A number of studies of pollutants and their effects on the fish fauna of Lake Molnbyggen outside Leksand, Sweden have been focused on leachate from the municipal landfill (Lindbodarna), which is located above the lake. The aim of the present study was to test the toxicity of sediment from Molnbyggen and some adjacent lakes, in order to see if this could explain the causes of the effects on the fish. The sampling sites were selected in co-operation with the project group at the Swedish Environmental Protection Agency to make the results more easily comparable with those from other studies on the fish fauna in the area. Both acute and chronic tests were made with the crustacean Ceriodaphnia dubia, and the sediment samples were equilibrated with standard reference water prior to exposure. The exposures were made under standard conditions in the laboratory to make them more readily comparable between the sampling locations and with previous studies. The effects on survival and reproduction were monitored during 8 days. After that a series of Toxicity Identification Evaluation manipulations of the tested waters were made in order to identify the cause(s) of the toxicity. The results from the Toxicity Identification Evaluation tests pointed towards heavy metals as the cause of toxicity, and the analytical results of heavy metals in the water phase showed that the concentrations of six metals (Cd, Co, Mn, Ni, Pb and Zn) were significantly correlated with toxicity. The concentrations of these metals were also correlated with each other making it hard to separate cause and effect among the metals. However, the concentrations of Cd, Ni, Pb and Zn were below the expected effect concentrations, but the measured concentrations of Co and Mn were high enough to be probable causes of the observed experimental toxicity. There was no indication that lipophilic compounds should have caused these effects. Therefore, the result of this study was rather surprising, showing that heavy metals like cobalt and/or manganese which are not generally considered as environmentally problematic may be of environmental concern. Potential effects of cobalt and manganese could be mediated through the olfactory system, because both these metals have been observed to affect this system in fish. However, the link between such effects and those observed on reproduction in Lake Molnbyggen is vague.
Landfill leakage has been suspected to cause reproductive disturbances in fish in some Swedish lakes, and especially in Lake Molnbyggen (Noaksson et al., 2000, 2001). The problem has been discussed at the Swedish Environmental Protection Agency (Naturvårdsverket, 1999, 2001). The leakage from the landfill Lindbodarna has been analysed for a large number of known pollutants and it has also been tested for acute and chronic toxicity to water fleas, Ceriodaphnia dubia (Dave and Nilsson, 2000a). One of the major toxicants in landfill leakage water is ammonia. The leakage from Lindbodarna has, therefore, been tested before and after degradation in order to investigate the possible contribution of ammonia to the observed toxicity (Dave and Nilsson, 2000b).
In spite of the physiological disturbances (decreased activity of brain aromatase, and decreased concentrations of testosterone and 17β-estradiol in blood plasma) of perch (Perca fluviatilis) found by Noaksson et al. (2000, 2001), no ecological effects (abundance, age class distribution, or species diversity) of the fish fauna were detected during standardised gill net sampling in August 1999 (Appelberg et al., 1999) and no toxicity (acute or chronic) to C. dubia was detected in the brook from the landfill or in the Lake Molnbyggen in December 1999 (Dave and Nilsson, 2000a). Further studies on the fish health in Lake Molnbyggen and adjacent lakes are in progress by Noaksson and co-workers (Noaksson, 2003) in order to investigate possible cause and effect relationships.
The aim of the present study was to investigate if the sediments in Lake Molnbyggen and adjacent lakes were toxic and if the cause of possible toxicity could be identified. A sediment bioassay with Ceriodaphnia dubia was used to measure acute and chronic toxicity of water equilibrated with sediment from the lakes and the same water was used for toxicity identification evaluation using standardised procedures.
Materials and methods
In addition to L. Molnbyggen, three adjacent lakes (L. Yxen, L. Djursjön and L. Styrsjön) and the major lake in the area (L. Siljan) were included in this study (Figure 1). The sites were selected after consultation by the steering group appointed by the Swedish EPA (Marklund, 2000). Methods for sampling and treatment of samples are described as well as methods for toxicity testing with sediments, methods for toxicity identification evaluations (TIE), chemical analysis and statistical treatment.
Sampling and treatment of samples
All sediment samples were taken 14–15 February 2001 when the lakes were covered with ice. The local environmental authority and its staff assisted in selection of sampling sites and the actual sampling of the sediments. Two sampling sites were used in L. Molnbyggen and L. Siljan and only one in each of the other three lakes (L. Yxen, L. Djursjön and L. Styrsjön). At each sampling site 4 holes were drilled in the ice within a radius of 5 m. From each hole a sediment sample was taken by a core sampler. The top 2 cm of each sample was transferred to a plastic sampling vessel. The samples were stored at about 4°C during transport, storage prior to testing. One of the authors (E. Nilsson) was responsible and present during the entire procedure as well at the testing of the samples. The sampling sites are described in Table 1.
Sediment bioassay procedures
The sediments were tested for chronic toxicity with Ceriodaphnia dubia using elutriates made with reconstituted test waters with a hardness of 250 mg l−1 and pH of 8.0 (SIS, 1996). From each sediment sample (7 lake sites and 4 replicate field samples from each) a sediment sample was taken and mixed with reconstituted water to make a 32% (wet wt) sediment-water mixture in a glass beaker. The sediment/water mixture was mixed for 30 s with a glass rod two times with 2 h interruption. Then the mixture was allowed to settle for 1 d. The overlaying water (supernatant without centrifugation) was used for the reproduction test directly. Each water phase was tested with 3 replicates containing 1 newborn (<24 h old) Ceriodaphnia dubia and 10 ml test water. All exposures were made in test plates (NUNC multidish) at a temperature of 25°C and a light period of 16 h light and 8 h darkness. Test solutions were renewed daily, and survival and reproduction was recorded daily as well. The negative controls contained only dilution water, and the sensitivity of the test organisms was tested for acute toxicity of the reference toxicant potassium dichromate.
The method used for determination of the sensitivity of C. dubia to toxicants in water is based on the method developed by Mount and Norberg-King (1984). The method has thereafter been slightly modified by Unger and Ek (1994). The method provides for determination of survival and reproduction of the test organism under defined conditions of test vessel (NUNC multidish with 10 ml test volume), age of test organisms (3–24 h at start), reconstituted dilution water matrix (SIS, 1996) with a hardness of 250 mg l−1 as CaCO3, test solution renewal (daily), feeding (daily with 20 μYTC + 7.6 × 105 cells of the green algae Rapidocellus (formerly Selenastrum capricornutum), ambient temperature (25°C and a light period of 16 h light and 8 h darkness via day-light tubes at 400–700 lux). All sediment samples were tested in parallel and the results were evaluated for survival and reproduction (total number of live neonates per female after 8 d).
Some of the test procedures used for TIE, which were developed by researchers at the United States Environmental Agency (USEPA; US EPA, 1991), were used to identify causes of observed toxicity in this study. These procedures were originally developed for wastewaters but they have also been used for sediments, usually with pore waters (interstitial waters) or elutriates (leakage waters).
In this study the same sediments used for the reproduction test were equilibrated once more with reconstituted water (at 32% wet weight) and allowed to settle for 48 h after mixing twice as before. The tested waters are thus called elutriates. They were tested as described in Table 2 with respect to acute toxicity (immobility) for C. dubia after 24 and 48 h exposure.
The columns (C18, QMA and CM) were conditioned by 15 ml reconstituted water. Before that the C18 column had been preconditioned with 15 ml methanol (Pro Analysi) and 15 ml deionised water. The Millex column was used directly. One column was used for each lake site and all four site replicates (elutriates) were passed in consecutive order (1–4) through the same columns using the same syringe. The addition of EDTA (10 μl 0.01 M EDTA per 10 ml test solution) was made directly into the well on the test plate at a temperature of 25°C and a light period of 16 h light and 8 h darkness. The negative controls contained only dilution water and the sensitivity of the test organisms was tested for acute toxicity of the reference toxicant potassium dichromate.
At the sampling event, water pH was measured on site, and during the reproduction tests pH and dissolved oxygen was measured in the old test solutions on days 1, 5 and 8.
The water phases after equilibration and toxicity testing (reproduction and TIE) were sent to an commercial laboratory (SGAB Analytica, Luleå, Sweden) for direct determination of the following elements with detection limits in μg l−1 given in parentheses: Al (1), Ba (0,2), Ca (200), Cd (0,05), Co (0,05), Cr (0,5), Cu (1), Fe (0,4), Hg (0,2), K (500), Mg (90), Mn (90), Na (120), Ni (0,5), Pb (0,2), S (160) and Zn (1). The analytical laboratory is accredited for the method (V-3a, Analysis without digestion) used for environmental water samples. All concentrations are given as total concentrations.
Differences between lake sites (N = 8) including controls and sediment sample (N = 5 including controls) were analysed using analysis of variance (ANOVA) and post hoc tests for establishment of differences between the lake sites (Bonferroni, Newman-Keuls, REGW, Scheffe, Sidak, T-test and Tuckey). These tests and Spearman rank correlation tests were made with a soft ware program (Crunch Ver. 4, Crunch Corp., California, USA).
The results are presented in the same order as they were generated (reproduction tests, TIE tests, chemical analysis of metals and correlations between measured effects and concentrations of metals.
Effects on survival and reproduction
The sensitivity of the test animals, determined as 24-h EC50for potassium dichromate was 1.9 mg l−1 for the animals used for the reproduction tests, and control survival (97%) and reproduction in the controls (23 ± 7 young per female). This means that the reproduction tests were valid. The results from the ANOVAs are given in Table 3 (survival) and Table 4 (reproduction). Mean values, standard deviations and significant differences are given in Table 5.
The ANOVAs, using Lake (site), sediment sample (site replicate) and elutriate replicate (taken in the laboratory) as sources of variance, showed that the Lakes (sites) were the major cause of variance with Sediment sample as a minor but significant cause of variance. The variance of the elutriate replicates was not significant, but there was a (minor) combined significant variance of Lake and Sediment sample for both survival and reproduction (Tables 3 and 4).
The results from the post hoc significance tests for differences between the lake sites are summarised in Table 5 together with the mean values and standard deviations.
The results in Table 5 show that there were large differences in survival and reproduction between the lake sites. L. Yxen, and L. Styrsjön had 100% survival and unaffected reproduction, but both L. Djursjön and one of the L. Siljan sites (Isunda) had no survival and no reproduction at all. This is remarkable because the latter two were selected to be reference sites for effects on fish health. The L. Molnbyggen and the other L. Siljan site (Fornby) had intermediate survival and reproduction. Significant effects on both reproduction and survival were also found in sediments from the deep-water site in L. Molnbyggen and in sediments from the Fornby site in L. Siljan. Only reproduction, but not survival, was affected at the inflow site in L. Molnbyggen.
The 24-h EC50 for the reference toxicant, potassium dichromate, was 1.7 mg l−1 for the C. dubia used in the TIE tests, and the immobility in the controls with dilution water was 0% for all columns after 24 h and 0% for all treatments except the CM and QMA columns, for which it was 5 ± 10% (mean ± SD; N = 4). Thus, the results from the TIE tests are considered to be valid. The results for all sediment elutriates and TIE treatments are shown in Table 6.
The TIE tests for L. Yxen and L. Styrsjön confirmed that the sediments from these two lakes were non-toxic. They also confirmed that the TIE manipulations were not affecting the test organisms. The results from the other lakes showed that several of them were affected positively (reduced immobility) by addition of EDTA (L. Molnbyggen inflow and deep station, Fornby station in L Siljan and L. Djursjön). Samples from the L. Molnbyggen inflow and deep station and L. Djursjön were also affected positively by the CM-column. For the Isunda station in L. Siljan, only the CM-column had effect, but it was very efficient in reducing the toxicity at this station.
The overall impression from the TIE-manipulations was that the toxicity was caused by heavy metals, because only the EDTA-addition and/or the CM-column reduced the toxicity. Ammonia and/or hydrogen sulphide was ruled out as a potential toxicant because all sediments were surficial (0–2 cm) and the lakes are oligotrophic. Therefore, only metals were considered as potential toxicants in these sediments, and the forthcoming chemical analyses were directed towards metals, and especially heavy metals. The chemical analyses were made on the same water phases that were used in the TIE-tests, in order to make statistical correlations between toxicity and exposure concentrations of potential toxicants possible.
Concentrations of heavy metals and correlations with toxicity
Metals were measured in the water phase after equilibration with sediment and reproduction and TIE tests from all replicates (7 lakes and 4 replicates). Chemical replicates were sent from 4 samples to evaluate the chemical analytical precision. These showed that the analytical precision was good. The results for all heavy metals (Cd, Co, Cr, Cu, Fe, Hg, Mn, Ni, Pb and Hg) and some other elements that were analysed simultaneously (Al, As, Ca, K, Mg, Na and S) are shown in Table 7.
An ANOVA for the metal concentrations (not presented) showed that there were significant differences between the lake sites investigated for all the analysed elements except for Cu and Hg. These two heavy metals have, therefore, been excluded as potential causes of the observed differences in toxicity. The measured concentrations of Cu, Zn, Cd, Pb, Cr, Ni and As in these elutriates have been compared with those in Swedish natural lake waters (Naturvårdsverket, 2000). Among these metals the concentration of Cd was high at the deep site in L. Molnbyggen and at the Isunda site in L. Siljan. The concentration of Zn was also high at the Isunda site. For the rest of these metals the concentrations were at most moderately high according to the Swedish EPA classification for metals in surface waters (Naturvårdsverket, 2000). One of the metals showing the highest variability was cobalt (Co), which has a background value of only 0.03 μg l−1 in Swedish lakes. The concentration of Co in the sediment elutriates was 2.5 μg l−1 in the L. Djursjön sample and 18.7 μg l−1 in the Isunda sample. The most variable metal was manganese (Mn), for which the average measured concentrations were 32.7, 13.9, 2.2 and 0.9 mg l−1 in the elutriates from L. Djursjön, Isunda in L. Siljan, the deep site in L. Molnbyggen and the Fornby site in L. Siljan, respectively, that is, those with the highest measured toxicity to C. dubia in this study. The only exception from this relationship was the inflow site in L. Molnbyggen, which had the lowest concentration of Mn among all sites in this study.
The measured concentrations of the metals were tested for correlation with toxicity of all sediment elutriates (Spearman 2-tailed rank correlation test, N = 28). Those with significant (p ≤ 0.05) correlations were Al, As, Cd, Co, Fe, Hg, Mn, Ni, Pb and Zn. These correlations are shown in Table 8.
In order to explain toxicity the correlation coefficient, R, must be positive between metal concentration and immobility in TIE tests (Direct 24 h and Direct 48 h), but negative between metal concentration and reproduction and survival. This was the case for Cd, Co, Fe, Mn, Ni and Pb for both the direct TIE tests (24- and 48-h immobility) and reproduction and survival. For Zn some of the TIE test effects but not the reproduction test effects were correlated, and for Hg only the reproduction effects were correlated. For As the significant correlations were contradictory with anticipated toxic cause and effect relationships. The significant correlations cannot be taken as proof for cause and effect relationships, because the concentrations of Cd, Co, Fe, Mn, Ni and Pb were also correlated. Therefore, it is not possible to tell which among these metals that was/were the most important toxicant(s).
A calculation of the decrease in immobility after both 24 and 48 h caused by the manipulations (non-manipulated—EDTA added, etc.) was also made. No significant correlations were found between the metal concentrations and the manipulations with EDTA, C18-column, QMA-column and Millex column. But for the CM column the reduction in immobility after 24 and 48 h was correlated with the concentrations of Cd, Co, Mn, Ni and Zn and after 48 h also for Pb (data not shown). These correlations are considered to be further proof that either one or some of these metals is/are (a) probable cause(s) of the observed toxicity.
The first ANOVAs (Tables 3 and 4) showed that the lake site was the most important cause of variability for both survival and reproduction, but also that sediment samples taken at the same site varied significantly. The significant co-variation between site and sediment sample also suggest that some sites were more variable than others (Tables 4 and 5). The water content (or dry weight) is often used as an indicator of sediment structure. Sediments from accumulation bottoms have high water contents (lower dry weight). The sediment water content (100% dry weight) is given in Table 1 together with typical values for accumulation and transport bottoms according to Håkanson (1992).
The values in Table 1 show that all sediments from the smaller lakes, including the shallow inflow site in L. Molnbyggen, were taken from accumulation bottoms, which was the intention in this study. Those from the L. Siljan sites (Fornby and Isunda) were taken from transport bottoms. L. Siljan is much bigger than the other lakes. However, no consistent pattern of correlations (Spearman 2-tailed tests) between sediment water content and effects or metal concentrations was found (results not presented).
The effects seen in the reproduction tests (survival and reproduction) were correlated (positively) with each other and (negatively) with the effects (immobility) observed in the non-manipulated TIE tests (Table 9). This indicates that there was no major change in the toxicity during the course of these experiments, and that all results obtained are comparable.
The results from the TIE manipulations showed that it was mainly the EDTA addition and the CM column filtration that reduced toxicity. There was no decrease in immobility after filtration by the C18 column, and, therefore, the toxicity from lipophilic toxicants can be excluded. This is of special interest because earlier studies on fish in Lake Molnbyggen (Noaksson et al. 2000, 2001) have indicated that the effects seen are such that can have been caused by lipophilic compounds. If lipophilic toxicants are present in these lakes they should also be present in the sediments. However, so far no significant concentrations of such organic pollutants have been found in the investigated lakes or in leakage from the landfill Lindbodarna (Öman, 1999; Naturvårdsverket, 2001).
The results from this study rather suggest that certain heavy metals can have caused effects on survival and reproduction of crustaceans and possibly also affected the fish fauna in some of the investigated lakes. The toxicity of the sediments was most severe in L. Djursjön and at Isunda in L. Siljan, but also significant in L. Molnbyggen and at Fornby in L. Siljan. In L. Molnbyggen the effects were stronger at the deepwater site than at the more shallow inflow site.
The toxic effects were correlated with the concentrations of Cd, Co, Mn, Ni, Zn and Pb, and the most remarkable findings from the metal analyses were the high concentrations of Mn and Co. The latter two metals are normally not included in environmental monitoring programs. The sites with toxic effects also had the highest concentrations of these two metals, except for the inflow site in L. Molnbyggen, which had low concentrations of all these metals. The concentrations of Cu and Hg were not significantly different between the lake sites, and the concentrations of Cu and Hg as well as those of Cd, Ni, Zn and Pb were generally not above the background levels in Swedish lake waters (Naturvårdsverket, 2000). Therefore, these metals are not potential toxicants for the effects seen in this study, and Co and Mn are the only potential toxicants remaining among those analysed in this study.
Manganese was identified as a major sediment toxicant in L. Champlain, Vermont, USA, using similar TIE procedures with C. dubia as we used in this study (Boucher and Watzin, 1999). In that study also As and Ni was present at high concentrations. The 48-h LC50 for Mn reported by Boucher and Watzin (1999) was 9.1 mg l−1, and they also reported an unpublished 48-h LC50 for C. dubia of 12.7 mg l−1. Lassier et al. (2000) demonstrated an increased sensitivity to Mn in both Hyalella azteca and C. dubia by decreasing hardness with LC50s ranging from 15.2 mg l−1 in hard water to 3.0 mg l−1 in soft water. This is consistent with Hockett and Mount (1996), who determined the 24- and 48-h LC50s for C. dubia to be 39 and 17 mg l−1, respectively in hard water, and these authors also demonstrated that the Mn2 + toxicity was reduced by 74 μM EDTA in hard water similar to our reconstituted water. The EDTA concentration in our TIE tests was only 10 μM, and this is probably insufficient for reducing the toxicity of Mn2 +, which has a much lower EDTA complex stability constant than most other heavy metals. Therefore, the effects in elutriates from L. Djursjön with 32.6 mg l−1 Mn and Isunda in L. Siljan with 14.0 mg l−1 Mn could have been caused by Mn, even if no or only a weak reduction of toxicity by EDTA addition was seen for these two lakes. However, the CM-column reduced toxicity at both sites. Unfortunately, we don't know how Co and Mn are affected by CM column filtration, and we don't know the speciation of Mn in our samples.
The Mn concentrations in L. Molnbyggen were much lower, especially at the shallow, inflow site. Addition of EDTA was more effective in reducing toxicity than CM-column filtration at the deep site but both EDTA addition and CM-column filtration was most effective at the shallow, inflow site. Therefore, the conclusion from the TIE tests that the primary toxicants in all toxic elutriates were metals is valid, but the identity of the toxic metal(s) in L. Molnbyggen is not proven, because the measured concentrations seem to be to low as single causes of toxicity.
The concentrations of Co were also high at the two most toxic sites, 2.49 μg l−1 in L. Djursjön elutriate water and 18.7 μg l−1 in Isunda elutriate water. The information on Co regarding aquatic organisms is rather limited. However, the most extensive study we found (Diamond et al., 1992) included data from both acute and chronic tests with C. dubia at various levels of water hardness, as well as a literature survey, and these authors suggested tentative acute water quality criteria of 288 μg l−1 in soft water and 873 μg l−1 in hard water, and corresponding chronic water quality criteria of 44 and 15 μg l−1. Furthermore mayflies (Ephemerella ignita) were affected at 32.6 μg l−1 in a 4 wk flow-through experiment (Södergren, 1976). Therefore, the measured concentrations of Co in the elutriate waters in this study were slightly below those that would be toxic to C. dubia in chronic tests and at least one order of magnitude below those that would be toxic in acute (TIE) tests. Therefore, Mn is a more probable toxicant in the L. Djursjön and Isunda sediment elutriates than Co.
The reasons for the high concentrations of Mn and Co are not understood. The concentration of Mn was much higher at the deep site of L. Molnbyggen and L. Djursjön elutriates (which were toxic) than in L. Yxen and L. Styrsjön elutriates (which were non-toxic). The concentration of Mn was highest (825 μg l−1) of all frequently analysed metals in the landfill leakage water at Lindbodarna (Öman, 1999), and some unidentified compound(s) from the landfill Lindbodarna is the major hypothetical cause of the observed biochemical and physiological disturbances in fish from L. Molnbyggen (Noaksson et al., 2000, 2001). In this context the identification of high concentrations of Mn and Co is no major clue, because these metals may originate from other sources than the landfill, and the concentrations were not elevated at the inflow site in L. Molnbyggen, where the brook from the landfill area runs into the lake. However, after the effects on fish were observed, measures were taken towards prohibiting leakage water from reaching the lake (Naturvårdsverket, 2001). Therefore, the shallow inflow site may have responded by reduced concentrations, while the deep site may still be affected. An alternative explanation is that the Mn at the deep site originates from other sources than the landfill. Thus, the linkage of exposure of fish in the L. Molnbyggen to Mn coming from the landfill is rather weak.
The linkage of effects of Mn causing reproductive disturbances in fish is also rather weak. But there is at least one study which suggest that Mn can affect fish in the environment, especially during periods of acid conditions (Nyberg et al., 1995). Manganese is also accumulated into several tissue of fish, including the olfactory system, and this uptake may be of toxicological significance in fish (Rouleau et al., 1996). Manganese has also been associated with effects of low oxygen in marine Crustaceans (Nephrops norwegicus) on the Swedish west coast, and high concentrations have been found in the brain of this species (Eriksson and Baden, 1998; Baden and Neil, 1998). However, we are not aware of any study on the chronic toxicity of Mn to fish which has demonstrated reproductive disturbances, and we don't know of any evidence of fish avoidance of Mn in water. Therefore, this linkage between high concentrations of Mn and reproduction disturbance in fish becomes pure speculation. Furthermore, the effects on the fish community have only been detected at the sublethal, biochemical and physiological level (Noaksson et al., 2000, 2001), and no major effects at the population and fish community level have been detected in L. Molnbyggen (Appelberg et al., 1999; Dahlberg, 2001). The overall assessment of the fish community based on standardised gillnet sampling was that the status was better in L. Molnbyggen and L. Styrsjön than in L. Djursjön (Dahlberg, 2001), which is more consistent with our sediment bioassays.
The most interesting findings from this study is, therefore, that the concentrations of Mn may be higher than previously known in bottom water close to sediments. This may be a more common phenomenon than previously known, because it was found in three out of the six deep water sites in this study (1, 2, 3, 5, 6 and 7). We don't believe that the Mn toxicity in this study is an artefact due to anoxic conditions (NH3 and/or H2S), because our sediment samples were taken 0–2 cm from the sediment surface. Therefore, we suggest that Mn should be included in the monitoring of water quality, especially of bottom water, and that the chronic toxicity of Mn, including effects on reproduction, should receive more attention. The chronic IC50 of Mn for C. dubia in soft, moderately hard and hard water was determined to be 3.9, 8.5 and 11.5 mg l−1, respectively (Lassier et al., 2000), and these concentrations may occur in many deepwater lakes.
The high concentrations of Co at two of the sites (L. Djursjön and Isunda in L. Siljan) were unexpected, but the concentrations were not considered to be high enough to be of immediate toxicological concern.
The toxicity of sediments from 7 lake sites around Leksand, Sweden, was determined using the reproduction test with C. dubia, and the toxicants were identified by TIE tests and metal analyses of the sediment elutriates used for exposure of C. dubia.
Sediment elutriates from 4 of the 7 sites (L. Djursjön, L. Molnbyggen, and 2 sites in L. Siljan (Fornby and Isunda) reduced survival and reproduction. These effects were most likely caused by metals and not by lipophilic compounds. The concentrations of Cd, Co, Mn, Ni, Sn and Pb were correlated with the toxic effects on C. dubia. The concentrations of the metals were also correlated, making it difficult to determine cause-effect relationships for the different metals. The concentrations of most metals are below natural background concentrations in Swedish lake waters, but the concentrations of Co and Mn were high in the most toxic water samples, except for those in the inflow to L. Molnbyggen. The concentration of Mn was high enough to explain toxicity in all the toxic elutriates, except for that from the inflow to L. Molnbyggen.
Most of the toxicity from the studied lake sediments was explained by the high concentration of Mn, but the cause of the high Mn concentrations is presently unknown. High concentrations of Mn in water have previously been related to anoxia and acidity, but leakage water from landfills may be an additional source. This means that the toxicity, and especially the chronic toxicity of Mn to aquatic organisms, should receive more attention.
This study was supported by the Swedish EPA, and it was planned in co-operation with the project group for Lake Molnbyggen studies at the Swedish EPA. Details relating to sampling studies and design were discussed in particular with Håkan Marklund at the Swedish EPA and Olle Bergfors at Leksand municipality. The staff from Leksand municipality assisted in the sediment sampling and localisation of the sampling sites. Constructive comments made by two anonymous referees improved the paper. The authors are grateful for the support of the persons and organisations mentioned above.