Effect of major cations (Ca 2+ , Mg 2+ , Na + , K + ) and anions (SO 42

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The effect of the counter-anion on a cation's effect on transparent disposable ..... If the uncharged NiSO4 In fact, the...

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Environmental Toxicology and Chemistry, Vol. 32, No. 4, pp. 810–821, 2013 # 2013 SETAC Printed in the USA DOI: 10.1002/etc.2116

EFFECT OF MAJOR CATIONS (Ca2þ, Mg2þ, Naþ, Kþ) AND ANIONS (SO42, Cl, NO 3 ) ON Ni ACCUMULATION AND TOXICITY IN AQUATIC PLANT (LEMNA MINOR L.): IMPLICATIONS FOR Ni RISK ASSESSMENT YAMINI GOPALAPILLAI,*yz BEVERLEY HALE,y and BERNARD VIGNEAULT§ ySchool of Environmental Sciences, University of Guelph, Guelph, Ontario, Canada zCANMET Mining, Natural Resources Canada, Ottawa, Ontario, Canada §Geological Survey of Canada, Natural Resources Canada, Ottawa, Ontario, Canada (Submitted 22 March 2012; Returned for Revision 28 August 2012; Accepted 13 November 2012) Abstract— The effect of major cation activity (Ca2þ, Mg2þ, Naþ, Kþ) on Ni toxicity, with dose expressed as exposure (total dissolved Ni

concentration NiTot) or free Ni ion activity (in solution Ni2þ), or as tissue residue (Ni concentration in plant tissue NiTiss) to the aquatic plant Lemna minor L. was examined. In addition, Ni accumulation kinetics was explored to provide mechanistic insight into current approaches of toxicity modeling, such as the tissue residue approach and the biotic ligand model (BLM), and the implications for plant Ni risk assessment. Major cations did not inhibit Ni accumulation via competitive inhibition as expected by the BLM framework. For example, Ca2þ and Mg2þ (sulfate as counter-anion) had an anticompetitive effect on Ni accumulation, suggesting that Ca or Mg forms a ternary complex with Ni–biotic ligand. The counter-anion of the added Ca (sulfate, chloride, or nitrate) affected plant response (percentage of root growth inhibition) to Ni. Generally, sulfate and chloride influenced plant response while nitrate did not, even when compared within the same range of Ca2þ, which suggests that the anion dominated the observed plant response. Overall, although an effect of major cations on Ni toxicity to L. minor L. was observed at a physiological level, Ni2þ or NiTot alone modeled plant response, generally within a span of twofold, over a wide range of water chemistry. Thus, consideration of major cation competition for improving Ni toxicity predictions in risk assessment for aquatic plants may not be necessary. Environ. Toxicol. Chem. 2013;32:810–821. # 2013 SETAC

Keywords—Ni toxicity

Ni accumulation

Lemna minor

Biotic ligand model

Ni risk assessment

the TRA tends to be more useful for organic compounds than metals, since the effective internal dose at the sites of toxic action is more complex for metals, often because of essentiality. A previous study suggested that when an element is essential, such as Ni2þ, detoxification mechanisms may not be initiated immediately, as would occur for exposure to a nonessential element, such as Cd2þ [9]. Nonetheless, in complex media, such as mining-impacted freshwaters, the difficulties of estimating bioavailable external dose may outweigh the disadvantages of the TRA. Hence, more research on the applicability of the TRA for metals is warranted. Exposure-based (e.g., BLM) or tissue residue–based (e.g., TRA) metal toxicity will benefit from a greater understanding of mechanisms of trace metal uptake. The BLM is based on the general assumption of competitive inhibition between cations, where the inhibitor (e.g., major cation) binds only to the site of toxic action or biotic ligand, thereby competing with the metal of interest [2]. In this case, high concentrations of the metal can overcome the competitor effect. Although not considered in the general BLM framework, when more than one type of binding site is present, other types of competitive interactions, such as noncompetitive, anticompetitive (also known as uncompetitive), and mixed-competitive interactions, may exist [10]. In such a case, it is possible to observe increased accumulation of the metal of interest in the presence of the competing cation or partial inhibition at high competitor concentrations [10]. In noncompetitive interaction, the added cation can bind to the free biotic ligand and the metal–biotic ligand complex. In this case, no amount of metal addition can completely overcome the competitor effect. In an anticompetitive interaction, the competing cation binds only to the metal–biotic ligand complex,

INTRODUCTION

The ability to predict the toxicity of a metal is an important tool for establishing relevant water-quality guidelines and for regulating industrial emissions. The U.S. Environmental Protection Agency has adopted the biotic ligand model (BLM) as the regulatory tool to establish water-quality guidelines for copper [1]. The BLM framework is used to predict site-specific metal toxicity to an organism by considering metal complexation and speciation in solution as well as interaction with major cations at the organism–water interface [2]. Metal speciation is defined as the occurrence of a metal in separate identifiable forms. The free metal ion M2þ is generally accepted as one of the most bioavailable forms, expressed as activity rather than concentration. The BLM has been successfully implemented for fish and invertebrates for several metals. In contrast, Ni is the only metal for which the BLM has incorporated an aquatic plant for the purposes of Ni risk assessment [3]. It is plausible that plants may not follow competitive inhibition of trace metal uptake by major cations, as is the case for some aquatic organisms [4–6]. The tissue residue approach (TRA), in which the bioaccumulated metal in tissue becomes the dose metric in the dose– response relationship, has been recently explored as an alternate approach to using estimates of bioavailable metal in the environmental medium for predicting metal toxicity [7,8]. However, All Supplemental Data may be found in the online version of this article. * To whom correspondence may be addressed ([email protected]). Published online 7 January 2013 in Wiley Online Library (wileyonlinelibrary.com). 810

Effect of major cations on Ni toxicity to Lemna minor

which is commonly referred to as a ternary complex [11]. Since Ni is a slow-reacting metal, uptake is more likely kinetically limited (i.e., nonequilibrium) and may result in increased bioavailability of Ni complexes and formation of ternary complexes on the organism’s surface [12]. Such discrepancies may be incorporated into a toxicity model or be deemed unnecessary for the purposes of risk assessment. Nickel is an important environmental contaminant, particularly in Canada, due to the presence of numerous past and present Ni mining, refinery, and smelter operations. The prediction of Ni toxicity to the floating aquatic plant Lemna minor was previously attempted using cross-species extrapolation of existing Ni BLMs [3]. Interestingly, the Ni toxicity data for L. minor were better explained by the Daphnia magna Straus BLM than by the algae BLM [3], which is, like Lemna, a photosynthetic organism. It would be valuable to have a predictive Ni toxicity model for L. minor as it was one of the species used in the national monitoring program, that is, Canadian Metal Mining Effluent Regulations under the Environmental Effects Monitoring Program (MMER-EEM) [13,14]. Lemna minor is suitable for metal biomonitoring because it is ubiquitous, small, easy to identify [15], and sensitive to metals. However, little information on predictions of metal toxicity for L. minor is available, although there have been a few studies on other Lemna species, such as L. paucicostata Hegelm [16]. In the present study, the effect of major cation activity (Ca2þ, Mg2þ, Naþ, Kþ) on the toxicity of Ni, expressed as total dissolved concentration (NiTot) or free Ni ion activity (Ni2þ), to L. minor was examined to assess the usefulness of the general BLM framework (i.e., exposure-based model with consideration of competition) for complex exposures. Toxicity based on tissue residue (NiTiss) and calculation of Ni accumulation kinetics, specifically the binding affinity constant (Kd) and binding capacity (Bmax), were examined to provide mechanistic explanations of the observed effects as well as to explore the use of the TRA as an alternate method for predicting Ni toxicity. The effect of the counter-anion on a cation’s effect on Ni toxicity was considered by testing different calcium salts (i.e., CaSO4, Ca(NO3)2, and CaCl2). Note that since preliminary results indicated that the counter-anion effect on Ni toxicity was not different among cations, the effect of the counter-anion was expected to be similar for the other cations (Mg, Na, and K) and was not studied. Analysis of variance (ANOVA) was used to identify the main effects of and interactions among experimental factors. Toxicity testing was performed in American Public Health Association (APHA) medium [17], which is hard water (149 mg CaCO3/L) at pH 8.3 and is relevant because of its use in MMER-EEM testing and applicability to mining-impacted waters [13,14]. METHODS

Reagents

All solutions were prepared with ultrapure water (resistivity 18.2 MV-cm) obtained from the Milli-Q Gradient A10 purification system (Millipore) designed to provide ultralow dissolved organic carbon (DOC) concentrations of between 1 and 5 mg/L. A standard solution of Ni(II) was prepared on the day of toxicity test solution preparation, using ultrapure Ni(NO3)2 powder (Puratronic 99.9985%; Alfa Aesar). Due to the slow equilibrium kinetics of Ni with organic matter [18], test solutions were made 3 d prior to test start. Solutions used for the cation test included CaSO4 (99%; Sigma-Aldrich), Ca(NO3)2 (Fisher Scientific), CaCl2  2H2O (Fisher Scientific), MgSO4  7H2O (Fisher

Environ. Toxicol. Chem. 32, 2013

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Scientific), KNO3 (Fisher Scientific), and NaNO3 (Fisher Scientific). Reagents used for the culture and test media [17] were all American Chemical Society–grade unless specified otherwise. Suwannee River natural organic matter (NOM; International Humic Substances Society) was the DOC and 0.1 M HCl (OPTIMA; Fisher Scientific) and 0.1 M NaOH (99.998% metals basis; Sigma-Aldrich) were used for all pH adjustments. The pH meter used throughout the study was a Hach HQ Series Portable Meter. Ultrapure HNO3 (OPTIMA; Fisher Scientific) was used for digestion of plant samples and preservation of test solutions prior to elemental analysis. Lemna minor culture

Plants of L. minor L. #8434 (CPCC 490 or formerly UTCC 490) were obtained from the Canadian Phycological Culture Centre at the University of Waterloo, Canada (formerly known as University of Toronto Culture Collection of Algae and Cyanobacteria [UTCC]). Lemna minor was cultured based on the standardized Environment Canada protocol [17]. An AirClean 600 (PCR Workstation; AirClean Systems) was used to perform L. minor transfers. The culture medium was modified Hoagland’s Eþ [17], which was sterilized by autoclaving, with pH adjusted to 4.6 using 0.1 M HCl/NaOH. Plants were aseptically inoculated into four 50-ml tubes weekly and incubated under 64 to 90 mmol m2 s1 white fluorescent lights at 25  28C. Lemna minor toxicity testing

The Environment Canada protocol [17] for toxicity testing using L. minor was followed. Plants were acclimated in the test medium (APHA) for 24 h prior to test start, and the test duration was 7 d. Each treatment (a single potentially competing cation concentration) was set up as an eight-level Ni concentration gradient as follows: six replicates of 0 mg/L Ni; four replicates each of 1.56, 3.13, and 6.25 mg/L Ni; and three replicates each of 12.5, 25, 50, and 100 mg/L Ni in 100 ml of solution in 210-ml transparent disposable plastic cups (Polar Plastic). At test end, root length, frond count, and dry weight were measured as endpoints. Plants were blotted dry with KimWipes to remove excess test water and then oven-dried at 608C for 24 h in a preweighed aluminum weighing dish (20 ml, Fisherbrand; Fisher Scientific). Dry weights were measured on a highprecision balance (Sartorius,  0.00006). Root length and frond count were the most sensitive and consistent endpoints; however, for the present study, root length was used for data analysis as it was the less variable of the two endpoints (data not shown). Toxicity thresholds (NiTot at which growth was inhibited by 25% [IC25NiTot] or 50% [IC50NiTot]) for frond count data are presented in Supplemental Data, Table S1, as it is currently the recommended endpoint for regulatory purposes [17]. Supplemental Data, Table S2, presents the same data when using root length as the endpoint. All cation ranges tested were chosen to represent conditions of mining effluents as closely as possible. The MMER-EEM database was assessed to determine the relevant test ranges. All containers used to prepare the test solutions were acid-washed by soaking them in approximately 15% HNO3 (trace metalgrade; Fisher Scientific) for a minimum of 24 h and rinsing three times with deionized water and once with ultrapure water. Modified APHA excludes ethylenediamine tetraacetic acid [17], which is typically included in growth media as a buffer and prevents formation of metal precipitates resulting from exceeding the solubility limits. However, it is a strong metal complexant which would influence the speciation of the test metal and,

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Y. Gopalapillai et al.

thus, affect its bioavailability. Alternatively, in the present study, 0.5 mg/L of DOC was added as a weak buffer to ensure nutrient availability, particularly Fe. A preliminary study showed that there was no substantial effect of addition 0.5 mg/L DOC on L. minor root length growth in the dose– response control solution (Supplemental Data, Table S3). In addition, 0.5 mg/L DOC had no effect on the effective Ni toxicity threshold with respect to no DOC added (Supplemental Data, Fig. S1). Treatment controls (i.e., a single Ni dose–response test with no major cation addition) are labeled control unless otherwise specified as the dose–response control (i.e., the 0 mg/L Ni treatment within the dose response). Control tests were conducted multiple times over the course of the present study, and the average measured concentrations of the major cations in the test solution were as follows: 12.0 mg/L Ca, 29.0 mg/L Mg, 11.0 mg/L K, and 140 mg/L Na. The cation treatment ranges (measured) were as follows: CaSO4 treatments 1 to 6 (25.6, 37.3, 53.5, 72.6, 108, and 133 mg/L Ca, respectively), Ca(NO3)2 treatments 1 and 2 (40.0 and 75.4 mg/L Ca, respectively), CaCl2 treatments 1 to 5 (39.3, 73.9, 152, 298, and 392 mg/L Ca, respectively), MgSO4 treatments 1 to 5 (40.4, 42.6, 49.1, 56.2, and 114 mg/L Mg, respectively), KNO3 treatments 1 to 3 (16.2, 23.8, and 37.9 mg/L K, respectively), and NaNO3 treatments 1 to 3 (188, 213, and 293 mg/L Na, respectively; Table 1).

Table 1. Water chemistry of treatment solutionsa Treatmentb

pHic pHfd Ca2þ Mg2þ

APHA 8.15 8.23 7.49 8.22 8.30 15.0e CaSO4 1 CaSO4 2 8.20 8.31 21.2e 8.11 8.18 29.3e CaSO4 3 CaSO4 4 8.10 8.34 38.2e 8.13 8.19 53.1e CaSO4 5 CaSO4 6 8.07 8.07 62.8e Ca(NO3)2 1 8.10 8.11 23.9e Ca(NO3)2 2 8.09 8.06 43.8e CaCl2 1 8.09 8.22 23.5e CaCl2 2 8.12 —f 42.8e CaCl2 3 8.14 — 83.0e CaCl2 4 8.16 — 149e CaCl2 5 7.96 8.19 187e MgSO4 1 7.97 8.18 6.96 MgSO4 2 7.97 8.22 6.88 MgSO4 3 8.11 8.24 6.69 7.96 8.16 6.51 MgSO4 4 MgSO4 5 8.11 8.18 5.47 KNO3 1 8.02 8.16 7.37 KNO3 2 8.04 8.28 7.34 KNO3 3 8.06 8.18 7.29 NaNO3 1 8.10 8.37 7.09 NaNO3 2 8.12 8.41 6.98 8.12 8.43 6.69 NaNO3 3 a



17.8 9.09 17.5 8.92 17.0 8.87 16.5 8.82 15.9 8.75 14.9 8.65 14.4 8.59 17.7 8.90 17.3 8.81 17.7 8.90 17.2 8.81 16.2 8.64 14.7 8.38 14.1 8.25 24.1e 8.89 e 25.1 8.88 28.3e 8.84 31.6e 8.80 54.6e 8.52 18.2 14.5e 18.1 21.3e 18.0 33.9e 17.5 8.87 17.3 8.83 16.6 8.72

Naþ SO2 NO 4 3

Cl–

107 98.0 97.5 96.9 96.2 95.2 94.5 97.7 96.8 97.7 96.7 94.8 92.0 90.6 97.7 97.5 97.1 96.7 93.8 98.5 98.4 98.3 167e 188e 254e

62.9 62.8 62.5 62.1 61.7 61.0 60.6 62.5 61.7 105e 158e 272e 477e 600e 62.6 62.5 62.3 62.0 60.3 62.3 62.0 61.2 63.1 63.1 62.9

33.6 52.1e 66.2e 84.4e 105e 139e 162e 31.5 28.9 31.6 29.0 25.0 20.3 18.4 60.5e 65.4e 77.7e 91.4e 184e 32.6 32.1 30.7 34.0 33.8 33.6

167 166 166 165 164 162 161 244e 338e 166 164 160 155 152 166 166 165 165 161 282e 339e 523e 175e 186e 206e

Ion activities were calculated by WHAM 6 (test cation inputs were based on measured concentrations, while the remaining were nominal); [dissolved organic carbon] ¼ 0.5 mg/L. All values other than pH are provided as milligrams per liter. b Treatments are defined in Methods (see L. minor toxicity testing). c Average pH of test solutions at test start. d Average pH of test solutions at test end. e Indicates ion activities changing due to the addition of the test cation. f Data not available. APHA ¼ American Public Health Association.

Plant digestion

Replicates of the dried plant samples from each Ni concentration were pooled and weighed before transfer into 30-ml glass tubes. Ultrapure concentrated HNO3 (0.5 ml) was added to the tubes and mixed well prior to placing them in a Tecator Digestion System 40 (1016 Digester) located inside a fume hood. Samples were digested at approximately 100 to 1108C for 1 to 2 h. The KIMAX filling funnels (short stem, 25 mm top diameter) were used to cover the glass tubes to reduce evaporation of acid during digestion. After digestion, the sample volume was made up to 5 ml with ultrapure water. Note that reported Ni accumulation includes surface-bound plus internal as no rinsing occurred prior to drying the plant tissues but excess test solution was removed by blotting the plants with KimWipes. Elemental analysis

Metal concentration in the test solutions at test start and in the digested plant tissue were measured by inductively coupled plasma mass spectrometry (XSeries 2; Thermo Fisher Scientific) or by inductively coupled plasma atomic emission spectroscopy (Vista RL; Varian) with an ultrasonic nebulizer to enhance detection limit. All Ni concentrations reported here were measured rather than nominal. Measured total cation concentrations (Ni, Ca, Mg, K, and Na) were not different from dissolved (filtered
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