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Sediment Quality Assessment and Management: Insight and Progress Edited by M. Munawar© 2003 Ecovision World Monograph Series Aquatic Ecosystem Health & Management Society Contamination in intertidal areas: risks of increased bioavailability P.J. den Besten Institute for Inland Water Management and Waste Water Treatment (RIZA),P.O. Box 17, 8200 AA Lelystad, The Netherlands. Keywords: sediment pollution, intertidal waters, aquatic oligochaetes Introduction As a result of river pollution, sediments in the delta of the rivers Rhine and Meuse have become polluted with heavy metals, PCBs, organochlorine pesticides and PAHs. Now that remedial action has become an option to reduce the risks for ecosystems (because present quality of suspended solids is much better), priorities should be given to high-risk situations (Den Besten et al., 2000). The ecologically important intertidal flats contain high levels of contaminants, causing high risks for wader birds that are dependent on these areas for their food. In addition, the bordering terrestrial ecosystems (river banks and floodplain meadows) have received a considerable input of contaminants, as a result of the frequent flooding during the winter season. The effects caused by sediment pollution depend on the chance that (benthic) organisms are exposed to the contaminants present. The intensity of exposure is determined by the bioavailability of the contaminants. As a general rule, for organic chemicals with log Kow 5, uptake from ingested sediment particles can contribute significantly to the accumulation (Belfroid et al., 1996). Also for metals, there are clear indications that at least soft-bodied organisms bioaccumulate directly from pore water (Janssen et al., 1997; Spurgeon and Hopkin, 1999). Sediments with a comparable degree of contamination (total of dissolved 194 and sediment-bound) but with different sorption or complexation characteristics will exhibit differences in the bioavailability of the contaminants. The aim of the presentstudywastocomparecontaminantbioavailabilityinsedimentfromsublittoral waters, littoral/intertidal zones and floodplains. Bioavailability was assessed by combiningasedimentbioassaywithchemicalendpoints:measuringtheaccumulation of contaminants in aquatic oligochaetes after exposure to sediment from different locations. Material and methods Collection of sediment samples Sediments samples were collected in 1995 (Meuse) or 1998 (other locations). Sediment samples for chemical analyses and bioassays were collected with a 0.2 x 0.3 m box-corer. At the most shallow sampling points, 18 grabs were taken with a 65 mm diameter corer and pooled to give the same sampling area. Sublittoral sampling locations were chosen for river sediment in the Amer, Hollandsch Diep and the Haringvliet at a depth of > 2m. Littoral zones (depth < 2 m, but permanently submerged) were sampled in the Haringvliet and in the creeks of the floodplain forest Sliedrechtse Biesbosch. Sediments from the littoral zones in the Haringvliet are more strongly influenced by the tidal movement of (fresh!) water and waves induced by wind than in the Biesbosch. At a floodplain of the river Meuse near Eysden samples were collected from sediment that stood clear of water after seasonal flooding. Clean reference sediment was collected at a deep location (3 m) in the Drontermeer (Northern part of the Netherlands). A map of the area of interest has been published elsewhere (Den Besten et al., 1995). Accumulation bioassays The accumulation bioassays with aquatic worms were performed according to the method described by Maas et al. (1993). A sediment-water system was prepared by mixing 1 litre of sediment with 4 volumes of artificial fresh water in a glass aquarium after which the sediment was allowed to settle for two days. The test was started by transferring 20 g of worms to the sediment-water system and exposing the organisms for 28 days at 20°C in darkness. Oligochaetes were obtained from commercially cultured stocks. The dominant species were Limnodrilus claparedeianus (family: Tubificidae) and two species of the family of Naididae: Pristina idrensis and Homochaeta naidina. Before starting the bioassays the worms were kept on clean reference sediment for one month. During the test, oxygen concentration, pH, NO2 - concentration, NH4 + concentration and conductivity 195 were measured routinely to check whether validity criteria were met (Maas et al., 1993). The worms were fed minimally by adding once per week 1 ml of 10% Trouvit. The bioassays with sediment from the floodplain of river Meuse were performed in a 600 ml: 2400 ml sediment/water system and by feeding the worms three times per week by adding 5 ml of 10% Trouvit. The bioassays were terminated by washing the worms on a 250 m mesh sieve in tap water, after which the worms were placed for 24h in water from the sediment-water test system to allow clearance of gut contents. The worms were then transferred to a 200 m mesh sieve in Dutch Standard Water. Adherent water was removed by drying the sieve for one minute on filter paper, after which the sieve with worms was weighed. For analysis of metals and organic contaminants, worms were transferred to polythene and glass vials, respectively, and stored at 20 °C. Bioassays were performed in duplicate; for contaminant analysis the worms from two bioassays were pooled. Also, at the start of the bioassays, unexposed worms (kept on reference sediment) were stored for chemical analysis. Contaminant analysis Contaminant analysis in biota (and in sediment from accumulation bioassays) includedheavymetals(Hg,Cd,Pb,Cu,Zn,Cr,Ni),polycyclicaromatichydrocarbons (PAHs, 16 standard compounds proposed by US-EPA), standard PCBs (IUPAC congeners nrs. 28, 52, 101, 118, 138, 153 and 180) and the following chlorinated pesticides or related transformation products: pentachlorobenzene, hexachlorobenzene, octachlorostyrene, hexachlorocyclohexane stereoisomers: αHCH , ß-HCH, γ-HCH (lindane), dieldrin, endrin, heptachlor, heptachlorepoxide, o,p’-DDE, p,p’-DDE, o,p’- DDD, p,p’-DDD, o,p’-DDTand p,p’-DDT.Trace metals were analysed in microwave acid-digested samples with Atomic Absorption Spectrometry (flame-, HGA-, and cold vapour-AAS) according to methods previously described in Van Hattum et al. (1991, 1996a). PAHs were analysed in acetone/hexane Soxhlet extracts, after clean up (deactivated alumina adsorption column), using High Performance Liquid Chromatography (HPLC) with fluorescence detection, as described in detail in Wegener et al. (1992) and Van Hattum et al. (1998). PCBs and chlorinated pesticides were determined in different fractions (deactivated silica column) of purified (deactivated alumina column) acetone/hexane Soxhlet extracts with gas chromatography with electron capture detection (GC-ECD; see Leonards et al., 1993) Quality control monitoring of the chemical analysis included analysis of procedural blanks, internal control samples and certified reference materials. 196 Calculation of BSAFs The results are presented as biota-sediment accumulation factors (BSAFs), defined as the ratio between the level of a contaminant in the oligochaetes and that in the sediment. Organic contaminants in sediment and worms were expressed on organic carbon (O.C.) and lipid basis, respectively: BSAF organic contaminant = [contaminant] in lipid in organism / [contaminant] in o.c. in the sediment. The BSAF for heavy metals was calculated using levels of heavy metals in sediment and worms on dry weight basis: BSAF heavy metal = [metal] in organism (dry weight based) / [metal] in the sediment (dry weight based). Levels of heavy metals in sediment were normalized to standard sediment containing 10% organic matter and a fraction of 25% of grains <2m, according to Dutch standard methods (CUWVO, 1990): [metal]normalized=[metal]measured* (A + B*25 +C*10) (A + B*%GS<2μm + C*%OM). With: A, B and C being metal specific constants related to binding properties to silt and organic matter (see Table 1); %GS<2μm: the fraction of grains < 2 μm (weight based %); %OM: percentage organic matter Results Table 2 summarises some characteristics of the sediments sampled. Based on an earlier published classification scheme (Reinhold-Dudok van Heel and Den Besten, 1999), the sediments from the littoral zones in the Haringvliet, as well as the sediments from the Amer and the Meuse floodplain were characterised as sand or silty sand. The other sediments were more silt-rich. The reference sediment and the sediments from the littoral zones in the Haringvliet were relatively clean, whereas the sediments from the Meuse floodplain and the Hollandsch Diep contained high levels of one or more contaminants. After 28 days of exposure to the sediment samples, recovery of the worms varied between 50-70% (on wet weight basis) of the original biomass at the start Table 1. Constants for normalization of levels of heavy metals in sediment. (CUWVO, 1990) Metal A B C Cd 0.4 0.007 0.021 Hg 0.2 0.0034 0.0017 Zn 50 3 1.5 Cu 15 0.6 0.6 Cr 50 2 0 Ni 10 1 0 Pb 50 1 1 Table 2. Characteristics of the sediments sampled in the Rhine-Meuse delta. Location Depth Sediment characteristics 2) (m) 1) (receiving water % % silt Cd Hg PCB-153 BbF(mg/kg) from river Rhine O.C. (fraction (mg/kg) (mg/kg) (μ/kg) and/or Meuse) <63 um) Amer (M) 7.6 1.8 29 3.1 0.2 11 0.4 Hollandsch Diep (R/M) 7.7 5.6 76 4.1 0.6 26 0.8 Haringvliet 2.0 1.0 12 0.4 0.1 1 0.04 (R/M): littoral zone with sandy sediment Haringvliet (R/M): 1.5 1.4 21 1.0 0.1 1 0.2 littoral zone with silty sediment Haringvliet (R/M): 8.4 4.4 74 6.1 0.7 25 0.6 deep parts Biesbosch (R/M): 2.0 4.8 67 5.3 1.5 28 0.6 creeks of floodplain forest Meuse floodplain - 1.8 5.2 13.6 0.4 16 1.6 Reference sediment 3.0 8.4 92 0.7 0.1 2 0.2 1) mean water depth at location where sediment was sampled. 2) mean percentages and levels on dry weight basis. 197 198 of the bioassay. No differences were found between the various locations in this respect (not shown). Oxygen levels varied between 4.4 and 8.7 mg/L, pH between 7.1 and 8.3, ammonia (NH4 +) between 0 and 30 mg/L, nitrite (NO2 -) between 0 and 10 mg/L, and conductivity between 62 and 105 μS/mm. For most contaminants measured it was found that the accumulation levels after the exposure to the contaminated sediments were significantly higher than those in the worms from reference sediment. In Figure 1 and 2 the results of the bioaccumulation assays are shown for a selection of contaminants: the heavy metals cadmium and mercury, and the organic contaminants PCB-153 and benzo[b]fluoranthene. The BSAFs of cadmium and mercury were lowest in the river sediments, and significantly higher in the sediments from the littoral areas in the Haringvliet and in the dry sediment from the floodplains of river Meuse. The location-specific variation of the BSAFs was up to a factor of 30 (Fig 1). Comparable results were found for PCB-153 and benzo[b]fluoranthene. The bioavailability of PCB-153 in silt-rich sediment from the littoral zone in Haringvliet was six times higher than in river sediment from deep parts. For benzo[b]fluoranthene the BSAF differed 20 fold between the dry sediment from the floodplain of river Meuse and the river sediments (Fig 2). For several other heavy metals (chromium, nickel, lead), for all other PCBs measured and most organochlorine pesticides and PAHs similar trends were observed as for the compound described above (not shown). Discussion Accumulation studies with aquatic or terrestrial oligochaetes have shown that for most heavy metals and organic compounds with log Kow up to 7, equilibrium of internal steady state concentrations in oligochaete worms is reached within two weeks (Peijnenburg et al., 1999; Olliver, 1987; Harkey et al., 1995). For cadmium a linear uptake in the terrestrial oligochaete Enchytraus cryptus was found for 2 weeks (Peijnenburg et al., 1999). According to the equilibrium-partitioning (EP) theory, for heavy metals a theoretical biota-sediment accumulation factor (BSAF) can be calculated when the Kd (coefficient of the distribution between sediment and water) and the BCF (the ability of a specific species to accumulate a compound from the water) are known (Van der Kooij et al., 1991). Using BCF and Kd data from literature, for cadmium and mercury theoretical values of 0.04 and 0.09 are calculated. The BSAFs of cadmium and mercury of the present study that were measured in river sediments are comparable to these values, suggesting that equilibrium was reached within 28 days. At the same time, for sediments in littoral zones and dry sediment from the floodplain of river Meuse much higher BSAFs were found. The high bioavailability of metals in the dry sediment from the floodplain Fig. 1. Bioavailability of cadmium and mercury in sediments from the delta of the rivers Rhine and Meuse. Bars represent means of three assays ± S.D. (when no S.D. is given, n=2). Fig. 2. Bioavailability of PCB-153 and benzo[b]fluoranthene in sediments from the delta of the rivers Rhine and Meuse. Bars represent means of three assays ± S.D. (when no S.D. is given, n=2). 199 200 of river Meuse was also found in bioaccumulation tests with terrestrial oligochaetes (Den Besten et al., 2001). For all organic contaminants the equilibrium-partitioning theory predicts a value of 1.7 (Van der Kooij et al., 1991). In the case of PCBs and some organochlorine pesticides, BSAFs close to 1.7 were found in the river sediments, whereas clearly elevated bioavailability was found in sediment from littoral zones and dry sediment from the floodplain of river Meuse. For the PAH benzo[b]fluoranthene and some organochlorine pesticides the results were somewhat different, in that the BSAFs in river sediment are well below the theoretical 1.7, while for sediment from littoral zones and dry sediment from the floodplain of river Meuse BSAFs were closer to expected bioavailability according to the equilibrium-partitioning theory. The fact that the bioavailability of organochlorine pesticides and PAHs is considerably lower than EP predicts has been reported by other investigators (Tracey and Hansen, 1996). For the accumulation of PAH in oligochaetes from sediment in the floodplain forest of Sliedrechtse Biesboch a better correlation was found with the Tenaxextractable fraction (extraction of the rapidly desorbing fraction by the method of Cornelissen et al., 2001) than with the total level in sediment (Postma and Den Besten, 2001). This underlines the idea that the sediment pore water is the dominant source of exposure. The present study demonstrates that there may be a considerable locationspecific variation in the bioavailability of contaminants. To some extend this may be explained by differences in the grain size distribution of the sediments, resulting in a higher bioavailability in sandy sediments as compared with silty sediments. These findings are in line with results from others (Kukkonen and Landrum, 1995; Harkey et al., 1994) and indicates that sorption of contaminants over different grain size classes is not solely dependent on the organic carbon content. Also selective feeding of the oligochaetes on silt-rich sediment fractions could explain the higher BSAFs found for sandy sediments. These explanations may account for the differences in BSAFs between river sediments and the more sandy sediments from littoral zones. However, comparing the measured BSAFs in dry sediment withthosemeasuredinthelittoralsediments,itseemslikelythatespeciallyfluctuating hydrologicalconditionsinfluencebioavailabilityofcontaminants.Theresultsobtained for recently deposited sediment that stood clear of water, are in line with earlier observations in a number of landfarming experiments. An increase in the bioavailability of metals, PCBs and the pesticide dieldrin was also found during the landfarming of polluted sediment from a harbour in Rotterdam (Den Besten, 1995; Den Besten et al., 2001). In this case, a peak in the bioavailability of cadmium was observed after 28 months. The effect is most likely related to long-term processes by which soil is formed from sediment, resulting in changes in the structure and sorption characteristics of the organic matter. The increased metal bioavailability 201 could not be explained by changes in the pH of the soil during the landfarming experiment (remained around 7.5). In aquatic oligochaetes collected in the Rhine-Meuse delta or exposed in the laboratory to contaminated sediments from this same area (like in the present study), accumulation levels of cadmium, mercury and PCBs are found to exceed “maximum tolerable risk levels” by more than a factor 10 (Van Hattum et al., 1996b; Den Besten et al., 2000). Therefore, in terms of risk, these contaminants are considered as critical for effects occurring via accumulation in foodchains. The results of the present study suggest that the risks of sediment contamination may be higher in intertidal areas and in floodplains than in sublittoral waters. Indeed, measurements and calculations of metal concentrations in pore water of floodplain soils indicate increased mobility of metals with fluctuating water levels (Van den Berg et al., 1998, 1999). The same may be the case in littoral water where there is a tidal movement of surface water, e.g. in shallow areas bordering the Hollandsch Diep and Haringvliet, and especially in the tidal floodplain forests in the delta of the rivers Rhine and Meuse, known as the Biesbosch. The question remains how risks of contaminants should be compared between different biotopes. Conventional sediment/soil quality assessment methods, like the Triad approach, have focused either on aquatic or terrestrial ecosystems, but these methods cannot be used to have a direct comparison in risk levels. In order to set the right priorities between possible sediment and/or soil remediations in polluted waters and bordering floodplains, a decision support system (DSS) is proposed as outlined in Figure 3. This first step consists of a risk characterisation performed on sediment or soil, involving aquatic or terrestrial tests and related ecological measurements. This step ideally already includes chemical or ecotoxicological techniques (such as the accumulation bioassay used in the present study) that quantify the bioavailable fractions of the contaminants. In addition attention is given to possible risks for human health and risks resulting from the mobility of contaminants. Once ecological (or other) risks have been demonstrated by these techniques, the next step would be to perform a comparison between risks in the different water system compartments, taking into consideration for example differences between bioavailability and differences in sensitivity of aquatic and terrestrial species (or other exposed objects). The following steps in the DSS involve the definition of a preferred alternative for remedial action, the effectivity and acceptability of which is then evaluated from different points of view (Fig. 3). After an eventual optimization of the plan for remedial action to obtain a higher efficiency, finally the government can be advised as to what decision should be made. Such a framework could be applied to sites with complicated patterns of pollution that cover not only aquatic sediments but also bordering shores and land. Fig. 3. Decision support system for remediation of sediment in polluted waters and bordering shores or floodplains. Summary The bioavailability of contaminants in sediment was compared for the following biotopes: sublittoral waters, littoral/intertidal zones and floodplains. Bioavailability was measured in bioaccumulation tests with aquatic oligochaetes and presented as biota-sediment accumulation factors (BSAFs). The bioavailability of several heavy metals, PCBs, organochlorine pesticides and PAHs in the more sandy sediments from the intertidal areas was higher than in the sublittoral river sediments. The BSAFs were more than ten times higher in newly formed sediment from the river 202 Meuse, that had been deposited on a floodplain and subsequently stood clear of water. On the basis of these results it is concluded that the risks of sediment contamination can be considerably higher in intertidal areas and river banks or floodplains than in sublittoral zones. A decision support system is presented that provides a step to step procedure for setting priorities for remedial action in sites where both aquatic and terrestrial ecosystems may be at risk. Acknowledgements This work was in part financed by Department of Zuid-Holland of the Ministry of Transport, Public Works and Water Management. Bioassays were carried out by AquaSense Laboratory (Amsterdam) and chemical analyses were performed by the Institute for Environmental Studies (Amsterdam), OMEGAM (Amsterdam) and Alcontrol (Raamsdonkveer). The author wishes to acknowledge dr. J.P.M. Vink, ir. G.M. Boks and drs. N.M. 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