# Fate, subcellular distribution and biological effects of rare earth elements in a freshwater bivalve under complex exposure

Nicolas Lachaux, Alba Otero-Fariña, Laetitia Minguez, Bénédicte Sohm, Julie Rétif, Amélie Châtel, Laurence Poirier, Simon Devin, Sandrine Pain-Devin, Elisabeth Gross, et al.

► To cite this version:

Nicolas Lachaux, Alba Otero-Fariña, Laetitia Minguez, Bénédicte Sohm, Julie Rétif, et al. Fate, subcellular distribution and biological effects of rare earth elements in a freshwater bivalve under complex exposure. Science of the Total Environment, 2023, 905, pp.167302. 10.1016/j.scitotenv.2023.167302. hal-04286295

HAL Id: hal-04286295

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Distributed under a Creative Commons Attribution 4.0 International License
Fate, subcellular distribution and biological
effects of rare earth elements in a freshwater
bivalve under complex exposure

Nicolas Lachaux*ab, Alba Otero-Fariñaac, Laetitia Minguezª, Bénédicte Sohmª, Julie Rétifcd,
Amélie Chatelc, Laurence Poirierd, Simon Devin ab, Sandrine Pain-Devin ab, Elisabeth M.
Gross ab, Laure Giamberini ab

<sup>a</sup>Université de Lorraine, CNRS, LIEC, F-57000, Metz, France

<sup>b</sup>LTSER-Zone Atelier Moselle, F-57000, Metz, France

cUniversité Catholique de l'Ouest, Laboratoire Mer, Biologie des Organismes, Stress, Santé, Environnement (BIOSSE), F-49000 Angers Cedex 01, France

dNantes University, Institut des Substances et Organismes de La Mer, ISOMer, UR 2160, F-44000, Nantes, France

1 Abstract

Rare earth elements (REE) are emerging contaminants due to their increased use in
diverse applications including cutting-edge and green-technologies. Their environmental
concerns and contradicting results concerning their biological effects require an extensive
understanding of REE ecotoxicology. Thus, we have studied the fate, bioaccumulation and
biological effects of three representative REE, neodymium (Nd), gadolinium (Gd) and
ytterbium (Yb), individually and in mixture, using the freshwater bivalve *Corbicula fluminea*.
The organisms were exposed for 96 h at 1 mg L⁻¹ REE in the absence and presence of dissolved
organic matter (DOM) reproducing an environmental contamination. Combined analysis of the
fate, distribution and effects of REE at tissue and subcellular levels allowed a comprehensive
understanding of their behaviour, which would help improving their environmental risk
assessment. The bivalves accumulated significant concentrations of Nd, Gd and Yb, which were
decreased in the presence of DOM likely due to the formation of REE-DOM complexes that
reduced REE bioavailability. The accumulation of Nd, Gd and Yb differed between tissues,
with gills > digestive gland ≥ rest of soft tissues > hemolymph. In the gills and in the digestive
gland, Nd, Gd and Yb were mostly (>90 %) distributed among metal sensitive organelles,
cellular debris and detoxified metal-rich granules. Gadolinium, Yb and especially Nd decreased
lysosome size in the digestive gland and disturbed osmo- and iono-regulation of *C. fluminea* by
decreasing Na concentrations in the hemolymph and Ca²⁺ ATPase activity in the gills.
Individual and mixed Nd, Gd and Yb exhibited numerous similarities and some differences in
terms of fate, accumulation and biological effects, possibly because they have common abiotic
and biotic ligands but different affinities for the latter. In most cases, individual and mixed
effects of Nd, Gd, Yb were similar suggesting that additivity approach is suitable for the
environmental risk assessment of REE mixtures.

**Keywords:** lanthanide; mixture; ecotoxicity; bioaccumulation; dissolved organic matter,
*Corbicula fluminea*
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# 1. Introduction

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Rare earth elements (REE) are a group of seventeen metals composed of scandium (Sc), yttrium (Y) and the lanthanide series, which include lanthanum (La), cerium (Ce), praseodymium (Pr), neodymium (Nd), samarium (Sm), europium (Eu), gadolinium (Gd), terbium (Tb), dysprosium (Dy), holmium (Ho), erbium (Er), thulium (Tm), ytterbium (Yb) and lutetium (Lu). These elements share similar physicochemical properties, including electronic structure, oxidation state (+III) and atomic radius size. Based on their molecular weight, REE are divided in three groups: light REE (LREE) from La to Nd; medium REE (MREE) from Sm to Gd and heavy REE (HREE) from Tb to Lu. The ionic radius size of lanthanides is relatively close but it decreases progressively from La to Lu due to a phenomenon called “lanthanide contraction”. Due to their specific physicochemical properties, REE production and uses in diverse applications (cutting-edge technologies, renewable energy production, industry, medicine...) are growing, which is expected to continue in the future (Gregoir, 2022). From mining extraction through production and use to e-waste management, REE are released into all compartments of the environment, which raised concerns because of their potential accumulation and toxicity, notably towards aquatic organisms (Gwenzi et al., 2018). Despite the increasing amount of ecotoxicological studies, several knowledge gaps remain to be filled concerning their fate, bioaccumulation and biological effects (Arienzo et al., 2022; Blinova et al., 2020), as the ones described below. It is not clear whether all REE have a common pattern of biological effects. Some authors reported a similar toxicity among REE (Blinova et al., 2018; Huang et al., 2022; Lachaux et al., 2022b; Morel et al., 2021; Tai et al., 2010), which may be related to their similar physicochemical properties that influence their chemical behaviour, bioaccumulation pathway and toxicity mechanisms. On the contrary, other studies showed a higher toxicity or ecotoxicological risk of HREE (González et al., 2015; Gu et al., 2023, 2022, 2020; Manusadžianas et al., 2020) or LREE (Blaise et al., 2018; Lin et al., 2022). Environmental
risk assessment (ERA) of REE on freshwater ecosystems is limited because ecotoxicological
data focus on standard species and because existing studies are poorly representative of
environmental complexity. For instance, exposure media of standard tests are exclusively
mineral omitting the ubiquitous dissolved organic matter (DOM), which is likely to interact
with REE and thus influence their speciation (Davranche et al., 2017; Tang and Johannesson,
2003), bioaccumulation and toxicity (Lachaux et al., 2022a).

In the present study, we analyse REE from their distribution in water and organisms until
their biological effects at different levels of organization. This integrated approach aims at a
comprehensive understanding of the mechanisms running REE bioaccumulation and toxicity
using the bivalve *Corbicula fluminea*. This widespread filter-feeding organism represents a
relevant part of freshwater ecosystems with often high biomass being essential for regulation
of phytoplankton density and water depuration among others, which make this species a
relevant model organism (Araujo et al., 1993). In addition, it is worldwide distributed allowing
for comparisons of biological responses at a global scale (Pigneur et al., 2014). Its filtration
activity leads to the bioaccumulation of environmental contaminants to which it is sensitive,
making *C. fluminea* especially suitable for biomonitoring and assessing pollutants effects (Guo
and Feng, 2018). Its use in this study aimed to provide new sublethal ecotoxicological data of
REE on freshwater bivalves and more globally on benthic species, which despite their role in
ecosystem processes have been rarely used for regulatory purposes. Thus, *C. fluminea* was
exposed to Nd, Gd and Yb, which are representative of LREE, MREE and HREE, respectively.
These three REE are of specific concern because of their increasing uses (Nd in permanent
magnet), their occurrence in freshwaters (Gd is worldwide distributed) and their lack of
ecotoxicological data (Yb is one of the least studied REE) (Blinova et al., 2020; Gregoir, 2022;
Trapasso et al., 2021). We have compared their fate, bioaccumulation and biological effects to
decipher whether REE represent a uniform group. The three REE have been tested individually
and in mixture, in the absence and presence of DOM to reproduce complex exposure being
closer to environmental conditions because REE and DOM occur together in natural
freshwaters. In addition, the use of DOM can modify REE speciation, which is important to
assess its influence on REE bioaccumulation and toxicity in an environmentally realistic way.

In most previous works, REE bioaccumulation was measured in the whole soft tissues of the marine bivalve *Mytilus galloprovincialis* (Akagi and Edanami, 2017; Cunha et al., 2022; Freitas et al., 2020b; Henriques et al., 2019; Pinto et al., 2019). In the present study, Nd, Gd and Yb accumulation was investigated at the tissue (gills, digestive gland, rest of soft tissues, hemolymph) and subcellular (five sub-compartments of the gills and digestive gland) levels.

This represents the most exhaustive description of REE distribution in a bivalve species so far.
At the same time, a battery of biomarkers was measured in the gills, digestive gland and in the
hemolymph in order to analyse energetic reserves and metabolism, osmo- and iono-regulation,
nervous function, antitoxic and antioxidant defences and cellular damages in *C. fluminea*.
## 2. Material and Methods

### 2.1. Preparation of test medium, rare earth elements (REE) and dissolved organic matter (DOM)

The test medium was a reconstituted water composed of : 294 mg L⁻¹ CaCl₂, 2H₂O; 123 mg L⁻¹ MgSO₄, 7H₂O; 65 mg L⁻¹ NaHCO₃; 6 mg L⁻¹ KCl. It was previously oxygenated with an air pump and the temperature was adjusted to 9 ± 1°C to stay close to the temperature measured *in situ* during the bivalve sampling. The pH was lowered and adjusted at 7.3 (using HCl 3 %) to enhance REE solubility and exposure while remaining in the range of pH values measured at the bivalve sampling location. In order to limit losses of REE during the experiment due to their possible adsorption on glass walls, the flasks used to prepare the test medium and the test flasks (glass beaker, 600 mL capacity) were previously filled with the test medium and REE during at least 12 h and then emptied (Lachaux et al., 2022a). At the test beginning, the flasks were filled again with the test medium spiked with REE and DOM as described below.

REE stock solutions (5 g L⁻¹ REE salt) were prepared by dissolution in MilliQ water of chemicals (purity > 99 %) purchased from Sigma Aldrich, precisely: neodymium nitrate (Nd(NO₃)₃, 6 H₂O), gadolinium chloride (GdCl₃, 6 H₂O) and ytterbium chloride (YbCl₃, 6 H₂O). REE exposure concentration was obtained by dilution of the stock solution in the test medium.

The used DOM had been prepared at Wageningen University, the Netherlands, by purification of local groundwater to be used as drinking water using membrane filtration. Composition of the DOM stock solution in fulvic acids (>80 %), humic acids (>2 %) and hydrophilic acids (≈10 %) was determined using the batch procedure described in van Zomeren and Comans (2007). The tested dissolved organic carbon (DOC) concentration was 8 mg L⁻¹ C
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because it is widely representative of rivers and lakes (Thurman, 1985). This concentration was
reached by dilution of the DOM solution in the test medium (Lachaux et al., 2022a).

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**2.2. Experimental design and conditions**

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*Corbicula fluminea* individuals (2-2.5 cm) were collected in the Meurthe river, in Rosières-aux-Salines (Northeast of France), in February 2019. They were gradually acclimatised to lab conditions (test medium, temperature) and fed with green microalgae (*Raphidocelis subcapitata*, *Chlorella vulgaris* and *Scenedesmus subspicatus* at approximately 1.5 x 10⁷ cells ind⁻¹ day⁻¹) for 10 days.

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Afterwards, the organisms were exposed for 96 h in a beaker filled with 500 mL of test medium spiked with 1 mg L⁻¹ REE. These exposure concentrations and durations are representative of acute REE contamination due to mining (Liang et al., 2014) or industrial (Kulaksız and Bau, 2011) activities. Nd, Gd and Yb were tested individually and in ternary mixture (1/3 mg L⁻¹ of each REE) in order to assess the potential interactions between REE. Two different media were tested: in the absence (0 mg L⁻¹ C) or presence (8 mg L⁻¹ C) of DOM in order to estimate the influence of DOM on REE fate, bioaccumulation and biological effects. Controls (no REE) with and without DOM were included to the experiment in order to verify any direct influence of DOM (alone) on the results because DOM may affect organisms (Steinberg et al., 2008). For each experimental group (C, COM, Nd, NdOM, Gd, GdOM, Yb, YbOM, Mix, MixOM where C = control, Mix = mixture and OM = organic matter = 8 mg L⁻¹ C), 30 individuals were divided into 6 beakers (5 individuals per beaker).

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The test medium was changed daily to renew dissolved oxygen concentration and ensure semi-static exposure because REE concentrations tend to decrease significantly over time (Blinova et al., 2018; González et al., 2015; Romero-Freire et al., 2019). The pH was monitored for 24 h and confirmed that it was maintained at 7.3 ± 0.1. The photoperiod was natural

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corresponding to 11 h light / 13 h dark (in February in North East France). The bivalves were

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not fed during the experiment.

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**2.3. Analysed physiological functions**

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After 96 h of exposure and depending on biomarkers between six and 12 bivalves (per

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experimental group) were taken from the six beakers (= six replicates) and dissected to collect

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the gills, the digestive gland and the rest of soft tissues, which were immediately stored at -

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80°C until analyses. REE biological effects were assessed by measuring a battery of biomarkers

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in the gills and in the digestive gland of the bivalves (Figure 1). The samples were prepared

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according to Koehlé-Divo et al. (2018). Briefly, the organs were thawed and crushed with

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metallic beads with a beadbeater (30 sec) in phosphate buffer (50 mM, pH 7.6) supplemented

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with 1 mM phenylmethylsulfonyl fluoride (PMSF) and 1 mM L-serine borate mixture as

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protease inhibitor at 8:1 volume:weight ratio. The preparation was then centrifuged for 5

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minutes (min) at 4°C at 250 g. A part of the obtained supernatant (denoted homogenate) was

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used to measure protein (PROTHom), triglyceride (TRIGLY) and lipid hydroperoxide (LOOH)

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concentrations, acid phosphatase (ACP) and electron transport system (ETS) activities. The

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remaining supernatant was firstly centrifuged at 1000 g for 20 min and then at 20000 g for 50

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min at 4°C. The resulting supernatant corresponds to the cytosolic fraction used to measure

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protein (PROTCyt) and total antioxidant capacity (TAC) concentrations, lactate dehydrogenase

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(LDH), total glutathione peroxidase (GPx), superoxide dismutase (SOD), glutathione-S-

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transferase (GST) and caspase-3 (CSP-3) activities. These analyses were performed using the

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automated spectrophotometer analyser Konelab 20 XTi (Thermo Scientific) adapted from the

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protocols detailed by Garaud et al. (2016) and Bertrand et al. (2016).
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### 2.3.1. Energetic reserves and metabolism and nervous system

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Reserves represented by protein and triglyceride concentrations were measured using Thermo-Scientific Konelab ready-to-use reagents. ETS activity was determined according to the method described by Owens and King (1975), by monitoring formazan production at 480 nm resulting to the reduction of p-iodonitrotetrazolium as final electron acceptor in mitochondrial respiratory chain instead of O₂. LDH was determined by monitoring lactate to pyruvate oxidation by LDH, which reduces NAD to NADH, the latter being measured at 340 nm. ETS and LDH are indicators of aerobic and anaerobic metabolism, respectively.

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Cholinesterase (ChE) enzymes catalyse the breakdown of some choline esters (neurotransmitters) after nervous influx, which is essential for the well-functioning of nervous cells. REE are likely to influence nervous influx by replacing calcium (Ca) in the voltage operated Ca channels in nervous cells due the similarities of ionic radii and coordination numbers between REE and Ca (Valcheva-Traykova et al., 2014). This can lead to an inhibition of ChE activity (Das et al., 1988) as previously observed in *Mytilus galloprovincialis* exposed to La (Pinto et al., 2019) and Gd (Henriques et al., 2019). Prior to the experiment, we measured the activity of different pseudocholinesterase enzymes (acetylcholinesterase, butyrylthiocholinesterase, propionyl-thiocholinesterase) in *C. fluminea* gills. A higher signal was detected for propionyl-thiocholinesterase (PrSChE) activity, which was then selected for the rest of analyses. PrSChE activity was measured in the absence and presence of several ChE inhibitors (Iso-OMPA, serine, BW) in order to check the specificity of measured ChE activity. PrSChE activity was determined in the gills by photometry according to a protocol adapted from Ellman et al. (1961). Briefly, the gills were thawed and crushed with metallic beads using a bead beater (30 sec) in phosphate buffer (0.1 M, pH 8). The preparation was centrifuged at 1000 g (20 min) and then at 10000 g (30 min). PrSChE activity was determined in the final supernatant by measuring the absorbance of dithiobisnitrobenzoate (DNTB)-thiocholine
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complex after propionylthiocholine cleavage by PrSChE. PrSChE activity was expressed
according to protein concentration determined using the method from Bradford (1976), which
consists in measuring the absorbance at 595 nm of Coomassie blue complexed to proteins.

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## 2.3.2. Osmo- and iono-regulation

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Na⁺-K⁺ ATPase and plasma membrane Ca²⁺-ATPase (PMCA) were measured on total protein extract in the digestive gland and on enriched membrane extracts in gills. Digestive glands were suspended in 6 mL mg⁻¹ of fresh weight (FW) of Tris buffer (25 mM Tris pH 7.4, 1 mM PMSF) and crushed with metallic beads in a bead beater (30 sec). Homogenates were centrifuged at 2000 g for 10 min at 4°C, and then supernatants were centrifuged at 10000 g for 20 min at 4°C. The S10000 fractions were used to measure straight away protein concentration and ATPase activities. Membrane extracts were done according to the protocol of Santini et al. (2011). Briefly, gills were suspended in 6 mL mg⁻¹ FW in HEPES buffer (10 mM HEPES pH 7.2, 250 mM Sucrose, 1 mM EDTA, 1 mM PMSF) and crushed with metallic beads in a bead beater (30 sec). Gill's homogenate was centrifuged at 2000 g for 10 min at 4°C, then S2000 fraction was centrifuged at 10000 g for 20 min at 4°C and S10000 fraction was ultracentrifuged at 76000 g for 1 h at 4°C. Final pellet containing enriched membrane fraction was resuspended in 4 mL mg⁻¹ FW of Tris buffer. Biomarkers were measured straight away on the automated chemistry analyser Konelab 20 Xti. Protein concentrations were measured using Thermo-Scientific Konelab ready-to-use reagent. PMCA activity was measured on a modified protocol from Santini et al. (2011). Membrane fraction (6 µL) were incubated in a final volume of 120 µL for 20 min at 37°C in a buffer containing 25 mM Tris (pH 7.4), 100 mM NaCl, 20 mM KCl, 4 mM Na₂-ATP, 0.5 mM NaN₃, 0.2 mM ouabain and 10 mM of CaCl₂ for PMCA activity or 2 mM of EDTA to chelate Ca²⁺ free ion. The amount of Pi released was quantified at 700 nm by adding 120 µL of a solution containing 560 mM de H₂SO₄, 8.1 mM ammonium molybdate, 176
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mM FeSO₄. A standard curve of Na₂HPO₄ dilution is used to assess the phosphate concentration. The enzyme specific activity was expressed in mM Pi mg⁻¹ protein h⁻¹.

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Na⁺-K⁺ ATPase activity was assessed as the difference of ATP hydrolysis in the presence and absence of potassium, according to a modified protocol from Mosher et al. (2012). Membrane fractions (6 µL) were incubated at 37°C for 20 min in a final volume of 120 µL reaction mix (10 mM HEPES pH 7.2, 50 mM NaCl, 2.5 mM MgCl₂, 5 mM Na₂ ATP) with and without 10.5 mM KCl. Phosphate released was assessed as described above for PMCA activity. The enzyme specific activity was expressed in µM Pi mg⁻¹ protein h⁻¹.

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Bivalve hemolymph was sampled using minicaps (5 µL) and stored at -20°C until analyses. Sodium (Na), calcium (Ca), potassium (K) and magnesium (Mg) were quantified by inductively coupled plasma mass spectrometry (ICP-MS, iCAP-TQ, ThermoScientific) after hemolymph thawing and dilution in Milli-Q water.

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### 2.3.3. Antioxidant, antitoxic defences and cellular damages

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Several antitoxic defences were analysed to highlight any potential toxic stress induced by REE. GST activity was determined by monitoring, at 340 nm, CNDB decrease resulting to its conjugation with GSH being induced by GST (Habig et al., 1974). ACP was determined by measuring p-nitrophenol complexation at 400-420 nm resulting to 4-nitrophenol phosphate hydrolysis by ACP.

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Metallothionein (MT) concentrations were measured by polarography using the apparatus 797 VA Computrace (Metrohm®) based on the protocol from Mouneyrac et al. (1998). Briefly, tissues were ground on ice in two times the volume with the grinding tampon (Trizma® base + NaCl, pH 8.6). Then, they were centrifuged at 14000 rpm for 55 min at 4°C. 350 µL of supernatant (S1) were collected for the metal assay while the unsuspended pellet is frozen with the remaining supernatant. Tubes with S1 were then placed in a water bath for 15 min at 75°C

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and immediately incubated on ice (between 10 min and 1 h). Samples were then centrifuged for
10 min at 13000 rpm, 4°C and the supernatant was transferred into a new microtube and frozen
until analysis.

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The digestive lysosomal system morphology was analysed by histochemical revelation
of β-glucuronidase in unfixed cryostat sections according to Giambérini and Cajaraville (2005).
Image analysis was performed on digestive tissue sections (8 µm thick) using a microscope.
Four stereological parameters were calculated: surface density (S = S<sub>L</sub>/V<sub>C</sub>), volume density (V
= V<sub>L</sub>/V<sub>C</sub>), surface to volume ratio (S/V = S<sub>L</sub>/V<sub>L</sub>, interpreted as the inverse of lysosome size)
and numerical density (N = N<sub>L</sub>/V<sub>C</sub>) of lysosomes, with lower S, V and N refer to surface,
volume and number respectively and L, C correspond to lysosomes and cytoplasm, respectively.

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Activities of antioxidant enzymes were measured to decipher any oxidative stress. GPx
activity was measured by monitoring NADPH consumption at 340 nm according to Lawrence
(2012). SOD activity was determined by monitoring formazan production (inversely
proportional to SOD activity). TAC activity was measured by monitoring AzBTS⁺ reduction
by all antioxidant species including enzymes (GSH reductase, catalase, peroxidase), small
molecules (ascorbate, uric acid, GSH, vitamin E...) and proteins (e.g. albumin, transferrin)
(Erel, 2004).

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Lipid peroxidation is a cellular damage provoked by oxidative stress. It was measured by
monitoring Fe²⁺ oxidation by lipid hydroperoxides according to Arab and Steghens (2004).
CSP-3 activity was determined by measuring the fluorescence (emission 400 nm; excitation
505 nm) of 7-amino-4-trifluoromethylcoumarin (AFC) cleaved from Asp-Glu-Val-Asp
(DEVD) by CSP-3. The latter indicates cellular death.

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Some biomarkers were not measured in the gills because of an insufficient biomass.
Figure 1 summarizes the design of performed analyses.

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256 **2.4. REE concentration measurements and analysis**

257 Nd, Gd and Yb were quantified in the test medium and in the bivalves by ICP-MS (Nexion
258 350x PerkinElmer). Details concerning ICP-MS analyses can be found in Lachaux et al.
259 (2022a).

260 For one renewal cycle of the test medium (24h), REE concentrations were measured in
261 water immediately (t0), 1 h (t1), 6 h (t6) and 24 h (t24) after the successive addition of DOM,
262 REE and bivalves. For each experimental group, the test medium of three beakers (replicates)
263 was filtered at 0.45 µm using a mixed cellulose ester filter to dissociate the particulate phase
264 kept on the filter and the dissolved phase representing the filtrate. Determination of measured
265 dissolved (<0.45 µm, denoted [REEdiss]) and particulate (>0.45 µm, [REEpart]) concentrations
266 of REE is described in Lachaux et al. (2022a). We considered the sum of dissolved and
267 particulate concentrations as the total REE concentration ([REEtot]).

268 To investigate REE accumulation and organ distribution in the bivalves, Nd, Gd and Yb
269 were quantified in gills, digestive gland, rest of soft tissues and in the hemolymph by ICP-MS.
270 After 96 h of exposure, the bivalves were depurated for 96 h (Figure 1) in clean water (test
271 medium without REE) in order to remove the potential REE that are weakly bound to the
272 tissues. After dissection, the gills, digestive gland and the rest of soft tissues were immediately
273 stored at -80°C until analysis. Before analysis, the tissues were thawed and digested with 1 mL
274 of HNO3 65 % at 80°C.

275 In addition, REE were quantified in five subcellular fractions in the gills and the digestive
276 gland. Cellular debris (membranes, tissue fragments), metal rich granules (MRG), organelles
277 (i.e. nuclear, mitochondria, microsomes, lysosomes), heat-sensitive proteins (including a large
278 variety of enzymes) and heat-stable proteins (MT and MT-like proteins) were isolated based on
279 the protocol from Wallace et al. (2003). To do so, each tissue from individual bivalves was
weighed in Eppendorf tubes and grinded (~1:5 heaviest tissue to buffer ratio,) using cold TRIS buffer. The obtained homogenized tissue was first fractionated by centrifugation at 1450 g for 15 min producing a pellet (P1) containing tissue fragments and other cellular debris (e.g. membranes, nuclear, MRG) and a supernatant (S1') containing the cytosol. MRG were isolated from P1 through digestion of tissue fragments with NaOH and centrifugation at 5000 g (P2). The resulting supernatant (S2) contained metal associated with dissolved tissues. The 1450 g supernatants (S1') were collected in special ultracentrifuge tubes and ultracentrifuged at 100 000 g for 1 h to produce a pellet (P3) containing various organelles (i.e. nuclear, mitochondria and microsomes); the 100 000 g supernatant (S3) contained the cytosol which is collected in special ultracentrifuge tubes. The cytosol was then heat-denatured (80°C) for 10-15 min and then ultracentrifuged at 30 000 g for 10 min. The final pellet (P4) contained heat-sensitive proteins ('enzymes'), while the supernatant (S4) contained heat-stable proteins (MT and MT-like proteins). Each fraction was stored at -80°C until measurement. Investigation of subcellular distribution of metals affords to assess the part of accumulated metals that is likely to induce toxicity (Otero-Fariña et al., 2022). Organelles and heat-sensitive proteins (denoted "enzymes") are considered as "metal-sensitive fractions" (MSF) because metal accumulation in these fractions can lead to harmful effects (Wallace et al., 2003). On the contrary, metals accumulated in MRG and heat-stable proteins (denoted "MT") are considered "biologically detoxified metals" (BDM) (Wallace et al., 2003). It is still not clear whether metals accumulated in cellular debris should be included in MSF or BDM fractions because they contain tissue fragments, cell membranes and other cellular components of unknown consequence (Cardon et al., 2019).

The bioconcentration factor (BCF) was calculated in order to compare bioaccumulation between the different experimental groups using the formula:

$$BCF = \frac{[REE]ORG}{[REEdiss]mean} \qquad (2)$$
with [REE]ORG being the total REE concentration measured in the bivalve tissues (gills + digestive gland + rest of soft tissues).

REE mass balance was calculated in order to determine the distribution (proportion) of REE among the water (dissolved and particulate phases) and the bivalves as a function of total REE mass measured at the end of the experiment (Lachaux et al., 2022b).

Recovery of REE was estimated by calculating the ratio between the total mass of REE measured at the end of the experiment and the total mass theoretically introduced at t0. We consider the difference between theoretical (nominal) and real (measured) concentrations as significant if recovery is lower than 80 % or higher than 120 % (ISO, 2012). In that case, exposure concentrations of REE were corrected by determining the geometric mean concentration in accordance with Romero-Freire et al. (2019):

$$ [REE]mean = \sqrt{(C_{t0} \times C_{t24})} \qquad (1) $$

with Ct0 and Ct24 being the concentrations measured in the water at the beginning and at the end of medium renewal cycle (24 h), respectively. When using total concentration to calculate the geometric mean, the latter is denoted “[REEtot]mean” and when using dissolved concentration, it refers to “[REEdiss]mean”.

## 2.5. Data analysis

Statistical analyses were conducted with R software (version 3.6.1). Data from the different experimental groups were compared performing analysis of variance (AOV) when normality and homoscedasticity conditions were respected (parametric test) otherwise by performing Kruskal-Wallis (KW) test (non-parametric-test) in order to discriminate REE and/or DOM effect on a given variable (biomarker). Tukey HSD (Miller, 1981) and the multiple comparisons test from Siegel and Castellan (1988) (for parametric and non-parametric tests, respectively) were conducted as post-hoc tests to identify exactly which groups differ from each
other. Data normality and homoscedasticity were checked with Shapiro-Wilk and Bartlett tests respectively. In addition, Hedge's g statistic was calculated based on the mean and standard deviation of each variable (biomarker) with the package "metafor" (Viechtbauer, 2010). Effect size represents the magnitude of difference between experimental groups and was thus used as a complement to AOV and KW tests to quantify the difference between groups (Sullivan and Feinn, 2012). Effect sizes were calculated between exposed (Nd, NdOM, Gd, GdOM, Yb, YbOM, Mix, MixOM) and control groups (C, COM) to analyse REE effects respect to the control. Effect sizes were also calculated to compare mixed (Mix, MixOM) and individual effects of Nd, Gd and Yb. A given effect size is considered as very small if $0 \le g < 0.2$; small if $0.2 \le g < 0.5$; moderate if $0.5 \le g < 0.8$; large if $0.8 \le g < 1.3$; very large if $1.3 \le g$ (Sullivan and Feinn, 2012). The effect size is significant when confidence interval of g does not include 0. Data concerning subcellular distribution of REE was analysed by comparing medians of the different experimental groups and calculating their respective 95 % confidence intervals (CI) (BCa method, R=1000), using bootstrap resampling (Canty and Ripley, 2016). This method is more suitable and accurate for small samples (n = 3) (Cline, 2019). Differences in median values between groups were considered statistically significant when the 95 % CI excluded zero. Relationship between two variables was evaluated by performing a linear regression test of Pearson (when data normality is respected) otherwise Spearman using the whole data set for each variable (n = 60-120 individuals). A difference was considered as statistically significant in case of a p-value < 0.05. A principal component analysis (PCA) was performed to compare biological effects observed in the different experimental groups. The criterion of Kaiser-Meyer-Olkin (Kaiser and Rice, 1974) was calculated ($KMO = 0.72$) and confirmed that the used dataset was suitable to run factorial analysis. The PCA was based on the mean of biomarkers exhibiting significant differences between experimental groups (in AOV, KW or effect size tests) in order to enhance the variance explained by PCA factors (axis).
353

# 3. Results

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## 3.1. REE in the water

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REE concentrations measured at 0 h and 1 h were not significantly different. Then, we considered REE concentration at the beginning of the renewal cycle (t0) as the mean of the concentrations measured at 0 h and 1 h.

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The measured concentrations of Nd, Gd and Yb in water were significantly lower (<80 %) than nominal ones (cf. section 2.4). Total (dissolved + particulate) concentrations measured at the end of medium renewal cycle (t24) represented less than 80 % of nominal concentrations (Table 1). REE exposure concentrations were thus corrected by calculating the geometric mean of measured concentrations (= REE mean) (cf. section 2.4).

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Total and dissolved concentrations of Nd, Gd and Yb (individually and in mixture) decreased by 12-43 % from t0 to t24, whereas particulate concentrations increased by 37-130 % (Figure S1). The three REE (individually and in mixture) were mainly (77-93 %) measured in the dissolved phase and in a lesser extent in the particulate phase (6-20 %) (Figure S2). Total and dissolved concentrations of Gd were significantly lower than those of Nd and Yb. Dissolved concentrations of Yb were significantly higher than those of Nd and Gd whereas particulate Yb was lower than Nd. These results were observed with REE tested both individually (Table 1) and in mixture (Table S1) in the absence and presence of DOM.

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Recovery, total and dissolved concentrations of Nd, Gd and Yb tested individually (Table 1) and in mixture (Table S1) were significantly higher in the presence than in the absence of DOM. Particulate concentrations of REE were lower in the presence of DOM respect to in its absence (significant for Nd) (Table 1).
375 **3.2. REE concentrations in C. fluminea**

376 Significant concentrations of Nd, Gd and Yb were measured in the exposed bivalves
377 compared to the controls (not exposed to REE) ranging from 58 to 85 µg g⁻¹ and from 33 to 58
378 µg g⁻¹ in the absence and presence of DOM, respectively (Table 1), which represent 1-4 % of
379 total REE measured at the end of the experiment (Figure S2). Bioaccumulation of the three REE
380 were significantly lower in the presence of DOM compared to its absence. Bioconcentration
381 factors of Nd, Gd and Yb tested individually (Table 1) and in mixture (Table S1) were 2-3-fold
382 lower in the presence of DOM.

383 In the absence and presence of DOM, Yb bioaccumulation was significantly lower than
384 the one of Nd (Table 1). No significant differences were observed between Nd and Gd uptake.
385 Bioaccumulated concentrations of REE in mixture (73 and 38 µg g⁻¹ without and with DOM,
386 respectively) were relatively close to the average of individual accumulations of Nd, Gd and
387 Yb (73 and 48 µg g⁻¹ without and with DOM, respectively).

388 **3.3. Tissue and subcellular distribution of REE in C. fluminea**

389 In the absence of DOM, Nd, Gd and Yb (individually and in mixture) were mostly
distributed in the gills (63-74 %), then in the digestive gland (11-15 %), in the rest of soft tissues
390 (14-25 %) and in a lesser extent in the hemolymph (<1 %) (Figure 2; Table 1). Similar results
391 were observed in the presence of DOM. In all studied tissues, the concentrations of the three
392 REE were lower with than without DOM (Figure 2; Table 1).
393

394 In the gills and in the digestive gland, Nd, Gd and Yb (individually and in mixture) were
395 mainly (>90 %) distributed among the organelles, MRG and cellular debris (Figure 3). This
396 was observed in both the absence and presence of DOM. Less than 10 % of REE were
397 accumulated in the enzymes and metallothioneins (MT).
398

The proportion of REE in the organelles was lower in the digestive gland compared to the gills whereas REE proportion in MT was increased. The three REE were more accumulated in MT and less in organelles (in the digestive gland) in the presence of DOM than in its absence (significant for Nd). In the presence of DOM, REE were less accumulated in MRG and more in cellular debris than in the absence of DOM (significant for Nd and Yb). Neodymium was significantly less distributed in cellular debris (in the gills and digestive gland), enzymes (gills) and more in MRG (digestive gland) compared to Gd and Yb. Ytterbium (in the absence of DOM) had significantly higher concentrations in MT (in the gills and digestive gland) and enzymes (gills) than Nd and Gd. In the gills, Gd accumulated significantly more in MRG (with and without DOM) and less in organelles (with DOM) than Yb and Nd. No significant differences were observed between Nd, Gd and Yb tested in mixture. REE accumulated more in organelles and less in MRG when tested in mixture than individually (significant for Gd).

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### 3.4. Biological effects of REE and DOM

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Energetic reserves and metabolism, osmo- and iono-regulation and antitoxic defences were modulated in the bivalves exposed to REE compared to the controls (not exposed to REE) (Figure 4).

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Protein concentrations were lower in the gills of bivalves exposed to Nd and Gd in the presence of DOM. Bivalves exposed to Nd without DOM had lower triglyceride concentrations and an increased LDH activity in the gills. On the contrary, LDH activity was decreased in the bivalves exposed to the mixture, both with and without DOM.

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Neodymium, Gd and Yb, individually and in mixture (with and without DOM), decreased Na concentration in the hemolymph by 14 % in average (Table S2). The latter was negatively correlated to the REE concentrations measured in bivalve tissues and in the organelles of the digestive gland. Concentrations of K, Mg and Ca in the hemolymph were positively correlated

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to REE concentrations measured in the hemolymph and in the tissues. Ca²⁺ ATPase activity was 3-fold higher in the gills than in the digestive gland and Na⁺-K⁺ ATPase activity was detected only in the gills. Ca²⁺ ATPase activity was significantly decreased by 17 % in the gills of the bivalves exposed to Nd compared to the controls.

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GPx activity and MT concentrations were lower in bivalves exposed to Gd without DOM. In the presence of DOM, Nd, Gd and Mix increased or decreased MT concentrations in the gills and in the digestive gland. Metallothionein concentrations were positively correlated to REE concentrations in bivalve tissues. Neodymium, Yb, Mix and to a lesser extent Gd decreased lysosome volume (V) and/or surface (S) in the presence of DOM. Surface to volume ratio of lysosomes (S/V) was significantly higher in the bivalves exposed to Nd and Gd indicating a reduced size of organelles. V and S were positively correlated to the number of lysosomes (N) and negatively correlated to REE concentrations measured in bivalve tissues and in the enzymes and organelles of the digestive gland.

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Neither neurological function (PrSchE), antioxidant defences (except GPx) nor cellular damages were significantly affected in the bivalves exposed for 96 h to Nd, Gd and Yb. All results concerning percentage of biological effect (Table S2), biomarker measurements, statistical (Table S3) and correlation tests (Tables S4-S6) are available in Supplementary material.

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The principal component analysis (PCA) shows that V, S, S/V of lysosomal system and Na concentration in the hemolymph are the biomarkers that most differentiate control bivalves from the ones exposed to REE, especially to Nd (Figure 5). The gap between Nd and control is enhanced by Ca²⁺ ATPase activity and TRIGLY concentrations along the horizontal axis that explains most of data variance. The difference between Nd and control in the presence of DOM is also visible along the vertical axis driven by PROT and MT concentrations (Figure 5).

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The difference between mixed and individual effects of Nd, Gd and Yb was very small, small, moderate, large and very large for 25, 32, 21, 14 and 8 % of the 174 calculated effect sizes. Ten percent of the latter were significant corresponding to a lower LDH activity, S/V ratio, MT concentration and higher Ca²⁺ATPase, GPx activities, triglycerides, MT concentrations and lysosome volume (V) in the bivalves exposed to the mixture compared to individual Nd, Gd and Yb (Figure 6).

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The presence of DOM influenced significantly biomarker responses by increasing ETS activity, PROT and Na concentrations and by decreasing GPx activity, S/V, Ca and MT concentrations respect to the absence of DOM (Table S3), which is illustrated by the PCA (Figure 5). Effects of REE on lysosomal system (V, S, V/S) and Na concentrations in the hemolymph (described above) were stronger in the presence of DOM (Figure 4).
457 **4. Discussion**

458 **4.1. REE in the water**

459 The concomitant decrease of dissolved and increase of particulate REE in the test medium
460 should be a consequence of REE precipitation as observed in numerous studies (Blinova et al.,
461 2018; González et al., 2015; Romero-Freire et al., 2019). The difference between nominal and
462 measured concentrations is probably related to REE precipitation but also to REE adsorption
463 on glass walls (Crémazy et al., 2014; Weltje, 2002) despite the flask preconditioning (Lachaux
464 et al., 2022a). These observations highlight the relevance to consider measured concentrations
465 to characterize REE exposure and to avoid misinterpretation of observed biological effects
466 (Lachaux et al., 2022b). Precipitation and adsorption of REE could happen due to the high
467 reactivity of trivalent REE and to the low solubility of their complexes with carbonates,
468 phosphates, or hydroxides (Johannesson et al., 1995; Liu and Byrne, 1997).

469 The higher dissolved concentrations of Nd, Gd and Yb in the presence of DOM was also
470 demonstrated in previous works using similar medium composition (Lachaux et al., 2022a,
471 2022b). It is explained by REE-DOM complexation that maintain REE in the
472 dissolved/colloidal phase of the water enhancing their solubility (Johannesson et al., 2004;
473 Vukov et al., 2016). The influence of DOM on REE fate in freshwater is described elsewhere
474 (Catrouillet et al., 2019; Gangloff et al., 2014; Marsac et al., 2011, 2010; Sonke and Salters,
475 2006).

476 The three REE showed a common fate in the water balanced by some differences mostly
477 between Nd and Yb, which may be related to their atomic number. The higher dissolved
478 concentration and lower particulate concentration of Yb compared to Nd can be linked to the
479 higher solubility of HREE-carbonates than LREE-carbonate complexes (Elderfield et al., 1990;
480 Sholkovitz, 1995).
481 **4.2. REE concentrations in C. fluminea**

482 The amount of REE accumulated by the bivalves after 96 h demonstrates that a part of
483 REE was bioavailable despite the potential REE complexation and precipitation, which are
484 known to reduce their bioavailability (González et al., 2015; Romero-Freire et al., 2019).
485 Bioaccumulated concentrations of Nd, Gd and Yb determined in this study are relatively close
486 to the concentrations of Gd (45 µg g⁻¹) and samarium (Sm) (47 µg g⁻¹) measured in the
487 freshwater bivalve, *Dreissena polymorpha*, exposed to 1.25 mg L⁻¹ of Gd or Sm for 14 days
488 (Hanana et al., 2018, 2017). Similar results were reported in *Mytilus galloprovincialis*, which
489 accumulated 119 µg g⁻¹ (BCF = 142 L kg⁻¹) of La after 28 days of exposure at 1 mg L⁻¹ La (Pinto
490 et al., 2019). These results show that bivalves accumulate concentrations of different REE in
491 the same order of magnitude when they are exposed to similar concentration. The similar
492 bioaccumulated concentrations observed after different exposure durations (4, 14 and 28 days)
493 show that bivalves quickly accumulate REE from the beginning of the exposure period.
494 However, it should be noted that the different bivalve species were not exposed in the same
495 conditions (test medium, etc.).

496 The presence of DOM mitigated Nd, Gd and Yb (individually and in mixture) uptake due
497 to the formation of REE-DOM complexes, which reduced REE bioavailability, as previously
498 demonstrated with *D. magna* (Lachaux et al., 2022a) and *D. rerio* (Lachaux et al., 2022b). To
499 our knowledge, no other studies investigated the influence of DOM on measured
500 bioaccumulation of REE in animal species.

501 The lower bioaccumulation of Yb respect to Nd supports the lower accumulation of
502 HREE compared to LREE and MREE reported in several species during lab experiments (Khan,
503 2017; Lachaux et al., 2022a; Qiang et al., 1994; Yang et al., 1999) and *in situ* (Amyot et al.,
504 2017; Li et al., 2016; MacMillan et al., 2017). This could be linked to the higher stability of
505 HREE complexes, especially HREE-carbonate complexes (Pourret et al., 2007), which would
506

decrease HREE bioavailability relative to LREE (Neira et al., 2022). Another hypothesis may
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imply a different mode of bioaccumulation (e.g. different transporters) of HREE compared to
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LREE. The investigation of the distribution of Nd, Gd and Yb at tissue and subcellular levels
509
could help for a better understanding of REE internalisation pathway.

510

**4.3. Tissue and subcellular distributions of REE in C. fluminea**

511

Neodymium, Gd and Yb have a similar organotropism in *C. fluminea* that suggests a
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common pattern of bioaccumulation among the different REE, which has also been reported in
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fish (Hao et al., 1996). Neodymium, Gd and Yb were preferentially accumulated in the gills
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because these organs are directly exposed to the medium *via* filtration activity. Water was the
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only route of exposure (because bivalves were not fed during the experiment), which enhanced
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metal accumulation in the gills (Marigómez et al., 2002). Most of studies with bivalve species
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measured REE accumulation in the whole soft tissues and we found only two that quantified
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REE in different organs. In the freshwater bivalve, *Hyriopsis cumingii*, all REE were mainly
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distributed in the gills (Xu et al., 2022) whereas *C. fluminea* and *Dreissena rostriformis
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bugensis* accumulated higher concentrations of Gd in the digestive gland (Perrat, 2017).
521
However, the different form of Gd tested in the latest study (contrast-agent for magnetic
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resonance imaging against GdCl₃ in the present study) has probably influenced its distribution
523
in bivalves. We have also detected significant concentrations of Nd, Gd and Yb in the rest of
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soft tissues (including mantle, muscles, foot, siphon, etc.). Among these tissues, REE were
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mainly located in the mantle of *H. cumingii* (Xu et al., 2022) probably because it is the tissue
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of shell formation and calcification (Marigómez et al., 2002) where REE are likely to displace
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Ca due to their similar ionic radius size (0.83-1.03 pm) with Ca (1 pm) (Valcheva-Traykova et
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al., 2014). REE may be accumulated more precisely in inorganic calcium concretions
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(carbonates and/or phosphates) used to store Ca for shell formation. Indeed, the presence of
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REE has been observed in the shell of *C. fluminea* (Merschel and Bau, 2015). In the present

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study, we have observed that REE also occur in the hemolymph of *C. fluminea* being, to our
knowledge, the first demonstration of the presence of REE in internal fluids of aquatic animal.
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533 Bivalves have an opened circulatory system where the organs are immersed in hemolymph.
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Metals are transferred from an organ to another and eliminated *via* the hemolymph and
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circulating cells (Marigómez et al., 2002) explaining the presence of REE in the latter. The
occurrence of REE in all studied compartments show that they are dispersed in the whole body
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of *C. fluminea*.
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At subcellular level, the dominant distribution among the organelles (MSF), MRG
(BDM) and cellular debris shared by Nd, Gd and Yb suggests that they have common biotic
ligands. It is likely that the three REE do not have the same affinity for these ligands explaining
some differences in their distribution. All REE belong to the same category (class “a”) from the
empirical classification based on the atomic properties and the solution chemistry of metal ions
(Nieboer and Richardson, 1980). This classification separates the periodical table into three
categories: class “a” elements (REE, Ca, Na, K, Mg etc.) having a higher affinity for O-
containing ligands (OH-, CO32-, HPO42-, SO42-, etc.), class “b” elements (Hg, Cu, Pb, Ag, etc.)
binding preferentially to S-containing ligands and the third class corresponds to borderline
elements (Ni, Zn, Cd, As, etc.). Due to their similar physicochemical properties, an element
belonging to class “a” is able to replace another class “a” element from its endogenous binding
site (Nieboer and Richardson, 1980). Then, in MRG, REE can substitute Ca2+ due to their
similar ionic radius size (Brown et al., 1990; Lansman, 1990; Pałasz and Czekaj, 2000;
Valcheva-Traykova et al., 2014). As Nd (0.98 pm) and Gd (0.94 pm) have an ionic radius closer
to the one of Ca (1 pm) than Yb (0.87 pm), they may have a greater affinity for Ca2+-receptors
in the MRG, which could explain why they tend to be more accumulated than Yb in MRG
causing their depletion in other fractions.
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Subcellular distribution of REE in a bivalve species has never been analysed before.

Cardon et al. (2019) showed in the crustacean *D. magna* that yttrium was mostly (>75 %) recovered in the detoxified MRG, 6-18 % in the organelles (mitochondria, lysosomes) and the remaining yttrium was located in cellular debris and MT. The fish, *O. mykiss* exhibited a different distribution with most of yttrium in MSF, among which more than 50 % in the organelles (mitochondria, lysosomes and microsomes) and around 20 % in the fraction relative to enzymes. Less than 15 % of accumulated yttrium were associated to the detoxified MRG and MT and the rest was recovered in cellular debris. It has been demonstrated that *D. magna* accumulates more REE than fish species (Cardon et al., 2019; Lachaux et al., 2022b). Crustaceans may not limit REE uptake as most of accumulated metal can be detoxified whereas fish might regulate REE uptake as low levels of accumulated metal can be toxic. In the present study, the freshwater bivalve exhibited an intermediate pattern between crustacean and fish in terms of REE accumulated concentrations (Lachaux et al., 2022b) and subcellular distribution (Cardon et al., 2019) showing that species have different strategies to handle REE. The analysis of internal distribution of metals is useful for understanding the way they are handle by organisms and for deciphering their potential toxic mechanisms (Otero-Fariña et al., 2022).

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## 4.4. Biological effects of REE and DOM

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### 4.4.1. REE detoxification

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A relevant part of Nd, Gd and Yb was detoxified being accumulated in MRG (BDM). Freshwater mollusks have numerous and diverse mineral granules mainly composed of carbonates, phosphates, Ca, K and Mg. They perform several functions such as metal detoxification and regulation of pH and ion homeostasis, especially the one of calcium which is an essential element in bivalves (Marigómez et al., 2002). The high accumulation of REE in

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578 MRG is probably due to the presence of strong affinity ligands (carbonates, phosphates, Ca²⁺ receptors).
579

580 The lower accumulation of REE in the organelles (MSF) of the digestive gland compared
581 to the gills is related to the higher detoxification capacity of the digestive gland, which can
582 implement diverse detoxification and elimination mechanisms preventing metal accumulation
583 in MSF (Marigómez et al., 2002). This would explain the higher proportion of REE linked to
584 metallothioneins (MT) in the digestive gland compared to the gills. However, less than 7 % of
585 Nd, Gd and Yb were accumulated in MT, demonstrating that REE have less affinity for thiols
586 groups (contained in MT) than class “b” metals (Marigómez et al., 2002). Metallothionein
587 exhibited a contrasting pattern of response being increased or decreased in the presence of REE.
588 It has been demonstrated that MT induction by C. fluminea can be related to metal exposure
589 (Doherty et al., 1987), which is supported by the positive correlation that we observed between
590 MT and REE concentrations. However, Hanana et al. (2017) reported that MT expression in D.
591 polymorpha was stimulated at low concentration (50 µg L⁻¹) and inhibited at high concentration
592 (1.25 mg L⁻¹) of Gd. The authors suggested that MT inhibition was due to a strong interaction
593 of REE (especially HREE and MREE) with GSH of MT, which would prevent GSSG formation
594 and thus MT induction (Hanana et al., 2021).

595 In the present study, other defence enzymes (against toxic and oxidative stress) were not
596 induced after 4 days of exposure. After 28 days of exposure, REE enhanced enzymatic defence
597 mechanisms such as GST, SOD, GPx and catalase in M. galloprovincialis (Andrade et al., 2023;
598 Freitas et al., 2020b, 2020a; Pinto et al., 2019) and in D. polymorpha (Hanana et al., 2018,
599 2017). These results suggest that REE sequestration in MRG is an early response to REE
600 exposure, which is implemented in bivalves before enzymatic responses. However, it seems
601 that this detoxification process did not prevent the occurrence of some biological effects.
602

4.4.2. <u>REE effects on osmo- and iono-regulation</u>

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Homeostasis and osmoregulation are tightly regulated functions that maintain the well-
functioning of organisms. Even slight deviations of these constants from normal values
represent a disturbance. The decrease of hemolymphatic concentrations of Na (dominant cation
of osmoregulation) and the decreased Ca²⁺ATPase activity in the gills of exposed bivalves are
probably related to the accumulation of Nd, Gd and Yb in the gills. Indeed, gills represent the
main organ of osmo- and iono-regulation in aquatic animals (Griffith, 2017), which was
illustrated by the higher activities of Ca²⁺ATPase and Na⁺-K⁺ATPase measured in the gills
compared to the digestive gland. The presence of REE in cell debris suggests that they can be
present in cell membranes. As cell membranes contain numerous and diverse transporters
(including ion exchangers, cotransporters, ionic channels, ATPase proteins), with which REE
may have interacted, it could explain the observed effects on iono- and osmo-regulation. The
presence of REE in cells membranes can also be explained by REE binding to phosphate
moieties in membrane phospholipids (Nieboer and Richardson, 1980).

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The decrease of Ca²⁺ATPase activity in the gills was statistically significant only for Nd
possibly because of its closer ionic radius size with Ca compared to Gd and Yb leading to a
stronger interaction of Nd with Ca²⁺ receptors. The decrease of Na concentration in the
hemolymph of exposed bivalves can be also related to the close ionic radius size of REE (0.86-
1.03 pm) with Na (1.02 pm) leading to a disruption of Na receptors. A potential mechanism for
the lower Na concentrations in the hemolymph is a downregulation of the Na⁺K⁺-ATPase
activity, yet our results showed a decrease of these enzymes but it is not statistically significant
due to a high data variability.

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Effects of REE on osmo- and iono-regulation deserve more attention because they can
lead to a disruption of the metabolism of essential elements (namely Na and Ca) and affect the
well-functioning of the organism as already described in mollusc species exposed to other metals (Griffith, 2017; Mosher et al., 2012; Vercauteren and Blust, 1999). However, such effects have been poorly studied in aquatic animals exposed to REE, thus mechanisms remain to be deciphered as numerous and diverse processes are involved (ion channel, exchanger, carbonic anhydrase and lysosome activities, membrane properties, etc.).

### 4.4.3. Biological effects potentially associated to REE in metal sensitive fraction

A significant part of Nd, Gd and Yb was accumulated in the metal sensitive fraction relative to organelles (including nuclei, mitochondria, lysosomes, ribosomes), which induce harmful effects in bivalves.

Lysosomes perform essential functions in cells such as cellular waste digestion and homeostasis. They respond to intra- and extra-cellular stimuli adjusting continuously their number, position and size. The present results indicated a decrease of lysosomal system due to smaller lysosomes, which can be explained by the accumulation of Nd, Gd and Yb in lysosomes (retrieved in the organelle fraction) and their subsequent effects. Indeed, class "a" metals are present in lysosomes of mollusks (Marigómez et al., 2002). REE have possibly interacted with Ca²⁺ channels located in lysosome membranes, which regulate lysosome size (de Araujo et al., 2020). Thus, interactions of REE with lysosomes should disturb their turnover and especially the lysosome fusion process affecting their well-functioning (de Araujo et al., 2020). Consequently, the fundamental functions of lysosomes including signal transduction, vesicular trafficking, autophagy, exocytosis and membrane reparation (Feng and Yang, 2016) could have been disrupted inducing cellular stress and damages. Lysosome sensitivity to contaminants, including metals, has been described elsewhere (Giamberini, 1993; Regoli, 1992; Yancheva, 2017) but never in aquatic animals exposed to REE.
649

The decreased protein concentration could be related to REE accumulation in endoplasmic reticulum, Golgi apparatus or nuclei (retrieved in organelles), which may disturb protein synthesis. Protein decrease by REE has been never reported before contrary to previous studies showing that La and Gd increased protein concentration in *M. galloprovincialis* (Cunha et al., 2022; Henriques et al., 2019; Pinto et al., 2019) due to the production of enzymatic defences (GST, GPx, SOD, catalase) that do not seem to be activated in the present study.

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Aerobic metabolism (ETS activity located in mitochondria membranes) was not affected but REE modulated anaerobic metabolism by increasing or decreasing LDH activity. An increase of LDH activity could be associated to valve closure in order to limit REE exposure and prevent any stress. An increase of LDH can also indicate cellular and tissue damages (Chan et al., 2013), which would be in line with protein concentration decrease.

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### 4.4.4. REE mixture effects

661

In most cases, the biological effects of Nd, Gd and Yb in mixture were similar to individual effects, which is related to their relatively close subcellular distribution when tested individually and in mixture. The higher and lower proportions of individual REE in MRG (BDM) and organelles, respectively, compared to those of REE in mixture, reflect an enhanced detoxification of individual REE, which can be linked to the higher concentrations of Nd, Gd and Yb tested individually (1 mg L⁻¹) than in mixture (0.333 mg L⁻¹). The very close subcellular distribution of Nd, Gd and Yb in mixture suggests that they have common molecular targets. This would explain the similarity between individual and mixture effects suggesting additive effects of REE in mixture, which have previously been observed on *D. rerio* (Kang et al., 2022), *D. magna* and *R. subcapitata* (Lachaux et al., 2022b).

662

663

664

665

666

667

668

669

670

671

Yet, our results show also some differences between individual and mixed effects. The higher GPx activity and MT concentrations indicate an increase of defences and stress in the

672
bivalves exposed to the mixture compared to individual Gd and Yb, which could mean synergistic effects. On the contrary, the increase of Ca²⁺ATPase activity, S/V ratio, and the decrease of lysosome volume in the bivalve exposed to the mixture compared to individual REE suggest a lower impact of the mixture on osmoregulation and lysosomal system. Moreover, the lowered LDH activity, MT concentration and the higher TRIGLY concentrations observed in the bivalves exposed to the mixture is linked with a lower metal exposure and stress. These results indicate mitigation of harmful effects of REE in mixture, mainly in the presence of DOM, which is notable on the PCA. Antagonist interactions can be caused by a competition between the different REE for binding to common biotic ligands, which is supported by the similar subcellular distribution of Nd, Gd and Yb in mixture. A similar hypothesis had been suggested to explain antagonist effects of REE towards Hydra vulgaris (Hanana et al., 2022) and C. reinhardtii (Morel et al., 2021).

#### 4.4.5. DOM effects

Our results on the morphology of the lysosomal system indicate that lysosomes were larger in the presence of DOM suggesting that DOM was taken up by the bivalves and digested by lysosomes. The differences in REE subcellular distribution in the absence and presence of DOM suggest that at least a part of the REE-DOM complexes was accumulated. This would be consistent with the lower accumulation of REE in MRG (in the presence of DOM) because REE-DOM complexes are not likely to bind to carbonates, phosphates and/or Ca²⁺ receptors as much as free ion REE. REE-DOM complexes have possibly formed ternary complexes in cell membranes (Zhao and Wilkinson, 2015), which would support the higher accumulation of REE in cellular debris (including cell membranes) in the presence of DOM.

The presence of DOM and its possible uptake by the bivalves had direct effects on measured biomarkers. Digestion of DOM could increase metabolism explaining the higher
protein concentration and ETS activity in the digestive gland in the presence compared to the
absence of DOM. The decrease of GPx activity in the presence of DOM can be related to its
capacity to isolate superoxide radicals and other reactive oxygen species outside and inside cells
(Cardenas et al., 2011). A similar mechanism could explain MT decrease in the presence of
DOM with the capacity of the latter to scavenge metals, which would reduce metal abundance
(compared to in the absence of DOM) and thus the concentration of MT. These protective
effects of DOM against metal toxicity and oxidative stress can thus be considered as beneficial
for the organisms. The scavenging capacity of DOM for REE and other metals may also explain
the decrease of Ca concentration in the hemolymph. Complexation between Ca and DOM might
decrease Ca bioavailability and uptake. Indeed, the strong affinity of Ca for DOM lead to a
competition between Ca and REE for DOM complexation (Kautenburger, 2009). The increase
in Na concentration with DOM may be related to DOM interaction with Na or Na receptors.
Such an increase in the presence of DOM was also observed in *D. magna* (Glover and Wood,
2005) and *D. rerio* (Al-Reasi, 2012) and has been explained by Na uptake stimulation due to
DOM interactions on Na membrane transporters.

Dissolved organic matter can have an indirect effect on the observed REE effects. The
stronger decrease of Na concentration and lysosomal system size observed in bivalves exposed
to REE in the presence of DOM is probably due to the increase of lysosome volume and Na
concentration in the hemolymph induced by DOM (without REE) rather than an enhancement
of REE toxicity directly caused by DOM.

## 5. Conclusion

The combined analysis of REE distribution in the medium, in the organisms and their
biological effects at organ and subcellular levels allowed a thorough understanding of the
underlying mechanisms of REE accumulation and biological effects. We demonstrated by using
media with and without DOM that medium composition influences REE bioavailability and bioaccumulation, which highlights the relevance to consider speciation of REE in order to understand their bioaccumulation. The detailed analysis of REE accumulation, organotropism and subcellular distribution in the organisms allowed to explain the observed biological effects.

The presence of REE in gills, hemolymph and more precisely in cell membranes explain the observed effects on osmo- and iono-regulation, illustrated by the decrease of both Na concentrations in the hemolymph and Ca²⁺ATPase activity in the gills. These effects can affect the metabolism of Na and Ca, which are essential elements in bivalve physiology. Further, REE accumulation in the organelles of the digestive gland is related to the decreased size of lysosomal system, which could disturb lysosome and cell functioning. These early physiological effects of REE have so far been little described and deserve more attention.

Cellular and molecular approaches would be consistent to decipher the exact mechanisms of toxicity and the potential impacts on organisms. Concerning REE detoxification, sequestration in MRG appears to be an early process implemented in bivalves.

Our study on the fate, bioaccumulation and effects of Nd, Gd and Yb revealed several similarities between LREE, MREE and HREE such as 1) a close distribution in water and in organisms (measured concentrations of the three REE did not vary by a factor > 2); 2) their fate and bioaccumulation were similarly affected by the presence of DOM; 3) they accumulate mainly in the gills, MRG, organelles and cellular debris; 4) they induced common biological effects on lysosomes and Na concentration in the hemolymph. These similarities in their fate, bioaccumulation and biological effects suggest that REE share common abiotic and biotic ligands (phosphates, carbonates, hydroxides, Ca receptors, etc.) which is supported by different studies (Blinova et al., 2020; Huang et al., 2022; Morel et al., 2021; Tang and Johannesson, 2003; Wood, 1990). Abiotic ligands of REE determine their speciation and thus their bioavailability whereas biotic ligands influence their uptake and internal distribution, which are
746 directly linked to their biological targets and subsequent effects. It is likely that Nd, Gd and Yb
747 have different affinities for these ligands, in relation with their physicochemical properties
748 (including their ionic radius size), which could explain the observed differences between the
749 three REE. The relatively close fate, bioaccumulation and biological effects between individual
750 and mixed REE suggest limited interactions between Nd, Gd and Yb. Occasionally, mixture
751 effects were lower than individual effects suggesting antagonism, which could be due to
752 competition between the three REE for their common biotic ligands. These results show that an
753 additivity approach should be appropriately protective in a context of environmental risk
754 assessment of REE mixture.

755

756 Acknowledgements

757 We gratefully acknowledge technical assistance from the Pôle de compétences en biologie
758 environnementale and the Pôle de compétences en chimie analytique environnementale,
759 ANATELo, LIEC laboratory, UMR 7360 CNRS – Université de Lorraine. The authors
760 specifically thank Maximilien Beuret, Danièle Pauly and Marie Zaffino for their contributions
761 and Bert-Jan Groenenberg for providing the DOM stock solution.

762

763 Funding

764 This work was supported by the French National Research Agency through the ECOTREE
765 project [ANR-16-CE34-0012] and the national program “Investissements d'avenir” [ANR-10-
766 LABX-21-01/RESSOURCES21] with financial support by the ZAM LTSER Moselle for the
767 thesis of Nicolas Lachaux.

768
769 Copyright

770 Licence CC BY: https://creativecommons.org/licenses/by/4.0/

<figure><img src="image_2.png" /></figure>

771
772 References

773 Akagi, T., Edanami, K., 2017. Sources of rare earth elements in shells and soft-tissues of bivalves from Tokyo Bay. Marine Chemistry 194, 55–62.
774 https://doi.org/10.1016/j.marchem.2017.02.009

775 Al-Reasi, H.A., 2012. Quantifying the Direct and Indirect Effects of Dissolved Organic Matter (DOM) on Aquatic Organisms: Interaction with pH and Quality Measures. McMaster University.

779 Amyot, M., Clayden, M.G., MacMillan, G.A., Perron, T., Arscott-Gauvin, A., 2017. Fate and Trophic Transfer of Rare Earth Elements in Temperate Lake Food Webs. Environ. Sci. Technol. 51, 6009–6017. https://doi.org/10.1021/acs.est.7b00739

782 Andrade, M., Soares, A.M.V.M., Solé, M., Pereira, E., Freitas, R., 2023. Threats of Pollutants Derived from Electronic Waste to Marine Bivalves: The Case of the Rare-Earth Element Yttrium. Environmental Toxicology and Chemistry 42, 166–177.
784 https://doi.org/10.1002/etc.5508

786 Arab, K., Steghens, J.-P., 2004. Plasma lipid hydroperoxides measurement by an automated xylenol orange method. Analytical Biochemistry 325, 158–163.
788 https://doi.org/10.1016/j.ab.2003.10.022

789 Arienzo, M., Ferrara, L., Trifuoggi, M., Toscanesi, M., 2022. Advances in the Fate of Rare Earth Elements, REE, in Transitional Environments: Coasts and Estuaries. Water 14, 401. https://doi.org/10.3390/w14030401

792 Bertrand, C., Zalouk-Vergnoux, A., Giambérini, L., Poirier, L., Devin, S., Labille, J., Perrein-Ettajani, H., Pagnout, C., Châtel, A., Levard, C., Auffan, M., Mouneyrac, C., 2016. The influence of salinity on the fate and behavior of silver standardized nanomaterial and toxicity effects in the estuarine bivalve Scrobicularia plana. Environmental Toxicology and Chemistry 35, 2550–2561. https://doi.org/10.1002/etc.3428

797 Blaise, C., Gagné, F., Harwood, M., Quinn, B., Hanana, H., 2018. Ecotoxicity responses of the freshwater cnidarian Hydra attenuata to 11 rare earth elements. Ecotoxicology and Environmental Safety 163, 486–491. https://doi.org/10.1016/j.ecoenv.2018.07.033

800 Blinova, I., Lukjanova, A., Muna, M., Vija, H., Kahru, A., 2018. Evaluation of the potential hazard of lanthanides to freshwater microcrustaceans. Science of The Total Environment 642, 1100–1107. https://doi.org/10.1016/j.scitotenv.2018.06.155

803 Blinova, I., Muna, M., Heinlaan, M., Lukjanova, A., Kahru, A., 2020. Potential Hazard of Lanthanides and Lanthanide-Based Nanoparticles to Aquatic Ecosystems: Data Gaps, Challenges and Future Research Needs Derived from Bibliometric Analysis. Nanomaterials 10, 328. https://doi.org/10.3390/nano10020328

807 Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Analytical Biochemistry 72, 248–254. https://doi.org/10.1016/0003-2697(76)90527-3

810 Brown, P.H., Rathjen, A.H., Graham, R.D., Tribe, D.E., 1990. Chapter 92 Rare earth elements in biological systems, in: Handbook on the Physics and Chemistry of Rare Earths. Elsevier, pp. 423–452. https://doi.org/10.1016/S0168-1273(05)80135-7

813 Canty, A., Ripley, B., 2016. boot: Bootstrap R (S-Plus) Functions. R package version 1, 3–18.

814 Cardenas, N., Coballase-Urrutia, E., Gertrudis, B., Chaverri, J., Barragan, G., 2011. Antioxidant activity of fulvic acid: A living matter-derived bioactive compound. Journal of Food, Agriculture and Environment 9, 123–127.

817 Cardon, P.-Y., Triffault-Bouchet, G., Caron, A., Rosabal, M., Fortin, C., Amyot, M., 2019. Toxicity and Subcellular Fractionation of Yttrium in Three Freshwater Organisms: *Daphnia magna*, *Chironomus riparius*, and *Oncorhynchus mykiss*. ACS Omega 4, 13747–13755. https://doi.org/10.1021/acsomega.9b01238
821 Catrouillet, C., Guenet, H., Pierson-Wickmann, A.-C., Dia, A., LeCoz, M.B., Deville, S., 822 Lenne, Q., Suko, Y., Davranche, M., Catrouillet, C., Guenet, H., Pierson-Wickmann, 823 A.-C., Dia, A., LeCoz, M.B., Deville, S., Lenne, Q., Suko, Y., Davranche, M., 2019. 824 Rare earth elements as tracers of active colloidal organic matter composition. Environ. 825 Chem. 17, 133–139. https://doi.org/10.1071/EN19159

826 Chan, F.K.-M., Moriwaki, K., De Rosa, M.J., 2013. Detection of Necrosis by Release of Lactate 827 Dehydrogenase Activity, in: Snow, A.L., Lenardo, M.J. (Eds.), Immune Homeostasis: 828 Methods and Protocols, Methods in Molecular Biology. Humana Press, Totowa, NJ, pp. 829 65–70. https://doi.org/10.1007/978-1-62703-290-2_7

830 Cline, G., 2019. Nonparametric Statistical Methods Using R. Scientific e-Resources.

831 Cunha, M., Louro, P., Silva, M., Soares, A.M.V.M., Pereira, E., Freitas, R., 2022. Biochemical 832 alterations caused by lanthanum and gadolinium in Mytilus galloprovincialis after 833 exposure and recovery periods. Environmental Pollution 307, 119387. 834 https://doi.org/10.1016/j.envpol.2022.119387

835 Das, T., Sharma, A., Talukder, G., 1988. Effects of lanthanum in cellular systems: A review. 836 Biol Trace Elem Res 18, 201–228. https://doi.org/10.1007/BF02917504

837 Davranche, M., Gruau, G., Dia, A., Bouhnik-Le Coz, M., Marsac, R., Pédrot, M., Pourret, O., 838 2017. Rare Earth Elements in Wetlands, in: Trace Elements in Waterlogged Soils and 839 Sediments. pp. 135–162.

840 de Araujo, M.E.G., Liebscher, G., Hess, M.W., Huber, L.A., 2020. Lysosomal size matters. 841 Traffic 21, 60–75. https://doi.org/10.1111/tra.12714

842 Doherty, F.G., Failla, L., Cherry, S., 1987. Identification of a metallothionein-like, heavy metal 843 binding protein in the freshwater bivalve, Corbicula fluminea 8.

844 Elderfield, H., Upstill-Goddard, R., Sholkovitz, E.R., 1990. The rare earth elements in rivers, 845 estuaries, and coastal seas and their significance to the composition of ocean waters. 846 Geochimica et Cosmochimica Acta 54, 971–991. https://doi.org/10.1016/0016- 847 7037(90)90432-K

848 Ellman, G.L., Courtney, K.D., Andres, V., Featherstone, R.M., 1961. A new and rapid 849 colorimetric determination of acetylcholinesterase activity. Biochemical Pharmacology 850 7, 88–95. https://doi.org/10.1016/0006-2952(61)90145-9

851 Feng, X., Yang, J., 2016. Lysosomal Calcium in Neurodegeneration. Messenger (Los Angel) 852 5, 56–66. https://doi.org/10.1166/msr.2016.1055

853 Freitas, R., Cardoso, C.E.D., Costa, S., Morais, T., Moleiro, P., Lima, A.F.D., Soares, M., 854 Figueiredo, S., Águeda, T.L., Rocha, P., Amador, G., Soares, A.M.V.M., Pereira, E., 855 2020a. New insights on the impacts of e-waste towards marine bivalves: The case of the 856 rare earth element Dysprosium. Environmental Pollution 260, 113859. 857 https://doi.org/10.1016/j.envpol.2019.113859

858 Freitas, R., Costa, S., D Cardoso, C.E., Morais, T., Moleiro, P., Matias, A.C., Pereira, A.F., 859 Machado, J., Correia, B., Pinheiro, D., Rodrigues, A., Colónia, J., Soares, A.M.V.M., 860 Pereira, E., 2020b. Toxicological effects of the rare earth element neodymium in 861 Mytilus galloprovincialis. Chemosphere 244, 125457. 862 https://doi.org/10.1016/j.chemosphere.2019.125457

863 Gangloff, S., Stille, P., Pierret, M.-C., Weber, T., Chabaux, F., 2014. Characterization and 864 evolution of dissolved organic matter in acidic forest soil and its impact on the mobility 865 of major and trace elements (case of the Strengbach watershed). Geochimica et 866 Cosmochimica Acta 130, 21–41. https://doi.org/10.1016/j.gca.2013.12.033

867 Garaud, M., Auffan, M., Devin, S., Felten, V., Pagnout, C., Pain-Devin, S., Proux, O., Rodius, 868 F., Sohm, B., Giambérini, L., 2016. Integrated assessment of ceria nanoparticle impacts 869 on the freshwater bivalve Dreissena polymorpha. Nanotoxicology 10, 935–944. 870 https://doi.org/10.3109/17435390.2016.1146363
871 Giamberini, L., 1993. Etude des mécanismes de transport et de détoxication des métaux lourds
872 chez la moule d'eau douce Dreissena polymorpha : rôle des hémocytes et des organes
873 du système excréteur (histologie, ultrastructure, microanalyse) (phdthesis). Université
874 Paul Verlaine - Metz.

875 Giambérini, L., Cajaraville, M.P., 2005. Lysosomal responses in the digestive gland of the
876 freshwater mussel, Dreissena polymorpha, experimentally exposed to cadmium.
877 Environmental Research 98, 210–214. https://doi.org/10.1016/j.envres.2004.11.003

878 Glover, C.N., Wood, C.M., 2005. The Disruption of *Daphnia magna* Sodium Metabolism by
879 Humic Substances: Mechanism of Action and Effect of Humic Substance Source.
880 Physiological and Biochemical Zoology 78, 1005–1016.
881 https://doi.org/10.1086/432858

882 González, V., Vignati, D.A.L., Pons, M.-N., Montarges-Pelletier, E., Bojic, C., Giamberini, L.,
883 2015. Lanthanide ecotoxicity: First attempt to measure environmental risk for aquatic
884 organisms. Environmental Pollution 199, 139–147.
885 https://doi.org/10.1016/j.envpol.2015.01.020

886 Gregoir, L., 2022. Metals for Clean Energy: Pathways to solving Europe's raw materials
887 challenge. KU Leuven, Eurometaux.

888 Griffith, M.B., 2017. Toxicological perspective on the osmoregulation and ionoregulation
889 physiology of major ions by freshwater animals: Teleost fish, crustacea, aquatic insects,
890 and Mollusca. Environmental Toxicology and Chemistry 36, 576–600.
891 https://doi.org/10.1002/etc.3676

892 Gu, Y.-G., Gao, Y.-P., Huang, H.-H., Wu, F.-X., 2020. First attempt to assess ecotoxicological
893 risk of fifteen rare earth elements and their mixtures in sediments with diffusive
894 gradients in thin films. Water Research 185, 116254.
895 https://doi.org/10.1016/j.watres.2020.116254

896 Gu, Y.-G., Huang, H.-H., Gong, X.-Y., Liao, X.-L., Dai, M., Yang, Y.-F., 2022. Application of
897 diffusive gradients in thin films to determine rare earth elements in surface sediments
898 of Daya Bay, China: Occurrence, distribution and ecotoxicological risks. Marine
899 Pollution Bulletin 181, 113891. https://doi.org/10.1016/j.marpolbul.2022.113891

900 Gu, Y.-G., Wang, Y.-S., Jordan, R.W., Su, H., Jiang, S.-J., 2023. Probabilistic ecotoxicological
901 risk assessment of heavy metal and rare earth element mixtures in aquatic biota using
902 the DGT technique in coastal sediments. Chemosphere 329, 138592.
903 https://doi.org/10.1016/j.chemosphere.2023.138592

904 Guo, X., Feng, C., 2018. Biological toxicity response of Asian Clam (Corbicula fluminea) to
905 pollutants in surface water and sediment. Science of The Total Environment 631–632,
906 56–70. https://doi.org/10.1016/j.scitotenv.2018.03.019

907 Gwenzi, W., Mangori, L., Danha, C., Chaukura, N., Dunjana, N., Sanganyado, E., 2018.
908 Sources, behaviour, and environmental and human health risks of high-technology rare
909 earth elements as emerging contaminants. Science of The Total Environment 636, 299–
910 313. https://doi.org/10.1016/j.scitotenv.2018.04.235

911 Habig, W., Pabst, M., Jakoby, W., 1974. Glutathione S-Transferases - The first enzymatic step
912 in mercapturic acid formation. The Journal of Biological Chemistry 249.

913 Hanana, H., Gagné, F., Trottier, S., Bouchard, P., Farley, G., Auclair, J., Gagnon, C., 2022.
914 Assessment of the toxicity of a mixture of five rare earth elements found in aquatic
915 ecosystems in Hydra vulgaris. Ecotoxicology and Environmental Safety 241, 113793.
916 https://doi.org/10.1016/j.ecoenv.2022.113793

917 Hanana, H., Kowalczyk, J., André, C., Gagné, F., 2021. Insights on the toxicity of selected rare
918 earth elements in rainbow trout hepatocytes. Comparative Biochemistry and Physiology
919 Part C: Toxicology & Pharmacology 248, 109097.
920 https://doi.org/10.1016/j.cbpc.2021.109097
921 Hanana, H., Turcotte, P., André, C., Gagnon, C., Gagné, F., 2017. Comparative study of the
922 effects of gadolinium chloride and gadolinium - based magnetic resonance imaging
923 contrast agent on freshwater mussel, Dreissena polymorpha. Chemosphere 181, 197-
924 207. https://doi.org/10.1016/j.chemosphere.2017.04.073

925 Hanana, H., Turcotte, P., Dubé, M., Gagnon, C., Gagné, F., 2018. Response of the freshwater
926 mussel, Dreissena polymorpha to sub-lethal concentrations of samarium and yttrium
927 after chronic exposure. Ecotoxicology and Environmental Safety 165, 662-670.
928 https://doi.org/10.1016/j.ecoenv.2018.09.047

929 Hao, S., Xiaorong, W., Zhaozhe, H., Chonghua, W., Liansheng, W., Lemei, D., Zhong, L.,
930 Yijun, C., 1996. Bioconcentration and elimination of five light rare earth elements in
931 carp (Cyprinus carpio L.). Chemosphere 33, 1475-1483. https://doi.org/10.1016/0045-
932 6535(96)00286-X

933 Henriques, B., Coppola, F., Monteiro, R., Pinto, J., Viana, T., Pretti, C., Soares, A., Freitas, R.,
934 Pereira, E., 2019. Toxicological assessment of anthropogenic Gadolinium in seawater:
935 Biochemical effects in mussels Mytilus galloprovincialis. Science of The Total
936 Environment 664, 626-634. https://doi.org/10.1016/j.scitotenv.2019.01.341

937 Huang, Z., Gao, N., Zhang, S., Xing, J., Hou, J., 2022. Investigating the toxically homogenous
938 effects of three lanthanides on zebrafish. Comparative Biochemistry and Physiology
939 Part C: Toxicology & Pharmacology 253, 109251.
940 https://doi.org/10.1016/j.cbpc.2021.109251

941 International Organization for Standardization, 2012. Water Quality Determination of the
942 Inhibition of the Mobility of Daphnia Magna Strauss (Cladocera, Crustacean). ISO
943 6341: 2012.

944 Johannesson, K.H., Lyons, W.B., Stetzenbach, K.J., Byrne, R.H., 1995. The solubility control
945 of rare earth elements in natural terrestrial waters and the significance of PO43- and
946 CO32- in limiting dissolved rare earth concentrations: A review of recent information.
947 Aquat Geochem 1, 157-173. https://doi.org/10.1007/BF00702889

948 Johannesson, K.H., Tang, J., Daniels, J.M., Bounds, W.J., Burdige, D.J., 2004. Rare earth
949 element concentrations and speciation in organic-rich blackwaters of the Great Dismal
950 Swamp, Virginia, USA. Chemical Geology 209, 271-294.
951 https://doi.org/10.1016/j.chemgeo.2004.06.012

952 Kaiser, H.F., Rice, J., 1974. Little Jiffy, Mark Iv. Educational and Psychological Measurement
953 34, 111-117. https://doi.org/10.1177/001316447403400115

954 Kang, S., Guo, C., Ma, C., Mu, H., Liu, Z., Sun, L., 2022. Assessment of the biotoxicity of
955 lanthanides (La, Ce, Gd, and Ho) on zebrafish (Danio rerio) in different water
956 environments. Ecotoxicology and Environmental Safety 246, 114169.
957 https://doi.org/10.1016/j.ecoenv.2022.114169

958 Khan, A.M., 2017. Accumulation, uptake and bioavailability of are Earth Elements (REEs) in
959 soil grown plants from ex-mining area in Perak, Malaysia. Appl Ecol Env Res 15, 117-
960 133. https://doi.org/10.15666/aeer/1503_117133

961 Koehlé-Divo, V., Cossu-Leguille, C., Pain-Devin, S., Simonin, C., Bertrand, C., Sohm, B.,
962 Mouneyrac, C., Devin, S., Giambérini, L., 2018. Genotoxicity and physiological effects
963 of CeO 2 NPs on a freshwater bivalve (Corbicula fluminea). Aquatic Toxicology 198,
964 141-148. https://doi.org/10.1016/j.aquatox.2018.02.020

965 Kulaksız, S., Bau, M., 2011. Rare earth elements in the Rhine River, Germany: First case of
966 anthropogenic lanthanum as a dissolved microcontaminant in the hydrosphere.
967 Environment International 37, 973-979. https://doi.org/10.1016/j.envint.2011.02.018

968 Lachaux, N., Catrouillet, C., Marsac, R., Poirier, L., Pain-Devin, S., Gross, E.M., Giamberini,
969 L., 2022a. Implications of speciation on rare earth element toxicity: A focus on organic
970

matter influence in Daphnia magna standard test. Environmental Pollution 307, 119554.

971

https://doi.org/10.1016/j.envpol.2022.119554

972

Lachaux, N., Cossu-Leguille, C., Poirier, L., Gross, E.M., Giamberini, L., 2022b. Integrated

973

environmental risk assessment of rare earth elements mixture on aquatic ecosystems.

974

Frontiers in Environmental Science 10.

975

Lansman, J.B., 1990. Blockade of current through single calcium channels by trivalent

976

lanthanide cations. Effect of ionic radius on the rates of ion entry and exit. J Gen Physiol

977

95, 679–696. https://doi.org/10.1085/jgp.95.4.679

978

Lawrence, R.A., 2012. Reprint of “Glutathione Peroxidase Activity in Selenium-Deficient Rat

979

Liver.” Biochemical and Biophysical Research Communications 425, 503–509.

980

https://doi.org/10.1016/j.bbrc.2012.08.016

981

Li, J.-X., Zheng, L., Sun, C.-J., Jiang, F.-H., Yin, X.-F., Chen, J.-H., Han, B., Wang, X.-R.,

982

2016. Study on Ecological and Chemical Properties of Rare Earth Elements in Tropical

983

Marine Organisms. Chinese Journal of Analytical Chemistry 44, 1539–1546.

984

https://doi.org/10.1016/S1872-2040(16)60963-5

985

Liang, T., Li, K., Wang, L., 2014. State of rare earth elements in different environmental

986

components in mining areas of China. Environ Monit Assess 186, 1499–1513.

987

https://doi.org/10.1007/s10661-013-3469-8

988

Lin, Y.-T., Liu, R.-X., Audira, G., Suryanto, M.E., Roldan, M.J.M., Lee, J.-S., Ger, T.-R.,

989

Hsiao, C.-D., 2022. Lanthanides Toxicity in Zebrafish Embryos Are Correlated to Their

990

Atomic Number. Toxics 10, 336. https://doi.org/10.3390/toxics10060336

991

Liu, X., Byrne, R.H., 1997. Rare earth and yttrium phosphate solubilities in aqueous solution.

992

Geochimica et Cosmochimica Acta 61, 1625–1633. https://doi.org/10.1016/S0016-

993

7037(97)00037-9

994

MacMillan, G.A., Chételat, J., Heath, J.P., Mickpegak, R., Amyot, M., 2017. Rare earth

995

elements in freshwater, marine, and terrestrial ecosystems in the eastern Canadian

996

Arctic. Environ. Sci.: Processes Impacts 19, 1336–1345.

997

https://doi.org/10.1039/C7EM00082K

998

Manusadžianas, L., Vitkus, R., Gylytė, B., Cimmperman, R., Džiugelis, M., Karitonas, R.,

999

Sadauskas, K., 2020. Ecotoxicity Responses of the Macrophyte Algae Nitellopsis

1000

obtusa and Freshwater Crustacean Thamnocephalus platyurus to 12 Rare Earth

1001

Elements. Sustainability 12, 7130. https://doi.org/10.3390/su12177130

1002

Marigómez, I., Soto, M., Cajaraville, M.P., Angulo, E., Giamberini, L., 2002. Cellular and

1003

subcellular distribution of metals in molluscs: Distribution of Metals in Molluscs.

1004

Microsc. Res. Tech. 56, 358–392. https://doi.org/10.1002/jemt.10040

1005

Marsac, R., Davranche, M., Gruau, G., Bouhnik-Le Coz, M., Dia, A., 2011. An improved

1006

description of the interactions between rare earth elements and humic acids by

1007

modeling: PHREEQC-Model VI coupling. Geochimica et Cosmochimica Acta 75,

1008

5625–5637. https://doi.org/10.1016/j.gca.2011.07.009

1009

Marsac, R., Davranche, M., Gruau, G., Dia, A., 2010. Metal loading effect on rare earth element

1010

binding to humic acid: Experimental and modelling evidence. Geochimica et

1011

Cosmochimica Acta 74, 1749–1761. https://doi.org/10.1016/j.gca.2009.12.006

1012

Merschel, G., Bau, M., 2015. Rare earth elements in the aragonitic shell of freshwater mussel

1013

Corbicula fluminea and the bioavailability of anthropogenic lanthanum, samarium and

1014

gadolinium in river water. Science of The Total Environment 533, 91–101.

1015

https://doi.org/10.1016/j.scitotenv.2015.06.042

1016

Miller, R.G., 1981. Simultaneous Statistical Inference, Springer Series in Statistics. Springer,

1017

New York, NY. https://doi.org/10.1007/978-1-4613-8122-8
1018 Morel, E., Cui, L., Zerges, W., Wilkinson, K.J., 2021. Mixtures of rare earth elements show
1019 antagonistic interactions in Chlamydomonas reinhardtii. Environmental Pollution 287,
1020 117594. https://doi.org/10.1016/j.envpol.2021.117594

1021 Mosher, S., Cope, W.G., Weber, F.X., Shea, D., Kwak, T.J., 2012. Effects of lead on Na+, K+-
1022 ATPase and hemolymph ion concentrations in the freshwater mussel Elliptio
1023 complanata. Environ. Toxicol. 27, 268-276. https://doi.org/10.1002/tox.20639

1024 Mouneyrac, C., Amiard, J., Amiard-Triquet, C., 1998. Effects of natural factors (salinity and
1025 body weight) on cadmium, copper, zinc and metallothionein-like protein levels in
1026 resident populations of oysters Crassostrea gigas from a polluted estuary. Mar. Ecol.
1027 Prog. Ser. 162, 125-135. https://doi.org/10.3354/meps162125

1028 Neira, P., Romero-Freire, A., Basallote, M.D., Qiu, H., Cobelo-García, A., Cánovas, C.R.,
1029 2022. Review of the concentration, bioaccumulation, and effects of lanthanides in
1030 marine systems. Frontiers in Marine Science 9.

1031 Nieboer, E., Richardson, D.H.S., 1980. The replacement of the nondescript term 'heavy metals'
1032 by a biologically and chemically significant classification of metal ions. Environmental
1033 Pollution Series B, Chemical and Physical 1, 3-26. https://doi.org/10.1016/0143-
1034 148X(80)90017-8

1035 Otero-Fariña, A., Rétif, J., Métais, I., Poirier, L., Châtel, A., 2022. Relevance of cell
1036 subcompartmentalization techniques to predict adverse effects of metals in bivalves and
1037 fish. Ecological Indicators 144.

1038 Owens, T.G., King, F.D., 1975. The measurement of respiratory electron-transport-system
1039 activity in marine zooplankton. Mar. Biol. 30, 27-36.
1040 https://doi.org/10.1007/BF00393750

1041 Pałasz, A., Czekaj, P., 2000. Toxicological and cytophysiological aspects of lanthanides action.
1042 Acta Biochim Pol 47, 1107-1114. https://doi.org/10.18388/abp.2000_3963

1043 Perrat, E., 2017. Impacts environnementaux des agents de contraste à base de gadolinium :
1044 situation locale, approche cellulaire et in vitro. Université de Lorraine.

1045 Pigneur, L.-M., Etoundi, E., Aldridge, D.C., Marescaux, J., Yasuda, N., Van Doninck, K., 2014.
1046 Genetic uniformity and long-distance clonal dispersal in the invasive androgenetic
1047 Corbicula clams. Molecular Ecology 23, 5102-5116.
1048 https://doi.org/10.1111/mec.12912

1049 Pinto, J., Costa, M., Leite, C., Borges, C., Coppola, F., Henriques, B., Monteiro, R., Russo, T.,
1050 Di Cosmo, A., Soares, A.M.V.M., Polese, G., Pereira, E., Freitas, R., 2019.
1051 Ecotoxicological effects of lanthanum in Mytilus galloprovincialis: Biochemical and
1052 histopathological impacts. Aquatic Toxicology 211, 181-192.
1053 https://doi.org/10.1016/j.aquatox.2019.03.017

1054 Pourret, O., Davranche, M., Gruau, G., Dia, A., 2007. Competition between humic acid and
1055 carbonates for rare earth elements complexation. Journal of Colloid and Interface
1056 Science 305, 25-31. https://doi.org/10.1016/j.jcis.2006.09.020

1057 Qiang, T., Xiao-rong, W., Li-qing, T., Le-mei, D., 1994. Bioaccumulation of the rare earth
1058 elements lanthanum, gadolinium and yttrium in carp (Cyprinus carpio). Environmental
1059 Pollution 85, 345-350. https://doi.org/10.1016/0269-7491(94)90057-4

1060 Regoli, F., 1992. Lysosomal responses as a sensitive stress index in biomonitoring heavy metal
1061 pollution. Marine Ecology Progress Series 84, 63-69.

1062 Romero-Freire, A., Joonas, E., Muna, M., Cossu-Leguille, C., Vignati, D.A.L., Giamberini, L.,
1063 2019. Assessment of the toxic effects of mixtures of three lanthanides (Ce, Gd, Lu) to
1064 aquatic biota. Science of The Total Environment 661, 276-284.
1065 https://doi.org/10.1016/j.scitotenv.2019.01.155

1066 Santini, O., Chahbane, N., Vasseur, P., Frank, H., 2011. Effects of low-level copper exposure
1067 on Ca2+-ATPase and carbonic anhydrase in the freshwater bivalve Anodonta anatina.
1068 Toxicological & Environmental Chemistry 93, 1826-1837.
1069 https://doi.org/10.1080/02772240903217598
1070 Sholkovitz, E.R., 1995. The aquatic chemistry of rare earth elements in rivers and estuaries.
1071 Aquat Geochem 1, 1-34. https://doi.org/10.1007/BF01025229
1072 Siegel, S., Castellan, N.J., 1988. Nonparametric Statistics for the Behavioural Sciences, 2nd
1073 edn McGraw-Hill Book Company. New York 144-151.
1074 Sonke, J.E., Salters, V.J.M., 2006. Lanthanide-humic substances complexation. I.
1075 Experimental evidence for a lanthanide contraction effect. Geochimica et
1076 Cosmochimica Acta 70, 1495-1506. https://doi.org/10.1016/j.gca.2005.11.017
1077 Steinberg, C.E.W., Meinelt, T., Timofeyev, M.A., Bittner, M., Menzel, R., 2008. Humic
1078 substances: Part 2: Interactions with organisms. Environ Sci Pollut Res 15, 128-135.
1079 https://doi.org/10.1065/espr2007.07.434
1080 Sullivan, G.M., Feinn, R., 2012. Using Effect Size—or Why the P Value Is Not Enough. J Grad
1081 Med Educ 4, 279-282. https://doi.org/10.4300/JGME-D-12-00156.1
1082 Tai, P., Zhao, Q., Su, D., Li, P., Stagnitti, F., 2010. Biological toxicity of lanthanide elements
1083 on algae. Chemosphere 80, 1031-1035.
1084 https://doi.org/10.1016/j.chemosphere.2010.05.030
1085 Tang, J., Johannesson, K.H., 2003. Speciation of rare earth elements in natural terrestrial
1086 waters: assessing the role of dissolved organic matter from the modeling approach.
1087 Geochimica et Cosmochimica Acta 67, 2321-2339. https://doi.org/10.1016/S0016-7037(02)01413-8
1088 Thurman, E.M., 1985. Organic Geochemistry of Natural Waters. Springer Netherlands,
1090 Dordrecht. https://doi.org/10.1007/978-94-009-5095-5
1091 Trapasso, G., Chiesa, S., Freitas, R., Pereira, E., 2021. What do we know about the
1092 ecotoxicological implications of the rare earth element gadolinium in aquatic
1093 ecosystems? Science of The Total Environment 781, 146273.
1094 https://doi.org/10.1016/j.scitotenv.2021.146273
1095 Valcheva-Traykova, M., Saso, L., Kostova, I., 2014. Involvement of Lanthanides in the Free
1096 Radicals Homeostasis. CTMC 14, 2508-2519.
1097 https://doi.org/10.2174/1568026614666141203123620
1098 van Zomeren, A., Comans, R.N.J., 2007. Measurement of Humic and Fulvic Acid
1099 Concentrations and Dissolution Properties by a Rapid Batch Procedure. Environ. Sci.
1100 Technol. 41, 6755-6761. https://doi.org/10.1021/es0709223
1101 Vercauteren, K., Blust, R., 1999. Uptake of cadmium and zinc by the mussel Mytilus edulis and
1102 inhibition by calcium channel and metabolic blockers. Marine Biology 135, 615-626.
1103 https://doi.org/10.1007/s002270050662
1104 Viechtbauer, W., 2010. Conducting Meta-Analyses in R with the metafor Package. Journal of
1105 Statistical Software 36, 1-48. https://doi.org/10.18637/jss.v036.i03
1106 Vukov, O., Smith, D.S., McGeer, J.C., 2016. Acute dysprosium toxicity to Daphnia pulex and
1107 Hyalella azteca and development of the biotic ligand approach. Aquatic Toxicology
1108 170, 142-151. https://doi.org/10.1016/j.aquatox.2015.10.016
1109 Wallace, W., Lee, B., Luoma, S., 2003. Subcellular compartmentalization of Cd and Zn in two
1110 bivalves. I. Significance of metal-sensitive fractions (MSF) and biologically detoxified
1111 metal (BDM). Mar. Ecol. Prog. Ser. 249, 183-197. https://doi.org/10.3354/meps249183
1112 Wood, S.A., 1990. The aqueous geochemistry of the rare-earth elements and yttrium 28.
1113 Xu, X., Pan, B., Shu, F., Chen, X., Xu, N., Ni, J., 2022. Bioaccumulation of 35 metal(loid)s in
1114 organs of a freshwater mussel (Hyriopsis cumingii) and environmental implications in
1115 Poyang Lake, China. Chemosphere 307, 136150.
1116 https://doi.org/10.1016/j.chemosphere.2022.136150
1117 Yancheva, V., 2017. Lysosomal membrane stability and respiration rate in zebra mussel

1118 (Dreissena polymorpha) as biomarkers for ex situ heavy metal exposure. Period. Biol.

1119 119, 229–237. https://doi.org/10.18054/pb.v119i4.4715

1120 Yang, X., Yin, D., Sun, H., Wang, X., Dai, L., Chen, Y., Cao, M., 1999. Distribution and

1121 bioavailability of rare earth elements in aquatic microcosm. Chemosphere 39, 2443–

1122 2450. https://doi.org/10.1016/S0045-6535(99)00172-1

1123 Zhao, C.-M., Wilkinson, K.J., 2015. Biotic Ligand Model Does Not Predict the Bioavailability

1124 of Rare Earth Elements in the Presence of Organic Ligands. Environ. Sci. Technol. 49,

1125 2207–2214. https://doi.org/10.1021/es505443s

1126