Economic Instruments for the Circular Economy in Italy

4. International practices on using economic instruments for a circular economy

Abstract

Economic instruments are widely used across OECD countries to achieve environmental policy goals cost-effectively. Economic instruments can support shifting the economy from linear to circular practices, by generating price signals that promote eco-design, reuse, recycling and material recovery, minimising resource use and waste.

This chapter reviews international examples of the use of economic instruments with links to the circular economy, such as virgin material taxes on raw materials, plastic taxes, household waste disposal charges (including pay-as-you-throw), landfill and incineration taxes, as well as Extended Producer Responsibility and public procurement approaches.

4.1. Introduction

Economic (or market-based) instruments seek to address the market failure of environmental externalities, either by incorporating the external cost of production and consumption activities through taxes1 on processes or products, or by creating property rights and facilitating the establishment of a proxy market for the use of environmental services (OECD, 2007[1]). The internalisation of externality costs ensures that the damage caused by pollution is better reflected in market prices (OECD, 2001[2]). The use of economic instruments, as well as their variety, has grown steadily in recent decades in OECD countries, and to a lesser extent in non-OECD countries.

The promise of economic efficiency is one key advantage of economic instruments. Economic instruments attach a price to undesirable behaviour, such as a tax that industry or consumers must pay to use a specific product or service. By allowing flexibility in the method and pattern of compliance, economic instruments can lower the overall economic costs of achieving a given level of environmental protection (Svatikova, Brown and Börkey, Forthcoming[3]). Environmental taxes, in particular, offer the advantage of generating dynamic incentives for pollution abatement: as environmental taxes generally require polluters to pay for residual emissions on top of abatement costs, this provides a continuous incentive to lower pollution abatement costs and reduce emissions further to avoid paying the tax.

Across OECD countries, economic instruments are used to help governments reach their environmental policy goals in a cost-effective manner, as well as promote sustainable growth and, potentially, generate revenues. One additional advantage of economic instruments is that they may strengthen incentives for innovation in green technologies, as firms along the value chain seek cost-effective abatement solutions. Well-designed fiscal policy can reinforce “green” stimuli and align traditional forms of stimulus with environmental objectives. In this vein, the European Green Deal acknowledges the crucial role of economic instruments, including taxation, in the transition to a greener and more sustainable economy. It encourages EU Member States to design appropriate green taxes that can raise revenue, remove environmentally harmful subsidies, and ensure the full implementation of the polluter pays principle, a fundamental principle for internalizing the costs of environmental protection (OECD, 1972[4]).

4.2. A typology of economic instruments for the circular economy

Economic instruments can help to re-shape the economy from a linear pattern of consumption to a circular pattern of recovery, reuse, recycling and overall lower resource use and waste, while avoiding excessive economic impacts (OECD, 2021[5]). By favouring practices that are more sustainable and circular, and penalising those that are more wasteful and harmful to the environment and human health, economic instruments help to reorient consumer choices and market forces towards lower resource use, and can boost innovation in green technologies.

Governments generally set prices or policies that indirectly impact the price signal for actors. Price-based economic instruments include taxes, charges, as well instruments such as deposit-refund systems (Svatikova, Brown and Börkey, Forthcoming[3]). Performance-based economic instruments assign businesses with the responsibility to meet circular economy performance targets or goals, which indirectly impact or set a price to motivate behavioural change (Svatikova, Brown and Börkey, Forthcoming[3]). For instance, green public procurement (GPP) provides economic incentives for bidders to compete on green criteria through the purchasing power of public authorities, leading firms to seek ways to reduce the total product or service life cycle costs.

Table ‎4.1 summarises the types of economic instruments related to the circular economy. Many of the listed economic instruments, including landfill taxes, packaging charges, deposit-refund systems and taxes on virgin materials, have become an established and effective part of the materials management policy landscape in OECD and EU Member States over the past decades (OECD, 2016[6]; OECD, 2019[7]; OECD, 2014[8]; Lifset, Atasu and Tojo, 2013[9]).

Table ‎4.1. Typology of economic instruments for a circular economy

Type	Short description
Price-based economic instruments
Taxes and charges	Taxes and charges on goods and services elicit behavioural changes by increasing costs. As a consequence, they discourage the consumption and production of targeted products or activities. The level of a tax or a charge can be based on the social cost of the activity or physical characteristics (e.g. weight of the consumable).
Subsidies	Subsidies encourage behavioural change by reducing costs for targeted products or increasing benefits for targeted activities. They directly increase the relative cost of polluting products and activities, or they can increase the use of products or activities that have a positive impact on the environment.
Deposit refund systems (DRS)	A system in which an initial payment (deposit) is made by a customer at the point of purchase, which is then refunded when the product or packaging is physically returned to the collection scheme by the customer. DRS set a price for participating in the collection scheme, creating an incentive for customers to participate. DRS correlate with high collection rates, quality of collection, and low littering levels.
Advance disposal fees (ADF)	ADF are charged on products at the point of sale based on the estimated waste management costs. Governments or producer responsibility organisations that collectively fulfill EPR obligations set an ADF that creates a cost/price for end-of-life collection and recovery services. These help to internalize the costs of these services to the producers and consumers of the products that become waste.
Performance-based economic Instruments
Tradeable permits and tradeable compliance certificates	Tradable permits allocate limited emission or resource exploitation rights to firms. They have two key characteristics: i) a “cap” (a quantity limit) on the pollution that is being regulated; and ii) a trading system that allows individual firms to split the cap among them. The overall cap ensures the desired overall outcome in a cost-effective way (via trading). The key requirement is the existence of a competitive market for such permits (including low transaction costs and a sufficient number of market players).
Extended Producer Responsibility (EPR) take-back schemes	The most common EPR policy instrument are mandatory take-back schemes, which oblige firms to meet targets for the collection and processing of the products they produce. This creates incentives for firms to find ways to reduce the end-of-life costs of their products and to meet their EPR requirements more efficiently.
Green Public Procurement (GPP)	GPP directs the purchasing power of public agencies to incentivise markets to innovate and to award suppliers who offer more environmental or circular modes of supply. In the first instance, GPP increases demand for targeted products, increasing their price in the short term. In the medium to long term, GPP can send a price signal that could encourage producers to increase the supply of more sustainable materials and products on the market.

Source: (Svatikova, Brown and Börkey, Forthcoming[3]).

Countries generally strive to implement economic instruments as part of a policy mix. New regulations help to remove regulatory barriers, set targets and obligations (e.g. extended warranty obligations) or restrictions (e.g. bans on harmful substances), introduce standards, create framework conditions to enable the use of secondary materials, or ease the administrative burden for circular business development and innovation. Regulatory instruments may also encompass the monitoring and enforcement framework relevant to circular economy measures and targets. Regulation and economic instruments are often supported by voluntary approaches, such as certification schemes, sustainability standards and commitments (e.g. for recycled content in products). They may also include information-based instruments, such as eco-labelling, and communication and awareness-raising initiatives to empower consumers, as well as behaviourally informed interventions. Additionally, policy makers may support education, training and research measures, as well as measures to facilitate co-operation between the public and private sector and across sectors (e.g. waste and packaging industries), or across stakeholders (e.g. via circular hotspots).

4.3. Economic instruments along the value chain

Economic instruments can help generate price signals oriented towards higher sustainability and circularity all along the value chain, from material extraction, product design to consumption and waste processing. Figure ‎4.1 provides a summary of key economic instruments for the circular economy, categorised by the key point of the value chain they target.

Figure ‎4.1. Selected economic instruments for the circular economy

Note: Economic instruments may influence multiple life cycle stages. For instance, public procurement approaches may promote sustainability and circularity during product design and manufacturing, at the point of purchase, during the use phase, as well as at end-of-life.

Source: Based on (Svatikova, Brown and Börkey, Forthcoming[3]).

The circular economy emphasises the importance of sustainable and circular choices made at the beginning of a product’s life cycle. Circular design features can preserve the functionality and usability of products, their components and embedded materials for as long as possible, thereby increasing value retention and preventing waste. These choices need to be incorporated into company decisions to support sustainable and circular production patterns, from the sourcing of material inputs to the management of by-products. Economic incentives are essential to support sustainable and circular behaviour upstream in the value chain, encouraging the use of secondary materials, linking product design to end-of-life disposal costs, and incentivising business research and development (R&D) and eco-innovation (Svatikova, Brown and Börkey, Forthcoming[3]).

Consumers and households have a central role to play in promoting new circular business models that reduce consumption levels, prevent waste generation and contribute to better waste management through good waste sorting practices. Economic instruments can influence consumer and household behaviour at the point of sale, during the product’s use phase and at the end of the product’s life cycle (Svatikova, Brown and Börkey, Forthcoming[3]). For instance, product taxes change the relative price of products and services at the point of sale and can encourage more circular consumption patterns. Pay-as-you-throw (PAYT) and deposit refund systems (DRS) set a price for households’ pre-sorting effort. A DRS imposes a fee (an unredeemed deposit) on households not participating in the separate collection system. In addition to economic instruments, it is worth noting that information and awareness measures, as well as interventions informed by behavioural sciences, also have an important role to play in informing consumers about the environmental consequences of their choices and in influencing their behaviour.

Although waste management, recovery and recycling are positioned lower in the waste hierarchy, they are critical steps in mitigating environmental impacts and closing material loops. As discussed in the previous chapter, the EU Circular Economy Action Plan sets ambitious targets in this field, including a recycling target of 60% of municipal waste by 2030 and a maximum landfill rate of 10% of municipal waste by 2035. Economic instruments that improve waste management practices generally provide economic incentives for waste disposal agents to change their practices towards the reuse, recycling and recovery of products and materials, as well as to increase recycling by waste operators.

4.4. Selected international practices on the use of economic instruments

This section reviews examples of economic instruments in place in other OECD countries. This section focuses on environmentally related taxes known to be particularly relevant for the circular economy and which can be employed by multiple countries, as based on the OECD database on Policy Instruments for the Environment (PINE). These include taxes on raw materials extraction, taxes on specific products, waste disposal taxes, and landfill and incineration taxes. Additionally, the section covers EPR and public procurement approaches, which are also extensively used across OECD countries to support sustainable development and the circular economy.

4.4.1. Virgin material taxes on raw materials

Virgin material taxes (or taxes on the extraction of raw materials) have been in use in Europe since the early 1990s, mainly to discourage the excessive use of virgin materials in production, and to lower the relative price of secondary raw materials. Up until now, taxes on raw materials have been used principally on quarried materials such as stone, sand and gravel (Söderholm, 2011[10]). It is also possible to apply virgin material taxes to other materials, such as oil and gas extractions, plastics (as done in some OECD countries, see also section ‎4.4.2), or chemical substances used in the agricultural sector.

Virgin material taxes are distinct from royalty payments, and are often associated with resource extraction. Royalties are levied in exchange for access to publicly owned resources and should, ordinarily, generate a stable revenue, whereas environmental taxes should decrease over time if they prove to be effective in changing producers’ behaviour (International Resource Panel, 2020[11]). Royalties are more widely used by countries than virgin material taxes. Royalty payments are typically fees for the use of mining land per tonne of extracted material, or a small percentage of the market value of the extracted materials.

According to the OECD PINE database, there are currently 88 different fees applied to extractive activities of aggregates, gypsum and salt in OECD countries. More than half of these (58%) are earmarked and 84% are based on a physical metric, such as volumes extracted (ad quantum), while the remaining 16% are based on a monetary metric such as sales price (ad valorem). The tax base also differs, with 64% targeting specific types of construction aggregates, 24% covering all minerals, and the remaining pertaining to gypsum and salt. Table ‎4.2 presents examples from selected OECD and non-OECD EU countries. The majority of countries impose revenue-raising levies, whereas countries such as Sweden and the United Kingdom have designed virgin material taxes aimed at affecting the quantities extracted. This is reflected by the higher fee rates, compared to royalties, and the presence of regular updates.

Table ‎4.2. Royalties and taxes on construction aggregates in selected OECD and EU countries

Country	Name of tax, charge or duty	Taxable materials	Year of introduction	Tax rates
Colombia	Mining royalties	Minerals	1995/1996	3% of production value
Croatia	Extraction charge	Sand, gravel, crushed stone, limestone and clay	2009	EUR 0.41 /m3 (sand) EUR 0.55 /m3 (gravel)
Czech Republic	Fee for extracted minerals	Aggregates	1993	Up to 10% of market price of mineral resources
Denmark	Duty on raw materials	Stone, sand, gravel, peat, clay and limestone	1990	Until 2022 : EUR 0.71 / m3 2023 : EUR 0.75 / m3
Estonia	Materials extraction charge	Dolomite, granite, gravel, sand, limestone, clay, peat, phosphate rock and oil shale	1991	EUR 0.75 – 1.42 /m3 (clay) EUR 0.60 – 2.43 / m3 (gravel) EUR 0.98 – 2.94 /m3 (limestone) EUR 0.42 – 1.64 /m3 (sand)
France	Tax on extracted minerals	Minerals (granulates)	2000	EUR 0.21 /t
Latvia	Natural resources extraction charge	Gravel, limestone, clay and sand	1991	EUR 0.21 /m3 (clay) EUR 0.21 – 0.36 /m3 (dolomite, sand, gravel) EUR 0.14 – 0.28 /m3 (limestone)
Lithuania	Minerals extraction charges	Minerals	1991	EUR 0.51 – 0.86 /m3 (clay) EUR 0.84 /m3 (limestone) EUR 0.38 – 0.48 /m3 (sand)
Sweden	Natural gravel tax	Gravel, sand, cobble and boulder	1996	1996: EUR 0.57 /t (SEK 5 /t) 2006: EUR 1.58 /t (SEK 15 /t)
United Kingdom	Aggregates levy	Sand, gravel and crushed rock	2002	2002: EUR 2.55 /t (GBP 1.60 /t) 2010: EUR 2.55 /t (GBP 2/t)
United States (state level)	Kansas: Sand royalty	Sand		EUR 0.1246 /t (USD 0.15 /t)
	Nevada: Minerals tax	Minerals		5% of net proceeds
	West Virginia: Severance tax	Sand, gravel, sandstone and other mineral products		5% of gross receipts attributable to natural resource production

Note: Bold type is for countries discussed in the text. Conversion factor of sand, gravel, crushed rock = 1.8 t/m³; and limestone = 2.8 t/m³.

Sources: (HM Revenue & Customs, 2022[12]; OECD, 2022[13]).

Denmark was one of the first countries to introduce a virgin material tax on construction aggregates, which applies to all domestic extraction and imports. The virgin material tax was implemented in conjunction with the waste disposal tax, first introduced in 1987 and later differentiated by disposal type. Söderholm (2011[10]) notes that the two taxes have caused a marked increase in recycling, in particular, of construction and demolition waste (CDW). In 1985, 82% of CDW was landfilled and only 12% was recycled, but in 2004 the recycling rate increased to 94% (European Environmental Agency, 2008[14]). There has been a slight decrease in the extraction of these materials since the introduction of the tax in 1977, although the tax has not led to a drop in consumption (Söderholm, 2011[10]). Furthermore, the introduction of mandatory separate collection at source of CDW is believed to have significantly contributed to improving the supply of secondary aggregates in the Danish market (Söderholm, 2011[10]).

In Sweden, a low tax rate on the extraction of natural gravel has been applied since 1996 to preserve groundwater, but it was raised in 2003. This increment suggests a significant drop in the consumption of this material. However, the decrease in gravel extraction was already significant before the tax was introduced and could be due to the higher demand for crushed rock because of its higher quality compared to natural gravel. The drop in gravel consumption led to an increase in alternative materials, with a greater impact on emissions. Therefore, while groundwater quality has improved, emissions have increased. This example highlights the need for careful analysis and, possibly, additional instruments to prevent burden shifting. The Swedish case also shows that the gradual tax increase helps producers organise themselves and contributes to increasing the elasticity of the demand, allowing for a better acceptance of the tax (Söderholm, 2011[10]).

The United Kingdom introduced the Aggregates Levy in 2002 as a nation-wide environmental tax on commercially exploited virgin rock, sand and gravel that have been quarried, dredged from UK waters or imported (HM Treasury, 2019[15]). The tax rate has gradually increased to reach the current rate of GBP 2 per tonne. Annual revenues amount to GBP 240-410 million (HM Treasury, 2019[15]). Although it is difficult to attribute environmental outcomes specifically to the tax, it is likely that the landfill tax and the Aggregates Levy together had an important impact on the overall reduction in the extraction of aggregates (Ettlinger, 2017[16]). Figure ‎4.2 shows the historical use of secondary aggregates in the United Kingdom. It suggests that the introduction of the landfill tax in 1996 and the Aggregates Levy in 2002 contributed to a steady increase in the use of recycled aggregates as a proportion of overall aggregate consumption. In 2019, secondary aggregates constituted 27% of the total aggregate market in the country, an estimate significantly above the EU 27 average (10%) (UEPG, 2019[17]). In a recent stakeholder consultation process, carried out as part of a government review of the levy, producers of recycled and manufactured aggregates reported that they considered the Aggregates Levy to be important to incentivise the use of alternatives in place of virgin aggregates (HM Treasury, 2020[18]).

Figure ‎4.2. Recycled and secondary aggregates use in the United Kingdom

Aggregates used, by type (in million tonnes, left axis) and share of recycled and secondary aggregates in total (as percentage of total, right axis)

Source: Elaboration of data from (Mineral Products Association, 2023[19]).

Note: The UK Landfill Tax applies a lower rate to most CDW, equal to GBP 3.15 /t (compared to the standard rate of GBP 98.60 /t) (HM Revenue & Customs, 2022[20]).

4.4.2. Taxes on consumer products and plastic taxes

Product taxes can encourage consumers to purchase products that are more sustainable and circular, provided that the price signal is passed forward to consumers in the form of higher product prices. If set at an appropriate level, these taxes can elicit changes in consumer purchasing behaviour away from disposable or non-recyclable products, for example, or from products that have high disposal costs. Product taxes could also act as advance disposal fees, reflecting the cost of the end-of-life treatment of the designated products. In turn, they may incentivise businesses to bring more sustainable products to the market, which would be tax exempt.

Across OECD countries, product taxes have been widely applied to single-use plastic bags, single-use packaging and disposable tableware, fertilisers, mineral oils and tyres (OECD, 2021[5]). For instance, the Danish packaging tax, levied since 1978, aims to reduce waste and increase packaging reuse and recycling rates. Denmark chose to internalise packaging waste management costs through this tax instead of setting up an industry-run producer-responsibility scheme as implemented by many EU countries.

A handful of countries have instituted taxes on products considered harmful to the environment. One example was the Belgian “picnic tax”, an environmental tax on single-use applications. The tax had differentiated rates across four categories of single-use products: plastic carrier bags (EUR 3/kg), plastic foil (EUR 2.70/kg), aluminium foil (EUR 4.50/kg) and plastic cutlery (EUR 3.60/kg) (Eunomia and IEEP, 2016[21]). It remained in place from July 2007 to January 2015, when it was dismantled on the grounds that it had successfully changed consumers’ behaviour and because the tax revenues barely covered the implementation costs (De Muelenaere, 2014[22]). Later studies suggested that a pre-existing voluntary agreement for mandatory pricing of single-use plastic carrier bags (instead of free provision) was likely the main driver for the drop in consumption, while for other targeted single-use goods demand reductions were limited (Cornago, Börkey and Brown, 2021[23]).

In light of the growing awareness of the negative externalities caused by plastic products along their life cycle, in recent years several OECD countries have explored the use of environmental taxation applied to plastics. These may vary widely in scope, design, implementation and enforcement mechanisms, and will depend on the intended environmental, behavioural and economic outcomes. It is possible that more EU countries will consider introducing plastic taxes to raise funds to pay the EU plastics own resource. Selected experiences include the following:

• The UK Plastic Packaging Tax became operational in April 2022. It is a national level plastics packaging tax (of GBP 200/tonne) levied on plastic packaging containing less than 30% recycled plastic.

• In 2023, Spain introduced a weight-based tax on non-reusable plastic packaging (EUR 0.45/kg) (Svatikova, Brown and Börkey, Forthcoming[3]). The Spanish plastic tax only applies to the content of virgin plastic packaging, similarly to the foreseen Italian plastic tax (discussed in Chapter 5 of this report).

• By 2024, at least 25 OECD member countries will have introduced a policy for a minimum recycled content requirement of plastic products or packaging, with the United States enforcing a non-compliance penalty of approximately USD 0.44/kg (Brown and Börkey, 2024[24]).

Due to the emerging nature of plastic taxes, and the limited empirical evidence, ex-post assessments of the economic impacts and environmental outcomes are not yet available. However, general considerations on the design and implementation of plastic taxes can be drawn from the experience of selected countries with taxes applied to single-use plastics, as well as from similar instruments, such as charges and EPR schemes (Cornago, Börkey and Brown, 2021[23]), as summarised in Table ‎4.3.

In terms of the “policy coverage” (materials or products) and “target life cycle stage” (intermediate materials or goods, products or waste), weight-based taxes on virgin material input primarily aim to encourage businesses to reduce virgin material use and create demand for recycled plastics. In contrast, taxes on selected plastic products primarily target consumers’ behaviour to reduce the use of single-use packaging (not only plastic packaging). The “price signal” should be high enough to provide a strong incentive for the desired behaviour (e.g. reducing resource use, shifting to recycled materials or investing in new technologies), while not excessively penalising economic actors or consumers.

Incremental implementation and stakeholder engagement can help overcome implementation barriers and reduce administrative costs, while ensuring that the tax remains sufficiently ambitious to encourage behavioural changes among firms and consumers. Although technically not taxes, EPR fees may reflect similar price signals to producers as plastic taxes and are generally easier to implement or adjust due to their higher acceptability by the public and firms. Taxes may have revenue-raising purposes (e.g. to pay for the EU plastics own resource) or they may be earmarked partially or fully, for instance, for environmental protection causes.

Table ‎4.3. Summary of key aspects to be considered in the design of plastic taxes

Aspect		Advantages	Disadvantages
Target life cycle stage	Material input and production	Can incentivise the shift to materials not targeted by the tax (e.g. recycled materials, substitution alternatives). Can incentivise changes in design and production that lower materials use. Lower administrative costs (where there is a limited number of plastic producers/importers).	Does not differentiate packaging from durables. Does not differentiate by application (e.g. food packaging vs medical uses). Higher likelihood of lobbying resistance.
	Consumption	Helps to target products and applications that are particularly harmful (e.g. packaging over durables). More salient to consumers (and behavioural change).	Higher administrative costs. Implications for trade and competitiveness of domestic industry.
	Waste	Creates incentives for waste prevention. Can incentivise behavioural change. Can be levied on consumers or producers. May bear lower implementation costs where implemented as part of existing instruments (e.g. EPR fees).	May involve implementation challenges upfront (e.g. PAYT schemes). May not easily translate into incentives for better eco-design.
Policy coverage	Materials	Provides incentives for reductions in virgin materials use. Can incentivise the shift to recycled materials. Can target the most harmful resins.	May create incentives to shift to materials with a higher environmental footprint.
	Products	Can better target particularly harmful applications (e.g. single-use products).	Can mitigate the adverse effects of material substitution.
Price signal	Estimation of economic cost of negative externalities	Can better prompt behavioural change from consumers. Can alter production patterns in the long term.	Involves higher estimation complexity. May not always be sufficient for behavioural change (if the cost of environmental externalities is lower than WTP).
	Estimation of consumer willingness to pay	Allows behavioural change when the price signal is above the WTP.	Consumer WTP varies over time and it may adapt to the tax over time (unless the tax rate is adjusted accordingly).
	Ad hoc basis	Easier implementation (does not require estimation of costs or consumer WTP).

Note: WTP stands for willingness to pay

Source: (Cornago, Börkey and Brown, 2021[23]; Zero Waste Europe, 2018[25]).

4.4.3. Household waste disposal charges, including PAYT

Pay-as-you-throw (PAYT) is a scheme for household waste charges where fees are modulated according to the amount of waste generated. Generally, waste fees are based on a fixed and a variable fee component to reflect the cost structure of waste management and thus align incentives for users (i.e. a lower fee when less waste is produced) and waste collectors (i.e. revenue stability from the fixed fee component). The overall aim of PAYT is to enact the “polluter pays principle” and lead to enhanced separate waste collection and reduced waste generation.

The experience of OECD countries with using PAYT generates numerous best practices in the design and implementation of this instrument, as discussed in Box ‎4.1. For instance, best practices include charging household waste fees based on a fixed and a variable, weight-based fee component, and to combine these with door-to-door separate collection of waste. Good performances are also achieved with prepaid bag schemes, which are generally less costly to implement than weight-based systems (Dijkgraaf and Gradus, 2004[26]; OECD, 2006[27]; Watkins et al., 2012[28]). Key factors for a successful implementation include the measurement of waste at individual user level, the definition of a unit price that effectively drives behavioural change, and the engagement of citizens to ensure a correct understanding of the scheme’s features and widespread buy-in.

Box ‎4.1. Experience of selected OECD countries with PAYT

Flanders, Belgium

Flanders has implemented PAYT-based charges for all municipal solid waste on a regional level since 1995. Flemish municipalities are responsible for the collection and treatment of municipal solid waste produced within their jurisdiction. The Public Waste Agency of Flanders (OVAM) manages the PAYT system in collaboration with the municipalities or associations of municipalities. The region mandates municipalities to set the variable price for residual waste collection between EUR 0.11 and EUR 0.33 /kg (i.e. for a single bag of residual waste of 10 kg, the cost could go up to EUR 3.3). The regional PAYT system also relies on the differential pricing of residual waste, recyclables and bio-waste, with higher rates for residual waste (OECD, 2006[29]).

Initially, the scheme was a pay-per-bag pricing scheme, using stickers or standardized waste bags. Later, Flemish municipalities gradually (and voluntarily) shifted to a more sophisticated weight-based pricing scheme, whereby each household receives a waste bin marked by an identification chip. The weight of the bin is measured and automatically recorded before and after emptying, so that households are charged for the actual amount of disposed waste.

As a result of this change, the region saw a significant increase in the rate of separate collection (71% in 2012) and a decrease of residual municipal solid waste generation (approximately 21.4% in one year), while incineration and landfilling were virtually abandoned (ACR+, 2014[30]; Sasao, De Jaeger and De Weerdt, 2021[31]). Evidence from quantitative analysis indicates that weight-based PAYT schemes had a significant impact on decreasing residual municipal solid waste generation immediately after the scheme’s introduction, but this impact tends to level off and disappear in the ensuing years (De Jaeger and Eyckmans, 2015[32]). Possible reasons behind this may include a sort of “policy fatigue”, relatively low charges, a difficulty for households to anticipate economic savings (as compared to the pay-per-bag system), and the fading of public communication campaigns over time (Sasao, De Jaeger and De Weerdt, 2021[31]).

Key success factors in the implementation of PAYT schemes in Flanders included:

1. A clear legal framework, covering the mandatory separate collection of municipal solid waste on a regional level, a strong enforcement framework with fines to deter illegal activity, and the removal of legal obstacles.

2. Financial support in the form of subsidies for municipalities to purchase the required infrastructure (at least EUR 50 million was subsidised over the years, and the costs of PAYT schemes were partly covered by general taxes).

3. Supporting policy measures, such as landfill and/or incineration taxes and bans, which make recycling financially more attractive.

4. Wide-scale information and awareness raising campaigns informing citizens on the principles and benefits of PAYT, and the gradual implementation of the scheme once waste sorting is accepted.

5. Continuous and active dialogue with municipalities and key actors along the value chain (ACR+, 2014[30]).

Republic of Korea

The Republic of Korea was the first country to adopt PAYT on a nationwide basis in 1995, using a volume-based scheme with pre-paid bags for residual waste and charges for the collection of bulky waste. Individual municipalities, however, set different charges that were gradually increased over time. As a result, residual waste amounts significantly declined, and recycling rates more than doubled between 1994 and 2004 (OECD, 2006[29]).

Preparation prior to the full-scale implementation of the system started years in advance, with feasibility studies and pilot projects undertaken in 1992-1994, The government announced the introduction of the PAYT system in 1993, and large-scale communication campaigns were carried out during the test period. The government also prepared guidelines to support municipalities in the implementation of the PAYT scheme, providing indications on how to address issues that emerged during the test period, and from consultations with stakeholders (Yu, n.d.[33]).

In Seoul, the PAYT scheme is based on different charges and measuring methods for both residual waste and food waste. Residual waste is measured by the size (volume) of the pre-paid bag, whereas food waste can be either measured by weight (through radio-frequency identification [RFID] based weighing systems) or volume (using bins with chips or stickers for standardised bags). Stickers must also be used for the collection of bulky waste, or it can be handed to a specialised waste collection agency. Recyclables (i.e. paper, glass, scrap metal and plastics) are collected separately, free of charge. Moreover, vulnerable (poor and low-income) groups are entitled to discounted rates or exemptions. The implementation of the scheme in Seoul resulted in a 11% decline in municipal solid waste generation within two years and clear changes in consumption patterns (favouring the use of recyclable material), which in turn were reflected in product design (Yu, n.d.[33]).

A major initial issue linked to the PAYT scheme was illegal waste dumping, which was addressed through a system of fines and incentives to report illegal activity, leading to the gradual and steady decline in the number of incidents. Another issue was the growing amount of collected recyclables, while demand levels remained unchanged. This was resolved with the 2003 regulation on EPR as well as government’s support for the development of recycling operating facilities for plastics and the public sector’s preferential purchase of recycled plastic products. Lastly, the initial increase in landfilled food waste was tackled by integrating food waste in PAYT-based charges, a landfill ban in 2005, and the scaling up of food waste treatment facilities. Initially, the landfilled food waste stream was not included under PAYT charges and the material used for the standardised bags contaminated its quality, making it unsuitable for compost (Yu, n.d.[33]).

4.4.4. Landfill and incineration taxes

Landfill and incineration taxes are commonly employed in OECD countries to help steer waste away from landfilling and incineration and to encourage investments in recycling infrastructure. This section looks specifically at lessons to be learned from the experience of the UK Landfill Tax that has been in place since 1996. The tax is charged by weight, with a standard rate and lower rate applying to inert or inactive waste, as shown in Table ‎4.4. A “tax escalator” was introduced in 2015 to regularly update tax rates and maintain the effectiveness of the tax. Tax rates are adjusted yearly for inflation, and the standard rate cannot fall below GBP 80/tonne.

The UK Landfill Tax had positive environmental outcomes overall. Between 1998 and 2014, HM Treasury increased the standard rate of the landfill tax by 700% in real terms, contributing to a 65% fall in total waste sent to landfill over the period (National Audit Office, 2021[34]). The revenue raised from the landfill tax was around EUR 277 million in 1997/1998 and rose steadily to a peak of around EUR 1.02 billion in 2013/2014, in line with increasing tax rates. As the tax became effective in reducing landfilling, revenues dropped to around EUR 667 billion in 2021/2022 (Eunomia, 2022[35]).

Table ‎4.4. UK Landfill Tax rates

	Rate from 1 April 2022	Rate from 1 April 2021	Rate from 1 April 2020
Standard rate	GBP 98.60 per tonne	GBP 96.70 per tonne	GBP 94.15 per tonne
Lower rate	GBP 3.15 per tonne	GBP 3.10 per tonne	GBP 3.00 per tonne

Source: (HM Revenues & Customs, 2022[36]).

Since 2003, most of the revenue from the landfill tax went to the government’s general budget. In addition, a part of the revenue was allocated to support environmentally and socially beneficial projects, such as closing down unauthorised landfill sites and boosting jobs in other areas of waste management. Additionally, about 10% of the total revenue was channelled to the Landfill Communities Fund to finance environmental regeneration and remediation projects, maintain public amenities and promote biodiversity, thus helping the United Kingdom deliver on its Biodiversity Action Plan (Eunomia, 2022[35]).

While the UK Landfill Tax reduced landfilling overall, it generated perverse incentives for illegal waste disposal (Eunomia, 2022[35]). The number of tax exemptions enabled the misclassification of waste at authorised landfill sites for the financial gain of the landfill site operator. Furthermore, the dual tax rates created an incentive for misclassification. HM Revenues and Customs estimated that the misclassification of waste at authorised landfill sites, as well as waste disposed at unauthorised sites, reduced revenues by around GBP 275 million (28% of tax due) in 2018-2019 (National Audit Office, 2021[34]). Additionally, increases in tax rates caused more waste crimes, such as fly-tipping, illegal dumping, exemption breaches and illegal exports.

One critical issue was insufficient enforcement and compliance (Eunomia, 2022[35]). Enforcement fines were considered low compared to the tax. Additionally, the Environment Agency, responsible for enforcing the tax in England, had limited funding and capacity to robustly monitor and enforce the tax. It was suggested that part of the revenue earned from the landfill tax could be ring-fenced for monitoring and enforcing the tax, in addition to investments in waste management.

In addition, the implementation of the landfill tax was not accompanied by an incineration tax (Eunomia, 2022[35]). The lack of an incineration tax incentivised the switch from landfill to incineration instead of the shift from landfill to recycling due to the lower costs generally of incineration compared to recycling. If an incineration tax had been implemented alongside the landfill tax, an even greater increase in recycling would have been expected. Additionally, supporting measures, such as the implementation of a PAYT scheme for household waste collection, could contribute to achieving circularity gains.

4.4.5. Extended Producer Responsibility

Extended Producer Responsibility (EPR) is an environmental policy approach whereby a producer’s responsibility for a product is extended to the post-consumer stage of the product’s life cycle. In practice, EPR involves producers taking responsibility for collecting end-of-life products, and for sorting them before their final treatment, ideally by recycling (Lifset, Atasu and Tojo, 2013[9]). More recently, producer responsibility organisations (PROs) in some OECD countries have experimented with the use of advanced fee modulation, whereby EPR fees are modulated on the basis of relevant eco-design criteria, such as durability or recyclability, for instance, through higher and lower fees or a system of bonus/malus adjustments. The greater specificity provided by advanced fee modulation could strengthen the incentives for the eco-design of products and packaging (Laubinger et al., 2021[37]).

Selected PROs have modulated EPR fees in line with recycled material content. Generally, producers are best placed to improve the eco-design of products and packaging in line with better recyclability (hence contributing to improving the availability and quality of recycled materials supply) and by using a higher content of secondary materials in production processes. Yet, the use of secondary materials is often not the preferred choice among producers owing to the added costs related to adjusting production methods, as well as barriers such as contamination, supply risks and demand uncertainty (Wijayasundara, 2021[38]). Modulating EPR fees in line with recycled materials content can provide an additional factor to promote demand for recycled content. For example, in France, a 50% fee reduction is provided for polyethylene (PE) and polypropylene (PP) packaging with at least 50% recycled content (CITEO, 2019[39]; Laubinger et al., 2021[37]).

Table ‎4.5. Examples of EPR fee modulation for packaging according to recycled content criteria

Country (State/Province)	Examples of recycled content criteria
France	EPR fees for packaging include: 10% fee reduction for cardboard and graphic paper (in publications) with > 50% recycled content. 5% fee increase for using primary fibres from forests without eco-management labels. 50% fee reduction for textiles and shoes with 15% recycled fibres/materials (EY, 2016[40]). 50% fee reduction for PE and PP containing at least 50% recycled content (CITEO, 2019[39]).
Germany	The 2019 Packaging Ordinance requires PROs to provide incentives for sustainable packaging design and to modulate EPR fees accordingly. PROs are required to design fees that include differentiating fees along criteria of, among others, recyclability (given existing technologies) and recycled content, and content of renewable materials (BMJV Germany, 2019[41]).
Canada (Quebec)	EPR fee modulation is applied to packaging, inspired by the French bonus/malus scheme system. This involves a 20% bonus for producers who entirely manufacture packaging with recycled content and who use at least 50% to 80% of recycled content for printed materials (e.g. magazines and other publications) (EEQ, 2020[42]).
United States (California)	State law establishing EPR for carpets requires a difference in fees for the presence of post-consumer recycled content (California Legislative Information, 2020[43]).
Chile	Collective management systems for packaging must modulate fees with bonus or malus based on recycled content, if the secondary material is derived from waste generated in Chile (Ministerio del Medio Ambiente Chile, 2021[44]).

Source: (Laubinger et al., 2021[37]).

4.4.6. Public procurement

Public procurement is a crucial pillar of strategic governance and services delivery for governments. It accounts for 12% of GDP across OECD countries and 14% of the EU GDP (OECD, June 2022[45]; European Commission, 2022[46]). Given its economic significance, a strategic use of public procurement offers considerable opportunities to promote and accelerate the transition to a more circular economy by creating demand-side pressure in markets for resource efficiency, waste reduction and longer product lifespans. Procurement strategies and practices can influence all value chain stages, including consumption choices, eco-design for circularity, and the end-of-life phase.

In OECD countries, public procurement is increasingly used as a policy lever to promote sustainable production and consumption. Procuring in an environmentally friendly way involves looking beyond short-term needs and considering the longer-term impacts of each purchase. Particular attention is paid to the potential of green public procurement, and increasingly to circular procurement approaches.

• Green public procurement (GPP) usually refers to how public procurement can support the implementation of environmental objectives by purchasing products, services and public works that are less environmentally damaging. In the EU, GPP is defined as “a process whereby public authorities seek to procure goods, services and works with a reduced environmental impact throughout their life cycle when compared to goods, services and works with the same primary function that would otherwise be procured” (European Commission, 2008[47]).

• Sustainable Public Procurement (SPP) has a broader scope than GPP as it promotes the integration of all three pillars of sustainable development: economic development, social development and environmental protection.

• Circular Public Procurement (CPP) is the process by which public authorities purchase works, goods or services that seek to contribute to closed energy and material loops within supply chains, while minimising and, in the best case, avoiding negative environmental impacts and waste creation along the life cycle (European Commission, 2020[48]).

While GPP and SPP can be considered as more product- and technology-oriented, CPP focuses on choices and behaviours that lead to the retainment of the maximum value of products, components and materials. A circular procurement process focuses on changing the functional needs of the users within the organisation and retaining the maximum value of products, components and materials. Circular procurement can be implemented through different approaches, including the inclusion of circularity criteria in GPP, the procurement of innovative circular products, the procurement of services, and new business concepts that support the circular economy (Alhola et al., 2017[49]).

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Note

← 1.  The use of the terms “tax” and “charge/fee” varies across countries. One distinction, which is reflected in OECD statistics on government revenues, is to reserve the word "charge/fee" for an instrument where the payment leads to a direct individual entitlement to something in return (e.g. water), while a "tax" is a compulsory levy which does not give the individual payer a specific entitlement in return (OECD, 2017[50]). No significance should be attached to any differences in the use of the terms in this report.