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Original Article
INVERTEBRATE COMMUNITY RESPONSES TO A PARTICULATE- AND DISSOLVED-COPPER EXPOSURE IN MODEL FRESHWATER ECOSYSTEMS
STEPHANIE GARDHAM, ANTHONY A. CHARITON, and GRANT C. HOSE
Environ Toxicol Chem., Accepted Article • DOI: 10.1002/etc.2728
Accepted Article
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Original Article
Environmental Toxicology and Chemistry DOI 10.1002/etc.2728
INVERTEBRATE COMMUNITY RESPONSES TO A PARTICULATE- AND DISSOLVED-COPPER EXPOSURE IN MODEL FRESHWATER ECOSYSTEMS
Running title: Effects of copper on freshwater invertebrate communities
STEPHANIE GARDHAM,*†‡ ANTHONY A. CHARITON,‡ and GRANT C. HOSE§
† Department of Environment and Geography, Macquarie University, New South Wales, Australia
‡ Centre for Environmental Contaminants Research, CSIRO Oceans and Atmosphere Flagship, Lucas Heights, New South Wales, Australia § Department of Biological Sciences, Macquarie University, New South Wales, Australia
*Address correspondence to
[email protected]
Additional Supporting Information may be found in the online version of this article.
© 2014 SETAC Submitted 5 June 2014; Returned for Revision 18 August 2014; Accepted 18 August 2014
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Abstract: Historical contamination has left a legacy of high copper concentrations in the sediments of freshwater ecosystems worldwide. Previous mesocosm studies have focused on dissolved-copper exposures in the overlying waters which, due to altered exposure pathways, may not accurately predict the effects of copper exposure on invertebrate communities at historically contaminated sites. The present study assessed the effects of copper on the establishment of invertebrate communities within a large outdoor pond mesocosm facility containing environmentally relevant copper-spiked sediments. High particulate copper concentrations (>400 mg/kg dry wt) caused a pronounced effect on the benthic community richness, abundance and structure in the mesocosms, but particulate copper concentrations below 100 mg/kg dry weight had no effect. Further, there were no effects of copper on the invertebrate communities within the water column, even in the highest copper treatment. The response of the benthic community to copper was influenced by inter-specific interactions, the stage of ecological succession and inter-species variation in sensitivity to copper. The present study demonstrates the importance of using environmentally realistic exposure scenarios which provide both particulate and dissolved exposure pathways. It also emphasizes that risk assessments for aquatic ecosystems should consider the influence of inter-specific interactions and inter-species variation in driving the biotic response to contamination. Keywords: Metal; Mesocosm; Inter-specific interactions; Inter-species variation; Benthic macroinvertebrates; Ecological risk assessment
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INTRODUCTION Within the last 100 years, there has been a sharp increase in the discharges of metals,
including copper, into aquatic ecosystems [1]. However, with better management of contamination sources, concentrations of dissolved metals are now generally low in surface waters worldwide [2]. In contrast, because sediments are a sink for metals and as the metals cannot be degraded, a legacy of historical contamination remains in the sediments of aquatic ecosystems [3–5]. Concentrations of copper in sediments are often orders of magnitude greater than those in the overlying waters [6]. A review of sediments at freshwater field sites worldwide found a maximum copper concentration of 740 mg/kg dry weight [7,8] (this does not include mine-impacted sites, where copper partitioning is often very different to other historically contaminated sites [9]). In the field, complex and unknown stressors may exert synergistic or antagonistic effects
on biota making it difficult to elucidate the impact of a specific toxicant [10]. Mesocosms and microcosms avoid many such issues by permitting some control over the experimental conditions and allowing direct tests of causality under realistic environmental conditions. Consequently, they provide a useful tool to assess direct and indirect effects of a contaminant on diverse biological communities [11,12]. From the community-level perspective, an increase in copper concentrations generally
leads to shifts in the composition of invertebrate communities and a decline in biomass and total abundance [13–17]. The uptake of metals by benthic organisms is strongly influenced by the partitioning of the metal in the sediment [18]. Organisms can take up metals, like copper, through exposure to the dissolved phase (for example, facilitated diffusion through permeable surfaces via filter feeding) or the particulate phase (for example, active uptake by deposit
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feeding) [19,20]. Thus, the dominant exposure route for an individual organism depends on its feeding behavior [19]. In manipulative studies, it is important to establish environmentally relevant partitioning
of copper between dissolved and particulate phases in order to create exposure pathways for organisms that are similar to those present in field-contaminated sites. However, the handful of lentic mesocosm studies that have considered the effects of copper on invertebrate communities have not demonstrated copper partitioning representative of historically contaminated sites [14– 17,21,22]. These studies exposed invertebrate communities to copper dissolved in the overlying waters at concentrations between 2–420 µg/L. In the context of historically contaminated sites (excluding mine-impacted sites), overlying water copper concentrations at the mid-upper end of this range are unrealistically high. Concentrations at the lower end of the range represent those likely to occur in overlying waters present at historically contaminated sites, so may be useful for determining effects on invertebrates which are predominantly exposed to copper dissolved in the overlying water. However, in the studies with low overlying water concentrations, sediment and porewater concentrations are likely to have been unrealistically low (although they were generally not measured). For example, Shaw and Manning [14] exposed their mesocosm systems to overlying water copper concentrations between 20–260 µg/L and recorded sediment concentrations of between 4.7–9.1 mg/kg wet weight at the end of their study. Thus, invertebrates that are predominantly exposed to copper via the particulate phase in sediments would not have been exposed to the concentrations of copper that they would be exposed to in field-contaminated sites with similar overlying water copper concentrations. Here, the effects of copper on the establishment of invertebrate communities within a
series of copper-contaminated lentic mesocosms is described. The mesocosms were manipulated
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to produce as realistic copper partitioning as possible [8]. The hypothesis was that, with increasing copper concentrations, the structure of invertebrate communities in the water column and benthos would be altered and a decline in abundance, richness and diversity of invertebrates would be observed relative to the controls. METHODS Study design Twenty pond mesocosms were established on the Macquarie University campus
(33.76946 S, 151.11496 E), in northern Sydney, NSW, Australia. The infrastructure and experimental design of the mesocosms are described in detail by Gardham et al. [8], and briefly below.
Each mesocosm had a volume of 1500 L, was sunk into the ground, shaded by 70% shade
cloth, aerated, and inputs from rainwater were sufficient to maintain water depth. Sediments were spiked in situ with copper and allowed to equilibrate for 2 months, to create environmentally relevant copper partitioning [8]. The mesocosms were opened to allow biotic colonization on the 1st November 2010 (0 d). The mesocosms were allowed to colonize naturally
via mobile (flying or crawling) biota, and some biota colonized the mesocosms following the introduction of a plant specimen of Vallisneria spiralis to each mesocosm after the first 6 months. The experimental design consisted of a control (C) and 4 sediment copper concentrations
(very low, VL; low, L; high, H; and very high, VH), each with 4 replicates (Table 1). The invertebrate communities and water quality parameters (see below) were monitored every month in the first 6 months (first spring/summer season; at 31, 72, 100, 135 and 161 d) and then at 12 (365 d), 13 (407 d) and 16 months (497 d) (second spring/summer season). The final sample was taken on 12th March 2012 (497 d). The concentrations of copper in the overlying waters,
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porewaters and sediments are described and discussed in detail by Gardham et al. [8]. The change in partitioning over time between treatments is also discussed within the results section of the present study, where additional statistical analyses are included. Copper concentrations and water quality Copper concentrations in the sediments, porewaters and overlying waters were measured
as per Gardham et al. [8]. Conductivity, pH, turbidity, dissolved oxygen and the redox potential (versus the standard hydrogen electrode) of the overlying water were recorded using a multimeter (HI 9828; Hanna instruments). Invertebrate community sampling For the benthos, 3 cores (6 cm diameter) were collected from the surficial layer of the
sediment (top 1–2 cm) and were mixed together. Then, 100 mL of that pooled sediment was retained and preserved in 100% ethanol for later processing. Samples were processed by a flotation procedure adapted from that described by Anderson [23]. In this procedure, each sample was drained of ethanol through a 106 µm mesh sieve then placed in a container and flooded with saturated glucose solution. The sample was stirred to assist organisms to float to the surface and then the glucose solution was decanted through the 106 µm mesh sieve to collect the floating organisms. Each sample was processed twice and soaked in water for 5 minutes between processing to remove any residue of the glucose solution. The animals collected from both extractions were pooled and identified. Water column invertebrates were collected using a plastic tube (5 cm diameter) that was
inserted vertically from the water surface to the sediment at a random location within the mesocosm. A stopper was placed at the bottom of the tube, then the tube was removed from the mesocosm and the water collected was retained. This process was repeated 4 times and the
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samples pooled. The pooled sample was mixed, then a sub-sample (~1 L) was filtered through a 106 µm sieve and the animals retained were preserved in 70% ethanol. At 16 months, standardized sweep net samples were also collected using a kick net
(0.12 m2 opening) with 500 µm mesh. The nets were swept across the surface of the sediment
and macrophytes in one-half of each mesocosm. This process was repeated 3 times. Large macroinvertebrates (i.e., those obvious to the eye without magnification) were live picked by 2 people until all visible animals were removed. Data analysis Water quality data (including copper concentrations) were analyzed to identify key
differences in water quality among treatments over time. A repeated measures (RM) ANOVA was performed on the data of each parameter (between-subjects factor: treatment, within-subjects factor: time) followed by least significant difference (LSD) pair-wise multiple comparison tests to assess the differences between individual treatments and time points using PASW Statistics [24]. Where a significant interaction between treatment and time was found, a one-way ANOVA followed by LSD pair-wise multiple comparison was performed on the data from within each time point to identify where a significant difference among treatments occurred, and a separate RM ANOVA was performed for the data of each treatment to identify changes in the parameter within each treatment over time. Data were transformed to meet assumptions of homogeneity of variance and normality as described in Supplemental Data (SD) Table S1. Where transformed data failed Levene’s test, the significance level () was 0.01, for all other comparisons was
0.05 [25]. If the assumption of sphericity (assessed by Mauchly’s test) was not met during the RM ANOVA analyses, the Greenhouse-Geisser correction was used.
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The effects of treatment and time on the invertebrate assemblages identified in the
benthic, water column and sweep net sample data were analyzed separately. For each data set, community indices (total abundance, taxonomic richness and diversity [H’]) were each analyzed using a RM ANOVA with post hoc tests (as described for the water quality data). The benthic and water column biota data sets were then analyzed using principal response curves (PRC), a multivariate method of analysis [26]. Recommended software settings in Canoco [27] were used on appropriately transformed data (SD Table S1). The PRC identifies differences in community composition between treatments and controls over time. The accompanying species weights indicate the effect of the treatment regime at the species level. Taxa with strong positive species
weights respond to the treatment regime similarly to the community response shown in the PRC. Taxa with strong negative weights respond to the treatment regime in the opposite pattern to the PRC. Those taxa with weights close to zero either do not respond to the treatment regime, or respond in a way unlike that shown in the PRC. The significance of the PRC was tested using Monte Carlo permutation tests (using 0.05). To further interrogate the pattern observed in the
PRC, the variation in community composition between treatments at each time point was analyzed using Permutational Multivariate Analysis of Variance (PERMANOVA) [28] followed
by post hoc pair-wise comparisons. The PERMANOVA analyses were based on Bray-Curtis similarities of transformed data (SD Table S1) and performed in Primer 6+ [29]. To understand the responses of individual taxa to copper treatment, a RM ANOVA was
performed (as described for the water quality data) on appropriately transformed (SD Table S1) abundance data of sensitive taxa, identified by the PRC, within each invertebrate data set. Where present, even if not identified in the PRC analysis, the response of Physa acuta (Gastropoda: Physidae) to copper treatment was characterized because field observations indicated that the
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gastropod was abundant in the C/VL/L treatments and rare in the H/VH treatment. For the sweep net data set, a one-way ANOVA (as there was only 1 time point) with LSD post hoc analysis was performed on the transformed (SD Table S1) abundance data of P. acuta. RESULTS Copper concentrations and water quality Concentrations of copper in the sediments, porewaters and overlying waters generally
increased from the C < VL < L < H < VH treatments (SD Figure S1). There was a significant interaction between treatment and time on copper concentrations in the sediments (F17,64 = 2.6,
p = 0.003), porewaters (F14,53 = 3.2, p < 0.001) and overlying waters (F13,48 = 4.2, p < 0.001) (SD
Table S2). In general, there was a significant difference between the copper concentrations in the
sediments among treatments, however, the VL and L treatments were only significantly different (p < 0.05) on 4 occasions (SD Figure S1A). In addition, at 161 d, there was no significant difference between sediment copper concentrations in the H and VH treatments (p = 0.12). Within the C, L, H and VH treatments, copper concentrations changed significantly over time (p < 0.05), however, post hoc analysis generally showed no difference between individual time points.
Porewater copper concentrations were similar among C, VL and L treatments and they
were usually significantly higher in the H and VH treatments (SD Figure S1B) (SD Table S2). Within the C, L, H and VH treatments, porewater copper concentrations changed significantly over time (p < 0.05); initially they decreased (up to 42 d), but then a plateau was reached (SD Figure S1B). In the overlying waters, the concentrations of copper in the C, VL and L treatments were
similar, although they were often significantly lower in the C than the VL treatment (SD Figure
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S1C, SD Table S2). The H and VH treatments were also generally significantly different from each other (70% of the time p < 0.01). Within each treatment, overlying water copper concentrations changed significantly over time but there was no underlying directional trend (SD Figure S1C). There was an interaction between treatment and time on pH (F16,61 = 4.6, p = 20 µg/L, respectively. This demonstrates that, in agreement with laboratory studies [18,31], the major route of exposure to biota in the benthos is via the porewaters and sediments. In terms of risk assessment, this is extremely important, as based on previous studies [14,17], it may be assumed from overlying water copper concentrations that there would be no effect of copper on the benthic community, when in fact concentrations in the porewaters and sediments could be causing an effect. As the copper partitioning in the present study is representative of historically
contaminated sites [8], it is concluded that chronic copper exposure may affect the invertebrates present in the benthos at such sites. However, because effects on the water column invertebrate
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community were not evident, the invertebrates in the water column at historically contaminated sites may not be affected. This lack of response in the water column invertebrate community is in line with previous research. Two studies considered the effects of copper on water column invertebrate communities during chronic (5- and 32-week) exposures with overlying water copper concentrations in the same range as those used here: Winner et al. [16] observed changes in the zooplankton community at concentrations of 20 µg Cu/L in the overlying waters, while Hedtke [17] observed changes occurring between 9.3 and 30 µg Cu/L. Therefore, in the present study, the concentrations of copper in the highest treatment (VH: average of 11± 0.9 µg/L, maximum of 15 ± 0.5 µg/L at 135 d), were between the ‘no-effect’ and ‘effect’ concentrations reported previously [16,17]. In addition to considering copper partitioning during exposure experiments, this study
also emphasizes the importance of carrying out truly chronic exposures to understand the longterm effects of a persistent contaminant in the field. The strong response of the benthic community to copper in the present study, carried out over 71 weeks, is similar to other longterm studies. Hedtke [17] also recorded severe effects of copper on snails (Viviparus, Physa sp.
and Helisoma campanulata) during a 32-week overlying water copper exposure and Shaw and Manning [14] observed a pronounced response of the benthic community to copper during 19week exposures. In contrast, during their 5-week lentic exposures, Winner et al. [16] observed a weak response of the benthic community to copper; the densities of small chironomids responded differently depending on season and the only consistent response in the benthic community to copper was a reduction in the densities of small caenid mayflies at 40 µg/L (compared to 0 and 20 µg/L treatments). Specifically Winner et al. [16] also noted that snails were unaffected by copper. Even based on overlying water exposures, Winner et al. [16] should
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have seen a response (as noted earlier, Hedtke [17] and Shaw and Manning [14] observed strong effects on the benthic community at overlying water copper concentrations of ≥30 µg/L (not
≤9.3 µg/L) and ≥20 µg/L respectively). Winner et al. [16] supposed that the weak response that
they observed in the benthos to copper could have occurred because the 5-week experiments did not incorporate effects on reproduction and recruitment, a theory supported by the present study and the long-exposure studies previously reported [14,17]. The importance of inter-species variation in sensitivity Within the benthic response, some taxa responded differently compared to previous
studies. For example, Tanypodinae were negatively affected by copper [14], but in the present study, no effect of copper was identified. Other studies that have considered the effects of copper on Tanypodinae in the field have also found species within the sub-family to be copper tolerant, which supports the lack of response in the present study [32,33]. There was also inter-species variation in copper susceptibility of Ostracoda within the present study, as 1 species of benthic Ostracoda was not affected by copper but another declined in abundance. In contrast, Shaw and Manning [14] showed an increase in benthic Ostracoda. This demonstrates that organisms with similar life histories may respond differently to contaminants and confirms that it is important to identify organisms to an appropriate taxonomic level to allow contaminant specific responses to be identified. The importance of inter-specific interactions Inter-specific interactions (interactions between individuals of different species) within
the mesocosms influenced the response of the benthic community to copper. In the first spring/summer season (up to 161 d), the abundances of Chironominae, the Ostracoda species and Cladocera were lower in the H/VH treatments compared to the control. However, in the second
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spring/summer season (from 365 d), the abundances of each taxa generally declined in the C/VL/L treatments to become more similar to the H/VH treatments. It is possible that the decrease in abundance of these taxa was due to inter-specific competition in the low copper treatments between these taxa and the gastropod P. acuta (which colonized the mesocosms and proliferated in those treatments between 161 and 365 d). Inter-specific competition between these taxa has been documented elsewhere [34–37]. For Ostracoda and Cladocera it may be due to competition for food, but for Chironominae, there is evidence to suggest that reproductive
interference may occur. Devereaux and Mokany [35] showed that Chironomus oppositus avoided P. acuta during site selection for oviposition. Clearly, the inter-specific interactions between taxa were important in driving the biotic response to copper, demonstrating that understanding these interactions is necessary when characterizing the potential effects a contaminant has, or will have, on a community. Stage of ecological succession The present study indicates that the effect of copper on taxonomic richness is most
prominent when benthic communities are establishing in new, or recently disturbed, freshwater habitats. An effect of higher taxonomic richness in C/VL/L treatments compared to H/VH treatments was only observed during the establishment of the benthic communities within the mesocosms. Consistent with this, Shaw and Manning [14] observed no effect on taxonomic richness on already established benthic communities. CONCLUSIONS There was a strong difference in the benthic communities that developed within the
highly copper-contaminated sediments compared to those in the controls. However, there was little effect of copper on the invertebrates in the water column. The present study has
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demonstrated the importance of using environmentally realistic exposure scenarios, including both realistic partitioning of contaminants between phases in the environment and environmentally relevant exposure times, the latter being particularly important with regard to persistent contaminants in the field. Further, there was evidence to suggest that communities may be particularly sensitive during colonization, with implications for environmental management when sites are disturbed. The present study has demonstrated the importance of considering inter-specific interactions and inter-species variation in sensitivity to contaminants in driving the community response. Ultimately, the present study has shown that different communities will respond to
contaminants in different ways through both direct and indirect effects. As such, although guidelines are useful, site-specific assessments on the health of aquatic ecosystems are paramount in understanding the effects of contaminants at those sites. SUPPLEMENTAL DATA Tables S1. Figures S1–S2. (166 KB PDF). Acknowledgment—Construction of the mesocosms was funded by a Macquarie University Infrastructure Grant. S Gardham was supported by a Macquarie University Research Excellence Scholarship. The authors thank A Michie, M Nagel and L Oulton, for their help in running the experiment, and two anonymous reviewers for their useful comments. We also thank the CSIRO’s Water for a Healthy Country flagship (partfunding and specialist time). The authors have no conflict of interest to declare.
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Figure 1. Benthic community (A) total abundance, (B) taxonomic richness and (C) diversity over time. Key: Control (white), Very low (dashed horizontal), Low (diagonal cross-hatching), High (dashed vertical) and Very high (black). The main results of the one-way ANOVAs performed at each time point on the taxa richness and diversity data (due to a significant interaction between time and treatment for those data) are displayed above each respective time point: ns = not significant, * = p < 0.05. Letters above each bar indicate the outcomes of LSD post hoc analyses, where treatments that do not share a letter within a time point are significantly different.
Figure 2. Principal response curves with species weights showing the effect of copper on establishment of the benthic invertebrate community. Key: Control (), Very low (), Low
(), High (), and Very high (); for species weights, red tick marks on the species weights axis indicate other taxa with calculated species weights (labels have been removed for clarity).
Figure 3. Population responses of individual benthic taxa over time: (A) Chironominae; (B) Ostracoda Sp. 2; (C) Cladocera; (D) P. acuta (which appeared in the mesocosms during the second spring/summer season). Key: Control (), Very low (), Low (), High () and Very high () treatments. Values are mean ± S.E.M., (n = 4). The main results of the one-way ANOVAs performed at each time point on the abundances of Chironominae, Ostracoda Sp. 2
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and Cladocera (due to a significant interaction between time and treatment for those data) are displayed above each respective time point: ns = not significant, * = p
