The stealthy journey of nanoplastics in bivalves: accumulation dynamics and toxic burden

Abstract

The strong filter-feeding capacity of bivalves makes them more prone to accumulating nanoplastic particles from their environment, posing a threat to aquaculture and food safety. Despite their inconspicuous size, nanoplastics embark on a stealthy invasion through bivalve tissues, evading conventional detection. Reliable detection methods for nanoplastics are essential for risk assessment. This paper provides a comprehensive review of nanoplastic detection techniques in biological tissues and suggests improvements in in situ detection and AI-based recognition methods. These advancements are critical to unveiling the hidden pathways of nanoplastics in biological systems. Next, we summarize the endocytic mechanisms and bioaccumulation patterns of nanoplastics based on particle size classification and realistic environmental scenarios, identifying gills and hepatopancreas as primary accumulation targets. This dynamic accumulation process highlights how nanoplastics progressively infiltrate key organs, escalating their toxic burden. Additionally, this paper offers a thorough overview of the pathways through which nanoplastics breach biological barriers and trigger cascading reactions, from cellular stress to organelle dysfunction, tissue damage, and ultimately organismal consequences. These cascading effects underscore the insidious yet pervasive toxic burden imposed by nanoplastics. Finally, this review identifies key research gaps, including the synergistic or inhibitory effects of coexisting marine pollutants on bivalve bioaccumulation and the unclear pathways and efficiency of nanoplastic accumulation in filter-feeding bivalves under eco-corona regulation. Unraveling these uncertainties is vital to mapping the full journey of nanoplastics and mitigating their ecological toll. This review aims to enhance the understanding of nanoplastic–bivalve interactions and guide mitigation strategies for their ecological effects.

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Yuntian Shi

Yuntian Shi is pursuing his master's degree in Aquaculture at Shanghai Ocean University. His research focuses on environmental physiology and toxicology, with particular emphasis on the effects of nanoparticles, climate change, and other stressors on marine organisms such as horseshoe crabs and mussels. He has published multiple papers in high-impact journals including Marine Environmental Research and Aquaculture, and has received honors such as the National Scholarship for Postgraduate Students. His technical skills include Scanning Electron Microscopy (SEM), Confocal Laser Scanning Microscopy (CLSM), and genomic DNA extraction.

Kangping Jiao

Kangping Jiao is pursuing his master's degree in the College of Fisheries and Life Science at Shanghai Ocean University under the supervision of Professor Menghong Hu. His research focuses on the aquatic ecotoxicology of environmental pollutants, with particular emphasis on the toxic effects and mechanisms of nanoplastics on aquatic organisms.

Youji Wang

Youji Wang is a Professor and doctoral supervisor at Shanghai Ocean University, specializing in marine ecophysiology and aquaculture. He is a recipient of the National Ten Thousand Talents Program—Young Top-Notch Talent award and serves as an editor for several SCI journals including Marine Environmental Research and Aquaculture. His research integrates environmental stress responses in shellfish, particularly mussels, under conditions such as ocean acidification, hypoxia, microplastics, and nanoparticles. He has published over 230 SCI papers, with significant contributions to understanding the physiological and ecological adaptations of bivalves to multifactorial environmental changes.

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Environmental significance

Nanoplastics (NPs) are increasingly threatening aquatic ecosystems and food safety, primarily through their uptake by filter-feeding bivalves. This critical review examines the “stealthy” bioaccumulation pathways and toxicological impacts of NPs in bivalves, revealing their potential to trigger oxidative stress, immune dysfunction, and metabolic disturbances. Special emphasis is placed on the role of eco-corona formation in influencing NP behavior and toxicity under environmentally relevant conditions. The study also presents advanced detection methods and proposes standardized frameworks for risk assessment. Our findings underscore the urgent need for effective mitigation strategies to safeguard marine ecosystems and ensure seafood safety, thereby supporting informed environmental policies and sustainable aquaculture practices.

1. Introduction

The current century has been aptly termed the “Plastic Age” due to plastics' profound influence on enhancing human lifestyles. However, as the plastic industry has expanded exponentially, the environmental consequences of plastic waste have become increasingly apparent.¹ The unsustainable growth of global plastic production has led to elevated levels of micro- and nanoplastic pollution across all ecological compartments. It's estimated that between 22% and 43% of plastic garbage is thought to be poorly handled, which leads to its buildup in landfills and eventual river transportation to the oceans.² Borrelle et al.³ estimated that between 19 and 23 million tonnes of plastic garbage entered aquatic environments; by 2030, this amount is expected to increase to 53 million tonnes per year.

This expanding plastic waste, particularly in aquatic environments, leads to its degradation through exposure to ultraviolet radiation, abrasion, oxidation, ocean waves, and microbial activity. These processes break down plastics into smaller particles, microplastics, and nanoplastics via diverse chemical, physical, and biological mechanisms. For example, a single 50 μm (50 000 nm) particle can fragment into approximately 1.25 × 10⁵ particles of 1000 nm or 1.0 × 10⁹ particles of 50 nm.⁴ An illustrative example of this is that the breakdown of a single-use polystyrene coffee lid, after 56 days of UV irradiation, can release up to 2.52 × 10⁹ nanoplastics into the surrounding aqueous environment.⁵ Additionally, rotifers have been observed to produce more than 7.14 × 10⁵ submicrometre particles per individual per day from photo-aged microplastics.⁶ Using nanoparticle tracking analysis, the concentration of nanoplastics in collected samples has been estimated at approximately 10⁸ particles per milliliter.⁷

Despite the growing evidence of nanoplastic pollution,^1,2 research on the physiological effects and bioaccumulation of nanoplastics remains limited, primarily due to challenges in detection methods. Optical microscopy fails below 200 nm, while SEM/TEM faces resolution limits and artifacts for particles under 20 nm. Fluorescence methods suffer from signal instability and interference, and Raman spectroscopy is hindered by environmental contaminants. Though PY-GC-MS enables quantification, it requires destructive sampling and misses real-time dynamics. No single technique provides complete morphological, chemical and dynamic data, forcing reliance on complementary methods that increase complexity. Environmental matrices further complicate detection through signal masking and particle aggregation. While emerging technologies like AI-enhanced spectroscopy show promise, current limitations in sensitivity and real-time analysis impede comprehensive risk assessment and monitoring.

One of the most critical attributes of nanoplastics is their exceptionally high specific surface area and intrinsic surface reactivity, which drive the formation of heteroaggregates with natural organic matter (NOM) and dissolved organic matter—collectively termed the “eco-corona”.⁸ These interactions involve a range of substances, including humic acids^9–12 and co-present environmental pollutants.^13–15 High-affinity organic components can form a dense, persistent “hard corona” directly on the nanoparticle surface, while lower-affinity molecules may constitute a more dynamic “soft corona” outer layer.¹⁵ This layered structuring not only promotes the creation of larger and more complex aggregates but also profoundly alters the environmental fate of nanoplastics, influencing colloidal stability, transport potential, and bioavailability.

The term “nano” signifies not merely a size range (typically 1–1000 nm), but more importantly, a regime where quantum and surface effects dominate over bulk material properties. Under controlled laboratory conditions, nanoplastics are often studied as monodisperse, well-characterized particles with specific surface chemistries (e.g., amine- or carboxyl-modified). However, in real environmental systems, they are highly heterogeneous in polymer type (e.g., polyethylene, polystyrene, PVC), shape (fragments, fibers, spheres), and surface morphology (porosity, roughness). These characteristics are further transformed by aging processes—such as photo-oxidation, mechanical abrasion, and biofilm formation—which can increase surface charge heterogeneity, introduce oxygen-containing functional groups, and enhance sorption capacity for pollutants. A key consequence of eco-corona formation is the conferral of additional negative surface charges onto nanoplastics via adsorbed anionic macromolecules (e.g., humic acids).¹⁶ When such molecules fully cover the particle surface, they impart colloidal stability through electrostatic repulsion and steric hindrance. The latter arises from entropy reduction and osmotic effects when corona layers overlap during particle approach.^16–18 For instance, Cai et al.¹⁹ demonstrated that humic acid supplementation prevented the aggregation of 100 nm polystyrene nanoparticles even in high-ionic-strength NaCl and CaCl₂ solutions. These complex interactions enable nanoplastics to remain dispersed and “stealthy” in aquatic systems, evading natural removal processes such as sedimentation and heteroaggregation. This prolonged mobility enhances their potential for long-range transport and bioaccumulation via non-classical pathways. Moreover, the eco-corona modifies the biological identity of nanoplastics, influencing their interactions with organisms and leading to unexpected tissue distribution and toxicological outcomes. Thus, the environmental and health risks associated with nanoplastics are not solely inherent to the plastic particles themselves, but are co-determined by their evolving surface chemistry and biogeochemical context.

Bivalves play a vital role in ecosystems through their filter-feeding activities that help purify water bodies and maintain the health of aquatic ecosystems.^20,21 They also serve as key bioindicators of plastic pollution and important seafood sources for humans, and exhibit distinct species-specific, habitat-dependent, and size-related patterns of microplastic accumulation.²² Filter-feeding species like scallops (Patinopecten yessoensis)²⁷ and cockles (Scapharca subcrenata)²³ demonstrate particularly high accumulation rates, especially in polluted urban coastal waters such as China's Yangtze River estuary, where concentrations can reach 40–55 particles per individual. The microplastics predominantly accumulate in gills and digestive glands, with fibers representing over 80% of accumulated particles,^24–26 a proportion significantly higher than in surrounding waters, suggesting selective retention of these problematic shapes. This bioaccumulation poses serious socioeconomic and health risks. First, the predominance of polyester (PET), polyethylene (PE) and polypropylene (PP) – polymers associated with food packaging and textiles – raises concerns about seafood safety, as these plastics may release additives and adsorb environmental pollutants. Second, farmed bivalves often contain 40–250% higher microplastic loads than wild counterparts due to plastic aquaculture infrastructure, directly impacting seafood production systems.^27,28 Most alarmingly, smaller, commercially valuable bivalves show particularly high microplastic concentrations, with studies documenting 1.2- to 8.3-fold greater accumulation in younger individuals,²⁹ suggesting potential impacts on fisheries productivity and food security. While the full health implications require further research, the demonstrated accumulation pathways through digestive and circulatory systems indicate potential for trophic transfer and human exposure through seafood consumption. These findings underscore the urgent need for better pollution control measures in both marine environments and aquaculture practices to safeguard ecosystem and human health.

Previous reviews have underscored the critical need to investigate the bioaccumulation dynamics of plastic particles, particularly nanoplastics, within bivalve species, as well as their potential for interactions with co-occurring toxic substances.^30–32 Furthermore, a significant limitation identified in prior research is the frequent neglect of nanoplastics' complex environmental behavior. Consequently, many laboratory-based studies exhibit limited environmental realism, failing to adequately represent real-world exposure scenarios and their implications.³² To address these key knowledge gaps, the present review specifically focuses on the bioaccumulation pathways and tissue distribution of nanoplastics in bivalves. We emphasize that understanding the unique physicochemical characteristics of nanoplastics – such as their environmental behavior, surface charge, and potential for aggregation or functionalization – is fundamental to elucidating their specific toxicological mechanisms and impacts on these ecologically and economically vital organisms. This focus aims to provide a more environmentally relevant and mechanistic understanding of nanoplastic risks to bivalves.

2. Overview of nanoplastic detection techniques

The identification of nanoplastics begins with visual characterization, which forms the foundation for subsequent analytical techniques. For particles larger than 500 nm, optical microscopy is effective; however, its utility diminishes for plastics smaller than 200 nm due to diffraction limits.³³ To overcome these limitations, Scanning Electron Microscopy (SEM) is frequently employed for high-resolution imaging of nanoplastics.³⁴ SEM works by directing an electron beam onto the sample surface and detecting signals from interactions between electrons and the particles, such as secondary and backscattered electrons.

Despite its advantages, SEM has notable limitations including the charging effect where non-conductive nanoplastics accumulate charges under electron irradiation which can degrade image quality or introduce artifacts, especially for smaller particles.³⁵ For instance, Skawina et al.³⁶ observed that while 20 μm polystyrene spheres were clearly visible with SEM, 20 nm plastics appeared blurred. Moreover, SEM provides primarily morphological information, requiring complementary techniques such as energy-dispersive X-ray spectroscopy (EDS) for chemical characterization.³⁷ Additionally, the use of conductive coatings during SEM sample preparation can obscure surface characteristics, necessitating alternative methods like Transmission Electron Microscopy (TEM) or Atomic Force Microscopy (AFM) for internal structural analysis.^38,39

Building on SEM and TEM's imaging capabilities, fluorescence-based methods offer an alternative for tracking nanoplastics within biological systems.^40,41 Fluorescent labeling, achieved by incorporating dyes or molecules into nanoplastics, allows researchers to visualize their bioaccumulation and distribution.^42,43 Techniques such as stereoscopic fluorescence microscopy or laser confocal microscopy can then detect these fluorescent signals.⁴⁴ However, fluorescence methods face challenges, including quenching effects caused by environmental factors like pH and temperature, as well as interference from other fluorescent substances.⁴⁵ These limitations complicate differentiation between target nanoplastics and background noise. Moreover, prolonged exposure to light can lead to fluorescence bleaching, reducing signal intensity during real-time monitoring.⁴⁶

While fluorescence methods primarily provide optical information, vibrational spectroscopy techniques, such as Raman microscopy, enable compositional analysis of nanoplastics.⁴⁷ Raman spectroscopy detects molecular vibrations by analyzing spectra generated from light scattering, making it effective for identifying chemical structures.⁴⁸ However, its precision can be affected by the presence of dyes, microorganisms, or other materials that obscure signals.⁴⁹ Surface-Enhanced Raman Scattering (SERS) addresses this limitation by significantly amplifying Raman signals through “hotspots” created at specific nanostructures on metallic substrates.^49,50 For example, metallic nanoparticles can enhance signal sensitivity, enabling the detection of nanoplastics at extremely low concentrations.⁵¹

SERS has proven particularly effective when combined with complementary techniques. For instance, combining SERS with Thermal Desorption-Gas Chromatography-Mass Spectrometry (TD-GC-MS) provides a powerful approach for nanoplastic detection.⁵⁰ TD-GC-MS allows for the non-destructive analysis of nanoplastic samples, followed by comprehensive quantification through mass spectrometry.⁵² Pyrolysis-GC-MS (PY-GC-MS), another method that has high performance in the quantitative analysis, breaks down samples at high temperatures and examines the breakdown products. PY-GC-MS is capable of detecting nanoplastics as small as 50 micrograms in mass, making it a preferred method for precise quantification.^43,53

Finally, while these methods provide valuable insights into nanoplastic detection, their effectiveness can be influenced by environmental factors. For example, nanoplastics often form heteroaggregates with NOM, which may hinder access to SERS hotspots, complicating detection.⁵⁴ Additionally, laser excitation during SERS analysis can induce localized plasmonic heating,^54,55 potentially altering the physical state of nanoplastics and affecting results. Combining these techniques with other methods, such as TD-GC-MS, provides a more comprehensive framework for detecting and analyzing nanoplastics. The field of nanoplastic detection is advancing through cutting-edge technologies such as machine learning-based image analysis⁵⁶ (e.g., convolutional neural networks) and advanced spectroscopic methods (e.g., surface-enhanced Raman spectroscopy and transfer learning-enhanced NIR). These innovations enable rapid classification and trace-level identification of nanoplastics in complex samples. Future research will focus on multimodal detection platforms integrating microfluidics, portable mass spectrometry, and AI algorithms to enhance in situ accuracy and throughput.⁵⁷ Additionally, specialized sensors and adaptive signal processing for high-interference environments (e.g., biological matrices, high-salinity media) are critical to achieving comprehensive nanoplastic monitoring, supporting pollution control and health risk assessment.⁵⁸Fig. 1 reveals the advantages and constraints of NP detection methodologies.

Fig. 1 Overview of the benefits and drawbacks of nanoplastic detection techniques.

3. Materials and methods

The comprehensive literature search in this study was conducted using the Web of Science Core Collection database, employing a systematic search strategy with the following parameters: Search Query: TS = (Bivalves OR Bivalve OR Clams OR Clam OR Mussels OR Mussel OR Scallop OR Scallops OR Oyster OR Oysters OR Cockle OR Cockles OR arcidae OR Scapharca OR cardioidal OR Corbicula OR Dreissena OR mercenaria OR Mya OR mugilidae OR Mytilus OR Perna OR ostreidae OR Crassostrea OR Ostrea OR pectinidae OR Pecten OR Pinctada OR spicula OR unionoida OR Anodonta OR unit) AND TS = (nanoplastic OR nanoplastics OR nano-plastics).

Timespan: January 1995 to 30 March 2025. Database Update: all years inclusive through March 30, 2025. Results: 472 publications retrieved.

This search syntax was designed to comprehensively capture studies investigating nanoplastic interactions with bivalve species and related taxonomic groups, while maintaining specificity through Boolean operators. The extended 30-year timeframe ensures inclusion of foundational research in both marine ecotoxicology and emerging nanoplastic studies. The search encompassed all document types including articles, reviews, and conference proceedings to maximize coverage of this developing research field.

Following initial screening based on titles and abstracts, articles meeting the following inclusion criteria were selected for this review: 1) employment of plastic particles with diameters <1000 nm; 2) investigation of nanoplastic bioaccumulation patterns or toxicological effects in bivalve species; and 3) implementation of reliable detection methodologies with appropriate quality control protocols.

4. Bioaccumulation of nanoplastics in bivalves

Nanoplastics internalized by bivalve cells may utilize various endocytosis pathways, including clathrin-mediated endocytosis, caveolin-mediated endocytosis, macropinocytosis, and passive transport.⁵⁰ Studies employing SERS with gold nanostars on Daphnia magna have shown that clathrin-mediated endocytosis and macropinocytosis are the primary mechanisms for nanoplastic translocation across intestinal barriers.⁵⁵ A recent study⁵³ identified clathrin and caveolin as the likely primary mechanisms for the endocytosis of polystyrene nanoplastics.^59,60 Gaspar et al. utilized deconvolution fluorescence microscopy and fluorescence imaging to demonstrate that hepatopancreatic cells in Crassostrea virginica can accumulate 50 nm PS nanoplastics.⁶¹ Sendra et al.⁶² investigated the internalization mechanisms of nanoplastics of varying sizes in Mytilus galloprovincialis, finding that hemocytes primarily internalized 1 μm PS nanoplastics through phagocytosis. In contrast, 50 and 100 nm PS nanoplastics were likely internalized via caveolae- and clathrin-mediated pathways, with internalization rates varying by particle size.⁶² Notably, the internalization rate for 50 nm nanoplastics was faster than that for 1 μm particles.⁶²

Table 1 presents a summary of current research on the bioaccumulation of nanoplastics in bivalves. In the broader context of nanoplastic bioaccumulation research, studies by Zhou et al.,⁶³ Wang et al.,⁶⁴ Liu et al.,⁶⁵ Meng et al.,⁶⁶ and Du et al.⁶⁷ consistently report significant accumulation of nanoplastics in the hepatopancreatic region. Additionally, research by Soubaneh et al.,⁶⁸ Gaspar et al.,⁶¹ Ventura et al.,⁶⁹ Metais et al.,⁷⁰ Chunmiao et al.,⁷¹ and Skawina et al.³⁶ has documented agglomerations of nanoplastics in gill tissues. The main findings from the studies are presented in Table 2.

Table 1 Studies on the bioaccumulation of nanoplastics in bivalves

Authors	Species	Nanoplastic	Diameter	Concentration	Exposure	Preparation of sample	Methods	Bioaccumulation
Merzel et al. 2020 (ref. 4)	Dreissena bugensis	COOH-modified fluorescent nanoplastic	200 nm	1 × 10^−12 M	24 h	KOH digestion	Fluorescence imaging combined with AFM-IR spectroscopy	Digestive tract
			1000 nm	1 × 10^−12 M				Siphon
Soubaneh et al. 2023 (ref. 68)	Placopecten magellanicus	^14C-radiolabelled	325 ± 121 nm	2.90 × 10^9 beads per L	6 h and 336 h	Not applicable	Whole-body autoradiography	None
		PS nanospheres
Gonçalves et al. 2023 (ref. 76)	Mytilus galloprovincialis	PS nanoplastics	50 nm	10 μg L^−1	672 h	Buffered mechanical homogenization method	Fluorescent quantification using DCVJ properties	Day 3: gills; days 7–14: gonads
								Day 14: digestive glands
Gaspar et al. 2018 (ref. 61)	Crassostrea virginica	PS nanoplastics	50 nm	10 and 100 μg L^−1	In vitro: 1, 2, 4 h	Not applicable	Deconvolution fluorescence microscopy, epifluorescence microscopy	Hepatopancreatic cells
			50 nm	50 μg L^−1	In vivo: 48 h			Hepatopancreas, gills
Ribeiro et al. 2022 (ref. 77)	Crassostrea gigas	Polystyrene nanoparticles doped with palladium (Pd)	Raspberry (151 ± 12 nm)	0.8 mg plastic per L	144 h	Acid digestion	Inductively coupled plasma mass spectrometry + TEM	Day 6: digestive gland
			Smooth (164 ± 11 nm)
Al-Sid-Cheikh et al. 2018 (ref. 78)	Pecten maximus	^14C-radiolabeled nanopolystyrene	24 ± 13 and 248 ± 21 nm	15 μg L^−1	6 h	Not applicable	Whole-body autoradiography	After 6-hour exposure, 250 nm plastics accumulate in the intestine, while 24 nm plastics distribute systemically
Seong et al. 2024 (ref. 79)	The D-shaped, umbo-veliger, and pediveliger stages of Crassostrea gigas	Fluorescent PS nanoplastics	550 nm	0.1, 1.0, 10.0, and 20 μg mL^−1	24 h	Not applicable	Fluorescence microscopy	Mantle
								Cavity, foot and velum
Chang and Wang 2024 (ref. 80)	Perna viridis	Fluorescent PS nanoplastics	49.1 ± 7.2 nm and 198.4 ± 12.4 nm	1 mg L^−1	8 h	Not applicable	Transmission electron microscopy and confocal microscopy	Granular and agranular hemocytes
Ventura et al. 2024 (ref. 69)	Ruditapes decussatus	PS nanoplastics	50 nm	10 μg L^−1	240 h	Buffered mechanical homogenization method	Fluorescent quantification using DCVJ properties	Gill and digestive gland
Zhou et al. 2023 (ref. 63)	Ruditapes philippinarum	Pd-doped polystyrene nanoparticles	139.5 ± 14.1 nm	0.02 mg L^−1 and 2 mg L^−1 and 20 mg L^−1	336 h	Acid digestion	TEM + inductively coupled plasma mass spectrometry	Digestive gland
		Fluorescent PS-NH2/PS-SO3H	66 ± 4.1 nm or 61.2 ± 3.9 nm			Histological section	Laser confocal microscopy	Digestive gland and gill
Wang et al. 2021 (ref. 64)	Mytilus coruscus	Fluorescent PS nanoplastics	70 nm	0.2 mg L^−1	87 h	KOH digestion	Multifunctional microplate reader	Gills: increased after 15-hour exposure, decreased after 87 hours, with no significant changes during recovery
								Digestive glands: significantly increased after 87-hour exposure, significantly decreased during recovery
Liu et al. 2021 (ref. 65)	Meretrix meretrix	Carboxylated polystyrene and amino-modified fluorescent polystyrene	100 nm PS-NH2, 200 nm PS-COOH	0.02, 0.2, and 2 mg L^−1	168 h	Histological section	Laser scanning confocal microscopy	Digestive gland
Cole and Galloway 2015 (ref. 81)	Crassostrea gigas	Fluorescent PS nanoplastics	0.07/0.16/0.87 μm/	1000 microplastics per mL	3, 10, and 24 d.p.f. juvenile exposed for one day	None	Fluorescence microscopy	160 nm PS beads in 100% of 3 d.p.f. larvae and >80% of 10 d.p.f. larvae, and 870 nm PS in >80% of 3 and 24 d.p.f. larvae
			0.87 μm fluorescent PS or 0.99 μm PS-NH2 or 0.94 μm PS-COOH		8 d.p.f. larvae	None		Aminated PS beads were present in a greater number of larvae than carboxylated and standard PS beads
Xin et al. 2022 (ref. 82)	Mytilus galloprovincialis	Fluorescent PS nanoplastics	101.83 ± 3.74 nm	0.5 mg L^−1, and 5.0 mg L^−1	168 h	KOH digestion	Multifunctional microplate reader	Digestive gland and gill
Moraz et al. 2021 (ref. 83)	Mytilus edulis	PSNP	49 nm and 100 nm and 300 nm	0.9–14.4 mg L^−1	168 h	10% KOH	Fluorescence microscopy	Gut
Metais et al. 2023 (ref. 70)	Scrobicularia plana	ENV NPs and PS NPs	ENV NPs: 235 ± 70 nm	5 mg L^−1	504 h	Concentrated nitric acid is heated to 250 °C in a microwave oven	Fluorescence microscopy	Gill and digestive gland
			PS NPs: 200 nm
Cid-Samamed et al. 2024 (ref. 84)	Mytilus galloprovincialis	PS NP-NH2	50 nm	0.01 μg mL^−1 and 2.5 μg mL^−1	168 h and 336 h and 504 h	None	Dynamic light scattering (DLS) + Scanning Electron Microscopy (SEM)	Gill and digestive gland
Kuehr et al. 2022 (ref. 85)	Corbicula fluminea	Fluorescent PS nanoplastics	47 and 100 nm	5 mg L^−1	72 h	Proteinase K digestion method	Multifunctional microplate reader	Muscle
Du et al. 2022 (ref. 67)	Ruditapes philippinarum	A nanoparticle model with gold as the substrate, Cy7 as the Raman reporter molecule, PS as the shell layer, and BSA coating	156.7 nm	0.2 mg L^−1	72 h	Acid digestion	Raman spectroscopy and inductively coupled plasma mass spectrometry (ICP-MS)	86.7% digestive gland followed by gill (5.2%), mantle (5.1%), foot (1.3%), exhalant siphon (1.1%), and adductor (0.6%)
Li et al. 2020 (ref. 86)	Corbicula fluminea	Fluorescent PS nanoplastics	80 nm	0.1 mg L^−1 and 1 mg L^−1 and 5 mg L^−1	96 h	Histological section	Laser confocal scanning microscopy	Liver, intestines, stomach, gills, and mantle
Wenger et al. 2012 (ref. 87)	Mytilus edulis	Fluorescent PS nanoplastics	30 nm	0.1 and 0.2 and 0.3 g L^−1	8 h	None	Ultraviolet-visible spectrophotometer converts turbidity to nanoplastic concentration	Foot
Chunmiao et al. 2023 (ref. 71)	Meretrix lyrata	Fluorescent PS nanoplastics	80 nm	1 mg L^−1	168 h	Direct imaging of fresh tissue	Laser confocal microscopy	Intestine, gill, mantle, foot, and siphon
Skawina et al. 2024 (ref. 36)	Unio tumidus	Fluorescent PS nanoplastics	15–18 nm	0.01 mg L^−1 and 1 mg L^−1 and 3 mg L^−1	48 h	None	Scanning electron microscopy, confocal microscopy, DXR Raman spectroscopy, FT-IR	Gut
			15–18 nm	3 mg L^−1	720 h
Meng et al. 2023 (ref. 66)	Ruditapes philippinarum	Fluorescent PVC and PMMA nanoplastics	98.6 ± 17.6 nm	1 mg L^−1 and 5 mg L^−1 and 15 mg L^−1 and 25 mg L^−1	24 h	Direct imaging of fresh tissue	Fluorescence microscopy	Mantle, gills, foot, liver, gonads, and intestines
			NP-PVC/
			111.9 ± 37.1 nm NP-PMMA

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Table 2 Overview of studies' main findings

Key findings	Supporting studies
Nanoplastics primarily accumulate in the gills, digestive gland, and hepatopancreas of bivalves	Wang et al., 2021;^64 Li et al., 2020;^86 Sendra et al., 2021 (ref. 140)
Smaller nanoplastics (e.g., 50 nm) exhibit higher bioaccumulation rates compared to larger particles	Gaspar et al., 2018;^61 Liu et al., 2021 (ref. 65)
Long-term exposure to nanoplastics leads to oxidative stress, immune suppression, and metabolic disorders	Capolupo et al., 2021;^95 Lebordais et al., 2021 (ref. 93)
Co-exposure with other pollutants (e.g., heavy metals, organic contaminants) exacerbates nanoplastic toxicity	Guo et al., 2021;^139 Arini et al., 2022 (ref. 138)
Nanoplastics disrupt larval development and reduce reproductive success in bivalves	Tallec et al., 2018;^120 Luan et al., 2019 (ref. 123)
Energy metabolism and growth are significantly affected by nanoplastic exposure, leading to reduced fitness	Gardon et al., 2024;^114 Wang et al., 2023 (ref. 137)

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Bioaccumulation patterns in bivalves demonstrate clear tissue specificity, with nanoplastics predominantly accumulating in filtration and digestive organs (Table 1). NPs smaller than 1 μm exhibit broader translocation, accumulating not only in the digestive gland and intestine but also in gills, visceral mass, muscle, foot, mantle, and even haemolymph. Internalization occurs via various endocytic pathways within specific cells (e.g., hepatopancreas cells, haemocytes, endothelial cells), with the dominant mechanism depending on NP size: phagocytosis for larger NPs (∼1 μm), and clathrin-mediated or caveolae-mediated endocytosis for smaller NPs (50–200 nm). Following cellular uptake, NPs, particularly negatively charged, environmentally derived ones, can disrupt critical cellular functions. A major mechanism of toxicity involves NP-induced mitochondrial dysfunction: they impair glycolysis and fatty acid oxidation (reducing ATP production), disrupt ion transport causing calcium overload, increase reactive oxygen species (ROS) formation, damage the electron transport chain, and alter mitochondrial membrane permeability. This triggers the expression of damage-associated molecular patterns (DAMPs), oxidation of NLRP3 inflammasome components, and opening of mitochondrial permeability transition pores, ultimately leading to cell death. Furthermore, NPs absorbed by intestinal epithelial cells can disrupt gut microbiota balance and compromise the intestinal barrier, increasing susceptibility to pathogens and inflammation (Fig. 2).

Fig. 2 Summary of the toxicity mechanism of nanoplastics.¹⁴¹

Probing deeper into the mechanisms, Fig. 3 and Table 3 show the many methods that viral nanoparticles (VNPs) and NPs use to enter the cells, emphasizing four different routes: direct penetration, endocytosis, macropinocytosis, and electrostatic adsorption.⁷² These pathways are influenced by the size of the nanoparticles, with examples shown for particles of 50 nm and 200 nm, and their interactions with the cell membrane (phospholipid bilayer) are detailed. Electrostatic adsorption is the process by which positively charged nanoparticles, such PS-NH₂, interact with negatively charged cell membrane components. The attachment of the nanoparticles to the membrane surface is made easier by this electrostatic attraction, which is very important for improving cellular uptake in a variety of biomedical applications, such as drug delivery systems. The nanoparticles' surface charge, which can be adjusted to maximize uptake efficiency, greatly influences the strength of this interaction. Endocytosis is subdivided into two major types: clathrin-mediated endocytosis and caveolin-mediated endocytosis. Clathrin proteins cover the inward budding membrane in a lattice-like structure during clathrin-mediated endocytosis, which ultimately results in the production of vesicles that swallow the nanoparticles. The majority of the nanoparticles that use this pathway are between 100 and 200 nm in size.⁷³ However, caveolin-mediated endocytosis requires caveolae, which are tiny invaginations of the plasma membrane that are rich in caveolin proteins. These caveolae are mostly in charge of internalizing nanoparticles that are larger than 200 nm or have particular surface characteristics.⁷³

Fig. 3 Potential mechanisms of nanoplastics infiltrating cells.

Table 3 Features of several endocytic processes

Endocytic mechanism	Characteristics	Particle size
Clathrin-mediated endocytosis	Dependent on clathrin-coated vesicles, primarily mediates the uptake of large molecules and particles	50–200 nm
Caveolin-mediated endocytosis	Dependent on caveolin, primarily mediates the uptake of small molecules and nanoparticles	50–200 nm
Macropinocytosis	Non-selective endocytic mechanism, mediates the uptake of large particles and fluids	>200 nm
Direct penetration	Particles cross the membrane directly without vesicle formation, often relies on chemical properties	<50 nm
Electrostatic adsorption	Particles attach to the cell membrane via electrostatic attraction, may trigger subsequent endocytosis but does not directly enter the cell	>50 nm

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The process by which the cell absorbs nanoparticles larger than 200 nm by large-scale membrane invagination is known as macropinocytosis. Immune cells like macrophages frequently use this technique to absorb particulate debris, including nanoparticles.⁷⁴ The process of macropinocytosis results in the development of bigger vesicles that non-specifically absorb external materials and nanoparticles into the cell, in contrast to other endocytic routes that produce smaller vesicles. In the context of delivering drugs and vaccine development, where bigger particles might be employed to boost cellular uptake in certain target cells, this pathway is very pertinent.

The delivery of small, functional molecules directly into the cytoplasm is made possible by direct penetration, which enables nanoparticles smaller than 50 nm, such as dendrimers, small metal clusters, and cell-penetrating peptides, to pass through the cell membrane without requiring vesicle-mediated endocytosis.⁷⁵ These nanoparticles take use of the cell membrane's natural permeability, frequently by temporarily creating hydrophobic areas that allow passage. These particles are perfect for gene therapy, drugs transport, and other intracellular applications because of their small size and functionalization, which are essential for breaking down cellular barriers.

Thus, NP bioaccumulation is not merely tissue-specific but a dynamic process linking cellular uptake routes (endocytosis, adsorption) to systemic oxidative, immune, and metabolic toxicity. This mechanistic nexus explains chronic impacts beyond detectable tissue presence.

5. Toxicity mechanism of nanoplastics on bivalves

5.1 Antioxidant system overload

The enzymatic antioxidant system in bivalves consists of key enzymes such as superoxide dismutase (SOD), catalase (CAT), and glutathione S-transferase (GST).⁸⁸ Upon exposure to nanoplastics, these particles interact with cell membranes and can trigger inflammatory responses through surface receptors (e.g., Toll-like receptors, TLR).⁸⁹ Alternatively, nanoplastics may be endocytosed by cells, disrupting the intracellular redox balance and promoting the mitochondria to produce adenosine triphosphate (ATP) via oxidative phosphorylation and the electron transport chain (ETC) to mitigate cellular damage.⁹⁰ However, this process also stimulates the production of superoxide anions,⁹⁰ which must be detoxified by antioxidant enzymes like SOD, CAT, and GST. Excessive production of reactive oxygen species (ROS), however, can lead to the generation of peroxidation products, such as malondialdehyde (MDA) and protein carbonyls,⁹¹ which increase cell membrane permeability and disrupt cellular function. If this oxidative stress is not adequately addressed, it can result in cellular damage or, ultimately, cell death.⁹²

Initial NP–cell interactions engage surface receptors (e.g., Toll-like receptors) to trigger inflammatory signaling,⁸⁹ while internalized particles disrupt mitochondrial electron transport chains (ETC).⁹⁰ Cell inflammation increases ATP demand, but damage to the ETC creates superoxide anions that consume antioxidants,⁹⁰ leading to uncontrolled reactive oxygen species (ROS) conversion into cytotoxic peroxidation products (e.g., malondialdehyde, protein carbonyls).⁹¹ The resulting membrane permeability increase and apoptosis activation⁹² exhibit tissue-temporal specificity: gill/digestive glands show acute compensatory CAT/GST/SOD upregulation within 7 days,⁹³ but chronic exposure (21 days) causes systemic collapse (↓SOD, ↑DNA damage, histopathology).^94–97

Digestive glands (high metabolic activity) suffer earlier mitochondrial failure and MDA surges^86,98 than gills with higher redox thresholds.⁶³ However, the most pronounced effects may occur in the gills after prolonged exposure. Trophic transfer of nanoplastics continues to activate catalase (CAT) activity even after 21 days.⁹⁶ It should be noted, though, that these experiments involved different species of bivalves and utilized nanoparticles from different sources, making direct comparisons unreliable. While both studies emphasize the environmental relevance of nanoplastics, neither considered the potential epigenetic alterations or modifications that can occur during exposure. Epigenetic regulation—which can occur both intergenerationally and as a reversible “wash-in and wash-out” effect in response to stress—involves methylation and acetylation of promoter regions in specific gene sequences, thereby influencing gene expression. Overlooking this mechanism may lead to an incomplete understanding of how bivalves modulate their responses to nanoplastic-induced stress.

Ultimately, the ROS–antioxidant imbalance progresses through three sequential pathological stages: (i) initial electron transport chain (ETC) disruption,⁹⁰ which leads to (ii) macromolecular peroxidation,^63,86,91 and finally culminates in (iii) tissue necrosis. This cascade's severity is modulated by three key factors: nanoparticle size, exposure duration, and specific biotic interfaces.

5.2 Subcellular organelle dysfunction

5.2.1 Mitochondrial failure. NP exposure directly compromises mitochondrial integrity in bivalve immune cells, triggering a cascade of metabolic dysfunctions. Charged NPs (e.g., PS-NH₂) adsorb electrostatically to cell membranes, penetrating tissues via endocytosis or paracellular pathway.^99,100 Once internalized, NPs disrupt mitochondrial membranes, reducing membrane potential (Δψ_m) and stability—as observed in Mytilus galloprovincialis hemocytes after 24-hour exposure to 10 μg L⁻¹ PS-NH₂.¹⁰¹ This damage impairs oxidative phosphorylation, forcing cells to increase ATP production despite escalating superoxide generation.⁹⁰ Concurrently, NPs induce phosphatidylglycerol peroxidation in the inner mitochondrial membrane,¹⁰² depleting ATP reserves while accelerating ROS accumulation. The resulting metabolic crisis activates apoptosis: NP-exposed hemocytes show upregulated Bcl-2-associated X protein (BAX) and caspase-3 expression,¹⁰³ culminating in phagocyte depletion. Critically, mitochondrial failure amplifies immune suppression; reduced Δψ_m correlates with diminished bacterial killing capacity in hemolymph,¹⁰¹ leaving bivalves vulnerable to pathogens.

5.2.2 Lysosomal & ER stress. NPs concurrently induce lysosomal destabilization and endoplasmic reticulum (ER) stress, crippling key immune functions. Lysosomal acidification and membrane permeabilization are hallmarks of NP toxicity—documented in Meretrix meretrix hemocytes exposed to PS-NH₂/PS-COOH (0.02–2 mg L⁻¹), where lysosomal instability reduced phagocytic efficiency by 40–60%.⁶⁵ This damage stems from NP-triggered cathepsin L downregulation⁶⁹ and sustained p38 MAPK phosphorylation,¹⁰² which dysregulate phagosome–lysosome fusion. ER stress further exacerbates dysfunction: transcriptomic analyses reveal NP-induced suppression of NF-κB inhibitor alpha and interleukin-1 receptor-associated kinase 4,⁶⁵ disabling Toll-like receptor and NF-κB signaling pathways essential for cytokine release. Protein corona formation intensifies these effects; corona-coated NPs exhibit 3-fold higher lysosomal membrane destabilization in hemolymph serum versus artificial seawater.¹⁰² Consequently, immune homeostasis collapses: granulocyte/hyalinocyte ratios shift,¹⁰⁴ antimicrobial peptide genes (e.g., Myticin C) are downregulated,⁶² and phagocytic recognition of pathogens fails even during co-exposure with Vibrio splendidus.⁶²

5.3 Tissue-specific pathologies

5.3.1 Gut-barrier breakdown & microbiota dysbiosis. NPs compromise intestinal health through three cascade mechanisms. First, they induce gut microbiota dysbiosis by altering microbial composition. PS-NPs, with a size of 500 nm and concentration of 0.6 mg L⁻¹, reduce beneficial Spirochaetes while elevating pathogenic Firmicutes, Proteobacteria, and Bacteroidetes in Perna viridis.¹⁰⁵ Charged NPs, such as Pd-doped PS with a size of 139.5 nm, increase Vibrionaceae abundance in Ruditapes philippinarum, with a parallel downregulation of carbohydrate metabolism and immunity.^63,106 High-concentration PS-NPs, at 10⁴ particles per L, promote opportunistic genera like Phaeobacter and Nautella in Crassostrea angulata.¹⁰⁷ Second, NPs disrupt intestinal barrier integrity through mucosal damage. NPs at concentrations of 0.1–1 mg L⁻¹ cause severe epithelial injury in Corbicula fluminea, exceeding the effects of microplastics.¹⁰⁸ They also dysregulate lipid transport at 1 mg L⁻¹, inducing biomembrane stress and vesicular NP transcytosis,^108–110 as evidenced by PS-COOH NPs at 50 μg L⁻¹ suppressing microbial lipid/nucleic acid synthesis and amino acid metabolism in Laternula elliptica.¹¹¹ Third, metabolic reprogramming occurs under environmentally relevant concentrations. NPs at 10–100 μg L⁻¹ enrich amino acid and carbon metabolism pathways in Magallana hongkongensis, while sublethal doses ranging from 500 to 2500 μg L⁻¹ trigger stress responses. These doses upregulate genes for ABC transporters, apoptosis, and lysosomes, while inhibiting ribosomal biosynthesis.¹¹²

5.3.2 Histological alterations in digestive gland/gills. NPs induce progressive structural damage in digestive tissues through concentration-dependent and size/charge-mediated pathological mechanisms, as demonstrated in multiple histopathological studies. For instance, in Meretrix meretrix, exposure to aminated (PS-NH₂, 100 nm) and carboxylated (PS-COOH, 200 nm) polystyrene NPs resulted in dose-dependent digestive gland atrophy at 2000 μg L⁻¹, characterized by luminal dilation, epithelial detachment, and tubule necrosis.⁶⁵ These alterations are attributed to direct physical membrane penetration, oxidative stress, and a reduction in transepithelial electrical resistance mediated by protein phosphorylation.

Similarly, in Mytilus galloprovincialis, treatment with 200 nm PS-NPs (20–2000 μg L⁻¹) led to significant epithelial thinning, loss of digestive cells, and lumen expansion, quantified through increased branch-to-lumen radius ratios (ABR/ALR) and MET ratios.¹¹³ Accompanying these morphological changes, molecular analyses revealed that at 200 μg L⁻¹, miR-34 promoted apoptosis and inflammation by inhibiting cell proliferation, while downregulated miR-281 mitigated cellular damage via nucleotide excision repair. At the higher concentration of 2000 μg L⁻¹, novel_mir63 was upregulated, disrupting circadian rhythm, neural excitability, and synaptic plasticity. Together, these molecular changes provide a mechanistic explanation for the observed histopathological alterations.

It should be noted that results obtained at 2000 μg L⁻¹ may lack direct environmental relevance due to the high exposure level, and thus should be interpreted with caution in ecological risk assessment. Collectively, these histopathological observations morphologically validate previously reported subcellular toxicities—such as mitochondrial and lysosomal dysfunction—resulting from NP accumulation, thereby establishing a direct link between nanoplastic-induced metabolic disruption and tissue-level damage.

5.4 Organismal consequences

5.4.1 Energy allocation trade-offs. NPs impose significant energetic constraints on bivalves by disrupting the delicate balance of energy allocation among maintenance, growth, and reproduction. Bivalve growth and physiological performance are governed by the equilibrium between energy intake and expenditure, which typically sustains essential processes such as basal metabolism, immune function, tissue development, and gamete production. Under NP exposure, however, bivalves are forced into a state of metabolic triage, reallocating energy away from growth and reproduction to support stress responses.

This metabolic shift occurs through several interconnected mechanisms. NPs activate MAPK and P450 pathways, mobilizing 30–50% of glycogen reserves for antioxidant defense, while concurrently suppressing biomineralization capacity—evidenced by the formation of thinner aragonite sheets in pearl oysters.¹¹⁴ In Meretrix meretrix, nanoplastics elevate energy expenditure via enhanced pancreatic secretion, protein digestion and absorption, tricarboxylic acid (TCA) cycle activity, and PPARα signaling.¹¹³ Although compensatory responses such as transient upregulation of succinate dehydrogenase (in the electron transport chain) and D-sorbitol accumulation occur in species like Crassostrea angulata,¹⁰⁷ chronic NP exposure ultimately leads to systemic metabolic failure.

High-dose polystyrene NPs reduce total lipid content by 40–60% in Sinonovacula constricta, while stimulating phospholipid synthesis to repair damaged membranes.¹¹⁵ These changes coincide with Na⁺/K⁺-ATPase inhibition, which disrupts ion homeostasis,¹¹⁹ and declined activity of TCA cycle enzymes, resulting in reduced ATP production.¹¹⁶ Similar to other environmental stressors—such as ocean acidification and pathogen infection—NPs elevate the basal energy demand for detoxification. This forces bivalves to either increase food intake (where feasible) or divert energy resources from critical biological functions.

The consequent energy reallocation has profound ecological implications, reducing gamete production by 22–38% (ref. 117 and 118) and impairing shell formation. These effects collectively diminish individual fitness and threaten population viability in NP-contaminated habitats. Thus, the NP-induced energy crisis represents a multi-level physiological threat, disrupting cellular metabolism while constraining organismal energy budgets essential for growth and reproduction.

5.4.2 Larval developmental defects. NPs exert profound detrimental effects on bivalve larval development through multiple interconnected pathways, beginning with direct impairment of reproductive success. Exposure to PS-NH₂ (50 nm) nanoparticles reduces oyster sperm motility by 79% and decreases swimming speed by 62% through membrane adhesion mechanisms,¹¹⁹ while concurrent ROS overproduction disrupts sperm capacitation processes.^120,121 Following fertilization, embryos exhibit charge-dependent vulnerability, with positively charged PS-NH₂ (0.15 μg mL⁻¹) arresting development in 90% of oyster embryos compared to PS-COOH (25 μg mL⁻¹) due to enhanced electrostatic endocytosis.¹²² The resulting developmental abnormalities, including characteristic umbrellar larva deformities in Meretrix meretrix (1–2 mg L⁻¹ PS-NH₂), stem from fundamental cellular damage through lipid peroxidation and DNA disruption.¹²³ Importantly, these impacts are significantly modulated by environmental factors, as protein and humic acid coatings forming eco-coronas alter NP–membrane interactions, either exacerbating or mitigating developmental arrest depending on their specific composition.^{122,124–127} These findings collectively demonstrate how NPs disrupt critical developmental processes across multiple life stages, from gamete function to larval morphogenesis, with potentially severe consequences for bivalve population recruitment and ecosystem stability.

5.4.3 Co-stressor sensitization. NPs interact synergistically with environmental co-stressors through three primary mechanisms that collectively enhance organismal vulnerability, though the precise nature of these interactions depends on specific physicochemical and biological contexts. The pollutant carrier effect demonstrates complex modulation of toxicity, where PS-NPs increase arsenic bioaccumulation by 45% in oysters through inhibition of methylation detoxification pathways,⁹³ while conversely, certain organic coatings may reduce bioavailability of other contaminants like benzo[a]pyrene, which nonetheless induces DNA damage at 1000-fold lower doses when combined with NPs.¹²⁸ Eco-corona formation may explain this; potential interactions between HA and nanoplastic particles (NPPs) could lead to the formation of an eco-corona, ultimately modulating the toxicological behavior of NPPs. The adsorption process is primarily driven by hydrophobic and π–π interactions, as demonstrated in studies on carbon-based nanoparticles.^129,130 Additionally, carboxylic groups on HA may facilitate its interaction with NPPs.¹³¹ Such strong intermolecular attractions within the adsorbed layer are capable of altering the surface characteristics of suspended NPPs, consequently affecting their toxicological responses. This effect, however, is seldom realized in bivalve research. When nanoplastics form eco-corona with other negatively charged molecular such as NOM, they may reduce oxidative stress by elevating antioxidant enzyme activities and exert proinflammatory and anti-inflammatory effects by modulating the immune system as they have a higher chance to interact with the positive charged animal protein.¹³² Immune–energy trade-off failures reveal consistent stress amplification, as evidenced by hypoxia tripling NP immunotoxicity in mussels through respiratory stress,⁸² and ciprofloxacin co-exposure (0.5 μg g⁻¹) with NPs (10 μg g⁻¹) depleting 70% of SOD/GSH antioxidant reserves via parallel metabolic burdens.¹⁰⁵ Transgenerational studies indicate unpredictable epigenetic outcomes, where norfloxacin–NP combinations upregulate HSP70/IRAK-4 markers suggesting chronic immunotoxicity,¹³³ while carbamazepine–NP coexposure paradoxically downregulates HSP70 alongside esterase elevation in gills, reflecting stress response pathway dysregulation.¹³⁴ The underlying of co-stressor mechanisms involve: (1) physical carrier effects altering contaminant bioavailability through surface adsorption/desorption kinetics; (2) shared metabolic pathways (e.g., CYP450 inhibition) creating competitive detoxification bottlenecks; and (3) cumulative oxidative damage exceeding cellular repair capacity. However, the net effect (toxicity enhancement or mitigation) depends critically on NP surface properties (e.g., NH₂vs. COOH), environmental matrix composition, and organismal life stage, explaining observed variations in co-stressor outcomes across studies.

5.5 Summary of toxicity mechanisms of nanoplastics

NPs initiate their toxic effects in bivalves primarily through mitochondrial disruption and antioxidant system overload, triggering a cascade of cellular and systemic disturbances. Following cellular internalization, these particles first impair mitochondrial function by simultaneously disrupting energy metabolism through inhibited glycolysis and fatty acid oxidation (reducing acetyl-CoA production and ATP synthesis), inducing calcium overload via Na⁺/K⁺-ATPase interference, and generating oxidative stress through electron transport chain dysfunction. This mitochondrial damage leads to three critical consequences: (1) structural compromise of intestinal epithelial cells through ROS-mediated tight junction degradation, (2) activation of inflammatory pathways (NLRP3 inflammasome via oxidized mitochondrial DNA), and (3) metabolic reprogramming that depletes antioxidant reserves (glutathione, SOD, CAT). The resulting oxidative stress propagates outward to disrupt gut microbiota equilibrium, characterized by a 40–60% reduction in beneficial symbionts and 2–3 fold increase in opportunistic pathogens, which further exacerbates intestinal barrier dysfunction. As this vicious cycle progresses, systemic effects emerge including impaired hemocyte phagocytosis (50–70% reduction), lysosomal destabilization, and chronic inflammation mediated through NF-κB and Toll-like receptor pathways. Crucially, the sustained energy demands for cellular repair and antioxidant defense force bivalves into metabolic triage, diverting resources from growth and reproduction – evidenced by 30–50% reductions in gametogenesis and shell deposition under chronic exposure. This mechanistic cascade, progressing from molecular oxidative damage to population-level fitness consequences, explains how nanoplastic accumulation ultimately threatens bivalve viability even at sublethal concentrations.

6. Future outlook

Future research must focus on simulating the effects of nanoplastic exposure on bivalves under realistic environmental conditions. As highlighted in earlier discussions, nanoplastics interact with biological and ecological systems in complex ways, such as forming protein coronas or eco-corona structures that significantly influence their behavior, bioavailability, and toxicity (Fig. 4). Addressing these interactions is critical for understanding the mechanisms of nanoplastic toxicity and bioaccumulation in bivalves. Laboratory experiments should prioritize simulating realistic environmental conditions, including variations in organic matter content, pH, salinity, and other ecological factors. The interaction of nanoplastics with natural organic matter, sediments, and colloids often alters their aggregation and toxicity. Future studies must evaluate these eco-corona formations to better predict nanoplastic behavior in natural habitats. Future research should focus on investigating the formation and characteristics of eco-coronas and protein coronas on nanoplastics under conditions that involve co-existing contaminants and natural organic matter. This can be achieved by using real environmental water samples (e.g., lake water, seawater) or by preparing simulated environmental media with standard natural organic matter, such as Suwannee River Humic Acid (SRHA). Additionally, artificially simulated bivalve hemolymph should be employed to mimic their bodily fluids. Techniques such as SDS-PAGE and mass spectrometry can be applied to analyze the composition of the eco-coronas and protein coronas formed on nanoplastics. Morphological changes can be observed using SEM and TEM, while alterations in zeta potential and particle size distribution can be assessed using laser particle analyzers. Additionally, nanoplastic morphology and polymeric diversity, such as fibers from aquaculture ropes or fragments from degraded consumer products, need further investigation. The differential toxicological impacts of various nanoplastic forms and compositions are significant. Experiments utilizing diverse plastic types such as polyethylene and polypropylene will provide more comprehensive insights into their toxicological and ecological consequences.

Fig. 4 The eco-corona and protein corona formed on nanoplastics (created with https://BioRender.com using data from the Protein Data Bank (PDB)).

One major limitation is the inconsistency in experimental methodologies. Future studies should adopt standardized protocols for nanoplastic concentration, exposure duration, and unit measurements. The use of mass concentration fails to reflect particle number, whereas particle number concentration does not adequately represent particle size or total mass. Therefore, it is highly recommended to employ both expressions concurrently in experimental studies. Furthermore, the application of surface area concentration (cm² mL⁻¹) is also critically important. The surface of nanoparticles serves as the primary site for chemical reactions, pollutant adsorption, and interactions with cells. Thus, quantifying the surface area is essential for investigating the “Trojan horse effect” mediated by adsorbed pollutants. Furthermore, improved analytical methods for detecting nanoplastics in biological tissues are essential. Fluorescence-based techniques, though widely used, are susceptible to photobleaching and quenching. Combining advanced methods like surface-enhanced Raman spectroscopy (SERS), gas chromatography-mass spectrometry (GC-MS), and atomic force microscopy (AFM) will improve detection accuracy and reliability. Future efforts should focus on integrating multimodal platforms that minimize environmental interference—such as NOM-induced aggregation and plasmonic heating—while leveraging machine learning and microfluidics to enhance specificity and throughput in complex matrices. Understanding the molecular mechanisms of nanoplastic toxicity including oxidative stress, immune disruption, and metabolic interference is paramount. Future research should focus on adverse outcome pathways (AOPs) to identify key molecular events linking nanoplastic exposure to physiological and ecological effects. This will enable more precise predictions of long-term impacts on bivalve populations and ecosystems. For instance, studies on the role of antioxidant enzyme systems (e.g., SOD, CAT, and GST) should extend beyond reactive oxygen species (ROS) generation to include lipid peroxidation and mitochondrial dysfunction. Transcriptomic and metabolomic analyses could reveal additional biomarkers for assessing nanoplastic-induced stress responses.

The synergistic effects of nanoplastics and other pollutants, such as heavy metals, pharmaceuticals, and organic contaminants, must also be evaluated. Future studies should simulate co-exposure scenarios to assess how these interactions influence bioaccumulation, toxicity, and ecological dynamics. Understanding these combined effects is crucial for developing comprehensive environmental management strategies. To mitigate the environmental impact of nanoplastics, aquaculture practices must be optimized. Replacing synthetic materials with biodegradable alternatives for nets and ropes could reduce nanoplastic generation. The significant contribution of aquaculture-derived fibers to nanoplastic pollution in marine ecosystems underscores the importance of developing and adopting sustainable materials to minimize long-term ecological footprints.

Future research should adopt an ecosystem-based management approach to address nanoplastic pollution. Integrating data on nanoplastic interactions with bivalve physiology will provide a holistic understanding of their ecological implications. Collaboration between researchers, policymakers, and industry stakeholders is essential to develop actionable strategies for mitigating nanoplastic contamination. By addressing these research gaps, future studies can improve our understanding of nanoplastic toxicity and develop effective mitigation strategies. A standardized, interdisciplinary approach will be essential for tackling the multifaceted challenges posed by nanoplastics to marine ecosystems and bivalve populations.

To reduce nanoplastic pollution in aquaculture, it is recommended that policymakers strengthen regulation of plastic waste management and promote the use of biodegradable materials. Industry stakeholders should optimize aquaculture equipment, reduce the use of plastic ropes and nets, and adopt environmentally friendly materials to replace traditional plastics. Microplastics and nanoplastics may persist in the environment, making them difficult to remove in the short term and thus becoming a long-term global problem.¹³⁵ Many experts believe that governments can effectively reduce or even control this pollution problem by formulating relevant laws and regulations. For example, some countries, such as the United Kingdom, have taken measures to restrict the production and consumption of plastics, such as banning the distribution and sale of plastic products.¹³⁶ Effective mitigation of nanoplastic pollution requires targeted interdisciplinary collaborations. Environmental chemists and toxicologists should work with public health experts to establish safety thresholds, while materials scientists partner with industry to develop biodegradable alternatives and advanced filtration. Data scientists and ecologists can integrate monitoring with modeling to improve risk assessment, and social scientists must collaborate with policymakers to design practical regulations. Additionally, urban planners and waste management specialists need to work with industry on implementing circular economy strategies. These coordinated efforts are essential to translate research into actionable solutions.

Conclusion

This review comprehensively examines the toxicological effects of nanoplastics on bivalves, highlighting their ability to induce oxidative stress, immune dysregulation, and metabolic disturbances. It explores the mechanisms of nanoplastic accumulation and their interactions with biological structures, emphasizing the role of protein and eco-coronas in influencing bioavailability and toxicity. The bioaccumulation pathways reveal species-specific differences in retention and elimination, with nanoplastics' small size facilitating tissue penetration and posing risks to bivalve physiology and food safety.

The analysis of toxicity mechanisms shows how nanoplastics impair antioxidant enzyme activity, disrupt immune functions, and alter gut microbiota, leading to cellular homeostasis disruptions, genotoxicity, and increased vulnerability to co-pollutants. The intricate relationship between particle characteristics, such as charge and size, and their biological impacts is also elucidated.

To address these challenges, future research must simulate realistic environmental conditions, evaluate the synergistic effects of co-pollutants, and refine detection methodologies. Standardizing experimental protocols and employing advanced analytical tools will be crucial for advancing the field. Additionally, ecosystem-based approaches and collaborative efforts among researchers, policymakers, and industry stakeholders are necessary to develop effective mitigation strategies.

By synthesizing current knowledge, this review identifies critical research gaps and lays the groundwork for addressing the multifaceted challenges posed by nanoplastics. These efforts are essential for protecting marine ecosystems and mitigating the long-term impacts of nanoplastics on bivalves and aquatic environments.

Conflicts of interest

There are no conflicts to declare.

Data availability

The data supporting the findings of this review are derived from previously published studies, which are cited throughout the manuscript. No new primary data were generated or analyzed during the preparation of this review. Readers interested in the original data sources can refer to the cited articles for access to the specific datasets.

Acknowledgements

This work was supported by the “14th Five-Year Plan” National Key Research and Development Program “Marine Agricuand Freshwater Fisheries Science and Technology Innovation” Key Special Project (2024YFD2402203), the Innovation Program of Shanghai Municipal Education Commission (2023ZKZD52), the Open Research Fund of State Key Laboratory of Estuarine and Coastal Research (Grant number SKLEC-KF202309), and the Foreign Experts Program of the Ministry of Science and Technology (S20240181).

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Footnote

† These authors have contributed equally to this work and share first authorship.

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