Policy measures to increase the use of recycled materials in … · 2020. 2. 28. · Policy measures to enhance increased use of recycled materials in production III In order to explore

IIIEE Report 2020:XX

Policy measures to increase the use of recycled materials in production

Achievement and challenges of existing policy measures and anaylsis of selected potential policy measures

Naoko Tojo

A report written for the research program Mistra Closing the Loop Lund, Sweden, February 2020

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ISSN 1401-9191

Policy measures to enhance increased use of recycled materials in production

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Acknowledgements I would like to thank the Mistra Closing the Loop research programme and Evalena Blomqvist for providing me with the assignment of writing this policy report, which gave me invaluable opportunities to think through some of the issues I have been contemplating on over some years, as well as to obtain new knowledge on a number of recent development. A cordial gratitude is directed to Lena Smuk for your guidance in framing this study, as well as your support, curiosity, flexibility, trust and patience. Meeting with leaders of the six projects under the second phase of Mistra Closing the Loop research programme gave me many insights on the six projects, as well as some of the policy instruments I am assigned to look at. Thank you also for your feedback provided to ideas of the report presented in the project meetings.

My involvement in this task was facilitated by Martin Kurdve, the project leader of CIMMREC, one of the six research projects which I was also fortunately part of. Thank you, Martin, for bridging this opportunity and your flexibility. My gratitude is directed also to Sasha Shahbazi and Anna Runa Kristinsdottir in the CIMMREC project for their excellent coordination.

Many ideas presented in this study are inspired by numerous discussions I have had with my previous and current colleagues at the International Institute for Industrial Environmental Economics (IIIEE) at Lund University over many years, and especially, those with Thomas Lindhqvist, Chris van Rossem, Jessika Luth Richter, Tomohiro Tasaki, Wakana Takahashi and Panate Manomaivipool. Thank you very much, all, for enriching my thinking in so many ways.

Last but not least, thank you, Axel and Takeo, for all the support and for allowing me to sacrifice some of my family time.

Lund, February 2020,

Naoko Tojo

Naoko Tojo, IIIEE, Lund University

II

Executive Summary Closure of material loops has been an aspiration of policy makers around the world over the last several decades. Economic cricis in the late 2000s as well as demand on critical raw materials needed for the production of various “smart” products, among others, revitalised the political attention worldwide. In the EU, it is manifested in the Circular Economy Action Plan adopted in 2015. Among the current efforts of closing the material loop includes the research programme Mistra Closing the Loop. The programme focuses on the enhancement of the effective and sustainable recycling of industrial waste. During the second phase of the programme (2016-2020), six research projects covering different types of manufacturing industry (e.g. cars, bagtteries, construction, pulp and paper, textiles) and waste type (e.g. mining waste, residual waste from production process) seek to improve material efficiency in these various settings.

One of the common findings from the six research projects has been the challenges of bringing recycled materials back in the production process. This study, as part of the Mistra Closing the Loop research programme, is conducted with the overall aim of providing the current knowledge and international experiences concerning policy measures to enhance the use of recycled materials in production. Taking into consideration the existence of a number of policy measures seeking to close material loops, as well as the assignment from the Mistra Closing the Loop research programme, the study takes the following two steps:

1. Analysis of the prevailing existing policy approaches used in closing the material loop,

focusing on their achievements, outcomes, challenges in increasing the use of recycled

materials in production, and potential causes of challenges.

2. Assessment of selected (potential) policy measures which are deemed particularly useful in

increasing the use of recycled materials in production, focussing in particular on its

practical implementation potential.

By looking into the achivements and challenges as well as the potential causes of challenges, the first step seeks to identify gaps in the present policy landscape. As this study is run in parallel with another study on drivers and barriers for the increased use of recycled materials in production, the focus of the discussion regarding the causes is concentrated on issues related to policy design. Findings from the first step feeds into the second step where existing knowledge and experiences on (potential) policy measures tailoured for increased use of recycled materials in production are examined. The focus of the assessment is prospect for introducing the respective policy measures in practice, taking into consideration, for instance, the potential and limitation pertaining to the characteristics of the instrument (issues addressed, level of governance, type of instruments), as well as the gaps identified in the first step.

Among the various policy measures in the field, given that all the six research projects were based in Sweden, the study focuses primarily on interventions in the EU and its Member States. Meanwhile, experiences from jurisdictions outside of the EU are also reffered to when relevant. The analysis of the study is centred around legal text analysis and (re-)construction of intervention theories of the respective policy measures, focusing primarily on how the policy measues in question is supposed to work so that it brings about the increased use of recycled materials in production. It is primarily a desk-top research concentrating on the review of exsiting studies, laws and empirical data in the relevant area. In addition, the author of the study participated in two meetings organised for the coordinators of the six research projects under the Mistra Closing the Loop.


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In order to explore potential policy measures, the study starts by analysing the prevailing existing policy measures pertaining to closure of material loops, focussing specifically on their effects on the increased use of recycled materials in production, as well as the underlying logics (intervention theories) that should lead the introduction of an intervention to the intended outcome. Table below summarises the 9 policy measures examined, as sell as their main achievements and otucomes so far.

Policy measures Achievements and outcomes

Collection/sorting requirements Tangible achievement in terms of sorting selected waste streams from

the rest of waste Collection targets

Recycling targets Tangible achievement in terms of bringing considerable portions of

sorted waste streams to recycling plants (packaging and EEE), and for

processing the incoming materials (batteries and cars)

Separation of components containing hardous

substances

Relevant clause available in national and EU laws

Landfill bans Tangible effects in promoting recycling/energy recovery experienced

for tyres

End-of-waste and by-products critera Very limited implementation except for Italy

Substance restriction Tangible effects upstream

Product design requirement related to closure of

material loops

General clauses available in national and EU laws, with some concrete

policy actions for implementation

EPR programme Potential design effects and IPR recognised and articulated in laws

The study reveals that, while existing policies succeded in several aspects essential for the closure of material loops – e.g. sorting of specific waste streams from the rest of the waste stream, restricting the use of specified hazardous substances in products – they have failed in several other aspects also essential for the closure of material loops. Most notably, the assumption that source separation of specific waste streams would lead to the development of cleaner and recycled material stream does not hold for the considerable parts of sorted waste streams (theory failure). The study also highlights that design requirements related to closure of material loop is currently limited, and the uptake of so-called end-of-waste criteria has been very low in most parts of Europe. It further points to the struggle of EPR-based approach in implementing individual producer responsibility, which in turn hinders the provision of incentives for producers to make upstream changes. The following figure visualise where in the policy interventions these failures occurred.


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A number of causes of challenges regarding why policy interventions have not been working well relate to target setting and their measurement (e.g. weight-based targets, lack of quality requirement for sorting and recycling, use of the input materials to recycling facilities as numerator when calculating recycling targets, instead of output materials). There are also issues related to the nature of material efficiency, which is difficult to compare quantitatively, as well as the inherent innovative nature of design process.

The study subsequently examines six selected (potential) policy measures, from the perspective of their potential effectiveness in increasing the use of recycled materials in production. Five of them was prioritised by the reseach projects under the Phase II of the Mistra Closing the Loop research programme, and the sixth one is derived considering the challenges of the existing policy measures. In addition to the content of the policy mesures, their intervention logic and how they might rectify the challenges facing existing policy measures, experiences of introducing similar policy measures have been reviewed.

The following table summarises the overall findings regarding the six policy measures focusing primarily on the prospect of introducing them in reality, at which level of governance.

Policy measures Prospects of introduction

Quality standards for recycled

materials

There seems to be a wider opening for EU Member States to introduce end-of-waste

criteria on their own on the streams not covered by the EU, including case-by-case

introduction

Requirements for functional

recycling

Rather slim possibility to introduce a direct policy measure. Perhaps utilise end-of-waste

criteria, or expect effects of EPR programmes

Minimum recycled material

content standards

Mandatory – some examples from US States, and EU introduced one on plastics in 2019,

which would pave the way for introduction for other material streams. Best to introduce at

the EU level.

Voluntary – various opportunities for individual Member States to continue integrating as a

criteria in, e.g. green public procurement, Type I eco-labels, guidelines

Clearer and expanded

requirements for design for

recycling

Mandatory – the first implementing regulation on ErP including material efficiency related

criteria came into force, which would enhance possibilities for introducing other material

efficiency related criteria for other products. Best to introduce at the EU level.

Voluntary – various opportunities for individual Member States to continue integrating as a

criteria in, e.g. green public procurement, Type I eco-labels, guidelines

Material recycling combined

with refundable security fee or

penalty

At the moment hypothetical, but examples of policy measures based on the similar

intervention logics exist. If there is political wills, it is possible to introduce at a Member

State level, but requires some more thinking for concrete design details.

Alternative collection and

recycling targets and their

measurement

For changes related to what to measure, best to introduce at the Member States level.

Overall, there is a good prospect for introducing four out of six of the policy measures examined. Depending on the level of coerciveness as well as the current legislative setting, some of these measures could be easily introduced at the Member States level, while when it comes to mandatory requirements related to properties of a product, it is best to introduce at the EU level.

Based on the analysis of both existing and potential policy measures, the study provides policy recommendation tailoured specifically for the Swedish policy makers, as listed below.


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• Actively explore the introdution of end-of-waste criteria tailoured for the Swedish context

and for the Swedish industry.

• Monitor output materials from recycling plants and their subsequent paths.

• Push the inclusion of material efficiency related criteria in the EU ErP Directive, while

continue enhancing these criteria within domestic voluntary policy measures.

• Continue supporting the further development of restriction of hazardous substances.

• Explore concrete potentials for implementing individual producer responsibility.


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Table of Contents

ACKNOWLEDGEMENTS ....................................................................................................................... I

EXECUTIVE SUMMARY ........................................................................................................................ II

LIST OF FIGURES ................................................................................................................................. VI

LIST OF TABLES ................................................................................................................................. VII

ABBREVIATIONS ................................................................................................................................ VII

1 INTRODUCTION ............................................................................................................................ 1

2 TYPOLOGIES OF POLICY MEASURES, INTERVENTION THEORIES ................................. 4

2.1 TYPOLOGIES OF POLICY MEASURES ......................................................................................................................... 4 2.2 INTERVENTION THEORIES ........................................................................................................................................ 5

3 ACHIEVEMENTS AND CHALLENGES OF PREVAILING POLICY APPROACHES IN

INCREASING THE USE OF RECYCLED MATERIALS IN PRODUCTION............................. 8

3.1 CHARACTERISTICS AND LOGICS OF PREVAILING POLICY APPROACHES ............................................................ 8 3.1.1 Collection/sorting requirements ............................................................................................................................ 13 3.1.2 Collection targets.................................................................................................................................................. 14 3.1.3 Recycling targets ................................................................................................................................................... 15 3.1.4 Separation of components containing hazardous substance .................................................................................... 16 3.1.5 Landfill bans ...................................................................................................................................................... 16 3.1.6 End-of-waste and by-products criteria .................................................................................................................. 17 3.1.7 Substance restriction ............................................................................................................................................ 18 3.1.8 Product design requirements related to closure of material loops ............................................................................ 19 3.1.9 Extended Producer Responsibiltiy (EPR) programme ......................................................................................... 19

3.2 INTERVENTION THEORIES OF PREVAILING POLICY APPROACHES.................................................................. 21 3.3 ACHIEVEMENTS AND OUTCOMES ......................................................................................................................... 21 3.4 CHALLENGES ............................................................................................................................................................ 26 3.5 CAUSES OF CHALLENGES PERTAINING TO POLICY DESIGN ............................................................................. 28

4 ANALYSIS OF SELECTED POTENTIAL POLICY MEASURES ............................................... 32

4.1 QUALITY STANDARDS FOR RECYCLED MATERIALS ............................................................................................ 32 4.2 REQUIREMENTS FOR FUNCTIONAL RECYCLING ................................................................................................. 33 4.3 MINIMUM RECYCLED MATERIAL CONTENT STANDARDS .................................................................................. 34 4.4 CLEARER AND EXPANDED REQUIREMENTS FOR DESIGN FOR RECYCLING................................................... 36 4.5 MATERIAL RECYCLING COMBINED WITH REFUNDABLE SECURITY FEE OR PENALTY ................................. 37 4.6 ALTERNATIVE COLLECTION AND RECYCLING TARGETS AND THEIR MEASUREMENT ................................ 38 4.7 SUMMARY OF SELECTED POTENTIAL POLICY MEASURES .................................................................................. 40

5 CONCLUSIONS AND RECOMMENDATIONS ......................................................................... 42

5.1 RECOMMENDATIONS FOR POLICY MAKERS ......................................................................................................... 42 5.2 SUGGESTIONS FOR FURTHER RESEARCH .............................................................................................................. 43

BIBLIOGRAPHY .................................................................................................................................... 45

SOURCES OTHER THAN LAWS ............................................................................................................................................ 45 EU, NATIONAL AND LOCAL LAWS .................................................................................................................................... 49

List of Figures Figure 1: Typologies of various environmental policy measures discussed in this study ............. 5


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Figure 2: Typologies of outcomes of a policy intervention (adopted from Mickwitz (2003) and Vedung (1997)) ........................................................................................................... 6

Figure 3: Outcomes that could be covered in (an) intervention theory(ies) of a policy measure, and fucos of this study ................................................................................................. 7

Figure 4: Intervention theories of existing policy measures pertaining to increased use of recycled materials in production ................................................................................................ 21

Figure 5: Areas of existing interventions that have not worked well, pertaining to increased use of recycled materials in production ................................................................... 28

List of Tables Table 1: A summary of existing EU laws stipulating requirements relevant for closure of

material loops for selected product/waste streams ................................................................ 11

Table 2: rate of packaging waste entering the recycling facilities, EU-28 and Sweden, 2012-2016, the recycling targets stipulated in Directive 94/62/EC for 2025 & 2030 and in Swedish packaging ordinance (SFS 2014:1073) before and after 1 January 2020 (unit: percentage) ................................................................................................................ 22

Table 3: Collection rate of batteries and accumulators, EU-27 and Sweden, 2011-2018, the collection targets stipulated in Direcive 2006/66/EC for 2016 and in Swedish ordinace for batteries (SFS 2008:834) (unit: percentage) ....................................................... 23

Table 4: Reuse and recycling rate of end-of life vehicles, EU-28 and Sweden, 2008-2017, the reuse and recycling targets stipulated in Directive 2000/53/EC for 2016and theSwedish ordinance for cars (SFS2007:185) (unit: percentage) ......................................... 23

Table 5: Per capita figures of put-on-the-market (POM), collection and reuse & recycling, EU -28 and Sweden, 2010-2017 (unit: kg) .............................................................. 24

Table 6: Summary of achievements and outcomes of existing policy measures pertaining to increasing the use of recycled materials in production ...................................................... 25

Table 7: Potential policy measures to enhance the use of recycled materials, and prospect of their introduction in practice ................................................................................ 41

Abbreviations EEE: Electrical and Electronic Equipment

ELV: End-of-life Vehicles

EPR: Extended Producer Responsibility

IPR: Individual Producer Responsibility

POM: Put on the Market

RoHS: Restriction of the use of Hazardous Substances

WEEE: Wasete Electrical and Electronic Equipment


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1 Introduction Closure of material loops has been an aspiration of policy makers around the world over the last several decades. Some examples of concrete policy measures developed earlier include the adoption of the Ecocycle Bill by the Swedish Parliament in 1993,1 the Circular Economy and Waste Law of 1994 in Germany2 and the Basic Law for the Promotion of Circular Society of 2000 in Japan.3 Although focus in environmental policy discourse was dominated by climate change issues during the 2000s, global economic crisis in the late 2000s among others revived some attentions back to material efficiency issues. The full re-recognisation of its significance is manifested in, among others, the development of the EU Circular Economy package in the 2010s (See, for example, European Commission, 2015). Among the concrete objectives to be achievered through the generational goal in Sweden, which is to guide environmental actions at various levels of society, include “Materials cycles are resource-efficient and as far as possible free from dangerous substances.” (Naturvårdsverket, 2019).

Among the current efforts of closing the material loop include the research programme Mistra Closing the Loop. The programme focuses on the enhancement of effective and sustainable recycling of industrial waste. During the second phase of the programme (2016-2020), six research projects covering different types of manufacturing industry (e.g. cars, batteries, construction, pulp and paper, textiles) and waste type (e.g. mining waste, residual waste from production process) seek to improve material efficiency in these various settings.4

One of the common findings from the six research projects has been the challenges of bringing recycled materials back in the production process. Researchers in the six research projects were asked to consider policy measures deemed useful in rectifying the problems. This resulted in a total of twelve policy measures, as listed below.

1. Quality standards for recycled materials

2. Requirements for functional recycling

3. Miniumum recycled material content standards

4. Clearer and expanded requirements for design for recycling

5. Material recycling combined with refundable security fee or penalty

6. Improved labelling of plastic parts

7. Requirements for sorting and recycling of waste

8. Trade of recycling certificates

9. Subsidised transport of secondary raw materials

10. Taxes and subsidies that influence consumption patterns of products adapted for

recycling

11. Waste incineration tax (or ban of incinerating recyclable materials).

12. Tax on unsorted waste.

Among the twelve policy measures, researchers in the six projects prioritised the first five as most important.

1 Riktlinjer för en kretsloppsanpassad samhällsutveckling, or simply, Kretsloppspropositionen 1992/93:180.

2 Gesetz zur Förderung der Kreislaufwirtschaft und Sicherung der umweltverträglichen Beseitigung von Abfällen

(Kreislaufwirtschafts- und Abfallgesetz – KrW-/AbfG)

3 Junkan gata Shakai Keisei Suishin Kihonhou.

4 For more information about the Mistra Closing the Loop Programme as well as the six research projects, see

https://closingtheloop.se/.

https://closingtheloop.se/


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Against this background, this study is conducted with the overall aim of providing the current knowledge and international experiences concerning policy measures to enhance the use of recycled materials in production. Taking into consideration the existence of a number of policy measures seeking to close material loops, as well as the assignment from the Mistra Closing the Loop research programme to pay special attention to the five short-listed policy measures, the study takes the following two steps:

3. Analysis of the prevailing existing policy approaches used in closing the material loop,

focusing on their achievements, outcomes, challenges in increasing the use of recycled

materials in production, and potential causes of challenges.

4. Assessment of selected (potential) policy measures which are deemed particularly useful in

increasing the use of recycled materials in production, focussing in particular on its

practical implementation potential.

By looking into the achivements and challenges as well as the potential causes of challenges, the first step seeks to identify gaps in the present policy landscape. As this study is run in parallel with another study on drivers and barriers for the increased use of recycled materials in production, the focus of the discussion regarding the causes is concentrated on issues related to policy design, and not on, e.g. market conditions, business models and the like. Findings from the first step feeds into the second step where existing knowledge and experiences on (potential) policy measures tailoured for increased use of recycled materials in production is examined. The focus of the assessment is prospect for introducing the respective policy measures in practice, taking into consideration, for instance, the potential and limitation pertaining to the characteristics of the instrument (issues addressed, level of governance, type of instruments, See Section 2.1), as well as the gaps identified in the first step.

Among the various policy measures in the field, given that all the six research projects were based in Sweden, the study focuses primarily on interventions in the EU and its Member States. Meanwhile, considering the potential of policy transfer and extrapolation (See, for example, Bardach, 2004, Dolowitz and Marsh, 2000), experiences from jurisdictions outside of the EU are also reffered to when relevant. In addition to the five policy measures short-listed by the researchers in the six projects under the Mistra Closing the Loop, other measures identified in literature and discussion with practitioners will be also discussed.

The analysis of the study is centred around legal text analysis and (re-)construction of intervention theories of the respective policy measures. A policy measure tends to have more than one intended outcomes, and (re-)construction of intervention theory(ies) often helps in identifying expected unintended outcomes (see Section 2.2). This study focuses primarily on how the policy measues in question is supposed to work so that it brings about the increased use of recycled materials in production. Discussions on other intended outcomes (e.g. avoidance of littering, improved product design) as well as expected unintended outcomes will be made only when there deems to be direct relvenace to the increased use of recycled materials in production.

It is primarily a desk-top research concentrating on the review of exsiting studies, laws and empirical data in the relevant area. In addition, the author of the study participated in two meetings organised for the coordinators of the six research projects under the Mistra Closing the Loop, which enhanced the author’s understanding on the respective projects as well as the content of the five short-listed policy measures.


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There have been many instances where what is stipulated in a law is different from what is happening in reality due to various reasons such as agreements made on the ground between different actors, lack of enforcement of laws and the like, and that is also true for at least some of the policy measures discussed in this report (see, for instance, Tojo, 2004; Watson et al., 2017). When discussing the content of the law, the discussion in this report is primarily based on what is written in the law, instead of what is happening in reality, unless otherwise mentioned.

The structure of this report follows the research steps mentioned above. Following this introductory chapter, Chapter 2 provides a concise description of approaches used in this study when categorising and analysing policy measures. Chapter 3 discusses the characteristics, achievement, outcomes and challenges of prevailing policy approaches relevant to the icreased use of recycled materials in production and contemplate on causes of challenges pertaining to policy design. Chapter 4 subsequently investigates potential policy measures to rectify the current policy situation. Chapter 5 concludes the paper, together with a concise set of recommendations for policy makers in Sweden and suggestions for future research.


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2 Typologies of policy measures, intervention theories Before discussing existing and potential policy measures this chapter provides a concise description on typologies of policy measures relevant for this study (Section 2.1), as well as introduces the concept of (an) intervention theory(ies) (Section 2.2), a crucial element of a theory-based policy evaluation. The respective sections also highlight the elements that are of particular relevance to this study.

2.1 Typologies of policy measures In terms of nature, policy measures discussed in this document can be categorised into administrative, economic and informative instruments. They could be also introduced together (often referred to as policy mix), either within one government intervention or in two or more separate intereventions introduced in parallel (Tojo & Lindhqvist, 2010).

Administrative instruments are policy measures that concern fulfilment of certain tasks. Examples include achievement of a specific target (e.g. recycling targets for spent batteries), elimination of the use of certain substances (e.g no mercury in electrical and electronic equipment) and prohibition of certain action (e.g. landfilling of construction waste). When required by law, the targeted entities (e.g. industry) have no choice but to make sure to act, in accordance with what is stipulated in the law (Vedung, 1998; van der Doelen, 1998).

Economic instruments (also referred to as market-based instruments) generally provide monetary incentives – such as subsidies and refunds – when the targeted entities fulfill tasks that the instrument wishes to promote (e.g. bringing empty cans to the collection points), or disincentives such as tax, when targeted entities do not act what is desired by the instruments (e.g. landfilling of construction waste) (Vedung, 1998; van der Doelen, 1998). The crucial difference between administrative instruments and economic instruments is that in the former, the addressee has no choice but to fulfil the task if mandated by law, while in the latter, the addressee has more freedom – they may be rewarded for taking actions prompoted by the intervention, or they may have to pay for the actions discouraged by the intervention.

Informative instruments, or information, concern the collection and provision of various types of information (e.g. labelling schemes, guidance document), and are used with the assumption that targeted entities would behave differently when they have better informed. Also referred to as “moral suasion”, it seeks to influence people “through the transfer of knowledge, the communication of reasoned argument, and persuasion” (Vedung, 1998, 33).

These three policy instruments can be categorised between mandatory and voluntary from the perspective of level of coerciveness. The targeted entities of a mandatory instrument are required to fulfil the tasks laid down in law, while the private actors can set up specific goals themselves and strive to achieve them in voluntary initiatives. Between these two spectrums exists, for instance, negotiated agreements. In a negotiated agreement, the government and private actors typically agrees on not introducing a law on condition that the private actors achieve a certain goal. The level of coerciveness might influence the level of governance through which the instrument can be introduced.

A policy measure can be categoised between local, national, regional or international, based on the level of governance as well. The level of governance can be a critical determinant of what type of measures could be introduced. For instance, there are limitations to introduce mandatory legislation that demands that a selected type of product has specific properties (e.g. it does not contain specified substances, it contains certain amount of recycled materials) at a national level in the EU, due to the so-called internal market requirement.


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Policy instruments can be also categorized based on their targeted entities (e.g. states under international law, organisations of different types, individuals), as well as their main domain.

Figure 1 summarises categorisation of policy measures used in this study.

Leve

l of

gove

rnan

ce

Types of policy measures

Voluntary - Mandatory

Administrative

Economic/ market-based

Informative

Target entities (e.g.

con

sum

ers,

pro

du

cers, state

s…)

Local

National

Regional

International

Environmental domain (e.g. energy, material efficiency...)

Figure 1: Typologies of various environmental policy measures discussed in this study

2.2 Intervention theories Theory-based evaluation5 as a way of systematically analyse a policy intervention has been advocated by a number of policy evaluation researchers (Coryn et al., 2011; Hansen & Vedung, 2010; Weiss, 1997, Chen, 1990). The term “theory” in this context has a broader meaning than what is normally used in social science. It is “grandiose” (Rossi et al., 1999, p198) if it is considered “to mean a set of highly general, locially interrelated propositions that claim to explain the phenomena of interest” (Weiss, 1999, p502). Intervention theory6 is rather “the set of beliefs and assumptions underlying an intervention (that) can be expressed in terms of a phased sequence of casues and effects” (Weiss, 1999, p501), or “A model, theory or philosophy about how the program works, and indicate the causal relationships supposedly operating in the program” (Fitz-Gibbon & Morris, 1996). In essence, it is “a chain of assumptions through which an intervention, step by step, is supposed to achieve what was originally designed to achieve” (Tojo, 2004). Important to note is that theory of intervention tends to be plural, as there are often more than one paths of logic assumed in achiving the desired outcomes (Weiss, 1999). The envisioned pathways might be also perceived differently from one actor to the other (Hansen and Vedung 2010).

An important function of (re-)constructing intervention theories is to make implicit assumptions explicit. Despite that an intervention is inevitably built on some assumptions, these assumptions may not be explicitly spelled out. Making these assumptions explicit provides an opportunity to examine the validity of such assumptions (Weiss, 1997). Furthermore, the re-constructed theories, with the clarification of how an intervention is supposed to be implemented, can serve as a blueprint against which the implementation of the actual intervention can be compared. In other words, in addition to understanding whether or not an

5 Other terminologies, such as theory-driven evaluation, theory-guided evaluation, theory-led evaluations have also been used

(see, for instance, Coryn et al., 2011), though meaning essentially the same thing – a policy evaluation based on a “theory”.

6 Other terminology such as “program theory” and “logic models” are also used (see, for instance Weiss, 1998, Rossi et al.,

1999, Hansen & Vedung 2010), but essentially means the same as intervention theory.


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intervention achieved its overall purpose, the construction of intervention theories allows an evaluator to get a better grip on how the intervention has worked (Coryn et al., 2011).

The re-construction of theories also helps identifying events and changes deemed essential for the attainment of intended outcomes, which in turn can serve as candidates of variables to be examined when conducting the evaluation (Coryn et al., 2011; Rossi et al., 1999; Vedung, 1997; Fitz-Gibbon & Morris, 1996). This is especially useful when evaluating an intervention whose ultimate intended outcome is difficult to evaluate due to, for instance, spatial and temporal reason – a characteristic often seen in environmental interventions.

If an intervention is based on an explicit theory, model or philosophy, the intervention theories could be reconstructed in line with such theory, model or philosophy. Other sources of information that could be used for reconstruction of intervention theories include documents pertaining to intervention (e.g. preparatory documents, intervention itself), the intuitions and experiences of those who designed the intervention, prior research, logical reasoning and the like (Fitz-Gibbon & Morris, 1996; Chen, 1997; Weis, 1997).

A policy intervention can produce multiple outcomes, both within and outside of the area the intervention seeks to address. The outcomes can be positive or negative. There are also potential unanticipated outcomes that the intervention brings about, within or outside of the targeted policy area and could be positive or negative (Mickwitz, 2003; Vedung, 1997). Figure 2 summarises these different categories of (potential) policy outcomes.

Figure 2: Typologies of outcomes of a policy intervention (adopted from Mickwitz (2003) and Vedung (1997))

A role of outling (an) intervention theory(ies) is also to identify various anticipated outcomes (highlighted in dotted line in Figure 3). For this study, the main focus is to examine the effects on the anticipated, positive outcome of increased use of recycled materials in production (bottom left corner of Figure 3).

The possibility of distinguishing between the examination of the validity of causal mechanisms or the assessment of the correctness of implementation also assists an evaluator to distinguish whether there is something wrong with the theory based on which an intervention was developed (theory failure), or whether something went wrong in the implementation process (implementation failure). This is especially useful in cases when desired outcomes did not occur.

In this paper, intervention theories are considered to visualise which assumptions behind the prevailing policy measures pertaining to closure of material loops was correct or faulty, or whether their concrete implementations have or have not worked.


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Figure 3: Outcomes that could be covered in (an) intervention theory(ies) of a policy measure, and fucos of this study


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3 Achievements and challenges of prevailing policy approaches in increasing the use of recycled materials in production

This chapter provides an overview of prevailing policy approaches relevant to the closure of material loops, including the icreased use of recycled materials in production. After a brief description of the general characteristics of some of the most typical policy approaches used to date, intervention theories of the relevant policy approaches will be laid out. The author then considers what these measures managed to achieve and what have not worked out well. It further discusses potential causes of challenges facing these exsiting policy measures, focussing on issues related to policy design.

3.1 Characteristics and logics of prevailing policy approaches Since the 1990s, a number of policy measures relevant to the closure of material loops have been implemented in the EU, its Member States and many other parts of the world. In the EU, for instance, in addition to the framework legislation on waste,7 a number of laws on specific product/waste streams such as packaging,8 cars,9 electrical and electronic equipment (EEE)10 and batteries11 seek to address closure of material loop both from end-of-life phase and production/design phase of these streams (see, for instance, Tojo & Thidell, 2018; Milios, 2018; Tojo & Lindhqvist, 2010; Krämer, 2007).

As detailed in Footnotes 7 to 11, all of the EU laws went through revisions, often including the amendments of specific measures to enhance collection and recycling of targeted product/waste streams. There are also laws on waste installation and operations that have implications to closure of material loops, such as the directive on ladfills which among others restricts the landfilling of certain waste streams.12,13 Concerning the easier utilisation of waste

7 The current Directive 2008/98/EC on waste replaced the original waste framework directive from 1975 (75/442/EEC).

The current Directive went through a number of revisions, and the latest rather substantial revision was made via Directive

(EU) 2018/851, reflecting the 2015 EU Circular Economy Action Plan (COM(2015) 614 final).

8 Directive 94/62/EC, with several revisions regarding, for instance, the recycling targets (Directive 2004/12/EC) and

inclusion of provisions addressing the consumption of lightweight plastic carrier bags (Directive (EU) 2015/720). The latest revision by Directive (EU) 2018/852 includes, among others, provisions on reuse as well as enhancement of recycling

targets.

9 Directive 2000/53/EC, which went through many revisions and was last amended by Directive (EU) 2018/849.

10 A directive on waste electrical and electronic equipment originally came into force in 2002 (Directive 2002/96/EC). Due

to substantial changes required, the Directive was replaced with the current Directive 2012/19/EU, as furthere amended by

Directive (EU) 2018/849.

11 The original directive regarding batteries (91/157/EEC) concens mostly on the restriction of hazardous substances within

batteries, with limited mentioning of collection and recycling of spent batteries. The newer directive (2006/66/EC) made further restricions regarding the use of hazardous substances in batteries, while introducing clear mandates regarding

collection and recycling of spent batteries.

12 Council Directive 1999/31/EC, which was last revised in 2018 by Directive (EU) 2018/850. The latest revision added

further items to be banned from landfilling.

13 There are other EU laws, such as on extractiveate industries (Directive 2006/21/EC), which is also relevant for closure of

material loop, but it is not waste directly generated from manufacturing industires and their products, thus outside of the scope of this study. Another law governing the procedure for waste shipment within the EU as well as between an EU Member State and countries outside (so-called Waste Shipment Regulation, (EC) No 1013/2006), could be also relevant as it affects the administrative burden given to industry to ship recyclable materials. In the context of this research, however, the author focuses on the development of end-of-waste criteria under the Waste Framework Directive (2008/98/EC) (see Section 3.1.6), as the intention of introducing the end-of-waste criteria is precisely to reduce unnecessary administrative burden from the provider of recycled materials by determining when waste is no longer considered as waste.


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generated from industrial operations, the Waste Framework Directive (2008/98/EU) also includes provision on end-of-waste (Article 6) and by-products (Article 5).

In addition, many of the product/waste stream specific laws,14 as well as the law governing safe management and use of chemicals – so-called REACH Regulation (EC) No 1907/2006 – restricts the use of certain hazaroud substances. Furthermore, there are provisions mandating or encouraging the closure of material loops by changing the design of products.

Many of these policy instruments find in these laws constitute an extended producer responsibility (EPR) programme, whose overall objective is to facilitate the closure of material loops by connecting the upstream and the downstream of products.

Table 1 provides a summary of prevailing policy measures relevant to the closure of material loops that are found in the exsiting EU laws, focusing on industry relevant to six projects under the Phase II of Mistra Closing the Loop research programme. The policy measures are grouped between downstream and upstream, based on the phase of product’s life upon which these measures are applied. Given the focus of this study of enhancing the recycled materials in production, reuse targets and preparation for reuse targets as well as other measures pertaining to the higher level of waste hierarchy, despite its strong relevance in closing material loops, are not included.

In the following subsections, the author highlights the characteristics and development regarding the respective policy measures briefly introduced above and listed in Table 1, including their intervention theories/underlying logics. Instead of providng a full account of the respective measures, the description focuses on dimensions relevant to the increased use of recycled materials in production. In addition to the situation in Europe, Some of the examples outside of Europe are also described when relevant.

As found in Footnotes 7 to 12 and 14, all the EU Directives referred to in the document went through some amendments. The article numbers referred to in the subsequenst sections are what is found in the latest amendments, unless otherwise mentioned.

14 Directives on end-of-life vehicles (2000/53/EC) and batteries (2006/66/EC) includes the provisions on the restriction of hazardous substances within them. Regarding EEE, a separate directive restricting the use of hazardous substances (RoHS) was developed in parallel to the one governing waste EEE (WEEE). Similarly to the WEEE Driective (see Footnote 10), the original RoHS Directive (2002/95/EC) was replaced by the revised Directive (2011/65/EU).


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Table 1: A summary of existing EU laws stipulating requirements relevant for closure of material loops for selected product/waste streams

Product/ waste

streams

Downsream policy instruments Upstream policy instruments Law governing both

upstream & downstream

Collection/ sorting

requirments

Collection targets

Recycling targets

Separation of components

containing hazardous substance

Landfill bans

End-of-waste and bi-product

criteria

Substance restrictions*

Product design requirement

related to closure of material loops

Extended Producer Responsibility

Packaging 94/62/EC,Art.7 94/62/EC,Art.6 1999/31/EC, Art. 5(3)(f)

2008/98/EC, Art. 5,6 Recital (22)

94/62/EC (Art. 9, 11, Annex II(1)),

94/62/EC, Art. 9, Annex II

Not at the EU level, but practiced by many MS

Electrical & electronic equipment

2012/19/EU, Art. 5

2012/19/EU, Art. 7

2012/19/EU, Art. 11, Annex V

2012/19/EC, Art. 8(2), Annex VIII

1999/31/EC, Art. 5(3)(f)

2008/98/EC, Art. 5, 6,Recital (22)

2011/65/EU, Art.6, Annex II, III

2009/125/EC, 2012/19/EC, Art. 4

2012/19/EU, Art 7, .8(3),11,12,13,15 Recital (6), (12), (22)

Cars 2000/53/EC, Art..5

2000/53/EC, Art.7

2000/53/EC, Art.6(3), Annex I

1999/31/EC, Art. 5(3)(d)

2008/98/EC, Art. 5,6 Recital (22)

2000/53/EC, Art.4, Annex II

2000/53/EC, Art.4 2000/53/EC, Art.5, 7, 9(2), Recital (7),(22)

Batteries 2006/66/EC, Art. 8

2006/66/EC, Art.11

2006/66/EC, Art. 12(4), Annex III Part B

2006/66/EC, Art. 12 2006/66/EC,Art.4 2006/66/EC, Art.11 2006/66/EC, Art.8,12, 16, 17, Annex IV, Recital (19)(28)

Construction & demolition waste

2008/98/EC, Art. 11(1)

2008/98/EC, Art 11(2)(b)

1999/31/EC, Art. 5(3)(f)

2008/98/EC, Art. 5,6 Recital (22)

Textiles 2008/98/EC, Art. 11(1)

1999/31/EC, Art. 5(3)(f)

2008/98/EC, Art. 5,6 Recital (22)

Sources: developed by the Author based on Directive 94/62/EC, Directive 1999/31/EC, Directive 2000/53/EC, Directive 2006/66/EC, Directive 2008/98/EC, Directive

2009/125/EC, Directive 2011/65/EC, Directive 2012/19/EC. Article numbers are based on the most updated version of the laws incorporating all the amendments as of February

2020.

* Specific substances restricted under the Regulation REACH (EC) 1907/2006 (Art. 67, Annex XVII) relevant to the respective waste streams can be found at, for instance,

https://echa.europa.eu/support/qas-support/browse/-/qa/70Qx/view/scope/REACH/Restrictions?

https://echa.europa.eu/support/qas-support/browse/-/qa/70Qx/view/scope/REACH/Restrictions?


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3.1.1 Collection/sorting requirements

Sorting of specific waste stream from the remaining waste stream is considered prerequisite for closure of material loops. In addition to other intended outcomes such as reduced littering, there is an assumption that sorting would provide a more homogeneous and cleaner stream of materials, which would – after some further steps in between – eventually lead to the use of recycled materials in production and elsewhere.

As found in Table 1, there exists legal mandate to sort all the six waste sterams focused in this study from the rest of the waste stream (when including the upcoming requirement for construction and demolition waste (C&D waste) and waste textitles). The EU laws mandate the EU Member States to ensure the establishment of a source separation system without, except for the cars and to an extent EEE,15 specifying actors who should carry out such responsibilities. This means the concrete allocation of responsibility and practical organizations are left in the hands of Member States and actors carrying out the responsibility allocated to them, leading to various solutions.

Regarding the actors involved in organizing collection, the actors most typically become responsible are municipalities, distributors and municipalities. A study on the transposition of the original WEEE Directive (2002/96/EC) found that Member States allocate responsibility to one or the combination of the three actors in varying ways (Sander et al., 2007). Similar situation is found in the case of packaging collection – some countries assign responsibility for collection to municipalities (e.g. Denmark, Belgium), while others to producers, packers and distributors (e.g. Germany, Austria, Sweden) (Spasova, 2014, Tojo et al, 2003).

Systems and infrastructures used for collection also vary between countries and between waste streams, with different levels of convenience and financial (dis)incentives provided to the consumers. Examples of such systems with financial (dis)incentives include deposit-refund systems for recyclables (most commonly beverage containers) and pay-as-you-throw approaches for residual waste, and those without financial incentives include kerb-side collection systems and collection centre (“bring”) systems (Tojo & Lindhqvist, 2010).

Similarly to allocation of responsibility for collection and sorting, there are also some rooms for Member States to choose coverage of concrete items to be sorted in the case of, for example, packaging, so long as the concrete collection targets (see Section 3.1.2) are archived. This has lead to diverse solution as to which fractions of plastics are collected. For instance, while all the plastic packaging are currently sorted and collected in Sweden (Viklund & Fråne, 2017), it is only plastic bottles and flakes that are collected in Belgium (Spasova, 2014). Though the details differ from one another, many municipalities in Japan establish much more detailed sorting system for plastic packaging, among others. Many supermarkets collect 1) white styrofoam, 2) transparate hard plastics and 3) PET bottles used for food packaging in three separate fractions, and accept only those that are washed and free from food deblis. Other plastics are collected separately by the municipalities, but they often also require that the sorted plastic fractions are clean. While the sorting most likely facilitates use of these

15 In the case of cars, Article 5 of the Directive 2000/53/EC stipulates that Member States shall see to it that the

“economic operators,” who are “producers, distributors, collectors, moter vehicle insurance companies, dismantlers, recoverers, recyclers and other treatment operators of end-o-flife vehicles” (Article 2(10)), set up systems for collection. In the case of EEE, despite that Article 5(2) of the Directive 2012/19/EU indicates the allocation of responsibility to distributors on one-for-one, old-for-new basis and Article 5(3) suggests collection responsibility on distributors with area more than 400 m2 for very small WEEE (external dimension no more than 25 cm) regardlesss of new purchase, the Directive keeps the possibility for Member States open to introduce other solutions under certain circumstances.


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fractions as recycled materials, the time and efforts it requires from citizens also face criticisms (Takahashi, Itoh & Tojo, 2013)

3.1.2 Collection targets

The collection targets provide quantitative measurement of level of collection/sorting achieved (Tojo & Lindhqvist, 2010). Similarly to the collection/sorting requirements (Section 3.1.1), the assumption behind setting such a target is, the higher the achieved collection rate is, the more recycled materials are saved for their potential use, including in production.

In addition to the collection/sorting requirements, the EU Directives for batteries and electrical and electronic equipment mandates specific collection targets to be met (see Table 1). In these two cases, the two primarily basis of the calculation is the total weight of products put on the market (POM) and the total weight of targeted products coming into the waste stream (Article 7(1), Directive 2012/19/EU; Article 3(17), Directive 2006/66/EC). In both cases, considering the longevity of the products, what comes to the denominator is the average of POM figures of three years.16 WEEE Directive (2012/19/EC) also provides an alternative where a Member State could calculate the collection rate by dividing the weight of WEEE separately collected by the weight of WEEE generated. It is worth noting that the collection target of the original WEEE Directive was absolute (4 kg per person per year of WEEE from private households) – its implication will be discussed under Section 3.5.

The Directive on packaging (94/62/EC) does not have any collection targets but have reuse/recycling targets. Meanwhile, the recycling rates is calculated by dividing the weight of what enters the recycling operation by the waste generated in the same year, which “may be deemed to be equal to the amount of packaging placed on the market in the same year” (Article 6a(1), Directive 94/62/EC). Therefore, the recycling rate calculation combines the calculation of collection rate in effect.

In the case of these three waste streams governed under the EU laws, the collection rate requirements have increased over time. In the case of WEEE, the formulation of the collection targets also changed, as mentioned above. As the legal basis of these Directives concerning the waste related provisions is so-called environmental clause (Article 192(1) of the Treaty on the Functioning of the European Union), some Member States have more strigent requirements than what is stipulated at the EU level.

Similarly to the situation regarding the collection/sorting requirement, although it is Member States who need to ensure that systems for achiving the stipulated collection rate are developed, they have the liberty to assign the responsibility to specific actors for the achievement of targets. It is typically the actors responsible for collection who also become responsible for achieving the targets.

The EU Directive on end-of-life vehicles (ELV) (2000/53/EC) does not include any collection targets, nor do some laws for WEEE outside of EU, such as Japan and Switzerland (Tojo, 2004).

16 In the case of EEE the POM figures of the preceding three years is used, while in the case of batteries, the POM figure

used is the figure of the same calendar year and the preceding two years.


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3.1.3 Recycling targets

The EU Waste Framework Directive (2008/98/EC) defines the term “recycling” as “any recovery operation by which waste materials are reprocessed into products, materials or substances whether for the original or other purposes.”, excluding energy recovery and backfilling operations (Article 3.17, 2008/98/EC). This definition is directly referred to in the Directives for packaging (94/62/EC) and EEE (2012/19/EC) while very similar formulation of “the reprocessing in a production process of the waste materials for the original purpose or for other purposes but excluding energy recovery” is found in the Directives for cars and batteries (Article 2(7), 2000/53/EC; Article 3(8), 2006/66/EC). Recycling targets thus should ideally measure the achievement of the rate of waste materials reprocessed for the original or other purposes.

Except for textiles, the EU laws have mandatory recycling targets for waste streams covered in this study (see Table 1). Concerning textiles, the Swedish Environmental Protection Agency, in their work commissioned by the Swedish Government to suggest policy options for the sustainable handling of textiles, included numerical targets for preparation for reuse and recycling (Naturvårdsverket, 2016).

The way the targets are calculated differ between products with relatively short life (packaging) and those with longer life (batteries, EEE, cars, textiles). For the former, as discussed in Section 3.1.2, the targets are calculated by dividing the weight of material stream that enters the reclycling operation by the weight of the same material stream put on the market. As the material stream that enters the recycling operations should have “undergone all necessary checking, sorting and other preliminary operations to remove waste materials that are not targeted by the subsequent reprocessing and to ensure high-quality recycling” (Article 6a(1)(b), Directive 94/62/EC), what is measures is essentially the collection rate of material streams that are good enough for recycling by weight.

One crucial difference of recycling rate calculation between packaging and others is that regarding the latter, it is based on what is collected, NOT what is put on the market. In other words, it does not concern how much waste which could be potentially recycled is collected, but examine the propotion of materials that could be recycled within the waste collected. Further differences exist among these latter products. In the case of end-of-life vehicles, the weight-base targets are set per vehicle which comes into the recycling operation (Article 7, Directive 2000/53/EC). Concerning batteries, the expression “recycling efficiencies” are used, and weigh-based targets with some qualitative requirements are set for three different types of batteries (Annex III, Directive 2006/66/EC). Regarding WEEE, similarly to packaging, the numerator is “the weight of the WEEE that enters the recovery or recycling/preparing for re-use facility, after proper treatment…” (Article 11(2), Directive 2012/19/EU). Despite that Member States are to ensure to keep records of “…materials when leaving (output) the recovery or recycling… facility” for the potential development of calculating the recycling rate based on output from the recycling facilities (Article 11(4) & 11(6), Directive 2012/19/EU), from what can be found from the European Commission website, as of February 2020, no further decision has been made. This means that concerning WEEE, the figures reported by Member States to the EU is NOT the percentage of actual materials that came out of the recycling process, but rather, what goes into the recycling plant.

Similarly to the collection targets, the recycling targets for WEEE and end-of-life vehicles have become more stringent over time, and some Member States set more stringent targets than what is stipulated under the EU Directives.


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The definition of recycling as summrised in the beginning of this section clearly indicate the intention that the recycled materials are to be used for the original or other purposes. The recycling rates should ideally indicate the proportion of products in the waste stream that are processed for such use. Examination of the EU law texts indicates, however, that recycling rate as is currently defined does not provide that measurement in the case of packaging and WEEE. Except for the Battery Directive, none of the relevant directives mention of the qualitative requirement pertaining to recycled materials. The implications of these issues will be discussed further under Section 3.5.

3.1.4 Separation of components containing hazardous substance

Despite measures to reduce hazardous substances contained in products from the beginning (See Section 3.1.5), many products currently entering the waste stream still contain some substances that have negative impacts on human and the environment. In addition to protecting the workers in the reuse/recycling plants as well as people living in the vicinity of these facilities, removal of these substances prior to further processing for reuse and recycling should help in keeping the material cycle free from contamination with these substances.

All the six product/waste streams covered in this study can potentially contain such substances. Among them, there exist EU laws mandating the sorting of such substances in end-of-life vehicles, EEE and batteries (see Table 1). Concerning end-of-life vehicles, Article 6 (3) of the Directive 2000/53/EC stipulates the necessity of stripping hazardous materials and components before further treatment, removing and segregating them, and that the operation should be carried out “in such a way as to ensure the suitability …. in particular for recycling”. Article 8(2) of the WEEE Directive (2012/19/EU), requires removal of fluids and various components listed in Annexi VII(1), and further stipulates specific treatment as described in Annex VII(2). Furthermore, the European Standardisation Organizations, upon the request from the European Commission, developed European standards for treatment and depollution of various components of WEEE (European Commission, 2019b).

In the case of batteries, in addition to minimum treatment requirements as stipulated in the Annex III Part A of the Directive 2006/66/EC, Article 12(1) also indicates the possibility of Member States to dispose the collected batteries containing cadmium, mercury or lead in landfills or underground storage, if there is no viable market for them. Recycling efficiency requirements in Annex III Part B also require, in addition to meeting the weight-based percentage target, the recycling of lead and cadmium “to the highest degree that is technically feasible while avoiding excessive costs”.

3.1.5 Landfill bans

As spelled out in, among others, Recitals (2), (8) and (10) of the amendments (Directive (EU) 2018/850) of the original Landfill Directive (1999/31/EC), restricting specific waste streams from landfilling is considered as a way of ensure the implementation of waste hierarchy as well as to enhance the closure of material use.17 Prohibition of the landfilling of materials that have good poteintial for recycling at least indirectly promote recycling by making it more difficult to dispose these materials in landfills.

17 For example, Recital (10) of the amendments states: “A progressive reduction of landfilling is necessary to… ensure that economically valuable waste materials are gradually and effectively recovered through proper waste management and in line with the waste hierarchy…”


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The original EU Landill Directive (1999/31/EC) bans the landfilling of, among others, use tyres and shredded used tyres (Article 5(3)(d)). The amendments to the landfill Directive that came into effect in 2018 added more waste streams (paper, metal, plastic, glass, textiles, wood, concrete, bricks, tiles and ceramics, stones and plasters) in the prohibition list (Article 5(3)(f)), reflecting the waste streams that have become subject to mandatory sorting in the latest revision of Waste Framework Directive (Article 11(1), Directive 2008/98/EC).

3.1.6 End-of-waste and by-products criteria

One of the most discussed items when the original Waste Frame Directive from 1975 underwent the revision process was how to distinguish waste from non-waste. There are differences in what rules to follow if the object being handled is waste or not: for instance, transportation of waste is subject to waste specific legislation both when the transport includes the movement acrsoss the national border and within the national border. It has been argued that compliance with waste legislation incurs additional cost. In addition, case-by-case decision regarding what is waste and non-waste by different EU Member States has caused uncertainties to the industries who handle the object (Levänen, 2015). All in all, distinction is considered useful in reducing the economic burden from the actors engaged in the transaction of the object, and therefore, enhance the used of recycled materials (European Commission. (2003).

After several rounds of negotiation between the European Commission, the European Parliament and the Council, the revised (and current) waste framework directive (2008/98/EC) includes the provision of both by-products (Article 5) and End-of-waste (Article 6). Although similar in character, by-products are “a substance or object resulting from a production process the primary aim of which is not the production” and if not considered as waste, should be able to “be used directly without any further processing other than normal industrial practice” (Article 5), whereas in the case of end-of-waste, substance or object goes through a recycling or other recovery condition (Article 6). In both cases, the Directive provides specific criteria to be met in order for the substance or object to be considered by-products or cease to be waste.

Since the entry into force of the Waste Framework Directive in 2008, there appears to be no court cases regarding by-products. Meanwhile, Lavänen (2015) provides an example of how the new criteria laid out in the revised Finnish Waste Act, in line with the provision in the Directive 2008/98/EC, has helped a steel making company to have one of its production residues to be considered as a by-product instead of waste.

Regarding end-of-waste criteria, the European Commission, as of February 2020, established criteria for three material streams in the form of Council Regulations – iron, steel and aluminium scraps ((EU) No 333/2011), glass cullet ((EU) No 1179/2012) and copper scraps ((EU) No 715/2013). The technical study for waste paper, bio-waste and waste plastics have been also published by the Insitute for Prospective Technological Studies of the Joint Research Center under the European Commission in 2011, 2014 and 2014 respectively (Villanueva & Eder, 2011; Saveyn & Eder, 2014; Villanueva & Eder, 2014). In the case of waste paper, the European Commission put together a proposal for the regulation (European Commission, 2013). However, with different discussions taking place in the middle, none of three have been finalised in the form of Council Regulation as of February 2020.

The latest amendments to the Waste Framework Directive (2008/98/EC) that came into effect in 2018 include substantial revisions to both Article 5 and 6, and especially Article 6. In


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both cases, the main change is to include a clearer responsibility on Member States to take appropriate measures so that by-products are to be recognised as by-products, and that a substance or object meeting end-of-waste criteria is to receive end-of-waste status (Article 5(1), Article 6(1)). Compared to the original wording of the 2008 Directive where the development at the EU level was somewhat prioritised, the wordings of the amendments regarding end-of-waste criteria (Article 6(2)) appear to encourage the development at the Member States level more. The revision further enhances the possibility for Member States to make case-by-case decision by, for instance, changing the wording of criteria (Article 6(1)(a)) and not requiring the Member States to report on such decision to the Commission (Article 6(4)) (See further discussion under Section 4.1).

3.1.7 Substance restriction

The most efficient way of avoiding the contamination of material loop with unwanted substances is not to include them in the products to start with. While there are different types of “contamination” that pose challenges to recycling, such as mixture of different materials, additives or waste, hazardous substances also hinder the circulation of materials. Substance restriction, thus, in addition to its primary function of protecting human and the environment from hazardous substances, facilitates the closure of material loop by prohibiting the inclusion of such substances at source.

In the EU, the laws on packaging, EEE, cars and batteries prohibit the use of selected heavy metals. In addition to mercury and lead whose use is prohibited above the theashold level stipulated in the respective laws, lead and hexavalent chromium are also listed in the law for packaging, cars and EEE (Article 11, Directive 94/62/EC; Article 4, Directive 2000/53/EC; Article 12, Directive 2006/66/EC; Article 4, Annex II, Directive 2011/65/EU). In the case of EEE, the prohibition is extended to two types of brominiated flame retardants and four types of phthalate as well (Annex II, Directive 2011/65/EU).

Meanwhile, the directives on end-of-life vehicles as well as restriction of the use of hazardous substances (RoHS) in EEE both contains a long list of exemptions for specific applications (Annex II, Directive 2000/53/EC, Annex III and IV of Directive 2011/65/EC). The amendments of RoHS Directive in 2017, via Directive (EU) 2017/210, included a few provisions to facilitate reuse of repair parts even when they contain substances prohibited by the Directive.

In addition to these directives specific to particular product groups, the so-called REACH Regulation (EC/1907/2006) governing the registration, evaluation, authorisation and restriction of chemicals in Europe restricts the manufacturing, placing on the market and/or use and sales of chemicals listed in Annex XVII of the Regulation, unless it complies with the conditions specified in the same Annex (Article 67). According to the Euroepan Chemical Agency, as of February 2020, seventy chemical substances are on the restriction list (Euroepan Chemical Agency, 2020c). Many of these restricted substances have relvance for six product groups covered under this study (Euroepan Chemical Agency, 2020b). Moreover, substances included on the REACH candidate list of substances of very high concern may be subject to aurhorisation requirements. Companies a substance that would require authorisation must provide information on an analysis of alternatives, technical and economic feasibility of substitutions and the like when applying for the authorisation (Article 62(4), Regulation (EC) 1907/2006). According to the European Chemical Agency, 205 substances are on the candidate list (Euroepan Chemical Agency, 2020a).


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3.1.8 Product design requirements related to closure of material loops

In addition to restricting the use of hazardous substances, there exist several other provisions mandating or encouraging “design for closure of material loops”, and more specifically, design for recycling. The overall logic is, material recycling is facilitated by taking into consideration the recyclability aspects at the design phase.

Concerning packaging, the EU Packaging Directive (94/62/EC) requires the fulfilment of essential requirements (Article 9) as laid out in Annex II. After the revision in 2018, Annex II(1) stipulates that “Packaging shall be designed, produced and commercialised in such a way as to permit its… recycling”. Annx II(3)(a) further requires that “Packaging must be manufactured in such a way as to enable the recycling of a certain percentage by weight of the materials used into the manufacture of marketable products, in compliance with current standards in the Community.” Regarding cars, Article 4(1)(b) of the EU Directive (2000/53/EC) requires Member States to encourage “the design and production of new vehicles which take into full account and facilitate the dismantling, reuse and recovery, in particular the recycling, of end-of life vehicles, their components and materials.” Article 4(1)(c) further requires Member States to encourage the use of recycled mateirals.

The EU battery law does not have any specification regarding design except for material restriction (See Section 3.1.7), but they set a design requirement for the manufacturers of products that use batteries for their functioning. Article 11 of the Directive 2006/66/EC requires that such products should be designed in such a way that waste batteries “can be readily removed by qualified professionals that are independent of the manufacturer”, unless the function of the products requires otherwise. Article 4 of the WEEE Directive (2012/19/EU) requires Member States to encourage “measures to promote the design and production of EEE, notably in view of facilitating re-use, dismantling and recovery of WEEE, its components and materials.”

As found, with the exception of battery removal requirement, the existing laws in the EU includes provisions that encourage design-for-recycling on general terms, but without any specific requirements.

With regard to energy-using, and later, energy-related products, such as EEE, there exist another EU law setting eco-design requirements for those products (Directive 2005/32/EC, as replaced by Directive 2009/125/EC). Despite its nickname of “eco-design directive”, however, the criteria set through the implementing regulations of different products have been concentrated mostly to use-phase energy efficiency (van Rossem & Dalhammar, 2010; Dalhammar et al., 2014; European Commission, 2015; Talens Peiróa et al., 2020). Some recent development of the Directive provides new prospects, however, as will be discussed further in Section 4.3.

3.1.9 Extended Producer Responsibiltiy (EPR) programme

The concept of extended producer responsibility (EPR) was first coined and defined by Lindhqvist & Lidgren (1990) in a report submitted to the Swedish Ministry of Environment (Miljödepartement, 1991), and has been widely used as a key concept governing waste and product over the last several decades. An international survey of the understandings of the concept engaging in total of 376 stakeholders (e.g. practitioners such as policy makers, producers, municipalities, actors engaged in the organisation of EPR systems as well as academia) reveals exsistence of various views on the concept (Tasaki, Tojo & Lindqhvist,


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2019). However, the majority of the respondents indicates that the multiple aims of an EPR programme covering both the influence on the upstream of the product (i.e. design of products and product systems) as well as downstream (i.e. end-of-life phase) (Tasaki et al., 2019). The Organisation for Economic Co-operation and Development (OECD) retained its definition found in its original guidance manual from 2001 and defines EPR as “an environmental policy approach in which a producer’s responsibility for a product is extended to the post-consumer stage of a product’s life cycle.” (OECD, 2016, p21).

EPR being a policy principle (Davis, 1994; Lindhqvist, 2000; Krämer, 2007), an EPR-based law typically consists of a multiple policy instruments, and the type of responsibility, as well as the extension of responsibility to producers, vary between product groups and between jurisdictions (Tojo, 2004). Subtantial number of laws on packaging, cars, EEE and batteries introduced in the EU and elsewhere have been based on EPR, 18 and they typically incorporate some or all of the instruments discussed in Sections 3.1.1 to 3.1.4, 3.1.7 and 3.1.8 (See Table 1).

While the intervention theories of an EPR programme incorporates the theories of individual policy instrments constituting the respective programme (Tojo, 2004), an essential logic of an EPR programme encompassing such instruments is the mechanisms of incentive provision to producers. That is, by making producers responsible for end-of-life management of their products, producers have opportunities to receive feedback regarding how their product design might influence downstream opearations, and are incentivised to make appropriate upstream changes so that their burden at the end-of-life phase is reduced.

A critical question among the practitioners as well as policy makers over the last several decades is, how to design and implement an EPR programme in such a way that it does provide design incentives to producers as intended in the concept. The discussion on this issue is shaped around how to implement a so-called individual producer responsibility (IPR) in practice. In short, a producer has an individual producer responsibility “when he/she takes repsonsibiltiy for the end-of-life management of his/her own products” (Tojo, 2004, p243). The importance of allocating IPR was recognised in the development of the original WEEE Directive (2002/96/EC), as manifested in the rather clear articulation of IPR for new WEEE (the waste from WEEE put on the market after the law comes into force). The Directive distinguishes the new WEEE from so-called historical WEEE (the waste from EEE put on the market prior to the entry into force of the original WEEE Directive) which a producer cannot influence its end-of-life fate by changing the design upstream (Article 8, Directive 2002/96/EC, as inherited to Aricle 12, Directive 2012/19/EC). There has been considerable disputes and discussions as to whether and how to make an IPR work in relation to the development and implementation of the original WEEE Directive (2002/96/EC), especially when the physical management of waste products are conducted collectively (Kalimo et al., 2012; van Rossem, 2008; Sander et al., 2007; Tojo, 2004). Reflecting upon the debate in light of moving forward with a circular economy, the latest amendments to the Waste Framework Directive (2008/98/EC) includes, among others, an incorporation of “modulation” of a fee in an EPR system operated jointly for multiple producers (Article 8(a)4(b)). The modulation

18 As reflected in the overall strategic decision made by the EU to introduce the principle of producer responsibility

(Krämer, 2007), the EU Directive on cars (2000/53/EC), EEE (2012/18/EU) and batteries (2006/66/EC) integrates the EPR principle. Though the Directive on packaging (94/62/EC) is not explicitly based on EPR, the majority of the EU Member States implement the Directive through an EPR-based system. Many OECD countries outside of the EU (e.g. Norway, Switzerland, Japan, Canada) also base many of their legislation governining specific waste streams on EPR, and its use is spreading to non-OECD countries. For more information, see, for instance, Tojo et al. (2003), Manomaivipool (2011), OECD (2016).


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should be done “by taking into account (products’) durability, reparability, re-usability and recyclability and the presence of hazardous substances…” (Article 8(a)4(b)).

3.2 Intervention theories of prevailing policy approaches Based on the descriptions of various existing policy measures addressing the closure of loop in Sections 3.1.1 to 3.1.9, Figure 4 depicts the overall, simplfied intervention theories of these measures. Tje figure concentrates on the outcome investigated in this study – increase use of recycled materials in production.

Figure 4: Intervention theories of existing policy measures pertaining to increased use of recycled materials in production

Source: own illustration based on Tojo (2004) and materials presented in Sections 3.1.1 to 3.1.9

3.3 Achievements and outcomes Starting from the instrument in the upper end of Figure 4, waste streams subject to landfill bans relevant for this study were mostly included in the law only in 2018, and no study regarding its effectiveness for these newly added waste streams was found. The only one stream whose landfill restriction has been implemented for some time and of relevance to this study is tyres, as part of cars. Although no recently study is found, the study conducted for the European Commission in 2005 (4 years after the landfill ban of whole used tyres, and 1 year after the landfill ban of shredded tyres came into force) reported that the landfilling of whole tyres had been noticeably declined, and that most countries appeared to increase energy recovery or recycling of waste tyres (Golder Associates, 2005). Although it is difficult to generalise the success of a policy for different types of product/waste streams, the study provides an indication that there was a tangible contribution of a landfill restriction to recycling or energy recovery. What is unclear, however, is the contribution to recycling as opposed to energy recovery.

The quantitative information of the achievement of the first three instruments under the EPR programme (collection/sorting requirements, collection targets and recycling targets) is


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available for four of the six waste streams covered in this study – packaging, EEE, batteries and cars.

Table 2 provides the change of collection/recycling rate of packaging waste (i.e. the weight of packaging waste that enters the recycling operation divided by the weignt of same category of packaging put on the market – there is no separate targets for collection, see Sections 3.1.2 and 3.1.3) in the EU-28 Member States and Sweden between 2012 and 2016. The targets set for the respective waste streams for 2025 and 2030 at the EU level, as well as those set in the Swedish national law (SFS 2014:1073) is also included. On one hand, compared to the targets set in the respective laws, the achievement has been fine. In fact, as of 2016, the achievement of the EU-28 countries concerning all but plastic packaging has exceeded the EU 2025 targets, and for metals and wood packaging, the EU 2030 targets. Looking at the Swedish situation, already in 2016 the national target set for all packaging and glass exceeds the national target for 2020 onwards. Given the already high achievement, the strigency of the target could be questioned. It should be also noted that reliability of the statistical figures, and in particular waste, is questionalable.

Table 2: rate of packaging waste entering the recycling facilities, EU-28 and Sweden, 2012-2016, the recycling targets stipulated in Directive 94/62/EC for 2025 & 2030 and in Swedish packaging ordinance (SFS 2014:1073) before and after 1 January 2020 (unit: percentage)

Packaging streams

Countries covered

2012 2013 2014 2015 2016 EU target Swedish target

2025 2030 Before 2020

After 2020

All packaging

EU-28 64.7 65.3 65.5 65.8 67.2 65 70 55 65

Sweden 69.6 71.9 70.5 71.8 68.2

glass EU-28 72.3 72.6 74 72.9 74.1

70 75 70 90 Sweden 88.2 89 94.7 93.6 92.8

plastic

EU-28 35 36.6 38.9 39.9 42.4

50 55 30 exl PET,

90 PET bottles

50 exl PET, 90

PET bottles Sweden

34.9 45.6 47.5 49 50.7

paper & cardboard

EU-28 83.9 84.7 82.5 83 85 75 85 65 85

Sweden 76.8 78.4 79.3 81.7 81.8

Metal

EU-28 72.3 74.2 75.2 76.1 78.3 70 ferrous metals,

50 aluminium

80 ferrous metals, 60 aluminium

70 exl beverages,

90 beverages

85 exl beverages,

90 beverages Sweden

74.4 77.4 76.5 77.7 81.5

wood EU-28 38.7 35.9 38.5 40.1 39.8

25 30 15 15 Sweden 79.8 59.9 22.3 21.5 30.9

Source: developed by the author based on Eurostat (2020), Directive 94/62/EC and SFS 2014:1073.

Table 3 presents the collection results of spent batteries for EU-27 countries19 and Sweden as found in the Eurostat, as well as the target sets in the EU law (Directive 2006/66/EC) and the Swedish national law (SFS 2008:834).20 As of 2016 the average collection rate of EU-27 countries exceeds the target set at the EU level. However, a study conducted for the

19 While the Eurostat provides data for EU-28 countries (including UK) for waste streams other than batteries, the figures

provided for batteries was limited to EU-27 countries. Considering the relative similar representativeness, the author uses the figure for batteries.

20 Regarding the way the collection rate is calculated for batteries, see Section 3.1.2.


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European Portable Battery Association in 2017 reports on the unreliability of the data, and questions the appropriateness of basing the calculation on POM (put on the market) figure (Perchards & SagisEPR, 2017). The problem with cheating on POM figures was also reported in relation to the packaging directive (Spasova, 2014) and could explain an apparent high level of collection/recycling rate achievement.

Table 3: Collection rate of batteries and accumulators, EU-27 and Sweden, 2011-2018, the collection targets stipulated in Direcive 2006/66/EC for 2016 and in Swedish ordinace for batteries (SFS 2008:834) (unit: percentage)

Countries covered 2011 2012 2013 2014 2015 2016 2017 2018

EU target 2016 Swedish target

EU - 27 35.2 36.8 37.3 39.7 42.9 46.8 45.5 47.6

45

95 automotive & industrial, 75

non-automotive & industrial Sweden 53 61 64 59 61 45.1 51.2 48.5


As of February 2020, Eurostat does not provide the recycling rate achieved for different types of batteries for the EU-27 countries. Concerning lead batteries and nickel-cadmium batteries in Sweden, the results available from 2014 onwards shows that the recovery of lead is above 97%, and cadmium, 100%. For the recycling of other types of batteries, figures available between 2009 and 2013 vary between 75 to 88% (Eurostat, 2020). No figure is available after 2014.

Table 4 provides the reuse and recycling rate of end-of-life vehicles achieved in EU-28 countries and in Sweden between 2008 and 2017, as well as the targets set at the EU level and at the Swedish level (the same percentage but the Swedish law introduced the mandate one year earlier than the EU requirement). From what we can find from Eurostat, the recycling of cars coming to car dismantlers seem to be progressing well.

Table 4: Reuse and recycling rate of end-of life vehicles, EU-28 and Sweden, 2008-2017, the reuse and recycling targets stipulated in Directive 2000/53/EC for 2016and theSwedish ordinance for cars (SFS2007:185) (unit: percentage)

Countries covered 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017

EU targets

after 2016

Swedish target after 2015

EU – 28 82.8 82.1 83.5 84.3 84.5 85.3 85.7 87 87 87.5 85

85

Sweden 83 86 84.4 84.4 85 84.6 84.4 84.6 86.7 88.2


Table 5 provides the figures for the amount of EEE put on the market, total WEEE collected as well as recycled, in EU-28 and Sweden on per capita basis, found in Eurostat. Though the collection rate figure is not currently available (for detailed calculation methods, see Section 3.1.3), the figures from the table indicates that there is a big difference between the amount of products put on the market and those collected. Considering the longevity of the products, however, it is difficult to assess the level of success regarding collection. Concerning reuse and recycling rate, although there is a significant difference between different types of EEE (see, for instance, Richter 2019), the overall recycling rate achievement in terms of weight has been relatively high.


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Table 5: Per capita figures of put-on-the-market (POM), collection and reuse & recycling, EU -28 and Sweden, 2010-2017 (unit: kg)

Countries covered categories 2010 2011 2012 2013 2014 2015 2016 2017

EU - 28

POM 18.76 18.17 17.94 17.24 18.15 19.25 19.77 20.55

collection 7 7.06 6.88 6.96 6.86 7.63 8.85 8.93

reuse & recycling 5.55 5.74 5.42 5.59 5.6 6.17 7.4 7.44

Sweden

POM 24.78 24.51 23.02 25.28 24.54 26.22 26.1 27.71

collection 17.21 18.69 17.71 18.39 14.94 14.69 16.45 14.06

reuse & recycling 14.45 15.63 14.93 15.4 12.57 12.29 13.72 11.74

Source: developed by the author based on Eurostat (2020)

All in all, despite taking into account the potential inaccuracy of data, it can be concluded that the collection/sorting and recycling of the four product groups have been going relatively well.

Concerning the separation of components containing hazardous substances, according to the implementation report of the WEEE Directive commissioned by the European Commission, Member States have been transposing the requirements specified in the Annex VII in the national laws (Eunomia, ENT, IVL & EPEM, 2018). Similarly, the latest implementation report of the ELV Directive commissioned by the European Commission indicates transposition of Article 6(3) into national legislation, although pointing out the unclarity of the situation in Sweden based on the answers received from the Swedish authority. The report also pointed to the concern on the prevailing exisence of illegal dismantling (Elliot et al., 2019).

Concerning the effects of material restrictions, interviews with producers conducted at different times indicate a very strong influence of both the RoHS Directive (2002/95/EC and 2011/65/EU) as well as the REACH Regulation ((EC) No 1907/2006) in making sure that appropriate design changes are made to avoid inclusion of (potentially) restricted substances (Tojo, 2004; Tojo & Thidell, 2012).

With regard to product design requirements, the afformentioned implementation report of the WEEE Directive concludes that, while the national laws contain elements to support producers to work on design-for-recycling in accordance with Article 4 of the Directive (2012/19/EU) and some good examples are identified (e.g. differentiated fee for producers with good eco-design practices, financial support provided to eco-design activities, development of eco-design guidelines), a prevailing problem is that manufacturers are often outside of their territory (Eunomia, ENT, IVL & EPEM, 2018). Concerning the Article 4(1)(b) and (c) of the the ELV Directive, the implementation report of the Directive reports that while most of Member States incorporate relevant clause in their national laws, concrete actions might not have been implemented (Elliot et al., 2019). Elliot et al (2019) also highlighted that there is a tendency among Member States to incorporate the clauses in the general waste law, instead of those specific to cars.

Concerning the overall effects on EPR programmes, in addition to the achievements mentioned above, several studies, such as Tojo (2004), as well as information from some producer responsibility organisations for packaging, point to its effects to upstream changes, especially at the initial phase of the introduction of EPR programmes. Meanwhile, the overall


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perception has been that they have been “seldom sufficient to serve as the triggering factor” (OECD, 2016, p23).

Finally, with regard to by-products/end-of-waste criteria, an evaluation of the first end-of-waste criteria regulation on aluminium and iron scraps reveals that the vast majority of companies cetified or generating end-of-waste compliant scraps are located in Italy, while in Sweden, similarly to many other countries, no company was registered as producing end-of-waste compliant scraps. The highest number of respondents participated in the study mentioned lack of supply of compliant scrap as not increasing the purchase of such. Meanwhile, one of the respondents, presumably from Sweden, was quoted “Swedish steel mills do not see any reason to start to use end of waste steel scrap”, indicating that the quality of the scrap available in the country is already of sufficiently high quality. Regarding the question on the benefits of complying with the criteria, the alternative that had the highest number of responses was “no benefit observed”. However, the second and the third highest alternatives were increased customers and increase sales. There are also some respondents commenting on the benefits of reduced administrative burden and lab tests (Saveyn et al., 2014).

With this very slow and limited uptake of end-of-waste criteria in practice, however, it is difficult to evaluate its effect on the further steps of interevention theories.

Table 6 summarises the main achievements and outcomes of existing policy measures pertaining to the increased use of recycled materials in production as discussed in this section.

Table 6: Summary of achievements and outcomes of existing policy measures pertaining to increasing the use of recycled materials in production

Policy measures Achievements and outcomes

Collection/sorting requirements Tangible achievement in terms of sorting selected waste streams from the rest of waste Collection targets

Recycling targets Tangible achievement in terms of bringing considerable portions of sorted waste streams to recycling plants (packaging and WEEE), and for processing the incoming materials (batteries and cars)

Separation of components containing hardous substances

Relevant clause available in national and EU laws

Landfill bans Tangible effects in promoting recycling/energy recovery experienced for tyres

End-of-waste and by-products critera Very limited implementation except for Italy

Substance restriction Tangible effects upstream

Product design requirement related to closure of material loops

General clauses available in national and EU laws, with some concrete policy actions for implementation

EPR programme Potential design effects and IPR recognised and articulated in laws


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3.4 Challenges Despite the positive outcomes mentioned in Section 3.3, the overall outcome of the six research projects under the Phase II of the Mistra Closing the Loop research programme reveals that there is a clear limitation in the use of recycled materials in production.

As found in the intervention theories of the existing policy measures (Figure 4), there is an assumption that sorting a specific waste stream from the rest of the waste stream would lead to the development of a more homogeneous and cleaner recycled material stream, which in turn would lead to the creation of more demand for the recycled materials generated from the sorted materials. There has been indeed some experiences on that, for example when the collection of aluminium cans started in the United States long time ago (Peck, 2003). Some electronic industry has been also using the PET bottles and CD casing as raw materials for their new products (Tojo & Thidell, 2012). However, as seen in these examples, especially aluminium cans and PET bottles, they are made of homogeneous materials themselves, and especially with the development of deposit-refund systems for beverage containers, the chances where unexpected inpurities come into the waste stream is relatively low.

The situation is quite different for more general packaging materials, and especially plastics. There are many different types of plastics, with different properties and additives, making it very challenging to establish a homogeneous waste stream. This is reflected, on the one hand, the relatively low achievement of collection/recycling of plastic packaging compared to other packaging materials. On the other hand, the measurement of recycling rate, as it stands now, is based on what is going into the recycling plants (see 3.1.2 and 3.1.3), and there are many uncertainties on the quality and destiny of what is coming out of the recycling plants. According to the Plastic Europe, as cited in European Commision (2019a), of 27.1 million tonnes of plastic waste collected in Europe in 2016, only 8.4 million tonnes (corresponding to 31.1%) went to recycling facilities. Compared to the total demand for plastics, the demand for recycled plastics remains less than 8 % (European Commission, 2019a).

As a way of dealing with that, some European countries, such as Belgium, collects only those relatively homogenous plastics with higher market value (Spasova, 2014). However, if only those plastics start to come under an EPR programme where producers have to finance, a potential consequence could be that producers start to use other materials so as not to have to be part of the system. There are other systems, such as those in Japan in which consumers are requested to carry out much more extensive sorting and cleaning than what is typically practiced in Europe (see Section 3.1.1). However, feasibility of implementing such systems in Europe is highly questionable considering the higher burden on consumers, what they have been used to and the like.

The challenges related to plastic packaging is a prevailing problem for other products using plastics (including cars, EEE, batteries), and any other products with mixed materials. It is one of the most highlighted challenges in relation to fibre-to-fibre recycling of textiles (Elander et al., 2017). While there are many sorting technologies developed, the challenge still prevails. All in all, there seems to be a fundamental failure in assuming that, by sorting to a level “reasonable” to engage consumers, or in the case of industry, floor workers, there will be a waste stream sufficiently clean and attractive to be processed into a recycled material stream.

In addition to the failure in assumptions mentioned above (“theory failure”), there has been also failure in implementation. That is, sorting is not always done as it is intended. It can be anything from a simple mistake of putting waste in a wrong bin to misunderstanding of the


27

meaning of waste (e.g. some citizens misunderstood that light bulbs should be sorted as glass, instead of EEE). The practical difficulties of sorting products made of multiple materials (e.g. glass bottles with plastic lids) also contribute to this failure.

As has been raised for decades as a primary logic for waste prevention, the highest in waste hierarcy, what can be done at the waste stream is limited, which brings us look at policy measures addressing upstreams of the products. Looking at upstream changes, there are some tangible policy effects on the reduced use of hazardous substances. However, unless the material mixed is one of the hazardous substances addressed in a policy (e.g. brominated flame retardants), the challenge of mixed materials is not solved. Policy measures to address issues that are not related to toxic substances are rather limited and are currently staying at a rather general level. Another challenge regarding especially the REACH regulation is its stringent requirements. There is on one hand a clear benefit of substance restriction to the closure of material loops by “cleaning” the material flow from the beginning. On the other hand, there has been criticisms, pointing to the situation where stringent requirements by the REACH on material safety may compromise the possibility to enhance the use of recycled materials (Alaranta & Turunen, 2017). A staff working document issued by the European Commission regarding the interface between chemical, product and waste legislation acknowledges this potential conflicts, including the potential of derogating parts of the requirements for recycled materials in a specific context for a limited period of time (European Commission, 2018b). As it stands now, however, no clear-cut decision has been made.

Regarding EPR programme, a prevailing challenge is a concrete form of implementing individual producer responsibility (IPR). After many years of discussion by industry actors, NGO groups and academia, in which the author herself participated, a viable IPR system for EEE has not been developed. Another problem that the original WEEE Directive faced was many Member States transposed the directive incorrectly in their national law, especially in relation to the parts concerning IPR (Van Rossem, Tojo & Lindhqvist, 2006; Sander, 2007). Here, implementation failure at different levels is observed.

Concerning the by-product/end-of-waste criteria, the limited uptake by the industry actors seem to be the main challenges. One of the reasons for the limited uptake seems to do with perceived lack of benefits due to the already high standard in the existing market. If the market actors see no difference in the quality with or without the certification of end-of-waste status, it would be difficult to motivate them to take the efforts to obtain the status.


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Figure 5 visualised the areas where existing interventions have not been working well in terms of increasing the use of recycled materials in production.

Figure 5: Areas of existing interventions that have not worked well, pertaining to increased use of recycled materials in production

Source: own illustration based on Tojo (2004) and materials presented in Sections 3.1.1- 3.1.9 and 3.4

3.5 Causes of challenges pertaining to policy design As found, while existing policy measures produced some tangible positive outcomes (Section 3.3), not everything worked out in the way as it was originally assumed or envisioned (Section 3.4). In this section, the author seeks to explore causes of challenges pertaining to policy design. As mentioned in Chapter 1, given that there is a parallel study exploring the drivers and barriers for the increased use of recycled materials in production, factors other than


29

policy design (e.g. market factors, interaction of actors, behaviour of addressees) will not be discussed.

Weight-based targets: a cricisism raised by many in relation to target setting for collection and recycling, especially regarding complex products such as cars and electronics, is that the achievement of target is measured with weight irrespective of the properties of different materials (Huisman, 2003; Kalimo et al., 2012; Richter, 2019). As found in Sections 3.1.2 and 3.1.3, existing collection and recycling targets as identified in this study are weight-based.

While weight-based targets has the clear advantage of the ease of measurement, they are problematic in many ways. First, it tends to induce actors responsible for collection to focus on products that are heavy, regardless of the relative concentration of materials that require attention from societal perspective (e.g. valuable materials, hazardous substances). This is especially the case when the one responsible for meeting the collection targets are different from those taking care of recycling. This was among the main causes of low collection rate of smaller and thus often lighter WEEE, such as mobile phones (see, for instace, Tojo & Manomaivipool, 2011). The revised WEEE Directive (2012/19/EU) sought to address this by setting a specific requiment for distributors to accept very small WEEE regardless of new purchase (see Footnote 15). However, even though the collection target was also changed from absolute to relative (See Section 3.1.2) at the same time, the Directive kept it as a weight-based target. In addition, instead of mandating different categories of WEEE meet the percentage-based target respectively, the revised Directive kept the calculation base on the entire WEEE. Therefore, there are still no incentives for the entity responsible for collection target achievement to strive to collect smaller and lighter WEEE. This leads to loss of various valuable materials contained in smaller and lighter WEEE, which could be used in production process, or hazardous substances that need to be collected. Similar situation exists also for batteries, where inclusion of heavy batteries compensate in achiving the collection rates (Perchards & SagisEPR, 2017)

Concerning WEEE, the same issue prevails also for recycling targets. Though the targets are differentiated between different categories of WEEE, it is still weight based, which means that recyclers could achieve the target without necessarily paying attention to materials within the respective WEEE. The recycler would naturally seek to capture materials with economic values (e.g. precious metals), but economic value would fractuate depending on the market situation (Richter, 2019). Materials that may not have (immediate) high economic value but might still have strategic value (e.g. some rare earth metals in LED lamps) may not be captured in the recycling process when it no longer makes economic sense for the recyclers to take them out from the rest of the materials, even when technological solution exists (Richter, 2019).

Lack of quality requirement for sorting and recycling: a further issue related to the weight-based targets is lack of quality requirement for sorting and recycling. Concerning sorting, when the collection is targeting relatively homogenious stream (e.g. PET bottles, aluminium cans, batteries of specific type), the quality of the waste stream can be maintained relatively high. Regarding packaging, sorting requirements as well as differentiated collection/recycling targets for different types of packaging materials addresses this at least to an extent. This becomes more complicated for mixed materials, especially plastics, and attempts to have further sorting, or sorting only specific materials, both have their limitation (see Section 3.4).

In practice, the quality requirement has been often put when the sorted waste streams are futher sold to the recycling plants. As the recycling plants put certain requirements for


30

acceptance of such mixed waste, presorting is done prior to the mixed waste enters the recycling facilities, at least to an extent.

Once the waste stream comes into the recycling plants, as of today, there is no clear government policy set on the quality of the output materials. Haupt and colleagues (2016), in their study on the management of selected streams of municipal solid waste in Switzerland, shows that while closed-loop recycling does take place, a good proportion of PET, tinplaste, aluminium and paper streams left for open-loop recycling. The study also highlights that there are many uncertainties regarding the fate of these flows. As indicative from the lengthy discussions regarding end-of-waste criteria and the fact that three out of six have not been actualised (See Section 3.1.6), it is deemed very challenging to introduce a policy measure regarding quality requirements for waste streams, considering among others the ever changing nature of flow of the materials, change in the market, political priority and the like.

Meanwhile, quality requirement in terms of prioritisation of specific materials is found to an extent. For example, regarding batteries, instead of recycling “targets”, recycling “efficiencies” requirement is set for at least two types of batteries that includes lead or cadmium, so that these two substances are recycled to the highest degree. Though the EU-wide overview is currently not available, it has resulted in a very high recycling achievement for lead and cadmium in Sweden (see Section 3.4). Regarding cars, a relatively high recycling target was set in the Directive from the beginning (the original recycling target was 80% to be achieved by 2006, which was increased to 85% in 2016), for the reason that roughly 75% of the cars are made of metals and there already existed a viable economic system to recycle them. Thus the target is set to enhance the material closure and environmentally sound management of the remaing 25% consisting of glass, plastics and the like which used to be simply shredded and discarded (Peck, 2003).

Timing of measurement: As explicitly mentioned in the directives on packaging and WEEE, the numerator used to calculate recycling rate is total weight of the waste stream in question that enters the recycling operation (see Sections 3.1.2 and 3.1.3). As quoted in Section 3.1.3, both the Packaging Directive and the WEEE Directive stipulate, in essence, that the numerator should be materials that are deemed to be of use for recycling of good quality. However, detailed studies of the actual recycling rate based on output materials from recycling operations, as opposed to input materials, indicate that there is a significant difference. A study scrutinising the recycling of sveral material streams (paper, cardboard, aluminum, tinplate, glass and PET) in Switzerland reveals that the actual recycling rate of these material streams are all considerably lower than the current measurement done at the entry point to the recycling facilities (Haupt, Vadenbo & Hellweg, 2016). The timing of measurement is an issue of discussion for the last several decades and one of the reasons for casting doubts to the reliability and comparability of waste statistics. While it is on one hand good that the latest amendment of the Packaging Directive in 2018 provided a clearance in the timing of the measurement, one could doubt the meaningfulness of using the input materials for recycling rate calculation. Even when it is difficult to have a qualitative requirement on the output materials, it could have been possible to have at least a weight-based target measured at the end of the recycling process.

Difficulties of quantitatively compare material efficiency: One of the fundamental challenges related to target setting concerning material efficiency – be it upstreams or downstreams – is the very nature of material efficiency. Unlike energy efficiency which is relatively easy to define and measure, determing material efficiency of a product is much more complicated, when considering the different choices of materials and their combination, long supply chain and


31

arising impacts from various parts of their life cycle and the like. The difficulty of agreeing on specific design parameters to differentiate the fee paid by producers participating in a joint WEEE management system was among the fundamental underlying challenges of implementing IPR in practice (van Rossem, 2008; Kalimo et al, 2012). This also makes it challenging to set design requirements pertaining to material efficiency, and measure progress in relation to material efficiency. There has been progress in this area, however, which will be discussed further in Section 4.4.

Nature of design process: relating to the the measurability of material efficiency, the nature of design process makes it challenging to put specific requirements regarding design. Design process is inherently innovative, and it is in fact quite dangerous that government “dictates” design, especially considering the information asymmetry between the government and those engaged in the design process. While it is necessary and important to regulate what should not be inside of a product from health, safety and environmental and ethical perspective, which is reflected in substance restriction and laws related to, for instance, conflict minerals, due to existence of numerous materials, their properties (including unknown ones) and potential, it has been difficult to “define” what is preferable quantitatively from material efficiency perspective.

Not very ambitious targets: Lastly, the review of what has been (reported to be) achieved in light of what the law stipulates (see Section 3.3) show that the targets set in the existing laws have not been very ambitious in general. Some of them have been already achieved when the new targets are set. The results shown in Section 3.3 are the average of the EU-27/28 and Sweden, which is considered to be among the most progressive states with regard to environmental measures, and there are certainly other countries who have struggles in achieving the targets. However, as long as the legal basis of a directive is on environmental clause, it is possible for Member States to introduce more stringent targets than what is stipulated at the EU level (see Section 3.1.2). In addition to the interests of stakeholders, issues such as domestic law making process and (lack of) political wills could create hurdles in introducing something more ambitious at the national level.


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4 Analysis of selected potential policy measures Chapter 3 examines the exixting policy measures pertaining to closure of material loops and especficially to the increased use of recycled materials in production, and by doing so, seeking to identify what has not been working well and understand why. This Chapter seeks to assess selected (potential) policy measures which are deemed particularly useful in increasing the use of recycled materials in production, taking into consideration the challenges of the prevailing intervetions and their causes, as well as their potential and limitation inherent to the characteristics of the respective instruments.

In line with the assignment from the Mistra Closting the Loop research programme, the author assesses the five potential policy measures prioritised by the six research projects under the second phase of the research programme (see Chapter 1). In addition, reflecting upon the analysis of the existing policy measures, the author seeks to explore alternative ways of setting collection and recycling targets.

Each section starts with a concise description of the instrument, together with a brief background of the necessity for the instrument in question and its underlying logics, focusing primarily on how it intends to increase the use of recycled materials in production. It subsequently provides examples of existing experience when available, and discusses the prospect of introducing the policy measure in practice.

4.1 Quality standards for recycled materials The necessity of estabilishing quality standards for recycled materials have been raised as a condition to ensure reduce transaction costs for industry and thereby they could more easily uitlise recycled materials in their production (Tam, Kotrayothar & Loo, 2009; Tojo & Thidell, 2012; Finnveden et al., 2013; Milios, 2018). A study of how manufacturing industry seeks to balance the two policy demands of enhancing the use of recycled materials while ensuring no hazardous substances are included in the products revealed that manufacturing industries who are using recycled materials in their production use the same criteria when purchasing materials whether virgin or recycled (Tojo & Thidell, 2012). The same study also reveals that costs of verifying the quality was among the very important reasons why some of the manufacturing industries are hesitant to use recycled materials.

The most relevant existing policy in this regard is end-of-waste and by-product criteria (see Section 3.1.6). In its origin the primary issue was to release non-waste from being bound by waste-related laws and thereby reduce administrative and cost burden from industries involved. Since its initroduction in the revised Waste Framework Directive (2008/98/EC), however, the development of criteria at the EU level is limited to only three waste streams, and the uptake of the existing criteria in, among other countries, Sweden has been very limited (see Sections 3.1.6 and 3.3).

With the revived attention in the closure of material loops, however, the utilisation of end-of-waste criteria is also further sought through in order to provide certainties to the market and provide both virgin and recycled materials a level-playing field (Recital (17), Directive (EU) 2018/851; European Commission, 2018b). The latest end-of-waste criteria read as follows:

(a) the substance or object is to be used for specific purposes;

(b) a market or demand exists for such a substance or object;


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(c) the substance or object fulfils the technical requirements for the specific purposes

and meets the existing legislation and standards applicable to products; and

(d) the use of the substance or object will not lead to overall adverse environmental or

human health impacts (Article 6(1), Directive 2008/98/EC)

In addition to changing the text where Member State’s responsibility to take measures to ensure the application of end-of-waste criteria is clearly stated, the latest legal texts also clarifies that the EU-wide end-of-life criteria is to be done by monitoring the development at the Member States level and assessing the necessity (Article 6(2)). With the more “bottom-up” approach where Member States is expected to take more leading roles, as well as possibility of applying the criteria on case-by-case basis more easily (see Section 3.1.6), there appears to be more possibility for Sweden to explore possibility to utilise end-of-waste criteria.

Given that one of the reasons for the low uptake in Sweden was the already existing higher (business) standard within the country (See Section 3.3), being able to provide end-of-waste criteria more specific to a given context seems to provide more opportunities to utilise this policy instrument. Considering the reported hesitation of the Nordic countries to the introduction of end-of-waste criteria for construction and demolition waste at the EU level (Hjelmar et al., 2016), the enhanced possibitility of introducing nation-based end-of-waste criteria is presumably a welcoming approach for policy makers as well. It should be noted, however, that there are more sceptical views on further potentials of end-of-waste criteria, as found in, for instance, Johansson & Forsgren (2020).

A potential issue of contention is its relation to the chemical policy, as highlighted in the staff working document of the European Commission (2018). The working document contemplates on, among others, how to balance the legacy substances inside of recycled materials and requirement put on new materials.

4.2 Requirements for functional recycling Requirements for functional recycling is a policy measure suggested and prioritised by three of the six research projects to promote use of recycled materials in production. It relates mostly to recycling of metals. High-strength steels containing small amounts of metals such as niobium is needed, for instance, for car manufacturing. When cars are recycled these metallic parts typically end up in a common flow of other metals which goes into re-melting process and the small amounts of alloys disappear in the flow. This is an example of non-functional recycling. The idea is to instead treat the high-strength steels separately, so that they would not lose their properties after re-melting.

The author could not identify any existing policy measure that stipulates it as a requirement. However, there have been examples of industry practices. Toyota, for instance, developed its own thermoplastics (Toyota Super Olefin Polymer) in 1998 with the intention to recycle the materials for the same purpose (Toyota Motor Corporation, 2000). In the area of textile, Teijin, a manufacturer of synthetic fibires, established a recycling plant for the high-end polyester and established a worldwide network named “Ecocircle” to collect and recycle products made from the specific polyester so that they could be used for producing the same products (Teijin 2013). Development of measures as such would require a collection and recycling network that allows the streamlined treatment of the specific material.

In terms of potential policy measures, a concrete form of implementing individual financial responsibility in an EPR system would lead to the use of streamlined recycling system for specific materials. As it stands now, the discussion is mostly on how to differentiate the cost


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to reflect the actual cost of recycling, and it is happening to an extent in the areas of battery recycling in some countries where mercury-free batteries are sorted from mercucry-including batteries prior to recycling and producers pay differentiated fee (Tojo, 2004). As illustrated here, however, the existing example is mostly to do with the ease of recycling, and not so much for functional recycling. Moreover, cost differention in a collectively organised infrastructure has been a major challenge for other products such as EEE (see Section 3.4).

Another potential policy measure could be the use of end-of-waste criteria, as discussed under Section 4.1. Though it is essentially the measure to ensure the quality of a recycled material stream, including its environmental quality, and does not require recycling companies to conduct functional recycling per se, establishment of a criteria may facilitate implementation of functional recycling.

4.3 Minimum recycled material content standards Mandating the use of a certain amount of recycled materials in new products is perhaps one of the most straightforward policy measures to enhance the use of recycled materials in production.

Although the EU Circular Economy Action Plan from 2005 (COM(2015) 614 final) sounded hesitant concerning an introduction of a legislative measure requiring minium recycled material content, the EU introduced a minimum recycled content requirements for recycled plastics in 2019. Article 6(5) and Annex F of the Directive Directive (EU) 2019/904 requires that minimum 25% of the beverage containers made mainly of PET and sized up to 3 litres must be made of recycled plastics by 2025, and 30%, by 2030.

Legislative measures have been also introduced in, among others, some states in the United States. For insntace, in the State of California, the “Recycled-content Newsprint Program” and “Recycled-Content Trash Bag Program”, based on the Division 30 of Public Resource Code and Title 14 of the Californian Code of Regulations, mandate the inclusion of postconsumer waste paper in newsprint, and actual postconsumer materials in plastic trash bags. The former, started as early as 1991, defines “recycled-content newsprint” as newsprint comprises at least 40% of postconsumer waste paper fibre, and set the minimum usage requirement of the recycled-content newsprint on commercial printer and publisher in California (Section 42756, Public Resource Code). The usage requirement increased from 30% in 1994 to 50% in 2000, which is still effective until now (Section 42761, Public Resource Code). The program for plastic trash bags, started in 1998, mandates manufacruers of plastic trash bags, unless they meet the exemption criteria, to certify each year that either 1) at least 10 percent by weight of the regulated trash bags (at or above 0.70 mil) intended for sale in California is made of Actual Postconsumer Materials, or 2) 30 percent of the weight of the material used in all the plastic products intended for sale in California composes of actual postconsumer materials (Section 42291, Public Resource Code). The wholesalers of plastic trash bags must report on, among others, the manufacturers and other wholesalers from who they purchase the bags and the total number and tonneage of the bags sold (Section 17979.5, Californian Code of Regulations)

The deadline for Member States to transpose the recycled content requirements for PET bottles is 3 July 2021, and the effect of the implementation of the provision is still unknown. Given that bottle-to-bottle recycling of PET already exists in some places, such as PET bottles subject to deposit-refund system in Sweden (Tojo 2011), the target does not seem to be very ambitious. However, the introduction of the mandatory minimum recycled material content standards in itself is a big step forward.


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Regarding the two programmes in California, the 2009 compliance report, the latest one available from the programme website, reveals that overall 50% of the newsprint sold in California in 2009 was made of recycled-content newsprint, which uses at least 40% or more of waste paper fibre. The compliance rate of the year was 70%, which was reported to be lowest since the introduction of the programme in 1991. The report commented that the general decline of the consumption of printed materials might have negatively affect the availability of the waste paper fibre (California Department of Resources Recycling and Recovery, 2010). Concerning the trash bag program, the reporting is solely on the list of entities who are either compliant or non-compliant to the law. The reports from 2017 (California Department of Resources Recyclign and Recovery, 2017a, 2017b) include a substantial number of non-compliant entities, but no figure is available concerning how that may be translated into the use of virgin versus post consumer materials. The non-compliant entities will become ineligible to make any contract or subcontract with state entities, including their renewal and extension (Section 42297, Public Resource Code), which is deemed useful as a sanction.

Aside from legislative measures minimum recycled material content standards have been incorporated in different non-mandatory policy measures, such as public procurement, mandates for the public sectors to use recycled materials, ISO Type I Eco-labels, guideline, voluntary pleadge and the like.

Examples of public procurement include use of certain percentage of recycled materials (granulate) in road planning (Netherlands), requirements on municipalities to use recycled paper (Denmark), packaging, paper and furniture (EU) and the like (Alhola et al., 2017; Tojo, 2006). The Finnish Waste Law requires the authorities to use products made of recycled materials as much as possible (Section 4 Paragraph 3). Eco-label Type I programmes, such as Nordic Swan and German Blue Angel Programme, include containment of recycled materials as an award criterion for different types of products (Suikkanen & Nissinen, 2017; Tojo & Lindhqvist, 2010). The Waste Management and Recycling Guideline in Japan, developed by the Industrial Structure Council under the Ministry of Economy, Trade and Industry and covering 35 product groups, also set recycled material content as a criteria for some product groups. Examples include use of 62% of waste paper fibre in the manufacturing of paper for various purposes (e.g. newsprint, office paper, beverage and other packaging, carton boxes) and glass bottles containing 91% of recycled glass cullets, by 2011 (Industry Structure Council, 2007).21 Finally, the EU, as part of the strategy for plastics in a Circular Economy, has been calling for voluntary pledges to both sellers and buyers of recycled plastics (European Commission, 2018a). An assessment by the European Comission (2019) regarding its progress shows that, compared to the situation in 2016 (see Section 3.4), according to the pleadge expressed, the demand would increase by 60%, an increase from 3.9 million tonnes to 6.4 million tonners. The pledge from the supply side, meanwhile, exceeds 11 million tonnes (European Commission, 2019).

As seen, though limited to specific product/material streams, there has been various non-mandatory policy measures introduced at a national level that include minimum recycled mateirals content standard for many years. There are also some mandatory measures, though limited. As mentioned in Section 2.1, minimum recycled material content standards being a requirement related to the property of a product, it makes it very difficult to introduce it at the national level within the EU. Different standards used in different Member States would

21 Although the guideline was updated several times before, the author could not identify the version after the 2007 version.


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geopardise the smooth functioning of the internal market function (Recital (17), Directive (EU) 2019/904). However, any voluntary measures, such as green public procurement, eco-labels, guidelines and the like, could be introduced at the national level.

An issue prevailing to both mandatory and voluntary measures is verification of recycled materials, especially for products with long supply chain (Dalhammar et al., 2014). In the case of Recycled-content Newsprint Program in California, the verification is done by requiring the manufacturers of the recycled-content newsprint (not the publishers but the manufacturer of the paper itself) to certify, among others, “the metric tons of postconsumer waste paper and/or deinked pulp received or produced” at the manufacturer’s paper mills (Section 17962, California Code of Regulations). As for the trash bag program, the supplier of recycled plastic postconsumer material must provide to the manufacturers of regulated trash bag a statement on, among others, the quantity and proximate prior usage of the post consumer materials, as well as the actual content of the postconsumer materials (Section 17980, California Code of Regulations). In the case of the new EU law, Aritcle 6(4) of the Directive (EU) 2019/904 mandates the European Commission to adapt the rules for calculation and verification of the targets.

4.4 Clearer and expanded requirements for design for recycling Given the limitation of measures that can be taken at the end-of-life phase of products, changing the design of products so that they are “born-to-be” easy to take care of and low in its environmental impacts during the waste phase is considered to be among the most effective way of enhancing the closure of material loop. Among different end-of-life oriented design strategies, of focus in this study is design for recycling.

Many manufacturers have been incorporating end-of-life aspects of their products at the design phase of the products for the last several decades, triggered, at least in part, by the introduction of EPR programmes (Tojo, 2004). Regarding existing policy measures, however, as found in Sections 3.1.7 and 3.1.8, except for the restriction of the use of certain hazardous substances, mandatory and concrete policy measures mandating design for recycling hardly exists.

Meanwhile, there exist a number of guidelines published by different levels of government as well as entities such as industry associations and academia, providing guidance on the overall approach as well as concrete issues. In addition, with the enhanced political interests on circular economy, the better utilisation of the directive seeking to set eco-design requirements for energy-related products (ErP Directive, 2009/125/EC) (See Sections 3.1.8 and 4.3) is among the measures highlighted in the EU’s action plan for a Circular Economy (European Commission, 2015).

In the latest eco-design implementing regulation introduced in March 2019 for servers and data storage products, several aspects related to material efficiency – design for disassembly, firmware availability, inclusion of data deletion functionality as well as information provision requirement regarding critical raw materials – was included (Annex II, 1.2, 3.3). Though this particular regulation does not include anything specific to design for recycling, there has been a gradual development of methodologies for evaluating material efficiency from life cycle perspective, which can be translated into concrete material efficiency requirements in a law (Talens Peiróa et al., 2020).

As it is difficult to introduce laws requiring specific properties of a product at a Member State level (See Sections 2.1 and 4.3), mandatory design requirement is best to be introduced


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at the EU level. Meanwhile, similarly to the mandate on miniumn recycled material content, as already practiced, Member States could utilise different types of voluntary policy measures, such as green public procurement, ISO Type I Eco-labels, development of the guidelines and the like.

4.5 Material recycling combined with refundable security fee or penalty

This instrument, as suggested by research projects in the Mistra Closing the Loop research programme related to construction industry, concerns large construction projects in order to provide financial incentives for a contractor to recycle construction waste. It could be arranged in two different ways. One is through the advance payment of a security fee which can be returned if the waste is recycled – if the waste is not recycled, the fee is not returned. A variant of this system is to mandate the contractor to pay some penalty fee if the recycling is neglected.

(Potential) policy mesures resembling the first alternative

The author did not come across any examples where a policy instrument similar to the first variant has been implemented in the field of material efficiency. However, a study assessing the potential policy to enhance fibre-to-fibre recycling of textile examines a similar hypothetical policy, named refunded virgin payments (Elander et al., 2017). The policy got the ispiration from an actual policy intervention, a refunded emission payment programme for nitrogen oxides (NOx), which has been used to reduce NOx emission from large combustion plants.

In the case of NOx programme, the overall intervention logic starts where the respective combustion plants pay into the system based on the amount of NOx they emit. Second, the money paid into the system is evenly distributed among the participating combustion plants based on the quantity of the “product” they produce (energy output). The plants surpassing their peers in terms of reducing NOx emission per produced energy unit become net receivers of the refund, while the opposite become net payers. This gives economic incentives to reduce the production or use of the amount of undesirable substances. As an economic instrument, the fact that the money paid in is refunded to the addressees is considered to make the instrument less politically contentious. It also helps reduce the distortion of competition between companies covered by the instrument and those that are not (Höglund-Isaksson and Sternar, 2009).

According to an evaluation by the Swedish Environmental Protection Agency (Naturvårdsverket, 2012), since its introduction in Sweden in 1992, the emission of NOx per unit of energy produced continued to decrease, despite that the overall NOx emission increased since 2008 due primarily to the increase of overall energy use.

Extrapolating the idea of NOx programme, the envisioned refunded virgin payments systems requires producers putting new textile products on the market to pay the fee based on the total weight of virgin fibres included in the new products. The producers then receive refund based on the weight of total textile products they put on the market. Similarly to the NOx programme, the more a producer manages to replace virgin fibres with recycling fibres, the less (s)he needs to pay into the system, thus giving incentives to the producer to replace virgin fibres with recycled fibres (Elander et al., 2017).


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The idea proposed by the research projects related to construction industry is somewhat similar, though not very well articulated. It would not be difficult to change what is encouraged by the system from recycling of waste to actual use of recycled materials – rather, it is more straight-forward in terms of policy objective and which fills in what is missing in the existing policies, and perhaps easier to calculate. Instead of the “security fee”, the payment could be based on the weight of input virgin materials.

Similarly to the textile industry, there are a number of practical policy design issues that needs consideration, such as scope in terms of addresses, scope in terms of materials, level of benchmark (what is the appropriate level of recycled materials that the policy should aim for, and how should that change over time), level of virgin payment and the like (Elander et al., 2017). Other complexities added in the case of construction industry is the varying timing of the project duration, which makes it difficult to anticipate the incoming payment, based on which the amount of payment made in proportion to the total material used should be decided. Nevertheless, considering the very successful implementation of the NOx programme, it may be worthwhile exploring the idea more.

Policy measures resembling the second alternative

Concerning the second alternative of construction industry paying penalties when neglecting recycling, similar policy package considered could be a combined use of deposit-refund system and beverage container tax found in Norway. According to Numata (2016) and Anker Andersen A/S (n.d.), on top of the deposit refund systems, there is environmental fee to be paid for beverage containers. The fee is the same until the return rate for the containers does not exceed 25%, but after that, up to the return rate of 95%, the rate of environmental fee per container is proportionately reduced. Once the return rate exceeds 95%, no fee needs to be paid.

According to Infinitum (2018), the organisation which runs the deposit refund system in Norway, as of 2018, the return rate for cans was 87.3% and bottles, 88.6%. Considering the fact that Sweden, which has a deposit-refund system but no tax, managed to achieve comparable result – 87% for metal beverage packaging and 82% for PET bottles in 2016 (Viklund & Fråne, 2017) – a further study to examine the effect of the tax in further enhancing the collection rate might be needed. Perhaps the tax incentives works not necessarily to achieve higher collection rates, but rather to encourage fillers to join a deposit refund system. However, the way the interaction of the two instruments (tax differentiated based on the collection rate achievement) could still provide a useful insights when designing a differentiated “penalty” (by tax payment) system depending on the recycling achievement of a construction project.

4.6 Alternative collection and recycling targets and their measurement

As found in Sections 3.3 to 3.5, the existing sorting/collection requirements, as well as collection targets have managed to achieve the very first important step to facilitate the closure of material loop – sorting targeted waste streams from the rest of waste. The achieved recycling rates also indicate that, for some waste streams (e.g. packaging and EEE), a good portion of sorted waste stream have been brought to the recycling plants, and for others (e.g. batteries and cars), has led to progress in actual recycling. However, as exemplified with the case of plastics, the measurement in the existing policy measures does not fully capture what comes out of the recycling facility. Achievement of the targets does


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not indicate anything about the quality of the recycled materials, nor does it provide information on the fate of these products.

Based on the challenges regarding the target settings as well as their potential causes identified in this study, some alterantive ways of setting and measuring the collection and/or recycling targets are discussed.

Measuring output materials

The recycling activity becomes meaningful for closure of material loop only when they produce recycled materials that could be used in one way or another. From this perspective, the recycling rate, as it is currently measured for packaging and EEE in Europe, does not serve as a good indicator for the closure of material loop (Haupt, Vadenbo & Hellweg, 2016). It would be very good if the EU starts counting the recycling rate of packaging and EEE based on what is coming out of the recycling plants. It is happening for cars and batteries in Europe already, and systems in other countries, such as recycling rate counting of large WEEE in Japan (see futher discussion below). Considering that Article 11(4) of the WEEE Directive (2012/19/EU) already mandates Member States to keep track of the amount of output from the recycling facilities for the purpose of setting output-based counting of recycling targets (Article 11(6)), this has been already a policy consideration. From that perspective, it is somewhat puzzling that the latest amendment of the Packaging Directive (94/62/EC) from 2018 still set input materials as the basis of calculating the recycling targets.

Differentiating output materials based on their quality

From the pespecitve of establishing better indicators for closure of material loops, and especially in relation to use of recycled materials in production, another important issue is to monitor the “fate” of materials coming out of the recycling plants. In addition to changing the way of measuring, it would be meaningful to distinguish between at least recycling for the original purpose (closed-loop recycling) and for other purposes (open-loop recycling, or down-cycling). That way, in addition to understanding the flow from demand side, it would first become easier to grasp how much of the materials the respective society manage to bring back to the original purpose. In addition, by gradually increasing the targets for closed-loop recycling, the policy could truly push the closure of material loop. This requires additional thinking, such as definition of closed-loop and open-loop recycling – in order words, what is meant by “original purpose”.

Mandate selling potential as a criteria for “recycled materials”

Another way of enhancing the use of recycled materials is to include selling potential of recycled products as part of the criteria for “recycled materials.” Such requirement has been in place when the EPR system for large WEEE came into force in Japan in 2001. Under the Specified Home Appliance Recycling Law, recycling is determined as:

- The act (of producers) to separate components and materials from equipment that

becomes waste, and use them as components or raw materials of a product

themselves;

- The act (of producers) to separate components and materials from equipment that

becomes waste, and bring them to the state in which (these components and

materials) can be sold or given free-of-charge to the entities who will utilise them

(Article 2(1), translation by the author, content of parenthesis added by the author).


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As found in the definition, in order for products to be considered recycled, either the components and materials taken from the waste products are used by the producers themselves in their new products, or these processed components and materials should have some market value, or it is at least possible to give them away free of charge for the potential user.

An evaluation specific to this particular element of the Japanese law is not available, and there are many other features of the system which is quite different from a typical EPR system in Europe, including much more direct engagement of manufacturers in recycling activities (Tojo, 2004; OECD, 2016). However, it is very clear that manufacturers have been putting substantial efforts for the development of various recycling technologies (Tasaki, 2006).

Differentiated collection targets for respective categories of WEEE and battery stream

Finally, a weight-based collection targets irrespective of different types of waste stream within WEEE and waste batteries, respectively, have a clear problem in inducing the collection of heavier WEEE regardless of materials contained and their value in society. Similarly to packaging, it is desirable to set differentiated targets for respective WEEE and batteries.

Setting these targets are typically under the so-called environmental clause of the Treaty establishing the European Union (See Section 3.1.2). As the target setting discussed above are more stringent than what is currently stipulated at the EU level, it is theoretically possible to set these targets at a Member State level. However, except for the last item on differentiated collection targets, these changes are to do with how to count the collection/recycling rates. If the EU requires one way of counting and a Member State suggests the inclusion of another in their national jurisdiction, the Member State, in order to meet the reporting requirement by the EU, would most likely end up having to count different types of recycling rate. This would be very confusing for various actors involved, and most likely lead to some problems with reporting. Therefore it would be ideal if at least the first three changes can take place at the Member States level.

4.7 Summary of selected potential policy measures Table 7 summarises the findings from the analysis of the potential policy measures examined under Chapter 4, focusing primarily on the prospect of introducing them in reality, at which level of governance.


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Table 7: Potential policy measures to enhance the use of recycled materials, and prospect of their introduction in practice

Policy measures Prospects of introduction

Quality standards for recycled materials

There seems to be a wider opening for EU Member States to introduce end-of-waste criteria on their own on the streams not covered by the EU, including case-by-case introduction

Requirements for functional recycling

Rather slim possibility to introduce a direct policy measure. Perhaps utilise end-of-waste criteria or expect effects of EPR programmes

Minimum recycled material content standards

Mandatory – some examples from US States, and EU introduced one on plastics in 2019, which would pave the way for introduction for other material streams. Best to introduce at the EU level.

Voluntary – various opportunities for individual Member States to continue integrating as a criteria in, e.g. green public procurement, Type I eco-labels, guidelines

Clearer and expanded requirements for design for recycling

Mandatory – the first implementing regulation on ErP including material efficiency related criteria came into force, which would enhance possibilities for introducing other material efficiency related criteria for other products. Best to introduce at the EU level.

Voluntary – various opportunities for individual Member States to continue integrating as a criteria in, e.g. green public procurement, Type I eco-labels, guidelines

Material recycling combined with refundable security fee or penalty

At the moment hypothetical, but examples of policy measures based on the similar intervention logics exist. If there is political wills, it is possible to introduce at a Member State level, but requires some more thinking for concrete design details.

Alternative collection and recycling targets and their measurement

For changes related to what to measure, best to introduce at the Member States level.


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5 Conclusions and recommendations This study seeks to explore policy measures that would effectively enhance the use of recycled materials in production, an essential building block of the closure of material loop. The starting point was, despite the existence of a number of policy measures seeking to enhance the closure of material loops for the last several decades, the use of recycled materials in production is still quite limited. This was also one of the common findings of the six research projects under the Phase II of Mistra Closing the Loop research programme.

In order to explore potential policy measures, the study starts by analysing the prevailing existing policy measures pertaining to closure of material loops, focussing specifically on their effects on the increased use of recycled materials in production, as well as the underlying logics (intervention theories) that should lead the introduction of an intervention to the intended outcome. The study reveals that, while existing policies succeded in several aspects essential for the closure of material loops – e.g. sorting of specific waste streams from the rest of the waste stream, restricting the use of specified hazardous substances in products – they have failed in several other aspects also essential for the closure of material loops. Most notably, the assumption that source separation of specific waste streams would lead to the development of cleaner and recycled material stream does not hold for the considerable parts of sorted waste streams (theory failure). The study also highlights that design requirements related to closure of material loop is currently limited, and the uptake of so-called end-of-waste criteria, has been very low in most parts of Europe. It further points to the struggle of EPR-based approach in implementing individual producer responsibility, which in turn hinders the provision of incentives for producers to make upstream changes. The study identified several potential causes of these challenges in the way targets are set and measured, and point to some difficulties that arise due to the inherent nature of material efficiency and design.

The study subsequently examines six selected (potential) policy measures, from the perspective of their potential effectiveness in increasing the use of recycled materials in production. Five of them were prioritised by the reseach projects under the Phase II of the Mistra Closing the Loop research programme, and the sixth one is derived considering the challenges of the existing policy measures. In addition to the content of the policy mesures, their intervention logic and how they might rectify the challenges facing existing policy measures, experiences of introducing similar policy measures have been reviewed. Overall, there is a good prospect for introducing four out of six of the policy measures examined. Depending on the level of coerciveness as well as the current legislative setting, some of these measures could be easily introduced at the Member States level, while when it comes to mandatory requirements related to properties of a product, it is best to introduce at the EU level.

Based on these findings, the next section provides a concise set of recommendations for policy makers in Sweden. The chapter concludes with a few remarks on future research needs.

5.1 Recommendations for policy makers In reviewing the existing laws as well as assessing potential policy measures, the author wishes to propose the following recommendations tailoured for policy makers in Sweden.

Actively explore the use of end-of-waste criteria tailoured for the Swedish context and for the Swedish industry: At the moment, there seems to be an opening where Member States, instead of the


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European Commission, takes the lead in putting together end-of-waste criteria for waste streams whose end-of-waste criteria has not been set at the EU level. The waste streams whose end-of-waste criteria already exists at the EU level is currently limited to iron, steel and aluminium scraps, copper scraps and glass cullets. Meanwhile, the very low uptake of the existing end-of-waste certified metals by, among others, Swedish industry, as well as strong voices wanting to have quality criteria for recycled materials to reduce transaction cost, indicate that existence of end-of-waste criteria tailoured for specific context is needed and desired also by the industry actors.

Monitor output materials from recycling plants and their subsequent paths: In order to put an end to a situation where achievement of the recycling rates is “praised” even when the vast majority of recycled materials are not entering back to production process, it is crucial to monitor the materials coming out of the recycling plants more carefully. In addition to gaining deeper understanding of the quality of materials coming out of the plants, it is also important to know how these materials are used. These knowledge, in turn, would facilitate a better way of setting and measuring recycling targets, as suggested in Section 4.6 of the study.

Push the inclusion of material efficiency related criteria in the EU ErP Directive, while continue enhancing these criteria within domestic voluntary policy measures: With the first implementing regulation including material efficiency related criteria in place, there is a momentum to bring more material efficiency related requirements at the EU level. Considering the influence the EU law has within as well as outside the EU (Bradford, 2012), it would be very good to keep this momentum going. A Nordic study of using ErP Directive for non-energy-related products, such as textiles and furniture, put together a set of potential material efficiency related criteria for the two product groups (Bauer et al., 2018). Such studies might also trigger further discussion in Brussels. Meanwhile, the EU ErP Directive sets a minimum standard to cut laggards, and in order to pull the front runners we need other policy instruments such as ISO Type I and other endorsement eco-labels, Green public procurement and the like (Tojo & Lindhqvist, 2010). It would be therefore equally important to continue integrating stringent, material-oriented standards.

Continue supporting the further development of restriction of hazardous substances: Substance restriction is among the best working instruments related to closure of material loops, producing tangible effects. Closure of material loops that is safer for the environment is best achieved by eliminating the problems at source. While careful assessment is needed concerning what needs to be taken away from the material cycle, it is still essential to avoid unnecessary contamination whenever possible.

Explore concrete potentials for implementing individual producer responsibility: A crucial shortcoming in regard to laws based on extended producer responsibility, especially for complex products, is lack of concrete means to implement individual producer responsibility when the industry organise the management of end-of-life infrastructure jointly. While there are some hints found in some existing laws (Kalimo et al., 2012), further investigation in this issue is needed to utilise the full potential of the concept. It is certainly a step forward that a requirement of “fee modulation” is established at the EU level, but how it is going to be operationalised in practice requires careful follow up. It is especially so when Sweden is in the process of introducing an EPR system for textiles.

5.2 Suggestions for further research The current way of setting and measuring collection and recycling rate seems to indicate that there is a gap in knowledge regarding the quality of materials coming into and coming out of


44

recycling facilities. It would be of great use to look into industry practices regarding how the quality of input materials are specified, how they are evaluated, as well as that of output materials. In addition to enhancing the understanding on the flow of different material flows, from policy perspective, it would facilitate the establishment of recycling targets based on output materials, as well as that of end-of-waste standards. It might be useful to explore how the knowledge established in the field of, for instance, material flow analysis could be utilised in this area.


45

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