Multi-Ratio Equilibrium Passive Sampling Method to Estimate

The freely dissolved concentration (Cw,0) in the pore water and the accessible (releasable) concentration in the sediment (Cas,0) are important parame...
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Multi-Ratio Equilibrium Passive Sampling Method to Estimate Accessible and Pore Water Concentrations of Polycyclic Aromatic Hydrocarbons and Polychlorinated Biphenyls in Sediment Foppe Smedes,†,‡,* L. Alexander van Vliet,§ and Kees Booij∥ †

Masaryk University, Recetox, Kamenice 753/5-A29, 62 500, Brno, Czech Republic Deltares, PO. Box 85467, 3508 AL Utrecht, The Netherlands § Ministry of Transport, Public Works and Watermanagement, Centre for Water Management, P.O. Box 17, 8200 AA, Lelystad, The Netherlands ∥ Royal Netherlands Institute for Sea Research, P.O. Box 59, 1790 AB Texel, The Netherlands ‡

S Supporting Information *

ABSTRACT: The freely dissolved concentration (Cw,0) in the pore water and the accessible (releasable) concentration in the sediment (Cas,0) are important parameters for risk assessment. These parameters were determined by equilibrating contaminated sediments and passive samplers using largely differing sampler− sediment ratios. This method is based on the principle that incubations at low sampler/sediment ratios yield the concentration in the pore water (minor depletion of the sediment phase) and incubations at high sampler/sediment ratios yield the accessible concentration in the sediment (maximum depletion of the sediment phase). It is shown that equilibration was faster in dense suspensions and at high sampler/sediment ratios when compared to low sampler/sediment ratios. An equilibrium distribution model was used to estimate Cw,0 and Cas,0 by nonlinear least-squares regression. The method was evaluated for three sediments (harbor, estuarine, marine). Accessible concentrations of 13 PAHs were 2 (low Kow) to 10 (high Kow) times lower than the total concentrations (three sediments). By contrast, the accessible concentrations of 15 PCBs were about 1.2 times lower than the total concentrations and displayed no trend with Kow (one sediment). Implications for risk assessment and considerations for application of multi-ratio equilibrium passive sampling with other sediments are discussed.

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INTRODUCTION The freely dissolved concentration of hydrophobic contaminants in pore water of sediment has shown to be a good predictor for the concentration in benthic organisms.1 A second parameter relevant for the availability of contaminants for uptake by organisms in sediment is the accessible, or releasable, concentration in the sediment. For systems that follow the equilibrium partitioning theory, Reichenberg and Mayer2 used two parameters [1] the chemical activity of a compound and [2] the accessibility to describe (bio)availability. The difference in chemical activity represents the driving force for contaminant transport between two compartments. The freely dissolved concentration in the aqueous phase is directly proportional to the chemical activity. Accessibility can be defined as the concentration in the sediment solids that is mobile and available for exchange with the water phase. The accessible concentrations in sediments can be much smaller than the total concentrations because, in addition to amorphous organic carbon, sediment also can contain different types of black carbon such as soot and coal fragments, which have much stronger sorption properties.3,4 © 2012 American Chemical Society

The accessible concentration expresses the capacity of the sediment to buffer concentrations of contaminants in the pore water following removal, for example by diffusion to the overlying water, biodegradation, or transfer into lipid pools of benthic organisms. In analytical terms, accessibility can be quantified by determining the concentration that is released from sediment in a depletive extraction where at the end point the freely dissolved concentration (Cw) is zero or at least negligible compared to the initial Cw. If Cw is zero, the concentration of analyte still present in the sediment is defined as the inaccessible concentration. It may be inaccessible because the analyte is sorbed orders of magnitude stronger, blocked for release (e.g., incorporated in a crystal structure), or because the time scale for release is extremely long. Extractions with Tenax are often applied to estimate the fraction that is easily released Received: Revised: Accepted: Published: 510

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to the aqueous phase,5 but this method does not yield the initial concentration in the pore water. The determination of Cw of hydrophobic organic compounds in water samples by active sampling is often associated with problems of incomplete phase separation between particlebound and dissolved analytes, and adsorption to sampling equipment such as filters and glass bottles.6,7 These effects are even more pronounced for pore waters, even if it would be possible to obtain the large pore water volumes needed for satisfactory detection of the analytes. An alternative is the application of passive samplers, which accumulate compounds predominantly via the water phase.8 When equilibrium is attained, the Cw can be estimated through the sampler−water partition coefficient and the measured concentration in the sampler.9 Passive sampling for estimating Cw in sediment suspensions has been applied using solid phase micro-extraction (SPME) fibers that are coated with polydimethylsiloxane (PDMS),10,11 or using polymer strip samplers such as polyoxymethylene (POM)12 and low density polyethylene (LDPE).13 Golding et al.14 applied vials coated with ethylene vinyl acetate copolymer to estimate bioavailability of PAHs in sediments. Recently, PDMS coated vials were used to estimate Cw of PCBs in sediments.15 Most authors assume negligible depletion11 or correct for depletion, assuming a linear sorption behavior.12 Reichenberg et al.16 confirmed nondepletive conditions by varying the size of the sampler. This article further explores a method suggested earlier,17 to incubate passive samplers and sediments at different sampler− sediment ratios, creating different levels of depletion. This allows estimation of the initial, nondepleted Cw in the pore water (low sampler−sediment ratios) as well as the accessible (water exchangeable) concentration in the sediment (high sampler−sediment ratios). Silicone rubber (SR) sheets were used as passive sampler. These samplers are easy to construct, robust, reusable, and of low cost with sufficiently high partition coefficients18 and low transport resistances.19

water volume of sampler and sediment together. The contaminant will distribute accordingly. Note that the water phase is not included in the denominator because of its negligible capacity compared to the other compartments. The amounts in the left term of eq 2 can be converted to concentrations by substituting Np = mpCp, Nas,0 = msCas,0, where Cp is the concentration in the polymer and Cas,0 is the initial accessible concentration in the sediment. Following eq 1, Kasw = Cas,0/ Cw,0, where Cw,0 is the pore water concentration before exposure of the sampler. Making these substitutions in eq 2 and rearranging yields Cp =

MATERIALS AND METHODS Chemicals. Acetone, acetonitrile, hexane, methanol, aluminum oxide, and silicon dioxide were of analytical grade and obtained from Boom BV (Meppel, The Netherlands). Calibration standards were from different producers and purchased through Boom BV (Meppel, The Netherlands). Silicone rubber (SR) sheets of 0.2 cm thickness were purchased from Rubber BV (The Netherlands) and sheets of 0.05 cm thickness from Vizo (The Netherlands). The SR samplers were made by cutting strips of 4 cm wide of various lengths, depending on the experiment. Typically pieces of 4 × 12.5 cm were used, resulting in a total surface area of 100 cm2 and a weight of about 3 g. Before use, the samplers were extracted with ethyl acetate for 72 h to remove possible additives and low-molecular weight polymers. About 20−30 mg g−1 nonevaporable matter was extracted from the SR. Sediment. A 25 kg fine-grained sludge sample was taken in Harlingen Harbor (H) and used for various tests. An estuarine sediment (E) was taken from the Western Scheldt (near Hansweert) and a marine sediment (M) from the middle part of the Wadden Sea (Figure S1 of the Supporting Information). The samples E and M were sieved over 63 μm, freeze-dried, homogenized by a ball mill and stored at room temperature in the dark. The H sample was already fine grained and no sieving or freeze-drying was done. This sample was stored at 4 °C. Exposure of samplers to sediment suspensions was carried out in glass bottles of various sizes (50−1000 mL) selected to be approximately 60% full to allow adequate mixing during shaking. Shaking was performed by placing the bottle in a horizontal position on an orbital shaker (Gerhardt, RO 500,

(1)

where Kasw (L kg ) is the sediment-water partition coefficient of the accessible pool. The addition of a polymeric sampler results in a reduction of both Cas and Cw. At equilibrium, the contaminant is distributed according to Nas,0

=

mpK pw msK asw + mpK pw

(3)

■

−1

Np

mp msCas,0

after which Cw,0, Cas,0 can be obtained by linear regression. However, weighted linear regression would be required in this case because the reciprocal from the highest Cp has the smallest influence on the regression outcome. Unweighted NLS using eq 3 is more straightforward.

THEORY For developing a sampler−sediment exchange model, we define an accessible concentration of contaminants in sediment (Cas) that follows a linear sorption isotherm. The concentration in the pore water can be expressed as Cas K asw

+

The equation shows that Cp is highest when the ratio mp/ms is lowest and vice versa. Negligible depletion occurs when mp/ ms → 0. In this case, Cp = Kpw Cw,0. Maximum depletion occurs when mp/ms → ∞. In this case, mpCp = msCas,0, that is, all accessible compounds are transferred to the polymer. Cw,0 and Cas,0 can be estimated by fitting Cp as a function of mp/ms using nonlinear least-squares regression (NLS). Eq 3 can also be linearized as mp 1 1 = + Cp Cw,0K pw msCas,0 (4)

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Cw =

1 1 Cw,0K pw

(2)

where Np is the contaminant amount in the sampler, Nas,0 is the initial accessible amount in the sediment, mp and ms are the masses (kg) of sampler and sediment respectively, and Kpw (L kg−1) is the sampler−water partition coefficient. The product of mass and partition coefficient (e.g., mpKpw) has units of volume, and may indeed be interpreted as an equivalent water volume, that is, mpKpw liters of water contain the same contaminant amount as mp kg polymer. Consequently, the numerator of the right-hand side of eq 2 represents the equivalent water volume for the sampler and the denominator represents the equivalent 511

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extract (and for Exposure A and C the whole extract), was azeotropically transferred to acetonitrile through addition of 10 mL acetonitrile followed by Kuderna Danish evaporation to 1− 2 mL. When necessary a gentle nitrogen stream was used to obtain the required final volume of 0.5−2 mL. Chromatographic analysis for PAHs was performed using a high performance liquid chromatograph (Agilent, 1100 series) equipped with two fluorescence detectors (Jasco, Model FP920) placed in series. Standard solutions and extracts (20 μL) were injected on a Vydac 201TP column and the separation was performed using a water/methanol gradient. For PCB quantification, two μL of extract was injected in splitless mode in a PerkinElmer gas chromatograph equipped with an electron capture detector (ECD). Quantification was based on peakheight using a 4 level point-to-point calibration. Total concentrations in sediment were determined by Soxhlet extraction20 of the freeze-dried sediment. Extracts were treated and analyzed for PCBs and PAHs as described above. The organic carbon content of the sediments was measured on an elemental analyzer (Flash EA 1112, Thermo Fischer Scientific, USA) by combustion at 1000 °C after removal of carbonates with hydrochloric acid.

Germany). The experiments were performed in a climate room at 20 ± 0.5 °C in the dark. After exposure, the samplers were taken from the bottle, rinsed with ultra pure water and wiped dry with a tissue. Exposure A. The influence of the sediment content (SC) in the slurry on the uptake rate was investigated by exposing SR samplers with low surface-volume ratio to ensure that measurements in the initial linear part of the uptake curve were possible for most compounds. Portions of 50 g wet sediment H (20 g dry weight) and different volumes of ultra pure water (50, 180, 550, 1600, and 4000 mL) were mixed to obtain SCs of 0.3, 0.1, 0.04, 0.01, and 0.005 g g−1 suspension. These five SC levels were four times identically prepared. Samplers of about 3 g (4 × 3.15 × 0.2 cm) were exposed for different time periods (6, 15, 40, and 144 h) under shaking at 160 rpm, after which they were cleaned, wiped dry, and analyzed for PAHs as described below. Exposure B. Uptake curves were recorded at largely different sampler−sediment ratios to study how this may affect the rate of equilibrium attainment. Using sediment H, exposures at three different sampler−sediment ratios (0.01, 0.1 and 1) were prepared by combining sampler (g), sediment (g dw), and ultra pure water (g) in amounts of 2.5/200/600, 2.5/20/60, and 5/5/15 respectively. Five replicates were prepared for each ratio, exposed for different time periods: 3, 6, 12, 24, and 48 days. Additionally exposures with amounts 2.5/100/300, 2.5/50/150, and 2.5/7/20 (duplicate) were prepared and exposed for 24 days. The surface area of the samplers was 40 cm2 g−1 (thickness 0.05 cm). All mixtures resulted in a SC of 0.25 g g−1 (dry sediment in suspension) but exposures with 7 g or less sediment required some additional water to allow proper mixing. The exposure bottles were shaken at 200 rpm. After exposure, samplers were cleaned, wiped dry and analyzed for 13 parent PAHs and 15 PCBs (Table S3 of the Supporting Information). For the highest phase ratio, two samplers in one bottle were used, which were analyzed separately. These duplicates showed no differences larger than the analytical variability. Exposure C. Following the concept outlined in the theory section, exposures with different sampler−sediment phase ratios were performed for sediment samples H, E, and M. Amounts of 4, 10, 20, and 70 (the latter only for H) g sediment (dry weight) were suspended in ultra pure water to obtain a SC of 0.1 g g−1 (9 mL per g dry sediment) and samplers (2.5 g, 4 × 12.5 × 0.05 cm) were added. An amount of 5 g sediment was used to prepare a similar suspension, but two samplers (5 g) were added. Bottle sizes were selected to achieve a filling level between 40 and 70%. Resulting sampler−sediment ratios were 1, 0.63, 0.25, 0.13, and 0.03 (the latter only for H). Exposure bottles were shaken at 160 rpm for 28 days, followed by analysis of PAHs in the samplers. Analysis. Samplers were extracted in a brown colored hot Soxhlet for 6 h with 75 mL of hexane/acetone (3/1). The extract was concentrated by Kuderna-Danish evaporation to less than 2 mL. Sulfur was removed by ultrasonic treatment for 30 min after addition of 100 mg copper powder. Then the extract was purified over 4 g aluminum oxide (deactivated with 10% water and prerinsed with 25 mL hexane), eluting with 25 mL hexane. The eluate was concentrated by Kuderna-Danish evaporation to two mL, and 1000 ng 2-methylchrysene and 50 ng PCB143 (only Exposure B) were added as internal standards. For samplers of Exposure B half of the extract was used to determine PCBs by GC-ECD. The other part of the

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RESULTS AND DISCUSSION Prior to this work, several trials were done showing that equilibrium between a SR sampler and sediment was attained for pyrene in about a day and for dibenz[a,h]anthracene (logKow = 6) in less than 20 days (Figure S2 of the Supporting Information). This is much less than passive sampler equilibration times in surface waters. Therefore, the exposures A and B were carried out to investigate which parameters were responsible for this higher equilibration rate. Equilibrium Attainment. Examples of uptake curves obtained from exposure A plotted in Figure 1 show that SC

Figure 1. Amount of pyrene and benz[ghi]perylene in passive samplers versus time, for different dilutions of 20 g (dw) sediment (H) (Exposure A). Sediment contents in suspension are indicated in the graph (g sediment per g suspension). mp/ms = 0.15. Drawn lines represent first order model fits.

has a large influence on the uptake rate. All samplers are expected to reach the same equilibrium concentrations for all SC (because mp/ms was the same, viz. 0.15; eq 3), but the rates at which these equilibrium concentration were attained increased with increasing SC (pyrene, Figure 1, left panel). The graph for benz[ghi]perylene even more clearly shows that the uptake rates rapidly increase when the sediment 512

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Figure 2. Concentrations of pyrene, benz[e]pyrene, benz[ghi]perylene, PCB52, PCB118, and PCB180 in passive samplers versus time (exposure B) at a sediment content of 0.25 g g−1 suspension. Different curves represent sediment amounts of 80 (Δ), 8 (○), and 1 (■) g g−1 sampler mass. Drawn lines represent first order model fits.

0.3 g g−1 often does not mix well, especially in small bottles, but SCs around 0.1 g g−1 do for most sediments. These conditions were used for the equilibrations at different phase ratios that were performed in exposure C to evaluate the equilibrium model presented. Evaluation of the Equilibrium Distribution Model. Using eq 3, Cp values of the equilibrations at multi-phase ratios in exposure C were fitted through nonlinear least-squares regression (NLS) as a function of mp/ms with Cw,0 and Cas,0 as adjustable parameters. The obtained values for Cw,0 and Cas,0, with their uncertainties as well as corresponding Kasw and Kaoc values, and the accessible fraction (Fa = Cas,0/Cs,0) are summarized in Table S4 of the Supporting Information. The latter parameter expresses the ratio of accessible sediment concentration, Cas,0 to the total sediment concentration, Cs,0 that was determined by Soxhlet extraction. The Kpw values required for the above calculation were adopted from the literature18 and listed in Table S3 of the Supporting Information. Note that the uncertainties listed in Table S4 of the Supporting Information are originating from the fit and indicate the adequacy of the conceptual approach presented here, but the further accuracy of Cw,0, Kasw, and Kaoc is proportional with the accuracy of the Kpw. However, the Cas,0 does not rely on Kpw, because in eq 3 the product Cw,0Kpw is actually the concentration in the sampler under nondepletive conditions. Plotting Cp versus mp/ms showed a convex curve and plotting 1/Cp versus mp/ms yielded straight line, in accordance with eq 4 (Figure S5 of the Supporting Information). However, plotting Cw versus the residual concentration in the sediment (Cs) better illustrates the environmental relevance of the data. In the individual incubations Cw equals Np/(mpKpw) where Np is the

suspensions were more dense, although this compound did not reach equilibrium within the 6 days. In each system, the amount of sampler and sediment was constant, and because it is not likely that sediments release contaminants slower in diluted suspensions the faster uptake must be attributed to the sampler uptake process. For passive sampling with SR samplers in water, the uptake is controlled by the diffusion process in the water boundary layer.19 Shaking in dense suspensions may disrupt the sampler’s water boundary layer and decrease the diffusion distance resulting in a faster transport of compounds from sediment to sampler.21 The results of exposure B in Figure 2 show examples of obtained uptake curves at three different sampler−sediment ratios (1/80, 1/8, and 1/1) at a constant SC (0.25 g sediment per gram suspension). The equilibrium concentration in the sampler decreases with increasing mp/ms (eq 3). For the highest sampler−sediment ratio (closed squares), equilibrium is attained fastest, particularly visible for PCB180. The reason for the fast equilibration is that the equilibrium concentrations in both the sampler and the sediment are low. Also both depletion of the sediment and uptake by the sampler promote the rate at which this equilibrium is attained. By contrast, the equilibrium concentrations are high at low sampler/sediment mass ratios and equilibrium attainment rates are solely due to uptake by the sampler and not by depletion of the sediment. The curves of 80 and 8 g g−1 sampler coincide for benz[ghi]perylene as they both represent a nondepletive condition due to the low Kpw. The equal curves further confirm that the uptake by the sampler controls the equilibrium attainment rate. For PAHs a SC of 0.1 g g−1 and an equilibration time of 28 days is sufficient, as also supported by Figure S2 of the Supporting Information. Practice also showed that a SC as high 513

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where Cs,0 is the original total concentration in the sediment without any added sampler. Figure 3 shows the decrease of Cw for fluoranthene as a function of Cs. The vertical line with the arrow on the right side indicates Cs,0. The dotted line intercepts the x axis at Cs,0−Cas,0 representing the strongly bound or inaccessible concentration. In the y direction the dotted line crosses Cs,0 at the level of Cw,0. In Figure 3, Cw decreases proportionally with the concentration extracted from the sediment and essentially shows the desorption isotherm with a slope equal to 1/Kasw. For fluoranthene, the extraction by the passive sampler causes Cw to decrease by more than a factor 10, whereas only about 30% of the fluoranthene was released from the sediment. The other 70% seems not to participate in the partitioning process. The line is extrapolated but likely will not be linear down to the x axis but may curve toward the origin at very low Cw values. Nevertheless, the Kasw can be used to model the decrease of Cw with decreasing concentration in the sediment. Similarly, Cas,0 may be a useful measure for an accessible concentration in comparing different sediments. The results were well repeatable, demonstrated by the fact that results for Cw,0 and Cas,0 for sediment H calculated from the three phase ratios equilibrations of 48 d in exposure B (Table S6 of the Supporting Information) were in good agreement with those in exposure C for the majority of the compounds in spite of the fact they were performed by different chemists in a different year. Exposure B also included PCBs for which the

Figure 3. Fluoranthene concentrations in the water (Cw) versus the residual concentration in the sediment (Cs) as calculated by eq 5. The arrow on the right indicates the original concentration in the sediment (Cs,0).

amount adsorbed by the sampler. Np is also the amount that was released from the sediment. The concentration in the sediment at equilibrium after exposure (Cs) can be obtained from Cs = Cs,0 −

Np ms

(5)

Figure 4. Accessible concentrations (Cas,0) in the sediment (shaded and hatched bars for exposure C and B respectively) and original total concentrations (Cs,0) in the sediment (wider open bars). Panel I shows data for PAHs and PCBs of sediment H, and panels II and III data for PAHs of sediments E and M, respectively. For PCB49 Cs,0 was not measured. Error bars represent standard deviations. Full names for abbreviations listed in Table S3 of the Supporting Information. 514

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Figure 5. Log Kaoc of PAHs (●) and PCBs (Δ) versus log Kow. Error bars (s) do not include the uncertainty in the organic carbon content. The drawn lines represent the Karickhoff relationship Koc = 0.63 Kow.

Implications for Risk Assessment. The method in this work yields two parameters that are important for the risk assessment of contaminated sediments. Pore water concentrations (Cw,0) represent the contaminant levels that aqueous organisms are exposed to, and accessible concentrations (Cas,0) represent the sediment’s capacity to buffer pore water concentrations when removal processes occur, such as release to uncontaminated overlying water, contaminant accumulation by organisms, or bioremediation. Applying the US EPA equilibrium partitioning sediment benchmarks,28 the measured parameters could all be expressed as a predicted sum of toxic units (PSTU). This could also be done for Cw,0 as these benchmarks were also listed for the dissolved phase. In Figure 6, ∑PSTU values for the 13 PAHs in

isotherms intercepted the x axis much closer to the origin (Figure S7 of the Supporting Information) indicating that values for Fa are close to unity. It may be tempting to estimate Cw,0 and Cas,0 by simple linear regression (Figure 3), but this is not correct because Cs is not error free and because Cw and Cs originate from the same Np measurement. Accessible concentrations for PAHs are 2 to 10 times lower than total concentrations (Figure 4). The accessible fractions are highest at H, which has of all three stations the most contaminated sediment and is probably also the most freshly contaminated station possibly explaining the larger Cas,0. At station M, most remote from pollution sources, the accessible concentration as well as the accessible fraction is lowest. For PCBs, only determined at H, the accessible fraction is much larger than for PAHs (part I of Figure 4). The accessible fraction for higher chlorinated PCBs ranges from 0.8 to 1. The exceptionally high Cas,0 of PCB187 is likely due to analytical error. Partition Coefficients. We speculate that the experimental logKaoc values in this work reflect sorption by the amorphous organic carbon fraction. Note that with just a single Cw measurement, the accessible concentration is unknown and using the total concentration in the sediment only an apparent K’oc can be calculated, which will be much higher, especially for PAHs. In Figure 5, the obtained logKaoc values are plotted versus logKow and compared with the relation for amorphous organic matter as suggested by Karickhoff.22 The fraction amorphous organic carbon is slightly overestimated (and Kaoc underestimated) because the fraction organic carbon determined by applying oxidation at 1000 °C includes also the black carbon.23 The logKaoc values of the PAHs are slightly higher for the industrial and polluted site (H and E) than at M, which is located further away from pollution sources. For E and M, the lighter PAHs have a relatively high logKaoc. Deviations from the Karickhoff relationship may partly arise from uncertainties in logKow values that were taken from Mackay et al.24 (PAHs) and Hawker and Connell25 (PCBs). Plotting the logKaoc values versus logKoc obtained from an empirical relation based on structural parameters26,27 the scatter is reduced and most data are closer to the 1:1 line with the lower PAHs showing the largest deviation (Figure S8 of the Supporting Information). Experimental logKaoc of the PAHs were highly correlated with molecular weight (MW), with the exception of 3-ring PAHs (Figure S9 of the Supporting Information). LogKaoc of PCBs showed also a good correlation with MW with trichlorobiphenyls being outliers (Figure S8 of the Supporting Information).

Figure 6. Predicted sum of toxic units (∑PSTU) for 13 PAHs in the three sediments calculated from freely dissolved (Cw,0), accessible (Cas,0), and total (Cs,0) concentration in the sediments.

all three sediments are displayed and indicate that results calculated from freely dissolved and accessible concentrations are in reasonable agreement, but using the total concentration in average a four times higher ∑PSTU is obtained. Practical Application. For the present experiments, a sediment content of 0.1 g sediment per gram water was optimal, and an equilibration time of 28 days was sufficient for PAHs. It cannot be excluded that the application of the multiratio passive sampling method with other sediments, other sampler phases, or compounds requires a different scaling. A sediment content of about 0.1 g g−1 is advisable for fine grained sediments to allow optimal mixing. For coarser sediments a higher SC can be considered. It is difficult to predict the necessary equilibration times for other sediments and other compounds than PAHs. Exposure B showed that for the harbor sediment (H) a longer incubation time is required for higher PCBs. Longer equilibration times will be required when the Kpw is higher. In future studies it may be wise to first carry out a 515

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(2) Reichenberg, F.; Mayer, P. Two complementary sides of bioavailability: Accessibility and chemical activity of organic contaminants in sediments and soils. Environ. Toxicol. Chem. 2006, 25, 1239−1245. (3) Jonker, M. T. O.; Koelmans, A. A. Sorption of polycyclic aromatic hydrocarbons and polychlorinated biphenyls to soot and soot-like materials in the aqueous environment mechanistic considerations. Environ. Sci. Technol. 2002, 36, 3725−3734. (4) Cornelissen, G.; Gustafsson, O.; Bucheli, T. D.; Jonker, M. T. O.; Koelmans, A. A.; van Noort, P. C. M. Extensive sorption of organic compounds to black carbon, coal, and kerogen in sediments and soils: Mechanisms and consequences for distribution, bioaccumulation, and biodegradation. Environ. Sci. Technol. 2005, 39, 6881−6895. (5) Cornelissen, G.; Rigterink, H.; ten Hulscher, D. E. M.; Vrind, B. A.; van Noort, P. C. M. A simple Tenax extraction method to determine the availability of sediment-sorbed organic compounds. Environ. Toxicol. Chem. 2001, 20, 706−711. (6) Hermans, J. H.; Smedes, F.; Hofstraat, J. W.; Cofino, W. P. A method for estimation of chlorinated biphenyls in surface waters: influence of sampling method on analytical results. Environ. Sci. Technol. 1992, 26, 2028−2034. (7) Smedes, F. Sampling and partition of neutral organic contaminants in surface waters with regard to legislation, environmental quality and flux estimations. Intern. J. Environ. Anal. Chem. 1994, 57, 215−229. (8) Huckins, J. N.; Petty, J. D.; Booij, K. Monitors of Organic Chemicals in the Environment: Semipermeable Membrane Devices; Springer: New York, 2006; pp 1−223. (9) Mayer, P.; Tolls, J.; Hermens, J. L. M.; Mackay, D. Equilibrium sampling devices. Environ. Sci. Technol. 2003, 37, 185A−191A. (10) Heringa, M. B.; Schreurs, R. H. M. M.; Busser, F.; van der Saag, P. T.; van der Burg, B.; Hermens, J. L. M. Toward more useful in vitro toxicity data with measured free concentrations. Environ. Sci. Technol. 2004, 38, 6263−6270. (11) Mayer, P.; Vaes, W. H. J.; Wijnker, F.; Legierse, K. C. H. M.; Kraaij, R.; Tolls, J.; Hermens, J. L. M. Sensing dissolved sediment pore water concentrations of persistent and bioaccumulative pollutants using disposable solid-phase microextraction fibres. Environ. Sci. Technol. 2000, 34, 5177−5183. (12) Jonker, M. T. O.; Koelmans, A. A. Polyoxymethylene solid phase extraction as a partitioning method for hydrophobic organic chemicals in sediment and soot. Environ. Sci. Technol. 2001, 35, 3742− 3748. (13) 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, 4213−4220. (14) Golding, C. J.; Gobas, F.; Birch, G. F. Characterization of polycyclic aromatic hydrocarbon bioavailability in estuarine sediments using thin-film extraction. Environ. Toxicol. Chem. 2007, 26, 829−836. (15) Jahnke, A.; Mayer, P.; McLachlan, M. S. Sensitive equilibrium sampling to study polychlorinated biphenyl disposition in Baltic Sea sediment. Environ. Sci. Technol. 2012, 46, 10114−10122. (16) Reichenberg, F.; Smedes, F.; Jonsson, J.; Mayer, P. Determining the chemical activity of hydrophobic organic compounds in soil using polymer coated vials. Chem. Cent. J. 2008, 2, 8. (17) Smedes, F. Methods using passive sampling techniques in sediment for the estimation of pore water concentrations and available concentrations for hydrophobic contaminants. ICES CM 2007/J:07 2007, http://www.ices.dk/products/CMdocs/CM-2007/J/J0707.pdf. (18) 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, 7047−7054. (19) Rusina, T. P.; Smedes, F.; Klanova, J.; Booij, K.; Holoubek, I. Polymer selection for passive sampling: A comparison of critical properties. Chemosphere 2007, 68, 1344−1351. (20) Smedes, F.; de Boer, J. Determination of chlorobiphenyls in sediments - analytical methods. TRAC-Trend. Anal. Chem. 1997, 16, 503−517.

pilot experiment at low mp/ms ratios because equilibration times for higher mp/ms ratios are always shorter. Although sediment suspensions were not poisoned, there was no indication of degradation as biodegradation would have developed different for individual samples and cause a poor fit with NLS. Addition of a biocide, or at least removing oxygen before exposure, should be considered. The equilibrium distribution model (eq 2) can be used to select the sampler−sediment ratios such that the results will cover the whole range from low to high depletion. Adopting the relationship of Karickhoff,22 Kasw can be approximated as 0.63 foc Kow, where foc is the organic carbon content in the sediment. Inserting this in eq 2 and rearranging, the sampler−sediment phase ratio can be calculated for a preset depletion D (equals Np/Nas,0) by mp ms

=

0.63foc Kow K pw

( D1 ′ − 1)

(6)

In the ideal case, multi-ratio passive sampling should be applied for a set of D’s ranging from 0.1 to 0.9. Using conditions selected accordingly, application of two different phase ratios in the very low D’ range will also allow verification of equilibrium attainment.16 If in specific cases only measurement of Cw,0 is required and a nondepletive situation (e.g., D’ < 0.05) is needed, eq 6 can also be applied to select an appropriate sampler−sediment ratio. Future work will focus on understanding the sediment− sampler exchange process and for faster equilibrium using thinner films coated on the inside bottle wall.15,29

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ASSOCIATED CONTENT

* Supporting Information S

Sampling locations, uptake curves of preliminary investigations, table with compounds and used logKpw, results of exposure C, examples of nonlinear versus linear regression of Cp and phase ratio, table of results of exposure B, Cw versus Cs for PCB101, experimental logKaoc plotted versus logKoc values from an empirical relationship, and experimental logKaoc plotted versus MW. This material is available free of charge via the Internet at http://pubs.acs.org.

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AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected], tel: +420 549 49 3097, fax: + 420 549 49 2840. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS The technical assistance of Audrey Luszezenac and Frida Nicolai is highly appreciated. This work was performed at the National Institute of Coastal and Marine Management/RIKZ and publication was further supported by Masaryk University, Deltares, and the Royal Netherlands Institute for Sea Research. Unknown reviewers are acknowledged for their critical examination and constructive comments.

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REFERENCES

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