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Environmental contaminants are known to have adverse effects on aquatic organisms. They impact on survival 1 and affect crucial behaviors such as the ability to effectively forage for food, avoid predation, find suitable habitat, compete for resources, and tolerate a number of biotic and abiotic interactions within the surrounding environment 2. Contaminant-induced behavioral changes can be linked to biochemical and physiological effects at the organism level, which can have flow-on effects at the population level 3. Therefore, significant changes to crucial behavioral traits of an organism (e.g., foraging behavior, social interaction, and locomotion) may pose a threat to the survival of a population 4, 5. Behavioral traits can be exploited as toxicity test endpoints and, in some cases, have been demonstrated to be sensitive to contaminant exposure 6, 7. Consequently, behavioral endpoints are increasingly being considered to have high ecological relevance as contaminant-induced changes to an organism's behavior offer insight into potentially major environmental impacts 3. Previous studies have reported contaminant-induced changes to a number of key behaviors, including predator–prey relationships 8, swimming and mobility 5, and contaminant avoidance 7, 9.
The concentration of toxicants in surface sediments will often vary spatially, and a number of organisms have demonstrated the ability to detect and actively avoid areas of contamination 10, 11. In the natural environment, individual organisms, and eventually populations, will migrate away from a contaminated site before the exposure results in the uptake of a harmful dose (sublethal or lethal). Hence, avoidance behavior is a significant factor in determining the extent of exposure 4, 6, the magnitude of the hazard, and the overall risk the sediment poses to ecosystem health. Avoidance is often a rapid and easily measurable response, with some suggesting that it is more sensitive than commonly used lethal and sublethal endpoints 6, 7. Behavioral endpoints, such as avoidance, also have the advantage of providing an early warning to potential contamination which may result in lethal impacts 12. Avoidance behavior has been investigated for a diverse range of aquatic organisms including fish 13, amphipods 9, cladocerans 7, aquatic snails 6, and bivalves 14. These studies have all reported the usefulness of contaminant avoidance as a behavioral ecotoxicological endpoint.
Contaminant avoidance is likely to be a defense mechanism against the effect of poor environmental conditions, eliminating the threat to survival caused by the presence of toxicants 6. Contaminant avoidance may also be an adaptive behavior 15. Lefcort et al. 6 observed that aquatic snails collected from contaminated environments could avoid contaminated sediment, while those from reference sites could not, inferring a potential genetic influence on avoidance behavior. Regardless of the mechanisms triggering the avoidance response, it is important to note that the relocation of a population of aquatic organisms may ensure their survival at a new site. On an ecological scale, the disappearance of the species from the previous location is equivalent to the loss of the entire population 5, 7. Consequently, large areas of sediments that are avoided due to contamination may be classified as high risk, although short-term effects to many species used for toxicity tests may not be observed.
Static toxicity tests rarely reflect the dynamic and heterogeneous natural environment 6 and, therefore, do not always adequately mimic an organism's exposure to contaminants, meaning that these methods may lack ecological significance 1, 16. Despite this, many researchers and regulatory guidelines rely on standard methods which confine test organisms to a contaminant, forcing a continuous static exposure 17, 18. The relevance of these common ecotoxicological methods is widely debated in the literature, with chronic test methods often preferred over acute toxicity testing 1, 7. Chronic toxicity tests are considered to offer greater ecological relevance, protection at the population level, increased sensitivity, better prediction of toxicity, and the ability to model population effects 19. However, chronic toxicity tests generally use long exposure periods that often exceed 20 d 20, 21. Considering the ability of mobile aquatic organisms to avoid contamination, the long test durations often associated with chronic tests do not adequately represent realistic exposure conditions in the natural environment. Understanding the avoidance response of mobile benthic organisms will shed light on the response of benthic invertebrates to contaminated sediments in the environment and allow the development of environmentally relevant methods for assessing sediment toxicity caused by contamination in heterogeneous field settings.
The present study investigated the suitability of contaminant avoidance as an endpoint in whole-sediment toxicity testing using three species of benthic estuarine organisms: an epibenthic amphipod (Melita plumulosa), a harpacticoid copepod (Nitocra spinipes), and a snail (Phallomedusa solida). All of these species are known to graze or ingest sediment particles, making them suitable candidates for sediment toxicity testing 22–24. For each species the influence of varying physicochemical properties of the sediment on the distribution of test organisms and the optimal exposure time to measure an avoidance response elicited by contaminated sediments were initially determined. This provided optimal methods for avoidance bioassays for assessing the ability of each species to detect and move away from contaminated sediment.
MATERIALS AND METHODS
Clean seawater was collected from Port Hacking (Sydney, Australia), membrane-filtered (0.45 µm), and acclimated to room temperature (21 ± 1°C). Where necessary, the salinity of the filtered seawater was adjusted to the test salinity of 30 practical salinity units (psu) using deionized water (18 MΩ/cm; Milli-Q Academic Water System).
Relatively clean silty sediments were collected from an estuarine location in Bonnet Bay on the Woronora River (Sydney, Australia). These sediments had low or negligible concentrations of metal and organic contaminants and have been demonstrated to not cause toxic effects to the organisms used in the study 24, 25. The surface layers (upper 2–4 cm) of sediments were collected using clean Teflon spatulas and press-sieved through a 1.1-mm mesh on-site to remove coarse materials. The sediment was transferred into clean plastic bags with minimal headspace and stored in a cool room at 4°C for no longer than one month. Clean sand that contained negligible contamination 23 was also used as a control material in sediment-avoidance tests. Sediments with 50 and 10% silt were created by mixing the control sediment with clean sand.
Contaminated sediments were chosen to provide exposure to a range of contaminants. Contaminated sediments were collected from estuarine field sites in New South Wales and Tasmania, Australia. Collected sediments were stored at 4°C in the dark, and toxicity testing was undertaken within eight weeks 10, 25. Analyses of physicochemical properties (pH, organic carbon, particle size, acid-volatile sulfide) and metal contaminants were made on all sediments. Concentrations of total petroleum hydrocarbons were <250 mg/kg and those of polycyclic aromatic hydrocarbons were <1 mg/kg in all sediments 10, 25, 26.
Copper-spiked sediments were prepared using clean silty sediment (collected from the Bonnet Bay reference site) and equilibrated as described previously 27. The preparation, manipulation, and equilibration of copper-spiked sediments were done in a nitrogen gas-filled glove box at room temperature. Sediments were thoroughly homogenized with a plastic spoon and a bottle roller for 2 h at least two times per week. Sediments were adjusted to pH 7.5 with 1 M NaOH one day after copper spiking and maintained at this pH using small additions of NaOH as required during the one-month equilibration period. Changes in pH, redox potential, and dissolved metals in the porewater were monitored during the equilibration period.
Diesel-spiked sediments were prepared in two stages. Diesel oil was added to suboxic Bonnet Bay sediment to create a 10% diesel stock sediment. Following a two-week equilibration period, an aliquot of stock sediment was combined with suboxic Bonnet Bay sediment to give a final concentration of 5% diesel sediment. Diesel-spiked sediments were maintained in glass containers with minimal headspace. Lids were securely fastened to avoid losses through volatilization. Containers were housed inside two polypropylene bags, which were sealed with duct tape. Both the stock and test sediments were homogenized by vigorous shaking of the containers, followed by 2 to 3 h on a bottle roller at least three times per week during the one-month equilibration period.
All three species used in the present study were epibenthic deposit feeders found in intertidal estuarine environments of southeastern Australia. Melita plumulosa (family Melitidae) is commonly found in estuarine tidal mudflats ranging from silty to sandy sediments in freshwater, estuarine, and marine environments throughout southeastern Australia 28. Adult specimens of M. plumulosa typically range from 8 to 10 mm in length. The harpacticoid copepod species N. spinipes (family Ameiridae) is known to adapt to a wide range of environmental conditions (including salinity and temperature) and, as such, has a worldwide distribution 29. Mature copepods of this species are approximately 400 µm long. Phallomedusa solida (family Amphiboloidea), formerly known as Salinator solida, is one of the most common deposit-feeding gastropods found to inhabit salt marshes and mangroves in southeastern Australia 22. Snails of a uniform size (∼10 mm along the ventral surface) were collected for avoidance tests.
Specimens of M. plumulosa and N. spinipes were obtained from laboratory cultures maintained as per King et al. 28 and Ward et al. 29, respectively. Those of P. solida were collected from natural populations occurring in the intertidal mangroves at Bonnet Bay.
Amphipod toxicity testing
Glass beakers and acrylic beaker lids used for toxicity tests were cleaned in a dishwasher (Gallay Scientific) programmed for a phosphate-free detergent wash, a dilute acid wash (1% HNO3), followed by thorough rinsing with Milli-Q water.
The 10-d whole-sediment toxicity tests with M. plumulosa were conducted to assess the effects of continuous exposure to the sediments and were performed in accordance with standard protocols 23. In brief, tests were performed at 21 ± 1°C in a constant environmental chamber (Labec Refrigerated Cycling Incubator) on a 12-h light/12-h dark cycle (light intensity = 3.5 µmol photons/s/m2). Dissolved oxygen (>85%), pH (7.5–8.2), salinity (30 ± 1 psu), and temperature (21 ± 1°C) were monitored and maintained according to Spadaro et al. 23. For whole-sediment toxicity tests, three replicate 250-ml beakers containing 20 g of test sediment, 200 ml seawater, and 15 adult M. plumulosa were used per treatment. Every 2 to 3 d, 80% of the overlying water was replaced with clean seawater. In all tests, water samples were collected before and after each water change for dissolved metal copper analysis. No food was added during the 10-d bioassays.
At the termination of the tests, the contents of each beaker were gently sieved through a 180-µm stainless steel mesh sieve and transferred to large amphipod counting trays. Live amphipods were identified by movement and counted. The sieved sediment was transferred to 120-ml polycarbonate vials with 100 ml of seawater, fixed with 4 ml of 10% v/v neutral phosphate-buffered formalin, and stained with 5 ml of rose bengal solution (0.1 g rose bengal sodium salt [Sigma] per 100 ml Milli-Q). Sediments were left for 72 h to enable any surviving amphipods missed in the initial count to take up the stain and be counted. Tests were considered acceptable if the physicochemical parameters in beakers remained within the limits of pH 7.7 to 8.2, 20 to 22°C, dissolved oxygen >80% saturation, and salinity 28 to 32 psu throughout the test and if survival of amphipods was on average ≥80% in the controls.
Contaminant-avoidance experimental design
Avoidance assays were conducted in test chambers specifically designed and constructed for each test species (Supplemental Data, Fig. S1). Each design utilized a plastic vessel divided into two chambers by a permanent barrier fixed across the bottom of the container. The top of this barrier was at the sediment–water interface and allowed the contiguous placement of two sediments in the same container with minimal mixing of the sediment samples. Removable barriers were constructed from plastic (polyethylene terephthalate plastic film) and used to divide the test chambers from the base of the sediment to the surface of the overlying water. This barrier was used during the initiation and termination of experiments to prevent the migration of test organisms between test chambers. The non-toxic nature of all materials and adhesives was assessed before use to construct the testing apparatus. Test chambers were set up with sediment and filtered seawater and equilibrated for 24 h before the initiation of the experiment. The overlying water was replaced immediately before the test organisms were added.
For M. plumulosa and P. solida, test chambers were constructed from 3-L polyethylene containers (20 × 15 cm wide, 10 cm high; Supplemental Data, Fig. S1A). A plastic barrier approximately 1 cm high was adhered to the base across the middle of the container with non-toxic aquarium silicone (Sellys Glass Silicone). Homogenized sediments were placed on the relevant sides of the test chamber, and clean filtered seawater was added to achieve a depth of approximately 8 cm for M. plumulosa and approximately 5 cm for P. solida. For P. solida experiments, gel (Lucas' Pawpaw Ointment) was applied to the chamber walls above the water line to prevent the snails from escaping during testing (this showed no toxic effect in control studies).
The test chambers used for N. spinipes were constructed from 1 × 1 cm wide, 5 cm high polyethylene cells usually used for ultraviolet-visible spectrophotometry (Supplemental Data, Fig. S1B). An opening (0.5 cm wide × 2.5 cm tall) was cut into one side of each cell. Two cells were placed together with the opening facing each other and secured in place with adhesive tape. A sediment slurry was created (10 g sediment in 2 ml seawater) and added dropwise to the cells to achieve a layer of sediment approximately 0.5 cm deep overlain by approximately 3.5 cm of filtered seawater. Teflon spatulas were cut down and used to partition the two cells of the test chamber.
Avoidance experiments with M. plumulosa and N. spinipes were initiated by seeding one side of the sediment chamber with 30 test organisms. In avoidance assays where contaminated sediment was utilized, M. plumulosa and N. spinipes were always seeded into the chamber containing contaminated sediment. The removable barrier that divided the overlying water and prevented organism movement between treatments was removed after 10 min, and the organisms were permitted to move around the test chamber as desired. Because the snails had a lower mobility than the amphipods and copepods, the avoidance tests with P. solida were initiated by placing 20 snails in a line across the middle of the test chamber directly above the permanent barrier (no removable barrier was necessary).
All avoidance experiments were conducted under ambient conditions in a temperature-controlled laboratory maintained at 21 ± 1°C. For M. plumulosa and P. solida tests, the overlying water was gently aerated to maintain dissolved oxygen concentrations >85%. Overlying water in N. spinipes experiments was not aerated as the bubbling disturbed the sediment. Previous experiments with this species indicated that aeration of the overlying water was not necessary over short time periods (<7 d) to maintain sufficient levels of dissolved oxygen 29. Subsamples of the overlying water were collected and analyzed for dissolved methods (as described above) every 2 to 3 d during the test and immediately prior to test termination. No food was added to the test chambers for the duration of the test. Tests were terminated by inserting the temporary barriers into the test vessels. The sediment on both sides of the barrier was gently sieved, and the number of animals in the seeded and non-seeded chambers was recorded. For quality-control purposes, an average organism recovery of ≥80% was required for a test to be considered acceptable.
New plasticware was used for all chemical analyses. All chemicals used were analytical reagent grade or of equivalent analytical purity. Measurements of pH, salinity, temperature, and dissolved oxygen were made in accordance with the instrument manufacturers' instructions. Deoxygenated waters were prepared by bubbling solutions with high-purity, oxygen-free nitrogen gas for >8 h to give dissolved oxygen concentrations <0.1 mg/L. Measurements of pH, particle size distribution (wet sieving and gravimetry), organic carbon (high-temperature total organic carbon analyzer), and particulate metals (2:1 concentrated HCl:HNO3, heated) were made as described previously 23, 30. Overlying water samples were rapidly filtered through acid-washed 0.45-µm membrane filters (Minisart; Sartorius) immediately following collection and acidified to 2% HNO3 (v/v) with concentrated HNO3 (Tracepure; Merck). Acid-volatile sulfide and simultaneously extracted metals were analyzed according to Simpson 31. Dissolved metal concentrations in water samples and digested sediments were determined by inductively coupled plasma-atomic emission spectrometry (Varian 730-ES; Varian Australia) calibrated with matrix-matched standards as described in Angel et al. 32. Analyses of filter and digest blanks, replicates for 20% of samples, analyte sample spikes, and the certified reference material (PACS-2; National Research Council of Canada) were made as part of the quality assurance; and recoveries were within 85 to 110% of expected values. The limits of reporting for the various methods were less than one-tenth of the lowest reported values. All sediment-related concentrations are reported on a dry mass basis.
All statistical analyses were carried out using the Microsoft Excel (2007) data analysis tool pack. The distribution of organisms in avoidance bioassays was reported as a percentage of recovered organisms to account for small differences in the recovery rate of organisms between replicates. The difference in the proportion of organisms on either side of the test chambers was determined by a t test, with p < 0.05 resulting in a significant difference being detected. In all cases, avoidance was significant when there was a >10% difference in the mean distribution of organisms (between the contaminated and non-contaminated sides of the test vessel) and a lack of overlap of standard errors. Where a difference in the mean distribution of test organisms was <10% (equivalent to a difference of fewer than three animals), it was considered that avoidance had not occurred. This variation was assumed to reflect natural variability in bioassay experiments, and t tests were not performed.
RESULTS AND DISCUSSION
Grain size and avoidance behavior in uncontaminated sediment
Sediments can vary greatly in physical and chemical properties, and these variations can influence the nature of the resident ecosystem structure 10, 11. Before the avoidance of M. plumulosa, N. spinipes, and P. solida was determined for contaminated sediments, we determined whether varying sediment grain size and organic carbon content influenced the behavior of organisms when exposed to uncontaminated sediments (Table 1).
In vessels with uncontaminated silty sediment in both chambers, each of the test species dispersed to create a relatively even distribution of the test organisms on both sides of the test vessel over 48 h (Fig. 1). This confirmed that M. plumulosa, N. spinipes, and P. solida could move freely around the test chambers and had a preference for spreading out rather than remaining in a high-density group. This behavioral characteristic has been shown for several estuarine species including the amphipod Eohaustorius estuarius9 and Rhepoxynius spp. 33. The even distribution of the test organisms also suggests that ambient laboratory conditions did not influence organism dispersal.
When exposed to two different types of uncontaminated sediments M. plumulosa, N. spinipes, and P. solida dispersed throughout the vessel, as they had when contained in vessels with the same sediment on both sides (Fig. 1). For all three species, there was no significant difference in the distribution of the test organisms between the silt reference sediment and the silt and sand sediment mixtures. For M. plumulosa, a minor preference for 50% silt over the reference sediment was observed (Fig. 1A). However, a 9.4% difference in the distribution of these organisms between the sediments was considered to be the result of natural variability and not to represent an avoidance response.
The sediment mixtures tested in these experiments encompassed a range of sediment characteristics including particle, acid-volatile sulfide, total organic carbon, and metal concentrations (Table 1). In past studies, estuarine snails (Ilyanassa obsoleta) exposed to coarse sand containing no organic carbon showed a preference for silty sediment. This preference was attributed to the availability of food (as organic matter) in the silty sediment 34, 35. Meadows 36 reported that the amphipod Corophium volutator also had a preference for burrowing in finer-grained sediments compared to sandy sediments. The contaminated sediments used in the present study (Table 2) all contained ≥10% silt (except Sediment 8) and had total organic carbon and acid-volatile sulfide levels similar to those tested in the grain size–avoidance study (Fig. 1). Therefore, the properties of the test and reference sediments were not expected to influence the avoidance behavior of M. plumulosa, N. spinipes, and P. solida in subsequent experiments.
Table 2. Sediment chemistry data for the test sediments used in the present study
Silt refers to percentage of particles <63 µm. All metal concentrations were dilute acid-extractable metals. Amphipod survival was in 10-d toxicity testing.
Sediment 2 was a 5% diesel-spiked sediment that contained 26 mg/kg total polycyclic aromatic hydrocarbons (the sum of 16 polycyclic aromatic hydrocarbons 1 and 11,500 mg/kg C10–C36 total petroleum hydrocarbons).
AVS = acid-volatile sulfide. TOC = total organic carbon. ND = not determined.
14 ± 12
Amphipod survival (% ± standard error)
12 ± 5
35 ± 30%
24 ± 6
32 ± 2
47 ± 10
59 ± 11
63 ± 4
93 ± 9
96 ± 6
Time to response: Dispersal and avoidance of uncontaminated sediment
The time required to observe an avoidance response of M. plumulosa, N. spinipes, and P. solida for contaminated sediment was tested for periods ranging from 3 to 240 h. Under control conditions, M. plumulosa distribution was variable for shorter test durations (36 ± 21 to 64 ± 21% at 6 h for seeded and non-seeded sides, respectively). However, for longer durations (24–240 h) an even distribution of organisms occurred (Fig. 2A). This indicated that M. plumulosa were actively dispersing and occupying available space in the test vessel within 6 to 24 h but that a 48-h exposure period induced an avoidance response with less variability.
The distribution of N. spinipes was also considerably variable within the first 24 h (Fig. 2B), with shorter experiment durations having a greater proportion of organisms remaining in the seeded side of the test vessel (85 ± 4 and 75 ± 4% for 1- and 3-h exposures, respectively). Exposure times of 24 h were sufficient for the copepods to disperse across both sides of the test vessel (Fig. 2B) and were used for the remaining avoidance studies with N. spinipes.
The snail P. solida dispersed evenly throughout the test vessel after 6 h (48 ± 4% on the seeded side at 6 h; Fig. 2C). Minimal movement was observed in the snails at shorter time periods; therefore, shorter test durations were not included in the remainder of the study. In contrast to the amphipod and copepod, the distribution of P. solida between the two chambers of the test vessel did not change significantly as the duration of the experiments was increased.
The difference in time required to achieve an even distribution of the test organisms is expected to be influenced by both the mobility of the species and the starting position at the beginning of the assay. While P. solida was much less mobile than M. plumulosa and N. spinipes (DJ Ward, personal observation) , the shorter dispersal time for P. solida is most likely the result of placing them along the central barrier dividing the two test chambers. As the copepods and amphipods were distributed throughout the seeded chamber, the organisms potentially had a greater distance to travel than the snails before sensory systems could trigger an avoidance response to the less desirable areas of contaminated sediment.
Zylstra 37 found that dermal sensory cells in the freshwater pulmonate snail Lymnaea stagnalis were prevalent on the surface of the tentacles as well as around the lips, the front edge of the foot, and the mantle edge. The concentration of sensory cells on the tentacles, mouth, and margins of the lips reached densities up to 5,000 cells/mm2, resulting in a greater number of peripheral sensory neurons than in the central nervous system 38. In light of this, it makes sense that aquatic snails are sensitive to their surrounding chemical environment and that chemoreception is an important factor controlling the behavior of these organisms. Aquatic pulmonate snails have been shown to be extremely sensitive to chemical cues in the environment. For example, Physella columbiana has demonstrated the ability to avoid soil solutions with a mixture of zinc and cadmium at concentrations <10 µg/L 6. Therefore, we speculate that the chemoreceptors of P. solida are more sensitive than those of N. spinipes and M. plumulosa and, as such, enhance the ability of this species to sense the surrounding environment and move away from unfavorable conditions.
Organism density within a population may be a factor contributing to the speed at which the organisms disperse; for example, Rosenberg et al. 39 found that a higher density of brittle star Amphiura filiformis resulted in a significant increase in the rate of dispersal. Estuarine snails have also been shown to exhibit behavioral changes in response to high organism density, including increases in the incidence of floating, emigration, and climbing behavior as well as a reduction in feeding and crawling rates 40, 41. If aquatic organisms prefer low spatial densities, then physically larger species should react to overcrowding more actively to disperse within available space. This behavior was previously observed for two species of estuarine snail which displayed an increase in dispersal among larger (mature) specimens when resources were limited 42. As P. solida was the largest of the test species used in the present study, it is reasonable to expect that population density–related effects may occur and facilitate the dispersal rates. If considering only the dispersal of the test organisms under control conditions, it is possible that a density-dependent effect may encourage the spreading out of organisms in the test vessel; and this may be more significant for P. solida. We did not undertake tests to determine whether this may be the case. However, based on the results from the contaminant-avoidance experiments, we found that, under the same density constraints, all three test species showed a preference to migrate away from the contaminated sediment and occupy the clean, habitable sediment despite the resulting higher density of organisms.
Avoidance of contaminated sediment
The ability of the three species to avoid contaminated sediments was tested using the sediments shown in Table 2. The percentage of survival in 10-d lethality tests using the amphipod M. plumulosa is also shown. These were field-collected sediments, except for the diesel-spiked Sediment 2 and copper-spiked Sediments 5 and 7.
To determine whether the behavior of the test organisms was altered by the presence of contaminants, contaminated sediments were placed on one side of the test containers and uncontaminated reference sediment on the other. If the organisms no longer dispersed throughout the test vessel but remained at a higher density on the clean reference sediment, then this would be sufficient evidence that the chemical stimuli (in this case mostly metal contaminants) altered the behavior of the test organisms (Fig. 2).
The times to achieve uniform distribution of M. plumulosa, N. spinipes, and P. solida in uncontaminated sediments under control conditions were 48, 24, and 6 h, respectively; and changes to these times would also be used to provide information on avoidance behavior. We placed M. plumulosa in the test vessels with the reference and copper-spiked sediment (Sediment 5). After just 6 h of exposure, the distribution of M. plumulosa was significantly lower in the contaminated, seeded side (Sediment 5) compared to the uncontaminated, non-seeded side of the test chamber, with a distribution of 24 ± 4 to 76 ± 4%, respectively (Fig. 2A). Exposure durations of 24, 48, and 240 h showed a similar distribution to the 6-h result, with at least 85% of amphipods residing in the uncontaminated sediment, despite the extended time period. However, significantly more M. plumulosa avoided Sediment 5 following a 24-h exposure (95 ± 2%) than after a 48- or 240-h exposure (88 ± 5 and 85 ± 4%, respectively). Overall, M. plumulosa maintained a preference to inhabit clean sediments over longer exposure times, despite a higher population density (number of organisms per given area) than observed under control conditions.
While M. plumulosa clearly avoided Sediment 5 after a 6-h exposure period (Fig. 2A), the variability of amphipod distribution at 6 and 24 h in uncontaminated (control) treatments made interpretation of the contaminated sediment results difficult. A longer exposure period of 48 h was chosen for future avoidance experiments as it allowed sufficient time for the organisms in the control sediments to disperse and interact with the sediment before the experiment was terminated.
The copepod N. spinipes had the fastest avoidance time of the three test species. When exposed to contaminated Sediments 1 and 3 (Table 2) >87% of copepods occupied the non-contaminated chamber of the test vessel after only 3 h of exposure (Fig. 2B). This is a highly significant proportion of the copepod population migrating away from the contaminated sediment, showing a clear preference for the non-contaminated sediment. For Sediment 3, 1-h exposure resulted in 89 ± 4% of copepods moving to the uncontaminated sediments in the test chamber. For both sediments, there was no significant difference between the number of avoiding organisms at the termination of the experiment for all exposure times; hence, the increased exposure duration did not elicit a greater avoidance effect.
When exposed to Sediment 7, P. solida had an avoidance time of ≥16 h, with a significant number of snails (61 ± 2%) avoiding the contaminated sediment (Fig. 2C). This avoidance increased over time, with 48 h resulting in 78 ± 10% of P. solida migrating to the uncontaminated sediment. When exposed to Sediment 6, avoidance occurred after a shorter time than for Sediment 7, with 85 ± 6% of snails migrating away from the contaminated Sediment after 6 h.
The avoidance response time of P. solida was similar to the 6- to 24-h avoidance time of the estuarine snail Ilyanassa obsoleta when exposed to clean and contaminated harbor sediments 34, 35. Despite extending the duration of exposure to as much as 48 h, P. solida maintained a preference for the control sediment. Overall, the results show that in the presence of an unfavorable chemical stimulus, all three species will migrate away from the source and will remain there for extended periods of time.
Validation of sediment avoidance of M. plumulosa
The amphipod M. plumulosa is commonly used in Australia for sediment quality assessment 18, 30. For nine contaminated sediments that caused varying degrees of toxicity to M. plumulosa in 10-d lethality tests (Table 2), the avoidance response after 48 h was determined (Fig. 3).
Six of the contaminated sediments resulted in a significant avoidance response in M. plumulosa within 48 h. The greatest avoidance occurred for Sediments 1, 2, and 3, which resulted in 97, 94, and 93% of M. plumulosa moving to the uncontaminated sediment, respectively. These sediments were highly toxic, with Sediments 1 and 3 resulting in >75% 10-d lethality (Table 2). While it was not possible to accurately determine the 10-d lethality of the diesel-spiked Sediment 2 (perhaps due to the volatility of the diesel changing during the exposure), the sediment was strongly avoided. Exposure of M. plumulosa to sediments that were moderately toxic (Sediments 5, 6, and 7) did not always result in a significant avoidance response. More than 75% of amphipods avoided Sediments 5 and 7; however, avoidance was not significant for Sediment 6 despite resulting in >40% lethality in 10-d toxicity tests. Sediments 5 and 7 contained similar concentrations of copper but differed significantly in grain size distribution (Table 2). When M. plumulosa was exposed to sediments that did not elicit acute toxicity (Sediments 8 and 9), avoidance was not observed.
In all exposure chambers the overlying water was continuously exchanged between the two sides and dissolved metal concentrations in the overlying waters were relatively low (e.g., <30 µg Cu/L for tests with sediments 5 and 7) and not significantly different for samples taken from the different sides of the test chamber (data not shown). As it is known that M. plumulosa is sensitive to sediment-bound copper 28, 43, avoidance of these sediments was believed to be due to differences in particulate copper or the flux of copper at the sediment–water interface, rather than differences in dissolved copper in the overlying water. Sediments 5 and 7 were copper-spiked sediments, and although they had been equilibrated and porewater copper concentrations were <5 µg/L (unpublished results), the copper in these sediments was likely to be more bioavailable than if present in field-collected sediments with similar copper concentrations 30. The greater avoidance of Sediments 5 and 7 compared to Sediment 4, which had high lead and zinc concentrations and greater toxicity, indicates that M. plumulosa may be more sensitive to copper than other metal contaminants. However, although dilute acid-extractable metal concentrations (Table 2) are expected to predict the bioavailability of metals better than total metal concentrations 25, the fact that Sediments 5 and 7 were copper-spiked may have resulted in this copper being easier for M. plumulosa to sense than the lead and zinc in the field-collected Sediment 4.
Despite the high lead and zinc concentrations in Sediment 9, this sediment was not avoided and was also not toxic to M. plumulosa. This sediment had a relatively high concentration of acid-volatile sulfide (14 ± 12 µmol/g), and the lead and zinc are predicted to be present in sulfide phases of low bioavailability 44. Interestingly, M. plumulosa appeared to show a slight attraction to this sediment, with 61% of amphipods remaining on the seeded portion of the test vessel, effectively choosing to inhabit the contaminated sediment. However, there was no statistical significance between the distribution of amphipods on the seeded and unseeded chambers at the end of the test.
Experiments with uncontaminated sediments found that sediment grain size did not influence the avoidance response for M. plumulosa, and this also appeared to be the case for the contaminated sediments. Marklevitz et al. 35 found that avoidance of contaminated harbor sediments by a mollusk was a balance between attraction to food and aversion to contamination. Given the high concentration of organic carbon present in Sediments 4 and 9 (6.8 and 6.0%, respectively, which exceeds the reference sediment), it is possible that the attraction to the added nutrition of the test sediments (presumably provided by the additional organic content of the sediment) may have affected the observed avoidance of these sediments. Sediment 3 also had a high organic carbon content (5.9%); however, the observation that it was both avoided and toxic (>75% mortality) indicates that if nutrition in the form of total organic carbon influences avoidance behavior, then the quality of the total organic carbon may also be important. The major contaminant in Sediment 3 was copper and, when considering the results for Sediments 5 and 7, may indicate that sediment copper is more readily detected and avoided by the amphipod.
Influence of hazard magnitude on avoidance behavior
The increasing hazard posed by contaminated sediments is thought to elicit a faster or greater response from exposed organisms 35, 45. For M. plumulosa, the results of the present study indicate a similar trend as the degree of avoidance was generally greater for sediment which elicited greater toxicity (Table 2, Fig. 3).
This indicates that the hazard imposed by the contaminated sediment (i.e., toxicity) influences the number of amphipods that avoid the unfavorable sediment over a 48-h period. The avoidance response of the copepods to Sediments 1 and 3 was very similar to the amphipod response (Fig. 2), but over a shorter time, with 3 h (87 ± 4 and 92 ± 5%, respectively) and 24 h (83 ± 5 and 82 ± 4%, respectively). Given the similar toxicity of these two sediment samples and the similarity between the avoidance response of the copepods and amphipods, the avoidance behavior of N. spinipes may also increase as toxicity increases. However, overall an insufficient range of contaminated sediments was tested to determine if this was true for N. spinipes and P. solida.
Comparing the distribution of M. plumulosa and N. spinipes exposed to control conditions at each time point and the corresponding contaminant avoidance provides strong evidence that the presence of sediment-bound contaminants influences the observed avoidance response (Fig. 2). In test chambers containing only control sediment, approximately 24 h was required for a significant percentage of the organisms to migrate away from the seeded portion of the container compared to <6 h when contaminants were present. This was despite possible density-related pressures that may be placed on the organisms due to seeding on only one side of the test container. It is particularly evident for N. spinipes, where after 1 h only 15 ± 4% of the copepods migrated away from the seeded control treatment compared to 89 ± 4% of copepods avoiding Sediment 3 (onto which they were seeded) in the same period of time. While the response was not as rapid or clear for M. plumulosa, the distribution after 6 h was highly variable for the controls; but 76 ± 4% of amphipods avoided the contaminated Sediment 5. The enhanced activity of M. plumulosa and N. spinipes was clearly avoidance behavior, and toxicity would have occurred if the species remained in the sediments.
The snail P. solida was slower-moving, but more rapid avoidance was observed for Sediment 6 (≤6 h) than for Sediment 7 (∼16 h; Fig. 2). While Sediments 6 and 7 caused similar acute toxicity (Table 2), accompanying studies showed that Sediment 6 caused much greater chronic toxicity to M. plumulosa (D Spadaro, CSIRO Land and Water, Lucas Heights, NSW, Australia, unpublished data). In addition, the degree of avoidance exhibited by P. solida exposed to Sediment 6 for 6, 24, or 48 h was not significantly different (<7% difference in the number of avoiding organisms). This suggests that density-related pressures did not influence the behavior of P. solida; however, tests were not undertaken to determine whether this was the case here. After 48 h, the snails preferentially maintained a higher density in clean sediments rather than inhabiting the contaminated Sediments 6 and 7.
Previous studies have demonstrated behavioral changes in benthic organisms resulting from the presence of environmental contaminants. The bivalve Macomona liliana was shown to crawl or drift away from sediments spiked with copper and zinc 14, 46. Thus, a behavioral endpoint for the bivalve was more sensitive than both burial/morbidity and mortality endpoints 46. Physella columbiana, an aquatic pulmonate snail, demonstrated the ability to detect cadmium and zinc at concentrations of 9 µg Cd/L, 56 µg Zn/L, and below 10 µg/L for mixtures 6. Oakden 33 reported the avoidance of metal- and sewerage-enriched sediments by phoxocephalid amphipods (Rhepoxynius spp.), while Kravitz et al. 9 found that Eohaustorius estuarius could avoid sediments with moderate to heavy polycyclic aromatic hydrocarbon contamination.
The present study demonstrated the ability of M. plumulosa, N. spinipes, and P. solida to respond to chemical cues in the environment and choose a habitat that provides the best opportunity for survival by avoiding contaminated sediment. The avoidance response of the test organisms indicates that static 10-d toxicity methods are likely to overestimate toxicity for species, which would avoid contamination in heterogeneous field settings. The sensitivity of the avoidance response of each species to sediment-bound contaminants indicates that they may be suitable for development as rapid screening methods to assess sediment quality. The study provided strong evidence that the avoidance response of M. plumulosa was related to the toxicity associated with the sediment. For this species the 48-h avoidance endpoint correctly identified as a potential hazard six of seven sediment samples that caused significant acute toxicity during 10-d exposures.
Figure S1. (10 KB PDF).
Funding for the project was provided by an Australian Postgraduate Award and a CSIRO top-up scholarship. D. Spadaro and I. Hamilton are thanked for their assistance in the field and laboratory.