Theory, Practice and Research Needs

Transcription

1 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, Atlanta, GA. Phone: , Fax: Luk N. Van Wassenhove INSEAD Social Innovation Center, Boulevard de Constance, 77300, Fontainebleau, France Phone: 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 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 1

2 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 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 2

3 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 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 3

4 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 4

5 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 endof-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 5

6 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 ( 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 6

7 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 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 and Huisman et al for a broader picture). Belgium France, Ireland UK, Germany Collection method Municipalities x x x Retailers x x x Producers can own collection systems 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. stateoperated as in Belgium), or competitive (with multiple non-state operated systems as in France). The 7 x

8 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 8

9 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) Collection method Retailers x Postal system Management Multiple competing collective systems x Individual producer operated systems allowed Recycling Producers own recycling facility x x Who pays End user x Producer x x Japan (PC) 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 9 x x

10 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 10

11 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 11

12 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 12 Cost under Return Share A $50 $75 B $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 policymaking, 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 takeback 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

13 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 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 nontraditional 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 13

14 for profit making (Fleischmann et al. 2001), the case of costly recycling enforced by take-back laws remains an open problem to be investigated 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 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 14

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 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 takeback 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 Moderating Factors: 15

16 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 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 manufacturerowned 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 takeback implementations and investigating their impact on stakeholders is a critical question to be answered 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 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 ). It creates additional jobs 16

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 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 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 stateoperated 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 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 17

18 impact of take-back legislation also needs to consider the dynamics in competition, legislative processes and policy objectives 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 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 ) and monitoring requirements (as discussed in section ). 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 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 18