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An Operations Perspective on Product Take‐Back Legislation for E-Waste:

Theory, Practice and Research Needs

Atalay Atasu

College of Management, Georgia Institute of Technology, 800 West Peachtree Street, 30308 Atlanta, GA.

e-mail: atalay.atasu@mgt.gatech.edu, Phone: +1 404 894 4928, Fax: +1 404 894 6030

Luk N. Van Wassenhove

INSEAD Social Innovation Center, Boulevard de Constance, 77300, Fontainebleau, France

e-mail: luk.van-wassenhove@insead.edu, Phone: +33 160 72 4266

Abstract

A growing stream of environmental legislation enforces collection and recycling of used electrical and

electronics products. Based on our experiences with producers coping with e-waste legislation, we find

there is a strong need for research on the implications of such legislation from an operations perspective.

In particular, as a discipline at the interface of systems design and economic modeling, operations focused

research can be extremely useful in identifying appropriate e-waste take-back implementations for

different business environments and how producers should react to those.

Keywords: WEEE, Product Take-Back, Recycling, Electronics

Original Submission: June 2010, Revisions: February 2011 and May 2011, Accepted: May 2011.

1. INTRODUCTION

Increased consumption and associated waste generation is an important item on regulators’ agendas.

Many countries around the world, under pressure from environmental activists, have acknowledged the

waste generation problem and enacted legislation to deal with it. Product take-back legislation based on

Extended Producer Responsibility (EPR) (Lifset 1993, Lindhqvist 2000) is a popular type of legislation.

The basic idea behind EPR is to hold producers physically and financially responsible for the

environmental impact of their products after the end-of-life. This has been enacted for many industries,

from automotive to packaging, and batteries to electrical and electronic waste (e-waste).

The exponential increase in e-waste generation in major economies is a growing concern. According to

Greenpeace (2010) the average lifespan of computers in developed countries has dropped from six years

in 1997 to just two in 2005 and mobile phones have a lifecycle of less than two years. 183 million

computers were sold worldwide in 2004 (Greenpeace 2010), versus 281 million units in 2009 and 384

million projected by the end of 2014 (RM 2010). For mobile phones, these numbers are 674 million in

2004, versus 1.15 billion in 2009 and 1.3 billion projected in 2010 (Berridge 2010). If these trends

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continue, the generated e-waste will reach significantly higher volumes. The expected amount of

computers disposed of annually in landfills in the US is equal to a pile the size of a football field and a

mile high (TRC 2010). Accordingly, the e-waste problem is highly present on regulator agendas (OECD

2010, Grid-Arendal 2010) and take-back legislation mandating producer responsibility for collection and

recycling of e-waste has recently been a popular form of environmental legislation. The Waste Electrical

and Electronic Equipment (WEEE) Directive (Directive 2003/108/EC) in Europe and The Specified

Household Appliance Recycling (SHAR) Law enacted in Japan in 2001 (see Tojo 2004, 2006) are early

examples of such legislation. Since 2004, twenty-two states in the US (ETC 2010) also passed e-waste

bills, mandating producer responsibility. The majority of these programs started operating in 2009. The

remaining states are expected to take action soon, following global trends.

The objective of these laws is to lower the environmental impact by reducing the amount of waste sent to

landfills and to provide producers with incentives to design greener products (Lifset 1993, Mayers et al.

2005). However, they naturally cause economic concerns. Collection and processing (e.g., recycling) of e-

waste generally results in a net additional cost to many stakeholders, including producers, consumers and

local governments. Therefore, policy makers have to carefully consider the impacts of their e-waste

policy choices on the economic efficiency of production systems. The potential cost increase caused by

these laws can not only influence the competitiveness of an industry but also create uneven playing fields

between continents. There is no doubt that policy makers should be looking to identify policy guidelines

for desired environmental benefits while minimizing economic drawbacks of e-waste legislation.

From an academic point of view, an obvious location to look for policy guidelines is the environmental

economics literature. Research in this literature typically investigates the social welfare impact of policy

and identifies socially optimal policy instruments. In practice, however, the findings in this literature are

not necessarily applied. The environmental economics literature demonstrates that producer take-back

mandates may not achieve desired legislative outcomes such as providing incentives to design more

recyclable products (see Walls 2006 for further discussion). The WEEE Directive, however, takes the

producer take-back mandate approach, which requires European States to enforce producer responsibility

to meet regulatory targets on product collection, recycling and recovery, despite the fact that a clear

objective of the directive is to create design incentives for producers.

In this paper we argue that such deviations from theoretically optimal policy instruments can be driven

by the lobbying influence from stakeholders (particularly producers) and externalities such as the

challenges associated with implementing policy objectives. Hence, there is an important need to

understand the stakeholder perspectives on, and the economic impact of, different implementations of e-

waste policies. This objective requires an operations-based look at the challenges associated with e-waste

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law implementations, particularly because implementation related problems are mainly operational. We

posit this based on our experiences with the practice of the WEEE Directive in Europe, which we

accumulated through multiple workshops (see www.insead.edu/weee) and 6 years of intense

communication with industry. As such, this paper is a position paper that aims to (i) demonstrate the

shortcomings of a high-level policy perspective, (ii) build an operations framework to highlight the policy

implementation challenges, and (iii) call for operations management research on e-waste legislation.

Figure 1 illustrates our perspective. They key observation in this figure is the presence of the “Grey

Zone” i.e., the likelihood that policy objectives are translated into working systems with adverse effects

on producers and economies.

Figure 1: Research Objectives: Defining the Grey Zone

Based on our experiences with the practice of e-waste legislation in Europe, we posit that there are three

phases in realizing its economic impact. The first phase is the choice of an e-waste policy instrument as

discussed in the economics literature. The next two steps contain important micro-level (operational)

decisions: The second phase is the implementation of these policy instruments, while the third phase is the

producer response to policy and implementation choices, which can accentuate or attenuate the economic

implications of such legislation.

The implementation phase consists of two main steps. The first is translation of policy instruments into

working systems. Consider the WEEE Directive: While this directive mandates European States to ensure

a given collection and recycling rate target for each category of e-waste, the states are sovereign to

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translate these targets into national laws, i.e., the result counts, not how it is obtained. This flexibility

results in significant variations in national laws. The second step of implementation is creating

infrastructures and practical rules. Many questions need to be answered at this step. How and where

should e-waste be collected? Who should operate (i.e., collect and recycle) the take-back system? How

should the associated costs be financed? What is the involvement of municipalities? Where and how

should one set up the recycling market? How do these choices affect other stakeholders such as

consumers and the environment?

Finally, it is crucially important to understand how the chosen implementation affects producers’

operational decisions. Product design choices, competitive forward and reverse supply chain decisions,

planning of e-waste networks, technology and business model choices are among those that are very

likely to be affected by the take-back legislation. These decisions, affected by the regulators’

implementation choices, surely impact the economics of production systems and associated stakeholders.

Consequently, measuring the efficiency of such legislation, as well as its effect on social welfare, requires

a systematic investigation of the whole process. Understanding how take-back legislation is translated

into practical laws, how it is operated and how it affects producers’ operational decisions is crucially

important for coming up with recommendations to policy makers and economic stakeholders. A high

level economic analysis may not be able to answer these questions. A micro level analysis, however, can

not only help identify superior implementation choices, but also shed light on how existing and

potentially upcoming product take-back legislation affects operational decisions in production systems.

Given that operations management is a field at the intersection of systems design and economic analysis,

we believe that the right tools to investigate such systems can be found in our discipline. Therefore, many

pressing issues related to product take-back legislation can be answered by a systematic operational look.

Our discussion starts with a description of policy choices from an environmental economics perspective,

providing a summary of instruments used for policy making in the product take-back context. Next, we

discuss a number of early examples of e-waste laws to illustrate the importance of the implementation

phase and how it affects different stakeholders. Observing the complexity and variety of take-back

implementations in practice, we then discuss the economic implications of take-back laws from an

operations perspective. As such, we first highlight the importance of implementation choices from a

producer perspective and discuss how take-back legislation influences key operational decisions. Finally,

we consider a social welfare perspective, identify potential stakeholders and welfare objectives, and

determine the operational implementation decisions a policy maker has to consider. This discussion not

only helps us construct a big picture of legislated take-back economics that can be used to link research to

practice, but also identifies critical research questions to be answered using standard operations

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management techniques. We note that while we focus specifically on legislation for e-waste, our

discussion would also apply to other industries facing take-back legislation.

2. THE POLICY PERSPECTIVE

Environmental economists have long investigated the economic impact of regulation. In general, the

discussion in this literature focuses on environmental taxation and the impact of such distortionary taxes

on economic welfare (see Bovenberg and Goulder 2001 and Goulder and Parry 2008 for a detailed

summary). The general topic of interest in this stream is finding welfare maximizing policy instruments

and allocating costs and revenues of green taxes to different parties affected by such regulation. A more

specialized stream of work focuses on EPR to determine the policy models to obtain socially optimal

waste generation and disposal (Palmer and Walls 1997, Fullerton and Wu 1998, Palmer and Walls 1999,

Walls and Palmer 2000, Calcott and Walls 2000, 2002, Walls 2003, 2006). These studies aim to identify

policy instruments that maximize social welfare and product recyclability in the product take-back

context. For future reference, a number of relevant alternative policy tools considered in this literature can

be summarized as follows.

An advance recycling fee (ARF) is a fee collected from consumers (producers) at the time of sale, to

recycle the products they purchase (sell), e.g., as in California (CWIMB 2004) or Taiwan (Lee et al.

2000). A disposal fee model charges the end-user for the cost of recycling (e.g., as in Japanese SHAR

Law (Tojo 2004)). These fees can be used to build funds to undertake the recycling operations when end-

of-life products arrive at disposal streams such as municipal junkyards. The difference between the two

approaches is the timing of fee charge; with the former (latter), the fee is charged at the moment of

purchase (disposal). With a recycling subsidy, the recycling party, which can be the producer or a third

party, is paid a subsidy per recycled item by the government. This instrument needs funding from the

social planner, which makes it harder to implement. In a deposit-refund model, a tax on production

and/or consumption is associated with a subsidy proportional to product recycling, where the financing of

subsidies can be handled through the taxes collected. Note that the deposit-refund model is more general

than the simple advance recycling fee, which typically uses collected fees to finance a state-controlled

recycling system. In a deposit-refund model, independent third parties can also undertake the recycling

and receive the refund (e.g., a recycling subsidy). A recycling target is a standard recycling objective set

by the policy maker and can be defined as the proportion of products that need to be recycled.

Palmer, Sigman and Walls (1997) state an important policy result—a deposit-refund policy is the least

costly and most favorable option for reducing waste. Fullerton and Wu (1998) and Walls and Palmer

(2000) extend this result by considering environmental externalities in their models to discuss the

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efficiency of various policies (e.g., disposal fees, recycling subsidies, deposit-refund models) to determine

the socially optimum level of product recyclability. They conclude that different policies can maximize

welfare depending on objectives, market failures and ease of implementation. Calcott and Walls (2000,

2002) conclude that policies that regulate the disposal stage alone (e.g., disposal fees) do not help

encourage product recyclability when recycling markets are inefficient. They find that deposit-refund type

policies can not only improve economic welfare but also help obtain socially optimal recyclability levels.

For a detailed analysis of this literature, we refer the reader to Walls (2006), who provides a

comprehensive overview of the modeling approaches, and compares a number of policy tools from an

economic perspective. In sum, the conclusions from this literature are that (i) multiple policy instruments

are necessary to achieve multiple environmental objectives (e.g., increased landfill diversion and

incentives for recyclable product design), (ii) deposit-refund models can be more cost effective in

achieving these goals. However, although such high level approaches create useful analyses of economic

systems, they seem to stop short of addressing important practical issues that result in significant

variations between the theory and e-waste law implementations around the world. We illustrate these in

the next section using a number of examples.

3. E-WASTE LAW IMPLEMENTATION VARIATIONS IN PRACTICE

In this section, we highlight the presence and large extent of variations between the existing e-waste law

implementations in practice. We illustrate those and their impacts on different stakeholders using

examples from Europe, Japan and the US. The choice of these examples and associated discussions are

based on our experiences with producers from the electronics industry (www.insead.edu/weee). Our

objective in this section, however, is not to provide detailed information about the myriad of e-waste laws

in practice. Rather, we focus on certain important aspects of a number of early and influential examples of

e-waste laws to lay down the fundamental motivations for our framework. A detailed analysis of the

specifics of these e-waste laws is provided in Dempsey et al. (2010). We also note that while we base our

discussion on examples from European, Japanese and US e-waste laws, other countries in different parts

of the world have also enacted or are in the process of enacting similar laws. For instance, China has

recently approved a WEEE plan to be financed by producers and run by a governmental agency as of

2011, although exact details have not yet been specified (CEL 2010).

3.1 EUROPEAN UNION

The WEEE Directive (Directive 2003/108/EC) enforces producer responsibility for end-of-life electrical

and electronic waste in 27 European States for eleven product categories. Through this directive, member

states are currently required to make producers physically and financially responsible for meeting

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predetermined recycling targets for e-waste, and to ensure that at least 4 kg of e-waste is collected per

capita per year. These objectives, however, are subject to change. A recent revision proposal for the

WEEE Directive considers expanding the scope of products covered by the directive, mandating a more

stringent collection objective to be imposed on producers rather than the member states, and increasing

recycling targets (EC 2008). The details of the revision are expected to be finalized by early 2012.

What is interesting from the perspective of this paper is that the transposition of the WEEE Directive into

national laws and the associated national implementation choices significantly differ between member

states. Table 1 uses examples of Belgium, France, Ireland, UK and Germany to show a number of

implementation differences between these member states (see Tables 5 and 6 in Appendix, Dempsey et al.

2010 and Huisman et al. 2008 for a broader picture).

Belgium France, Ireland

UK, Germany

Collection method Municipalities x x x

Retailers x x x

Producers can own collection systems x

Management Single collective system x Multiple competing collective systems x x Individual producer operated systems allowed x

Who pays End user x x Producer x x x Recycling Costs Recycling Fees x x Costs split according to Market Share x

Table 1: Examples of Implementations in EU. (Excerpt from Tables 5 and 6 in Appendix.) The x’s

in the table stand for the presence of the options.

A major source of variation between the WEEE Directive implementations in Europe concerns the

concept of Individual Producer Responsibility (IPR). IPR is a policy principle based on the notion that

every producer should be responsible only for its own products. In practice, it is widely confused with

individual producer operated systems, i.e., where a producer collects and recycles its products

individually. The IPR idea however, is a principle of autonomy, which allows a producer to determine the

faith of its own products and incur costs associated only with those.

Whether IPR is achieved or not depends on how e-waste is managed. In practice, there exist collective

and individual producer operated e-waste management systems. In collective systems, products from a set

of producers are collected and recycled jointly. Collective systems can be monopolistic (e.g. state-

operated as in Belgium), or competitive (with multiple non-state operated systems as in France). The

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major drawbacks of these systems in Europe are that (i) producers are required to join the collective

systems, (ii) collective take-back costs are shared on the basis of producers’ market shares, no matter

what their actual product return volumes are, and (iii) they ignore recycling cost variations between

brands. Hence, these collective systems do not follow the IPR principle. This is because the faith of a

producer’s products cannot be determined by the producer, and the take-back cost a producer incurs

depends on other producers’ products as well.

Under individual producer operated systems, producers collect and recycle their own products only.

Hence, there is no doubt that individual producer responsibility can be achieved in such systems. The

problem, however, is that they may not be cost effective due to loss of scale economies, given that the

producer has to set up an individual logistics system to collect its products as well as facilities to recycle

them. To the best of our knowledge, there are currently no individual producer operated systems in

Europe. Yet, this does not mean that the individual producer responsibility principle cannot be achieved.

IPR can be achieved under collective systems, as long as producers have the freedom to operate their

systems independently (as in the UK and Germany) and they share the total system costs based on their

actual cost contributions (i.e., processing cost variations between brands are taken into account).

Achieving IPR in these collective systems, however, is challenging. This is because of the extremely

broad scope of the WEEE directive. It applies to thousands of product types (e.g., small household

appliances), with tens of thousands of individual models belonging to hundreds of brands. One could take

a conceptual (some would say naive) approach and assume that with the help of technology (e.g., RFID),

this myriad of products could be identified by brand and model, and their processing costs could therefore

be tracked and assigned to the respective producers under collective systems to exercise the IPR principle.

Or one could assume that every producer collects and recycles its own products (e.g., individually

operated systems). The reality on the other hand is very different. Individually operated systems can be

prohibitively expensive as discussed above. Similarly, if e-waste collection is mixed (e.g., electric

toothbrushes are collected and recycled with cell phones) brand sorting is very expensive. To-date, few

electrical or electronic products in the e-waste streams contain RFID tags, and even if they did, the

technology is not accurate enough. Supposing RFID tags would be placed on newly sold products, many

of those would enter in waste streams in 8-10 years (e.g., washing machines) and some may return even

after 20 years (e.g. refrigerators). Meanwhile, manual sorting is too expensive in most instances.

Therefore, in the practice of WEEE, collected e-waste is often a mixed bag of product types, brands and

models. While some crude sorting according to broad recycling categories (such as the eleven categories

in the Directive) is done, recycling also takes mixed e-waste as input, and an average cost is charged per

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unit weight. Given the above realities, how can one make a producer responsible for its own product only

and expect the producer to design products for recycling?

Another source of variation regards the nature of collection operations. While the UK and Germany allow

individual producers to develop their own collection systems, this is not allowed in Belgium and France,

where producers are obliged to use municipal collection points or retailers for the collection of e-waste.

Although using municipalities seems to be an easy and efficient approach, misaligned incentives may

make life harder for producers. We witnessed several discussions where producers complained about

municipalities seeing e-waste collection as a revenue generator, charging producers excessive fees for

access to their waste.

Finally, consumers may also be affected by implementation choices. Although the WEEE Directive

recommends that e-waste be recycled at no charge to customers, some countries (e.g., Belgium) can

charge consumers recycling fees at the moment of purchase. Whether the use of such fees will be allowed

in the future is part of the current debate regarding the revision of the WEEE Directive (EC 2008).

3.2 JAPAN

Two Japanese directives regulate the recycling of household appliances (e.g., TV sets, cooling devices,

washing machines and air conditioners) and computers, respectively. The Specified Household Appliance

Recycling (SHAR) Law assures that end-users are charged an end-of-life management fee by the

producer upon disposal, while the computer recycling law holds producers responsible for recycling (Tojo

2006, Dempsey et al 2010).

Japan (SHAR)

Japan (PC)

Collection method Retailers x Postal system x

Management Multiple competing collective systems x

Individual producer operated systems allowed x

Recycling Producers own recycling facility x x

Who pays End user x

Producer x x Financing End-of-Life Consumer Fee x Split according to actual recycling cost x x

Table 2: Implementations in Japan. (Excerpt from Tables 5 and 6 in Appendix). The x’s in the table

stand for the presence of the options.

Table 2 summarizes some specifications of the Japanese models and illustrates sharp contrasts with the

European models. First, Japanese policy makers use retailers for collection of appliances and the postal

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network for computers. This simple differentiation has a purpose since smaller items like computer

hardware are easy to transport by postal networks, while for bulky products, retailer collection is more

convenient to consumers than going to municipal collection points or to producers directly. Second, the

Japanese model holds both producers and consumers responsible for the costs of appliance recycling,

while operational responsibility is on producers. Finally, and most importantly, the Japanese model is an

example of individual producer responsibility based collective systems, where a number of Japanese

appliance producers only collect and recycle their own used products. These systems are capable of

distinguishing brands and properties of products through a barcode system, allowing the operators to

identify the producer of each product, applicable collection points and recycling plants according to the

brand and category of the products. Such a system has been reported to create incentives for greener

designs (Tojo 2004, 2006) because of extensive producer control on the recycling operations. The system

also guarantees fair cost allocation between producers (when they collaborate) since the recycling costs of

products are differentiable and can be correctly assigned to the right producer. Further operational details

of these systems are provided in Dempsey et al. (2010).

It is important to realize that the scope of Japanese laws is different from the WEEE Directive. While the

Japanese laws focus on a relatively narrow class of products (e.g., washers or PCs), the WEEE Directive

covers eleven broad categories (from TVs to fridges, from vending machines to lighting equipment). The

narrower scope of the Japanese laws allows product separation (e.g., brand sorting) at much higher cost

efficiency and allows exercising the IPR principle more effectively.

3.3 UNITED STATES

In the US, there is no unified federal level e-waste legislation, but twenty-three individual states have

adopted e-waste laws, 22 of which are based on producer responsibility (see ETC 2010 for details). The

content and implementation of these laws vary significantly between states, similar to Europe. We

illustrate these using the examples of California, Washington State and Maine.

WA ME CA

Who pays Consumers x

Producers x x

Collection method Municipalities x x x

Retailers x x Producers own take back systems x x Management Multiple competing collective systems x x x

Individual producer operated systems allowed x x

Recycling Producers own recycling facility x x

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State operated x x x Table 3: Examples from the US. (Excerpt from Tables 5 and 6 in the Appendix). The x’s in the table

stand for the presence of the options in the state e-waste laws.

California was the first to establish an e-waste recycling program (CIWMB 2004). Consumers are

charged an advance recycling fee for the purchase of a product that contains a screen. The fee applies to

all transactions to which the California sales tax applies, including leases and to Internet and catalog sales

for purchasers who take possession in California. These fees are then used to finance collection and

recycling operated by a state-controlled system (i.e., the California model is an example of the advance

recycling fee model defined in Section 2). Washington State on the other hand, requires producers to

participate in an approved recycling plan (DOEW 2009). Producers may join a collective system called

the standard plan operated by a state-controlled authority (see wmmfa.net). They can also operate

individual producer responsibility based systems (individually or collectively in collaboration with other

producers), as long as their plan conforms to the standards in the legislation. However, no individual

producer responsibility based systems are operational to-date (Jackson 2010). Cost allocation for the

current collective system is based on return shares of the producers, which are calculated by the

Department of Ecology (DE) in collaboration with the National Center for Electronics Recycling

(NCER). Finally, in Maine, the collection task is assigned to municipalities, who then pass the waste to

one of seven previously assigned consolidators (DEPM 2009). Two options are allowed for producers:

They can collect a proportion of waste (based on their return share) and recycle it, or they can have a

consolidator recycle their share, and pay for it.

Table 3 illustrates that the US laws are quite different from the examples in the EU and Japan. First, the

scope of these laws is even narrower than those in EU and Japan; they typically focus on TVs, monitors,

and IT products. Compared to the examples of Europe and Japan, US take-back laws seem to have higher

flexibility. A dominant policy choice in the US appears to be mandating producer responsibility (with the

exception of California.). Although producers are technically allowed to operate individual producer

operated systems, collective systems are common in the US, where producers pay average collection and

recycling costs per volume of e-waste to a state-operated plan. An important difference is that producer

cost sharing based on return shares is more common in the US, while in Europe market share based cost

allocation models are favored. Unlike market share models, producers under return share models do not

pay proportional to their sales volume but rather proportional to their collected product volumes. This cost

allocation differentiation is made on the basis of sampling to identify individual producers’ waste

ownership in Washington, whereas in Maine a count of producer brands in the waste streams is used to

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determine return shares. While the choice between return share and market share based cost allocations

seems to be trivial, this is one of the most crucial issues in practice. Companies with lower return volumes

do not want to share costs with companies that have higher return volumes.

To see why, assume two companies with different characteristics as in Table 4. Company A has low sales

volume, but high return volume. The other company (B) has higher sales volume, but lower returns

volume. Under the market share based collective system mandated by some European Countries,

company B indeed subsidizes recovery costs of company A. The return share based cost allocation model

improves fairness of cost sharing under collective systems.

Company Sales Volume

Return Volume

Cost under Market Share

Cost under Return Share

A 100 75 $50 $75 B 200 75 $100 $75

Table 4: Return Share versus Market Share (assuming a unit take back cost of $1)

All these examples cited above show that there are additional complexities embedded in EPR legislation

beyond the policy instrument choice. A high level economic analysis of policy choices is not sufficient to

identify the impacts of such implementation decisions. While similar tools may be used for policy-

making, such as recycling targets, disposal fees or advance recycling fees, the implementations in

different countries or states vary significantly and these variances heavily impact the behaviors of

stakeholders. As one would expect, implementation related differences may lead to different outcomes,

cause disturbance in competition and create fairness concerns. Our experiences with managers from

companies such as HP, Nokia, Samsung, Electrolux and Sony suggest this is very much the case.

4. AN OPERATIONS PERSPECTIVE

The examples in the previous section highlight the importance of a detailed operational look at the take-

back problem. As illustrated by Figure 1, the two phases beyond the policy choice –implementation and

associated producer responses – are critical. Hence, their economic implications should be investigated. In

this context, there are two relevant big-picture questions from an operations perspective:

1- Given the variety of existing take-back implementations, how should manufacturers react to a set

of take-back rules, i.e., policy instrument and implementation choices?

2- How should a social planner design take-back rules by anticipating manufacturer responses?

In what follows, we provide a framework that can help answer these questions by laying down the

fundamental trade-offs in the take-back legislation context. Rather than providing a generic economic

modeling perspective, we focus on the current practice of take-back legislation based on our interactions

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with producers. This allows us to highlight the practically relevant problems that require a systematic

operational analysis and the meaningful questions underlying modeling exercises.

4.1 ANTICIPATING PRODUCER RESPONSES

Since operational efficiency of production systems is at the core of operations management research,

understanding how product take-back laws affect producer behavior is a key research objective.

Answering this question would not only provide a roadmap for industries but also help policy-makers

anticipate manufacturer responses and their impact on welfare. Most decisions producers face evolve

around achieving take-back cost efficiency through product design, network design, technology, and

business model choices. Below, we discuss these issues and highlight important research problems that

can help e-waste laws achieve desired outcomes such as green design incentives and economic efficiency.

4.1.1 Network Design. Producers may have to set up infrastructures and design collection and

recycling networks to comply with e-waste laws. They have to make location decisions in addition to

managing waste flows to achieve desired policy targets (e.g., collection and recycling targets). This is

indeed a traditional operations management problem, but it faces a new set of assumptions. The e-waste

network design problem focuses on reverse flows and aims to minimize compliance costs under non-

traditional constraints. These constraints include several restrictions such as collection point limitations

(DOEW 2009), recycling technology standards, landfill bans (ETC 2010), waste export restrictions such

as the Basel Convention (see ban.org), collection and recycling targets, and other regulatory

implementation choices discussed previously. The impacts of these non-traditional constraints need to be

discussed in relation to EPR legislation.

Identifying conditions for efficient collection and recycling infrastructures is an important research

direction. For instance, the European Recycling Platform (ERP) is an e-waste system operator (practically

referred to as a producer responsibility organization-PRO) founded by manufacturers (HP, Sony, P&G

and Electrolux) as a response to monopolistic collective systems in Europe and has been able to reduce

average take-back costs significantly (Guilcher 2005). Given that similar alliances are not yet available in

the US, it would be interesting to know under what conditions such producer-operated collective systems

arise. It is also important to understand how such systems can be operationalized, e.g., how ERP-like

alliances should choose collection and recycling locations and determine transportation of waste flows

from several producers between states. The allocation of costs between producers in such a system is also

a critical problem to be solved, which requires collaborative game theory applied in the e-waste recycling

network framework. Finally, forward supply chain networks can sometimes be used for reverse flows as

well. While this problem has been considered for a setting where the producer transports used products

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for profit making (Fleischmann et al. 2001), the case of costly recycling enforced by take-back laws

remains an open problem to be investigated.

4.1.2 Product Design: Product design implications of different take-back implementations are yet to

be discovered. It has been argued (Lifset 1993, Lifset and Lindhqvist 2008) that take-back laws would

create green product design incentives. Yet, with the exception of the recyclability improvements reported

in Japan (Tojo 2004) and TV redesigns in Europe (Huisman 2006) for recycling cost reductions, very few

studies have investigated the design implications of take-back laws. The impact of different take-back law

implementations on producer design choices is one of the most interesting research problems.

Take-back laws can influence the design for reuse, recyclability and durability. Design for reuse is

encouraged if take-back laws credit product or component reuse. To-date, most take-back laws do not

promote product reuse since the focus is on product recycling for e-waste recovery. Design for

recyclability is inherent to take-back laws, because this can achieve recycling cost reductions for

producers. However, the implementation structure can improve or deteriorate these incentives. For

instance, market or return share based cost sharing under collective systems can indeed reduce design

incentives, because these cost allocation heuristics do not consider the recyclability of products. Yet,

producers argue that return share based cost allocation mechanisms can provide better incentives for

durable designs as opposed to market share based cost allocation, because durability investments can

potentially extend the life cycle of products and reduce product returns. Indeed, Plambeck and Wang

(2009) have shown that producers can find incentives to extend the duration of consumer use of

electronics depending on the choice of policy instrument. The individual producer responsibility based

systems on the other hand can create better incentives for more recyclable product designs because they

allow producers to directly benefit from their own recyclability investments (see Section 3.2). All these

examples point to one critical observation: A general investigation of the implementation choices on

product designs remains an important problem which can be researched using traditional product

development or product line choice models.

4.1.3 Closing the Loop: Product take-back legislation has a direct connection to closed-loop

production systems. As such, mandated take-back and for profit take-back and reuse (e.g.,

remanufacturing) are likely to be confused, although most take-back legislation focus solely on material

recycling or energy recovery. With the exception of the UK, who considers remanufacturing as an

acceptable form of mandated take-back in the scope of the WEEE Directive revision (Calliafas 2010), no

other e-waste law implementation considers remanufacturing as a preferred form of handling product

take-back. In fact, we observe the opposite in practice, where product remanufacturing can result in

additional product take-back costs to producers. Consider the market share based cost sharing in several

15 

 

European countries. This heuristic determines product take-back financing obligations of producers based

on the number of products they put on the market. Thus, depending on the accounting of market shares,

selling a product once as new and once as remanufactured may double the product take-back costs these

producers incur. Although they generate one unit of waste, they may have to pay for its recycling twice,

because remanufacturing is not a well-defined form of product take-back in the WEEE Directive.

Consider another scenario with advance recycling fees such as those in California. Assume the producer

(or buyer) is charged a fee at the moment of purchase and this fee is collected by the regulatory body. If

the producer actually recovers the product at the end-of-life and remanufactures it for resale in another

country, the funds generated for recycling are never used for that purpose and the producer may face

similar obligations in the country of export.

4.1.4 Technology and Business Model Choice: Research has demonstrated that technology choices of

producers and their investments (Gray and Shadbegian 1998, Ovchinikov and Krass 2009) are

significantly affected by environmental laws and alternative policy implementations have different

impacts on technology choice (Jaffe and Stavins 1995). Similarly, alternative implementations of take-

back laws are likely to have different effects on how producers choose production technologies. For

instance, take-back cost sharing, or collection and recycling target choices can influence production or

recycling technology choices. Thus, the impact of different implementation choices on producers’

technology decisions is an important problem to be considered, both empirically and analytically.

Producers’ business models can also be affected by take-back laws. Take a producer whose business

model includes remanufacturing durable products (e.g., Kodak or Xerox). Consider a scenario where this

producer is required to join a monopolistic collective system, which mixes used products from a variety

of producers and recycles them for material recovery. This means the producer pays the recycling costs

for valuable products that could have been remanufactured and sold for profit. Is it possible to continue

remanufacturing when it competes with recycling obligations? Similarly, leasing as a business model can

suffer from such practices, where take-back costs enforced by legislation can affect producer choices with

respect to leasing versus selling.

4.2 THE SOCIAL PERSPECTIVE

Given that the policy and implementation choices affect producer responses, a policy maker needs to

determine the set of rules that maximize its objective, i.e., social welfare. The first step is to determine

which stakeholders are affected by social planner choices and how. Once this is clearly understood, the

right policy instruments and implementation choices can be determined.

4.2.1 Moderating Factors:

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The typical stakeholders in this context include producers, consumers, take-back operators (e.g., waste

collectors, transporters and recyclers), local governments, and the environment. We posit that six key

factors moderating these stakeholders’ individual objectives need to be taken into account.

4.2.1.1. Cost structure of take-back operations. Cost efficiency in product take-back is perhaps the most

important factor from all stakeholder perspectives. Higher take-back costs reduce profitability for

producers, and may result in higher prices being reflected on consumers. Cost efficiency would also

benefit take-back operators (e.g., collectors, transporters and recyclers) by increasing their margins.

Therefore it is important to understand how different implementation structures affect the cost efficiency

of product take-back. For instance, although municipal collection may eliminate the fixed costs of

building a collection infrastructure, appears to be very cost efficient and generates local revenue for

municipalities, it works against producers who want to separate their own brand products and results in

additional separation costs for these manufacturers. Municipalities may not appreciate manufacturer-

owned collection points since e-waste collection can be a source of extra revenue. This tension is readily

observable in practice (Guilcher 2005): A major European electronics producer has complained in a

meeting that a French municipality had once priced e-waste at 700 Euros per ton, since they knew the

producer had an obligation to take the e-waste from that specific location. Such opportunistic behavior

aiming at municipal revenue generation can distort the potential cost efficiency of product take-back from

producer perspectives. In other words, stakeholder preferences may not be aligned and an efficient system

for one stakeholder may not be preferred by the others. Identifying the cost structures of existing take-

back implementations and investigating their impact on stakeholders is a critical question to be answered.

4.2.1.2 Environmental Impact of Take-Back. The environmental hazard level of a product should

determine policy objectives (e.g., collection and recycling targets in the WEEE Directive) and

implementation choices. With higher product hazard levels, collection and recycling targets need to be

stricter and collecting environmentally hazardous products at municipal junkyards may not necessarily be

the best channel. Products with higher environmental impacts are also likely to cost more to collect and

recycle. Thus, there is an inherent conflict between cost efficiency and environmental impact. Balancing

the economic and environmental impacts is a challenging task and requires a precise measurement of

environmental benefits from recycling and its value to the society. To the best of our knowledge, this is

still an open problem in the context of product take-back legislation. The value of diverting a product

from landfill through recycling needs to be measured to balance the economic and environmental impacts

of product take-back.

4.2.1.3 Local Perspectives. While take-back typically results in a net cost to producers or consumers, it

can be a value generating activity from a local perspective (see section 4.2.1.1). It creates additional jobs

17 

 

(e.g., collection sites, local transportation and recycling workforce) and is therefore often popular with

local authorities. However, most stylized models of product take-back legislation focus on producer and

consumer surplus only (e.g., Atasu et al. 2009, Plambeck and Wang 2009). The economic

benefits/drawbacks of take-back on municipal revenue generation remain an open problem to be explored.

4.2.1.4 Competitive externalities. Take-back legislation can create fairness concerns in certain markets

(as illustrated by Table 4) and distort competition. While collective systems seem to be cost efficient, they

may allow some producers to free-ride on take-back costs when they are not based on individual producer

responsibility. To producers, such fairness concerns could outweigh cost efficiency considerations and

individual producer operated systems may be preferred. Thus, developing an understanding of how

alternative take-back implementations affect competition is important.

4.2.1.5 Monitoring and Controlling Costs. In addition to stakeholder perspectives, regulators need to

take into account the additional social cost of monitoring and controlling take-back systems. These costs

may be shifted to producers but their extent will depend on the implementation. For instance, a state-

operated collective system (e.g., with advance recycling or disposal fees) may incur lower monitoring

costs, because the regulator undertakes take-back operations and does not need to monitor producers or

control their systems, while collection and recycling mandates imposed on producers as in the WEEE

Directive do imply additional monitoring costs. The number of producers and variety of product types

may also affect costs of monitoring producer run systems. This is an important factor to be considered

while choosing the take-back implementation. The need for empirical research to determine the impact of

monitoring and control costs is evident.

4.2.1.6 Dynamics. Finally, dynamics of business and legislative environments should be considered

when designing and implementing take-back legislation. Political status and economic parameters such

as resource prices are inherently dynamic. Consider the situation in the US, for instance. While a number

of states have e-waste systems in place, a future federal law can change the objectives of existing laws,

and hence their implementations. The competitive landscape can also change over time. Producers

currently active in a market may disappear in the future. Given that product sold today will be arriving in

waste streams in the future, it is important to take the potential absence of producers from future markets

into account. To-date, financing of orphan products through financial guarantees is a big debate both in

Europe and in the US (Dempsey et al. 2010), which is mainly caused by the anticipation of competitive

dynamics in the electronics industry. Similarly, although it has not been finalized to-date, the potential

revision of the WEEE Directive (EC 2008) considers increasing collection and recycling rates to higher

levels and expanding the scope of product categories to be covered. Hence, research on the economic

18 

 

impact of take-back legislation also needs to consider the dynamics in competition, legislative processes

and policy objectives.

4.2.2 Implementation Decisions

Once the stakeholder perspectives and moderating factors are clearly understood in a given business

environment, the criterion for policy and implementation choices can be determined. Based on our

experiences with the practice of product take-back in Europe, we posit that take-back implementation

requires decisions on eight dimensions.

4.2.2.1 What is the policy instrument? As discussed in Section 2, several take-back policy instruments

exist. Collection/recycling rate targets, advance recycling fees or disposal fees are the commonly used

instruments in practice but differ substantially on how they affect implementation choices. For instance,

collection/recycling rate mandates require producer operational responsibility, while advance recycling

fees or disposal fees can be used in a state-operated system. This choice may affect the cost structure of

take-back operations (as discussed in section 4.2.1.1) and monitoring requirements (as discussed in

section 4.2.1.5). At the same time, while advance recycling fees are charged for each sold product,

disposal fees are charged for each collected product. If sales volumes do not match collection volumes,

this practically means that the former model creates a higher burden on consumers and producers. Indeed,

some producers highly criticize advance recycling fees, because governments can potentially use them for

other purposes than e-waste recycling (HP 2010). This is one of the major concerns for the upcoming

Chinese WEEE Directive. Disposals fees also have an important disadvantage. End-users are discouraged

to return used products to appropriate collection points, because they have to pay recycling fees. Tojo

(2006) reports that this is the case in Japan, where a significant proportion of used household equipment

ends up in illegal dumping.

For a broad discussion of policy instrument choice in the take-back context, we refer the reader to Walls

(2006). At the same time, we reiterate that the policy instrument choice coupled with other

implementation decisions (discussed below) can significantly affect the benefits and drawbacks of

product take-back for stakeholders. Understanding the social welfare implications of these different

policy instruments, while taking implementation related externalities into account is therefore an

important research topic.

4.2.2.2 Which waste management model? As discussed in section 3.1, e-waste can be managed by:

collective (monopolistic or competitive) or individual producer operated systems. A monopolistic

collective system (as in Belgium), appears to provide the highest possible scale economies. However,

producers argue that competitive collective systems (such as those in France and Ireland) can further

19 

 

reduce recycling costs, because the competitive pressure creates better incentives to reduce costs to attract

more business. At the same time, individual producer responsibility based systems (be it collective or

individual producer operated) can improve product designs further and result in higher cost effectiveness

(see sections 3.2 and 4.1.2). Empirical and analytical research comparing the cost efficiency of these

waste management models, as well as their impact on stakeholders, can help significantly improve our

understanding of efficient take-back legislation.

4.2.2.3 How to collect? The choice of the collection network is another important decision for the

implementation of take-back systems. Practical examples suggest using municipalities, producer owned

collection centers, retailers, postal networks or local consolidators as collection centers (see section 3.2

for details in the context of Japanese laws). Using municipal collection points can be very cost efficient in

principle. However, our observations in practice also suggest that municipalities can abuse this to

generate income and provide employment (as discussed in section 4.2.1.1). A data driven comparison of

these collection options in different business environments is an interesting research topic. In particular, a

cost efficiency comparison between existing collection infrastructures in Europe, Japan and the US would

help better understand the benefits of each option and provide valuable insights for policy-making.

4.2.2.4 Operational Responsibility and Infrastructure. Who should undertake collection and recycling

operations: regulators or producers? Producers highly criticize state-run programs, because of cost

inefficiency and monopolistic pricing strategies (e.g., the ERP example in section 4.1.1). On the other

hand, government run collective systems can eliminate the disadvantage of the monitoring and controlling

costs of producer run programs (as discussed in section 4.1.2.5) and possibly provide scale economies to

reduce recycling costs. The cost of building an infrastructure for a producer operated system is also an

important consideration (as discussed in section 4.2.1.1). Understanding the externalities that drive the

efficiency of these two approaches is another important research avenue.

4.2.2.5 Who should have the financial obligation: the end-user, the buyer or the producer? In practice,

we observe that all three models are being used. California charges an advance recycling fee at the

moment of purchase to the buyer, the Japanese SHAR Law charges a disposal fee to the end user at the

moment of disposal, while most remaining laws hold producers responsible for the financing of e-waste

systems. An important research question is how the financing choice affects different stakeholders.

Stylized economic models (e.g., Atasu et al 2009) would anticipate that producers always find it

profitable to share these costs with consumers. No matter who is held responsible for the financing of e-

waste, these costs will somehow be reflected both on producers and consumers. More specifically, with

price sensitive linear demand models, the producer would reflect part of his recycling costs to the

consumers. At the same time the producer would take over part of the consumers’ recycling costs

20 

 

(through price reduction) when the consumer is responsible. Thus, this financial obligation choice may

not have a significant effect on welfare in such stylized economic models. Reality on the other hand can

be different. Consumers may not associate recycling fees with the price of the product and recycling fees

may not have an impact on their purchasing behavior. From a practical point of view, this suggests that

producers may not really share consumers’ recycling costs when they make their pricing decisions.

Furthermore, when consumers are responsible for financing e-waste systems, producers may not find

incentives to improve the greenness of their products or make their recycling easier. While this seems to

be a minor issue in stylized analytical models, it may be a concern with respect to the economic efficiency

of take-back legislation and remains a question to be answered. The consumer perception of recycling

fees and resulting producer behavior require a detailed empirical investigation.

4.2.2.6 Cost Sharing. A major problem in collective systems is the allocation of costs between

producers, particularly when waste streams are mixed (e.g., as in Europe, see section 3.1 for details). To-

date, no cost sharing mechanism guarantees fair cost sharing between producers in collective e-waste

systems and this remains one of the most important problems to be solved for the efficient

implementation of e-waste laws. In practice, two heuristics are used: market share or return share based

cost allocation (see section 3.3 for details). Market share appears to be favored due to its simplicity. It is

based on the assumption that all products will eventually end up in waste streams and the only data

required to allocate costs between producers is their market shares. However, producers argue that market

share models force unfair cost sharing between producers because return rates differ significantly by

producer and not all products end-up in waste streams.

Although return share is believed to be fairer, it too is criticized for not differentiating recycling costs

between products. Such differentiation can however be extremely challenging in practice. It can be

achieved at alternative levels with varying costs associated with the level of differentiation. The first level

is recognizing individual products, e.g., model x of brand y from producer z, with its specific design

characteristics and associated collection and recycling costs, which heavily depend on the recycling

technology used. One can imagine that this level of differentiation can be extremely expensive to achieve.

The second level is to define categories of product types like mobile phones, i.e., a mix of models and

brands, already containing thousands of products with potentially very different design characteristics and

recycling costs. The third level is to define product categories, e.g., small appliances, which can obviously

contain largely different product types and recycling costs, especially if these product categories are very

broad and based on consumer categories (as in the WEEE Directive). While this level of differentiation

does not cost much, it is clear that it would not be particularly helpful in achieving fair cost sharing

because recycling costs would vary significantly between products within such broad categories, and this

21 

 

would not be reflected in the average recycling cost per ton charged for this product category. However, a

reasonable compromise solution could be achieved if e-waste would be dynamically categorized based on

the recycling technology required. Products that can be recycled at about the same cost should be

recycled together and belong to the same category. An investigation of how processing cost based product

differentiation in collective e-waste systems can be achieved is one of the most important practical

questions to-date.

4.2.2.7 How to set fees and targets? A major decision in take-back laws is how targets or fees are set.

Consider the WEEE Directive: it applies to eleven categories of products and each category has different

recycling rate targets. We believe the choice of these targets has been made on an ad-hoc basis and may

not adequately reflect the economic or environmental impact of recycling these products. A recent report

to the European Commission highly criticized the WEEE Directive for not taking into account these

issues in determining recycling objectives (Huisman et al. 2008). Interestingly, the upcoming revision of

the WEEE Directive (EC 2008) considers the collection and recycling target choice to be one of the most

important themes (see section 3.1 for details). Yet, it is still not clear how the new targets are to be

chosen. Similarly, it is not clear on what basis the calculation of advance recycling fees in California (i.e.,

$6-$10 depending on the size of the monitor) were made. We believe these targets or fees must be set to

reflect a balance of the economic and environmental impacts of required take back levels. Tracking the

actual collection and recycling costs, as well as the environmental impact of landfill, is crucially

important for managing this balance, and requires analyzing the take-back collection and recycling

networks, and differentiating products in waste streams according to their environmental properties.

Furthermore, network externalities such as capacity based costing and scale or scope economies can

significantly affect actual costs of collection and recycling. There is a need for identifying these effects

for an efficient calculation of fees and targets.

4.2.2.8 Incentives for Green Design. Finally, incentives for the design and production of greener

products are important in determining the social impact of take-back legislation. It is important to realize

different implementations can result in alternative design choices for producers (as discussed in section

4.1.2). There is a strong need for research on how green design incentives can be generated by take-back

legislation, and how particular implementations affect such incentives. This remains an important

research avenue to be explored.

In sum, many different factors affect social implications of take-back legislation. Furthermore, the

implementation choices affect different components of welfare in different ways. Policy choice should

carefully consider how different components (which are tightly interlinked) are affected by different

22 

 

implementation choices. The links between multiple welfare components and their dependence on policy

and implementation choices forms a broader set of questions, most of which are yet to be answered.

5. CONCLUSION

In an era where the diffusion of product take-back legislation has gained momentum, several major

problems related to take-back policy require attention. In this paper, we highlight pressing problems and

research needs in the context of product take-back legislation for e-waste. Using examples from existing

take-back laws, we show that operational factors significantly influence efficiency of take-back

legislation in achieving desired policy objectives. Essentially, we identify a grey zone between theory and

practice of take-back policy choices, wherein several implementations related decisions take place. We

posit that there is need for operations management research investigating such take-back implementation

choices, their effect on production systems and resulting impact on different stakeholders.

 

Figure 2: The Grey Zone Revisited

Figure 2 explains our perspective by revealing the practical content of the grey zone. This figure is useful

especially for one particular reason. It illustrates the major causes of variations between existing

implementations: 8 dimensions of implementation choices, 4 dimensions of potential producer responses

to these choices, along with impact on 6 different stakeholders that create a very complex landscape. It

would be unfair to expect policy makers to come up with a best case solution under such complexity. It is

also clear that a one-size-fits-all best case solution would not apply in different parts of the world and to

different business environments. In other words, a successful policy model and implementation in one

23 

 

country may not necessarily work in another; not only because welfare perspectives may differ but also

because operational constraints vary significantly between countries.

Figure 2 is also useful for highlighting research needs. While it is almost impossible to identify an

optimal policy choice and associated implementation decisions in such a complicated framework, we

strongly believe it provides a practical perspective such that correct assumptions can be made for research

on take-back legislation. Empirical and normative research revealing the impact of different policy and

implementation choices on different stakeholders is clearly needed as take-back laws continue to spread

around the globe. Focusing on such practically relevant problems not only helps the regulating bodies or

decision makers in actual production systems, but also increases the credibility of our discipline.

It is our hope that the framework in this paper will open a new stream of research. Policy economics tend

to design policy instruments ignoring implementation choices and their effects on stakeholder reactions.

Such high-level perspectives however may not work in practice, because the devil is in the (operational)

details. This paper not only makes a call upon our operations management discipline to make a useful

contribution, but also provides a framework for understanding how operations and take-back legislation

interact.

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Acknowledgements: We thank the INSEAD IPR Network participants for the inspiring and informative

discussions that led to many of the ideas in this paper. We also thank Prof. Kalyan Singhal and three

anonymous reviewers for their valuable comments.

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Market Share with No Competition and ARF

Market Share with Competition and ARF

Market Share with Competition and No ARF

Return Share by Sampling

Return Share by Full Brand Count

Return Share with Fee Differentiation

Individual Brand Responsibility by Brand Segregation

Individual Brand Responsibility by Individual Collection System

Country examples Belgium, California

France, Ireland

UK, Germany, WA,RI ME, NL None Japan (SHARL)

Japan (PC Recycling)

Collection method Municipalities x x x x x x

Retailers x x x x x x x

Producers own take back systems x x x x

Postal system x

Management Single compliance system x x x

Multiple competing compliance systems x x x x x x

Individual producer operated system x x x x x

Recycling Producers own recycling facility x x x x x

Collective recycling system x x x x x x

Who pays End user x x x

Producer x x x x x x

Design incentive No incentive x x x

Design for durability x x x x x

Design for reuse/refurbishment x x x

Design for recycling x x x

Table 5: Comparison of Various E-waste Take-Back Laws (Developed by INSEAD IPR Network Participants)

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Market Share with No Competition and ARF

Market Share with Competition and ARF

Market Share with Competition and No ARF

Return Share by Sampling

Return Share by Full Brand Count

Return Share with Fee Differentiation

Individual Brand Responsibility by Brand Segregation

Individual Brand Responsibility by Individual Collection System

Country examples Belgium, California

France, Ireland

UK, Germany, WA, RI ME, NL None Japan (SHARL) Japan (PC)

Recycling Cost

Sharing According to products sold

Advance Recycling Fee x x

Costs split according to Market Share x

According to products returned

End-of-Life Consumer Fee x

Costs split according to Return Share x x x

Costs split according to real costs of recycling x x

Recycling Cost

Differentiation No differentiation x x x x x

Differentiation determined by recycler per producer x

Differentiation determined by recycler per product x

Differentiation based on real cost of recycling x x

Table 6: Comparison of Various E-waste Take-Back Laws (Developed by INSEAD IPR Network Participants)