A review of microplastic ﬁ bres: generation, transport, and vectors for metal(loid)s in terrestrial environments †

The laundering of synthetic fabrics has been identi ﬁ ed as an important and di ﬀ use source of microplastic (<5 mm) ﬁ bre contamination to wastewater systems. Home laundering can release up to 13 million ﬁ bres per kg of fabric, which end up in wastewater treatment plants. During treatment, 72 – 99% of microplastics are retained in the residual sewage sludge, which can contain upwards of 56 000 microplastics per kg. Sewage sludge is commonly disposed of by application to agricultural land as a soil amendment. In some European countries, application rates are up to 91%, representing an important pathway for microplastics to enter the terrestrial environment, which urgently requires quanti ﬁ cation. Sewage sludge also often contains elevated concentrations of metals and metalloids, and some studies have quanti ﬁ ed metal(loid) sorption onto various microplastics. The sorption of metals and metalloids is strongly in ﬂ uenced by the chemical properties of the sorbate, the solution chemistry, and the physicochemical properties of the microplastics themselves. Plastic – water partition coe ﬃ cients for the sorption of cadmium, mercury and lead onto microplastics are up to 8, 32, and 217 mL g (cid:1) 1 respectively. Sorptive capacities of microplastics may increase over time, due to environmental degradation processes increasing the speci ﬁ c surface area and surface density of oxygen-containing functional groups. A range of metal(loid)s, including cadmium, chromium, and zinc, have been shown to readily desorb from microplastics under acidic conditions. Sorbed metal(loid)s may therefore become more bioavailable to soil organisms when the microplastics are ingested, due to the acidic gut conditions facilitating desorption. Polyester (polyethylene terephthalate) should be of particular focus for future research, as few quantitative sorption studies currently exist, it is potentially overlooked from density separation studies due to its high density, and it is by far the most widely used ﬁ bre in apparel textiles production.


Introduction
The annual global total production of plastics exceeded 400 million metric tonnes (MMT) per year in 2015. 1 In 2015, the seven plastics with the highest demand in the EU (excluding bres) were polypropylene (PP) > low-density polyethylene (LDPE) > high-density polyethylene (HDPE) > polyvinyl chloride (PVC) > polyurethane (PUR) > polyethylene terephthalate (PET) > polystyrene (PS). 2 Plastics offer many advantageous properties such as corrosion resistance, low cost of raw materials and production, and general durability, meaning they have become increasingly favourable over traditional materials (metal, paper, wood) since the 1950s. 1 Plastics are now routinely manufactured for a wide variety of end uses, including food packaging, synthetic textile bres, building insulation, and protective coatings.The most commonly manufactured plastics are not readily biodegradable, and thus accumulate in the environment.The rate of plastic waste recycling to the original product is below 10%. 3 These factors, together with the continuously increasing global plastic demand at 8.7% annual growth rate, 1 have resulted in a global plastic pollution issue.
0][21] Published literature on the environmental impact microplastics has increased exponentially since 2010.However, the majority of studies have focused on the impacts to marine environments. 22From January 2004 to June 2018, only 4% of published literature on microplastics focused on terrestrial sinks such as soil and sludge. 22Consequently, current knowledge on the scale, environmental fate, and ecological impacts of microplastic pollution in terrestrial environments is limited.
This knowledge gap is concerning as it has been estimated that, in the EU, terrestrial environments could receive 4-23 times more microplastic pollution than oceans. 23It has also been estimated that up to 48% of this microplastic pollution is due to the direct application of contaminated sludge to agricultural soils. 24The shedding of textiles during laundering is thought to be a considerable source of microplastic pollution. 25ynthetic fabrics, such as polyethylene terephthalate (PET), and nylon (polyamide-6,6) shed recalcitrant, non-biodegradable bres during laundering; the majority of which are 0.20 to 2.75 mm in length. 25,26Between 72% and 99.9% of microplastics (by number) are removed during the sewage treatment process, 27 and approximately 78% are retained in the semi-solid sludge fraction. 28This sludge is collected and disposed of in a number of ways, including application to agricultural land, incineration or landll.Recycling to agricultural land is oen a preferred option across much of Europe, USA, and China Harrison Frost is a PhD student at the University of Surrey, studying the interactions between microplastics and pollutants in wastewater.He is interested in the environmental fate and behaviour of heavy metals and microplastics, particularly textiles bres.Prior to his PhD, Harrison completed a BSc in environmental science, and an MSc in environmental pollution at the University of Reading.
Dr Tom Sizmur is an Associate Professor in Environmental Chemistry at the University of Reading, UK.His research explores the biogeochemistry of potentially toxic elements and soil organic matter in natural and contaminated environments.He has undertaken research projects on the impact of earthworms on pollutant mobility, the biogeochemistry of mercury in intertidal mudats, and the use of soil amendments and cover crops to improve soil health.He is the Vice Chair of the Environmental Chemistry Group of the Royal Society of Chemistry and an Editor for Environmental Toxicology and Chemistry.
because treated sludge poses little risk to human and animal health, and contains essential plant nutrients and organic matter which improve soil fertility and physicochemical properties. 29Sludge use in agriculture is heavily regulated.Before application to agricultural land, sludge must be treated to reduce pathogen content, odour, and attraction of potential disease vectors such as rats.This treatment may be biological (e.g.aerobic or anaerobic digestion), chemical (e.g.lime stabilisation), or physical (e.g.thermal drying), or any combination of the three.Sludge that has been treated and stabilised for land application purposes is referred to as biosolids. 29,30n the UK, approximately 79% of municipal sewage sludge is recycled to agricultural soils as biosolids. 31In other European countries, application rates vary from 0-91%. 24Therefore, the application of biosolids may constitute a very signicant route for the entry of microplastic bres derived from laundered fabrics, into agricultural soils.This conceptual pathway is illustrated in Fig. 1 (Section 2).Moreover, during sewage treatment, synthetic bres are exposed to elevated levels of metals and metalloids, 32 and organic contaminants, including antibiotics, endocrinedisruptors and polycyclic aromatic hydrocarbons, 33,34 which may sorb to the surfaces of the bres.The study of the sorption of metals and metalloids to microplastics is still in its infancy.Nevertheless, metals including cadmium (Cd), copper (Cu), nickel (Ni), and lead (Pb) have been shown to sorb to both virgin and beached microplastic pellets. 35,36The bioavailability and environmental fate of microplastic-bound metals and metalloids is poorly understood.However, microplastics may act as vectors for metals and metalloids that would otherwise have been discharged in the effluent.Therefore, metal and metalloid sorption to microplastics may ultimately increase their exposure to important soil organisms such as earthworms. 37hroughout this review, microplastic particles (#5 mm in length) with brous morphologies are referred to as 'microplastic bres', and the uniform adoption of this term is suggested for future research.9][40] A denier is a unit of linear density equal to the mass in grams of 9000 m of bre. 38In environmental science however, the term 'microbre' has not been appropriately dened and is sometimes used to refer to only synthetic bres, or to synthetic and natural bres. 7,25,41,42We suggest that 'microplastic' and 'microbre' are not used interchangeably, and that the distinction between natural and synthetic bres should be retained, given that bres are likely to have different sources, environmental partitioning behaviours, and environmental impacts than natural bres or microplastics with other morphologies.
This review aims to (i) synthesise and critically evaluate recent qualitative research on fabric shedding as a source of microplastic bre pollution; (ii) provide a conceptual framework for, and comment on the environmental signicance of, the transfer of microplastic bres from laundry wastewater to agricultural soils through the application of sewage sludge as a soil amendment, and nally (iii) critically evaluate existing data concerning the sorption of metals and metalloids to microplastics.

Microplastic fibres from synthetic textiles
The laundering of synthetic textiles was rst evidenced as a diffuse source of microplastic bre pollution by Habib et al.
(1998), 14 who used polarised light microscopy to qualitatively identify synthetic bres in dewatered sewage sludge, biosolid pellets, and wastewater effluent from a wastewater treatment plant (WWTP) in Long Island, New York.These bres were hypothesised to come from the shedding of apparel textiles during laundering (Fig. 1a). 14Shedding refers to the detachment and release of bres from the surface of the fabric and primarily occurs during laundering, where the rotational force and mechanical action of the washing machine drum cause bres to break and enter the water. 42Zubris and Richards (2005) 19 detected synthetic bres in soil samples up to 15 years aer the application of biosolids, implicating sludge disposal to soil as an important pathway for the terrestrial transport of bres.Browne et al. (2011) 4 found that sediments from sewage sludge disposal sites and wastewater treatment plant effluent contained proportions of synthetic bres which resembled those used in apparel textiles (78% polyester, 9% polyamide, 7% polypropylene, and 5% acrylic), 43 suggesting that the bres were largely derived from the shedding of clothing during laundering.
7][48] Fabric shedding is also inuenced by laundering conditions, such as washing machine type (front-loading or top-loading), water temperature, and the presence of surfactants and fabric soeners. 49,50e highlight 25 research papers, published between 2011 and 2021, that report quantitative data on fabric shedding and summarise key data in Table 1. 4,25,41,42, Quanifying the number of bres shed from a fabric during laundering is practically difficult due to the small size and vast numbers of bres generated, and the lack of standardised methodologies.Zambrano et al. (2019), 25 Haap et al. (2019), 58 Kelly et al.
(2019), 61 Frost et al. (2020), 48 Raja Balasaraswathi and Rathinamoorthy (2021), 65 Cai et al. (2020), 52 and Özkan and Gündogdu (2021), 63 all used standard laundry testing apparatus, while the remaining studies used commercially available washing machines to generate bres.Fabric shedding varied from 900-110 000 bres per garment, 4,42 although the sizes and masses of the garments were unspecied.On a number per mass basis, bre shedding ranged from 8809-72 000 000 per kg of fabric, 52,66 and on a mass per mass basis, 7-1507 mg bres per kg of fabric. 54,65he extreme range in literature values highlights the need for standardisation in the quantication of fabric shedding during laundering to make meaningful comparisons between fabrics.Pirc   66 found shedding propensity signicantly increases with fabric thickness.This is thought to be due to an increased density of bre ends per unit of surface area. 65A higher stitch density (number of stitches per unit area) results in less bre release, as friction between constituent bres is increased. 65Dalla Fontana et al. (2021) 54 compared the shedding of two 100% polyester fabrics with differing constructions, and observed signicantly different bre release during conventional laundering experiments.Differences were attributed to the differing linear densities of the constituent bres, which inuences tensile properties, and the stitching used to nish the fabric edges. 54Fabrics composed of natural bres, such as cotton, generally have a higher shedding propensity then fabrics constructed with synthetic bres, such as polyester. 25This may be due to the lower tensile strength of natural bres, or the shorter bre length of cotton, resulting in more bre breakages overall. 25,66It is important to note that chemical identication of the shed fabrics with spectroscopic techniques (IR/Raman), was only performed in 5 of the 25 studies (Table 1).In the remaining studies, bres were counted and/or weighed without conrming their composition.10 studies investigated the shedding of fabrics with mixed bre compositions, and of these, only 3 conrmed the chemical composition of the shed bres with FTIR.Chemical conrmation of shed bres is of particular importance in studies using mixed composition fabrics, so that the relative proportions of shed bres can be assigned to each bre composition.Haap et al. (2019) 58 investigated the shedding of bres from a 50% polyester, 50% cotton woven fabric.Aer quantication, chemical separation was performed by using sulphuric acid to digest the cotton bres.1 found that increasing the water volume in accelerated laundering experiments, from 300 mL to 600 mL, resulted in an increase in bre shedding, from 54 mg kg À1 to 120 mg kg À1 .Several studies in Table 1 do not report the total volume of water used, and the fabric weight, density and surface area.It is recommended in future that these parameters are quantied and reported, as they also inuence shedding propensity.
Microscopy and manual or computational counting from micrographs was by far the most common method for bre characterisation, used in 18 of the 25 studies included in Table 1.However low spatial resolution of optical microscopes and image analysis techniques mean that underestimations are likely.McIlwraith et al. (2019) 62 used ImageJ (image processing soware) to quantify bres from a series of micrographs.However, the limit of detection was 100 mm in length, resulting in the exclusion of bres <100 mm.Hernandez et al. (2017) 50 also analysed micrographs with ImageJ; ascertaining a 40 mm limit of detection from the minimum visible number of covered pixels in each micrograph (between 2 and 5 pixels).In this study, it was found that shed bres from a PET single-knit jersey and interlock fabrics were typically 100-800 mm in length.However, size distributions revealed a general increase in bre frequency as bre length decreased.Moreover, bres above 1 mm in length represented only 2-5% of the total shed bre    25 for bres shed from PET, cotton, and PET-cotton blend fabrics also revealed this trend -bre frequency increased as bre length decreased, until the lower limit of detection for bre length (200 mm) was reached. 25Since the widths of synthetic bres are usually very uniform and have a typical mean diameter of 11-16 mm, 25,46 it could be reasonably assumed that the majority of shed bres would be captured by a lter with a pore-size of 10 mm or below.In Table 1, 9 of the 25 studies used lters with a pore-size above 10 mm, meaning bres may have been lost even before analysis.Furthermore, it is important to highlight that only six studies provided a lower limit of detection for the measurement of bre length.This is an essential parameter to evaluate the suitability of the methodology, and to contextualise results in ecotoxicology, since bre size inuences environmental fate and transport, 67 ingestion rates by organisms, 68 and specic surface area. 69he laundering method and apparatus used also likely contributed to the variation in shedding rates reported in the literature.Laundering speed varied from 40-1600 rpm, 50,61 and was unspecied in 10 of the 25 studies listed in Table 1.Laundering speed is thought to greatly inuence bre shedding because it determines the mechanical action exerted on the fabric. 42,49Accelerated laundering methods were used in 9 studies, whereas commercial, or laboratory-scale washing machines were used in the other 8 studies.Accelerated laundering refers to the laboratory-scale simulation of home laundering, by placing the fabric and water/detergent solution under continuous agitation, oen with the addition of metal beads to increase mechanical abrasion.This allows multiple experimental treatments and replicates to be performed simultaneously and reduces the total volume of water to be ltered and analysed.Cai et al. (2020) 52 reported that up to 72 000 000 bres per kg of fabric could be released, but accelerated laundering was used in their study.Moreover, Zambrano et al. (2019) 25 found accelerated laundering generated approximately 40 times more bres per unit mass of fabric, compared to conventional washing machines.This likely accounts for some variation in shedding rates between studies adopting conventional and accelerated laundering, and highlights that results from accelerated laundering studies should not be used to estimate bre emissions during home laundering.
Control strategies at various levels of intervention have been proposed to reduce microplastic bre emissions (Ramasamy and Subramanian, 2021). 70At the individual level, commercially available capture devices may be used during laundering to reduce microplastic bre emissions at the source.McIlwraith et al. (2019) 62 compared the microplastic bre reduction efficiencies of two such devices; the Cora ball (a plastic ball with hooked arms) and the Lint LUV-R (retrotted lter), when laundering a 100% polyester eece.Where no device was used, 4800 bres per litre were released.Fibre release decreased to 3580 bres per litre with the use of the Cora ball, and only 648 bres per litre with the Lint LUV-R lter, 62 representing bre capture rates of 25.4% and 86.5% respectively.Napper et al.
(2020) 71 tested the efficiency of several capture devices, nding the XFiltra retrotted lter to be the most effective, reducing bre emissions by 78% compared to a control where no devices were used.This was attributed to the ne pore-size of the lter (60 mm).However, two mesh bags, the Guppyfriend and the Fourth Element washing bag, resulted in reductions of 54% and 21% respectively, despite both having a mesh pore-size of 50 mm, so other design variables appear to inuence the efficiency of these devices. 71Using lower water volume laundering cycles has also been shown to signicantly reduce the mass of shed bres. 61Fabric production may also be altered to reduce bre shedding propensity.Generally, bres with higher tensile strength and tenacity result in yarns with a lower hairiness (number of protruding loops and ends), meaning the nal fabric has a higher abrasion resistance and therefore a lower shedding propensity. 25,44Increasing the number of yarn twists per unit length, and the stitch density of fabrics can also decrease their shedding propensity. 42,65Fleece fabrics, which are mechanically cut aer construction, have an increased shedding propensity compared to similar, non-eece knitted fabrics. 42Current legislation aimed at reducing microplastic bre emissions is non-existent in most countries. 70,72Laws implemented in New York, and California, state that clothing containing more than 50% synthetic bres must be labelled as a contributor to microplastic bre pollution, aim to impart consumer knowledge to facilitate gradual consumer behavioural changes. 70

Microplastics in sewage sludge
Shed bres are typically carried through municipal drainage systems to a wastewater treatment plant (WWTP) (Fig. 1b).
Processes of wastewater treatment vary between facilities, however they typically begin with primary screening and sedimentation to remove coarse grit and suspended solids.This stage is followed by secondary treatment, which involves aerobic or anaerobic microbial incorporation to remove suspended or dissolved organic matter, oen aided by the addition of occulants in a secondary sedimentation tank.Secondary treatment usually also involves a disinfection stage to remove pathogens. 17Tertiary treatment involves additional specialised mechanisms to improve effluent quality before discharge into the environment, such as additional ltration or the removal of nitrates and phosphates. 73The solid residue, or sewage sludge, is collected and typically dewatered to reduce its volume, before being chemically, aerobically or anaerobically stabilised. 74,75In the UK, the majority of sewage sludge is then utilised as a fertiliser (biosolids) in agriculture (79%), incinerated for energy recovery (18.4%), or disposed of in landll (0.6%). 32Throughout Europe, biosolids application rates to land vary greatly between countries, from over 90% in Ireland and Lithuania, to less than 5% in the Netherlands, Slovenia and Malta. 76][79][80] Several studies have quantied microplastic contamination through the different stages of the wastewater treatment process at specic WWTPs.Comparison of microplastic concentrations in the inuent to the effluent has revealed that 72-99.9% of microplastics (by number) are removed during the wastewater treatment process. 81,82For example, Murphy et al.
(2016) 16 sampled concentrations of microplastics at four progressive stages of wastewater treatment, at a WWTP in Scotland, nding that microplastic concentrations decreased from 15.7 microplastics L À1 in the inuent, to 0.25 microplastics L À1 in the nal effluent; a reduction of 98%.Microplastics were identied visually using a dissection microscope and characterised using FT-IR (Fourier Transform Infrared) spectroscopy, although specic details of spatial resolution and limits of detection for size were not reported.It has been estimated that approximately 78% of microplastics (by number) entering a WWTP are removed during wastewater treatment and are present in the sewage sludge. 29Primary wastewater treatment has been shown to remove 92-93% of textiles bres (both natural and synthetic bres), with secondary treatment resulting in only a further 0.2% reduction. 83In the same study, microplastics removal aer primary treatment was 32%, with 76% of remaining microplastics being removed aer secondary treatment.
While there is some published data on the fate of microplastics in WWTPs, there is a lack of information concerning the fate of microplastic bres specically.The differing behaviours of textiles bres and microplastics may be explained by differences in density, which affects the settling velocity of the particles. 84Polyethylene terephthalate (polyester) is the most widely used synthetic polymer in the manufacturing of apparel textiles, 44 and has a density of 1.32-1.41g mL À1 , 85 which is higher than that many other polymer types commonly identi-ed in sewage, including polyethylene (r ¼ 0.89-0.97g mL À1 ) and polypropylene (r ¼ 0.85-0.92g mL À1 ). 86It is yet unclear how microplastic morphology inuences wastewater separation efficiency, although for spherical microplastics, a larger diameter will increase settling velocity. 84Nonetheless, methods of sampling, microplastic extraction from sludge, and characterisation are still currently in development and oen differ widely between studies.
Chemical oxidation is oen employed to digest organic matter, allowing an easier separation of microplastics from the solid sewage sludge fraction. 27It also aids the removal of any organic substances impregnating/coating the surfaces of the microplastics, which would hinder the spectral characterisations and chemical classication with techniques such as FT-IR or Raman spectroscopy.Mason et al. (2016)  17 used 30% hydrogen peroxide (H 2 O 2 ) to digest sludge samples from 17 WWTPs, before analysing microplastics visually under a dissection microscope.The dominant shape fraction (bres, fragments, lms, pellets, foams) in this study was found to be bres; accounting for 46% and 80% of total microplastics in the 125-355 mm and >355 mm size fractions respectively. 17However, sampling bias and misidentication of microplastics can occur where only visual identication is employed. 27Talvitie et al.
(2017) 28 found that only 34% of bres separated from sludge were composed of synthetic polymers (PET -33%; polyacrylic -1%) when analysed with FT-IR, with the remainder being natural bres such as cotton, or regenerated bres such as rayon.Eriksen  Nylon is a common synthetic bre used in apparel textiles, 44 and the shedding propensity of nylon garments during laundering is comparable to that of polyester (Table 1). 42,49,51It is possible, therefore, that the relative proportion of nylon bres is underestimated when H 2 O 2 is used as a chemical oxidant in sludge processing.Other methods used for processing of microplastics from environmental or biological samples include ultrasonic extraction and the use of alkaline (e.g.KOH or NaOH) dissolution and/or acid (e.g.][92][93][94] However, these methods have not been tested for the analysis microplastic bres in sewage sludge.Microplastic extraction from sludge in more recent studies typically involves a density separation step to remove inorganic debris such as sand, grit and aluminosilicate minerals. 27This separation involves the agitation and prolonged settling of the sludge matrix in a high density, saturated salt solution, such as sodium chloride (NaCl) (r ¼ 1.2 g mL À1 ) or zinc chloride (ZnCl) (r ¼ 1.6 g mL À1 ), 85,94 or mixtures of water, sucrose, and ethanol. 95Plastics typically have a density of 0.89-1.2g mL À1 .However, PET and PVC can have densities up to 1.41 and 1.70 g mL À1 , respectively. 86Li et al. (2018) 96 extracted microplastics by density separation from sewage sludge from 28 WWTPs in China using saturated NaCl solution, followed by H 2 O 2 oxidation to digest remaining organic matter.Microplastics were visually sorted by morphology and subsamples were qualitatively analysed by FT-IR and SEM microscopy.Li et al. (2018)  96 identied between 1565 and 56 386 microplastics per kg of sludge (37 mm-5 mm in size), 62.5% of which were bres.The total number, and proportion of brous microplastics were potentially underestimated in this study due to the higher density of PET (r ¼ 1.32-1.41g mL À1 ) than the NaCl solution (r ¼ 1.2 g mL À1 ). 85This underestimation is a signicant limitation of the methodology because PET is by far the most widely used synthetic bre in apparel textiles production. 43During quality control experiments by Liu et al. (2018), 20 soils were spiked with microplastics including polypropylene (PP), polyethylene (PE), nylon (polyamide), PET and PVC, before density separation with NaCl.Recovery of both PET and PVC (r ¼ 1.32-1.7 g mL À1 ) 86 from the soil samples was 0%, highlighting the importance of using higher density salts such as ZnCl (r ¼ 1.6 g mL À1 ). 94here is general agreement between studies investigating the mass balance of microplastics entering WWTPs that microplastics are removed from wastewater very effectively during treatment.Carr et al. (2016), 15 Magnusson and Norén (2014), 82 Gies et al. (2018) 97 and Leslie et al. (2017) 81 reported removal rates of 99.9%, 99.9%, 97-99%, and 72%, respectively, by quantifying the number of microplastics in the inuent, effluent and sludge.][100][101][102][103][104][105][106] Despite practical limitations of microplastic extraction from solid media such as sludge and soils, reported concentrations in sewage sludge samples have ranged from 1-56 386 microplastics per kg of sludge. 17,19,82,96,107Their small size and high specic surface area, coupled with the consistently elevated concentrations of metals and metalloids in sludge, means that sorption of metals and metalloids on microplastic surfaces may occur during the wastewater or sludge treatment processes.

Metal and metalloid sorption onto microplastics
9][110][111] Many studies have quantied the sorption of a variety or organic pollutants such as polycyclic aromatic hydrocarbons (PAHs), 112 antibiotics, 113 and phthalate esters 114 onto microplastic surfaces, and this topic has been extensively reviewed. 115,116owever, the sorption of metals and metalloids to microplastic surfaces is far less studied.Furthermore, the sorption of metals and metalloids to microplastic bres has rarely been studied, and data are extremely scarce.As a result, inferences on the sorption of metals and metalloids to microplastic bres need to be made, based on observations made on the sorption of metals and metalloids to microplastic pellets (Table 2).There are several limitations to this inference that should be acknowledged.The surface area to mass ratio of microplastic bres is likely to be greater than pellets.Therefore, a fairer comparison of the sorption of metals and metalloids to microplastics should be based on mass of adsorbate to surface area of adsorbent.However, because the surface area of microplastics is rarely reported in sorption experiments, this is currently not possible with the data available in Table 2.
Reported plastic-water partition coefficients (K d values), representing the distribution of the respective metal or metalloid between the plastic-bound phase and the aqueous phase, range widely between metal or metalloid type, plastic type, plastic aging, and aqueous matrix type (Table 2).For example, the partitioning of aqueous Cu to virgin plastics decreases in the order PVC > PS > PE, 36,118 and is higher in seawater than in freshwater. 35,36Cr is sorbed more strongly to aged, rather than virgin, PE microplastics by an order of magnitude. 35Cu sorption was not signicantly affected by plastic age in seawater whereas, in freshwater, sorption was greater on the aged, rather than virgin, PE microplastics. 35 Adsorption of metals to microplastic pellets under various conditions (K d ¼ plastic-water partition coefficient; PS and aged PS and PE microplastics, nding aged microplastics to have a higher sorption capacity (Table 2).This was partly attributed to the increased specic surface area, and more negative zeta potentials of the aged microplastics. 1209][120] Previous studies investigating metal and metalloid sorption onto microplastics have mainly focussed on microplastic types commonly found in marine environments, such as PE, PS and PVC.Data for metal and metalloid sorption onto PET microplastic surfaces are scarce in the literature in comparison with other polymer types.Cojocariu et al. (2017) 122 quantied the sorption of Pb and Cu to recycled PET bres, reporting Pb and Cu sorption to be 48.9 mg g À1 and 30.7 mg g À1 respectively.These sorption values are several orders of magnitude higher than others reported thus far in the literature, 35,36,[118][119][120] although it should be noted that a physicochemical characterisation of the PET bres was not provided in this study, so bre size and specic surface area are not known.Recently, Han et al. (2021) 123 investigated the sorption of Pb, Cu and Cr onto PET microplastics of varying size fractions (<0.9 mm; 0.9-2 mm; 2-5 mm).Sorption increased in the order Cr < Cu < Pb, and decreased with increasing microplastics size.Kinetics experiments revealed that at equilibrium the sorbed concentrations of Pb, Cu and Cr on the PET bres, calculated by tting the pseudo-second order model, were 1.04, 0.488 and 0.385 mg g À1 respectively.These values are similar to those previously reported for virgin and beached PE pellets. 35,36ET is oen underrepresented in density separations due to its high density, 85 yet is the most commonly used material in the manufacturing of apparel textiles, 43 which have been shown to shed large quantities of bres into the wastewater system (Table 1).The laundering of synthetic textiles therefore represents a considerable diffuse source of PET microplastics into sewage sludge where they are exposed to elevated concentrations of metals and metalloids (Table S1 †).Unlike the previously investigated plastic types (PE, PVC and PS), PET contains hydroxyl (-OH) groups at the ends of the polymer chains, which are approximately 100 monomer units in length. 124These hydroxyl groups may increase the sorption capacity of PET for metals and metalloids, as they become deprotonated at circumneutral pH, which is oen characteristic of sewage sludges. 125Deprotonated hydroxyl groups (-OH-) are negatively charged and therefore may facilitate the sorption of cationic metals, such as Cu 2+ , Pb 2+ , and Cd 2+ . 126,127Zou et al. (2020) 117 found that the sorption capacity of chlorinated-PE for Cd, Cu and Zn, was at least one order of magnitude higher than HDPE and LDPE microplastics, which was partly attributed to the high electronegativity of chlorine.The high electronegativity the oxygen atoms in PET may therefore increase its sorption capacity, particularly for cationic metals.Further study of PET is required to understand the role of its unique functional groups in the sorption of metals and metalloids.
Environmental water samples oen contain humic substances; high molecular weight organic macromolecules, with heterogeneous branching structures and oxygen- View Article Online containing functional groups, 36,126 which may also inuence the sorption of metals and metalloids onto microplastics.Li et al.
(2019) 128 reported decreased sorption of Cd to PP and PE microplastics with increasing concentrations of humic substances.This observation was attributed to the complexation of Cd by the acidic functional groups of the humic substances, and the subsequent decrease in free Cd 2+ ions in solution. 128Turner and Holmes (2015), 36 however, hypothesised that the observed increase in the sorption capacity of beached PE pellets over virgin PE pellets could be in part due to the gradual accumulation of organic matter on the surfaces of the beached pellets.It was suggested that this organic matter provided an increased number of charged functional groups, and therefore increased the sorption capacity of the plastic surface. 36These conicting results are difficult to resolve, because the concentrations of humic substances in the water samples used by Turner and Holmes (2015) 36 were not quanti-ed.Biolm formation may also alter the sorption capacity of microplastics over time.Biolms are formed by the extracellular polymeric substances, such as polysaccharides, of colonising bacteria on surfaces.Biolms may change the surface properties, and therefore sorption capacity of microplastics by introducing new functional groups, such as amines, hydroxyls, and carboxylic acids. 129Biolms may also enrich the microplastic surfaces with particular elements, for example strontium and sulphur, present in the radiolarian protozoa Acantharea. 130,131he accumulation of metals and metalloids to microplastics represents an important knowledge gap, and more work is necessary in order to provide a more systematic quantication of the retention of metals and metalloids onto microplastic surfaces in general, and microplastic bres in particular.Nonetheless, the current data show that microplastics are capable of sorbing considerable amounts of metals and metalloids which may in turn inuence the biogeochemical cycling of these metals and metalloids in agricultural soils when the microplastics are applied in biosolids.

Sewage sludge application and impacts to terrestrial environments
Sewage sludge, the solid residual product of wastewater treatment, contains elevated concentrations of microplastics, and is widely applied to agricultural soils as biosolids.Using density separation, followed by ltration and visual identication with a stereo microscope, Corradini et al. (2019) 21 quantied microplastic concentrations in agricultural soils with differing biosolids application rates.The median estimated mass of microplastics in the soil increased with each successive biosolids application, from 1.37 to 4.38 mg kg À1 in soils that had received 1 and 5 biosolids applications, respectively, and 97% of these microplastics were categorised as bres.1 provide strong in situ evidence that the application of sewage sludge derived biosolids provides an important pathway for the transfer of microplastics, particularly microplastic bres, to agricultural soils (Fig. 1d).
Like most other widely manufactured polymers, PET is highly resistant to degradation under typical environmental conditions, however, is susceptible to hydrolytic degradation due to its ester bonds, particularly at extreme pH values. 86,134ver time, water reacts with the ester bonds of the PET backbone, forming two shorter polymer chains ending in alcohol and carboxylic acid groups.UV radiation from sunlight exposure will also initiate photooxidation of PET, where the ester bonds are cleaved, leaving carboxylic acid groups on the polymer surface.Degradation products include CO, CO 2 , terephthalic acid and other carboxylic acids. 134Environmental exposure of PET can signicantly decrease tensile strength aer approximately one year, depending on UV intensity, temperature, and precipitation. 135Shape also inuences degradation rate, as highlighted by Chamas et al. (2020), 134 who estimated that HDPE lms, bres and spheres weighing 2.75 g each, will take approximately 1.8, 465 and 2000 years respectively to completely degrade.Microplastic bres are likely to degrade in the environment more rapidly than lms, but let rapidly than spheres, as degradation is controlled largely by the size of the surface area exposed. 134Biological degradation of PET bres is typically slow.Zambrano et al. (2019) 25 assessed the biodegradation of various yarns, by measuring the total oxygen demand in a closed respirometer over 243 days, aer inoculation with aerobic microorganisms from activated sludge.Biodegradation of the polyester yarn was only 4.1% compared with 75.9% for the cotton yarn, revealing that PET is a relatively unavailable carbon source for microorganisms. 25Yoshida et al. (2016) 136 isolated a bacterium, Ideonella sakaiensis 201-F6, that was capable of biodegrading PET by adhering to its surface and secreting two hydrolytic enzymes.The bacterium was capable of almost completely degrading a 60 mg PET lm aer 60 days, and catabolised 75% of the total carbon aer 15 days. 136The main biodegradation product was mono(2-hydroxyethyl) terephthalic acid, which was rapidly metabolised further into the two monomers of PET, terephthalic acid and ethylene glycol. 136,137The co-polymer poly(ethylene terephthalate-colactate) can be effectively degraded under laboratory composting conditions, 138 although further research is needed before bioremediation strategies are implemented for PET bres in sludge and soils.
The impacts of microplastics on agroecosystems are still relatively unknown.However, emerging research is indicating that microplastic contamination may inuence the physical, chemical, and biological properties of soils.Machado et al.
(2018) 139 reported a reduction in bulk density, and an increase in water holding capacity for soils spiked with polyester bres, that was not observed in soils spiked with polyethylene fragments and polyamide beads.This observation is thought to be due to the more efficient entanglement of soil aggregates by the bres, creating more air-lled pore spaces in the soil.Kim and An (2019) 140 observed marked behavioural changes and a signicant reduction in mobility of springtails in soils due to the presence of PS microplastics, even at relatively low concentrations (8 mg kg À1 ).PS nanoplastics at concentrations of 10 mg L À1 were shown to induce toxicity in nematodes (Caenorhabditis elegans). 141Earthworms (Lumbricus terrestris) were observed ingesting PE microplastics (150-2800 mm in diameter) in mesocosm experiments, 67,142,143 however the ecotoxicological implications of this were not studied.
Sewage sludge contains elevated concentrations of metals and metalloids (Table S1 †), which have been shown to sorb onto microplastic surfaces (Table 2).In the UK and EU, legislation stipulates the maximum concentrations of metals and metalloids such as Cd, As, Hg, Pb and Zn in agricultural soils and maximum permissible loadings in sludge applications (Table S2 †).However, there are currently no regulatory limits on microplastic additions to agricultural soils.As a result, the occurrence and accumulation of microplastics in agricultural soils has gone largely unmonitored in recent years.In the EU, it is estimated that up to 4.7 times more microplastics per year are released into agricultural soils than to surface waters, 24 despite terrestrial microplastics research continuing to lag behind marine and freshwater microplastics research. 22etals and metalloids sorbed onto the surfaces of ingested microplastics may have an altered bioavailability and bioaccessibility to soil organisms, compared to dissolved ions.Pollutants sorbed onto the surfaces of ingested microplastics may readily desorb in acidic gut environments. 37,144,145Microplastics may therefore act as a vector, facilitating an increased exposure of the sorbed pollutant to the organism (Fig. 1e). 144For example, desorption of persistent organic pollutants from PE and PVC microplastics was shown to be up to 30 times greater under simulated gut conditions, than in seawater alone. 145ynthetic earthworm gut extraction tests of HDPE microplastics and soil aggregates by Hodson et al. (2017), 37 revealed that Zn desorption was 4-30 times greater from the microplastics than from the soils.Despite this, in the same study, Zn-loaded microplastics induced no statistically signicant effects in Zn bioaccumulation, weight or mortality on earthworms.Exposure studies have shown that ingestion can facilitate the fragmentation of microplastics, creating smaller particles.In a study by Kwak and An (2021), 146 earthworms (Eisenia andrei) were exposed to polyethylene microplastics (180-300 mm diameter) for 21 days, before microplastics were extracted from earthworm intestines and casts.Using SEM, the researchers identi-ed smaller microplastics on the surfaces of the ingested and excreted microplastics, measuring as small as 182 nm in diameter. 146This is thought to be primarily due to microbial degradation by bacteria in the gut microbiome of earthworms. 147Similar results were reported by Dawson et al.  (2018), 148 who observed that microplastics ingested by Antarctic krill (Euphausia superba) were on average 78% smaller in diameter (7.1 mm) than the original microplastics (31.5 mm).Ingestion of microplastics by soil organisms may therefore result in an increase in the specic surface area of the microplastics as they are fragmented.Experiments by Khan et al.
(2017), 144 using polyethylene microplastics loaded with silver (Ag), revealed that desorption of Ag was signicantly higher under lower pH conditions, reaching 98% aer 3 hours at a pH of 4.1.However, such experiments have not been performed with polymers commonly used in textiles, such as PET, nylon, and polyurethanes.These polymers are likely to contribute signicantly to the microplastics entering agricultural soils that receive municipal sewage sludge.

Conclusions
Microplastics are ubiquitous, highly recalcitrant emerging contaminants of concern.Current data suggest microplastics, particularly bres, are pervasive in sewage sludge and agricultural soils.Further work is required to develop more robust, efficient, and consistent analytical procedures to quantify their abundance.The laundering of synthetic bres provides an important diffuse source of microplastic bres, which are transported into wastewater systems.Microplastic bres from laundering have a typical diameter of approximately 20 mm.An inverse relationship between frequency and length is oen observed.Values reported in the literature for the number of bres shed during conventional laundering experiments are 1900-110 000 bres per garment, 22 600-13 100 000 bres per kg of fabric, and 7-1240 mg of bres per kg of fabric.During wastewater treatment, 78% of microplastics are retained in the sewage sludge.Municipal biosolids can reportedly contain up to 56 000 microplastics per kg, although values reported in the literature vary greatly.This sludge is commonly applied to agricultural land as a soil amendment, creating a pathway for microplastic bres generated during laundering to contaminate agricultural soils.Therefore, microplastic bres may cause deleterious effects to the soil biophysiochemical environment which require further investigation.
The surfaces of microplastics may also be loaded with elevated concentrations of metals and metalloids due to sorption during the wastewater treatment process.Research on the sorptive properties of microplastics is still in its infancy.However, available data reveals that sorption of metals and metalloids is largely dependent on solution chemistry (pH, ionic strength, presence of humic substances), and the physicochemical properties of the microplastics (functional groups, degree of surface oxidation, specic surface area, point of zero charge).Therefore, the application of sewage sludge to agricultural land may represent an important vector for metals and metalloids, sorbed on the surfaces of microplastics, to soil organisms.
Microplastic bres are an important subcategory of microplastics, due to their prevalence in sludge and agricultural soils, and high specic surface area.However, the sample processing method used to isolate microplastics from soils and sludge should be carefully considered, particularly during density separation to avoid excluding high-density microplastics, such as PET, from being quantied.Research on the sorptive properties of synthetic polymers commonly used in textiles, such as PET and poly(amide), is extremely scarce, representing an important knowledge gap.This gap requires immediate attention because textiles-derived microplastic bres are likely to represent a signicant proportion of microplastics applied to agricultural soils, in biosolids produced from municipal sewage sludge.

Fig. 1
Fig. 1 Conceptual pathway visualising the transport of microplastic fibres from laundry effluent to agricultural soils through the production and application of sewage sludge (Me ¼ metal(loid) ion).
Tang et al. (2021) 119 investigated the sorption of Cu, Ni and Zn onto nylon microplastics collected from the environment.Langmuir modelling of isotherm data revealed maximum sorption capacities for Cu, Ni and Zn were 16.7 mg g À1 , 10.6 mg g À1 and 12.7 mg g À1 respectively, however, the data are not reported for the virgin nylon microplastics.Chen et al. (2021) 120 compared the sorption of Cu onto virginTable 2 56,57ndez et al. (2017), 50 Almroth et al. (2018), 42 De Falco et al. (2018)56,57 et al. (2016) 47 reported among the lowest number of bres per mass of fabric, but also used the largest lter size (200 mm).Zambrano et al. (2019) 25 reported that shed bres 25-200 mm in length were more numerous than bres 200 mm-2.75mm, shed from PET, cotton, and PET-cotton blend fabrics, suggesting that the shedding data from Pirc et al. (2016) 47 may have underestimated total bre release, due to smaller bres passing through the lter.Raja Balasaraswathi and Rathinamoorthy (2021) 65 and Vassilenko et al. (2021) Kelly et al. (2019)

Table 1
Review table of fabric shedding data from laboratory studies.Fabric type, fibre shedding, and general experimental parameters are shown.NS machine Speed unspec.; 30 min; temp.Unspecied; 43 L capacity, front-load and top-load machines; articially aged and new garments; Â 20 cm swatches; no detergent; with bio/ non-bio detergent; with/without Spain 1000 rpm; 15 min; ambient temp.; 22 L water, commercially available detergent used as specied; water hardness of 349 mg L À1

Table 1 (
Contd. ) C; various wash cycles including cotton short cycle, cold express cycle and delicate cycle; 30-69 L; with/ without 35 mL of commercial liquid detergent; 1.5 kg of T-shirts 89 al.(2013) 87 and Lenz et al. (2015)88found 20% and 32% of particles were visually identied erroneously as SEM) and Raman microscopy, respectively.Mason et al.(2016)17may therefore have overestimated the presence of microplastic bres in sewage sludge samples, as only visual identication based on morphology was employed.The use of H 2 O 2 may also result in the oxidation and subsequent destruction of polymers such as nylon in the digested sample.89
As highlighted by Corradini et al. (2019) 21 microplastic morphologies were categorised at the operator's discretion, further necessitating a need for a strict and appropriate denition of microplastic bres.Nevertheless, Corradini et al. (2019)