Estimation of Water Sampling Rates and Concentrations of PAHs in

Jiries, A.; Hussain, H.; Lintelmann, J. Determination of polycyclic aromatic hydrocarbons in sewage, sediments, sludge and plants in Karak Province, J...
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Environ. Sci. Technol. 2007, 41, 5044-5049

Estimation of Water Sampling Rates and Concentrations of PAHs in a Municipal Sewage Treatment Plant Using SPMDs with Performance Reference Compounds LIJANA AUGULYTE* AND PER-ANDERS BERGQVIST Department of Chemistry, Environmental Chemistry, Umeå University, SE-901 87 Umeå, Sweden

Semipermeable membrane devices (SPMDs) were exposed at ten sampling points, each representing a different stage in the treatment process, in a municipal sewage treatment plant. Differences in SPMD uptake kinetics of polycyclic aromatic hydrocarbons (PAHs) due to variations in conditions at the sampling sites were evaluated by using five performance reference compounds (PRCs) with log Kow values of 4.20 to 6.34. PRC release rate constants (ke,PRC values) were calculated for PRCs for which 5098% of the initial amounts were lost during the sampling period. The ke,PRC values were high, ranging from 0.08 to 0.11 day-1 for the studied PRCs, at sampling site W1 (raw sewage), the only sampling site where significant amounts of the PRCs with log Kow values >5 were released from the SPMDs. At the other sampling sites, only PRCs with log Kow values between 4.20 and 4.50 were released in significant amounts. The release rates at these sites were lowest (0.04 day-1) at sampling site W9 (the secondary clarifier) and highest (0.18 day-1) at W8 (the active sludge aeration basin). Differences between sampling rates (Rs) obtained using published laboratory-calibrated data and PRCcorrected Rs values were visualized by principal component analysis (PCA). The water concentrations of 24 studied PAHs fell substantially during the course of the sewage treatment process. However, low molecular weight PAHs were more effectively removed than high molecular weight PAHs. Significant deviations between actual and estimated water concentrations may arise unless PRCcorrected Rs values are applied.

Introduction In a conventional municipal sewage treatment plant (STP), physical, chemical, and biological treatments are applied to municipal sewage, which mainly consists of domestic and industrial effluents and urban runoff (1). The flow rates of the process streams, the concentrations of various pollutants (including polycyclic aromatic hydrocarbons, PAHs), and the pollution profiles vary widely between STPs and can also fluctuate substantially diurnally (2-4). PAHs comprise a class of compounds with widely varying mutagenic and tumorigenic properties and are found at low levels in the sewage inputs originating from households, industrial sites, urban * Corresponding author phone: +46-90-786-9241; fax: +46-90128-133; e-mail: [email protected]. 5044

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runoff, oil leakage, and other diffuse sources in the environment (5). PAHs can be bioconcentrated by certain invertebrates that lack effective biotransformation systems, but fish and other aquatic vertebrates readily biotransform PAHs (6). Recently, PAHs have been included in the Water Framework Directive (2000/60/EC) in a list of 33 “priority substances” that pose a threat to the environmental quality of surface waters in Europe (7). Integrative sampling methods are more suitable than instantaneous or grab sampling methods for monitoring sewage water because they collect samples over long periods, thus providing data on average pollutant profiles rather than “snapshots” that are heavily biased by the diurnal fluctuations that characteristically occur at STPs. The amount of an analyte extracted by an integrative sampler depends on the analyte concentration in the sewage and the exposure time (8). Thus, using appropriate sampling rate and analyte concentration data, time-weighted average dissolved concentrations of the analyte(s) of interest can be determined. Passive semipermeable membrane devices (SPMDs) have been shown to be suitable tools for sampling sewage process streams in earlier studies (9-11), and they are among the most thoroughly established devices for passively sampling compounds in water. SPMDs have high surface-to-volume ratios and sample between 0.5 and 15 L of water per day. As a result of their time-integrative mode of sampling, they can detect highly hydrophobic compounds, for which sampling rates are high, at low nanogram per liter levels (12) or even lower. Water sampling rates (Rs) of specific analytes by SPMDs depend on a complex set of interacting environmental variables, including (inter alia) temperature, water flow/ turbulence, sorption of the compounds to dissolved organic carbon (DOC), biofouling, photodegradation, and the geometry of the mounting cages (13). For instance, variations in temperature and water flow rates can reportedly cause up to 4-fold to 10-fold differences, due to variations in analyte uptake rates and facial velocity-turbulence effects, especially for compounds with log Kow values >4.5 (14, 15), because their uptake by SPMDs is heavily influenced by the external water boundary layer (WBL). For the in situ calibration and evaluation of site-specific variables, a method involving use of SPMDs containing performance reference compounds (PRCs) has been introduced (13, 14, 16, 17). PRCs are labeled or unlabeled compounds that have moderate to relatively high fugacity, do not interfere with analysis of the target analytes, and are added to the SPMDs before they are deployed. Estimation of water sampling rates from the PRCs’ dissipation parameters provides a means to evaluate the influence of the exposure variables on the uptake kinetics of the analytes by each sampler and thus calculate the sampling rates of the analytes by each SPMD. In this study, the dissipation of PRCs from SPMDs and the uptake kinetics of analytes by SPMDs at each of ten sites in the STP at Umeå, Sweden, were evaluated. PRC-corrected Rs values were calculated using empirical and WBL uptake models and compared with published laboratory-calibrated Rs values. Estimates of PAH concentrations in the sampled process streams were then derived by applying traditional and full water concentration (Cw) estimation models using both laboratory-calibrated and PRC-corrected Rs values. As well as presenting the acquired data, factors influencing the uptake of PAH compounds by the SPMDs in the sewage process streams are discussed. 10.1021/es070054+ CCC: $37.00

 2007 American Chemical Society Published on Web 06/15/2007

Experimental Section Sampling Site. Umeå STP is located on an island in the river Umea¨lven, which receives the treated sewage water. This treatment plant is located indoors, so the sewage treatment processes occur under controlled-temperature conditions in the presence of little or no UV light. Therefore, photodegradation of the photosensitive PAHs was not considered to be a significant factor during the study. The total capacity of the STP was around 116 000 p.e. (population equivalents) during the study period (January 19, 2003, to February 9, 2003) and it received approximately 28 000 m3 day-1 of raw sewage (18). Three types of processes are used to treat the received sewage: mechanical, chemical, and biological. During the course of the treatment, water levels of BOD7 (the biochemical oxygen demand during 7 days), COD (the chemical oxygen demand), phosphorus, nitrogen, and various measured metals (including mercury, cadmium, lead, copper, zinc, chromium, and nickel) all declined by more than 95% (18). The temperature at the sampling sites varied only from 11 °C to 13 °C; therefore, differences in temperature are unlikely to have had a significant influence on the analytes’ uptake. The pH varied among the sites from 7.2 to 9.2. Sampling Method. A standard size SPMD containing PRCs (“ExposMeter lipophilic PRC” obtained from ExposMeter AB, Tavelsjo¨, Sweden) was deployed at each of 10 sampling points (W1-W10) in the STP, representing different stages in the treatment process for a 3-week sampling period. A schematic diagram of the STP showing the sampling sites is presented in Supporting Information Figure S1. Flow conditions and visual observations of the surfaces of the membranes’ retrieved from the STP are described in Supporting Information Table S1. There were indications that the SPMDs had been fouled, presumably due to the complex nature of the pollution entering the STP. Commercially available SPMDs were used in a standard configuration (2.54 cm × 91.4 cm, 75-90 µm wall thickness) with a thin film of 1 mL of triolein (99%) sealed in a low-density polyethylene (LDPE) layflat tube (14). The lipid in the SPMDs was spiked with five PRCs: the PAHs acenaphthene-d10, fluorene-d10, and phenanthrened10 (10 µg mL-1 triolein of each, obtained from Chem Service, West Chester, U.S.A.) and 13C-labeled PCB congeners 37 and 54 (55 ng mL-1 triolein of each, obtained from Wellington Laboratories, Ontario, Canada). The PRCs were used to evaluate the influence of the exposure conditions at the different sampling sites on the uptake of the studied compounds by the SPMDs. Unlabeled octachloronaphthalene (OCN) was also added to the SPMDs. OCN tends to dissipate from SPMDs very slowly; therefore it was used to check whether mechanical damage had occurred to the membranes. PRC-containing SPMDs were mounted on deployment racks and placed in stainless steel canisters to avoid mechanical damage to them and to minimize the influence of variations in flow/turbulence at the membrane surface during the sampling. To monitor contamination of the SPMDs during their deployment/retrieval, transport and to evaluate initial PRC concentrations in the SPMDs, two single SPMDs exposed to the air for the time it took to deploy one of the other SPMDs were used as field controls (FCs). In addition, a laboratory control (LC), exposed solely to the organic solvents and glassware used during the analytical procedures was used to quantify possible contamination in the laboratory. Chemical Analysis of SPMDs. After the 3-week exposure period, the SPMDs were collected and separately stored in tightly closed, acetone-washed tin cans in a freezer awaiting analysis. After washing and dialysis of the SPMDs, the extracts were cleaned by gel permeation chromatography and passage through an open silica column then analyzed by GC/MS as detailed by So¨derstro¨m and Berqvist (19). Labeled PAHs,

naphthalene-d8, anthracene-d10, fluoranthene-d10, and benzo[ghi]perylene-d12 (Cambridge Isotope Laboratories, U.S.A.), were added as internal standards (ISs) to evaluate losses during the SPMD cleanup steps. Dibenzofuran-d8 (Promochem, Kungsbacka, Sweden) was also added as a recovery standard (RS) at the end of the cleanup. Samples were analyzed using a Fisons GC 8000 gas chromatograph with a 30 m × 0.32 mm DB-5 capillary column (0.25 µm film thickness, J&W Scientific, CA) coupled to a Fisons MD 800 mass spectrometer. Target compounds in the extracts were each quantified using a PAH reference standard mixture containing 24 PAHs, including methyl-PAHs (Promochem, Kungsbacka, Sweden), see Supporting Information Table S2. The most abundant ion of each unlabeled compound and labeled standard was quantified in selected ion monitoring mode. Possible contamination during handling was checked by comparing the analytical results obtained from each of the samples and the appropriate FC. The method limit of detection (LOD) for each analyte was calculated as the quantity of the compound giving a response three times higher than the baseline noise of the resultant chromatogram at the expected retention time. Recoveries for the surrogate ISs were calculated and used to adjust the results. The recovery values of all the ISs were in the range 60-95%, except for naphthalene-d8, for which recoveries varied from 34 to 60%, presumably because of its high tendency to evaporate during handling. PRC-Corrected Water Sampling Rates. The amounts of the PRCs released from the SPMDs at each of the sampling sites representing various stages of the treatment process provide information about the influence of exposure conditions during the sampling. The in situ SPMD calibration approach is based on the hypothesis that the rate of PRC losses is proportional to the rate of analyte uptake (14, 20). The dissipation of each PRC at each sampling site can be described by the following equation:

ke,PRC ) -

ln(NPRC/N0,PRC) t

(1)

where ke,PRC is the PRC’s release rate constant (day-1); NPRC is the measured amount of the PRC after the exposure period (ng SPMD-1); N0,PRC is the measured amount of the PRC at the beginning of the exposure period, that is, when t ) 0 (ng SPMD-1); and t is the exposure period (here 21 days). The kinetic PRC sampling rates can be calculated using experimentally determined ke,PRC values and the following equation:

Rs,PRC ) VsKsw,PRCke,PRC

(2)

where Rs,PRC is the PRC sampling rate (L day-1); Vs is the SPMD volume (0.005 L for a standard size SPMD); Ksw,PRC is the PRC’s SPMD-water partitioning coefficient (L‚L-1); and ke,PRC is the PRC’s release rate constant (day-1). Because experimentally determined Ksw,PRC values are not available, we can calculate them from Kow values (octanolwater partitioning coefficients) using the following quadratic equation for nonpolar organic compounds with an intercept (a0) for PAH compounds at -2.61 (13):

log Ksw,PRC ) a0 + 2.321 log Kow - 0.1618(log Kow)2

(3)

PRC-corrected water sampling rates for target analytes (here 24 PAHs) can be calculated using either of two models for estimating Rs values: empirical and WBL-controlled. In the former, empirically determined constants are used (13):

Rs,target ) Rs,PRC

Rtarget RPRC

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where Rs,target is the sampling rate of the target analyte (L day-1); Rs,PRC is the PRC sampling rate (L day-1); Rtarget is the target analyte compound-specific effect; and RPRC is the PRC compound-specific effect, which can be modeled using the following equation (13):

log R ) 0.0130 log Kow3 - 0.3173 log Kow2 + 2.244 log Kow (5) The second model for evaluating PRC-corrected water sampling rates, the WBL-controlled uptake model (13, 15), applies when the exchange kinetics of PRCs and analytes are under WBL control:

Rs,target ) Rs,PRC

( ) VPRC Vtarget

0.39

(6)

where Rs,target is the sampling rate of the target analyte (L day-1); Rs,PRC is the PRC sampling rate (L day-1); VPRC is the PRC’s LeBas volume (cm3 mol-1); and Vtarget is the LeBas volume of the target analyte (cm3 mol-1). Principal Component Analysis (PCA). PCA was used to visualize trends in the sampling rates associated with between-site differences in sampling conditions and between laboratory-calibrated and PRC-corrected Rs values, obtained using the empirical and WBL models. A multivariate data table was used, containing 21 rows (sampling sites) and 24 columns (water sampling rates), for the 24 individual PAHs to reduce the dimensions of the data set (22). Prior to the PCA, the data were scaled to unit variance and meancentered. Each observation was designated according to the sampling site (W1-W10) and the model used to calculate the Rs (cal, e, and w, for laboratory calibration data, empirical, and WBL, respectively). Water Concentrations. Depending on the physicochemical properties of the analytes of interest, the environmental conditions and duration of the sampling, the analytes sequestered by the SPMDs may be in the linear uptake (integrative) sampling, curvilinear, or equilibrium partitioning sampling phases (13). For several decades, linear (eq 7) and equilibrium (eq 8) uptake rate models were used together in combination with laboratory-calibrated Rs and SPMD-water partitioning coefficients (Ksw) for estimating water concentrations (13):

Cw )

NSPMD Rst

(7)

Cw )

NSPMD KswVs

(8)

where Cw is the aqueous concentration of the analyte (ng L-1); NSPMD is the amount of the analyte absorbed by the SPMD (ng SPMD-1); Rs is the water sampling rate of the analyte (L day-1); Ksw is the analyte’s SPMD-water partitioning coefficient (L L-1); and t is the exposure period (days). PAHs with log Kow values < 4.0 were considered to be in the equilibrium uptake phase after the 3-week sampling period. Later, the full water concentration estimation model was introduced, which could be used regardless of whether the analyte’s uptake is in a linear, curvilinear, or equilibrium phase (13, 21):

Cw )

(

NSPMD

VsKsw 1 - exp

( )) -Rst VsKsw

(9)

Compound Ratios (CRs). CRs were calculated to evaluate the proportional changes in concentrations of each PRC 5046

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FIGURE 1. Release ratios (CRs) for the PRCs (a) PRC-corrected sampling rates for the 24 PAHs according to empirical (b) and WBLcontrolled (c) uptake models vs log Kow values at each of the sampling sites (W1-W10). For log Kow values* and laboratory-calibrated water sampling rates** see Supporting Information Table S2. during the treatment processes from initial concentrations assumed to be equal to the mean values obtained from two FCs, which were set to 1.0 for each PRC:

CRWi )

NWi NFC

(10)

where NWi is the measured amount of the PRC at step i of the sewage treatment (i ) 1, 2, ..., 10; ng SPMD-1) and NFC is the mean value of measured PRC amounts from the FCs (ng SPMD-1).

Results and Discussion Evaluation of PRC Release. Dissipation of the five PRC compounds, for which log Kow values ranged from 4.20 (acenaphthene-d10) to 6.34 (13C PCB 54), depended on the exposure conditions in the STP. The amounts of the PRCs released from the SPMDs at sampling site W1 (raw sewage) were extremely high, and their CRs ranged from 0.09 to 0.20 (Figure 1a). At W1, 85% of the initial amount of OCN remained in the SPMD after the exposure period, indicating that little or no mechanical damage had occurred to the SPMD membrane, although fouling could have reduced dissipation

FIGURE 2. PCA score plot for water sampling rates of the 24 PAHs based on the laboratory-calibration data (W-cal; Huckins et al. (23)) and PRC-corrected sampling rates calculated by the empirical (W1-e to W10-e) and WBL (W1-w to W10-w) uptake rate models. of the PRCs from it. Exposure variables that could have influenced sequestration at W1 include flow/turbulence at the membrane surface and (possibly) the formation of a fatty layer from the raw sewage on the surface of the membrane, with consequent reductions in the WBL. In contrast, dissipation of all PRCs was lowest at sampling site W9 (the secondary clarifier), where the CRs for PRCs with both low and high log Kow values changed very little (at most to 0.46 for acenaphthene-d10). Slow sewage water flow and fouling of the membrane surface could have reduced dissipation of the PRCs and uptake of the studied PAHs by the SPMD at this sampling site. To obtain PRC-corrected water sampling rates using the PRC release data, we chose to use PRCs for which losses were in the range g50 to 98% or 0.02 < CR e 0.5 (Figure 1a), to avoid including potential “burst effects” during the release of PRCs from the SPMDs but still have sufficient amounts to quantify validly by the analytical method. At W1 the CRs of all PRCs were in the selected range, but only data for acenaphthene-d10, fluorene-d10, and phenanthrene-d10 were used in the calculations, because the CRs of the PCB PRCs were in the same range as those of the PAH PRCs and thus would have given similar release rate (ke,PRC) values. For the other sampling sites, the following PRCs were selected: acenaphthene-d10 for W2 and W9; acenaphthene-d10, fluorene-d10, and phenanthrene-d10 for W3, W4, W8, and W10 and acenaphthene-d10 and fluorene-d10 for W5, W6, and W7. When several PRCs were selected for the same sampling site, PRC-corrected Rs values were calculated using data for PAHs with the most similar log Kow values. PRC Release Rate Constants. PRC release rate constants (ke,PRC) at each of the sampling sites were calculated using eq 1 and are presented in Supporting Information Table S3. The ke,PRC values at all sampling sites were lower than 0.2 day-1. The PRCs proved to be sensitive indicators of the widely varying exposure conditions in the STP. Release rate constants for the PRCs with the lowest log Kow values (acenaphthened10 and fluorene-d10) were highest at sampling site W8 (the active sludge aeration basin): 0.18 and 0.11 day-1, respectively. The ke,PRC values for all PRCs were high, similarly at sampling site W1 (ca. 0.1 day-1), which was unexpected, because there was no significant release of the high log Kow PCBs at any of the other sampling sites. A possible explanation for this finding is that a fatty layer that formed on the surface of the membrane from the raw sewage at this site may have replaced the WBL and exerted a “cleaning” effect on high log Kow compounds from the SPMD.

PRC-Corrected Water Sampling Rates. Laboratory calibration data for the water sampling rates (ranging from 0.5 L day-1 for naphthalene to the highest 2.5 L day-1 for pyrene) were selected from the laboratory calibration study (23) see Supporting Information Table S2. For high molecular weight (HMW) PAHs with log Kow values between 5.60 and 6.90, the laboratory-calibrated Rs values ranged from 1.4 to 1.8 L day-1. The same laboratory calibration data were used as reference data for both estimating water concentrations of the analytes and the comparative water sampling rate calculations. PRC-corrected water sampling rates for the 24 PAHs obtained using the empirical uptake model (eq 4) were negatively correlated to the Kow values of the PAHs (Figure 1b). The Rs values were highest for fluoranthene, pyrene, and 1-methylphenanthrene (log Kow 5.2-5.3) and varied from 2.4 to 21 L day-1 at the different sampling sites. The exposure conditions at the various sampling sites clearly differed, because there were significant between-site variations in PRC losses. Estimated in situ water sampling rates based on these losses varied from 2 to 4.8 L day-1 and were lowest at sampling point W9. The sampling rates for all 24 PAHs were very similar at W2 and W5, suggesting that exposure conditions were similar at these sites. The PRC-corrected Rs profiles of the studied PAHs were similar for sampling sites W3, W4, W6, and W7, and the values were slightly higher than those for sampling sites W2 and W5. Water sampling rates were higher at sampling site W8, ranging from 8 to 19 L day-1 for low molecular weight (LMW) PAHs and 7 to 12 L day-1 for HMW PAHs. However, the rates were highest for sampling site W1, where PRC-corrected Rs values for the PAHs ranged from 11 L day-1 for benzo[ghi]perylene (the compound with the highest log Kow value) to 21 L day-1 for fluoranthene and pyrene. The WBL-controlled uptake model yielded results that were generally similar to those obtained using the empirical uptake model; the estimated sampling rates for PAHs were lowest at sampling site W9, highest for LMW PAHs at sampling site W8, and highest for HMW PAHs at sampling site W1 (Figure 1c). However, the results differed from those obtained using the empirical model in that the estimated sampling rates for the HMW PAHs range were less strongly negatively correlated with their log Kow values. PRC-corrected sampling rates varied between 3 and 18 L day-1 among the sampling sites. However, while the empirical model indicated that Rs values fell sharply (by 2-10 L day-1, depending on the exposure conditions and the analytes’ log Kow values of 5-7) the corresponding reductions estimated VOL. 41, NO. 14, 2007 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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using the WBL model were much more minor, 1-2 L day-1 (Figure 1b,c). Principal Component Analysis (PCA). The sampling rates obtained using laboratory-calibrated data, W-cal, appear close to the PRC-corrected rates for sampling site W9, implying that the exposure conditions in terms of water flow, temperature, and fouling of the SPMD at this site were similar to those in the laboratory calibration study (Figure 2). Two main groups of sampling sites were recognized (Figure 2) among the PRC-corrected water sampling rates for the studied PAHs in the PCA plot. The first (Group 1) comprises sampling sites W2, W5, W6, and W9. At all of these sites the sewage water flow was slow, for example, in the clarifiers (W5, W6, W9), except at sampling site W2 where the FeSO4 added in large quantities may have had a strong influence. The ke values derived from the PRC release data were low for all of these sampling sites. The second (Group 2) includes three sampling sites (W3, W4, and W7), where the exposure conditions were similar and minor fouling was observed on the membrane surfaces of the SPMDs (Supporting Information, Table S1). Sampling sites with turbulent sewage water flow (W1, W8, and W10), and consequently high ke values, were located to the right side of PCA plot. At sampling site W8, in the active sludge basin, there was high bacterial activity and powerful aeration, which increased the mixing. The differences in PRC release rates indicate that there were clear differences in the exposure conditions among the sampling sites within the STP, demonstrating the importance of PRC calibration. Estimation of PAH Water Concentrations. The amounts of PAHs collected by the SPMDs (ng SPMD-1), presented in Supporting Information Table S4) are traditionally converted to water concentrations (ng L-1). The sums of PAH concentrations decreased during the course of the treatment (Figure 3a). However, the sum of the PAHs calculated using the laboratory-calibrated Rs values were higher and sometimes lower for all of the studied sampling sites, except sampling sites W2, W5, W6, and W9, where conditions were probably close to those in the selected laboratory calibration study. The sum of PAH concentrations calculated using PRCcorrected Rs values showed similar trends and similar reductions in total PAH concentrations during the sewage treatment process, suggesting that the process reduces the total dissolved PAH concentrations from approximately 130 ng L-1 to approximately 15 ng L-1. However, use of summed PAH concentrations masks differences in results obtained for individual compounds using the laboratory-calibrated and PRC-corrected sampling rates, especially the HMW compounds, for which water concentrations estimated using PRC-corrected rates were three-to-fourfold lower than those estimated using laboratory-calibrated rates. Concentrations of LMW PAHs generally decreased between successive stages of the sewage treatment process, implying that the process effectively removes LMW PAHs from the sewage water (Figure 3b), although use of laboratorycalibrated Rs values yielded higher concentration estimates than use of the PRC-corrected Rs values. Differences in concentrations calculated by the traditional (eqs 7 and 8) and full model (eq 9) equations varied up to 60%, and the latter model yielded significantly higher estimated higher summed PAH concentrations than the latter. The differences were largest for the HMW PAH concentrations (Figure 3c). HMW PAH concentrations calculated using laboratory calibration data were on average approximately five times higher than those obtained using in situ Rs values. Estimates of the HMW PAH concentrations obtained using the empirical and WBL models only slightly differed. The differences between the results obtained by these two models are due to the use of different compoundspecific effects, which are more important for compounds 5048

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FIGURE 3. Illustrative changes in (a) summed PAH concentrations and concentrations of (b) selected LMW and (c) selected HMW PAHs during the sewage treatment process (sampling sites W1W10) according to laboratory-calibrated traditional (eqs 7 and 8), and full laboratory calibrated (eq 9) and PRC-corrected Rs values (full empirical and full WBL).

with HMWs. In addition, the DOC content in the water has especially strong effects on HMW PAHs with high log Kow values, because it decreases the bioavailability or uptake of these compounds by the SPMDs by binding them and thus reducing their dissolved fractions (24). The introduction of compound-specific effects in the recently suggested calculation methods (empirical and WBL) avoid this “carbon effect” on laboratory calibrated Rs values, but how well they predict “true” dissolved concentrations is still under discussion. HMW PAHs did not show a clear decrease during the investigated sewage treatment process, indicating that the process does not successfully remove dissolved HMW PAHs from the sewage. In addition, fluctuations during the process were detected, especially associated with aeration of the sewage water. Large deviations among the PAH concentrations derived by the models used to estimate concentrations were seen, especially for HMW PAHs. Using laboratory-calibrated Rs values would result in overestimation of dissolved PAH concentrations in the sewage water. Thus, significant deviations between actual and estimated concentrations may arise unless appropriate equations and PRC-corrected Rs values are applied.

The sewage treatment process is complex, involving several kinds of treatments, all of which influence the quality and composition of particles and cause dramatic changes in flow conditions between sampling sites and partitioning effects at the surfaces of the membranes. These variables seem to influence the uptake of compounds by SPMDs, but variations in their effects can be accounted for by using PRCs. An important variable of this kind is fouling, which was observed in this study and can be caused not only by biological formations but also by sludge particles, fat, and salt particles. Such fouling can influence the thickness of the WBL on the surface of the membranes, and further studies on the potential effects of these phenomena on the uptake kinetics of pollutants by SPMDs are required.

Supporting Information Available A schematic diagram of the sewage treatment plant showing the sampling sites, descriptions of SPMD membranes’ surfaces after exposure, data on the physical properties of analytes, PRC release rate constants, and amounts of analytes collected by each SPMD. This material is available free of charge via the Internet at http://pubs.acs.org.

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Received for review January 9, 2007. Revised manuscript received April 23, 2007. Accepted May 4, 2007. ES070054+

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