Comparison Of Five Methods For Measuring Sediment Toxicity Of

Regulation, Sacramento, California. Received July 31, 2007. Revised manuscript received October. 16, 2007. Accepted October 22, 2007. Sediment toxicit...
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Environ. Sci. Technol. 2007, 41, 8394–8399

Comparison Of Five Methods For Measuring Sediment Toxicity Of Hydrophobic Contaminants Y I P I N G X U , †,‡ F R A N K S P U R L O C K , * ,§ ZIJIAN WANG,‡ AND JAY GAN† Department of Environmental Sciences, University of California, Riverside, California 92521, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing, China, and California Department of Pesticide Regulation, Sacramento, California

Received July 31, 2007. Revised manuscript received October 16, 2007. Accepted October 22, 2007.

Sediment toxicity from hydrophobic organic compounds (HOCs) is complicated by chemical partitioning among multiple phases and sediment-specific bioavailability. In this study, we used three hydrophobic pyrethroid insecticides as test compounds and derived 10-d median lethal concentrations (LC50s) for Chironomus tentans in three different sediments. The LC50s were expressed using HOC concentrations on a bulk sediment basis (CS), organic carbon (OC)-normalized sediment basis (CS-OC), porewater basis (CPW), dissolved organic carbon (DOC)-normalized porewater basis (CPW-DOC), and freely dissolved porewater basis (C free). The bulk phase CS and CPW yielded highly variable LC50s across sediment types, whereas the use of normalized concentrations CS-OC and CPW-DOC generally reduced variability due to sediment type but not that due to aging. In contrast, LC50s based on C free were essentially independent of sediment conditions. The sediment porewater samples contained approximately 20-90 mg L-1 DOC, and the C free expressed as a percentage of the total bulk pore water concentration ranged from 9 to 28% for fenpropathrin (mean ) 19%), 8 to 18% for bifenthrin (mean ) 13%), and 3 to 8% for cyfluthrin (mean ) 6%) across the different sediments. These results indicate that the use of C free reduces uncertainties caused by sediment variables such as OC properties and aging effects.

Introduction Hydrophobic organic contaminants (HOCs) have a strong affinity for sediments, and sediment toxicity from HOC contamination has been the focus of numerous studies and regulatory guidelines (1–8). It is generally recognized that the bulk sediment concentration CS is a poor indicator for benthic organism exposure due to variation in contaminant bioavailability, with the result that CS-based HOC median lethal concentrations (LC50s) are variable among sediments (9–11). Di Toro et al. (10) introduced the concept of equilibrium partitioning (EqP) to account for the effect of phase distribution on bioavailability. At equilibrium, an organism’s exposure is related to the contaminant activity in any of the equilibrated phases, regardless of the con* Corresponding author e-mail: [email protected]. † University of California, Riverside. ‡ Chinese Academy of Sciences. § California Department of Pesticide Regulation. 8394

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taminant uptake route. Therefore, if sediment effect concentrations are expressed in concentration units that are proportional to chemical activity, those effect concentrations should be generally applicable to other sediments. That explains, for example, why median lethal concentrations (LC50s) expressed in units of organic carbon (OC)-normalized sediment concentration C S-OC are often less variable across different sediments as compared to LC50s expressed in terms of CS (10). Alternative units for expressing sediment LC50s also include porewater concentration CPW (10). However, most recent studies express effect concentrations in terms of C SOC (2–5). This preference is attributable to the generally easier analytical determination of HOCs in bulk sediment as opposed to porewater (12, 13). The use of C S-OC implicitly assumes a constant KOC for a given HOC across different sediments and conditions (10). However, various properties of sediment organic matter (e.g., black carbon content) and other factors (e.g., contaminant residence time) affect HOC sorption, so that KOCs for a single HOC may be quite variable (11, 14, 15). From a practical standpoint, this means that C S-OC-based effect concentrations determined in one sediment may be only approximate for other sediments and/or conditions. For instance, Amweg et al. (5) and Maund et al. (6) observed substantial differences in Hyalella azteca and Chironomous tentans CS-based LC50s for pyrethroid compounds among different sediments. For four of six pyrethroids, OC normalization reduced LC50 variability in the study of Amweg et al. (5), but substantial variation still occurred in Maund et al. (6) where high OC sediments were included. These and other studies demonstrate that other factors beyond OC content influence the sediment toxicity of HOCs. Consequently other approaches for predicting sediment toxicities should be explored. For porewater-based effect concentrations, the direct use of CPW may also result in high variability, especially for strongly hydrophobic HOCs. This occurs because the freely dissolved concentration (C free) can deviate greatly from CPW due to complexation with dissolved organic carbon (DOC) in porewater (16, 17). The U.S. Environmental Protection Agency has recognized the importance of accounting for porewater DOC in their draft methods for deriving site-specific equilibrium partitioning sediment guidelines (8). Those guidelines recommend either direct measurement of C free, or use of KDOC to estimate C free. In practice five phase concentrations may be used to describe sediment exposure and toxicity: CS, C S-OC, CPW, CPW-DOC (CPW normalized by DOC), and C free. Although finding the most suitable phase concentration has been an issue of discussion for several years (8, 10, 18), to our knowledge no study has experimentally compared all five methods for describing median lethal concentrations of highly hydrophobic organic chemicals in sediment. Therefore, the merits and limitations of the different approaches are somewhat unclear. In this study we compared the use of the five phase concentrations for expressing Chironomus tentans LC50s in different sediments using pyrethroid insecticides as test compounds. Pyrethroids are acutely toxic to benthic invertebrates at sediment concentrations in the ppb range (5), and they are strongly hydrophobic with log Kow ranging from 4.53 to 7.0 (19). The overall goal was to identify the most reliable and practical approach to obtaining LC50s for HOCs that are independent of sediment conditions. 10.1021/es071911c CCC: $37.00

 2007 American Chemical Society

Published on Web 11/13/2007

TABLE 1. Selected Physical-chemical Characteristics of the Three Sediments Used in This Study sediment SDC SR BM

a

OC (%)

soot C (%)

lipids (g kg-1)

soot C (% OC)

lipids (% OC)

1.44 1.88 5.03

0.03 0.04 0.02

0.7 1.8 2.6

2.29 2.13 0.42

5.00 9.61 5.08

a SDC ) San Diego Creek sediment; SR ) Santa Rosa sediment; BM ) Black Mountain sediment.

Experimental Section Pyrethroids. Bifenthrin (98.0% pure) and fenpropathrin (99.5% pure) standards were purchased from Chem Service (West Chester, PA). Cyfluthrin standard (>93.3% pure) was obtained from Bayer CropScience (Stillwell, KS). Other chemicals and solvents were purchased from Fisher and were of GC or analytical reagent grade. Sediments. Three California sediments were selected to span a range in OC. San Diego Creek sediment (SDC) was collected from Orange County, Santa Rosa sediment (SR) was collected from Sonoma County, and Black Mountain sediment (BM) was collected from San Luis Obispo County. For each, 0–10 cm of surface sediment was collected using a hand shovel, and the sediments were transported to the laboratory and stored at 4 °C. All sediments were wet sieved through a 2-mm sieve prior to use. Aliquots of sediments were air-dried at room temperature and passed through a 250-µm sieve for the property characterization. Sediment OC as a percentage of total dry sediment mass was determined by elemental analysis after acidification to remove carbonates on a nitrogen/carbon analyzer (Thermo Finnigan, Woods Hole, MA). Lipids were extracted from the dry sediment samples (50 g) by using 24-h Soxhlet extraction with 300 mL of chloroform/methanol mixture (2:1, V/V). The total carbon content of the sediment sample was measured on the nitrogen/carbon analyzer before and after the extraction, and the difference in total carbon content after Soxhlet extraction was estimated as the lipid content (20). The sediment soot carbon content was measured following the procedure described by Gustafsson et al. (21). Sediment characteristics are summarized in Table 1. Sediment Spiking and Equilibration. Sediments were spiked with the test compounds using a rolling-jar method following U.S. EPA guidelines (22). Briefly, 10 g of silica sand in glass jars was spiked with 0.25–1.0 mL of stock solution (10–1000 µg mL-1 in acetone) of test compound or 1.0 mL of acetone (solvent control), to obtain six concentrations of pesticide-dosed sediments and one solvent control sediment. Solvent was evaporated in a fume hood for 1 h, wet sediment (400 g dry weight equivalent) was added to each jar, and the sample jars were tumbled on a rolling mill overnight to ensure complete mixing. The jars were then stored in the dark at 4 °C for 30 d to allow pyrethroid phase equilibration. During this time, jars were rolled once a week for 2 h. A previous study indicated apparent equilibrium of pyrethroids in the different phases using this protocol (23). To evaluate the influence of contact time, a set of SDC sediment samples was spiked as above and kept in the dark at room temperature for 90 d before toxicity testing and analysis. Toxicity Testing. Sediment toxicity tests were conducted with Chironomus tentans for each sediment-pesticide combination using the USEPA 10-d static exposure method for benthic invertebrates acute toxicity testing (22). C. tentans were initially obtained from Aquatic Biosystems (Fort Collins, CO) and cultured in aquaria with reconstituted hard water (RHW), and silica sand at 23 ( 1 °C and under a 16:8-h light/ dark cycle. Second to third-instar C. tentans larvae from the laboratory culture were used in the test. Test systems

consisted of 500-ml glass jars containing 30 g (dry weight equivalent) of sediment and 200 mL of RHW. Four replicate jars were used for each concentration point in the mortality assay, and each jar contained 10 organisms. In addition, three replicate jars without organisms were identically prepared and used for analysis of pesticide phase distribution. The toxicity tests were conducted at 23 ( 1 °C with a 16:8-h light/dark cycle. Test organisms were fed daily with 6.0 mg of Tetrafin goldfish food (Tetra, Blacksburg, VA). During the test, an 80% water change was performed every day. Partitioning calculations showed that the mass of pyrethroid removed in each water change was negligible, on the order of 0.1%-0.3% of total pyrethroid mass in each jar. Water samples were taken prior to water renewal on day 2 and day 10 for measurement of water quality characteristics including alkalinity, ammonia, conductivity, hardness, and pH. The dissolved oxygen (DO) level in the overlying water was kept above 2.5 mg L-1 during the entire test. At the end of the 10-d exposure, the mortality of C. tentans was recorded. The water and sediment were filtered through a #40 (425 µm) sieve and the number of live organisms was counted. In each assay, quality control tests were performed using the untreated sediment. The survival rate of C. tentans in the control jars was greater than 90% in every case, and the mean C. tentans survival over the entire study was 93%. Analysis of Pesticide Concentrations in Different Phases. Two days into the toxicity testing, the overlying water from the organism-free jars was removed by vacuum and discarded. The sediment was quantitatively transferred to a 200mL polyethylene centrifuge tube and centrifuged at 10,000 rpm for 30 min to separate pore water from the sediment phase. The supernatant was collected with a pipet and transferred into a 50-mL brown glass vial with a Teflon lined cap. The sediment was collected from each centrifuge tube and transferred into a 100-mL glass jar. Both porewater and sediment samples were stored at 4 °C in the dark if not analyzed immediately. Preliminary experiments showed that the test compounds were stable under the storage conditions, and the loss due to sorption to the container walls was negligible (23). Mass-based sediment concentrations CS were determined using a sonication extraction procedure and analysis by gas chromatography with electron capture detection (GC-ECD). Recoveries of the three pyrethroids from sediment ranged from 74.7 to 94.8% using the above conditions. OC-normalized sediment concentration C S-OC was calculated as CS/ [sediment OC]. Total porewater concentration CPW was determined by liquid–liquid extraction (LLE) using ethyl acetate, reconstitution in hexane, and GC-ECD analysis. The recoveries of pyrethroids ranged from 87.6 to 114.9% using this method. Porewater DOC was determined by analyzing an aliquot (4.0 mL) of every porewater sample on an Apollo 9000 Carbon Analyzer (Teledyne Instruments, Mason, OH). CPW-DOC was calculated as CPW/[porewater DOC]. Further analytical details of the sediment and porewater analyses are available in the Supporting Information (SI). C free was determined using a previously reported SPME method (23, 24). Briefly, a Supelco 30-µm polydimethylsiloxane (PDMS) fiber SPME assembly was used to sample a 10-mL aliquot of each porewater sample. The SPME fiber extraction time was 25 min, fiber immersion depth was 2 cm below the surface, and a small Teflon-coated magnetic bar was used to stir the sample solution at 600 rpm. After sampling, the fiber was directly injected into the GC-ECD under the same chromatographic conditions described in the SI for sediment. External standards with known pesticide concentrations were similarly analyzed and used to quantify C free. In a few low concentration cases, cyfluthrin or fenpropathrin C free were lower than the method detection limits VOL. 41, NO. 24, 2007 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 2. Sediment-water partition coefficients of pyrethroid compounds in sediments (average ± standard deviation; n = 8–12) sediment

Kd (× 103)

KOC (× 105)

SDC SR BM SDC-90b

9.3 ( 1.6 13.1 ( 2.1 18.9 ( 5.4 14.1 ( 3.4

6.5 ( 1.1 7.0 ( 1.1 3.8 ( 1.0 9.8 ( 2.3

SDC SR BM

6.3 ( 0.4 7.1 ( 2.0 16.4 ( 3.1

4.4 ( 0.2 3.8 ( 1.1 3.3 ( 0.6

SDC SR BM

6.6 ( 1.2 8.8 ( 1.0 18.8 ( 3.5

4.6 ( 0.8 4.7 ( 0.5 3.7 ( 0.7

DOC (mg L-1) bifenthrin 37.8 ( 5.7 29.7 ( 5.1 66.8 ( 7.1 22.8 ( 3.5 cyfluthrin 33.8 ( 8.7 33.2 ( 7.3 79.6 ( 19.4 fenpropathrin 31.6 ( 7.9 34.1 ( 6.7 85.9 ( 14.7

KDOC (× 105)

Ffreea

1.9 ( 0.8 1.7 ( 0.6 1.8 ( 0.4 3.5 ( 1.3

0.139 ( 0.044 0.185 ( 0.058 0.078 ( 0.016 0.119 ( 0.032

3.5 ( 0.4 4.0 ( 0.7 4.4 ( 0.8

0.078 ( 0.009 0.072 ( 0.013 0.029 ( 0.004

0.9 ( 0.3 0.8 ( 0.1 1.2 ( 0.2

0.271 ( 0.061 0.275 ( 0.027 0.093 ( 0.014

a Fraction of solute in bulk solution present as “freely dissolved”, ) Cfree/CPW. b SDC sediment was equilibrated at room temperature for 90 days, while the regular spiked sediments were equilibrated at 4 °C for 30 days.

(MDLs). C free values were calculated assuming linear partitioning between the DOC and the freely dissolved phases.

Results and Discussion Phase Distribution of Sediment-Associated Pyrethroids. In the test sediment-water systems, the pyrethroids were primarily associated with the solid phase. Simple partition coefficients Kd, defined as CS/C free, ranged from approximately 6000 to 18,000 L kg-1 (Table 2). For cyfluthrin and fenpropathrin, OC-normalized partition coefficients KOC were less variable than Kd. KOCs measured here for all three pyrethroids were in the range of 105 – 106, similar to literature-cited KOC values for pyrethroids (19, 25). The water-DOC partition coefficient KDOC (L kg-1) was calculated as follows: KDOC )

(CPW - Cfree) ⁄ [DOC] Cfree

(1)

The KDOC obtained here (Table 2) varied by a factor of about 5 across the three pyrethroids and three sediments. To our knowledge KDOC values have not been previously reported for cyfluthrin and fenpropathrin. Yang et al. (26) reported bifenthrin KDOCin dilute suspensions of sediments SDC and SR (DOC ) 2-5 mg L-1), and their results were comparable to the bifenthrin KDOC obtained here for much higher DOC porewaters of SDC and SR (Table 2). In addition, we also found no consistent evidence of correlation between KDOC and C free in this study, so we conclude that KDOC is independent of concentration over at least 1-2 orders of magnitude in these sediments. The KDOCs for a single HOC are typically highly variable among different sources of DOC (27, 28). Surprisingly, in this study there was no significant effect of source sediment on KDOC based on analysis of variance (ANOVA) of the 84 KDOC data points obtained for the three pyrethroids in pore waters of the three sediments (p ) 0.077). That same two-way ANOVA demonstrated a modest but significant difference in mean KDOC between each of the three pyrethroids (p < 0.001), following the order fenpropathrin < bifenthrin < cyfluthrin as shown in Table 2. The 90-d SDC bifenthrin mean KDOC was almost twice that of the 30-d SDC sediment (Table 2), and that difference was significant (2-sample t test, p < 0.001). It is most likely that this effect was attributable to changes that occurred to the DOC during the extended incubation, as noted in other studies (28). Porewater DOC concentrations ranged from approximately 20 to 90 mg L-1 and the measured KDOCs were all on the order of 105 L kg-1 (Table 2). Consequently C free constituted 8396

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only a small fraction of total pyrethroid in the bulk solution. C free expressed as a mean fraction of the total bulk pore water concentration ranged from 9 to 28% for fenpropathrin (mean ) 19%), 8 to 18% for bifenthrin (mean ) 13%), and 3 to 8% for cyfluthrin (mean ) 6%) across the different sediments (Table 2). Median Lethal Concentrations. Concurrent measurement of chemical concentrations in the different phases allowed construction of five different concentration–response curves for each pyrethroid/sediment combination as shown in Figure 1 for bifenthrin. Similarly, five different types of LC50 for each sediment are reported (Table 3). To our knowledge the only pyrethroid for which published C. tentans sediment LC50 data are available is cypermethrin (6). The reported C S-OC-based cypermethrin LC50s in that study ranged from 13 to 67 µg g-1 (6), comparable to bifenthrin LC50s measured in this study (18 to 49 µg g-1, Table 3). Overall, both fenpropathrin and cyfluthrin were generally more toxic to C. tentans than bifenthrin regardless of which LC50s were compared. Differences among mortality-concentration–response curves for a pyrethroid in different sediments depended on how concentrations were expressed (Figure 1). From the standpoint of EqP, the measured LC50 for a particular pyrethroid should correspond to the chemical activity, regardless of sediment. Therefore, large sediment-to-sediment differences in LC50 result when a particular type of concentration unit does not accurately reflect the chemical activity in different systems. For example, it is evident from the LC50 data in Table 3 that the CS- and CPW-based LC50s for each pyrethroid are the most variable. We conclude that CS and CPW provide an inconsistent measure of the pyrethroid activity that C. tentans was exposed to in the different sediment systems. When the bulk concentrations were normalized to organic carbon, concordance between the nonaged treatments (e.g., C S-OC and CPW-DOC; Figure 1 B and D) and LC50s (Table 3) improved markedly, but the 90-d SDC bifenthrin mortality curve and LC50 did not conform well with those for the nonaged bifenthrin sediments. We expected the LC50s expressed on a C S-OC or CPW-DOC basis to be less variable than those expressed on a bulk phase basis, but the consistency among nonaged sediments was surprising. We attribute the high level of agreement among C S-OC- or CPW-DOC-based LC50s to the similarity in partitioning behavior among the sediments. The OC-normalized partition coefficients for each pyrethroid were quite similar when excluding the 90-d SDC bifenthrin treatment (Table 2) so that a single mean value of KOC or KDOC adequately described partitioning between freely dissolved chemical and particulate organic matter or

FIGURE 1. Chironomus tentans mortality curves in bifenthrin-spiked sediments, with bifenthrin concentrations expressed in different phase concentrations: (A) bulk sediment concentration CS; (B) OC-normalized sediment concentration C S-OC, (C) total pore water concentration CPW, (D) DOC-normalized pore water concentration CPW-DOC, and (E) freely dissolved pore water concentration C free (SDC ) San Diego Creek sediment; SR ) Santa Rosa sediment; BM ) Black Mountain sediment; SD-90d ) San Diego Creek sediment with 90 d incubation after pesticide treatment). dissolved organic matter, respectively, regardless of sediment. However, this is unusual. KDOCs typically display a high level of variability (27, 28), and OC-normalized sorption coefficients KOC for a single HOC may vary by a factor of 5 or greater (29). We suspect that C S-OC- or CPW-DOC-based LC50s determined in a greater number of sediments would display greater variability because OC normalization does not account for the known HOC partitioning variability across sediments. For example, Amweg et al. (5) and Maund et al. (6) report various pyrethroid LC50 data for Hyalella azteca in multiple sediments, with the highest CVs in each study exceeding 50%. When LC50s were calculated on a C free basis the bifenthrin mortality curves were nearly congruent (Figure 1E) and LC50s for each pyrethroid were essentially equal, regardless of sediment. We attribute this result to C freeproviding a consistent measure of the bioavailable pyrethroid activity in these sediment pore waters, so that the SPME

method was providing a direct measure of pyrethroid activity experienced by C. tentans in the test systems. Overall the LC50s expressed on a bulk phase basis displayed coefficients of variation (CV) ranging from 30 to 40%, while LC50s based on C S-OC, CPW-DOC, or C free generally yielded much less variable LC50s for nonaged treatments (Table 3). The nonaged LC50 mean CVs for C S-OC, CPW-DOC, and C free were 11.9%, 7.0%, and 6.5%, respectively. However, inclusion of the 90-d SDC bifenthrin treatment resulted in overall mean CVs for C S-OC, CPW-DOC, and C free of 17.2%, 19.2, and 6.2%, respectively. Application Considerations. Additional factors should also be considered in selecting the best approach to determining and expressing sediment toxic concentrations. Measurement of CS (and subsequently C S-OC) requires extraction and analysis of whole sediment samples. The analytical procedure is generally labor intensive and timeVOL. 41, NO. 24, 2007 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 3. Sediment Median Lethal Concentrations (LC50) for Three Pyrethroids Using Standard 10-d Sediment Toxicity Tests with Chironomus tentans, Expressed Using Various Phase Concentrations LC50 (95% CI) sediment

Cs (µg kg-1)

C S-OC (mg kg-1)

SDC SR BM SDC-90 %CVa

418 (291–588) 560 (367–955) 924 (611–1448) 692 (420–1718) 33.2%

29.0 (20.1–40.8) 29.8 (19.5–50.8) 18.3 (12.1–28.7) 49.4 (30.0–122.7) 41.0%

SDC SR BM %CV

34.4 (17.7–57.8) 41.4 (21.8–70.7) 122.2 (65.9–214.2) 73.9%

2.39 (1.23–4.01) 2.21 (1.17–3.77) 2.43 (1.31–4.25) 5.0%

SDC SR BM %CV

34.0 (20.5–52.4) 47.0 (24.7–87.6) 112.5 (66.7–181.1) 65.2%

2.36 (1.43–3.64) 2.50 (1.31–4.66) 2.23 (1.32–3.59) 5.7%

a

CPW (µg L-1) bifenthrin 0.314 (0.239–0.424) 0.258 (0.161–0.469) 0.608 (0.446–0.867) 0.402 (0.255–0.915) 38.8% cyfluthrin 0.119 (0.072–0.173) 0.123 (0.079–0.180) 0.301 (0.175–0.495) 57.4% fenpropathrin 0.021 (0.012–0.032) 0.023 (0.013–0.038) 0.071 (0.045–0.106) 73.8%

Cfree (µg L-1)

8.31 (6.32–11.21) 8.68 (5.42–15.79) 9.10 (6.68–12.98) 17.63 (11.18–40.13) 41.0%

0.048 0.053 0.048 0.051 4.6%

(0.041–0.056) (0.034–0.051) (0.041–0.058) (0.039–0.072)

3.52 (2.13–5.12) 3.71 (2.38–5.42) 3.78 (2.20–6.22) 3.7%

0.0087 (0.0070–0.0106) 0.0089 (0.0073–0.0106) 0.0087 (0.0069–0.0101) 0.9%

0.67 (0.36–0.95) 0.67 (0.39–1.15) 0.83 (0.57–1.33) 12.8%

0.0064 (0.0053–0.0077) 0.0078 (0.0065–0.0093) 0.0061 (0.0049–0.0076) 13.0%

Coefficient of variation.

consuming, often requiring the of use significant amounts of solvents. For sediments with potential matrix influences (e.g., sulfur), multiple cleanup steps may be required, further increasing time and resource inputs. In comparison, CPW and subsequently CPW-DOC are relatively easier to measure, especially if techniques such as solid-phase extraction are used. An earlier argument against the use of porewater for toxicity assessment was the need to prepare a large volume of porewater to afford detectable concentrations. However, analytical methods with much improved sensitivities, such as GC-mass spectrometer (MS) in selective ion monitoring mode, GC-MS-MS, and liquid chromatography-MS-MS, are now commonly used, so that a small amount of porewater (10–30 mL) may be sufficient to derive CPW and CPW-DOC. Our finding that the use of CPW-DOC reduced sediment dependence of LC50 was consistent with the EPA recommended approach for using KDOC to estimate C free when assessing sediment toxicity (8). However, the EPA method would require the determination of CPW and DOC, as well as knowledge of reliable KDOCs that are usually nonexistent. In comparison, measurement of C free does not require knowledge of KDOC, thus reducing uncertainties. When EqP was first introduced, there were few methods for measuring C free, and most of the known methods were tedious (e.g., membrane dialysis), required special devices (e.g., ultracentrifugation), or were compound specific (e.g., fluorescence quenching) (30–32). The most referenced method in earlier studies on detecting porewater C free was the use of reversephase C18 cartridges (31), which was later found to give inconsistent results for some HOCs due to potential retention of DOC on the adsorbent phase (33). Consequently, EPA guidelines cautioned that the use of such methods should include a mass balance check (8). In contrast, over the past few years SPME has evolved as a rapid and reliable method for measuring C free of HOCs in a range of matrices (34) including freshwater and marine sediment porewater (35). Furthermore, many studies have demonstrated that SPMEdetected HOC concentrations are highly correlated with bioaccumulation or toxicity (17, 36, 37), providing evidence that SPME is essentially sampling the C free that reflects the bioavailable portion of HOC in sediment/water/DOC systems. A volume of porewater as small as 1.0 mL is adequate for SPME analysis (16). One of the most significant advantages of using C free for expressing sediment toxicity may be that benchmarks (e.g., LC50s) derived from water-only exposures 8398

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can be directly used for predicting sediment toxicity via calculation of toxicity units, as recently demonstrated for PAHs (38). It should be noted that although partitioning to DOC clearly influences sediment toxicity of the highly hydrophobic pyrethroids here, for chemicals of lesser hydrophobicity such as such as 2- and 3- ring polycyclic aromatic hydrocarbons, differences in the ability of CPW, CPW-DOC, and C free to predict toxicity may be less important. For such chemicals Hawthorne et al. (39) concluded that toxicity in different sediments could be adequately predicted by porewater concentrations alone because CPW = C free. However, for strongly hydrophobic organic chemicals such as pyrethroids we conclude C free is the most suitable method for measuring and reporting LC50s because (a) of low variability, with a mean CV here of 6.2% across all treatments, (b) the technique provides a direct measurement of bioavailable concentration in short-term acute toxicity bioassays, and (c) the method has several practical advantages, including ease of analysis, small sample size requirement, and no need for ancillary data such as KDOC.

Supporting Information Available Analytical details on extraction and analysis of pyrethroids in bulk sediment and pore water via GC-ECD. This material is available free of charge via the Internet at http:// pubs.acs.org.

Literature Cited (1) Weber, D. E.; McKenney, C. L.; MacGregor, M. A.; Celestial, D. M. Use of artificial sediments in a comparative toxicity study with larvae and postlarvae of the grass shrimp, Palaemonetes pugio. Environ. Pollut. 1996, 93, 129–133. (2) Weston, D. P.; You, J.; Lydy, M. J. Distribution and toxicity of sediment-associated pesticides in agriculture-dominated water bodies of California’s Central Valley. Environ. Sci. Technol. 2004, 38, 2752–2759. (3) Maund, S. J.; Hamer, M. J.; Warinton, J. S.; Kedwards, T. J. Aquatic ecotoxicology of the pyrethroid insecticide lambda-cyhalothrin: Considerations for higher-tier aquatic risk assessment. Pestic. Sci. 1998, 54, 408–417. (4) Amweg, E. L.; Weston, D. P.; You, J.; Lydy, M. J. Pyrethroid insecticides and sediment toxicity in urban creeks from California and Tennessee. Environ. Sci. Technol. 2006, 40, 1700– 1706. (5) Amweg, E. L.; Weston, D. P.; Ureda, N. M. Use and toxicity of pyrethroid pesticides in the Central Valley, California, USA. Environ. Toxicol. Chem. 2005, 24, 966–972.

(6) Maund, S. J.; Hamer, M. J.; Lane, M. C. G.; Farrelly, E.; Rapley, J. H.; Goggin, U. M.; Gentle, W. E. Partitioning, bioavailability, and toxicity of the pyrethroid insecticide cypermethrin in sediments. Environ. Toxicol. Chem. 2002, 21, 9–15. (7) U.S. Environmental Protection Agency. Methods for collection, storage and manipulation of sediments for chemical and toxicological analyses; Technical manual EPA-823-B-01-002; U.S. EPA: Washington, DC, 2001. (8) U.S. Environmental Protection Agency. Methods for the Derivation of Site-Specific Equilibrium Partitioning Sediment Guidelines (ESGs) for the Protection of Benthic Organisms; Nonionic Organics (Draft); EPA-822-R-02-042; U.S. EPA: Washington, DC, 2002. (9) Adams, W. J.; Kimerle, R. A.; Mosher, R. G. Aquatic safety assessment of chemicals sorbed to sediments. In Aquatic Toxicology and Environmental Fate, Cardwell, R. D., Purdy, R., Bahner, R. C., Eds.; American Society for Testing and Materials: Philadelphia, PA, 1985; pp479–493. (10) Di Toro, D. M.; Zarba, C. S.; Hansen, D. J.; Berry, W. J.; 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, 1541–1583. (11) Ehlers, G. A. C.; Loibner, A. P. Linking organic pollutant (bio)availability with geosorbent properties and biomimetic methodology: A review of geosorbent characterisation and (bio)availability prediction. Environ. Pollut. 2006, 141, 494–512. (12) Clarke, J. U.; McFarland, V. A. Uncertainty analysis for an equilibrium partitioning-based estimator of polynuclear aromatic hydrocarbon bioaccumulation potential in sediments. Environ. Toxicol. Chem. 2000, 19, 360–367. (13) Fredrickson, H. L.; Furey, J.; Talley, J. W.; Richmond, M. Bioavailability of hydrophobic organic contaminants and quality of organic carbon. Environ Chem. Lett. 2004, 2, 77–81. (14) Alexander, M. Aging, bioavailability, and overestimation of risk from environmental pollutants. Environ. Sci. Technol. 2000, 34, 4259–4265. (15) Semple, K. T.; Morriss, A. W. J.; Paton, G. I. Bioavailability of hydrophobic organic contaminants in soils: fundamental concepts and techniques for analysis. Euro. J. Soil Sci. 2003, 54, 809–818. (16) Hawthorne, S. B.; Grabanski, C. B.; Miller, D. J.; Kreitinger, J. P. Solid-phase microextraction measurement of parent and alkyl polycyclic aromatic hydrocarbons in milliliter sediment pore water samples and determination of KDOC values. Environ. Sci. Technol. 2005, 39, 2795–2803. (17) Ramos, E. U.; Meijer, S. N.; Vaes, W. H. J.; Verhaar, H. J. M.; Hermens, J. L. M. Using solid-phase microextraction to determine partition coefficients to humic acids and bioavailable concentrations of hydrophobic chemicals. Environ. Sci. Technol. 1998, 32, 3430–3435. (18) U.S. Environmental Protection Agency. Technical Basis for the Derivation of Equilibrium Partitioning Sediment Guidelines (ESGs) for the Protection of Benthic Organisms Nonionic Organics; EPA-822-R-00-001; U.S. EPA: Washington, DC, 2000. (19) Laskowski, D. A. Physical and chemical properties of pyrethroids. Rev. Environ. Contam. Toxicol. 2002, 174, 49–170. (20) Kukkonen, J. V. K.; Landrum, P. F.; Mitra, S.; Gossiaux, D. C.; Gunnarsson, J.; Weston, D. Sediment characteristics affecting desorption kinetics of select PAH and PCB congeners for seven laboratory spiked sediments. Environ. Sci. Technol. 2003, 37, 4656–4663. (21) Gustafsson, O.; Haghseta, F.; Chan, C.; MacFarlane, J.; Gschwend, P. M. Quantification of the dilute sedimentary soot phase: Implications for PAH speciation and bioavailability. Environ. Sci. Technol. 1997, 31, 203–209. (22) U.S. Environmental Protection Agency. Methods for Measuring the Toxicity and Bioaccumulation of Sediment-associated Con-

(23) (24) (25) (26) (27) (28)

(29)

(30)

(31)

(32)

(33)

(34) (35) (36)

(37) (38)

(39)

taminants with Freshwater Invertebrates, 2nd ed.; EPA-600/R99/064; U.S. EPA: Washington, DC, 2000. Bondarenko, S.; Putt, A.; Kavanaugh, S.; Poletika, N.; Gan, J. Time dependence of phase distribution of pyrethroid insecticides in sediment. Environ. Toxicol. Chem. 2006, 25, 3148–3154. Liu, W.; Gan, J.; Lee, S.; Kabashima, J. N. Phase distribution of synthetic pyrethroids in runoff and stream water. Environ. Toxicol. Chem. 2004, 23, 7–11. Gan, J.; Lee, S. J.; Liu, W. P.; Haver, D. L.; Kabashima, J. N. Distribution and persistence of pyrethroids in runoff sediments. J. Environ. Qual. 2005, 34, 836–841. Yang, W. C.; Gan, J.; Hunter, W.; Spurlock, F. Effect of suspended solids on bioavailability of pyrethroid insecticides. Environ. Toxicol. Chem. 2006, 25, 1585–1591. Burkhard, L. P. Estimating dissolved organic carbon partition coefficients for nonionic organic chemicals. Environ. Sci. Technol. 2000, 34, 4663–4668. Krop, H. B.; van Noort, P. C.; Govers, H. A. J. Determination and theoretical aspects of the equilibrium between dissolved organic matter and hydrophobic organic micropollutants in water (Kdoc). Rev. Environ. Contam. Toxicol. 2001, 169, 1–122. Wauchope, R. D.; Linders, J. B.; Yeh, S.; Kloskowski, R.; Tanaka, K.; Rubin, B.; Katayama, A.; Kordel, W.; Gerstl, Z.; Lane, M. Pesticide soil sorption parameters: theory, measurement, uses, limitations and reliability. Pestic. Manage. Sci. 2002, 58, 419– 445. Escher, B. I.; Schwarzenbach, R. P. Partitioning of substituted phenols in liposome-water, biomembrane-water, and octanol-water systems. Environ. Sci. Technol. 1996, 30, 260– 270. Landrum, P. F.; Nihart, S. R.; Eadle, B. J.; Gardner, W. S. Reversephase separation method for determining pollutant binding to aldrich humic acid and dissolved organic carbon in natural waters. Environ. Sci. Technol. 1984, 18, 187–192. Backhus, D. A.; Gschwend, P. M. Fluorescent polycyclic aromatic hydrocarbons as probes for studying the impact of colloids on pollutant transport in groundwater. Environ. Sci. Technol. 1990, 24, 1214–1223. Ozretich, R. J.; Smith, L. M.; Roberts, F. A. Reverse-phase separation of estuarine interstitial water fractions and the consequences of C18 retention of organic matter. Environ. Toxicol. Chem. 1995, 14, 1261–1272. Mayer, P.; Tolls, J.; Hermans, J. L. P.; Mackay, D. Equilibrium sampling devices. Environ. Sci. Technol. 2003, 37, 184A–191A. Bondarenko, S.; Spurlock, F.; Gan, J. Analysis of pyrethroids in sediment porewater by solid phase microextraction. Environ. Toxicol. Chem. 2007,in press. Van der Wal, L.; Jager, T.; Fleuren, R. H. L. J.; Barendregt, A.; Sinnige, T. L.; Van Gestel, C. A. M.; Hermens, J. L. M. Solidphase microextraction to predict bioavailability and accumulation of organic micropollutants in terrestrial organisms after exposure to a field-contaminated soil. Environ. Sci. Technol. 2004, 38, 4842–4848. Yang, W.; Spurlock, F.; Liu, W.; Gan, J. Inhibition of aquatic toxicity of pyrethroid insecticides by suspended sediment. Environ. Toxicol. Chem. 2006, 25, 1913–1919. 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, 287–296. Hawthorne, S. B.; Azzolina, N. A.; Neuhauser, E. F.; Kreitinger, J. P. Predicting bioavailability of sediment ploycyclic aromatic hydrocarbons to Hyalella azteca using equilibrium partitioning, supercritical fluid extraction, and pore water concentrations. Environ. Sci. Technol. 2007, 41, 6297–6304.

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