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Environ. Sci. Technol. 2002, 36, 1791-1797

Calibration and Field Verification of Semipermeable Membrane Devices for Measuring Polycyclic Aromatic Hydrocarbons in Water DREW R. LUELLEN* AND DAMIAN SHEA North Carolina State University, Department of Environmental and Molecular Toxicology, P.O. Box 7633, Raleigh, North Carolina 27695

The use of semipermeable membrane devices (SPMDs) has become common in environmental sampling of nonpolar organic contaminants, yet few data exist for the uptake or sampling rates of polycyclic aromatic hydrocarbons (PAH). Two separate laboratory calibration experiments were conducted to determine the sampling rates of 28 individual PAH and 19 homologues. PAH with a log Kow > 4.5 remained in the linear uptake phase for 30 days, but PAH with a log Kow < 4.5 began to approach steady state within 15 days. Sampling rates, corrected for dissolved organic carbon, ranged from 2.11 to 6.06 L d-1. Shear flow across the membrane had no statistically significant effect on rates over the range of 0.01-0.50 cm s-1. Field verification of these sampling rates yielded agreement within about a factor of 2 for most PAH and a factor of 4 for all PAH. The worst agreement was for the most hydrophobic PAH, where partitioning into dissolved and particulate organic carbon pools are more important and less certain. These SPMD sampling rate data will allow quantitative estimations of freely dissolved concentrations of 47 compounds that are commonly used for PAH and petroleum product source identification and allocation.

Introduction To understand the transport, fate, and potential toxicity of a contaminant in any system, it is essential to have some measure or estimate of the contaminant concentration in the environmental medium of interest. In aquatic systems, environmentally relevant concentrations of many contaminants can be extremely low and highly variable, and often it is the freely dissolved or bioavailable concentration that is of greatest importance. Membrane-based passive sampling devices (PSDs), such as the semipermeable membrane device (SPMD), have shown great promise as a tool to help estimate the freely dissolved concentration of persistent hydrophobic chemicals in water (1-7). Their ability to concentrate trace quantities from the water in-situ and to integrate exposure over time offers tremendous advantage over discrete water sampling. The most common use of PSDs requires them to be calibrated by measuring uptake rate constants for the device under controlled conditions and then estimating the freely dissolved contaminant * Corresponding author phone: (804)684-7749; fax: (804)684-7793; e-mail: [email protected]. Corresponding author address: P.O. Box 1346, Gloucester Point, VA 23062. 10.1021/es0113504 CCC: $22.00 Published on Web 03/09/2002

 2002 American Chemical Society

concentration in the field using the following equation

CW,fd ) CPSD/k1t

(1)

where CW,fd is the concentration of the chemical freely dissolved in water (ng/L), CPSD is the concentration of the chemical in the PSD (ng/g), k1 is the first-order sampling or uptake rate constant (L g-1 d-1), and t is the time period the PSD is deployed (d). Equation 1 assumes linear uptake. The derivation of eqs 1 and 2, and other theoretical aspects of SPMD design and function, have been described extensively by others (2-7). It is usually more convenient to express uptake as the apparent volume of water that is quantitatively extracted by the device per unit time, an effective sampling rate, RS (L d-1)

CW,fd ) NPSD/RSt

(2)

where NPSD (ng) is the amount of chemical sorbed by the PSD. Thus RS values can be determined in the laboratory by rearranging eq 2 to solve for RS (at a fixed CW). One implicit assumption in the use of PSDs is that only the freely dissolved form of a chemical is available to partition into the membrane (2-7). Although this assumption has not been tested rigorously, it is reasonable to incorporate a correction for partitioning into dissolved organic carbon (DOC) and/or particulate organic carbon (POC), as was done recently for PCBs (5). Equation 2 can be modified to correct for the presence of DOC and estimate the total dissolved concentration of a contaminant in water (CW,td) using

CW,td ) (1 + CDOC KDOC)NPSD/RSt

(3)

where CDOC is the DOC concentration and KDOC (L g-1) is the partition coefficient of the chemical between water and the DOC. This correction can be made for total organic carbon (DOC + POC) if the sample is not filtered, either by including an additional term for POC with a corresponding partition coefficient (KPOC) or combining the two organic carbon pools and using a single organic carbon partition coefficient (KOC). This latter approach was used by Meadows et al. (5) to correct uptake rate constants for PCBs into SPMDs. However, KDOC have been reported to be lower than KPOC values in a recent review (8), and therefore this simplification can lead to an overestimation of partitioning into organic carbon. Regardless of the approach taken, one must recognize the uncertainty inherent in these calculations. Equation 1 can be modified in a similar manner to calculate uptake rates that are corrected for the DOC (and/or POC) in the laboratory exposure studies or to correct measured concentrations in the field. These equations apply to SPMDs and other similarly constructed PSDs (7, 9). However, sampling rates are proportional to the size of the SPMD, so it is necessary to express RS normalized to a standard configuration (3-5), currently defined as lay-flat polyethylene tubing sealed and lined with 0.91 g (1 mL) of the neutral lipid triolein (1,2,3tri[cis-9-octadecenoyl]glycerol) and having a 4:1 membraneto-triolein (w/w) ratio, a membrane surface thickness of 7590 µm, a total mass of 4.55 g, and a membrane surfacearea-to-triolein volume ratio of ≈450 cm2/mL (2-5). Use of eq 2 to estimate exposure in the field requires estimates of RS that are representative of uptake during field deployment of the SPMD. Sampling rates are usually estimated from laboratory flow-through experiments (2-4, 7), though static systems and field deployments have sometimes been used (5, 9). Given the central importance VOL. 36, NO. 8, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Summary of Sampling Rates for PAHs and Related Heterocyclic Compoundsb SPMD sampling rates Rs (L d-1) compound

symbol

MW

log Kow

exp 1

exp 2

mean DOC corrected

Huckins (4)

naphthalene biphenyl dibenzofuran C1-naphthalenes C2-naphthalenes C3-naphthalenes C4-naphthalenes 1-methylnaphthalene 2-methylnaphthalene 2,6-dimethylnaphthalene 2,3,5-trimethylnaphthalene acenaphthylene acenaphthene fluorene C1-fluorenes C2-fluorenes C3-fluorenes 1-methylfluorene dibenzothiophene C1-dibenzothiophenes C2-dibenzothiophenes C3-dibenzothiophenes phenanthrene C1-phenanthrenes/anthracenes C2-phenanthrenes/anthracenes C3-phenanthrenes/anthracenes C4-phenanthrenes/anthracenes 1-methylphenanthrene anthracene fluoranthene pyrene C1-fluoranthenes/pyrenes benz[a]anthracene chrysene C1-chrysenes C2-chrysenes C3-chrysenes C4-chrysenes benzo[b]fluoranthene benzo[k]fluoranthene benzo[a]pyrene benzo[e]pyrene perylene indeno[1,2,3-c,d]pyrene dibenz[a,h]anthracene benzo[g,h,i]perylene coronene

N BIP DBF N1 N2 N3 N4 1MN 2MN DMN TMN Acy Ace F F1 F2 F3 1MF D D1 D2 D3 P P1 P2 P3 P4 1MP AN FL PY FP1 BaA C C1 C2 C3 C4 BbF BkF BaP B eP PER IDP DBA BgP COR

128 154 168 142 156 170 184 142 142 156 170 152 154 166 180 194 208 180 184 198 212 226 178 192 206 220 234 192 178 202 202 216 228 228 242 256 270 284 252 252 252 252 252 267 278 276 300

3.37 3.90 4.12 3.86 4.37 4.90 5.30 3.86 3.86 4.37 4.90 4.07 3.92 4.18 4.97 5.2a 5.5a 4.97 4.38 4.8a 5.5a 5.7a 4.46 5.14 5.60 5.85 6.5a 5.14 4.54 5.22 5.18 5.7a 5.91 5.61 6.20 6.50 6.80 8.0a 5.80 6.0 6.04 6.40 6.50 7.0 6.75 7.23 7.64

3.46 ( 0.19 4.49 ( 0.26 4.75 ( 0.30 5.13 ( 0.15 4.76 ( 0.23 5.72 ( 0.39 4.46 ( 0.41 5.04 ( 0.24 4.91 ( 0.16 5.68 ( 0.29 5.94 ( 0.20 3.65 ( 0.20 3.92 ( 0.26 4.60 ( 0.10 5.82 ( 0.26 6.02 ( 0.21 5.69 ( 0.28 5.57 ( 0.08 3.86 ( 0.27 4.10 ( 0.34 3.82 ( 0.30 2.73 ( 0.35 4.96 ( 0.23 5.87 ( 0.45 4.60 ( 0.51 3.48 ( 0.46 3.05 ( 0.47 5.63 ( 0.21 4.71 ( 0.46 5.50 ( 0.30 6.20 ( 0.77 4.45 ( 0.45 4.88 ( 0.62 4.70 ( 0.40 4.29 ( 0.71 3.15 ( 0.55 2.36 ( 0.46 0.77 ( 0.19 2.90 ( 0.60 3.07 ( 0.56 3.11 ( 0.60 2.38 ( 0.33 2.47 ( 0.35 2.19 ( 0.27 1.85 ( 0.35 1.32 ( 0.28 1.04 ( 0.40

3.04 ( 0.17 4.37 ( 0.10 5.24 ( 0.05 5.40 ( 0.15 7.27 ( 0.15 6.27 ( 0.64 4.33 ( 0.51

3.25 ( 0.19 4.43 ( 0.26 5.00 ( 0.24 5.27 ( 0.15 5.53 ( 0.21 6.04 ( 0.55 4.47 ( 0.47 5.04 ( 0.24 4.91 ( 0.16 5.69 ( 0.29 5.96 ( 0.20 3.65 ( 0.20 3.92 ( 0.26 4.88 ( 0.10 5.66 ( 0.31 5.85 ( 0.40 5.21 ( 0.51 5.59 ( 0.08 3.43 ( 0.38 3.75 ( 0.47 3.20 ( 0.50 2.11 ( 0.46 4.73 ( 0.32 5.78 ( 0.69 4.56 ( 0.47 3.36 ( 0.72 3.21 ( 0.80 5.66 ( 0.21 4.72 ( 0.46 4.82 ( 0.25 6.06 ( 0.54 4.37 ( 0.68 4.46 ( 0.57 4.23 ( 0.47 3.34 ( 0.67 2.48 ( 0.55 2.19 ( 0.48 4.69 ( 2.87 2.43 ( 0.51 3.19 ( 0.59 3.24 ( 0.63 2.31 ( 0.66 2.77 ( 0.39 3.04 ( 0.37 2.25 ( 0.42 2.60 ( 0.78 2.79 ( 1.07

0.5 ( 0.22

5.14 ( 0.09 5.41 ( 0.34 5.51 ( 0.50 4.45 ( 0.73 2.98 ( 0.47 3.35 ( 0.52 2.42 ( 0.64 1.34 ( 0.53 4.47 ( 0.36 5.55 ( 0.78 4.21 ( 0.55 2.86 ( 0.82 2.06 ( 0.91 4.01 ( 0.15 5.75 ( 0.23 3.92 ( 0.81 3.49 ( 0.49 3.48 ( 0.51 1.73 ( 0.58 0.98 ( 0.54 0.76 ( 0.49 0.37 ( 0.25 1.74 ( 0.40 1.48 ( 0.42

0.88 ( 0.31

1.7 ( 0.07 2.4 ( 0.17 2.8 ( 0.03

5.0 ( 0.60

4.6 ( 1.4 6.8 ( 0.95 7.6 ( 0.91 4.7 ( 0.80 7.6 ( 0.76

3.3 ( 1.1 5.5 ( 1.0 5.4 ( 0.54 4.7 ( 0.38 3.4 ( 0.58 2.4 ( 0.22

a Log K a b Sampling rates are estimated from ow data are from Mackay (21) except for those denoted with which are from Neff and Burns (22). eq 2 using data from each time point beyond day 1 and within the linear portion of the uptake curve. Mean DOC corrected rates are the mean of exp 1 and exp 2, corrected for DOC using eq 3 and assuming KDOC ) 0.11Kow.

of uptake rate constants to estimating CW and the wide use of SPMDs, there have been surprisingly few peer-reviewed reports of uptake rate data. This is particularly true of perhaps the most common class of contaminants that accumulate in SPMDs, the polycyclic aromatic hydrocarbons (PAH). For the purpose of this study, the term PAH refers to all PAH, alkyl homologues, and related heterocyclic compounds listed in Table 1. One laboratory study reported uptake rates for the 16 priority-pollutant PAH (4) and another reported on just few PAH (7). Others have examined the uptake of a full suite of PAH in the marine environment (10, 11). It is now well established in the scientific literature and in environmental forensic cases (12-17) that a much larger suite of PAH must be measured to adequately understand the sources, fate, and toxicity of PAH in the environment. Several of the PAH exist as alkyl homologues (Table 1), with the parent nonalkylated compound (C0) and monoalkylated (C1), dialkylated (C2), trialkylated (C3), and tetraalky1792

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lated (C4) compounds. The relative abundance of these homologues for naphthalene (N), fluorene (F), dibenzothiophene (D), phenanthrene (P), and chrysene (C) is indicative of the source of the PAH (e.g., petrogenic versus pyrogenic, creosote versus diesel exhaust, etc.) and the degree of weathering (12-17). For example, fresh crude oil typically has an alkyl homologue distribution (AHD) dominated by the alkylated homologues (C1-C4). A maxima at the C1 homologue or higher indicates a petrogenic source of the hydrocarbon contamination (16). As the oil weathers, the AHD becomes depleted in C0 and the lower alkylated homologues because they are preferentially lost due to their higher aqueous solubility, vapor pressure, and biodegradation rates. Highly weathered oils often exhibit the profile C0 < C1 < C2 < C3 < C4, though sometimes C3 > C4. Pyrogenic PAH typically are dominated by C0 with very little alkylated homologues, though lower temperature combustion sources (e.g., diesel engines) can retain a substantial portion of the

alkylated signature in exhaust particles (12, 14, 17). By measuring a broad suite of PAH (Table 1), rather than just the 16 priority-pollutant PAH, one gains tremendous power to discriminate (allocate) among possible PAH sources and better understand complex transport, fate, and weathering processes. It also allows for more accurate risk assessments because the concentrations (and thus toxicity) of the alkylated homologues (C1-C4) are usually much higher than that of the parent PAH (C0) for petroleum derived sources. The ability to integrate exposure in water over time by measuring these PAH in SPMDs would be an important addition to the tools researchers use to study PAH in the environment. The purpose of this paper is to present the first peer-reviewed report of uptake rate constants for a broad suite 28 individual PAH and 19 alkyl homologues using a laboratory flow-through exposure system and to provide the first field verification of SPMD sampling rates for PAH.

Experimental Section Materials. SPMDs were either purchased (EST, St. Joseph, MO) in the standard configuration: 90 cm long, 2.5 cm wide, and 1 mL (0.91 g) of triolein, surface area ≈450 cm2, and ≈4.6 g or constructed as described by Hofelt and Shea (6), using 3-mil (approximately 86 µm) “virgin” low-density polyethylene (LDPE) tubing (Brentwood Plastics, Inc., St. Louis, MO). The LDPE tubing was extracted with hexane for 48 h prior to use and was filled with 95% triolein (Sigma, St. Louis, MO). SPMDs constructed for laboratory studies were 2.5-cm wide, 10 cm long, and contained 0.1 g (0.11 mL) triolein, resulting in a surface area of 50 cm2 and a mass of 0.40 g, thus conforming to the same relative dimensions of the standard configuration as defined by Huckins et al. (2, 4). The SPMDs used in experiment 1 and in the field were fortified with 1 µg/SPMD fluorene-d10, anthracene-d10, and pyrene-d10 for use as permeability/performance reference compounds (PRC). All solvents were Ultra Resi-Analyzed grade (J. T. Baker Inc., Phillipsburg, NJ). Aluminum foil was baked (400 °C) overnight before use, and glassware was solvent rinsed and baked prior to use. PAH standards were obtained from Accustandard Inc. (New Haven, CT); deuterated PAH standards were obtained from Cambridge Isotopes, Andover, MA). Alaska North Slope crude oil was a gift from Kevin McCarthy, Battelle, Duxbury, MA. Calibration Experiments. Two separate uptake experiments were conducted. Experiment 1 (exp 1) utilized a flowthrough diluter system that pumped a mixture of Alaska North Slope crude oil (ANS crude) and additional PAH standards (Table 1), at 10 L/h to a 40-L aquarium containing 18 standard SPMDs. The nominal ANS crude concentration dispersed in the water was 100 µg/L. The SPMDs were removed in triplicate on days 0, 0.5, 1, 3, 5, 7, 13, 19, and 29; single SPMDs were removed on days 9, 11, 15, 17, 21, 25, and 27. This system utilized a separate pump to induce mixing and create effective linear velocities ≈0.50 cm s-1, measured with a flow meter. Experiment 2 (exp 2) used a flow-through system consisting of a 200-L reservoir containing ANS crude and water mixture (100 µg/L nominal concentration) that was pumped to four 20-L glass aquaria, each holding up to twelve constructed (10-cm long) SPMDs. Tygon fuel and lubricant tubing was used to transfer the oil and water mixture from the reservoir to the aquaria. The flow rate was 11.5 mL/min to allow for one complete renewal of water each day. The linear velocity of the water was less than 0.01 cm s-1, estimated from the volume rate of flow and the cross-sectional area of the aquaria (7). The SPMDs were removed on days 1, 3, 5, 7, 9, 11, 13, 15, 17, 19, 21, 23, 25, 27, and 29. Triplicate samples were collected on days 1, 3, 5, 7, 13, 19 and 29; single samples were collected all other days.

In both experiments, the temperature of the exposure chambers was maintained at 25 ( 1 °C, and the flow-through systems were allowed to equilibrate for 48 h before adding the SPMDs. Copper sulfate (3-5 ppm) was added to both systems to inhibit the growth of microorganisms and algae; biofouling was not observed during any experiments. Both systems used tap water; mean dissolved organic carbon was 0.345 ( 0.074 mg L-1 and 1.29 ( 0.11 mg L-1 in experiments 1 and 2, respectively. After collection, the SPMDs were individually wrapped in aluminum foil and kept frozen at -20 °C until analysis. Water samples were collected twice a day during the course of the experiment and composited over 2-5 day periods before analysis. In most cases, the measured concentrations did not deviate more than 10% from the nominal or expected concentrations. Field Sample Collection and Sample Processing. Standard SPMDs were suspended in copper mesh cages with solid tops that provided shade from the sun and deployed from a dock near the New England Aquarium in Boston Harbor, an area known to have high PAH concentrations (18). Each cage contained two SPMDs that were later composited for an individual sample. Nine cages were deployed to provide triplicate samples at days 8, 15, and 29. The SPMDs were removed briefly approximately every 2 days, wiped clean, and dipped in a mixture of copper sulfate and an aquatic herbicide; no biofouling was visible at any time. Discrete 20-L water samples were collected almost every 2 days at about 0700 and 1800 h, returned to the laboratory, and filtered through a 0.7-µm glass-fiber filter. The filtrate was fortified with surrogate standards and serially extracted with dichloromethane (DCM) as described by Hofelt and Shea (6). Water extracts were treated in the same manner as SPMD extracts. Field-deployed SPMDs were cleaned (3) prior to extraction; only a light brown coloration was observed. Individual SPMDs from both the field and lab experiments were cut into small strips and serially extracted three times in Teflon bottles on a shaker table using a total of 75 mL of dichloromethane (DCM). Total extraction time was 24 h. Previous studies in our laboratory have shown this extraction method yields analyte recoveries equivalent to those of a 1-L 48-h dialysis (9). Concentrated extracts were fractionated using high performance gel permeation chromatography to remove polyethylene waxes and triolein. The samples were solvent exchanged into hexane and then fractionated on a 3-g silica column using a modified method (DCM was used instead of benzene) of Wang et al. (19) to split the extract into F1 (sterane and hopane biomarkers, data not presented) and F2 (PAH) prior to analysis. Instrumental Analysis. The F2 fraction was analyzed for PAH using an HP5890 Series II GC equipped with electronic pressure control connected to an HP5970 or HP5972 MSD utilizing a Restek 30 m × 0.25 mm Rtx-5 (film thickness 0.25 µm) MS w/Integra-Guard column. The pressure was ramped to 40 psi before injection with a 1-min hold time. The flow was then dropped to give a constant flow of 1 mL/min for the duration of the run. The temperature program for PAH analysis was as follows: initial temperature 40 °C for 1 min with a ramp of 6 °C/min to 290 °C and a final hold time of 30 min; injector temperature 300 °C, detector temperature 280 °C. Selected ion monitoring (SIM) was used for analysis. The method is similar to that described elsewhere (12, 15, 17, 20). Data Quality Assessment. Procedural blanks, polyethylene blanks, and triolein blanks were run with the samples to determine what background contamination was present in the materials or introduced during extraction and cleanup. All of the blanks were extremely clean, with only small amounts of naphthalene and phenanthrene being detected ( 0.20) in the average recoveries measured between the SPMD and water samples, in both the laboratory and the field. Therefore corrections for surrogate recovery would influence both sides of eq 2 equally, and the data were not corrected for recovery.

Results and Discussion Uptake Curves. The 47 PAH and related heterocyclic compounds investigated here are listed in Table 1. Uptake curves for four representative PAH ranging from 2 to 5 rings are shown in Figure 1; uptake curves for all of the PAH are given in Luellen (23). All PAH with log Kow > 4.5 remained in the linear uptake phase through the 30 day period, but PAH with lower Kow began to approach steady state during this period. Deviations from linearity began after 7 days for the 6 PAH with log Kow < 4.0 and began after 13-15 days for the 7 PAH with log Kow between 4.0 and 4.5. This is an important verification of the appropriateness of a linear model (eq 1) to derive SPMD sampling rates for nearly all PAH (41 of 47) over about a 2-week period. However, this also illustrates the potential problem of assuming linear uptake during longer deployment times especially since deployment periods are often about 4 weeks (3, 6). The use of eqs 1 or 2 to estimate Cw will not be valid for these longer deployment periods for PAH with log Kow < 4.5. There are several factors that influence the time window that SPMDs can be deployed while still allowing the assumptions inherent in eqs 1 and 2 to remain true. A nonzero intercept would require deployments be long enough to reduce the contribution from this initial uptake to an insignificant fraction. We generally observed very low 1794

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intercepts (Figure 1), and thus deployment times as short as a few days should be sufficient to reduce effects from nonzero intercepts. This is consistent with the data of Huckins et al. (2, 4) but is in contrast to substantial intercepts observed by Booij et al. (7). Booij et al. (7) utilized a suspended sediment system for their exposures that might have allowed particles to sorb to the surface of the membrane yielding higher initial amounts sorbed, though their calculations indicated this effect should be small. Short exposure times result in the accumulation of less chemical so detection sensitivity and variability at lower amounts can become important. Our experience with the full range of PAH studied here is that exposures of at least 5 days are required to provide sufficient mass to yield reproducible measurements with negligible contribution from any initial sorption. This is true for even the high molecular weight PAH as shown in Figure 1 for benzo[a]pyrene. Sampling Rates. Sampling rates (RS) for individual PAH were calculated using eq 2. Rates were calculated for each time point beyond day 1 and within the linear uptake region, and the mean values and standard deviations are reported in Table 1. Note that calculation of sampling rates based on only a single time point (triplicate analyses) within the linear uptake range would not have affected the mean or variance significantly. As discussed above, sampling rates were normalized to the standard SPMD configuration (3, 4). The use of eq 2 to estimate RS assumes constant aqueous concentrations of PAH (Cw). This was verified by collecting exposure water twice daily and compositing the samples for 2-day (experiment 1) and 5-day (experiment 2) averages; average measured exposures were within 11% of the nominal concentration. Despite the different experimental designs, there is excellent agreement between experiments 1 and 2 (Table 1). There was a slight negative bias for higher Kow PAH in

experiment 2, but this could mostly be accounted for by the nearly 4-fold higher DOC concentrations in experiment 2. Using eq 3 to correct for the DOC made differences in the sampling rates statistically insignificant for most PAH. In this paper we are using the relationship derived by Burkhard (8) where KDOC ) 0.11Kow. The mean DOC-corrected sampling rates are listed in Table 1. The amount of POC in both experiments was below the detection limit of 0.1 mg/L so no correction for POC was made. If POC concentrations approached 0.1 mg/L in these experiments, then it is possible that we underestimated sampling rates of the highest Kow PAH (e.g. C4-chrysene and coronene). However, assuming a POC concentration of 0.05 mg/L (1/2 the detection limit) and KPOC ) 0.41 Kow (8), a POC-corrected sampling rate would still be within one standard deviation of the mean DOCcorrected sampling rate for all 47 PAH. There is reasonably good agreement (Table 1) between our normalized RS values and those of Huckins et al. (4), with our values being somewhat higher for lower Kow PAH and lower for higher Kow PAH. The lower RS values reported by Huckins et al. (4) for naphthalene, acenaphthylene, acenaphthene, and fluorene are likely a result of approaching steady state in their study, where linearity was observed for acenaphthylene and fluorene for only 4 and 7 days, respectively. One difference between our work and Huckins et al. (4) was their periodic (10 min cycles) mixing of the water that might have increased shear flow across the membranes. This could reduce the aqueous boundary layer thickness and thus increase sampling rates for PAH whose uptake is controlled by diffusion through the aqueous boundary layer. Previous work has indicated that uptake of chemicals with log Kow > 4.5 is controlled by the aqueous boundary layer and that increasing shear flow across the SPMD surface increased uptake rates (2, 3, 7). It is these higher Kow PAH that have the higher sampling rates reported by Huckins et al. (4). In our work, shear flow was approximately 50 times higher in experiment 1 compared to experiment 2 (0.50 cm s-1 vs 0.01 cm s-1), yet we observed only slightly higher sampling rates for the higher Kow PAH. This is consistent with a previous study (3) that found only a 1.5-fold increase in uptake rates with a 50-fold increase in flow rate (0.0040.20 cm s-1), but it is inconsistent with theory (2, 3, 7, 24) and with the most recent experimental work (24). However, this latest study (24) included only a single PAH. Others have reported that uptake rates of PAH begin to decrease when log Kow exceeds about 5.0. This is presumably due to steric hindrance in the membrane as there is some evidence that the molecular dimensions of the higher Kow PAH sterically hinder movement through the pores of the membrane (4, 23). All three data sets in Table 1 show a consistent trend of RS increasing with log Kow, reaching a maximum near log Kow 5.0-5.5, and then decreasing again at higher Kow. Other chemical classes, such as PCBs and pesticides, do not exhibit the decrease at higher Kow (5, 7). Thus the assumption that the aqueous boundary layer is controlling uptake for compounds with log Kow > 4.5 may not be true for PAH, and this would explain the lack of a significant influence of flow rate on PAH uptake rates. Theoretical considerations have shown (2, 4, 25), and empirical data confirm (4), that exposure concentration should not effect the sampling rate. Thus, although there were substantial differences in exposure concentrations in the two studies, this should not influence the resulting sampling rates. The specific effect of temperature on sampling rates will depend on whether uptake is controlled by diffusion through the aqueous boundary layer or by permeability through the membrane (2, 4, 25). A previous study has shown small but measurable increases in PAH uptake rates with increasing temperature (4). Thus in Table 1, we compare our PAH sampling rate data for 25 °C to those of Huckins et al.

(4) at 26 °C. Extrapolation to lower temperatures can be made using the temperature-sampling rate response functions published by Huckins et al. (4), but water temperatures would have to approach 10 °C to have a statistically significant effect on our sampling rates listed in Table 1. Field Verification. The primary purpose of SPMDs is to estimate average concentrations of persistent hydrophobic chemicals in water. SPMDs are typically deployed for 2-4 weeks at a site, contaminant residues are measured, and average concentrations in the water estimated from eqs 2 or 3 and from published sampling rates. With the new data in Table 1, we can now estimate concentrations of 47 PAH and related heterocyclic compounds present in both petroleum and combustion sources. Field verification of these estimates was performed at a moderately contaminated site in Boston Harbor. Concentrations of PAH dissolved in water were estimated using eqs 2 and 3, the mean DOC-corrected sampling rates (Table 1), and PAH residues in the field-deployed SPMDs. SPMD residue data were taken from 8-day and 15-day deployments for PAH with log Kow < 4.0 and log Kow > 4.0, respectively, ensuring that uptake remained in the linear phase. These predicted concentrations are compared to measured dissolved PAH concentrations in Figure 2, where observed concentrations are presented as uncorrected for DOC (Figure 2a) and corrected for DOC measured at the site (Figure 2b,c). Corrections were not made for POC because the water samples were filtered prior to analysis. In both cases, most PAH concentrations are underpredicted and the greatest difference from unity occurs with the higher molecular weight PAH. The average predicted:observed ratio is 0.44 ( 0.36 and 0.53 ( 0.34 for uncorrected and DOCcorrected measurements, respectively. The most significant difference is with the highest molecular weight PAH with predicted values as low as 4% and 24% of observed values for uncorrected (directly measured) and DOC-corrected values, respectively. For DOC-corrected values, over half of the predicted values are within a factor of 2 of measured PAH. A regression of measured dissolved PAH versus DOCcorrected (KDOC ) 0.11Kow) predicted PAH yields a slope of 0.59, a near-zero intercept (1.14) and R2 ) 0.61. Removal of two statistical outliers (N and N3) does not change the slope (0.59) or intercept (-0.91) appreciably and R2 increases to 0.94. One obtains an even better relationship by assuming KDOC ) Kow when making the correction for DOC as shown in Figure 2c. The mean predicted:observed ratio is 1.26 ( 0.83. However, there is substantial evidence that our initial assumption of KDOC ) 0.11Kow is a much better representation of the quality of natural organic carbon (8) and thus this better agreement is probably fortuitous. The predicted:observed comparisons in Figure 2 were based on field measurements using the mean of seven duplicate water samples collected over a 15-day period. This represents a 7-fold increase in effort to obtain mean PAH concentrations that are very near the SPMD-derived predicted PAH concentrations. This is a substantial cost advantage that SPMDs have over collecting discrete water samples. The integrative nature of SPMDs also might provide a substantial advantage in better representing the true mean when fewer samples can be collected due to cost constraints. This was investigated by comparing the mean and range of the two duplicates collected at each sampling event to the mean SPMD-derived PAH concentrations (Figure 3). Considering the discrete sampling first, one can see the large range possible within a single day (7-fold on day 19). On this day a sample collected in the morning would have provided a very different picture compared to sampling in the afternoon. A sample collected on day 22 would be very different from a sample collected on day 8. This illustrates the potential problem of bias and poor representation of true conditions VOL. 36, NO. 8, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 2. Ratio of freely dissolved PAH concentrations derived from SPMD residues (predicted) to that measured in filtered water in Boston Harbor (observed). Vertical lines represent standard deviations based on triplicate predicted values: (a) uncorrected for DOC in Boston Harbor water, (b) corrected for DOC assuming KDOC ) 0.11Kow, and (c) corrected for DOC assuming KDOC ) Kow. when extrapolating data from a single or a few time points over larger time periods. And without continuous sampling during a day, we do not know how the concentrations changed during the time we did not sample. In contrast, SPMD-derived PAH concentrations integrate exposure over the time of deployment. Minima, maxima, and other shorter-term temporal features will be obscured, but average concentrations based on SPMD residues will likely be a better representation of true conditions in nature unless a substantial number of water samples are collected. The SPMD residues indicate that PAH concentrations decreased slightly with length of deployment, and this is consistent with the discrete water sampling. Longer deployments would begin to approach steady state for the lower molecular weight PAH (Figure 1), yielding artificially low estimates of dissolved PAH. Nearly all of the PAH remained in the linear uptake phase during the 1-month deployment (Figure 1 and ref 23); however, maximum deployment times of about 2 weeks would be more appropriate to retain the integrative nature of the SPMDs for all PAH. As discussed by many others (1-7, 9, 23), SPMDs also offer many advantages 1796

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over the use of bivalves for estimating aqueous exposures. One of the more important advantages also is demonstrated in Figure 1; the slower response time of SPMDs relative to bivalves provides an infinite sink (i.e., linear uptake) over a much longer period than is typical for bivalves (several weeks versus a few days). Estimating average dissolved concentrations of these 47 PAH in water by deploying SPMDs will provide a substantial improvement in our ability to understand PAH source, fate, and effects in aquatic systems. Performance Reference Compounds. Several authors have shown that temperature, flow-rate, and biofouling can affect the uptake rate of chemicals into SPMDs (2-7, 24, 26, 27). Differences in these parameters between laboratory calibration studies and environmental exposures could restrict the use of laboratory-derived sampling rates (Table 1) to a very narrow range of environmental conditions. Huckins et al. (2) proposed using permeability or performance reference compounds (PRCs) as an in situ calibration of SPMDs to correct for differences between laboratory calibrations and field deployments. PRCs are chemicals that would not be encountered in nature that are fortified in the SPMD

Harbor site and using the same baffled cages) we measured flow rates ranging from 0.05 to 0.09 cm s-1, which is within the range of our laboratory calibration experiments and is probably too small to cause a measurable difference in sampling rates (24, 26). Our results are consistent with recent theoretical descriptions of PRC usage (24, 26), though more field verification is necessary under environmental conditions that cause substantial differences in PRC release rates.

Literature Cited

FIGURE 3. Measured total dissolved PAH concentrations over time in Boston Harbor. Values based on direct measurements: average ([) and range (vertical lines) of duplicate discrete water samples. Average water concentration over 8 days (9), 15 days (b), and 29 days (2). Values predicted from SPMD residues: corrected for DOC (assuming KDOC ) 0.11 Kow) deployed for 8 days (s s), 15 days (s), and 29 days (- - -). Total dissolved PAH calculated from the sum of PAH listed in Table 1 minus 1MN, 2MN, DMN, TMN, 1MF, and 1MP.

TABLE 2. Comparison of Performance Reference Compound (PRC) Recoveries (%)a 8 days PRC

exp 1

field

15 days exp 1

field

29 days exp 1

field

anthracene-d10 72 ( 6 76 ( 5 68 ( 6 61 ( 5 54 ( 5 59 ( 6 fluorene-d10 57 ( 5 52 ( 6 43 ( 4 47 ( 5 27 ( 3 22 ( 3 70 ( 4 71 ( 8 65 ( 5 61 ( 4 51 ( 5 56 ( 4 pyrene-d10 a SPMDs were fortified with each PRC during assembly and residues were measured at three time points in both experiment 1 (exp 1) and in the field. Values represent mean recovery ( standard deviation (n ) 3). The loss follows first-order decay, and there is no statistical difference between paired data (exp 1 and field data).

during assembly, and their rate of loss during field deployments are compared to that measured during laboratory calibrations. This approach is based on theory and empirical evidence that both the uptake and release of nonpolar organic chemicals are controlled by the same molecular processes and undergo isotropic exchange kinetics (24, 26, 27). We fortified SPMDs used in exp 1 and the field with three perdeuterated PAH, fluorene-d10, anthracene-d10, and pyrened10, and found no statistically significant difference (p > 0.10) in their rates of release (Table 2). The implication is that for this study, sampling rates measured in the laboratory (Table 1) are similar to those in the field and no correction for temperature, flow rate, or biofouling is necessary. The excellent agreement in the PRC data that we found is probably not likely in most studies. The average temperature during field deployment was 20.8 °C, only a 4-degree difference from the calibration experiment and too small to have a statistically significant effect on RS (4). Using the temperature-sampling rate function derived by Huckins et al. (4), we estimate that changes in RS ranged only from 1 to 8% due to this 4-degree temperature difference. In addition, the SPMDs were removed every 2 days, wiped clean, and dipped in a mixture of copper sulfate and an aquatic herbicide to reduce the potential for biofouling, and no biofouling was visible at any time. Although flow rates were not measured in the field for this study, in subsequent experiments (at this same Boston

(1) Huckins, J. N.; Tubergen, M. W.; Manuweera, G. K. Chemosphere 1990, 20, 533-552. (2) Huckins, J. N.; Manuweera, G. K.; Petty, J. D.; Mackay, D.; Lebo, J. A. Environ. Sci. Technol. 1993, 27, 2489-2496. (3) Huckins, J. N.; Petty, J. D.; Lebo, J. A.; Orazio, C. E.; Prest, H. F.; Tillet, D. E.; Ellis, G. S.; Johnson, B. T.; Manuweera, G. K. In Techniques in Aquatic Toxicology; Ostrander, G. K., Ed.; CRCLewis Publishers: Boca Raton, FL, 1996; Chapter 34. (4) Huckins, J. N.; Petty, J. D.; Orazio, C. E.; Lebo, J. A.; Clark, R. C.; Gibson, V. L.; Gala, W. R.; Echols, K. R. Environ. Sci. Technol. 1999, 33, 3918-3923. (5) Meadows, J. C.; Echols, K. R.; Huckins, J. N.; Borsuk, F. A.; Carline, R. F.; Tillitt, D. E. Environ. Sci. Technol. 1998, 32, 1847-1852. (6) Hofelt, C. S.; Shea, D. Environ. Sci. Technol. 1997, 31, 154-159. (7) Booij, K.; Sleiderink, H. M.; Smedes, F. Environ. Toxicol. Chem. 1998, 17, 1236-1245. (8) Burkhard, L. P. Environ. Sci. Technol. 2000, 34, 4663-4668. (9) Hofelt, C. S. Ph.D. Dissertation, North Carolina State Univeristy, Raleigh, NC, 1998. (10) Røe, T. I.; Durell, G. S.; Johnsen, S. Mar. Pollut. Bull. 1999, 38, 977-989. (11) Røe, T. I.; Johnsen, S. Environ. Sci. Technol. 1999, 33, 19631969. (12) Page, D. S.; Boehm, P. D.; Douglas, G. S.; Bence, A. E. In Exxon Valdez Oil Spill: Fate and Effects in Alaskan Waters; Wells, P. G., Butler, J. N., Hughes, J. S., Eds.; ASTM: 1995; pp 41-83. (13) Wang, Z.; Fingas, M.; Sergy, G. Environ. Sci. Technol. 1994, 28, 1733-1746. (14) Benner, B. A.; Wise, S. A.; Currie, L. A.; Klouda, G. A.; Klinedinst, D. B. Environ. Sci. Technol. 1995, 29, 2382-2389. (15) Douglas, G. S.; Bence, A. E.; Prince, R. C.; McMillen, S. J.; Butler, E. L. Environ. Sci. Technol. 1996, 30, 2332-2339. (16) Yunker, M. B.; Snowdon, L. R.; Macdonald, R. W.; Smith, J. N., Fowler, M. G.; Skibo, D. N.; Mclaughlin, F. A.; Danyushevskaya, A. I.; Petrova, V. I.; Ivanov, G. I. Environ. Sci. Technol. 1996, 30, 1310-1320. (17) Burns, W. A.; Mankiewicz, P. J.; Bence, A. E.; Page, D. S.; Parker, K. R. Environ. Toxicol. Chem. 1997, 16, 1119-1131. (18) Shea, D.; Kelly, J. R. Transport and fate of toxic contaminants discharged by MWRA into Massachusetts Bay; Report ENQUAD 1992-04; Massachusetts Water Resources Authority: Boston, MA, 1992; 78 p. (19) Wang, Z.; Fingas, M.; Li, K. J. Chromatogr. Sci. 1994, 32, 361382. (20) Stout, S. A.; Uhler, A. D.; McCarthy, K. J. J. Environ. Forensics 2001, 2, 87-98. (21) Mackay, D.; Shiu, W. Y.; Ma, K. C. Physical-Chemical Properties and Environmental Fate Handbook; CRC Press: Boca Raton, FL, 1999. (22) Neff, J. M.; Burns, W. A. Environ. Toxicol. Chem. 1996, 15, 22402253. (23) Luellen, D. R. Ph.D. Dissertation, North Carolina State University, 2000. (24) Vrana, B.; Schuurmann, G. Environ. Sci. Technol. 2002, 36, 290296. (25) Gale, R. W. Environ. Sci. Technol. 1998, 32, 2292-2300. (26) Huckins, J. N.; Petty, J. D.; Lebo, J. A.; Almeida, F. V.; Booij, K.; Alvarez, D. A.; Cranor, W. L.; Clark, R. C.; Mogensen, B. B. Environ. Sci. Technol. 2002, 36, 85-91. (27) Booij, K.; Smedes, F.; van Weerlee, E. M. Chemosphere 2002, 46, 1157-1161.

Received for review October 10, 2001. Revised manuscript received January 17, 2002. Accepted January 28, 2002. ES0113504

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