Effects of lead-spiked sediments on freshwater bivalve, Hyridella australis: linking organism metal exposure-dose-response

Aquatic Toxicology 149 (2014) 83–93

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Effects of lead-spiked sediments on freshwater bivalve, Hyridella australis: linking organism metal exposure-dose-response Chamani P.M. Marasinghe Wadige ∗ , Anne M. Taylor, William A. Maher, Rodney P. Ubrihien, Frank Krikowa Ecochemistry Laboratory, Institute for Applied Ecology, University of Canberra, Canberra, ACT 2601, Australia

a r t i c l e

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Article history: Received 17 October 2013 Received in revised form 19 January 2014 Accepted 22 January 2014 Keywords: Biologically active lead Biologically detoxified lead Biomarkers Hyridella australis Oxidative stress Sub-cellular partitioning

a b s t r a c t Lead entering aquatic ecosystems adsorbs to sediments and has the potential to cause adverse effects on the health of benthic organisms. To evaluate the freshwater bivalve Hyridella australis as a bioindicator for sediment toxicity, their exposure-dose and response to lead contaminated sediments (< 0.01, 205 ± 9 and 419 ± 16 ␮g/g dry mass) was investigated in laboratory microcosms using 28 day exposures. Despite high concentrations of lead in the sediments, organisms accumulated low concentrations of lead in their tissues after 28 days of exposure (low treatment: 2.2 ± 0.2 ␮g/g dry mass, high treatment: 4.2 ± 0.1 ␮g/g dry mass), however, accumulated lead concentrations in lead exposed organisms were two fold (low treatment) and four fold (high treatment) higher than that of unexposed organisms (1.2 ± 0.3 ␮g/g dry mass). Accumulation of lead by H. australis may have occurred as analogues of calcium and magnesium. Labial palps accumulated significantly more lead than other tissues. Of the lead accumulated in the hepatopancreas, 83%–91% was detoxified and stored in metal rich granules. The proportions and concentrations of lead in this fraction increased with lead exposure, which suggests that lead detoxification pathway plays an important role in metal tolerance of H. australis. The biologically active lead was mainly present in the mitochondrial fraction which increased with lead exposure. Total antioxidant capacity of H. australis significantly decreased while lipid peroxidation and lysosomal membrane destabilation increased with lead exposure. This study showed a clear lead exposure-dose-response relationship and indicates that H. australis would be a good biomonitor for lead in freshwater ecosystems. © 2014 Published by Elsevier B.V.

1. Introduction Metals are mined in many countries worldwide and significant amounts are deposited into aquatic ecosystems causing adverse effects to biota (Beltman et al., 1999; Besser et al., 2009; Brumbaugh et al., 2005; Farag et al., 1998; Khozhina and Sherriff, 2008). Mine accidents such as collapses of tailings dumps and dam spills may exacerbate metal contamination in aquatic environments (Grimalt et al., 1999; Korte et al., 2000; Pérez-López et al., 2009; Soldán et al., 2001). The Molonglo River, New South Wales, Australia is contaminated with metals originating from the Captains Flat mining activities, in particular, two separate collapses of tailings dumps in 1939 and 1943 (Weatherley et al., 1967). The inputs of metals to the river and contamination of river biota have been studied on several occasions since the late 1960s (Brooks, 1980; Graham et al., 1986; Nicholas and Thomas, 1978; Norris, 1986;

∗ Corresponding author Tel.: +61 469 786 330. E-mail address: [email protected] (C.P.M. Marasinghe Wadige). 0166-445X/$ – see front matter © 2014 Published by Elsevier B.V. http://dx.doi.org/10.1016/j.aquatox.2014.01.017

Weatherley et al., 1967) but none of these studies have integrated the uptake of metals with responses of benthic biota. The most recent study conducted in 1996 (Sloane and Norris, 2003) recorded concentrations of cadmium – 11.8 ␮g/g, copper – 748 ␮g/g, lead – 4137 ␮g/g and zinc – 3534 ␮g/g dry mass in Molonglo River sediments. These values exceed the high trigger value of the ANZECC/ARMCANZ (2000) interim sediment quality guidelines (ISQG) for cadmium – 10 ␮g/g, copper – 270 ␮g/g, lead – 220 ␮g/g and zinc – 410 ␮g/g dry mass. The sediment bound metals have the potential to be released into the water column under suitable physicochemical conditions, leading to further contamination of aquatic systems (Arakel, 1995; Macklin et al., 1997) thus posing a risk to benthic organisms (Chapman et al., 1998; Lamoureux and Brownawell, 1999; Tessier et al., 1993). In this context, the fate and effects of metals in Molonglo River sediments needs to be investigated. The exposure-dose-response framework provides a means of linking chemical exposure to biological responses to assess the health of organisms and ecosystems (Salazar and Salazar, 1997, 2000; Taylor and Maher, 2012a, b, c). A suitable sentinel organism is required for use within this frame work. Sentinel organisms map the bioavailable metal fraction by

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retaining the contaminant in tissues and should show a simple correspondence between tissue and ambient bioavailable contaminant concentrations over an area/time interval (Beeby, 2001). In this context, a number of freshwater bivalves (Anodonta anatina, Corbicula fluminea, Dreissena polymorpha, Unio tumidus and Velesunio angasi) have been used in metal toxicity studies (Bollhöfer, 2012; Bollhöfer et al., 2011; Doyotte et al., 1997; Falfushynska et al., 2013; Marie et al., 2006). In the environment, as recorded in the Molonglo River sediments, contaminants are present as mixtures (Stewart, 1999). Before applying a new sentinel organism to monitor environmental contamination, sufficient knowledge of metal accumulation, handling strategies and biologically significant effects need to be investigated under laboratory controlled condition (Mersch et al., 1993; Rainbow et al., 1990). Metals may exert effects on organisms through interference in the redox cycling pathway, resulting in the production of reactive oxygen species (ROS) which react with critical cellular targets such as DNA, proteins and lipids, potentially leading to DNA damage, enzyme inactivation, lipid peroxidation and, ultimately, cell death (Winston and Di Giulio, 1991). Antioxidant defence systems in organisms have evolved to detoxify ROS, as well as repair oxidized components (Alves de Almeida et al., 2007). Changes in antioxidant defence enzymes such as superoxide dismutases, catalase and glutathione peroxidases have been measured as metal related oxidative stress biomarkers in freshwater bivalves (Cossu et al., 2000; Doyotte et al., 1997; Guidi et al., 2010). The objective of the present study was to evaluate the use of the freshwater bivalve Hyridella australis as a sentinel organism to assess the bioavailability and toxicity of lead in freshwater sediments. For this purpose, effects of lead-spiked sediment on H. australis under laboratory controlled conditions were investigated by examining the exposure-dose-response relationships. Organism external exposure was measured as the total lead concentration in the spiked sediments. Total lead burden in soft body tissues and sub-cellular distribution of lead in hepatopancreas tissue was measured as organism internal dose. The sub-cellular distribution provides evidence of metabolically available and detoxified forms of lead. Total antioxidant capacity, lipid peroxidation and lysosomal membrane stability were measured as enzymatic and cellular biomarkers of lead effects. Lead was investigated as it has the highest recorded metal concentrations in Molonglo River sediments (Sloane and Norris, 2003). It is a non-essential metal and has no known physiological function (Johannesson, 2002), and even in trace amounts can be toxic to aquatic organisms (Angelo et al., 2007; Bollhöfer, 2012; Krause-Nehring et al., 2012). Lead toxicity stems from its ability to mimic biologically essential metals such as calcium, iron and zinc (Company et al., 2011; Company et al., 2008). Lead is a redox inactive metal but has a high affinity for sulfhydryl groups of biologically important enzymes such as ␦-aminolevulinic acid dehydrase, which are involved in heme synthesis (Astrin et al., 2006; Bechara, 1996); superoxide dismutase and glutathionine peroxidase that are involved in the removal of ROS (Dafre et al., 2004; Ercal et al., 2001). In this context, free radical-induced damage by lead is generally achieved via depletion of the cellular antioxidant pool (Ercal et al., 2001; Jomova and Valko, 2011), therefore, increased production of ROS may indirectly lead to lipid peroxidation of cell membranes and the formation of secondary products such as malondialdehyde (Daniel et al., 2004; Ercal et al., 2001; Sandhir and Gill, 1995). H. australis was selected as a bioindicator organism in this study because it meets the requirements of a good biomonitor (Phillips and Rainbow, 1994). H. australis is a sediment burrowing organism, has a restricted mobility, is a filter feeder and is representative of Australian freshwater ecosystems. Moreover, it is hardy and has an ample amount of tissue for metal and biomarker

analysis. The related unionid bivalve Hyridella depressa has been used extensively in water-borne metal accumulation studies in Australia (Byrne and Vesk, 2000; Markich et al., 2001; Markich and Jeffree, 1994), however, H. australis has not previously been used in toxicological studies. 2. Materials and Methods 2.1. Study Design H. australis was exposed to concentrations (< 0.01, 220 and 440 ␮g/g dry mass) of lead-spiked and control freshwater sediments in laboratory microcosms for 28 days. The lowest lead concentration (220 ␮g/g) used for the sediment spiking represents the high effects trigger value concentration of lead in the ANZECC/ARMCANZ (2000) ISQG and the highest concentrations (440 ␮g/g) represents two times this value. Three replicate microcosms per lead treatment were used. Dose was measured by total lead content in labial palps, gill, mantle, visceral mass and muscle tissues of two individuals from each treatment replicate (n = 6) at day 7, 14, 21 and five individuals from each treatment replicate (n = 15) on day 28. Markich and Jeffree (1994) and Markich et al. (2001) have shown that the uptake of lead in unionid bivalve H. depressa occurred as an analogue to calcium uptake, and metal toxicity most likely results from the competitive binding of calcium and magnesium via the same channels of the cell membrane. Therefore, in the present study the relationship between tissue lead accumulation and tissue calcium and magnesium concentrations were determined for each treatment at day 28. Sub-cellular distribution of lead was examined in hepatopancreas tissues at day 28 to evaluate the biologically active and detoxified lead fractions (n = 3 per treatment). Hepatopancreas tissue was selected for sub-cellular lead determination as it acts as a repository for metal ions and a high level of metabolic impairment occurs in this tissue (Das and Jana, 1999). Enzymatic biomarkers (total antioxidant capacity and lipid peroxidation) were measured in six individuals from each treatment at day 7, 14, 21 and nine individuals after 28 days exposure. A cellular biomarker of lead toxicity (lysosomal membrane stability) was measured in hepatopancreas tissues of exposed organisms after 28 days exposure (n = 9 per treatment). 2.2. Sediment, water and H. australis collection Freshwater sediments and water used in bivalve acclimation tanks and microcosm experiments were collected from the uncontaminated Warrambucca Creek, New South Wales, Australia. Sediment was collected from the surface layer (2–4 cm) using a clean stainless steel shovel. Collected sediment was placed into 20 L plastic containers and allowed to settle for 30–60 min; surface water then was carefully removed prior to transportation to the laboratory. Sediments were sieved through a 2 mm stainless steel sieve to remove large pieces of organic debris, rocks and organisms and stored in a cool room at 4 ◦ C until use. H. australis (55 ± 5 mm shell length) were collected by hand from the minimally polluted site at the Nepean River, near Menangle, south-west of Sydney, NSW, Australia, where the unionid bivalve H. depressa has been collected for previous studies (Byrne and Vesk, 2000; Markich et al., 2001). The measured sediment lead concentrations of the bivalve collection site was 5.7 ± 0.1 ␮g/g dry mass (Mean ± SE; n = 3) which is significantly lower than the low effects trigger value concentration of lead (50 ␮g/g dry mass) in the ANZECC/ARMCANZ (2000) interim sediment quality guidelines. >After collection, mussels were placed into a clean plastic cooler with water and sediment collected from the same site. Overlaying water was aerated during transportation to the laboratory. Mussels were maintained for two weeks in uncontaminated sieved sediments, a sediment depth

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of 8–10 cm with overlaying water (20 cm) in glass aquaria in a temperature-controlled room for acclimation before experimentation. Overlaying water in glass aquaria was continuously aerated at a temperature of 20 ± 0.1 ◦ C to match mean water temperature in the Nepean River system with a day/night light cycle of 12/12 h (Markich et al., 2003). During this period, mussels were fed once every three days with a commercially available freshwater mussel food (unicellular green algae – Nannochloropsis – Nanno 3600, Instant algae, USA.) at 1% (v/w) of total body mass. Half water changes were performed every three days. The mean total tissue lead concentrations measured in unexposed organism was 1.2 ± 0.3 ␮g/g dry mass (Mean ± SE; n = 6). 2.3. Sediment lead spiking Sediments were homogenized and spiked with lead following the method developed by Besser et al. (2011) adapted from Simpson et al. (2004) for freshwater sediments. Briefly, lead as Pb (NO3 )2 (M & B, England) was added to 2 kg of wet sediment in a 2 L glass container to obtain a final lead concentration of 440 ␮g/g dry mass. Water was added to obtain a sediment: water ratio of 4: 1 (v/v) and spiked sediments homogenised using a plastic spatula. Sediment pH was measured immediately after metal addition and weekly during the equilibration period. pH was maintained at 7.2–8.0 by adding 1 M NaOH (AR grade BDH, Australia), prepared in deoxygenated freshwater. The sediment slurry was purged with nitrogen for 2 h. Head space of the containers was filled with nitrogen before closure and rolled (Cell-production Roller) for 2 h before being returned to a dark cupboard at room temperature (22–25 ◦ C). Containers were subsequently mixed on the roller for two hours per day. Control sediments were treated identically with spiking of lead replaced by spiking with KNO3 (AnalaR®, BDH, England) to give the same concentration of NO3 − as in Pb (NO3 )2 -spiked sediments to eliminate the confounding effects of added NO3 − . Pore water samples were taken weekly by centrifugation of a sediment subsample (2500 g for 20 min) at 4 ◦ C and lead measured to determine when the added Pb (NO3 )2 was completely bound to sediment particles. Collected pore water samples were filtered (0.45 ␮m) and acidified to 1% (v/v) with nitric acid (Suprapur® Nitric Acid, Merck, Germany) and analysed using an ELAN DRC-e ICP-MS (PerkinElmer, SCIEX, USA) (Maher et al., 2001). Once pore water lead concentration had fallen below 7 ± 1 ␮g/L (after 6–7 weeks), a half volume of sediment from spiked sediment was mixed with the same volume of uncontaminated sediments to produce low lead-spiked sediment (220 ␮g/g dry mass). The initial (at day 0) and final lead concentrations (at day 28) of the low-spiked sediments were 207 ± 13 and 196 ± 13 ␮g/g dry mass, respectively. In the high-spiked lead treatment, initial and final lead concentrations of the sediment were 446 ± 21 and 393 ± 15 ␮g/g dry mass, respectively. As the initial and final lead concentrations in the spiked sediments were not significantly different, the reported exposure concentrations are the mean value of day 0 and day 28 sediment lead concentrations (control: < 0.01, low: 205 ± 14 and high: 419 ± 16 ␮g/g dry mass). 2.4. Laboratory exposure of H. australis to lead-spiked sediments H. australis were exposed for 28 days to each of the three lead treatments (control: < 0.01, low: 205 ± 14 and high: 419 ± 16 ␮g/g dry mass) following the methods described in ASTM (2010) and USEPA (2000) for sediment bioaccumulation tests. Each treatment consisted of three treatment replicates in 12 L polystyrene aquariums. 2000 g of wet sediment was added to each aquarium and filled with 10 L of water. Each aquarium was left for 24 h to settle and the temperature to equilibrate to 20 ± 0.1 ◦ C. Again a day/night light cycle of 12/12 h was used. Overlaying water was aerated without agitating settled sediments through air flow

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controllers to maintain ≈ 100% air saturation. Twelve H. australis were added to each aquarium. H. australis feeding and water changes were as previously described. 2.5. Lead measurements 2.5.1. Sediments Approximately 0.2 g of freeze-dried sediment or certified reference material (NIST- RM 8704, Buffalo River sediment) was digested using 2 mL HNO3 (Suprapur® Nitric Acid, Merck, Germany) and 1 mL of HCl (Suprapur® Hydrochloric Acid, Merck, Germany) in a MARS microwave oven using a two-step time and temperature programme (Telford et al., 2008). Lead concentrations were measured using an ELAN DRC-e ICP-MS (PerkinElmer, SCIEX, USA) (Maher et al., 2001). Measured lead concentration (mean ± SD; n = 5) from NIST- RM 8704 certified reference materials, was 132 ± 4 ␮g/g dry mass and in fair agreement with certified value 150 ± 17 ␮g/g dry mass. 2.5.2. Biota Tissues were microwave digested as described by Baldwin et al., (1994). Freeze-dried tissue samples were finely ground to a homogenous powder in an acid-cleaned mortar. Approximately 0.07 g of tissue sample or certified reference materials (TORT- 2, lobster hepatopancreas tissue and NIST 1566b, oyster tissue) were digested in 1 ml of nitric acid (Suprapur® Nitric Acid, Merck, Germany) in a microwave oven (CEM MDS-2000, USA) for 2 min at 630 W, 2 min 0 W and 45 min at 315 W. Lead, calcium and magnesium concentrations were measured using an ELAN DRC-e ICP-MS (PerkinElmer, SCIEX, USA) (Maher et al., 2001). Lead concentrations measured in the certified reference material- TORT- 2 and NIST 1566b were 0.47 ± 0.12 ␮g/g and 0.312 ± 0.004 dry mass, respectively (mean ± SD; n = 20) and in agreement with certified value (TORT- 2 0.35 ± 0.13 ␮g/g dry mass; NIST 1566b – 0.308 ± 0.009 ␮g/g dry mass). Calcium and magnesium concentrations measured in oyster tissue NIST 1566b were (mean ± SD; n = 20) 0.082 ± 0.001 and 0.1103 ± 0.0015 mass fraction, %, respectively and in agreement with certified values (Ca: 0.0838 ± 0.0020 mass fraction, % and Mg: 0.1085 ± 0.0023 mass fraction, %). 2.5.3. Sub-cellular partitioning of lead The sub-cellular distribution of lead in hepatopancreas tissues of H. australis was measured using the procedure developed by Taylor and Maher (2012b). Subsamples of tissues were homogenised in Ca2+ /Mg2+ free saline buffer on ice using an IKA® Labortechnick Ultra-turrax-T10 homogeniser equipped with an S10N-5G dispersing tool (Janke & Kunkel, Germany) at 3500 rpm for 30 sec as three 10 sec pulses (Graham, 1997). >Six different fractions; granules (P2), heavy organelle fraction (P3), light organelle and particulate fraction (P4), heat sensitive proteins (P5), nuclei and cellular debris (S2), and heat stable metallothionein like protein fraction (S5) were obtained after differential centrifugation using a 5804R centrifuge (Eppendorf, Germany) and Himac CP90WX preparative ultracentrifuge (Hitachi, Japan) (Fig. 1). Each sub-cellular fraction was acidified to 1% (v/v) with nitric acid (Suprapur® , Merck, Germany) and placed in a water bath set at 80 ◦ C for 8 h. Lead concentration was then analysed by AAnalyst 600 Atomic Absorption Spectrometer (PerkinElmer, USA). 2.6. Biomarkers of lead toxicity 2.6.1. Enzymatic biomarkers – Total antioxidant capacity and lipid peroxidation 2.6.1.1. Tissue preparation. Dissected hepatopancreas tissue was homogenised on ice in 500 ␮L of a 5 mM potassium phosphate

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Fig. 1. Summary of the differential centrifugation procedure used to isolate lead in different sub-cellular fractions of hepatopancreas tissues of Hyridella australis. After homogenization, differential centrifugation and tissue digestion steps were used to obtain six different fractions (Four pellets P2, P3, P4, P5 and two supernatants S2 and S5).

buffer containing 0.9% (w/v) sodium chloride and 0.1% (w/v) glucose, pH 7.4 (1:5, w/v) using a motorised microcentrifuge pellet pestle, sonicated on ice for 15 s at 40 V and centrifuged (Eppendorf, 5804R centrifuge – Germany), at 10, 000 g for 15 min at 4 o C (Cayman, 2011). Total antioxidant capacity (TAOC) and lipid peroxidation (malondialdehyde, MDA) were carried out immediately after tissue preparation. Sub samples of tissue lysates were stored at – 80 ◦ C for protein analyses and analyses were completed on the following day.

2.6.1.2. Total antioxidant capacity assay. The TAOC of tissue lysates were measured using a Cayman Chemical assay (Sapphire Bioscience # 709001). The assay is based on the ability of the antioxidant in the sample to inhibit oxidation of 2, 2’-Azino-di[3-ethylbenzthiazoline sulphonate] (ABTS) to radical cation (ABTS. + ) by metmyoglobin in the presence of hydrogen peroxide. 10 ␮L of supernatant, 10 ␮L of metmyoglobin and 150 ␮L of Chromogen were pipetted into a 96 well plate. The reactions were initiated by adding 40 ␮L of 441 ␮M hydrogen peroxide. The (ABTS. + )

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2.6.1.3. Lipid peroxidation assay. Lipid peroxidation was determined by thiobarbituric acid reactive substances assay (Oxitek® TBARS assay- Zeptometrix corporation, # 0801192). It measures the MDA concentration, which is a by-product of lipid peroxidation in tissue lysates. The MDA present in the sample forms a 1:2 adduct with TBARS and produces a colour which can be read spectrophotometrically at 532 nm. 100 ␮L of tissue lysate and 100 ␮L of sodium dodecyl sulphate (SDS) were added into glass tubes. 2.5 mL of TBA/Buffer reagent was then added and incubated at 95 o C for 60 min. After cooling samples were centrifuged at 3000 rpm for 15 min and absorbance was read at 532 nm on a Bio Rad Benchmark plus microplate spectrometer and compared with MDA standard curve. 2.6.1.4. Protein assay. The protein concentration in tissue lysates used for TAOC and TBARS assays were analysed using the FluoroProfileTM (Sigma #FP0010, Sigma–Aldrich, USA) protein assay. It is a fluorescent assay based on Epicocconone, a biodegradable natural product. Tissue lysate were thawed at room temperature and diluted at 1:10 (v/v) with deionised water. 50 ␮L of diluted sample and the same volume of fluorescent reagent were pipetted into an F 96 MicrowelTM plate. The fluorescence intensity was read at 485 nm excitation and 620 nm emission, on a Luminoskan Ascent Fluorescence Plate Reader (Thermo Electrical Corp., USA). Bovine serum (BSA) calibration curve standards were made up in sample buffer. 2.6.2. Cellular biomarker – Lysosomal membrane stability The lysosomal stability test used was based on a procedure of Ringwood et al. (2003) developed for oysters with some modifications. Minced hepatopancreas tissues in 500 ␮L of Ca+2 , Mg+2 free saline (CMFS) buffer (20 mM HEPES, 450 mM NaCl, 12.5 mM KCl and 5 mM tetrasodium EDTA adgested to pH 7.35 with 6N NaOH) were shaken on a reciprocating shaker at 110 rpm for 20 min. Samples were shaken for a further 20 min after adding 500 ␮L of collagenase in Mg+2 free saline (MFS) buffer (20 mM HEPES, 450 mM NaCl, 12.5 mM KCl and 5 mM CaCl2 adjusted to pH 7.53 with 6N NaOH). Sheared samples were then filtered through 40 ␮m screen and centrifuged at 200 g at 15 o C for 8 min to collect hepatopancreas cells. After discarding supernatant, cells were again resuspended in 1 mL CMFS and centrifuged at 200 g at 15o C for 5 min. Cells were then incubated in 0.04 mg/mL of neutral red (Sigma, USA) in CMFS for 1 h and one hundred cells per slide were counted under the light microscope at 400 X magnification. Dye present in the cytoplasm was scored as cells with unstable lysosomes. Healthy cells retain the dye in the lysosomes. 2.7. Statistical analysis All statistical tests were performed using SPSS 20 and visually represented by Microsoft Excel 2010. Tests of normality of data were verified using Shapiro–Wilk test, and homogeneity of variance of the data was checked by Levene’s test before a comparison of means. If data were normally distributed and met the assumptions of homogeneity of variance, parametric tests, one way analysis of variance or factorial analysis of variance (ANOVA) was performed. If data showed significant differences (significant at p ≤ 0.05), post hoc comparisons for means were made by TukeyKramer’s test. If data were not normally distributed logarithmic transformation was applied and assumptions were tested. If data

5

Tissue Pb µg/g dry mass

was measured at 25 o C and absorbance read at 750 nm on a Bio Rad Benchmark plus microplate spectrometer. The capacity of the antioxidants in the sample to prevent oxidation of ABTS was compared with the antioxidant capacity of a standard, 6-hydroxy2,5,7,8-tetramethylchromane-2-carboxylic acid (Trolox).

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c c

4

bc bc

b 3

b b

a 2

a

a

a

1

0 Control

Low

High

Sediment Pb µg/g dry mass Day 0

Day 7

Day 14

Day 21

Day 28

Fig. 2. Lead concentrations in whole tissues of Hyridella australis at weekly intervals over 28 days of exposure to lead-spiked sediments: control < 0.01; low 205 ± 9; high 419 ±16 ␮g/g dry mass. Mean ± SE, all treatments n = 6 at days 0, 7, 14, 21and n = 15 at day 28. Different letters denote significant differences between lead accumulation over the exposure period within individual treatments and differences between treatments within the same exposure period (Bonferroni; p ≤ 0.05).

was still not normally distributed and Levene’s test showed significance, a non-parametric Kruskall–Wallis test was performed. If Kruskall–Wallis test showed significant differences (significant at p ≤ 0.05), post hoc analysis were formed by Mann–Whitney U-tests with Bonferroni correction for multiple comparisons. When data met assumptions, a factorial analysis of variance (ANOVA) test was performed to analyse the effects of time and treatment on whole tissue lead accumulation of H. australis to lead-spiked sediments. Lead accumulation between different tissues within each time and lead accumulation in each tissue over the exposure period of the each treatment was compared by ANOVA or Kruskall–Wallis test. Differences between the treatments for total accumulated lead in the hepatopancreas tissues, lead concentrations in the nuclei + cellular debris, biologically active metal (BAM) and biologically detoxified metal (BDM) from sub-cellular fractionation were compared by Kruskall–Wallis test. When data were significant multiple comparisons were conducted using Mann–Whitney U-tests with Bonferroni correction for multiple comparisons. ANOVA or Kruskall–Wallis test was used to analyse the effect of treatment (n = 3) on the effect measurement variables; TAOC, lipid peroxidation and lysosomal stability. Regression analysis was used to relate: whole organism metal accumulation and lead concentration in sediment; whole organism tissues lead with calcium and magnesium concentrations in tissue at day 28. Relationship between biomarkers (lipid peroxidation and TAOC; destabilized lysosomes % and TAOC; destabilized lysosomes % and lipid peroxidation) were analysed by linear regression. 3. Results 3.1. Exposure-dose relationships 3.1.1. Lead accumulation H. australis accumulated lead in both lead exposure treatments (Fig. 2, Supplementary Table 1a–d). In control organisms, whole organism tissue lead concentration showed no significant differences over the exposure period. After 28 days of exposure H. australis accumulated lead from both the low (2.2 ± 0.2 ␮g/g dry

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Table 1 Linear regression of Hyridella australis whole organism tissue lead with whole organism tissue calcium and magnesium concentrations after 28 days exposure to lead-spiked sediments (< 0.01 (control), 205 ± 9 (low) and 419 ± 16 (high) ␮g/g dry mass). Mean ± SE. n = 15. Treatment

Ca

Control Low High

Mg

Equation

r2

p value

Equation

r2

p value

y = 0.0002x + 0.165 y = 0.003x + 0.7614 y = 0.0005x -0.1714

0.169 0.424 0.837

0.144 0.012 0.000

y = 0.0016x -0.6059 y = 0.0056x -2.7106 y = 0.0033x -0.0454

0.123 0.313 0.932

0.123 0.037 0.000

y = Whole organism tissue lead concentration (␮g/g dry mass). x = Whole organism tissue calcium/magnesium concentration (␮g/g dry mass).

Table 2 Total tissue lead concentration, percentage of total recovered lead from sub-cellular fractionation, lead concentration and percentage of distribution in sub-cellular fractions (nuclei + cellular debris, biologically active and biologically detoxified) in hepatopancreas tissues of Hyridella australis after 28 days exposure to lead-spiked sediments. Total tissue lead and total lead in sub-cellular fractions were expressed in ␮g wet mass. Pb control (< 0.01), Pb low (205 ± 9) and Pb high (419 ± 16) ␮g/g dry mass. Mean ± SE, n = 3.

Total Tissue lead (␮g/g) Total Recovered Lead. (␮g/g) Proportion of total recovered in all fractions (%) Sub-cellular Distribution of Lead Nuclei + Cellular debris (␮g/g) Nuclei + Cellular debris (%) Biologically Active Metal (␮g/g) Biologically Active Metal (%) Biologically Detoxified Metal (␮g/g) Biologically Detoxified Metal (%)

Pb control

Pb low

Pb high

0.28 ± 0.09a 0.27 ± 0.12a 97

0.48 ± 0.24a 0.47 ± 0.15a 97

1.08 ± 0.50b 0.85 ± 0 .43b 79

0.01 ± 0.01a 2 0.02 ± 0.01a 7 0.25 ± 0.1a 91

0.02 ± 0.02a 4 0.06 ± 0.01b 13 0.39 ± 0.18a 83

0.03 ± 0.03a 3 0.12 ± 0.01c 14 0.73 ± 0. 38a 83

Letters denote significant differences for lead concentrations between treatments for total tissue lead and between treatments within the same sub-cellular fraction. Bonferroni; p ≤ 0.05.

mass) and high treatments (4.2 ± 0.1 ␮g/g dry mass) but at levels lower than the lead concentrations in the sediment. At each exposure period (except day 0) lead exposed organisms contained more lead than controls and were in the order high > low > control (Fig. 2). At day 28 accumulated lead concentrations in whole body tissues reflected the sediment lead exposure (p ≤ 0.001; r2 = 0.97) (Fig. 3). Lead accumulation patterns for both lead treatments were similar (Fig. 2). A rapid accumulation of lead was observed from day 0 to day 7 followed by a slight decrease on day 14 and then a slight but not significant increase over the next 14 days. Lead accumulation did not vary with the gender or organism body mass. The accumulation of lead in the different tissues at each time in each treatment was significantly different (Supplementary Table 2a and b) and in the order labial palps > mantle > gill > visceral mass > muscle (Fig. 4). The highest concentration of the lead was always found in the labial palps whatever the length of the exposure period. 3.1.2. Lead uptake versus calcium and magnesium tissue concentrations The accumulation of lead in whole soft body tissues of H. australis exposed to lead contaminated sediment for 28 days

was significantly associated with the tissue calcium and magnesium concentrations (Table 1). In control organisms, lead tissue concentration was more or less similar in all organisms and not related to the tissue calcium and magnesium concentrations (Table 1).

3.1.3. Sub-cellular lead distribution In control and low lead exposure treatments, 97% of lead was recovered while in high exposure lead treatment, 79% of lead was recovered in the fractions (Table 2). A high percentage of lead (83%–91%) was accumulated in the BDM fraction after 28 days of lead exposure. Lead concentration in metal rich granules (MRG) also increased with lead exposure (Fig. 5). In contrast, 13%–32% of lead sequestered in the metallothionein like protein (MTLP) fraction decreased with increased lead exposure (Table 3). The BAM fraction significantly increased with increased lead exposure (p ≤ 0.001). The highest percentage of lead in BAM was in the mitochondrial fraction (Table 3). In both low and high lead treatments, the percentage of lead in lysosomal + microsomal fractions was higher than the heat sensitive protein (HSP) fraction. In control organisms the lysosomal + microsomal and HSP fractions contained the same percentage (15%) of lead (Table 3).

6

Tissue Pb µg/g dry mass

R2 = 0.9736 5

Table 3 Mean percentage of lead in sub-cellular fractions (nuclei + cellular debris; mitochondria, lysosomes + microsomes, heat sensitive proteins – biologically active metals; metal rich granules, heat stable MT like proteins – biologically detoxified metals) of Hyridella australis after 28 days exposure to lead-spiked sediments. Pb control (< 0.01), Pb low (205 ± 9) and Pb high (419 ± 16) ␮g/g dry mass. n = 3.

4 3 2 1 0 0

100

200

300

400

500

Sediment Pb µg/g dry mass Fig. 3. Regression of Hyridella australis whole organism tissue lead concentration against sediment lead concentrations at day 28. Mean ± SE, n = 15.

% of Nuclei + Cellular debris in total % of BAM in total % of BAM in mitochondria % of BAM in lysosomes + microsomes % of BAM in heat sensitive proteins % of BDM in total % of BDM in metal rich granules % of BDM in heat stable MT like proteins

Pb control

Pb low

Pb high

2 7 70 15 15 91 68 32

4 10 64 27 9 86 80 20

3 11 77 17 6 86 87 13

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Labial Palps

Tissue Pb µg/g

14

ns

ns

GilI b

ns a

10

b b

5

b

12

6 Tissue Pb µg/g

16

8 6 4

4 3

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b

b

a

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a

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a

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0 7

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7

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14

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c

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10 9 8 7 6 5 4 3 2 1 0

89

a

c 4

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3

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a a

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1 0 7 Control

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7 Control

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Muscle 3.5

b

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Tissue Pb µg/g

3 2.5 2

1

c

b

1.5

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b b

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a

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0 7 Control

14 Days 21 28 Low High

Fig. 4. Lead concentrations in visceral mass, gill, muscle, mantle and labial palps of Hyridella australis at weekly intervals for 28 days exposure to lead-spiked sediments: control < 0.01; low 205 ± 9; high 419 ± 16 ␮g/g dry mass. Mean ± SE, n = 6 at days 7, 14, 21 and n = 15 at day 28. Different letters denote significant differences between treatments within the same exposure period (Bonferroni; p ≤ 0.05). ns – not significantly different.

3.2. Dose-response relationships Total antioxidant capacity in hepatopancreas cells was significantly reduced with lead exposure relative to the control organisms (Fig. 6A and Supplementary Table 3). The organisms exposed to low concentrations of lead had a significantly higher level of TAOC than those exposed to high concentration of lead. MDA levels increased with lead exposure but were not significantly different between treatments (Fig. 6B). Lysosomal stability in hepatopancreas cells significantly decreased in both lead treatments, relative to the control organisms (Fig. 6C, p ≤ 0.001; Supplementary Table 3). The regression analysis of MDA versus TAOC and destabilized

lysosomes % versus TAOC did not show significant relationships (r2 = 0.03 and r2 = 0.2) and relationships between destabilized lysosomes % and MDA also were not significant (r2 = 0.06). 4. Discussion 4.1. Exposure-dose relationships 4.1.1. Lead accumulation Lead accumulation by H. australis varied with lead exposure and tissue type (Figs. 2 and 4). H. australis accumulated relatively low lead in the body tissues suggesting low bioavailability of

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b

A)

0.7

10

0.6

9

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TOAC mmol/mg protein

Lead µg / g wet mass

0.8

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a

0.3 0.2

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0.1

Low

a a

0 N + CD

HSP LYSO + MITO MTLP MICRO

MRG

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a b

8 7

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6 5 4 3 2 1 0 Control

N+CD = Nuclei + Cellular debris

Biologically Detoxified Lead

MITO = Mitochondria HSP = Heat sensitive proteins LYSO + MICRO = Lysosomes + microsomes

B)

100

MDA nmol/mg protein

Biologically Active Lead

MTLP = Metallothionein like proteins MRG = Metal rich granules

Fig. 5. Lead concentration in the sub-cellular fractions of the hepatopancreas tissues of Hyridella australis exposed to lead-spiked sediments (control < 0.01; low 205 ± 9; high 419 ±16 ␮g/g dry mass) at 28 days exposure. Lead in metal rich granules and metallothionein like proteins are classified as biologically detoxified lead and lead in mitochondria, lysosomes + microsomes and heat sensitive fractions are classified as biologically active lead. Different letters denote significant differences between treatments within the same sub-cellular fraction (Bonferroni; p ≤ 0.05).

Low

a

90 80

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70 60 50 40 30 20 10 0 Control

C)

Low

High

80

Destabilised Lysosomes%

sediment bound lead. Lead has a strong binding affinity to sediments (Benjamin and Leckie, 1981; Gadde and Laitinen, 1974) reducing the bioavailability of lead in the spiked sediments to the benthic biota. In a study where the freshwater bivalve, Unio pictorum was transplanted into the metal contaminated (As, Fe, Ni, Pb) Cecina River basin only moderate amounts of metals were accumulated compared to the elevated concentrations of these metals in the river sediments (Guidi et al., 2010). The whole body tissue data in the present study revealed a significant increase in lead concentration in lead exposed bivalves from day 0 to day 7 compared to the unexposed organisms (Fig. 2). Rapid uptake was followed by a slight decrease of lead accumulation at day 14 in both lead exposed organisms. This would suggest after the initial exposure and uptake of lead, the induction of regulatory mechanisms in the organism occurs, probably via excretion of accumulated lead in faeces or pseudofaeces a process which has been previously measured with cadmium in oysters Crassostrea gigas (Strady et al., 2011). A similar pattern of lead uptake has been observed in the marine bivalve Tellina deltoidalis exposed to lead-spiked sediment (Taylor and Maher, 2013). The present study showed that lead accumulation is different in different tissues/organs and that labial palps accumulated the highest concentrations of lead when compared to other tissues. This may be attributed to the different metabolic activities and the variations of the distribution of granules between tissues of exposed organisms (Canli, 2000; Jones and Walker, 1979). A study on the freshwater bivalve Corbicula fluminea has demonstrated that the distribution of lead is organ specific (Labrot et al., 1999). The high percentage of accumulated lead in labial palps and gills of the exposed organisms in the present study suggests that the majority of lead accumulation occurs via dissolved lead exposure. Exposure of H. australis to dissolved lead probably relates to sediment burrowing activities which may facilitate the release of sediment bound lead into pore water. In this study, the lowest lead accumulation occurred in the less metabolically active muscle tissues of H. australis. Jones and Walker (1979) found that the highest concentrations of essential metals such as Fe, Mn and Zn also accumulated

70

b

b

Low

High

60 50 40 30

a

20 10 0 Control

Fig. 6. Biomarker responses in Hyridella australis exposed to lead-spiked sediments: control < 0.01; low 205 ± 9; high 419 ±16 ␮g/g dry mass. Mean ± SE. Enzymatic biomarkers n = 27. 6A – Total antioxidant capacity (TAOC). 6B – Lipid peroxidation (malondialdehyde [MDA]). 6C – Cellular biomarker (lysosomal membrane stability) n = 9. Different letters denote significant differences between treatments (Bonferroni; p ≤ 0.05).

in gills + labial palps and lowest concentration of these metals in the foot of the freshwater mussel Velesunio ambiguus. They concluded that differences in accumulated metals between different organs largely depended on the distribution of storage granules within this organism. The freshwater mollusc Elliptio complanata also accumulated high concentrations of Pb, Cu, Zn, Fe and Mn in gill and mantle tissues, lowest concentration of these metals in muscle tissues such as foot and adductor muscle and intermediate levels in hepatopancreas tissues (Tessier et al., 1984). Our study confirmed that the analysis of separate tissues is important to identify the major exposure route of metal exposed organisms when they are evaluated as a biomonitor in sediment toxicity assessments. Lead accumulation in various tissues of the freshwater bivalve

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C. fluminea (gill, foot, visceral mass and remaining tissues) was studied when Labrot et al. (1999) were evaluating its use as a sentinel organism for environmental quality survey programmes. Their results also indicated the importance of analysing the specific tissues of a bioaccumulating species, rather than the whole organism metal accumulation, as discussed by (Morgan and Morgan, 1990). 4.1.2. Lead uptake versus calcium and magnesium tissue concentrations Lead accumulation was significantly related to tissue calcium and magnesium concentrations in lead exposed organisms (Table 1). Our results suggest that the uptake of lead may be due to the calcium and magnesium requirements of H. australis. In unionidae bivalves, calcium is required for shell formation and other physiological process and is also sequestered in calcium phosphate granules (Byrne and Vesk, 2000). These granules have the potential to sequester other metals, particularly Pb, Fe, Mg, Mn, Al, Cu, P, S, Ba and Zn (Adams and Shorey, 1998; Adams et al., 1997; Byrne and Vesk, 1997, 2000; Vesk and Byrne, 1999). Markich and Jeffree (1994) investigated the accumulation of divalent nonessential trace metals (Pb, Mn, Cd and Co) in two freshwater unionid bivalves, H. depressa and V. ambiguus. They showed that calcium tissue concentrations could be used to predict the variation of Cd, Co, Cu, Mn, Ni, Pb and U concentration in individual bivalves. 4.2. Sub-cellular lead distribution Most of the lead accumulated in the hepatopancreas tissues of H. astralis was sequestered into the granule fraction and the amount of lead in this fraction increased with lead exposure (Fig. 5 and Table 3). In contrast, the relative proportion of lead sequestered in the MTLP fraction decreased with lead exposure suggesting that the granules play an important role in lead homeostasis and detoxification of the unionid bivalve H. australis. Byrne and Vesk (2000) have also shown the importance of granules in sequestering metals in the tissues of H. depressa. The biologically active metal fraction in the hepatopancreas of exposed organisms in the present study was very low (0.02–0.12 ␮g/g wet mass) relative to the lead accumulated in the biologically detoxified fraction (0.25–0.73 ␮g/g wet mass), therefore, detoxification of lead was actively occurring in H. australis. Sub-cellular partitioning of cadmium in the freshwater bivalve Pyganodon grandis has also shown that cadmium detoxification occurs due to sequestration of cadmium in granules and the resultant low concentration of cadmium in biologically active metal pool in gill tissue (Bonneris et al., 2005; Cooper et al., 2013; Cooper et al., 2010). Bonneris et al. (2005) further emphasized that the constituent elements in the granules of unionid gills act as non-inducible metal sinks at the cellular level. The present study has shown that the concentration of lead in the biologically active pool increased with lead exposure (Table 2). When the accumulation of metals exceeds the detoxification capacity of the organism, non-specific binding of metals by biologically important molecules such as DNA, proteins and cellular organs results and causes deleterious metal-induced toxicity at the cellular level (Brown and Parsons, 1978; Mason and Jenkins, 1995; Perceval et al., 2006).

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increased lead exposure, indicating that oxidative stress increased with increasing lead exposure and sequestration (Fig. 6A). Lead is a non-essential element, therefore, even at small body burdens, impairment of biochemical and cellular pathways were observed. The depletion of antioxidant capacity increased the formation of ROS in lead exposed organisms and this resulted in an increase of MDA in exposed organisms compared to controls (Fig. 6A and B). Previous studies have indicated similar responses. Giguère et al. (2003) showed that increased MDA concentrations in gill tissues of P. grandis resulted from an increase in levels of cytosolic Cd concentrates. Cossu et al. (2000) reported a sharp depletion in antioxidant enzymes (glutathione reductase, selenium-dependent glutathione peroxidase), resulting in reduced glutathione, and an increase in lipid peroxidation in digestive glands and gills of transplanted freshwater bivalve U. tumidus, after 15 days when exposed to different contaminated sediments. Doyotte et al. (1997) found that the depletion of antioxidants and a slight enhancement of lipid peroxidation occurred in gill tissues of U. tumidus after one week of exposure to copper and the fungicide thiram. The present study demonstrated a reduced ability of the bivalve antioxidant system to compensate for an increase in ROS formed on lead exposure and that the balance between ROS production and removal was disrupted. Destabilisation of lysosomal membranes was related to sediment lead exposure (Fig. 6C). As TAOC went down there was a corresponding increase in lipid peroxidation and lysosomal destabilisation (Fig. 6 A–C). Alterations in lysosomal membrane permeability facilitate the leakage of hydrolase enzymes into the internal environment of the cell which will ultimately lead to cell death. Increased lysosomal membrane destabilisation in marine bivalve Mytilus galloprovincialis (Domouhtsidou et al., 2004; Regoli and Orlando, 1993), Anadara trapezia and Tellina deltoidalis (Taylor and Maher, 2012b, 2013) has been shown to be a general response to exposure and uptake of lead. 5. Conclusion This research has demonstrated a significant exposure-doseresponse relationship of H. australis to lead-spiked sediments under laboratory controlled conditions. In response to increasing lead concentrations in the sediments, increased tissue lead dose and biologically active lead burdens were measured in exposed organisms. This resulted in significant impairment of a total antioxidant capacity which may lead to oxidative damage as measured by lipid peroxidation and lysosomal destabilisation. These findings indicate that these enzymatic and cellular biomarkers are useful to evaluate oxidative stress and to assess the effects of lead accumulated from contaminated sediments. Based on these observations, H. australis should be a good bioimonitor for use in sediment risk assessment in freshwater environments. Conflict of Interest The authors declare that there are no conflicts of interest. Acknowledgements

4.3. Dose-response relationships Lead in the biologically active metal pool increased significantly with lead exposure and was considerably higher in the mitochondrial fractions than in the lysosomes + microsomes and HSP fractions (Fig. 5, Table 3). The effect of this metal pool on H. australis was evident from the two enzymatic biomarkers measured in this study. Total antioxidant capacity of organisms decreased with

We wish to thank: Patrick Ceeney, Tom Long and Peter Ogilvie for assistance with water and sediment collection, Jamie Potts with bivalve collection, Maria Byrne, for help with H. australis identification and distribution and Julio Romero for assistance with statistical analysis. We gratefully acknowledge Chitral Siriwardana, Chamoda Siriwardana and Bhagya Siriwardana for their assistance with aquarium setup, labelling and sample preparation.

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Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at http://dx.doi.org/10.1016/j.aquatox. 2014.01.017.

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