# Molecular and genotoxic effects of carbamazepine and methylmercury on the gonads of Dreissena polymorpha

Clément Baratange, Tainá Rocha de Almeida, Jean Armengaud, Isabelle Bonnard, Elise David, Laurence Delahaut, Véronique Gaillet, Emmanuel Guillon, Mélodie Kielbasa, Jean-Luc Loizeau, et al.

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## To cite this version:

Clément Baratange, Tainá Rocha de Almeida, Jean Armengaud, Isabelle Bonnard, Elise David, et al. Molecular and genotoxic effects of carbamazepine and methylmercury on the gonads of Dreissena polymorpha. Aquatic Toxicology, 2025, 287, pp.107493. 10.1016/j.aquatox.2025.107493. hal-05252406

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# Molecular and genotoxic effects of carbamazepine and methylmercury on the gonads of Dreissena polymorpha

Clément Baratange<sup>a,b,*</sup>, Tainá Rocha de Almeida<sup>a</sup>, Jean Armengaud<sup>c</sup>, Isabelle Bonnard<sup>a</sup>, Elise David<sup>a</sup>, Laurence Delahaut<sup>a</sup>, Véronique Gaillet<sup>a</sup>, Emmanuel Guillon<sup>e</sup>, Mélodie Kielbasa<sup>c</sup>, Jean-Luc Loizeau<sup>d</sup>, Nicolas Maurin<sup>e</sup>, Stéphanie Sayen<sup>e</sup>, Claudia Cosio<sup>a</sup>

## ARTICLE INFO

**Keywords:**
Bivalve
Co-exposure
Proteomics
Genotoxicity
Toxicity pathways
Spermatogenesis
Gonads

## ABSTRACT

Carbamazepine (CBZ) and methylmercury (MeHg) are pervasive contaminants that share molecular toxicity pathways (e.g., neurotoxicity, oxidative stress, energy metabolism) and, in addition to their main toxicity target, are both suspected of reprotoxicity. Yet, the precise mechanisms behind these putative effects remain a subject of ongoing investigation. *Dreissena polymorpha* holds great potential for freshwater biomonitoring due to its wide distribution and filter-feeder lifestyle, but the influence of the reproductive stage for the biological response to pollutants is still understudied. Here, *D. polymorpha* males in the proliferation and maturation reproductive stages were exposed for 24 h to CBZ, MeHg and CBZ+MeHg. We combined the comet assay, proteomics and targeted RT-qPCR approaches to evaluate the impact of those exposures on the gonads. In the proliferation stage, both single and co-exposures modulated transcripts involved in antioxidant defenses. In the maturation stage, no significant modulation was observed with targeted approaches, most likely due to a higher background level masking the effect of pollutants. At both stages, differentially abundant proteins were involved in similar metabolic pathways, including the energy metabolism, antioxidant defenses and gonadal cell development. Proteomics also revealed modulation of proteins involved in genotoxicity pathways exclusively in the proliferation stage, aligning with the observation of an increase in genotoxic damage. MeHg exposure caused more alterations than co-exposure, in contrast with previous studies on other organs during similar exposure in *D. polymorpha*. This underlined the importance of assessing organ-specific responses to pollutants. Moreover, data confirmed a newly identified reprotoxic effect of MeHg, reinforcing the need to better consider the reproductive stage in biomonitoring programs using this species, and species with similar characteristics.

## 1. Introduction

Anthropogenic activities release a wide variety of pollutants into the aquatic environment, posing a threat to the health of biota. The impact of pollutants is assessed through biomonitoring using biological responses as biomarkers to evaluate their effects and determine ecosystem health. Indeed, the development of early-warning tools to anticipate risk for biota is essential for mitigating pollution before detrimental effects on ecosystems occur (Løkke et al., 2013; Spurgeon et al., 2010).

Here, we investigated the molecular toxicity pathways of two problematic and ubiquitous pollutants in aquatic ecosystems: carbamazepine (CBZ) and mercury (Hg) (e.g., Zhang et al., 2017). Carbamazepine is a persistent pharmaceutical so widely detected that it has been proposed as a marker of anthropogenic activities in aquatic systems (Dvory et al., 2018; Oldenkamp et al., 2019; Verlicchi et al., 2012; Wilkinson et al., 2022). Carbamazepine is an anti-epileptic drug measured in concentrations ranging from ng L⁻¹ to µg L⁻¹. In Europe, no dedicated regulation exists, but recommendations for an environmental quality standard

<footnote><sup>a</sup> Université de Reims Champagne-Ardenne, Université Le Havre Normandie, INERIS, Normandie Univ, UMR-I 02 SEBIO, Reims, France</footnote>

<footnote><sup>b</sup> present address: Nantes Université, Institut des Substances et Organismes de la Mer, ISOMer, UR 2160, Nantes, France</footnote>

<footnote><sup>c</sup> Université Paris-Saclay, CEA, INRAE, Département Médicaments et Technologies pour la Santé (DMTS), SPI, F-30200 Bagnols-sur-Cèze Cedex, France</footnote>

<footnote><sup>d</sup> Department F.-A. Forel for Environmental and Aquatic Sciences, and Institute for Environmental Sciences, University of Geneva, Boulevard Carl-Vogt 66, 1211, Geneva 4, Switzerland</footnote>

<footnote><sup>e</sup> Université de Reims Champagne-Ardenne, CNRS, ICMR UMR 7132, 51097 Reims, France</footnote>

<footnote>* Corresponding author.
E-mail address: clement.baratange@univ-nantes.fr (C. Baratange).</footnote>
(EQS) was proposed at 500 ng L⁻¹ CBZ (ETOX, 2011; Kase, 2010). Hg is a global priority pollutant with a long legacy and persistence in the aquatic environment that poses a threat to the public health, as recognized by European legislation, Agency for Toxic Substances and Disease Registry Priority List (2022) (https://www.atsdr.cdc.gov) and the Minamata Convention (Directive 2008/105/EC; Reg. EC 1881/2006). Methylmercury (MeHg) is formed from inorganic mercury (IHg) through microbial activity in aquatic environment, mostly under anoxic conditions (Capo et al., 2023). IHg and MeHg are the two main forms of Hg to which aquatic organisms are most frequently exposed, with the latter having a greater propensity to bioaccumulate in organisms and biomagnify through food webs, in addition to its higher toxicity when compared to IHg. Nonetheless MeHg toxicity is poorly studied in invertebrates, although they are important intermediates of food webs (Branfireun et al., 2020; Cosio et al., 2013; Yang et al., 2020). In Europe, United States, Brazil, and Japan, the minimum threshold established for total Hg (THg = IHg+MeHg) ranged from 70 to 2000 ng L⁻¹ in freshwater (Directive 2008/105/EC; Conselho Nacionaldo Meio Ambiente (CONAMA), 2025; United States Environmental Protection Agency U.S. EPA, 2025; Japan Ministry of The Environment, 2025). In contrast, measured THg concentrations in natural environments vary widely—from below 1 ng·L⁻¹ in non-contaminated sites (Constantino et al., 2022) to as high as 7900 ng·L⁻¹ in heavily polluted areas (Zhou et al., 2018).

Furthermore, CBZ and Hg have toxic effects on aquatic biota; however, studies on invertebrates at environmental concentrations are scarce yet (Baali and Cosio, 2022; Baratange et al., 2022, 2023). According to the literature, CBZ and Hg triggered similar toxicity pathways in bivalves leading to oxidative stress, genotoxicity, and alterations in energy metabolism (Almeida et al., 2015, 2020; Brandts et al., 2018; Chen et al., 2014; Coppola et al., 2017; Jiang et al., 2019; Morosetti et al., 2020; Pytharopoulou et al., 2013). Nevertheless, the toxicity of MeHg is less studied than that of IHg (Kershaw and Hall, 2019), and studies with more environmental relevance (e.g. lower concentrations, exposure with other classes of pollutant) are lacking. In this line, 1 and 7 d of CBZ and MeHg co-exposure resulted in more effects on antioxidant responses, targeted metabolites and cellular alterations, than single exposures in *D. polymorpha* during resting stage and post-spawning stages (Baratange et al., 2022, 2023). In addition to the expected toxicity targets of CBZ (primarily the nervous system) and MeHg (mainly the nervous and oxidative stress pathways), reprotoxic effects were also suspected. Indeed, CBZ delayed spermatogenesis in *D. polymorpha* (Magniez et al., 2018), while IHg modulated 13 genes involved in reproduction in the gonads of *Crassostrea angulata* (Zhang et al., 2013) and MeHg modulated proteins involved in reproduction in *Gammarus fossarum* (Cosio et al., 2021), suggesting a putative reprotoxicity of those pollutants.

In this context, the freshwater bivalve *Dreissena polymorpha* is identified as a potential species of interest in both ecotoxicology and freshwater biomonitoring programs, due to its wide distribution, sessile nature and high-filtering activity which promotes pollutant accumulation (Baratange et al., 2023; Binelli et al., 2015; Hani et al., 2021; Louis et al., 2021). *D. polymorpha* provides essential ecosystem services such as water purification and nutrient cycling (Burlakova et al., 2023) and serves as food sources for numerous species (Kobak and Kakareko, 2011; Molloy et al., 1997). As such, pollutants that affect their reproductive health may have broader ecological consequences. Nonetheless, the limited understanding of its intrinsic physiology hampers the interpretation of biomonitoring data (Amiard-Triquet et al., 2012). For example, it is known that responses to pollutants may be influenced by the reproductive stage (González-Fernández et al., 2016; Louis et al., 2021; Palais et al., 2011), but studies of the impact of pollutants on gonadic tissues are lacking. Thus, a current research priority is to enhance our understanding of *D. polymorpha* physiology during its reproductive cycle and in reproductive organs to improve subsequent data interpretation in biomonitoring programs. Indeed, molecular toxicity pathways in the gonads were rarely studied in bivalves, despite their critical role in

reproductive success, population dynamics and overall ecological balance. In this context, investigating the toxicity in the gonads is essential for gaining a better understanding of the mechanisms through which pollutants exert their harmful effects and for developing more effective mitigation strategies and pollution control measures. Using more relevant environmental scenarios will enhance our understanding of the impacts of pollution on aquatic ecosystems (e.g. Almeida et al., 2018; Baratange et al., 2022; Jiang et al., 2019).

Here, our goal was to identify and compare the genotoxic effect and the early molecular toxicity pathways in gonads of male *Dreissena polymorpha* during both the development of gonadic cells (proliferation) and maturation of gonadic cells until spermatozoa (maturation) reproductive stages, under environmental concentrations (10x EQS) of CBZ (5000 ng L⁻¹) and MeHg (700 ng L⁻¹) single and co-exposures, to mimic a contaminated area. We aimed to address the following questions: (i) Does co-exposure result in more alterations than single exposures? (ii) Does the reproductive stage influence the response to CBZ, MeHg and the co-exposure? (iii) What are the molecular toxicity pathways associated with CBZ, MeHg, and co-exposure in the male gonads?

## 2. Materials and methods

### 2.1. Mussel collection and acclimation

To obtain *D. polymorpha* individuals exhibiting different reproductive stages, bivalves were collected in February and June 2021 from the Lake of Der, France (48° 36' 22.02" N, 4° 46' 34.0" E). This artificial lake is located in Champagne area, a temperate climate zone, with an annual mean temperature of 10.7 ± 0.3 °C and 752 ± 20 mm of precipitation. The absence of urban or industrial influences contributes to its low pollutant concentrations. CBZ and MeHg concentrations in the Lake water were below the limit of detection (LOD). For both dates, individuals measuring 18–22 mm in length were sorted, and males were selected by gamete identification using microscopy (200x) on 0.1 to 0.2 mL of the gonad collected with a syringe. In the laboratory, to avoid bias due to temperature, mussels were gradually (1 °C/day) acclimated from field temperatures (8 and 16 °C, respectively) to the exposure temperature (13 °C) and were kept for one week in an acclimatation aquarium filled with 9 L of spring water (Cristaline Aurèle Spring, Table S1). Mussels were then placed in exposure beakers containing 45 mL of spring water per individual at 13 °C for one week to allow for byssus attachment before exposure. During acclimation, twice a week, water was renewed and feeding at 2 million algal cells per day per mussel performed with a 50/50 (v/v) mixture of *Nannochloropsis* spp. (Nanno 3600, Instant Algae) and Shellfish Diet 1800 (Planktowie), following the standard protocol routinely used in our laboratory.

After acclimation, gonads were collected and immediately fixed in Bouin's aqueous solution for 24 h to precisely assess through histology sexual maturity on both dates. Dehydration, tissue permeabilization, and paraffin embedding were performed as described previously (Baratange et al., 2022; Paris-Palacios et al., 2000, 2003). Sections of 5 µm were obtained using a Leica microtome, stained with nuclear fast red and pico-indigo carmine, and mounted on glass slides for histological examination (100x). Two sections containing the follicles were observed and analyzed per individual. For each follicle, the reproductive stage was determined to calculate a sexual maturity index for each individual, as described elsewhere (Gist et al., 1997; Louis et al., 2021; Palais et al., 2011). The gonadic index for animals in February (n = 10) and June (n = 5) was 2.46 ± 0.07 and 2.94 ± 0.06, respectively, corresponding to the gamete proliferation and maturation stages (Fig. 1).

### 2.2. Exposure and sampling

A short-term exposure was conducted to identify early toxicity pathways and compare the impact in gonadic tissue with previous data available on gills and the digestive gland (Baratange et al., 2022, 2023).
<figure><img src="image_4.png" /><figcaption>**Fig. 1.** Histological analyzes (100x) of gonads of *D. polymorpha* in February (A), and June (B), showing follicles in the proliferation and maturation stages, respectively.</figcaption></figure>

Exposure was conducted for 24 h using 5 beakers per condition. Spring water was renewed, and beakers were left untreated as a control (Ctl) or spiked with CBZ (Sigma-Aldrich), MeHg (CH₃HgCl; VHG Labs Manchester; 1 g L⁻¹), or with their co-exposure CBZ+MeHg. Immediately after spiking, 10 mL of water was filtered (0.45 µm, Sterivex) and effective concentrations of CBZ (8400 ± 600 ng L⁻¹) and MeHg (270 ± 16 ng L⁻¹) were determined, most likely due to glass adsorption. No mortality or spawning was observed during this study.

After exposure, animals were dissected. Five gonads were immediately used for Comet assay. Fifteen gonads from other individuals were immediately frozen in liquid nitrogen and stored at -80 °C for quantitative polymerase chain reaction (qPCR; *n* = 10) and label-free shotgun proteomics analyzes (*n* = 5).

## 2.3. Pollutant measurements

### 2.3.1. Determination of CBZ

Water samples were frozen at -20 °C until CBZ analysis by LC-MS/MS. The rests of the soft tissues of mussels were freeze-dried, weighed and ground by an 8-mm stainless steel bead during 8 min at 30 Hz (MM400, Retsch) and suspended in 10 mL of acetonitrile (ACN; Carlo Erba, France), 200 µL of heptane (Carlo Erba) and 2.5 mL of ammonium acetate 2 mol L⁻¹ (Sigma Aldrich) in ultrapure water (ALPHA Q 18.2 MΩ cm⁻¹). The suspension was centrifuged for 10 min at 927 g. The organic fraction was transferred in a 20 mL glass tube and submitted to centrifugal vacuum evaporation (5 mbar) at 30 °C and 500 g using a Genevac HT Series (SP Scientific, Warminster, PA, USA). The pellet was resuspended in 500 µL of ultrapure water/ACN 90/10 % (v/v) and vortex-mixed for 1 min before analysis. Each sample was analyzed by LC-MS/MS consisting of an Acquity H-Class UPLC with a quaternary pump and a Synapt G2-Si electrospray mass spectrometer (Waters, Milford, MA, USA). Each sample (10 µL) was eluted at 20 °C and at a flow rate of 0.45 mL min⁻¹ on an ACQUITY UPLC BEH C18 Column (2.1 × 50 mm, 1.7 µm, Waters) using a mobile phase containing ultra-pure water with 0.1 % formic acid (Sigma-Aldrich) (A) and ACN (B) in gradient mode from 90/10 % to 10/90 % (v/v) A/B in 3 min followed by a 3 min equilibrium at 90/10 % (v/v). Carbamazepine was detected in positive ion mode and MS/MS mode using the precursor ion *m/z* 237 and quantified using the product ion *m/z* 194 at a retention time of 2.8 min. The following settings for the Synapt G2 mass spectrometer were applied: source temperature 150 °C, desolvation 550 °C, nitrogen flow 1000 L h⁻¹, sampling cone 40 V, capillary voltage 3 kV, dwell 0.2 s and trap collision 20 V. Analyzes were performed in triplicate.

### 2.3.2. Determination of MeHg

Samples of water were acidified with 0.5 % v/v HCl (Suprapur, Merck) and stored at 4 °C until analyzed by cold-vapor atomic fluorescence spectroscopy (MERX-M, Brooks Rand Instruments). The rests of

soft tissues were freeze-dried, weighed and ground by an 8-mm stainless steel bead during 8 min at 30 Hz (MM400, Retsch). Tissues were digested for 12 h at 60 °C in 30 % v/v HNO₃ (Suprapur, Merck) following the MeHg procedure (US EPA 1630, 2001). The quality of the measurements was validated by analyzing blanks and certified reference materials (DORM-4, TORT-2; NRC - CNRC).

## 2.4. Targeted analyzes

### 2.4.1. Comet assay

Freshly dissected gonads (*n* = 5) were suspended in 500 µL of phosphate-buffered saline (PBS 0.1 M, pH 7.4), containing disperse II (0.8 M) previously activated at 37 °C to favor cellular dissociation. Gonadic cells were kept at 20 °C for 30 min in darkness and homogenized every 10 min. The larger cellular debris was removed with a plier, and gonadic cells were conserved at 4 °C under inactivin light to prevent additional DNA damage. An aliquot (10 µL) was mixed with 40 µL of PBS and 50 µL of 1 % agarose type VII (Low Melting Point) and maintained at 37 °C. Two samplings (30 µL) were deposited on a microscopic slide (Starfrost), previously recovered with 0.8 % agarose type I (Normal Melting Point) and a 18 × 18 mm coverslip. Slides were left on a frozen inox plate for 15 min to allow agarose solidification. Coverslips were then removed, and the cellular lysis was performed for 2 h at 4 °C in a solution at pH 10 (2.5 M NaCl, 100 mM Na₂EDTA, 10 mM Tris, 1 % Triton X-100, 10 % DMSO). Slides were rinsed with demineralized water and kept at 4 °C for 40 min. Denaturation was performed at 20 V, 300 mA for 24 min in an electrophoretic chamber filled with NaOH 10 N, Na₂EDTA 200 mM (pH > 13), and was efficient to distinguish DNA fragments from the intact DNA. Slides were washed 2 × 10 min in neutralization solution of Trizmabase 0.4 M (pH = 7.5) for dehydration, in absolute ethanol for 5 min, and dried at room temperature in darkness. The gels containing DNA were colored for 5 min with SYBR Green 1X solution (Sigma-Aldrich) prepared following manufacturer recommendations, rinsed in demineralized water for 5 min, and kept in the dark. The software Comet Assay IV (Perceptive Instruments) was run on automatic imagery EVOS to score and identify colored nucleotides. The percentage of tail DNA to total DNA was used as a proxy of DNA damage. For each individual, two gels were analyzed, and a total of 100 nucleoids per gel were scored.

### 2.4.2. Extraction of RNA, reverse transcription and quantitative PCR

Total RNA was extracted from the gonads using TRI Reagent following the manufacturer protocol (Euromedex). The absorbance at 260 and 280 nm (GenovaNano, Jenway) and electrophoresis in 0.5 % Tris Borate EDTA buffer on 1 % agarose gel determined the quantity, purity and quality of RNA. A Reverse Transcription (RT) was performed with 400 ng RNA, the verso cDNA synthesis kit (Thermo Scientific) and oligodT primers using a PCR Mastercycle (Eppendorf) for 1 h at 42 °C,
followed by 3 min at 95 °C. Finally, 3 µL of 1/10 diluted cDNA was used for qPCR using Absolute Blue qPCR SYBR Green kit (Thermo Scientific) and a CFX96 device (BioRad) for 15 min at 95 °C, followed by 40 cycles at 95 °C for 10 s and 60 °C for 45 s. Primers amplifying genes involved in defense responses, energy metabolism and reference genes were used (Table S2). Relative transcript levels were calculated by the 2^ΔΔCt^ method (Schmittgen and Livak, 2008). Two individuals were sampled per beaker and outliers were filtered using the interquartile range method to provide a more accurate reflection of the range of individual responses (Mramba et al., 2024). This strategy was chosen because biological variability was greater than exposure variability, allowing for a more robust representation of the data.

## 2.5. Shotgun proteomics

### 2.5.1. Data acquisition

Gonads were freeze-dried, ground by two 2-mm steel beads and a Precellosy instrument (Bertin) operated for 3 cycles of 20 s at 5000 rpm. The powder was pelleted by centrifugation for 1 min at 1500 g, weighted, dissolved in 12.5 µL of LDS1X (Thermo) per mg of powder and ground again for 6 cycles of 20 s at 7200 rpm. The solution was ultrasonicated for 30 s, with a Hielscher UP50H sonicator at 90 % amplitude, and then bead-beated for 4 cycles at 6500 rpm. Debris was eliminated in each sample by centrifugation for 3 min at 10,000 g, the resulting supernatant was diluted by 7 vol of LDS1X, and proteins (30 µL) were denatured for 5 min at 99 °C. They were subjected to short electrophoresis on a NuPAGE 4–12 % gradient gel ran in MES buffer (Thermo) as previously described (Hartmann et al., 2014). For each proteome, a single polyacrylamide gel slice and proteolysis with trypsin (Roche) in presence of 0.01 % ProteaseMAX surfactant (Promega) was performed, as described by Hartmann et al. (2014). Data acquisition was applied as in Baratange et al. (2023). For this, 200 ng of the resulting peptides were analyzed with a Q-Exactive HF mass spectrometer (Thermo) operated in data-dependent mode with the same parameters as previously described (Klein et al., 2016). Peptides were resolved on an Acclaim PepMap100 C18 (3 µm, 100 Å, 75 µm id x 50 cm) at 0.2 µL per min as flow rate along a 90 min gradient of acetonitrile and 0.1 % formic acid (3.2 %–20 % for 75 min, then 20 %–32 % for 15 min). A dataset of 2327,845 high-resolution MS/MS spectra were recorded. The dataset was queried with the MASCOT software (MatrixScience) against an RNA-seq derived database named Dreissena_HJ1_pep_20,170,801.fasta to select proteins for creating a second database, MS21–038-CARMA_Cosio300seq_20, 210,830.fasta in which 300 well annotated protein sequences were added. This proteogenomic cascade strategy allowed the interpretation of 1420,639 spectra. These peptides-to-spectrum matches (PSMs) listed 39,724 peptide sequences, and to validate the presence of 4761 proteins in total, containing at least 2 different peptides, including at least one unique among all proteins. Proteins were quantified by their spectral counts (Table S3). For each date of exposure, four comparisons were applied according to the TFOLD procedure as described (Carvalho et al., 2012): CBZ vs Ctl; MeHg vs Ctl; CBZ+MeHg vs Ctl.

The mass spectrometry proteomics data was deposited to the ProteomeXchange Consortium via the PRIDE partner repository with the dataset identifier PXD054018 and 10.6019/PXD054018.

### 2.5.2. Protein annotation and enrichment analysis

We identified 3539 and 3623 proteins detected with at least 5 spectral counts in the proliferation and maturation stages, respectively. These proteins were annotated by sequence similarity search with the EggNOG v5.0 software, applying default parameters (auto adjust per query, transfer annotation from any ortholog, E-value threshold set at 10⁻³, minimum of 40 % identity and of 20 % of query and subject coverage). Enrichment analysis was performed for each exposure on annotated differentially abundant proteins (DAPs) compared to the reference list. This was conducted using the WEB-based Gene Set Analysis Toolkit (WebGestalt), Kyoto Encyclopedia of Genes and

Genomes (KEGG) database and gene ontology (GO) terms for identified proteins. Redundancy reduction was then applied to identify the main KEGG pathways and GO terms, using the weighted set cover method (DAPs ≥2 and p-value <0.05; Table S4). Non-relevant KEGG pathways and GO-terms (e.g. ko05016 Huntington's disease, ko05012 Parkinson's disease) for *D. polymorpha* were excluded. To identify and discuss the most significant effects, KEGG pathways presenting ≥3 DAPs and GO-terms with ≥5 DAPs were selected. The 30 pathways and terms with the highest Enrichment Ratio (ER), calculated as the number of observed DAPs divided by the expected number of DAPs for each pathway or term, were retained.

## 2.6. Data analyzes

Data were presented as arithmetic mean ± standard error of the mean (SEM). The bioaccumulation factor (BAF; L kg⁻¹) for CBZ and MeHg was calculated by dividing the pollutant concentration in mussels (µg kg⁻¹) by the effective concentration of the pollutant in water (µg L⁻¹). A two-way ANOVA was performed to assess the effects of exposure and stage on bioaccumulation, qPCR and proteomics data, followed by TukeyHSD post-hoc analysis with Benjamini-Hochberg corrections (α = 5 %). Homoscedasticity and normality of residuals (α = 5 %) were verified using Bartlett and Kolmogorov-Smirnov tests in Rstudio. Differences were considered significant compared to respective Cts in the proliferation and maturation stages when *p*-value was <0.05 and fold change (FC) was ≥1.40 or <0.71. *InteractiVenn* was used to create Venn diagrams of DAPs (Heberle et al., 2015). Mixed linear model was used for genotoxicity data, and Dunnett's post hoc analysis with Benjamini-Hochberg corrections were applied through *lsmeans* package (α = 5 %). Principal component analysis (PCA) on log₂ FC was performed in Rstudio using the *ade4* and *factoextra* packages, and bubble plots were generated with the *ggplot2* package. PCA results were validated through permutation analysis.

# 3. Results

## 3.1. Bioaccumulation

Carbamazepine was significantly bioaccumulated in single and co-exposures at both stages (*p* < 2e⁻¹⁶; Fig. 2). Carbamazepine bio-accumulation was significantly 1.5 x higher in the maturation vs proliferation stage (*p* = 0.002); however, BAFs were similar (*p* ≥ 0.05), indicating that it was directly related to the effective concentration of CBZ in the water. In the maturation stage, both a significant 1.5 x higher CBZ bioaccumulation and BAF were measured in the co-exposure vs single exposure (*p* < 0.0001; Fig. 2), suggesting an increased uptake of CBZ in the presence of MeHg, a trend not observed in the proliferation stage.

Methylmercury in single and co-exposures was significantly bio-accumulated (*p* < 2e⁻¹⁶) with similar concentrations and BAF observed in both stages (*p* ≥ 0.05; Fig. 2). As such, MeHg bioaccumulation was not modulated by the presence of CBZ at both dates.

## 3.2. Targeted responses

In the proliferation stage, MeHg single and co-exposures significantly increased DNA damage 2.3 (*p* = 0.0001) and 2.1 x (*p* = 0.0008), respectively, while CBZ single exposure had no significant effect. In the maturation stage, CBZ single and co-exposures significantly reduced DNA damage 1.5 x (*p* = 0.0044) and 2.1 x (*p* < 0.0001), respectively, while MeHg single exposure had no significant effect (Table 1; Figure S1).

In the proliferation stage, only the exposure to CBZ downregulated the transcript levels of catalase 2.4 x (cat; *p* = 0.013), glutathione-s-transferase 1.7 x (gst; *p* = 0.041) and 2.4 x citrate synthase (cs; *p* = 0.033) (Table 1; Figures S2, S3), suggesting ROS modulation, and
<figure><img src="image_5.png" /><figcaption>**Fig. 2.** Bioaccumulation of CBZ (A) and MeHg (B) in rests of soft tissues of male *D. polymorpha* during CBZ, MeHg and their co-exposure in the proliferation (filled green) and maturation (red stripes) stages. Values are presented as mean ± SEM. Significant differences (*p* < 0.05; *n* = 5) compared to respective controls are indicated by an asterisk (*). BAF: bioaccumulation factor; LOD: limit of detection; dw: dry weight.</figcaption></figure>

<caption>**Table 1**</caption>

Tail DNA (%), relative transcript levels in the proliferation (Prolif.) and maturation (Matur.) stages in gonads of male *D. polymorpha* exposed to CBZ, MeHg and their co-exposure (*n* = 8 to 10). Significant differences (*p* < 0.05) compared to respective Ctls are indicated by an asterisk (*).

<table><thead><tr><th rowspan="2">Gene</th><th colspan="2">Ctl</th><th colspan="2">CBZ</th><th colspan="2">MeHg</th><th colspan="2">CBZ+MeHg</th></tr><tr><th>Prolif.</th><th>Matur.</th><th>Prolif.</th><th>Matur.</th><th>Prolif.</th><th>Matur.</th><th>Prolif.</th><th>Matur.</th></tr></thead><tbody><tr><td>%Tail<br>DNA</td><td>11.8 ± 1.2</td><td>32.6 ± 1.9</td><td>19.6 ± 1.3</td><td>21.2 ± 1.3*</td><td>26.8 ± 2.4*</td><td>27.6 ± 1.44</td><td>24.9 ± 1.4*</td><td>15.4 ± 1.4*</td></tr><tr><td>cat</td><td>1.00±0.17</td><td>0.70±0.15</td><td>0.42±0.07*</td><td>0.94±0.24</td><td>1.41±0.22</td><td>0.97±0.22</td><td>1.72±0.18</td><td>1.10±0.24</td></tr><tr><td>gst</td><td>1.00±0.07</td><td>0.91±0.15</td><td>0.59±0.05*</td><td>0.98±0.12</td><td>0.99±0.07</td><td>1.00±0.15</td><td>0.96±0.13</td><td>1.08±0.14</td></tr><tr><td>mt</td><td>1.00±0.18</td><td>3.18±1.03</td><td>0.60±0.28</td><td>3.66±0.85</td><td>1.16±0.21</td><td>3.52±1.23</td><td>1.63±0.49</td><td>3.87±1.20</td></tr><tr><td>sod</td><td>1.00±0.14</td><td>1.62±0.17</td><td>0.57±0.08</td><td>1.29±0.33</td><td>1.41±0.17</td><td>1.39±0.19</td><td>1.19±0.23</td><td>1.20±0.21</td></tr><tr><td>ampk</td><td>1.00±0.10</td><td>3.29±0.44</td><td>1.40±0.39</td><td>4.13±1.00</td><td>0.76±0.25</td><td>3.85±0.97</td><td>1.02±0.25</td><td>3.33±0.46</td></tr><tr><td>aox</td><td>1.00±0.18</td><td>2.38±0.34</td><td>0.83±0.29</td><td>2.38±0.77</td><td>0.83±0.31</td><td>2.05±0.82</td><td>1.05±0.20</td><td>1.87±0.28</td></tr><tr><td>atpase</td><td>1.00±0.13</td><td>0.84±0.13</td><td>0.87±0.15</td><td>0.95±0.16</td><td>1.03±0.11</td><td>0.89±0.13</td><td>0.74±0.08</td><td>0.81±0.13</td></tr><tr><td>ccox</td><td>1.00±0.16</td><td>1.29±0.22</td><td>0.93±0.18</td><td>1.69±0.24</td><td>1.05±0.11</td><td>1.92±0.49</td><td>1.26±0.15</td><td>1.40±0.47</td></tr><tr><td>cs</td><td>1.00±0.24</td><td>0.78±0.14</td><td>0.42±0.08*</td><td>0.74±0.22</td><td>1.44±0.13</td><td>0.73±0.16</td><td>1.52±0.19</td><td>0.92±0.17</td></tr><tr><td>ldh</td><td>1.00±0.21</td><td>1.08±0.22</td><td>0.95±0.20</td><td>0.97±0.16</td><td>1.21±0.21</td><td>0.92±0.15</td><td>0.99±0.15</td><td>0.87±0.15</td></tr></tbody></table>

aerobic energy metabolism alterations. No significant modulation was measured in the maturation stage. Transcripts of metallothionein (mt), superoxide dismutase (sod), AMP kinase (ampk), alternative oxidase (aox), ATP synthase (atpase), Cytochrome C oxidase (ccox) and lactate dehydrogenase (ldh) were not significantly changed by exposures.

### 3.3. Shotgun proteomics

#### 3.3.1. Proliferation stage

Carbamazepine, MeHg single and co-exposures respectively resulted in 123, 157 and 76 DAPs in gonads, identifying MeHg single exposure as the most impactful on the proteome under our experimental conditions. In the gonads, DAPs showing the highest modulation in single and co-exposures were associated with lipid metabolism, RNA processing, cellular development, cytoskeletal organization, and DNA processes

(Table S5). Most DAPs were specific to each exposure (Fig. 3), with 32 DAPs common across all treatments. MeHg exposure increased “spermatogenesis-associated protein 4” abundance.

Carbamazepine, MeHg, and their co-exposure altered 13, 13 and 7 KEGG pathways and GO-terms, respectively, in line with the higher number of DAPs in single exposures (Fig. 4). In all exposures, DAPs involved in protein turnover, DNA processes, response to stress and cell cycle were identified (Table S4). In single exposures at both stages, DAPs were involved in energy metabolism, including lipid metabolism, and the cytoskeleton. Carbamazepine exposure specifically showed DAPs involved in “PPAR signaling pathway” (ER = 3.6; DAPs = 4), “nucleus” (ER = 1.3; DAPs = 63), and “cell cycle” (ER = 4.7; DAPs = 5). Methylmercury-exposure specifically showed DAPs involved in DNA repair, such as “base-excision repair” (ER = 6.2; DAPs = 5) and “nucleic acid binding” (ER = 1.6; DAPs = 44), but also in “cell cycle” (ER = 4.2;

<figure><img src="image_6.png" /><figcaption>**Fig. 3.** Venn diagrams of differentially abundant proteins (DAPs) in male gonads of *D. polymorpha* exposed to CBZ, MeHg and their co-exposure in the proliferation (A) and maturation (B) stages (*n* = 5).</figcaption></figure>
<figure><img src="image_7.png" /><figcaption>**Fig. 4.** Bubble plot of the 30 most enriched KEGG Pathways (KP, blue), biological processes (BP, green), molecular functions (MF, brown) and cellular components (CC, yellow) GO-terms (C) in male gonads of *D. polymorpha* exposed to CBZ, MeHg and their co-exposure in the proliferation (A) and maturation (B) stage (n = 5). ER: enrichment ratio. DAPs: differentially abundant proteins.</figcaption></figure>

DAPs = 7), “TCA cycle” (ER = 3.0; DAPs = 5), and “vasopressin-regulated water reabsorption” (ER = 4.2, DAPs = 3), suggesting potential osmotic stress. In the co-exposure, DAPs were also involved in “nucleus” (ER = 1.4; DAPs = 39) and “cell cycle process” (ER = 1.8; DAPs = 14; Fig. 4, Table S4). Overall, the KEGG pathways and GO terms were predominantly exposure-specific, as illustrated in the Venn diagram (Fig. 3).

### 3.3.2. Maturation stage

Carbamazepine, MeHg and their co-exposure respectively resulted in 59, 210 and 79 DAPs (Fig. 3), indicating that MeHg exposure had the most significant impact on the proteome under our experimental conditions. As observed in the proliferation stage, each exposure resulted in a largely specific proteome, with 18 common DAPs across all exposures (Fig. 3). Differentially abundant proteins showing the highest modulations in single and co-exposures were mainly associated with the “myosin heavy chain”. Single and co-exposures also altered “glucose
dehydrogenase" (Table S5).

Carbamazepine, MeHg, and their co-exposure enriched 6, 19 and 10 KEGG pathways and GO-terms, respectively (Fig. 4). In all exposures, identified DAPs were involved in response to stress, cellular development and behavior. In both single exposures, identified DAPs were also involved in protein turnover. In MeHg single and co-exposures, common DAPs were involved in extracellular interactions and the cytoskeleton. In CBZ exposure, specific DAPs were involved in "adrenergic signaling in cardiomyocytes" (ER = 5.2; DAPs = 4). In MeHg exposure, DAPs involved in inflammation, i.e. "systemic lupus erythematosus" (ER = 4.6; DAPs = 5), but also "cAMP signaling pathway" (ER = 2.3; DAPs = 8) and energy metabolism were identified (Fig. 4, Table S4). The co-exposure showed DAPs involved in "multicellular organism reproduction" (ER = 1.8; DAPs = 21) and "cellular developmental process" (ER = 1.3; DAPs = 27) (Table S4). Overall, the enrichment analyzes of KEGG pathways and GO terms indicated that the proteomic responses were largely specific to each exposure, as illustrated in the Venn diagram (Fig. 3).

# 4. Discussion

## 4.1. Carbamazepine and MeHg exposures vs their co-exposure

At both stages, MeHg exposure caused similar or more alterations than the co-exposure, which contradicts our initial hypothesis based on previous studies of other organs in *D. polymorpha* under similar exposure scenarios. Indeed, in the whole soft tissues of *D. polymorpha* during the resting stage, exposure to 6100 ng L⁻¹ CBZ, 460 ng L⁻¹ MeHg and the co-exposure for 24 h resulted in the modulation of 98, 112 and 177 DAPs, respectively, suggesting that the co-exposure was more stressful (Baratange et al., 2023). Similarly, in *D. polymorpha* during the post-spawning stage, a 7-d exposure to 3900 ng L⁻¹ CBZ and 280 ng L⁻¹ MeHg modulated 25 metabolites in whole soft tissues, while single exposures did not affect metabolites (Baratange et al., 2022). These findings emphasize the importance of considering organ-specific responses and sexual maturity stage to fully characterize the molecular pathways of stress. Besides, here the genotoxic damage and the number of molecular modulations did not correlate with bioaccumulation, consistent with findings from previous studies under similar experimental conditions on *D. polymorpha* (Baratange et al., 2022; 2023) and other bivalves (Almeida et al., 2018; Jiang et al., 2019). Similarly, in *Mytilus edulis*, 14-d exposure to 50,000 ng L⁻¹ IHg also showed a lower bioaccumulation and higher TBARS concentration in the gonad compared to the gills and digestive glands (Amachree et al., 2024). This suggests that other factors may drive the modulation of toxicity pathways. These discrepancies highlight the need to better understand the influence of confounding factors to more accurately predict toxicity in environmental conditions.

## 4.2. Reproductive stage as confounding factor

Globally, gonads of *D. polymorpha* exhibited greater sensitivity to pollutant exposure in the proliferation compared to the maturation stage. In the proliferation stage, exposures significantly increased the abundance of tail DNA, indicating reprotoxicity after single and co-exposure to MeHg. Exposures also modulated targeted transcripts associated with antioxidant defenses and energy metabolism when compared with Ctl. In contrast, in the maturation stage, we observed a significant decrease of the abundance of tail DNA and no transcript modulation compared to Ctl.

Proteomic analyses revealed different DAPs in the maturation compared to the proliferation stage, indicating varied impacts of exposures depending on the date and highlighting the effectiveness of proteomics in identifying molecular toxicity pathways. In addition, an increased abundance of DAPs associated with energy metabolism, "apoptosis" and stress response, and a decreased abundance of DAPs associated with DNA processes, were measured in the maturation vs

proliferation stage (Table S4). This is also consistent with observations in *D. polymorpha* of a higher electron transport system activity and aerobic metabolism observed in summer (i.e., the spawning stage) compared to winter (i.e., the resting stage; Louis et al., 2020; Palais et al., 2011). In vertebrates, mature spermatozoa showed a reduced capacity to repair DNA damage compared to non-mature gonadic cells (Smith et al., 2013; Aitken and De Iuliis, 2007), and in *Gammarus fossarum*, spermatozoa showed lower DNA repair capacities than hemocytes (Lacaze et al., 2010).

Although, these aspects need to be studied in *D. polymorpha* to conclude, overall, data indicated a higher basal level of stress in Ctl in the maturation than in the proliferation stage. Thus, we hypothesize that it may have masked the effects of pollutant exposures on targeted biomarkers in the maturation stage, likely because energy allocation is attributed to reproduction rather than defense responses during this stage (Sokolova et al., 2012).

In bivalves, several studies evidenced that the RedOx homeostasis, defense responses, energy metabolism, metabolome and the transcriptome fluctuate throughout the reproductive cycle (Dheilly et al., 2012; Elia et al., 2020; González-Fernández et al., 2016; Guerra et al., 2012; Kronberg et al., 2021; Meistertzheim et al., 2009; Navarro et al., 2013; Popović et al., 2020; Storhaug et al., 2019), including in *D. polymorpha* (Louis et al., 2020, 2021), with sometimes contradictory observations. These discrepancies may arise from variations in experimental design, species differences, and the sensitivity thresholds of different organs, as well as seasonal effects and reproductive stages. However, data underscore the necessity of better characterizing the basal physiological status and reproductive stage to improve the interpretation of analyzed endpoints.

## 4.3. Gaining insight into molecular toxicity pathways of CBZ, MeHg and their co-exposure

### 4.3.1. Energy metabolism

In gonads, proteomics highlighted alteration of the lipid metabolism. Lipids are known to be essential for reproduction (Birsoy et al., 2013), and may serve as the main energy reserves in gonads. Similarly, in digestive glands of male *Mytilus galloprovincialis*, exposure 3 d to 80 ng L⁻¹ CBZ also altered proteins and metabolites involved in lipid metabolism (Dumas et al., 2022). Here, in CBZ exposure in the proliferation stage, cs transcript level was downregulated by 2.4x. Citrate synthase is involved in the TCA cycle, suggesting an alteration in aerobic metabolism. Similarly, in gills of *Ruditapes philippinarum*, exposure 2 d to 2000 ng L⁻¹ IHg altered several metabolites involved in energy metabolism, suggesting increased anaerobic metabolism (Liu et al., 2011).

Here, no modulation of lhd gene expression, an anaerobic metabolism biomarker, was observed after pollutant exposure, which may be due to the lower concentration used. Other exposures in the proliferation and exposures in the maturation stage showed no significant impact on the other selected genes involved in energy metabolism. However, in the proliferation stage, proteins involved in the TCA cycle were modulated by MeHg. Discrepancies between proteomic and transcriptomic expression levels may result from regulation at both transcriptional and post-transcriptional levels, as well as feedback regulation across different levels of biological organization (Bradley et al., 2009; Gedeon and Bokes, 2012).

### 4.3.2. Redox homeostasis

In the gonads of *D. polymorpha* in the proliferation stage, CBZ exposure down-regulated cat and gst transcripts, while MeHg exposure increased the abundance of "NADH cytochrome b5 reductase". Both suggest ROS modulation during single exposures and linked defense responses; however, no such pathway was observed in the co-exposure. Previous studies on *D. polymorpha* in the resting stage revealed that single and co-exposure to 3900 ng L⁻¹ CBZ and 280 ng L⁻¹ MeHg for 1 d and 7 d modulated antioxidant enzymes at both the transcript and
protein activity levels in digestive glands and gills (Baratange et al., 2022). In contrast, a similar exposure during the post-spawning stage for 24 h and 72 h did not modulate antioxidant responses in digestive glands, gills and gonads (Baratange et al., 2023). These findings further suggested that biological responses could be influenced by the reproductive stage and linked to the basal metabolism.

### 4.3.3. Cell cycle and DNA processes

Here, in the proliferation stage, CBZ and MeHg single and co-exposures enriched DAPs involved in “cell cycle process” and “cell cycle”. Concomitantly, single and co-exposures also modulated proteins involved in DNA processes and was accompanied by DNA damage in MeHg single and co-exposures, suggesting genotoxicity. Maintenance of DNA integrity is critical for cell proliferation and maturation (Kanellou et al., 2020). These findings suggested that CBZ, MeHg and their co-exposures may lead to disruptions in the cell cycle regulation, including potential genotoxic effects. In contrast, in the maturation stage, proteins involved in DNA processes and cell cycle were not modulated despite a higher bioaccumulation of CBZ. Both pollutants have been shown to cause DNA damage, particularly at higher concentrations and longer exposure duration. For example, in soft tissues of *Linnoperna fortunei*, exposure to 100,000 ng L⁻¹ IHg for 15 d increased DNA damage 2x, while shorter exposure and lower IHg concentrations did not (do Amaral et al., 2019); in hemocytes of *M. galloprovincialis*, exposure 4 d to 6300 ng L⁻¹ CBZ increased DNA damage 5x (Brandts et al., 2018). Here, data obtained from a 1-D-long exposure suggested that the gonad of *D. polymorpha* was sensitive to both pollutants, particularly in the proliferation stage.

### 4.3.4. Spermatogenesis

Differentially abundant proteins identified in CBZ and MeHg single and co-exposures confirmed a putative reprotoxicity. Proteins potentially involved in gamete production and maturation were modulated by both pollutants at both stages. For example, in the maturation stage, CBZ single exposure enriched “adrenergic signaling”, MeHg single exposure enriched “cAMP signaling pathway”, while the co-exposure enriched “multicellular organism reproduction”. These pathways are known to be involved in gametogenesis in vertebrates and are suspected of playing a similar role in bivalves (Klouche et al., 2015; Ni et al., 2020; *Strączyńska et al.*, 2022). Furthermore, CBZ and MeHg single and co-exposure modulated “myosin Va” in the maturation stage. This protein is an F-actin dependent molecular motor essential for acrosome formation during spermiogenesis in mice (Sun et al., 2012). In this line, in the proliferation stage, CBZ and MeHg single exposures modulated DAPs involved in “cilium” and “cilium movement”, which are important for the cellular development of germinal cells during spermatogenesis in vertebrates (Kim and Tsiokas, 2011). This function was also hypothesized in bivalves, as suggested by histological studies in *M. galloprovincialis* (Bardales et al., 2009). In addition, in the proliferation stage, CBZ and MeHg single exposures enriched “PPAR signaling pathway”, and “spermatogenesis-associated protein 4”, respectively, both involved in the proliferation and differentiation of male gametes (Jiang et al., 2013; Yao et al., 2015). Overall, data suggested molecular alterations of DAPs involved in gametogenesis due to CBZ and MeHg exposures at both stages, consistent with the delayed spermatogenesis measured in this species when exposed 15 w to 5000 ng L⁻¹ CBZ in previous work (Magniez et al., 2018). Findings here supported the need for further studies on the potential reprotoxicity of CBZ and MeHg (e.g. gamete fertility, gamete fecundity) as alteration in energy metabolism and spermatogenesis could affect the fitness of *D. polymorpha*.

## 5. Conclusion

The results of the present study confirmed the value of combining targeted and untargeted endpoints to assess molecular toxicity pathways and highlighted the efficiency of proteomics in this context. It also

confirmed that reproductive stage modified responses to pollutants. MeHg emerged as the most stressful for gonads of *D. polymorpha* in comparison to CBZ single exposure at both stages, triggering genotoxicity and the highest number of DAPs. The exposure-specific proteome further confirmed the difficulty to anticipate co-exposure toxicity based solely on responses observed in single exposures. As EQS are determined after single exposures to contaminants, this study underlined the need to consider pollutant mixtures for environmental biomonitoring. Future research should aim to characterize the link between identified DAPs and deleterious effects, particularly concerning gametogenesis and reproductive success, for their future use as early-warning signature of reprotoxicity and the applicability of these DAPs as biomarker for biomonitoring programs. However, no DAP was found to be commonly modulated by single and co-exposures at both stages, suggesting the importance of wisely selecting the organ to be analyzed, as well as biomarkers that are not or only weakly modulated by reproductive stage. Moreover, in this context, it is important to better document developmental periods and determine whether exposure during sensitive time windows could increase the risk for biota.

### CRediT authorship contribution statement

**Clément Baratange:** Writing – original draft, Visualization, Validation, Methodology, Investigation, Formal analysis, Conceptualization.
**Tainá Rocha de Almeida:** Writing – review & editing, Methodology, Investigation, Formal analysis. **Jean Armengaud:** Writing – review & editing, Validation, Resources, Formal analysis. **Isabelle Bonnard:** Investigation. **Elise David:** Writing – review & editing, Validation, Resources. **Laurence Delahaut:** Investigation. **Véronique Gaillet:** Investigation. **Emmanuel Guillon:** Validation. **Mélodie Kielbasa:** Investigation. **Jean-Luc Loizeau:** Writing – review & editing, Validation. **Nicolas Maurin:** Investigation. **Stéphanie Sayen:** Writing – review & editing, Validation, Investigation. **Claudia Cosio:** Writing – review & editing, Validation, Supervision, Resources, Project administration, Methodology, Investigation, Funding acquisition, Formal analysis, Conceptualization.

### Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

### Acknowledgements

This research was funded by Grand Reims and URCA through Aquasurv and Base Chairs, and the French national program EC2CO (Ecosphère Continentale et Côtière) through CARMA project. We are also grateful to the Mobicyte platform for access to qPCR facilities and support, and Dr Marc Bonnard for Comet assay analyzes. Authors would like to thank Nicolas Borie, Clémence Carouge, Andrea Gallorini and Kilian Kavanagh for technical help, and Maxime Leprêtre for analytical advice. Authors also benefited from the French GDR “Aquatic Ecotoxicology” framework which aims at fostering stimulating scientific discussions and collaborations for more integrative approaches.

### Supplementary materials

Supplementary material associated with this article can be found, in the online version, at doi:10.1016/j.aquatox.2025.107493.

### Data availability

Data will be made available on request.
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