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. 2022 Jun 27;56(14):10020–10029. doi: 10.1021/acs.est.2c00312

Equilibrium Porewater Measurement of PCBs and PAHs Using Direct Water Extraction and Comparison with Passive Sampling

Songjing Yan 1, Mandar Bokare 1, Upal Ghosh 1,*
PMCID: PMC9302434  PMID: 35759616

Abstract

graphic file with name es2c00312_0005.jpg

The freely dissolved concentration of hydrophobic pollutants in sediment porewater (Cpw) is a critical driver for exposure to aquatic organisms, bioaccumulation, toxicity, and flux across interfaces. In this research, we compared direct porewater extraction and passive sampling for Cpw measurements of a range of PCBs and PAHs in field-collected sediments. The direct water extraction method provided accurate quantification of Cpw for low to moderately hydrophobic PCB and PAH compounds (log Kow < 6.5) that compared well with independent measurements performed using four passive sampling methods. Direct water extraction was adequate to assess narcosis toxicity of PAHs to benthic organisms that is driven by the concentrations of low to moderately hydrophobic PAHs (naphthalene to chrysene), even for a hypothetical sediment that had a tenth of the PAH concentrations of the study sediments and was assessed to be nontoxic. Prediction of PCB bioaccumulation in benthic organisms agreed within 50% for all measurement methods, but it was apparent that for less contaminated sediments, the direct water extraction method would likely have detection limit challenges, especially for the strongly hydrophobic PCBs. To address the uncertainty of the Cpw measurement of the strongly hydrophobic compounds and naphthalene, a new extrapolation approach is demonstrated that can be applicable for both direct water extraction and passive sampling methods.

Keywords: sediment porewater, direct water extraction, passive sampling, PCBs, PAHs

Short abstract

Porewater concentration of hydrophobic pollutants in sediments drives environmental transport and exposure. This study compares direct measurement with passive sampling and demonstrates extrapolation for difficult to measure compounds.

Introduction

Hydrophobic organic compounds (HOCs) such as polychlorinated biphenyls (PCBs) and polycyclic aromatic hydrocarbons (PAHs) are often the key drivers of human health and ecological risk at sediment sites contaminated with legacy pollutants.1 The freely dissolved concentrations of HOCs in sediment porewater (Cpw) are used as the critical metric to assess bioavailability, toxicity, and flux of PCBs and PAHs.2−6 Two main approaches exist for the measurement of Cpw in sediments: 1) direct extraction of porewater or water equilibrated with the sediments7,8 and 2) passive sampling of the sediment porewater using a third, well characterized phase,9,10 that is often a polymer brought to equilibrium with the sediment/porewater.

A major challenge with direct porewater extraction is the separation of colloidal particles that can greatly bias the measurement of dissolved concentrations of strongly hydrophobic compounds. Centrifugation to remove colloidal particles is feasible but is often limited by the large volume requirements for analytical detection. Filtration to remove suspended particles and colloids faces the drawback of potential sorption of some dissolved constituents in the filtration system and breakthrough of very fine particles. Work by Ghosh et al.7 used alum flocculation followed by centrifugation to remove colloids and measure freely dissolved PCBs in sediment equilibrated water. Subsequent work adapted this approach for the measurement of PCBs and PAHs in a range of field samples.8,11−14 However, the low aqueous concentrations of strongly hydrophobic pollutants make it difficult to extract enough water volume to allow detection in the sub-ng/L range of concentration often required for regulatory goals.15

Passive sampling has been applied in multiple research projects as a practical approach to determine PAH and PCB porewater concentrations.16−22 By using a hydrophobic polymer, passive sampling devices can accumulate HOCs from the environment to levels that are adequate for achieving low concentration regulatory requirements. Low-density polyethylene (PE) sheets and solid-phase microextraction (SPME) polydimethylsiloxane (PDMS) fibers are two of the most commonly used passive sampling devices that have been used to determine Cpw in several recent research projects.9,19,23−25 While there are several advantages of using passive samplers for sediment porewater measurement, there can be several disadvantages including the need for correction of nonequilibrium in the polymer that can get increasingly error-prone with increasing compound hydrophobicity26 and high detection limits for less hydrophobic compounds due to weak polymer sorption.

Very few studies have performed intercomparisons of direct water measurement and passive sampling for sediment porewater and determined the real trade-offs for each method across the hydrophobicity spectrum for PCBs and PAHs. Work by Gschwend et al.27 compared three passive sampling polymer-based porewater PCB concentrations with independently measured concentrations using an air-bridge approach and centrifugation/direct water extraction. They showed a general agreement of the measurements within a factor of 2 across methods. However, that study measured only tetra- and higher chlorinated PCB congeners and did not use corrections for nonequilibrium using performance reference compounds (PRCs).28 Similar work by Fernandez et al. found measurements of PAHs made using polyethylene strips, and those made using liquid–liquid extraction of porewater agreed within a factor of 2.29 Measurement of freely dissolved PAH and PCB concentrations in a water column using various types of passive samplers such as PE, PDMS, and semipermeable membrane devices and those measured using grab sampling by Burgess et al. showed similar agreements within a factor of 2.30 The comparison of direct surface water extraction and the polyoxymethylene passive sampling approach were in good agreement, within a factor of 2–3 in research by Cornelissen et al.31 Agreement with a factor of 2 to 4 for low-molecular-weight PAHs and PCBs between directly extracted water column samples and PE measurements was also reported by Lohmann et al.32 Comparison studies are required to frame the conditions where direct water sampling may be feasible and advantageous to use and where passive sampling is the most appropriate approach to achieve project specific data quality objectives. Comparison studies are also needed to address any remaining lack of confidence in and limited regulatory acceptance of the passive sampling technologies.10,33,34

The present study measured the freely dissolved porewater concentrations of 20 PCB congeners and 16 PAH compounds in contaminated sediment using direct water extraction and passive sampling with two different polymers deployed in mixed and unmixed conditions. Passive sampling was evaluated under unmixed exposures and “actively” mixed exposures on a rolling table as parallel experiments compared with the direct water extraction. The results from direct water extraction were corrected for association with dissolved organic carbon (DOC) before comparison with passive sampling results that were corrected for nonequilibrium using PRCs when necessary. A new extrapolation approach was tested for estimating concentrations of compounds that were difficult to measure by each of the methods. Finally, the results were used to compute bioaccumulation in benthic organisms assuming equilibrium between organism lipids and HOCs in porewater and to calculate PAH toxicity using the narcosis model35 to assess the differences across measurement methods in typical data use scenarios.

Materials and Methods

Chemicals and Materials

A low-density polyethylene (PE) sheet with a thickness of 25 μm (1 mil) was purchased from Husky (Bolton, ON). The solid-phase microextraction polydimethylsiloxane (SPME-PDMS) fiber, with a 35 μm thickness coating layer and a core diameter of 486 μm, was obtained from Polymicro Technologies (Phoenix, AZ). Sodium hydroxide (NaOH), hydrogen chloride (HCl), and alum solution (10%, w/v) were purchased from Fisher Scientific (Waltham, MA). PAH standards were purchased from Ultra Scientific (now Agilent, Santa Clara, CA) and Crescent Chemicals (Islandia, NY), while deuterated PAHs used as internal standards and surrogate standards were purchased from AccuStandard (New Haven, CT) and Cambridge Isotope Laboratories (Tewksbury, MA). PCB calibration standards were purchased from AccuStandard, and PCB surrogates and internal standards were purchased from Ultra Scientific. Performance Reference Compounds (PRCs) for PAHs and PCBs, labeled with 13C, were purchased from Cambridge Isotope Laboratories. Pesticide grade solvents, hexane, acetone, and methanol, were purchased from Fisher Scientific (Waltham, MA). Deionized (DI) water (resistance between 17 and 18 MΩ) was used throughout the study.

Sediment

Sediment samples contaminated with PAHs, PCBs, and metals originated from Indiana Harbor, Chicago, IL, and were stored at 4 °C at the U.S. Army Engineer Research Development Center (ERDC) Environmental Laboratory in Vicksburg, MS.36 Sediments were homogenized at ERDC, apportioned into 18 1 L glass jars, and shipped in a cooler to the University of Maryland, Baltimore County (UMBC). Sediments were stored at 4 °C until use. Six of the jars were used for direct water extraction, while the remaining 12 jars were used for unmixed and mixed passive sampling.

Direct Water Extraction

The direct water extraction method was adapted from Ghosh et al.7 and Khalil et al.8 Wet sediment (860 g) was transferred to 2 L glass jars with about 800 mL of DI water to provide sufficient water volume for analytical measurements. The 2 L jars with sediment slurry were placed on an orbital shaker at the 110 rpm setting for 28 days for sediment-water equilibrium at room temperature. The samples were allowed to settle for 1 day following the mixing period. About 6 mL of alum solution to achieve 0.001 M was added slowly to each jar and gently mixed into the overlying water to flocculate the remaining suspended colloids. A few drops of a 1 M NaOH solution were used to bring the pH of the overlying water back to neutral. The supernatant water was mixed slowly without disturbing the settled sediments using a glass pipet, and the alum flocs formed in minutes (Figure S1A, Supporting Information). The jars were capped, and the alum floc was allowed to settle down by gravity for 6 h to produce a clear supernatant. The supernatant water was then transferred to a 1 L separatory funnel using a 100 mL glass pipet (Figure S1B, Supporting Information). The water sample volume recovered for direct water extraction was 800 ± 40 mL. Twenty milliliters of hexane was used to rinse the pipet and extract the PAH and PCB compounds from water. To ensure complete extraction, a total of three extraction steps were performed. For each extraction, the collected water sample and hexane were placed in a separatory funnel and shaken by hand for at least 1 min. After allowing 4 min for separation of water and hexane layers, the water layer was drained from the bottom of the funnel, after which the hexane solution was also drained from the bottom of the funnel. The collected hexane solution was then concentrated to 1 mL under nitrogen blowdown and subjected to silica gel cleanup based on EPA SW-846 3630C37 prior to analysis. The volume of collected water was measured for each sample and used to calculate the final concentrations.

PE and PDMS Preparation for Passive Sampling

The PE sheets were cut into 2 cm × 5 cm (approximately 25 mg) for PCB analysis and 1 cm × 2 cm (approximately 5 mg) for PAH analysis. The PDMS fibers were cut into 5 cm lengths; and a 10 cm total length was extracted for each PCB sample, while 2 cm lengths were used for PAH measurement. The different lengths and masses of passive samplers were designed to accommodate different anticipated concentrations of PAHs and PCBs (based on concentrations in sediments). Prior to use, the passive samplers were cleaned with hexane, methanol, and acetone for 24 h to eliminate any residual oligomers and contaminants. The samplers were then soaked in deionized water overnight to remove the solvent. PRCs were used to assess the attainment of equilibrium between the passive samplers and sediment. The cleaned PE strips were loaded with PRCs in DI water spiked at a known concentration of the PRCs.38 A mixture of methanol/water (20/80 in v/v) was used as a PRC loading solution for PDMS fibers. Four 13C6 labeled PAHs (phenanthrene, fluoranthene, chrysene, and indeno[1,2,3-cd]pyrene) were selected as PAH PRCs. Six 13C12 labeled PCBs (PCB #37, PCB #47, PCB #54, PCB #111, PCB #138, and PCB #178) were used as PRCs for PCB congeners. The passive samplers were preloaded with the PRCs for 28 days on an orbital shaker for equilibrium. After PRC impregnation, the PE strips and PDMS fibers were wiped with Kimwipes to remove the loading solution before deployment. Six pieces of PE samplers (1 cm × 2 cm) and six pieces of PDMS samplers (2 cm) loaded with PAH PRCs were used to measure the initial PAH PRC concentrations. Similarly, six PRC-impregnated PE (2 cm × 5 cm) and PDMS (10 cm) samplers were measured for initial PCB PRC concentrations. The initial measurements were also used to check for any background contamination.

Passive Sampler Exposure and Extraction

The PRC-impregnated PE strips were introduced directly into each homogenized sediment sample. To allow for easier retrieval without damaging the fiber, the PDMS fibers were enveloped in a metal mesh bag before deployment. All the sediment sample vessels were covered with aluminum foil to minimize potential photodegradation of the target compounds and PRCs. Six sampling jars were set aside for “unmixed” sampling, while the remaining six jars were actively “mixed” on an orbital shaker. Each sampling jar contained two PE samples and two PDMS samples for PAH and PCB analysis. Samplers were allowed to equilibrate for 2 months at room temperature. At the conclusion of the exposure period, the PE strips and PDMS fibers were retrieved, rinsed with deionized water, and cleaned with a damp lint-free tissue to remove sediment residue. The 5 cm PDMS fibers were sectioned into small pieces (about 0.5 cm), pooled together in a 2 mL GC vial with prefilled hexane, and extracted at 4 °C for at least 12 h prior to measurement to ensure complete extraction. PE strips were extracted with 30 mL of hexane on an orbital shaker for 24 h, and this step was repeated three times. After extraction, the PE strips were dried and weighed. The extraction solvent was concentrated by nitrogen blowdown to 1 mL for PE samples and 250 μL for PDMS samples. PCB extracts from PE were cleaned using deactivated silica gel following a modified procedure based on EPA SW-846 3630C.37 The PRC initial samples were analyzed in the same way and measured prior to exposure. The passive sampling study was part of a larger interlaboratory validation study of passive sampling performed in collaboration with other research and commercial laboratories;36 however, the results presented here for comparison with direct water extraction are from our laboratory only.

Chemical Analysis

Extraction surrogates (PCB #14, PCB #65, and phenanthrene-d10) were added prior to the extraction of water samples and passive samplers. PCB #30 and PCB #204 were used as internal standards and added to all samples before analysis. Both the surrogates and the internal standard compounds do not typically occur in contaminated sediments and have been used as internal standards in several previous studies.26,39,40 PCB analysis was performed on a Shimadzu TQ8030 gas chromatograph–mass spectrometer (GCMS) equipped with a fused silica capillary column (Rtx-5MS, 60 m × 0.25 mm id, 0.25 μm film thickness). EPA method SW8270D41 was used to determine the PCB surrogates, PRCs, and 20 target PCB congeners, which were NOAA 18 congeners (PCB#8, #18, # 28, #44, #52, #66, #101, #105, #118, #128, #138, #153, #170, #180, #187, #195, #206, #209) plus PCB #126 and PCB #169. Peak identification and integration were performed in the Selected Ion Monitoring (SIM) mode. Detection limits for individual PCB congeners in samples range from 0.4–5 ng in the final extract. A five-point calibration was performed with R2 > 0.989 for all compounds.

The PAH concentrations were determined by GCMS in the Multiple Reaction Monitoring (MRM) mode using a method modified from EPA Method 8270. Four deuterated PAH compounds (purchased from Cambridge Isotope Laboratories) were used as internal standards: 1-fluoronaphthalene, p-terphenyl-d14, benzo[a]pyrene-d12, and dibenzo[a,h]anthracene-d14. Deuterated phenanthrene was used as a surrogate spike. In addition to the PRCs and surrogate compounds, the concentrations of 16 priority pollutant PAHs (Table S1) were quantified in this study. Detection limits ranged from 0.1 to 1 ng in the final extract. A six-point calibration was performed with R2 > 0.986 for all PAH compounds.

Dissolved organic carbon in direct water extraction samples was measured using a Shimadzu Total Organic Carbon Analyzer (TOC-V CPH model) using the Non-Purgeable Organic Carbon (NPOC) mode and detection performed with a nondispersive infrared (NDIR) detector. Forty milliliters of supernatant water after flocculation and the settling step was collected for DOC measurement. After collection, samples were immediately filtered with 0.45 μm nylon filters to remove any remaining suspended solids, acidified with HCl to a pH ∼ 3, and then stored in a refrigerator until analysis. All samples were analyzed within 28 days of collection. The DOC analyzer was calibrated before each set of measurements with standard solutions of potassium hydrogen phthalate, with concentrations of 1, 5, 10, 20, and 50 mg/L carbon. A DOC reference standard (ERA Organic Carbon, PotableWatR), obtained from ERA, a Waters Company, was run with every set of measurements. The sample results were considered valid only if the DOC of the reference standard was within the Quality Control (QC) performance acceptance limits (6.88 to 8.99 mg/L DOC).

Results and Discussion

Direct Water Extraction Results and Koc Estimation

The PCB surrogate recoveries were 85 ± 4% for surrogate PCB #14 and 92 ± 6% for surrogate PCB #65. The direct water extracted concentration is a combination of the freely dissolved and the dissolved organic carbon (DOC) associated concentrations. Therefore, the measurements were corrected for DOC-association using measured DOC concentrations in porewater and estimated partition coefficients for DOC using eq 1

graphic file with name es2c00312_m001.jpg 1

where KDOC is the partition coefficient of the chemical between the DOC and freely dissolved phases. The partition coefficients for PCBs and PAHs, between DOC and sediment porewater, were estimated by a KDOC-Kow correlation (log KDOC = 0.99 × log Kow – 0.88) from Burkhard et al.42 Octanol–water partition coefficients (Kow) from Hawker et al.43 and Ma et al.44 were used for PCBs and PAHs, respectively. The average DOC concentration in all extracted porewater samples was 35 ± 1.2 mg/L.

One limitation of the direct water extraction method for determining porewater concentrations is that the required water volume is usually large to meet the detection limits, especially for the high-MW strongly hydrophobic compounds with low aqueous concentrations. For direct water extraction, 6 out of 14 detected PCBs and 4 out of 16 detected PAHs were below our lowest calibration levels.

To address this challenge, we evaluated a relationship between the organic carbon normalized partition coefficient (Koc) and the Kow based on the detected compounds and extrapolated for the high-MW compounds and Nap (see Table S1 for PAH abbreviations), assuming that a single parameter linear free energy relationship adequately describes partitioning for the same class of compounds for a specific sediment. Koc values for low-MW compounds can be estimated by sediment concentration Csed, fraction organic carbon in sediment foc, and measured Cpw. Csed values for the same batch of sediment sample were taken from Michalsen et al.36 and are provided in Table S1. The foc value for the study sediments was 0.07.

Log Koc was plotted against log Kow for PCBs (Figure S2A) and PAHs (Figure S2B) and compared to previously published Koc-Kow correlations presented in Hansen et al.,45 Baker et al.,46 and Nguyen et al.47 The calculated Koc values were higher than the literature estimated ones for both PAHs and PCBs potentially from the presence of black carbon (not measured), as seen in other studies.27 The extrapolated Cpw for Nap, the high-molecular-weight PAHs, and PCBs along with the estimated DOC-associated concentrations are shown in Figure 1. Figure S3 shows the % freely dissolved and % associated with DOC, along with total concentrations for each compound. The extrapolated Nap concentration was an order of magnitude higher than the measured data (Tables S2 and S6), which is potentially due to loss of this volatile compound during measurement. Therefore, Nap concentration was also extrapolated using the Koc-Kow correlations.

Figure 1.

Figure 1

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Freely dissolved porewater concentrations (Cpw) of PCBs (A) and PAHs (B) measured by the direct water extraction approach with dissolved organic carbon corrections. The red bars indicate concentrations of compounds directly measured, while the blue bars indicate estimated concentrations based on an extrapolated Koc-Kow relationship. The DOC-associated fraction is visualized in percent plots in Figure S3.

The DOC correction was found to have an increasing impact with increasing hydrophobicity for both PCBs and PAHs, with DOC-associated concentrations accounting for less than 5% for the lower MW PAHs (Nap, Acy, Ace, Flo) and more than 90% for the BbF and BkF (Figure 1, Tables S2 and S3). The DOC-associated concentration portion for PCBs also increased from 32% (PCB #8) to 95% (PCB #118) with increasing MW. Dichloro- (PCB #8, 10.1 ng/L) and trichloro- (PCB #18, 6.0 ng/L; PCB #28, 6.6 ng/L) biphenyls were the most abundant PCBs in the equilibrated water for which the impact of DOC was small compared to that for the more chlorinated homologues. The sediment also had less high-MW compounds, and the concentrations decreased with an increase in MW.

Due to the high concentration of PAHs in the extracted porewater samples, the extracts had to be diluted before analysis, and PAH surrogates fell below the instrument detection limits. However, the good recoveries of the PCB surrogates described above provide confidence of adequate recovery in the simultaneously extracted PAHs. As shown in Figure 1, the low-MW compounds dominated in the water phase with Nap being the most abundant in the porewater. Total PAH concentration in porewater was not substantially impacted by the DOC correction.

Passive Sampling Results

Passive samplers exhibited good surrogate recoveries in both mixed and unmixed samplings. For surrogates PCB #14 and PCB #65, the recoveries were within 84% to 97% in the unmixed sampling and 91% to 100% in the mixed sampling. The measured extract concentrations were normalized by the PE mass or by the PDMS volume (Csampler), converted to equilibrium concentrations measured in the respective polymer, and further converted to corresponding PCB aqueous concentrations (Cpw), as shown in eq 2

graphic file with name es2c00312_m002.jpg 2

where feq is the fraction to equilibrium value, estimated from loss of PRCs from the sampler. The analyte-specific feq values for the unmixed PE exposures were calculated based on the concentration of PRCs in the sampler before and after deployment, assuming that loss of PRCs during deployment followed a first-order decay, using the method from Perron et al.48 Additional details on feq calculations are presented in the Supporting Information. For PDMS, the PRC correction method from Yan et al.19 was used. This method includes both the internal and external resistances in the PDMS and has been applied previously for PDMS data from field deployments.49

Polymer–water partitioning coefficients (KSampler-Water) were estimated using the regression relationship with Kow based on correlations from Ghosh et al.16 Log Kow values were taken from Hawker et al.43 for PCB congeners and from Ma et al.44 for PAHs. PRC corrections were performed for all samples except for mixed PE samples. In the mixed PE sampling, the final PCB PRC concentrations were below our detection limits after a two-month deployment, which indicates that the compounds reached equilibrium and the feq was 1. The estimated feq values for the unmixed PE samplers were similar to PDMS samplers for low-MW PCB congeners (log Kow < 6) but deviated for higher MW PCB congeners (log Kow > 6). The PCB concentrations determined by PE and PDMS passive sampling are shown in Figure S4A (and Table S4). The replicates gave (n = 6) consistent concentrations for PCBs with an average coefficient of variation (CV) less than 15% for all sampling batches (Table S5). The lower chlorinated PCB congeners had relatively higher concentrations and less CV among measurement methods compared to the higher chlorinated PCBs; for example, PCB #8 and #28 both have a CV value of 9%, while PCB #105 has a CV value of 24% among various passive sampling approaches. In total, 14 out of 20 PCB congeners were detected in the polymers, but 6 of them (PCB #128, #138, #153, #170, #180, and #187) were “J flagged (the value is less than the minimum calibration level)” because their extracted concentrations were 24%–64% lower than the lowest calibration level (3 ng/mL), which can lead to inaccuracy in results. Hence, the Koc-Kow relationship used for direct water extraction can also be for passive sampling technologies.

The PAH surrogate recoveries ranged from 84% to 95% for unmixed sampling and 76% to 86% for mixed sampling. The sediment porewater concentrations were determined in the same way as described above for PCBs and are shown in Figure S4B (and Table S6). Similar to PCBs, the calculated feq was 1 for the mixed samples, and equilibrium corrections were not necessary. The PDMS unmixed samplers had also obtained equilibrium with feq equal to 1 for all target compounds, while some of the high-MW PAHs (log Kow > 5) on PE samples had not reached equilibrium. The replicates (n = 6) gave consistent concentrations for PAHs as well. The average coefficient of variation (CV) was 12% in mixed samples and 18% in unmixed samples (Table S5). The measured Cpw values of PAHs were relatively high compared to PCB concentrations, with all 16 compounds detected, although several compounds were measured at a level below the lowest calibration point. The concentration of Nap determined by passive sampling was 2 orders of magnitude lower than from direct water extraction and about 1 order of magnitude lower than the lowest calibration point (RL) with large relative standard deviation. Possible reasons include weak partitioning of Nap (the least hydrophobic compound) to the passive samplers and post-exposure volatilization losses during sampler cleaning and sectioning steps. The lighter PAHs, like Acy and Flo, were the dominant compounds in the water phase, and the concentrations were 3 to 4 orders of magnitude higher than the heavier compounds like BaP, DahA, IcdP, and BghiP. The concentrations of these four high-MW PAHs determined by passive sampling were lower than half of the lowest calibration point (2 ng/mL). Similar to results for PCBs, the CV values for PAH compounds with lower concentrations were higher than those for other compounds. The porewater concentrations of these higher MW compounds also have large differences among the different passive sampling approaches (Table S6).

To estimate the target compounds with concentrations lower than our lowest calibration point, we applied our extrapolation approach to passive sampling as well. As shown in Figure S5, the log Koc values for low- to mid-molecular weight PCB and PAH compounds generally follow a linear trend with compound log Kow values and show greater scatter for the high Kow compounds: PCB #128, #138, #153, #170, #180, and #187 and PAHs BaP, DahA, IcdP, and BghiP. In addition, Nap was difficult to measure using passive sampling and deviated from the trend in the Koc-Kow relationship. After excluding the higher MW compounds and Nap, a linear relationship was derived between log Koc and log Kow for each passive sampling approach separately for PCBs and PAHs. The R2 of log Koc-log Kow linear trend lines was 0.78 ± 0.03 for PAHs and 0.93 ± 0.03 for PCBs. Subsequently, the Koc values of higher MW compounds and Nap were derived from the Koc-Kow correlations and further applied as a practical way to estimate the Cpw for the difficult to measure compounds as per eq S1 (Tables S7 and S8). Figure S4 shows the PCB and PAH porewater concentrations obtained by the four passive sampling methods compared with the direct water extraction method. For the well-detected compounds (i.e., measured concentrations above the lowest calibration point), the Cpw are the measured data; while for the high-molecular-weight compounds and Nap, the Cpw values are extrapolated using the correlation of Koc as indicated by the red boxes in Figure S4. The extrapolated PAH concentrations were generally lower than the measured values, while they were higher for PCBs. The difference among different passive sampling approaches decreases for high-MW PAHs, PCBs, and Nap for the extrapolated data, as shown in Figure S4 and Tables S7 and S8. This hybrid approach of combining measured and estimated values makes optimal use of reliably measured concentrations in sediments and site-specific partitioning characteristics based on directly measured porewater concentrations.

Comparison of Directly Measured Cpw with Passive Sampling Determined Cpw

A comparison across measurement methods in Figure 2 indicates that after extrapolating the concentrations of compounds difficult to measure, the mixed passive sampling approaches and direct water extraction method gave consistent results for all the compounds. The extrapolated Nap concentrations from passive sampling were comparable with the direct water extracted results, which is a marked improvement from the nearly 2 orders of magnitude difference observed before extrapolation in Table S6.

Figure 2.

Figure 2

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Comparison of porewater PCB and PAH concentrations measured by direct water extraction and different mixed passive sampling approaches. A: PCBs measured by the PE sampler; B: PCBs measured by the PDMS sampler; C: PAHs measured by the PE sampler; D: PAHs measured by the PDMS sampler.

Figure 2 shows individual comparisons of the direct water extracted concentrations with each of the PE and PDMS passive sampler data. Since the unmixed and mixed sampling approaches gave similar results, only mixed sampling data were used for comparison here. The unmixed data are provided in Figure S6. For PE samples, results from mixed sampling and direct extraction methods were in very good agreement for most compounds, especially the ones with low MW. The slope of the regression for the plots in Figure 2 was 1.1 for PCBs and 1.0 for PAHs. The porewater concentrations derived from PDMS samples were also consistent with the direct water extraction results, with a regression slope of 0.97 for PAHs and 1.06 for PCBs. The intercept for each of these correlations ranged from −0.30 to 0.05 log units, indicating very low bias. The average biases of measured porewater concentration via passive sampling approaches and the direct water extraction method were 1.2 ± 0.11 for various PAHs and 1.4 ± 0.29 for PCBs, as shown in Table S9. Thus, the direct porewater extraction results agreed very well with both passive sampling approaches and for a large range of PCB and PAH compounds. Therefore, any of these approaches (after required extrapolations) can be used to determine the porewater concentrations of PAHs and PCBs in equilibrium with sediments when detection limits allow.

The propagated uncertainties in the measured and extrapolated concentrations were also calculated based on uncertainties in a) partitioning coefficients, b) concentrations of analytes in the passive samplers and in the extracted porewater, c) DOC concentrations in the porewater, and d) extrapolated Koc values. Additional details of the uncertainty propagation calculations are provided in the Supporting Information. Average concentrations of analytes and their 95% confidence intervals for each method based on the propagated uncertainties are presented in Tables S10 and S11.

Predicting Bioaccumulation and Toxicity in Benthic Organisms

Measured porewater concentrations are frequently used to predict bioaccumulation of PCBs and PAHs in benthic organisms, as well as being used to interpret toxicity of PAHs to benthic organisms based on toxic units.35Figure 3 shows a comparison of the predicted bioaccumulation of PCBs and PAHs in a model benthic organism (Corganism,i, ng/g) calculated using eq 3 based on equilibrium partitioning of compound i between organism lipids and porewater (Cpw,i, ng/L). The full comparison including unmixed passive sampling results is shown in Figure S8. The organism was assumed to be a benthic worm with 0.6% lipid fraction (fLipid), and the compound-specific Kow was used as a surrogate for the lipid–water partitioning coefficient, KLW,i (L/kg).

graphic file with name es2c00312_m003.jpg 3

Figure 3.

Figure 3

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Comparison of predicted bioaccumulation of PCBs (A) and PAHs (B) in a model benthic organism across the mixed passive sampling and direct water extraction methods. Predictions are based on equilibrium partitioning of PCBs and PAHs between organism lipids and porewater. Bioaccumulation predictions include extrapolated compounds.

For PCBs (Figure 3A), PE gave the highest predictions of bioaccumulation (103 ± 8.9 and 110 ± 7.0 ng/g tissue for unmixed and mixed deployments, respectively), with average PDMS and direct water extractions giving predictions that were smaller by 21% and 35%, respectively, compared to the average of PE predictions. The differences between predictions from the different methods were small but statistically significant (p < 0.05, single factor ANOVA). Predicted bioaccumulation of PAHs across the methods ranged from 28,000 ± 1600 ng/g for PE mixed samples to 40,000 ± 1500 ng/g for PDMS mixed samples (Figure 3B). Differences between bioaccumulation predictions from direct water measurements (average: 39,000 ng/g) and from passive sampling methods (average: 35,000 ng/g) were small but statistically significant (p < 0.05, two-sample t test).

Toxic units (TU) for 16 individual PAHs were calculated using porewater concentrations from each method (Cpw,i, ng/L) and individual final chronic values (FCVi, ng/L) for water-only exposure from EPA.35 Individual TU values were used to obtain total TU (TU16PAHs) for each method, as per eq 4, where TU16PAHs > 1 indicates adverse impacts to benthic communities from porewater PAH concentrations.

graphic file with name es2c00312_m004.jpg 4

For all methods, TU16PAHs was greater than 1 indicating potential adverse effects on benthic organisms (Figures 4 and S9). The TU16PAHs values based on PE, PDMS, and direct water extraction ranged from 7.0 for mixed PE to 11 for mixed PDMS samplers. Differences between methods were statistically significant (p < 0.05, single factor ANOVA). Both PDMS methods predicted higher toxic units (10 for unmixed, 11 for mixed) than those from other methods. PE provided the lower range of measurements (7.8 for unmixed and 7.0 for mixed), while direct water extraction provided an intermediate value (9.9). Across all the methods, PAH compounds Nap to Chr contributed the most toward TU16PAHs with 99% contribution across all methods, with Nap contribution to respective TU16PAHs ranging from 75% (mixed and unmixed PE) to 80% (mixed and unmixed PDMS). Thus, passive sampling alone without the extrapolation approach presented here would have greatly underpredicted toxicity to benthic organisms for the study sediments. In applications where the goal is to use the Cpw measurement for TU calculations, the chosen analytical approach should provide accurate measurements of PAH compounds Nap to Chr also observed in earlier work by Hawthorne et al.50

Figure 4.

Figure 4

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Comparison of total and individual PAH toxic units calculated based on porewater concentrations measured by mixed passive sampling and direct water extraction methods. Toxicity calculations include extrapolated compounds.

Detection Limit Issues

The Indiana Harbor sediments used in this study had a high total PAH concentration of 1000 mg/kg and a corresponding high Cpw value of 1400 μg/L of total PAHs (average concentration from all methods). To assess the applicability of the direct water extraction method to sediments with lower PAH levels, an evaluation was performed assuming PAH concentrations in water to be 1/10th of those measured by direct water extraction. Based on an average extraction volume of 800 mL water, concentrations of 11 out of the 16 PAHs were still above the lowest calibration level, with BbF, BkF, DahA, IcdP, and BghiP falling below the lowest calibration level. The TU resulting from this assessment was 0.98, indicating that direct water extraction can be a viable method for delineating toxic sediments (TU > 1) from nontoxic ones (TU < 1). Similar evaluations assuming PAH concentrations in water to be 1/10th of those measured by PE and PDMS resulted in 3 out of 16 PAHs (Acy, DahA, and BghiP) being below the lowest calibration level for PE, while Acy and all PAHs from BaA to BghiP were below the lowest calibration point for PDMS. Toxic units predicted from these evaluations were 0.74 for PE and 1.01 for PDMS. Thus, both direct water extraction and passive sampling can be used to determine narcosis toxicity of PAHs in sediments; however, measurement of sediment concentrations and extrapolation of Nap values will likely be required for all methods.

Total PCB concentrations were comparatively much lower compared to PAHs (3.3 mg/kg for the measured 18 congeners). The total PCB concentration in porewater based on average PE measurement was 47 ng/L. Under the same scenario of sediment PCB concentrations at 1/10th the present value (resulting in a Cpw value of 4.7 ng/L), predicted concentrations for all PCB congeners were below the lowest calibration level for direct water extraction. Similar evaluations for the PE and PDMS measurements resulted in all congeners being below the lowest calibration point. Given that regulatory requirements for dissolved PCBs in water are much lower (e.g., USEPA ambient water quality criteria for human health protection at 10–5 cancer risk of 0.64 ng/L for PCBs), direct water extraction is unlikely to be adequate for the measurement of PCBs in sediment porewaters unless 10s of liters of porewater are extracted or an alternate analytical method with lower detection limits is used (e.g., a high resolution MS). For the passive sampling, PCB detection limits can be more easily improved by increasing the mass of polymer and sediment used for the measurement.

We show that the direct water extraction method provided accurate quantification of porewater for low to moderately hydrophobic PCB and PAH compounds (log Kow < 6.5) that compared well with independent measurements performed using four passive sampling methods. Direct water extraction was adequate to assess narcosis toxicity of PAHs to benthic organisms that is driven by the concentrations of low to moderately hydrophobic PAHs (Nap to Chr) but would have difficulty meeting detection limit requirements of PCBs for the use in bioaccumulation assessments. We also demonstrate an innovative but simple extrapolation approach applicable for both direct water extraction and passive sampling to estimate difficult to measure compounds based on the accurately measured concentrations of a set of similar compounds.

Acknowledgments

This study was supported by the US Department of Defense, Environmental Security Technology Certification Program, ER-201735 as part of a study to standardize passive sampling methods through an interlaboratory testing program. The authors acknowledge valuable input through the project from Dr. Mandy Michalsen (USACE), Dr. Philip M. Gschwend (MIT), and Dr. Danny Reible (Texas Tech University).

Supporting Information Available

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acs.est.2c00312.

  • Tables and figures of PCB and PAH concentrations; direct water extraction process; log Koc plotted against log Kow for PCBs and PAHs; freely-dissolved porewater concentrations (Cpw); and comparison of log Koc values of PCBs and PAHs, porewater PAH and PCB concentrations, average freely dissolved concentration of PCB congeners and PAH compounds, predicted bioaccumulation of PCBs and PAHs, and total and individual PAH toxic units (PDF)

The authors declare no competing financial interest.

Supplementary Material

es2c00312_si_001.pdf (1,007.4KB, pdf)

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es2c00312_si_001.pdf (1,007.4KB, pdf)

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