# High microplastic exposure affects survival and health of Dreissena spp. mussels: implications for freshwater pollution monitoring

Adèle Wolinski, Miléna Ferret, Isabelle Calvès, Loïc Tettling, Audrey M. Pruski, Edouard Lavergne, Anne-Leïla Meistertzheim, Franck Lartaud

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## ► To cite this version:

Adèle Wolinski, Miléna Ferret, Isabelle Calvès, Loïc Tettling, Audrey M. Pruski, et al. High mi-croplastic exposure affects survival and health of Dreissena spp. mussels: implications for freshwater pollution monitoring. Environmental Pollution, 2025, 385, pp.127123. 10.1016/j.envpol.2025.127123. hal-05384018

HAL Id: hal-05384018

https://hal.sorbonne-universite.fr/hal-05384018v1

Submitted on 26 Nov 2025

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# High microplastic exposure affects survival and health of *Dreissena* spp. mussels: implications for freshwater pollution monitoring☆

Adèle Wolinski<sup>a,b</sup> ◐, Miléna Ferret<sup>a</sup>, Isabelle Calvès<sup>a</sup>, Loïc Tettling<sup>a</sup>, Audrey M. Pruski<sup>b</sup>, Edouard Lavergne<sup>a</sup> ◐, Anne-Leïla Meistertzheim<sup>a</sup>, Franck Lartaud<sup>b,*</sup> ◐

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## ARTICLE INFO

**Keywords:**

Freshwater
Polyethylene
Pollution
bivalves
Toxicity
Bioindicator

## ABSTRACT

The occurrence and concentrations of microplastics in freshwater environments are far less studied than in marine ecosystems, despite rivers being the receptacle of plastic pollution from terrestrial activities. Since microplastics impact aquatic organisms at multiple levels, the development of reliable and efficient methods for their quantification is crucial to improve ecological risk assessment and support regulatory efforts. In this study, we investigated the potential of *Dreissena* spp. mussels as bioindicators for microplastic pollution by exposing them to fluorescent polyethylene microbeads (10–20 µm) for 2 months at two environmental concentrations (200 and 2000 MP L⁻¹). Both microplastic uptake and health indicators (survival rates, condition index, and energy reserves) were assessed. Our results confirmed microplastic uptake by dreissenids under both exposure concentrations. Exposure to 2000 MP L⁻¹ led to a decrease in condition index and energy reserves within 15 days, resulting in increased mortality, whereas exposure to 200 MP L⁻¹ had no significant effect. These findings support the use of *Dreissena* spp. mussels as bioindicators for microplastic pollution in freshwater ecosystems, highlighting the bioaccessibility of MPs to filter-feeders and allowing the assessment of both microplastic accumulation in tissues and associated health effects.

## 1. Introduction

Plastics are synthetic materials composed of polymers combined with additives that influence their mechanical, thermal, and chemical properties (Kale et al., 2015). Their advantageous characteristics such as lightness, durability, versatility, and affordability (Ivleva et al., 2017; Lechthaler et al., 2020) have made them widely used across industries and promising for future technological and medical advancements (Thompson et al., 2009). Nevertheless, these same qualities, particularly their durability and ubiquity, also make plastics highly persistent in the environment. Despite growing awareness among scientists and citizens, plastic production continues to rise, from 335 million tons in 2016 to more than 400 million tons in 2023 (Plastics Europe, 2024). Combined with short product lifespans, improper disposal practices, and inadequate waste management (Diggle and Walker, 2022), this surge has resulted in the release of large quantities of persistent wastes into the environment, estimated at 22 million tons in 2019 (OECD, 2022).

Plastic wastes can be found under several sizes in natural habitats. Distinction is made between macroplastics (>25 mm) and mesoplastics (between 5 and 25 mm), large items representing the major visible pollution, and microplastics (MPs), which are defined as particles with a diameter smaller than 5 mm (Van Cauwenberghe et al., 2015b). MPs can be directly manufactured as microparticles (Coyle et al., 2020) used in resin pellets, textiles, medicines, and cosmetics (Cole et al., 2011; Chang, 2015), or they can derive from the fragmentation of larger plastics through physical (e.g., UV exposure, heat, abrasion) and biological (e.g., microbial degradation) processes (Andrady, 2011).

Rivers receive waste directly from terrestrial activities and can retain it for decades (Van Emmerik et al., 2022), leading to MP concentrations that are often higher than those in estuarine and coastal waters (Luo et al., 2019). However, far less studies focused on monitoring MP occurrence and concentration in freshwater environments compared to marine ones (Wagner et al., 2014; Li et al., 2018). The primary sources of MPs in freshwater systems include wastewater treatment plants

<footnote><sup>a</sup> SAS Plastic At Sea, 17 avenue Pierre Fabre, 66650, Banyuls-sur-Mer, France</footnote>

<footnote><sup>b</sup> Sorbonne Université, CNRS, Laboratoire d'Ecogéochimie des Environnements Benthiques (LECOB), UMR 8222, Observatoire Océanologique de Banyuls, F-66650, Banyuls-sur-Mer, France</footnote>

<footnote>* This paper has been recommended for acceptance by Prof. Wen-Xiong Wang.</footnote>

<footnote>* Corresponding author.
E-mail address: franck.lartaud@obs-banyuls.fr (F. Lartaud).</footnote>
(Murphy et al., 2016; Ruffell et al., 2023), urban and industrial dusts, fisheries, terrestrial runoff (Wagner et al., 2014; Dris et al., 2016; Ding et al., 2021), and direct waste disposal. Research has documented MPs in freshwater ecosystems across continents (Eerkes-Medrano et al., 2015; Dusaucy et al., 2021), even in protected areas (Free et al., 2014; Naying et al., 2025), and sometimes in quantities exceeding the abundance of fish larvae (Lechner et al., 2014). MP concentrations in freshwater ecosystems vary significantly between regions (Jolaosho et al., 2025), with their spatial and temporal distributions influenced by seasonal and hydrodynamic conditions, and typically showing higher prevalence in urban areas (Rodrigues et al., 2018; Jolaosho et al., 2025). Polystyrene (PS), polypropylene (PP), and polyethylene (PE) are the most common polymer types identified in freshwaters (Rodrigues et al., 2018; Semmouri et al., 2023).

Over the past decade, research on the toxicity of MPs has surged, highlighting their hazardous effects across terrestrial (Dissanayake et al., 2022), freshwater (Wong et al., 2020; Sarijan et al., 2021), and marine species (Leistenschneider et al., 2023). However, studies remain predominantly focused on marine ecosystems (Wagner et al., 2014; Magni et al., 2018; Li et al., 2020), with freshwater habitats receiving limited attention (Li et al., 2018; Wagner et al., 2014; Horton et al., 2017), accounting for less than 4 % of research related to MPs (Wagner and Lambert, 2018). This gap is particularly concerning, as freshwater systems may be more vulnerable to MP pollution due to their proximity to human activities (Free et al., 2014; Baldwin et al., 2016; Li et al., 2020). Nevertheless, ingestion of MPs has been demonstrated across all freshwater taxa, spanning different feeding modes and trophic levels (Eerkes-Medrano et al., 2015), with MPs accumulating in the gut lumen and infiltrating in the circulatory system and tissues (Ding et al., 2018; Magni et al., 2019). Adverse effects of MPs on freshwater species have been documented, including reduced feeding and increased mortality in zooplankton (Galir et al., 2025), altered antioxidant enzyme activities and depleted energy reserves in molluscs (Magni et al., 2018, 2019; Vidal et al., 2023), and gastrointestinal impairments and metabolic disruption in fish (Qiao et al., 2019; Zhao et al., 2020). This underscores the urgent need for MP pollution monitoring in freshwater ecosystems as part of ecological risk assessments, while also recognising potential risks to humans through trophic transfer (Nair and Perumal, 2022).

Current MP quantification methods in freshwaters were originally developed for marine research (Yao et al., 2020), involving net deployments and water samplings. However, rivers are highly heterogeneous systems with various factors influencing MP distribution (Skalska et al., 2020). These methods, often limited by brief sampling durations and small spatial coverage, do not accurately reflect in situ pollution. Furthermore, particles smaller than 25 µm are rarely collected and quantified due to methodological constraints (Imhof et al., 2016), while they are the most bioavailable and potentially hazardous to biota (Li et al., 2020). Given these limitations, complementary tools like bioindicators (i.e., organisms providing information on the quality of the environment) should be developed to assess MP pollution levels in freshwaters and their potential toxic effects on biota (Cho et al., 2019).

Bivalves such as clams, oysters, and mussels, are widely used as sentinel species (Hamza-Chaffai, 2014) for their adequate size, broad distribution, abundance in global aquatic ecosystems, sedentary lifestyle, and relative tolerance to pollutants (Sericano et al., 1995; Waykar and Deshmukh, 2012; Farrington et al., 2016). Several studies have confirmed the presence of MPs with varying sizes, shapes, and polymer types in freshwater bivalves in both controlled experiments and environmental monitoring (Berglund et al., 2019; Moreschi et al., 2020; Martyniuk, 2022). Among them, dreissenid mussels (the zebra mussel *Dreissena polymorpha* and the quagga mussel *Dreissena bugensis*) are of particular interest for MP biomonitoring. These mussels are common across numerous European and North American water reservoirs (Hoellein et al., 2021; Binelli et al., 2022; Karatayev and Burlakova, 2025), and they filter large volumes of water (up to 400 mL h⁻¹ per individual (Baldwin et al., 2002)), allowing them to accumulate

contaminants in their tissues even at low ambient concentrations (Minier et al., 2006; Binelli et al., 2020; Atamanalp et al., 2023). Although dreissenids typically ingest particles around 40 µm (Winkel and Davids, 1982), they can also capture larger particles due to their 1 mm inhalant opening (Magni et al., 2020; Binelli et al., 2020). Finally, zebra mussels are already used as heavy metal bioindicators (Camusso et al., 1994) for their easy and inexpensive sampling and deployment.

This study aimed to evaluate the potential of freshwater mussels from the genus *Dreissena* as bioindicators for MP pollution by assessing MP bioaccessibility and associated biological effects. Both MP accumulation in tissues and health parameters at the individual level were assessed, including survival, condition index, and energy reserves (i.e., glycogen, total lipids, and proteins). These parameters collectively reflect the mussel ability to maintain their biological functions and tolerate the stress caused by MP exposure (Palais et al., 2011). Glycogen serves as the primary energy source mobilised under stress in bivalves (Haag et al., 1993), while total lipids may also decline due to triglyceride consumption (Su et al., 2021), and protein catabolism only occurs under extreme energy stress (Shang et al., 2021). Since exposure duration and concentration are critical factors influencing the biological responses of species, a two-month chronic exposure experiment was conducted using environmental concentrations of PE microbeads at 200 and 2000 MP L⁻¹.

## 2. Material and methods

### 2.1. Mussel sampling and acclimation

Hundreds of dreissenid mussels (*Dreissena* spp., shell length: 2.3 ± 0.5 cm, range: 1–3.2 cm) were collected in September 2023 by a scientific fisherman from a depth of 12 m at the Bourget Lake (45°43′46.842″ N – 5°52′10.484″ E), Savoie, France. Although most of the mussels collected were *D. bugensis*, we refer to them as *Dreissena* spp. since distinguishing them from *D. polymorpha* can be challenging. The mussels were transported to the laboratory in a bucket filled with lake water, which was oxygenated using an air stone to maintain adequate dissolved oxygen levels. They were individually cleaned with a brush to remove epiphytes, then placed in a 100 L aquarium filled with lake water for the first day of acclimation. During the first three days, water from the lake was progressively replaced by a 50:50 mixture of deionized and 48h-dechlorinated water (1/3 vol per day) (Guidi et al., 2020; Louis et al., 2021). This operation was repeated the two following days to gradually replace all lake water. The mussels underwent then 15 days of acclimation in a room set to 18 °C with a 12:12 photoperiod (with lighting from 7 a.m. to 7 p.m.) and were fed daily with a suspension of *Chlorella* sp. (10⁶ cells per day per mussel, provided by Greensea, France (Evariste et al., 2017);).

### 2.2. Model microbeads

The MPs used were fluorescent PE microspheres, ranging from 10 to 20 µm in diameter (density: 1.00 g cm⁻³, peak emission of 525 nm when excited at 485 nm) purchased as a dry powder from Cospheric (USA). Although fluorescent PE microbeads are not the most common MPs found in freshwater environments, they were selected here because their fluorescence facilitates accurate uptake quantification. Hereafter, the term 'plastic uptake' is used in the general sense given by Kolandhasamy et al. (2018), that integrates both the processes of ingestion and external adhesion to the organs. Moreover, this particle size was chosen based on the typical range filtered by dreissenid mussels (i.e., from 0.7 to 40 µm (Winkel and Davids, 1982; Sprung and Rose, 1988)), while falling within the size range of MPs that are typically not retained by conventional net-based sampling methods (<25 µm). PE was selected as it is the predominant polymer found in freshwater ecosystems, including water, sediments, and biota (Cera et al., 2020). MP concentrations in freshwater bodies vary widely in the literature, ranging from 0.016 to 2867
MP L⁻¹ (Li et al., 2020), so two concentrations were used: a moderate environmental concentration (L-MP) with 200 MP L⁻¹ and a high environmental concentration (H-MP) with 2000 MP L⁻¹. To facilitate aquarium contamination, a MP stock solution was prepared by adding the microbeads in a glass bottle with 96 % ethanol, for a final concentration of 100 000 MP mL⁻¹. Ethanol was used to ensure effective dispersion of the microbeads, as recommended by Karami (2017). Before each contamination, the suspension was homogenised by a vigorous shaking of the bottle. At each water change, 100 µL and 1 mL of the suspension were delivered to the 50 L aquaria for L-MP and H-MP groups, respectively, to obtain the desired MP concentrations.

## 2.3. 3,5-Dichlorophenol exposure

To compare the potential toxic effects of MPs on dreissenid health, 3,5-dichlorophenol (3,5-DCP, 97 % purity, Sigma-Aldrich, USA) was used as a positive control. This compound is frequently used as a reference chemical pollutant in toxicity studies (Ertürk and Saçan, 2012; Tato et al., 2018; Xie et al., 2018). Two concentrations were used: 2100 µg L⁻¹ (H-DCP), corresponding to the 48h EC₅₀ of an embryo-larval bioassay in *Platynereis dumerilii* (Hutchinson et al., 1995), and a 10-fold lower concentration of 210 µg L⁻¹ (L-DCP). At each water change, 10.5 and 105 mg of 3,5-DCP were added to the L-DCP and H-DCP 50 L aquaria, respectively.

## 2.4. Experimental set up

After the acclimation period, dreissenid mussels were transferred into PlastiCell® structures (patent BT8167PC00, Plastic At Sea, France), which ensure optimal exposure to MPs through a continuous flow of water and MPs (see Fig. S1 and Text S1 for more details). Six mussels were placed in each PlastiCell®, with 8 PlastiCell® per structure (n = 48), and a total of 5 structures (n = 240), one for each treatment level: a control without pollutants, a low environmental concentration of MPs (L-MP), a high environmental concentration of MPs (H-MP), a low concentration of 3,5-DCP (L-DCP), and a high concentration of 3,5-DCP (H-DCP). The water in the structures was changed twice a week with a 50:50 mix of deionized and 48h-dechlorinated tap water, with contaminants added immediately afterwards. Mussel filtration likely reduced contaminant levels between water changes, so concentrations were not maintained continuously, and exposure occurred in pulses. Temperature, photoperiod, and feeding were kept the same as during acclimation. The temperature was checked daily, and physicochemical parameters of the water (NO₂, NO₃, NH₄, Ca²⁺, pH) were monitored weekly using colorimetric tests (ProAquaTest, JBL, Germany) to ensure optimal water conditions. Control mussels were sacrificed at the start of the experiment, then three mussels of each PlastiCell® and each treatment level were sampled after 2 weeks of exposure (S1), and the remaining three mussels were sampled after 2 months (S2). For each group of three mussels, one was used for MP uptake analysis, while the other two were pooled for health analysis to ensure sufficient biological material. After measuring their size with a calliper, the mussels were dissected using metal scalpels and tweezers to separate the tissues from the shell. The tissues were rinsed with distilled water, placed in 25 mL glass tube for freeze-drying, and weighted on a microscale to determine their dry weight (DW). Special care was taken to minimise contamination with external MPs during exposure, sampling, and sample preparation in the laboratory. Cotton laboratory coats were worn, glass or metal equipment was used after calcination at 450 °C, all solutions (KOH, water, ethanol) were pre-filtered through GF/F filters before use, and samples and glassware were kept covered with aluminium foil during analyses.

## 2.5. Survival rates

Mortality among dreissenids was monitored daily throughout the

experiment by inspecting each individual. Mussels were considered dead when their shells were wide open, and they did not react to mechanical stimulation. Dead mussels were immediately removed to prevent contamination of the others. Daily survival rates for each treatment level were calculated following the formula: Number of live mussels × 100/Total number of mussels.

## 2.6. Microplastic quantification

To quantify the fluorescent MPs incorporated in dreissenid tissues, an alkaline digestion of whole organisms was carried out. To each tube containing freeze-dried mussel tissues, 20 mL of 10 % potassium hydroxide solution (10 % KOH, Sigma Aldrich) were added. After 24h of digestion in an oven at 60 °C with orbital agitation, the samples were filtered onto a 1.2 µm fiberglass filter (GF/C, diameter 47 mm, Whatman® Sigma-Aldrich) using a glass filter unit coupled with vacuum pump. Each tube was rinsed twice with ultrapure water and 70 % ethanol to ensure that all MPs were collected on the filter. The filtration unit was also carefully rinsed between samples to avoid cross-contamination. The filters were then placed in small Petri dishes and left to dry for a few hours. Each filter was entirely analysed under an epifluorescence microscope (BX-UCB, Olympus, Japan) with a FITC filter (excitation peak = 490 nm) to count the fluorescent MPs. Results are expressed both as the number of MPs per dry weight (DW) and per individual. The values normalised to DW were obtained by dividing the number of MPs per individual by the corresponding mussel dry weight.

## 2.7. Condition index

The condition index of dreissenid mussels was calculated following the formula: Dry weight/Shell length, as used in previous studies (Shoults-Wilson et al., 2015; Blaise et al., 2017; Hoellein et al., 2021).

## 2.8. Total lipid content

Mussels sampled for health analysis were milled into a homogeneous powder using a bead mill (TissueLyser II, Qiagen, Germany). The total lipid content was determined using the sulfophosphovanillin colorimetric method described by Barnes and Blackstock (1973) with minor modifications (Pruski et al., 2017). Briefly, total lipids were extracted from 5 mg of mussel powder by adding of 1.5 mL of a chloroform/methanol mixture (2:1 v/v) (Carlo Erba, France) and agitating for 20 min. Following centrifugation (2000 rpm, 10 min, 4 °C), 1 mL of the supernatant was transferred to a new tube and evaporated at 90 °C. The lipid layer was then hydrolysed with 1.5 mL of sulphuric acid and heated for 10 min at 90 °C. For quantification, 100 µL of the hydrolysate was mixed with 2.5 mL of ortho-phosphovanillic reagent (Honeywell, USA), and absorbance was measured at 520 nm using a spectrophotometer (DS-11 DeNovix, USA). Total lipid concentrations were estimated using cholesterol as a standard (Fisher Scientific, USA).

## 2.9. Total protein content

Total proteins were extracted from 5 mg of mussel powder in the presence of 500 µL of extraction buffer (NaCl 150 mM, NaH₂PO₄ 10 mM, phenylmethylsulfonyl 1 mM, pH = 7.2) in a lysing tube (Lysing Matrix A, MP Biomedicals, USA). After a 20-sec agitation using a homogeniser (FastPrep®-24, MP Biomedicals), the samples were centrifuged (10 000 rpm, 10 min, 4 °C) to separate the proteins in the supernatant from the cellular debris. The quantification of total proteins was performed using the DC protein assay from Bio-Rad (USA). Following the addition of both reagents, absorbance was measured at 685 nm using a microplate reader (Victor Nivo, PerkinElmer, USA). Total protein concentrations were estimated using bovine serum albumin as a standard (Sigma-Aldrich).
## 2.10. Glycogen content

Carbohydrates were extracted from 5 mg of mussel powder using 500 µL of 0.6 N perchloric acid (Sigma-Aldrich). After homogenisation, the samples were centrifuged (14 000 rpm, 5 min, 4 °C) and the pH was then adjusted to 6.5 by adding K₂CO₃ (Sigma-Aldrich). After a second centrifugation (14 000 rpm, 1 min, 4 °C), the samples were diluted by half with 0.1 M sodium citrate buffer (Sigma-Aldrich). The samples were split into two aliquots; one of which received 5 µL of amyloglucosidase (120 U mL⁻¹, LIBIOS) and was incubated at 33 °C for 2 h, followed by 1 h at 4 °C. After a final centrifugation (14 000 rpm, 30 min, 4 °C), the glucose content in both aliquots was quantified using the GOPOD assay KIT (LIBIOS) with glucose as the standard. Absorbance was measured at 480 nm, and the glycogen content was calculated by subtracting the glucose concentration in the aliquot that received the enzyme from the concentration in the aliquot that did not.

## 2.11. Statistical analysis

All statistical analyses were performed using R v4.3.1 (R Core Team, 2025) with a significance threshold of p ≤ 0.05. Survival of the mussels from the different treatment levels was analysed using a Cox regression model. Hazard ratios (HR) were estimated for each treatment relative to the control. For condition index, lipid, protein, glycogen, and MP contents, the assumptions for parametric testing were formally checked on the residue for normality, using the Shapiro-Wilk's-W test, and on raw data for homoscedasticity, using the Bartlett test. As normality assumptions were not met, treatment differences were determined using a non-parametric Kruskal-Wallis test, followed by a Dunn post-hoc test for pairwise comparisons with the Bonferroni correction.

# 3. Results

## 3.1. Survival rate

The Cox regression revealed strong treatment effects on mussel survival (Supplementary Table S1). Compared to the Control, exposure to H-DCP dramatically increased mortality risk (HR = 92.3, p ≤ 0.05). H-MP and L-DCP also significantly elevated mortality risks (HR = 6.6 and 3.7, respectively, p ≤ 0.05), whereas L-MP did not affect survival (HR = 1.0, p > 0.05).

Survival curves were generated to visualise temporal survival patterns (Fig. 1). After 2 weeks of exposure, survival in the H-DCP group dropped below 40 %, while survival in the other four treatment levels remained above 80 %. At the 2-week sampling point, all surviving mussels in the H-DCP group were subsequently sampled and sacrificed for health analyses. In the Control and L-MP groups, dreissenids exhibited low mortality, with survival rates of 96 % in both treatment levels after 2 months. In contrast, the survival rate of mussels exposed to L-DCP treatment decreased gradually, reaching 83 % at S2. Exposure to H-MP treatment led to a more pronounced decline in survival, which dropped to 42 % after 2 months of exposure.

## 3.2. Microplastic uptake

MPs were detected in the soft body of dreissenid mussels exposed to both MP concentrations (200 and 2000 MP L⁻¹) after 2 weeks and 2 months (Fig. 2). After 2 weeks, MP content did not differ significantly between the two MP exposure levels, with average values of 1.5 ± 0.6 and 2.4 ± 0.7 MP mg⁻¹ (dry weight) for L-MP and H-MP groups, respectively (p > 0.05, Kruskal-Wallis test followed by a Dunn test, Supplementary Table S2), corresponding to 48.5 ± 23.1 and 77.4 ± 54.9 MPs individual⁻¹. After 2 months, no significant differences were observed either, with average values of 2.5 ± 3.5 and 1.6 ± 0.8 MP mg⁻¹ for L-MP and H-MP groups, respectively (p > 0.05, Kruskal-Wallis test followed by a Dunn test), corresponding to 8.9 ± 7.2 and 30.8 ±

<figure><img src="image_5.png" /></figure>

**Fig. 1.** Survival rate of dreissenid mussels exposed to control (blue), L-MP (pink), H-MP (red), L-DCP (yellow), and H-DCP (green) groups along the experiment. S1 and S2 represent the first and second samplings, respectively. Letters indicate significant differences (Chi-squared test, p < 0.05). All mussels in H-DCP treatment were sacrificed at S1. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)

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**Fig. 2.** Number of MPs per mg of DW for dreissenid mussels exposed to control (blue), L-MP (pink), and H-MP (red) groups. Boxplots show the median, 25th and 75th percentiles. Red diamonds represent the mean values and black dots represent outliers. Letters indicate significant differences between groups for each sampling time (Kruskal-Wallis test followed by a Dunn test with Bonferroni correction, p < 0.05). (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)

39.2 MPs individual⁻¹.

## 3.3. Health status

The condition index of dreissenids at the start of the experiment (0 days) was 12.1 ± 3.3 mg DW cm⁻¹ (Fig. 3). After 2 weeks of exposure, mussels in the Control and L-MP groups exhibited condition indices about 1.5–2.5 times higher than those in the H-MP, L-DCP, and H-DCP groups (p ≤ 0.05, Kruskal-Wallis test followed by a Dunn test). By 2
<figure><img src="image_7.png" /><figcaption>**Fig. 3.** Condition indices (mg DW·cm⁻¹) of dreissenid mussels exposed to control (blue), L-MP (pink), H-MP (red), L-DCP (yellow), and H-DCP (green) groups. The H-DCP treatment is absent at 2 months due to mortality. Boxplots show the median, 25th and 75th percentiles. Red diamonds represent the mean values and black dots represent outliers. Letters indicate significant differences between treatments for each sampling time (Kruskal-Wallis test followed by a Dunn test with Bonferroni correction, *p* < 0.05). (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>

months, the condition indices of dreissenids in the Control and L-MP groups remained twice higher than those in H-MP group (p ≤ 0.05, Kruskal-Wallis test followed by a Dunn test), although a significant decrease between 2 weeks and 2 months was observed (p ≤ 0.05, Kruskal-Wallis test followed by a Dunn test, Supplementary Table S3).

Lipid, protein, and glycogen contents normalised by DW are shown

in Fig. 4, while their corresponding values per individual are provided in Supplementary Table S4.

After 2 weeks of exposure, mussels in the H-MP and L-DCP groups exhibited lipid contents about 1.5 times lower than those in the Control and L-MP groups (p ≤ 0.05, Kruskal-Wallis test followed by a Dunn test, Fig. 4A–Supplementary Table S5). After 2 months, mussels exposed to L-

<figure><img src="image_8.png" /><figcaption>**Fig. 4.** Total lipid (A), protein (B), and glycogen (C) contents (µg·mg⁻¹ DW) of dreissenid mussels exposed to control (blue), L-MP (pink), H-MP (red), L-DCP (yellow), and H-DCP (green) groups. The H-DCP treatment is absent at 2 months due to mortality. Boxplots show the median, 25th and 75th percentiles. Red diamonds represent the mean values and black dots represent outliers. Letters indicate significant differences between groups (Kruskal-Wallis test followed by a Dunn test with Bonferroni correction, *p* < 0.05). (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>
DCP group displayed the highest lipid content (77.0 ± 5.8 µg mg⁻¹ DW), while lipid levels in other treatment levels were slightly lower, ranging from 62.8 ± 7.0 to 69.8 ± 7.1 µg mg⁻¹ DW.

Dreissenids exposed to the H-MP group showed the lowest protein levels at both 2 weeks and 2 months, about 1.5 times lower than other treatments (Fig. 4B). However, this difference was statistically significant only when comparing H-MP to L-DCP groups at 2 months (p ≤ 0.05, Kruskal-Wallis test followed by a Dunn test, Supplementary Table S5).

After 2 weeks, mussels in the H-MP and H-DCP groups had significantly reduced glycogen contents, about 6–26 times lower than those in the Control and L-MP groups (p ≤ 0.05, Kruskal-Wallis test followed by a Dunn test, Fig. 4C–Supplementary Table S5). By 2 months, glycogen content remained the lowest in mussels exposed to the H-MP group, about 2–5 times lower than in the other treatment levels, although this difference was not statistically significant (p > 0.05, Kruskal-Wallis test followed by a Dunn test, Supplementary Table S5).

# 4. Discussion

## 4.1. Microplastic uptake without accumulation in dreissenid mussels

Dreissena spp. mussels were selected as model species due to their suspension-feeding mode and their key bioindicator traits such as a sedentary lifestyle, high abundance, sufficient lifespan, and ease of sampling. Consistent with previous laboratory studies (Pedersen et al., 2020; Weber et al., 2021) and investigations on wild populations (Pastorino et al., 2021; Atamanalp et al., 2023), our findings confirm the MP uptake by dreissenids, supporting their use for MP pollution monitoring in rivers and lakes (Magni et al., 2018). Although the number of MPs detected in mussel soft bodies in our experiment exceeded the range reported in field studies (0.03–2 MPs individual⁻¹ (Pastorino et al., 2021; Atamanalp et al., 2023)), it is in line with results from Weber et al. (2021), who observed similar MP burdens (31 and 157 MPs individual⁻¹) in zebra mussels exposed to 300 and 3000 MP L⁻¹, respectively.

Two concentrations of MPs were tested in our study: 200 MP L⁻¹, representing a moderate level in freshwater environments, and 2000 MP L⁻¹, simulating heavy pollution conditions already reported in some lakes and rivers (Li et al., 2020). However, increasing MP concentrations in the water did not lead to significantly higher MP levels in mussel tissues, highlighting their limitations as quantitative bioindicators in heavily polluted environments. Similarly, Weber et al. (2021) reported higher, but statistically insignificant, MP uptake in *D. polymorpha* exposed to 3000 MP L⁻¹ compared to 300 MP L⁻¹ after 12 h of exposure. Together, these findings suggest the existence of a maximal MP level in dreissenids, as observed for *M. edulis* (Woods et al., 2018). These laboratory results align with field observations, where MP levels in dreissenid mussels remained low despite variable MP concentrations in surrounding waters (Hoellein et al., 2021).

We also assessed the MP content in mussels after two exposure durations (2 weeks and 2 months), but longer exposure did not lead to MP accumulation in mussel tissues. Similar results were reported over shorter exposure durations: little to no bioaccumulation within 3 days in quagga mussels (Merzel et al., 2020), and even a decrease in MP content between 12 and 48 h in zebra mussels (Weber et al., 2021). Likewise, a field study found no evidence of MP accumulation in dreissenids over 30–60 days (Hoellein et al., 2021). Collectively, these findings suggest that dreissenids do not accumulate MPs during either short- or long-term exposures, potentially due to physiological mechanisms limiting MP uptake. For example, several bivalve species have been observed to close their valves in response to MP exposure (Bringer et al., 2021; Fu et al., 2022), thereby reducing filtration and MP uptake (Zhong et al., 2025). Another possible explanation is enhanced pseudofaeces production, as overloaded gill cilia may redirect excess particles for rejection (Ward et al., 1993; Weber et al., 2021). Indeed, Ward et al. (2019) demonstrated the rapid elimination of several types of MPs in pseudofaeces by marine bivalves. Future studies incorporating pseudofaeces analysis are

therefore needed to clarify the absence of MP in dreissenids.

## 4.2. Adverse effects of high microplastic concentrations on dreissenid mussels

Exposure of dreissenid mussels to 10–20 µm PE MPs at a moderate concentration (200 MP L⁻¹) caused no detectable health impairment after either 2 weeks or 2 months, based on all measured parameters. Survival, condition index, and energy reserves remained comparable between the Control and L-MP groups, suggesting similar metabolic responses under low environmental MP exposure and MP-free treatment. The decline in condition index observed over time in both groups may reflect natural variation or stress related to laboratory conditions (Haskell et al., 2024). For example, Kilgour and Baker (1994) previously reported progressive weight loss in dreissenids held in laboratory facilities.

In contrast, dreissenids exposed to the high environmental concentration of MPs (2000 MP L⁻¹) exhibited significant adverse effects as early as 2 weeks into exposure. The reduction in condition index and energy reserves in this group was comparable to that observed in mussels exposed to the highest level of the reference toxicant (i.e., 3,5-DCP at 2100 µg L⁻¹, H-DCP). After 2 months, survival, condition index, and energy reserves in this H-MP group declined even further, showing more severe effects than those induced by the lowest concentration of 3,5-DCP (i.e., 210 µg L⁻¹, L-DCP). The observed lipid depletion after 2 weeks likely reflects insufficient food intake, as lipids are among the first energy reserves to be mobilised during starvation (Palais et al., 2011; Said and Nassar, 2022). Likewise, the reductions in glycogen at 2 weeks and in protein content at 2 months reveal a state of severe energy deficiency. Glycogen, the primary storage carbohydrate in bivalves (Haag et al., 1993), is rapidly broken down under stress (Louis et al., 2021), while protein catabolism typically occurs only during conditions of prolonged energy stress (Shang et al., 2021). This depletion of macromolecular energy stores was manifested as a decline in the condition index, a widely used indicator of environmental stress in bivalves, as it reflects the consumption of energy reserves to maintain basic physiological functions (Taylor et al., 2017; Said and Nassar, 2022).

These metabolic disturbances observed in dreissenids exposed to high MP concentrations may arise from increased energy demand, reduced energy supply, or a combination of both. Elevated energy expenditure under MP exposure has been documented in marine mussels (*Mytilus edulis*) exposed to polystyrene (PS) microspheres (Van Cauwenberghe et al., 2015a), and may reflect efforts to maintain homeostasis (Smolders et al., 2002) or respond to MP-induced inflammation (Auguste et al., 2023), as antioxidant production and repair of oxidative damage are known to be energy-intensive processes (Said and Nassar, 2022). Consistent with this, oxidative stress is frequently reported in mussels exposed to MPs (Huang et al., 2021; Li et al., 2022; Wang et al., 2024), often accompanied by increased activities of antioxidant enzymes (Li et al., 2022; Zhong et al., 2024). Conversely, energy supply may be reduced due to decreased ingestion and/or assimilation of phytoplankton. As previously discussed, valve closure and reduced clearance rates have been observed in bivalves exposed to MPs (Bringer et al., 2021; Fu et al., 2022), potentially limiting phytoplankton intake (Wegner et al., 2012; Xu et al., 2017; Weber et al., 2021). Additionally, MPs may impair digestion, as energy metabolism is disturbed under MP exposure (Gu et al., 2020; Zhong et al., 2024, 2025), or compete with microalgae in the water column (Gardon et al., 2018), further reducing assimilation efficiency.

These disruptions likely compromise metabolic function in dreissenids, leading to weight loss, depletion of energy reserves, and reduced condition index (Said and Nassar, 2022). These physiological impairments likely contributed to the increased mortality observed in the H-MP group, which occurred more rapidly than in mussels exposed to the L-DCP group. Similarly, Pedersen et al. (2020) found that 10–45 µm high-density PE (HDPE) microbeads impacted quagga mussel feeding by
reducing clearance rates, resulting in a 50 % higher mortality rate for mussels exposed to 0.8 g L⁻¹ compared to controls, although the difference was not statistically significant.

### 4.3. Use of dreissenid mussels as bioindicators for microplastic pollution

Our study shows that high MP concentrations (2000 MP L⁻¹) negatively affect dreissenid mussels, whereas moderate concentrations (200 MP L⁻¹) induced no significant effects on survival, condition index, or energy reserves. Consistent with our findings, previous studies have also reported adverse effects on dreissenids under high MP exposure at the individual, cellular, and molecular levels, including alterations in antioxidant capacity (Magni et al., 2018; Weber et al., 2020; Binelli et al., 2020), modulation of protein expression (Magni et al., 2019; Brehm et al., 2022), and changes in clearance rates (Weber et al., 2021). These findings support the existence of a physiological threshold of MP tolerance in dreissenid mussels, beyond which key metabolic functions are impaired. This response suggests that dreissenids may serve as bioindicators, capable of distinguishing between low and high MP pollution in freshwater ecosystems. Rapid changes in the health status or population density could serve as early warning signals of high MP contamination. In particular, the condition index appears to be a promising biomarker: it is easy and rapid to measure, cost-effective, requires minimal equipment, and has already been used to monitor other pollutants in mussels (Orban et al., 2002; Rogers et al., 2024).

However, since our results were obtained with fluorescent PE microbeads, they represent only a first step that should be complemented by further experiments using more environmentally relevant MPs, varying in size, polymer type, shape, and degree of weathering. Field studies are also needed to assess how dreissenid responses evolve under confounding factors such as co-occurring pollutants, temperature, or food availability (Digka et al., 2024). Finally, further research should aim to distinguish the specific impacts of MPs from those of natural suspended particles, as some studies indicate similar effects (Putri et al., 2021; Hamm et al., 2022), while others demonstrate higher impacts of MPs compared to silt, clay, or kaolin (Harris and Carrington, 2020; Yap et al., 2020; Chen et al., 2025).

A recognised limitation in the use of dreissenids as bioindicators lies in their invasive nature, which restricts their application in active biomonitoring to habitats where they are already established, or in passive biomonitoring with indigenous populations (Binelli et al., 2015). Nonetheless, their broad geographic spread and successful use in monitoring various freshwater pollutants, including heavy metals (Kutluyer Kocabas et al., 2024), pesticides (Riva et al., 2010), polycyclic aromatic hydrocarbons (PAHs) (Parolini and Binelli, 2014), and radio-nuclides (Kinney et al., 1994), underline their practical relevance for assessing environmental quality.

## 5. Conclusion

Our results highlighted that chronic exposure to high environmental concentrations of MPs induces energy stress and increases mortality in dreissenids, whereas lower concentrations do not cause measurable adverse effects. Although MPs were detected in mussel tissues at both exposure levels, no clear relationship was found between MP burden and the water MP concentration. These findings suggest that dreissenids may serve as effective bioindicators for MP pollution through the combined assessment of their health status and MP content. Physiological changes, such as decline in condition index, depletion of energy reserves, and increased mortality, can signal high MP concentrations. To improve the accuracy of MP pollution assessments in freshwater ecosystems, we recommend that biomonitoring programmes integrate both MP content and health parameters in dreissenid mussels, providing a more comprehensive evaluation of biological impacts alongside contaminant loads.

## CRediT authorship contribution statement

**Adèle Wolinski:** Writing – review & editing, Writing – original draft, Visualization, Supervision, Methodology, Investigation, Formal analysis. **Miléna Ferret:** Investigation, Formal analysis. **Isabelle Calvès:** Writing – review & editing, Project administration, Methodology. **Loïc Tettling:** Methodology. **Audrey M. Pruski:** Writing – review & editing, Supervision, Methodology, Investigation, Conceptualization. **Edouard Lavergne:** Writing – review & editing, Supervision, Methodology, Investigation, Conceptualization. **Anne-Leïla Meistertzheim:** Writing – review & editing, Project administration, Methodology, Investigation, Funding acquisition, Conceptualization. **Franck Lartaud:** Writing – review & editing, Supervision, Project administration, Methodology, Investigation, Funding acquisition, Conceptualization.

## Funding

This work was supported by the French Ecological Transition Agency (MusselPlast RIVER project). This work was part of the PhD thesis of A. W. supported by the CIFRE contract with the Plastic At Sea company, the National Research and Technology Agency and Sorbonne University.

## Declaration of competing interest

The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: Meistertzheim reports financial support was provided by The French Agency for Ecological Transition. Meistertzheim has patent PlastiCell issued to BT8167PC00. If there are other authors, they 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

The authors would like to thank the professional fishermen Cédric Giroud for supplying us in dreissenid mussels. We kindly thank Pascal Romans and the aquariology department of the Oceanological Observatory of Banyuls-sur-Mer (OOB) for access to experimental ecotoxicology laboratory conditions. We thank Gilles Vétion for his help in developing the lipid assay method. We acknowledge the Biology Platform of Imaging and Cytometry (BioPIC) for access to microscope facilities and the Biodiversité & Biotechnologies Marines (Bio2Mar) platform for access to technical equipment.

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.envpol.2025.127123.

## Data availability

Data will be made available on request.

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