# Plastic pollution and marine mussels: Unravelling disparities in research efforts, biological effects and influences of global warming

Marine Uguen, Sylvie M Gaudron, Laurent Seuront

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Marine Uguen, Sylvie M Gaudron, Laurent Seuront. Plastic pollution and marine mussels: Unravelling disparities in research efforts, biological effects and influences of global warming. Science of the Total Environment, 2025, 959, pp.178078. 10.1016/j.scitotenv.2024.178078. hal-05025195

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## Review

# Plastic pollution and marine mussels: Unravelling disparities in research efforts, biological effects and influences of global warming

Marine Uguen<sup>a, *,</sup> , Sylvie M. Gaudron<sup>a, b</sup> , Laurent Seuront<sup>a, c, d</sup>

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## HIGHLIGHTS

* Existing research on the effects of plastics on marine mussels was reviewed.

* Research efforts are concentrated on specific countries, species and plastic types.

* Impacts of plastics on marine mussels were found at every organisational level.

* Nanoplastics and leachates, less studied, showed greater impacts than microplastics.

* Interaction of plastics with temperature and other stressors should be investigated.

## GRAPHICAL ABSTRACT

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## ARTICLE INFO

Editor: Damià Barceló

**Keywords:**
Pollutant
Chemical
Heat stress
Bivalve
Reproduction
Toxicity

## ABSTRACT

The ever-growing contamination of the environment by plastics is a major scientific and societal concern. Specifically, the study of microplastics (1 µm to 5 mm), nanoplastics (< 1 µm), and their leachates is a critical research area as they have the potential to cause detrimental effects, especially when they impact key ecological species. Marine mussels, as ecosystem engineers and filter feeders, are particularly vulnerable to this type of pollution. In this study, we reviewed the 106 articles that focus on the impacts of plastic pollution on marine mussels. First, we examined the research efforts in terms of plastic characteristics (size, polymer, shape, and leachates) and exposure conditions (concentration, duration, species, life stages, and internal factors), their disparities, and their environmental relevance. Then, we provided an overview of the effects of plastics on mussels at each organisational levels, from the smaller scales (molecular, cellular, tissue and organ impacts) to the organism level (functional, physiological, and behavioural impacts) as well as larger-scale implications (associated community impacts). We finally discussed the limited research available on multi-stressor studies involving plastics, particularly in relation to temperature stress. We identified temperature as an underestimated factor that could shape the impacts of plastics, and proposed a roadmap for future research to address their combined effects. This review also highlights the impact of plastic pollution on mussels at multiple levels and

<footnote><sup>a</sup> Université de Lille, CNRS, Université du Littoral Côte d'Opale, IRD, UMR 8187, LOG, Laboratoire d'Océanologie et de Géosciences, Station marine de Wimereux, F-59000 Lille, France</footnote>

<footnote><sup>b</sup> Sorbonne Université, UFR 927, F-75005 Paris, France</footnote>

<footnote><sup>c</sup> Department of Marine Resources and Energy, Tokyo University of Marine Science and Technology, 4-5-7 Konan, Minato-ku, Tokyo 108-8477, Japan</footnote>

<footnote><sup>d</sup> Department of Zoology and Entomology, Rhodes University, Grahamstown 6140, South Africa</footnote>

<footnote>* Corresponding author.
E-mail address: marine.uguen@univ-lille.fr (M. Uguen).</footnote>
emphasises the strong disparities in research effort and the need for more holistic research, notably through the consideration of multiple stressors, with a specific focus on temperature which is likely to become an increasingly relevant forcing factor in an era of global warming. By identifying critical gaps in current knowledge, we advocate for more coordinated interdisciplinary and international collaborations and raise awareness of the need for environmental coherence in the choice and implementation of experimental protocols.

# 1. Introduction

The numerous advantageous characteristics of plastic, e.g. its versatility, resistance, lightness and affordability, have elevated this material to the third most produced man-made material worldwide (behind steel and cement; Geyer et al., 2017). Since the onset of its large-scale production in the 1950s, the total amount of plastic produced worldwide has been exponentially increasing, reaching 400.3 million tonnes in 2022 (PlasticsEurope, 2023). However, this production has profound disadvantages. The ever-growing production, overuse and subsequent rapid disposal and poor management of plastics have resulted in the global accumulation and distribution of plastic wastes (Geyer et al., 2017; Geyer, 2020; Hurley et al., 2020; Porta, 2021). Plastic is subsequently carried by several forces, such as the wind, wastewaters and rivers, to the ocean, which is the ultimate receptacle of these wastes. Although the quantity of plastic in the oceans each year is difficult to accurately quantify, it is estimated that ca. four to 12 million tonnes of plastic enter the marine environment yearly; this figure is expected to increase by an order of magnitude by 2050 (Jambeck et al., 2015), making plastics the most abundant and widespread marine waste globally (Auta et al., 2017).

Plastics typically vary in size, shape, colour, polymers and additive content and are mainly categorised by size. The largest ones (> 5 mm) are referred to as macroplastics, but the most abundant by far are those of smaller sizes (Eriksen et al., 2014), microplastics (1 µm – 5 mm; MPs) and nanoplastics (< 1 µm; NPs). Beyond their physical characteristics, plastic acts as a vector of numerous chemical compounds, which include (i) the large variety of additives incorporated into a plastic polymer during its manufacture to improve its performance (e.g. plasticisers, antioxidants, flame retardants) and (ii) the wide range of environmental contaminants sorbed from the surrounding environment (e.g. persistent organic pollutants and inorganic contaminants; Delaeter et al., 2022; Rochman, 2015). Through desorption, both of these types of chemical compounds associated with the plastic polymer may cause chemical damage to marine organisms, which adds to the physical damage caused by plastic particles (e.g. entanglement and obstruction of the digestive and/or respiratory tract; Auta et al., 2017; Delaeter et al., 2022; Seuront et al., 2022).

As filter feeders, mussels can filter several litres per hour; e.g. a 1 g dry weight *M. edulis* filter >7 l per hour (Riisgård, 2001), making them highly exposed to plastic pollution. Mussels may ingest large amounts of plastic particles; indeed, a review on plastic contamination in bivalves have shown that mussels from European shores have contaminant levels ranging from 0.04 to 4.44 MPs g⁻¹ of tissue wet weight (Hermabessiere et al., 2019). After ingestion, both MPs and NPs can migrate to mussel cells, tissues or the circulatory system, where MPs have been shown to persist for >48 days (Browne et al., 2008; Sendra et al., 2020a). Plastic particles can also be transferred through the food web, e.g. from mussels (*M. edulis*) to crabs (*Carcinus maenas*; Farrell and Nelson, 2013; Watts et al., 2014). Beyond the abovementioned absorption, translocation, accumulation and transfer phenomena, plastic particles and their leachates may cause numerous deleterious effects on mussels from subcellular to population levels (e.g. Avio et al., 2015; Capolupo et al., 2021a; Choi et al., 2022; Delaeter et al., 2022; Seuront et al., 2021; Uguen et al., 2024).

Adverse effects on intertidal mussels are likely to act on larger scales, given the valuable ecological and economic roles of these species. As key ecosystem engineers, mussels form unique habitats that enhance local

biodiversity as biodiversity hotspots (Borthagaray and Carranza, 2007; Buschbaum et al., 2009). Mussels also have important economic and heritage value due to professional and recreational fishing, e.g. mussels account for one-third of European Union aquaculture production, i.e. around USD 465 10⁶ (Avdelas et al., 2021). In this context, any direct and/or indirect effects that plastics have on mussels are likely to cascade through this ecosystem, as well as mussel goods and services, and may be further affected by higher temperatures.

As intertidal species, mussels are subject to a wide range of anthropogenically induced stressors, such as thermal stress and acidification (Kunze et al., 2021), that introduce an additional level of complexity to research fields devoted to assessing the impacts of plastic. Temperature is a particularly important stressor for intertidal mussels, as these organisms live close to their tolerance limits due to the emersion process (Somero, 2002). In this context, the combined impact of temperature and plastic pollution may amplify the adverse effects experienced by intertidal mussels, with synergistic effects representing the prevalent form of multistress interaction observed in marine species (e.g. see Dinh et al. (2022), Gunderson et al. (2016) and Holmstrup et al. (2010) for reviews). In an era of global change, temperatures rise and heat waves are expected to become more frequent and intense (IPCC, 2022). As such, there is a critical need to determine the combined effects of these stressors, which have predominantly been studied independently (Ford et al., 2022). In this context, this review provides a state-of-the-art analysis of the impacts of plastic pollution on marine mussels. By synthesising the available literature, we aim first to shed light on the disparities present in research efforts, encompassing aspects such as plastic characteristics and exposure conditions used. Second, we aim to relate the disparities to their environmental relevance and highlight potential similarities between exposure conditions in the laboratory and those in nature. We subsequently examined the plastic-induced effects on mussels across diverse organisational levels, ranging from cellular to organism levels. We finally discuss the role of temperature as a central factor affecting mussel during both immersion and emersion, emphasising the gap in the literature regarding the combined effects of thermal stress and plastic exposure on mussels, despite the significant potential of temperature to influence the impacts of plastics. Thus, we offer a comprehensive review of plastic pollution in marine mussels, emphasising our current knowledge on the complexity of the factors faced by these key organisms. We then conclude by highlighting areas in which research efforts need to be intensified to better develop and implement effective conservation and management strategies for intertidal ecosystems.

# 2. Methodology

The Web of Science database (assessed on May 25, 2023) was used with a search in Topics of the following terms in 3 distinctive search categories: (i) "microplastic*" OR "nanoplastics*" (ii) AND "effect*" OR "impact*" OR "affect*" (iii) AND "mussel*". The document type restriction "Article" was added. A total of 368 articles were returned, among which a preliminary screening was performed based on a close examination of their titles and abstracts. The articles excluded were those that (i) belonged to research domains other than biology, ecology, or toxicology (e.g. physics, chemistry), (ii) focused on species other than marine mussels (other taxa or freshwater mussels), (iii) were not related to plastic pollution and (iv) were review papers. A secondary screening excluded the literature focusing on mussel plastic ingestion without
evaluating their effects (53 papers), modelling (2 papers) and other aspects that lie well beyond the scope of the present work, i.e. how mussels enhance the sinking of plastic particles (5 papers). This approach resulted in the selection of 106 studies that are described in detail in the present work (Fig. 1).

The main information of each article, i.e. reference, publication year, mussel species, life stage, plastic polymer, shape and size, qualitative and quantitative assessment of the chemical compounds attached or bound to the particle and/or in the leachate solution, minimum concentration, maximum exposure duration, location, presence or absence of plastic particles during exposure, and consideration of the temperature factor as well as their main effects are provided in Supplementary Material S1.

The research effort was estimated for each parameter (e.g. species, location, polymer). In case where a single study investigated more than one variable for a parameter, each variable was counted separately as one study for this parameter. For instance, in a study that examined multiple polymer types (e.g. Polyethylene Terephthalate, Polypropylene, Polystyrene, Polyvinyl Chloride, and Rubber; see Capolupo et al. (2020)), each experiment involving one polymer type was considered as a separate research item, resulting here in five individual counts. Consequently, the denominator when assessing the research effort may exceed 106, depending on the parameter studied, this applies to location (n = 112), species (n = 113), polymer (n = 141) and shape (n = 112). This approach has been deliberately chosen to ensure an accurate representation of the research effort for each parameter. Note that, the same method was applied for the section “Main effects of plastics on mussels: an overview.” Details on the percentage calculation are provided in Supplementary Material, S2.

### 3. Disparities in research effort and their environmental relevance

While plastic research is among the most prolific fields of the last 50 years (see Seuront et al. (2022) for a review), the effects of plastics on marine mussels have only recently been addressed, starting with an article published in 2008 (Browne et al., 2008). The research area has

noticeably been growing rapidly, as >90 % of the 106 papers selected have been published since 2017, with 60 % within the last three years only (2021–2023; Fig. 2), without including the papers published from June to December 2023.

Despite the growing interest in this issue, research efforts have not been evenly distributed between the categories described above, leading to clear disparities. Many studies have concentrated on specific parameters, e.g. one plastic polymer, shape, and concentration, that may not be consistent with *in situ* observations. In this section, we have highlighted these disparities and provided recommendations to steer future research in a more comprehensive and effective direction.

#### 3.1. Spatial and dominant species disparities

Among the 106 studies that directly addressed the effects of plastics on marine mussels, experiments were conducted in 32 different countries. However, the research effort is far from being evenly distributed (Fig. 3A). Among the 32 countries represented, four collectively accounted for 50 % of research efforts, with China displaying the highest publication record (18 %), followed by Italy (16 %), Spain (9 %), and France (7 %). On a global scale, Europe (54 %) and Asia (28 %) dominated the research effort by far. Although the African continent is noticeably acknowledged for its high levels of plastic pollution (Lebreton and Andrady, 2019), this continent accounted for <5 % of the global research effort on this topic, as only 5 papers were spread among three countries, Tunisia (3 papers), Cape Verde (1 paper), and South Africa (1 paper). In addition, we did not find any study relevant to the issues discussed in the present work in Central America or most South American countries, with the exception of Chile (2 papers) and Brazil (4 papers), where research efforts were comparable to some North American countries.

Fourteen mussel species belonging to five different genera were studied, including *Brachidontes*, *Choromytilus*, *Mytilus*, *Perna* and *Semi-mytilus*. The genus *Mytilus* accounted for a substantial 88 % of the total

<figure><img src="image_5.png" /><figcaption>**Fig. 2.** Temporal trends in the total number of publications included in this review (n = 106). Studies exploring the effects of microplastics and nanoplastics on marine mussels are indicated in blue and yellow, respectively, while those focusing on the effects of leachates (i.e. specifically in the absence of MPs and NPs) are in orange. Studies examining multiple categories are denoted by cross-hatched patterns, corresponding to the respective colours. Asterisk indicates that this review was conducted in mid-2023 (May 25, 2023), and as such the data do not include the papers published from June to December 2023.</figcaption></figure>

<figure><img src="image_6.png" /><figcaption>**Fig. 1.** Flow chart of literature selection. Values (n) are the number of papers excluded at each stage of our screening process.</figcaption></figure>
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(caption on next page)
**Fig. 3.** (A) Research effort based on the number of publications per country that discuss the impacts of plastic pollution on marine mussels, from the start of the research (2008) to 25 May 2023. (B) Relative distribution of (i) the mussel genera (outer pie chart), *i.e.* *Mytilus spp.* (blue), *Perna spp.* (green), *Semimytilus spp.* (orange), *Brachidontes spp.* (grey), *Choromytilus spp.* (purple) and (ii) species (inner pie chart), with the value for the three most studied species, *i.e.* *M. galloprovincialis* (47 %), *M. edulis* (18 %), and *M. coruscus* (12 %). Note that *Mytilus spp.* refers to mussels of the genus *Mytilus* not identified at the species level potentially containing several species. (C) Global distribution of two mussel genera, *Mytilus spp.* (blue) and *Perna spp.* (green); darker green areas indicate regions where both genera are present, adapted from Gardner et al., 2016; Hilbish et al., 2012.

research effort (Fig. 3B). This dominance is primarily due to the use of *Mytilus galloprovincialis* as a model species, which accounts for almost 50 % of the total research effort (Fig. 3B), followed by *M. edulis* (18 %) and *M. coruscus* (12 %). *Mytilus* and *Perna spp.* dominate the majority of the world's coastlines, with *Perna spp.* prevalent in the tropical zone and *Mytilus spp.* in temperate zones (Fig. 3C). However, even though *Perna* was the second most studied genus, it only accounts for 8 % of the research effort. The relatively limited research effort on *Perna spp.* (8 %) compared to *Mytilus spp.* (47 %) likely results from disparities in ecological distribution and species abundance in their respective habitats, as well as, in the spatial distribution of research efforts (Fig. 3). Research efforts are concentrated in certain countries, primarily in temperate zones (95 % vs. 5 % in tropical zones; Supplementary Material, S1), resulting in more studies on *Mytilus spp.* found in these areas, compared to *Perna spp.* located in tropical zones (Fig. 3). Note that no studies have been carried out in the polar zones.

The concentration of research efforts towards specific countries and

specific mussel species may lead to skewed knowledge on plastic pollution in other coastal regions and different mussel species. Despite being underrepresented, tropical regions are not exempt from plastic pollution. Due to the lack of proper regulations, adequate waste management practices, and the import of plastic waste from developed countries, tropical countries are significant contributors to plastic waste in the ocean (Lebreton and Andrady, 2019; Meijer et al., 2021). In addition, the consumer demand for plastic is projected to increase, especially in developing countries, *e.g.* by 375 % from the current demand in Africa by 2060 (Lebreton and Andrady, 2019).

Research in the tropics is relatively scarce (5 %) despite the global presence of *Perna spp.* (Fig. 3A; Fig. 3C), possibly because research groups in many parts of Asia (excluding China) and Africa may not have the necessary infrastructures and financial resources to purchase the needed tools, such as Fourier transform infrared spectrometer, Raman spectrometer or pyrolysis-gas chromatography-mass spectrometry packages (see Seuront et al. (2022) for a discussion on this topic),

<figure><img src="image_10.png" /></figure>

**Fig. 4.** Relative distribution of the plastic size (A) polymers (B) and shapes (C) used in plastic impact studies on marine mussels. MPs: Microplastics, NPs: Nano-plastics. PS: polystyrene, PE: polyethylene, PVC: polyvinyl chloride, PP: polypropylene, PET: polyethylene terephthalate, PA: polyamide, PMMA: poly(methyl methacrylate), PEST: polyester fossil based, PLA: polylactic acid, PBS: polybutylene succinate, PHB: poly-β-hydroxybutyrate, PEVA: ethylene-vinyl acetate, paraffin and rubber. Asterisk indicates other polymers used, *i.e.* PEST: polyester fossil based, PEVA: ethylene-vinyl acetate, paraffin and rubber. NA indicates no available data.
despite the growing concern of the issues related to plastic pollution in these areas (Akindele and Alimba, 2021; Kapinga and Chung, 2020; Marnn et al., 2021; Matsuguma et al., 2017). The situation may be similar in Central America and South American countries, in which plastic pollution is a growing and pressing issue (e.g. Ardusse et al., 2021; Blettler et al., 2020; Denuncio et al., 2017; Ita-Nagy et al., 2022; Ivar do Sul and Costa, 2007). To address these challenges and achieve a more balanced representation of plastic pollution's impact on marine mussels worldwide, it is crucial to promote international collaborations, increase research capacity in underrepresented regions, and foster awareness of the importance of studying different mussel species in different geographical areas as well as their inter- and intraspecific variability in plastic pollution (Nicastro et al., 2023a). Notably, only a limited number of cross-border studies investigated the same parameter across multiple mussel species (Hamm et al., 2022; Seuront et al., 2021) or in the same species across different locations (Kolarević et al., 2023; Provenza et al., 2023).

## 3.2. Plastic characteristics: Size, polymer, shapes and leachates

### 3.2.1. Size

Most of the studies considered here primarily focused on MPs (1 µm – 5 mm), accounting for 82 % of the research effort (Fig. 4A). NPs (< 1 µm) studies represented 28 % of the research effort and 6 % of the research effort was specifically dedicated to studying the impact of plastic leachates, a topic that is further explored in detail hereafter (Section 3.2.4); notably, leachate solutions were consistently prepared from MP but not NP particles. Note that the total percentage of research effort exceeds 100 % as 16 % of the articles investigated more than one type of plastic. Specifically, 1 % (1 paper) examined the impact of both MPs and leachates in the same study and 15 % (16 articles) investigated MPs and NPs either individually or in combination. Most studies focused on the individual impacts of MPs and NPs on mussels, and only six articles examined the effects of a mixture of both particle sizes. The use of a NP and MP mixture provides a more realistic scenario, given the ubiquity of these two particle size classes in the marine environment (Eriksen et al., 2014; Ter Halle et al., 2017). Although the presence of MPs in wild mussels has often been reported (Li et al., 2019), the presence of NPs remains more elusive, even though their retention and accumulation have been demonstrated in the laboratory; see Zaki and Aris (2022) for a review. However, given that the number of particles in the environment is known to increase as the size decreases (Eriksen et al., 2014), NPs are expected to be more widely abundant than MPs. NPs are emerging contaminants and constitute a relatively new research field, indeed the impact of NPs on mussel was first studied relatively recently (Wegner et al., 2012) in contrast to studies on MPs (Browne et al., 2008; Fig. 2). In addition, the methods and technologies currently available to detect and quantify NPs in environmental matrices remain in their infancy (Zaki and Aris, 2022).

To thoroughly assess the overall impact on marine ecosystems, researchers must understand the influence of particle size on the toxicity and ecological consequences of plastic pollution. Particle size may induce different toxicity, as it influences the interaction mode between the particle and the organism, e.g. ingestion, retention time, translocation, and biomagnification (Zaki and Aris, 2022). Notably, compared to larger particles, smaller MPs have a higher potential to be translocated across the epithelial membrane of the gut to the mussel haemolymph and haemocytes, where they can persist for more than a month (Browne et al., 2008). In addition, these small MPs can be transferred through the food web, e.g. from mussels to crabs (Farrell and Nelson, 2013). The size of particles also affects their surface-to-volume ratio; thus, compared to larger particles, smaller particles exhibit a greater adsorption capacity, which can result in higher toxicity due to the adsorption of toxic compounds on their larger surface (Rodrigues et al., 2019).

### 3.2.2. Polymers

With the exception of one paper, polymer types were consistently indicated in the papers considered here. Specifically, a range of 14 different polymer types were used to assess the effect of plastic on marine mussels (Fig. 4B), with polystyrene (PS) being the most widely studied, accounting for ca. 50 % of the papers. The remaining polymer types ranked as follows: polyethylene (PE; 20 %, including low and high density), polyvinyl chloride (PVC; 6 %), polypropylene (PP; 6 %), polyethylene terephthalate (PET; 6 %), polyamide (PA; 4 %) and poly(methyl methacrylate) (PMMA; 4 %). These types were followed by polyester fossil-based (PEST; <1 %) and biobased, i.e. polylactic acid (PLA; 3 %), polybutylene succinate (PBS; <1 %), poly-β-hydroxybutyrate (PHB; <1 %). Finally, <1 % of the research effort involved ethylene-vinyl acetate (PEVA), paraffin or rubber.

Conducting experiments with representative polymer compositions in a given region is necessary to obtain valuable insights into the real ecological impact of plastic pollution on marine mussels. The dominant use of PS in marine mussel impact studies (Fig. 4B) conflicts with the typical polymer composition commonly observed in the natural environment. Specifically, a recent review (Bajt, 2021) shows that for several different mussel species and locations, PE is the predominant polymer found in marine mussels, followed by PP, e.g. for M. edulis from the French Atlantic coast (Phuong et al., 2018) and for M. galloprovincialis in the Adriatic Sea (Gomiero et al., 2019). PE and PP are the most produced polymers worldwide, with 26.9 % and 19.3 % of the global production in 2021, respectively (PlasticsEurope, 2022), and are the most frequently observed polymers in the sea-surface environment (Erni-Cassola et al., 2019). As mussels are filter feeders, the polymer composition inside mussels is representative of the polymer composition found in their environment (Li et al., 2019). Mussels typically filter not only one type of polymer but also a variety of polymers depending on their region and the related level of contamination (Bajt, 2021; Li et al., 2019). The ecological consequences of this diverse polymer variety have, however, only recently attracted attention, and a mere fraction of the research (2 %) investigated the effects of mixed polymer exposures (Bandini et al., 2021; Revel et al., 2019; Romdhani et al., 2022).

In addition, there is growing interest in the so-called 'biodegradable plastics' (i.e. PLA, PBS and PHB), which were first used in experiments with mussels in 2017 and only accounted for 4 % of the research effort; Fig. 4B). While these biodegradable plastics accounted for ca. 0.5 % of the global plastic production in 2022 (PlasticsEurope, 2023), they are often considered a potential solution to plastic pollution. However, their biodegradable properties are subject to considerable debate due to their restricted decomposition in natural environments; see Paul-Pont et al. (2023) for a recent discussion on this topic. These plastics are also not risk free, as they induce a range of alterations at various levels in mussels (Green et al., 2019; Green et al., 2017; Khalid et al., 2021; Magara et al., 2019; Phothakwanpracha et al., 2021), although exceptions exist (Joyce and Falkenberg, 2022).

### 3.2.3. Shape

Various particle shapes were used to investigate the effect of plastics on marine mussels, in which the following main shapes can be defined: spherical (e.g. beads and pellets), irregular (e.g. fragments and films) and fibrous (e.g. fibres and filaments; Fig. 4C). Among these, spherical shapes constitute the highest proportion of the research effort, i.e. 56 %. Irregular and fibrous shapes represent 16 % and 6 % of the research effort, respectively. Notably, 22 % of the studies did not specify the shape of the plastics used.

While spherical particles are the most commonly used in laboratory experiments, fibrous shapes are highly predominant, followed by irregular shapes in field investigations on plastic inside mussels; see Bajt (2021) and Li et al. (2019) for reviews. These plastics are defined as secondary plastics (i.e. they result from the degradation of larger plastic debris) and are the most prevalent plastic in marine environments (Ajith et al., 2020; Paul-Pont et al., 2018; Shim et al., 2018). Secondary plastics
are different from primary plastics, which are mainly spherical and directly produced in this form. However, plastic toxicity is intrinsically shape-dependent, as shape leads to different levels of potential tissue irritability (Ziajahromi et al., 2017) and to different mussel plastic uptake, retention and elimination (Li et al., 2021). These observations highlight the significance of providing accurate information on the shape of the plastic used, particularly in experiments focusing on the impact of plastic ingestion.

As fibres and filaments are the predominant plastic shapes found in marine mussels but the least studied, it is critical to intensify efforts to clarify their impact. Additionally, investigating a combination of shapes that mirrors the proportions found in the environment is essential to ensure the ecological coherence of this research area. In addition, although most studies showed a dominance of fibrous shapes in mussels followed by irregular and spherical shapes, some studies found opposite results (Ibryamova et al., 2022). Noticeably, local waste sources and accidental events, e.g. the X-press Pearl shipwreck in Sri Lanka, which released 70 billion pellets into the environment (James et al., 2022), can influence the proportion of the different plastic shapes. As such, it is highly recommended to systematically conduct in situ measurements before laboratory experiments to obtain environmentally relevant and meaningful results.

### 3.2.4. Plastic-associated chemicals

#### 3.2.4.1. A neglected aspect: absence of chemical composition analyses.
Plastics are intimately associated with a diverse array of chemical compounds that can have profound and a wide-range effect. Numerous additives (e.g. plasticisers, antioxidants and flame retardants), are incorporated into plastics during manufacturing process improving the functionalities, performances and/or appearances of the final item (Wiesinger et al., 2021). A recent inventory identified over 16,000 chemicals associated with plastics, with >4200 identified as having hazardous properties, and only 980 of these being subjected to international regulations (Wagner et al., 2024). In addition, during their residence in the aquatic environment, plastics typically sorb and accumulate a wide range of chemicals from the environment, e.g. heavy metals and persistent organic pollutants (POPs; Fries and Zarfl, 2012; Rodrigues et al., 2019; Yeo et al., 2020). Plastics can carry a complex mixture of chemicals that would not naturally co-occur, accumulating

compounds at concentrations up to 1 million times higher than their surrounding environment (Mato et al., 2001), and transporting them to remote areas (Yamashita et al., 2019). Thus, plastics act as sinks, vectors and also sources for a wide range of chemicals. Indeed, chemicals can desorb from plastic particles—known as plastic leachates— be transferred to aquatic organisms, bioaccumulated, biomagnified, and lead to numerous harmful effects (Delaeter et al., 2022; Hermabessiere et al., 2017; Liu et al., 2019; Peng et al., 2017). Incorporating the study of leachates in plastic studies is essential for determining whether observed changes result from the particles themselves, associated chemicals, or their interactions. However, despite their significances, the chemical impacts of plastics remain largely underexplored, and chemical analyses are rarely performed (Delaeter et al., 2022; Seuront et al., 2022).

Here, among the 106 papers analysed, only 8 % performed—at least partially—a chemical content analysis (Fig. 5). Therefore, most of the literature addressing the effect of plastic pollution on mussels (92 %) did not provide any information about the potential chemical composition of the MP particles that were employed. Notably, among them, 15 % considered implementing measures to avoid chemical compounds by using ‘additive-free’ plastics (7 %) or by ‘washing’ the particles to allegedly release their additives (8 %). However, none of the studies conducted a chemical analysis to validate the absence of additives or the efficacy of the washing process. In the 8 % of articles that conducted chemical content analyses, a specific list of compounds (e.g. bisphenol A, bis(2-ethylhexyl) phthalate) from particular chemical classes (e.g. additives, mineral contaminants) was targeted. For instance, the lists of screened additives are often very limited compared to the ever-growing number of additives used in the plastic industry (i.e. 481 so far as listed under the mapping exercise of the European Chemical Agency plastic additives initiative; www.echa.europa.eu, accessed June 2023), which *de facto* potentially overlooks important chemicals of the particles (e.g. toxic components). In addition, these studies essentially carried out qualitative rather than quantitative analyses, potentially overlooking the role of additive doses.

Chemical compounds can constitute a substantial part of the plastic, with plasticisers up to 60 % of the polymer's weight (Net et al., 2015). These chemical compounds can be released (i) inside the organisms after particle filtration or/and (ii) in seawater and interact directly or/and indirectly with the organisms. High quantities of chemical contaminants, including plastic additives, such as phthalates and bisphenols, have been recorded in mussels in nature (Castro et al., 2022;

<figure><img src="image_11.png" /><figcaption>**Fig. 5.** Relative distribution of chemical composition analyses in the 106 studies reviewed (inner pie chart) and origin of the effect found in those studies (outer pie chart).</figcaption></figure>
Hermabessiere et al., 2019; Rios-Fuster et al., 2022) and are known to have negative effects on organisms (Delaeter et al., 2022). Thus, the dearth of chemical screening analyses represents a significant knowledge gap that can hinder progress in research areas involving the assessment and clarification of the biological effects of plastic pollution. To address this gap, interdisciplinary collaborations are crucial, as a wide range of analytical methods are available to qualify and quantify these additives; see Seuront et al. (2022) for a statement on this specific aspect of plastic research.

### 3.2.4.2. Physical or chemical impact?
Among the 106 studies that examined the impacts of plastic on mussels, only 6% knew the origin of the impact, i.e. physical or/and chemical impact, exerted on organisms (Fig. 5). Thus, five papers specifically assessed the impact of plastic leachates by removing plastic particles —note that only MPs were used to make the plastic leachate solutions—from the solutions to which the organisms were exposed (Capolupo et al., 2020, 2021a; Gandara e Silva et al., 2016; Seuront et al., 2021; Uguen et al., 2022), and one paper assessed the exposure to particles and leachates individually, showing that the effect resulted from the physical impact of plastic particles (Wang et al., 2023b). In contrast, in 94% of the screened literature, the experimental design did not reveal whether the observed effects resulted from the physical presence of plastic particles, the chemical impact of the particles' leachates, or a combination of both factors. To address this concern, future studies should focus on isolating the impacts of plastic particles from those of leachates to clarify the relative contribution of each factor to the overall harm caused by plastics on marine organisms. Interestingly, 22 papers investigated the effect of plastic particles alone or in combination with an added pollutant (e.g. benzo(a)pyrene, fluoranthene, cadmium; Supplementary Material, S1) on marine mussels. However, only two of these 22 papers screened the chemical composition of their plastic. Rather than adding a specific type of pollutant, leachates should be used from plastic particles collected from (or incubated in) the field to clarify the environmental chemical cocktail that plastic might accumulate.

One study compared the effects of leachates from raw industrial plastic pellets (virgin pellets) with those from plastic pellets found on beaches (beached pellets) in marine mussels (Gandara e Silva et al., 2016), showing that leachates from beached pellets exert a higher impact than that of virgin pellets. However, the lack of leachate chemical screening in the study of Gandara e Silva et al. (2016) prevents direct comparison with other studies with an opposite pattern, e.g. on oysters at early life stages (Gardon et al., 2020; Tallec et al., 2022a). Gandara e Silva et al. (2016) were nevertheless the first authors to investigate the impact of leachates on marine mussels, and most (80%) of the subsequent papers that addressed this topic conducted chemical screenings. However, the methodological approaches to prepare leachate solutions are quite different. For instance, the particle incubation time varied from 24 h (3 papers) to 14 days (2 papers), while the leaching temperatures varied from 10 °C (1 paper) to 25 °C (2 papers; the remaining papers provided no information about these parameters) and were not always aligned with the incubation temperature of the model organism. Although they do not jeopardise the central point of these papers, these discrepancies nevertheless prevent the results from being unambiguously compared. Notably, a protocol has been recently proposed to standardise the methodology for obtaining MP leachates (Almeda et al., 2023).

Compared to the abundant literature on MP and NP physical effects, the impact of leachates is generally neglected (Delaeter et al., 2022; Seuront et al., 2022); however, leachates are a major contaminant transfer route. A study that compared two contaminant bioaccumulation pathways in mussels found that uptake through the aqueous phase (leachates) was a more significant pathway than plastic particle ingestion (Jang et al., 2021). By isolating the effects of plastic particles from leachates, considering the environmental chemical composition, using a

standardised protocol and conducting chemical screenings, future studies can provide more comprehensive and accurate insights into the complex interactions between plastics and marine ecosystems.

In addition to plastic chemical and physical impacts, the colonisation of plastic by organisms, referred to as the “plastisphere”, may also induce biological impacts; see Bowley et al. (2022) and Sooriyakumar et al. (2022) for reviews. The surfaces of plastics are colonised by a succession of organisms, including bacteria that form a biofilm and then more complex eukaryotic organisms; these plastic-associated communities generally differ from those in the surrounding environment (Bowley et al., 2022). Plastic facilitates the transport of their associated communities over long distances, thereby accelerating the dispersal of organisms and potentially leading to biological invasions (Bowley et al., 2022; Carlton et al., 2017; Shen et al., 2019). The transport-associated organisms may also include pathogenic species, e.g. Vibrio spp. and Escherichia spp., which may induce disease upon ingestion of plastic particles; see Bowley et al. (2022) for a review. In marine mussels, only one study investigated the effects of plastic exposure with bacterial biofilm. This study found that an exposure to plastic with bacterial biofilm induced a significant immune response and affected the microbial composition of their gills, compared to mussels exposed to sterile plastics (Bandini et al., 2021). In addition, plastic particles enhance gene exchange among attached biofilm communities and between biofilms and surrounding environment, fostering the transfer of pathogenic and antibiotic resistance genes (Shen et al., 2019). Given the multifaceted potential biological impacts of plastic particles outlined above, future studies must delve deeper into their exact role as biological vectors (Sooriyakumar et al., 2022).

## 3.3. Exposure conditions

### 3.3.1. Plastic concentration

All studies provided information about the concentrations of plastics they used. However, these concentrations were expressed in a range of units, with the most common being mass/volume (e.g. ng mL⁻¹, µg L⁻¹ or mg L⁻¹) and items/volume (e.g. NP mL⁻¹, MP mL⁻¹, NP L⁻¹, MP L⁻¹, MP mm⁻³), while the least frequently used unit was volume/volume (e. g. mL L⁻¹). According to the plastic characteristics (polymer, size and shape), these three units of measurement may vary considerably. Due to their complementarity, reporting the concentration in mass/volume, item/volume and volume/volume would provide a more straightforward overview of the concentration used and thus improve the reproducibility of research.

For a representative overview of the concentration used in the plastic impact on marine mussel studies, we focus on studies that used mass/volume units (66% of the studies; 70 papers). Among them, 52% (36 papers) used a range of different concentrations. To study the environmental relevance of the concentration used, we only examined the minimum concentration employed in each study. By considering the minimum concentration, we increase the likelihood of identifying whether at least one concentration used could be compared to environmental relevant levels, thus ensuring some degree of environmental applicability. All concentrations were converted to µg L⁻¹ to ensure comparability.

The minimum concentrations used varied from $10^{-4}$ to $10^8$ µg L⁻¹, with almost 50% (49%, 34 papers; Fig. 6A) of the research effort in the 1–1000 µg L⁻¹ range (median = 200 µg L⁻¹). Based on the highest plastic concentration reported (i.e. 1890 µg L⁻¹; see Isobe et al., 2015 in Beiras and Schönemann, 2020), 71% (Fig. 6A) of the research papers reviewed used environmentally relevant concentrations. Notably, environmental plastic mean concentrations were generally far below 1890 µg L⁻¹, ranging from 0.01 to 34.09 µg L⁻¹, depending on the geographical region (Beiras and Schönemann, 2020). However, these measurements were taken at the sea surface, with a defined mesh size and far from the coast, which may bias the estimation of the plastic concentration faced by mussels, especially those living in coastal waters. Coastal systems are
<figure><img src="image_12.png" /><figcaption>**Fig. 6.** (A) Distribution of the minimum concentration used in plastic impact studies on marine mussels. The dashed line represents the highest concentration recorded in the environment, i.e. 1890 µg L⁻¹ (= 1.89 mg L⁻¹; see Isobe et al., 2015 in Beiras and Schönemann, 2020). The data presented here are based only on concentrations in µg L⁻¹, available in 70 studies. (B) Relative distribution of the mussel life stages used to assess the impact of plastics on marine mussels, pelagic stages include gametes, embryos and larvae, and benthic stages include juveniles and mature mussels.</figcaption></figure>

among the most heavily plastic-polluted ecosystems (Serra-Gonçalves et al., 2019; Wu et al., 2022), and the plastic concentration faced by benthic species at the water-sediment interface can be even higher (Besseling et al., 2014). Additionally, mussel beds, with their spatial structure and important filtration rate, alter the water flow, increase turbulence, and act as local hotspots of benthic MP pollution (e.g. Lim et al., 2020; Van Colen et al., 2021); this effect can be significantly enhanced by mussel shell endolithic infestation, which further increases the rugosity of beds (Nicastro et al., 2022). As a consequence, mussels may be exposed to concentrations greater than expected, and the plastic concentration used in the experiment may not be as unreasonable as previously thought (e.g. Delaeter et al., 2022; Paul-Pont et al., 2018; though these figures were obtained outside the mussel literature).

The highest minimum concentration found in this review, i.e. a

concentration of 1.2 × 10⁷ µg L⁻¹ (= 1.2 × 10⁴ mg L⁻¹), was used to study the impact of plastic leachate on marine mussels (Seuront et al., 2022). In this experiment, particles were incubated in seawater for 24 h before being removed. As (i) the particles are not intended to be ingested by mussels and (ii) the desorption of additives may be a lengthy process that can take several weeks (Paluselli et al., 2019), the short incubation time used potentially counterbalances the high plastic concentration employed. Although the resolution of this specific issue lies well beyond the scope of the present review, these observations warrant the need for further work to deciphering the relative effect of particle concentrations, the chemical components of particles, the leaching time and the duration of exposure to the resulting leachate solutions.
### 3.3.2. Exposure duration

When an experiment used several exposure durations (11 %), only the maximum duration was considered here for clarity. Each experiment covered a wide range of exposure times, i.e. from a few hours to a maximum of 91 days, with a median exposure duration of 7 days. Most studies (ca. 50 %) used standard exposure durations in line with OECD guidelines, i.e. 7 days (15 %), 4 days (12 %), 21 days (12 %) and <24 h (10 %). Long-term exposures are typically chosen to reflect environmental conditions, involving lower plastic concentrations than in shorter experiments. However, no significant correlation was found between the maximum duration of the experiment and the minimum concentration used (p = 0.6618).

Finally, despite the growing interest in (i) full life cycle tests (i.e. stressor exposure from early-life stages to mature individuals), and (ii) multigenerational tests (i.e. stressor exposure across multiple generations; Guimarães et al. (2023)), neither approach has been explored to examine the impact of plastics on marine mussels. Plastics and their associated chemicals induce several adverse intergenerational effects in terrestrial and aquatic species (see Junaid et al. (2023) for a review), including the marine bivalve *Magallana gigas* (Bringer et al., 2022; Sussarellu et al., 2016). Parental exposure to a different type of plastic was found to result in motility and development impairment of their (nonexposed) offspring with parents exposed to concentrations ranging from 100 to 10,000 µg L⁻¹ in Bringer et al. (2022) and 23 µg L⁻¹ in Sussarellu et al. (2016). Using full life cycle and multigenerational tests in mussel studies would provide an opportunity to clarify the impacts of plastics on these organisms.

### 3.3.3. Mussel life stage and other internal factors

The sensitivity to a contaminant has been shown to vary with the life stage of an organism (Mohammed, 2013). Mussel life stages were defined here as (i) pelagic early-life stages (gametes to larvae) or (ii) benthic life stages (juvenile to adult). Among the 106 papers screened, nine did not contain specific information on mussel size or life stages. However, through contextual information in these nine papers, we could classify the individuals in the benthic mussel life stage category. In the plastic impact on marine mussel fields, benthic life stages received much more attention than pelagic life stages, accounting for 97 % and 7 % of the research effort, respectively (Fig. 6B). Note that the total percentage of research effort exceeds 100 % as 4 % of the articles examined the plastic impact on both stages (Auguste et al., 2021; Capolupo et al., 2020, 2021b; Franzellitti et al., 2019; Fig. 6B).

Morphological and physiological differences between pelagic and benthic life stages can lead to variations in their exposure and vulnerability to contaminants, with early life stages generally being more sensitive to toxicants (Mohammed, 2013). This phenomenon was also recently described for MPs in a global pattern analysis of benthic species (Mason et al., 2022), highlighting the importance of considering the full life cycle of mussels when assessing the effects of pollutants.

Note that other factors internal to an organism, such as its sex, sexual maturity, size, and genotype, can also influence the bioaccumulation of a pollutant (Boening, 1999) and its potential effects. These factors, however, remain underexplored, representing <4 % of the research effort. Specifically, out of the 106 studies reviewed, only 3 took into account the mussel sex (Abidli et al., 2021; Choi et al., 2022; González-Soto et al., 2022), 1 the sexual maturity (Laubscher et al., 2023), and none the mussel size and genotype (although see recent studies by Uguen et al. (2023) and Cozzolino et al. (2024)). Sex is also a critical parameter in pollutant impact studies, although this parameter has received very little attention, especially in ‘invertebrate’ studies (Burger, 2007). Sex-specific effects may be caused by plastic additives, as some can mimic hormones and act as endocrine disruptors; a typical example is polybrominated diphenyl ethers (PBDEs), flame retardants that induce mussel sex-specific effects (Ji et al., 2013). Sexual maturity can also modulate the plastic effects, e.g. MP impact on clearance rate was more pronounced in mussels that had recently spawned (Laubscher et al.,

2023). This highlights the importance of sex and sexual maturity when evaluating the effects of pollutants on organisms and underlines the need for further research in this area to clarify the potential sex-related differences in pollutant sensitivity.

## 4. Main effects of plastics on mussels: An overview

This section aims to examine the state-of-the-art knowledge and understanding on the influence of plastic on marine mussels. The impact of plastics on mussels has been studied in depth at different organisational levels, from small-scale impacts encompassing molecular, cellular, tissue and organ levels to organism-level effects encompassing functional, physiological, and behavioural traits, and, more generally, in terms of their large-scale effects, e.g. on associated communities. Most research effort has focused on small-scale impacts (57.2 %), followed by effects at the organism level (41.4 %), with limited exploration of larger-scale impacts (1.4 %; Fig. 7A; Supplementary Material, S2). For more details regarding the presence or absence of effects at any organisational levels for each study, and the impact percentage calculation, see Supplementary Materials S2 and S3. Note that in order to improve the readability, a large number of references in this section has been moved and summarised in Supplementary Material, S3. This Supplementary Material indicates for any specific studied responses (e.g. DNA damage, changes in gene expression) which articles investigated it and whether they found an impact depending on the type of treatment (e.g. MPs, NPs, Leachates).

Among the 106 studies we examined in the present work, the vast majority (91 %) found plastic-induced effects on at least one mussel organisational level. Notably, MPs have been the most extensively studied in relation to mussels, although NPs and leachates have shown more pronounced effects on these organisms (Fig. 7B). The effects found following the addition of pollutants to plastics (see Section 3.2.4; e.g. benzo(a)pyrene, fluoranthene, cadmium), their combined effects with plastic particles and effects alone are not considered in this review.

It is stressed here that, while observed effects span multiple organisational levels, they likely produce cumulative impacts on ecosystem structure and function due to the key role mussels play as ecosystem engineers. The potential for these various effects to interact synergistically in this key species highlights the need for future systematic studies to better understand these complex interactions.

### 4.1. Small-scale impacts

The small-scale effects of plastics range from molecular alterations to tissue and organ impairments. Molecular impact was the most studied of the small-scale effects (58.1 %), followed by cellular impacts (26.5 %) and tissue and organ impacts (15.4 %). Among the actual number of papers that focused on small-scale effects, 92 % found an effect (Supplementary Material, S2). These effects were predominantly observed at the molecular level (91 %; 72 out of 79 papers), followed by the cellular level (83 %; 30 out of 36 papers) and finally in tissues and organs (71 %; 15 out of 21 papers; Supplementary Material, S2). The molecular level is often the first level affected by toxicants, leading to disruptions in DNA, gene expression, enzymatic activities, oxidative stress and biochemical content (see Section 4.1.1 for more details). The cellular level may be affected through changes in organelles, haemocytes, and directly in cell viability (see Section 4.1.2 for more details). Finally, these alterations may subsequently extend to the tissue and organ levels, with histo-pathological effects on different tissues and organs, and disrupt mussel microbiota (see Section 4.1.3 for more details). The effects and studies are detailed in the sections below.

#### 4.1.1. Molecular impact

##### 4.1.1.1. DNA disruption and changes in gene expression. DNA damage, i.
<figure><img src="image_13.png" /></figure>

<figure><img src="image_14.png" /><figcaption>**Fig. 7.** (A) Relative distribution of the research effort across impact levels, i.e. small scales (57.2 %), organism level (41.4 %) and large-scales (1.4 %) and the main effects found. (B) The relative distribution of research efforts and their impacts across various organisational levels (Molecular, Cellular, Tissue & Organ, Organism, and Ecology) and types of plastics studied: Microplastics (MP), Nanoplastics (NP), mixed plastic treatments (Mix; e.g. MP + NP), and plastic leachates. The size of each circle represents the number of publications, while the colour intensity indicates the percentage of studies reporting an impact (% Impact). N/A indicate the absence of studies for this specific combination of plastic type and organisational level. Details on the percentage impact calculation are provided in Supplementary Material, S2.</figcaption></figure>
e. DNA fragmentation and the presence of micronuclei, was investigated in 21 articles, with 17 of them reporting significant increase following exposure to MPs and NPs (Supplementary Material, S3). MPs and NPs also induce changes in DNA content in one study and several changes in gene expression in mussels (Supplementary Material, S3). Thus, among the 22 studies that determined gene expression, 18 found alterations following exposure to MPs and NPs (Supplementary Material, S3). This encompassed a wide variety of genes, such as those linked to oxidative stress responses, immune system function, apoptosis regulation, byssus protein production, shell matrix formation, and various metabolic pathways. However, some studies (n = 4; Supplementary Material, S3) did not find that MPs, NPs, or plastic leachates affect mussel gene expression.

#### 4.1.1.2. Change in enzymatic activity.
A wide range of enzymes involved in various physiological processes, such as digestion, neurotransmission, metabolism, detoxification, and antioxidant defense, were significantly influenced by plastic exposure. The activities of digestive enzymes (i.e. amylase, cellulase, lipase, protease, trypsin and pepsin) were altered by MPs and NPs in four studies, with two exceptions (Supplementary Material, S3). Following exposure to MPs, NPs and plastic leachates, neurotoxic effects were induced in seven studies, such as changes in the activity of the neurotransmitter enzyme cholinesterase, although seven studies reported no plastic-induced effects on this enzyme (Supplementary Material, S3). Metabolism-associated enzymes, i.e. lactate and isocitrate dehydrogenase, acid and alkaline phosphatase, esterase, aspartate aminotransferase, alanine transaminase and acyl-CoA oxidase were also affected by MPs and NPs in 8 studies with six out of the fourteen studies reporting no significant effects on these enzymes (Supplementary Material, S3). Moreover, plastic exposure led to changes in enzymatic activity, which can subsequently alter cellular defense mechanisms. Activity alterations were observed in antimicrobial enzymes, mainly lysosomes, following exposure to MPs and NPs in eight studies (Supplementary Material, S3). Plastic exposure also affects xenobiotic detoxification enzymes in one study and antioxidant enzyme production (see the next part on oxidative stress).

In addition to enzyme-related effects, plastic exposure led to an increase in heat shock proteins (HSP70) in one study (Supplementary Material, S3). One study showed that plastics also disrupted the regulation of signaling proteins, including p38 MAPK and PKC, and another one showed that plastics resulted in a decrease in the production of male and female hormones, namely, testosterone and estradiol (Supplementary Material, S3).

#### 4.1.1.3. Oxidative stress.
Oxidative stress, the most studied parameter, was examined in 56 studies, with 43 showing an impact of plastics. Reactive oxygen species (ROS) generation was quantified in 17 studies, in 15 of them the production of ROS was altered (mainly through an increase) when exposed to MPs or NPs (Supplementary Material, S3).

To counteract the potential damage caused by ROS, the cells produce antioxidants. The content of antioxidant enzymes, i.e. catalase, superoxide dismutase, peroxidase, glutathione S-transferase, glutathione peroxidase and/or glutathione reductase, was affected by MPs, NPs and plastic leachates in 28 studies (Supplementary Material, S3). Nonenzymatic antioxidant molecule content, such as glutathione, was also modified by exposure to MPs and NPs in all four studies (Supplementary Material, S3). Thus, out of the 41 articles that examined the antioxidant system, only 12 reported no observed effects caused by plastic (Supplementary Material, S3). In addition, the total antioxidant capacity was impaired after exposure to MPs and NPs in four out of eight studies (Supplementary Material, S3). However, several methods were used (e.g. ferric reducing antioxidant potential, total oxyradical scavenging capacity towards peroxyl and hydroxyl) and a universal protocol for estimating total antioxidant capacity is lacking (Flieger et al., 2021).

If the antioxidant defense system cannot adequately counterbalance

ROS production, oxidative stress may occur. Oxidative stress induces cellular damage, such as oxidative modifications of lipids, proteins, and DNA. Notably, of the 28 studies examining lipid peroxidation (LPO), 18 showed an alteration by MPs, NPs and plastic leachates and 10 reported no discernible plastic-induced effects (Supplementary Material, S3). Although lipofuscin and protein carbonylation are less extensively studied than LPO, the accumulation of these compounds was associated with oxidative stress. Increased levels of lipofuscin content in the cells were measured after exposure to MPs, NPs and plastic leachates in 6 studies; although one did not find any effect on this parameter (Supplementary Material, S3). Changes in protein carbonyl concentrations were also observed in two studies following NP and MP exposure (Supplementary Material, S3).

#### 4.1.1.4. Other changes in biochemical content and cellular budgets.
Biomolecule content is affected by plastics, which may lead to cellular energy budget impairment. All metabolomics studies (n = 8; Supplementary Material, S3) revealed the effects of plastics, showing numerous changes in metabolites, e.g. amino acids, carbohydrates, and nucleotides, following exposure to MPs and NPs. Neutral lipid content was assessed in 13 studies, among which six recorded an alteration (mainly through an increase), i.e. by MPs (n = 4), NPs (n = 1) and plastic leachates (n = 1; Supplementary Material, S3). MPs did not impair neutral lipid content in the remaining studies. Other changes, such as protein abundance and sugar content, were also identified in two studies (Supplementary Material, S3). A decrease in cellular energy allocation by MPs was found in two studies, although no effect was found in two others (Supplementary Material, S3). These changes in the cellular energy budget may compromise vital cellular processes and overall organism health and performance.

### 4.1.2. Cellular impact

#### 4.1.2.1. Organelle impact.
Among the effects reported at the cellular level, lysosomal membrane stability (LMS) was the most extensively studied effect (22 studies). While 9 studies did not observe any impact, a decrease in LMS was reported following exposure to MPs, NPs, both MPs and NPs and plastic leachates in 13 studies (Supplementary Material, S3). Other impacts on lysosomes were observed, with one study showing changes in lysosome abundance, two an increase in the lysosome-cytoplasm volume ratio and two others an impact on other organelles, e.g. damage to the rough endoplasmic reticulum (RER) and mitochondria (Supplementary Material, S3). Mitochondrial damage can impair energy production, cellular metabolism and the regulation of various signaling pathways (Lee et al., 2022), while RER damage can disrupt protein synthesis by affecting the proper coordination of protein folding and maturation (Koksal et al., 2021).

#### 4.1.2.2. Haemocyte impairment.
Alteration of cell phagocytic activity was examined in 19 studies, and 15 studies reported an effect. Most of these studies (n = 13) found a decrease in mussel cell phagocytic activity upon exposure to MPs and NPs (Supplementary Material, S3). The remaining studies found a more nuanced response, with some indicating an increase (n = 2; Katsumiti et al., 2021; Pittura et al., 2018) or no significant alteration in the phagocytosis response (n = 4; Supplementary Material, S3). Plastic also influences haemocyte motility in one study (Supplementary Material, S3). These changes in motility and phagocytosis could be due to the decrease in haemocyte cytoskeleton actin filaments (F-actin) as shown in a study following an exposure to NP (Supplementary Material, S3). Moreover, the granulocyte-hyalinocyte ratio (two distinct types of haemocytes) was found to decrease in two out of four studies, affecting the number of cells involved in the phagocytic process.

#### 4.1.2.3. Global cell viability.
Cell morphology was altered in all studies
that examined this parameter (n = 4) following exposure to MPs, NPs and plastic leachates (Supplementary Material, S3). The observed changes included reduced membrane integrity, cell vacuolisation or loss of filopodia. All these alterations might have implications for cell survival. In seven studies, exposure to MPs and NPs led to an increase in apoptotic and necrotic cells, while four studies reported reduced cell viability and four others reported shifts in haemocyte numbers (Supplementary Material, S3). No studies observed an absence of effects on these parameters.

Notably, all these cellular effects were modulated by various parameters, such as the polymer type (Avio et al., 2015; Capolupo et al., 2020), concentration (Auguste et al., 2021; Canesi et al., 2015, 2016; Huang et al., 2022; Katsumiti et al., 2021; Pavičić-Hamer et al., 2022), particle size (Auguste et al., 2021; Capolupo et al., 2021c; Choi et al., 2021; Cole et al., 2020; Jong et al., 2022), exposure duration (Auguste et al., 2021; Pavičić-Hamer et al., 2022; von Moos et al., 2012), or the type of cell studied (Sendra et al., 2020b).

### 4.1.3. Tissue and organ impact

#### 4.1.3.1. Histopathological changes.
Mussel histopathology was assessed in 17 studies, among which 12 found significant changes (Supplementary Material, S3). The gills, the digestive gland and other tissues from the digestive system (e.g. stomach and gut) were the most extensively studied tissues. Significant abnormalities in these tissues, e.g. hypoplasia, hypertrophy, atrophy, inflammation, haemocyte infiltration, ceroids and necrosis, were found after exposure to MPs and NPs. Abnormalities generally increase with plastic concentration (Alnajar et al., 2021) and were more severe in the presence of NPs than MPs (Qi et al., 2023). Interestingly, no tissue recovery was observed after one week of depuration (Hariharan et al., 2021). Although less studied, exposure to MPs also led to histological alterations in other tissues, such as the foot and adductor muscles (Hariharan et al., 2021). Gonadal inflammation of both sexes and oocyte degeneration were also found after MP particle exposure (Choi et al., 2022; Ciocan et al., 2020; González-Soto et al., 2022; Paul-Pont et al., 2016). No histopathological anomalies were found in the five remaining studies following exposure to MPs, NPs, a mix of both or plastic leachates (Supplementary Material, S3).

#### 4.1.3.2. Changes in microbiota composition.
There has been a recent and growing interest in studying microbiota due to the essential role played by these complex microbial communities in maintaining tissue integrity and contributing to the overall health of their host. As a relatively new research field, the impact of plastic on mussel microbiota has only been investigated by five studies. Of these, four studies found significant effects of plastics (Supplementary Material, S3). The gill (Bandini et al., 2021), digestive gland (Ferguson et al., 2022) and gut (Li et al., 2020b) microbiota compositions were altered by MP exposure. NPs also affected microbiota composition in mussel haemolymph (Auguste et al., 2020), predominantly towards a reduction in microbial abundance and diversity. Notably, the severity of the impact was enhanced by higher plastic concentrations, longer experiment durations, and the presence of biofilms on plastic particles (Bandini et al., 2021; Ferguson et al., 2022; Li et al., 2020b). Interestingly, an 8-day depuration period was insufficient to restore the mussel microbiota after plastic exposure (Li et al., 2020b). Although one study (Yang et al., 2021) found that MPs did not significantly affect the mussel gut microbiota composition, it is noteworthy that plastic pollution may impact the overall microbiome; the observed alterations in microbial communities indicate the potential for broader ecological consequences.

### 4.2. Organism-level impacts

At the organism level, the impacts of plastic were categorised through their effects on (i) functional traits (41.7 %), (ii) physiological

traits (41.7 %), and (iii) behavioural traits (16.6 %). Functional traits encompassed parameters related to growth, survival, and reproduction. Within the reproduction category, we included the effects of different organisational levels when they were exerted on mussels at early-life stages. Physiological endpoints comprised filtration, respiration, adsorption, excretion, scope for growth, heart rate, and shell regeneration. Finally, the behavioural traits category included byssus production, aggregation, displacement, valve movement and antipredator behaviours. Plastics were found to affect mussels in 73 % of the studies considering organism-level responses (Supplementary Material, S2). Interestingly, among the papers considering behavioural traits, 79 % (11 out of 14 papers) identified significant impacts of plastics despite being the less frequently studied category. In comparison, among the papers considering physiological traits, 66 % (23 out of 35 papers) found significant effects, while 54 % (19 out of 35 papers; Supplementary Material, S2) of those examining functional traits reported significant impacts. The effects and studies are detailed in the sections below.

#### 4.2.1. Functional traits

##### 4.2.1.1. Growth and condition index.
The effect of plastics on the mussel growth rate, which is typically estimated from length, width, or weight, was assessed in seven studies. With the exception of one study, in which a decrease in mussel growth rate was observed following exposure to MPs (Détrée and Gallardo-Escárate, 2018), no effect was ever reported after exposure to MPs or to a combination of MPs and NPs (Supplementary Material, S3). To the best of our knowledge, there are no studies on the impact of NPs or plastic leachates on mussel growth. Larval growth was also altered by plastic; see Section 'Reproduction and larval development' for further details.

Based on mussel morphological parameters, the mussel condition index was assessed in 18 studies, and five studies identified a significant impact of plastics (Supplementary Material, S3). Specifically, exposure to MPs led to a decrease in the condition index (Supplementary Material, S3), except for one study, in which an increase was observed (González-Soto et al., 2019). The response to plastic exposure is complex and influenced by several factors, such as the mussel feeding regime (Shang et al., 2021a), plastic size (González-Soto et al., 2019; Pavičić-Hamer et al., 2022), its concentration (Pavičić-Hamer et al., 2022) and its polymer composition (Barkhau et al., 2022), e.g. the condition index was impacted by PVC but not PMMA particles (Barkhau et al., 2022). In addition, differences in research methodologies (e.g. used of dry weight, ash-free dry weight, mussel shell length) may contribute to the diverse findings reported in the literature. No significant effects on the condition index were detected in the remaining studies, regardless of whether the exposure involved MPs, NPs, or a combination of both (Supplementary Material, S3).

##### 4.2.1.2. Survival rate.
The survival rate following plastic exposure was investigated in 12 papers, among which four studies found an effect (Supplementary Material, S3). MPs were found to decrease the mussel survival rate in two species, the green mussel *P. viridis* (Phothakwanpracha et al., 2021; Rist et al., 2016) and the Mediterranean mussel *M. galloprovincialis* (Pavičić-Hamer et al., 2022; Provenza et al., 2023). Mussel mortality was influenced by several factors, including polymer, colour and size as well as plastic concentration, duration of the experiment, presence or absence of food and the site of mussel collection (Pavičić-Hamer et al., 2022; Phothakwanpracha et al., 2021; Provenza et al., 2023; Rist et al., 2016). Compared to a lack of mortality in control experiments, *M. galloprovincialis* reached a mortality maximum of 66.7 % and 63.3 % in unfed and fed conditions, respectively, when exposed to $5.0 \times 10^4 \mu g L^{-1}$ PE for 7 days (Provenza et al., 2023). Exposure to a concentration of $2.16 \times 10^6 \mu g L^{-1}$ of PVC led to total mortality in *P. perna* in the first ca. 50 days (Rist et al., 2016). Compared to control conditions in which no mortality occurred, small, medium and
large MPs induced ca. 100 % mortality in a 4-day experiment on *P. viridis* at concentrations ranging from 66 to 1333 MP L⁻¹ for PP and at concentrations of 333 to 1333 MP L⁻¹ for PBS and PS (*Phothakwanpracha et al., 2021*). Finally, Pavicić-Hamer et al. (2022) assessed mussel mortality through a stress-on-stress experiment, showing that mortality occurred more quickly in mussels contaminated by PMMA (5 to 10 days earlier than in the control) depending on the concentration used (0.0001 to 0.01 µg L⁻¹) and size of the particles used. For the remaining 8 papers, no effect of MPs or NPs on mussel survival was recorded (Supplementary Material, S3). Notably, no studies investigated the impact of plastic leachates on mussel survival. Mussel early-life stage survival was also affected; see Section 'Reproductive success and larval development' for further details.

### 4.2.1.3. Reproductive success and larval development.
Mussels are broadcast spawners and gonochoric species, in which fertilisation and larval development occur in the water column (pelagic stage) until the veliger larvae settle to the benthic hard substrates at the pediveliger stage; then metamorphosis is completed and the larvae reach the juvenile stage, starting their benthic life (Bayne, 1976). Ten scientific articles investigated parameters related to reproductive success, e.g. fertilisation rate, embryo and larval survival rates, or reproductive system impairment of adult mussels. All these papers showed that plastics affect at least one of these parameters.

The mussel fertilisation rate was investigated in MPs and leachate exposures but not under NP exposure. While there was no observed effect of PS MP particles on the *M. galloprovincialis* fertilisation rate (Capolupo et al., 2021b), leachates (no particles) from all the tested polymers (PS, PP, PET, PVC, rubber) were found to induce a decrease in this parameter (Capolupo et al., 2020). In contrast, mussel embryos and larvae were consistently impacted by MPs, NPs and leachates. Larval motility was only analysed under plastic leachate exposure, revealing a polymer-dependent impact. A decrease in larval motility was found with rubber and PVC leachates but not PS, PP, or PET leachates (Capolupo et al., 2020). Larval development was impaired (i.e. developmental delay, shell formation, malformation and/or mortality) by both plastic leachates (Capolupo et al., 2020; Gandara e Silva et al., 2016) and NPs (Auguste et al., 2021; Rist et al., 2019). Except in one study (Rist et al., 2019), MPs did not impact mussel larval development (Capolupo et al., 2018, 2021b) or larval food consumption (Capolupo et al., 2018). While MPs did not greatly impact the mussel early life stages at the individual level, they exhibited several effects at lower-scale levels. MPs may impact mussel cellular defense mechanisms at early life stages, as they induce a change in the multixenobiotic resistance system (Franzellitti et al., 2019). Changes in the expression of genes involved in shell biogenesis, immunomodulation, neuroendocrine signaling, and antioxidant and lysosomal responses were shown after exposure to MPs; see Capolupo et al. (2018) for more details. Other lower-scale effects, as a decrease in lysosomal membrane stability was found when exposed to plastic leachates (Capolupo et al., 2020). These sublethal processes may weaken the early life stages of mussels and potentially compromise their survival and ability to reach the juvenile stage, which coincides with the settlement stage.

As previously discussed, in the context of exposure duration (Section 3.3.2), knowledge of the impact of plastics on mussel reproductive output following parental exposure remains limited. MPs can affect gonads, gametes and sex hormones, i.e. estradiol and testosterone, of adult mussels (Choi et al., 2022; Ciocan et al., 2020; González-Soto et al., 2022). Thus, mussel reproductive success is affected by plastic, which directly affects mussel offspring from gametes to larvae and the adult reproductive system, potentially impairing mussel recruitment on a broader scale.

### 4.2.2. Physiological traits

The impact of plastic on physiological traits was assessed in 35

studies through several parameters, such as filtration, respiration, absorption, excretion, scope for growth, heart rate and shell regeneration. Among these studies, 66 % found an effect in at least one parameter, with the magnitude of the effects being dependent on several factors, such as MP properties (e.g. polymer, size) and exposure mode (e.g. concentration and duration; Supplementary Material, S2). However, the research effort is uneven, as most studies focused on MPs, only a few addressed the potential impact of NPs (2 papers), and none investigated the impact of plastic leachates on these parameters. As such, increased research efforts in these plastic categories would be advisable, as demonstrated in another bivalve, the Pacific oyster *Magallana gigas*, with the impact of plastic leachates on different physiological parameters (Tallec et al., 2022b).

#### 4.2.2.1. Filtration rate.
The impact of plastic on the mussel filtration rate has been extensively studied (29 papers), and almost 50 % of the studies reported an effect (14 papers; Supplementary Material, S3). Specifically, with the exception of one paper (Fraissinet et al., 2023) that reported an increase in mussel filtration rate with increasing MP concentrations, MPs consistently decreased filtration rates in 12 papers, as well as when exposed to a mix of MPs and NPs (one study; Supplementary Material, S3). The remaining papers (n = 15) reported no significant effect on the filtration rate when exposed to MPs and NPs (Supplementary Material, S3). The disparities in the mussel filtration rates are epitomised by differences in experimental protocols, such as the polymer used and their concentrations. For instance, the mussel filtration rate was impacted by PVC but not by PS MPs (Hamm and Lenz, 2021). In turn, the effect of PE on the mussel filtration rate was only significant at high concentrations (Harris et al., 2021). These observations further stress the need to diversify the range of polymers, concentrations, and experimental parameters examined; in addition, the impact of plastic leachates on this parameter should be emphasised, as this aspect has not been studied in marine mussels.

#### 4.2.2.2. Respiration rate.
The respiration rate was studied in 17 papers, and only 7 papers found that plastic affected this parameter (Supplementary Material, S3). Oxygen consumption decreased in four studies after exposure to MPs (Huang et al., 2022; Tang et al., 2022; Rist et al., 2016) or NPs (Wang et al., 2022). In contrast, mussel respiration was found to be enhanced in three other studies after exposure to MPs (Sui et al., 2022, 2023; Wang et al., 2021) and NPs (Wang et al., 2021). Respiration impairment is noticeably dependent on several factors, such as plastic concentration and experiment duration. For instance, in the studies by Sui et al. (2022, 2023), respiration was significantly impacted at the highest plastic concentration tested (1000 MP L⁻¹) after 3 d of exposure, but not at the medium concentration (10 MP L⁻¹), while the opposite was found after 21 d of exposure. The remaining studies (n = 10) did not observe any significant impact on the respiration rate due to MP, NP exposures or a combination of both (Supplementary Material, S3).

#### 4.2.2.3. Absorption efficiency.
The absorption efficiency of food organic content (based on the proportions of organic dry weights in food and in faeces; Conover, 1966) was studied in 10 papers; among these, 6 papers found that plastic impacted this parameter (Supplementary Material, S3). Specifically, absorption efficiency decreased in three studies following exposure to MPs (González-Soto et al., 2022; Huang et al., 2022) and NPs (Wang et al., 2021) but increased in three other studies following exposure to MPs (Sui et al., 2022, Sui et al., 2023) and NPs (González-Soto et al., 2019). Among them, two studies compared the effects of MPs and NPs, and both showed that exposure to NPs led to a disruption in absorption efficiency (González-Soto et al., 2019; Wang et al., 2021), highlighting the significant role of particle size as a crucial determining factor. The remaining studies (n = 4) did not observe any impact of plastic on the absorption efficiency, either after exposure to
MPs or a mixture of NPs and MPs (Supplementary Material, S3).

#### 4.2.2.4. Excretion.
The excretion rate, measured through ammonia concentration in seawater, was assessed in five studies (Supplementary Material, S3). Although one study did not observe any plastic-induced impact on excretion (Opitz et al., 2021), the others showed conflicting effects. Some studies found an increase in ammonia concentration in the presence of MPs (Sui et al., 2022, 2023; Wang et al., 2021) or NPs (Wang et al., 2021), while one study reported a decrease in ammonia concentration following MP exposure (Huang et al., 2022). The effect of plastic on the excretion rate was not constant throughout the experiment and depended on the timing and concentration (e.g. Sui et al., 2022, 2023). An increase in the amount of pseudofaeces was also observed in three studies following an exposure to MPs (Foekema et al., 2022), NPs (Wegner et al., 2012), and a mixture of both (Santana et al., 2018; Supplementary Material, S3).

#### 4.2.2.5. Scope for growth.
The scope for growth (SFG)—the difference between the acquisition of energy by filtration and/or assimilation and the loss of energy by respiration and/or excretion to produce somatic tissue and gametes (Widdows and Johnson, 1988)—is estimated from the physiological rate data mentioned above and provides a proxy of the mussel overall condition and energy budget. However, SFG calculation methods vary among studies, potentially leading to discrepancies in reported values; for instance, SFG was estimated as SFG = Absorption - (Respiration + Excretion) in Wang et al. (2022) but as SFG = (Filtration × Absorption) - Respiration in Fernández et al. (2022). Among the eight papers that estimated SFG, four found that plastic affects this parameter (Supplementary Material, S3). Thus, a decrease in SFG was found following MPs (Huang et al., 2022; Wang et al., 2021) and NP exposures (Wang et al., 2021). In contrast, an increase in SFG was found following NPs (González-Soto et al., 2019) and MPs (Sui et al., 2023) exposures, although this effect was not consistently observed throughout the duration of the experiment. Size, concentration, and exposure duration were the factors that mainly impacted this parameter (González-Soto et al., 2019; Sui et al., 2023). The remaining studies (n = 4) did not observe any impact of MPs on the ability to grow (Supplementary Material, S3). More recently, the Dynamic Energy Budget model (DEB; Kooijman, 2010) has emerged as a powerful tool that surpasses SFG, providing holistic insights into an organism's energy dynamics and life cycle; promising results have been obtained, as shown for M. edulis (e.g. Larsen et al., 2014). Additionally, DEBtox, an extension of DEB, was designed to simulate the effects of toxic substances on an organism's physiology, energy allocation, and life cycle (Jager and Ashauer, 2018; Jager, 2019), making it a valuable tool for studying the environmental impact of plastics and other pollutants on organisms, such as mussels.

#### 4.2.2.6. Others.
Two other studies focused on other physiological parameters, i.e. mussel heart rate and mussel shell regeneration rate under MP exposure (Kolarević et al., 2022; Yu et al., 2022; Supplementary Material, S3). When used in combination with fluoranthene, MPs act on the mussel heart rate, but MPs alone do not induce an effect (Kolarević et al., 2023). A shell study showed that the regeneration rate of mussel shell was significantly inhibited after an exposure to MPs (Yu et al., 2022). Notably, plastic also affects shell formation (the prodissoconch) of the veliger larvae during the larval stage; see Section 'Reproductive success and larval development' above.

### 4.2.3. Behavioural traits

The effect of plastics on mussel behaviour was assessed in 14 studies, and 79 % of studies found at least one effect among the following studied parameters: byssus production, displacement, aggregation, valve movement, and antipredator behaviour (Supplementary Material, S2, S3).

#### 4.2.3.1. Byssus.
The majority (75 %) of studies that assessed the impact of plastic on byssal thread properties found a significant effect (9 out of the 12 articles; Supplementary Material, S3). Byssus (i.e. the proteinic filaments produced by the byssal gland at the base of the foot) production was consistently reduced after an exposure to MPs (Hamm and Lenz, 2021; Hariharan et al., 2021; Huang et al., 2022; Rist et al., 2016; Rodrigues et al., 2022; Webb et al., 2020) or a combination of NPs and MPs (Green et al., 2019). Other byssal properties, such as the adhesion strength, decreased after exposure to plastic particles (Green et al., 2019; Shang et al., 2021a). Only one study showed an increase in byssus production, and interestingly, this was the only one that examined the impact of plastic leachates (Seuront et al., 2021). Byssus properties were not affected by MPs (Barkhau et al., 2022; Yap et al., 2020) or a combination of MPs and NPs (Hamm et al., 2022) in the remaining studies. Notably, the plastic effects were dependent on the mussel species studied (Seuront et al., 2021), the plastic concentration (Hamm and Lenz, 2021) and the polymer used (Green et al., 2019; Hamm and Lenz, 2021), e.g. PE induced an effect in M. edulis byssus properties, but not PLA (Green et al., 2019).

#### 4.2.3.2. Other behavioural changes.
A single study focused on mussel valve movements, revealing that valve opening decreases after exposure to NPs (M. edulis; Wegner et al., 2012). Movement behaviour has been studied more extensively, as three studies showed that mussel movement decreased after exposure to MPs (n = 1) and MP leachates (n = 2). Specifically, Seuront et al. (2021) showed that MPs impact mussel locomotion and aggregation in a species-specific way (Seuront et al., 2021). M. edulis and C. meridionalis shared the same response, i.e. a decrease in locomotion and aggregation in MP leachate seawater, while this parameter remained unaffected in M. galloprovincialis and P. perna. Finally, the mussel antipredator response was also impaired by MP leachates, as they induced a loss in their negative chemotaxis to crab cues (Uguen et al., 2022).

### 4.3. Large-scale impacts and implications for mussel populations, communities and ecosystems

Mussels play a crucial role as ecosystem engineers, and the effect of plastic pollution may extend from the species to the biodiversity and functioning of the ecosystem they support. Especially given the ability of mussel beds to trap MP particles, indeed, through their high filtration ability, their spatial arrangements and shell structure mussels enhance turbulence and plastic retention (Lim et al., 2020; Nicastro et al., 2022; Van Colen et al., 2021). However, only two studies have assessed the influence of plastic on mussel bed-associated communities (Foekema et al., 2022; Green et al., 2017). Interestingly, exposure to high-density polyethylene (HDPE) and PLA NPs and MPs did not result in significant effects on ecosystem functioning or the community assemblages of associated invertebrates (Green et al., 2017). This greatly conflicts with the same experiment conducted on flat oysters (Ostrea edulis), in which plastic exposure caused effects on their associated community assemblage (e.g. benthic cyanobacteria, polychaetes, oligochaetes; Green et al., 2017). Species diversity was also unaffected by exposure to PS MPs, although an increase in the abundance of the benthic community was observed (Foekema et al., 2022).

These investigations were conducted in mesocosms, with study periods ranging from 50 to 56 days. Importantly, mesocosm studies, which offer a more realistic environment than laboratory settings, still oversimplify natural conditions; for example, the effect of wave actions is removed. While MPs alter byssus production (see Section 4.2.3), the potential influence of MPs on mussel dislodgment and the subsequent cascading effects at larger ecological scales may be underestimated when the wave action in this mesocosm is absent. Additionally, no top-down effects were measured in the mesocosm set-ups, oversimplifying a system without predators and potentially causing the effects of plastics
on the mussel bed ecosystem to be underestimated. Finally, the impairment of mussel reproduction and early life stages may act on mussel recruitment and underscore the potential far-reaching consequences of plastic pollution on the population dynamics and resilience of these marine organisms and their ecosystems. Mesocosm studies are effective tools for incorporating multiple stressors and increasing the similarity of the results to those obtained under environmental conditions. Considering the changing climate that is increasingly disrupting ecosystem structures and functions, multistress studies are very important in understanding the toxicity of plastic pollution on these keystone species in the environment.

# 5. Multistress experiment: The case of global warming

In natural environments, organisms face multiple, simultaneous stressors, conditions that diverge from controlled laboratory settings, where toxicological studies are typically conducted under optimal conditions (Holmstrup et al., 2010). However, these laboratory tests may overlook the potential risks posed by toxicants, as synergistic interactions are the most prevalent between environmental stressors and toxicants (Holmstrup et al., 2010). Thus, for a realistic understanding of the ecological consequences of plastic pollution, it is essential to consider multiple environmental factors, particularly those related to the Anthropocene (e.g. heat stress, acidification, hypoxia) in toxicological studies.

Among the range of Anthropocene-driven stressors, we focused on temperature due to its role as a major driver of climate change and the substantial impacts it may have on marine ecosystems (IPCC, 2022). Temperature, may influence the toxicity of a pollutant by directly affecting the nature of the pollutant, its bioavailability, and/or the interplay of an *a priori* toxicity of a chemical compound and its actual effect, which depends on the target organisms and the thermal conditions; thus, its toxic effects either increase or decrease (Mohammed, 2013). Although plastic pollution and climate change are fundamentally linked (Ford et al., 2022; Sharma et al., 2023; Shen et al., 2020), their combined effects have received limited attention; see also, e.g. Nicastro et al. (2023b) for an illustration of the research gaps in the literature that has focused on climate change and plastic pollution. However, the potential interaction between these stressors fundamentally introduces an additional level of complexity and irregularity in how organisms and ecosystems respond. In this context, it is necessary to consider the effect of temperature in plastics research to improve our knowledge of the complex interactions that may arise between these two major global change stressors.

## 5.1. Temperature: A research gap in plastic studies

### 5.1.1. Heat stress in intertidal mussel research

Plastic pollution does not occur alone but fundamentally interacts with various other environmental stressors. Among them, global warming stands out as a critical factor that affects various biological systems. The rapid increase in greenhouse gases, which resulted from human activities since the Industrial Revolution, has led to a dramatic rise in CO₂ levels, triggering an increase in global temperature. This rapid change in climate also led to an increase in the frequency, severity and duration of extreme weather events, such as terrestrial and marine heatwaves (IPCC, 2022). For example, the frequency of marine heat waves doubled between 1982 and 2016 and is expected to increase by a factor of 16 following a global warming of 1.5 °C compared to preindustrial levels (Frölicher et al., 2018).

Mussels from intertidal coastal systems inhabit one of the most challenging and stressful habitats on the planet (Harley, 2008), as mussels must address challenging fluctuating conditions caused by immersion and emersion cycles (e.g. temperature, salinity, oxygen and food availability). Mussels, as intertidal molluscs, experience body temperature close to that of the substrate during emersion (Seuront et al., 2018).

The substrate can exhibit strong spatial and temporal variability in temperature (Seuront et al., 2018) and reach temperature exceeding 50 °C during low tide in tropical (Liu and Morton, 1994; Seuront and Ng, 2016) but also in polar regions (Sokolova et al., 2000). Thus, organisms may face these extreme conditions, e.g. the use of 'robomussels' have shown that mussel body temperature can reach up to 41.7 °C in temperate region during the summer (eastern English Channel; Seuront et al., 2019). While the temperature experienced by the mussel at emersion can strongly fluctuate within a single tidal cycle, changes at immersion occur over longer time scales. However, both temperatures may occasionally exceed the tolerance range of mussels and cause mass mortality of mussels.

### 5.1.2. Emersion: An overlooked factor in temperature studies

The tidal cycle plays a role in the context of climate change studies that focus on intertidal species and ecosystems. However, as recently stressed for the fraction of plastic research focusing on plastic leachates (Delaeter et al., 2022; Seuront et al., 2022), emersion represents only a tiny fraction of climate change research (Fig. 8). To illustrate this, a Web of Science search (accessed on October 28, 2023) returned 2745 articles using a search in Topics of the following keywords in 3 distinctive search categories: (i) "effect*" OR "impact*" OR "affect*" (ii) AND "mussel*" (iii) AND "temperature*" OR "heat*". However, when a row containing the keywords: "aerial*" OR "emersion*" OR "air*" was added, only 235 articles emerged (Fig. 8). Emersion thus represents <10 % of the research effort on the impacts of climate warming on intertidal mussels. Furthermore, despite the ever-growing research efforts devoted to clarifying the global impacts of temperature on mussels, research into the specific effects of emersion shows a more erratic trend (Fig. 8).

### 5.1.3. Temperature as an overlooked stressor in plastic studies

Despite the significance and growing interest in plastic pollution and climate change, a notable research gap persists regarding their combined impact. Thus, among the marine mussel studies that examined the effects of plastic and temperature, only one paper addressed their combined effect (Pittura et al., 2022; Fig. 8B). This study revealed the complexity of plastic and temperature interactions. It estimated a higher molecular and cellular impact with plastic following marine heat stress for one polymer (PES) and the opposite for another (PLA; Pittura et al., 2022). Thus, plastic and temperature interactions could yield unpredictable outcomes for organisms and ecosystems. As these two stressors potentially amplify each other's impact, expanding the research is crucial for obtaining comprehensive knowledge on the multifaceted interplay between these stressors and their potential implications. For instance, there is a gap in studies that explore the combined effects of NPs or plastic leachates with temperature, as well as investigations into the effects of thermal stress during emersion. Moreover, an encompassing approach to research involving a wider variety of polymers, mussel species, endpoints, and temperatures is essential to fathom the effects of plastic in our changing world.

## 5.2. Plastic pollution in a warming world: A roadmap for future research

### 5.2.1. Temperature: A significant stressor

Mussels are ectothermic poikilothermic organisms, i.e. their temperature varies with the environmental temperature; thus, temperature is among the most important stress factors (Guschina and Harwood, 2006). Although mussel body temperature is close to that of the surrounding water, their body temperature can be significantly warmer than the surrounding air and substrate temperatures at low tide (10 to 17 °C and 5 to 14 °C warmer, respectively), as recorded by M. edulis biomimetic loggers (maximum biomimetic mussel body temperature = 41.7 °C; Seuront et al., 2019). In addition, mussels have a limited capacity of movement and, depending on the season, live close to the upper limit of their thermal tolerance window; therefore, mussels are particularly susceptible to extreme weather conditions (Mislan and
<figure><img src="image_15.png" /><figcaption>**Fig. 8.** (A) Temporal trends in the total number of publications on the impact of plastic (blue), temperature (light orange) and specifically emersion temperature on mussels (dark orange) between 1991 (first study) and mid-2023. (B) Distribution of the research effort, represented by the number of studies, in two marine mussel research fields, i.e. plastic impacts (blue) and temperature impacts (orange). The intersection area shows the studies that examined the combined effect of both factors, while the distinct sections highlight the lack of combined research between these research fields. Details of how the research fields and the number of articles were defined are described in the review. Specifically, for the plastic impact on mussels, 368 articles were initially returned from Web of Science (accessed May 25, 2023) with the following keywords in 3 distinctive rows: (i) "microplastic*" OR "nanoplastics*" (ii) AND "effect*" OR "impact*" OR "affect*" (iii) AND "mussel*", of which 106 articles were sorted (dark blue; Fig. 8A, B), 6 of which focused on leachates (light blue; Fig. 8B) and only 1 on the combined effect of temperature and plastics (red number; Fig. 8B). For the temperature impact on mussels (light orange; Fig. 8A, B) the keywords used in Web of Science (accessed October 28, 2023) were organised into three distinct rows: (i) "effect*" OR "impact*" OR "affect*" (ii) AND "mussel*" (iii) AND "temperature*" OR "heat*"). For the emersion temperature search (dark orange; Fig. 8A, B) the additional keywords: "aerial*" OR "emersion*" OR "air*" were added in an additional row. The numbers of articles in the temperature and emersion impact fields are noted as an approximate count (≈) in Fig. 8B, as they did not undergo additional sorting when returned from Web of Science.</figcaption></figure>

Wethey, 2015; Somero, 2002). Thermal stress can be buffered by behavioural responses, such as gaping behaviour or interaction with other individuals through aggregation (Lathlean et al., 2016; Nicastro et al., 2012). As the temperature is directly linked to metabolism, if the behavioural responses do not sufficiently limit the thermal stress,

exposure to extreme heat events may induce subcellular damage and alterations in functional traits, such as growth, reproduction and survival; see Zippay and Helmuth (2012) for a review. In addition to this direct effect, temperature affects mussels indirectly through altering the feeding performances of their predators, e.g. sea stars and dogwhelks
(Contolini et al., 2020; Pincebourde et al., 2008). Similar to plastic pollution, temperature affects mussels across all levels of organisation; as mussels are ecosystem engineers, temperature may affect the thermal, humid and protective microhabitats formed that other species depend upon (Jurgens et al., 2022).

Mussel sensitivity to thermal stress is dependent on several factors. First, sensitivity to thermal stress is species-dependent, e.g. *M. galloprovincialis* exhibits a higher tolerance temperature than that of *M. edulis* (Thomas and Bacher, 2018). This adaptation to thermal stress is transferred from female mussels to their offspring. For instance, *M. galloprovincialis*-*M. edulis* hybrid larvae from *M. galloprovincialis* females exhibit lower sensitivity to heat stress than those from *M. edulis* females (Mlouka et al., 2020). In addition to interspecific variability, sensitivity to thermal stress is influenced by intraspecific diversity and by the type of temperature exposure (i.e. temperature immersion or emersion), as shown recently with *P. perna* mussels (Nicastro et al., 2023b). Heat stress exposure conditions (e.g. duration, acclimation) are also an important parameter to consider. For instance, exposure to marine heatwaves may affect mussel survival (e.g. Venter et al., 2023; although Ducker et al., 2023, did not find any effect), a response modulated by the temperature to which they have been acclimatised (e. g. Brownlee, 2022). Although studies have demonstrated that mussels can adapt to aerial thermal stress (Moyen et al., 2020; Schneider, 2008), others found that multiple exposures to high aerial temperature greatly decreased the thermal tolerance of mussels (Jones et al., 2009; Seuront et al., 2019; Sorte et al., 2018). The mussel response to aerial heat stress is also typically modulated by emersion time (Seuront et al., 2019), exposure month (Jones et al., 2009; Sorte et al., 2018), relative humidity (Delorme et al., 2021), aggregation and gaping behaviour (Nicastro et al., 2012).

### 5.2.2. Temperature may increase the impact of plastics on mussels

Stressor interactions can fundamentally be additive, antagonistic or synergistic (Crain et al., 2008). Warming and contaminant stressors—mainly metals and pesticides, which are much more studied than plastics—showed synergistic effects in most studies; see Dinh et al. (2022) for a review. Specifically, in marine mussels, many studies have observed synergistic effects between temperature and contaminants (Andrade et al., 2022; Bordalo et al., 2023; Leite et al., 2020; Lopes et al., 2022; Moleiro et al., 2022), including plastic pollution (Pittura et al., 2022), although our literature review shows that this area of research remains in its infancy. Thus, each stressor could alter an organism's resistance to the other stressor (Sokolova and Lannig, 2008); for instance, compared to mussels from 'pristine' environments, mussels from polluted environments showed a significantly lower thermal tolerance (Hellou and Law, 2003; Smaal et al., 1991; Thomas et al., 1999).

As previously discussed, mussels have shown species-specific responses to plastic pollution (e.g. Seuront et al., 2021) and to thermal stress (e.g. Mlouka et al., 2020), which highlights the need to gather more information on the combined impacts of these factors on each species individually. Some species may be more vulnerable than others to the combined exposure of climate-related stressors and toxic substances, this at the individual, population and or community levels, which can lead to large-scale shifts (Moe et al., 2013). Such shifts could have broader implications, such as undermining mussel beds' ecosystems, their biodiversity and, their ability to protect coastlines from erosion (Borsje et al., 2011), to ensure water quality (Newell, 2004) and nutrients cycle (Lindahl et al., 2005). Beyond these ecological impacts, human food security and economic stability can also be disrupted, as mussels offer numerous nutritional benefits, representing a more environmentally sustainable animal protein source (Suplicy, 2020) as well as a significant part of the economy in some countries (e.g. China, Chile, and Spain), that collectively produce over 70 % of the world's mussels (FAO, 2023). Understanding how these stressors interact across different species will enhance our ability to predict and mitigate ecological and

economic impacts. Additionally, although these two stressors can directly act on mussels, the stressors can also impact mussel food sources, predators or competitors and indirectly affect mussel populations and their ecosystem. Thus, these stressor interactions emphasise the complexity of potential consequences for organisms and ecosystems. Understanding these implications remains essential for effectively evaluating chemical risks and improving our ability to protect ecosystems.

### 5.2.3. Temperature may impact plastic physical and chemical properties

Temperature may also directly influence the physical, chemical, and mechanical properties of any plastic item. Elevated temperatures can accelerate the degradation of plastics through abiotic (e.g. thermal degradation and oxidation processes; Chowdhury et al., 2022; McKeen, 2014; Zhang et al., 2021) or biotic processes (e.g. increasing enzymatic activity; Ge et al., 2023). Increasing temperature can induce the removal of a hydrogen atom from the polymer, generating an unstable and reactive radical that may lead to material weakness and the acceleration of its breakdown (McKeen, 2014). The breakdown of polymer chains induced by these processes may impact the structural integrity of plastics and result in the formation of smaller and coarser particles (Villamizar and Morillas, 2018). Increasing temperatures strongly promote the release of MPs (Jiang et al., 2022) and NPs although this depends on the polymer type, with a higher release observed in PS, HDPE, and LDPE (low-density polyethylene) than in PP (Hernandez et al., 2023).

High temperatures can also influence the interactions between plastic particles and other environmental components, potentially enhancing the adsorption and leaching of chemicals; as a result, the transport and fate of plastic-associated chemicals in ecosystems are altered. At high temperatures, molecular mobility in plastics increases enhancing polymer chains to open and the leaching of diverse chemicals (Avio et al., 2015; Gunaalan et al., 2020; Li et al., 2024). Several studies have shown an increase in the leaching of plastic additives (e.g. phthalates ester plasticisers and flame retardants; Dhavamani et al., 2022; Kida et al., 2022; Kida and Koszelnik, 2021; Li and Tang, 2023; Ye et al., 2020; Zhang et al., 2022) and other environmental chemicals adsorbed on plastic particles (e.g. polycyclic aromatic hydrocarbons (PAHs), metals and antibiotics; Godoy et al., 2020; Kida and Koszelnik, 2021; Peng et al., 2023) with increasing temperature. For instance, a 15 °C temperature increase raises the metal leaching rate from plastics by 1.6 to 8.5 times, depending on the polymer type and metal studied (Peng et al., 2023). In turn, the adsorption of chemical compounds (e.g. perfluorinated compounds, pharmaceutical drugs, plasticisers, flame retardants, oils and heavy metals) by plastic is enhanced by temperature up to a certain optimal temperature (Guan et al., 2022; Khumalo et al., 2022; Liu et al., 2018, 2022; Oz, 2019; Peng et al., 2023; Tang et al., 2021; Xu et al., 2019; Yap and Tan, 2021). High temperatures modify polymers at the structural level, increasing polymer segment mobility and reducing crystallinity, which in turn increases the number of active surface sites, enhancing chemical sorption by plastics (Agboola and Benson, 2021; da Costa et al., 2024). The adsorption and desorption optimal temperatures are dependent on the chemical compound (Khumalo et al., 2022; Kida et al., 2022) and the plastic polymer (Du et al., 2022; Zhang et al., 2022).

Note that while certain chemical compounds might migrate faster at elevated temperatures, high temperatures may also accelerate the degradation of the compound, potentially impacting its bioavailability (Kang and Kondo, 2005; Manzano et al., 1999; Saeed et al., 2011; Yang et al., 2018). However, elevated temperatures might also modify the transformation of contaminants into more toxic metabolites; for example, although the half-life of polychlorinated biphenyls (PCBs) in rainbow trout decreases with increasing water temperature, this compound was biotransformed into toxicologically more active hydroxylated PCBs with increasing temperature (Buckman et al., 2007). Overall, in efforts to understand and mitigate the environmental consequences of
plastic pollution, the impact of temperature on plastic materials, a complex and multifaceted issue, should be carefully considered. Thus, future knowledge on the role that plastics play as direct or indirect carriers of molecules with different levels of toxicity, and their influence on organisms and ecosystems, partially depends on our efforts to investigate how temperature affects the adsorption, desorption, and degradation dynamics of various contaminants in or on plastic particles and within the surrounding environment.

## 6. Conclusion

This review highlights the profound and widespread impact of plastic pollution on marine mussels, as keystone engineer species. In addition, the urgency of more holistic research that considers the challenges posed by multiple stressors, particularly in an era of global warming, is emphasised. While plastic pollution affects mussels at multiple levels, strong disparities in research efforts persist; thus, our efforts to clarify the role of plastic in the biology and ecology of mussels, as well as marine invertebrates in general, remain in their infancy. The asymmetric concentration of research effort in specific countries, mussel species, life stages and specific plastic types (e.g. MP size, PS polymer and sphere shape) raises concerns about the possibility of generalising the findings. Awareness of environmental consistency in the choice of experimental protocols should be increased. In addition, to address regional differences in research, we advocate for more coordinated interdisciplinary and international collaborations along with the standardization and harmonization of methodologies (e.g. in plastic chemical compound analyses) to facilitate data comparability across studies. Furthermore, it is essential to recognise that plastic pollution interacts with numerous stressors in the Anthropocene era, resulting in intricate and often unpredictable effects. However, multistressor studies are still in their infancy within the literature addressing the biological effects of plastic on mussels. To address this issue, we must assess plastic pollution in conjunction with temperature, as discussed here, and extend the framework to other stressors, such as acidification, hypoxia, food availability, and salinity. Implementing the above recommendations is imperative for fostering comprehensive knowledge of the ecological consequences of plastic pollution and promoting effective conservation strategies to ensure the health and sustainability of our oceans in our changing world.

### CRediT authorship contribution statement

**Marine Uguen:** Writing – review & editing, Writing – original draft, Visualization, Methodology, Investigation, Formal analysis, Conceptualization. **Sylvie M. Gaudron:** Writing – review & editing, Supervision, Funding acquisition. **Laurent Seuront:** Writing – review & editing, Supervision, Funding acquisition.

### Data statement

This review is based on an analysis of 106 articles. All data used for the analysis are available in Supplementary Materials. The full list of the 106 articles includes: Abidli et al., 2021; Alnajar et al., 2021; Auguste et al., 2020, 2021; Avio et al., 2015; Bandini et al., 2021; Barkhau et al., 2022; Brandts et al., 2018; Bråte et al., 2018a, 2018b; Browne et al., 2008; Canesi et al., 2015, 2016; Capolupo et al., 2018, 2020, 2021a, 2021b, 2021c; Cappello et al., 2021; Chelomin et al., 2022; Choi et al., 2021, 2022; Christoforou et al., 2020; Ciocan et al., 2020; Cole et al., 2020; Détrée and Gallardo-Escárate, 2017; Détrée and Gallardo-Escárate, 2018; Dovzhenko et al., 2022; Ferguson et al., 2022; Fernández et al., 2022; Foekema et al., 2022; Fraissinet et al., 2023; Franzellitti et al., 2019; Gonçalves et al., 2019, Gonçalves et al., 2022; González-Soto et al., 2019, 2022; Green et al., 2017, 2019; Gu et al., 2020; Hamm and Lenz, 2021; Hamm et al., 2022; Han et al., 2021; Hariharan et al., 2021; Harris and Carrington, 2020; Harris et al., 2021; Huang et al.,

2021, 2022; Jong et al., 2022; Jaouani et al., 2023; Joyce and Falkenberg, 2022; Katsumiti et al., 2021; Khalid et al., 2021; Kolarević et al., 2023; Laubscher et al., 2023; Leroux et al., 2022; Li et al., 2020a, 2020b; Lu et al., 2023; Magara et al., 2018, 2019; Masiá et al., 2021; Nunes et al., 2020; O'Brien et al., 2021; Opitz et al., 2021; Paul-Pont et al., 2016; Pavičić-Hamer et al., 2022; Phothakwanpracha et al., 2021; Pitura et al., 2018, 2022; Provenza et al., 2020, 2022, 2023; Qi et al., 2023; Revel et al., 2019; Rist et al., 2016, 2019; Rodrigues et al., 2022; Roman et al., 2023; Romdhani et al., 2022; Santana et al., 2018; Sendra et al., 2020a, 2020b; Seuront et al., 2021; Shang et al., 2021a, 2021b; Gandara e Silva et al., 2016; Sui et al., 2022, 2023; Sun et al., 2023; Tang et al., 2022; Trestrail et al., 2021; Uguen et al., 2022; Van Cauwenberghe et al., 2015; von Hellfeld et al., 2022; von Moos et al., 2012; Wang et al., 2020, 2021, 2022, 2023a, 2023b; Webb et al., 2020; Wegner et al., 2012; Wei et al., 2021; Yang et al., 2021; Yap et al., 2020; Yu et al., 2022.

### Funding

M. Uguen was funded by a PhD fellowship from the University of Lille and the Region-Hauts-de-France. This research was also financially supported by both the European Maritime and Fisheries Fund (FEAMP) and France Filière Pêche through the research project SOLACE (« Resilience of the mussel *Mytilus edulis* to anthropogenic and climatic stressors ») and by the French National Program EC2CO. This paper was edited professionally by highly qualified native English speaker.

### 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.

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2024.178078.

### Data availability

Data are available in Supplementary Materials.

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