Microplastics and Nanoplastics
REVIEW	Open Access
Beyond plastics: occurrence and ecological 2 risks of non-plastic microfibres in aquatic organisms
Daniel J. Jolly1*, Eoin J. O'Gorman2, Dannielle Senga Green3, Richard J. Cooper1, Adil Bakir4 and Trevor J. Tolhurst1
Abstract
Non-plastic microfibres, here defined as anthropogenically manipulated fibres of cellulosic or animal origin such as wool, cotton and rayon, have been increasingly recognised as a significant component of anthropogenic microparticle pollution in aquatic environments. Emerging attention has also been brought to the ecological ramifications of non-plastic microfibres and their associated chemical additives. However, environmental anthropogenic microparticle surveys have often omitted both plastic and non-plastic microfibres, with greater attention being placed on microplastic fragments, films, and spheres. This lack of attention has potentially led to the underestimation of non-plastic microfibre pollution in aquatic environments and biota. Through a comprehensive systematic review, we collate and analyse published literature (2011-2024) on non-plastic microfibre occurrence in aquatic animals and the associated ecological impacts. This review demonstrates that plastic and non-plastic microfibres, when looked for, are a significant component of anthropogenic microparticle (AMP) loading in aquatic biota across environments, habitats and feeding strategies. Greatest loading appears in freshwater environments with some cases accounting for 100% of detected anthropogenic microparticles, despite relatively limited study focus. The ecological impacts of non-plastic microfibres may also elicit varied effects on biota, depending on the context and nature of exposure, although targeted experiments are scarce in the literature. This review highlights the underestimation and potential misidentification of non-plastics due to methodological limitations, inconsistent reporting, and lack of focus. Here we emphasise that future research should develop standardised anthropogenic microparticle survey methodologies that incorporate non-plastic particles and microfibres, with greater effort placed on understanding microfibre pollution in aquatic biota. Further exploration into the ecological impact of non-plastics is crucial to understanding and mitigating the risks associated with these pollutants.
Keywords Natural fibres, Microplastic, Semi-synthetic fibres, Plastic alternative, Ecological impacts, Methodological limitations, Anthropogenic microparticle

Correspondence:
Daniel J. Jolly
d.jolly@uea.ac.uk
'School of Environmental Sciences, University of East Anglia, Norwich
Research Park, Norwich NR4 7TJ, UK
School of Life Sciences, University of Essex, Colchester CO4 3SQ, UK
3Applied Ecology Research Group, School of Life Sciences, Faculty of
Science and Engineering, Anglia Ruskin University, Cambridge
CB1 1ΡΤ, UK
4Centre of Environment Fisheries and Aquaculture Science, Lowestoft NR33 OHT, UK

0 Springer Open

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Introduction
Microfibres are some of the most prevalent micropollutants in aquatic environments, accounting for 80-90% of anthropogenic microparticle (AMP) compositions [1-3]. Microfibres, under the definition proposed by Liu et al. [4], are “any natural or artificial fibrous materials of threadlike structure with a diameter smaller than 50 um, length ranging from 1 μm to 5 mm, and length to diameter ratio greater than 100:1". Globally, fibre production is rapidly expanding, increasing from 92 million tonnes in 2013 to 124 million tonnes in 2023 [5, 6]. Of these fibres, 67% were composed of plastic polymers such as polyethylene terephthalate, polypropylene, and polyamide [5]. The remaining 33% originated from natural (e.g. cotton and wool) or semi-synthetic materials (e.g. rayon, lyocell, modal). Aquatic microfibre pollution stems from woven textiles (e.g. clothing), nonwoven textiles (e.g. wet wipes and other personal hygiene products), and cigarette filters, often through wastewater discharge [7-9]. For example, from a single 7 kg load of domestic clothes washing over 700,000 microfibres could be released into wastewater systems [10, 11]. Wastewater treatment systems in countries such as the United Kingdom are considered highly efficient, retaining up to 99.9% of all wastewater AMPs (including microfibres).However, mismanagement of sewage outputs and high loading of microfibres into the wastewater system causes considerable amounts of microfibre pollution to enter natural aquatic environments [12-17]. It has been well established that due to the morphology and size of microfibres, organisms across trophic systems ingest them [18-23]. Due to microfibres and AMPs being persistent and pervasive in the natural environment, with no current evidence of full biodegradability, this makes them materials of particular concern.
Non-plastic microfibres
Since their first reported presence, research focus has mainly been targeted on plastic microfibres, those that are composed of anthropogenically derived polymers that are persistent in the environment and are imbued with chemical additives, some of which are toxic [15, 24-28]. However, limited attention has been given to natural or semi-synthetic microfibres, despite their being a significant proportion of global fibre production [23, 29-33].
Despite established definitions, such as those used by the OSPAR (Convention for the Protection of the Marine Environment of the North-East Atlantic) commission, the definitions and categorisation of microfibres are often confused or misrepresented, leading to fibres being inconsistently or mis-labelled as plastic, natural, bioplastic and others [29, 30, 34]. Clearer definitions should be utilised within the field of environmental sciences. For the purpose of this investigation, we define plastic microfibres as microfibres being composed of purely anthropogenic components, such as synthetic polymers and their associated chemical additives (e.g., polypropylene, polyethylene terephthalate, polyamide (nylon)) [27, 35]; and non-plastic microfibres are defined as being composed either purely from natural sources (e.g., cotton, wool, asbestos), or semi-synthetic fibres, those derived from natural materials that have been chemically or mechanically reconstructed with the addition of manufactured synthetic substances, that do not include plastic polymers (e.g., rayon, lyocell, modal) [36, 37]. For this investigation, it was decided that cellulose acetate would also be considered with plastic polymers as an exception despite being a semi-synthetic material. This is due to the preparation of cellulose acetate, where acetylation alters the chemical structure of the material so that it is no longer similar to natural cellulose or regenerated cellulose materials [38]. This alters its behaviour so that it acts more like conventional plastic polymers in that it persists in the environment longer than other non-plastic materials (a period of years rather than weeks or months), causing similar negative effects [38, 39].
Methodological limitations in the accuracy of characterisation of microfibres may cause mislabelling or omission of non-plastic microfibres in favour of AMPs, driving their lack of attention [29-31]. As plastic microfibre pollution has been the focus and concern for the last decade, most methodologies used in plastic and nonplastic microfibre characterisation are designed to suit the extraction and identification of plastic particles [29, 31]. This commonly includes digestion protocols such as potassium hydroxide and hydrogen peroxide under high temperatures. Both digestants, as well as high temperatures, are known to degrade non-plastic materials such as microfibres, potentially fragmenting or destroying them entirely during the process [29, 40-46], . Other common methods, such as the use of sodium hypochlorite, have been demonstrated to degrade plastic microfibres, likely causing even greater damage and underestimation of non-plastic microfibres [47]. The common material characterisation methods of FTIR or Raman spectroscopy are also known to be challenging when examining microfibres of diameter < 30 pm, especially when samples may be degraded by digestion procedures [29, 46]. Therefore, it is highly likely that non-plastic fibres, as a component of microfibre and overall AMP pollution, are underestimated.
Non-plastic microfibres can also be omitted or overlooked from AMP surveys in favour of plastics due to assumptions that, as they are cellulosic or animal in nature, or are listed as biodegradable, they should degrade into harmless constituents and have little to no environmental impact [30, 48-50]. The assumption that natural and semi-synthetic materials are harmless and rapidly degrade is being contested [51-56]. Several studies in geographically isolated locations and aquatic organisms, such as the arctic, deep-sea trenches and the diet of king penguins, have noted a significant presence of natural and semi-synthetic microfibres [57-60]. This calls into question their assumed rapid rates of degradation and short residency time, as they can withstand long-range transport and accumulate in the environment and are ingested by organisms [31].
In addition, an increasing number of studies have proposed that non-plastic microfibres may have negative ecological or ecotoxicological impacts [51-56], and their fate may be altered by the chemical modifications they receive during manufacture.
Ecological concerns
Industrially manufactured non-plastic microfibres are exposed to various chemical dyes, additives, and surface finishes to improve their functional performance and durability [32, 49, 51, 58, 61-63]. A proportion of the utilised chemical additives are known to be toxic or elicit carcinogenic effects on exposed organisms [51, 58, 64, 65]. The morphology of microfibres may also cause physical damage to exposed organisms [59, 66]. Multiple studies and reviews have highlighted concerns for the ecological damage microfibres may cause, whilst others have demonstrated little or no effects of these materials, making the determination of their impact challenging to conclude (e.g [22, 23, 47, 52, 66, 67]). , . However, few studies have yet focussed on the potential ecological effects specific to non-plastic microfibres and their additives (for examples see [52, 53, 68]), or how they behave in natural systems. Therefore, the threat of non-plastic microfibres to aquatic ecosystems, relative to their synthetic counterparts, is unknown.
Research aims
When considering microfibre and AMP loading in aquatic animals, studies often use the number of items per individual. While this can be a useful value when examining similar taxa, using items per individual does not accurately allow for comparison across taxa (e.g., 10 microfibres per mussel, versus 10 microfibres per bluefin tuna) [22]. The relative number of items per gram of weight thus provides a more comparable understanding of the loading within animals across taxa and could help identify taxa that are more vulnerable to microfibre exposure. However, it must be noted that when examining invertebrates, often the entire soft tissue is utilised, whereas vertebrate samples often only include specific body parts such as the digestive tract. Despite the limitations, for the purpose of this investigation, the number of items per gram of wet weight was utilised as a more accurate comparison between different species, which is further discussed in Sect. 4.1.
Reviews have also focussed on taxonomic groups rather than potentially more environmentally meaningful habitats or functional groups (e.g [23, 69]). Whilst taxonomic groups can provide insight into which evolutionary lineages may be more vulnerable to microfibre pollution, they do not account for the high variability in life history and function of evolutionarily related species. For example, species within the class Polychaeta, despite often being grouped together in AMP research, inhabit a variety of environments, and employ completely different feeding and living strategies, making their vulnerability and exposure to AMPs highly variable [70]. Here we assert that it is more important to understand how animals with different feeding strategies and different habitats may be affected by microfibre loading, as this is more likely to determine their vulnerability and exposure. Furthermore, whilst the presence of microfibres has been reviewed, the ecological impacts of non-plastic microfibres have not yet been considered in relation to all environments. Therefore, here we summarise the current academic understanding of non-plastic microfibres and consider the issues highlighted above for aquatic animals by focussing on four main questions:
1.		What levels of AMP concentrations are found in wild aquatic animals?
2.		What proportions of AMPs are microfibres?
3.		What proportion of microfibres are non-plastic?
4.		What are the ecological consequences of non-plastic microfibres?
We address these questions through a systematic review and meta-analysis of the current literature, identifying studies that have examined the presence of non-plastic microfibres. The average loading of microparticles per gram of wet weight was calculated; and, for each species, the proportion of those particles that were plastic and non-plastic microfibres. The results are discussed in the context of current understanding of the ecological impact of non-plastic microfibres, to highlight vulnerable groups of organisms that may be at risk and guide future research.
Methods
Publication reference material
Analysis of the current understanding of (plastic and non-plastic) microfibre loading in aquatic animals and the ecological impact of non-plastic microfibres was conducted, following Walkinshaw et al. [22]. Focus for this aspect of the investigation centred on aquatic meso- and macrofauna as AMP loading of microfauna is scarcely investigated. Aquatic macrophytes and microphytes were
excluded, due to the scarcity of investigations characterising microfibre materials and AMPs, in addition to the different phylogeny and pathways of AMP contamination of macro and microphytes compared to animals making analysis and comparison challenging.
Peer-reviewed literature was examined using a specific set of terms separated by Boolean operators (Table 1), based on the guidance outlined by Foo et al. [71]. Only studies that explicitly observed the presence of nonplastic microfibres were considered in this investigation. While limiting the overall representation of AMP pollution, this approach allows a more comprehensive analysis of non-plastic microfibre presence and proportions in aquatic animals. Due to the apparent limitations of many AMP surveys and the potential omission of microfibres and non-plastics, this prevents the overinflation of null observations and highlights a need for future research to utilise methods that allow the recovery of non-plastic and plastic microfibres.
Three academic search engines were utilised for resource collation: Web of Science, Scopus and PubMed. The AI tool Research Rabbit was also utilised to perform forward and backward searches of literature that cited or had been cited by the collected publications. Research articles published between January 1st 2011 (the earliest identified year of publication for microfibres in aquatic environments by spectroscopic methods) and December 31st 2024 were included in the data analysis of this review. The collated publications were then screened and assessed for relevant results, scope, and appropriate methodology to ensure quality control, using the software Rayyan.
Source material quality control and assessment
The literature that was extracted from each database was first screened for relevancy and duplications. Only
Table 1 Search terms entered into Scopus, web of science and pubmed research databases for collecting publications on the occurrence of non-plastic microfibres in aquatic animals

microfibre(s) OR
microfibre(s)

"natural microfibre(s)"OR cellulose OR cellulosic OR "animal microfibre(s)"OR "anthropogenic microfibre(s)"OR wool OR rayon OR viscose OR

marine OR freshwater OR aquatic OR water OR ocean OR lake OR river OR Brackish OR Estuary

lyocell OR modal OR cotton

Term 1 Term 2 AND Term 3 AND Term 4 AND invertebrate OR fish OR mollusc OR mammal OR bird OR seabird OR gut OR ingestion OR bivalve OR crab OR barnacle OR holothurian OR egestion OR digestive OR tissue OR organism OR echinoderm OR polychaete OR ecotoxicology OR species OR cetacean OR seal peer-reviewed publications were included to ensure academic quality. As the focus of the first part of this investigation is on aquatic animals in natural systems, any studies involving farmed animals or forced exposures were omitted. Studies that involved the examination of fish market produce were also omitted, as potential loading from the preparation, transport and storage processes could impact the reliability of results. Basic methodological standards were also assessed, excluding any species that had less than three replicates within a study, and only including studies where suitable methodologies for AMP and non-plastic microfibre identification such as FT-IR or Raman spectroscopy, SEM EDX analysis, LC-MS or pyrolysis GC-MS were employed, and loading control measures such as field and laboratory blanks were used to ensure greater reliability of results. After screening, 102 papers were retained for further analysis of AMP and microfibre data (See supplementary materials for included publications and summary table).
Organism grouping, AMP content and proportions data analysis
Once screened, individual species from the included literature were recorded and placed into groupings for three distinct variables. These included one of three environmental categories: “freshwater” (e.g. lakes, rivers, ponds), “brackish” (e.g. estuaries), and “marine” (e.g. coastlines, enclosed seas, ocean basins). Animals were also assigned to one of five different habitats: “reef” “benthic”, “pelagic” “benthopelagic” and “demersal”. Finally, animals were assigned to one of six feeding types: “grazers” (consuming biofilm from substrate surfaces), "suspension feeders” (actively or passively filtering the water column for planktonic food sources or organic matter), “detritivores” (primarily consuming dead and decaying organic matter), “omnivores” (utilising a range of food sources), “herbivores” (primarily consuming aquatic plants or macroalgae), and “predators” (primarily consuming live prey). If the grouping data was not present in the study, classifications for each species was decided using the World Register of Marine Species (WoRMS) [72], the Marine Life Information Network (MarLIN) [73], or FishBase [74].
Within the AMP literature, studies may present their results either as AMP per individual or per gram of wet weight (AMP g 1 w.w.). When this was unavailable, AMP per gram wet weight was calculated by dividing the average AMP per individual by the average wet weight of the organism in grams. When no data was explicitly available in the study or it’s supplementary materials, the corresponding author was contacted. If the data were still not obtained, the study was excluded. This occurred in 34 cases, leaving a final dataset of 68 studies for AMP loading analysis. Furthermore, in studies examining AMP loading and material characterisation in vertebrates such as fish and mammals, specific organs are analysed such as the gastro-intestinal tract, whereas in invertebrates such as molluscs, the whole soft tissue is analysed. For the purposes of this investigation, vertebrate and invertebrate species are included to provide an indication of which environments, habitats and feeding strategies may be more vulnerable than others, but the AMP loading between these groups is not directly comparable. The limitations of the different methodological approaches used for these two groups is discussed in Sect. 4.1.
Data on microfibre and non-plastic proportions of AMP loading in aquatic animals were also collected. Due to inconsistencies in reporting procedures and presentation of results by different studies, it was impossible to accurately isolate non-plastic microfibres from other non-plastic microparticles in the literature for individual species as they were often pooled with other materials. Therefore, to provide an estimation, this review pooled all known material types not relating to purely plastic polymers as non-plastic particles, to indicate how prevalent non-plastic particle loading is in aquatic animals when detected. The observed proportions of material composition were collated for each species, including overall microfibre proportions and non-plastic particle proportions. Where no data was explicitly available from the study, and no response was received from the corresponding author, the research was omitted. This occurred in 6 cases, leaving a final dataset of 97 studies for microfibre and non-plastic particle proportions.
The mean AMP loading, microfibre proportions and non-plastic particle proportions for each species were then assessed against the three different grouping variables separately, using a one-way ANOVA and Tukey test. The analysis was conducted using R version 4.3.1 [75]. The following packages were used: ggplot2 for
Table 2 Search terms entered into Scopus, web of science and pubmed research databases for collecting publications on the ecological impact on non-plastic microfibres
Term 1	Term 2 AND	Term 3 AND	Term 4 AND	
microfibre(s) OR	"natural microfibre(s)"	marine OR	eco-	
microfibre(s)	OR cellulose OR	freshwater OR	toxicol-	
	cellulosic OR "animal	aquatic	ogy OR	
	microfibre(s)" OR		ecotox-	
	"anthropogenic		icOR	
	microfibre(s)"OR		toxic	
	wool OR rayon OR		OR ex-	
	viscose OR lyocell OR		posure	
	modal OR cotton		OR	
behaviour OR health
OR function
visualisation, readxl and dplyr for general coding and analysis [76-78]. Explanatory variables were assessed for normality and homogeneity of variance. Due to non-normal distribution of AMP loading, data was loglO transformed to reveal any potential relationships between grouping factors.
Assessing the ecological impacts of non-plastic microfibres When considering the ecological consequences of non-plastic microfibres, a traditional literature review approach was adopted due to the scarcity of publications. As for AMP loading and non-plastic microfibre proportions, Boolean operators were utilised to collate all relevant studies between January 1st 2011 to December 31ST 2024 (Table 2). Publications were then screened for relevance and overall quality, such as number of replicates, appropriate controls, and relevant test parameters. The resulting studies and related investigations were reviewed to establish an overview of current understanding of the ecotoxicological effects that non-plastic microfibres have on aquatic organisms (see supplementary materials for summary table).
Results
Publication bias
Within the available literature there was a clear bias and popularity towards marine environments, where 75 of the included 102 studies were carried out, compared to only 14 in brackish and 12 in freshwater. Benthic animals were the most studied habitat grouping within the studies (n = 132), followed by demersal (n = 78) and pelagic (n =71) animals. Reef animals were the least commonly studied habitat grouping (n = 13), followed by benthope- lagic (n = 37) animals. The most studied feeding groups were predators (n = 131), followed by omnivores (n = 97) and suspension feeders (n = 50). Herbivores and grazers were the least represented feeding groups (n = 9), followed by detritivores (n = 22).
Prevalence of AMPs in aquatic animals
A total of 219 animals belonging to 183 species were examined for the presence of AMPs, including microfibres and non-plastic particles, including bivalves, cetaceans, elasmobranchs, and reptiles. AMP loading ranged from 28.85 AMP g 1 w.w. in the common prawn (Palaemon sp.), to 0.0001 AMP g-1 w.w. in the blue jack mackerel (Trachurus picturatus). The mean AMP loading across all animals was 0.65 AMP g-1 w.w.
Environment
No significant difference among environments (freshwater, brackish, or marine) was detected in the observation of AMP loading (P: 0.807, F: 0.215, df: 2; Fig. 1A).

Fig. 1 Mean anthropogenic microparticle contamination in aquatic animals per gram of wet weight. Grouped by (A) Environment, (B) Habitat, (C) Feeding strategy. The lettering above each factor denotes statistically significant differences between each grouping by Tukey's test, with a being statistically different from b and c, with each grouping having the sanae letter demonstrating no statistically significant differences from each other. (Susp feeder: Suspension feeder)
Habitat
Habitat significantly influenced the loglO AMP loading of animals (Ρ: 2.4e 5, F: 7.046, df: 4; Fig. 1B). AMP loading of benthic animals was statistically different (Tukey test, P < 0.05) from all other habitats except reef animals (Tukey test, P> 0.05), with a back: transformed median loading of 0.27 AMP g 1 w.w. (n: 72, IQR: 0.98) and 0.19 AMP g-1 w.w. (n: 6, IQR: 0.18). Pelagic (0.07 AMP g-1 w.w., n: 50, IQR: 0.35), benthopelagic (0.04 AMP g-1w.w., n: 30, IQR: 0.06) and demersal (0.03 AMP g-1 w.w., n: 61, IQR: 0.17) animals demonstrated no statistically significant differences in AMP loading. All residual values were examined for homogeneity of variance and normality of distribution, demonstrating normally distributed data after loglO transformation (Figure S1).
Feeding strategy
Feeding strategy significantly influenced the loglO AMP loading of animals (P: 1.72e 12, F: 14.84, df: 5; Fig. IC). Grazers and suspension feeders demonstrated the highest back-transformed median loading (0.8±1.5 AMP g_1 w.w., median±IQR, n = 9, and 0.46± 1.26 AMP g-1 w.w., n = 27 respectively) statistically different (Tukey test, P < 0.05) to other feeding strategies except herbivores (0.18 ± 0.32 AMP g-1 w.w., n =7) and detritivores (0.11 ± 1.29 AMP g-1 w.w., n=8) (Tukey test, P>0.05). Omnivores (0.09 ± 0.45 AMP g w.w., n =73) and predators (0.02 ± 0.07 AMP g-1 w.w., n =95) demonstrated the lowest loading (Tukey test, P < 0.05) statistically different to all feeding strategies except herbivores (Tukey test, P> 0.05). All residual variance was found to be normally distributed (Figure S2).
Proportion of plastic and non-plastic microfibres in aquatic animals
From the same pool of studies reviewed, the overall (plastic and non-plastic) microfibre proportion of identified AMPs was examined in 253 animals from 220 different species. When detected, the proportion of AMPs that were microfibres in organism gastro-intestinal tracts and tissues ranged from 100% in 24 species (animals such as the bivalves; Amarillesma mactroides, Brachidontes rodriguezii and Pacific herring Clupea pallasii) to 0.7% in the giant freshwater shrimp (Machrobrachium rosenber- gii). The mean proportion of AMPs that were microfibres across all species was 72.8%. In all but 10 species the proportion of microfibres was greater than 50%.
Environment
Environment influenced microfibre proportions of AMP loading (P: 0.005, F: 5.355, df: 2; Fig. 2A). Freshwater animals exhibited the greatest proportion of microfibres (84.9±4.9%, mean± SE, n = 29) (Tukey test, P< 0.05), compared to brackish (71.3 ±2.4%, n = 41) and marine (71.3 ± 1.6%, η = 183) animals which were not statistically different to each other (Tukey test, P> 0.05). All residual variance was found to be normally distributed (Figure S3).
Habitat
Habitat significantly influenced microfibre proportions (P: 0.0002, F: 5.872, df: 4; Fig. 2B). Benthic animals exhibited the greatest microfibre proportions (79.3 ± 2%, n = 99) compared to demersal animals (67 ± 2.9%, n = 65) and reef animals (49.3 ± 12%, n = 6) (Tukey test, P < 0.05). However, despite pelagic (71.9 ±2.3%, n = 62) and benthopelagic animals (69.5 ±4.7%, n=21; Fig. 3B) demonstrating lower mean proportions than benthic animals, these groups were not statistically different from other habitats (Tukey test, P> 0.05; Fig. 3B). All residual variance was found to be normally distributed (Figure S4).
Feeding strategy
Feeding strategy significantly influenced microfibre proportion (P: 0.041, F: 2.352, df: 5; Fig. 2C). However, when comparisons were made between groups, no statistically significant differences were detected (Tukey test, P> 0.05; Fig. 3C). All residual variance was found to be normally distributed (Figure S5).
Proportion of non-plastic particles in aquatic animals
When examining non-plastic materials as proportions of AMP loading, 255 animals from 226 different species were examined from the literature. The proportion of AMPs determined as non-plastic ranged from 100% in 10 different species (freshwater crustaceans and insects such as Asellus sp., Caenis sp., and Limnius sp.) to 1.2% in Bogue (Boops boops). The mean non-plastic particle proportion across all studied animals was 54.6%.
Environment
Environments significantly influenced the proportion of non-plastic particles observed in animals (P: 7.91e-6, F: 12.31, df: 4; Fig. 3A). Freshwater environments demonstrated the greatest proportion (75.3 ± 5%, n = 25), followed by marine (54.4 ± 1.9%, η = 188) and brackish environments (43.2 ± 3.6%, n = 42). All environments were statistically different from one another (Tukey test, P <0.05). All residual variance was found to be normally distributed (Figure S6).
Habitat
Habitat significantly influenced non-plastic particle proportions (P:4.74e-6, F: 7.954, df: 4; Fig. 3B). Benthic habitats demonstrated the highest mean proportion (62.5±2.4%, n =116) but were not statistically significant different from reef (60.2 ± 5.6%, η = 10), benthopelagic (56.1 ± 7.6%, η = 14) and pelagic habitats (52. ± 3.3%,

Fig.2 Mean microfibre proportion or recorded anthropogenic microparticle contamination in aquatic animals. Grouped by (A) Environment, (B) Habitat, (C) Feeding strategy. The lettering above each factor denotes statistically significant differences between each grouping by Tukey's test, with a being statistically different from bande, with each grouping having the same letter demonstrating no statistically significant differences from each other. (Susp feeder: Suspension feeder)

Fig.3 Mean non-synthetic particle proportion or recorded anthropogenic microparticle contamination in aquatic animals. Grouped by (A) Environment, (B) Habitat, (C) Feeding strategy. The lettering above each factor denotes statistically significant differences between each grouping by Tukey's test, with a being statistically different from b and c, with each grouping having the same letter demonstrating no statistically significant differences from each other. (Susp feeder: Suspension feeder)
n = 62) (Tukey test>0.05). Demersal habitats demonstrated the lowest proportions (39.1 ± 3.3%, n = 53), statistically different from benthic and pelagic environments (Tukey test, P <0.05). All residual variance was found to be normally distributed (Figure S7).
Feeding strategy
Feeding strategy demonstrated no significant influence on non-plastic particle proportion in animals (P: 0.144, F: 1.664, df: 5; Fig. 3C).
Ecological impacts of non-plastic microfibres on organisms Once an organism is contaminated through ingestion, inhalation or adhesion, plastic and non-plastic microfibres can cause detrimental ecological impacts via three main effects: physical damage, chemical damage, and exposure to harmful pathogens as vectors of disease. Less than 30 studies were identified as having explicitly explored the ecological impacts of non-plastic microfibres. Here we collate the main ecological threats to biota currently associated with non-plastic microfibres.
Physical effects of microfibres
When exposed to plastic and non-plastic microfibres, organisms exhibit a wide variety of responses. Organisms such as the three spined stickleback (Gasterosteus aculeatus) and the freshwater shrimps Gammarus due- beni and Gammarus pulex demonstrated no detrimental effects associated with microfibre exposure [79-82]. When exposed to relatively low, environmentally relevant concentrations (up to 2 mg L 1), organisms were noted to take up both plastic and non-plastic microfibres in their digestive system, with no detected effects on growth, mortality, behaviour, immune response, or histology [79-82].
When concentrations are increased (e.g. 3, 10, and 10000 microfibres L-1 and 75-2000 mg L- 1), other organisms have exhibited detrimental responses to nonplastic microfibres. Microfibre exposure (including polyester, nylon, cotton and hemp) after seven days reduced growth rate, altered swimming behaviour and increased oxidative stress in the mysid shrimp Americamysis bahia [83]. The pacific oyster Crassostrea gigas also exhibited reduced enzyme activity and heightened oxidative stress when exposed to polyester, acrylic, nylon, wool and cotton [53]. Microfibre ingestion both plastic and non-plastic, caused gut damage, reduced growth, and mortality at high concentrations (1000 and 2000 mg L’ 1) to brine shrimp (Artemia fransiscana) and reduced reproductive output, oxidative stress and mortality in daphnia (Daphnia magna) under chronic exposures [54, 55, 84]. Juvenile organisms have also demonstrated heightened vulnerability, with cotton microfibres inhibiting growth and development in blue mussels (Mytilus edulis) and mysid shrimp (A. bahia), with effects also seen on the movement and swimming behaviour of estuarine shrimp (Menidia beryllina) [52, 56].
Chemical leachates and environmental contaminants
Harmful additives are leached from plastic and non-plastic microfibres as an exposure pathway to aquatic biota, eliciting harmful effects. Non-plastics such as wool and cotton have been identified as containing Polychlorinated biphenyls (PCBs) and Polcyclic Aromatic Hydrocarbons (PAHs), known toxic compounds that are carcinogenic and negatively impact reproductive health and immune responses [53, 85, 86]. The relatively rapid degradation of non-plastic microfibres releases additives in a shorter timeframe compared to plastic, posing a short-term high concentration risk of exposure, especially in aquatic systems and wastewater treatment outflows, where heightened biodegradation can release these additives in detrimental quantities [87-89]. Non-plastics can also adsorb toxic compounds from the water column and expose organisms to them through ingestion and inhalation [68]. Released leachates of non-plastic microfibres have induced gill damage in exposed zebrafish (Danio rerio), and rainbow trout (Onchorhynchus mykiss) and oxidative stress in juvenile brown trout (Salmo trutta), with high concentrations causing mortality [90]. Pacific oysters (C. gigas) have exhibited oxidative stress and increased immune response when exposed to non-plastic and plastic microfibre leachates. Greater detrimental impacts were found with a combination of the microfibres, and leachates compared to leachates alone [53]. Non-plastic microfibres and their leachates have also been shown to impact parasite host interactions, by causing lethal effects on the parasite Gyrodactylus turnbulli and oxidative stress in the fish Poecilia reticulata [91, 92]. Thick shelled mussels (Mytilus coruscus) have exhibited enhanced gut damage, altered gut microbiome and oxidative stress when exposed to a combination of rayon microfibres and PCBs [68].
Leaching into the water column also affects aquatic plant, algae and microalgal species. High concentrations (100-500 mg L 1) of non-plastic microfibres with associated leachates inhibit growth, reduce photosynthetic capacity and increase oxidative stress in common duckweed (Lemna minor) and microalgae (Chlorella pyre- noidosa) [93, 94]. Common duckweed has also shown successive growth and photosynthesis inhibition with generations chronically exposed to non-plastic microfibres [93].
Vectors of disease
Plastic and non-plastic microfibres can act as vectors for harmful pathogens and parasites. Microbial colonisation is especially prevalent when microfibres pass through wastewater treatment systems, where they become host to pathogenic communities such as Vibrio spp. or E. coli [95, 96]. Specific textile sources of microfibres, such as wet wipes and cotton buds, both plastic and non-plastic, have been demonstrated to host E. coli communities for up to 15 weeks, risking exposure to aquatic organisms and facilitating the development of antibiotic-resistant bacteria in waterways through increased exposure to other contaminants within wastewater [95, 97].
Discussion
Methodological challenges
Inconsistent reporting, variable methodologies, and display of results within the literature present challenges to understanding the distribution of AMP and microfibres in aquatic animals and the potential vulnerability of certain communities or taxonomic groups. It was apparent from this investigation that many studies, when presenting the characterised materials detected in animals, presented a pooled result of the average particles found in all animals, even when multiple different species were examined. This makes it impossible to distinguish any variation between different animals or species, or where the examined particles originated. A lack of consistent reporting was also prevalent, with non-plastic microfibres being labelled in various forms or not being distinguished from one another, for example being called “natural”, or “animal”, or all labelled under “cellulosic” “semi-synthetic” or “anthropogenic”. In some cases, non-plastic microfibres are labelled as plastic, whereas other studies distinctly separate non-plastic from plastic microfibres. This unclear and inconsistent characterisation drives potential underestimation and misunderstanding of non-plastic fibre pollution [30, 33]. The clear bias towards marine investigations also skews AMP loading data, as freshwater and brackish animals are underrepresented.
Furthermore, the differing target methodologies for vertebrates and invertebrates makes comparison difficult as many studies of aquatic vertebrates do not give a representation for total body burden as only the digestive system and in some cases the respiratory system are included in analysis of material types and quantities. This prevents accurate comparisons between invertebrate and vertebrate species. However, as ingestion is the most likely way that aquatic animals may interact with and become contaminated by AMPs, followed by inhalation, it is probable that the majority of AMPs within the body would be found in the digestive and respiratory systems [98, 99]. While this does not account for AMPs that may have transferred to other tissues and organs, it allows for a rough comparison and evaluation of AMP loading and material characterisation in aquatic animals.
The lack of standardised, reliable sample identification and analysis, therefore, complicates understanding of non-plastic microfibre loading in aquatic animals. It is difficult to accurately establish the extent of loading across animals and environments. However, despite these caveats, the data that does exist on the proportions of non-plastic particle pollution provide insight as to what material types are more common in aquatic animals across the globe, and whether non-plastic microfibres are a material of concern for future monitoring and research. This review highlights the need for better understanding of non-plastic microfibre loading in aquatic animals and indicates several potentially vulnerable environments, habitats, and life strategies.
The distribution of microplastics, microfibres and nonplastic particles
Environment
Despite an apparent focus on marine systems (previously viewed as the major global AMP sink [100]), no significant difference was found in AMP loading of animals from different aquatic environments. AMP loading appears ubiquitous. Considering a significant proportion of AMP pollution originates from terrestrial sources, this investigation suggests that biota may be exposed to AMPs at similar levels from source to sink. The proximity to pollution sources, such as wastewater treatment and sewage outflows, industrial waste, and litter from urban centres, coupled with the relatively confined nature of freshwater bodies could counterbalance the increased exposure of biota in marine environments that are the final sink for these materials, albeit over a relatively much greater area [98].
Microfibres and non-plastic particles, however, account for a greater proportion of AMPs observed in animals from freshwater environments, despite their relatively limited coverage by research. This may be related to the proximity of terrestrial sources, as microfibres are a significant pollution source from domestic wastewater from the washing of clothes and flushing of wet wipes and sanitary products, as well as potential wastewater outflow [2, 8, 10, 29, 30]. The degradable nature of non-plastics could also account for their higher proportions in freshwater systems, as many may degrade during transportation and never reach the marine environment. It has been observed that cellulosic fibres more readily degrade in freshwater than brackish or marine water, due to the different microbial community assemblages, salinity, and water chemistry [101-103]. Despite this, our investigation shows non-plastic fibres are still a major component of brackish and marine environments. Fibre chemical structure and material finishes play key roles, with crystallinity and the presence of chemical additives increasing the resistance of fibres to degradation [104-106]. This allows for more resistant non-plastics to be deposited in oceans and estuaries, and non-plastic microfibres also show persistence in deep sea sediments in remote regions such as the poles, despite the literature suggesting the relatively rapid degradation of non-plastics [2, 58, 60]. Degradation studies mainly focus on controlled condition experiments, whereas understanding of the persistence of non-plastics in-situ in natural aquatic environments is currently limited. The loading of freshwater organisms is understudied and the life cycle of non-plastic microfibres from source to sink is also poorly understood.
Habitat
Of the specified groupings, benthic habitats consistently demonstrated higher levels of AMP loading, and microfibre and non-plastic particle proportions; followed by reef habitats and varying levels in demersal, pelagic and ben- thopelagic habitats. Benthic habitats are a known major sink for AMPs and microfibres, as their relatively greater density causes them to be deposited from the water column to the benthic zone, often intermixed with sediment or organic matter [9, 106]. Benthic sediments are considerable stores of AMP pollution, including plastic and non-plastic microfibres, with over 60% of deep-sea sediment AMPs being non-plastic microfibres [29, 58, 60, 107, 108]. Microfibres can also adhere to biotic and abiotic components within benthic habitats, including rocks and sedentary species, often aided by biofilms, where they can accumulate [109]. This increases the likelihood of exposure for species in these habitats, that often prescribe to deposit-, grazing- and filter-feeding strategies. Reef habitats are also a known AMP sink, due to the ingestion by and adhesion to coral species, deposition in reef sediments, and accumulation in coral skeletons [108, 110].
The habitats that encompass the water column generally demonstrated lower AMP loading but similar proportions of plastic and non-plastic particles. Surface waters of freshwater and marine systems are a major transport mechanism for AMPs and microfibres [1]. Non-plastic microfibres account for the majority of microfibres in ocean surface waters [3]. They have also been recorded as significant components of river water [29, 111]. Therefore, exposure of organisms to these anthropogenic microparticles can be high as they traverse the pelagic zone, especially when anthropogenic microparticles are mistaken for food sources, or are trapped in gills during inhalation [79, 80, 112-115].
Feeding strategy
When considering AMP loading, suspension feeders, grazers, and detritivores, demonstrated higher levels than herbivores, predators, and omnivores. However, when examining microfibre and non-plastic particle proportions, animals across feeding strategies demonstrated statistically similar values. Suspension feeders, detritivores and algal grazers are likely to be exposed to greater levels of AMP loading, as their feeding methods and focus on microscopic food particles may cause them to mistake AMPs for planktonic or detrital food sources [19, 20, 79]. The strategy of filtering the water column, sifting through organic matter and grazing biofilms all heighten the exposure risk of animals to AMPs as they ingest them, especially when AMPs are adhered to actual food sources [69, 79, 116]. The formation of biofilms on the surface of AMPs and microfibres also increases their likelihood of exposure to grazers and detritivores enticed by their organic coating [69, 79, 116, 117]. It has been observed that organisms at lower trophic levels are likely to exhibit higher levels of AMP loading per gram than their predators [22]. These organisms tend to prescribe to filter- and deposit- and algal-feeding strategies as they act as primary consumers and secondary producers.
Higher trophic organisms may display lower loading due to reduced exposure via feeding. Prey food sources act as the primary loading point for predators, transferring AMPs that they have ingested or are adhered to [109, 118], . Aquatic herbivore loading may be related to the adhesion of AMPs to the surface of aquatic macrophytes, or the ingestion of algal biofilms that colonise these food sources [119]. Due to the mixed feeding strategy of omnivores, their consumption of algae, plankton and organic matter may heighten their loading compared to predators. It has been suggested that AMPs could biomagnify in trophic systems as predators consume contaminated prey [120]. However, this investigation indicates the contrary, that bioaccumulation occurs within trophic levels but reduces across trophic levels, as has been suggested by other systematic reviews [22, 121, 122]. While AMP ingestion is prevalent, animals may be able to egest most AMPs, reducing the likelihood that they are transferred through trophic systems and retained [19, 20, 123]. Thus, the loading of AMPs in animals may be prevalent but transitory.
In the case of microfibres however, both plastic and non-plastic are prevalent across all trophic levels and feeding strategies, accounting for >90% of AMP loading in species belonging to every grouping except grazers and herbivores for non-plastics. Microfibres are often retained by organisms once ingested or inhaled, due to their morphology allowing them to become easily trapped or caught and persist in the digestive tract [83, 114, 124-127]. They can also resist egestion compared to other AMP types [19, 20, 113]. These factors could increase their likelihood of them being transferred from prey to predator [118]. Preferential ingestion by animals due to the similarity of the anthropogenic microparticles to natural, cellulosic food sources [79, 80, 112], would not explain the presence of microfibres in carnivorous organisms. Loading could also be influenced by an overall excessive loading of microfibres both plastic and nonplastic in aquatic environments, increasing the likelihood of exposure for all organisms [1, 31, 60, 108].
The potential ecological concerns and risks to aquatic organisms
The results of this study highlight species-level and context-dependent variability of responses to microfibre exposure, with organisms in certain cases responding similarly to plastic or non-plastic microfibres, and others responding more severely to non-plastic microfibres than plastic microfibres, or vice-versa. Some species exhibit no detected effects of these microfibres at all, others exhibited increased oxidative stress, reduced growth, altered behaviour, gut damage, altered enzyme activity and gut microbiome [52, 54-56, 79-84]. However, non-plastic microfibres often demonstrated sub-lethal and nonsignificant effects in lab experimentation, such as those observed by Walkinshaw et al. [52]. These results support the statements made by Parker et al. [128] in their meta-analysis of aquatic microfibre ecotoxicology, where they concluded that non-plastic and plastic microfibres caused context-dependent effects, with relatively greater numbers of non-significant effects associated with nonplastic fibres. Chosen microfibre exposure levels also influence detected results, as studies that demonstrated the most detrimental effects often used extremely large, environmentally unrealistic concentrations. For example, Suaria et al. [3] detected a maximum of 25.8 fibres L 1 in global marine water samples, whereas exposure studies often exceed 1000 fibres L 1 (e.g [54, 55, 83, 84]). While it is important to understand concentrations at which fibres cause detrimental effects, more studies at environmentally relevant concentrations are key to understanding their real-world impact [52, 129].
Another clear issue is determining whether microfibres act as a physical environmental hazard due to their morphology, or the chemical additives associated with their manufacture, or a combination of factors, regardless of their origin. Many studies do not differentiate between these factors; when they do, however, it is evident that chemical leachates from non-plastic microfibres and textiles can have detrimental effects on exposed organisms [53, 90-92]. As it is often not disclosed by manufacturers exactly what chemical additives are used during the manufacture of fibres, this acts as a significant barrier to advancing understanding [33]. Therefore, it is vital to establish what chemicals are used in the manufacture of plastic and non-plastic fibres, their leaching behaviour, and assess their toxicity and ecological impacts in realistic environmental contexts.
Due to their transport through wastewater treatment systems and being a component of aquatic pollution, microfibres, able to adsorb harmful components and harbour pathogens, could be an underestimated source of disease outbreaks and contaminant exposure [68, 95, 97]. However, this has only been examined a handful of times and it is also currently unclear as to whether plastic and non-plastic microfibres interact with contaminants and pathogens in similar ways. This aspect requires further research to fully understand and quantify.
Furthermore, the drivers of these ecological impacts have not yet been explored fully in aquatic systems, making it difficult to ascertain their true threat [128]. Greater representation in the literature is crucial in understanding whether emphasis needs to be placed on the regulation of the microfibre materials, or the chemical additives used by manufacturers, or both.
In conclusion, the ecotoxicological risks associated with non-plastic microfibres are understudied, with several indicators suggesting similar or even greater negative impacts than their plastic counterparts.
Recommended areas of future research
As discussed throughout this review, non-plastic microfibres are present in aquatic organisms across habitats and trophic systems, with an unclear and underexplored potential threat to the health and functioning of organisms and ecosystems. There are three key areas of research that we suggest being explored.
1.		Dedicated investigations to accurately assess the presence of both plastic and non-plastic anthropogenic microparticles in aquatic organisms are needed. These studies need to adhere to a standardised extraction and analysis protocol that considers the entire anatomy of the organism and maximises the recovery rate of particles and minimises degradation by avoiding harsh digestants and high temperatures. The presentation of results should also be standardised, with raw data availability for each species studied and their corresponding levels of loading. This would ensure accurate and representative reports of loading, identifying organisms and environments that are vulnerable, whilst allowing the data to be usable in other, large-scale investigations.
2.		Plastic and non-plastic microfibre loading affects biota across trophic levels, particularly those in benthic and demersal environments. Greater focus, therefore, is needed to quantify and understand loading of these pollutants across aquatic habitats and species. Particular focus should be placed on benthic organisms, as a known vulnerable group to AMP loading [22]. Freshwater and brackish environments also need greater research focus, as this review suggests species from freshwater
environments may have greater vulnerability to plastic and non-plastic microfibre exposure. The mechanisms or causes of this are not yet fully understood as data is limited.
3.		Research should also focus on the ecological impact and risk of microfibres and their associated chemical additives, and their interaction with the environment and biota, regardless of their material origin. As non-plastic microfibres have been relatively neglected, if they do cause negative ecological effects, their damage to aquatic environments may not be fully realised. Governments are attempting to tackle the plastic pollution challenge by encouraging nonplastic material alternatives such as natural or semi-synthetic fibres, as they are viewed as “environmentally friendly” and sustainable [88]. However, the scarcity of information on their environmental impacts makes it vital that future research focuses on non-plastic materials as well as plastic microfibres, to provide a sufficient evidence base for developingstrategies to mitigate the impact of all harmful anthropogenic microparticles. It is vital to understand the mechanisms of degradation of these materials in natural environments and processes such as ingestion and passage through tissues and the gastro-intestinal tract or their retention. The mechanisms for the release of additives, their persistence, and the real-world risks of organisms being exposed to them also needs to be explored to recognise the true threat of these materials.
Abbreviations
AMP	Anthropogenic microparticle
PCBs	Polychlorinated Biphenyls
PAHs	Polcyclic Aromatic Hydrocarbons
OSPAR	(Oslo and Paris conventions, also known as: Convention for the Protection of the Marine Environment of the North-East Atlantic)
Supplementary Information
The online version contains supplementary material available at https://doi.or g/10.1186/S43591-025-00153-6.
Supplementary Material 1
Supplementary Material 2
Supplementary Material 3
Supplementary Material 4
Author contributions
**DJJ: ** Conceptualisation, data curation, formal analysis, methodology, writing - original draft, writing - review and editing, visualisation. **EJO‘G: ** funding acquisition, supervision, writing - review and editing, formal analysis, visualisation. **DSG: ** funding acquisition, supervision, writing - review and
editing, formal analysis, visualisation. **RJC: ** supervision, writing - review and editing, formal analysis, visualisation. **AB: ** funding acquisition, supervision, writing - review and editing, formal analysis, visualisation. **TJT: ** funding acquisition, conceptualisation, supervision, writing - review and editing, formal analysis, visualisation.
Funding
This research has been partly funded by the UKRI Natural Environment Research Council through the ARIES DTP, as well as the University of East Anglia and the Centre of Environment, Fisheries and Aquaculture Science.
Data availability
All data generated or analysed during this study are included in this published article [and its supplementary information files].
Declarations
Competing interests
The authors declare no competing interests.
Received: 3 June 2025 / Accepted: 20 October 2025
Published online: 01 December 2025
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