4
Direct Costs and Financing of Recycling Programs

Local governments and private entities bear the primary financial burden of recycling in the United States, yet they receive only a fraction of the broader environmental and economic benefits. Spending on curbside and other recycling systems involves costs of collection, such as trucks, labor, and fuel, and costs of processing, sometimes at capital-intensive materials recovery facilities (MRFs). This spending often exceeds the revenue earned from selling recyclable commodities, straining municipal budgets and discouraging program expansion. Emerging financing approaches, such as extended producer responsibility (EPR), show promise in redistributing costs and stabilizing material quality. This chapter examines the costs and financing sources for recycling programs nationwide, with particular emphasis on curbside recycling as the dominant mode of collection (see Figure 2-1 in Chapter 2).

4.1 CURBSIDE RECYCLING PROGRAMS

Many different waste and recycling systems operate throughout the United States, which entail multiple collection and processing stages. The most common recycling system for households is curbside recycling. An estimated 53 percent of the U.S. population—about 176 million persons living in approximately 69.8 million households—has curbside recycling services provided to their homes automatically as

4.1.2 Curbside Recycling Processing Costs

Recyclables collected at the curb from single-family residences are typically processed at MRFs. Figure 4-1 shows recent trends of processing costs per ton on waste from MRFs in the Northeast region (10 states).

A review of commingled recycling costs completed for the Oregon Department of Environmental Quality audited a small number of MRFs in Oregon using their actual 2022 fixed and operating costs (Crowe LLP, 2024). These MRFs ranged from small, labor-intensive facilities to larger MRFs with investments in some newer technologies, although none are as advanced as the newest MRFs being built in the United States (see Box 2-5 in Chapter 2). The average cost of processing commingled recyclables through the single-stream MRFs in Oregon in 2022 was $129/ton (Crowe LLP, 2024).

Processing prices can also vary in response to international policy changes. China and other countries have increasingly imposed restrictions on the levels of contamination allowable in imported recyclables (Resource Recycling, 2022). The private company WM estimated that these restrictions have increased processing costs by about 15 percent—or about $13 per ton—across its 43 single-stream MRFs. These increased costs cover labor needs and investments in sorting equipment to meet the new contamination standards.

Investments in new technology have large initial costs that enable MRFs to process each additional ton of recycling less expensively and more efficiently (see Chapter 3 for more on economies of scale). Over

4.1.4 Overall System Costs for Curbside Recycling

Ignoring external costs for now, the direct costs above include collection plus processing costs minus revenue received from recycling. Putting these direct costs together provides an overall picture of the net financial costs of curbside recycling. A review of rates in several states highlights the variability of these overall costs across the country.

National Estimates

Anshassi and Townsend (2024) estimated government expenditures (net of commodity revenues) on waste and recycling systems in regions across the United States (see Figure 4-3). They found that these expenditures vary based on MRF processing fees, revenue-sharing structures with MRFs (described in the figure by different contract letter types), and market conditions for recyclables. The lowest annual household waste management expenditures were in 2011, during peak recycling markets, while the highest expenditures occurred in 2020, during a market low. Expenditures ranged from $124 to $241 per household per year in 2011, $166 to $267 in 2020, and $154 to $243 in 2021 (Anshassi and Townsend, 2024). Across all years and regions, the lowest expenditures were associated with Contract A, which featured a high revenue share of nearly 97 percent and a processing fee of $85 per ton (see Figure 4-3). The Northwest and Northeast regions had the lowest expenditures, while the Southeast region had the highest. In these estimates, garbage and recyclables collection costs accounted for 53–78 percent of total household waste management costs (excluding recyclables revenue), while landfill and waste-to-energy disposal made up 19–34 percent, and MRF processing (once commodity revenue in subtracted) ranged from less than 1 percent to 23 percent (Anshassi and Townsend, 2024).

California. The state of California recycles 20 percent of the volume of recyclables collected in the entire United States. Regulations require separate diversion of waste, recycling, and organics, which means that multiple trucks service each house each week. Recycling requirements and fees on waste disposal increase the overall cost to the customers. With few exceptions, all costs are bundled into a monthly garbage charge that includes recycling at no additional visible cost to customers.

The cost of waste and recycling in California is higher than in other areas of the country because of this regulatory environment. These costs appear in fees for consumers. For example, in San Jose, households are charged $53 per month, or $641 per year, for residential solid waste management and recycling.² In San Mateo, this service costs $42 per month.³

Washington. Washington is one of only two states that regulate waste and recycling rates (West Virginia is the other). While incorporated cities can contract with private companies for services or provide municipal services, customer rates in unincorporated areas of the state are set by Washington’s Utilities and Transportation Commission. The annual cost for all waste and recycling services in Washington State ranges from $406 to $500/year across various urban and suburban areas of the state. The recycling portion of these rates is highly scrutinized, and regulated rates range from $103 to $128 per year per household.

Although, existing literature does not provide a systematic comparison of the costs of recycling and disposal. Table 4-1 shows the per ton costs for disposal and for recycling in North Carolina (not per household monthly costs as above). As reported in the table, collection accounted for most of the cost of both recycling (73 percent) and disposal (80 percent) (EREF, 2024; NERC, 2024). In North Carolina in 2023, average costs per ton recycled exceed the cost of disposing of a ton, even when revenues from end markets were high for the year. These cost differences do not indicate the savings from shifting materials from recycling to disposal (or the costs of shifting them from disposal to recycling), because they do not capture possible increasing or decreasing returns to scale for either activity. As with the per-household costs, they only capture average costs of current levels of recycling and disposal.

These cost estimates are context and time dependent and do not apply to all curbside recycling programs. In some areas of the United States, such as the Northeast, where landfill costs are high, curbside recycling may have lower average costs than landfilling. In addition, as Figure 4-2 shows, recycled commodity revenue is sometimes much higher than in 2023, which would make the net per-ton cost of recycling more favorable. However, even when average costs (net of commodity revenues) are lower per ton for recycling than for disposal, curbside recycling may not save towns money because of the duplication involved in sending both a garbage and a recycling truck to households every week or every other week.

TABLE 4-1 Costs of Curbside Recycling and Residential Refuse Collection in North Carolina, 2023

SOURCES: Data from North Carolina Benchmark Project 2.0 for collection and processing costs; NERC (2024) for 2023 revenues; EREF (2024) for disposal costs in North Carolina.

The cost discussion thus far considers only the “on-budget” costs of recycling. Even single-stream recycling requires input of household time and effort not included here, although it is a real resource cost to consider in assessing these programs. Table 6-1 in Chapter 6 presents estimates that suggest household time costs may be either on par with or several times higher than the local government’s collection costs for recycling. Although households may also have time costs for setting out waste, those costs are likely much lower. In addition, none of these costs are full social costs of recycling or waste management because they recognize market costs from energy and materials use and do not account for other external environmental impacts (see Chapter 7). Box 4-1 describes an effort to calculate this broader definition of costs by the State of Oregon.

The discussion of recycling costs above concerns curbside programs, which are not universally available to households. When households have access only to drop-off programs, the on-budget costs of the programs are probably much lower, although estimates are not available. However, households’ time and effort costs are much higher for drop-off programs than for curbside recycling, so the overall costs of collection under these programs may be higher and participation lower.

Recycling costs fluctuate over time and vary over space for many reasons, including the cost of infrastructure, labor, and fuel, and variations in housing density and transportation distances. In addition, different municipalities offer different services and impose varying participation requirements, that also generate heterogeneity in costs. The regulatory environment may also be an important source of variation in costs, because some cities keep recycling costs low while burdening disposal with fees. Expanding household participation in recycling can help reduce the costs per ton of material recycled and thus help make recycling more competitive with landfilling.

While the data presented in this chapter thus far display recent, contemporary, and potential near-term costs of recycling programs, it is important to note potential future changes (such as developments in renewable energy seen in Box 4-2) that could have implications for direct costs of recycling and overall waste management systems.

4.2 TRADITIONAL FINANCING APPROACHES

Financing of the recycling system in the United States comes from both private and public sources. Typically, local governments, households, and commercial establishments pay for recycling collection and processing. MRFs and other recycling processors receive revenue from end markets for recycled materials and from gate tipping fees (fee per ton paid for processing by governments and collectors). This section discusses traditional financing approaches to cover the costs of recycling collection and processing that exceed the money earned from end markets, which covers only a small share of the total costs of recycling and processing.

4.2.1 Public Financing

Almost all government financing of residential recycling in the United States comes from local governments (cities, towns, and counties), with at most small contributions from higher levels of government. Some local governments use a “general fund” approach, where recycling does not have a dedicated revenue source and is funded along with other categories of expenditure. Other municipalities rely on an “enterprise fund” approach, whereby the municipality collects fees for recycling and garbage collection, perhaps as an item on property tax bills or utility bills or as an explicit charge for businesses. For residential recycling, the general fund approach is more common than the enterprise fund approach (Sheahan, 2024). Governments may use general funds for residential collection but collect fees from commercial establishments.

Local governments pay for recycling at all stages of the process. For recycling collection, they may use their own employees or contract with a recycling collector. They pay gate tip fees for processing the recyclables they collect and implicitly or explicitly pay these fees as part of their contract with a recycling collector. In addition, when recycling drop-off programs are available, either in place of or as supplement to curbside collection, they are likely to be funded by local governments. Recent data are lacking, but in 1995, governments funded the programs in 77 percent of communities with drop-off services (Walls et al., 2003). Finally, about 20 percent of MRFs were publicly owned in 2020, often by county governments (Anshassi and Townsend, 2024). These public MRFs often collect processing fees from private collectors and other governments that use their services. However, the governments that own them may use general funds to cover some costs.

In the United States, state and the federal government provide little financial support for recycling. For all solid waste management, average state spending in 2021, was $4 per capita compared with $86 per capita for local government spending (U.S. Census Bureau, 2024). Figures for recycling specifically are not available but are likely to have a similar breakdown. Recent years have seen some of the first significant federal expenditures on recycling (see Box 4-3).

4.2.2 Private Financing

Commercial establishments and households may also pay privately to have their waste collected and processed. This private funding for recycling collection falls into two categories: (1) With a “franchise” provision, the municipality requires households or establishments to subscribe directly with private firms that provide recycling services at rates and service characteristics that follow a contract with the government. Household participation in these services may be mandatory or voluntary (subscription), depending on city policy. (2) In other communities, some households may subscribe to private recycling services independently, without a publicly negotiated contract. In 2015–2016, 19 percent of residences with curbside collection had subscription-based services, which could include franchise or individually contracted services (Sustainable Package Coalition, 2016).⁴ Walls and colleagues (2005) found that residential private collection arrangements become much less common as population density rises.

Private collection arrangements are much more common for commercial establishments than for residences. In 1995, 47 percent of communities relied on private recycling arrangements for commercial establishments and another 6 percent had franchise arrangements (Walls et al., 2003). Commercial waste

4.3.3 Management of Risk

The volatility of market prices for recyclables presents challenges for all approaches to funding recycling. MRF and other recycling processing facilities sell sorted recyclables for prices that vary dramatically over time. As discussed above, sometimes the recovered material revenue may cover MRF processing costs; at other times, it falls short.

Sometimes MRFs share this risk with municipalities. Contracts between municipalities and MRFs include provisions for sharing of revenues from the sorted materials (The Recycling Partnership, 2020b). A revenue-sharing component is typically combined with base processing fees or may enter the contract through a processing fee with a sliding scale that depends on prices (Anshassi and Townsend, 2024). Local governments typically enter into contracts of 3, 5, 7, or even 10 years with MRF owners to process and market recyclables. These contracts vary in scope and requirements but generally specify contamination rate limits, cost allocation for rejected loads, processing fees paid by the government, and share of revenue from recyclables paid to local governments after processing fees are paid. In Florida, for example, the current trend is toward higher processing fees paid by local governments and lower revenue shares to local governments (Anshassi and Townsend, 2024). Historically, MRFs charged about $50–$60 per ton to local governments, although sometimes no fee was assessed. Now, fees can reach up to $210 per ton, with an average of $107 per ton for Florida communities (Anshassi and Townsend, 2024).

Key parameters in these contracts include contamination levels and market values of materials (Anshassi and Townsend, 2024). Some contracts include minimum values for per-ton market revenue; if revenue from end markets falls below the minimum, the local government may make supplemental payments to the MRF. Lower fees are typically found when the materials are processed by a municipally owned MRF instead of a privately owned facility. In Florida, for example, MRF contracts rarely use a fixed processing fee; instead, fees (and sometimes revenue shares) are adjusted based on the fluctuating value of materials (Anshassi and Townsend, 2024). Outside of Florida, new MRF contracts sometimes have a flat tipping fee (that includes profit and annual escalators) that ensures that MRFs cover their cost and/or turn a profit. Then, in a separate agreement, the MRF might share with the municipality a portion of the end-market revenue from sale of materials (“revenue-sharing”).

When a municipality agrees to greater revenue-sharing, it accepts more of the price risk and thus may reduce average overall costs (fees less revenues). Large municipalities may be able to manage these risks by dipping into contingency funds or by reordering immediate spending. They may also have good access to credit because of favorable tax treatment of municipal bonds, which eases the difficulty of covering short-term shortfalls. When states allow municipalities to engage in short-term borrowing, residents are less likely to have to pay privately for recycling (Walls et al., 2005). However, even large governments may sacrifice other priorities when recycling markets are unfavorable, and smaller municipalities have even more trouble managing this variability without cuts to services. Thus, imposing these risks on cities and towns may be a substantial disadvantage of the U.S. traditional system for funding recycling.

These risks probably raise the prices paid by customers for private recycling services and may reduce reliability. Residential subscription services pricing does not depend on commodity revenue, so private parties that collect and process the recyclables must bear the risk of commodity price movement. They likely charge higher prices on average to compensate for the risks, raising costs for households relative to a situation in which the government accepts some of the risk. Private companies offering subscription services may also be vulnerable to closure when markets are unfavorable and may be difficult to restart once closed.

4.3.4 Distribution of Financial Burdens

Funding approaches share the costs of recycling in different ways. In particular, public financing can reduce costs for vulnerable households. When a government uses general funds for recycling, it can take advantage of revenue sources such as property taxes that impose more burden on higher-income households. Alternatively, with subscription services or an enterprise fund, every household typically pays the

The outcomes are equivalent: either the consumers must pay social costs of disposal, and so they demand products with less packaging and with product designs that minimize disposal costs (see Chapter 6), or the firms must pay social costs of disposal and thus design products to require less packaging and easier recycling. Thus, a “perfect” individual EPR policy would achieve the socially optimal decisions about product designs, packaging, and consumer recycling decisions.

From the very beginning of EPR rules in the 1990s, however, European countries quickly realized that the perfect individual EPR policy would not be practical. It would require each firm to collect its own packaging and keep track of each product for years until the final consumer is done with it. Having many individual firms collect their waste in disparate locations is unnecessarily expensive relative to having one hauler collect waste at every house along each street. To address these collective EPR policies take advantage of these economies of scale in collection by having a “producer responsibility organization” (PRO) can collect all packaging and used products for disposal and then charge each firm for the weight of its waste collected.

A collective EPR policy, however, is very different from the ideal individual EPR. The collective EPR forces each producer to raise their sales price to cover the cost of disposal, so it effectively becomes a different policy called an “advance disposal fee,” a charge for the purchase of a product that will ultimately require recycling or disposal. But it does not encourage the firm to increase recyclability, nor the consumer to recycle. The fee collects significant revenue, which replaces local government financing of recycling collection.

Thus, a collective EPR suffers from two significant problems not found in the “perfect” individual EPR policy. First, private firms pay the PRO’s marginal costs of collection but not the external damages from waste disposal. Thus, incentives do not necessarily align the full costs of recycling with its benefits. For example, the extent of recycling is not sensitive to whether full direct and indirect (external) local disposal costs are high (i.e., where recycling is beneficial). Second, private firms pay the cost of collecting their own total waste (i.e., average cost per ton), but this payment is based only on weight and not on recyclability. The firm therefore has some incentive to reduce the total weight but not to design for recyclability. In only rare cases does the PRO’s fee vary based on the recyclability of the materials it collects (i.e., “eco-modulation).

The economic impact of EPR varies depending on the scope of the law. For example, Canada’s first packaging EPR policy was implemented in the province of Ontario and was established as a “shared responsibility” law, where producers and municipalities share the cost of recycling programs. However, the law resulted in financing disputes and lawsuits, with the system ultimately transitioning to a packaging EPR policy in 2019 with full responsibility on producers (with full implementation by 2025).

EPR policies may have both advantages and disadvantages in the way they redistribute the burden of paying for recycling. On the one hand, the need to pay producer responsibility organization fees will likely raise certain product prices, although prices may not rise by the full amount of PRO fees. These price increase pass along the costs of recycling to consumers and perhaps discourage consumption of these goods (source reduction of MSW that might be desirable). On the other hand, low-income households spend a larger share of their income on goods and thus may end up shouldering more of the burdens of recycling than they would if recycling were financed by local property tax, which can be designed to fall more on higher income households. Thus, the fact that an EPR follows a “polluter pays” principle does not necessarily make it fairer, because lower-income households may end up more burdened than they did under local government finance.

4.4.1 EPR Policies Outside the United States

Since the 1991 German “Green Dot” program, use of EPR policies has extended throughout the globe. EPR as a policy lever has grown, from voluntary programs and lightly enforced laws in Southeast Asian countries, to shared responsibility programs in the Canadian provinces of Manitoba and Saskatchewan, and to far-reaching producer responsibility programs in British Columbia and Quebec.

Empirical research supports the concern that collective EPR policies do not create the desired effect on product design. Joltreau (2022) studied the effects of European EPR policies, using data on the annual cost to comply with EPR from the organization PRO Europe (Packaging Recovery Organisation Europe) for 25 European countries and four packaging materials over 18 years (1998–2015). EPR rules vary in their stringency about the percentage of each firm’s total cost that must be paid and what materials can be recycled. Joltreau (2022) argues that compliance cost is an effective overall measure of stringency and scope and observes the causal effects of EPR stringency on the amount of packaging and on substitution between packaging materials. Joltreau (2022) found that European EPR rules have a statistically significant but very small effect on the weight of packaging and no effect on substitution toward a type of material that could more easily be recycled.

What can be done to reestablish the original goal of using EPR policies to create incentives for recyclability? A recent innovation of EPR legislation in the European Union and North America is the development of “eco-modulation” of fees on producers. Eco-modulated fees vary by product characteristics, and they reward firms that design products to be more recyclable or to be more environmentally friendly. Lifset and colleagues (2023) discuss ideas for restoring the incentives for eco-design by using eco-modulation within EPR systems; they considered choices around products to be included in the fee structure, objectives to be pursued, criteria to be employed, differentiation within the structure of fees, and the amount of each fee to be charged. These choices all depend on the goal of the EPR system, such as (1) to achieve selected policy targets for eco-design, for recycling, or for all product-life management, including product durability and the ability to repair or to reuse it; (2) to account for the presence and toxicity of hazardous substances; (3) to charge for external costs of disposal and thus achieve socially optimal disposal; or, more comprehensively, (4) to cover external costs of production (Lifset et al., 2023).

Ultimately, Lifset and colleagues (2023) propose a fee structure based on each product’s impact on the environment as measured through a life cycle assessment. They also recommended requiring better data from PROs about product characteristics and better use of those data for ex post policy evaluation. However, even eco-modulated charges do not vary with all the dimensions of heterogeneity in cost of recycling by location and over time, so they fall short of the ideal of imposing a tax that reflects the external costs of disposal for specific material types—a tax that would provide incentives for consumers to demand products with optimal durability, eco-design for reusability, and recyclability.

4.4.2 EPR for Packaging in the United States

For over 2 decades, many states have used EPR policies to fund and manage “hard to handle” materials. According to the Product Stewardship Institute (2025), 33 states and the District of Columbia have enacted 141 laws across 20 product categories. These policies have mostly addressed electronics, mercury thermostats, batteries, pharmaceuticals, paint, fluorescent lighting, and mattresses. What is more, the U.S. Environmental Protection Agency (2024) has published guidance on EPR policies for preventing plastic pollution (see Box 4-4).

Responsibility for waste and recycling management is decentralized in the United States, with authority delegated to individual states. This structure complicates efforts to implement federal recycling legislation, resulting in a patchwork of state and local recycling policies. Maine led the United States in passing a packaging EPR policy in 2021, followed by Oregon in 2021, Colorado and California in 2022, and Minnesota in 2024. Each has different financing and regulatory requirements (see Box 4-5). These states have used the policy as an opportunity to address other related packaging and recycling goals, as the new laws included regulatory requirements that are not part of the EPR policies themselves. In each of these five states, the enacted EPR legislation for packaging reflects the varying recycling gaps and perceived needs for that state. California and Oregon’s bills were born of frustration about the growing use of plastic packaging and its low recycling rates, while Colorado’s legislation (see Box 4-6) was driven by the high cost of recycling and low recycling rate. The need for economic support is also clear in the structure of Maine’s law, while Minnesota is looking for future increases in its recycling program investments.

While some producers view EPR as a regulatory burden, others see it as an opportunity to gain greater control over packaging standards and materials, aligning compliance efforts with their sustainability goals. To achieve high recycling rates, producers need households to recycle more, resulting in more recyclable feedstock.

4.4.3 Deposit-Return Systems

In a deposit-return (or deposit-refund) systems (DRSs), consumers pay a small amount when buying a product—often beverages in single-use cans and bottles. This amount is reimbursed to them when they bring the empty beverage container to a collection point with a vending machine or manual handheld scanning readers.

A DRS replaces the local government as the party responsible for funding and managing the recycling of the items it covers. The collection or redemption center manages their collection and is usually funded from the DRS. This system can encourage consumers to recycle containers rather than dispose of

4.4.4 Benefits of Deposit-Return Systems

A unique feature of DRSs is that they provide incentives directly to consumer and commercial establishments to return their containers for recycling. Not surprisingly, recycling rates of these containers are much higher with a DRS than without one. For example, states that implemented a DRS for polyethylene terephthalate (PET) bottles saw a 56 percent recycling rate, compared with 18 percent in states with no such system (Container Recycling Institute, 2024). Globally, recycling rates exhibit a strong positive association with the deposit fee and refund amounts, (see Figure 4-4).

To determine the causal effect of DRSs and to address place-based selection bias, Viscusi and colleagues (2012) conducted a difference-in-difference analysis, comparing the change over time in two states that extended their DRS to include plastic bottles with changes in states that did not. After the extension, the number of households that recycled increased by a statistically significant 8.5 percent. Likewise, Ashenmiller (2009) assessed the material brought in for rebates to California recycling centers and concluded that 36–51 percent of the material would not otherwise have been collected. In 2017, Oregon increased its deposit from 5 cents to10 cents, because of a provision that increases the deposit when the recycling rate falls below a certain threshold, and saw an immediate increase in the recycling rate of the bottles covered in its program from 64 percent to 81 percent. Their recycling rate for covered bottles is estimated at 90.5 percent for 2023, according to the Oregon Beverage Recycling Cooperative (2023).

DRSs may have other benefits as well. First, given the high degree of sorting involved in returning beverage containers to recycling systems, a DRS can provide a cleaner recycling supply stream with less contamination relative to material collected at the curbside. This source-separated waste has higher value for end uses and is less likely to be diverted to landfill.

Second, the initial justification for using these systems for beverage containers included litter reduction; empirical research finds that they have been successful in this area. Levitt and Leventhal (1986) observed that New York’s bottle bill reduced litter near highway exits and railways by a statistically significant 44 percent. And, using data from the United States and Australia, Schlyer and colleagues (2018) found that containers constitute a 40 percent lower share of coastal debris surveys in states with container DRSs. Similarly, Critchell and colleagues (2023) found that adopting a DRS significantly reduced plastic marine debris in Australia.

Third, these policies may have broader social benefits. Ashenmiller (2009) found that the beverage container DRS provide a significant source of income for a group of very-low-income “professional” recyclers. Extending this logic, she found that adopting beverage container DRS reduces petty crime by 11 percent (Ashenmiller, 2010).

4.4.5 Disadvantages of Deposit-Return Systems

Despite their many demonstrated advantages, deposit-return systems may have important disadvantages as well. First, while these programs increase the recycling of beverage containers through monetary incentives, they also remove high-value materials such as aluminum and PET bottles from the curbside stream. Studies have shown that aluminum and PET bottles are a small percentage of the recycling stream by weight but contribute disproportionately to MRF revenue because of their higher market value (Container Recycling Institute, 2022). Diverting these materials from the curbside stream reduces revenue for MRFs and could negatively impact the financial sustainability of curbside programs (unless MRFs were to receive some of the net revenue from the DRS). In fact, curbside recycling programs sometimes need to enforce against theft of containers from curbside bins (Lange, 2012).

Removing these materials from the curbside stream under a national DRS could reduce sales revenue to MRFs by more than the cost-saving from handling less curbside material. In addition, states with a DRS have struggled to invest in sophisticated MRF technology without the revenue associated with aluminum cans and PET bottles (Basuhi et al., 2024). Basuhi and colleagues (2024) estimate that implementing a national DRS would require increases in MRF processing fees net of commodity revenues by 13 percent and that $373 million more would be needed to offset these losses. Using unclaimed deposits to help fund MRFs, as California does, might help preserve this system. To date, DRSs have not confronted these issues because most were in place before curbside recycling programs were implemented.

A second potential drawback of DRSs is their high cost to redemption centers and consumers. Redemption centers, often in retail stores, must allocate space for collection and storage for the materials. Manual return systems rely on workers, and even automatic systems such as “reverse vending machines” require labor to maintain and manage, in addition to the capital costs of the machines. Estimates of handling costs vary. In Connecticut in 2018, handling costs were estimated at $0.0284 per container, which is within the range of handling fees paid by states that offer explicit reimbursement to redemption centers (Reloop, 2021). Handling fees offered by states range from 1 cent in Iowa to 3.5 cents for commingled containers in New York, Maine, and Vermont. An earlier estimate pegged the costs of California’s system at 0.2 cents per container (Ackerman et al., 1995), which is an inflation-adjusted 0.41 cents per container today. About 50 percent of redemption centers in California, where the handling fee was 0.86, closed in the last decade, implying that centers may experience higher costs in many locations.

These estimates suggest relatively high costs relative to curbside recycling programs. For a PET bottle that weighs 9.25 grams (Recycling Today, 2015), a range of cost from $0.0041 to $0.0284 per bottle collected for refund implies a cost range from $400 to $2800 per ton of PET bottles collected, which may typically be higher than average costs for curbside collection and MRF processing. Adding a DRS to an area with curbside collection for non-deposit materials such as paper probably does not lower the costs for curbside collection dramatically. Thus, the combined costs of the two systems may be high.

Third, these redemption center costs exclude the substantial time and energy that households spend to sort, transport, and return materials for redemption. Household costs are likely lower for curbside recycling. Thus, the social costs of having a DRS may exceed curbside recycling costs substantially, suggesting that they are suited to collecting materials for which legal and illegal disposal is most harmful.

Finally, as with an EPR program, collecting materials through a DRS does not ensure socially valuable uses for the collected materials. The problem may be particularly acute for plastics. Virgin plastic resin is very inexpensive; manufacturers have little incentive to pay for U.S. Food and Drug Association–approved postconsumer resin without a policy that provides either a strong incentive or a requirement to do so. As a result, recycled plastic is more commonly used in nonfood applications (PET to textiles, high-density polyethylene [HDPE] to piping, low-density polyethylene [LDPE] to decking, polypropylene to paint cans). Even in these nonfood applications, however, demand may be low. A DRS may need to be coupled with policies to drive demand for the materials collected (e.g., a recycled content standard).⁹

4.4.6 Role of Unredeemed Deposits

Imperfect return rates in DRSs can generate net revenue via unredeemed deposits (i.e., when a customer forfeits deposit value by not returning the PET bottle). Handling of unredeemed deposits varies widely in the United States:

• Connecticut, Massachusetts, and Maine use it for general funds.
• New York gives producers a fraction to cover system costs.
• Iowa allows producers to keep it and operate the DRS.
• Michigan uses it to support retailers that handle deposits.
• California and Vermont earmark it for specific purposes such as a beverage container recycling fund and clean water programs.

If a 10-cent deposit fee were implemented nationwide, close to $1 billion in unredeemed deposits could be available. Policy design is needed for effective use of this revenue without introducing counterproductive or perverse incentives. For example, although perhaps politically expedient, allowing producers or retailers to keep forfeited deposits may undercut their incentives to encourage recycling (Calcott and Walls, 2005). Policymakers designing a DRS need to carefully consider the degree to which unredeemed funds are allocated back to DRS operations.

4.4.7 Considerations for Policy Implementation

In addition to unredeemed deposits, other considerations for implementing a DRS include the systems’ limited scope and issues of end-market demand.

DRSs invite consumers to recycle, so expanding such a system nationwide could increase the amount of material collected for recycling. But it is important to note that the materials collected in these systems would account for only about 20 percent of the volume of recyclables in the United States. New deposit-return policies could be coupled with EPR policies to address the remaining 80 percent.

Additionally, increasing the supply of plastic and aluminum recyclables may have unintended market consequences, especially because demand for recyclables, particularly plastics, is always uncertain. A demand-side postconsumer recycling policy could balance the additional supply. Policies that address gaps in financing of plastic recycling are particularly important because the cost of virgin plastic is less than the cost of processing postconsumer plastic for recycling (see Chapter 5 for relevant recommendations).

4.4.8 Deposit Legislation

Deposit-return legislation is sometimes thought to be a form of EPR, based on the role that producers play in the recycling process. However, EPR policies tend to require “invisible” fees built into the cost of the product, while DRSs require a visible fee to be paid by the purchaser (to be refunded only when that product is redeemed). Ten states in the United States have DRSs—also known as “bottle bills” (Figure 4-5; Container Recycling Institute, n.d.). These states tend to have high recycling rates for materials covered under the DRS; for example, bottle recycling returns are as high as 80 percent in Oregon.

Except for Hawaii’s program, all deposit-return programs in the United States were enacted before curbside recycling programs were implemented. Deposit-return policies include aluminum cans and PET bottles that could instead have been collected and sold as part of curbside recycling, reducing revenue for these programs unless the deposit-return policy provides compensation for them. Without compensation, the foregone revenue may make curbside recycling less affordable for local governments.

4.5 KEY POLICY OPTIONS AND RECOMMENDATIONS

The following section outlines key policy options and recommendations for improving recycling systems, focusing on strategies centered around the costs and approaches of recycling systems.

4.5.1 Extended Producer Responsibility

As discussed extensively in this chapter, EPR can be a policy strategy for decreasing the total environmental impact from a product by making the manufacturer of the product responsible for its entire life cycle and especially for the take-back, recycling, and final disposal of the product (Lindhqvist, 2000).

While EPR policies have already been implemented for hard-to-handle materials in the United States (e.g., e-waste, tires, mattresses, paint), EPR policies focused on plastics and other packaging and recyclables are less common but growing in prominence (France Ministry of Economy, 2025). As of 2024, five U.S. states have passed EPR policies for packaging.¹⁰ While each state law is quite different, with language that reflects each state’s unique recycling needs, four of the policies (all but Maine’s) incorporate requirements for packaging reduction, recycling, use of postconsumer content, packaging design, and “responsible end-market” requirements.

The general structure of an EPR policy requires producers to cover the cost of the entire life cycle of packaging, including disposal, and to pay fees based on specified criteria. Additions or subtractions from the producer’s disposal fee are intended to encourage changes in covered materials that would increase their recyclability or reduce their environmental harm. The fees are then used to pay for the cost of collection, processing, and recycling of the covered materials. In this way, EPR policies also establish a financing mechanism for waste management systems that is different, and likely more stable and reliable, than traditional financing schemes. This alternative funding scheme can improve recycling access and increase supply of recycled materials.

EPR policies for packaging typically allow for a nonprofit PRO to manage the program efficiently on behalf of the producers whose products are covered in the regulation. The organization collects fees from the producers and uses those funds to support and manage the recycling programs on their behalf to achieve the requirements of the law cost-effectively. By using the revenue generated by those EPR fees to support waste and recycling systems, EPR policies reduce the financial burden on local governments and recyclers, facilitate investment in recycling, and reduce the cost of recycling. Producer responsibility toward recycling programs increases access to recycling, ultimately increasing the supply of recycled materials. In addition, as discussed in this chapter, fee requirements under an eco-modulation scheme in the legislation can provide incentives for packaging design choices to help achieve the EPR policy goals, and disincentives for packaging design choices counter to those goals. Cost implications of EPR policies vary, but initial ranges can be observed in the states that have already enacted EPR for plastics and packaging. For example, Oregon has a designated PRO, the Circular Action Alliance, that has established packaging fees in a proposed program plan, ranging from 1 to 273 cents per pound depending on the packaging type (ODEQ, 2024).

EPR policies have trade-offs as well. First, at least a portion of the EPR fees imposed on manufacturers would be passed along to consumers in the cost of goods (e.g., groceries) and thus could be considered a regressive tax, as these goods make up a larger-than-average percentage of spending by low-income households. Second, EPR policies have not historically been designed to address end-market demand for recyclables. Adding supply of material without addressing demand has the potential to erode market prices,

Key Policy Option 4-2: State governments could enact extended producer responsibility (EPR) policies, informed by any minimum standards provided by the U.S. Environmental Protection Agency. State-level needs assessments should identify gaps in current services and programs and serve as a basis for setting EPR fees. Within an EPR framework, state governments could consider policies, such as recycled content standards, to enhance end markets for recyclable materials.

4.6.2 State-Based Landfill Surcharges for Recycling

MSW landfills charge a “tipping fee” for the waste that they accept. This fee is typically charged as dollars per ton of solid waste (or in some cases, dollars per cubic yard). The tipping fee comprises (1) a baseline fee paid to the owners of the facility to cover their capital and operating costs and (2) a surcharge to address the social and environmental impacts associated with landfilling (i.e., external costs relative to other means of MSW management). A landfill tipping fee raises the cost of disposal and reduces its cost advantage over recycling. Furthermore, the funds collected can be used to support statewide or local recycling programs, further shifting MSW management from landfills toward recycling. Along with these incentives, reducing landfill use confers benefits by reducing the social and environmental externalities that they generate (see Chapter 6). Costs are also reduced by lengthening the lifetime of available landfill capacity.

The landfill tipping fee surcharges is an example of a Pigovian tax or fee intended to internalize the social and environmental external costs of a management option (as discussed in Chapter 3). Under ideal competitive market assumptions, a Pigovian tax induces reduction of costly polluting behavior in an economically efficient manner, by setting the tax rate equal to the marginal external cost (Arrow, 1970; Burrows, 1979; Cremer et al., 1998). The process for determining an appropriate externality fee involves two major steps: first, environmental impact studies such as life cycle analyses calculate the social or environmental damage expected from each disposal option (discussed below); second, economic studies are necessary to value this impact. These economic studies can employ behavioral (revealed preference), or attitudinal (stated preference) methods (see Chapter 6; see also Mendelsohn and Olmstead, 2009). Examples of social and environmental costs that have been explored for use in cost-benefit studies include Matthews and Lave (2000), Sovacool and colleagues (2021), and Comineti and colleagues (2024). Applications of externality pricing in cost-benefit analysis for solid waste management are demonstrated by Goddard (1995), Fullerton and Kinnaman (1995, 1996), Nahman (2011), Paes and colleagues (2020), Medina-Mijangos and colleagues (2021), Matheson (2022), and Massarutto (2024).

Landfills are regulated and permitted by state governments; however, their baseline tipping fees are set by the landfill owners based on market conditions (private landfills) or costs (public landfills). These fees can be paid at the landfill or at a transfer station that feeds into a particular landfill. At present, many but not all states have a landfill tipping fee surcharge (essentially, a tax), which is collected by landfill operators and paid by the individual or entity disposing of MSW at that landfill. In urban areas, this entity is often a hauler of MSW, but in rural areas the fee may apply to others transporting their own household waste to a landfill. Regardless of the entity directly paying this surcharge at the landfill, the surcharge would mostly be passed along to the original waste generator, raising the cost of waste disposal relative to other management options (e.g., recycling). State governments can also tailor fee levels to their state’s conditions and can choose the appropriate local governments and programs to receive the funds raised.

Current Surcharge Rates Across the United States

By 2009, 30 states had imposed tipping fee surcharges, with an average fee of about $2 per ton (Jenkins and Maguire, 2012). In response to a survey from the National Conference of State Legislatures (2021), 12 states reported surcharges as of 2021, as shown in Box 4-10. These states have tipping fee surcharge rates that range from $0.25 to $6.00 per ton, with a mean of $2.40 per ton, a median of $2.00 per ton, and a standard deviation of $1.70 per ton.

The Environmental Research & Education Foundation (2024) reported total tipping fees in 2023 had a mean of $58 per ton, with regional averages ranging from $43 to $83 per ton, the $2 per ton surcharge represents 3.4 percent of the average total tipping fee (and 2.4–4.7 percent of the low and high regional averages). With a proposed additional surcharge of $2 per ton, the decrease in the financial advantage of landfilling relative to recycling is likely to be minor. However, the very presence of a landfill surcharge and its use to support recycling awareness and infrastructure may be enough to motivate public and stakeholder awareness, support, and action to begin and to continue the process of transition from landfill disposal toward recycling. While the most appropriate landfill surcharge rate for a state will be determined by its legislature with input from its citizens, industry, and environmental and economic advisors, the results of our estimates and analysis suggest that a $2 per ton surcharge can serve as a moderate and representative point of departure to initiate and support state-specific recycling legislation.

Appropriate (Moderate) Rates for a Tipping Fee Surcharge?

While states may select tipping fee surcharge rates based on the environmental and social externalities associated with disposal at a landfill, external damages are difficult to calculate in a consistent manner across sites and are subject to high measurement uncertainty. Furthermore, calculated rates may exhibit wide variation across a state, due to both physical and demographic heterogeneity and the use of differing methods and modeling assumptions. Guidance is needed for states to identify a range of tipping fees to be deemed as moderate and appropriate. For a total tipping fee to be considered moderate, it should encourage systems to begin recycling, due either to the modified cost differential or the new funding for recycling efforts generated by the fees. However, if they are too high, they present a threat to an MSW system’s financial viability while it undergoes the transition from landfill dominated MSW service toward more recycling, nor to penalize communities that lack the necessary conditions and resources for this transition. A first test for a state or system is to ask, “How does a proposed surcharge compare with tipping and surcharge rates at other landfills, nationally or in their state or region, especially those with similar waste, economic, and environmental conditions?”

While surcharges affect costs only minimally, it is reasonable to assume they are passed along to the original waste generator, thereby raising the cost of waste disposal relative to other management options, including recycling and composting. This may help address difference in cost between landfill disposal (which are usually lower) versus recycling and composting (which are usually higher). Addressing this difference is especially important when revenues from the sale of recovered recyclables are in a market downswing. Increasing surcharges for landfill disposal could increase economic pressures and incentives for recycling and composting.

Surcharges can further reduce the price differential between landfilling and recycling if the revenues generated are allocated by state governments to support local recycling efforts (or other MSW treatment, including composting or other treatments of organics). These funds could offset the cost of local recycling operations by being allocated to recycling enterprise funds, or could support grants for local recycling infrastructure, fund local social modeling programs (i.e., locally organized programs that promote recycling norms and behaviors [see Section 6.3 in Chapter 6]). The funds could even support research efforts that inform recycling systems and promote their cost efficiency.

The economic pressure and incentive to recycle from landfill tipping fee surcharges would be felt by consumers differentially. For example, businesses and rural households that pay for private waste hauling (and rural households that transport their own waste for disposal) would see the direct effects of this policy and have greater incentives to substitute recycling for disposal, when possible, as well as reducing their overall waste generation. Local governments that pay for MSW collection will likewise have stronger economic incentives to provide curbside and recycling drop-off programs and municipal composting programs. Households served by these government-funded systems may see an increase in taxes or waste management fees, but not a change in costs that would motivate changes in their waste management behavior. Exceptions would include households covered by pay-per-bag or subscription programs that are sensitive to the waste quantities. Thus, a limitation of this policy is that it may not encourage appropriate or desired changes in behavior for these households, which make up most of the population.

Additional disadvantages of this policy may arise but can be managed by keeping landfill tipping fee surcharges at an appropriate level. For waste generators who experience greater direct costs of disposal, surcharges may encourage not only reuse and recycling, but also illegal dumping of waste or greater contamination of recycled materials. The unintended adverse effect of a landfill surcharge on illegal dumping is likely to vary by location and the associated demographic characteristics of that locality. Residents of rural areas may have easy access to unmonitored locations for dumping waste illegally; urban locations could also see illegal dumping in vacant lots or in commercial dumpsters, if those locations are not monitored. Revenues raised from surcharges could be spent on efforts to monitor and/or discourage illegal disposal and “wish-cycling” behavior (i.e., the act of placing non-recyclable items in recycling bins under the mistaken assumption that they can be processed, leading to contamination and inefficiencies in recycling systems).

The distribution of burdens of landfill tipping fee surcharges may also be a concern. As described above, direct costs of waste management will rise for households that transport waste to landfills themselves or pay for private waste hauling. It will also raise costs for local governments who will see an increase in their waste management fees. When local governments use enterprise funding for MSW and impose uniform charges across all households, these increased costs will place a higher relative burden on low-income families than on high-income families. Some of these effects might be mitigated by focusing the redistribution of surcharge revenue to address recycling costs for the households most burdened by the surcharges. Lastly, for local government-owned and -operated landfills, landfill tipping fees are often used to support general funds for other government-provided public goods. If the tipping fee surcharges work as intended and decrease the amount of recyclable material entering a landfill, these landfills may see a decline in revenue that reduces funding for local governments.

Landfill tipping fee surcharges to encourage and support recycling need to be mandated at the state rather than local level. Some local governments own and operate their own landfill and therefore have different incentives for the volume of waste that is landfilled. Furthermore, residents or business could bypass local surcharges could be bypassed by transporting waste to landfills in localities without surcharges. State governments can tailor surcharge levels to their appropriate state-level conditions and can choose the appropriate local governments and program to receive the funds raised.

Especially as compared with other policy options for recycling, the time frame necessary to implement surcharges on landfill tipping fees could be short. If a state legislature adopts or increases a landfill surcharges, changes can likely begin within 1–2 years. Funding for recycling programs arising from these surcharges can thus likely be available within 2–3 years (i.e., 1 fiscal year after the initiation of the surcharge). To measure its effectiveness as a policy, states could monitor both funds raised and MSW diversion

While tribal and rural recycling often run into similar issues—such as lack of infrastructure; long distances to access waste disposal; recycling, or composting; and lack of funding−−tribal recycling programs face many unique challenges. If a reservation-based tribe is not checkerboarded with non-tribal residents (i.e., if the land is solely tribal land), then they often are responsible for their own solid waste responsibilities as they are not tied to a municipality or county collection system. Funding is seriously lacking to address this problem, since tribal programs are not locally, state or federally funded. These locations may have reduced populations without the ability to collect taxes to cover costs, substantially raising the fees to address open dumping of solid waste. Cultural differences also arise between tribes and rural citizens. Tribal citizens who live on reservations solely inhabited by tribal members believe that the U.S. government will provide for them based upon treaty rights that are in place. Non-tribal citizens who live in rural areas may try to do the right thing but many simply do not want to pay to dispose of their waste and therefore dump it wherever they can—to avoid paying fees.

Unique challenges are also present for each tribe. Of the 574 sovereign tribal nations across the United States and Alaska, 229 are Alaskan Native Villages. Nearly half of those Alaskan Nations must send out their waste by barges or planes that may only be able to operate for 6 months of the year. For example, the Alakanuk Traditional Tribe of Alakanuk, Alaska, does not have good roadways in and out of the Village, so waste is addressed through barges and planes. Conversely, the Chickaloon Village of Palmer and Sutton, Alaska, relies on its roadways to dispose of waste. The Fort Belknap Indian Community, the fourth largest reservation in Montana is only 40 miles south of the Canadian border. This community have challenges with finding resources for waste disposal because of the distance from larger communities. The Mille Lacs Band of Ojibwe, located in East Central Minnesota approximately 2 hours north of Saint Paul, Minnesota, has developed a successful hub and spoke system but funding remains a challenge.

Conclusion 4-3: Relying on local government financing limits access to recycling programs, particularly for residents of rural areas, where recycling costs are high. It also limits access for many small business owners, who often face difficulties participating in municipal recycling programs. Alternative funding mechanisms, such as state or federal grants or EPR programs, would help distribute recycling costs across a broader population.

Key Policy Option 4-4: The U.S. Congress and state legislatures could authorize and appropriate funds for rural and tribal recycling. These funds could help communities overcome transportation distances and achieve economies of scale, through purchase of infrastructure such as trucks, drop-off and transfer facilities, and processing facilities. In parallel, the U.S. Environmental Protection Agency could continue to provide Solid Waste Infrastructure for Recycling grants for rural and tribal communities. State government funding could be derived from revenues generated from extended producer responsibility policies, landfilling tipping fee surcharges, or other state-based revenues.

REFERENCES