Ecotoxicological Effects of Activated Carbon Addition to Sediments

Jul 8, 2009 - Cynthia C. Gilmour , Georgia S. Riedel , Gerhardt Riedel , Seokjoon .... Hans Peter H. Arp , Caroline Raymond , Göran Samuelsson , Jenn...
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Environ. Sci. Technol. 2009, 43, 5959–5966

Ecotoxicological Effects of Activated Carbon Addition to Sediments MICHIEL T. O. JONKER,* MARTIN P. W. SUIJKERBUIJK, HEIKE SCHMITT, AND THEO L. SINNIGE Institute for Risk Assessment Sciences, Utrecht University, P.O. Box 80177, 3508 TD Utrecht, The Netherlands

Received February 19, 2009. Revised manuscript received June 18, 2009. Accepted June 19, 2009.

Activated carbon (AC) addition is a recently developed technique for the remediation of sediments and soils contaminated with hydrophobic organic chemicals. Laboratory and field experiments have demonstrated that the addition of 3-4% of AC can reduce aqueous concentrations and the bioaccumulation potential of contaminants. However, one aspect of the technique that has hardly received any attention is the possible occurrence of secondary, eco(toxico)logical effects, i.e., effects of AC addition on the health, behavior, and habitat quality of local organisms. In the present study, several ecotoxicological effects were investigated in AC-water and ACenriched (0-25%) sediment systems. It was demonstrated that (i) powdered activated carbons can be toxic to aquatic invertebrates (Lumbriculus variegatus, Daphnia magna, and Corophium volutator) based on different mechanisms and preferably should be washed prior to application; (ii) Asellus aquaticus and Corophium volutator may physically avoid ACenriched sediments; (iii) exposure of Lumbriculus variegatus to AC-enriched sediments lead to a time and dose-dependent reduction in the worms’ lipid content, which was most probably caused by the observation that (iv) worm egestion rates decreased drastically upon AC addition, indicating that the presence of AC disturbed feeding behavior; and (v) there were no obvious effects on the microbiological community structure. All in all, these results suggest potential ecotoxicological effects of powdered AC addition and stress the need for a detailed further investigation of secondary effects of the technique, prior to any large-scale field application.

Introduction As a spin-off of the recent discovery that condensed carbon materials (i.e., soot, charcoal, coal) cause strongly enhanced sorption of hydrophobic organic chemicals (HOCs) in sediments and soils (1-3), environmental researchers and engineers have started to deliberately add anthropogenic condensed carbon (i.e., activated carbon; AC) to HOCcontaminated environments. The objective of such additions is to reduce freely dissolved HOC concentrations, thereby diminishing ecotoxicological and human health risks, as these concentrations are considered available for uptake into organisms (4). Based on this principle, AC addition techniques are presently being developed as in situ alternative for the more expensive and invasive dredging and disposal approach currently applied. So far, several laboratory studies have * Corresponding author phone: +31 30 2535338; fax: +31 30 2535077; e-mail: [email protected]. 10.1021/es900541p CCC: $40.75

Published on Web 07/08/2009

 2009 American Chemical Society

supported the remediation potential of AC addition, by showing strongly reduced aqueous and bioaccumulated concentrations of PCBs, PAHs, and DDT upon the addition (5-13). Very recently, AC addition was applied in situ, and preliminary results indicated decreased bioaccumulation under field conditions as well (14). From a chemical point of view, AC addition therefore seems promising. Nevertheless, in situ application may have some cons, and several physical/ chemical objections have been forwarded (15), including mobilization of highly contaminated, light AC particles and their uptake into the aquatic food chain. Additionally, indications exist that the technique might have unwanted secondary ecological effects. A priori one might not expect interactions between AC and organisms due to the carbon particles’ inert character. Still, Jonker et al. (16) observed that the lipid content of aquatic worms (Limnodrilus sp.) was reduced by about 28 and 38% upon a 4-week exposure to sediment enriched with 1.5% of powdered coal and charcoal, respectively. The authors hypothesized that the carbonaceous materials might be able to adsorb nutritious compounds (dissolved organic matter, bacteria) strongly, thereby limiting food availability or quality for specific organisms. Furthermore, Millward et al. (17) reported a 50% growth rate reduction for the polychaete Neanthes arenaceodentata, when exposed for 28 d to sediment enriched with 3.4% of powdered AC. Finally, McLeod et al. (8) observed a reduction in the condition index (i.e., wet weight divided by shell length) of the clam Corbicula fluminea with increasing concentration of the same AC in sediment. On the other hand, in their study, Millward et al. (17) did not observe a significant decrease in survival, growth rate, fecundity, and lipid content of the amphipod Leptocheirus plumulosus; no effects on survival and lipid content of Neanthes arenaceodentata; and no effects on digestive enzymes and digestive fluid surfactancy in both organisms. Also, Cornelissen et al. (6) reported no reduction in lipid content of two benthic species (a polychaete and a gastropod) exposed to three different sediments enriched with 2% of powdered activated charcoal. Finally, Ho et al. (18) observed 100% survival of an amphipod and a mysid shrimp in sediment enriched with 15% of powdered coconut charcoal. In summary, the information on possible effects of AC on biological parameters is limited, and the available data are contradictory. Possibly, the occurrence and extent of effects depend on the sediment, the AC, and/or the organism investigated. If AC addition were to be considered a serious in situ remediation technique however, a thorough assessment of the existence, nature, and seriousness of such effects would be a prerequisite. After all, the objective of remediating sediments should (in at least most of the cases) not only be to remove contaminants or reduce their concentrations but also to restore a healthy environment for (benthic) organisms. Therefore, in the present study we made a next step toward an eco(toxico)logical risk assessment of AC addition to sediments by investigating the effects of the addition of various powdered ACs to different sediments on several biological parameters.

Experimental Design The experimental design of the present study is summarized in Table 1. First, in order to investigate whether ACs are toxic to aquatic organisms, survival of three different, field relevant aquatic invertebrates for which standard toxicity assays are available (Lumbriculus variegatus, Daphnia magna, and Corophium volutator) was studied in AC-only assays, emVOL. 43, NO. 15, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Overview of Experiments

ploying five different types of powdered AC. The ACs were selected on the basis of their particle size (powdered ACs are mostly used for remediation experiments) and to represent different manufacturers/source materials (see Table S1). Next, one type of AC was added to one marine and four fresh water sediments (ranging in organic carbon content) to obtain AC concentrations of 0, 1, 2, 4, 7, 10, 15, 20, and 25 (dry weight) %, and the following series of ecological tests was performed: (i) survival of C. volutator and L. variegatus as a first tier, crude response; (ii) avoidance/preference behavior of Asellus aquaticus and C. volutator to test whether organisms physically accept AC-containing habitats. So far, avoidance tests have mainly been applied for studying the effects of chemical contamination (e.g. refs 19 and 20), but tests with C. volutator have been used recently to study the effects of ‘physical’ matrix constituents as well (21); (iii) possible changes in lipid content of L. variegatus. The occurrence of such changes would indicate an impairment of food uptake, either passively or actively; (iv) egestion rates of L. variegatus. Egestion rates are proportional to ingestion rates and thus can be used to judge food uptake and to explain possible effects observed during lipid content determinations; and (v) bacterial community structure (analyzed by terminal restriction fragment length polymorphism, T-RFLP). AC is capable of adhering bacteria (22), possibly leading not only to an initially reduced viability (23) but also to colonization and biofilm formation (24). Depending on changes in substrate availability or the percentage of surface-attached bacteria, changes in the community composition might thus occur. Finally, one sediment was selected to which 4% of the five different ACs was added, after which the ecological tests with worms were repeated. Although the higher AC concentrations (10-25%) may not seem realistic from an application point of view, it should be stressed that the present study was a mechanistic one searching for trends, which legitimizes the broad range tested. Also, when adding AC in the field and aiming at a concentration of about 4%, insufficient mixing or water movements may cause local concentrations to exceed the intended percentage (25). Most experimental effort in this study was put on investigating 5960

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only one (instead of more types of) AC but combined with several types of sediment and species as the occurrence of effects probably depends on the type of sediment and/or organism rather than on the type of AC. After all, the same type of AC (i.e., Calgon TOG) added to one type of sediment did induce effects in one, but not in another organism (17), and caused effects (i.e., reduced condition index) to a different extent when added to different sediments (8, 13).

Materials and Methods Chemicals and Activated Carbons. n-Hexane and acetone (Pestiscan grade) were obtained from Lab-Scan (Dublin, Ireland). Ottawa sand (general purpose grade; 20-30 mesh) was purchased from Fisher Scientific (Loughborough, UK). Activated carbons used were the following: Organosorb 200-1 (Desotec Activated Carbon; Roeselare, Belgium), activated charcoal (C3345; Sigma-Aldrich; Steinheim, Germany), Norit SAE Super (Norit Activated Carbon; Amersfoort, The Netherlands), Carbopal MB 4 (Donau Carbon; Frankfurt, Germany), and Hydraffin P 800 (Donau Carbon). Available physicochemical characteristics of the carbons are presented in Table S1. Sediments. Five sediments were collected from different locations in The Netherlands: one relatively uncontaminated marine sediment from location Oesterput (OP; Eastern Scheldt estuary) and four freshwater sediments from locations Leidschendam (A4; a PAH and oil-contaminated canal close to the highway A4), Langbroek (LB; a PAH-contaminated rural ditch), and Gendt (GE and GC; sediments sampled from the edge and the center, respectively, of a relatively uncontaminated floodplain lake along the river Waal). GC sediment was sampled using a boat and a box corer; the other sediments were sampled with a custom-made sampling bucket from the shallow water systems. All sediments were sieved on site through a 500 µm sieve and were stored at 4 °C until use. Organic carbon contents of all sediments were determined as described before (16) by CHN analysis on freeze-dried and decalcified (HCl) samples and measured 1.00, 2.80, 3.47, 4.10, and 4.30% for OP, GE, GC, LB, and A4, respectively.

Organisms. Aquatic oligochaete worms (Lumbriculus variegatus) were reared in the laboratory in 45 L aquaria at 24 °C. Chlorine-free cellulose served as substrate, and the aquaria were continuously flushed with preheated, copperfree tap water. Once a week, the deposit-feeding organisms were fed with pulverized flake fish food. Prior to experiments, organisms were separated from the substrate and allowed to clear their guts overnight in gently flowing water. Water fleas (Daphnia magna) were obtained from the Aquasense laboratory (Amsterdam, The Netherlands), where they were reared in copper-free tap water at 20 °C. Asellus aquaticus were sampled at the Duno country estate (Doorwerth, The Netherlands), transported back to the laboratory, and kept at 20 °C for 3 days for acclimatization purposes in an aerated 30 L aquarium, containing birch leaves collected on site. Marine mud shrimp (Corophium volutator) were collected from location OP and acclimatized to the testing temperature (20 °C) in aerated buckets with OP sediment and seawater for 3 days prior to the experiments. AC Addition to Sediments. Washed ACs (see below for washing procedure) were added to the sediments to obtain AC concentrations in all sediments of either 0, 1, 2, 4, 7, 10, 15, 20, or 25 dry weight %, or 4 dry weight % (see Table 1). About 1.5 L of wet, homogenized freshwater sediments was put into 2 L polypropylene jars and based on measured dry weight contents AC was added on a weight basis. For the marine sediment (OP), AC was added likewise to about 0.6 L of wet sediment in 1 L polypropylene jars. The jars were closed, shaken manually, and degassed by removing the lids again (AC additions caused increased pressure). Subsequently, all jars were closed again, lids were sealed with tape, and all systems were homogenized on a roller couch for one week. Finally, they were stored at 4 °C until use. AC-Only Toxicity Tests. Increasing amounts (1.5 to 4000 mg) of all 5 ACs were separately added to 50 mL vials, together with 40 mL of water (tap water for L. variegatus, copper-free water for D. magna, and filtered Eastern Scheldt water for C. volutator). Vials were shaken and left for a day before either 10 worms, 5 water fleas, or 5 mud shrimp were added. Exposure lasted for 14, 2, or 3 days, respectively. All exposures were performed at 20 °C and a 12:12 h light:dark cycle. Each individual exposure (organism-AC-concentration) was performed singular (water fleas), in triplicate (mud shrimp), or in quadruplicate (worms). After exposure, living organisms were counted by sieving media through a 500 µm sieve. Experiments were performed with the untreated (as received) AC materials and with ACs that had been washed as follows: about 400 g was heated to about 80-100 °C for 30 min in 2.5 L of tap water while stirring. After particle settlement (about 30 min), the water was decanted, and the procedure was repeated twice. Finally, the coals were dried at 85 °C for 3 days. Tests with Corophium volutator were performed only for all unwashed ACs and washed Organosorb AC, due to limited availability of these organisms. Survival in AC-Containing Sediments. Subsamples of about 20 mL of the (AC-containing) freshwater sediments were put in 50 mL glass vials, and 25 mL of tap water was added. The sediments were allowed to settle for one day after which 10 worms (L. variegatus) were added to each vial. Worms were exposed to the sediments for 28 days without the addition of food and oxygen. However, 2/3 of the overlying water phase was refreshed twice a week. Survival of C. volutator exposed to AC-containing OP sediment was tested by adding about 50 mL to 200 mL jars. The water phase consisted of filtered Eastern Scheldt water, and 10 organisms were added per jar. During exposure (10 days), the systems were aerated, but organisms were not fed. Both worm and mud shrimp exposures were performed in triplicate. Upon

finishing all exposures, survival was determined by pouring the vial/jar contents in a white tub and counting living organisms. Avoidance/Preference Tests. Sediment avoidance/preference tests were performed with A. aquaticus and C. volutator on sediments containing 0, 4, 7, 15, or 25% of AC, according to a setup based on the one described in ref 19. Tests with A. aquaticus were performed in triplicate in randomly placed plastic boxes (16 × 11 × 6 cm), which were divided in half by a thin plastic insert. Each side was simultaneously filled with a 2 cm layer of freshwater (A4, LB, GE, or GC) sediment. The one side always contained AC-free (0%) sediment; the other half either 0% or AC-enriched sediment. A layer of three cm of tap water was carefully added to both sides, and the systems were left for a day for particle settlement. Then, the plastic insert was carefully removed, and the small crack remaining was leveled with a spoon. Seven organisms were subsequently placed in the center of the box. After 3 days at 20 °C and a 12:12 h light:dark cycle, the inserts were placed back, and the number of organisms at each side was recorded. Tests with C. volutator were performed in a similar manner; however, they were executed in quadruplicate in 250 mL glass beakers, containing a 3 cm layer of OP sediments and filtered Eastern Scheldt water. The beakers were continuously aerated with Pasteur’s pipettes, placed exactly in the middle-top 1 cm of the overlying water. Results of the avoidance/preference tests were interpreted statistically with a binomial test (R ) 0.1). Effects of AC Addition on Worm Lipid Content. About 50 mL of the freshwater sediments was added to 200 mL glass jars, resulting in a 3-4 cm-thick sediment layer. Subsequently, 100 mL of tap water was added, and sediments were allowed to settle for 1 day. Then, about 1.6 g of wet weight L. variegatus was added to each jar. All sediments were tested in triplicate, worms were not fed during exposure, and 2/3 of the overlying water was refreshed twice a week. Exposure duration was 4 weeks. Upon finishing the testing period, worms were separated from the sediments, allowed to clear their guts for 8 h (during which the water was refreshed twice), frozen at -18 °C, and freeze-dried prior to lipid extraction. Lipids were extracted with a hexane/acetone (3:1) mixture as follows: About 100 mg of dry material was weighed into 10 mL glass tubes, and 4 mL of the solvent mixture was added. The tubes were vortexed for 1 min, sonicated for 5 min, and centrifuged at 1500 rpm for 7 min. The supernatant was transferred to a weighed aluminum cup, and the extraction procedure was repeated twice. The three extracts were pooled and evaporated to dryness, and lipid contents were determined gravimetrically. Determination of Worm Egestion Rates. Egestion rates of L. variegatus exposed to the four (AC-containing) freshwater sediments were determined according to the method developed by Leppa¨nen and Kukkonen (26). In short, about 20 mL of sediment and 20 mL of tap water were successively added to 50 mL glass jars, yielding a sediment layer of about 3-3.5 cm. The systems were left to settle for one day, after which 5 worms were added to each jar. Small and active individuals were selected, as the chance on reproduction during exposure is the smallest for these animals (27). Worms were allowed to dig themselves in for 1 h, after which a 2 mm thick layer of Ottawa sand was carefully put on top of the sediment. Based on the results in ref 28, worms were allowed to reach a constant feeding rate first, for which they were given 5 days. During this period, fecal pellets were removed with a pipet from the top of the Ottawa sand layer once a day. After removal of the pellets, the water phase was cautiously replenished with fresh tap water, using a pipet. From day 6 on, all pellets produced were transferred to precombusted and weighed aluminum cups. Pellets were collected for 5 days, after which the cups were dried overnight VOL. 43, NO. 15, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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at 85 °C and weighed. Worms were collected from the exposure jars as well, allowed to clear their guts overnight in small beakers with tap water, dried in aluminum cups at 85 °C, and weighed. Results of the egestion rate experiments were then expressed as mg egested material per mg of worm per h. All exposures were performed in triplicate under the same light regime and temperature as described above. Effects on Bacterial Community: T-RFLP Measurements. To screen for changes in bacterial community structure, the A4 and OP sediments containing 0, 2, 4, 10, and 20% of AC were subjected to Terminal Restriction Fragment Length Polymorphism (T-RFLP) analyses, based on ref 29. DNA was extracted from 400 mg of the upper 2 mm layer of sediments that had been left for a few weeks at 4 °C using a beadbeating procedure (FastPrep Spin Kit for Soil, Bio101, Vista, US). After dilution of the DNA extract with filter-sterilized and autoclaved Millipore water to reduce possible PCR inhibition, eubacterial 16S genes were amplified by PCR using the primers 27F (AGAGTTTGATCCTGGCTCAG -FAM) and 1392R (ACGGGCGGTGTGTRC), Expand long template polymerase (Roche, Almere, The Netherlands), and BSA addition. The PCR reactions were performed under the following conditions: 94 °C 2 min initial denaturation followed by 25 cycles of 94 °C for 15 s, 55 °C for 30 s, and 68° for 2 min, followed by an extension at 68 °C for 7 min. Five µL of the first PCR product was used for a reconditioning PCR (30) of 94 °C 2 min initial denaturation followed by 3 cycles of 94 °C for 15 s, 55 °C for 30 s, and 68° for 2 min, with a final extension at 68 °C for 7 min. After purification of the PCR products (Qiaquick, Qiagen, Venlo, The Netherlands) and quantification by UV absorption, 15-40 ng of the PCR products were digested for 3 h at 37 °C with 5 U of HhaI (Promega, Leiden, The Netherlands) and precipitated. Analysis of fragments was performed on a genetic analyzer (ABI PRISM 3130 xl; Applied Biosystems, Foster City, CA) with a GeneScan ROX 2500 size standard (Applied Biosystems, Foster City, CA). Final results (peak areas retrieved in GeneScan, Applied Biosystems, Foster City, CA) were normalized to proportions of the total peak area and subjected to a noise filtering and binning routine (IBEST, University of Idaho). Principal component analysis (PCA) was performed on the final data matrix with the software R (R, Vienna, Austria).

Results and Discussion AC-Only Toxicity Tests. Dose-response assays with L. variegatus, D. magna, and C. volutator performed on the five ACs demonstrated that at high dosings all ACs were lethal to the worms and the water fleas (Figures S1-S3). The mud schrimp could stand the Hydraffin and Organosorb coal over the full dose range applied, but the other ACs appeared to be toxic to this organism as well. Some examples of dose-response curves are presented in Figure 1. Lethal concentrations (LC50 values) were within the range of 0.5-4 g/L for Daphnia magna and 4-60 g/L for Lumbriculus variegatus and Corophium volutator (Table S2). The observed toxicity of the ACs may have a physical or/and chemical cause. In the first case, effects could hypothetically be caused by e.g. obstruction of the gastrointestinal tract, impediment of movement, or adsorption of skin constituents (e.g., mucus) that may be essential for the organism’s survival. In the second case, suspending AC in water would lead to a change in solution chemistry (e.g., by release of AC-associated chemicals), thereby inducing toxic responses. Unraveling the exact cause of toxicity was beyond the scope of the present study and should be the subject of future experiments. Still, in order to obtain some initial indications, dose-response assays were repeated with washed ACs, ACs were analyzed for organic and inorganic contaminants, and dose-pH and dose-O2 relationships were determined. In the Supporting 5962

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FIGURE 1. Dose-response curves for Daphnia magna (], solid line), Corophium volutator (9, dashed line), and Lumbriculus variegatus (∆, dotted line) exposed to untreated Sigma-Aldrich activated carbon. Markers represent averaged values; for clarity standard deviations are omitted. Information a detailed description of the methods, results, and discussion is provided; here only the major results and preliminary conclusions are summarized: (i) washing the ACs increased the survival of all organisms (Figures S1-S3), (ii) the investigated ACs did not contain (bioavailable) toxic organic chemicals; (iii) the ACs did contain µg-mg/kg levels of a broad range of inorganics, and washing overall reduced the levels (Table S3); and (iv) suspending the ACs in water caused a dose-dependent decrease in oxygen levels (Figure S4) and a dose-dependent pH change, showing a decrease for the Sigma-Aldrich AC and an increase for the other ACs (Figure S5). From these results it can be inferred that even though AC is often considered a carbon-only, relatively inert material, the present samples are toxic to specific aquatic invertebrates, contained up to 3.5% of metals, represent an oxygen sink, and are capable of altering the aqueous pH. The cause of toxicity seems to be different for different organisms, as also suggested by the different dose-response curve shapes in Figures S1-S3. The present, preliminary results suggest physical effects and possibly metals or other inorganic chemicals for Daphnia magna; nontolerable pH values for Lumbriculus variegatus; and nontolerable pH values or possibly metals/other inorganic chemicals for Corophium volutator. To exclude bias by toxic effects in the next-stage experiments, the AC used (Desotec Organosorb 200-1) was washed prior to adding it to the sediments. Survival on AC-Containing Sediments. No significant mortality of L. variegatus and C. volutator was observed upon exposure to sediments containing up to 25% of Organosorb AC (see Figures S7 and S8). These results are in line with expectations (see above), although it should be noted that the current exposure period was 2-3 times as long as in the AC-only tests. Also, the data agree with previously reported observations of 100% survival of different organisms on ACenriched sediments (13, 14, 17, 18). This simple, first tier response thus suggests the absence of harmful effects of AC addition. However, a closer look at the exposure systems revealed an indication for subtle effects, i.e., a decreased turbidity of the overlying water phase with increasing AC % for all sediments/both organisms. This might point to e.g. decreased biological activity or increased binding of dissolved organic matter, as will be discussed below. Avoidance/Preference Behavior. Avoidance/preference tests with reference systems (boxes containing AC-free

sediments on both sides) showed that A. aquaticus distributed randomly on all four freshwater sediments, indicating a robust testing system. In systems containing AC-free sediment on the one and AC-enriched sediment on the other side, avoidance of the latter sediment was observed in several cases (see Table S5). Although the organisms appeared to accept AC in LB sediment (no avoidance behavior observed on this sediment), statistically significant avoidance was observed for the two highest AC concentrations (15 and 25%) in GC sediment, the highest AC % in GE sediment, and all but the 15% AC concentration in A4 sediment. Tests with C. volutator also demonstrated random distribution of the organisms in the reference systems. Obvious avoidance was observed for all AC-enriched sediments, except the one with the highest AC concentration (see Table S5). Although this last result and some of the results for A. aquaticus may be difficult to explain, overall the data indicate that organisms should not a priori be expected to accept habitats containing AC. The level of acceptance/avoidance however probably depends on the organism, the AC concentration, and the sediment. Previously, Hellou et al. (21) also observed avoidance of sediment containing (>30% of) charcoal and coal by C. volutator and suggested that the texture or chemistry of the coal (color, composition, smell, or taste) rather than the physical state may be involved in the avoidance behavior, since grinding the coal did not affect their results. Because in the present study the AC particles were smaller than particles generally occurring in both organisms’ natural habitat, our results support this hypothesis as well as do the results to be discussed below. Worm Lipid Contents. In Figures 2A and S9, lipid contents of L. variegatus after a 4-week exposure to the four freshwater sediments are plotted as a function of the AC % in sediment. Lipid contents are expressed relative to that of worms exposed to AC-free sediment (set at 100%), thereby focusing on the relative change in lipid content induced by the addition of AC. The figures show that for all sediments there was a decrease in lipid content upon AC addition, which was statistically significant in 60% of the cases (see Figure S9). For A4, LB, and GE sediment, contents seemed to stabilize beyond AC levels of 4-10%, whereas for GC sediment rather a gradual decrease over the full AC range tested seemed to occur. Reductions in lipid content of up to 13% (A4 and GC), 20% (LB), and 22% (GE) were observed, and for the AC addition percentage frequently applied (i.e., ∼4%) (9-11, 13, 14, 17), decreases of 13, 20, 4, and 10%, respectively, occurred. These reductions are less than previously reported in ref 16, but it should be stressed that both the testing species and coals applied in that study were different from the present ones. The fact that no reduction in lipid content was observed by Cornelissen et al. (6) might be explained by either the limited data resolution or the species investigated in that study. At least, the present study and the one reported in ref 16 indicate that significant lipid reductions are possible upon a 4-week exposure. Still, the results might reflect the specific type of AC used, and other ACs might perhaps not cause reductions. An additional small-scale test was therefore performed, in which A4 sediment was enriched with 4% of either of the five different ACs (see Table 1), including the one from Sigma-Aldrich used in (6). Very similar lipid reductions were however observed for the different ACtreated systems in this test, with a percentual standard deviation of only 3.5% (results not shown). As the source materials, acidity, particle size, surface area, and metal concentrations of the ACs differ, this result suggests a nonspecific response to the presence of AC particles in sediment. From a cost-benefit perspective, one might consider the above lipid reductions acceptable, because the organisms after all do survive. However, it should be stressed that the

FIGURE 2. Relative lipid percentage of worms exposed (A) for 4 weeks to LB (9) or GC (]) sediment, containing varying percentages of Organosorb AC; (B) to LB sediment containing either 0% (() or 10% (0) of Organosorb AC, as a function of time. Data for the 0% treatment in graph (B) do not extend beyond 10 weeks due to limited sample availability. Lines serve to guide the eye and do not represent model fits. results in Figure 2A provide a snapshot after a relatively short exposure period only. When exposed chronically, lipid contents might decrease further to less acceptable levels. To investigate this issue, the above-described experiment was repeated in a kinetic way: worms were exposed (singular) to LB sediment without and with 10% of Organosorb AC for different lengths of time, varying between 1 and 16 weeks. The results are presented in Figure 2B and demonstrate that the lipid content of L. variegatus indeed gradually decreased with time. After 4 months of exposure, the lipid pool was reduced to 74% of its initial level but to about 50% of that of worms exposed to sediment without AC, because the latter worms’ lipid percentage increased (the sediment apparently provided a more nutritious environment than cellulose). The question remains why worm lipid contents are reduced upon exposure to AC-enriched sediments. Excluding VOL. 43, NO. 15, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 3. Relative egestion rate (%) of worms exposed to GE sediment as a function of sedimentary AC percentage. Data points are averaged values ( standard deviation (n ) 3). The line serves to guide the eye. toxicity as an explanation (see above; the ACs were washed), hypothetically, there are four possibilities: (i) as oligochaete worms have a slight preference for small, low-density, organic-rich particles (31), L. variegatus might selectively feed on the added AC fraction. AC however has a presumed very low nutritional value, and the worms might consequently start metabolizing their lipid reserves in order to stay alive; (ii) due to its high binding capacity, AC might strongly sorb organic molecules/phases that serve as a food source for the worms (DOC, bacteria). As such, food availability may be limited in AC-enriched sediments (16), again forcing the worms to metabolize their lipid reserves. In particular for the high AC dosages, food availability may also simply be reduced by dilition; (iii) (selective) feeding on AC particles might somehow cause clogging of the worms’ feeding tract, thereby impairing food uptake and inducing lipid metabolization; and (iv) A. aquaticus and C. volutator tend to avoid AC-enriched sediments, a response most probably induced by ‘chemically’ sensing the particles (see above). In case L. variegatus would likewise be able to sense AC, it could try to avoid the material by reducing or ceasing ingestion. The egestion rate experiments discussed below were performed in order to shed more light upon this issue. Worm Egestion Rates. During the egestion rate experiments, most of the worms in all systems protruded their tails from the added sand layer, and no reproduction was observed. Averaged egestion rates determined for the ACfree sediments measured 0.15, 0.17, 0.11, and 0.12 mg dry feces/mg dry worm/h for A4, LB, GC, and GE sediment, respectively. These values are a factor of 3-5 lower than those reported by Leppa¨nen and Kukkonen (26, 28) for L. variegatus feeding on natural sediments at 20 °C. However, because the presently used worms were about three times smaller than those used by Leppa¨nen and Kukkonen (28) and egestion rate is positively correlated with worm size (27), it can be concluded that the worms in the control systems displayed normal feeding behavior. In Figures 3 and S10, the feeding rates of Lumbriculus variegatus in AC-enriched sediments are presented relative to those in the respective AC-free sediment. Remarkably, egestion rates in all ACenriched sediments declined by on average 92% to ‘background’ rates of only 0.004-0.025 mg dry feces/mg dry worm/ h. Hardly any differences were observed for different sediments, and the presence of only 1% of AC already appeared to disturb normal feeding behavior. As selective feeding on AC particles in the low AC treatment systems and 5964

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subsequent gut obstruction by these particles is not very likely, an actual reduction of the particle intake most probably explains these results. In line with the suggestion above, the worms may dislike the AC’s smell, taste, or appearance or may sense its high metal levels or acid/base surface groups and simply stop ingesting. In order to survive, they will need to metabolize their lipid reserves, which accounts for the gradual decrease in lipid contents (Figure 2B). Still, the data presented in Figures 3 and S10 do not fully explain the curve shapes in Figure 2A, as lipid fractions would be expected to drop to plateau levels already at 1% of AC added and certainly not to gradually decrease as observed for GC sediment. On the basis of the present experiments, a conclusive explanation for this discrepancy cannot be provided. In contrast to the current results, Figure 1 in ref 12 does show Lumbriculus variegatus feces on top of AC-enriched sediment. Sun and Ghosh (12) however placed the AC as a layer on top of the sediment, and worms were therefore still able to feed on pristine sediment particles. Although a similar way of AC application may preserve initial feeding behavior (note that Figure 1 in ref 12 represents a snapshot after 2 days), a prolonged exposure may result in feeding obstruction as well, since the AC will be worked into the sediment in time (12). On the other hand, in case chemical properties (metal levels, acidity, smell) of the AC would be involved in the feeding response, there might be ACs that are accepted by worms and whose addition would not lead to a similar large reduction in particle ingestion. To test this hypothesis, an additional egestion rate experiment was performed (see Table 1) in which LB sediment was enriched with either no or 4% of the five other ACs and homogenized for a week. The results (not shown) indicated that all ACs reduced egestion rates to a similar extent (with 93% for Organosorb and Sigma-Aldrich AC to 98% for Carbopal AC). This agrees with the lipid content reductions being similar for the different systems and thus again suggests a nonspecific response to the presence of AC particles in sediment. Bacterial Community Structure Tests. Principal component analysis (PCA) of the 16S gene profiles resulting from T-RFLP data indicated that the different sediments investigated inhabited bacterial communities with different structures (see Figure S11). The data however did not indicate differences in structure for samples relating to the same sediment, i.e., no significant trends in bacterial community structure could be identified going from 0 to 20% of AC addition, apart from a community shift in the OP sediment at an intermediate AC concentration (4%). Hence, the AC had no effect on the structure of the bacterial community detectable by T-RFLP. It has been reported that initial adherence of bacteria to AC can reduce their viability through effects on membrane integrity (23). However, because in the present study the AC had been equilibrated with sediment for several weeks, regrowth and formation of bacterial biofilms may have occurred, obscuring these direct effects. Furthermore, no major changes in bacterial genera in raw drinking water before and after granular AC treatment have been found with cultural techniques (32). Given the T-RFLP results, it can thus be assumed that there was no difference between the bacterial communities on sediment particles and communities that colonized the AC. It should be stressed that the present approach focused on community structure only and could not discriminate effects on community size. Also, effects on processes mediated by bacterial activity (e.g., nutrient cycling) in sediments cannot be precluded as the functionally active bacteria do not necessarily belong to the abundant species detectable in T-RFLP analyses. Perspective for Field Application of Activated Carbon. Nowadays, polluted sediments are remediated if necessary by dredging and disposal, a costly and invasive approach. Recent scientific literature suggests that activated carbon

addition is a promising, less expensive and less invasive alternative that at least from a chemical point of view is effective. However, field applications of the technique may be considered premature, as the present data indicate that adding AC may result in a variety of ecotoxicological effects. First of all, although activated carbons may be expected free of toxic (levels of) organic contaminants, this study demonstrated that considerable amounts of metals can be present, probably originating from the AC source material. Adding AC to sediments may thus increase local metal concentrations, although it should be noted that metal bioavailability was not studied in detail. Moreover, when applying an acidic coal (such as the Sigma-Aldrich AC), pH values may locally decrease, resulting in mobilization of native metals. As such, the one chemical stress factor (HOC contamination) would be replaced by another. Other effects on metal fate that theoretically might be expected relate to redox conditions, as AC was also demonstrated to represent an oxygen sink. Obviously, this property may also affect the general ecological status of sediments, but considering the excess of water in field situations, both oxygen reductions and pH changes may be expected small, although not absent on a microscale. By washing ACs prior to addition, the chance on pH changes may be reduced. As such, toxic effects caused by pH and oxygen changes as observed for specific organisms in the present study may not occur during field application, all the more since the AC dose generally applied is small compared to the bulk sediment. Still, toxic effects by physical stress as observed for Daphnia magna may occur in the field. Also, ecological responses may be expected in situ, in particular ones related to avoidance behavior of organisms. Even at a relatively low dose of 4%, which is generally applied (9-11, 13, 14, 17), benthic invertebrates may face a reduced habitat quality and deposit-feeding organisms like worms may drastically diminish their feeding rates and start living on their fat reserves. As demonstrated above, this can cause their lipid content to gradually decrease, presumably leading to impairment of growth and reproduction, and eventually even mortality. As this response was demonstrated in the present study and in ref 16 for different ACs and different sediments, it probably concerns a general phenomenon. Note that apart from ecological implications, a gradual decrease in lipid contents will also complicate the interpretation of bioaccumulation experiments with HOCs, as previously observed and discussed in ref 16. One might consider chemical status more important than a good ecological sediment quality. However, it should be mentioned that benthic species form the basis of the aquatic food web, and in case these organisms should become extinct, the aquatic ecosystem may be at risk. A remediation method should therefore not only remove ecotoxicological, human health, and leaching risks but also safeguard a good ecological status. AC addition does reduce the risks mentioned but possibly does not meet the last criterion. On the other hand, dredging removes all three risks as well but also destroys existing habitats and will remove the majority of organisms present. Still, dredging leaves the possibility of recolonization, whereas AC-containing sediment might perhaps be avoided or escaped by benthic fauna (note that AC addition is invasive to some extent as well, in case it is mixed mechanically with the sediment). The present data add to the growing body of evidence that AC addition might have ecotoxicological effects on sediment-dwelling organisms (8, 16, 17). Still, it should be recalled that some authors did not observe similar effects (6, 17, 18). Also, the present study solely focused on powdered ACs, whereas sometimes granular materials may be used, for which ecotoxicological effects might be different or absent. Therefore, the AC addition remediation technique should not be discarded, but additional research on secondary effects

is certainly advisible. Furthermore, because possible effects (may partly) depend on the type of AC and the specific sediment to be remediated, it is recommended to implement a standard test battery to be applied to each combination of sediment/AC and evaluating at least (i) metal concentrations and pH of the AC to be applied (acidic ACs are not recommended for use), (ii) the leaching potential of metals, and (iii) egestion rates of worms as affected by the addition of the intended percentage of AC. Egestion rate determinations (26, 28) are simple but sensitive and, as argued above, represent an ecologically relevant response parameter (27). Finally, washing AC prior to application could be beneficial as it moderates the pH and will increase particle settlement by removing air from pores (5, 9). Results of the suggested tests, perhaps combined with those of future tests to be selected or developed, should preferably be part of a thorough cost (ecotoxicological/financial)-benefit (chemical/risks) analysis prior to AC field application.

Acknowledgments We kindly acknowledge the practical assistance, advice, and/ or materials provided by John Beijer, Elfriede Burger, Marco Dubbeldam, Frits Gillissen, Gerdit Greve, Stephan van der Heijden, Stefan Kools, Matti Leppa¨nen, Barry Muijs, Frank van Steenbeek, Manon Vos-Loohuis, and Helen de Waard.

Supporting Information Available Additional text, tables, and figures. This material is available free of charge via the Internet at http://pubs.acs.org.

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