The effect of water hardness on the toxicity of uranium to a tropical freshwater alga (Chlorella sp.)

Aquatic Toxicology 60 (2002) 61 – 73

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The effect of water hardness on the toxicity of uranium to a tropical freshwater alga (Chlorella sp.) Amanda L. Charles a,b,*, Scott J. Markich a, Jennifer L. Stauber c, Lou F. De Filippis b a b

En6ironment Di6ision, Australian Nuclear Science and Technology Organisation, PMB 1, Menai NSW 2234, Australia Department of En6ironmental Sciences, Uni6ersity of Technology, Sydney, PO Box 123, Broadway NSW 2007, Australia c Centre for Ad6anced Analytical Chemistry, CSIRO Energy Technology, PMB 7, Bangor NSW 2234, Australia Received 17 May 2001; received in revised form 23 October 2001; accepted 15 November 2001

Abstract Uranium (U) derived from mining activities is of potential ecotoxicological concern to freshwater biota in tropical northern Australia. Few data are available on the effects of water hardness (Ca and/or Mg), which is elevated in U mine wastewaters, on the toxicity and bioavailability of U to freshwater biota, particularly algae. This study determined the effect of water hardness (8, 40, 100 and 400 mg CaCO3 l − 1, added as calcium (Ca) and magnesium (Mg) sulphate) on the toxicity (72 h growth rate inhibition) of U to the unicellular green alga, Chlorella sp., in synthetic freshwater, at constant pH (7.0) and alkalinity (8 mg CaCO3 l − 1), similar in chemical composition to sandy coastal streams in tropical northern Australia. A 50-fold increase in water hardness resulted in a 5-fold decrease (P5 0.05) in the toxicity of U to Chlorella sp. (i.e. the 72 h EC50 increased from 56 to 270 mg U l − 1). Possible explanation for the ameliorative effect of water hardness includes: (i) competition between U and Ca and/or Mg for binding sites on the cell surface; and (ii) a change in U speciation, and hence, bioavailability. Results showed that extracellular (cell-surface) and intracellular U concentrations significantly (PB 0.05) decreased (2 – 5-fold) as water hardness increased from 8 to 400 mg CaCO3 l − 1. Calculation of U speciation using the geochemical model HARPHRQ showed that there were no significant (P \0.05) differences in the predicted speciation (% distribution) of U amongst the four water hardness levels. The reduction in U toxicity with increasing water hardness was most likely due to competition between U and Ca and/or Mg for binding sites on the algal cell surface. The minimum detectable effect concentrations of U were approximately 3 and 24 times higher (at 8 and 400 mg CaCO3 l − 1 hardness, respectively) than the national interim U guideline value (0.5 mg l − 1) for protecting aquatic ecosystems. Overall, the results reinforce the need for a more flexible U guideline based on a hardness-dependent algorithm, which may allow environmental managers to relax the national guideline for U on a site-specific basis. © 2002 Elsevier Science B.V. All rights reserved. Keywords: Water hardness; Toxicity; Uranium; Freshwater; Algae; Speciation

* Corresponding author. Tel.: + 61-2-9717-3331; fax: + 61-2-9717-9260. E-mail address: [email protected] (A.L. Charles). 0166-445X/02/$ - see front matter © 2002 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 6 - 4 4 5 X ( 0 1 ) 0 0 2 6 0 - 0

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1. Introduction Fresh surface waters in tropical northern Australia, typically contain less than 0.2 mg U l − 1 (Markich, 1998; Klessa, 2000). Uranium may exist in freshwater environments in a variety of soluble forms, including the dissolved uranyl ion (UO22 + ) and uranyl complexes with inorganic (e.g. sulphate, carbonate) and organic (e.g. humic and fulvic acids) ligands (Gascoyne, 1992). Redox potential, pH and the presence of various complexing ligands influence the chemical behaviour and speciation of U in aquatic systems (Bernhard et al., 1998). Due to the high solubility and mobility of U in aquatic systems (Morse and Choppin, 1991), the contamination of fresh surface waters from U mine wastewaters is a potential hazard to freshwater biota in tropical northern Australia (Markich and Camilleri, 1997). Calcium and Mg concentrations in U mine wastewaters are typically 25– 250-fold higher than receiving waters. However, few data are available on the effects of water chemistry variables, such as true water hardness (Ca and/or Mg concentration), on U toxicity and bioavailability to freshwater biota, particularly algae. Knowledge of the relationship between water chemistry variables (e.g. water hardness) and U toxicity is needed for predicting the potential ecotoxicological effects of U in freshwater ecosystems. Several studies (Parkhurst et al., 1984; Poston et al., 1984; Barata et al., 1998; Riethmuller et al., 2000) have reported that the toxicity of U to freshwater biota (fish, crustaceans and cnidarians) typically decreases as water hardness increases. The study by Riethmuller et al. (2000) is the only one that has not confounded the effects of true water hardness (Ca and/or Mg concentration) with alkalinity and pH. The authors determined the effect of true water hardness and alkalinity independently, at a constant pH (6.0), on the toxicity of U to a tropical freshwater green hydra (Hydra 6iridissima). They found that a 50-fold increase in hardness (Ca and Mg concentration) halved the toxicity of U to H. 6iridissima (i.e. the 96 h EC50 increased from 114 mg U l − 1 at 6.6 mg CaCO3 l − 1 to 219 mg U l − 1 at 330 mg CaCO3

l − 1). Riethmuller et al. (2000) postulated that U toxicity was most likely reduced by Ca2 + and/or Mg2 + ions competing with UO22 + for binding sites at the cell membrane surface, without directly altering U speciation in solution. Studies that have investigated the effects of true water hardness (Ca and/or Mg) on the toxicity of trace metals to unicellular algae (Rai et al., 1981a; Rai and Raizada, 1985; Issa et al., 1995) have generally indicated that metal toxicity is reduced by Ca and/or Mg, due to either competition for cell surface binding sites or solution speciation changes resulting from metal complexation and/or coprecipitation. Rai et al. (1981b) suggested that total metal concentrations in solution are decreased through the formation of insoluble carbonate and/or hydroxide complexes, leading to a reduction in metal bioavailability and toxicity. This typically occurs when increased Ca and/or Mg concentrations are associated with concomitant increases in pH (7–9), and hence, increased carbonate and hydroxide concentrations. The influence of true water hardness on the toxicity of U to freshwater algae has not been investigated. Franklin et al. (2000) report the only available data on the toxicity (growth rate inhibition) of U to freshwater algae. They found that the 72 h EC50 for a tropical Chlorella sp. (a different isolate to that used in the present study) ranged from 44 mg U l − 1 at pH 6.5 to 78 mg U l − 1 at pH 5.7, at a hardness of 4 mg CaCO3 l − 1. Increasing the pH from 5.7 to 6.5 increased the formation of polymeric uranyl species, including (UO2)3(OH)+ 5 , (UO2)3(OH)− and (UO2)2(OH)3CO− 7 3 , with the proportion of free uranyl ion (UO22 + ) being insignificant (B 1%) (Franklin et al., 2000). Increased U toxicity at pH 6.5 was due to increased binding of U to the algal cell surface, and higher concentrations of intracellular U relative to pH 5.7. The aim of this study was to determine whether true water hardness (8, 40, 100 and 400 mg CaCO3 l − 1, added as Ca and Mg sulphate) affects the toxicity (72 h growth rate inhibition) of U to a tropical freshwater alga, Chlorella sp. (isolate 12), in a synthetic freshwater, at a constant pH and alkalinity, similar in chemical composition to sandy coastal streams in tropical northern Aus-

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tralia. Uranium bound to the algal cell surface (i.e. extracellular U) and intracellular U were also measured to determine whether changes in U toxicity at different water hardness levels were due to competition between U and Ca and/or Mg at the algal cell surface. In addition, U speciation in solution was calculated to assess whether speciation changes could explain changes in U toxicity with varying water hardness.

2. Materials and methods

2.1. General analytical procedures All solutions, including the culture medium, were prepared with high purity Milli-Q water ( B 18 MV cm − 1 resistivity). To avoid metal contamination, all laboratory ware was soaked in 10% (v/v) nitric acid (HNO3) overnight and thoroughly rinsed with Milli-Q water prior to use. All reagents used were analytical grade, except for ultra-pure HNO3 (Normatom).

2.2. Algal stock cultures and preparation of inoculum All toxicity tests were performed using the freshwater tropical unicellular green alga, Chlorella sp. (isolate 12) isolated from Lake Aesake, Strickand River, Papua New Guinea. It was not possible to identify the Chlorella isolate to species level (W. Vyverman, pers. commun.). Chlorella sp. was cultured axenically in a 1/5 strength Jaworki’s medium (Thompson et al., 1988) on a 12-h light:12-h dark cycle (Philips TL 40W cool white fluorescent lighting, 120 mmol photons PAR m − 2 s − 1) at 2791 °C. Exponentially growing Chlorella sp. cells, harvested from a 4- to 5-day-old stock culture, were centrifuged in 30 ml glass centrifuge tubes at 2500 revolutions per minute (rpm) at 20 °C in a Jouan CR4.11 refrigerated centrifuge for 7 min. The centrifugation and washing procedure was repeated three times to remove the high nutrient culture medium, which could potentially ameliorate toxicity by complexing trace metals such as U (Stauber and Florence, 1989).

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2.3. Toxicity test procedure The toxicity of U to Chlorella sp. was assessed using standard protocols described in Stauber et al. (1994) and Franklin et al. (2000), and summarised in Table 1. The test medium used was a synthetic freshwater (Table 2), lacking in organic chelating agents which may complex metals and reduce their toxicity to freshwater biota (Stauber and Florence, 1989). Based on preliminary experiments, four water hardness levels were selected—8, 40, 100 and 400 mg CaCO3 l − 1, which are representative of fresh surface waters from tropical northern Australia (Riethmuller et al., 2000). The hardness of the synthetic freshwater (8 mg CaCO3 l − 1) was increased by adding Ca (CaSO4 · 2H2O) and Mg (MgSO4 · 7H2O) as sulphates, rather than nitrates, to prevent the confounding effects of nitrate added as a nutrient for algal growth. Water hardness was calculated from measured Ca and Mg concentrations using the method described in Table 1 Summary of the toxicity test protocol for Chlorella sp. 1. Test type 2. Temperature 3. Light quality

Static 27 91 °C Cool white fluorescent lighting 4. Light intensity 115–125 mmol photons PAR m−2 s−1 5. Photoperiod 12-h light:12-h dark cycle 6. Test chamber size 250 ml 7. Test solution volume 50 ml 8. Renewal of test solutions None 9. Age of test organisms 4–5 days 10. Initial cell density in 2–4×104 cells ml−1 test chambers 11. Number of replicate 3 chambers/concentration 12. Shaking rate Once daily by hand 13. Test medium Synthetic freshwater 14. Metal concentrations Minimum of five and a control 15. Test duration 72 h 16. pH 7.0 90.3 17. Test endpoints Growth (cell division) rate 18. Test acceptability Control cell division rate 1.3 90.3 doublings day−1; CV in controls B20%

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64 Table 2 Composition of synthetic freshwater Parameter

Concentration (mg l−1)a

Na K Ca Mg Cl SO4 HCO3 NO3 PO4 U Organic carbon

8.9 9 0.7 0.36 90.03 1.690.1b 1.0 9 0.1b 0.289 0.02 7.89 0.2 8.79 0.3 15.09 0.3 0.149 0.10 0.159 0.05c B0.2

Mean 9S.D.; n =12. Hardness, 8 mg CaCO3 l−1 (i.e. 2.47 [Ca]+4.11 [Mg]) (APHA et al., 1998). c Concentration expressed in mg l−1. a

b

APHA et al. (1998). At each hardness level, a 1:1 mole ratio of Ca and Mg was used. Range finding tests were initially carried out using a wide range of U concentrations at each water hardness level. Concentrations for the definitive tests (0.1– 700 mg U l − 1) were selected with the aim of producing no effect at the lowest concentration of U and 100% effect at the highest concentration of U, with a range of partial effects at the intermediate U concentrations. A minimum of three definitive toxicity tests was performed at each water hardness level. The test containers used for all toxicity tests were acid-washed 250 ml Erlenmeyer flasks coated with a silanising solution, Coatasil (Ajax), to reduce U adsorption to the flask walls. To 54.5 ml of filtered synthetic freshwater, 0.5 ml of 26 mM sodium nitrate (15 mg NO3 l − 1) and 0.05 ml of 1.3 mM potassium dihydrogen phosphate (0.15 mg PO4 l − 1) were added as nutrients to maintain exponential growth of the algae over the 72 h test. Controls, together with five test concentrations, each in triplicate, were inoculated with 2– 4 ×104 cells ml − 1 of prewashed Chlorella cells. Flasks were incubated at 2791 °C under the same conditions used to culture the algae (Section 2.2). All flasks were swirled once daily by hand to minimise CO2 limitation. The pH of the test solutions was measured in one replicate flask from each

treatment on the initial and final day of the test. The cell density and mean cell size in each flask were determined daily using a Coulter Multisizer II Particle Analyser (70 mm aperture) with background correction.

2.4. Chemical analyses A 174 mg l − 1 U stock solution was prepared from analytical grade uranyl sulphate (UO2SO4 · 3H2O) in Milli-Q water and refrigerated at 4 °C until required. Immediately after the addition of U to the test flasks, two 5 ml sub-samples were taken from each flask. One sub-sample was used for the determination of anions, while the other was acidified (pHB 2) for metal analysis. Non-acidified sub-samples were analysed for nitrate (NO3), phosphate (PO4), chloride (Cl) and sulphate (SO4) using ion chromatography (Dionex). Bicarbonate concentrations were measured using an automated potentiometric titration facility (Brown et al., 1992). Organic carbon was measured using photocatalytic oxidation (Matthews et al., 1990). All water samples were refrigerated in darkness at 4 °C prior to analysis. Uranium was measured using inductively coupled plasma mass spectrometry (ICPMS) (Hewlett–Packard 4500). Calcium and Mg concentrations, required for calculating the hardness of the test waters (using the APHA et al., 1998 method), as well as Na and K, were measured using inductively coupled plasma atomic emission spectrometry (ICPAES) (Varian Vista). A multimetal calibration standard and a reagent blank were analysed with every ten samples to monitor signal drift. For all samples and all metals, the signals changed by less than 8%, but typically 3–5%. Where ICPMS was used, gallium and rhenium were employed as internal standards to correct for non-spectral interferences. A duplicate sample and standard reference materials (National Institute of Standards and Technology, Trace elements in water 1643c and National Research Council of Canada, Riverine water for trace metals SLRS-2) were analysed with each batch of ten samples to determine method precision and accuracy, respectively. The

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mean concentrations of U in the standard reference materials were always within their certified concentration ranges. The mean coefficient of variation (CV) for duplicate samples was 7%.

2.5. Intracellular and extracellular uranium determination To investigate the mechanisms by which water hardness may alter U toxicity, intracellular and extracellular U concentrations were determined after 72 h exposure to U at the lowest (8 mg CaCO3 l − 1) and highest (400 mg CaCO3 l − 1) water hardness levels. Based on results from the definitive toxicity tests, four U concentrations (15, 60, 300 and 500 mg U l − 1) and a control (0.1 mg U l − 1) were prepared in triplicate at both water hardness levels. At the completion of the toxicity tests (72 h) the remaining test medium was retained after cell counting. In a Class 100 clean room, 40.09 0.1 g of solution from each flask was transferred into acid washed 50 ml teflon tubes. Samples were centrifuged at 3500 rpm at 20 °C in a Jouan CR4.11 refrigerated centrifuge for 20 min. The supernatant (20 ml) was pipetted into polycarbonate vials and acidified (pHB 2) with 40 ml of concentrated HNO3 (Normatom). These samples were analysed for dissolved U by ICPMS (‘dissolved U’ fraction). The algal pellet was re-suspended in 20 ml of 0.02 M ethylenediaminetetraacetic acid (EDTA) and shaken for approximately 30 s. EDTA has been shown to remove the U bound to the outside of the cells (Franklin et al., 2000) without disrupting cell integrity. These samples were centrifuged for a further 20 min and 20 ml of supernatant retained for U analysis by ICPMS (‘extracellular U’ fraction). An EDTA blank was also analysed. The remaining 0.5 ml algal pellet was dried, 2 ml of concentrated HNO3 was added, and the samples allowed to stand for 30 min. The samples were acid-digested in a microwave oven (Panasonic Model NN-FS 769 WA) on a power setting of 90 W for 5 min. After cooling to room temperature, the samples were made up to 20 ml with Milli-Q water and analysed for U by ICPMS (‘intracellular U’ fraction).

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The binding constant (i.e. distribution coefficient, Kd) of U to the algal cell surface was calculated from a plot of extracellular U concentration (× 10 − 6 ng cell − 1) versus equilibrium dissolved U concentration (mg l − 1) after 72 h, for both 8 and 400 mg CaCO3 l − 1. The slope of the linear portion of the curve was equivalent to the Kd value. To calculate the mass balance of U between solution, cells and the flask surface, the emptied flasks were filled with 50 ml of 0.03 M HNO3 and left overnight. Uranium was analysed using ICPMS to obtain the ‘flask bound U’ fraction.

2.6. Geochemical speciation modelling of uranium The thermodynamic geochemical speciation code HARPHRQ (Brown et al., 1991) was used to calculate the speciation of U in the test solutions. The input parameters for HARPHRQ were based on physico-chemical data measured in the test solutions (pH and total metal and anion concentrations; Table 2). Stability constants were taken from Markich and Brown (1999).

2.7. Statistical analysis The algal growth (cell division) rate was calculated for each flask over 72 h. For growth rate, simple linear regression analysis was used to fit log10 cell density vs. time (h) for each flask — the regression slope was equivalent to the cell division rate per h (m) for each treatment. Daily doubling times were calculated by multiplying this value m×24× 3.32 (i.e. 2.303/ln 2; a constant). Concentration–response curves were fitted using a logistic regression model (Seefeldt et al., 1995). Using the model, the EC50 (i.e. the concentration of U giving 50% inhibition of algal growth rate over 72 h compared with the controls) and its 95% confidence interval (CI) were calculated at all water hardness levels. Measured, rather than nominal, concentrations of U were used. The 10% bounded effect concentration (BEC10), an alternative to the no observed effect concentration (NOEC), was estimated using the approach described by Hoekstra and van Ewijk (1993). The BEC10 has been defined as the highest concentra-

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tion, with 95% confidence, at which the biological effect does not exceed 10% of the observed effect. The minimum detectable effect concentration (MDEC), an alternative measure to the lowest observed effect concentration (LOEC), was estimated using the approach described by Ahsanullah and Williams (1991). The MDEC was calculated from a regression model and is defined as the U concentration at which the response becomes significantly (P 5 0.05) lower than the control. The advantages of using a regressionbased model to estimate low toxic effects (e.g. MDEC), rather than a hypothesis-testing approach (e.g. LOEC), are described elsewhere (Bruce and Versteeg, 1992; Moore and Caux, 1997). Differences in the predicted speciation of U amongst the water hardness levels were tested using the generalised F-test (Ratkowsky, 1990). Tests for significance between treatments (e.g. comparison of intracellular and extracellular U concentrations both between and within water hardness levels) were carried out using a pairwise t-test (Sokal and Rohlf, 1995). Significance levels were tested at the P= 0.05 level.

3. Results A summary of measured physico-chemical data for the synthetic freshwater (i.e. 8 mg CaCO3 l − 1) is given in Table 2. For all experiments, the mean measured concentrations of U in the test waters were all within 20%, but typically 10%, of their nominal concentrations. Similarly, Na, K, Ca, Mg, SO4, Cl, HCO3, NO3 and PO4 were always within 10% of their nominal concentrations. The pH in all test solutions remained constant at pH 7.0 90.3. The concentration– response relationships for Chlorella sp. exposed to U for 72 h at 8, 40, 100 and 400 mg CaCO3 l − 1 hardness are shown in Fig. 1a–d, respectively. Values for each growth test endpoint (BEC10, MDEC and EC50) are also shown. All mean growth rates (doublings day − 1) fell within the range 1.0– 1.6, with CVs B20%, indicating test acceptability. There was no significant (P\0.05) difference in the control growth

rates of Chlorella sp. amongst the different water hardness levels (i.e. 1.29 0.2, 1.69 0.3, 1.59 0.2 and 1.290.3 doublings day − 1 at 8, 40, 100 and 400 mg CaCO3 l − 1). The growth rate of Chlorella sp. decreased with increasing U concentration at all water hardness levels tested. Uranium toxicity decreased with increasing water hardness. Increasing the hardness of the test water from 8 to 400 mg CaCO3 l − 1 produced a five-fold reduction in U toxicity over 72 h (i.e. EC50 values increased from 56 to 270 mg U l − 1; Fig. 1). The BEC10 and MDEC values also increased with increasing water hardness (Fig. 1). To explain the significant (P5 0.05) decrease in the toxicity of U to Chlorella sp. with increasing water hardness, extracellular and intracellular U were determined at the completion of the 72 h growth inhibition tests for both the lowest (8 mg CaCO3 l − 1) and highest (400 mg CaCO3 l − 1) water hardness. Previous work has shown that metals, such as Cu and U, may increase the cell size of unicellular algae (Franklin et al., 2001). Therefore, the effect of U on Chlorella sp. cell size was also determined in each bioassay. There was no significant (P\ 0.05) difference in the median cell diameter (mm) of Chlorella sp. between the 8 and 400 mg CaCO3 l − 1 hardness levels, after 72 h exposure to U. However, the median cell diameter (mm) significantly (P5 0.05) increased from 2.7 to 3.6 mm at 8 mg CaCO3 l − 1, and 2.8 to 3.2 mm at 400 mg CaCO3 l − 1, as the U concentration increased. Since the median cell size of Chlorella sp. increased with increasing U concentration, it was necessary to calculate intracellular U in terms of cell volume (mm3), rather than per cell. Similarly, extracellular (surface-bound) U was calculated in terms of cell surface area (mm2). Concentrations of intracellular and extracellular U at 8 and 400 mg CaCO3 l − 1, at varying U concentrations, are shown in Table 3. At both water hardness levels, the concentrations of extracellular and intracellular U increased with increasing U concentration. Increasing the water hardness of the test solutions from 8 to 400 mg CaCO3 l − 1 substantially reduced both intracellular and extracellular U in Chlorella sp. A larger proportion of the total U was located extracellu-

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Fig. 1. Growth rate of Chlorella sp. exposed to U for 72 h at (a) 8; (b) 40; (c) 100; and (d) 400 mg CaCO3 l − 1 water hardness. Data points represent the mean 9 95% confidence intervals, from at least three definitive experiments at each water hardness.

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Table 3 Intracellular and extracellular (surface-bound) U for Chlorella sp. at 8 and 400 mg CaCO3 l−1 water hardness after 72 h growth Hardness (mg CaCO3 l−1)

Nominal U (mg l−1)

Measured U (mg l−1)

Intracellular U (×10−6 ng mm−3)

Extracellular U (×10−6 ng mm−2)

Extra:Intra U

8

15 60 300 15 60 300 500

12 53 270 17 62 310 520

0.60 9 0.10 0.82 90.15 4.6 9 0.1 0.07 9 0.02 0.10 9 0.02 0.15 90.02 1.1 9 0.4

0.41 90.03 0.80 90.09 15 9 1 0.03 9 0.02 0.22 9 0.04 0.36 90.07 5.0 9 2.5

0.68 9 0.16 0.98 9 0.28 3.3 9 0.3 0.43 9 0.41 2.2 9 0.8 2.4 9 0.8 4.5 9 4.0

400

Values are reported as the mean 9 S.D.; intracellular and extracellular U concentrations for controls (0.1 mg U l−1) were several orders of magnitude lower than test U concentrations.

larly at both water hardness levels (Table 3 Fig. 2a), except at the lowest U concentration (15 mg U l − 1). Intracellular U concentrations decreased with increasing water hardness by a factor of 9, 8 and 30 at 15, 60 and 300 mg U l − 1, respectively (Table 3 Fig. 2b). The ratio of extracellular/intracellular U increased with increasing U concentration (Table 3). No significant (P \ 0.05) differences in the 72 h extracellular/intracellular ratios were observed between the 8 and 400 mg CaCO3 l − 1 hardness waters. The Kd values (expressed on a per cell basis) were 2× 10 − 9 and 3 × 10 − 10 ng cell − 1 at 8 and 400 mg CaCO3 l − 1 hardness, respectively. Less U partitioned from solution to the algal cell surface at the highest water hardness, in agreement with the extracellular U and toxicity results, which resulted in U being less toxic at the highest hardness value. Similar Kd values (1− 3× 10 − 9 ng cell − 1) have been found for another sensitive tropical Chlorella sp. exposed to U in a synthetic freshwater (4 mg CaCO3 l − 1 hardness) (N. Franklin, pers. commun.). The mass balance for U in each treatment was \ 90%. The majority of the U at the end of the test was present in solution (ca. 75%), with 10% bound to the cell surface. A moderate amount of U (ca. 15%) was adsorbed to the walls of the flasks, despite silanisation. The intracellular and extracellular U fractions each represented less than 5% of the total U added. The predicted speciation (% distribution) of U

at 8, 40, 100 and 400 mg CaCO3 l − 1 hardness (pH 7.0) is given in Fig. 3. No significant (P\ 0.05) differences were found in the speciation of U amongst the four water hardness levels. The major U species at pH 7.0 for all hardness lev2− els were (UO2)2(OH)3CO− and 3 , UO2(CO3)2 UO2CO3. As the concentration of U increased, the percentages of UO2(CO3)22 − , UO2CO3, deUO2OH+, UO2(OH)2, and UO2(OH)− 3 creased, whilst the percentage of (UO2)2(OH)3CO− 3 increased (Fig. 3).

4. Discussion This study provides the first data on the effects of water hardness on the toxicity of U to a tropical freshwater alga. A 50-fold increase in water hardness (from 8 to 400 mg CaCO3 l − 1) substantially reduced the toxicity of U (up to 10-fold) to Chlorella sp. (Fig. 1). These results are consistent with those of previous studies using other freshwater biota, where U toxicity was found to decrease with increasing water hardness. (Parkhurst et al., 1984; Poston et al., 1984; Barata et al., 1998; Riethmuller et al., 2000). In the only other comparable study, Riethmuller et al. (2000) found that a 50-fold increase in water hardness (from 6.6 to 330 mg CaCO3 l − 1) reduced the toxicity of U to a tropical freshwater hydra (H. 6iridissima) by a factor of two (i.e. the EC50 increased from 114 to 219 mg U l − 1).

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Parkhurst et al. (1984) found that a six-fold increase in water hardness (from 35– 208 mg CaCO3 l − 1) decreased the acute (96 h) toxicity of U to brook trout (Sal6elinus fontinalis) by a factor of four. Poston et al. (1984) reported that the acute (48 h) toxicity of U to Daphnia magna decreased 6-fold as water hardness increased from 70 to 195 mg CaCO3 l − 1. Similarly, Barata et al. (1998) found that an increase in water hardness from 90 to 180 mg CaCO3 l − 1 resulted in a three-fold reduction in the acute toxicity (48 h) of U to D. magna. However, in the latter three studies, the effect of true water hardness (Ca and/or Mg) was confounded to some extent by changing alkalinity and pH. Potential mechanisms to explain the reduction in the toxicity of U to Chlorella sp. with increasing water hardness include: (a) changes in the speciation and bioavailability of U due to increased concentrations of Ca and/or Mg, and/or

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(b) competition between U and Ca and/or Mg at cell surface binding sites. Modelling with HARPHRQ showed that the speciation (% distribution) of U in solution did not significantly (P\ 0.05) change with increasing water hardness (Fig. 3). This suggests that U speciation changes are unlikely to be responsible for the observed decrease in U toxicity with increased water hardness. Further evaluation of U speciation could be achieved through chemical measurement techniques, such as the use of time-resolved laser induced fluorescence spectroscopy (Markich, 2002). Thus, the reduction in U toxicity with increased water hardness was most likely due to competition between U and Ca and/or Mg for binding sites on the algal cell surface. Both extracellular and intracellular U concentrations were lower at the highest water hardness (400 mg CaCO3 l − 1), indicating that the decreased toxicity of U to Chlorella sp. was primarily due

Fig. 2. Extracellular (a) and intracellular (b) U concentrations for Chlorella sp. at 8 and 400 mg CaCO3 l − 1 water hardness.

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Fig. 3. The predicted speciation (% distribution) of U at 8, 40, 100 and 400 mg CaCO3 l − 1 water hardness (pH 7.0). Ionic strength ranged from 0.002 M at 8 mg CaCO3 l − 1 to 0.008 M at 400 mg CaCO3 l − 1.

to a decrease in U binding at the cell surface (Fig. 2a), leading to a decreased uptake of U into the cells (Fig. 2b). Calcium has been shown to have a greater protective effect than Mg on the toxicity of trace metals to freshwater organisms (Carroll et al., 1979; Part et al., 1985; Markich and Jeffree, 1994; Welsh et al., 2000). For example, Carroll et al. (1979) reported that 7-fold more Mg was needed to exhibit the same ameliorative effect as Ca on the toxicity of Cd to brook trout (S. fontinalis). The stronger protective effect of Ca, rather than Mg, was also observed by Welsh et al. (2000), who found that increasing the Ca:Mg ratio within a standard hardness (40 mg CaCO3 l − 1) substantially ameliorated the toxicity of Cu to rainbow trout (Oncorhynchus mykiss) and chinook salmon (O. tshawytscha). Markich and Jeffree (1994) suggested that only in surface waters where the con-

centration of Mg considerably exceeds that of Ca, will the joint hardness (Ca+Mg) contribute to amelioration of metal toxicity. Further experimental work to define the individual protective effects of Ca and Mg on the toxicity of U to Chlorella sp. is proposed. The tropical Chlorella sp. used in this study was sensitive to U in a synthetic freshwater similar in chemical composition to sandy coastal streams in tropical northern Australia. Comparing the sensitivity of Chlorella sp. with other test organisms (Table 4) to U is difficult due to the varying test conditions (e.g. differences in pH, hardness, conductivity). The only data reported on the toxicity of U to a freshwater alga (c.f. a different species of Chlorella) was obtained at pH 6.5 (Franklin et al., 2000), rather than the pH 7.0 used in this study. As U bioavailability to algae is known to change with increasing pH (Franklin et al., 2000), it is difficult to directly compare the two studies. However, the Chlorella sp. used in this study appears to be the more sensitive of the two. The only other chronic test data available are for the green hydra (H. 6iridissima) (Markich and Camilleri, 1997; Riethmuller et al., 2000). The BEC10 values for H. 6iridissima (14–47 mg U l − 1) were similar to the BEC10 value reported by Franklin et al. (2000) (11 mg U l − 1) using another tropical Chlorella sp. at similar pH and water hardness (Table 4). However, the green hydra was an order of magnitude less sensitive to U than the Chlorella sp. used in this study (Table 4). The data for a freshwater mussel (Velesunio angasi ) and a fish (Mogurnda mogurnda) were based on acute toxicity tests with U (Table 4), and were found to be less sensitive to U than the chronic tests with algae and the green hydra. An interim guideline for U of 0.5 mg l − 1 was proposed in the recent revision of the Australian and New Zealand water quality guidelines (ANZECC and ARMCANZ, 2000). Using the data from Table 4, based on studies carried out in synthetic freshwater at different hardness levels and pH, a U guideline of 0.7 mg U l − 1 was calculated with the same statistical procedure used to derive the interim U guideline (ANZECC and ARMCANZ, 2000). This value is only slightly higher than those measured at minimally dis-

72

96

(Chlorella sp.12)

Cnidaria Green hydra (Hydra 6iridissima)

96

48

a 95% confidence interval. All studies carried out in synthetic freshwater.

Chordata ( fish) Purple-spotted Gudgeon (Mogurnda mogurnda)

Mollusca Mussel (Velesunio angasi )

48

(Chlorella sp.12)

96

72

Exposure time (h)

Algae (Chlorella sp.)

Species

BEC10 LC50 BEC10 LC50 BEC10 LC50

BEC10 EC50 BEC10 EC50

BEC10 EC50 BEC10 EC50 BEC10 EC50

BEC10 EC50 BEC10 EC50 BEC10 EC50 BEC10 EC50 BEC10 EC50

Endpoint

Table 4 Comparative toxicity of U to Australian tropical freshwater biota

6.6

6.0 330

4

4

6.0

6.0

4

4

330

6.6

400

8

400

8

2–4

Hardness (mg CaCO3 l−1)

5.0

6.0

6.0

7.0

7.0

6.5

pH

1270 1570 (1510–1630) 900 1965 (1600–2325) 860 1770 (1570–1970)

92 117 (114–120) 416 634 (606–662)

14 114 (107–121) 47 219 (192–246) 56 108 (102–114)

11 44 (39–49)a 0.9 23 (15–31) 3.5 230 (200–260) 0.7 56 (52–60) 4.5 270 (230–300)

U toxicity (mg U l−1)

Riethmuller et al. (2000)

Markich and Camilleri (1997)

Markich et al. (2000)

Markich and Camilleri (1997)

Riethmuller et al. (2000)

This study

Charles (2000)

Franklin et al. (2000)

Reference

A.L. Charles et al. / Aquatic Toxicology 60 (2002) 61–73 71

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A.L. Charles et al. / Aquatic Toxicology 60 (2002) 61–73

turbed sites (0.1–0.2 mg l − 1) in northern Australia (Markich, 1998; Klessa, 2000). Given the demonstrated hardness dependence of U toxicity to Chlorella sp. reported in this study, together with H. 6iridissima from a previous study (Riethmuller et al., 2000), it seems possible that a guideline for U based on a hardness algorithm, similar to that proposed for Cd, Cr(III), Cu, Ni, Pb and Zn (ANZECC and ARMCANZ, 2000), may be calculated with further affirmative data.

Acknowledgements The authors would like to thank Merrin Adams, Monique Binet and Natasha Franklin (CSIRO Energy Technology), and Henry Wong, Atun Zawadski and David Hill (ANSTO Environment Division) for technical assistance.

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