Polyethylene–Water Partitioning Coefficients for Parent- and Alkylated

Mar 14, 2013 - Department of Civil and Environmental Engineering, Stanford University, Stanford, California 94305-4020, United States. Environ. Sci. ...
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Polyethylene−Water Partitioning Coefficients for Parent- and Alkylated-Polycyclic Aromatic Hydrocarbons and Polychlorinated Biphenyls Yongju Choi, Yeo-Myoung Cho, and Richard G. Luthy* Department of Civil and Environmental Engineering, Stanford University, Stanford, California 94305-4020, United States S Supporting Information *

ABSTRACT: We report polyethylene (PE)−water partitioning coefficients (KPE) for 17 parent-polycyclic aromatic hydrocarbons (PAHs), 22 alkylatedPAHs, 3 perdeuterated parent-PAHs, and 100 polychlorinated biphenyl (PCB) congeners or coeluting congener groups. The KPE values for compounds in the same homologue group are within 0.2 log units for alkylated-PAHs but span up to an order of magnitude for PCBs, due to the greater contribution of the position of the substituents (i.e., chlorines for PCBs and alkyl groups for alkylated-PAHs) to the molecular structure. The KPE values in deionized water for parent- and alkylated-PAHs show a good correlation with a regression model employing the number of aromatic carbons (CAR) and aliphatic carbons (CAL) in each compound: log KPE = −0.241 + 0.313CAR + 0.461 CAL. The regression model is useful for the assessment of freely dissolved aqueous concentrations of alkylated-PAHs, which comprise a significant fraction of the total in petroleum-derived PAHs and in some pyrogenic PAH mixtures. For PCBs, experimentally determined octanol−water partitioning coefficients are the best predictor of the KPE values among the molecular parameters studied. The effect of salinity up to 20 or 30 parts per thousand is found to be relatively insignificant on KPE values for PAHs or PCBs, respectively.

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water.12,13 Materials tested as passive sampling devices include semipermeable membrane devices (SPMDs), solid-phase microextraction (SPME) fibers, and simple polymeric materials such as polyoxymethylene (POM), polydimethylsiloxane (PDMS) and low-density polyethylene (PE).14−17 Among those, PE passive samplers have several advantages over others: they are readily available, inexpensive, robust, and easy to deploy.18,19 To estimate the freely dissolved aqueous concentrations using a passive sampler such as PE, the partitioning coefficient between the passive sampler and the aqueous phase must first be known. At equilibrium, the aqueous concentration can be determined as Cw = CPE/KPE, where Cw (mol L−1 water) and CPE (mol kg−1 PE) are the concentrations in water and PE, respectively, and KPE (L water kg−1 PE) is the PE−water partitioning coefficient.20 In a laboratory experiment, the equilibrium aqueous concentration can be determined by exposing the PE to a well-mixed solution or sediment slurry until equilibrium. In the field under quiescent conditions, such as for sediment pore-water, the aqueous concentration can be

INTRODUCTION Many studies show that the freely dissolved aqueous concentration is a useful indicator for the risk to biota and human health from exposure to hydrophobic organic contaminants (HOCs) in aquatic environments.1−3 This calls for a reliable method to measure the freely dissolved aqueous concentration in a water column or sediment pore-water.4−7 Although equilibrium partitioning (EqP) principles for sediment−water partitioning are used most frequently to estimate HOC freely dissolved aqueous concentrations, an increasing body of evidence suggests that this approach may not give the desired accuracy for estimation of environmental fate and risk.5,6,8 For example, the partitioning coefficients between water and sediment organic carbon (Koc) in field sediments may range up to 2 orders of magnitude for polycyclic aromatic hydrocarbons (PAHs) including both parent- (i.e., nonsubstituted) and alkylated- (i.e., alkyl-substituted) PAHs.8 Direct measurement of aqueous concentration is problematic, since it requires a large volume of water, is labor intensive, and presents many analytical challenges.9,10 Passive sampling techniques show promise as a simple and relatively accurate assessment tool for the measurement of freely dissolved aqueous concentrations of HOCs in natural waters and sediments.10−13 It is reported that the use of passive samplers allowed the measurement of aqueous concentrations of PAHs and polychlorinated biphenyls (PCBs) down to the picogram per liter level without requiring a large volume of © 2013 American Chemical Society

Special Issue: Rene Schwarzenbach Tribute Received: Revised: Accepted: Published: 6943

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were purchased from Ultra Scientific (N. Kingstown, RI). A total of 30 alkylated-PAH standards, either in solution or powder, was obtained from Ultra Scientific, SPEX Certiprep (Metuchen, NJ), Sigma-Aldrich (St. Louis, MA), and Alfa Aesar (Ward Hill, MA). The alkylated-PAH standards were evaluated by gas chromatography-mass spectrometry for peak separation and 22 compounds were finally selected as study compounds. PE−Water Equilibria. The tests for the determination of PE−water partitioning coefficients (KPEs) were conducted by an initial step of compound impregnation into PE followed by a PE−water equilibration step. A cocktail solution of PAHs in methylene chloride (5 μg for each PAH) or PCBs in hexane (188 ng as total PCBs; mixture of Aroclor 1232, 1248, and 1262 and PCB congener nos. 29, 69, 103, 155, and 192) was spiked into a 1 L bottle. The bottle was gently swirled until the organic solvent was completely evaporated. After filling with water and adding three pieces of PE cut in 2.5 × 2.5 cm squares, the bottles were rolled at 2 rpm for 3 months for the impregnation of the compounds into the PE. Sodium azide (0.1 g L−1) was added to prevent microbial activity. A preliminary measurement of the concentration in the PE and water after this impregnation period indicated that the system contained nondissolved organic compound aggregates in water. Therefore, for reliable estimate of aqueous concentration, the three pieces of PE were collected from the bottle, wiped clean with Kimwipes, and each was placed into a separate 1 L bottle filled with 950 mL deionized (DI) water to set up a new equilibrium experiment without any organic-phase aggregates. To study the effect of salinity, tests with saline water were set up in parallel by dissolving sea salt (Red Sea, Houston, TX) into DI water. The salinity of the water was adjusted to 20 parts per thousand (ppt) for PAHs and 30 ppt for PCBs. These salinities were chosen for application of the KPE values at our study sites in a coastal channel and an estuarine area, respectively. Sodium azide was added for tests with both DI and saline water (0.1 and 1 g L−1, respectively). The bottles were shaken at 150 rpm for 11 weeks at 20 °C to ensure equilibrium. Studies on thin polymeric passive samplers including PE show that the passive sampler-water equilibrium is established within four weeks for PAHs and PCBs in mixed systems.12,13,28,29 After the equilibration period, both the PE and water were collected for PAH or PCB analysis. PE and Water Extraction. The PE was extracted twice for 24 h with 40 mL hexane for each extraction. More than 99.9% of PAHs and PCBs could be recovered by this extraction procedure as verified by extracting the remaining PE for two weeks by hexane and another two weeks by 1:1 (v:v) hexane:acetone. The weight of the PE was determined after the extraction. The water was extracted with three sequential aliquots of 40 mL hexane. The PE and water extracts were passed through a silica gel column to remove organic interferences following USEPA Method 3630C prior to analysis. PAH and PCB Analysis. The analysis of parent- and alkylated-PAHs was performed on a gas chromatograph−mass spectrometer (GC-MS; Agilent model 6890 and 5973N, respectively) using an HP-5MS capillary column (30 m × 0.25 mm i.d.). Details on the analytical method are described in Choi et al.30 The GC separation was performed by following an oven temperature program developed by Wang et al.:31 60 °C for 2 min, heated to 258 °C at 6 °C min−1, then to 300 °C at 2 °C min−1, and hold for 4 min at 300 °C. Six internal standards, d8-naphthalene, d10-acenaphthene, d10-anthracene, d10-pyr-

estimated after calibrating for nonequilibrium conditions using performance reference compounds18,21 or mass transfer modeling.22,23 Although several studies have reported laboratory-determined KPE values for parent-PAHs and PCBs (listed in Lohmann19), most such studies have focused on several representative members of the compound groups. A study by Smedes et al.24 is an exception, which reported KPE values for 15 parent-PAHs, 11 deuterated parent-PAHs, and 41 PCB congeners. A critical review of experimentally determined KPE values among different laboratories shows large variations in the values for some members of PAHs and PCBs.19 Therefore, reliable data are still needed for KPE values determined for a comprehensive set of parent- and alkylated-PAHs and PCBs. Alkylated-PAHs account for a substantial portion of total PAH (i.e., sum of parent- and alkylated-PAHs) toxicity to biota and thus are drivers of regulatory concern in many cases.25−27 Alkylated-PAHs are major risk drivers in petroleum-impacted areas where they predominate over parent-PAHs both in terms of abundance and toxicity to biota.25,26 Hawthorne et al.26 reported that the toxicity for alkylated-PAHs accounted for 98% and 99% of the total PAH toxicity in diesel fuel and crude oil, respectively, whereas parent-PAHs contributed only 2% and 1%. Alkylated-PAHs contributed approximately 60% of the total concentrations and toxicity of total PAHs in sediments collected from manufactured gas plant sites.26 The U.S. Environmental Protection Agency’s (USEPA) guideline to protect benthic organisms from PAH mixtures recommends summing the toxicological contribution of so-called 34-PAHs in sediment, including 18 parent-PAHs and 16 groups of alkylated-PAHs ranging from C1- to C4-PAHs (C no. refers to the number of aliphatic carbons attached to the aromatic structure).3 Despite the importance of alkylated-PAHs for the characterization and reduction of risk at regulated sites, little attention has been paid to the use of passive samplers to measure the freely dissolved aqueous concentrations of those compounds. For POM, Hawthorne et al.13 reported POM− water partitioning coefficients (KPOMs) for the 34-PAHs.13 However, to our knowledge, no studies have reported the PE− water partitioning coefficients (KPEs) for a comprehensive set of alkylated-PAHs. The goal of the present study is to enable an accurate determination of the freely dissolved aqueous concentrations for PAHs (including alkylated-PAHs) and PCBs using polyethylene (PE) at equilibrium by providing a comprehensive data set and regression models that apply to a wide range of PAHs and PCBs. The KPE values for 17 parent-PAHs, 3 deuterated parent-PAHs, 22 alkylated-PAHs, and 100 PCB congeners or coeluting PCB congener groups are measured experimentally in both fresh and saline water. A variety of molecular properties are used for correlation with the measured KPE values to identify the most appropriate regression model for each compound group.

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MATERIALS AND METHODS Test Materials. Additive-free low-density polyethylene (PE) with 51 μm thickness (0.92 g cm−3 density) was obtained from Brentwood Plastics (St. Louis, MO). The PE was precleaned with a series of solvents (hexane, acetone, and deionized water) prior to use as described in Tomaszewski and Luthy.18 All solvents were pesticide grade and used without further purification. Standard solutions for PCBs, parent-PAHs, deuterated parent-PAHs, 2-fluorobiphenyl, and d14-terphenyl 6944

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Figure 1. Logarithm of the PE−water partitioning coefficients (log KPE) determined for 17 parent-PAHs and 22 alkylated-PAHs at 20 °C in DI water. Solid triangles (▲) and hollow squares (□) represent the experimental results for parent- and alkylated-PAHs, respectively, and the solid line represents the regression model employing the number of aromatic carbons (CAR) and aliphatic carbons (CAL).

ene, d12-chrysene, and d12-benzo[a]pyrene, were spiked into each sample prior to analysis. PCBs were analyzed on an Agilent model 6890 GC with a DB-5MS capillary column (60 m × 0.25 mm i.d.) and a micro electron-capture detector (μECD). The analytical procedure followed USEPA Method 8082 with minor modification as described in Janssen et al.32 The GC temperature program was as follows: heat from 100 to 270 °C at 1.5 °C min−1, then to 280 °C at 15 °C min−1, and hold for 15 min at 280 °C. Two internal standards, PCB nos. 30 and 204, were spiked into each sample prior to analysis. The analytical method allowed determination of 83 PCB congeners and 21 coeluting congener groups. Correlation of PE−Water Partitioning Coefficients (KPE) with Molecular Parameters. A variety of molecular parameters were used to correlate with the experimental KPE values for PAHs and PCBs. For parent- and alkylated-PAHs, the number of aliphatic and aromatic carbons, octanol−water partitioning coefficient (Kow), hexadecane−water partitioning coefficient (Khdw), molar volume, and molecular weight were used for the correlation. For PCBs, number of chlorines, Kow, Khdw, and molar volume were used as predictors for the KPE values. The molecular parameters for the study PAHs and PCBs are provided in Tables S1 and S2 of the Supporting Information (SI). The Kow values for PAHs were obtained from the USEPA’s EPI suite (Estimation Programs Interface Suite, v. 4.10). The calculated values were used only when experimental values were not found in the database. Experimental Kow values for PCBs were obtained from Hawker and Connell.33 For the Khdw values and molar volumes for PAHs and PCBs, the SPARC online calculator (http://archemcalc.com/sparc, v. 4.6, accessed July 2012) was used as a basic approach because it is easily accessible, widely used, and applicable for any organic compounds without the need for further determination of the input parameters.34,35 The values were calculated at 20 °C, 1 atm using SMILES strings as the only input. The Khdw values were also calculated by polyparameter linear free energy relationships (pp-LFER) for PCBs using coefficients for a dry hexadecane−water system:36

where capital letters are compound descriptors for various types of molecular interactions, E is the excess molar refraction; S is the dipolarity/polarizability parameter; A and B are the overall solute H-bond acidity and basicity, respectively; and V is the McGowan characteristic molar volume divided by 100.37 The compound descriptors for each PCB congener were obtained from an updated set by van Noort et al.,37 which accounts for the effect of ortho-substituted chlorines on the descriptors.

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log Khdw(pp − LFER) = 0.087 + 0.667E + 1.617S − 3.587A − 4.869B + 4.433V

RESULTS AND DISCUSSION

PE−Water Partitioning Coefficients (KPE) for PAHs and PCBs. The KPE values (L kg−1) are determined for 17 parentPAHs, ranging from 2-ring to 6-ring compounds, 22 alkylatedPAHs, ranging from methyl-naphthalenes to a tetra-methylnaphthalene and to a dimethyl-benz[a]anthracene, and 3 deuterated parent-PAHs. For PCBs, the KPE values are determined for 79 PCB congeners and 21 coeluting PCB congener groups ranging from mono- to nona-chlorobiphenyls. The log KPE values for the PAHs and PCBs in deionized (DI) water are plotted in Figures 1 and 2 and the numerical values in DI and saline water are presented in Tables S3 and S4 of the SI along with the literature values for comparison. For most of the alkylated-PAHs and for many PCB congeners, the KPE values in the current study are the first to be reported in the literature. The PAH log KPE values range from 3.23 for naphthalene to 6.50 for indeno[1,2,3-cd]pyrene and the PCB log KPE values range from 3.87 for PCB no. 3 (4-chlorobiphenyl) to 7.56 for PCB no. 206 (2,2′,3,3′,4,4′,5,5′,6-nonachlorobiphenyl). The standard deviations of the triplicate measurements are within 0.1 log units for most compounds, substantiating the reproducibility of the measurement. For most PAH compounds and PCB congeners with literature values of KPE, the KPE values measured in the current study are within or close to (i.e., ± 0.3 log units) the range of the previously reported values.10,24,28,38−41 However, the KPE values measured in this study for some high molecular weight PAHs (e.g., benzo[ghi]perylene) are smaller than the values measured in Smedes et al.24 by up to one log units although the PE was obtained from the same manufacturer. This indicates that an interlaboratory study is needed for the agreement on the KPE values for those high molecular weight compounds. Generally, the addition of a methyl group to the ring structure results in a 0.4 to 0.5 log unit increase in KPE values

(1) 6945

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structure is unlikely to significantly affect the sorption affinity of alkylated-PAHs to PE. KPE Regression Models for PAHs. A linear model employing the number of aliphatic carbons (CAL) and aromatic carbons (CAR) is applied for the regression of the measured PAH KPE values. This approach was suggested by Hawthorne et al.13 and validated for polyoxymethylene (POM)−water partitioning coefficients for 18 parent-PAHs and 16 alkylatedPAHs . The following equation gives a good estimation for the KPE values determined at 20 °C in DI water (Figure 1): log KPE = − 0.241 + 0.313CAR + 0.461CAL

(2)

R2 = 0.982, n = 39

The regression equation suggests that the addition a methyl group to a PAH molecule results in an average of 0.46 log units increase in the KPE value. The coefficient for CAR is 47% greater than the coefficient for CAL, indicating that the addition of a carbon at the alkyl branch increases PAH KPE values by a greater extent than does the addition of carbon at the aromatic ring structure. This can be attributed to the relatively greater contribution of an alkyl carbon to the size of an alkylated-PAH molecule than that of an aromatic carbon. The correlation of the PAH KPE values with molecular weights, SPARC-calculated molar volumes (Vm(SPARC)), octanol−water partitioning coefficients (Kow), and SPARCcalculated hexadecane−water partitioning coefficients (Khdw(SPARC)) is shown in Figure 3. The correlation coefficients (R2) for the regression models using the four parameters are slightly smaller than the correlation coefficient for eq 2. However, all parameters are relatively good predictors of the PAH KPE values with correlation coefficients greater than 0.94.

Figure 2. Logarithm of the PE−water partitioning coefficients (log KPE) determined for 79 PCB congeners at 20 °C in DI water. The symbols represent the experimental results with different numbers of ortho-substituted chlorines, and the solid line represents the regression model using the number of chlorines in the molecules (NCl).

for PAHs. The KPE values of alkylated-PAHs in the same group (i.e., same molecular weight) are within 0.2 log units. This indicates that the position of the methyl substituents in the ring

Figure 3. Correlation between the logarithm of PE−water partitioning coefficients (KPE) and the (a) molecular weights (MW), (b) SPARCcalculated molar volumes (Vm(SPARC)), (c) logarithm of octanol−water partitioning coefficients (log Kow), and (d) logarithm the SPARC-calculated hexadecane−water partitioning coefficients (log Khdw(SPARC)) for 17 parent-PAHs and 22 alkylated-PAHs. Solid triangles (▲) and hollow squares (□) represent the experimental results at 20 °C in DI water for parent- and alkylated-PAHs, respectively, and the solid line represents the regression curve. 6946

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Figure 4. Correlation between the logarithm of PE−water partitioning coefficients (log KPE) and the (a) SPARC-calculated molar volume (Vm(SPARC)), (b) logarithm of octanol−water partitioning coefficients (log Kow), (c) logarithm of the SPARC-calculated hexadecane−water partitioning coefficients (log Khdw(SPARC)), and (d) logarithm of hexadecane−water partitioning coefficients calculated by polyparameter linear free energy relationships (log Khdw(pp‑LFER)) for 79 PCBs. Symbols represent experimental results at 20 °C in DI water, and the solid line represents the regression curve.

The KPE values for PCBs vary much more within the same homologue group than those for alkylated-PAHs. Data in Figure 2 show that the PCB KPE values within the same homologue group vary by up to an order of magnitude. This can be attributed to the relatively greater effect of chlorine substituents on the structure of a PCB molecule than the effect of a methyl substituent on that of a PAH. While all methylsubstituted alkylated-PAHs are planar, the steric nature of PCBs depends on the positioning of chlorines in the biphenyl moieties. The increasing number of ortho-substituted chlorines generally results in decreased hydrophobicity of a PCB molecule, as the ortho-substituted chlorines contribute to the restriction of the rotation of the two benzene rings relative to each other.35,37,43 Data in Figure 2 clearly show the general trend of decreased KPE values with the increased number of ortho-substituted chlorines. Taking into account this ‘ortho’ effect, a better correlation between the KPE values and the number of chlorines in PCBs is obtained:

The regression equations are not provided separately for parent- and alkylated-PAHs because the correlation among the KPE values and the molecular parameters is almost identical for the two compound groups. The Khdw(SPARC) values show a better correlation with the PAH KPE values than the Kow values: log KPE = 0.849log Khdw(SPARC) − 0.242

(3)

R2 = 0.971, n = 39

Since PE consists of repeating methylene units (−CH2−)n, hexadecane should be a better surrogate for PE partitioning than moderately polar solvents such as octanol.42 Therefore, the Khdw values are expected to show a better correlation with KPE than Kow once accurate values are obtained. The relatively good correlation between Khdw(SPARC) and KPE for the 40 parent- and alkylated-PAHs provides an indirect evidence that the SPARC online calculator can provide relatively accurate estimates of Khdw (or at least, adequately account for the difference in the values) for parent- and alkylated-PAHs. Lohmann’s review19 showed that the Khdw(SPARC) values were good predictors of KPE values for parent-PAHs. KPE Regression Models for PCBs. The KPE values for the 79 PCB congeners measured as single congeners in this study are plotted against the number of chlorines (NCl) in Figure 2. The general increase in KPE values with increasing number of chlorine substituents is evident for PCBs. The dependence of the KPE values and the number of chlorines (NCl) in PCBs is found as log KPE = 3.828 + 0.432NCl

log KPE = 3.691 + 0.533NCl − 0.181NCl − ortho

(5)

R2 = 0.920, n = 79

where NCl is the total number of chlorines and NCl‑ortho is the number of ortho-substituted chlorines. The correlation of PCB KPE values with SPARC-calculated molar volumes (Vm(SPARC)), experimental values of octanol− water partitioning coefficients (Kow) obtained from Hawker and Connell,33 SPARC-calculated hexadecane−water partitioning coefficients (Khdw(SPARC)), and Khdw values calculated by polyparameter linear free energy relationships (Khdw(pp‑LFER)) is shown in Figure 4. Among the four parameters, the Kow

(4)

R2 = 0.898, n = 79 6947

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values are the best predictor of PCB KPE values with the following relationship: log KPE = 1.02log Kow − 0.451

(6)

R2 = 0.926, n = 79

The SPARC online calculator did not give much variation in Vm(SPARC) and Khdw(SPARC) values for congeners with same degree of chlorination. As a result, data in Figure 4a and c are lumped together for each homologue group. The SPARC online calculator predicts any physicochemical properties of an organic compound based on the underlying phenomena (e.g., resonance, electrostatic, induction, dispersion, H-bonding interactions, etc.), which are quantified explicitly from the molecular structure of a compound.34 The 2D molecular structure is the only input for the calculation of the physicochemical properties of a compound for the SPARC online calculator.44 Therefore, it is likely that the SPARC online calculator does not adequately take into account the variation in 3D molecular structure due to positions of chlorines for PCBs that causes variable molecular properties for compounds within a homologue group. For example, the KPE value for the monoortho-substituted tetra-chlorinated biphenyl, PCB no. 60 (log KPE = 5.79), was significantly greater than the KPE value for the tri-ortho-substituted tetra-chlorinated biphenyl, PCB no. 46 (log KPE = 5.10). The log Kow values experimentally determined by Hawker and Connell33 were 6.11 for PCB no. 60 and 5.53 for PCB no. 46, also indicating the significant difference in hydrophobicities for the two PCB congeners. On the other hand, the log Khdw(SPARC) values were almost identical for the two congeners (7.12 for PCB no. 60 and 7.17 for PCB no. 46). The inability of the SPARC online calculator to predict reliable Kow values for PCBs was reported by Niederer and Goss.45 Note that for the parent- and alkylated-PAHs studied, which are all planar except for 2-ethylanthracene, the Khdw(SPARC) values correlated much better with the experimental KPE values than for the PCBs. The Khdw(pp‑LFER) values calculated using the compound descriptors that take into account the “ortho” effect showed better correlation with the KPE values than the Khdw(SPARC) values (Figure 4d). Still, the variation of the Khdw(pp‑LFER) values (up to 0.3 log units) for compounds within a PCB congener group was relatively smaller than the variation of the experimental KPE and Kow values (up to 1.0 log units), suggesting that the pp-LFER should be further improved to adequately account for the effect of chlorine positions on the estimated Khdw values. Several researchers suggested using the commercial software COSMOtherm to estimate Khdw values, which is based on quantum chemical calculations using the 3D molecular structure.42,45−47 However, the software is not publicly accessible and needs to be validated by a comprehensive data set, which currently limits its applicability for general use.44,48 Effect of Salinity on KPE Values. The KPE values for PAHs and PCBs in saline water (KPE,SW) are compared with those in DI water (referred to as KPE,DIW in this section only) in Figure 5. The data show that the presence of salinity of 20 or 30 parts per thousand (ppt) does not significantly affect the PE−water equilibrium. The plot of KPE,SW versus KPE,DIW fall close to the 1:1 line and the differences between the two values are not statistically significant (p > 0.05, Student’s t test) for most PAH and PCB compounds.

Figure 5. Logarithm of PE−water partitioning coefficients measured at 20 °C in saline water (log KPE,SW) plotted against the experimental log KPE values in DI water (log KPE,DIW) for (a) 17 parent-PAHs and 22 alkylated-PAHs and (b) 79 PCB congeners. The solid line represents the 1:1 curve for the two coefficients, and the dashed line represents the correlation between the log KPE,SW and the log KPE,DIW values applying the salting out effect with Setchnow constants (Ks) estimated in Endo et al.48.

Generally, it has been postulated that the addition of salt results in an increase in the KPE value as the compound aqueous solubility is reduced by the salting out effect.19 Based on Setchnow constants (Ks) estimated in Endo et al.48 for PAHs and PCBs using the pp-LFER, approximately 0.12 and 0.18 log unit increase in KPE values are expected for PAHs with 20 ppt salinity and for PCBs with 30 ppt salinity, respectively. Data in Figure 5 show that the slight difference between the KPE,DIW and the KPE,SW can be well described by the salting out effect with Ks values in Endo et al.48 for compounds having KPE,DIW values smaller than 5.5. For PAHs and PCBs with greater KPE,DIW values, however, the salting out effect is not clearly observed. For some high molecular weight PAHs and highly chlorinated PCBs, the KPE,SW values are slightly smaller than the KPE,DIW values. This suggests that the effect of ionic strength on the PE−water equilibrium may not be simply explained by the salting out effect for PAHs and PCBs with KPE,DIW values greater than 5.5. Since the effect of salinity on KPE values is relatively minor (i.e., less than 30%), salinity correction may not significantly improve the quality of data for the estimation of freely dissolved concentration in saline water using PE samplers. If needed, the salinity correction according to the salting out effect is applicable for PAHs and PCBs with KPE,DIW values smaller than 5.5. The salinity applied in this study (i.e., 20 ppt for PAHs and 30 ppt for PCBs) is relevant to coastal channels for PAHs and estuarine areas for PCBs, and for such cases the 6948

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and stipend for Yongju Choi was supported by Samsung Scholarship. We thank Dr. William R. Gala and Dr. Thomas P. Hoelen at Chevron for technical and academic support.

experimental values shown in Tables S3 and S4 of the SI can be applied. Recommendations. For alkylated-PAHs, hundreds of isomers may exist within each homologue group and the concentrations of alkylated-PAHs in environmental media are often determined by the sum of compounds that belong to each group. The determination of environmental concentrations for a certain alkylated-PAH homologue group is typically performed by gas chromatography or gas chromatography− mass spectrometry by baseline integration of the peaks that correspond to the compounds that belongs to the group.31 In these cases, the regression model shown in eq 2 is particularly useful for the determination of the freely dissolved alkylatedPAH concentrations. The equation provides a representative KPE value for a certain alkylated-PAH homologue group that agrees well with the experimental values simply by taking the number of aromatic and aliphatic carbons as variables. The variation of the KPE values for alkylated-PAHs within a homologue group are found to be relatively minor. At equilibrium, the sum of aqueous concentrations for the alkylated-PAH group j (∑iCjw,i) can be estimated from the sum of PE concentrations for the group (∑iCjPE,i) by

∑ Cw ,i j = ∑ CPE,i j/KPE,rep j i

i j

j

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(1) Landrum, P. F.; Reinhold, M. D.; Nihart, S. R.; Eadie, B. J. Predicting the bioavailability of organic xenobiotics to Pontoporeia hoyi in the presence of humic and fulvic materials and natural dissolved organic matter. Environ. Toxicol. Chem. 1985, 4 (4), 459−467. (2) Di Toro, D. M.; Zarba, C. S.; Hansen, D. J.; Berry, W. J.; Swartz, R. C.; Cowan, C. E.; Pavlou, S. P.; Allen, H. E.; Thomas, N. A.; Paquin, P. R. Technical basis for establishing sediment quality criteria for nonionic organic chemicals using equilibrium partitioning. Environ. Toxicol. Chem. 1991, 10 (12), 1541−1583. (3) Procedures for the Derivation of Equilibrium Partitioning Sediment Benchmarks (ESBs) for the Protection of Benthic Organisms: PAH Mixtures, EPA/600/R-02/013; Office of Research and Development; U.S. Environmental Protection Agency: Washington, DC, 2003. (4) Kraaij, R.; Mayer, P.; Busser, F. J.; van het Bolscher, M.; Seinen, W.; Tolls, J.; Belfroid, A. C. Measured pore-water concentrations make equilibrium partitioning work–a data analysis. Environ. Sci. Technol. 2003, 37 (2), 268−274. (5) ter Laak, T. L.; Barendregt, A.; Hermens, J. L. Freely dissolved pore water concentrations and sorption coefficients of PAHs in spiked, aged, and field-contaminated soils. Environ. Sci. Technol. 2006, 40 (7), 2184−2190. (6) Hawthorne, S. B.; Azzolina, N. A.; Neuhauser, E. F.; Kreitinger, J. P. Predicting bioavailability of sediment polycyclic aromatic hydrocarbons to Hyalella azteca using equilibrium partitioning, supercritical fluid extraction, and pore water concentrations. Environ. Sci. Technol. 2007, 41 (17), 6297−6304. (7) Lohmann, R.; Muir, D. Global Aquatic Passive Sampling (AQUAGAPS): Using passive samplers to monitor POPs in the waters of the world. Environ. Sci. Technol. 2010, 44 (3), 860−864. (8) Hawthorne, S. B.; Grabanski, C. B.; Miller, D. J. Measured partitioning coefficients for parent and alkyl polycyclic aromatic hydrocarbons in 114 historically contaminated sediments: Part 1. KOC values. Environ. Toxicol. Chem. 2006, 25 (11), 2901−2911. (9) ter Laak, T. L.; Agbo, S. O.; Barendregt, A.; Hermens, J. L. Freely dissolved concentrations of PAHs in soil pore water: Measurements via solid-phase extraction and consequences for soil tests. Environ. Sci. Technol. 2006, 40 (4), 1307−1313. (10) Fernandez, L. A.; MacFarlane, J. K.; Tcaciuc, A. P.; Gschwend, P. M. Measurement of freely dissolved PAH concentrations in sediment beds using passive sampling with low-density polyethylene strips. Environ. Sci. Technol. 2009, 43 (5), 1430−1436. (11) Booij, K.; Hoedemaker, J. R.; Bakker, J. F. Dissolved PCBs, PAHs, and HCB in pore waters and overlying waters of contaminated harbor sediments. Environ. Sci. Technol. 2003, 37 (18), 4213−4220. (12) Hawthorne, S. B.; Miller, D. J.; Grabanski, C. B. Measuring low picogram per liter concentrations of freely dissolved polychlorinated biphenyls in sediment pore water using passive sampling with polyoxymethylene. Anal. Chem. 2009, 81 (22), 9472−9480. (13) Hawthorne, S. B.; Jonker, M. T.; van der Heijden, S. A.; Grabanski, C. B.; Azzolina, N. A.; Miller, D. J. Measuring picogram per liter concentrations of freely dissolved parent and alkyl PAHs (PAH34), using passive sampling with polyoxymethylene. Anal. Chem. 2011, 83 (17), 6754−6761. (14) Huckins, J. N.; Tubergen, M. W.; Manuweera, G. K. Semipermeable-membrane devices containing model lipid - a new approach to monitoring the bioavailability of lipophilic contaminants and estimating their bioconcentration potential. Chemosphere 1990, 20 (5), 533−552. (15) MacRae, J. D.; Hall, K. J. Comparison of methods used to determine the availability of polycyclic aromatic hydrocarbons in marine sediment. Environ. Sci. Technol. 1998, 32 (23), 3809−3815.

(7) −1

where Cw,i and CPE,i are the aqueous concentration (mol L ) and PE concentration (mol kg−1) for isomer i in group j, respectively, and KPE,rep j is the representative PE−water partitioning coefficient (L kg−1) for group j. On the other hand, when an individual parent- or alkylatedPAH compound is studied, we recommend using the experimental value of KPE. Equation 2 does not account for the compound structure and conformation, which results in prediction errors. For example, the equation predicts the same KPE value for benzo[b]fluoranthene and perylene (C20H12), for which experimental values are different by a factor of 1.7. For PCBs, the KPE values for almost all PCB congeners that significantly occur in the aquatic environment are reported in this study. In rare cases where the experimental values are not available for specific PCB congeners, we recommend using the correlation with the octanol−water partitioning coefficients (Kow) for the estimation of the KPE values.

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ASSOCIATED CONTENT

S Supporting Information *

Numerical values of PE−water partitioning coefficients (KPE) determined in DI and saline water and molecular properties for the suite of PAHs and PCBs studied. This material is available free of charge via the Internet at http://pubs.acs.org.

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REFERENCES

AUTHOR INFORMATION

Corresponding Author

*Phone: 650-721-2615; fax: 650-725-9720; e-mail: luthy@ stanford.edu. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS Funding for this research was provided by the Chevron Energy Technology Company (contract number: CW786669) and the Department of Defense Strategic Environmental Research and Development Program (SERDP), ER-1552 Add-On. Tuition 6949

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Environmental Science & Technology

Article

(16) Rusina, T. P.; Smedes, F.; Klanova, J.; Booij, K.; Holoubek, I. Polymer selection for passive sampling: A comparison of critical properties. Chemosphere 2007, 68 (7), 1344−1351. (17) Mayer, P.; Vaes, W. H. J.; Wijnker, F.; Legierse, K. C. H. M.; Kraaij, R.; Tolls, J.; Hermens, J. L. M. Sensing dissolved sediment porewater concentrations of persistent and bioaccumulative pollutants using disposable solid-phase microextraction fibers. Environ. Sci. Technol. 2000, 34 (24), 5177−5183. (18) Tomaszewski, J. E.; Luthy, R. G. Field deployment of polyethylene devices to measure PCB concentrations in pore water of contaminated sediment. Environ. Sci. Technol. 2008, 42 (16), 6086− 6091. (19) Lohmann, R. Critical review of low-density polyethylene’s partitioning and diffusion coefficients for trace organic contaminants and implications for its use as a passive sampler. Environ. Sci. Technol. 2012, 46 (2), 606−618. (20) Adams, R. G. Polyethylene devices and the effects of sediment resuspension on the cycling of PAHs and PCBs in the lower Hudson Estuary. Ph.D. Dissertation, Massachusetts Institute of Technology: Cambridge, MA, 2002. (21) Fernandez, L. A.; Harvey, C. F.; Gschwend, P. M. Using performance reference compounds in polyethylene passive samplers to deduce sediment porewater concentrations for numerous target chemicals. Environ. Sci. Technol. 2009, 43 (23), 8888−8894. (22) Cho, Y. M.; Werner, D.; Choi, Y.; Luthy, R. G. Long-term monitoring and modeling of the mass transfer of polychlorinated biphenyls in sediment following pilot-scale in-situ amendment with activated carbon. J. Contam. Hydrol. 2012, 129−130, 25−37. (23) Hale, S. E.; Werner, D. Modeling the mass transfer of hydrophobic organic pollutants in briefly and continuously mixed sediment after amendment with activated carbon. Environ. Sci. Technol. 2010, 44 (9), 3381−3387. (24) Smedes, F.; Geertsma, R. W.; van der Zande, T.; Booij, K. Polymer-water partition coefficients of hydrophobic compounds for passive sampling: Application of cosolvent models for validation. Environ. Sci. Technol. 2009, 43 (18), 7047−7054. (25) Boehm, P. D.; Douglas, G. S.; Burns, W. A.; Mankiewicz, P. J.; Page, D. S.; Bence, A. E. Application of petroleum hydrocarbon chemical fingerprinting and allocation techniques after the Exxon Valdez oil spill. Mar. Pollut. Bull. 1997, 34 (8), 599−613. (26) Hawthorne, S. B.; Miller, D. J.; Kreitinger, J. P. Measurement of total polycyclic aromatic hydrocarbon concentrations in sediments and toxic units used for estimating risk to benthic invertebrates at manufactured gas plant sites. Environ. Toxicol. Chem. 2006, 25 (1), 287−296. (27) Zeigler, C. D.; Robbat, A. Comprehensive profiling of coal tar and crude oil to obtain mass spectra and retention indices for alkylated PAH shows why current methods err. Environ. Sci. Technol. 2012, 46 (7), 3935−3942. (28) Adams, R. G.; Lohmann, R.; Fernandez, L. A.; MacFarlane, J. K.; Gschwend, P. M. Polyethylene devices: Passive samplers for measuring dissolved hydrophobic organic compounds in aquatic environments. Environ. Sci. Technol. 2007, 41 (4), 1317−1323. (29) Jonker, M. T.; Koelmans, A. A. Polyoxymethylene solid phase extraction as a partitioning method for hydrophobic organic chemicals in sediment and soot. Environ. Sci. Technol. 2001, 35 (18), 3742−3748. (30) Choi, Y.; Cho, Y. M.; Gala, W. R.; Luthy, R. G. Measurement and modeling of activated carbon performance for the sequestration of parent- and alkylated-polycyclic aromatic hydrocarbons in petroleumimpacted sediments. Environ. Sci. Technol. 2013, 47 (2), 1024−1032. (31) Wang, Z.; Li, K.; Lambert, P.; Yang, C. Identification, characterization and quantitation of pyrogenic polycylic aromatic hydrocarbons and other organic compounds in tire fire products. J. Chromatogr., A 2007, 1139 (1), 14−26. (32) Cho, Y. M.; Smithenry, D. W.; Ghosh, U.; Kennedy, A. J.; Millward, R. N.; Bridges, T. S.; Luthy, R. G. Field methods for amending marine sediment with activated carbon and assessing treatment effectiveness. Mar. Environ. Res. 2007, 64 (5), 541−555.

(33) Hawker, D. W.; Connell, D. W. Octanol water partitioncoefficients of polychlorinated biphenyl congeners. Environ. Sci. Technol. 1988, 22 (4), 382−387. (34) Verification and Validation of the SPARC Model, EPA/600/R-03/ 033; Office of Research and Development, U.S. Environmental Protection Agency: Research Triangle Park, NC, 2003. (35) Arp, H. P.; Breedveld, G. D.; Cornelissen, G. Estimating the in situ sediment-porewater distribution of PAHs and chlorinated aromatic hydrocarbons in anthropogenic impacted sediments. Environ. Sci. Technol. 2009, 43 (15), 5576−5585. (36) Abraham, M. H.; Ibrahim, A.; Zissimos, A. M. Determination of sets of solute descriptors from chromatographic measurements. J. Chromatogr., A 2004, 1037 (1−2), 29−47. (37) van Noort, P. C.; Haftka, J. J.; Parsons, J. R. Updated Abraham solvation parameters for polychlorinated biphenyls. Environ. Sci. Technol. 2010, 44 (18), 7037−7042. (38) Booij, K.; Hofmans, H. E.; Fischer, C. V.; Van Weerlee, E. M. Temperature-dependent uptake rates of nonpolar organic compounds by semipermeable membrane devices and low-density polyethylene membranes. Environ. Sci. Technol. 2003, 37 (2), 361−366. (39) Cornelissen, G.; Pettersen, A.; Broman, D.; Mayer, P.; Breedveld, G. D. Field testing of equilibrium passive samplers to determine freely dissolved native polycyclic aromatic hydrocarbon concentrations. Environ. Toxicol. Chem. 2008, 27 (3), 499−508. (40) Hale, S. E. Reducing the Availability of Sediment Pollutants through Activated Carbon Amendments. Ph.D. Dissertation, Newcastle University, Newcastle, U.K., 2009. (41) Perron, M. M.; Burgess, R. M.; Ho, K. T.; Pelletier, M. C.; Friedman, C. L.; Cantwell, M. G.; Shine, J. P. Development and evaluation of reverse polyethylene samplers for marine phase II wholesediment toxicity identification evaluations. Environ. Toxicol. Chem. 2009, 28 (4), 749−758. (42) Hale, S. E.; Martin, T. J.; Goss, K. U.; Arp, H. P.; Werner, D. Partitioning of organochlorine pesticides from water to polyethylene passive samplers. Environ. Pollut. 2010, 158 (7), 2511−2517. (43) Bucheli, T. D.; Gustafsson, O. Soot sorption of non-ortho and ortho substituted PCBs. Chemosphere 2003, 53 (5), 515−522. (44) Endo, S.; Escher, B. I.; Goss, K. U. Capacities of membrane lipids to accumulate neutral organic chemicals. Environ. Sci. Technol. 2011, 45 (14), 5912−5921. (45) Niederer, C.; Goss, K. U. Effect of ortho-chlorine substitution on the partition behavior of chlorophenols. Chemosphere 2008, 71 (4), 697−702. (46) Goss, K. U.; Arp, H. P.; Bronner, G.; Niederer, C. Nonadditive effects in the partitioning behavior of various aliphatic and aromatic molecules. Environ. Toxicol. Chem. 2009, 28 (1), 52−60. (47) Sacks, V. P.; Lohmann, R. Development and use of polyethylene passive samplers to detect triclosans and alkylphenols in an urban estuary. Environ. Sci. Technol. 2011, 45 (6), 2270−2277. (48) Endo, S.; Pfennigsdorff, A.; Goss, K. U. Salting-out effect in aqueous NaCl solutions: Trends with size and polarity of solute molecules. Environ. Sci. Technol. 2012, 46 (3), 1496−1503.

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NOTE ADDED AFTER ASAP PUBLICATION There was an error in equation 5 in the version of this paper published April 3, 2013. The correct version published April 23, 2013.

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