Mercury tissue residue approach in Chironomus riparius: Involvement of toxicokinetics and comparison of subcellular fractionation methods

Aquatic Toxicology 171 (2016) 1–8

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Mercury tissue residue approach in Chironomus riparius: Involvement of toxicokinetics and comparison of subcellular fractionation methods Frédéric Gimbert a,∗ , Alain Geffard b , Stéphane Guédron c , Janusz Dominik d,1 , Benoit J.D. Ferrari e a Department Chrono-Environment, UMR UFC/CNRS 6249 UsC INRA, University of Bourgogne Franche-Comté, 16 route de Gray, F-25030 Besanc¸on Cedex, France b Université Reims Champagne Ardenne, Unité Mixte de Recherche-Ineris (UMR-I02) Stress Environnementaux et Biosurveillance des milieux aquatiques, Unité de Formation et de Recherche Sciences Exactes et Naturelles, Moulin de la Housse, BP 1039, F-51687 Reims Cedex 2, France c ISTerre, Université Grenoble-Alpes, IRD-UMR 5275 (IRD/UJF/CNRS)-BP 53, F-38041 Grenoble, France d Institute F.-A. Forel, University of Geneva, 10 route de Suisse, CP 416, CH-290 Versoix, Switzerland e Centre Ecotox, Eawag/EPFL, EPFL-ENAC-IIE-GE, Station 2, CH-1015 Lausanne, Switzerland

a r t i c l e

i n f o

Article history: Received 5 August 2015 Received in revised form 23 November 2015 Accepted 27 November 2015 Available online 30 November 2015 Keywords: Uptake Excretion Bioaccumulation Toxicity Metal partitioning

a b s t r a c t Along with the growing body of evidence that total internal concentration is not a good indicator of toxicity, the Critical Body Residue (CBR) approach recently evolved into the Tissue Residue Approach (TRA) which considers the biologically active portion of metal that is available to contribute to the toxicity at sites of toxic action. For that purpose, we examined total mercury (Hg) bioaccumulation and subcellular fractionation kinetics in fourth stage larvae of the midge Chironomus riparius during a four-day laboratory exposure to Hg-spiked sediments and water. The debris (including exoskeleton, gut contents and cellular debris), granule and organelle fractions accounted only for about 10% of the Hg taken up, whereas Hg concentrations in the entire cytosolic fraction rapidly increased to approach steady-state. Within this fraction, Hg compartmentalization to metallothionein-like proteins (MTLP) and heat-sensitive proteins (HSP), consisting mostly of enzymes, was assessed in a comparative manner by two methodologies based on heat-treatment and centrifugation (HT&C method) or size exclusion chromatography separation (SECS method). The low Hg recoveries obtained with the HT&C method prevented accurate analysis of the cytosolic Hg fractionation by this approach. According to the SECS methodology, the Hg-bound MTLP fraction increased linearly over the exposure duration and sequestered a third of the Hg flux entering the cytosol. In contrast, the HSP fraction progressively saturated leading to Hg excretion and physiological impairments. This work highlights several methodological and biological aspects to improve our understanding of Hg toxicological bioavailability in aquatic invertebrates. © 2015 Elsevier B.V. All rights reserved.

1. Introduction The critical body residue (CBR) is the concentration of a chemical species bioaccumulated in an organism that corresponds to a defined onset of toxicity, regardless of how it is measured (e.g., mortality, growth inhibition. . .) (McCarty and Mackay, 1993). This approach has recently evolved into a more general concept, which considers tissue residues as the dose metric when characterizing dose-response relationships, evaluating mixture toxicity, developing guidelines to protect organisms, and conducting risk

∗ Corresponding author. Fax: +33 381 665 797. E-mail address: [email protected] (F. Gimbert). 1 Present address: ISMAR-CNR, Arsenale—Tesa 104, Castello 2737/F, 30122 Venezia, Italy.

http://dx.doi.org/10.1016/j.aquatox.2015.11.027 0166-445X/© 2015 Elsevier B.V. All rights reserved.

assessments (Sappington et al., 2011). The so-called tissue residue approach (TRA) relies on the toxicological principle that a toxic effect is not observed unless the chemical reaches the site of action, i.e., only a fraction of the contaminant taken up is metabolically available to interact with sites of toxic action (Rainbow, 2007). Concerning metals, both essential elements in excess of metabolic requirements and non-essential elements must be detoxified either via excretion or sequestration (bound to cytosolic proteins or low-molecular weight compounds and to ‘insoluble’ deposits) processes. The cytosolic proteins involved are generally metallothioneins (MT) or metallothionein-like proteins (MTLP), which can specifically bind certain trace metals (e.g., Ag, Cd, Cu, Hg, Zn) (Amiard et al., 2006). Insoluble deposits include relatively heterogeneous lysosomal residual bodies or commonly discrete metal-rich granules (MRG) (Mason and Jenkins, 1995; Marigómez

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F. Gimbert et al. / Aquatic Toxicology 171 (2016) 1–8

issue regarding its efficiency to accurately define the cytosolic partitioning of Hg is still pending. The main objectives of the present research were to study Hg bioaccumulation and fractionation in larvae of the midge Chironomus riparius. First, we compared the efficiency and accuracy of the two methodologies previously mentioned for subcellular fractionation of Hg both in terms of Hg balance recoveries and partitioning fluxes. Second, we used a modeling approach, using a toxicokinetic model, to assess environmental bioavailability (i.e., uptake and bioaccumulation processes) and toxicological bioavailability (i.e., the portion of assimilated chemical that reaches and interacts with the sites of toxic action). Finally, ecotoxicological implications are discussed.

2. Material and methods Fig. 1. Mercury toxicokinetics in different subcellular fractions in C. riparius exposed to spiked water and sediments. Each point represents a replicate of 20 pooled larvae.

et al., 2002). According to Wallace et al. (2003), MT and MRG can be grouped as biologically detoxified metal (BDM) and the cytosolic organelles and target molecules such as heat-sensitive proteins (HSP, mostly enzymes) can be grouped as metal-sensitive fraction (MSF), which are of particular interest when considering TRA for toxicity assessment. In 2011, the TRA Special Series of Integrated Environmental Assessment and Management (volume 7, issue 1, pages 1–154) demonstrated that the TRA remains intrinsically limited if it is targeted at determining steady-state, whole-organism concentrations but may further advance if it is focused on target site concentrations and if toxicokinetic (what the organism does with the toxicant) and toxicodynamic (what the toxicant does to the organism) characteristics are integrated (Adams et al., 2011). The knowledge of the subcellular fate of metals is therefore a prerequisite to better understand physiological processes underlying their bioaccumulation and toxicity at various biological levels (Perceval et al., 2004; Campbell et al., 2005) as well as throughout the food web (Cheung and Wang, 2005; Seebaugh and Wallace, 2009). One of the most elaborate protocols of subcellular fractionation of metals was developed by Wallace and Lopez (1996) based on heat treatment and centrifugation (referred as HT&C method) allowing the partitioning of metals in the cytosolic fraction according to the heat-stable and heat-denaturation characteristics of MTLP and HSP, respectively. However, potential methodological artifacts may occur during this critical step, which are globally linked with the stability of the sulphur binding driving the association of metals with thiol compounds of the proteins (Bragigand and Berthet, 2003). Hence, alternative approaches using size exclusion chromatographic separation (referred to as SECS method) have been used to assess metal partitioning in the cytosolic fraction (Fraysse et al., 2006; Perceval et al., 2006). A comparative study of subcellular fractionation of cadmium (Cd), nickel (Ni) and lead (Pb) in Gammarus fossarum using both SECS and HT&C methods emphasized important differences in the cytosolic partitioning of Cd particularly but also of Ni and to a lesser extent of Pb between HSP and MTLP (Geffard et al., 2010). Among studied metals known to have significant toxicological impact on biota, mercury (Hg) has been identified as a potent neurotoxicant for animals and humans, especially under its organometallic form (i.e., methylmercury) which strongly binds with sulphydryl groups in proteins and is therefore readily accumulated and retained in tissues (Clarkson, 1992). Up to date, only few data are available regarding its subcellular distribution in aquatic invertebrates (Seebaugh and Wallace, 2009; Tsui and Wang, 2009; Xie et al., 2009; Pan and Wang, 2011). Moreover, these studies exclusively used the Wallace-based fractionation protocol and the

2.1. Organisms We used fourth instar larvae of C. riparius cultured in the laboratory prior to the experiments according to standard methods (details given in the Supplementary material I).

2.2. Sediment spiking We spiked artificial silica sediment (TerraSand, JBL, Germany) previously sieved through a 400 ␮m mesh. The particle size distribution was 50% between 250 and 400 ␮m, and 50% below 250 ␮m (Coulter® LS-100, Beckman Coulter, Fullerton, CA, USA). Seven kilograms of artificial sediments were first maintained with 6 L of Lake Geneva water and a small amount of food (0.2 g Tetramin® fish food) for three weeks to allow bacterial development. Then, about 2 kg of wet sediment were disposed into 2 L jars together with the required quantity of metal (HgCl2 , Sigma–Aldrich, purity 99.99%) dissolved in 0.5 L lake water to reach nominal concentrations of 0 and 1.5 ␮g g−1 (control and exposure conditions, respectively). This exposure design (artificial sandy sediment spiked with a unique reactive Hg species) and concentration were chosen to favor (i) Hg availability and bioaccumulation and (ii) the onset of potential toxic effects (Chibunda, 2009). Jars were then shaken end-over-end during 6 h, kept at test temperature and swirled manually each day for 10 days.

2.3. Experimental design Larvae of C. riparius were exposed in static conditions for four days in 0.6 L beakers, filled with 0.1 L control or spiked sediments and 0.4 L of lake Geneva water (pH 7.5 and conductivity 300 ␮S cm−1 ). Beakers were set in a water bath at 21 ◦ C with a 16:8 h light:dark photoperiod and left to stabilized (sediment-water partitioning of Hg) for 48 h before the introduction of larvae. Water was oxygenated through a continuous air injection in each beaker and daily checked for pH, conductivity and temperature. For each exposure condition, 15 replicate beakers containing 20 individuals were used and fed ad libitum (0.6 mg Tetramin® .larvae−1 .day−1 ) (Péry et al., 2002). After 12 hours, 1, 2, 3 and 4 days of exposure, the larvae were sampled in triplicate beakers by gently sieving the sediment to recover organisms. Then, living larvae were dried on an absorbent paper sheet, individually measured for their length, then pooled per replicate in cryogenic tubes, weighed and finally sacrificed by introducing the tubes into liquid nitrogen. Tubes were kept frozen (−20 ◦ C) until fractionation and Hg analysis. At each sampling date, overlying water and sediments were sampled for grain-size and Hg analysis.

45.9 2.98 0.16 40.3 0.78 – 7.27 – 28.9 31.2 2.85 0.16 33.4 0.70 12.2 6.54 19.5 17.6 0.97 0.88 0.78 0.95 0.95 0.90 0.92 0.97 0.94 <0.001 0.022 0.023 <0.001 <0.001 0.608 <0.001 0.108 0.005 0.717 5.938 5.785 0.603 1.864 0.313 2.542 0.024 0.635 28.52 17.93 0.866 20.77 1.394 6.639 17.23 6.775 14.53 19.89 7.580 0.270 13.73 0.967 3.586 9.831 4.610 9.524 2.280 4.421 0.143 2.114 0.131 0.925 1.920 0.633 1.534 23.75a 12.37b 0.424c 16.93b 1.154d 4.926a 12.70b 5.613a 11.65b

d

a ␮g g C(4) ± SD ␮g g

32.9 ± 1.5 a 2.72 ± 0.44 b 0.17 ± 0.01 c 30.8 ± 2.8 a 0.63 ± 0.07 d 11.0 ± 2.9 a 6.39 ± 0.81 b 17.5 ± 3.7 ac 16.8 ± 0.6 c

Fraction

Homogenate Debris Granules Cytosol (total) Organelles MTLP HSP MTLP HSP

Sample

H S2 P2 S3 P3 S4HT&C P4HT&C S4 SECS P4 SECS

Within total (upper part) and cytosolic (lower part) fractionations, values that share similar letters are not significantly different (GLRT).

0.350 2.369 1.502 0.260 1.156 −0.135 1.248 −0.005 0.207 0.097 1.643 1.032 0.103 0.219 0.137 0.334 0.017 0.126 0.517a 4.155b 2.585b 0.420a 1.464c 0.072a 1.746b 0.009a 0.403c

ke d p-Value Max SE

Min

<0.001 0.013 0.009 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001

C (ss) ␮g g−1 C (4) ␮g g−1 R2 SE

Min

Max

p-Value

Modeled CI 95%

−1

CI 95%

−1 −1

3

2.4. Subcellular fractionation

−1

Measured

Table 1 Toxicokinetic parameters of subcellular fractionation of mercury in exposed Chironomus riparius according to heat treatment and centrifugation (HT&C) and size exclusion chromatography separation (SECS) methods.

F. Gimbert et al. / Aquatic Toxicology 171 (2016) 1–8

The whole body fractionation procedure was adapted from the method described by Wallace et al. (2003). From the homogenate (H), a total of four subcellular fractions were obtained: P2 = granules, S2 = debris (i.e., exoskeleton, gut content and cellular debris), P3 = organelles and S3 = total cytosol. Details are given in the Supplementary material II. The Hg fractionation of cytosolic compounds was carried out according to two methods based on heat treatment and centrifugation (HT&C method) and size exclusion chromatography separation (SECS method). After aliquot sampling for total Hg analysis, the cytosolic fraction (S3) was therefore divided into two parts. The first part was heated at 80 ◦ C for 10 min, subsequently cooled on ice for 1 h and centrifuged at 50,000 × g for 30 min at 4 ◦ C. This resulted in a pellet (P4) containing heat-sensitive proteins like enzymes (HSP) and a supernatant (S4) containing heat-stable metallothionein-like proteins (MTLP). The second part of the cytosol (S3) was partitioned according to the methodology of Perceval et al. (2006) and Geffard et al. (2010). The sample was filtered (0.2 ␮m, PES membrane) and fractionated by high-performance liquid chromatography (HPLC) in a steric exclusion column (BIOSEP-SEC-S 2000; 30 × 0.75 cm) preceded by a guard column. The column was eluted with a degassed mobile phase of 10 mM Tris–HCl (Sigma–Aldrich, purity 99%), 100 mM NaCl (Sigma–Aldrich, purity 99.5%), adjusted to pH 6.8, at a flow rate of 0.5 mL min−1 . The HPLC column was washed periodically with an injection of 1 mM EDTA (TraceSELECT® , Sigma–Aldrich) in the same mobile phase. The column was calibrated for molecular weight estimation using thyroglobulin (660 kDa, V0 = 6 mL), cglobulin (158 kDa), bovine albumin (66 kDa), cytochrome C (17 kDa), vitamin B12 (1355 Da), and tryptophan (204 Da) as standard markers. Eluting fractions (0.5 mL) were collected automatically up to a volume of 18 mL and acidified at 0.5% (v/v) with suprapure HCl (Ultrex, Fisher scientific) and stored at 4 ◦ C until Hg analysis. Three metal-ligand pools were defined: a high-molecular weight pool (HMW; 255–18 kDa), a metallothionein-like protein pool (MTLPs-MW; 18–1.8 kDa) and a low-molecular weight pool (LMW; <1.8 kDa). Using this protocol, the HSP pool is assumed to correspond to the sum of HMW and LMW fractions (Bragigand and Berthet, 2003; Geffard et al., 2010). 2.5. Mercury analysis For the entire procedure, all material in contact with the samples were acid-washed (5 days in 20% HNO3 v/v followed by 5 days in HCl 10% v/v) and rinsed several times with MQ water before use. Reagents were controlled by analyzing blanks which were always below 0.1 ng L−1 .Mercury analyses in sediments, overlying waters and all subcellular fractions were performed by cold vapor atomic fluorescence spectrometry (CVAFS) followed by detection using a Tekran® (Model 2500) mercury detector after conversion of all mercury species into Hg0 . The principles of the method used are from the gold amalgamation method (Bloom and Fitzgerald, 1988). Analysis of Hg in pellet fractions and sediments were performed after HCl/HNO3 (Ultrex, Fisher scientific) digestion (10 h at 70 ◦ C) in PFA Teflon reactors (Coquery et al., 1997). The accuracy of analyses was checked using CRM ORMS-4 (22.0 ± 1.6 ng L−1 –National Research Council of Canada). Recoveries were usually within 10% of the certified value. 2.6. Statistical analyses The larval growth (in terms of individual fresh mass and length) during the exposure was assessed using an exponential growth model allowing a growth rate constant (kg , d−1 ) to be estimated.

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A one-compartment model (Gimbert et al., 2006, 2008) was used to describe Hg bioaccumulation kinetics in each subcellular fraction Eq. (1). C(t) = C(0) +

a (1 − e−(kg +ke )t ) (kg + ke )

(1)

where C(t) is the Hg concentration in the subcellular fraction (␮gHg × gchiro −1 ) at time t(d), a is the uptake rate (␮gHg × gchiro −1 × d−1 ) into the subcellular fraction, kg is the previously estimated growth rate constant (d−1 ) of the whole larva and ke is the excretion rate constant (d−1 ) from the subcellular component. C(0) is the initial Hg concentration (␮gHg × gchiro −1 ) measured in the various fractions at the beginning of the experiment. The growth, uptake and excretion parameters were estimated by fitting the models with a nonlinear least squares procedure (nls, Bates and Chambers, 1992). For all models, when residuals were skewed, variance functions (power and exponential) were applied and the best model was selected according to Akaike’s Information Criterion (AIC, pgirmess R package). Differences in parameter estimates between exposure modalities and subcellular fractions were checked using a generalized likelihood ratio test (GLRT). Differences in Hg concentrations, distributions and recoveries between control and exposed organisms and between methodological protocols were verified using analysis of variance (lm, anova, p < 0.05). The post-hoc test of Tukey was thereafter applied to determine the statistical differences for pairwise comparisons (p < 0.05). All statistics were performed with the free statistical software package R (ver 3.1.0) (R Development Core Team, 2013). 3. Results and discussion 3.1. Whole body and partitioning toxicokinetics of mercury For C. riparius larvae exposed to control conditions (1.56 ± 0.06 ng g−1 dry weight and 4.93 ± 0.13 ng L−1 in sediments and overlying water, respectively), no increase of total Hg concentration was observed and internal concentrations never exceeded 0.05 ␮g g−1 (fresh weight) in the entire body (homogenate (H) fraction, data not shown). Under the exposure conditions (1.35 ± 0.08 ␮g g−1 dry weight and 6.99 ± 0.30 ␮g L−1 in sediments and overlying water, respectively), total Hg concentrations in the homogenate (H) fraction increased rapidly (Fig. 1) before they tended towards a plateau (C(ss) = 45.9 ␮g g−1 fresh weight) at the end of the exposure period (C(4d) = 32.9 ± 1.5 ␮g g−1 fresh weight; Table 1). The one-compartment toxicokinetic model adequately described the variations of total Hg concentrations over time in C. riparius larvae and allowed the estimation of significant uptake rates (a) and excretion rate constants (ke ) in whole larvae and in selected fractions (Table 1). We observed high Hg accumulation capacities of exposed C. riparius larvae, with a total uptake rate of 23.7 ␮gHg × gchiro −1 d−1 (homogenate, H). This value testified an elevated environmental bioavailability of Hg to C. riparius. This is not surprising considering the cumulative exposure to contaminated sediments and water and their Hg concentrations were certainly elevated, but quite realistic, for instance, in the vicinity of gold mining sites (Olivero-Verbel et al., 2014) or chlor-alkali plant (Ullrich et al., 2007). To allow inter-specific comparison of exposure, an uptake rate constant (ku ) can be derived by dividing the uptake rate (a) by the exposure concentration. For C. riparius, we calculated a Hg uptake rate constant from water of 3.4 L g−1 d−1 which is higher than reported values for other aquatic Hexapoda (from 0.3 to 0.9 L.g−1 d−1 ) (Xie et al., 2009) or Daphnia magna (0.35 L g−1 d−1 ) (Tsui and Wang, 2004), but in the same range as values reported for bivalves (from 3.5 to 32.8 L g−1 d−1 ) (Pan and Wang, 2011). For various species, elimination processes

may counteract entering fluxes leading to the levelling of the Hg accumulation towards a plateau, as observed in this study. Excretion has been clearly separated from the dilution effect due to growth (kg = 0.199 d−1 , Fig. 2) which nevertheless contributed to 23% of the total Hg elimination. With a value of 0.54 d−1 , the Hg excretion rate constant of whole C. riparius was much higher than those reported for other Hexapoda (0.09–0.16 d−1 ) (Xie et al., 2009), D. magna (0.05–0.06 d−1 ) (Tsui and Wang, 2004) or bivalves (0.02–0.06 d−1 ) (Pan and Wang, 2011). Although excretion may rely on species-specific physiological processes, it can also be modulated according to the rate of uptake and the efficiency of detoxification strategies that condition the internal distribution and toxicodynamics of trace metals (Wang and Rainbow, 2008). The study of Hg fractionation allowed the identification of subcellular compartments involved in its distribution within larvae (Table 2). The first step was the verification of the efficiency of the whole body fractionation protocol. Hence, the Hg mass balance was evaluated by comparing Hg concentrations measured in the homogenate (H) to the sum of Hg concentrations in the four main subcellular compartments (granules, debris, organelles and cytosol). Results appeared very satisfactory with Hg recoveries (at day 4) of 97% and 104% in control and exposed larvae, respectively. To refine whole body residue observations, we applied an original tissue residue approach of Hg based, as previously advocated (Adams et al., 2011), on kinetic modeling within the different subcellular compartments investigated. The variability in ecotoxicological outcomes and species sensitivity is due in part to differences in toxicokinetics, which consist of several processes, including absorption, distribution, metabolism, and excretion (ADME), that influence internal concentrations and also toxicity. According to our results, the granular fraction (P2), also called metal-rich granules (MRG), was not a sink for Hg (slow accumulation pattern reaching 0.51% of the accumulated Hg after 96 h of exposure; Table 2), probably because of its low affinity with the carbonate, sulphate or phosphate constituents of the MRG (Tai and Lim, 2006). Similarly, organelles (P3) did not accumulate elevated Hg concentrations (0.6 ␮g g−1 , about 2% of total Hg) but the kinetics was quite different with higher uptake rate (1.15 ␮g g−1 d−1 ) and lower excretion rate constant (1.464 d−1 ; Table 1). Within this fraction, lysosomes, cellular structures involved in protein degradation, may incorporate metal although mechanisms are not well-known. Nevertheless, several authors have observed the concomitant presence of sulphur and particular trace metals in lysosomes, which probably results from the incorporation of MT into these cellular organelles (Marigómez et al., 2002). The debris fraction (S2) accounted for 15% of the total Hg accumulated (Table 2). Within this compartment, Hg concentrations increased rapidly (with an uptake rate of 12.4 ␮g g−1 d−1 ) and the steady state was reached already after only 12 h of exposure (ke = 4.15 d−1 , Table 1). This almost instantaneous phenomenon likely corresponded to the amount of Hg adsorbed onto the exoskeleton and the particles present in the digestive tract of the chironomids. The relative proportion of Hg in these subcellular fractions is generally lower than was observed in bivalves (Pan and Wang, 2011) or other aquatic insects (Xie et al., 2009). Finally, most of the accumulated Hg was found in the cytosolic fraction (S3) (78% and 92% in control and exposed larvae, respectively; Table 2) and accounted for more than 70% of the total Hg uptake rate in the larvae (Table 1). This is consistent with the general idea that Hg can be sequestered by proteins (sulfhydryl groups) or larger peptides such as glutathione or Hg-cysteine complexes (Tai and Lim, 2006). 3.2. Comparison of subcellular fractionation methods To examine Hg bioaccumulation in C. riparius larvae regarding detoxification strategies and toxicity induction, the Hg fractionation within the cytosolic fraction (S3) was performed following

F. Gimbert et al. / Aquatic Toxicology 171 (2016) 1–8

5

Fig. 2. Growth of Chironomus riparius expressed as (a) length (mm) and (b) fresh mass (mg) in both control and exposed conditions. The curves and growth rate constants (kg ) are provided by fitting exponential growth model to the data (see text for details). Different letters indicate significant difference (GLRT).

Table 2 Average (± standard deviation) mercury distributions (in %) and recoveries (percentage relative to homogenate, H) in whole body subcellular fractions of control and exposed C. riparius (4 days). Distribution

Control Exposed

Recoveries

P2 Granules

S2 Debris

P3 Organelles

S3 Total cytosol

1.25 ± 0.86 0.51 ± 0.03 ns

15.3 ± 2.6 9.34 ± 2.62 ns

3.88 ± 1.03 2.10 ± 0.04 ns

78.2 ± 21.5 91.9 ± 11.3 ns

97.5 ± 18.6 104 ± 14 ns

ns = not significantly different from control (lm, anova, p < 0.05).

both HT&C and SECS protocols. For the latter, an example of an elution profile and the Hg concentrations in the three molecular weight-based protein pools is presented in Fig. 3. Although the biomass of control and exposed larvae significantly differed, the quantity and the weight-based distribution of proteins were quite similar and their elution profiles overlapped. The highest Hg concentrations were found in the MTLP-MW pool, followed by HMW pool and finally the LMW pool where relatively low Hg concentrations were found (Fig. 3). To allow comparison between the two methodological approaches, HMW and LMW pools have been summed as a HSP fraction (Bragigand and Berthet, 2003; Geffard et al., 2010) that will be only presented subsequently in the paper. The efficiency of both protocols to determine Hg fractionation in cytosol was evaluated using Hg recoveries (differences between Hg concentrations measured in the total cytosol (S3) and the sum of Hg concentrations in MTLP and HSP fractions) (Table 3). The results showed that the HT&C protocol led to lower Hg recoveries compared to the SECS methodology. Although the differences

were not significant for the control (78%), they were particularly marked for the exposed organisms (56%). A first explanation is Hg volatilization and/or adsorption onto vessel, occurring during the Table 3 Mean (± standard deviation) mercury distributions (in %) and recoveries (percentage relative to total cytosol, S3) in the cytosolic fraction of control and exposed C. riparius (4 days) according to heat treatment and centrifugation (HT&C) and size exclusion chromatography separation (SECS) methods. Cytosolic distribution P4 HSP HT&C SECS

Control Exposed Control Exposed

Recoveries S4 MTLP

40.3 20.8 75.9 49.7

± ± ± ±

15.0 3.2* 10.5◦ 3.6*◦

38.0 35.6 21.8 60.7

± ± ± ±

9.5 6.9 6.7◦ 4.7*◦

78.3 56.4 97.6 110

± ± ± ±

17.2 3.7* 10.6 8◦

Asterisk (*) symbolises significant difference from control and circle (◦ ) symbolises difference between the two methods (lm, anova, p < 0.05).

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Fig. 4. Kinetics of mercury cytosolic fractionation between metallothionein-like proteins (MTLP, circles and dashed curve) and heat-sensitive proteins (HSP, crosses and solid curve) in exposed C. riparius according to heat treatment and centrifugation (HT&C, in black) and size exclusion chromatography separation (SECS, in grey) methods. Each point represents a replicate of 16–20 pooled larvae.

Fig. 3. Elution profile (a) and total mercury concentrations (␮g g−1 ) (left axis for exposed and right axis for control larvae) (b) of cytosolic proteins in C. riparius at the end of exposure (4 days). Three metal-ligand pools were defined: a highmolecular weight pool (HMW), metallothionein-like protein pool (MTLP-MW) and a low-molecular weight pool (LMW) (see text for details).

methodology for Hg subcellular fractionation, unless improvements of the HT&C heating step are provided. The use of closed PTFE centrifuge tubes during the heat treatment step could possibly improve the yield of this methodology.

3.3. Cytosolic fractionation of mercury heating phase and leading to an Hg loss of about 30% in the HSP fraction (Table 3). Moreover, this phenomenon has probably been facilitated by a concomitant Hg transfer from HSP to MTLP compartments, as already observed for various metals and organisms (Bragigand and Berthet, 2003; Geffard et al., 2010). It is possible that heat denaturation of HSP released Hg into the solution, which could then bind to MTLP, as long as they were not saturated with Hg. Published experiments regarding subcellular fractionation of Hg in aquatic invertebrates has employed the HT&C protocol (Tsui and Wang, 2006; Seebaugh and Wallace, 2009; Xie et al., 2009; Pan and Wang, 2011) and one can regret the frequent absence of information regarding Hg recoveries. The accumulation patterns of Hg in HSP and MTLP fractions for both methodologies are presented in Fig. 4. Although each protocol provided quite similar accumulation patterns over the first hours of exposure, differences became more and more evident with time and the increase of internal Hg concentrations. This was also shown by the kinetic parameter estimates (Table 1). The observed loss of Hg resulted in an increase of the elimination rate constant modeled for the HSP fraction using HT&C protocol (1.75 d−1 vs 0.40 d−1 using the SECS method) which prevented Hg cytosolic toxicokinetics to completely overlap (Fig. 4). However, it is interesting to note that this methodological artifact did not influence uptake rates that reached values of about 5 and 12 ␮gHg × gchiro −1 d−1 for MTLP and HSP fractions, respectively, regardless of the methodology employed. This aspect may be of particular importance for toxicological bioavailability consideration in TRA, especially when regarding the growing body of evidence suggesting that chronic and sublethal effects depend rather on fluxes towards a site of action than internal concentration levels (Rainbow, 2007; Wang and Rainbow, 2008). However, the Hg loss and the related artifacts concerning its fate, especially when it concerns physiologically reactive fraction such as HSP, may lead to serious misinterpretations of results. For that reason, we advocate the use of SECS

According to the SECS protocol, we highlighted two contrasting patterns of Hg fractionation in the cytosol of control and exposed larvae (Table 3). While in control MTLP bound Hg represents about 20% of the cytosolic accumulation, it reached 60% in exposed organisms. Comparing to the available literature data, Hg proportions in the MTLP fraction were higher than those reported in D. magna (about 41%) exposed in the lab to comparable Hg concentrations in water (Tsui and Wang, 2006), underlining the Hg detoxification capacities of C. riparius. The proportions of Hg accumulated in the heat-sensitive fraction found in D. magna (about 43%) were quite similar to the ones we reported here for C. riparius and also related to sublethal toxicity. However, the threshold lethal concentration (i.e., 24 h LC50 ) regarding Hg concentration in the metal-sensitive fraction (MSF = organelles + HSP) in D. magna was 3.8 ± 1.9 ␮g g−1 fresh weight (Tsui and Wang, 2006). In comparison, we measured 17.4 ␮g Hg g−1 in the MSF fraction of exposed larvae without mortality, only a growth inhibition of about 20% (Fig. 2), indicating the high tolerance of C. riparius. One added value of the present paper is the kinetic assessment of Hg partitioning between the sensitive (HSP) and detoxified (MTLP) components of the cytosol (Wallace and Luoma, 2003). We highlighted two differential Hg metabolisms: a linear accumulation pattern in the MTLP fraction while Hg bound-HSP tended to a plateau if the duration of exposure was increased. The linear accumulation in the MTLP fraction indicated detoxification processes (average uptake rate close to 5.6 ␮g Hg g−1 d−1 ) and no significant excretion processes (0.009 d−1 , p = 0.108) probably in relation with slow MTLP turnover and degradation in organelles (lysosomes). However, although not overloaded, MTLP could not sequester all the Hg entering the cytosol (about 33%) and an important part of the taken up Hg reached the HSP fraction (Fig. 4), which was close to saturation at the end of the exposure period (17 ␮g g−1 fresh weigh, 60% of the modeled steady-state concentration, Table 1).

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3.4. Ecotoxicological implications

Acknowledgements

The benefit of determining fractionation of metals relies on a better consideration of the relationships between their bioaccumulation in organisms and the potential biological effects, considering that only a part of the accumulated metals is toxicologically active (Rainbow, 2007). Metals bound to HSP (mostly enzymes) can be considered as toxicologically bioavailable since they may exert adverse effects at the molecular and individual levels (Vijver et al., 2004). Here, the potential effects of Hg on C. riparius larval growth was assessed considering the time course of both body length (mm) and individual fresh mass (mg) (Fig. 2). These two biological traits gave similar results with significantly lower lengths and masses (p = 0.027 and p = 0.005, respectively) in larvae exposed for 96 h to spiked sediments and water compared to control conditions. The modeling indicated a growth inhibition of 17% on the length basis and 23% on the individual fresh mass basis with growth constants (kg ) estimates of 0.260 d−1 and 0.199 d−1 in control and exposed conditions, respectively. Considering the combined exposure to contaminated water (6.9 ␮g Hg L−1 ) and sediments (1.3 ␮g Hg g−1 ), these results are in agreement with the literature data since toxicity bioassays with C. riparius reported EC20 for growth ranging between 7.5 and 50.3 ␮g Hg L−1 in water and between 1.8 and 3.2 ␮g Hg g−1 in sediments (Chibunda, 2009; Azevedo-Pereira and Soares, 2010). This may be interpreted as the result of a toxic effect of Hg boundHSP affecting, for instance, lipid peroxidation (Arambourou et al., 2013) or protein functioning and synthesis (Barka et al., 2001). However, regarding the intense Hg trafficking at the subcellular level, this growth impairment could also be attributed, referring to the DEBtox theory (Kooijman et al., 2009), to a change in the energetic allocation between somatic maintenance (excretion and/or detoxification processes) and growth. Although quite moderate, this impact on growth could be particularly important for insects with short adult stages, because fecundity is determined by the size of larva (especially female) upon metamorphosis (Péry et al., 2002). As chironomid larvae constitute an important food reserve for macro-invertebrates, fishes and birds, knowledge of the internal distribution of Hg is a valuable tool for interpreting its ecotoxicological hazard through trophic transfer. Indeed, Wallace and Luoma (2003) proposed the various subcellular compartments to be arranged, according to their biological significance, in metalsensitive fractions (MSF = organelles + HSP), biologically detoxified metal (BDM = granules + MTLP) and trophically available metal (TAM). The composition of the TAM fraction has to be determined in function of the predator and more precisely, the strength of its digestive and assimilative processes (Rainbow et al., 2011). Considering the trophic position of chironomid larvae, we can estimate that, at least, the cytosol fraction (HSP + MTLP) could be trophically available to most, if not all, predators. In our case, it represented about 90% of the total Hg taken up by the exposed organisms.

This study was supported by the Swiss National Science Foundation (grant no. 200020-117942). The authors sincerely thank Hervé Sartelet (Laboratoire Signalisation des Récepteurs Matriciels SiRMa, Université de Reims Champagne-Ardenne) for technical assistance in SECS and Benjamin Pauget (Department Chrono-Environment, University of Bourgogne Franche-Comté) for fruitful discussions about the manuscript.

4. Conclusions The proposed toxicokinetic approach appears as an interesting tool, along with toxicodynamic considerations, for the development of more mechanistic approaches regarding sublethal toxicity and understanding species sensitivity. For this purpose, fractionation methods containing a heating step are not recommended, especially for trace metals subject to volatilization such as Hg or arsenic (As). Separation using size exclusion chromatography appears to be an accurate methodology and allows to emphasize the important detoxification and tolerance abilities of C. riparius towards Hg, especially through its sequestration in the MTLP fraction.

Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.aquatox.2015.11. 027.

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