Anthropogenic Molecular Markers: Tools To Identify the Sources and

Chapter 12

Anthropogenic Molecular Markers: Tools To Identify the Sources and Transport Pathways of Pollutants 1

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HideshigeTakada ,FutoshiSatoh ,Michael H.Bothner ,Bruce W. Tripp , Carl G.Johnson ,and John W. Farrington 3

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Downloaded by COLUMBIA UNIV on December 10, 2014 | http://pubs.acs.org Publication Date: July 1, 1997 | doi: 10.1021/bk-1997-0671.ch012

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Faculty of Agriculture, Tokyo University of Agriculture and Technology, Fuchu, Tokyo 183, Japan U.S. Geological Survey, Woods Hole, MA 02543 Woods Hole Oceanographic Institution, Woods Hole, MA 02543 2

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The activities of modem civilization have released to the oceans a wide variety of both mobilized natural compounds and synthetic compounds not found prior to modem times. Many of these com pounds provide a means of identifying sources of inputs and path ways of movement of chemicals through oceanic ecosystems and serve as molecular markers of human activities. A coastal ocean (Tokyo Bay) and a deep ocean (Deep Water Dump Site 106 in the Western North Atlantic Ocean) example are presented. In the deep ocean study, the correlation between potential sewage marker, i.e. linear alkylbenzenes (LABs), and polychlorinated biphenyls (PCBs) concentrations indicates a contribution of sewage sludge PCBs to the dump site sediments. The analysis of organic molecular markers to acquire information about environ mental processes has been an important concept in organic geochemistry during the past three to four decades. The absolute and relative concentrations of organic com pounds in an environmental sample reflect both the original sources of the organic matter and the alteration processes which have occurred in the environment. Signifi cant advances in analytical methodology have been accompanied by an increase in the number of organic compounds yielding important information about environmental samples of all types (7, 2). Human activity continues to add an increasing variety of organic compounds into the environment or has changed the ratios and amounts of naturally occurring com pounds. Both anthropogenic and naturally occurring compounds are found mixed together in recent environmental samples (Figure 1), and several of these compounds may be used as tracers to study natural processes affecting the fate and effects of chemical contaminants in the ocean. Analyses of anthropogenic compounds and their post-discharge degradation products in the environment concurrent with analyses of natural compounds provide information which can be used to more fully determine the history of pollutant release and the transport pathways and fates of anthropogenic compounds in the ocean. Thetimeand space scales of environmental processes that can be studied using molecular marker compounds is dependent on the number of sources, magnitude of each source, and persistence in the environment. 178

© 1997 American Chemical Society

In Molecular Markers in Environmental Geochemistry; Eganhouse, R.; ACS Symposium Series; American Chemical Society: Washington, DC, 1997.

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Classification and Properties of Anthropogenic Markers. Many anthro pogenic markers have been investigated in recent decades. Table I shows one possi ble scheme for classifying markers based on properties such as hydrophobicity and on the markers' major sources. Chemical structures of many of these markers are shown in the Appendix. The criteria for selecting a compound to be used as an an thropogenic marker include source specificity, amount used and released to the envi ronment, and persistence in the environment. An ideal anthropogenic marker should be contributed from specific sources so that an unambiguous pathway from input to deposition can be demonstrated. A marker should be used and released in quantities sufficient to permit detection after dilution in the environment. Finally, a prospective marker should be resistant over time to environmental alterations which might result in loss to unidentifiable or common compounds. Table I. Classification of Anthropogenic Markers Hydrophobic markers

Source SEWAGE

Natural products

Water-soluble markers

(I) Coprostanol (III) Urobilin (II) a-Tocopheryl acetate (TV) Caffeine CV) Aminopropanone

Synthetic (VI) Linear alkylbenzenes (X) Linear alkylbenzenedetergents (LABs) sulfonates (LAS) (VH) Tetrapropylene-based (XI) Tetrapropylene-basedalkylbenzenes (TABs) alkylbenzenesul fonates (ABS) (VIII) Trialkylamines (XU) Dialkyltetralin(TAMs) sulfonates (DATS) (IX) Nonylphenols (XIII) 4,4'-bis(2-sulfo(NPs) styryl)biphenyl (DSBP) ;

Street Runoff

(XIV) 2-(4-morphoIinyl)benzothiazole (24MoBT)

Multiple Sources

Polychlorinated biphenyls (PCBs) Polycyclic aromatic hydrocarbons (PAHs) Pol y organosiloxanes (silicones) ^ *Roman numbers in parentheses correspond to numbers in Appendix, which can be found on page 192. Compounds with multiple sources, widespread industrial application, and resis tance to degradation have been used to study local, regional and global processes. These include polycyclic aromatic hydrocarbons (PAHs; e.g. ref. 3) and polychlori nated biphenyls (PCBs; e.g. ref. 4). Polyorganosiloxanes have also been used as indicator of inputs from modem human activities (5). Sometimes, labile compounds have been utilized in combination with persistent markers from the same source where comparison between labile and stable compounds provides data on the degree


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MOLECULAR MARKERS IN ENVIRONMENTAL GEOCHEMISTRY

Sources Natural Biological Activities

Human Activities


Environmental Alteration ^ /

Environmental samples

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Organic Analysis

Information \ •Origin & Alteration of organic matter •sources of contaminants •transport pathways and fates of contaminants Figure 1. Concept of anthropogenic marker approach.

Hydrophoblclty Low

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Coprostanol LAS LABs NPs Silicones PCBs PAHs 01

2 3 4 5 6 7 8 9 10 log K Q W

Figure 2. Octanol-water partition coefficient (K ) of the molecular markers. References for the data source are as follows, coprostanol: (13), LAS: (14), LABs: (9), NPs: (75), silicones: (16), PCBs: (17), PAHs: (17). AT for copro stanol is estimated from water solubility data of cholesterol (75) using an equation in ref. 7 7 p. 139 (Table 7.2) ow

0W



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of environmental alteration or degradation, or about the residence time of the con taminants (e.g. ref. 6). We discuss two examples of the use of linear alkylbenzenes (LABs; VI in Ap pendix) as molecular markers for sewage later in this paper. LABs having Cio - Q4 normal alkyl chains are sulfonated in the industrial production of linear alkylbenzene sulfonates (LAS; X in Appendix), which are widely used anionic surfactants. How ever, the sulfonation of LABs into LAS is not complete, and the unsulfonated residue is present in LAS-type synthetic detergents (6, 7). Use of LAS-detergents and subse quent disposal, thus, brings LABs into aquatic environments (6, 8)\ in most instances by way of sewage discharges. Since LABs are highly hydrophobic (9) they sorb to particles and become useful for tracking sewage particles and sewage-derived hydro phobic pollutants (70, 77). Differences in hydrophobicity between "hydrophobic" markers (e.g. comparing PCBs and LABs) and within a class of hydrophobic markers (e.g. comparing PCB congeners) are important to consider when using these compounds as tracers of transport processes. Measurements of differences in environmental distributions cor related with measures or estimates of hydrophobicity provide clues to processes in fluencing the fate of the compounds. For example, less hydrophobic markers such as nonylphenols (IX in Appendix) could selectively desorb from sediment particles rela tive to more hydrophobic markers such as LABs (72). The octanol-water partition coefficient (AT ), is a surrogate measure of the hydrophobicity of a compound. Fig ure 2 presents K s for several groups of markers. The use of markers as tools for contaminant source discrimination also requires consideration of the similarity and/or differences in hydrophobicity and sorption mechanism between the markers and particular media. Sorption processes have been described using the concept of compounds partitioning between organic matter in particles and water depending on the hydrophobicity of the compound. During the 1980's, it was demonstrated that the sorption of organic compounds to particles is more complex than simple two-phase particle-water equilibrium (77, 18 and refer ences therein). The association of contaminants with organic colloids is seen to en hance the mobility of hydrophobic compounds in sediments (18, 79). Combustionderived PAHs are strongly associated with soot particles and are not easily available for equilibrium partitioning (20-22). Each of these complex mechanisms requires careful consideration in the quantitative application of anthropogenic markers. 0W

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Field Application of The Molecular Marker Approach We provide two examples, studies in Tokyo Bay and at Deep Water Dump Site 106, from our own research as illustrations of the more general approach of using mo lecular markers by researchers in the field. Tokyo Bay Sewage Effluent Contributions to Organic Contaminants in Sediments. There are considerable horizontal gradients of LAB concentrations in surface sediments of Tokyo Bay as shown in Figure 3 (10, 23). In Tokyo, some un treated wastewater is discharged directly into streams and rivers which ultimately flow into Tokyo Bay. The highest concentrations of LABs are clearly seen 5-10 km off the mouths of the major inflowingrivers.The embedded graph in Figure 3 further documents that the deposition ofriverdischarged hydrophobic contaminants are con centrated in the northern Bay. Knowledge of this depositional pattern based on mo lecular markers is beneficial to other studies such as a harbor-wide monitoring pro gram. For example, if we were concerned about inputs of chemicals of environmental concern such as PAHs or PCBs, inputs from the major rivers accumulating in the sediments and also contaminating the animals in the benthic ecosystem, we could use the LAB data to guide our selection of stations and samples for detailed analyses.




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Figure 3. Horizontal distribution of LABs in surface sediments of the Tokyo Bay. Contour lines based on data for 24 stations in the northern bay (10). Bar graph showing 2LAB concentrations along with the transect from S-0 to S-l I (23).



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Anthropogenic compounds have been deposited since the onset of industrial syn thesis and/or the increase in human activities and can, thus, serve as recent (usually annual to decadal scale resolution) geochronometers in sediments. Data from a core in Tokyo Bay illustrates a simple use to estimate the date of a depth interval in the sediment. For this core, excess lead-210 shows constant values throughout the upper 20 cm (Katoh, Y., Tokai University, personal communication, 1994), probably be cause of strong sediment mixing due to physical or biogeochemical processes. Thus, the excess lead-210 procedure for estimating sediment accumulation rates (24) does not provide a reliable sediment accumulation rate. In this type of situation, anthro pogenic markers with a known production record may provide time reference points in sediment cores. We can use the historic record of alkylbenzene production in Japan (Figure 4a) to provide a time reference point for Tokyo Bay sediments. Tetrapropylene-based alkyl benzenes (TABs; VII in Appendix) have branched alkyl chains which result in poor biodegradability of the sulfonated surfactants (XI in Appendix) and have caused some environmental problems. As a result, LABs were introduced and ultimately replaced TABs by the late 1960s. Figure 4b shows the vertical distribution of LABs and TABs in a sediment core collected from Tokyo Bay. Depth profiles in this core provide an approximate time for the depth at which compounds initially entered the ecosystem. The TAB and LAB concentration vs. depth profiles would also be influenced by sediment mixing processes. The decreasing TAB concentration vs. sediment depth from a maximum at about 30 cm to the lower concentrations at the sediment-water interface can be explained by a reduction of environmental loading accompanied by regular dilution through mixing with more recently deposited sediments. A date of no older than approximately 1958 can be assigned for sediment at the bottom of the "mixed zone" - approximately 35 cm since TABs were not produced prior to 1958. By similar reasoning, using the depth profiles of LAB concentrations, the sediment at 30 cm depth is no older than approximately 1965. If there had not been a disturbance of the sediment by either physical or biological mixing, then the depth profile could have provided us with a more detailed historical record of TAB and LAB releases to Tokyo Bay in the manner that historical records of other contaminants such as PAHs, PCBs, DDT have been measured in sediment cores from other aquatic ecosystems (25). Use of the historic record approach of measuring molecular markers in undisturbed sediments to document the release of LABs to coastal areas has been reported by Eganhouse and Kaplan (26). Deep Water Dumpsite (DWDS) 106 Study. DWDS 106 is located 106 miles off the coast of New Jersey, USA, in 2300-2900 m water depth (Figure 5) and, while operational, was the world's largest deep water sewage sludge dump site (27, 28). Approximately 42 million wet tons (metric) of sewage sludge were dumped between 1986 and 1992 when all dumping stopped (28). Sewage sludge, sediment trap and bottom sediment samples obtained from the site in 1989 and 1990 were ana lyzed for molecular markers in order to obtain evidence of transport and accumula tion of dumped sludge particles to the deep sea floor (11, 27). We summarize our results reported in detail earlier for LABs and coprostanol (77) and use these to con strain our results for PCBs reported in this paper. Sewage sludge contains PCBs, and the dumped sludge could contribute to PCBs in DWDS-106 sediments. PCBs are known to have diverse sources of input (e.g. atmospheric transport or lateral transport from coastal zones) in addition to sludge disposal. We have attempted to estimate a specific contribution of sewage-derived PCBs to DWDS-106 sediments. This estimation is based on the assumptions that LABs are sewage-specific, the K s of LABs and PCBs overlap significantly (9), and that the sorption mechanisms of the two compound classes are similar. ow



Production (x 10 ton/y) 3


o

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Concentrations (ng/g-dry sediment) 0

500

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Figure 4. (a) Historic record of long chain alkylbenzene production in Japan; open circle: LABs; closed triangle :TABs; (b) Vertical distribution of LABs and TABs in a sediment core collected from Tokyo Bay in 1993 (Satoh & Takada, unpublished data); open circle: LABs excluding peaks coeluting with TABs (i.e. 6-, 4-, 3- and 2-Cn); closed triangle: TABs excluding peaks coeluting with LABs. Sampling location is indicated in Figure 3 as open square.

Figure 5. Sampling location for the Deep Water Dump Site 106 Study. Closed circles indicate surface sediment locations in and around the dump site.



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Sampling and Analytical Methods. Surface (0-1 cm) sediment samples were collected from the dump site and the control site by DSRV Alvin in September 1989 and August 1990 (Figure 5). The control site is ca. 100 km northeast (i.e. upcurrent) of the dump site. Sinking particles were also collected by deploying sediment traps from September 1989 to August 1990 in the lower 110 m of water column at two locations along the western edge of the dump site. The details of sampling were described previously (27). Two sewage sludge samples from New York City were also analyzed. The previously described analysis of LABs and coprostanol (11) is summarized in Figure 6. The sum of all 26 secondary LAB congeners is represented as "2LAB". The detection limit of 2LAB was 0.2 ng/g-dry sediment. Coprostanol, epicoprostanol and cholesterol were separated from each other and quantified by capillary GC-FID after acetylation. The detection limit of coprostanol was 1 ng/g-dry sedi ment. Procedural blanks for LABs and coprostanol were below the limit of detection. The extraction and isolation method for PCBs is outlined in Figure 6. Individual concentrations of PCBs were determined on a Hewlett Packard 5890 gas chroma tograph equipped with a 30-m DB-5 capillary column (0.25 mm i.d.; 0.25 pirn film thickness; J&W Scientific) and an electron capture detector. Seventy nine chroma tographic peaks were identified and quantified using a standard PCB mixture (1: 1: 1: 1 mixture of Kanechlor 300:400:500:600, kindly supplied by Dr. Tanabe, Ehime University). The peak identification of the standard was performed by GC-MS, re tention index (29), and previously reported identification of Kanechlor (30). The congener composition of the standard mixture was determined by GC-FID analyses. The sum of the 79 chromatographic peaks is expressed as 2PCB. Replicate analyses (n=4) of a sediment sample showed less than 5% relative standard deviation for indi vidual PCB concentrations. Although PCB concentrations were not corrected for re covery, analyses of sediment samples spiked with the PCB standard showed a re covery of greater than 90%. Procedural blanks for PCBs through the entire procedure correspond to less than 0.06 ng £PCB/g-dry sediment. Transport of Sewage Particles to the Deep Sea Floor. LABs were de tected in the dump site sediments and sediment trap samples, whereas no LABs were detected in the control site sediments. This fact clearly indicates that despite horizontal dispersion and dilution immediately after dumping (28) the dumped sludge particles were transported vertically through the water column of over 2300 m, and accumu lated in the deep sea sediments. The distribution pattern of LABs is consistent with the sewage sludge deposition model of Fry and Butman (31) and is well correlated with those for other anthropogenic markers: coprostanol (I in Appendix), silver, and Clostridium perfringens, a bacterium found in human fecal material (27). The coprostanol detected at the control site at a concentration of 12 ng/g (significantly above the analytical blank) is attributed to contributions from the feces of birds (52), marine mammals such as whales (55), or in situ microbial or biogeochemical alterna tion of cholesterol to coprostanol (34). The isomeric distribution of LABs indicates the absence of significant microbial attack during transport. Microbiological incubation experiments (55, 56) and envi ronmental observations (8, 37) have demonstrated the selective depletion of isomers having phenyl substitution positions near the end of alkyl chains (i.e. external iso mers) relative to isomers having phenyl substitution positions near the center of alkyl chains (i.e. internal isomers). To quantitatively express the ratio of the internal to ex ternal isomers, an I/E ratio ([6-Cj AB + 5-C, AB]/[4~Ci AB + 3-Q AB + 2-Q AB]) is defined; m-C ; m : phenyl substitution position, n : alkyl carbon number (56). I/E ratios in untreated wastewater are approximately 0.7, a higher I/E ratio indicates a greater depletion of external isomers and, thus, more extensive degradation (56). As is obvious from Figure 7b, within the dump site samples, no increasing trend in I/E 2

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Solvent Extraction I Ultrasonic Extraction I with isopropanol, methanol, chloroform

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Figure 7. (a) LAB and coprostanol concentrations, (b) I/E ratio of LABs, and (c) a ratio of coprostanol relative to cholesterol in the sewage sludge, sediment trap, and sediment samples collected from the DWDS 106 site (77).


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ratio was observed from sludge (1.7 ± 0.2), to sediment trap (1.3 ± 0.3), to sediment samples (1.7 ± 0.2). This suggests that no significant biodegradation of LABs occurs in the water column or at the sediment-water interface. This lack of LAB degradation in the deep sea is probably caused by rapid sinking of dumped sludge particles, or incorporation of sewage particles into rapidly sinking fecal material. Low water tem perature and (36) and high pressure (38) in the deep sea might also suppress micro bial activity leading to LAB degradation. Dilution Factors of Sewage Particles. LAB concentrations in the sedi ment trap samples are about two orders of magnitude lower than those in sludge (Figure 7a). In the sediment trap, sludge particles are probably diluted by naturally produced particles and less contaminated resuspended sediment. This interpretation is supported by a lower ratio of coprostanol to cholesterol in sediment trap samples than in sludge (Figure 7c). Because phytoplankton and zooplankton produce cholesterol but not coprostanol, the addition of naturally produced sterols to sewage-derived sterols decreases the ratio. Using an average LAB concentration of sewage sludge of 152 x 10 " g/g dry weight, and assuming conservative behavior for LABs during transport through the water column, we estimate a sewage sludge dilution factor of between 1:1 and 1:500 from our sludge and sediment trap data (/7). Hunte* al. (28) have reported on analyses of seven sludge samples from Sewage Authorities contrib uting sludge to the material dumped at DWDS-106. The range of LAB concentrations was 39.8 to 150 x 10 g/g dry weight sludge with a mean of 123 x 10' g/g dry weight sludge. In addition to the absence of significant microbial degradation of the LABs as evi denced by the I/E ratio discussed above, dissolution of LABs from the particles into the aqueous phase should be negligible due to the extremely high hydrophobicity of LABs (log Kq : 6.90 - 9.29; ref. 9). Using the coprostanol concentrations, dilution factors in sediment trap samples ranging between 1:11 and 1:700 can be estimated. (Figure 7a). These dilution factors are somewhat higher than those calculated using LABs perhaps due to faster degradation and/or dissolution of coprostanol relative to LABs. LAB concentrations in the sediments from the dump site are roughly two orders of magnitude lower than those for the sediment trap samples (Figure 7a). This de crease is probably due to mixing with sediments containing no LABs which had been deposited before the onset of sludge dumping. The mixing hypothesis is supported by the vertical distribution of LABs in a sediment core collected from the dump site in which LABs were present to depths of greater than 5 to 6 cm (77). Because the sedi ment accumulation rate for the continental slope is reported to range from 0.7 cm/100 yr. to 2.2 cm/100 yr. (39), and the sewage disposal to this site began about 4 years before the sampling, deeper penetration suggests rapid sediment mixing in the dump site. As noted above, the sewage sludge dilution factors are estimated at between 1:800 and 1:10000 based on LABs data and 1:3200 to 1:32000 based on coprostanol data. The differences are probably due to lower stability of coprostanol as compared to the LABs. 6

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Contribution of Sewage Sludge PCBs to Dump Site Sediments. We summarize here a more detailed presentation of our PCB data for the 1989 and 1990 samples from DWDS-106 (Takada et al., in preparation). ZPCB concentrations were 11 to 21 ng/g dry sediment for the dump site, similar to concentrations reported for sediments sampled in the DWDS-106 area in 1992 (40). These values are signifi cantly higher than the value of 3 ng/g for the control site, suggesting that the dump site sediments were contaminated by PCBs derived from ocean dumping or other processes. There is a good correlation (r^O.93) between ZLABs and ZPCBs with a positive intercept (Figure 8a). A similar correlation was observed for sediment sam-



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ILABs(ng/g)

ILAB(ng/g)

Figure 8. Relationships of PCBs with ILABs in the deep-sea sediments, (a): IPCBs, (b): CB101 (filled symbols) and CB201 (open symbols). Circle sym bols: the dump site; square symbols: the control site. The solid and broken lines represent least squares linear regressions for the dumpsite data. The equations and the regression coefficients are as follows, IPCB = 0.056ZLAB + 10.0 (1^=0.93), CB101 = 0.0022ILAB + 0.118 (^=0.97), CB201 = -0.000852LAB + 1.164 (^=0.14). *CB101 may also contain CB90.



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pies collected from the same area in 1992, although the slope for the correlation is different (40). Figure 8a indicates that PCBs are contributed by sludge dumping but that additional sources must exist as well. The positive intercept, ~10 ng/g, is a somewhat higher concentration than PCB concentrations at the control site (-3 ng/g) but not much different than the 5.1 and 7.5 ng/g dry sediment reported for the 1992 control/reference stations (40). Both the control site and the dump site are approxi mately at the same water depth and distance from shore and would be expected to receive similar inputs of PCBs from atmospheric and non-point coastal sources. A possible additional source of PCBs at DWDS 106 could be the chemical waste dumping which occurred prior to sewage sludge disposal (31). This may be consis tent with the unusual PCB congener distribution at the dump site. Compared to the control site sediment and the sewage sludge, higher chlorinated congeners (especially octa- and nona-chlorinated biphenyls) are enriched in the dump site sediments as seen in Figure 9. This difference in relative distribution of PCB congeners suggests dif ferent sources. Within the dump site we observed different spatial distributions of PCBs. The stations which are more heavily affected by sludge show a higher con centration of LABs and higher proportion of lower-chlorinated biphenyls relative to the less-affected dump site stations. Lower-chlorinated PCB congeners were abun dant in the sewage sludge samples we analyzed as well as the samples analyzed by Hunt et al (28). These data are consistent with the scenario that dump site sediments initially acquired higher-chlorinated biphenyls from an independent source, followed by sewage sludge-associated PCBs rich in lower-chlorinated congeners. The relative contribution of sludge-derived PCBs to the dump site sedimentary environment can be estimated by two approaches (Estimation-1 and Estimation-2). Estimation-1 is based on the correlation between LABs and PCBs shown in Figure 8. Assuming a constant contribution rate for non-sludge-derived PCBs, the increment in PCB concentrations in the highest-sludge-impacted station vs. y-intercept in the LABs-PCBs diagrams can be considered as sludge-derived PCBs. For ZPCBs, the contribution of sludge-derived PCBs is estimated at ~ 50% (Figure 8a). The greater contribution is calculated for the lower-chlorinated congeners (e.g. - 80 % for CB101) and lesser contribution of the higher-chlorinated congeners (e.g. no contri bution for CB201), as shown in Figure 8b. Table II summarizes the estimate of sludge-derived PCBs (as a percentage of the total) for selected congeners. Table II. Percent contribution of sewage sludge PCBs to . the dump site sediment* Degree of Estimation-2 Estimation-1 Chlorination 5 CB101 26 ±13 80 6 CB138 80 14± 8 7 CB180 30 8± 4 8 CB201 0 2± 1 9 CB206 0 0 For a station (Dive#2165) where is most heavily affected by sewage sludge sludge-PCBs is estimated as increment in PCBs concentrations vs. y-intercept in the LABs-PCBs diagrams. sludge-PCBs is estimated multiplying PCB/LAB ratios for sludge by the LAB concentrations in the sediment sample. CB90 may also be contained. CB163 and CB164 may also be contained (41, 42). b

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Another approach (Estimation-2) uses LAB data and the ratio of PCBs-to-LABs in sewage sludge samples to estimate the sludge contributions of PCBs to the dump site sediments. Averaged PCBs/LABs ratios (CB101/2LAB=0.74 10 , CB138/ZLAB=0.60x KT, CB180/ZLAB=0.36 x 10' , CB201/ZLAB=0.10 x 10" ; CB206/£LAB=0; Takada et al. in preparation) for sludge are multiplied by the LAB concentrations in dump site sediments. The estimates are summarized in Table II. We find greater contributions of the lower-chlorinated congeners and minimal con tribution of the higher-chlorinated congeners. However, the estimates of sludgecontribution between two approaches are significantly different (i.e. higher contribu tions are estimated by Estimation-1 than Estimation-2). Several factors may cause this difference. We have individual chlorobiphenyl congener data available for only the two sludge samples. Hunt et al. (28) in their analyses of six sludge samples re ported total PCB concentrations, not individual congener data, in their analyses of seven sludge samples. Uncertainty is introduced by the small number of sludge sam ples analyzed and the potential for varying compositions among sludge samples for mixtures of congeners. The assumption of constant background PCBs in Estimation1 may not be valid, and the higher sludge-impacted stations may have received greater contributions of non-sludge (i.e. background) PCBs. This may lead to an overestimation for Estimation-1 using the LAB-PCB correlation. Also, the calculation using PCB/LAB ratio may be underestimated if preferential loss of LABs relative to PCBs occurs. Lamoureux and co-authors (40) observed significantly lower concentrations of LABs (more than a factor of 3) in sediment samples collected in the same area in 1992 compared to our 1989-90 samples. They proposed a hypothesis that LABs are selectively assimilated by deposit-feeders relative to inorganic markers (i.e. Ag). If selective processing occurs between LABs and PCBs, it would lead the Estimation-2 approach to underestimate sewage sludge contributions of PCBs to sediments. Phase associations of LABs and other molecular markers in sludge and their compound-specific environmental behaviors should be examined in future studies. Our studies of the molecular markers as tracers of sludge particles at DWDS-106, and those of others (28, 40), are hampered by the fact that the compliance monitoring analyses of sludge samples had relatively high detection limits. Thus, we have limited data for concentrations of PCBs, PAHs and LABs in the sludge material prior to dumping. Furthermore, we agree with Hunt et al. (28), that studies of the particle size distributions within the sludge and the analyses of molecular markers within size classes are important for future studies of this type. Nevertheless, the molecular markers studies that have been conducted at the DWDS-106 site have provided the basis for longer term studies of biogeochemical processes in the deep sea such $s biological and physical mixing of the hydrophobic sludge organic chemicals into the sediments and the resuspension, transport and dispersion of these chemicals in the deep sea epibenthic and benthic environments. x


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Conclusions and Future Directions for Research One important application of anthropogenic markers during the past two decades has been the identification of sources of pollutants and their fate in coastal and open ocean sediments. Estimates of the contribution made by a specific source is important, espe cially in cases where contaminants are derived from multiple sources, because this information can be used in decisions about where to concentrate management efforts for controlling sources of significant input. Analyses of multiple molecular markers have great promise in sorting out contributions from multiple sources. The recent ad dition of analytical capabilities to measure ratios of stable isotopes of carbon and carbon-14 content of individual molecular markers greatiy enhances the discriminatory power of molecular markers as source identifiers (43, 44) and for revealing biogeo-



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chemical processes acting on contaminants released to all environments, including the ocean. Even as discharges of several chemicals of environmental concern are being re duced or eliminated in many areas of the world, we are challenged by the legacy of past human activities. Large amounts of chemicals of environmental concern have accumulated for decades in coastal sediments which are a potential source for con tinuing contamination of the benthic ecosystems and overlying water after reduction or elimination of discharges of these chemicals. Molecular markers can be used to understand sedimentary processes such as resuspension, deposition, burial, biologi cal mixing, and bioavailability thereby contributing to the knowledge base needed to reduce or eliminate the pollution threat associated with these sediments. Urban sewage discharges, urban runoff, agricultural runoff or atmospheric transport to the ocean continue to be significant sources of chemicals which are of environmental concern either because of potential human health impacts as a result of transfer through the food web back to humans, or because of potential adverse im pacts on valuable living natural resources. These inputs may increase because of human population growth and development in coastal areas around the world. Studies to date have provided valuable guidance for management of contaminant and pollut ant inputs. Studies of a wider variety of molecular markers specific to sewage dis charges, urban runoff, agriculture related runoff, and atmospheric inputs in varying ocean regimes, particularly different types of coastal ecosystems, are urgently needed to improve our quantitative knowledge of how biogeochemical processes influence the fate of chemicals of environmental concern. As demonstrated by the research re ported in the references cited and discussed in this paper, sufficient documentation exists for using several molecular markers in monitoring programs to meet environ mental management needs. APPENDIX. Structures and Sources of the Anthropogenic Molecular Markers Natural Products

5p-cholestan-3p-ol (coprostanol) •Human feces •Animal feces •In-situ reduction of cholesterol

a-Tocopheryl acetate (cc-TA) •chiral carbon •Industrially-synthesized Vitamin E

(IV)

Caffeine •Urine from coffee-drinker

Urobilin

(V)

•Human urine

Aminopropanone •Human urine


12.

193


TAKADA ET AL.

Synthetic products (VII)

(VI) 2n+1

H

1

n+m = 9 - 1 3

^m 2m+1 H

Linear alkylbenzenes (LABs) •Anionic surfactants

^C H 12

25

Tetrapropylene-based alkylbenzenes (TABs) •Anionic surfactants

(VIII) Cn 2n+i C H H

m

2 m +

i NCH

(IX)

n = 1,14,16or18 m = 14,16, or 18


OH

3

C9H19'

Trialkylamines (TAMs)

Nonylphenols (NPs)

•Cationic surfactants

•Nonionic surfactants

(X)

(XI)

so 3

nH2n+i

;

'C H 5

n 2m+l

12

H

n+m = 9 - 1 3

Linear alkylbenzenesulfonates (LAS)

2

Tetrapropylene-based alkylbenzenesulfonates (ABS) •Anionic surfactants

•Anionic surfactants

(XII)

4,4'-bis(2-sulfostyryl)biphenyl (DSBP) •Fluorescent Whitening Agents (FWAs)

(XIV)

OK) 2-(4-morpholinyl)benzothiazole (24MoBT) •Tire rubber debris


194


Acknowledgments We appreciate the stimulating discussions with Drs. Robert P. Eganhouse and Paul Sherblom about molecular markers. We thank Dr. J. Frederick Grassle for organiz ing the research project of DWDS 106; Mr. Hovey Clifford and Ms. Rose Petracca for collecting samples; Dr. Damian Shea for kindly supplying the sludge samples; Dr. Susan McGroddy for stimulating discussions and cooperation in the laboratory. The officers and crews of RV Oceanus, RV Atlantis II, and DSRV Alvin provided es sential sampling assistance. Financial support was provided by U.S. NOAA Na tional Undersea Research Program and by Ministry of Education of Japan. This is Contribution Number 9404 of Woods Hole Oceanographic Institution.


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Anthropogenic Molecular Markers: Tools To Identify the Sources and

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