EU Waste Framework Directive, What's Next? : A costbenefit analysis of an extended producer responsibility for textiles in the European Union

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EU Waste Framework Directive;

What’s Next?

A Cost-Benefit Analysis of an Extended Producer

Responsibility for textiles in the European Union

Amanda Gerbendahl Madeleine Johansson

Handledare: Pernilla Ivehammar

Linköpings universitet SE-581 83 Linköping, Sverige 013-28 10 00, www.liu.se

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Abstract

The objective of our thesis was to conduct a pilot study to evaluate if an Extended Producer Responsibility (EPR) for textiles in the EU could be a socioeconomically beneficial policy to complement the EU Waste Framework Directive’s amendment of separately collected textiles. The aim was to investigate if the policy could achieve increased circular design of textiles as well as if it could work as a management plan for the increased collection rates. The evaluation was made with a Cost-Benefit Analysis, using the French EPR-system for textiles as a base. It was further complemented with previously conducted research of EPR-systems for other waste streams in the EU, as well as by previously conducted investigations for other national implementations of producer responsibilities for textiles. In additional support, we used data for differences between the member states in the European Union and conducted an expert interview. The EPR was compared to a situation where the municipalities in the member states would instead be responsible for the separate collection of textiles.

The result of our investigation illustrates how both alternatives generate a net-loss, the Municipal Responsibility with - €7,611,410,291 and the Extended Producer Responsibility with - €6,012,109,341 during the first year of implementation. The EPR alternative generates a lower net-loss during the first three years of implementation. The producer responsibility is however the less beneficial alternative four years after implementation, since the decreased opportunity cost of labour generated through the hiring of unemployed assumed under the producer responsibility is deducted. The producer responsibility does however generate benefits through clearly defined responsibility of the textiles placed on the European market and gives incentives for increased fibre-to-fibre recycling and for increased durability of textiles. The initiative therefore generates both higher quantifiable-and non-quantifiable,

environmental benefits than the alternative. We conclude that an Extended Producer

Responsibility should be further examined as a complement to the regulation of separate collection of textiles, to reach an increased circular textile industry.

Keywords: EU Waste Framework Directive, Extended Producer Responsibility, circular

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Sammanfattning

Syftet med vår uppsats var att genom en pilotstudie utvärdera om ett utökat producentansvar (EPR) för textilier i EU skulle vara en samhällsekonomisk lönsam policy för att hantera Avfallsdirektivets tillägg om separat insamling av textilier. Undersökningen hade för avsikt att utvärdera om policyn skulle kunna leda till ökad cirkulär design av textilier samt fungera som en hanteringsplan för den ökade insamlingen av textilier. Utvärderingen genomfördes med en kostnad-nyttoanalys med den franska implementeringen av ett EPR-system som grund. Den kompletterades med tidigare utförda studier på EPR-system för andra avfallsflöden inom EU samt med tidigare utvärderingar av nationella implementeringar av producentansvar för textilier. Som ytterligare underlag användes även data över skillnader mellan medlemsstaterna i den Europeiska unionen samt en expertintervju. EPR-systemet jämfördes sedan mot att ländernas kommuner istället skulle vara ansvariga för den ökade insamlingen av textilier. Resultatet av vår undersökning visar att båda alternativen genererar en nettoförlust, kommunansvaret med - €7,611,410,291 och producentansvaret med - €6,012,109,341 under implementeringsåret. Det utökade producentansvaret genererar en mindre förlust under de första tre åren av implementering. Efter fyra år blir dock producentansvaret det mindre lönsamma alternativet, och genererar en nettoförlust mot alternativet på - €181,811,951. Detta eftersom den minskade alternativkostnaden för arbetskraft som antas genereras av att anställa arbetslösa under det utökade producentansvaret försvinner. Det utökade producentansvaret genererar dock nyttor i form av tydlig ansvarsfördelning för de textilier som placeras på den europeiska marknaden, incitament för fiberåtervinning samt incitament till ökad livslängd på textilier. Detta gör att initiativet får både högre kvantifierbara-och icke-kvantifierbara,

miljömässiga nyttor än alternativet. Utökat producentansvar för textilier bör därför fortsätta

utredas som en policy för en ökad cirkulär textilindustri, samt som en komplimenterande policy till EU:s avfallsdirektiv. Vi drar slutsatsen att ett utökat producentansvar för textilier bör fortsätta utredas som ett komplement till bestämmelsen om separat insamling av textilier, för att uppnå en ökad cirkulär textilindustri.

Nyckelord: EU:s avfallsdirektiv, utökat producentansvar, cirkulär ekonomi, fast-fashion,

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Acknowledgement

We would like to thank our supervisor, Pernilla Ivehammar, for her support and feedback throughout the process of writing our thesis. We would further like to thank our opponents and seminar group for interesting discussions and helpful comments. A special thanks is further given to Birgitta Losman and Yvonne Augustsson for their participation in our interview.

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Tables

Table 1. EU policy tools for textiles addressing different aspects of circular economy (European

Commission 2019a)... 3

Table 2. The social costs and benefits from implementing an EPR-scheme... 35

Table 3. Summary of rates: today and after an implementation of separate collection of textiles ... 41

Table 4. Net environmental benefits per kg of additional post-consumer textiles collected... 42

Table 5. Values €/kg CO2-eq (Noring vs Stern) ... 43

Table 6. Saved costs in € for less usage of CO2-eq due to increased reuse and recycling ... 43

Table 7. Saved CO2-eq and water usage as a result of longer durability of textiles ... 50

Table 8. Summary of the operational costs for the different stakeholders in the EU for the two alternatives, for the year of implementation ... 52

Table 9. Summary of the quantifiable environmental benefits for the two alternatives ... 52

Table 10. Net effect of EPR, using the quantifiable operational effects and the quantifiable environmental benefits ... 52

Table 11. Effect of EPR from the second and third year ... 53

Table 12. Effect of EPR from the fourth year and forward... 53

Table 13. Summary of side effects of an EPR implementation for both alternatives, and the effect of EPR ... 55

Table 14. New net effect from changes in EOCL ... 59

Table 15. Environmental savings in € from decreased CO2-eq in Municipal Responsibility ... 59

Table 16. Environmental savings in € from decreased CO2-eq in Extended Producer Responsibility 59 Table 17. Avoided CO2-eq- and water usage per kg of additional post-consumer textiles where 30% goes to reuse and 30% goes to recycling ... 60

Table 18. Cost of CO2-eq for different displacement factors-and reuse rate, Stern values ... 60

Table 19. Cost of CO2-eq for different displacement factors and reuse rate, Noring values ... 61

Table 20. New net effect from using different opportunity costs of sorting textiles ... 61

Table 21. Difference in net effect from deceased sorting time ... 62

Table 22. Net effect from having sorting activities placed in CEEC which results in difference in salaries and economic opportunity cost ... 63

Table 23. New net effect from increased DfE ... 64

Table 24. Difference in net effect from decreases in generated waste ... 65

Figures

Figure 1. Image based on Bukhari et al. (2018), material flow of post-consumer textiles in France .. 24

Figure 2. Effect on the textile value chain, as a result of an implementation of EPR, inspired by Krüger, Plannthin, Dahl, and Hjort (2012) ... 49

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Definitions

CEEC: Central- and Eastern European Countries CHF: Clothes, household textiles, and footwear

Circular economy: “In a circular economy, the value of products and materials is maintained for as

long as possible. Waste and resource use are minimised, and when a product reaches the end of its life, it is used again to create further value.” (European Commission n.d.)

CO2-eq: Carbon dioxide equivalent

DfE: Design for Environment. “Design for Environment is a design in which environmentally

destructive factors included in every stages of product life cycle are eliminated as much as possible and shifted into recyclable condition, resulting in reducing depletion.” (Kim 2010, p. 816)

EEE: Electrical and Electronic Equipment ELV: End-of-life vehicles

EPR: Extended Producer Responsibility. “An environmental approach in which a producer’s

responsibility for a product is extended to the post-consumer stage of a product’s life cycle.” (OECD 2001a, p. 9)

EU: European Union of 27-member states (without the UK)

Post-consumer textile waste: Products such as used clothing, footwear, accessories and home textiles

that have been discarded by the consumer after use (Maldini et al. 2017)

Pre-consumer waste: “Pre-consumer waste is manufacturing waste that has not reached the consumer”

(Redress 2017, p.27)

PRO: Producer Responsibility Organisation. It is an entity formed by producers, usually set up by

legislation, with the purpose to help the individual producers to reach the the goals of the EPR-Scheme, as well as to govern the function of the system (European Commission 2014a; OECD 2004)

Producer definition for EPR: The manufacturers of CHF, those who place orders for manufacturing

of CHF and those who imports CHF to the French- (EU) market (Eco TLC n.d.)

SSE: Social and Solidarity Economics

WEEE: Waste Electrical and Electronic Equipment (Cahill et al. 2010) WFD: Waste Framework Directive

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1. Introduction

1.1 Background

As global warming is increasing due to climate change, several global actions, such as the Kyoto Protocol and the Paris Agreement, have been made to prevent climate change from aggravate. The European Union (EU) has itself decided on several initiatives to prevent climate change, one of them being the European Union’s 2030 sustainability agenda for Europe, with an initiative to move towards a circular economy (Kettunen et al. 2019). A transition towards a circular economy is in line with key EU priorities, since businesses can be protected from scarce resources, and new business opportunities may spur innovation and enhance job opportunities (European Commission 2015; Moula et al. 2017; Kalmykova et al. 2018). It is also said to create environmental benefits such as energy savings, lower carbon dioxide emissions, as well as to create a solution to avoid irreversible damages from using up non-renewable resources, since the usage of secondary material will increase (ibid.).

The world is today relying on a linear economy, where the tradition is to buy products, use them, and then to throw them away, with usable material being wasted through incineration or landfill (Platform for Accelerating the Circular Economy (PACE) 2020). To cope with a transition towards a circular economy, and to handle the existing resources on the European Market, the EU Waste Framework Directive 2008/98/EC of the European Parliament and of the Council of 19 November 2008 on waste and repealing certain Directives, was introduced in 2008. It sets the basic concept for waste management with the purpose of making both the resource use more efficient, and to protect the environment from the negative effects of waste generation. It further introduces concepts such as the “polluter pays principle” and the “Extended Producer Responsibility (EPR)”.

EPR is an economic instrument to incorporate a polluter pays principle, and it extends the producers’ responsibility to cover for the products’ entire life cycle, thus from resource extraction all the way to the end-of-life management (OECD 2001a; Monier et al. 2014). In the Directive 2008/98/EC, EPR is suggested as an economic instrument to achieve the objectives of increased reuse and recycling and is since then mandatory in the waste streams for batteries, end-of life vehicles (ELV) and electrical and electronic equipment (WEEE). EPR-systems are also used as a complement to the Packaging and Packaging Waste Directive, and several other country-level-initiatives use EPR-schemes to cover additional waste streams (Monier et al. 2014). EPR will further be mandatory for packaging waste from 2024 according to Directive (EU) 2018/852 of the European Parliament and of the Council of 30 May 2018.

There are however no requirements for the EPR concept in the textile industry, even though the European Commission has identified the textile industry, and especially the apparel industry, to be of priority to convert into a circular one (Manshoven et al. 2019). Increased circularity for textile is further targeted in von der Leyen’s (2019) agenda for Europe, where

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she states that increased focus on sustainable resource usage in the textile industry is the focus of the new circular economy action plan. Certain actions towards a more circular management of textiles is implemented in the extended Directive 2008/98, namely Directive (EU) 2018/851 of the European Parliament and of the Council of 30 May 2018 amending Directive 2008/98/EC on waste, that establishes that all textiles must be collected separately in the member states no later than 2025. The directive further establishes a minimum preparing for reuse/recycling rate of at least 60% for municipal waste, including textiles, by 2030. It also states that separately collected material cannot be sent for incineration1_{unless it generates the} highest economic benefit. Directive (EU) 2018/850 of the European Parliament and of the Council of 30 May 2018 amending Directive 1999/31/EC on the landfill of waste further states that by 2030, all member states shall make sure to not landfill waste that is suitable for reuse, recycling or other recovery.

Increased circularity for the textile sector is promoted due to its severe environmental impact (Manshoven et al. 2019). Ellen MacArthur Foundation (2017) argues that the greenhouse gas emissions aligned with textile production in 2015 released more CO2-eq than emissions released from both flights and shipping combined. Worldwide, apparel and footwear together account for around 8% of all greenhouse gases emissions released (European Commission 2019a). In 2017 alone, 675 million tons of virgin material are estimated to have been used in the production of shoes, household textiles and apparel (Manshoven et al. 2019). The production of clothes further requires large amounts of water and chemicals for dyeing (Šajn 2019), leading to overall challenges regarding resource limitations and deterioration of human health (Ellen MacArthur Foundation 2017; Sandin & Peters 2018). The greatest deterioration from this production, is further seen outside the EU boarders. This goes both for its effects on human health through production where certain usage of azo dyes for dyeing has long-run consequences, with some showing indications of causing cancer (KEMI 2014; Ellen McArthur 2017), as well as the use of CO2-eq, virgin material, land and water (Manshoven et al. 2019). Even though some of the environmental damages burden the globe overall, such as greenhouse gas emissions (Brännlund & Kriström 2012), the textiles produced for the EU market mainly affects countries outside the EU. The production of textiles thus needs to be targeted both to reach circularity, but as well to decrease the inequality and unbalance of who bears the utility from the production, and who bears the burden.

A transformation to a circular economy and increased recycling of textile products could however be a potential solution to these challenges, according to Roos et al. (2016). Sandin and Peters (2018) argue that the environmental benefits from a circular economy in the textile industry originate both from less usage of primary resources in production, as well as decreased impact in their after-life and waste management. Implemented policies for circularity in the textile industry therefore need to cover both the recycling and reuse of clothes, as well as the production itself.

1_{If this covers all types of incineration (e.g. also incineration to energy recovery) is however not clear. Sandin} and Peters (2018), do for example highlight how recovery is sometimes included in the term “recycling”.

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The textile value chain is however argued to be one of the most challenging production models to handle when it comes to sustainable production, mainly due to its complexity in material composition and its long production chain (European Commission 2017a). To fulfil a circular economy in the clothing industry, Ellen MacArthur Foundation (2017) argues for more environmentally friendly production and easier recycling possibilities of the clothes, as well as higher quality and longer durability of the apparel. Effective usage of recycled clothes is further promoted, where all the material can be used to create inputs, either as fibres for new textiles or as e.g. insulation. The European Environment Agency (2019) promotes the Extended Producer Responsibility as a policy that can reduce the textile sector’s impact on the environment and convert today’s linear production system into a circular one.

Until today, France is the only country in Europe that has implemented an EPR-scheme for textiles (Watson et al. 2014; Bukhari et al. 2018), but several other countries, such as Ireland (Filho et al. 2019), the Netherlands (ten Wolde 2019) and Sweden (Regeringen 2019), have over the last couple of years started investigating in similar systems. The Swedish environment agency has however highlighted how the possibilities of forming an EPR for textiles are very limited on a national level, due to lack of support from the EU bodies. They argue for the need of a legislation on an EU-level (Naturvårdsverket 2016), and this study aims to work as a pilot for such implementation.

1.2 Problem formulation

The EU has already established some policy tools addressing different aspects of a circular economy in the textile industry. For instance, the REACH regulation that regulates the usage of chemicals in production (European Commission 2019a), the EU Eco label which ensures that products have a limited impact on polluting waters, air and affecting health (Šajn 2019), and the textile regulation, implemented to offer consumer relevant information of the composition of the product (Norstedts Juridik 2019). The WFD for waste handling and recycling will be implemented by 2025 with the separate collection of textiles, but the EU is still lacking a policy regarding minimum requirements for circular design, see Table 1. Table 1. EU policy tools for textiles addressing different aspects of circular economy (European Commission 2019a)

EU policy tools on safety, including sectoral legislation (non-exhaustive) EU policy setting minimum requirements for circular design EU policy tools promoting Sustainable production and/or consumption EU Policy tools on waste handling/recycling Textiles REACH Regulation GPSD - Textiles Regulation EU Ecolabel GPP WFD (from 1/1 2025)

As McCarthy, Dellink, and Bibas (2018) evaluated the macroeconomic impacts of a transition to a circular economy, they found that circular economy policy instruments do lead to lower natural resource extraction, making policy instruments an important part of this transition. In a survey conducted on both European consumers and producers it was however stated that the

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EU’ s existing environmental policies targeting the textile industry were insufficient, and 60% of the respondents found the different number of environmental labels confusing (European Commission 2019a). Over 83% of the respondents also stated that they would be willing to pay more for a product they knew were more sustainable (ibid.). One way to avoid large number of confusing labels, as well as increasing both durability and sustainability of textiles, could be to fill the gap in Table 1, ensuring consumers that textiles bought within in the Union have a circular design. As no such policy have been found, this study tries to link the regulation of separate collected textiles, with a directive for circular design.

As we move in to the 2020’s, the Extended Waste Framework Directive of separate collection of textiles is approaching. No policy has however been found on responsibility for the management of the textiles, or how to handle the increased collection rates. Germany has one of the world’s highest collection rates of textiles today but the majority of the collected textiles ends up as textile mountains in different warehouses; a risk the entire union might face if no clear plan for the increased collection of textiles is presented (Henkel 2019). In Germany, the producers are not responsible for the textiles placed on the market (ibid.) and there is thus no one responsible for the textiles all the way from resource extraction and design for circularity, to disposal of the textiles.

The initiative “European Clothing Action Plan”, increased the focus on the clothing supply chain and covered several areas such as influencing design for longevity, implementation of more sustainable fibres, influencing more sustainable production practices, improving collection rates of used textiles, and the integration of more recycled fibres in the textile industry (ECAP n.d.). This initiative, that ended in 2019, hence worked to improve some of the areas that an EPR-scheme would. As an EPR for textiles is already implemented in France, and that it is commonly used for other waste streams, and is investigated by several other countries in the textile sector; an EU-policy regarding EPR for textiles could be contemplated as a complementary policy to fill the gap of what to do with the increased collection of textiles. An EPR policy could further work to promote a more circular design of textiles (Walls 2003; Smith 2005; Filho et al. 2019).

An investigation of such policy’s effectiveness is important to prepare for an eventual implementation, and this can be done with a Cost-Benefit Analysis (Boardman et al. 2018). As France is the only European country with an EPR-system today, this study aims to evaluate with a CBA if the French EPR-system could work as a complement to the Extended Waste Framework Directive on textiles to achieve circular design, and if it is a socio-economic beneficial way of handling the separately collected textiles.

1.3 Aim

This study aims to work as a pilot study for evaluating if the French EPR-scheme on textiles could be socioeconomically beneficial to implement on the EU Member States. It is evaluated as a complement to Directive (EU) 2018/851 to achieve increased circular design of the textiles, and as a management plan for the increased collection rates. Due to the Subsidiarity and

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proportionality principle2_{set out in the Treaty on the European Union (Consolidated Version} of the Treaty of the European Union 2012), the policy evaluation is made from a European Union perspective and leaves the precise implementation methods to the national experts. A socioeconomic Cost-Benefit Analysis (CBA) is made with support from theory, literature, data, and an expert interview, where the cost-efficiency and spill-over-effects of such policy is assessed and put in relation to achieve a circular economy in the textile industry in the European Union. The CBA compares the EPR of waste management, with a Municipal Responsibility that is assumed to be the option if no actions are taken.

1.4 Limitations

At first, this study aimed at investigating how trade would be affected by introducing requirements for recycled materials in the production of textiles, and thus fill the gap of policies for circular design. We did however find the technical solutions for this to be very limited today (Östlund et al. 2015; Sandin & Peters 2018; Filho et al. 2019) and that a requirement of recycled material in production might be non-realistic. Even though there exist targets for reuse/recycling for municipal waste, there is no target of how much of the recycled material should be used further in production, and due to technical limitations, such policy seemed far ahead. As an EPR-scheme however could enhance recycling activities (OECD 2001a; Walls 2006; Watson et al. 2014), this aim was chosen instead.

This study originates from the assumption of 100% collection rate of textiles, and thus that the implementation of EPR is introduced after the separate collection of textiles has been implemented in 2025. The Directive (EU) 2018/851 states that all textiles should be collected separately, but Bartl (2020) argues that the actual collection rate target has not yet been set. This study thus originates from a perspective that might be an overestimation of the number of textiles collected. It is accordingly analysing the consequences of a full collection rate and which opportunities and obstacles that are aligned with it, which could be relevant if the separate collection of textiles is fully achieved. The study is thus not analysing whether this type of policy could enhance collection rates, which is a change from former EPR-systems where the aim often is to achieve higher collection (Monier et al. 2014).

Since only France has previously implemented an EPR-scheme for the textile industry, the impacts in the specific member states are not possible to assess properly. There is further limited data on the policy's actual effectiveness, and limitations in data for all posts of our CBA. The numbers that are quantified are done so by cost estimates based on the French EPR-system as well as Naturvårdsverket’s, (2016) and Soutukorva and Wallentin’s (2016) hypothetical EPR-programme in Sweden, modified to an EU price level. The results of the CBA are hence corresponding to an “EU average”-country rather than an actual estimate ready to be used by individual member states. In addition, as the Cost-Benefit Method uses an ex-ante approach,

2_{Decisions should be taken as close to the individual EU citizen as possible and the EU institutions shall not} interfere with more actions than necessary to achieve the objectives of the treaties (Eurlex, n.d.a; Eurlex, n.d.b).

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where both studied alternatives originate from a future state when WFD is in place, complementary studies will have to be made both before, during, and after implementation. This study is in addition written from an economic perspective, without expertise in the areas regarding legal implementation and technical limitations of the recycling industry. Laws aligned with the trade of secondary material that might need changes from an EPR, as argued by OECD (2001), and the possibilities for implementation and harmonisation with national regulations in the member countries, are hence areas that are not covered in this study. A clear end-of-waste criteria for textiles is according to EURATEX (n.d.) also missing, and the Waste Framework Directive hence induces legal uncertainty. As the interpretations of end-of-waste criteria are different, including for textiles, EURATEX (n.d.) argues that there is a need for a common definition on an EU-level of which textile waste ceases to be waste in order to provide legal certainty to national authorities. This is further not discussed or targeted within the framework of this study. The technical limitations of recycling are however briefly discussed in section 3.1.1.1 Consideration is equally not taken to differences in fibre composition nor to differences in environmental impact/possibilities for recycling for different types of textile material.

For a transition towards a circular economy to be possible, both production and consumption patterns must change to adjust from a linear trend of producing and consuming, to a circular one. This study is focusing on regulations that affect the production of textiles and therefore not on the consumption and consumer behaviour, which limits the result. Nonetheless, if the Extended Producer Responsibility can work as an incentive for producers to use more circular design, then more sustainable options for the consumers are available in the first place. We therefore consider a regulation on the producers to be of priority.

1.5 Methodology

McCarthy et al. (2018) argue that ex-ante, economy-wide and quantitative models are the best kind of tools for evaluating macroeconomic effects of a transition to a circular economy. As spill-over-effects are likely to take place across sectors and regions due to a circular economy, it is essential to make an economy-wide modelling. Further, the authors mean that the quantitative aspect is to be able to get a comprehensive picture of the interactions from implementing a circular economy. Policy instruments that have been suggested to achieve a circular economy have however not been implemented historically and are aligned with uncertainty regarding which ones that are the most effective (ibid.).

We chose an ex-ante Cost-Benefit Method for analysing this policy, since it identifies all possible social costs and benefits, meaning the costs and benefits to society as a whole (Boardman et al. 2018). A CBA-approach is effective when evaluating a policy’s implication on an economy since it includes possible spill-over-effects, and the European Commission does promote the usage of CBA-analyses for larger infrastructure projects within e.g. waste management (European Commission 2014b). The information for our CBA was gathered in four different ways; through theory, a literature review, data collection of the EU member

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states, and through an interview, in order to find necessary information to evaluate the performance of previously implemented EPR-systems, and which implications that are aligned with an implementation of EPR in the textile industry. The policy evaluation was based on the French EPR-system, but the other mandatory EPR-systems implemented on an EU-level, as well as other evaluations of national implementations of EPR-systems for textiles, were used to understand which obstacles and possibilities that are aligned with this type of policy. To understand the differences between the EU’s member states, data was collected on the aspects necessary for our implementation. Data has as well been collected on the performance of earlier implemented EPR policies in the union. To calculate the countries’ performance in the already implemented EPR-schemes, a system of “differences-to-average” was conducted. We did calculations on collection rates, recycling, reuse, and recovery of different waste-streams under EPR-schemes and compared each member states’ performance compared to the EU average. When all these differences were summarized, we got an “EPR-scoreboard” for the member states, to see which countries that, on average, perform more successfully than others. This is important for our analysis to investigate the overall functioning of the system, and to see if the implementations of EPR-systems are equally successful everywhere. A table of the EU member states’ differences in terms of inhabitants, unemployment rates, price levels, areas, and average salary per hour can be found in Annex 4 (A3). A summary of how well the targets for other EPR-implementations has been reached by the different member states can be found in Annex 5 with the results of the differences-to-average investigation.

One of our methods for collecting empirics, was a 60-minute-long interview with Birgitta Losman3_{, responsible for the investigation of an eventual EPR-scheme for textiles in Sweden,} and Yvonne Augustsson4_{from Naturvårdsverket, who assists Losman with textile expertise} during the investigation. The two were chosen due to their knowledge of policy implications aligned with implementing an EPR-system in the textile sector, as well as the benefits that are most likely to occur from it. Due to their expertise within the area, we decided to conduct semi-structured interviews where our primary interest still could be answered, whilst their knowledge within the area could still present itself without strict limitations (Bell et al. 2019). As only one interview was conducted, and as the questions were very specific to experts in the area, no pilot interview was made, and the questions could therefore not be modified to be more specific and/or relevant. Our interview was based on an interview guide5_{, but the questions} were slightly modified to fit the discussions in the interview and to ask follow-up-questions. The interview was recorded, with the participants approval, to make sure that all elements of the situation were captured, but notes were taken during the mean time to make sure we could separate the two interviewees during transcription. Due to the spread of COVID-19 during the spring of 2020, the interview was conducted over Skype from three different locations. It was held at each participants’ homes, which was limiting disturbance that could be aligned with

3_{Birgitta Losman, samordnare för hållbar utveckling, Högskolan i Borås, telephone interview on the 17th of} April 2020.

4_{Yvonne Augustsson, handläggare, Naturvårdsverket, telephone interview on the 17th of April 2020.}

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public places. The interviewed agreed to have their names and roles published, and their approval has been received for the results of the interview.

Furthermore, our CBA was conducted in a way that it identified, quantified, and valued which consequences the EU must consider if an implementation of the system would be in question, even if quantification of all costs and benefits might not be possible. Since we wanted to tackle an environmental problem, we had to handle the uncertainty aspects of long-term investments and future benefits (Bohm & Henry 1979). The results were therefore examined with a sensitivity analysis. As the CBA was conducted for the primary year of implementation, with investment costs and costs that are on a yearly basis, and only some consideration given to the three following years, no discounting of the results was made.

1.5.1 Ethics

This study aims at presenting the data, information, and assumptions with high transparency to ease for future studies to complement our work. Vetenskapsrådet (2017) has published ethic guidelines for good research practices, and these are considered during our work. For our interviews, the respondents had full information about the aim of the study, and their consent was given for publishing the results, their names, and the recording of our discussion. The respondents were aware of the opportunity to end their participation in the study at any time, and that the discussions in the interview was confidential and would not be published without the respondent's approval.

1.5.2 Methodological criticism

The Cost-Benefit Method is built on the rationale that all goods can be valued; even the ones without a market price such as the environment (Mattsson 2006). This basis has been questioned and criticised for the usage of e.g. human preferences as valuation method, with the arguments that humans are not always rational and that the answers are rather measures for attitudes rather than economic preferences (Kahneman et al. 1999; Chee 2004). Other criticisms are how results without a monetary value easily could be ignored (Mattsson 2006). Yet, Mattsson argues that the qualitative approach to CBA has emerged to cope with the issue of not ignoring unquantifiable posts, hence the above-mentioned identification of costs and benefits, rather than only quantification of results. This gives an opportunity to weigh all possible outcomes against each other, and to determinate if the identified outcomes without a monetary value could increase the benefits enough to cover the costs. The method is further considered useful for this study as Chee (2004) states that the valuation of goods without market prices, can help to inform the policy makers with essential information for decision making. Matson (2006) lastly opines that the usage of people’s preferences of the eco-systems, are the best practice for valuation of environmental changes, despite the lack of full rationality. The ground for our CBA is to a large extent built on articles about the French EPR-system, and these articles are mainly written in French, which give certain difficulties in exact translation of the information. As much information regarding the French EPR-system is presented in French, we still find it necessary to use these references. Yet, English articles are used to some

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extent, as well as the translated articles about the French PRO, to compare with the translated French articles and see if the information matches. There might also be a lack of overall information due to language barriers, as reports written in other local languages have not been researched within this study.

Certain difficulties in summarizing and comparing statistics and data of sectors where EPR is applied has further been highlighted by Monier et al. (2014). There has been a lack of transparency in the data reporting, as well as problems with data being reported in different ways (ibid). These difficulties may affect the results of the performed studies, as well as the data presented in our differences-to-average method. The extended WFD (2018/851) targeted the lack of transparency in the data reporting, as new rules for data collection and reporting for waste management was implemented. The methods are both simplified and intended to harmonize the reporting from the member states (European Environment Agency 2018). A large part of the communications regarding circular economy issues from public- and private organisations, such as Ellen MacArthur Foundation (Moreau et al. 2017). Referring to non-publicised sources, especially private organisations, might be biased due to underlying agendas from the authors of certain reports. Since much of the research within the circular economy area is conducted by such organisations after all, and since the European Commission uses e.g. Ellen MacArthur Foundation in their publications, we choose to use that type of sources within this study as well: with consideration taken to eventual biased results. A large part of our literature is conducted on a general level by studies made by EU and OECD, mainly because country-specific reports cannot not be found. Yet, since our evaluation is made for the EU, we consider these reports sufficient.

McCarthy et al. (2018) further argue that EPR is a ‘soft policy’ that does not directly affect prices, which makes it among the most difficult policy to model. This is mainly due to lack of data on effectiveness and costs, and that it tends to be adjusted for local circumstances as well as with different designs. The data presented to the EU on the Waste Management in the member states is in fact not collected and reported similarly (Monier et al. 2014; Maldini et al. 2017), and there might be misleading information in the data on waste management. The data does further not always seem to be updated, or presented at all, which can be seen in the case of collected textiles in Sweden. According to the data presented by Eurostat (2020h), Sweden has collected 41 tons of textiles, while Watson et al. (2018) present a collection rate of 23,000 tons. The usage of the French and Swedish implementation of an EPR could further be argued to differentiate from other countries, both in costs and in adjustments for local circumstances. The analysis is however made with the latest data we can find for the EU as a whole and is as well adjusted to an average EU country rather than the specific countries themselves. The expenses are modified to equal an EU average cost, and the discussion highlights some local circumstances that might hinder an EU average implementation.

Concerning the interview, as the interviewed chose to not be anonymous and as they represent the Swedish government in the EPR-investigation, it may have affected the responses they gave. Due to their expertise in the area, we nevertheless argue that it is beneficial to present

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their names, to strengthen the results of our study. Experts are moreover only represented from one country, which could lead to missing results regarding what others would experience as difficult with an implementation. Since previous investigations of EU-wide EPR-systems, as well as data on waste management from other waste streams with EPR are analysed, we do however believe that a large part of the EU’s assembled challenges are represented in our results.

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2. Theoretical framework

The motivation behind the selected theories presented in this section is to create a better understanding of the concepts behind our study and the effects an implementation of the French EPR-scheme on textiles could have on the EU’s member states. Theories that are covered are EPR as an economic instrument in environmental economics, circular economy, and CBA.

2.1 Economic Instruments in Environmental Economics

The rationale for government intervention in a market comes from market failures in terms of reaching economic efficiency in the presence of externalities. In an efficient market, the term Pareto Optimality is used to describe optimal usage of resources; referring to a state where no re-allocation of goods can make anyone better off, without making anyone else worse off (Boardman et al. 2018). The original criteria have however been modified to increase its feasibility, and the Kaldor and Hicks criterion is today more commonly used to describe the perfect market economy. Kaldor (1939) and Hicks (1939) argue that a policy should be implemented even if someone is worse off by the implementation, if those who gain from it can compensate the losers while still being better off. Efficient resource usage can however not be reached in the presence of externalities, where the market prices do not reflect the true cost of production or consumption (Brännlund & Kriström 2012).

Coase (1960) argues for the need of defining property rights to reach an optimal solution for these externalities, as property rights defines who is responsible for what in a production and consumption process. He further argues that the market cannot always define these property rights itself, due to high transaction costs aligned with bargaining arrangements and several market players involved in the market. He argues for government intervention to reach efficient solutions in such market situations (ibid.). One way to define the property rights with government intervention, is to induce the polluter pays principle. The principle builds on the rationale that the polluter should bear the full costs of its activity, both in a preventing stage to minimize the damage, and to compensate for the damages caused (OECD 1995). Several costs are however aligned with government intervention as well, ranging from administrative costs, to the possibilities for dynamic adjustments of the policy and uncertainty of the outcomes (Fullerton 2001).

Brännlund and Kriström (2015) argue that in terms of finding the most effective environmental policies, the marginal utility from the improvement in environment should be equal to the marginal cost of making that improvement, and the policy should as well be the most cost efficient. However, to make sure that these criteria are met, there is need for perfect information about costs and environmental impacts. In reality though, real costs and damages are often difficult to find. As governments intervene with the market to internalize the externalities, several policies can be used. Command-and-control instruments, such as regulations- and standards (Stavropoulos et al. 2019), are the most widely used instruments of governmental interventions today (Fullerton & Muehlegger 2019), but market-based instruments like taxes and tradable permits are getting more popular among OECD countries (OECD 2010). Jacobs

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(1997) argues that more efficiency is achieved with market-based instruments, since consumers and producers are given financial incentives to change the behaviour that is causing damaging effects on the environment. The usage of market-based instruments also leaves room for the producers to abate pollution at the most cost-efficient level, without the need for governments to have the necessary knowledge about the abatement costs within the production chain (Fullerton 2001).

An Extended Producer Responsibility is an economic instrument used in environmental economics and can be implemented through both market-based- and command-and-control instruments; with a mix of both being common (OECD 2016). What separates the policy from an implementation of a market based- or command-and-control policy alone, is how it incorporates the polluter pays principle (Ferrão 2002; Nnorom & Osibanjo 2008). EPR extends the polluter pays principle in such a way that others in the production chain may be incorporated as the “polluter”, since the definition of producer refers to both importers, distributors and manufacturers of products (OECD 2016). As governments rarely have the necessary knowledge about the abatement costs within the production chain, and the market price of the textiles does not reflect the true cost of production due to e.g. the fast-fashion phenomenon, a market failure arises. Government intervention in the form of an EPR-policy could hence work as a possible policy to reduce these market failures, where producers are being faced with the responsibility of taking waste management into account in the production process.

There are two goals of an implemented EPR-policy, a physical and/or economical shift of responsibility towards producers and away from the municipalities, and increased incentives for producers to take the environmental impacts into consideration in the design of their product (OECD 2001a; European Commission 2019b). An EPR thus prolongs the responsibilities of the manufacturers, distributors and importers and creates a situation where they are financially responsible for the entire cycle of the products they create and sell (Smith 2005). The municipalities do thus no longer have to carry the cost burden for the waste disposal. If the producers are to be responsible for the financial burden of the end-of-life waste management instead, they are argued to be encouraged to reduce the costs of it and hence to affect the upstream activities and Design for Environment (DfE). Reduced costs may result in higher profits or higher market share (ibid.).

EPR is widely used as a policy instrument to achieve a more circular economy, but the most efficient design of the policy is still under consideration (OECD 2018). There are several policies which adheres to EPR, such as tradable recycling credits, product take-back-schemes, advanced disposal fees, deposit-refunds and recycled content standards etc (Walls 2003; OECD 2004). However, one needs to consider the different conditions for the products for which the EPR is to be implemented to find out which policy is the most appropriate, as the EPR concept itself gives little guidance about which policy is the best (Walls 2003; Walls 2004). Most EPR-systems have an impact on recycling rate, as well as an impact on the usage of virgin material, but not all EPR-system are likely to have an impact on product design according to Walls (2006). Kaffine & O’Reilly (2015) argue that DfE is better reached if policies aimed for recyclability are used and promotes market-based policies such as a deposit/refund system.

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Another market-based instrument, the usage of eco-modulations, is argued by Calleja (2019) to be most efficient to ensure a circular design and recyclability of the material. This study evaluates the French scheme, and therefore assesses take-back-schemes as an EPR-policy, together with a decreased fee paid by those producers who manage to DfE.

2.1.1 Producer Responsibility Organisation (PRO)

An Extended Producer Responsibility policy states individual obligation on the producers, but cooperation between different producers is usual in order to fulfil their tasks from the EPR-policy. A Producer Responsibility Organisation (PRO) is an example of a type of cooperation that can be formed (OECD 2004; Monier et al. 2014). In order to understand how EPR works, it is essential understanding what a PRO is. A PRO is an entity formed by producers, usually set up by legislation, with the purpose to help the individual producers to reach the goals of the EPR-scheme, as well as to govern the functioning of the system (OECD 2004; Monier et al. 2014). Kunz et al. (2018) argue that a transfer of the waste responsibility from municipal authorities to producers might be complex as it requires producers to collect, sort and treat the waste they place on the market. OECD (2004) argue that to ease this activity, and to facilitate the requirements from an EPR, companies can work together through collective initiatives and by creating a PRO. A PRO can utilize the existing infrastructure and negotiate with involved parties that are engaged in the collection, transportation, sorting and recycling, and ease the burden on the concerned producers (OECD 2004). An EPR further needs governance, and a PRO, or several competing ones, can work as a governance body as well (OECD 2016).

2.2 Circular Economy

As EPR could be seen as a policy aiming for circular design, and increase the circularity of the textile industry, understanding the concept of a circular economy is essential. The circular economy concept has grown over the last decades and is a method to cope with today’s production and consumption patterns (Ghisellini et al. 2016; Moreau et al. 2017). Its primary goal is to decouple economic growth from harmful environmental impact, and instead create a balance between our economy, environment and society through more efficient resource usage (Ghisellini et al. 2016; Moreau et al. 2017). There are several definitions of the concept (Rizos et al. 2017; McCarthy et al. 2018), but we are in this study using the European Union’s definition of a circular economy (found in Definitions), with full respect to the large amount of other definitions for the concept.

The social-and institutional dimensions (not to be confused with the social costs and benefits that are measured in the CBA), are, besides the resource efficiency, highlighted by e.g. Moreau et al. (2017), Ghisellini et al. (2016), and Geissdoerfer et al. (2017) as important for the concept of circular economy. The institutional dimensions are, according to Moreau et al. (2017), crucial in distributing the costs of social- and environmental externalities that arises with economic activities, and they hence play a large part in how well a transition like this would work in practice. These dimensions, further on included in the term “Social and Solidarity Economy” (SSE) as done by Moreau et al. (2017), underlines the need for reduced usage of hazardous materials in production, an increase of job opportunities, and increased equality in

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the transition to a circular economy (Ghisellini et al. 2016; Geissdoerfer et al. 2017; Moreau et al. 2017).

The major changes that need to be done to reach a circular economy are according to Moreau et al. (2017) how the institutions should clearly define who is responsible for the costs of different economic activities in society and further create an inclusive economy with less social inequalities. In the case of reducing the usage of materials, increasing recycling priorities and reuse activities, the SSE concept includes that collective decisions should be considered, where the economic aspects are not in focus, but rather the social aspects of what is best for society as a whole. Parting from economic profitability could also lead to increased employment from people with employment difficulties and decreased social inequalities (ibid.). The social dimension of circularity in the textile industry, is further highlighted in this study as it goes in line with EU’s action plan for a more sustainable garment value chain, where issues such as gender equality and wages in the industry are targeted (European Commission 2019a). Kettunen et al. (2019) have identified both reasons to- and challenges with- a transition to an EU circular economy. The first one is an expected change in the demand for primary resources, which is mainly imported from low-income countries to the EU. This trend might harm the countries who have earlier secured an income from exports of primary raw materials. The transition further requires a change in recycling infrastructure within the union itself, to be able to take advantage of the resources that are already imported. The second one is the interplay with trade restrictions on raw materials. Some low income-countries have started to restrict the exports of primary resources, mainly to secure the domestic supply induced by increasing production. This makes the demand for cost-efficient sorting more needed in the EU. The third is that policies set domestically have global implications, and that integration between circular policies might be needed in the global economy (ibid.). A main finding of McCarthy et al. (2018) is that if activities in upstream extractive sectors are decreased, sectors such as mining, agriculture, forestry and material transformation, can end up being worse off. However, transitioning to a circular economy can also expand other sectors such as manufacturing and waste management. Depending on how these sectors are distributed over the globe, we might therefore see different results from a transition to a more circular economy (ibid.). As EPR can work as a way towards a transition to a circular economy, all these aspects need to be considered when evaluating possible impacts from implementing an EPR-system.

2.3 Cost-Benefit Analysis

Coase (1960) argues for the need to consider all costs in operating social arrangements and to take these into consideration when choosing between them, including the costs involved with transitioning into a new system. The Cost-Benefit Methodology can therefore be motivated from the perspective of market failures, where the market prices do not reflect the true costs. The CBA includes externalities, collective goods and may account for unemployment and therefore considers more accurate costs and benefits of an investment than can be shown through market prices (Kriström & Bergman 2014). Boardman et al. (2018) argues that the motivation behind conducting a CBA-method is to find the social costs and the social benefits,

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in order to find the costs and benefits to society as a whole. It thus aims at identifying the total resource usage associated with the analyzed alternatives (Trafikverket 2018).

A CBA-method can be used as a policy assessment tool to quantify the consequences of a policy in monetary terms. The value of the consequences can be named as the net social

benefits, or the net social costs (Boardman et al. 2018). A CBA measures the efficiency of

policy implementations and originate from the socioeconomic efficiency aligned with the Kaldor-Hicks criterion. If a CBA results in net social benefits being larger than the net social costs, then a re-allocation of goods, or a policy implementation, are resulting in increased welfare for at least one person with a large enough amount to compensate anyone being worse-off, and is thus argued to be pareto improving (Boardman et al. 2018). There are however also uncertainties aligned with a CBA, since it is often used to predict the future. By making a sensitivity analysis and creating a plausible range of the different outcomes, one can find how the net benefits would differ (ibid).

As costs and benefits from implemented policies are not always quantifiable, especially in terms of environmental impacts, the monetary value has to be estimated using alternative methods. Two approaches are common to price non-market goods such as effects on environment, namely revealed preferences and stated preference methods. Stated preference methods reveal the non-market price for e.g. environmental quality through a hypothetical market where the respondent is offered binary choices of accepting or not accepting certain amounts they would be willing to pay (WTP) or willing to accept (WTA) as compensation for changes in environmental quality. Commonly used methods are the Contingent Valuation, and the Choice Experiments. The revealed preferences method instead analyses associated markets, and prices the environmental effects based on those (Atkinson & Mourato 2008). This method does not, however, capture the non-use values of environmental effects, since it is based on e.g. the correlation between the environmental quality of a resurrection area and the amount of money the visitors are willing to pay to enter it6_{. There might be people who are not willing to} visit the park, but who still values the environmental quality within it, and such methods could hence miss out on certain values placed on non-market goods. Stated preferences methods are more likely to capture such values but could instead over-value the amount they would be willing to pay for e.g. increased environmental quality, as the market is hypothetical (Brännlund & Kriström 2012). This study aims at using estimated WTP from earlier studies, to calculate the decreased environmental effects from more reuse and recycling.

Not to mention, is the job creation that is aligned with an EPR-scheme. This post can be viewed from a cost-perspective as well as a from a solidarity-perspective. In France, the implementation of an EPR-system is argued to be beneficial both from a social, economic-and ecologic point of view, where the social parts originate from the increased employment opportunities; especially from those who have suffered from difficulties on the job market (Eco TLC 2018a). However, from a CBA perspective, the increase in employment is not considered a benefit, but entails a cost for society since the employed is thought to give up other job options

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for the one evaluated, i.e. the opportunity cost of employment. Several member states in the EU do, as France, experience unemployment due to lack of opportunities rather than transitional unemployment which change the way they are calculated in a CBA.

Boardman et al. (2018) argue that if the unemployment rate is between 5-10%, the policy most likely targets a larger majority of the unemployed in surplus, rather than the ones in the transition phase, which reduces the costs for labour. If the unemployment rate is lower, the job-creation most likely re-allocates the workforce from current jobs, creating a cost for society as these can no longer contribute at their original employment. However, the opportunity cost of hiring an unemployed is still argued to be above zero in a CBA, as the unemployed might be engaged in other valuable activities, such as childcare and leisure which do as well create value. The benefits from being employed is described by Bartik (2012) as the increase in income from employment subtracted by the reservation wage of the unemployed. The reservation wage can be described as the lowest wage an unemployed is willing to accept to leave unemployment, and to instead take a job (Brown & Taylor 2013). This reservation wage includes the unemployment benefits, as well as the leisure and time spent on other valuable activities (Brown & Taylor 2013; Bartik 2012).

Bartik (2012) does however argue that the opportunity cost of hiring unemployed can be zero, since there are several stigmas aligned with being unemployed. Frey and Stutzer (2002) highlight further issues aligned with unemployment, and states that being unemployed affects the well-being of the affected individuals and may lead to mental issues such as anxiety and depression. We originate from the perspective that there is a social gain from larger employment rates in terms of the increased happiness being aligned with people being employed (Frey & Stutzer 2002; Bartik 2012). This approach is further used since the French implementation of EPR is motivated from an SSE perspective, where decreased unemployment is desirable (Eco TLC n.d.). The opportunity cost of unemployed labour is accounted as zero in our CBA, and the social gain is thus presented in the CBA as a decreased cost of hiring

unemployed. In the sensitivity analysis, the alternative approach to calculate the opportunity

cost of unemployed is used and the value for that is larger than zero.

The value to the economy of the set of activities that are given up is called the economic opportunity cost of labour (EOCL), and these include the non-market costs or benefits (Jenkins et al. 2011). In the case of a project hiring labour, the project must pay a wage equal or higher than the gross-of-tax supply price (Wgs), to attract labour force. The higher wage results in some formerly employed workers entering the new labour force (Hd), but new labour opportunities could also lead to one proportion of the labour (Hs) comes from the newly induced supply. For (Hd), there are no regional migrations or distortions in the labour market such as taxes and protected labour (unions) between the markets that the new hired labour shifts between. Hence, the economic opportunity cost is equivalent to the local market wage (W) or the supply price of labour (Ws) as expressed in equation 1:

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2.3.1 Cost-Benefit Method for EPR-systems

OECD (2005) propose a framework for assessing the costs and benefits of EPR specifically. This approach can help identify the characteristics of EPR-schemes and find which are the most cost-efficient; yielding the highest social benefits in relation to the social costs involved. However, Smith (2005) argues that there might be aspects of costs and benefits that cannot be precisely quantified, but that are still important for the result. These still needs to be included and discussed. Smith defines three groups of costs and benefits; (1) operational effects, (2) environmental effects, and (3) side effects.

Smith (2005) states that the collection and treatment of the end-of-life products can either be run by a PRO, by individual firms or by municipalities, which induces costs for the system. However, there may be counterpart savings if municipalities face reduced collection and disposal costs. This may result in that the net effect is zero, as the waste treatment is diverted from a municipal-run system to a system run by a PRO. If one system is more cost-effective than another, however, Smith argues that the unit cost may differ. In addition, if EPR is appropriately designed and is working as intended, it may reduce the total volume of wastes compared to a situation when EPR is not in system. As EPR-schemes usually require more recycling than in conventional municipal waste management, the volume of waste disposal in landfills and incinerators usually changes, creating environmental benefits of the operation (ibid.).

The side effects may be other costs and benefits of the implementation, indirect effects and externalities involved with waste streams Smith (2005). The first step is thus to identify all costs and benefits aligned with an implementation. The second step is the quantification of each of the elements identified, and the fourth, final step, is to assess the net social benefit of the EPR-programme compared to the specified alternative. As the side effects are difficult to quantify, they should mainly be identified and later on considered in relation to the net benefit of the EPR-scheme (ibid.).

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3. Empirics

To conduct our CBA information about the textile supply chain and the benefits that can follow from reuse-and recycling, a review investigating earlier implemented and evaluated EPR-policies, and our expert-interview with Birgitta Losman7_{and Yvonne Augustsson}8_{is used. All} three sources are considered necessary to get a full understanding of the different aspects of this type of implementation.

3.1 Textiles

This section covers the textile supply chain and which benefits that are aligned with an increased focus on reuse and recycling of the material. It further presents the waste management of textiles in the EU today, to create an understanding of the differences an EPR-scheme for textiles would induce.

3.1.1 The textile supply chain

The textile supply chain can be divided into both upstream-and downstream activities, where the upstream activities cover all parts included in production, i.e. from design to the point when the item is sold. Downstream activities are instead related to the activities involved with the customer, i.e. from sale to end-of-life management (Watson et al. 2014). The extended producer responsibility policy evaluated, focuses mainly on targeting the downstream effects, but has the potential to include upstream effects as well (Watson et al. 2014). EPR, apart from other systems that involves take-back schemes, tries to create a feedback mechanism between the upstream- and downstream activities (Tojo et al. 2012) and thus affects both the production part and end-of-life management.

The textile productions’ environmental impact and large usage of virgin material presented in the background, has to a large extent been impacted of the “fast fashion” phenomenon (Šajn 2019). This business model is based on low prices and an increased focus on different fashionable styles. The EU inhabitants are assumed by Manshoven et al. (2019) to consume 25.9 kg of textiles per person and year and since the 2000’s, clothing production has almost doubled while consumers keep the products for around half of the time (Remy 2016). The phenomenon has further led to poorer quality of garments (Manshoven et al. 2019; Šajn 2019). The fast-fashion phenomenon thus affects both the upstream activities, with lower salaries in the production, large usage of virgin material and high production rates, as well as the downstream activities where the textiles are more quickly discarded due to poor quality. Due to the low prices, the fast fashion phenomenon has also led to a reallocation of production to countries that pay the lowest wages, such as Asia and Eastern Europe where the workers receive salaries below living wage (Luginbühl and Musiolek 2014).

7_{Birgitta Losman, samordnare för hållbar utveckling, Högskolan i Borås, telephone interview on the 17th of}

April 2020.

EU Waste Framework Directive, What&apos;s Next? : A cost­benefit analysis of an extended producer responsibility for textiles in the European Union