The Menace of Microplastics and Their Physiological Implications in Aquaculture

Abstract

The rapid expansion of aquaculture has increased its reliance on plastic-based infrastructure and materials, creating opportunities for microplastic (MP) contamination from culture systems, surrounding waters, and formulated feeds. This review synthesises current evidence on MP sources, exposure pathways, biological fates, and toxicological effects in farmed fish, with emphasis on their interconnected implications for fish health and aquaculture performance. MPs enter aquaculture systems through degraded equipment, contaminated water, atmospheric deposition, and feed ingredients, including fishmeal derived from MP-exposed fish. Ingestion represents the principal exposure route, with particle size and shape, polymer composition and concentration, exposure duration, life stage, and species influencing retention, translocation, and biological responses. Although gastrointestinal accumulation is common, smaller particles may translocate to internal organs, while evidence for trophic transfer does not yet establish consistent biomagnification. Across studies, MP exposure is associated with intestinal and gill damage, altered digestive enzyme activity, impaired feeding and feed utilisation, metabolic disturbances, haematological and histopathological alterations, and reduced growth. Oxidative stress emerges as a central, but not exclusive, mechanistic pathway, linking reactive oxygen species imbalance and altered antioxidant defences with inflammation, immunotoxicity, neurotoxicity, behavioural alterations, and endocrine and reproductive disturbances. These responses are highly heterogeneous, reflecting differences in species, polymer type, particle characteristics, exposure concentration, duration, life stage, environmental conditions, and experimental design. Interactions between MPs and co-occurring contaminants may further modify their biological effects. Overall, the evidence indicates that MP exposure represents a multidimensional stressor capable of connecting environmental contamination with physiological dysfunction and impaired aquaculture performance. Understanding these interconnected responses is essential for strengthening MP risk assessment and supporting effective feed and farm management, fish health protection, food safety, and sustainable aquaculture.

1. Introduction

Aquaculture has become a central component of global aquatic food production. Farmed aquatic animals surpassed capture fisheries output in 2022 and now account for approximately 51% of global aquatic animal production, with output concentrated in Asia [1]. Alongside this expansion, plastic pollution has emerged as a transboundary environmental and socio-economic problem affecting waterways, coastal systems, seafood security, and coastal livelihoods in aquaculture-producing regions worldwide [2,3]. Within this contamination landscape, microplastics (MPs)—conventionally defined as plastic particles smaller than 5 mm—are of particular concern, being heterogeneous in size, shape, density, and polymer composition and bioavailable across the full range of feeding guilds present in cultured systems [4,5,6].

Aquaculture is distinctive among food production sectors in that it both receives and generates this contamination. Modern hatchery, pond, cage, and land-based systems depend on nets, cages, ropes, liners, tanks, pipes, floats, and packaging [7], which shed fibres and fragments under ultraviolet exposure, mechanical abrasion, weathering, and biofouling [8,9]. Facilities simultaneously receive MPs from source waters draining rivers, estuaries, coastal environments, and wastewater pathways, so environmental and farm-derived contamination cannot be cleanly separated; in intensively farmed regions such as Southeast Asia, MPs have been reported across the rivers, estuaries, coasts, and biota connected to aquaculture water supply and receiving environments [10,11]. Routine production practices add further inputs and can re-suspend legacy particles settled in pond and cage sediments, while fishmeal and shrimp meal derived from wild-caught organisms introduce MPs directly through formulated diets [8]. The resulting exposure environment is therefore cumulative rather than incidental, and differences in water exchange, infrastructure, and containment govern how particles are distributed and retained within it [8,12].

The biological consequences of this exposure depend on the interaction between particle characteristics and fish physiology, and extend well beyond the site of contact. Ingested and branchially retained particles have been associated with epithelial injury, altered digestive enzyme activity, and gut microbial disturbance, with downstream effects on redox balance, immune and metabolic regulation, endocrine and reproductive function, and behaviour [12,13,14,15,16]. Three limitations nonetheless restrict the usefulness of this literature to aquaculture. First, most available evidence derives from short-term exposures at concentrations exceeding those reported from production systems, so the extent to which such findings translate to farmed fish under commercial conditions remains unresolved. Second, although sources, uptake, and toxicity have each been reviewed in detail [8,12,13,14], they have generally been treated as separate subjects, so the pathway from a contaminated input to a measurable production outcome has not been traced end to end. Third, the dietary route—MPs carried into culture systems through fishmeal and formulated feed—has been treated as peripheral, despite being the pathway most directly amenable to control at farm level.

This review traces MP exposure in cultured finfish along a single continuous pathway: from sources and entry routes in aquaculture infrastructure, culture waters, and formulated feeds; through ingestion, branchial and dermal uptake, bioaccumulation, and co-contaminant transport; to physiological consequences at the digestive, metabolic, oxidative, haematological, histopathological, immune, neuro-behavioural, endocrine, and reproductive levels; and to their expression as impaired growth, feed efficiency, disease resistance, and reproductive performance. The modifying influence of polymer type, particle characteristics, exposure dose, and duration is considered throughout. Evidence is drawn from studies conducted globally and concerns cultured and culture-relevant finfish, with invertebrate data cited only where they establish a mechanism for which finfish evidence is unavailable. The review concludes by identifying priority knowledge gaps and evaluating the extent to which the current evidence base can support risk assessment and practical management of MP contamination in aquaculture.

2. Types and Sources of Microplastics

2.1. Classification by Origin

Microplastics are conventionally distinguished by origin. Primary microplastics are manufactured at microscopic scale for incorporation into products such as cosmetics, cleaning agents, and industrial abrasives, and enter the environment directly through use and disposal; microbeads and pre-production resin pellets are the principal examples [17,18]. Secondary microplastics arise from the progressive photodegradation, mechanical weathering, and fragmentation of larger debris, including packaging waste, abandoned or damaged fishing gear, antifouling paint flakes, and synthetic microfibres shed during domestic laundering [18,19,20]. The distinction is retained here because the two classes differ in the properties that govern biological interaction: primary particles are comparatively uniform in size and shape, whereas secondary particles are irregular and surface-weathered.

2.2. Morphology and Polymer Types of MPs Detected in Fish

Microplastics recovered from fish and other marine organisms, including sharks, are typically identified in the form of fragments, films, foams, flakes, and fibres, which are either ingested directly or adhere to gill surfaces [21,22]. The polymer composition identified across studies varies considerably [21], though spectroscopic techniques such as Fourier-transform infrared spectroscopy (FTIR) have most frequently identified polypropylene (PP) [23], polyethylene (PE) [21], nylon-6 (PN6) [24], polyamide (PA) [21], polyethylene terephthalate (PET) [21], polyvinyl chloride (PVC) [25], and polystyrene (PS) [21] as the dominant polymer types.

These morphological and polymer classes are not interchangeable descriptors; each maps onto a different point in the exposure sequence examined in the remainder of this review. Density determines whether a particle remains suspended in the water column or settles into pond and cage sediment, and therefore which species and feeding guilds encounter it: the low density of PE and PP, for example, underlies both their transport into aquatic systems and their persistence in surface waters [26], while pelagic and filter-feeding species have been reported to ingest microplastics more readily than demersal or selectively feeding species [27,28]. Morphology and size determine physical fate within the organism, governing whether a particle is retained in the intestinal lumen, egested, or translocated across the epithelium. Polymer identity determines chemical exposure, since microplastics carry manufacturing additives and can sorb environmental contaminants that are subsequently released within the tissue [29,30]. These three properties recur throughout Section 4, Section 5, Section 6 and Section 7 as the principal modifiers of toxicological outcome. The size, morphological, colour, and polymer classes by which microplastics recovered from aquatic organisms are conventionally described are summarised in Figure 1.

Figure 1. Descriptive characteristics conventionally used to classify microplastics recovered from aquatic environments and organisms: size class (<1 µm, 1–10 µm, 10–100 µm, 100–500 µm, >500 µm), shape (fibre, fragment, film, sphere, foam), colour, and polymer type (polyethylene, polypropylene, polyvinyl chloride, polyethylene terephthalate, polystyrene, polyamide, polyester).

3. Environmental Pathways Facilitating MP Entry into Aquatic Systems

Modern aquaculture depends on polymer-based materials such as liners and tarpaulins, nets and ropes, feed bags, floats, pipes, aeration components, and plastic packaging [31]. Under ultraviolet exposure, heat, mechanical abrasion, and biofouling, these items shed microplastics, predominantly fibres and fragments, into the water column and settled sludge [32]. Aquaculture facilities additionally receive microplastics from the surface waters that supply them—rivers, lakes, shallow marine zones, and open ocean systems—through multiple interconnected pathways shaped by both anthropogenic activity and natural transport processes [33]. These pathways differ between freshwater and marine systems, although several, notably atmospheric deposition, contribute to both.

Microplastics reaching aquaculture facilities from the surrounding catchment derive from a limited number of recurring sources, summarised in Table 1. In inland systems, treated municipal sewage, agricultural runoff, and domestic laundering dominate reported inputs, delivering polypropylene (PP), polyvinyl chloride (PVC), polystyrene (PS), polyethylene (PE), and polyethylene terephthalate (PET) fragments together with synthetic fibres [19,34,35]. In coastal systems, urban stormwater, incompletely retained wastewater effluent, and the use of polyamide (PA), PP, and PE fishing and recreational gear are the principal reported inputs [36,37,38]. Two features of this input mixture are consequential for aquaculture rather than for receiving environments generally. First, the low density of PE and PP favours their retention in the surface layer, where fed fish take their ration, whereas denser polymers tend to settle towards sediment and become available principally to benthic feeders [26]. Second, catchment hydrology governs the loading regime that a facility experiences: rivers dilute and transport particles downstream, whereas lakes and impoundments act as accumulation zones, so facilities drawing from standing water may face a more persistent burden than those on flowing water at comparable upstream plastic loads.

Table 1. Principal pathways of microplastic entry into aquaculture systems.

Pathway	Dominant Polymers Reported	Typical Morphology	Systems Most Exposed
Municipal sewage/WWTP discharge	PE, PP, PS, PVC, PET, PA	Fragments, fibres	Flow-through and open-intake systems
Agricultural runoff	PE	Films, fragments	Inland ponds, lake-fed systems
Urban stormwater runoff	PE, PP	Fragments	Peri-urban coastal cages
Domestic laundering	PA, PN6, PET	Fibres	All systems on receiving waters
Fishing and recreational gear	PA, PP, PE	Fibres, fragments	Coastal cage and pen culture
Atmospheric deposition	PA, PS, PN6	Fibres	All open systems, including inland
On-farm infrastructure	PVC, PET, PE, PP	Fibres, fragments	All plastic-dependent systems
Formulated feed	PE, PP, PET	Fibres	All fed systems

Polymer and morphology entries are those reported in the sources cited in Section 3.

Two further inputs differ in kind rather than degree. Atmospheric deposition delivers fibres and fragments derived from urban activity, tyre wear, and construction over considerable distances, so inland and non-coastal facilities are not insulated from loading by location alone [39]. On-farm infrastructure is the only input generated within the production unit itself: liners, nets, ropes, floats, pipes, and packaging shed fibres and fragments continuously under ultraviolet exposure, mechanical abrasion, and biofouling, and these losses accumulate in the water column and in settled sludge [31,40]. The distinction is operationally important, because infrastructure-derived losses respond directly to material selection and replacement scheduling, whereas catchment- and atmosphere-derived inputs do not.

These pathways are not equivalent in magnitude or, more importantly for aquaculture, in the degree to which they can be managed at farm level. Ambient inputs from source water and atmospheric deposition are determined largely by catchment characteristics and lie outside operational control, whereas particles shed from on-farm infrastructure and those delivered in formulated feed are, in principle, both measurable and controllable. The available evidence does not, however, support a quantitative ranking. Pathway-specific loading rates into aquaculture systems have rarely been measured; sampling, extraction, and polymer-identification protocols differ substantially between studies; and no study identified in this review has partitioned microplastic loading within a single production system among its constituent sources. Polymer signatures offer only partial discrimination, since PE, PP, and PET recur as dominant polymers across aquaculture water, infrastructure, and feed alike [26,40,41]. Resolving the relative contribution of each pathway is therefore a prerequisite for source-directed mitigation and remains an open question.

Owing to their size relative to natural feedstuffs, fish frequently ingest microplastics after mistaking them for prey [27], and this is not restricted to carnivorous or planktivorous species: grass carp (Ctenopharyngodon idella), a herbivorous aquaculture species, has been reported to ingest particles resembling similarly sized aquatic weeds [42]. Ingestion is not necessarily selective, however, since fibres may be taken up passively during ventilation rather than actively targeted, and ingested abundance increases in the presence of food [43]. In fed aquaculture, these ambient routes operate alongside a dietary route that is specific to the production system and that is considered separately below.

3.1. Fishmeal and Formulated Feed as a Dietary Source of Microplastics

Microplastics have been detected in every commercial aquafeed and fishmeal sample examined in independent surveys conducted in Bangladesh, the United Kingdom, and multi-country collections, at reported abundances spanning approximately 124 to 14,270 particles kg⁻¹ (Table 2). Contamination is not confined to marine-derived ingredients: soybean meal and poultry by-products also carry anthropogenic particles [44,45], indicating multiple entry points along the feed supply chain. Fibres dominate the morphotype profile in most compound-feed surveys, typically accounting for 80–90% of recovered particles, although fragment dominance has been reported in some fishmeal datasets, so morphology reflects source material and analytical method rather than any universal pattern [41,46,47,48]. Reported size distributions are concentrated below approximately 1.5 mm, most commonly between 100 and 1500 µm, and PE, PP, and PET recur as the polymers most frequently identified [41,46,49].

Table 2. Reported microplastic abundance in commercial aquafeeds and feed ingredients.

Matrix	Abundance (Particles kg−1)	Region	Reference
Commercial fish feed, all grades	550–11,600	Bangladesh	[50]
Commercial fish feed (starter, grower, finisher)	500–2200	Bangladesh	[46]
Commercial fish feed (starter, nursery, grower)	9230–14,270	Bangladesh	[49]
Commercial fishmeal, ten producing countries	~5500 (mean)	Global	[41]
Aquaculture feeds, six countries, three continents	~1130 (mean)	Global	[47]
Commercial fishmeal, two products	~124	United Kingdom	[51]
Soybean meal	~1230	—	[44]
Fishmeal- and soybean-meal-based formulations	1070–2000	—	[44]

Reported abundances span nearly two orders of magnitude. Because extraction protocols, digestion reagents, minimum detectable particle size, and the extent of polymer confirmation differ between studies, this range reflects methodological heterogeneity as well as genuine variation in contamination, and surveys that do not recover particles in the low-micrometer range will underestimate precisely the size fraction most likely to cross biological barriers. Inter-study comparison should therefore be treated as indicative rather than quantitative.

Contamination appears to arise at two distinct stages of the feed chain. Fishmeal manufactured from wild-caught pelagic fish carries particles already ingested in the marine environment [52,53], whereas concentrations reported in processed fishmeal that exceed those measured in the captured fish themselves indicate augmentation during processing, handling, and packaging [45,51]. Consistent with the latter, anthropogenic particles in plant-derived feedstocks have been attributed to post-harvest contamination from equipment, transport, and airborne deposition rather than to uptake by the plant [44], and fishmeal produced from dried fish has been reported to carry higher loads than that produced from fresh material, plausibly reflecting prolonged open-air exposure during sun-drying and handling [48]. These stages are not mutually exclusive, and their relative contribution to the final feed burden has not been quantified.

Occurrence in feed translates into a calculable exposure estimate, but not directly into a measured body burden. Combining reported feed concentrations with feeding rates, intake over a production cycle has been estimated at the order of 10² particles for an individual tilapia over twenty weeks [50] and 10³ particles for an Atlantic salmon over its commercial lifespan [44], with fishmeal-based diets yielding the largest modelled intakes among cultured taxa assessed globally [41]. These figures are exposure estimates derived from feed concentration and ration size; they assume complete consumption of the ration and uniform particle distribution within it, and they should not be read as measured tissue burdens.

Direct evidence that dietary particles reach the animal is more limited but consistent in direction. In cultured common carp, microplastic abundance in commercial fishmeal was positively related to abundance in the gastrointestinal tract and gills, with overlapping polymer signatures in meal and tissue [53]. In recirculating aquaculture systems, polymer profiles overlap between feed, culture water, and fish, a pattern compatible with dietary transfer but incapable of isolating it from waterborne uptake [54]. Gastrointestinal presence must further be distinguished from tissue accumulation, since particles recovered from fish are found predominantly in the gut and gills rather than in muscle, and retention is strongly size-dependent [55,56].

The relative contribution of this dietary route to total exposure has not, however, been established. Occurrence in feed, a calculable particle intake, and recovery of particles from the gastrointestinal tract together establish the dietary route as real and quantifiable, but they do not establish that it exceeds waterborne, sediment-associated, or trophic exposure. No study identified in this review has partitioned intake among these routes in a cultured fish under production conditions, and the polymer overlap between feed and culture water precludes apportionment by polymer signature alone [54]. The comparison is further limited by uneven evidence: waterborne exposure is well characterised toxicologically but rarely quantified at concentrations measured in production systems; sediment- and detritus-associated exposure remains largely unexamined in cultured species, despite the accumulation of legacy particles in pond and cage sediments [8]; and trophic transfer, although demonstrated from zooplankton to larval fish with associated reductions in growth and elevated mortality [57], is of uncertain relevance in intensive systems where formulated feed displaces most natural prey. Formulated feed is therefore emphasised in this review not as the dominant route, but as the one whose contamination can be measured at source and reduced through ingredient specification, processing hygiene, and packaging control—a consistently documented, quantifiable, and farm-controllable dietary exposure pathway of uncertain relative magnitude.

Biological consequences of dietary exposure are reported inconsistently, and the inconsistency is itself informative. Adverse outcomes span growth, digestive function, and tissue integrity, and include reduced weight gain, poorer feed conversion, depressed digestive enzyme activity, and gut microbiota shifts in juvenile tilapia fed PE [58]; impaired growth, reduced survival, haematological abnormalities, and intestinal lesions in juvenile striped catfish fed PA [59]; hepatic and intestinal alteration with oxidative and immune responses in European sea bass fed PVC or PE [60]; and mortality, DNA damage, and altered swimming behaviour in medaka larvae fed environmental microplastics at low dietary inclusion [61]. Other trials report little or no whole-animal effect, including low gastrointestinal retention and no clear pathology in gilt-head seabream over 45 days [62], minimal hepatic stress and no deterioration of fillet quality in rainbow trout fed comparatively large PS particles [63], and no effect on growth or survival in Atlantic salmon fed realistic dietary PP concentrations for 90 days, with oxidative-stress markers responding only at higher doses [64]. Across these studies, effect strength tends to increase with decreasing particle size, increasing dose, and increasing exposure duration, whereas larger and more inert particles are largely egested; these relationships are examined further in Section 6. Because the trials differ in polymer, particle size, inclusion rate and duration, and because few employ particles matched to those actually recovered from commercial feeds, the magnitude of dietary hazard under production conditions remains poorly constrained.

3.2. Small Pelagic Fish as a Primary Entry Point

Lower-trophic and small pelagic species have been identified as among the taxa at highest risk of microplastic contamination, plausibly owing to filter-feeding and non-selective foraging behaviour [42], and are frequently landed as bycatch and diverted into fishmeal and feed production. Aquafeed manufacture draws heavily on small pelagics, bycatch, and processing by-products [65], with species commonly used including Peruvian anchoveta, blue whiting, lesser sand eel, menhaden, Atlantic herring, mackerel, capelin, and Pacific sardine. Occurrence data from these source species are consistent with this routing: microplastics were detected in 72%, 72%, and 57% of Sardinops sagax, Etrumeus whiteheadi, and Engraulis encrasicolus, respectively, at 1.58, 1.38, and 1.13 items individual⁻¹, with fibres the predominant morphological class [52]. Feeds formulated with a higher proportion of material derived from carnivorous species have been reported to carry elevated plastic content relative to those based on lower-trophic-level material [66]. Considered together with the direct transfer evidence summarised in Section 3.1, the overlap between microplastic-burdened wild stocks and feed-ingredient sourcing identifies fishmeal as a mechanistically plausible and partially demonstrated conduit into farmed species. The proportion of the feed-borne burden attributable to wild-stock origin rather than to processing, handling, and packaging has not, however, been resolved, and partitioning these contributions would materially improve the targeting of mitigation within the feed supply chain.

4. Key Factors Governing MP Uptake in Fish

MPs are distributed across nearly all marine habitats at a size range comparable to sediments and planktonic organisms, rendering them highly bioavailable to aquatic organisms, including fish [67]. Owing to their small size (<5 mm), fish frequently mistake MPs for natural prey items such as plankton, or ingest them incidentally when MPs are already embedded in or adhered to prey organisms. The small size, attractive colouration, and buoyancy of MPs collectively contribute to this misidentification and subsequent ingestion [27]. Uptake of MPs into fish is governed by several interacting biological and ecological factors:

Feeding behaviour is a primary determinant of MP exposure. Predatory species may accumulate MPs indirectly through consumption of MP-contaminated prey, creating potential for bioaccumulation at higher trophic levels [40], whereas filter and deposit feeders are generally considered more directly vulnerable to MP ingestion than predatory species [28]. Notably, Mizraji et al. [68] found that omnivorous intertidal fish ingested more MPs than either herbivorous or carnivorous fish when examining the relationship between feeding habit and MP uptake likelihood.

Habitat type also influences MP burden, with pelagic fish generally exhibiting higher MP ingestion than fish occupying other habitats, irrespective of predatory status [27]. This was substantiated by Phaksopa et al. [69], who reported greater MP ingestion in pelagic species (14.47%) compared to demersal species (12.63%). Additional variability in MP aggregation is attributable to species-specific traits, temporal factors, geographic distance, and exposure conditions [70].

Particle size and shape further determine MP uptake patterns. Myctophid fish, for instance, most commonly ingested MPs within the 1–2.79 mm range, closely matching the size profile of plankton, their principal food source [71]. Similarly, Oliveira et al. [72] detected MPs ranging from 1–3 mm in the stomach contents of neotropical omnivorous fish. Collectively, feeding behaviour, habitat preference, and MP size/shape emerge as the principal determinants of MP accumulation within fishmeal-producing species.

5. Routes of Exposure and Mechanisms of Toxicity

Microplastic (MP) toxicity in fish involves three distinct processes: routes of entry, mechanisms of toxicity, and in vivo fate. Routes of entry describe how MPs reach the fish, mechanisms of toxicity describe how they cause biological effects, and in vivo fate describes their subsequent distribution within the organism. These processes are therefore discussed sequentially below.

5.1. Routes of Entry

5.1.1. Ingestion

Ingestion is the principal route of MP exposure in fish. Particles taken up with food or water enter the gastrointestinal tract, where they may be retained or egested, or interact with the intestinal epithelium [27]. Consequently, ingestion has been associated with intestinal obstruction, tissue injury, histological alterations, disturbed lipid metabolism, and behavioural changes [27].

5.1.2. Branchial and Dermal Uptake

MPs may also contact fish through the gills and skin during exposure to contaminated water. Branchial retention is influenced by gill morphology, filtration mechanisms, and particle size [73]. In wild fish, MPs have been detected in the gastrointestinal tract (35%), gills (36%), and dorsal muscle (32%), with 49% of individuals showing detectable MPs [56]. The occurrence of MPs across multiple tissues indicates that branchial exposure can occur concurrently with ingestion.

5.2. Mechanisms of Toxicity

The routes above describe how MPs enter fish, whereas the following mechanisms describe how they produce biological effects. These mechanisms may act concurrently.

5.2.1. Physical Interaction and Mechanical Injury

MPs can cause physical damage through direct interaction with epithelial surfaces. Hard or angular particles may abrade the gastrointestinal or branchial tissues, while retained particles can contribute to intestinal obstruction [27]. These effects are primarily related to particle size and morphology.

5.2.2. Leaching of Polymer Additives

MPs may release additives incorporated during polymer manufacture following ingestion or branchial exposure. Compounds such as polybrominated diphenyl ethers (PBDEs), nonylphenol, bisphenol A (BPA), and octylphenol have been associated with adverse effects in fish [29]. This mechanism is therefore related primarily to polymer composition and additive characteristics.

5.2.3. Vector Transport of Sorbed Contaminants

MPs can also adsorb contaminants from the surrounding environment before biological uptake. These sorbed compounds may subsequently be delivered to fish together with the particle [30]. MPs may either enhance contaminant exposure by acting as vectors or reduce bioavailability by acting as sinks, resulting in synergistic or antagonistic effects depending on particle and contaminant properties [74]. Environmental contaminant concentrations can further influence this process [75]. These interactions may contribute to oxidative and inflammatory responses in exposed fish [76].

5.3. In Vivo Fate and Translocation

Following entry, the fate of MPs determines their distribution within the fish. MPs may remain at the initial site of exposure, pass through the gastrointestinal tract, or translocate to other tissues. Evidence indicates that MPs can be detected beyond the initial site of entry, including in the liver following intestinal exposure [27,77]. Thus, in vivo fate represents a post-entry process that determines tissue distribution rather than a route of exposure or a mechanism of toxicity. Tissue-specific distribution and accumulation are discussed in Section 6.

6. Bioaccumulation and Biomagnification of Microplastics in Farmed Fish

6.1. Bioaccumulation Pathways and Tissue-Specific Distribution

Microplastics (MPs) present in aquatic environments pose a well-recognised threat to the quality and safety of aquaculture products [78,79]. Farmed fish are primarily exposed to MPs through two principal pathways: ingestion of contaminated prey, such as plankton and benthic organisms, or feed; and direct uptake of suspended particles from the surrounding water column, with a further route of entry occurring via respiratory uptake at the gills [27,80].

Following ingestion, MPs may cross the gastrointestinal epithelium and reach other tissues, although the extent of translocation appears to depend on both particle characteristics and biological factors [27,78]. Particle size is among the more influential of these, yet the available evidence does not support universal size thresholds for intestinal retention or cross-tissue transport. Particles in the lower micrometre range appear to have the greatest potential to cross the intestinal barrier and enter systemic circulation, but the size range over which translocation has been observed varies among species and may also reflect particle shape and chemistry [81]. On this basis, larger particles—typically those above approximately 100 µm—have more often been reported to remain within the intestinal lumen and to be egested, whereas smaller particles have more frequently been associated with epithelial translocation; these represent tendencies observed across studies rather than definitive biological cut-offs [27,79,80]. Ribeiro et al. [81] further noted that ingested particles may be modified in vivo through gastrointestinal processes such as maceration and digestion, potentially altering their physical characteristics and, in turn, their capacity for translocation, although this possibility remains insufficiently investigated. Translocated MPs have been recovered from organs including the liver, kidney, muscle, and brain [27,79,80], indicating that gastrointestinal retention and systemic distribution are not mutually exclusive processes. Su et al. [55] reported MPs in the intestine of 22–100% and the gills of 22–89% of fish examined, with gill-associated particles generally smaller than those recovered from the gut; these data describe tissue-specific occurrence rather than epithelial transfer, and reflect differences in the site and mechanism of retention.

Susceptibility to bioaccumulation varies considerably across life stages, with embryos, larvae, and juveniles showing particular vulnerability owing to their small body size, elevated metabolic rates, underdeveloped physiological barriers, and immature detoxification mechanisms, all of which facilitate rapid particle uptake, tissue distribution, and amplified toxicological risk from both the particles themselves and their adsorbed contaminants [82,83]. Such early-life exposure has been linked to translocation into vital organs, including the gills, liver, and muscle, resulting in chronic effects such as developmental abnormalities, reduced organismal fitness, lipid peroxidation, neurotoxicity, stunted growth, and diminished survival [82]. Emerging evidence further points to the possibility of transgenerational transfer, with microplastics shown to translocate from maternal tissues to offspring in sea turtles [84], and nanoplastics observed to penetrate the chorionic pores of fish embryos, directly affecting early developmental processes [85].

Collectively, the bioaccumulation behaviour of MPs in farmed fish reflects the interaction of species-specific traits, such as trophic level and feeding ecology, individual behavioural patterns, life stage, and the physicochemical properties of the particles themselves, including size, polymer composition, and ambient environmental concentration [78,86]. Tissue-specific retention patterns, shaped by particle size, shape, and polymer type, therefore underscore varying degrees of physiological disruption and ecological risk across species and developmental stages.

6.2. Biomagnification Across Aquatic Trophic Levels

Distinct from bioaccumulation, which describes particle retention within a single organism, biomagnification refers to the progressive increase in contaminant concentration as it moves up successive trophic levels of a food web [87,88]. Given the ubiquitous distribution of MPs across aquatic ecosystems and their consumption by a wide range of organisms, concern has grown regarding their potential translocation through food webs, particularly along piscine food chains [87,88]. However, experimental evidence on this question remains heterogeneous, with some studies reporting trophic transfer and others finding negligible or no biomagnification, depending on species and ecosystem context [89]. The conceptual distinction between these two processes—progressive retention within an individual over its life history, and progressive concentration across successive trophic levels—is represented schematically in Figure 2.

Figure 2. Conceptual distinction between bioaccumulation and biomagnification of microplastics. Left: bioaccumulation, the progressive increase in particle burden within a single individual with time and growth (fingerling → juvenile → adult) following uptake across the gills, the body surface, and the intestinal lumen. Right: biomagnification, the postulated increase in microplastic concentration across successive trophic levels, from phytoplankton and zooplankton through small forage fishes to higher predators. Principal sources entering the culture environment—industrial waste, aquaculture infrastructure, formulated feed, and fishing nets and ropes—are indicated at the top. Biomagnification is shown as a conceptual pathway; as discussed in Section 6, it is inconsistently observed in field studies, with several food webs instead exhibiting trophic dilution.

Evidence from marine systems illustrates this inconsistency. In the Persian Gulf, Akhbarizadeh et al. [90] reported that Penaeus semisulcatus exhibited the highest mean muscle MP concentration among assessed species (0.360 items/g), while Epinephelus coioides showed the lowest (0.158 items/g), with MP burden generally higher in gill tissue than muscle, particularly among filter feeders and scavengers such as Liza klunzingeri, Portunus armatus, and P. semisulcatus. Despite these tissue-level differences, calculated trophic magnification factor (TMF) and biomagnification factor (BMF) values indicated no significant biomagnification of MPs within the edible tissues of this food web. By contrast, biomagnification has been documented in several higher-trophic marine predators: in the Mediterranean Sea, elevated MP concentrations consistent with biomagnification have been reported in swordfish (Xiphias gladius), albacore tuna (Thunnus alalunga), and Atlantic bluefin tuna (Thunnus thynnus), while trophic transfer of MPs from Atlantic mackerel (Scomber scombrus) to grey seal (Halichoerus grypus) has also been demonstrated [91,92]. Evidence from freshwater systems remains comparatively limited. Bhatt et al. [93] examined MP burden across feeding guilds and reported a mean of 6.09 and 5.85 particles in carnivorous and omnivorous fish, respectively, compared with 1.88 in herbivorous fish, indicating a general upward trend in MP load from herbivorous to omnivorous and carnivorous species, although the authors note that differences in geography and diet composition across studies limit direct comparison.

Where biomagnification does not occur, this is generally attributed to trophic dilution, whereby rapid excretion, poor tissue assimilation, and selective feeding behaviour reduce MP concentrations at higher trophic levels [90,94]. Taken together, systematic reviews indicate that while MPs are consumed across a broad range of taxa and can be transferred to predator species, a consistent, progressive increase in concentration across trophic levels is not a general occurrence; rather, current evidence more strongly supports bioaccumulation and trophic transfer as the dominant patterns, with true biomagnification occurring only under specific ecological conditions [35,93]. Research in this area remains constrained by methodological inconsistencies, sampling biases, and limited geographic scope; addressing these gaps will require standardised assessment protocols, broader ecosystem-level monitoring, and integrative studies linking bioaccumulation and biomagnification data across trophic levels.

7. Impact of Microplastics on Fish

7.1. Gill-Mediated Effects

The gills represent an additional entry and injury site for MPs [95]. MPs trapped within gill filaments can impair respiratory function and induce hypoxia [96], and may further compromise gill integrity through filament rupture, leading to infection, respiratory distress, and mortality [56,97].

7.2. Gastrointestinal Tract

Following ingestion, a portion of the MPs is retained within the gastrointestinal (GI) tract, while the remainder is presumably egested. Retained MPs can induce false satiation and consequently reduce food intake [56], trigger digestive blockages [67], and cause structural and functional injury, with downstream consequences for nutrition and growth [97,98]. Certain MPs with hard, pointed edges cause direct mechanical damage and ulceration of the intestinal lining. Mbugani et al. [99] demonstrated that dose-dependent MP exposure significantly altered small intestinal morphology in Oreochromis urolepis larvae, including changes in villus height-to-width ratio, epithelial cell height, and the functional characteristics of epithelial, goblet, and crypt cells, alongside leucocytic infiltration and vascular congestion. Similarly, Pedà et al. [98] reported histopathological alterations in the distal intestine of MP-exposed fish, including lamina propria expansion, separation of the mucosal epithelium from the lamina propria, villus contraction and swelling, vacuolation of absorptive cells, goblet cell proliferation, and serosal structural loss. In zebrafish specifically, polystyrene (PS) exposure has been shown to rupture intestinal villi [100], while exposure to PA, PE, PP, and PVC MPs produced villus dehiscence and enterocyte lysis [101]. Reduced food intake following PE ingestion in goby fish has similarly been attributed to digestive tract obstruction [102].

7.3. Systemic Organ Pathology

Beyond the GI tract, comprehensive histological assessment has revealed MP-induced damage across multiple organ systems [103]. In the kidney, this includes capillary enlargement, inflammatory cell infiltration, basophilic cytoplasm depletion, tubular vacuolation, convoluted tubule constriction, intertubular space expansion, glomerular atrophy and vacuolation, and tubaritis. Hepatic lesions include vacuolar hydropic degeneration, necrotic foci, pyknotic nuclei, hepatocyte deformation, and dilation or obstruction of blood sinusoids. Pancreatic tissue exhibits compressed and deteriorated acini with pyknotic nuclei, haemorrhagic and necrotic areas, inflammatory infiltrate, fatty tissue deposition, and vascular engorgement. Muscle tissue shows fibre segmentation, oedema, necrosis, and inflammatory infiltration, while gill tissue exhibits vascular dilation and congestion, lamellar fusion, epithelial lifting, secondary lamellar degeneration, hyperplasia, and interlamellar MP deposition. Spinal column effects include meningeal breakdown, notochord deviation and degeneration, and spinal cord displacement. Because these lesions arise concurrently in tissues with very different functions, the organ-level distribution of microplastic-induced effects, together with the behavioural and reproductive endpoints considered later in this section, is summarised in Figure 3.

Figure 3. Physiological cascade of microplastic toxicity in fish, summarising organ-level and systemic responses reported across the reviewed literature: gill (swelling, hyperplasia, lipid peroxidation, impaired gas exchange, filament rupture, and respiratory distress); brain (acetylcholinesterase inhibition, oxidative DNA damage, reactive oxygen species-mediated apoptosis, neurological dysfunction, and behavioural change); muscle (lipid peroxidation, fibre segmentation, oedema, necrosis, and inflammatory infiltration); liver (hepatocyte oedema, lipid accumulation, vacuolisation, glycogen depletion, single-cell necrosis, and oxidative stress); intestine (villus damage and increased mucosal permeability); gonads (delayed sexual maturity, reduced fecundity, altered gonadosomatic index, impaired gamete quality, and delayed hatching); and behaviour (hyperactivity, aggression, altered ingestion, olfactory-driven responses, and modified swimming activity). Endpoints are compiled from studies employing diverse polymer types, particle sizes, exposure routes, and concentrations; they do not all occur in any single species or exposure scenario.

7.4. Gastrointestinal Transfer and Hepatic Translocation

Although gastrointestinal MP retention is often transient, particles are frequently translocated to the liver [40]. MP-affected livers commonly display glycogen deficiency, lipid starvation, and single-cell necrosis [104]. Hepatotoxicity may result from the intestinal agglomeration of smaller MPs followed by endocytic or phagocytic uptake [73]. In a zebrafish model, seven days of PS exposure induced hepatic inflammation, oxidative stress, and disturbances in lipid and energy metabolism, including lipid accumulation and significant alterations in triglycerides and fatty acid profiles (monounsaturated fatty acids, linoleic acid, and related fatty acyl chains). This exposure also disrupted phospholipid synthesis and transport and impaired lipid metabolism via altered choline, phosphorylcholine, and cholesterol levels [100], while separately reducing branched-chain amino acids (BCAAs), which are important for fatty acid metabolism regulation [105].

7.5. Reduced Feed Utilisation Efficiency

Feed utilisation efficiency is a key determinant of both aquaculture profitability and long-term production sustainability. A growing body of evidence suggests that microplastics (MPs) compromise this efficiency by interfering with feeding behaviour, digestive processes, and metabolic function [16,106,107].

7.6. Locomotor Effects

Locomotor performance represents a further, mechanistically distinct axis of MP toxicity with direct consequences for feeding success. Swimming activity is closely tied to feed intake, as fish that swim more actively and sustain higher activity levels generally consume more feed [106,108]; accordingly, any MP-induced decline in swimming capacity would be expected to propagate into reduced feed acquisition and, ultimately, impaired growth. Consistent with this link, African catfish (Clarias gariepinus) exposed to polyethylene microplastics (PE-MPs) at 500–2000 mg/L showed significant declines in swimming speed, distance covered, and movement complexity relative to unexposed controls [107]. Locomotor and foraging capacity are similarly affected in black rockfish (Sebastes schlegelii), which exhibited reduced feeding and swimming ability under microplastic exposure [109]; a related study using fine polystyrene microplastics (PS-MPs, 0.015 mm) in the same species reported that exposure additionally triggered shoaling behaviour and reduced both swimming speed and range of activity, ultimately limiting foraging and exploratory behaviour [110]. Polyethylene exposure has likewise been shown to impair swimming performance in the common goby, compounding the predatory deficits noted above [111]. Collectively, these findings demonstrate that microplastics consistently impair swimming performance across fish species by reducing swimming speed, distance travelled, movement complexity, and behavioural range. These alterations diminish feed intake and feed utilisation efficiency, providing a mechanistic basis for the growth impairments.

7.7. Behavioural Effects

Growth impairment following MP exposure is frequently preceded by, or accompanied by, disruptions in feeding-related behaviour, indicating that behavioural change is an early and sensitive indicator of MP toxicity rather than merely a downstream consequence of it. Retained particles in the gut can generate a false sense of satiety that lowers voluntary feed intake, or, in more severe cases, causes outright gastrointestinal blockage that physically prevents normal feeding [112]. Chronic, long-term exposure produces comparable outcomes through slower-acting pathways, blunting the food-seeking response and lowering feeding rate in goldfish [16]. Sensory function underlying these feeding behaviours is also compromised. Microplastics have been shown to interfere with odorant detection in goldfish through several overlapping mechanistic pathways, thereby disrupting olfactory-guided behaviours such as foraging and predator avoidance [113]. Impairment of chemosensory and foraging cues in turn has consequences for higher-order behaviours such as predation. Juvenile common goby exposed to MPs display altered predatory feeding behaviour, raising the possibility of knock-on effects for trophic interactions and food-web dynamics [102], while polyethylene exposure has separately been shown to impair predatory success in the same species [111]. More broadly, MP exposure has been linked to disruption of both courtship and feeding behaviours across multiple species, with implications for ecological stability and food-chain integrity at the community level [114]. Collectively, these findings indicate that microplastics disrupt multiple behavioural domains, including feeding, sensory perception, predator avoidance, and social interactions, with cascading effects on trophic dynamics and ecosystem functioning.

7.8. Neurotoxic Effects

A key downstream consequence of MP-induced oxidative stress is lipid peroxidation of cell membranes, which has been shown to damage the gills, muscle, and brain [115]. In agreement with these observations, exposure to larger microplastic particles (60 μm) produced direct injury to the branchial epithelium of the African catfish, Clarias gariepinus [116]. In muscle tissue, this peroxidative damage impairs cellular energy production and neuromuscular coordination, manifesting as energy deficiency, poor movement coordination, and reduced swimming performance.

In the brain, much of this toxicity converges on acetylcholinesterase (AChE), the enzyme responsible for terminating cholinergic neurotransmission by inactivating acetylcholine at the neuromuscular junction and synapse [117]; because of this central role, AChE activity is widely used as a biomarker of neurotoxicity in ecotoxicological studies [118]. Findings on the direction of AChE disruption are not entirely consistent; Barboza et al. [56] found that MP-induced oxidative stress elevated AChE activity in fish brain, producing disorientation, elevated energy demand, and visual impairment outcomes likely to compromise individual fitness and increase susceptibility to disease. In contrast, Oliveira et al. [111] reported that pyrene-associated MPs impaired neuromuscular function through AChE inhibition rather than overactivation, while simultaneously suppressing intracellular energy metabolism via inhibition of NADP⁺-dependent isocitrate dehydrogenase (IDH). This apparent discrepancy likely reflects differences in plastic type, associated chemical additives, and exposure regime, and underscores that AChE dysregulation in either direction is a sensitive marker of neurotoxic insult.

The severity, and even the site, of neurological injury also appears to depend strongly on particle size. Nanoscale microplastics (24–27 nm) impair brain development in fish, producing developmental delay and structural damage [119,120] alongside reduced AChE activity [121], with particles of this size able to lodge directly within blood vessels and tissue surrounding the brain [122]. Larger particles, by contrast, tend to accumulate in the conjunctiva and perivascular tissue adjacent to the brain, where they exert a comparable, if more indirect, effect on neurological function [122]. Particles exceeding 500 nm appear to act through a still more indirect route: rather than penetrating neural tissue, they provoke immune responses and disrupt metabolism and intestinal microbiota, with these systemic disturbances thought to contribute secondarily to neurological impairment [123].

Despite this variability, inhibition of AChE remains the predominant outcome reported in the literature, having been recorded in European seabass [121], Nile tilapia (Oreochromis niloticus) following a 14-day exposure at 1, 10, and 100 μg/L [70], common carp [124], zebrafish larvae [125], juvenile common goby (Pomatoschistus microps) [126], African catfish [117], the Amazonian cichlid Symphysodon aequifasciatus [127], and goldfish [128]. The occurrence of AChE inhibition across such a wide range of fish species suggests that this response is not restricted to a particular taxon and may represent a broader neurotoxic effect. Mechanistically, this inhibition allows acetylcholine to accumulate at the synapse [70], disrupting neurotransmission and producing motor dysfunction such as muscle weakness, tremors, and impaired coordination [123].

This neurochemical disruption is mirrored at the transcriptional level. Prolonged PS exposure downregulates genes involved in glucose, lipid, and amino acid metabolism in zebrafish [129] and alters expression of both neurotransmitter-related genes (ACh, AChE, ChAT, ChE) and oxidative stress-related genes (SOD, CAT, GST) [130]. Polyethylene exposure produces a comparable phenotype of abnormal lipid metabolism and impaired neurotransmitter transmission [111]. At the subcellular level, PS nanoplastics accumulate directly within the zebrafish brain, where they induce oxidative DNA damage and apoptosis through ROS generation [131], consistent with broader evidence linking MP-induced ROS to elevated malformation rates in zebrafish embryos and juveniles [132]. Beyond AChE, MNP exposure has also been implicated in the broader suppression of several neurotransmitters and neurohormones, including dopamine, γ-aminobutyric acid, melatonin, vasopressin, serotonin, oxytocin, and kisspeptin, pointing to a wide neuroendocrine footprint of exposure.

These molecular and biochemical disruptions translate into measurable behavioural change. Exposure to 51 nm fluorescent PS nanoplastics induces oxidative stress alongside bradycardia and hyperactivity in zebrafish, impairing normal behavioural adaptation [133], while early-life PS-MP exposure produces hyperactivity that persists into adulthood, linked to altered expression of neurological (nrbf2, slc6a4b, npy) and oxidative stress-related (nrf2a, sod2) genes [134] directly connecting behavioural outcomes back to the oxidative and neurogenetic pathways described above. The severity of these effects is strongly dependent on exposure duration and concentration: brief exposure (1–10 mg/L for five days) in embryo-to-larval zebrafish produces only minor changes in swimming activity, whereas prolonged dietary exposure (1% w/w for five months) in marine medaka and zebrafish reduces growth and reproduction, causes mild premature mortality, and increases swimming activity [135], illustrating that chronic, low-dose exposure regimes may pose a distinct, and potentially greater, risk than short-term, high-dose exposure.

As summarised in Table 3, behavioural and neurophysiological endpoints respond to microplastic exposure across a broad range of species, polymer types, and exposure regimes, but the direction of response is not uniform. Effects have been reported in salmonids, cyprinids, cichlids, and medaka following both waterborne and dietary exposure, at nominal concentrations spanning three orders of magnitude and over durations from 7 to 60 days [136,137,138,139,140,141,142]. The consistent finding is that behaviour is altered, rather than that it is altered in a particular direction: hyperactivity and increased aggression are reported in some studies [136,138], whereas anxiety-like responses, reduced shoaling, and diminished foraging predominate in others [137,139,142]. Where mechanisms have been examined, they converge on cholinergic signalling, brain monoamine levels, and the brain–gut microbial axis [138,142], with nanoscale particles additionally reported to cross the blood–brain barrier [139]. Particle size in these studies ranges from the nanoscale to the upper microplastic boundary, and exposure conditions are incompletely reported in some cases, so size- and route-dependence cannot be separated. Collectively, these findings identify behaviour as an early and sensitive integrative endpoint, whose diagnostic value is presently constrained by the absence of standardised behavioural assays.

Table 3. Behavioural and neurophysiological responses to microplastic exposure in fish.

Behavioural and Neurological Responses
Fish Species	MP Polymer/Types	Exposure Conditions	Major Effects	References
Oncorhynchus mykiss	PS	52.5 ± 11.5 µm; 0.5, 2, 5% of diet, 6 wk	Reduced feeding behaviour; individuals more aggressive and active; linked to skin injury, serum ammonia changes, oxidative stress and gill/liver lesions.	[136]
Danio rerio	Naturally weathered MPs	5 mm; 0.1 and 1 mg/L, 21 d	Anxiety-like behaviour (altered zone preference); enhanced brain catalase but suppressed lactate dehydrogenase at 1 mg/L; reduced mitochondrial complex II/IV activity.	[137]
Danio rerio	PE (foodborne + waterborne)	40–47 µm water (0.1–10 mg/L); food 0.01–0.1 µg or particles/mg, 7 d	Enhanced hyperactivity associated with cholinergic system disruption and altered brain dopamine and serotonin (5-HT) levels.	[138]
Carassius carassius	Nanoplastics (not reported)	Not reported	Nanoparticles penetrated the blood–brain barrier, altering foraging behaviour.	[139]
Chindongo demasoni	PET	5 mm; 30 d	Altered swimming behaviour.	[140]
Carassius gibelio; Carassius carassius	PE + tyre-wear particles	0.1%, 60 d	Altered swimming behaviour.	[141]
Oryzias latipes	PS	2 µm; 0.1 mg/L/2.5 × 107 particles/L, 42 d	Reduced shoaling behaviour and social behaviour disorders, associated with brain–gut bacterial system alteration.	[142]
Danio rerio (larvae)	naturally weathered MPs	Not reported	Neurotoxicity associated with oxidative stress and disrupted neurotransmission.	[130]

7.9. Growth Effects

Across teleost species, microplastic (MP) exposure has been consistently linked to impaired growth performance, although the severity of this impairment is far from uniform and varies with species identity, polymer chemistry, particle size, exposure dose, and duration [143]. At a mechanistic level, ingested particles damage the gut epithelium, physically obstruct the intestinal lumen, alter feeding patterns, and interfere with nutrient absorption—a combination of effects that collectively suppresses somatic growth [143]. Because particles resist digestion, they also impose direct strain on the digestive system, reducing overall enzymatic activity while paradoxically elevating trypsin and chymotrypsin levels, a dysregulation that further compromises nutrient assimilation and, by extension, growth efficiency [144].

These physiological disruptions translate into measurable losses in energy balance and somatic growth. Sebastes schlegelii exposed to microplastics showed markedly reduced energy reserves [109], while polyethylene terephthalate (PET) particles raised metabolic energy demand and slowed growth in edible mussels [145]. In Manila clams, accumulated polystyrene particles lowered respiration, excretion, feeding, and absorption efficiency in tandem, ultimately constraining the energy budget available for growth [146]. A comparable pattern was reported in loach juveniles, in which polystyrene exposure significantly depressed both weight gain and specific growth rate [36].

Dose and exposure duration emerge as consistent modulators of severity across taxa. Leuresthes tenuis exposed for 14 days to LDPE and HDPE fragments (75–250 µm) at concentrations exceeding 1 × 10³ particles/mL exhibited pronounced growth suppression driven by both direct ingestion and trophic transfer, a finding echoed in Ctenopharyngodon idella [143,147]. Longer exposure windows appear to compound this damage: a 95-day trial exposing Ambassis dussumieri to HDPE, PVC, and PS produced severe growth impairment, consistent with cumulative, chronicity-driven toxicity [148]. In line with this dose–duration relationship, PVC particles (100–200 µm) at concentrations up to 136.65 µg/L significantly reduced weight gain in Cyprinus carpio at both 30 and 60 days of exposure [149]. At the tissue level, dietary PE-MP exposure in Oreochromis urolepis caused structural intestinal damage that impaired digestion and nutrient assimilation, manifesting as measurable growth deficits [99]. Reductions in weight gain and specific growth rate (SGR) have similarly been documented in Neogobius melanostomus [150] and Oreochromis niloticus [58,151], with the latter study additionally reporting that dietary MP inclusion of up to 10% elevated the feed conversion ratio (FCR), indicating poorer feed utilisation efficiency.

Not every study, however, reports growth impairment following MP exposure. Clarias gariepinus and Symphysodon aequifasciatus showed no significant change in growth performance, pointing to possible species- or life-stage-specific resilience [127,152], and a 150-day exposure of Oryzias latipes to polystyrene particles likewise produced no detectable effect on growth [153]. Growth suppression can also be delayed rather than immediate: Cyprinus carpio maintained stable body weight during the initial phase of polystyrene ingestion, with significant reductions emerging only after particle excretion, suggesting a latent, chronic mode of toxicity [154]. Taken together, and notwithstanding these exceptions, the weight of evidence indicates that microplastic exposure disrupts normal fish growth, acting primarily through mechanical intestinal injury, reduced feeding and digestive efficiency, lowered specific growth rate, gut microbiota disturbance, and broader systemic metabolic stress.

The studies compiled in Table 4 indicate that microplastic exposure impairs growth and feed utilisation in fish, but they also show that this impairment is registered principally through its proximate determinants rather than through direct measurement of somatic growth. Reduced growth and weakened feed utilisation have been reported in Nile tilapia after 63 days of dietary polyethylene exposure and in swordtail following acute polystyrene exposure, in both cases accompanied by suppressed digestive enzyme activity, altered energy metabolism, and reduced antioxidant capacity [58,155], while short-term exposure of African catfish to low-density polyethylene reduced swimming speed, an endpoint with indirect consequences for feed acquisition [107]. The mechanistic thread common to these studies is energetic rather than directly cytotoxic: particles compromise nutrient acquisition and assimilation, and the resulting energy deficit is expressed as reduced growth. The evidence summarised here is nonetheless limited in extent, dominated by exposures far shorter than a production cycle, and reports feed conversion only qualitatively, so the quantitative relationship between dietary microplastic load and production performance remains unresolved.

Table 4. Growth performance and feed utilisation responses to microplastic exposure in fish.

Growth and Feed Utilisation
Fish Species	MP Polymer/Types	Exposure Conditions	Major Effects	References
Oreochromis niloticus	PE	27 & 63 µm; dietary, 63 d	Intestinal accumulation of fine particles, with suppressed digestive enzyme activity, impaired growth, and metabolic disturbances.	[58]
Xiphophorus helleri (swordtail)	PS	1 µm; acute exposure	Decreased antioxidant function, immunity, energy metabolism, and growth performance; weakened feed utilisation.	[155]
Clarias gariepinus	LDPE	50–500 µm; 0.5–2.0 g/L, 4 d	Reduced swimming speed, indicating metabolic and neuromuscular disturbances affecting locomotor activity.	[107]

7.10. Effects on Endocrine and Reproductive Function

Reproductive activity in fish, as in all vertebrates, is governed by a complex neuroendocrine system involving the hypothalamus–pituitary–gonad (HPG) and hypothalamus–pituitary–renal (HPR) axes [156]. Within this system, HPG–liver (HPG-L) signalling coordinates gonad development, gametogenesis, vitellogenesis, steroidogenesis, and, in some species, gonadal sex change [157]. Briefly, hypothalamic GnRH stimulates pituitary release of FSH and LH, which in turn drive gonadal production of the dominant sex steroids 17β-oestradiol and 11-ketotestosterone in fish via receptor-mediated signalling [158].

Microplastics (MPs) disrupt this axis primarily by acting as vectors for endocrine-disrupting chemicals (EDCs), either adsorbed from the surrounding environment or incorporated as manufacturing additives. These compounds can competitively bind sex-steroid hormone receptors, interfering with normal reproductive signalling [159,160]. A well-established biomarker of this disruption is the upregulation of vitellogenin (Vtg) in male fish: EDCs can either mimic oestrogen directly or promote the conversion of testosterone to oestradiol via the CYP19A1 aromatase enzyme, driving Vtg induction and feminisation of males [161].

This mechanism has been demonstrated across several species. Dietary PE exposure in Japanese medaka (Oryzias latipes) over two months impaired endocrine function and caused abnormal proliferation of spermatogonia and spermatocytes, alongside broader shifts in gene expression [162]. A separate 60-day dietary PE exposure in the same species downregulated transcript levels of vitellogenin and choriogenin precursor proteins for egg yolk and egg envelope formation, respectively, again pointing to disrupted reproductive signalling [104]. Exposure to microfibres similarly altered HPG axis-related gene expression and promoted vitellogenesis in male medaka, indicating endocrine disruption and irregular maturation [163]. In African catfish, brief (96-h) exposure to polyethylene MPs (<60 μm) altered transcription of reproductive-axis genes [116], while longer dietary exposure (500 mg/kg for 15 days) caused testicular damage, reduced sperm quality and viability, and suppressed hormone profiles [164]. PE-MP exposure in male tilapia at 10 mg/L likewise triggered hormonal disruption, testicular damage, and induction of testis-ova [165]. In marine medaka, PE-MPs interfered with HPG axis gene expression, producing compromised seminiferous lobule structure, reduced seminal fluid in males, fewer mature spawning follicles, and an increased proportion of early vitellogenic oocytes in females [166]. Gonadal growth is affected in invertebrates, too: in whiteleg shrimp, environmentally relevant MP doses alone or in combination with bisphenol A (BPA) inhibited gonadal development, likely through combined disruption of endocrine and metabolic regulation [167].

Beyond gene- and tissue-level effects, MP exposure can impair reproductive performance, with reported effects including altered maturation and spawning, reduced fecundity, gonadosomatic index, fertilisation rate, and gamete quality [168]. Fertilisation success has been shown to decline to 56–58% following PS-MP exposure [169], consistent with evidence from broadcast-spawning bivalves showing impaired gamete interaction following MP exposure [170]. Similarly, PS-MP exposure adversely affected energy allocation and reproductive health in oysters [171], while BPA leached from MPs disrupted gonadal and reproductive organ development in carp [172]. In male medaka, MP exposure increased mortality and reduced body length, courtship displays, sperm bundles, and sexual activity [173]. Developmental effects have also been reported, including delayed hatching in fish [174]. Collectively, these findings indicate that MP exposure can disrupt fish reproduction across multiple stages, from gamete function and reproductive behaviour to gonadal development and early-life development.

Oxidative stress appears to be a key mechanism underlying these reproductive impacts, given its well-established role in ovarian and testicular dysfunction through disruption of the balance between reactive oxygen species (ROS) and antioxidant capacity [175]. In zebrafish, 21-day exposure to 100 µg/L PS-MPs thinned the testis basement membrane via elevated ROS production [176]. A related sub-chronic exposure trial using 1 µm PS particles found that ROS generation increased in both ovary and testis at 100 and 1000 µg/L, with spermatocyte apoptosis rising at the higher concentration; this was accompanied by elevated expression of pro-apoptotic markers (p53, Bax, and multiple caspases) alongside further thinning of the basement membrane [176]. A longer, 35-day PS-MP exposure in zebrafish similarly altered metabolic processes and gene regulation, with the effects again attributed to ROS generation [125]. These reproductive effects are not necessarily confined to the exposed generation. Toxic effects can be transmitted to offspring via germ cells, producing intergenerational, multigenerational, or transgenerational impacts [177], including increased embryonic mortality, delayed hatching, malformations, and developmental or behavioural abnormalities [178]. Parental exposure of marine medaka to PS-MPs, for instance, was associated with elevated heart rate, premature hatching, and reduced growth in F1 offspring, linked to disrupted sex steroid biosynthesis and cytochrome P450 signalling [179]. However, evidence on this point is not fully consistent: some studies suggest that environmentally realistic levels of virgin MPs induce physiological stress in parental fish without necessarily affecting the reproductive output of their offspring [180], and this remains a notable gap in the transgenerational toxicology literature.

As shown in Table 5, reproductive and developmental endpoints are among the most consistently affected of those reviewed, and the reported effects resolve into three distinct levels of biological organisation. Disruption of the hypothalamic–pituitary–gonadal axis is evident transcriptionally as altered gonadotropin-releasing hormone, luteinising hormone receptor, vitellogenin, and steroidogenic gene expression [163,181]; histologically as gonadal degeneration, oocyte vacuolation, testicular necrosis, and testis–ova formation [182,183]; and functionally as reduced circulating sex steroids, impaired sperm quality, and diminished reproductive output [164,184]. This convergence holds across polyethylene, polystyrene, polyvinyl chloride, and microfibre exposures and across cyprinids, cichlids, catfish, and small model species, indicating a response that is neither polymer- nor taxon-specific. Gonadal oxidative injury is the mechanism most frequently invoked and is supported by direct measurement of reactive oxygen species and apoptosis within the gonad [176,184]. Notably, effects are reported at concentrations as low as 10–100 µg/L [176,181], well below those eliciting growth or haematological responses, identifying reproduction as a comparatively sensitive endpoint of direct relevance to broodstock and hatchery performance.

Table 5. Reproductive and developmental effects of microplastic exposure in fish.

Reproductive Effects
Fish Species	MP Polymer/Types	Exposure Conditions	Major Effects	References
Paramisgurnus dabryanus	PE	8–15 µm; 1 and 10 mg/L, 15 & 30 d	Disrupted gonadal antioxidant defence, pathological lesions and apoptosis, with impaired semen quality and reduced F0 reproductive performance.	[184]
Oreochromis niloticus	PE	100 µm; 10 mg/L, 15 d	Reduced serum LH and testosterone (T), accompanied by degenerative testicular alterations and testis–ova formation.	[182]
Clarias gariepinus	PE	100 µm; 500 mg/kg, 15 d	Reduced serum LH, FSH, and sex steroid levels (T, E), accompanied by decreased sperm count, spermatocrit, viability, and motility, with degenerative testicular alterations.	[164]
Oryzias latipes	PS microfibres	157.68 µm; 500 and 1000 microfibres/L, 21 d	Upregulation of vitellogenin (VTG) mRNA, disruption of GnRH and luteinising hormone (LH) receptor mRNA expression, and increased GnIH mRNA levels.	[163]
Cyprinus carpio (larval)	PVC	100–200 µm; 50, 100, 150 mg/L, 60 d	Oocyte vacuolation and necrosis, along with testicular necrosis, cavity formation, reduced sperm cell abundance, and decreased Sertoli cell density, accompanied by alterations in HPG-axis, apoptotic, and sex-differentiation gene expression.	[183]
Danio rerio	PS	1 µm; 100 µg/L, 60 d	Altered hypothalamic–pituitary–gonadal (HPG) axis gene expression, characterised by the upregulation of cyp19a1b, 3βhsd, and 17βhsd, alongside gnrh3 downregulation in females and the downregulation of multiple reproductive genes in males.	[181]
Danio rerio	PS	1 µm; 10 and 100 µg/L, 21 d	Elevated ROS levels in both male and female gonads, increased testicular apoptosis accompanied by reduced basement membrane thickness, and histopathological alterations in both sexes.	[176]

7.11. Digestive and Enzymatic Effects

A substantial body of work confirms that MPs accumulate within the fish gut, where they can cause physical obstruction, suppress digestive enzyme secretion, and impair digestive function more broadly [78,185]. In large yellow croaker (Larimichthys crocea), a 14-day exposure to PS-MPs (488–518 nm; 10⁴–10⁶ particles/L) significantly reduced intestinal lipase and trypsin activity, although α-amylase activity was unaffected [185]. A comparable enzymatic effect was reported in blue mussels (Mytilus galloprovincialis), where exposure to 0.01 mm PS-MPs significantly lowered amylase activity [186]. At the metabolite level, zebrafish exposed to PS-MPs showed disrupted amino acid metabolism, with significant shifts in intestinal leucine, proline, threonine, lysine, glutamine, phenylalanine, alanine, and tyrosine [187]. In black rockfish, PS-MP exposure additionally produced structural abnormalities in the bile, liver, and intestinal lumen, together with a significant rise in ammonia excretion—evidence of physiological stress during nutrient assimilation [110].

The evidence summarised in Table 6 identifies the intestine as an early and convergent target of microplastic exposure, with injury manifesting simultaneously as structural damage, oxidative imbalance, and microbial dysbiosis rather than as any one of these in isolation [147,188,189]. Effects are reported at concentrations from 100 µg/L upward and after exposures as short as 48 h, indicating rapid onset, and the accompanying shifts in digestion-related transcript abundance [188] provide a plausible link between intestinal injury and the impaired nutrient utilisation considered in Section 7.5. Two constraints qualify this evidence. All three studies employed polystyrene, so polymer-dependence cannot be assessed from this dataset; and one used co-exposure with oxytetracycline [189], in which the contribution attributable to microplastics alone cannot be isolated. The emerging issue represented by this table is therefore the coupling of mechanical and microbial injury within the gut, together with the need for exposure studies employing the polymer types actually recovered from aquaculture feeds and culture waters.

Table 6. Digestive and intestinal effects of microplastic exposure in fish.

Digestive and Intestinal Effects
Fish Species	MP Polymer/Types	Exposure Conditions	Major Effects	References
Ctenopharyngodon idella	PS	0.5 and 15 µm; 100 and 500 µg/L	Induced oxidative stress and liver congestion; altered gut microbiota and severe intestinal damage.	[147]
Common carp (Jian), grass carp, largemouth bass	PS	8 µm; 0.02, 0.2, 2 mg/L, 48 h	Intestinal abnormalities with fluctuating mRNA abundance of digestion-related transcripts.	[188]
pelteobagrus fulvidraco	PS + oxytetracycline (co-exposure)	100 µg/L combined	Elevated intestinal oxidative stress and tissue damage; gut-flora dysbiosis inhibited juvenile growth.	[189]

7.12. Metabolic and Molecular Effects

Beyond digestion and absorption, MPs interfere with energy metabolism, driving dysfunction in both liver and pancreas. In zebrafish, polystyrene (PS) exposure suppressed hepatic glucose and energy metabolism, reflected in reduced glucose, α-ketoglutarate, and isocitrate dehydrogenase levels. Key glycolytic genes phosphoenolpyruvate carboxykinase (pepckc), glycerol kinase (gk), and pyruvate kinase (pk) were largely downregulated, with the exception of hexokinase (hk1), which was upregulated at lower exposure concentrations [129]. Similarly, lipid biosynthesis was inhibited, marked by declining total lipid and cholesterol levels and reduced expression of the key lipogenic genes carnitine palmitoyltransferase 1 (cpt1), peroxisome proliferator-activated receptor alpha (pparα), acetyl-CoA carboxylase 1 (acc1), fatty acid synthase (fas), sterol regulatory element-binding protein 1 alpha (srebp1α), and peroxisome proliferator-activated receptor gamma (pparγ), while genes governing lipid oxidation and transport, including acyl-CoA oxidase (aco) and fatty acid-binding protein 6 (fabp6), were upregulated [129]. The above evidence indicates that MPs reduce feed utilisation efficiency in farmed fish through three interconnected pathways: suppression of feeding behaviour via impaired swimming capacity and reduced feed intake; inhibition of digestive enzyme activity, leading to intestinal obstruction and disrupted nutrient processing; and disturbance of glycolytic and fatty acid metabolism through altered gene expression and metabolite profiles.

As compiled in Table 7, microplastic exposure perturbs hepatic energy metabolism and the associated cellular stress machinery across several species and polymer types [163,168,190,191,192]. The reported changes span lipid metabolism and the expression of genes governing digestion, transport, and absorption [168], respiratory function and tricarboxylic acid cycle intermediates [190], xenobiotic biotransformation through cyp1 induction [191], and oxidative and apoptotic signalling accompanied by genotoxic damage [163,192]. Taken together, these observations indicate that metabolic disturbance is not a discrete effect but the biochemical expression of a generalised stress response, with induction of biotransformation enzymes suggesting that the associated chemical burden, rather than the particle alone, contributes to the outcome. The studies differ by three orders of magnitude in exposure concentration and employ both waterborne and dietary routes, yet report broadly comparable directions of change, indicating a response that is qualitatively robust but not yet quantitatively predictable. Its practical significance for aquaculture lies in the diversion of energy from production towards maintenance and repair.

Table 7. Metabolic and molecular disturbances following microplastic exposure in fish.

Metabolic and Molecular Disturbances
Fish Species	MP Polymer/Types	Exposure Conditions	Major Effects	References
Danio rerio	PE	32.5 µm; 1, 10, 20 mg/L, 10 d	Disrupted lipid metabolism and altered expression of genes involved in digestion, transport, and absorption.	[168]
Oreochromis niloticus	PS	2 and 20 µm; 100 µg/L, 28 d	Upregulation of egln3 and nadk, reduced isocitrate levels, and impaired respiratory function.	[190]
Danio rerio	PS	4.5 µm; 0.05 mg/L, 21 d	Upregulated cyp1 expression and hepatic vacuolation.	[191]
Oreochromis mossambicus	PP	100, 500, 1000 mg/kg	Disruption of cellular homeostasis, increased reactive oxygen species (ROS) generation, enhanced apoptosis, genotoxic effects, and histopathological alterations.	[192]
Oryzias latipes	Microfibre-type MPs	157.68 µm; 500 and 1000 microfibres/L, 21 d	Elevated CAT, SOD, MDA, and caspase-3 levels, accompanied by hepatic oxidative stress, DNA damage, and apoptosis.	[163]

7.13. Impact of Microplastics on Fish Haematology

Fish, like other aquatic organisms, are highly susceptible to anthropogenic contaminants such as microplastics (MPs) [79]. Haematobiochemical profiling is a cornerstone diagnostic tool in fish health assessment, routinely used in the laboratory, and these parameters also function as sensitive bioindicators of exposure to a range of environmental pollutants [193]. Once MPs enter systemic circulation, they can disrupt blood chemistry and trigger toxicity across multiple physiological levels, altering haemoglobin (Hb), haematocrit (Hct), mean corpuscular volume (MCV), mean corpuscular haemoglobin (MCH), mean corpuscular haemoglobin concentration (MCHC), and thrombocyte (TC) counts, collectively signalling compromised physiological and pathological states [79]. In particular, MP-induced damage to erythrocyte membranes, haemolysis, and impaired erythropoiesis may compromise oxygen-carrying capacity and contribute to altered erythrocyte-related indices [193,194]. Beyond the particles themselves, MPs and their adsorbed chemical contaminants entering the bloodstream have also been shown to alter immune function and gene expression in exposed fish [194].

7.13.1. Red Blood Cell Indices

Several studies converge on a pattern of MP-induced haematopoietic disruption, oxidative stress, inflammation, and chemical toxicity, manifesting as significantly reduced red blood cell (RBC) counts, Hb, and Hct, alongside shifting proportions of lymphocytes, eosinophils, and neutrophils [139,143,195]. MCV and MCH, by contrast, have been reported to rise with increasing MP concentration [195], a pattern attributed to MP-driven iron deficiency, nutritional stress, impaired erythropoiesis, and oxidative damage to red blood cells. A notable exception is a sharp decline in Hb observed when Nile tilapia (Oreochromis niloticus) was exposed to polyamide (PA) across multiple dose levels [196]. MCHC, meanwhile, tends to fall following MP exposure, attributed to red cell damage and Hb breakdown, while packed cell volume (PCV) similarly declines as a result of red cell destruction coupled with impaired production [195].

7.13.2. White Blood Cell and Immune Indices

Exposure to polystyrene (PS), polyethylene terephthalate (PET), low-density polyethylene (LDPE), and polyamide (PA) across varying doses has been linked to significant elevations in white blood cell (WBC) counts and blood glucose, consistent with immune activation, inflammatory signalling, and physiological responses to oxidative stress [195,196,197]. The elevation in WBC counts may reflect the ability of microplastics (MPs) to translocate across biological barriers and reach the circulatory system, where they can disturb physiological homeostasis and stimulate oxidative and inflammatory responses, thereby promoting leukocyte mobilisation. Concurrent alterations in the gut microbiota may reflect microbial dysbiosis and impaired intestinal homeostasis, potentially increasing susceptibility to inflammation and disease development [198]. In contrast, prolonged or high-dose exposure in common carp (Cyprinus carpio) and pearl spot (Etroplus suratensis) exposed to polyethylene (PE) and polyvinyl chloride (PVC) resulted in marked reductions in WBC counts, which have been attributed to immunosuppression, haematopoietic tissue toxicity, and oxidative-stress-induced cellular damage [199,200]. These contrasting haematological responses suggest that the effects of MPs on immune status may depend on polymer type, exposure concentration and duration, and the physiological capacity of the exposed fish to compensate for persistent stress.

7.13.3. Biochemical Parameters

MP exposure also disturbs broader biochemical indices, including blood urea, glucose, and cholesterol. In yellow catfish (Pseudobagrus fulvidraco), polyethylene (PE) exposure raised all three, reflecting underlying kidney dysfunction, hepatic damage, and oxidative stress [143,201,202]. Plasma electrolyte balance was likewise significantly disrupted in African catfish (Clarias gariepinus) following PE exposure [203]. In rockfish (Sebastes schlegeli), PE exposure at concentrations of 400, 800, 1600, and 3200 µg/L significantly lowered total protein, calcium, and magnesium levels, consistent with liver damage, kidney dysfunction, gill injury, disrupted homeostasis, and hormonal imbalance [204]. Beyond these indices, MP-induced inflammation can elevate circulating markers such as C-reactive protein (CRP) and cytokines. MPs have also been shown to disrupt lipid metabolism, altering the blood lipid profile through increased total cholesterol, triglycerides, and low-density lipoprotein (LDL) cholesterol, alongside decreased high-density lipoprotein (HDL) cholesterol. Endocrine interference is a further route by which MPs alter blood biochemistry: exposure has been linked to disruption of the thyroid hormone system, with corresponding changes in circulating thyroid hormone levels [205]. Liver function is similarly affected, with downstream consequences for blood biochemical parameters [118,206]. With a few exceptions, the collective evidence indicates that MP exposure in fish lowers RBC count, Hb, Hct, thrombocytes, MCHC, and PCV, while raising WBC, MCH, MCV, glucose, and total protein a pattern consistent with combined stress and immune-activation responses to microplastic contamination.

As summarised in Table 8, the haematological evidence is limited in extent but internally consistent: polyethylene exposure reduced erythrocyte counts, haemoglobin, and haematocrit in both species examined, accompanied by declining leucocyte populations and antioxidant capacity in Nile tilapia and by altered plasma protein and electrolyte status in yellow catfish [143,201]. The concordant direction of these responses—suppression of erythroid indices together with evidence of oxidative and osmoregulatory disturbance—supports the interpretation. Microplastics compromise oxygen-carrying capacity through combined haemolytic, haematopoietic, and oxidative mechanisms rather than through a single pathway. The dataset is nonetheless too small to resolve polymer-, dose-, or duration-dependence, and exposure conditions were not reported in one of the two studies. Haematological indices accordingly remain a promising but under-characterised set of biomarkers for routine health surveillance of microplastic-exposed stocks.

Table 8. Haematological and biochemical alterations following microplastic exposure in fish.

Haematological and Biochemical Alterations
Fish Species	MP Polymer/Types	Exposure Conditions	Major Effects	References
Pseudobagrus fulvidraco	PE	Not reported	Decreased RBC, Hb, haematocrit, total protein, calcium, and magnesium; increased SOD, CAT, and GST; decreased GSH.	[143]
Oreochromis niloticus	PE	10 mg/L, 15 d	Reduced red blood cells, neutrophils, lymphocytes, and antioxidant capacity.	[201]

7.14. Histopathological Impact of Microplastic on Fish

The gastrointestinal tract is consistently identified as a primary target of microplastic (MP) toxicity in fish, with damage ranging from villous injury to inflammatory and degenerative changes in the intestinal wall. In adult zebrafish (Danio rerio), a mixed-polymer exposure comprising polyamide, polyethylene, polypropylene, and polyvinyl chloride particles (~70 μm) for 10 days produced cracking of intestinal villi and splitting of enterocytes [101], establishing that intestinal injury can occur rapidly and across multiple polymers types. Comparable, concentration-dependent degeneration of the small-intestinal wall has been documented in Wami tilapia (Oreochromis urolepis) following polyethylene MP exposure, and a related trial in the same species found that after a 65-day exposure followed by a 60-day depuration period, intestinal wall degeneration was still evident, indicating that recovery from MP-induced injury is slow and may be incomplete [99]. Longer-term exposure appears to deepen this injury rather than resolve it. In European sea bass (Dicentrarchus labrax) fed a PVC-MP diet for up to 90 days, the distal intestine showed the most pronounced alterations, including detachment of the mucosal epithelium from the lamina propria, fusion of villi, and hyperplasia of mucus-secreting goblet cells [98]. Using a semi-quantitative lesion-index approach, Ahrendt et al. [207] further showed that intestinal damage in the intertidal fish Girella laevifrons, including hyperemia, leukocyte infiltration, and loss of crypt and villous cells, increased in severity with MP concentration, reinforcing a dose–response relationship across taxonomically distinct species.

Beyond the intestine itself, MP exposure has been linked to histopathological injury across multiple organs and to functional consequences for the fish. In juvenile Nile tilapia (Oreochromis niloticus), a 15-day exposure produced lesions not only in the intestine but also in the liver, kidney, gills, pancreas, and muscle, and a subsequent 15-day clearance period resolved only some of this damage, leaving other lesions persistent [103]. This partial reversibility, together with the incomplete recovery reported in O. urolepis after depuration, suggests that once a threshold of exposure is exceeded, MP-induced tissue injury can outlast the exposure period itself. Because the intestine mediates nutrient absorption and energy acquisition, this damage has been proposed to compromise digestive efficiency and, ultimately, fish welfare [187]. These findings collectively frame the intestine as a sensitive and slow-recovering target of MP toxicity—one whose vulnerability may be compounded when MPs co-occur with other contaminants. Recent studies in zebrafish indicate that co-exposure to microplastics (MPs) and copper can intensify gastrointestinal damage, producing more severe effects than exposure to either stressor alone [208]. Such interactions are particularly relevant to aquaculture because MPs can act as carriers or co-stressors in contaminated aquatic environments. Overall, the available evidence indicates that MP exposure can disrupt intestinal architecture through epithelial injury, villous degeneration, inflammatory responses, and cellular degeneration, with the magnitude and persistence of these effects depending on MP characteristics, exposure concentration and duration, and the presence of additional environmental stressors.

The studies presented in Table 9 indicate that microplastic-induced histopathology is multi-organ rather than confined to the site of entry, with lesions reported in liver, gill, and intestine following exposure to polyvinyl chloride and polyamide particles [149,209]. Two features distinguish this dataset from the others compiled here. First, the particle sizes involved are the largest reviewed, extending from 100 µm to several millimetres, indicating that structural injury does not require epithelial translocation and can be produced by particles that remain within the lumen. Second, hepatic vacuolation and inflammatory infiltration occur alongside altered antioxidant gene expression [149], linking the histological changes to the oxidative mechanisms considered in Section 7.14. Exposures ranged from 15 to 60 days and encompassed both dietary inclusion at percentage levels and waterborne exposure, so the lesions reported reflect sustained rather than acute insult. Collectively, the table indicates that histopathological assessment provides an integrative record of cumulative exposure, complementing biochemical markers that respond more rapidly but less durably.

Table 9. Histopathological alterations following microplastic exposure in fish.

Histopathological Alteration
Fish Species	MP Polymer/Types	Exposure Conditions	Major Effects	References
Cyprinus carpio	PVC	100–200 µm; 10, 20, 30% of weight, 30 & 60 d	Altered antioxidant gene expression, accompanied by hepatic cytoplasmic vacuolation and inflammatory cell infiltration.	[149]
Oreochromis niloticus	PA	500 µm–4 mm; 10 mg/L, 15 d	Nuclear and cellular abnormalities, accompanied by histopathological damage to the gills and intestine.	[209]
Danio rerio (larvae)	Naturally weathered MPs	Not reported	Neurotoxicity associated with oxidative stress and disrupted neurotransmission.	[130]

7.15. Oxidative Stress and Its Downstream Toxic Effects in Fishes

Microplastics have been widely detected in aquaculture organisms, including fish, mussels, shrimp, and crabs [210]. Their accumulation has been associated with diverse toxicological responses, with oxidative stress being one of the most prominent effects, along with disturbances in behaviour, growth, and reproductive performance. Collectively, these effects may compromise animal health and productivity, potentially resulting in reduced economic returns from aquaculture production. Microplastics (MPs) are readily taken up across the intestinal epithelium of fish, from where they translocate into internal organs and systemic circulation and ultimately infiltrate individual cells; this cellular infiltration is now recognised as the principal route through which MPs exert cytotoxicity, largely by triggering oxidative stress [78]. Under normal physiological conditions, reactive oxygen species (ROS) produced in response to endogenous or external stressors are effectively controlled by the organism’s antioxidant defence system. However, when ROS production exceeds the capacity of these protective mechanisms, redox homeostasis is disrupted, leading to oxidative stress and subsequent inflammatory responses [211]. Polystyrene microplastics (PS-MPs), in particular, interfere with intracellular metabolic processes such as mitochondrial electron transport, driving excessive ROS production. The resulting radicals attack lipids, proteins, and nucleic acids, producing widespread cellular damage [125,212]. Elevated ROS levels are regarded as an early, sensitive biomarker of MP induced tissue injury in fish [79]. This oxidative cascade is a major mechanism contributing to the neurotoxic, immunotoxic, and behavioural effects associated with microplastic (MP) exposure.

7.15.1. Induction of Oxidative Stress and Antioxidant Enzyme Responses

MP exposure elevates classical oxidative stress markers, most notably superoxide dismutase (SOD) and catalase (CAT) activity [100]. Redox homeostasis is maintained jointly by enzymatic antioxidants catalase (CAT), glutathione peroxidase (Gpx), and superoxide dismutase (SOD), and non-enzymatic components such as glutathione (GSH) and glutathione S-transferase (GST), which together scavenge free radicals and limit oxidative damage [125]. The magnitude and direction of this antioxidant response, however, are highly tissue- and species-specific. In the Korean bullhead (Pseudobagrus fulvidraco), increasing MP concentrations produced dose-dependent increase in SOD, CAT, and GPX activity in the gills, liver, and intestine, while Gst activity declined over the same range [213]. Yellowtail kingfish (Seriola lalandi) and Ambon damselfish (Pomacentrus amboinensis) both showed coordinated increases in CAT, GPx, SOD, GSH, and GST activity relative to controls, accompanied by elevated lipid peroxidation [214,215]. A broadly similar pattern of raised Cat/Sod activity, together with elevated malondialdehyde (MDA), a lipid peroxidation end-product, has been reported in the liver and brain of Nile tilapia, and the liver and intestine of blunt snout bream (Megalobrama amblycephala) [216,217].

Not all species exhibit a straightforward dose-dependent response. In goldfish and crucian carp (Carassius carassius) exposed to polyethylene microplastics, antioxidant responses followed a biphasic pattern, with SOD, CAT, and GST activities increasing at lower concentrations but declining at higher exposures. In contrast, histological tissue damage continued to intensify with increasing exposure [86,139,218]. This pattern suggests that low levels of MP exposure act as a mild stressor capable of stimulating antioxidant defences up to a physiological threshold, and, beyond this threshold, ROS production outpaces the capacity of the antioxidant system, leading to enzymatic suppression and progressively worsening oxidative injury [211,218].

As shown in Table 10, oxidative stress is the most broadly replicated response considered in this review, reported across six polymer types—including ethylene, propylene, polystyrene, polyacrylamide, polyethylene and polypropylene—and in marine and freshwater species spanning flatfish, sparids, cichlids, cyprinids, and percids [219,220,221,222,223,224]. Elevated reactive oxygen species and malondialdehyde are near-universal, and are accompanied in several studies by DNA damage, lipid peroxidation, protein ubiquitination, and apoptosis [222,223], indicating that oxidative injury propagates to macromolecular damage rather than remaining a transient biochemical signal. The antioxidant response itself is not unidirectional: superoxide dismutase, catalase, and glutathione are elevated in some exposures and depressed in others [220,221], a divergence consistent with the concentration-dependent threshold behaviour described above, in which defences are induced at low exposure and overwhelmed at higher loads. Tissue distribution is likewise uneven, with gills more affected than liver in at least one species [219].

Table 10. Oxidative stress and antioxidant responses to microplastic exposure in fish.

Oxidative Stress and Antioxidant Responses
Fish Species	MP Polymer/Types	Exposure Conditions	Major Effects	References
Scophthalmus maximus	Ethylene propylene	50–200 µm	Oxidative stress response, more concentrated in gills than liver.	[219]
Sparus aurata	PS	1–20 µm; 25 and 250 mg/kg	Increased ROS and MDA; altered SOD, CAT, and GSH; upregulated HSP70 and HSP90 in intestine.	[220]
Oreochromis niloticus	Polyacrylamide	0.1–0.4 mm; 0.018–0.09 g/L	Reduced CAT and GSH; increased MDA and lipid peroxidase levels in gills, liver, and intestine.	[221]
Danio rerio & Perca fluviatilis	PE	10–45 and 106–125 µm	Induced oxidative stress, DNA damage, lipid peroxidation, and ubiquitination in liver and gills.	[222]
Danio rerio & Perca fluviatilis	PP	8–10 µm; 1 and 10 mg/g	Induced oxidative stress, genotoxicity, and apoptosis in the liver and gills.	[223]
Cyprinus carpio	PP	0.9 mm; 1.0–2.5 g/L (food), 7 d	Oxidative damage.	[224]

Microplastic exposure exerts broader effects on fish metabolism, influencing lipid metabolism, carbohydrate metabolism, and toxin excretion, of which oxidative stress induction remains the most extensively studied [225]. Exposure has been shown to shift the overall metabolic state of fish and promote ROS production, thereby triggering oxidative stress responses [226]; the resulting ROS elevation adversely affects circulating cholesterol levels and the activity of enzymes governing lipid metabolism [119,227]. This response is not limited to polystyrene or polyethylene, as polyvinyl chloride (PVC) and polyethylene terephthalate (PET) particles (40–150 μm) also induced oxidative cellular damage in sea bream and sea bass [60], suggesting that oxidative stress is a common response across different plastic polymers.

7.15.2. Immunotoxic Effects

Substantial evidence indicates that MP-induced behavioural and growth impairments are underpinned, at least in part, by oxidative-stress-mediated immunotoxicity. Oxidative stress arising from microplastic (MP) exposure is not confined to neural pathways; it also drives inflammation, genotoxicity, and immunotoxicity, and is frequently accompanied by shifts in gut microbiota composition and corresponding changes in swimming activity and feeding behaviour [228], linking the immunotoxic and behavioural axes of MP toxicity mechanistically rather than merely correlatively.

The pathway connecting oxidative stress to immune activation has been characterised at the molecular signalling level. In fish hepatocytes, MP-elevated mitochondrial reactive oxygen species (ROS) upregulate Toll-like receptor 2 (TLR2) signalling, initiating a cascade through myeloid differentiation factor 88 and TNF receptor-associated factor 6 that culminates in NF-κB phosphorylation [229]. A parallel pathway operates in common carp (Cyprinus carpio), in which PS-MP exposure induces gill apoptosis and immune-inflammatory responses via oxidative stress, NF-κB pathway activation, and NLRP3 inflammasome triggering [230]. In zebrafish, PS exposure similarly induces oxidative stress, activates innate immune responses, disrupts gut microflora, and provokes inflammatory reactions [231], reinforcing NF-κB and inflammasome signalling as conserved mechanistic hubs of MP-induced immunotoxicity across finfish species.

These immunotoxicological effects are not restricted to finfish but extend across cultured aquatic taxa more broadly, underscoring their relevance to aquaculture health and productivity as a whole. In mussels, microplastic exposure has been linked to inflammation of the digestive system [232], while in sea urchins, it destabilises phagocyte lysosomal membranes and induces apoptosis [144]. Elevated oxidative stress markers, including superoxide dismutase (SOD) activity and lipid peroxidation, have similarly been recorded in mussels (Perna canaliculus) co-exposed to microplastics and the antimicrobial compound triclosan [233], indicating that MPs can potentiate immunotoxic responses when co-occurring with other environmental contaminants. This potentiating effect is further corroborated in the blood clam Tegillarca granosa, where microplastics were shown to enhance the immunotoxicity of the pharmaceutical sertraline [113]. Collectively, this evidence indicates that oxidative-stress-mediated immune disruption spanning TLR2/NF-κB and NLRP3 inflammasome-mediated inflammatory signalling, gill and digestive tissue inflammation, apoptosis, and contaminant-potentiated immunotoxicity is a broadly conserved response to microplastic exposure across cultured finfish and shellfish alike, with direct implications for disease susceptibility, aquaculture productivity, and economic sustainability.

The studies compiled in Table 11 indicate that microplastic exposure elicits active immune engagement rather than simple immunosuppression, at least over the short exposures examined [230,234,235]. Reported responses include gill apoptosis and inflammation mediated through oxidative stress and NLRP3 inflammasome activation in common carp [230], accumulation of particles in spleen and kidney accompanied by increased melanomacrophage activity in spotted seabass [234], and upregulation of immune-related genes with elevated hepatic metabolic activity in rockfish after only three days of exposure [235]. Two points follow. First, the recovery of particles from spleen and kidney links immune activation directly to the tissue distribution considered in Section 6, indicating that the response is mounted against internalised material and not solely at epithelial surfaces. Second, the activation reported here contrasts with the leucocyte depletion after prolonged or higher-dose exposure, a discrepancy best interpreted as a biphasic response in which early inflammatory activation gives way to functional exhaustion. Distinguishing these phases is central to determining whether microplastic exposure increases disease susceptibility under production conditions.

Table 11. Immune responses and immunological alterations following microplastic exposure in fish.

Immune Responses
Fish Species	MP Polymer/Types	Exposure Conditions	Major Effects	References
Cyprinus carpio	PE	Not reported	Induced gill apoptosis and inflammation through oxidative stress and NLRP3 inflammasome activation.	[230]
Lateolabrax maculatus	PP	1–6 and 10–30 µm; 50, 500, 5000 µg/L, 14 d	Particle accumulation in the spleen and kidney, accompanied by increased melanomacrophage activation.	[234]
Sebastes schlegeli	PET	1.84–810 mg/kg, 3 d	Upregulated immune-related gene expression and enhanced hepatic metabolic activity.	[235]

8. Conclusions

Microplastic contamination in aquaculture represents a production concern rather than a peripheral environmental issue. Cultured fish may be exposed through contaminated source waters, plastic-based infrastructure, and, importantly, feed ingredients, including fishmeal derived from MP-exposed fish. Following ingestion, MPs can impair intestinal integrity, digestive enzyme activity, feeding, and nutrient utilisation, while smaller particles may translocate to internal tissues. Oxidative stress emerges as an important mechanistic pathway linking these effects with inflammation, immune and haematological disturbances, neurotoxicity, endocrine and metabolic disruption, and altered reproductive function, with potential consequences for growth, feed efficiency, disease resistance, and aquaculture productivity. However, responses remain highly variable among species and exposure conditions, and the extent of tissue translocation and biomagnification remains uncertain. Chronic exposure, environmentally relevant low concentrations, and interactions with co-occurring contaminants further complicate interpretation of MP toxicity. These uncertainties are compounded by methodological heterogeneity and differences in particle characteristics, exposure regimes, and experimental design. Overall, the evidence supports an integrated view of MPs as a multidimensional stressor affecting both fish health and production performance, highlighting the importance of consistent assessment approaches, monitoring of feed and water inputs, and management of plastic sources within aquaculture systems to strengthen risk assessment and support sustainable aquaculture.