2. A typology of microplastics released from textiles and tyres
Microfibre shedding and tyre wear are processes regularly occurring during the use of textile products and vehicle tyres. As textiles are worn and washed, mechanical abrasion occurring in their structure causes the detachment and loss of fibres. Similarly, during normal transport activity, the friction between vehicle tyres and the road surface results in the abrasion of the tyre tread and the emission of particles. In general, the emission of microfibres and tyre wear particles may occur during (and be influenced by) all stages of the lifecycle of products, as summarised in Figure 2.1.
This Chapter summarises current knowledge on the characteristics, environmental fate, and environmental and human health impacts of textile-based microfibres (Section 2.2) and tyre-based microplastics (Section 2.3). Where the data is available, it provides an assessment of where in the lifecycle of products emissions occur and which are the key influencing factors.
Source: Author
The textile industry is considered one of the most polluting in the world. Harmful chemicals, high-energy use, water consumption, textile waste generation, transportation and the use of non-biodegradable packaging materials are responsible for the resource heavy and polluting lifecycle of textiles and clothing. The European Environment Agency (EEA) (2019[1]) estimates that, in the EU, supply chain pressures of clothing, footwear and household textiles are the fourth highest pressure category for the use of primary raw materials and water, second highest for land use and the fifth highest for greenhouse gas emissions. Overall, the apparel and footwear industries contribute to 8% of global GHG emissions (Quantis, 2018[2]).
In particular, the stages of textile manufacturing (detailed in Section 2.2.2) may bear high environmental consequences. About 3500 substances are used in textile production, of which 750 have been classified as hazardous for human health and 440 as hazardous for the environment (KEMI, 2014[3]). Fibre production and wet textile processing especially are associated with environmental pressures from high consumption of energy, non-renewable feedstock to make synthetic fibres, fertilisers to grow cotton, chemicals employed in dyeing and finishing treatments and water, as well as from land use (UNEP, 2020[4]).
The high and growing demand for resource input into textile manufacturing raises concerns over the environmental impacts that continued increases in production and consumption may have. Annual clothing sales are projected to more than triple and reach 160 Mt by 2050 (EMF, 2017[5]). The use of and demand for polyester-based clothing in particular has been growing exponentially since its creation and synthetic fibres currently account for two thirds of overall fibre input into textile and apparel production, as presented in Table 2.1. Approximately 59 Mt of plastics (15% of total global production) were employed in the textile manufacturing sector in 2015 (Geyer, Jambeck and Law, 2017[6]). Annual production of plastic-based clothing is expected to more than double between 2015 and 2050 (EMF, 2017[5]).
In this context, the release of microfibres from synthetic clothing is one emerging reason of concern. Synthetic microfibres have been reported in significant quantities at all depths of the marine environment (Browne et al., 2011[7]; Desforges et al., 2014[8]; Obbard et al., 2014[9]; Thompson et al., 2004[10]; Woodal et al., 2014[11]) as well as in marine organisms (Lusher, McHugh and Thompson, 2013[12]). In addition to the washing, wear and tear of synthetic clothing, microfibres sampled in the oceans may originate from a variety of other sources, such as the disintegration of fishing gear, ropes and packaging materials. Yet, the laundering of synthetic textile products alone is estimated to account for 7-35% of total microplastics releases (see Table 1.4).
2.2.1. Characteristics, fate and environmental and human health risks
Fibre shedding is a natural propensity of all fabrics. As textiles are produced and used, mechanical abrasion occurring in their structure causes the detachment and loss of fibres from fabrics. Fibre shedding may occur at (and be influenced by) all stages of the lifecycle of textile products, as follows:
Manufacturing: it is likely that the emission of microfibres starts at the materials sourcing and manufacturing stages, although the extent of microfibre emissions is difficult to quantify with currently available data. Additionally, the choice of manufacturing practices is largely responsible for determining the tendency of fabrics to emit microfibres at later stages of their lifecycle.
Use: wearing, washing, drying of textiles and other stages of maintenance and care may deteriorate the textile structure and contribute to microfibre shedding.
End-of-life: it is possible that textiles also release microfibres at the end-of-life phase, if mismanaged into the environment, or possibly following reuse and recycling practices.
The mechanism and location of emission may determine the fate of the microfibres, as illustrated in Figure 2.2. Microfibres released from textiles enter marine and freshwaters mainly via municipal and industrial wastewaters and via dry and wet deposition. In OECD countries, conventional wastewater treatment technologies can be fairly effective at capturing a large percentage of the emitted fibres, yet, the sheer volumes of wastewaters processed imply that significant amounts of microfibres make their way into aquatic bodies. Once in the environment, synthetic microfibres are known to persist and accumulate, potentially leading to a number of ecological risks, as already discussed in Chapter 1. Microfibres (both synthetic and cellulose-based) have been largely sampled in oceans, freshwaters (Driedger et al., 2015[14]; Lahens et al., 2018[15]; Suaria et al., 2020[16]), as well as in soils where wastewater sludge has been applied (Liu et al., 2019[17]; Zhang and Liu, 2018[18]) and in air (Brahney et al., 2020[19]; Dris et al., 2016[20]).
Source: Adapted from (Henry, Laitala and Klepp, 2019[21])
The presence of airborne microfibres that can be inhaled also adds to total human exposure. Textile microfibres, both cellulose-based and synthetic, have been sampled both in indoor (1-60 fibres/m3) and outdoor (0.3-1.5 fibres/m3) environments (Dris et al., 2017[22]). Recent simulations indicate that inhalation may commonly occur (Vianello et al., 2019[23]), potentially leading to inflammation and health problems (Gasperi et al., 2017[24]; Pauly et al., 1998[25]; Prata, 2018[26]). Chronic exposure to microfibres has shown to lead to a higher prevalence of respiratory irritation, chronic respiratory symptoms, restrictive pulmonary function abnormalities and possibly also to reproductive toxicity and carcinogenicity (Gasperi et al., 2017[24]; Goldberg and Thériault, 1994[27]; Zuskin, Valic and Bouhuys, 1976[28]; Pimentel, Avila and Lourenco, 2008[29]). Yet, available evidence mainly comes from research carried out in industrial settings and may not be representative of ordinary exposure to textile microfibres. Further research is required in order to close the persisting knowledge gaps, in particular to identify critical exposure levels at which adverse health effect may occur (Gasperi et al., 2017[24]).
A major reason of concern with regards to microfibre pollution relates to the potential for microfibres, both synthetic and cellulose-based, to act as transport media for chemical substances employed in textile manufacturing into the environment. These chemicals, and especially those employed during wet processing stages (e.g. finishing treatments, dyeing), bring substantial advantages to apparel products, such as increased durability and a larger range of dyeing colours (EMF, 2017[5]). Yet, certain chemicals employed in the industry are known or suspected to be associated with adverse health effects, such as carcinogenicity, hormone disruption and reproductive toxicity. Textile/apparel manufacturing practices, regular washing, as well as microfibre leakage may release these substances into the environment, potentially posing risks to aquatic ecosystems and human health.
The EU REACH Regulation classifies as Substances of Very High Concern (SVHCs) several chemicals that may be employed in textile, apparel and footwear manufacturing, such as polycyclic aromatic hydrocarbons (PAHs), chlorinated aromatic hydrocarbons, phthalates, azo-dyes and chlorinated and/or brominated flame retardants, perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) (EU, 2006[30]; Istituto Superiore di Sanità, 2020[31]). In recent years, several regulatory efforts and industry-led initiatives have emerged to report the use of hazardous substances in textile manufacturing and minimise their releases, as detailed in Box 2.1. Yet, persisting gaps in transparency over the chemicals utilised during manufacturing create challenges for the adequate evaluation of the health hazards posed to ecosystems and human health (EMF, 2017[5]).
Fashion brands selling products in Europe and North America are subject to a number of restrictions on the chemical content of products. REACH Regulation 1907/2006 on the Registration, Evaluation, and Authorisation and Restriction of Chemicals restricts the use of substances identified as harmful (Substances of Very High Concern) in products manufactured in or imported into Europe. Similarly, in the US, the Toxic Substances Control Act (TSCA) requires companies to report, record and carry out testing on the chemical content of products placed on the market.
Further, as a response to emerging legislation and increasing public pressure, several voluntary industry-led initiatives were developed. Notably, the textile and apparel industry has developed the Zero Discharge of Hazardous Chemicals (ZDHC) Manufacturing Restricted Substances List (MRSL) to define a harmonised approach to managing harmful and hazardous chemicals in the sector. This evidence-based document provides a list of priority chemicals to be phased out and specifies a maximum concentration for each substance and serves as an industry-wide reference in multiple initiatives (ZDHC, 2015[32]). The OECD Due Diligence Guidance for the Garment and Footwear sector, for instance, recommends that companies adopt and implement an evidence-based common MRSL to address the risk of harmful chemicals in their products and supply chains (OECD, 2018[33]).
Cellulose-based fibres and the relevance of a holistic approach to microfibre pollution
Although cellulose-based (i.e. natural and semi-synthetic) fibres are expected to biodegrade quickly if released into aquatic media, emerging evidence suggests that these are commonly present in aquatic habitats (Dris et al., 2018[34]; Sanchez-Vidal et al., 2018[35]; Stanton et al., 2019[36]) and wildlife species (Compa et al., 2018[37]; Remy et al., 2015[38]; Lusher, McHugh and Thompson, 2013[12]). Recent studies suggest that past research may have largely overlooked their presence of cellulose-based microfibres in the environment, potentially also leading to an overestimation of the contribution of synthetic textiles to marine microplastics pollution (Suaria et al., 2020[16]). In fact, there seems to be a considerable mismatch between the share of fibres in textile production (of which over two thirds are synthetic) and the types of microfibres polluting the environment, with 60-80% of microfibres sampled in the oceans and in marine organisms being of cellulosic origin (Suaria et al., 2020[16]).
The widespread occurrence of cellulose-based fibres in the environment calls for further research with reliable characterisation of the polymer in order to assess the occurrence, degree of persistence and toxicity of different types of microfibres present in the environment. At the same time, current evidence (and uncertainties) may justify taking a holistic approach to microfibre mitigation, as characterised by two elements: a) a comprehensive and life-cycle assessment of the environmental impacts of textiles and b) a focus on finding solutions to mitigate risks associated with microfibre shedding (for all textile types). This is for several reasons:
It is possible that the risks associated with microfibre pollution may not be limited to synthetic fibres. As indicated above, knowledge gaps persist with regards to the degree of persistence and accumulation of non-plastic fibres in the environment. Further, cellulose-based fibres may also act as a transfer media for harmful chemicals. For natural fibres, it has been speculated that a more rapid biodegradation may increase the bioavailability of chemical additives once microfibres are ingested by aquatic organisms (Zhao, Zhu and Li, 2016[39]). For airborne microfibres, it has also been suggested that the adverse effects of chronic exposure to cellulose-based fibres may not be significantly different than for synthetic ones (Prata, 2018[26]).
Determining whether clothing sheds microfibres of synthetic content is not straightforward in practice. Blended textiles (e.g. polyester/cotton blends) are very common (for instance to enhance certain characteristics of the final product) and fabrics made out of natural fibres are often treated with synthetic coatings during wet processing.
More broadly, there is a strong case for finding ways to work with synthetic materials, as substitution away from synthetic fibres in textile and apparel manufacturing may not be a viable microplastics mitigation solution at scale. Natural alternatives are limited and cannot always provide the same performance capabilities of synthetic materials. Further, the lifecycle of textiles produced from natural fibres also bears significant adverse consequences on the environment, in particular in terms of high energy and water consumption, land use and the release of chemicals harmful to the environment (UNEP, 2017[40]).
2.2.2. Industrial emissions occurring during manufacturing
The stages of textile and garment manufacturing are associated with high risks for environmental and climate impacts. Industrial emissions from textile manufacturing plants have been long scrutinised, in particular with regards to the release of potentially harmful chemicals into the environment, such as certain flame retardants and chemical coatings applied to textile products during manufacturing. As the issue of microplastics pollution gained increasing scientific and policy attention, recently concerns have also emerged over the contribution of industrial emissions to microfibre pollution.
The stages of textile and apparel manufacturing are detailed in Table 2.2. Fabrics are manufactured from fibres or yarns, i.e. continuous strands of fibres, via different technologies. Several wet processing activities may be performed on fabrics to enhance the appearance and performance of the final product. These include preparatory treatments, dyeing processes and functional mechanical or chemical finishing treatments. The make-up is the last step before selling in retail or whole trade and consumer use.
Several stages of textile and garment manufacturing may contribute to the emission of microfibres into sewage waters or into the surrounding aerial environment. In particular, the processes involved in fibre processing, yarn manufacturing and fabric construction are known to lead to fibre mechanical stress and material losses (WRAP, 2019[41]). Fibre emission may also occur during the production of garments (e.g. during cutting, sewing and the application of finishing treatments), as a result of the removal of impurities and sizing, although this is less documented. WRAP (2019[41]) estimates that in the UK 168 thousand tonnes of material is lost each year during the production of clothing (per 1.1 Mt of clothing consumed annually), although it is unclear what percentage of this material loss is emitted as microfibres.
Recently, research has also been undertaken to quantify microfibres released into sewage waters. Available evidence is very limited, but suggests that textile manufacturing plants regularly emit microplastics into wastewaters. A study conducted in Sweden detected concentrations of 100-450 microfibres per litre of industrial effluent from five textile production facilities (Jönsson and Landin, 2018[42]). The detected microfibres were mainly of synthetic origin, although cotton and viscose fibres were also reported in large quantities for certain production plants. Research conducted in China found average concentrations of 16-334 (synthetic and natural-based) microfibres per litre in wastewaters discharged from textile printing and dyeing facilities (Xu et al., 2018[43]).
Although the contribution of the textile and apparel manufacturing stages to overall microfibre releases is difficult to assess due to a lack of reliable data and monitoring, there are concerns that this might be substantial. Firstly, considering the magnitude of the textile and apparel industry and the amounts of wastewaters being discharged during manufacturing, it is likely that even modest amounts of microplastics being released per litre of industrial effluent could result in significant amounts of fibres entering the environment. Secondly, the majority of textile and apparel production takes place in emerging economies, where the lower rates of connectedness to wastewater infrastructure and the lower levels of treatment (relatively to OECD countries) might potentially imply that a higher share of the industrial microfibre emissions reach water bodies.1
2.2.3. Emissions occurring during use
The use phase has been identified as a major source of microfibre emissions. Several stages of the use phase – i.e. wearing, washing and drying – may contribute to mechanical abrasion occurring in the structure of fabrics and lead to the detachment of fibres. Current research has focused on the laundering of synthetic garments, where the series of mechanical and chemical actions aimed at cleaning textile products contribute to the generation and emission of loose fibres.2 Several series of laboratory washing tests have been carried out in recent years to measure the degree of microfibre shedding from garments with different characteristics and under different washing conditions. These generally tend to employ a variety of test conditions and methods for microfibre measurement, which limits the comparability of studies and the generalisation of findings (Jönsson et al., 2018[45]). Yet, test washings have allowed the identification of certain factors that may lead to a higher or lower fibre shedding during laundering processes. These can be grouped in two categories:
Textile and garment characteristics. The microfibre shedding rate during use is dependent on the degree of fibre strength and resistance to abrasion of the product. These are influenced by a variety of design and manufacturing factors, including textile composition and fibre characteristics, yarn and textile structures and garment manufacturing processes. For instance, polyester fleece and microfleece fabrics are known for being particularly prone to fibre shedding: a single fleece jacket may shed up to 250,000 fibres per laundry wash (Hartline et al., 2016[46]).
Product maintenance and care. In general, laundering methods that minimise the degree of mechanical abrasion (e.g. low-temperature laundry washes and the use of softener liquid) are associated with a preservation of the integrity of textile yarns and a lower fibre shedding. The type of washing machine may also influence the degree of mechanical stress occurring in the textile structure. Drying practices, and tumble drying in particular, are likely to also influence the emission of microfibres.
Although knowledge gaps persist with regards to the relative importance of factors driving microfibre release, several mitigation options implementable at the production and use stage of textile products can already be drawn based on the available knowledge. These are presented and assessed in Chapter 3.
Microplastics may be emitted at all stages of the tyre lifecycle, as follows:
Manufacturing: although it is possible that microplastics are generated and released as by-products during the manufacturing of tyres, there is a lack of data to verify whether this is the case. Also, manufacturing practices influence the tendency of tyres to undergo abrasion during regular use.
Use: Tyre and Road Wear Particles (TRWP) are emitted during regular vehicle use due to the friction occurring between tyres and the road surface;
End-of-life management: the mismanagement of tyres into the environment may potentially lead to microplastics generation and leakage. Also, certain recycling options for end-of-life tyres (e.g. the use of tyre rubber granulate used as infill in artificial sport turfs) potentially constitute a further source of microplastics into the environment.
2.3.1. Tyre and Road Wear Particles
Characteristics, fate and environmental and human health risks
During normal transport activity, the friction between vehicle tyres and the road surface results in the abrasion of the tyre tread and the emission of particles. As road pavement materials tend to also agglomerate within the tyre material, the emitted particles are generally referred to as Tyre and Road Wear Particles (TRWP). In general, TRWP are composed of a complex mixture of tyre tread material (e.g. synthetic and natural rubber, silica, oil, carbon black, sulphur compounds, zinc oxide), road pavement material (e.g. polymer modified bitumen), road marking3 particles, brake wear particles and other airborne elements that commonly deposit on pavements (Kreider et al., 2010[47]).
Recent studies have attempted to quantify emissions of TRWP from road vehicles occurring during road transport activity, based either on emission factors for different vehicle categories and road transport activity data, or from average tyre wear rates and data on the number of tyres in use (Kole et al., 2015[48]; Lassen et al., 2016[49]; Wagner et al., 2018[50]; Magnusson et al., 2016[51]). Although estimates of the contribution of tyre wear to microplastics pollution differ, approximately 0.81 kg of emissions per capita are released from vehicle tyres annually, with the highest per capita releases occurring in the United States (Kole et al., 2017[52]). National emissions may differ based on the local context: for instance, in Germany the largest contributions to TRWP emissions come from heavy vehicles (trucks, buses) and driving on highways, while in the United States total emissions from passenger cars and trucks are roughly equivalent, and two-thirds of emissions occur in urban environments, mainly due to the higher urban travel distances in North America compared to European countries (Wagner et al., 2018[50])
At the point of emission, TRWP may become suspended in air or deposit on road surfaces and nearby soil. Additionally, the action of rain events may disperse or flush emitted TRWP into nearby water streams. The physical characteristics of the emitted particles, and in particular their size, may be important determinants of their environmental fate (Unice et al., 2019[53]). TRWP are generally elongated in shape (i.e., cigar-shaped) and are well below 1 mm in length4 (Unice et al., 2019[54]). A portion of TRWP (1-10% in mass) is emitted in the fine particulate matter size range (< 10 μm) and contributes to ambient air pollution (see Box 2.2). Larger particles are typically deposited on the road surface or on nearby soil. The majority of TRWP tend to be heavier than water (particles have an average density of 1.8 g/cm3) (Unice et al., 2019[54]) and so they may be prone to sedimentation if dispersed into aquatic environments (Parker-Jurd et al., 2019[55]; NIVA, 2018[56]). It is also relevant to note that, following release into the environment, TRWP may undergo ageing processes that affect their physical and chemical properties and ultimately their fate. Recent studies suggest that further research is required to understand the extent of these changes in the composition and properties of TRWP in order to accurately model their environmental fate (Klöckner et al., 2020[57]; Unice et al., 2015[58]).
Tracing the fate of the emitted TRWP is crucial in order to assess exposure routes and the associated health risks, as well as to identify potential hotspots where the implementation of end-of-pipe capture solutions could be prioritised. A number of factors may influence how the particles will spread into different environmental media following emission. Airborne particles can either be deposited on the road surface, or be transported via wet and dry deposition, potentially far away from point sources (Parker-Jurd et al., 2019[55]; Magnusson et al., 2020[59]). For instance, it has been suggested that atmospheric transport may significantly contribute to the long-distance transport of airborne TRWP and other non-exhaust emissions into the marine environment and remote regions such as the Arctic, where the particles may possibly pose additional climatic risks of increased light-absorption and enhanced snow and ice melting (Evangeliou et al., 2020[60]).
Available modelling estimates of the spatial distribution of TRWP emissions suggest that a large portion of the emitted particles is expected to deposit on roads or in nearby soil (Figure 2.3). Road runoff, wind and street cleaning may contribute to the removal of these larger particles from the road and their potential dispersal into the environment. Where roads are not connected to stormwater systems, TRWP will drain off with rain into adjacent land or water streams (Andersson-Sköld et al., 2020[61]). Where stormwater systems are present, the fate of TRWP will depend on the specific treatment technologies in place (i.e. direct discharge into a recipient, stormwater treatment facilities, or a WWTP). The amount of TRWP reaching surface waters largely depends on the local conditions (e.g. presence of drains for road runoff, the type of road, the intensity of rainfall). Where urban surface runoff is collected and treated prior to discharge, approximately 11-22% of TRWP is expected to reach surface waters directly or via the sewerage system (Verschoor et al., 2016[62]; Wagner et al., 2018[50]). The type of road infrastructure may also affect the fate of the emitted particles: for instance, in Netherlands half of the emitted particles remain incorporated into porous asphalt, a type of road pavement widely employed in Dutch highways that is prone to absorbing particles (Verschoor et al., 2016[62]).5
Only a limited number of studies have looked at the environmental presence of TRWP, typically in road dust and stormwater runoff. A key barrier to larger and more reliable environmental quantification of TRWP is the availability of appropriate analytical methods. Conventional methods used for the sampling and characterisation of microplastics are not easily adaptable to TRWP, while methods well-adapted to TRWP are costly and time-consuming (Andersson-Sköld et al., 2020[61]). There are concerns that inadequate and different analytical methods for sampling and characterisation may be underestimating the amount of (or falsely confirming presence of) TRWP in the natural environment and their overall contribution to microplastics pollution (Parker-Jurd et al., 2019[55]). Harmonised methods for sampling, sample preparation and analysis of TRWP are required to allow for further environmental sampling and for better consistency and comparability between different studies.
Notes: Comparisons are difficult to make as different studies included different sets of environmental compartments and/or had different objectives. All studies modelled annual emissions of TRWP. Sieber et al. (2020[63]) also calculated the accumulation of tyre-based microplastics for Switzerland over a period of 30 years (1988–2018) and included the emissions and fate of microplastics lost from artificial sports turfs. Baensch-Baltruschat et al. (2021[64]) did not include airborne emissions (which they considered to be 5% of total emissions). Wagner et al. (2018[50]) proposed two scenarios: scenario (A) assumes that 50% of TRWP deposited on road surfaces are mobilised via surface runoff, while scenario (B) assumes that only 15% are mobilised.
Source: Author’s elaboration based on (Baensch-Baltruschat et al., 2021[64]; Sieber, Kawecki and Nowack, 2020[63]; Unice et al., 2019[54]; Verschoor et al., 2016[62]; Wagner et al., 2018[50])
Available microplastics surveys indicate that tyre wear may be a significant contributor to the emission of microplastics into surface waters, potentially to a larger extent than previously estimated. A study completed around the San Francisco Bay area found that nearly half of all microplastics contained in stormwater discharge were suspected TRWP (Sutton et al., 2019[65]). A recent study conducted in the United Kingdom found a large presence of TRWP at key entry points into the marine environment (wastewater treatment effluent, stormwater runoff and wind), possibly several orders of magnitude greater than that of synthetic microfibres (Parker-Jurd et al., 2019[55]). Overall, further field data is needed to improve our understanding of the transport processes and sinks of TRWP and to validate and complement the available model estimates.
Only a limited number of studies have assessed the potential environmental and human health impacts of TRWP and further research is required to adequately assess risks. Some of the chemicals used in the manufacture of tyres, road marking products and polymer modified bitumen are hazardous to human health and the environment, however there is limited knowledge about the extent to which these substances are released from microplastics (Andersson-Sköld et al., 2020[61]). Research that informs toxicological considerations is based on the use of TWP, i.e. tyre wear particles that are created in laboratory conditions, rather than particles sampled from the environment.6 The majority of available studies assessed the (acute and chronic) toxicity of leachates from TWP on aquatic organisms: while some showed no toxicity on freshwater and sediment dwelling species (Marwood et al., 2011[66]; Panko et al., 2013[67]), others observed adverse health effects (Halle et al., 2020[68]; Tian et al., 2021[69]). As with other microplastics, the ingestion of TRWP is a key exposure route for aquatic wildlife (Khan, Halle and Palmqvist, 2019[70]; Redondo-Hasselerharm et al., 2018[71]; Wik et al., 2009[72]), however large knowledge gaps persist with regards to the potential health hazards posed. A recent study by Halle et al. (2021[73]) showed that TRWP in the aquatic environment may affect acute mortality and long-term growth.
Overall, further research is required both to assess the toxicity of the ingested particles and to improve our understanding of the associated hazards in realistic environmental scenarios (Halle et al., 2020[68]). With regards to risks for human health, the most researched exposure route for adverse health impacts is the inhalation of non-exhaust emissions, as outlined in Box 2.2. However, little is known with regards to the risks posed to human health by TRWP via ingestion, relatively to other microplastics.
Air pollution is a major environmental and human health risk, to which road transport emissions significantly contribute. Emissions of particulate matter (PM) from motor vehicles originate from two main sources: tailpipe exhaust and the degradation of vehicle parts and the road surface (OECD, 2020[74]). The latter are defined as non-exhaust PM emissions and comprise all airborne particulate emissions generated by the wear of vehicle parts (mainly tyres and brake pads) and of the road surface, as well as by the resuspension of road dust.
Tyre wear significantly contributes to non-exhaust emissions and air pollution. Estimates of the contribution of tyre tread wear to total particulate matter range from 0.1 to 10% for PM10 and 1–7% for PM2.5 (Andersson-Sköld et al., 2020[61]; Panko, Kreider and Unice, 2018[75]). Additional contributors to non-exhaust emissions are the wear of road surfaces and of brake pads and the resuspension of particles on the road surface (OECD, 2020[74]). Brake wear particles are emitted as a result of the abrasion occurring between stationary brake pads and the vehicle rotor during braking. They tend to be smaller in size and approximately 50% of brake wear particles become airborne at the point of emission (Grigoratos and Martini, 2015[76]).
It is now well established that the inhalation of fine PM and the associated metals and combustion products (PAHs) negatively affects human health. Exposure to PM, and in particular to PM2.5, is associated with increased risks of cardiovascular, respiratory and developmental conditions, as well as an increased risk of overall mortality (OECD, 2020[74]). The oxidative stress induced by the metals and organic compounds found in PM emissions is considered to be a main biological mechanism responsible for these negative health impacts (OECD, 2020[74]). In light of the hazards posed by certain polycyclic aromatic hydrocarbons (PAHs), the EU has placed a restriction on the use of 8 PAHs in tyres and extender oils via the REACH regulation (Annex XVII.50).
Influence factors
Tyre abrasion can cause an overall mass loss of up to 10% during the lifetime of a tyre (Grigoratos et al., 2018[77]). A variety of local factors may influence the amount of tyre tread material lost per kilometre travelled. These can be grouped in four categories (ETRMA, 2018[78]):
tyre characteristics: size, tread depth, construction, tyre pressure and temperature, contact patch area, chemical composition, accumulated mileage;
vehicle characteristics: weight and size, distribution of loads, location of driving wheels, wheel alignment, engine power, mechanical/electronic braking system, suspension type and conditions;
driving behaviour: speed, acceleration/deceleration, frequency and extent of braking, cornering;
road surface characteristics: pavement type, porosity, maintenance, weather conditions.
While current knowledge does not allow for a precise estimate of tyre wear rates, some general trends can be derived with regards to the influence of different factors on tyre wear. For instance, it is estimated that these are highest for heavier vehicles (e.g. buses, trucks and lorries) than for passenger cars. Further research is required in particular to assess and quantify the relative impact of each influence factor on tyre wear in real-life conditions. Several mitigation options implementable at the production and use stage of tyres can already be drawn based on the available knowledge over the drivers of TRWP emission. These are discussed in Chapter 3.
2.3.2. Management of end-of-life tyres (ELTs)
Tyres are typically replaced when they are no longer suitable for use due to wear or damage. Tyres may be re-used when they have been only partially worn and sufficient residual tread depth remains, or otherwise may be retreaded into new tyres. When neither reuse nor retreading is possible, scrap or End-of-Life Tyres (i.e. tyres which can no longer be used for their original purpose) may be employed for material recovery and civil engineering applications, or incinerated for energy recovery (WBCSD, 2019[79]).
Dumping and improper disposal of used tyres remain an issue in several countries. In general, the degree of recovery and the performance of ELT management is dependent on the existence and level of maturity of formal management systems. Landfilling of old tyres is illegal in several OECD countries (e.g. in the European Union, the US State of California)7. Generally, landfilling is considered an undesirable disposal option for tyres due to their slow degradation, the potential to cause damage to landfill liners and the intrinsic value of tyre materials. Yet, it is likely that in several emerging economies where formal management schemes are not in place, significant amounts of tyres are abandoned, landfilled, or stockpiled. In addition to wasting potentially valuable resources, the mismanagement of tyres contributes to several local environmental and human health risks, such as the risk of stockpile fires, the potential for old tyres to act as a breeding ground for disease-carrying mosquitos and hazards associated with chemical leachate.
Several OECD countries have introduced ELT management schemes to facilitate the separate collection and environmentally sound handling of used tyres, such as Extended Producer Responsibility systems and take-back obligation schemes. These resulted in an overall improvement of collection rates for used tyres, as well as fostered the development of the ELT recycling industry and the proliferation of solutions to close material loops in the sector. For instance, in the Flanders, the EPR system in place has contributed to decreasing the amounts of dumped tyres almost to zero (OECD, 2016[80]). Further, the regular flow of used tyres guaranteed by the management scheme in place has allowed for the development of a market for recycling tyres and tyre materials and a reduction of total tyre materials disposed via incineration from energy recovery.
In recent years, concerns emerged with regards to the potential for microplastics to leak from certain material recovery applications for ELTs. This is discussed below.
Leakage of rubber granulate from artificial sport turfs
A common method for material recovery from end-of-life tyres is shredding for the production of rubber granulate, i.e. small particles to be used in a variety of industrial applications. Rubber granulate can be manufactured from ELTs as well as from rubber derived from other sources (e.g. virgin elastomer alternatives such as EPDM rubber and TPE) and usually has a size between 0.5 and 2.5 mm (Eunomia, 2018[81]). A common application of rubber granulate is use as infill for artificial sport turfs. The use of rubber granulate as infill material offers several advantages compared to natural alternatives, such as durability, resistance to varying weather conditions, good shock absorbance and safety characteristics, low costs, as well as a lower need for virgin materials (Magnusson et al., 2016[51]).
Some recent studies have pointed to artificial turfs as an additional source of microplastics discharge into surrounding soil and surface drains, due to the emission of rubber granulate mainly caused by transport off the pitch during use (e.g. by athletes) or during maintenance and the effect of weather events (RIVM, 2018[82]). Initial estimates for Sweden found that approximately 2-3 tonnes of microplastics per football field may be lost yearly, suggesting that rubber infill may constitute a major source of microplastics pollution (Kole et al., 2017[52]). It is now recognised that several factors influence the overall volume of infill material (e.g. compaction) and that past figures may have largely overestimated the extent of microplastics leakage from artificial sport turfs. Still, a more recent study conducted in Denmark estimated the infill material loss (due to contact with athletes, snow clearance and rain water discharges) to be 300-730 kg/year per field (Løkkegaard, Malmgren-Hansen and Nilsson, 2018[83]). For Sweden, new calculations estimate that around 550 kg/year from an average football field, which would imply yearly national losses of 475 tonnes of microplastics (Swedish EPA, 2019[84]).
In response to recent findings, several OECD countries have mandated research projects and calls for evidence to fill knowledge gaps on the composition, leakage, exposure pathways and potential hazards of rubber granulate used in artificial sport pitches. A recent mass flow study in Switzerland demonstrated that about 3% of rubber-based particles entering the environment is released as granules (and 97% as TRWP) (Sieber, Kawecki and Nowack, 2020[63]). Further research is required to better assess the environmental risks associated to the use of rubber granulate as infill material in sport pitches, and in particular to further investigate the potential for release of hazardous substances via ELT-derived rubber granulate (ANSES, 2018[85]). An additional source of microplastics pollution which also requires further investigation is the use of rubber granulate in moulded rubber granule surfaces, such as fall protections and multicourts present in playgrounds.
References
[61] Andersson-Sköld, Y. et al. (2020), Microplastics from tyre and road wear - A literature review, Swedish National Road and Transport Research Institute (VTI).
[85] ANSES (2018), Scientific and technical support on the possible risks related to the use of materials derived from the recycling of used tyres in synthetic sports grounds and similar uses.
[64] Baensch-Baltruschat, B. et al. (2021), “Tyre and road wear particles - A calculation of generation, transport and release to water and soil with special regard to German roads”, Science of The Total Environment, Vol. 752/141939, https://doi.org/10.1016/j.scitotenv.2020.141939.
[19] Brahney, J. et al. (2020), “Plastic rain in protected areas of the United States”, Science, Vol. 368/6496, p. 1257, https://doi.org/10.1126/science.aaz5819.
[7] Browne, M. et al. (2011), “Accumulation of microplastic on shorelines woldwide: Sources and sinks”, Environmental Science and Technology, Vol. 45/21, pp. 9175-9179, https://doi.org/10.1021/es201811s.
[37] Compa, M. et al. (2018), “Ingestion of microplastics and natural fibres in Sardina pilchardus (Walbaum, 1792) and Engraulis encrasicolus (Linnaeus, 1758) along the Spanish Mediterranean coast”, Marine Pollution Bulletin, Vol. 128, pp. 89-96, https://doi.org/10.1016/j.marpolbul.2018.01.009.
[8] Desforges, J. et al. (2014), “Widespread distribution of microplastics in subsurface seawater in the NE Pacific Ocean”, Marine Pollution Bulletin, Vol. 79/1-2, pp. 94-99, https://doi.org/10.1016/j.marpolbul.2013.12.035.
[14] Driedger, A. et al. (2015), Plastic debris in the Laurentian Great Lakes: A review, https://doi.org/10.1016/j.jglr.2014.12.020.
[22] Dris, R. et al. (2017), “A first overview of textile fibers, including microplastics, in indoor and outdoor environments”, Environmental Pollution, Vol. 221, pp. 453-458, https://doi.org/10.1016/j.envpol.2016.12.013.
[34] Dris, R. et al. (2018), “Synthetic and non-synthetic anthropogenic fibers in a river under the impact of Paris Megacity: Sampling methodological aspects and flux estimations”, Science of The Total Environment, Vol. 618, pp. 157-164, https://doi.org/10.1016/j.scitotenv.2017.11.009.
[20] Dris, R. et al. (2016), “Synthetic fibers in atmospheric fallout: A source of microplastics in the environment?”, Marine Pollution Bulletin, Vol. 104/1-2, pp. 290-293, https://doi.org/10.1016/j.marpolbul.2016.01.006.
[1] EEA (2019), Textiles in Europe’s circular economy.
[5] EMF (2017), A New Textile Economy: Redesigning Fashion’s Future, Ellen Macarthur Foundation, http://www.ellenmacarthurfoundation.org/publications.
[78] ETRMA (2018), Way Forward Report.
[30] EU (2006), REACH Regulation 1907/2006 on the Registration, Evaluation, and Authorisation and Restriction of Chemicals.
[81] Eunomia (2018), “Investigating options for reducing releases in the aquatic environment of microplastics emitted by (but not intentionally added in) products - Interim Report”, Report for DG Environment of the European Commission, p. 335, https://doi.org/10.1002/lsm.22016.
[60] Evangeliou, N. et al. (2020), “Atmospheric transport is a major pathway of microplastics to remote regions”, Nature Communications, Vol. 11/1, p. 3381, https://doi.org/10.1038/s41467-020-17201-9.
[24] Gasperi, J. et al. (2017), “Microplastics in air: Are we breathing it in?”, Current Opinion in Environmental Science & Health, Vol. 1, pp. 1-5, https://doi.org/10.1016/j.coesh.2017.10.002.
[6] Geyer, R., J. Jambeck and K. Law (2017), “Production, use, and fate of all plastics ever made”, Science Advances, Vol. 3/7, p. e1700782, https://doi.org/10.1126/sciadv.1700782.
[27] Goldberg, M. and G. Thériault (1994), “Retrospective cohort study of workers of a synthetic textiles plant in quebec: II. Colorectal cancer mortality and incidence”, American Journal of Industrial Medicine, doi: 10.1002/ajim.4700250613, pp. 909-922, https://doi.org/10.1002/ajim.4700250613.
[77] Grigoratos, T. et al. (2018), “Experimental investigation of tread wear and particle emission from tyres with different treadwear marking”, Atmospheric Environment, Vol. 182, pp. 200-212, https://doi.org/10.1016/j.atmosenv.2018.03.049.
[76] Grigoratos, T. and G. Martini (2015), “Brake wear particle emissions: a review”, Environmental Science and Pollution Research, Vol. 22/4, pp. 2491-2504, https://doi.org/10.1007/s11356-014-3696-8.
[73] Halle, L. et al. (2021), “Tire wear particle and leachate exposures from a pristine and road-worn tire to Hyalella azteca: Comparison of chemical content and biological effects”, Aquatic Toxicology, Vol. 232, p. 105769, https://doi.org/10.1016/j.aquatox.2021.105769.
[68] Halle, L. et al. (2020), Ecotoxicology of micronized tire rubber: Past, present and future considerations, Elsevier B.V., https://doi.org/10.1016/j.scitotenv.2019.135694.
[46] Hartline, N. et al. (2016), “Microfiber Masses Recovered from Conventional Machine Washing of New or Aged Garments”, Environ. Sci. Technol., Vol. 50, p. 11532−11538, https://doi.org/10.1021/acs.est.6b03045.
[21] Henry, B., K. Laitala and I. Klepp (2019), “Microfibres from apparel and home textiles: Prospects for including microplastics in environmental sustainability assessment”, Science of The Total Environment, Vol. 652, pp. 483-494, https://doi.org/10.1016/j.scitotenv.2018.10.166.
[31] Istituto Superiore di Sanità (2020), Rapporti ISTISAN 20/10 - Chimica, moda e salute.
[42] Jönsson, C. and R. Landin (2018), Report no. 18004. Investigation of the occurrence of microplastics from the waste water at five different textile production facilities in Sweden, Swerea IVF.
[45] Jönsson, C. et al. (2018), “Microplastics Shedding from Textiles—Developing Analytical Method for Measurement of Shed Material Representing Release during Domestic Washing”, Sustainability, Vol. 10/7, p. 2457, https://doi.org/10.3390/su10072457.
[3] KEMI (2014), Chemicals in textiles – Risks to human health and the environment, http://www.kemi.se/files/8040fb7a4f2547b7bad522c399c0b649/report6-14-chemicals-in-textiles.pdf (accessed on 26 November 2019).
[70] Khan, F., L. Halle and A. Palmqvist (2019), “Acute and long-term toxicity of micronized car tire wear particles to Hyalella azteca”, Aquatic Toxicology, Vol. 213, p. 105216, https://doi.org/10.1016/j.aquatox.2019.05.018.
[57] Klöckner, P. et al. (2020), “Characterization of tire and road wear particles from road runoff indicates highly dynamic particle properties”, Water Research, Vol. 185, p. 116262, https://doi.org/10.1016/j.watres.2020.116262.
[48] Kole, P. et al. (2015), Autobandenslijtstof: een verwaarloosde bron van microplastics?.
[52] Kole, P. et al. (2017), “Wear and Tear of Tyres: A Stealthy Source of Microplastics in the Environment.”, International journal of environmental research and public health, Vol. 14/10, https://doi.org/10.3390/ijerph14101265.
[47] Kreider, M. et al. (2010), “Physical and chemical characterization of tire-related particles: Comparison of particles generated using different methodologies”, Science of The Total Environment, Vol. 408/3, pp. 652-659, https://doi.org/10.1016/J.SCITOTENV.2009.10.016.
[15] Lahens, L. et al. (2018), “Macroplastic and microplastic contamination assessment of a tropical river (Saigon River, Vietnam) transversed by a developing megacity”, Environmental Pollution, Vol. 236, p. 661−671, https://doi.org/10.1016/j.envpol.2018.02.005.
[49] Lassen, C. et al. (2016), Microplastics Occurrence, effects and sources of releases to the environment in Denmark, Danish Environmental Protection Agency, Copenhagen.
[17] Liu, X. et al. (2019), “Transfer and fate of microplastics during the conventional activated sludge process in one wastewater treatment plant of China”, Chemical Engineering Journal, Vol. 362, pp. 176-182, https://doi.org/10.1016/j.cej.2019.01.033.
[83] Løkkegaard, H., B. Malmgren-Hansen and N. Nilsson (2018), Mass balance of rubber granulate lost from artificial turf fields, focusing on discharge to the aquatic environment. A review of literature., https://www.genan.eu/wp-content/uploads/2020/02/Teknologisk-Institut_Mass-balance-of-rubber-granulate-lost-from-artificial-turf-fields_May-2019_v1.pdf.
[12] Lusher, A., M. McHugh and R. Thompson (2013), “Occurrence of microplastics in the gastrointestinal tract of pelagic and demersal fish from the English Channel”, Marine Pollution Bulletin, Vol. 67/1-2, pp. 94-99, https://doi.org/10.1016/j.marpolbul.2012.11.028.
[51] Magnusson, K. et al. (2016), Swedish sources and pathways for microplastics to the marine environment A review of existing data. Revised in March 2017, https://www.ivl.se/english/ivl/publications/publications/swedish-sources-and-pathways-for-microplastics-to-the-marine-environment.html.
[59] Magnusson, K. et al. (2020), Atmosfäriskt nedfall av mikroskräp, http://urn.kb.se/resolve?urn=urn:nbn:se:naturvardsverket:diva-8436.
[66] Marwood, C. et al. (2011), “Acute aquatic toxicity of tire and road wear particles to alga, daphnia, and fish”, Ecotoxicology, Vol. 20/2079, https://doi.org/10.1007/s10646-011-0750-x.
[56] NIVA (2018), Microplastics in road dust – characteristics, pathways and measures. Revised in 2020., Norwegian Institute for Water Research.
[9] Obbard, R. et al. (2014), “Global warming releases microplastic legacy frozen in Arctic Sea ice”, Earth’s Future, Vol. 2/6, pp. 315-320, https://doi.org/10.1002/2014EF000240.
[74] OECD (2020), Non-exhaust Particulate Emissions from Road Transport: An Ignored Environmental Policy Challenge, OECD Publishing, Paris, https://dx.doi.org/10.1787/4a4dc6ca-en.
[44] OECD (2019), Due Diligence on Upstream Production, https://mneguidelines.oecd.org/OECD-Garment-Forum-2019-session-note-Due-diligence-on-upstream-production.pdf.
[33] OECD (2018), OECD Due Diligence Guidance for Responsible Supply Chains in the Garment and Footwear Sector, OECD Publishing, Paris, https://dx.doi.org/10.1787/9789264290587-en.
[80] OECD (2016), Extended Producer Responsibility: Updated Guidance for Efficient Waste Management, OECD Publishing, Paris, https://doi.org/10.1787/9789264256385-en.
[67] Panko, J. et al. (2013), “Chronic toxicity of tire and road wear particles to water- and sediment-dwelling organisms”, Ecotoxicology, Vol. 22, pp. 13–21, https://doi.org/10.1007/s10646-012-0998-9.
[75] Panko, J., M. Kreider and K. Unice (2018), “Review of Tire Wear Emissions”, in Non-Exhaust Emissions, Elsevier, https://doi.org/10.1016/b978-0-12-811770-5.00007-8.
[55] Parker-Jurd, F. et al. (2019), Investigating the sources and pathways of synthetic fibre and vehicle tyre wear contamination into the marine environment, Report prepared for the Department for Environment Food and Rural Affairs (project code ME5435).
[25] Pauly, J. et al. (1998), “Inhaled cellulosic and plastic fibers found in human lung tissue”, Cancer Epidemiology Biomarkers and Prevention, Vol. 7/5, pp. 419-428.
[29] Pimentel, J., R. Avila and A. Lourenco (2008), “Respiratory disease caused by synthetic fibres: a new occupational disease.”, Thorax, Vol. 30/2, pp. 204-219, https://doi.org/10.1136/thx.30.2.204.
[26] Prata, J. (2018), “Airborne microplastics: Consequences to human health?”, Environmental Pollution, Vol. 234, pp. 115-126, https://doi.org/10.1016/j.envpol.2017.11.043.
[2] Quantis (2018), Measuring fashion - Insights from the Environmental Impact of the Global Apparel and Footwear Industries study, https://quantis-intl.com/measuring-fashion-report-2018/.
[71] Redondo-Hasselerharm, P. et al. (2018), “Ingestion and Chronic Effects of Car Tire Tread Particles on Freshwater Benthic Macroinvertebrates”, Environmental Science and Technology, Vol. 52/23, pp. 13986-13994, https://doi.org/10.1021/acs.est.8b05035.
[38] Remy, F. et al. (2015), “When Microplastic Is Not Plastic: The Ingestion of Artificial Cellulose Fibers by Macrofauna Living in Seagrass Macrophytodetritus”, Environmental Science & Technology, Vol. 49/18, pp. 11158-11166, https://doi.org/10.1021/acs.est.5b02005.
[82] RIVM (2018), Verkenning milieueffecten rubbergranulaat bij kunstgrasvelden, Verschoor, A. J., Bodar, C.W.M., Baumann, R.A., https://doi.org/10.21945/RIVM-2018-0072.
[35] Sanchez-Vidal, A. et al. (2018), “The imprint of microfibres in southern European deep seas”, PLOS ONE, Vol. 13/11, pp. e0207033-, https://doi.org/10.1371/journal.pone.0207033.
[63] Sieber, R., D. Kawecki and B. Nowack (2020), “Dynamic probabilistic material flow analysis of rubber release from tires into the environment”, Environmental Pollution, Vol. 258, p. 113573, https://doi.org/10.1016/j.envpol.2019.113573.
[36] Stanton, T. et al. (2019), “Freshwater and airborne textile fibre populations are dominated by ‘natural’, not microplastic, fibres”, Science of the Total Environment, Vol. 666, p. 377−389, https://doi.org/10.1016/j.scitotenv.2019.02.278.
[16] Suaria, G. et al. (2020), Microfibers in oceanic surface waters: A global characterization, Oceanography, http://advances.sciencemag.org/.
[65] Sutton, R. et al. (2019), Understanding Microplastic Levels, Pathways and Transport in the San Francisco Bay Region, San Francisco Estuary Institute, SFEI-ASC Publication No. 950., https://www.sfei.org/sites/default/files/biblio_files/Microplastic%20Levels%20in%20SF%20Bay%20-%20Final%20Report.pdf.
[84] Swedish EPA (2019), Microplastics in the Environment 2019, http://www.naturvardsverket.se/Om-Naturvardsverket/Publikationer/ISBN/6900/978-91-620-6957-5/.
[13] TextileExchange (2020), Preferred Fiber Materials - Market Report 2020, TextileExchange, https://textileexchange.org/wp-content/uploads/2020/06/Textile-Exchange_Preferred-Fiber-Material-Market-Report_2020.pdf.
[10] Thompson, R. et al. (2004), “Lost at Sea: Where Is All the Plastic?”, Science, Vol. 304/5672, p. 838, https://doi.org/10.1126/science.1094559.
[69] Tian, Z. et al. (2021), “A ubiquitous tire rubber–derived chemical induces acute mortality in coho salmon”, Science, Vol. 371/6525, p. 185, https://doi.org/10.1126/science.abd6951.
[4] UNEP (2020), Sustainability and Circularity in the Textile Value Chain: global stocktaking, https://wedocs.unep.org/20.500.11822/34184.
[40] UNEP (2017), Exploring the potential for adopting alternative materials to reduce marine plastic litter, United Nations Environment Programme.
[58] Unice, K. et al. (2015), “Experimental methodology for assessing the environmental fate of organic chemicals in polymer matrices using column leaching studies and OECD 308 water/sediment systems: Application to tire and road wear particles.”, Science of the Total Environment, Vol. 533, pp. 476–487, https://doi.org/10.1016/j.scitotenv.2015.06.053.
[54] Unice, K. et al. (2019), “Characterizing export of land-based microplastics to the estuary - Part I: Application of integrated geospatial microplastic transport models to assess tire and road wear particles in the Seine watershed”, Science of The Total Environment, Vol. 646, pp. 1639-1649, https://doi.org/10.1016/j.scitotenv.2018.07.368.
[53] Unice, K. et al. (2019), “Characterizing export of land-based microplastics to the estuary - Part II: Sensitivity analysis of an integrated geospatial microplastic transport modeling assessment of tire and road wear particles”, Science of the Total Environment, Vol. 646, pp. 1650-1659, https://doi.org/10.1016/j.scitotenv.2018.08.301.
[62] Verschoor, A. et al. (2016), Emission of microplastics and potential mitigation measures. Abrasive cleaning agents, paints and tyre wear, RIVM Report 2016-0026.
[23] Vianello, A. et al. (2019), “Simulating human exposure to indoor airborne microplastics using a Breathing Thermal Manikin”, Scientific Reports, Vol. 9/1, p. 8670, https://doi.org/10.1038/s41598-019-45054-w.
[50] Wagner, S. et al. (2018), “Tire wear particles in the aquatic environment - A review on generation, analysis, occurrence, fate and effects.”, Water Res 139, pp. 83-100, https://doi.org/doi:10.1016/j.watres.2018.03.051.
[79] WBCSD (2019), Global ELT Management – A global state of knowledge on regulation, management systems, impacts of recovery and technologies.
[72] Wik, A. et al. (2009), “Toxicity assessment of sequential leachates of tire powder using a battery of toxicity tests and toxicity identification evaluations”, Chemosphere, Vol. 77/7, pp. 922-927, https://doi.org/10.1016/j.chemosphere.2009.08.034.
[11] Woodal, L. et al. (2014), “The deep sea is a major sink for microplastic debris.”, Royal Society Open Science, Vol. 1, p. 140317−140317.
[41] WRAP (2019), Textile derived microfibre release: Investigating the current evidence base, Prepared by Resource Futures.
[43] Xu, X. et al. (2018), “Pollution characteristics and fate of microfibers in the wastewater from textile dyeing wastewater treatment plant”, Water Science & Technology, Vol. 78/10, pp. 2046-2054, https://doi.org/10.2166/wst.2018.476.
[32] ZDHC (2015), Manufacturing Restricted Substances List, Zero Discharge of Hazardous Chemicals Programme, http://www.roadmaptozero.com/fileadmin/pdf/MRSL_v1_1.pdf (accessed on 12 May 2021).
[18] Zhang, G. and Y. Liu (2018), “The distribution of microplastics in soil aggregate fractions in southwestern China”, Science of The Total Environment, Vol. 642, pp. 12-20, https://doi.org/10.1016/J.SCITOTENV.2018.06.004.
[39] Zhao, S., L. Zhu and D. Li (2016), “Microscopic anthropogenic litter in terrestrial birds from Shanghai, China: Not only plastics but also natural fibers”, Science of The Total Environment, Vol. 550, pp. 1110-1115, https://doi.org/10.1016/j.scitotenv.2016.01.112.
[28] Zuskin, E., F. Valic and A. Bouhuys (1976), “Byssinosis and airway responses due to exposure to textile dust”, Lung, Vol. 154/1, pp. 17-24, https://doi.org/10.1007/BF02713515.
Notes
← 1. As illustrated in Figure 1.4, releases into the environment of microfibres emitted during product use are particularly high in emerging economies (including major textile manufacturing countries such as China and India), mainly due to the lower rates of connectedness and treatment of wastewaters and the larger population sizes.
← 2. All washing methods are expected to contribute to fibre release, but there is limited knowledge on fibre release occurring during practices such as hand washing, steaming, or dry cleaning.
← 3. Road markings consist of plastic polymers, pigments, fillers and additives (Andersson-Sköld et al., 2020[61]).
← 4. The size range for TRWP was estimated to span from 4µm to 280 µm, with the mode centred around 50 µm (Kreider et al., 2010[47]).
← 5. This is not the rule in most other OECD countries. Porous asphalt is used in 95% of Dutch roads but only in 1% of roads in most other EU countries (Eunomia, 2018[81]). To maintain its functionality, porous asphalt pavements require regular street sweeping, which removes debris and pollutants (including TRWP).
← 6. Some studies have investigated risks associated with rubber granulate used as infill material. This is discussed in Section 2.3.2.
← 7. Council Directive 99/31/EC of 26 April 1999 on the landfill of waste (“Landfill Directive”) introduces a ban on the disposal in landfills of shredded and whole waste tyres, excluding tyres used as engineering material. The California Code of Regulations, Title 14, establishes that waste tyres may not be landfilled in a solid waste disposal facility, unless they are permanently reduced in volume prior to disposal.