Environ. Sci. Technol. 2004, 38, 194-201
Dissipation of Fragrance Materials in Sludge-Amended Soils ANGELA M. DIFRANCESCO,† P E I C . C H I U , * ,† L A U R E L J . S T A N D L E Y , ‡ HERBERT E. ALLEN,† AND DANIEL T. SALVITO§ Department of Civil and Environmental Engineering, University of Delaware, Newark, Delaware 19716, Stroud Water Research Center, 970 Spencer Road, Avondale, Pennsylvania 19311-9514, and Research Institute for Fragrance Materials, Inc., 50 Tice Boulevard, Woodcliff Lake, New Jersey 07677
A possible removal mechanism for fragrance materials (FMs) in wastewater is adsorption to sludge, and sludge application to land may be a route through which FMs are released to the soil environment. However, little is known about the concentrations and fate of FMs in soil receiving sludge application. This study was conducted to better understand the dissipation of FMs in sludgeamended soils. We first determined the spiking and extraction efficiencies for 22 FMs in soil and leachate samples. Nine FMs were detected in digested sludges from two wastewater treatment plants in Delaware using these methods. We conducted a 1-year die-away experiment which involved four different soils amended with sludge, with and without spiking of the 22 FMs. The initial dissipation of FMs in all spiked trays was rapid, and only seven FMs remained at concentrations above the quantification limits after 3 months: AHTN, HHCB, musk ketone, musk xylene, acetyl cedrene, OTNE, and DPMI. After 1 year, the only FMs remaining in all spiked trays were musk ketone and AHTN. DPMI was the only FM that leached significantly from the spiked trays, and no FMs were detected in leachate from any unspiked tray. While soil organic matter content affected the dissipation rate in general, different mechanisms (volatilization, transformation, leaching) appeared to be important for different FMs.
Introduction Fragrance materials (FMs) are a group of over 3000 structurally diverse compounds (1) that are widely used in consumer products, such as laundry detergent, soap, and shampoo. Most FMs are used at low concentrations and have global industry volumes less than 1 metric ton per year (t/yr), while a small number of FMs have production volume exceeding 3000 t/yr (2, 3). Through down-the-drain disposal of consumer products, FMs are discharged into sewage and enter municipal wastewater treatment plants (WWTPs). In recent years, there has been growing concern about the potential impact of organic micropollutants, such as pharmaceuticals and personal care products, in the environ* Corresponding author phone: (302)831-3104; fax: (302)831-3640; e-mail:
[email protected]. † University of Delaware. ‡ Stroud Water Research Center. § Research Institute for Fragrance Materials, Inc. 194
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ment (4), and WWTPs effluent is considered the primary source of input. Several FMs, particularly polycyclic and nitro musks and their metabolites, have been detected in surface waters downstream of WWTP discharge points as well as in aquatic environments and aquatic organisms. Standley et al. (5) used the polycyclic musks AHTN (7-acetyl-1,1,3,4,4,6hexamethyl-1,2,3,4-tetrahydronaphthalene) and HHCB (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-γ-2benzopyran) as molecular tracers to track the transport of WWTP natural organic matter in surface waters receiving WWTP discharge within 19 watersheds in New York. Bester et al. (6) measured AHTN and HHCB in areas in and around the North Sea in 1990 and 1995. FMs were also found in aquatic organisms, including lobster, eel, flounder, trout, cod, thornback ray, haddock, clams, and mussels (7-13). In addition, FMs have been detected in human blood, adipose tissue, and breast milk (12, 13). However, these FMs are believed to be absorbed through the skin as a result of dermal application of products containing the FMs (14). Several studies have investigated the removal of FMs within WWTPs. Gatermann et al. (15) detected musk xylene (1-(1,1-dimethylethyl)-3,5-dimethyl-2,4,6-trinitro-benzene) and musk ketone (3,5-dinitro-2,6-dimethyl-4-tertbutyl-acetophenone) along with their degradation (amino) products in the influent and effluent of a WWTP in Hamburg, Germany. Butte et al. (16) showed that several nitro musks, including musk ketone and musk xylene, can degrade photochemically. Herren et al. (17) and Berset et al. (18) also detected the monoamino metabolites of musk ketone and musk xylene in municipal WWTP biosolids following anaerobic digestion. Simonich et al. (3) measured 16 FMs in the influent, primary effluent, and secondary effluent of an activated sludge and a trickling filter WWTP in the United States. The overall removal efficiency varied from 92% to 100% in the activated sludge plant and from 81% to 99% in the trickling filter treatment plant. The study was followed up by another study that compared the removal of FMs from wastewater in six types of treatment plants in the United States and Europe (19), including activated sludge, trickling filter, oxidation ditch, lagoon, carousel, and rotating biological contactor. Although the overall removal efficiencies for FMs in WWTPs are generally high, biological transformation may not be the dominant removal mechanism for FMs (4, 15, 19). In the studies by Simonich et al. (3, 19) and Gatermann et al. (15), adsorption of FMs to activated sludge was considered a removal mechanism, and both groups acknowledged the importance of adsorption. Adsorption to the sludge may be a particularly important mechanism for hydrophobic FMs. Simonich et al. (19) indicated that the removal of adsorptive, nonbiodegradable FMs was correlated to the removal of suspended solids from the treated effluent. Following wastewater treatment, sludge is digested and dewatered and the resulting sludge (or biosolids) is commonly applied to agricultural land. If the adsorbed FMs are not degraded during sludge digestion, land application of biosolids may be a major route through which the residual FMs enter the terrestrial ecosystem. Land application has been a concern due to the presence of heavy metals and organic pollutants in biosolids (20). If the residual FMs dissipate slowly in soil and their concentrations in soil build up over repeated sludge applications, then the toxicity of FMs to soil organisms as well as their potential entry into the food chain through uptake by soil organisms and plants will need to be considered. To date, data on FM concentrations in digested sludge and sludge-amended soil are scarce and the rate at 10.1021/es034618v CCC: $27.50
2004 American Chemical Society Published on Web 11/26/2003
which FMs dissipate in soil following sludge application remains unknown. In addition, the fate processes that contribute to the removal of FMs in soil are unclear. Leaching is a process of particular interest since it is a potential path through which FMs may enter surface runoff and groundwater. With limited information, it is difficult to assess the level of exposure of soil organisms to FMs and to evaluate the importance of sludge application as an exposure pathway. Tas et al. (21) conducted a risk assessment study for two nitro musks in surface water and soil and concluded that the reliability of predicted soil concentrations can be greatly improved by obtaining experimental data on fate and behavior of these FMs in digested sludge and soil. The main objective of this study was to better understand the dissipation behavior of FMs in sludge-amended soils and the fate processes involved. This was accomplished by conducting a 1-year dissipation study using four different soils amended with digested sludge with and without spiking of 22 FMs.
Materials and Methods Chemicals and Materials. The 22 FMs selected for this study were research-grade materials supplied by Quest International (Mount Olive, NJ), International Flavors and Fragrances (Union Beach, NJ), PWF Aroma Chemicals (Milford, PA), Givaudan Roure (Vernier, Switzerland), Haarmann & Reimer (Teterboro, NJ), and Firmenich (Princeton, NJ): acetyl cedrene (cedar ketone), AHTN (>98%), benzyl acetate (phenylmethyl ester acetic acid, >98%), p-tert-bucinal (4(1,1-dimethylethyl)-R-methylbenzenepropanal, >98.4%), diphenyl ether (99.9%), DPMI (Cashmeran, 6,7-dihydro1,1,2,3,3-pentamethyl-4(5H)-indanone, 99%), eugenol (2methoxy-4-(2-propenyl) phenol, 99.9%), hexyl salicylate (2hydroxyhexyl ester benzoic acid), R-hexylcinnamic aldehyde (2-(phenylmethylene)octanal, 98.4%), HHCB, isobornyl acetate (1,7,7-trimethylacetate bicyclo[2.2.1]heptan-2-ol, 94.9%), D-limonene (dipentene, 99%), linalool (3,7-dimethyl-1,6oceadien-3-ol, 97.7%), methyl dihydrojasmonate (3-oxo-2pentylmethyl ester cyclopentane acetic acid, >96%), γ-methyl ionone (3-methyl-4-(2,6,6-trimethyl-2-cyclohexen-1-yl)-3buten-2-one, 98%), methyl salicylate (2-hydroxymethyl ester benzoic acid), musk ketone, musk xylene, OTNE (7-acetyl1,2,3,4,5,6,7,8-octahydro-1,1,6,7-tetramethyl naphthalene, 97%), phenethyl alcohol (2-phenyl ethanol), β-pinene (6,6dimethyl-2-methylene bicyclo[3,1,1]heptane, 96.7%), and terpineol (4-trimethyl-3-cyclohexene-1-methanol). These FMs were selected to cover a broad range of physicalchemical properties. The structures and properties of the FMs are provided in the Supporting Information (Figure S1 and Table S1). As shown in Table S1, the octanol-water partitioning coefficient (Kow), vapor pressure (P°), and water solubility (Cw) of the 22 FMs vary over 4, 7, and 4 orders of magnitude, respectively. Spiking cocktails were made by dissolving weighed amounts of the FMs in ethanol (99.5% Aldrich, WI). The cocktails were stored at 5 °C in clear glass bottles capped with Teflon-lined closures and sealed with low-permeability vinyl tape (3M, MN). FM standard solutions for gas chromatography-mass spectrometry (GC/MS) calibration and the internal standard solution were prepared by dissolving a preweighed amount of each chemical in methanol. The standard stock solution was stored in amber borosilicate vials with Teflon-lined lids, sealed with Teflon tape, and stored at -30 °C. Calibration standards were made through dilution of the stock solution using methanol and were stored in a similar fashion. The internal standard solution contained terpineol-d3 and phenanthrene-d10, both obtained from Cambridge Isotope Laboratories, Inc. (MA). Care was taken throughout the study to minimize contamination from external sources and to minimize the
loss of FMs through sorption. Laboratory personnel were required to use fragrance-free detergent and hand soap and to abstain from the use of perfume or cologne. All materials that came into contact with the sludge, soil, and leachate samples during all stages of the experiment and analysis were glass, metal, or Teflon. The die-away study was set up using four soils chosen to give a range of texture and organic matter content: a sandy agricultural soil from Georgetown, DE, a silty midwestern agricultural soil from Illinois, a clayey, high-organic carbon soil from Newark, DE, and a highly weathered, oxide-rich soil from Aiken, SC. The organic matter content for the Georgetown, midwestern, Newark, and southern soils was 1.55%, 2.63%, 7.01%, and 0.61%, respectively, as determined by chromic acid oxidation (i.e., the conventional Walkley Black method) at the University of Delaware’s Soil Testing Laboratory. The midwestern soil was steam-sterilized 1 year prior to use in the experiment. Anaerobically digested and dewatered sludge was obtained from two activated sludge plants: Georgetown WWTP (100% domestic, approximately 10% solid) and Wilmington WWTP (70% domestic, approximately 17% solid), both in Delaware. Sludge was dewatered using a gravity separator at Georgetown WWTP and a belt filter press at Wilmington WWTP, and samples were collected fresh from the conveyer belts. Both WWTPs added polymers to assist in sludge dewatering, and the Georgetown facility also used aluminum chloride for phosphorus removal. A total of 20 trays were set up. These include 16 trays containing the four soils amended with either raw (unspiked) or FM-spiked sludge from the Georgetown or Wilmington WWTP plus four duplicate trays: the Georgetown soil with raw Georgetown sludge, the Newark soil with spiked Georgetown sludge, the Midwest soil with spiked Wilmington sludge, and southern soil with raw Wilmington sludge. Each tray was given a name describing the sludge-soil mixture, and the naming convention is as follows. The first letter denotes the sludge type (i.e., G for Georgetown sludge and W for Wilmington), the second letter denotes the soil type (G for Georgetown, M for midwestern, N for Newark, and S for southern), and the third letter indicates whether the tray was spiked (S for spiked, U for unspiked). The name is followed by I or II if the tray is a duplicate. For example, GGUII is the second tray of a duplicate pair containing unspiked Georgetown sludge and Georgetown soil. Tray Setup and Sludge Spiking. Each tray contained 24 L of soil and 1 L of sludge. This ratio was chosen based on an average sludge application rate of 7000 wet gallons per acre in Georgetown, DE, in 1999 and a 6-in. (15-cm) plow depth. The trays were stainless steel with dimensions 0.53 m (L) × 0.32 m (W) × 0.15 m (H). The trays were exposed to environmental weather conditions on the roof of DuPont Hall at the University of Delaware (UD). A wooden frame constructed in-house was used to support the trays and also to lift the trays to allow leachate collection. A photograph of the trays is shown in Figure S2(a). For trays with spiked sludge, 1 L of sludge was spiked with a cocktail of the 22 FMs using the wall-coating method (22, 23). A known amount of fragrance cocktail in ethanol was transferred to a 4-L jar. The amount of cocktail was chosen to give a target initial FM concentration approximately 10 times the concentration of HHCB in the July 2000 Georgetown sludge sample (see Results and Discussion), based on an assumed 50% spiking efficiency. The ethanol was allowed to evaporate while the jar was rolled to establish a uniform coating of the FMs on the inner walls of the jar. Upon complete evaporation of ethanol, 1 L of sludge was added to the jar and the Teflon-lined jar lid was sealed with low-permeability vinyl tape. Each jar was rolled at 4 rpm for at least 1 hour on a custom-built rolling apparatus assembled VOL. 38, NO. 1, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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on site (Figure S2(b)). The wall coating efficiency, wall-tosludge transfer efficiency, and overall sludge spiking efficiency of the 22 FMs were determined as described in the Supporting Information. Due to the large volume of soil involved, mixing of soil and sludge was done in a cement mixer (Gilson, IL) with a capacity of approximately 0.12 m3 (Figure S2(c)). The soil and sludge were mixed for 30 min. Care was taken to break up aggregates in the soil and clumps of sludge to ensure mixture uniformity. A time zero sample was taken immediately after the soil-sludge mixture was transferred into the tray. Following transfer of the mixture to the tray, the cement mixer was washed with fragrance-free cleanser and allowed to dry before the next mixing. All unspiked trays were prepared before the spiked trays to avoid cross contamination of the FMs. The drain for leachate collection was assembled before the mixture was added to each tray. It consisted of a Teflon bulkhead connector (Cole Parmer, IL) affixed through a 1.3cm drain hole with a Teflon washer on the inside and outside of the tray. One end of a 0.5-m long Teflon tubing was attached to the bulkhead connector and the other end inserted into a 4-L glass jar (Fisher, PA) through a Teflon-lined lid. Glass wool was placed over the drain hole to filter out small particles and prevent clogging. Sample Collection. Soil samples were taken from the trays at predetermined times over 1 year, based on the assumption of annual land application of sludge. Soil cores were taken from each tray at 1 and 2 weeks and 1, 3, 6, and 12 months, using a brass soil corer (0.84 cm × 14 cm). Immediately following sampling, the hole was plugged with a glass rod of the same diameter to minimize soil disturbance. Each core was stored in an amber glass bottle and kept at -10 °C until extraction. Leachate samples were collected at the end of each rain event for the first 3-5 months of the dissipation experiment. A few early leachate samples, collected prior to September 1, 2000, contained 0.1 g of Tenax TA (Supelco, MO) to adsorb leached FMs and prevent volatilization. This was discontinued because Tenax TA was found to be soluble in dichloromethane (see Analysis section). Following a rain event, the leachate sample was removed from the roof and a formaldehyde solution in methanol (37%, Fisher, PA) was added to the leachate (3%, v/v) to minimize biological activity during storage. All leachate samples were stored at 5 °C in glass jars with a Teflon-lined lid for less than 2 weeks before extraction. When the volume of leachate produced from a tray exceeded 4 L during periods of extensive rain, a 10% subsample of leachate was taken from each tray and subsamples from the same tray were composited for extraction. Analysis. Soil cores and sludge samples were extracted by accelerated solvent extraction (ASE) with dichloromethane (DCM, HPLC grade, Burdick & Jackson, Muskegon, MI) based on the method reported by Simonich et al. (3) with some modifications. An ASE 200 Extraction System (Dionex, CA) was used. The total (wet) weight of each sample was between 7 and 10 g. The sample was mixed thoroughly using a steel spatula, and a subsample of approximately 2 g was extracted. The water content of each sample was determined by drying a different subsample at 105 °C overnight (15 h). Depending on its water content, the sample was mixed with 2-4 g of hydromatrix (Varian, CA) to create a free-flowing mixture before extraction. Thirty-three-milliliter extraction cells (Dionex) were used, with 3 g of silica gel (activated for 15 h at 105 °C) added to the cell as a cleanup step. Following sample addition to the extraction cell, the surface of the sample-hydromatrix mixture was spiked with 21 µL of an internal standard solution containing phenanthrene-d10 and terpineol-d3 (48 ng/µL). The remaining volume of the cell 196
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was filled with additional hydromatrix, and the content was packed tightly. Samples were extracted at 60 °C and 2000 psi (138 bar) in static mode for 15 min two times. The final volume of DCM extract was 60 mL. The DCM extract was solvent-exchanged to n-hexane (HPLC grade, Burdick & Jackson, Muskegon, MI) by cold rotary evaporation, and the sample was run through a sodium sulfate column (precleaned with methanol and a 1:1 n-hexane-acetone mixture) to remove water. The column was rinsed with approximately 1 mL of n-hexane after the addition of the sample. All of the n-hexane that eluted from the column was collected in a new vial. The sample was concentrated to 0.5 mL with a gentle stream of nitrogen and cleaned using a silica gel column. Following sample addition, FMs were eluted from the column with 2 mL of n-hexane and 5 mL of DCM. The sample extract was concentrated with nitrogen to a final volume of 50-200 µL, depending on whether the sample was spiked or unspiked. The solid extraction efficiency for the 22 FMs using this procedure was determined by spiking a known amount of FMs to triplicate sludge-amended soil samples immediately followed by extraction. Additional information about extraction efficiency for soil can be found in the Supporting Information, and the results are shown in Table S2. Leachate samples were spiked with 5 mL of methanol per liter of leachate and the deuterated standards prior to extraction. Leachate was extracted using a C-18 speed disk (Baker-bond, NJ) (3). Prior to extraction, the speed disk was conditioned with 20 mL each of methanol, DCM, and Nanopure water. The sample was added to the disk, and the disk was allowed to run dry. The FMs were then eluted from the speed disk using 10 mL of acetone and 20 mL of DCM. The liquid extraction efficiency using this procedure was determined by extracting triplicate FM-spiked Nanopure water samples, and the results are shown in Table S2. Samples that contained Tenax TA were extracted using the method described earlier along with the solvent exchange procedure below to remove Tenax TA from the extract. This was necessary due to the fact that Tenax TA is soluble in DCM. The solvent was exchanged to acetone, in which Tenax TA is not soluble, and then exchanged back to DCM following the removal of all particulate Tenax TA. All samples were run through a sodium sulfate column prior to extraction. All extracts of solid and liquid samples were analyzed by GC/MS using selected ion monitoring. With the exception of some sludge samples, GC/MS analysis of all samples was performed at Stroud Water Research Center (SWRC) using an Agilent 6890 GC-5973 MS equipped with a J&W DB-1701 capillary column (30 m, 0.25 mm i.d., 0.25 µm film thickness) and an Agilent 7683 autosampler. Each FM was identified and quantified based on a quantification ion and one or two conformation ions. The quantification and confirmation ions used for all the FMs and internal standards are given in Table S3. The ionization mode was electron impact (70 eV), the ion source temperature was 250 °C, and the dwell time was 50 ms/ion. Only data with all matching confirmation ions and a signal-to-noise ratio greater than 5 (corresponding to a detection limit of approximately 1 ng/µL) were accepted. The GC temperature program used was as follows: isothermal at 35 °C for 2 min, 5 °C/min to 200 °C, 20 °C/min to 285 °C, and isothermal at 285 °C for 5 min. Analysis at UD was preformed using an HP5890 GC coupled with an HP5970 MSD and a modified temperature program: isothermal at 35 °C for 2 min, 15 °C/min to 135 °C, held at 135 °C for 1 min, 8 °C/min to 245 °C, and isothermal at 245 °C for 1 min.
Results and Discussion Extraction and Spiking Efficiencies. Results of the method validation tests indicate that the spiking and extraction methods are not appropriate for all the FMs. As shown in
TABLE 1. Background Concentrations of FMs in Two Digested and Dewatered Sludges Wilmington sludge (µg/dry g) fragrance material acetyl cedrene AHTN diphenyl ether hexyl salicylate R-hexylcinnamic aldehyde HHCB γ-methyl ionone musk ketone musk xylene OTNE a
d
b
July 2000a
Oct 2002b
17
(n ) 7)c 9.0 ( 1.6 8.1 ( 1.6 99.6 ( 55.1 d 21.8 ( 4.3 1.1 ( 0.2
43
7.3 ( 1.4
Georgetown sludge (µg/dry g) July 2000a
51 d 86
March 2002b (n ) 3)c 31.3 ( 5.1 17.7 ( 2.2 1.5 ( 0.2 4.1 ( 1.5 37.6 ( 4.5 3.8 ( 0.5 1.3 ( 1.0 30.7 ( 3.7
c
Samples extracted at SWRC. Samples extracted at UD. Number of replicate samples collected and analyzed (individually, not composited). Trace amount detected (not quantified).
Table S2, the soil extraction efficiency for the 22 FMs using the ASE procedure ranged from 17% to 123%. The procedure had an efficiency greater than 80% for 11 FMs. However, it was ineffective to recover methyl salicylate, terpineol, Dlimonene, and β-pinene from soil. Second and third extractions of the FM-spiked, sludge-amended soil did not yield noticeable enhancement in recovery for any FM, suggesting that the incomplete recovery was most likely due to losses through processes such as volatilization rather than incomplete extraction. Due to the variable recovery, only results for the FMs with soil recovery greater than 70% will be shown. The liquid extraction efficiency using the procedure described above ranged from 0 to 86%. The method gave an efficiency of approximately 70% or higher for eight FMs but was inadequate to extract benzyl acetate, D-limonene, linalool, methyl salicylate, phenethyl alcohol, β-pinene, and terpineol from water. The low recovery is most likely due to the high volatility and high water solubility of these FMs, as shown in Table S1. Table S2 also gives the wall coating, wall-to-sludge transfer, and overall sludge spiking efficiencies. D-Limonene and β-pinene, the two FMs with the highest vapor pressure, were minimally coated on the walls. Four FMs that could be wallcoated were transferred to the sludge very poorly. The resulting sludge spiking efficiency ranged from 0 for the most volatile FMs to about 90% for AHTN and HHCB and was greater than 50% for 11 of the 22 FMs. Collectively, the method validation results suggest that the spiking and extraction methods used were suitable for approximately one-half of the 22 FMs. FMs in Digested Sludge. Background concentrations of FMs in anaerobically digested sludge from Georgetown and Wilmington municipal WWTPs were determined in July 2000 at SWRC and again in March (Georgetown) and October (Wilmington) 2002 at UD. The concentrations of the detected FMs are shown in Table 1. Analysis in July 2000 showed that there were significant concentrations of HHCB and AHTN in both sludges and traces of diphenyl ether in the Georgetown sludge. None of the other FMs were detected in these sludge samples. The July 2000 sludges were subsequently used to set up the soil dissipation experiment. Analysis of the Georgetown sludge collected in March 2002 and the Wilmington sludge collected in October 2002 resulted in the detection of six and four additional FMs, respectively. Low concentrations of γ-methyl ionone, hexyl salicylate, R-hexylcinnamic aldehyde, and musk ketone were found, whereas the concentrations of acetyl cedrene and OTNE were comparable to that of AHTN and HHCB. Diphenyl ether concentrations in the October 2002 Wilmington sludge was conspicuously high, approximately an order of magnitude greater than the major FMs (AHTN, HHCB, acetyl cedrene,
FIGURE 1. Concentration of FMs in the GGS tray over time. For clarity, only the recoverable FMs that did not completely dissipate within 2 weeks are shown. and OTNE). Because diphenyl ether has nonfragrance applications (e.g., as an industrial solvent) and its volume of use as fragrance in the United States is small (24), the high levels of diphenyl ether found in the Wilmington WWTP sludge was probably derived from industrial sources. The reason for the difference between the 2000 and 2002 samples is unclear but might be due to variations in FM usage, in removal efficiency in the WWTPs (volatilization, aerobic and anaerobic degradation, and sorption to sludge), and in analytical procedures between SWRC and UD. More monitoring of sludge from different regions, seasons, and WWTP types is necessary to better characterize FMs in wastewater sludge. Soil Samples. Because of the large numbers of soil (142) and leachate (135) samples and because the Wilmington WWTP sludge is not land-applied (thus comparison with field samples was not possible), only results from the trays that contained the Georgetown sludge will be shown. Figures 1-4 show the soil concentrations of 10 or 11 FMs in trays containing spiked Georgetown sludge with Georgetown (GGS), midwest (GMS), Newark (GNS), and southern soil (GSS), respectively. The FMs that are not shown either were absent in time zero samples due to low spiking and/or VOL. 38, NO. 1, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 2. Concentration of FMs in the GMS tray over time. Only the recoverable FMs that did not completely dissipate within 2 weeks are shown.
FIGURE 3. Concentration of FMs in the GNS trays over time. Only the recoverable FMs that did not completely dissipate within 2 weeks are shown. extraction efficiencies or dissipated rapidly and were present for only up to 1 week. The latter includes linalool, isobornyl acetate, methyl salicylate, hexyl salicylate, and phenethyl alcohol. The soil concentrations of the shown FMs in all spiked trays decreased through the year. Dissipation was rapid in the first month as the freshly spiked sludge-soil mixtures stabilized and poorly incorporated FM mass was removed. Diphenyl ether, R-hexylcinnamic aldehyde, and γ-methyl ionone continued to dissipate and were below the limit of detection within 3 months. This suggests that these 198
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FIGURE 4. Concentration of FMs in the GSS tray over time. Only the recoverable FMs that did not completely dissipate within 2 weeks are shown. FMs will not persist in soil even though they are present in digested sludge (Table 1). After 3 months, only seven FMs were present in all spiked trays at concentrations significantly above the quantification limits: musk ketone, musk xylene, AHTN, HHCB, acetyl cedrene, OTNE, and DPMI (and methyl dihydrojasmonate was present in GNS only). Dissipation of these FMs was insignificant between 3 and 6 months. This period corresponds to from November 2000 to February 2001, during which the soil was frozen the majority of the time. The FM concentrations decreased further between 6 and 12 months. At the end of the 12 months, the only FMs that remained were the nitro and polycyclic musks. The dissipation of AHTN and musk ketone was particularly slow. The processes that contributed to the observed FM dissipation in soil may include leaching, volatilization, abiotic reactions (e.g., hydrolysis), and biological transformation (aerobic and anaerobic). Although the importance of these fate processes is unknown for most FMs, several inferences can be made based on the dissipation behavior and the physical-chemical properties of the FMs. First, the retention of relatively hydrophobic and nonvolatile FMs in soil appeared to increase with soil organic matter content. For example, comparing the 12-month removal efficiencies of AHTN and musk ketone in the four soils, the Georgetown and southern soils showed a removal of approximately 50% whereas less than 15% was removed for the Newark soil (Figure 5). It should be noted that most of the organic matter was derived from soil rather than sludge, as soil analysis showed that sludge addition did not increase the soil organic matter contents appreciably (by approximately 0.1%). Second, in contrast to the seven more strongly retained FMs mentioned above, R-hexylcinnamic aldehyde, methyl dihydrojasmonate, diphenyl ether, and γ-methyl ionone dissipated almost completely within the first 3 months regardless of the soil organic matter content. Diphenyl ether and γ-methyl ionone have the highest vapor pressures of the 11 FMs (Table S2) and most likely dissipated through volatilization. In contrast, volatilization is probably not an important removal mechanism for methyl dihydrojasmonate and R-hexylcinnamic aldehyde, since these FMs are even less
FIGURE 5. Percent removal of musk ketone and AHTN in four spiked soils after 1 year. The organic matter contents of the soils are given in parentheses. volatile than AHTN and HHCB (Table S2). As will be discussed in the Leachate Samples section, R-hexylcinnamic aldehyde and methyl dihydrojasmonate were not detected in the leachate. It thus appears that biological and/or chemical transformation was the chief dissipation mechanism for these FMs. Indeed, R-hexylcinnamic aldehyde and methyl dihydrojasmonate have been classified as inherently and readily biodegradable, respectively (19), according to the Research Institute for Fragrance Materials database. Third, musk xylene and musk ketone were most likely removed through transformation as well. These nitro musks have the lowest vapor pressures (by orders of magnitude) of all the FMs. They also have very low water solubility and did not leach to any significant extent (see Leachate Samples). Because these nitro musks are not biodegradable aerobically (19, 21), the most plausible removal process was reductive transformation to their amino counterparts, as reported previously (15, 17, 18). The faster reaction of musk xylene than musk ketone (Figures 1-4) may be attributed to the less negative one-electron reduction potential (Eh1′) of musk xylene than musk ketone. Musk xylene is structurally analogous to trinitrotoluene, which has been shown to have a higher Eh1′ value and undergo reduction more rapidly than nitroaromatic compounds containing two nitro functions meta to each other (25), as in the case of musk ketone. The loss of AHTN and HHCB was also monitored in soil trays containing unspiked sludge, as shown in Figures 6 (GGU) and 7 (GNU). Consistent with the result from the spiked trays, HHCB dissipated more rapidly than AHTN. The concentration of HHCB, but not AHTN, decreased below the quantification limit within 12 months in both soils. The concentrations of these FMs (especially AHTN) also decreased more rapidly in the Georgetown soil than in the Newark soil, most likely due to the different organic matter contents. For comparison, we collected soil samples from an agricultural land in Georgetown, DE, before and immediately after application of sludge from the Georgetown WWTP. For each sampling, 20 soil samples from the top 15 cm were collected and composited for FM analysis and moisture measurement. AHTN and HHCB were not detected in the soil prior to sludge application, suggesting that the FM mass from previous applications had dissipated below quantification limits. The concentrations of AHTN and HHCB after sludge application were similar to their initial (time zero) concentrations in GGUI and GGUII trays, indicating that the sludge-to-soil ratio chosen for the soil tray study was representative of the field conditions. Comparing the results from GGU, GNU, and the field, it appears that whether AHTN accumulates over repeated
FIGURE 6. AHTN and HHCB concentration in the GGU trays over time. Concentrations below the quantification limit are included but should only be regarded as estimates.
FIGURE 7. AHTN and HHCB concentration in the GNU tray over time. Concentrations below the quantification limit are included but should only be regarded as estimates. sludge applications may depend on the organic matter content of the soil. The comparison also shows that AHTN dissipated more slowly in the trays than in the field. This is probably because the soil trays were static whereas the field soil was subject to plant, animal, and human activities, which presumably contributed to the volatilization and biodegradation (via aeration) of FMs. Therefore, the dissipation rates observed in the soil trays may be regarded as conservative estimates of the removal rates in the field. Leachate Samples. Leachate samples were collected for the first 3-5 months striving for at least five samples from each tray, as leachate volume and leaching frequency varied greatly between soil types. Up to 12 FMs were detected in leachate collected from the spiked trays, whereas no FMs were found in leachate from any of the unspiked trays. Figure 8 shows the cumulative percent mass of FMs leached, based on the initial soil concentration of each FM, as a function of cumulative volume leached from the GGS tray. The total percent mass leached was small (1% or less) for all the detected FMs except DPMI. This was also true for leachate VOL. 38, NO. 1, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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storms, and hence contained lower concentrations of desorbed FMs. DPMI has not been included in studies that monitored FMs in surface waters (3-5, 7, 19). On the basis of the leachate data, DPMI is expected to remain in the aqueous phase to a large extent and its removal through adsorption to activated sludge may be limited (DPMI was not detected in the two sludges). Therefore, although DPMI is relatively leachable, it is not likely to reach deep soil and groundwater through sludge application to soil.
FIGURE 8. Cumulative percent initial FM mass leached vs cumulative leachate volume in the GGS tray. The data were corrected with the extraction efficiency.
The conspicuously high leachability of DPMI is interesting. Comparing DPMI with other FMs with similar properties, DPMI, diphenyl ether, and musk ketone have similar log Kow values of 4.5, 4.1, and 4.3, respectively. Diphenyl ether has a higher water solubility and higher vapor pressure than DPMI, whereas musk ketone has a lower solubility and lower vapor pressure. As discussed earlier, diphenyl ether is one of the most fast-dissipating FMs and was probably removed through volatilization. In contrast, musk ketone is one of the two most lasting FMs in soil and is not biodegradable (19). Although biodegradation studies have not been performed on DPMI to the authors’ knowledge, if one makes the conservative assumption that DPMI did not undergo transformation, the comparison suggests that musk ketone did not leach because its solubility is too low, whereas diphenyl ether did not leach because it volatilized too quickly. DPMI leached substantially more than the other FMs because it is moderately soluble but not highly volatile or degradable. The limited degradability of DPMI can also be seen by comparing DPMI and OTNE. These two compounds dissipated very similarly in all spiked trays (Figures 1-4), but OTNE has higher log Kow, lower vapor pressure, and lower solubility than DPMI. Thus, abiotic or biological transformation was probably a more important dissipation mechanism for OTNE than for DPMI. The comparison of DPMI, diphenyl ether, and musk ketone also shows that Kow (Kd) and soil organic matter content alone cannot predict FM dissipation rates in soil. That is, the fate of FMs in soil is controlled by processes other than hydrophobic sorption and different fate processes may be important for different FMs. Unlike PCBs and PAHs, FMs are not a series of structural homologues. Rather, they are a much larger collection of compounds with diverse physical-chemical properties and structural features. As a consequence, the distribution and fate of different FMs in the environment may differ significantly. For the same reason, analysis of different FMs in environmental samples will require different analytical methods, as suggested by our method validation result.
FIGURE 9. Cumulative percent initial DPMI mass leached vs cumulative leachate volume in five spiked trays. DPMI concentrations were corrected with the extraction efficiency. from spiked midwest, Newark, and southern soil (data not shown). The cumulative percent mass leached for all the FMs in all the spiked trays was below 3%, except DPMI. Figure 9 shows the cumulative percent mass of DPMI versus cumulative leachate volume from five spiked trays. The percent mass of DPMI leached ranged from 1.4% for the Newark soil trays (GNSI and GNSII) to 26.5% for the southern soil tray (GSS). The amount leached generally increased with decreasing organic matter content of the soil, although slightly more DPMI leached from the GMS tray (2.63% organic content) than from the GGS tray (1.55% organic content). This is due to the lower FM concentration in leachate from the GGS tray than that from the GMS tray. The Georgetown soil is sandy and produced significantly more leachate from each rainstorm than the midwest soil, which is silty. Rainwater in the GGS tray had shorter retention times, especially when leachate was generated continually during consecutive 200
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Acknowledgments The authors thank the Research Institute for Fragrance Materials, Inc., for their support of this study. We also thank Domy Adriano, Peter Vadas, and Roger Bowman for providing the soils and Wilmington and Georgetown wastewater treatment plants for providing the sludge samples. We are grateful to Doug Baker for constructing the wooden frame and rolling apparatus, John Dykin for his assistance with the GC/MS, and Karen Jansson and Shucha Zhang for their effort in sample analyses.
Supporting Information Available Structures and properties of the FMs used, pictures of the soil trays and spiking and mixing devices, procedures and results of extraction and spiking efficiency measurements, and additional information on GC/MS analysis. This material is available free of charge via the Internet at http:// pubs.acs.org.
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Received for review June 17, 2003. Revised manuscript received October 14, 2003. Accepted October 16, 2003. ES034618V
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