterns (modified from EGANHOUSEand PONTOLILLO,2000). Location of sediment coring stations is shown in Fig. 5.
152
R.P. Eganhouse Table 2. Estimatedsedimentation (cm/yr) and mass accumulation (g/cm2-yr)rates for cores from stations 3C (1981) and 522 (1992) based on molecular stratigraphy Station/date
1955-1965
1965-1971
1971-1981
3C/1981 cm/yr g/cm2-yr
0.8 0.6
1.3 0.7
0.7 0.3
522/1992 cm/yr g/cmZ-yr
0.9 1.1
1.3 1.4
0.8 0.9
1981-1992
2.1 2.2
constituent of the LABs (see TAKADA and EGANHOUSE, 1998), and TAB3 is a highly persistent member of the tetrapropylene-based alkylbenzenes (TABs: see TAKADA and EGANHOUSE, 1998), a class of surfactant markers related to the LABs. Dates have been assigned according to known and/or estimated historical discharge/usage patterns for these compounds as discussed in EGANHOUSE and PONTOLILLO (2000). Based on the date assignments, average sedimentation (cm/yr) and mass accumulation (g/cm 2 yr) rates were computed by dividing the differences in 2 depth (cm) and mass accumulation ( g / c m ) by the time interval between collection of the two cores (11.2 yr). These results are given in Table 2. The estimated sedimentation and mass accumulation rates for three time periods (1955-1965, 1965-1971, 1971-1981) at station 3C are reasonably consistent with the historical trends in solids emissions from the outfall system (cf Table 2, Fig. 6). The sedimentation rates for station 522 during the same time periods are similar to those for station 3C. However, mass accumulation rates are systematically higher at station 522 than at station 3C, and there is a marked increase in both sedimentation rate and mass accumulation rate in the 1981-1992 period based on molecular stratigraphy (station 522 only). The latter observation is significant because emissions of wastewater effluent solids from the LACSD outfall system during the post-1981 period continued to decline (see Fig. 5 in EGANHOUSE and PONTOLILLO, 2000). Thus, the increase in sedimentation for the period 1981-1992 near station 522 (and presumably 3C) cannot be ascribed to an overriding influence of the outfall system. As discussed by EGANHOUSE and PONTOLILLO (2000), the cause of the increased sedimentation in the 1980s and early 1990s was most likely mobilization of coarser sediment from the Portuguese Bend Landslide, which resulted from large winter storms in 1982-1983 and 1988. This hypothesis is supported by the observation of sedimentological and mineralogical indicators of PBL sediment in the vicinity of stations 3C/522 (DRAKE, 1994; WONG, 2002; REYNOLDS, 1987). Prior to this work, it was presumed that transport of PBL debris primarily would be to the southeast via littoral currents (Fig. 8; KAYEN et al., 2002). Offshore transport and subsequent entrainment by poleward-moving currents could be envisioned as a means of bringing PBL debris to the vicinity of stations 3C/522. However, a more plausible explanation for the observed increase in sedimentation near stations 3C/522 in the 1980s and early 1990s and the systematically higher sediment accumulation rates at station 522 than station 3C for all time periods is that PBL debris was mobilized and transported offshore near to its source. Offshore transport would be favored in the vicinity of the PBL because poleward moving subthermocline currents are deflected to the west at Long Point owing to coastal bathymetry (Fig. 8; DRAKE, 1994). Under these circumstances, sediments deposited at station 522 would be expected to ,receive PBL inputs earlier and to a greater extent than those at station 3C. This case study provides an example of how data from molecular markers combined with sediment data can be used to elucidate sedimentation processes and transport pathways that might not have been identified using other techniques.
Bemidji, MN oil-spill site In 1979, a high-pressure pipeline carrying a light paraffinic crude oil ruptured near the town of Bemidji, MN. The nearly instantaneous release of approximately 1670 m 3 of oil resulted in its
Molecular markers and their use in environmental organic geochemistry
153
FIG. 8. Map of the Palos Verdes Shelf showing locations of wastewater outfall system, Portuguese Bend Landslide (PBL), sediment coring stations (3C, 522) and possible transport routes (arrows) of PBL debris.
accumulation within topographic lows and spraying of oil to the southwest of the pipeline (Fig. 9). After cleanup operations were completed, approximately 410 m 3 of oil remained unaccounted for. A portion percolated through the unsaturated zone and formed an irregularly distributed body of residual oil that came to rest on the water table. Soluble constituents of the oil entered ground water (forming a contaminant plume), and degradation of these substances resulted in production of a complex mixture of organic transformation products, as well as alteration of the geochemical conditions in the aquifer (EGANHOUSE et al., 1993a; BAEDECKER et al., 1993; BENNETT et al., 1993). Within a few years after the spill, degradation of the oil led to development of a lens of highly contaminated anoxic ground water beneath and 7 0 - 8 0 m downgradient from the oil body. Although a complex mixture of volatile aliphatic, aromatic, and alicyclic hydrocarbons was found in ground water near the residual oil body, roughly 85% of the volatile organic compounds was composed of MAHs (EGANHOUSE et al., 1993a). A number of these compounds are known or suspected carcinogens. Therefore, a study was undertaken to understand processes controlling the transport and fate of the MAHs within the anoxic plume. Because the source of these aromatic hydrocarbons was known with certainty, they were considered molecular markers of the crude oil and were used as process probes. Paired oil and ground-water samples were collected along the axis of the oil body in the general direction of ground-water flow; ground-water samples were also collected upgradient and downgradient from the oil body (Fig. 9). These samples were analyzed for benzene and all alkylbenzenes having 1 - 4 alkyl carbon substituents (Cl_4-benzenes; 36 compounds in all) using purge-and-trap high-resolution gas chromatography/mass spectrometry (EGANHOUSE et al., 1993b). Analysis of the paired oil and ground-water samples revealed that oil from the trailing edge of the oil body was depleted in MAHs compared with oil from the leading edge of the oil body (see EGANHOUSE et al., 1996). The extent of depletion of each MAH was found to be directly related to its aqueous solubility. A sequential batch equilibration simulation showed that most of the depletion could be explained by water washing of the oil as ground water passed beneath and through the residual oil body. At the same time, ground-water samples collected
154
R.P. Eganhouse
FIG. 9. Map of the Bemidji, MN crude oil-spill site showing approximate location of the residual oil body, the oil spray zone, and ground-water sampling wells along the general direction of ground-water flow. Main sampling transect is depicted as the bold line in general direction of ground-water flow (modified from EGANHOUSEet al., 1996).
along the axis of the oil body in the direction of ground-water flow (Fig. 9) showed increasing concentrations of MAHs. Together, these observations indicated that MAHs were dissolving from the oil into the ground water and that this process was leading to minor, but measurable, changes in composition of the residual oil. In order to assess the causes of observed downgradient changes in MAH composition, it was necessary to determine if ground water had reached saturation with respect to MAHs prior to being transported away from the oil body. Oil-water equilibration experiments were carried out using several oil samples collected from the field site (EGANHOUSEet al., 1996). The laboratory results were then compared to theoretical calculations based on the aqueous solubilities of individual MAHs. Good agreement was found between the laboratory results and theoretical calculations. This set the stage for computing the degree of saturation of MAHs in ground water within the anoxic plume downgradient from the oil. The 21 MAHs for which data were available fell into two groups. Compounds that clearly were undersaturated ( - 30 to - 90%; e.g. toluene, xylenes) tended to be less persistent within the anoxic plume downgradient from the oil. Compounds that were near saturation (_+ 30%; e.g. benzene, ethylbenzene) were more persistent within the anoxic plume downgradient from the oil. These findings are summarized in Fig. 10, where the minimum apparent disappearance rate (in units of % change in concentration/meter) within the anoxic plume is plotted versus the % difference between measured and equilibrium concentrations. Additional data on the temporal variability of MAH concentrations in ground water immediately downgradient from the oil showed the same type of compound grouping (EGANHOUSE et al., 1996). The more persistent
Molecular markers and their use in environmental organic geochemistry
155
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MAHs (e.g. benzene, ethylbenzene) showed little temporal variation (< about 20%), whereas the less persistent MAHs (e.g. toluene, xylenes) showed greater temporal variability (20-139%). These findings suggest that compounds that are persistent under the conditions prevailing within the anoxic plume approach or attain saturation and show little variation in concentration with time because dissolution rates exceed removal rates. On the other hand, compounds that are subject to rapid removal under the prevailing environmental conditions are undersaturated because removal rates exceed dissolution rates. To gain some understanding of the removal processes responsible for attenuation of MAHs at the Bemidji site, a number of homologous series (e.g. the n-alkylbenzenes) and isomer groups were measured. The physical-chemical properties of homologous series differ systematically. For example, the log octanol-water partition coefficient (log Kow), which is a predictor of the tendency of a compound to sorb to particulate matter, increases from a value of 2.16 for benzene to 3.71 for n-propylbenzene (see Fig. lla; data from SHERBLOM and EGANHOUSE, 1988). If sorption was the principle mechanism responsible for the decreasing concentration of MAHs with distance downgradient from the oil, it is expected that benzene would be transported farthest and n-propylbenzene least. As shown in Fig. 1 l a, this clearly is not the case at the Bemidji oil-spill site. Here, the distributions of benzene + n-C]_3-benzenes are plotted against distance. The rapid attenuation of toluene and, by comparison, the persistence of npropylbenzene indicate that sorption does not control the transport and fate of these compounds. Because chemical degradation is not likely to be important, biological degradation is the most plausible explanation for attenuation of these MAHs within the anoxic plume. By the same reasoning, isomers having similar physical-chemical properties would be expected to behave the same if physical processes were dominant. However, three C3-benzene isomers having near identical log Kow values (SHERBLOM and EGANHOUSE, 1988) show strikingly different distributions (Fig. l lb). This result, again, indicates that structure-specific biological degradation, not sorption, exerts a dominant control on the transport and fate of these compounds. Although ancillary supporting evidence for biodegradation of MAHs at this site exists (BAEDECKERet al., 1993; COZZARELLIet al., 1994), this case study is a good illustration of how molecular markers can serve as process probes.
156
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M o l e c u l a r markers and their use in environmental organic geochemistry
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CONCLUSIONS Molecular markers can be valuable tools for environmental investigations. With appropriate selection of marker compounds that are related to specific sources of organic matter and/or contamination, many types of information can be obtained. This information includes a better understanding of sources in complex multi-input systems as well as insights into the transport and fate of markers and associated contaminants. Unfortunately, the lack of data on the behavior of these compounds in the environment limits their utility for quantitative source apportionment. Studies are needed to ascertain the susceptibility of molecular markers to photolytic, chemical, and biological degradation processes. In addition, there is an ongoing need for reliable physicalchemical property data and systematic studies of the behavior of marker compounds vis-gt-vis non-source-specific organic contaminants of concern. These data are essential for successful application of modeling approaches to predicting the transport and fate of hazardous materials in air, land, and sea. Hydrophilic markers increasingly are being discovered as analytical limits are overcome (e.g. pharmaceuticals by electrospray high-performance liquid chromatography/mass spectrometry/mass spectrometry). Such developments and more widespread adoption of the molecular marker approach should advance our understanding of Man's impact on the environment. A c k n o w l e d g e m e n t s - - T h e author wishes to thank Dr Ian Kaplan, to w h o m this v o l u m e is dedicated, for providing the excellent scientific e n v i r o n m e n t at the University of California, Los Angeles, where the author c o m p l e t e d his doctoral and post-doctoral studies. The author also would like to a c k n o w l e d g e former students and research associates who have contributed to the data and interpretations discussed in this paper: P. Sherblom, C. Phinney (Boston Harbor), J. Pontolillo, T. Leiker (Palos Verdes, CA), M. J. Baedecker, I. Cozzarelli, T. Dorsey, C. Phinney, S. Westcott (Bemidji, MN). Finally, the author appreciates the helpful c o m m e n t s and suggestions of M. J. Baedecker, I. Cozzarelli, and Angel Martin who r e v i e w e d an early draft of this paper.
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