This chapter explores how synergies and trade-offs between responses to climate change, biodiversity loss and pollution manifest themselves at the design and implementation stages at various geographical scales and how they can be managed. These are explored through four deep-dives on: (i) renewable energy expansion, (ii) management and expansion of protected areas, (iii) air pollution control, and (iv) nutrient management.
Environmental Outlook on the Triple Planetary Crisis
Stakes, Evolution and Policy Linkages
Abstract
The interlinkages between climate change, biodiversity loss and pollution play out at multiple scales. Preceding chapters have provided a conceptual overview of synergies and trade-offs between policy objectives to tackle these challenges (Chapter 4) and explored how these considerations are currently reflected in national documents (Chapter 5). Yet, some of the key trade-offs and synergies between different policy objectives are fundamentally local, even if they ultimately affect national or even global outcomes. Thus, policy effects have to be assessed at the relevant levels, to allow for all major synergies and trade-offs to be appropriately accounted for and measured.
To grasp these interactions at different geographical scales, this chapter provides deep-dives into the existing tools and policies that can help harness synergies and manage trade-offs during design, implementation, and operation by exploring policy responses in four key areas (Figure 6.1): (i) renewable energy expansion, (ii) management and expansion of protected areas, (iii) air pollution control and (iv) nutrient management. Each of these cases examined has a primary objective (climate change mitigation, biodiversity conservation and pollution control1 for the latter two, respectively), but with significant interactions with the other objectives. By adjusting existing modalities and policy design, the primary policy objective can be aligned with the other objectives to reduce trade-offs and reap synergies.
Source: Authors’ own elaboration.
These four topics have been selected because their implementation is inherently local in nature, yet they span key policy responses across the three objectives, with significant interaction effects on the other objectives.2 Nonetheless, they differ crucially in how they are interlinked. Each can serve as a clear example of the different types of interlinkages, providing additional insight into the various core trade-offs and synergies. Together, the policy responses in the four areas highlight that a core response to one objective may risk trade-offs with other objectives, but may also spur synergies, depending on how they are implemented.
First, the large-scale deployment of wind and solar photovoltaic (PV) energy generation is key to climate change mitigation. However, despite their clear climate advantages the deployment of these technologies can negatively affect biodiversity and exacerbate pollution during (i) the extraction and processing of materials and manufacturing, (ii) the installation and operation stage, and (iii) the decommissioning and waste disposal stage. Anticipating these locally specific impacts as well as the upstream and downstream impacts can help alleviate the risks that the renewable energy expansion – critical for reducing climate-induced pressures on biodiversity and reducing pollution associated with fossil fuels – is stalled due to unintended adverse environmental impacts or due to lengthy permitting procedures to evaluate and mitigate such impacts.
Second, protected areas are a cornerstone of biodiversity conservation policy and are, via their positive spillovers on climate change mitigation, adaptation and pollution control, a key tool to address the interconnected environmental challenges together. At risk of declining, degrading or becoming less effective, and more vulnerable to climate change and pollution, integrating these linkages in the selection, design and management of protected areas is essential to preserve their conservation goals.
Third, air pollution control policies have been widely implemented and are fundamental to improve air quality at the local, national and global scales. Air pollutants have extensive impacts on terrestrial and aquatic biodiversity. Conversely, the loss of the pollutant filtration function due to ecosystem degradation leads to increased air pollution. Moreover, air pollutants can have warming (or cooling) effects through their impact on cloud properties, precipitation, and radiative forcing, while climate change affects both natural and anthropogenic sources of air pollutants and can lead to either lower or higher air pollution depending on climatic variables. Thus, policies aiming at preventing or reducing air pollution must consider the feedback loops of climate change and biodiversity in their design and implementation.
Fourth, releases of excess nutrients – not least nitrogen and phosphorus – to the environment can contribute to climate change, pollution in air, water, and soil, as well as affect ecosystems and drive biodiversity loss. Regulating excess nutrients requires an integrated approach across sinks as well as sources, considering the full nitrogen and phosphorus cycles. This includes considering the different nitrogen compounds, such as nitrate (NO3-) leaching in the soil and nitrous oxide (N2O) emissions in the air. Furthermore, as phosphorus is a non-renewable resource, it links directly to resource management.
Often, the ex-ante ambitions of policies when designed are not fully achieved upon implementation, for a variety of challenges and barriers. This “implementation gap” implies that while synergies between the various dimensions can be well-covered in policy design, they may be muted in reality. Specific attention is therefore paid to the available tools for effective implementation of synergistic approaches, including through measurement of impacts. The chapter concludes with key takeaways that can be distilled from the detailed analyses of the four deep-dives.
The accelerated deployment of renewable energy, in particular wind and solar power, constitutes one of the core responses to climate change. In light of the United Nations Framework Convention on Climate Change (UNFCCC) COP28 pledge for tripling the capacity of renewable energy by 2030, many governments have announced their targets to expand the share of renewable energy by 50% or more between 2023 and 2030 (IEA, 2024[1]). To meet these targets, countries are adopting a suite of market-based (e.g. feed-in tariffs) and non-market-based (e.g. renewable portfolio standards) policies. Under current policies, solar and wind power is projected to represent 37% of global electricity in 2050 (up from 9% in 2020; see Chapter 2).
Renewables expansion is integral to reducing greenhouse gas (GHG) emissions, which in turn can help alleviate climate-induced pressures on biodiversity. To the extent that it replaces fossil fuels, it is also critical for addressing pollution. Nonetheless, wind and solar power can incur new environmental pressures upstream (e.g. high material requirements) and downstream (e.g. limited recyclability of components), while posing threats to biodiversity through their spatial requirements and operation.
6.2.1. Synergies and trade-offs of renewable energy expansion, biodiversity conservation and pollution control
Synergies between renewable energy expansion, biodiversity conservation and pollution control
To the extent that renewables can reduce the reliance on fossil fuels as the primary source of energy, expanding renewables is key to avoiding the adverse environmental impacts associated with the lifecycle of fossil fuels (see Box 6.1 for more details). The transition away from fossil fuels can significantly reduce GHG emissions and address anthropogenic climate change. Renewables expansion also contributes to biodiversity conservation indirectly (by mitigating climate change, the fastest growing driver of biodiversity loss; see also Chapter 3) and directly (by reducing pressures on biodiversity from of fossil fuel extraction) (OECD, 2024[2]). This is an important consideration, particularly given that the terrestrial and marine environments in which oil and gas infrastructure are currently located in areas which have higher species richness and range rarity compared to where no exploitation was taking place (Harfoot et al., 2018[3]).
Renewables expansion can alleviate pressures from the fossil fuels extraction, which can degrade, fragment and destruct natural habitats. From clearing of forests for mining to disruptions of deep-sea ecosystems through seismic survey and oil and gas drilling, the biodiversity impacts of fossil fuels span across various types of ecosystems. Coal mining is associated with an especially high biodiversity loss compared to other energy sources (Holland et al., 2019[4]) because it can substantially alter the land surface and subsurface by disrupting deep soil and lithologic structures, resulting in a near-complete loss of terrestrial and wetland habitats for mammals and amphibians in the area (McManamay, Vernon and Jager, 2021[5]; Wickham et al., 2013[6]).
Wind and solar energy expansion can also indirectly reduce the distinct risks associated with transport of fossil fuels. For instance, historical oil spills have resulted in mortality of species, with enduring ecological impacts due to the persistence of toxic compounds of oil and impacts of restoration actions on marine and terrestrial wildlife (Barron et al., 2020[7]). Through tidal currents and wind, oil spills also affect coastal shorelines that provide erosion protection and typically serve as habitats for rich assemblage of species (Asif et al., 2022[8]).
Shifting from fossil fuels to renewables can also help reduce various forms of pollution, such as air pollution. From the extraction and refining processes to combustion, fossil fuels are responsible for nitrogen oxides (NOx), sulphur oxides (SOx), volatile organic compounds (VOCs) and particular matter (PM). Oil, gas and coal are also associated with high methane (CH4) emissions, both a potent GHG and a precursor to tropospheric ozone, also known as ground-level ozone (O3). Of the total 130 million tonnes (Mt) of CH4 emissions3 associated with the energy sector (more than a third of total emissions associated with anthropogenic activities), oil operations (50 Mt), coal (40Mt) and natural gas supply chain (30 Mt) contribute most significantly to CH4 emissions of the sector (IEA, 2024[9]).
Renewable energy expansion can also deliver synergies by reducing the adverse impacts of fossil fuels on water pollution. For example, coal mining incurs the risk of land subsidence and potentially alters the topography as well as groundwater quality (Rouhani, Skousen and Tack, 2023[10]). Acid mine drainage from surface runoffs, seepage from stockpiles, dumps and open pits contaminate water as well as soil (Simate and Ndlovu, 2014[11]). Fly ash and sludge from coal can contain heavy metals including mercury, arsenic and lead (Munawer, 2018[12]), while flowback water from hydraulic fracturing for oil and gas can be highly saline unless treated appropriately (Caldwell et al., 2022[13]).
Fossil fuels can cause adverse impacts on biodiversity and result in pollution across different environmental media throughout their lifecycle from extraction, processing and refining, transport and storage, combustion and waste management (Table 6.1). Like-for-like comparisons of the environmental impact of different types of fossil fuels and wind and solar energy are challenging. While the total environmental impacts may not be directly comparable due to different material and process requirements across energy systems, methodologies such as lifecycle assessments (LCAs) and material flow analyses can provide quantitative estimates of (a subset of) environmental impacts in comparable units.
Various studies attest to the lower environmental impacts of renewables on several metrics. A study analysing the impact of doubling the renewable electricity generation between 2005 and 2018 in the EU suggests that impact potentials of renewables are lower for eutrophication, PM formation as well as acidification (European Environment Agency, 2021[14]). The requirements of the surface and ground water withdrawal and consumption for solar and wind power are also significantly lower than those of fossil fuels (Jin et al., 2019[15]). For instance, lifecycle water consumption of oil per megawatt hour are estimated to be about 10 times larger than solar PV and 75 times larger than wind (Jin et al., 2019[15]). Importantly, adverse human health impacts are also considerably lower for renewables (UNECE, 2022[16]).
However, these studies also suggest that renewables carry some environmental risks that need to be carefully managed. These include higher freshwater ecotoxicity and land occupation impacts (European Environment Agency, 2021[14]). Although there is a large variation in the estimates of the land-use impacts, existing studies highlight that spatial requirements can be significantly higher for wind and solar power than for fossil fuels per unit of electricity produced. For instance, a study comparing the land-use intensity (measured by hectares per terawatt-hour of electricity) of a large sample of energy facilities in operation concludes that land-use intensity is higher for concentrated solar power and ground-mounted PV than for coal, despite the large variability (Lovering et al., 2022[17]). The potential land-use impacts underscore the importance of leveraging various spatial planning and land-sharing measures (see Section 6.2.2).
Source: Authors’ own elaboration based on the cited references.
Trade-offs between renewable energy expansion and biodiversity conservation
While the benefits of renewables expansion discussed in the previous section are paramount, the deployment of renewables is not entirely without environmental cost. Without adequate consideration, it risks posing threats to biodiversity and causing pollution at different stages (upstream, construction and operation and decommissioning) and across different geographical scales (Kiesecker et al., 2019[28]) (key impacts are summarised in Figure 6.2). There is also a cross-border dimension to the impacts of renewables expansion if the part of the supply chain for renewable energy equipment – such as mining, extraction, processing, and manufacturing – is not located in the same country or region where the renewable energy is used.4
Note: The most significant impacts are reported here. Environmental impacts not included in the figure include light pollution, shadowing from turbines and the potential introduction of invasive alien species due to movement of people, construction material and equipment.
Source: Authors’ own elaboration.
Although the trade-offs between renewable energy expansion and biodiversity are highly site-specific and vary significantly depending on the local climate and management (IUCN, 2022[29]; Carvalho et al., 2024[30]), there is a wide range of risks for adverse impacts on biodiversity (OECD, 2024[2]).5 Collision with wind turbines can result in mortality of certain avian species (Thaxter et al., 2017[31]). Soaring birds (e.g. raptors and other large-winged species) can be particularly vulnerable to collision due to their flight altitude in the rotors’ swept zone, reduced flight manoeuvrability for avoidance due to their size, as well as their reliance on thermal updrafts and orographic uplift (and hence the preference for windy slope regions such as mountain ridges) (Santos et al., 2022[32]).
While bird mortality due to wind turbines is considered small in magnitude relative to other anthropogenic causes (Loss, Will and Marra, 2015[33]), even a small number of fatalities can affect the population dynamics of long-lived species with low reproductivity. For example, increased fatalities of raptors due to wind turbines have been documented around the world (Estellés‐Domingo and López‐López, 2024[34]). However, there are also some concerns that the impacts of smaller passerine birds may be underestimated due to difficulties in identifying fatalities through methods commonly deployed such as carcass surveys (Nilsson et al., 2023[35]). Beyond direct impacts on fatalities through collision, bird species are also affected indirectly through displacement and habitat alterations and degradation (Shaffer and Buhl, 2016[36]; Reusch et al., 2022[37]). For instance, key foraging and breeding habitats for white-tailed eagles often overlap with optimal locations for setting up wind turbines. Nebel et al. (2024[38]) suggest that installing them within 5 km of white-tailed eagle nests in Finnish coastlines has resulted in 7.6% lower survival rates.
Wind turbines are identified as the leading cause of multiple mortality events for bats (O’Shea et al., 2016[39]). High incidences of bat fatalities have been recorded across continents, and available estimates suggest that incidents of migratory bat fatalities by wind turbines could lead to their drastically smaller population size (Frick et al., 2017[40]). Beyond direct habitat loss on the turbine sites and mortality impacts, recent evidence also shows that bats are avoiding large areas around wind turbines because of the generated wake turbulences and noise, restricting their habitats further (Frick et al., 2017[40]). For example, Barré et al. (2018[41]) estimated that the activity of some types of bat ensembles declined by over half on European farmland up to 1 km in distance from wind turbines. The direct and indirect impacts of wind energy on bats pose a serious challenge to local and global bat diversity (Voigt et al., 2024[42]) and to biodiversity more broadly, given the ecological importance of bats and the key roles they play in pest control, pollination, and seed dispersal across many ecosystems (Jones et al., 2009[43]).
Offshore wind farms are also associated with varying impacts on biodiversity. While the impacts of construction are predominantly negative, operational impacts can depend on the site and species, and there remains much uncertainty (Watson et al., 2024[44]). Studies have documented some positive impacts in terms of species abundance as the submerged infrastructure can create artificial reefs for certain benthic and soft-sediment fish species, and can help prevent bottom trawling in the vicinity (Buyse et al., 2022[45]; Degraer et al., 2020[46]). However, existing studies also highlight the risk of colonisation of windfarm infrastructure by invasive alien species (Watson et al., 2024[44]). While offshore wind farms have often been set up as fixed structures in shallow waters, expansion further offshore through floating technologies are also gathering pace (Bennun et al., 2021[47]). As offshore floating structures are anchored on the deep sea floor, they may also have uncertain impacts on deep sea ecosystems on which there is limited knowledge (Danovaro et al., 2024[48]).
Land-based solar PV also affects biodiversity through changes in habitats, although their effects are not as well understood as for wind power. Installing solar PV farms on crop lands or wooded areas alters or destructs natural habitats and provokes changes in hydrology, water availability and quality (Hernandez et al., 2015[49]). Fences around solar PV farms create physical barriers, further contributing to habitat fragmentation (McInturff et al., 2020[50]). Solar PV farms installed on natural land also generate shading that alters the composition of plant and animal species as well as the provision of ecosystem services such as carbon sequestration in the soil and pollination (Walston et al., 2018[51]; Grodsky and Hernandez, 2020[52]; Graham et al., 2021[53]). In addition to the impacts on habitats, solar PV farms can also alter the behaviour of animal species by drawing birds and bats to the large concentration of insects that mistake panels for water surfaces due to their reflective effects on light (Horváth et al., 2009[54]). Finally, bird and bat mortality from collision with solar PV infrastructure and exposure to concentrated sunlight has been documented, although it is likely lower than mortality from fossil fuel facilities, wind turbines and power lines (Walston et al., 2016[55]).
Importantly, both wind and solar power require electricity transmission, distribution and energy storage infrastructure to meet the needs of urban areas, typically distant from renewable energy sources. The impacts of transmission and distribution lines on biodiversity loss are well documented, particularly for birds (OECD, 2024[2]). Power lines are estimated to kill a large number of birds every year by collision and electrocution (Loss, Will and Marra, 2015[33]), with vulnerability factors including their body weight, size of wings, flight behaviour, and flight routes (Bernardino et al., 2018[56]). Electrocution of birds can also ignite wildfires (Guil et al., 2018[57]). Vegetation is cleared in the safe zone below transmission lines (i.e. “right of way”), and can become colonised by invasive plant species (Biasotto and Kindel, 2018[58]). However, knowledge gaps remain, particularly for impacts on non-avian species (e.g. amphibians), as well as for disruption of habitats and ecosystems from distribution lines and underground cables (OECD, 2024[2]).
Trade-offs between renewable energy expansion and pollution control
Among the most significant pollution challenges associated with solar PV and wind turbines is waste management of decommissioned equipment at their end of life. Cumulative solar PV panel waste could increase from 250 kt in 2016 to 8 Mt in 2030 and 78 Mt in 2050 under a scenario of early attrition (IRENA and IEA-PVPS, 2016[59]). Crystalline polysilicon panels that account for 98% of the global market share consist around 70% of glass (IEA, 2022[60]), which is a relatively low value material that makes solar panel recycling relatively expensive, increasing the risk of unsustainable disposal.
Disposal of end-of-life panels in landfills can be problematic since they can contain materials such as lead and cadmium that pose risk both to human health and the environment (Li et al., 2020[61]; Kwak et al., 2020[62]). Disposal in landfills also constitutes a missed opportunity to recover materials such as aluminium, copper, silver and zinc, further driving demand for primary materials and the environmental damages associated with mining and processing. Mismanagement of end-of-life panels is already a prominent issue in some developing countries, where a growing number of small standalone solar devices with relatively short lifetimes are being improperly disposed of due to a lack of appropriate collection and recycling infrastructure (ACE, 2021[63]). The mismanagement of solar panel waste in countries without adequate waste management systems can also be amplified by illegal trade in discarded solar panels, which remains difficult to quantify (Nyffenegger et al., 2024[64]; Huisman et al., 2015[65]).
Recycling of end-of-life solar panels is technically complex due to the heterogeneity in size and composition (IEA, 2022[60]). Solar panels have typically been designed for their durability rather than recyclability, with sandwich-like structure that makes sorting and separation of materials challenging (IEA-PVPS, 2021[66]). Economic viability can be compromised due to high recycling costs and limited recovery of homogeneous and high-purity content of materials.
Similarly, the development of wind power is projected to generate 43 Mt of cumulative wind turbine blade waste by 2050 (Liu and Barlow, 2017[67]). Wind turbines are made of three main parts: the tower (primarily made of steel), rotor (a hub and three blades made of composite materials), and nacelle (contains gearbox, generators, controller and brakes). Most components of a wind turbine such as the foundation, tower, gear box and generator are already recyclable. However, due to their heavy weight and heterogeneity, wind turbine blades are not currently widely reused or recycled (Khalid et al., 2023[68]). Furthermore, developments in blade designs have resulted in lightweighting with the use of composites such as fibre-reinforced plastics that cannot be easily recycled (Shen, Guo and Feng, 2023[69]).
Landfilling turbine blades has land use implications as it requires space to store large composite components that are non-biodegradable. Blades that are not landfilled or recycled are thermally recycled or incinerated. The pyrolysis or thermal recycling of blades separates the resin from the fiberglass or carbon fibres producing gas, oil and char. However, this process is highly energy intensive and can contaminate the recycled fibres with char (ACP, 2023[70]). With the average life of 25-30 years, limited recyclability of these technologies poses an imminent challenge, as their rapid deployment in the last two decades can result in a large quantity of decommissioned solar panels and wind turbines that end up in landfills (Lichtenegger et al., 2020[71]).
In addition to end-of-life management, wind power and solar PV generate pollution during their construction and operational phase, including visual and noise pollution that has significant effects on public perception of renewable energy. The visual and aesthetic impacts of wind energy facilities on landscapes are one of the main challenges faced by wind energy development at the local level (Krekel and Zerrahn, 2017[72]), although its importance for acceptability relative to other factors is highly context specific (Rygg, 2012[73]).
Closely related to the impacts on biodiversity, heightened demand for minerals used in renewable energy technologies can result in increased pressures for water- and land-intensive mining and processing activities, pollution consequences of which extend across various ecosystems. Wind turbines are associated with substantial material requirements and intensity, both in terms of bulk (e.g. steel and cement) and critical materials. Onshore and offshore wind infrastructure use large amounts of concrete and steel (Farina and Anctil, 2022[74]); production of which is currently associated with high CO2 emissions. Furthermore, manufacturing of rare earth metals or alloys for generators entails many steps including mining and refining that emit toxic chemicals into the soil, water and air such as PM and hydrogen fluoride (Zapp et al., 2022[75]). The extraction and processing of copper used in both solar PV and wind turbines generate slags, dusts and aerosols that are contaminated with metals and metalloids, including toxic elements such as arsenic, cadmium and lead (Izydorczyk et al., 2021[76]).
Manufacturing of solar PV also involves the use of different hazardous materials to extract and purify the raw materials such as silicon or copper and to manufacture the modules. These include ammonia (NH3), arsenic, and cadmium, some of which are associated with cancer and other chronic diseases (Aman et al., 2015[77]). Some of the materials used for improving the efficiency of solar panels can also carry the risk of release of pollutants. For instance, fluoropolymer containing per- and polyfluoroalkyl substances (PFAS) is used for coating in solar panels for front sheets (to increase the among of light reaching the panel) and as back sheets (to protect the modules from moisture and degradation) (OECD, 2022[78]).
6.2.2. Integrating biodiversity and pollution control in siting of renewable energy infrastructure
Procedures informing and deciding the siting of wind and solar infrastructure, including by considering their current and alternative use of the siting areas, are critical for avoiding the risks of biodiversity loss and pollution. Countries can actively identify appropriate sites for renewable energy infrastructure based on specific site characteristics such as the proximity to existing power infrastructure or steer them away from certain areas that are important for biodiversity. Environmental Impact Assessments (EIAs) as part of the licensing and permitting procedures can also help elucidate the broader environmental impacts of a project, although there is a need to reconcile comprehensive consideration of the environmental risks with the speed and predictability of permitting process so as to avoid well-intended safeguards from becoming an impediment to the accelerated deployment of renewable energy.
Spatial planning
Aided by tools such as Strategic Environmental Assessment (SEA), a range of “analytical and participatory approaches that aim to integrate environmental considerations into policies, plans and programmes and evaluate the interlinkages with economic and social considerations” (OECD, 2006[79]), spatial planning can help identify appropriate sites for renewables projects and streamline the permitting processes for the renewable infrastructure. Spatial planning can also inform land-sharing strategies to reduce the spatial footprint of the renewable energy infrastructure. While the site-specific impacts limit the generalisability of the findings, some existing studies suggest that land-sharing can deliver synergies for biodiversity conservation and pollution control while managing the risks of trade-offs. These mechanisms include installing solar PV on degraded land to prevent habitat loss and soil degradation in natural ecosystems (Gómez‐Catasús et al., 2024[80]). For instance, France mandates that solar panels installation prioritises areas with limited biodiversity concerns such as parking, buildings, abandoned roads or railways (France, 2024[81]). Projects involving the clearing of over 25 hectares in natural, agricultural and forest areas have been systematically rejected by regional authorities, and as of 2024, it is no longer possible to apply for a permit if the clearing exceeds this threshold.
Combining solar PV setup and conventional agriculture (“agrivoltaics”) is another land-sharing mechanism that has the potential to offer synergies in terms of pollution control and management. For instance, agrivoltaics in drylands is suggested to enhance water use efficiency (by decreasing the evaporation from soil) and enhance agricultural productivity (by reducing heat and light stress on food production) while the crops or other infrastructure underneath can provide cooling effects for the panels to improve their efficiency of electricity generation (Barron-Gafford et al., 2019[82]). Harnessing these potential synergies, a project (“Ecolume”) in the semi-arid regions of Brazil aims to promote water reuse through rainwater collected on solar panel surfaces with an aquaponic system located beneath photovoltaic panels (Lacerda et al., 2020[83]). In addition to energy generation, it also contributed to the cultivation of 17 types of vegetables for local consumption as well as seedlings that were utilised for local reforestation (Vidotto et al., 2024[84]).
Agrivoltaics can also provide a means for income diversification for farmers (Mamun et al., 2022[85]). For instance, against the backdrop of aging labour force of the agricultural sector, Korea provides support for small-scale agrivoltaics projects to prevent farmers from abandoning their land and to provide an additional source of income (Schindele et al., 2020[86]). Some countries in Europe are also taking steps towards the development of technical specifications (Germany), guidelines (Italy) and a definition of agrivoltaics (France) (Chatzipanagi, Taylor and Jaeger-Waldau, 2023[87]).
However, practical policy and technical challenges for widespread deployment of agrivoltaics still remain. In particular, the dual use of land for electricity generation and agriculture brings uncertainty over the land characterisation, which may have implications for other policy domains including agricultural support (Chatzipanagi, Taylor and Jaeger-Waldau, 2023[87]). Furthermore, optimising both agricultural and power generation can be challenging, especially as the impact of solar PV on crop growth (e.g. due to shading) is highly site- and crop-specific (Asa’a, 2024[88]) and synergies cannot always be assumed (Mamun et al., 2022[85]).
Co-locating different types of renewable infrastructure with other economic activities can also help realise synergies and manage the risks of trade-offs (OECD, 2024[2]). Well-recognised synergies include the co-location of different types of renewables for resource complementarity of the infrastructure and existing grid connection. A recent study on the European Atlantic also suggests that co-locating offshore wind and solar farms, for instance, may also confer additional benefits by contributing to lower power output variability, particularly in the areas where there is a negative relation between the seasonality of solar irradiance and wind resources (Martinez and Iglesias, 2024[89]). Spatial planning and early engagement with stakeholders can also help assess the viability of co-locating renewable infrastructure with other economic activities. For instance, offshore wind farms can also provide synergies for aquaculture to improve resource use and operational efficiency (Manolache and Andrei, 2024[90]) – the research on the optimal co-location sites is expanding to aid marine spatial planning (e.g. (Stockbridge, Brown and Kuempel, 2025[91])).
Environmental permitting and use of environmental assessment as integrative mechanisms
There is an important balance to be struck between accelerating renewables expansion and ensuring the risks for adverse impacts are managed during licensing and permitting processes, including EIAs. The processes, scope and timeline of an EIA can vary significantly across countries (see Annex 6.A for more details and examples of requirements for renewable energy projects). Existing regulations can require additional steps for permitting. In the European Union, the EU Habitats Directive prohibits the destruction or disturbance of the certain animal species and their habitats, although derogations can be made under certain circumstances by authorities in member countries.
Similarly, in Ontario, Canada, under the province’s Environmental Protection Act, a renewable energy generation facility cannot be installed in a provincial park or a conservation reserve, unless permitted by the Provincial Parks and Conservation Reserves Act. There are prohibitions on transmission and distribution lines, transformer and distribution stations and other activities within the distance 50 to 120 meters from natural habitats and cultural and heritage sites. To obtain derogation from these restrictions, applicants must submit an environment impact study that meets specific requirements and guidelines including: (i) assessment of environmental impacts, (ii) identification of mitigation measures, (iii) description of monitoring, (iv) explanation as to why installing at an alternative location is infeasible, as well as written confirmation from the Ministry of Natural Resources and Forestry that the study has been prepared in accordance with the Natural Heritage Assessment Guide.
Governments can implement various measures to help streamline and expedite the process, while maintaining the rigour of the assessment and ensuring compliance with (and where necessary review of) existing regulations. Improving predictability of the timelines can allow for greater cost- and time certainty for project proponents (McMaster et al., 2021[92]). In this context, some jurisdictions streamline the process by centralising the permitting process within one authority. For instance, in Scotland, offshore wind development licenses are issued by one central authority, the Marine Directorate Licensing Operations Team, which oversees all steps of the permitting process including Habitats Regulations Appraisals (Vasconcelos et al., 2022[93]).
Countries can also provide detailed guidance to support developers of wind and solar projects conduct an EIA to prevent practitioner discretion, which can arise from the lack of clear definitions and facilitate a shared understanding of what constitutes a significant environmental impact (Gasparatos, Ahmed and Voigt, 2021[94]). For instance, Australia has developed guidance on key environmental factors to consider for offshore windfarms, such as displacement of marine fauna and invasive marine species (Department of Climate Change, Energy, the Environment and Water, 2023[95]).6 Similarly, the European Commission provides guidance on wind energy developments, detailing relevant legal provisions for biodiversity conservation as well as highlighting good practices (European Commission, 2020[96]). More generally, the International Union for Conservation of Nature (IUCN) and the Biodiversity Consultancy have developed guidelines for project developers to mitigate biodiversity impacts associated with solar and wind energy development across lifecycle (Bennun et al., 2021[47]).
Enhancing availability and access to relevant information can also help alleviate burdens on project proponents. In France, for instance, the Renewable Energy and Biodiversity Observatory was established in 2024. In collaboration with research institutes, energy industry and NGOs, the Observatory synthesises and disseminates findings from existing studies and data to inform stakeholders, to reconcile the objectives of Energy Sovereignty and Carbon Neutrality 2050 with Zero Net Loss of Biodiversity and Zero Net Artificialisation.
As spatial planning facilitates the identification of the areas in which renewable energy projects can be set up with relatively limited environmental impacts, it can also help streamline the permitting process. For instance, the recently amended Renewable Energy Directive in the EU, which requires designating renewables acceleration areas stipulates that the permit granting procedure for renewable energy projects in the designated areas shall not exceed 12 months for renewable energy sites on land and 2 years for offshore wind projects.7
While renewables result in significantly less pollution during operation compared to fossil fuels, there are risks of pollution associated with production and end-of-life management (see Section 6.2.1). Consideration of these risks is mandated in the regulations governing the EIA in some countries. For instance, the Infrastructure Planning Regulations 2017 in the United Kingdom requires that the Environmental Statements submitted by project proponents provide information on pollutants emissions as well as the management plan for the disposal and recovery of waste, among others (United Kingdom, 2017[97]).
Opportunities for improving EIA processes
Establishing and assessing the baseline (initial state), such as via the landscape, species diversity and abundance, is fundamental to a robust quantification of impacts associated with renewable energy projects. Given the different impacts of solar and wind power at various geographical scales, it is also important to clearly define geographical perimeters. In France, for instance, considerations of (i) potential siting zone, (ii) biological study area (200 m around siting zone), (iii) close study area (6 km radius) and (iv) the remote study area (18 km radius) are facilitated by the guideline that collates various steps and data sources to understand the initial states of these areas (MTE, 2020[98]).
Examining alternative scenarios, including alternative siting zone, technology and the sourcing of the equipment, can also help project proponents examine the impacts more holistically, although these efforts need to be carefully balanced with the objective of accelerating renewable energy expansion. It is important to consider the temporal dimension of biodiversity, as new risks can arise from climate-induced range shifts of species even if initial siting has considered and avoided the most vulnerable areas; an important consideration given that the renewables infrastructure is typically installed for decades (Ashraf et al., 2024[99]). In some jurisdictions, it is mandatory to consider a “no project” scenario. For example, the EU amended its EIA Directive in 2014 to require including “a description of the reasonable alternatives studied by the developer” and indicating the main reasons for the chosen option, rather than just “an outline of the main alternatives” as stipulated in the previous version of the Directive of 2011. In Austria, the “zero-alternative” is often discussed, comparing the planned situation with a baseline scenario and representing the development of the state of the environment without the project (Jiricka-Pürrer, Bösch and Pröbstl-Haider, 2018[100]).
Considering the cumulative impacts of renewable projects, i.e. both contemporaneous and future impacts is also conducive to management of trade-offs. Such considerations can be mandated by legislations or informed by non-binding guidelines. The European Union mandates the assessment of cumulative impacts through the Environmental Impact Assessment Directive, the Strategic Environmental Assessment Directive, and the Marine Strategy Framework Directive (Willsteed et al., 2018[101]).
Protocols for impact monitoring can also detailed in EIA. These protocols can mandate specific impacts to be monitored (e.g. bat and bird populations and fatalities) as well as the frequency and set the different thresholds that trigger response actions. In some jurisdictions, regular monitoring is mandatory. For example, in France, Article 12 of the Decree of 26 August 2011 (as amended by the Decree of 10 December 2021) specifies that the wind power facility operator must set up environmental monitoring within the 12 months of installation (unless exempt in special circumstances) and conducted every decade at minimum to estimate the mortality of birds and bat species due to the presence of the wind turbines. The first criminal case against wind farm operators (Box 6.2) is illustrative of the potential cost of the failure to account for the implications of renewable energy projects on biodiversity.
In December 2024, the French Council of State cancelled a prefectorial order that had rejected a request for a “protected species” exemption, on the grounds that environmental permits issued for the wind farms were considered final. In the Prefect’s view, it precluded any subsequent derogation request. In April 2025, the Montpellier Correctional Court (France) ordered a one-year shutdown of the Bernagues wind farm in Hérault, holding its operators responsible for the death of a golden eagle, a protected bird.
The operating company of seven turbines was fined EUR 200 000 while its manager faced a EUR 40 000 fine. This decision was issued just after a similar order that mandated a four-month suspension of the 31 turbines at the Aumelas wind farm, located in a Natura 2000 site (also in Hérault), during the nesting period of the lesser kestrel (Falco naumanni). The ten operating companies were found responsible for the illegal killing of 160 birds from 20 protected species, in addition to bats from 7 protected species. Each of the ten companies involved was fined EUR 500 000 (with EUR 250 000 suspended) and the former CEO of one of the operating companies was sentenced to six months of prison with a suspended term and fined EUR 100 000 (of which EUR 30 000 was suspended). The judicial decisions by the Court of Montpellier mark a first in France in penalising wind energy operators for wildlife destruction.
Source: FNE-OCMED (2025[102]).
Incorporating the impacts of renewables throughout the lifecycle can also help broaden the consideration. The information regarding the type and the estimated quantity of waste generation during construction and operations are typically provided as part of EIA. However, with some exceptions (e.g. the United Kingdom), the impact of decommissioned renewables equipment at their end-of-life tends to be inadequately considered in EIA, possibly owing to the fact that waste management is prescribed by national and municipal laws rather than being specific to a site or a project.8
Procedural improvements and institutional arrangement such as public consultations at an early stage and independent review can further enhance the integration of a broader set of objectives. While these procedures can be time-consuming, anticipating the environmental risks at an early stage of project planning can reduce the risk of delays and cancellations of projects. In England, public consultations are mandatory at a pre-application stage for onshore wind projects involving more than two turbines or the hub height of any turbine exceeding 15 m (Rankl, 2024[103]). Some countries also provide specific resources to improve public consultations and bring broader environmental impacts into consideration. In Germany, the KNE ( dedicated to biodiversity and renewable energy, founded by the Ministry of the Environment, Climate Action, Nature Conservation and Nuclear Safety, provides a team of mediators to help resolve conflicts relating to nature protection in the context of renewable energy developments (OFB, 2023[104]).
In France, the French Environmental Authority and regional agencies are responsible respectively for reviewing authorisation applications of large- and small-scale renewable energy project and for providing an independent opinion. The review is intended to help improve the design of the project, as well as informing the public and enabling their participation in the decision-making. While the opinion is not intended to determine the appropriateness of the project, the competent authority takes the opinion into the consideration for the decision on authorisation. These environmental authorities also help highlight the common issues such as the inadequate consideration given to the connection of wind farms and solar PVs to electricity grid (IGEDD, 2023[105]). India’s EIA Process & Procedures legislation similarly mandates that the Ministry of Environment, Forest and Climate Change form an appraisal committee to review the project (Ministry of Environment, Forest and Climate Change, 2024[106]). Additional consultations with experts and can also be held when necessary and the appraisal committee is granted 90 days to review the EIAs (Ministry of Environment, Forest and Climate Change, 2024[106]).
6.2.3. Additional safeguards to minimise the impacts on biodiversity and pollution
While licensing and permitting processes are important regulatory tools, they can be complemented by other policies that provide additional safeguards. In line with the “mitigation hierarchy”, there is a need to minimise the risks to the extent that their avoidance is not feasible. Similarly, upstream and downstream management of wind and solar power equipment constitutes an important lever for minimising the impact on air, water and soil quality.
Safeguards for minimising the impact on biodiversity
Safeguards can range from physical (e.g. diverters on power lines) and operational controls (e.g. shutdown of wind turbines) as well as various “curtailment” strategies for wind turbines, such as preventing their blades from turning at low wind speeds, although only a small portion of wind turbines currently operate with curtailment (Voigt et al., 2024[42]). Creating and conserving microhabitats for preserving pollinator biodiversity can help reduce the risks associated with utility-scale PV farms such as the use of herbicides to reduce panel shading and fire risks (Vaverková et al., 2022[107]; Blaydes et al., 2021[108]). The need for these safeguards can be specified to provide legal certainty for developers, clarify the expectations of administrative authorities at an early stage and simplify the appraisal process.
Better management of upstream and downstream risks of renewables technologies
There can be extensive upstream impacts of renewable technologies due to the quantity and the type of materials required to produce them (see Section 6.2.1). In this context, improving traceability of inputs to these technologies can help identify and address the upstream impacts. For example, the European Union recently introduced the Critical Raw Materials Act, which aims to promote sustainability in raw material sourcing by requiring mineral supply chain mapping until the point of extraction while empowering the European Commission to establish rules for the environmental impact of material extraction (European Union, 2024[109]). To improve traceability of domestic mining supply chains, Colombia has also developed Minerals Traceability Platform, which requires and verifies information regarding all stages of the domestic mining and transaction processes (Republic of Colombia, 2024[110]).
Relatedly, policies that mandate biodiversity protection for mining activities can also help alleviate the risk of pollution. For instance, the Administrative Order 2022-04 by the Ministry of Environment and Natural Resources in the Philippines requires mining companies to conduct biodiversity assessments before exploration in accordance with national regulations, guidelines and technical orders that identify the appropriate monitoring entity along with specific criteria to include in their reports. Further, the administrative order stipulates that the mining company maintain at least 85% survival rate along with similar diversity of species after rehabilitation (Republic of The Philippines, 2022[111]).
To support better management of components of solar and wind technologies with llimited reusability and recyclability, downstream provisions can be applied. For instance, several countries around the world prohibit disposal, landfilling and incineration of wind turbine blades and facilitates recycling of end-of-life solar panels. In Germany, a large proportion of the decommissioned wind turbine blades is directed towards energy recovery in cement co-processing, contributing to the recovery of materials and energy generation (CGEDD and CGE, 2019[112]), although there are some concerns that this constitutes downcycling and is associated with PM2.5 emissions (Engie, 2021[113]; ACP, 2023[70]).
Decommissioning policies can also entail financial guarantees to cover the cost and to ensure appropriate dismantling and restoration of the site. For example, in Scotland, offshore renewable energy installations decommissioning policies may require a financial assurance for decommissioning, and compliance with reporting and records requirements over the lifetime of the project. Under the Energy Act 2004, financial security can be provided prior construction in the form of an upfront cash security paid into an escrow account, or through performance bonds (Scottish Government, 2022[114]).
Protected areas serve as a cornerstone of biodiversity conservation policy and are, via their positive spillovers on climate change mitigation, adaptation and pollution reduction, a key tool to address the triple planetary crisis. Without protected areas, global loss of biodiversity would be even higher (Laurance et al., 2012[115]; Gill et al., 2017[116]; Geldmann et al., 2018[117]; UNEP-WCMC, IUCN and NGS, 2018[118]). Protected areas are also key policy instruments to manage land- and sea- use systems at local, national and international levels. The importance of protected areas as a core response to biodiversity loss is reflected in the ambitious target under the Kunming-Montreal Global Biodiversity Framework (KMGBF) to expand the global protection and conservation to at least 30% of terrestrial and inland water areas and 30% of marine and coastal areas by 2030.
The IUCN defines a protected area as “a clearly defined geographical space, recognised, dedicated and managed, through legal or other effective means, to achieve the long-term conservation of nature with associated ecosystem services and cultural values” (Dudley, 2008[119]). The IUCN has developed six categories to classify protected areas based on their management objectives (see Annex 6.B) which regulate human activities with varying levels of stringency. The types of protected areas employed are influenced by national land contexts, particularly in relation to land tenure and unmanaged areas (OECD, 2020[120]).9 In countries with secure land tenure and relatively little unmanaged land, land-use restrictions tend be controlled through protected area designations that overlap with or are entirely within private land. In countries with large areas of unmanaged land, protected areas can play a key role in preventing the conversion of forests and other ecosystems (Nolte et al., 2013[121]; Pfaff, Santiago-Ávila and Joppa, 2016[122]; Gaveau et al., 2012[123]).
Protected areas are often instated with biodiversity conservation as the primary objective. They can also include benefits for climate change mitigation and adaptation and pollution control, not least through providing a diverse set of ecosystem services. Protected areas mitigate climate change mainly through carbon absorption, sequestration and avoided carbon release. They also constitute nature-based solutions to climate change adaptation by mitigating the risks and impacts of extreme events and natural disasters. Moreover, protected ecosystems can filter water and air-borne pollutants. Last, protected areas are also key policy instruments of sustainable land-use to align biodiversity, climate, and food policies (OECD, 2020[120]).
6.3.1. Synergies and trade-offs between protected areas, climate change and pollution
Synergies between protected areas, climate change and pollution
Protected areas, through their biodiversity conservation objectives and effects on ecosystem services, can provide multiple synergies related to climate change and pollution reduction. The most salient synergies to climate change mitigation are carbon absorption, sequestration and avoided carbon release. In 2020, protected areas accounted for 11% of the global forested area and stored 26% (61.4 Giga tonnes (Gt) carbon (C)) of the estimated above ground carbon (Duncanson et al., 2023[124]). Of this total of 61.4 Gt, around 9.65 Gt of additional stored carbon can be attributed to protected area status. These higher carbon stocks are mainly from (i) avoided emissions from deforestation and (ii) a combination of enhanced growth and or avoided degradation in protected forests compared to unprotected ones. In addition, protected areas and other effective area-based conservation measures are estimated to store 21% of the belowground biomass, 15% of the soil organic carbon and 7% of the marine sediment carbon (Secretariat of the Convention on Biological Diversity, 2020[125]). Protected areas also store about 23% of the irrecoverable carbon in Earth systems, and about half is concentrated in just 3.3% of world’s land area (Noon et al., 2021[126]). For example, at the national level, Parks Canada has estimated that the average carbon stock contained in 31 national parks10 between 1990-2020 was almost 1.5 Gt C, and that the average carbon density for forested ecosystems within national parks was a little more than 250 t of C per hectare (Sharma, 2023[127]).
Protected areas play a critical role in climate change adaptation by mitigating the risks and impacts of extreme events and natural disasters, such as wildfire. Protected areas can create thermal buffers (Xu et al., 2022[128]), mitigate the effects of floodings and storms (e.g. service provided by mangroves and coral reefs ecosystems (Salem and Mercer, 2012[129])), stabilise shoreline or control the erosion of riverbanks, stabilise sediments and protect from landslides (Dudley et al., 2015[130]).
Protected areas have in the past not been fully integrated in Nationally Determined Contributions (NDCs) under the Paris Agreement. In the 168 latest NDCs submitted by Parties (as of 9 September 2024), only 50% of Parties reported measures related to conserving, protecting, and restoring nature and ecosystems – including protected areas – within their conditional mitigation commitments (UNFCCC, 2024[131]). Thus, the third round of NDCs due in 2025 represents a decisive opportunity to integrate nature-based solutions including protected areas, to reduce global emissions and strengthen the resilience of ecosystems (Nature4Climate Coalition, 2024[132]).
Protected areas contribute to water quality improvements by filtering, sequestering, or storing waterborne pollutants. Their role in providing clean freshwater is particularly significant as they deliver 20% of the global total continental runoff and supply freshwater to nearly two-thirds of the global population living downstream (Harrison et al., 2016[133]). Moreover, protected areas offer water provisions less exposed to threats compared to unprotected areas. Over a quarter of water supplied by protected areas faces low threat levels, compared to only 10% of water provisions outside protected areas. In contrast, less than 10% of water provisions from protected areas are exposed to high threat levels, compared to a quarter outside of protected areas (Harrison et al., 2016[133]). However, challenges remain in ensuring comprehensive protection as about 70% of global river reaches (by length) have no protected areas in their upstream catchments, and only 11% (by length) achieve full integrated protection (Abell et al., 2016[134]).
By reducing ecosystem loss and degradation, especially in forest ecosystems, protected areas provide significant air purification services by absorbing pollutants – including PM2.5 generated by wildfires (Prist et al., 2023[135]). In the Brazilian Amazon, a quasi-experimental evaluation found that doubling upwind protected areas decreases PM2.5 by 10% and respiratory hospitalisations by 7% in the months of most active biomass burning (Sheehan et al., 2023[136]).
Trade-offs between protected areas, climate change and pollution
While protected areas provide significant benefits for biodiversity conservation, their use can entail trade-offs with climate change mitigation and pollution reduction objectives. Two types of trade-offs can arise: (i) pressures from climate change on biodiversity and the management of protected areas, and (ii) conflicts between climate change, pollution and biodiversity objectives in protected areas.
Renewable energy production such as bioenergy avoids the use of fossil fuels but can create competing land and forest uses and exert pressures on protected areas. Burning biomass also leads to air pollution. Bioenergy based on woody biomass – especially in combination with carbon capture, utilisation and storage – can potentially contribute to climate change mitigation but needs to reconcile biodiversity conservation with sustainable exploitation of forests and other ecosystems. The BioEUParks project in the European Union accompanied five European Parks in Austria, Hungary, Slovenia, Italy and Greece to develop models of small-scale supply chain for solid biomass harvested from sustainably managed forests (BioEUParks, 2016[137]). Among the project’s main recommendation is the call for a legal framework to use waste material from conservation actions, such as alien species removal, for energy production so as not to be considered as regular waste material.
Renewable energy projects, such as wind and solar power farms, contribute to reducing global pressures on biodiversity through climate change mitigation. However, when these projects are located within or near protected areas, they may exert direct pressures on local biodiversity and ecosystems, including impacts on bird and bat populations or pollution, as outlined in the previous section. The location of renewable energy projects may then affect the efficiency of protected areas to reach their biodiversity conservation goals. At the global level, in 2018, out of the 12 658 large-scale renewable energy facilities, 2 206 were operating inside important conservation areas, including 169 in protected areas and 122 inside strictly managed protected areas (IUCN categories I-IV) where no development activity should occur (Rehbein et al., 2020[138]). Furthermore, 252 renewable energy facilities were under development within protected areas, including 100 in strictly managed ones.
Renewable energy projects such as hydropower facilities can exert pressure on biodiversity in or near protected areas by changing the connectivity and river flows inside or upstream of protected areas. Dams affect sediment and thermal regimes and alter the water quality and the ecological functions downstream. Globally, between 1900 and 2016, at least 1 249 large dams were located within protected areas with two-thirds built before their designation (Thieme et al., 2020[139]). The primary uses of these dams located inside protected areas were hydroelectricity (22%), irrigation (22%); water supply (15%) and flood control (7%).
Efforts to maximise carbon storage can conflict with objectives to conserve endemic species, and adaptation to changing climate. In forested areas – including protected forest areas – planting fast-growing exotic tree species to enhance carbon sequestration and resilience to climate change may alter local ecosystems and displace native species. The expansion of tree plantations between 2000-2012 affected 4% of protected areas across the humid tropics, most frequently in southeast Asia, West Africa and Brazil (Fagan et al., 2022[140]).
The increasing frequency and severity of wildfires, exacerbated by climate change, necessitate wildfire prevention measures that may conflict with biodiversity conservation. Healthy forests are generally less susceptible to wildfire ignition and spread. Ecosystem protection, restoration and their adaptive management can therefore constitute part of wildfire prevention (OECD, 2023[141]). Nevertheless, fuel-reduction treatments, such as clearcutting corridors and removing dead biomass to create tree discontinuities, can reduce carbon storage and harm species dependent on deadwood such as insects, fungi and bacteria (Campbell and Ager, 2013[142]). Fuel breaks or prescribed fires can also create invasive plant risk by letting the invasion of exotic species along wildland corridors (Keeley, 2006[143]). For example, the British Columbia Parks requires different fuel management across designation types within and across protected areas, from Intensive Recreation zones where management can be more intensive to Wilderness Conservation zones where land management is avoided (British Columbia Wildfire Service, 2024[144]). In the aftermath of wildfire, “salvage logging” can also reduce species richness and negatively impact soil and water quality (Thorn et al., 2018[145]; Leverkus et al., 2020[146]).
Protected areas often overlap with high-risk zones for natural disasters. Infrastructure to mitigate disaster risks, such as tide walls, coastal levees, or sand-control dams, can exert direct pressures on biodiversity. For example, large coastal levees were constructed in Sanriku Fukko National Park (Tohoku, Japan) after the 2011 Great East Japan Earthquake, creating tensions between disaster risk reduction and ecosystem preservation (Tanaka and Takashina, 2023[147]).
6.3.2. Dealing with risks to the synergistic effect of protected areas from negative climate change and pollution impacts
Protected areas face significant risks from climate change and pollution, which threaten their ability to deliver biodiversity and ecosystem services benefits. As climate conditions – including temperature and hydrological systems – have evolved since many protected areas were designated, these areas must adapt to remain effective. Climate change directly alters natural habitats, leading to shifts in species distribution and range. Fauna and flora may move poleward, upslope, or to deeper waters in search of suitable conditions, while vulnerable species face heightened risks of extinction. Climate change can modify the timing of biological events, such as flowering or breeding, disrupting established interactions between species. Protected areas may also experience an influx of new exotic species, some of which could become invasive and disrupt existing ecosystems. Changing climate conditions may alter the patterns of disease vectors (organisms that transmit infectious diseases), further threatening biodiversity.
As discussed in Chapter 3, while climate change is a main driver of biodiversity loss, pollution also contributes significantly to biodiversity loss and threatens the functioning of protected areas. Protected areas are vulnerable to pollutants from agricultural fertilisers, pesticides, heavy metals, animal waste, and sewage, which can originate upstream, in contiguous zones, or within the areas themselves (Jaureguiberry et al., 2022[148]). These pollutants threaten air, soil, and water quality, directly endangering the flora and fauna of protected ecosystems (Gross et al., 2017[149]). The adverse effects of pollution add another layer of complexity to the management of protected areas. Approaches to pollution control and mitigation vary across protected areas, reflecting differences in local contexts and capacities.
Accounting for climate change in designating new protected areas and expanding their network
Biodiversity conservation strategies need to consider the current and upcoming risks related to climate change in the design and expansion of protected areas. The scientific literature provides five categories of recommendations (Ranius et al., 2022[150]): (i) ensure sufficient connectivity to enable species to shift distributions or recolonise patches in habitat networks; (ii) protect areas that remain relatively preserved from the effects of climate change over time (climate refugia) and that provide high habitat heterogeneity to capture variability in micro-climatic conditions; (iii) protect a few large rather than many small areas; (iv) protect areas predicted to become important for biodiversity and with the most suitable environmental conditions as predicted by projected climate change; (v) complement permanently protected areas with temporary protection as a response to species’ distributions and shifting habitat suitability.
In prioritising areas to protect, governments may consider the overlap between hotspots for carbon and areas with both high biodiversity and intactness. At the global scale in 2015, protected areas represented only 12% of the 38% overlap between carbon and biodiversity hotspots (Soto-Navarro et al., 2020[151]). Prioritisation can also target specific ecosystems. For example, peatlands, which cover only 3% of global land but store around 600 gigatons of carbon in their soil, store more carbon than all world’s forest biomass (Xu et al., 2018[152]; Kopansky, 2022[153]; Pan et al., 2024[154]). Despite their importance, only 17% of peatlands worldwide were within protected areas in 2023 – significantly less than other high-value ecosystems such as mangroves (42%), saltmarshes (50%), and tropical forests (38%) (Spalding, 2021[155]; UNEP-WCMC, 2025[156]; WRI, 2024[157]). Moreover, within protected peatlands, just over half benefited from strict protection (Austin et al., 2025[158]). The assessment of protected areas’ ecosystem services or the expected consequences of land use changes on species can help the mapping and prioritisation of other areas of importance to protect as well as areas to strictly protect (Watson et al., 2023[159]; Chen et al., 2025[160]).
Protected areas that are well-designed and managed more effectively retain cold-adapted species than unprotected sites (Gillingham et al., 2024[161]). Protected areas designation can also have a positive effect on species' persistence at trailing-edge warm range margins, although with a decreased magnitude at higher latitudes and altitudes (Gillingham et al., 2015[162]). Marine protected areas contribute significantly to ecosystem resilience. For example, coral reefs within Australia's Great Barrier Reef marine protected areas have demonstrated greater resilience to climate stressors, such as bleaching, diseases, outbreaks of Acanthaster planci, and storms (Mellin et al., 2016[163]). Similarly, the Maria Island Marine Reserve in Australia showed enhanced resistance of fish communities to climate warming, with greater stability in species richness and diversity over both annual and decadal timescales (1992–2012). Benefits also included the recovery of large-bodied temperate species and resistance to colonisation by subtropical species (Bates et al., 2013[164]). However, smaller and more isolated protected areas may be more vulnerable to climate change impacts. These areas are often more exposed to range-expanding species and may lack the capacity to support species persistence under shifting climatic conditions (Loarie et al., 2009[165]; Gillingham et al., 2015[162]).
A broad range of adaptation measures has been developed to enhance the persistence of species and ecosystem functions in the face of climate change. However, selecting appropriate measures requires a tailored and local approach that considers the specificities of each species, habitat, and ecosystem within a protected area. Managers must balance climate-targeted, interventionist approaches to enhance carbon storage with the preservation or restoration of natural and wild areas. This trade-off is particularly relevant when deciding between active management interventions and allowing ecosystems to adapt naturally, emphasising the need for careful, context-specific decision-making.
Environmental assessment as integrative tools
Environmental assessment instruments are vital to identify the main direct and indirect environmental impacts of economic activities inside or adjacent to protected areas. For example, in the United Kingdom, in 2013, the Marine Management Organisation determined that the proposed deepening of the channel to Falmouth Harbour would not guarantee the protection of the Fal and Helford Special Area of Conservation, which is home to a sensitive maerl bed. The conducted EIA identified potential impacts such as sediment quality degradation and water contamination from tributyltin and loss of maerl and benthic communities (Hoppit et al., 2022[166]; Marine Management Organisation, 2010[167]).
Strategic Environmental Assessments (SEAs) can be conducted at the ecosystem level going beyond the boundaries of a protected area. For example, Parks Canada, responsible for the management of national parks, national historic sites, and national marine conservation areas across the country, has developed a layered system of planning and impact assessment framework which involves assessing environmental changes, including those caused by climate change, human activities, invasive species, and habitat removal. It applies to specific species or habitats and considers a broader zone than a protected area, typically at the subnational scale (habitats and considers a broader zone than a protected area, typically at the subnational scale (Parks Canada, 2021[168]). For instance, the assessment of cumulative effects of building and development projects was conducted on the major components of the Jasper National Park in Alberta: glaciers, wetlands, terrestrial birds, carnivores, and mountain goats. Strategic mitigations were conceived to protect each of these species or habitats: from the description of the current status to a vulnerability and risks assessment, and the outline of the corresponding strategic and project-level mitigation measures (Jasper National Park, 2022[169]).
SEAs can also be conducted by focusing on a cross-cutting component such as water quality. In Australia, the Queensland government conducted a Strategic Environmental Assessment of the Great Barrier Reef coastal zone - spanning over 2 300 kilometres – and analysed direct, indirect, and cumulative impacts on Matters of National Environmental Significance (MNES) (State of Queensland, 2014[170]). Dovetailing this, in 2015, the Australian and Queensland governments launched the Reef 2050 Long-Term Sustainability Plan, a collaborative framework to guide management of the Great Barrier Reef in the short, medium, and long term. Part of this plan, the Reef 2050 Water Quality Improvement Plan (2017-22) guides how industry, government and the community can co-ordinate and work together to improve the quality of water flowing to the Great Barrier Reef (Australian and Queensland governments, 2018[171]). The plan sets targets for improving water quality and land management practices upstream, such as reducing loads of pesticides, sediments, and nutrients (OECD, 2019[172]).
Complementary measures to environmental assessment
The implementation of additional regulatory measures such as permitting procedures, play a critical role in preventing or mitigating pollution that may threaten protected areas. In the European Union, permitting procedures under the Habitats Directive11 are required for any new plan or project, including renewables, likely to have a significant impact on a Natura 2000 site.12
National programmes including pollution reduction targets can play a key role in the integration of protected areas policies. For instance, the Nitrogen Reduction and Nature Improvement Programme adopted in 2021 in the Netherlands aimed at reducing nitrogen deposition on at least 50% of the surface area of the nitrogen-sensitive nature areas within the Natura 2000 network to less than the critical deposition load by 2030.
Economic instruments, such as Payments for ecosystem services (PES) (see Box 6.3), can be leveraged to expand protected areas, enhance biodiversity conservation, and support climate change adaptation. PES can address the impacts of natural disasters and extreme events. For example, in France’s National Park of Forests (Parc National de Forêts), some local farmers receive payments for flood risk reduction. In exchange, agricultural land is left uncultivated to function as prairie reservoirs, which are intentionally inundated during flooding events to mitigate downstream impacts. The valuation of ecosystem services and the final agreement of payment schemes are dependent on the natural and institutional characteristics of each site, and require building strong capacities among protected areas’ administrators (Gaglioppa, 2016[173]).
PES can be described as voluntary transactions between users and service providers that are conditional on agreed rules of natural resources management for generating offsite ecosystem services (Wunder, 2015[174]). For example, Costa Rica has been a global leader in PES with its Payment for Environmental Services Program established in 1997. The programme pays landowners for reforestation, forest conservation, and agroforestry practices (OECD, 2023[175]), which helps maintain biodiversity corridors and supports protected areas’ connectivity. Between 1997 and 2021, the programme covered 1.3 million hectares, with over 560 000 hectares added between 2011 and 2021, 90% of which were under forest conservation contracts. For example, under the reforestation scheme, landowners agree to reforest abandoned cleared lands and maintain the trees for 15 years. The programme resulted in two-thirds of the new area to be located within biological corridor and 28% is in national wildlife refuges that include private land. Overall, between 1997 and 2021, the programme has paid more than USD 600 million to small and medium producers (OECD, 2023[175]).
When habitat protection alone is insufficient, additional classification policies can be implemented to reduce threats affecting biodiversity, such as climate change or pollution. For instance, certain species may be granted priority status. Norway's 2009 Nature Diversity Act allows authorities to prohibit damage or extraction of such species, reinforcing conservation efforts.
Voluntary mechanisms can strengthen biodiversity conservation by reducing sources of pollution within protected areas. These approaches involve commitments between governmental authorities and private stakeholders or among private entities themselves (OECD, 2024[176]). For example, in France’s Cévennes national park, the 2013 Chart involved 40 localities that adopted stricter environmental commitments than required by law (Parc national des Cévennes, 2018[177]). The “zero chemical pesticides” action consisted in conducting diagnoses of the common weeding practices, providing advice and training to local stakeholders and designing action plans to use alternative weeding means such as mechanical weeding, sowing ground cover plants, eco-grazing). At the national level, the use of phytosanitary products by public authorities for the maintenance of green spaces, forests and walks accessible to the public has been banned only after (in 2017). The national park management raised awareness of higher environmental value and engaged local inhabitants and stakeholders to reduce pollution. Voluntary protection schemes can also increase the robustness of protected areas, and reduce the threats induced by human activities.
Monitoring and evaluation of protected areas to ensure synergistic outcomes
The integration of climate change and pollution considerations into protected area planning involves several iterative steps (Figure 6.3): (i) a vulnerability and risk assessment to understand the effects of climate change and pollution on the protected area; (ii) the design of an action plan that includes targeted adaptation measures; (iii) the implementation and experimentation to operationalise the plan and test the measures; (iv) monitoring and evaluation to assess the outcomes and refining the adaptation plan based on observed results. These steps should be repeated periodically, after several years of implementation and monitoring, to adjust and update the adaptation strategy. The effects and effectiveness of adaptation measures, along with other desired or unintended outcomes, often take years or even decades to become fully apparent. Robust monitoring frameworks are essential to track these outcomes over time. Monitoring should involve the collection of data on key indicators, which must be systematically integrated at subnational and national levels. This vertical integration ensures the coherent tracking of individual protected areas and the overall network, enabling the identification of synergies, gaps, and areas for improvement. In France, the National Strategy for Protected Areas (2020–2030) explicitly emphasises the need for protected areas to incorporate climate change considerations into their management plans (Measure 7, Actions 9 and 10) (Ministère de la Transition écologique; Ministère de la mer, January 2021[178]). The Natur’Adapt project (2018–2023), co-ordinated by Réserves Naturelles de France—the association responsible for France’s nature reserves—has been a key initiative in translating this strategy into action and integrated climate change adaptation considerations in the French national strategy for protected areas. Methodological tools have been developed and experimented in pilot sites and now need to be scaled-up and applied across all protected areas at the national level, in accordance with the Third National Adaptation Plan to Climate Change.
Source: Authors’ own elaboration based on Gross et al. (2017[149]) and Coudurier, Petit and Tissot (2023[179]).
6.3.3. Tools and indicators to assess environmental effectiveness of protected areas
Assessing the effectiveness of protected areas at project, national, and global levels presents numerous challenges, and is essential to enhance synergies with climate change mitigation, adaptation and pollution reduction policies. These challenges include the site-specific design and management of protected areas, the connectivity and ecological representativeness of the protected area network, and the appropriateness of the management processes employed (UNEP-WCMC and IUCN, 2024[180]; Rodrigues and Cazalis, 2020[181]).
Protected areas’ effective management as well as representativeness, consistency, and connectivity need to be taken into account to tackle the triple planetary crisis efficiently. There is currently no systematic and mandatory standardised assessment methodology to evaluate the effectiveness of protected areas, but several tools exist, at the project level, subnational, national and at a larger scale. Impact evaluation studies remain relatively scarce, and hint toward more studies to be conducted. Existing tools and indicators have the potential to enhance synergies between climate change mitigation, adaptation and pollution reduction by providing regular and comparable data and indicators across protected areas, accounting for their ecosystem services and monitoring their evolution.
An important advancement of the KMGBF compared to the Strategic Plan for Biodiversity 2011–2020 is the inclusion of a robust monitoring framework to track and assess the effectiveness of biodiversity conservation efforts, including the progress on protected areas and other effective area-based conservation measures.13 The framework includes indicators on protected areas’ extent, fragmentation, intactness, species richness, management effectiveness and their connectivity. The monitoring framework aims at evaluating the contributions of biodiversity conservation to multiple global challenges, among which climate change and pollution.
At the supranational level, the eight member countries14 of the Amazon Cooperation Treaty Organization implemented a guiding framework for the development and implementation of strategic actions for regional co-operation on biodiversity in the Amazon Region. The Regional Program on Biological Diversity for the Amazon Basin/Region was launched in 2021 to achieve the objectives of the Convention on Biological Diversity (CBD). In 2023, the Amazon Cooperation Treaty Organization assessed that nearly 23% of the Amazonian basin was protected (OTCA, ANA, ABC, COBRAPE, 2023[182]).
The effectiveness of protected areas at the site level lacks a systematic and mandatory standardised assessment methodology, although many assessments are based on the IUCN World Commission on Protected Areas (WCPA) framework for Protected Areas Management Effectiveness (PAME) (see Box 6.4). Among these tools, the Management Effectiveness Tracking Tool (METT), a questionnaire-based approach, is the most widely used globally (UNEP-WCMC, 2017[183]). The latest version of METT the tool laid greater emphasis on threat assessment, including severity and management responses (Stolton, 2016[184]). Assessments of protected areas are compiled in the Global Database on Protected Area Management Effectiveness, which includes records from 177 countries using 75 different methodologies. While this diversity reflects varied management contexts, it also complicates the vertical integration of data and highlights the need for harmonisation at the international level. However, the coverage of effectiveness assessments remains limited. Only 6.8% of protected areas listed in the World Database on Protected Areas have been assessed. These assessments encompass 28 969 evaluations for 20 603 protected areas, covering 4.8% of the world’s terrestrial protected areas, and 1.3% of the world’s marine protected areas (UNEP-WCMC and IUCN, 2024[180]). Therefore, more effort is needed to conduct assessment and to report at a supranational or global scale.
Each protected area has its specific species, habitats, and ecosystems to protect, and the effectiveness of conservation can be assessed through observation, tracking and recording biodiversity indicators. The dynamics of so-called keystone species and threatened species (i.e. those belonging to the IUCN Red List of Threatened Species) can be tracked, but also more ordinary species. For example, in France, the charter of national parks spans 15 years and includes a monitoring and evaluation framework. These charters are assessed midway and at the end of their duration to ensure alignment with conservation objectives and to adapt management plans as needed. At the national level, the LIFE Biodiv’France project (2024-2032) supports skills development among managers of protected areas, such as management planning and management effectiveness evaluation (OFB, 2024[185]). This structured approach illustrates how to integrate assessment into long-term protected area governance while addressing local and national conservation priorities.
Existing tools to assess protected areas at the site level
The IUCN WCPA framework for Protected Areas Management Effectiveness (PAME) covers three main topics: (i) design matters referring to the individual sites and the protected areas networks, (ii) capability and suitability of management strategies, and (iii) achievement of protected areas objectives and biodiversity conservation values (Hockings, 2006[186]), (Pulido-Chadid, Virtanen and Geldmann, 2023[187]). A Management Effectiveness Tracking Tool (METT) has been developed to operationalise the PAME framework. Initially created by the World Bank/WWF Alliance for Forest Conservation and Sustainable Use in 2002, this open-source tool has been modified and extended by users to fit specific site and ecosystem contexts. The most recent version, METT-4 (2024), has been published by Protected Planet, the authoritative source of data on protected areas and effective area-based conservation measures.
Protected areas' impacts on climate change and pollution can be assessed using ecosystem services assessments, though no standardised implementation exists. The IUCN proposes a participatory method where managers, stakeholders, and local communities engage in consensus-building workshops. These workshops assess the benefits provided by a protected area and classify the types of benefits (Ivanic et al., 2020[188]).
Existing Tools to report Protected Areas ecosystem services at the national level
Reporting on ecosystem services provided by protected areas can be achieved through Natural Capital Accounting. This method measures changes in the stock and condition of natural capital (ecosystems) across various scales. It integrates the flow and value of ecosystem services into standardised accounting and reporting systems. Natural capital accounting can comply with the international standards of the System of Environmental-Economic Accounts Ecosystem Accounting (SEEA EA).
The SEEA EA is an internationally agreed-upon statistical framework developed under the United Nations Statistical Commission. It provides a consistent methodology for measuring ecosystems and the services they supply, aligned with the System of National Accounts (United Nations et al., 2021[189]). The SEEA EA offers information on ecosystem extent, condition, species ecosystem services and associated economic activities from different data sources and of different data types to be unified through the same framework. This allows trends in ecosystem assets and species, trends in ecosystem services and benefits and land or sea use activities to be compared with one another. This provides crucial information for decision-makers who need planning sustainable development and protected areas management (King, 2022[190]).
Existing tools can be more systematically relied on to improve the integration of the triple planetary crisis in protected areas policies. First, the METT explicitly includes questions related to climate change, pollution and ecosystem services. The METT-4 tool asks through a multiple-choice question whether the protected area is consciously managed to adapt to climate change and whether the protected area is consciously managed to prevent carbon loss and to encourage further carbon capture. It also specifies pollution threats to the protected area to be recorded such as industrial, mining and military effluents and discharges, agricultural and forestry effluents air-borne pollutants. Moreover, a specific question is dedicated to qualitatively assess whether land and sea use planning outside of the protected area recognise the protected area and contribute to the achievement of its management objectives. Finally, the METT tools also covers whether the management of the protected area considers ecosystem service provision. This investigates both whether existing or potential ecosystems services are even known about and, if so, whether some or all of them are being managed.
Second, conducting ecosystem services assessments and reporting results across vertical (local to national) and horizontal (across – networks of – protected areas) levels is essential to integrate the synergies of protected areas. These assessments require strengthened monitoring and reporting capacities at project, network, and national levels. Key synergies of protected areas include regulating services such as climate mitigation with carbon absorption, sequestration or avoided carbon release, climate adaptation that is strengthening resilience and adaptive capacity to climate-related hazards and natural disasters and pollution reduction with the provision of clean freshwater and the purification and pollutants removal of water. The EU LIFE Programme mandates the use of the Mapping and Assessing Ecosystems and their Services framework, based on the Common International Classification of Ecosystem Services, to ensure comparability and facilitate reporting through the LIFE KPI Webtool (European Climate, Infrastructure and Environment Executive Agency, 2021[191]).
Third, Natural Capital Accounting can improve the integration and mainstreaming of protected areas at higher levels, while achieving several major goals: (i) help illustrate and make more visible the manyfold benefits provided by protected areas provide; (ii) contribute to promoting their establishment and maintenance (United Nations Department of Economic and Social Affairs, 2020[192]); (iii) boost the return on investment in protected areas to meet government objectives for the environment, economic growth and social wellbeing (Varcoe, 2015[193]); (iv) support the decision-making process related to protected areas.
Fourth, the adoption of the SEEA EA can facilitate the mainstreaming of protected areas into economic planning and monitoring processes. This framework ensures consistency and can be integrated directly with national accounting workflows. Adopting SEEA EA also translates into providing a coherent set of statistics on the environment-economy nexus that can easily be integrated into policy analysis. Although, not widely implemented yet for protected areas (King et al., 2023[194]), SEEA EA has the potential to provide regular and consistent information on the state, evolution and benefits from protected areas such as ecosystem services, climate change mitigation, climate change adaptation, pollution reduction and contributions to economic activity and human well-being.
Air pollution has significant effects on human health, with impacts of outdoor air pollution (OECD, 2016[195]) and ozone-related mortality rates (see Chapter 3) projected to increase significantly. These health impacts are by themselves part of the interlinkages between the different pillars of the triple planetary crisis. Other direct links are the effects of air pollutant emissions on radiative forcing – and thus climate change – and the pollutant filtering services provided by ecosystems – which are threatened by biodiversity loss. Air pollution policies can affect these linkages but also interact more indirectly with policy objectives for climate mitigation and biodiversity conservation in other ways.
6.4.1. Synergies and trade-offs of air pollution control for biodiversity conservation and climate change mitigation
Synergies and trade-offs between air pollution control and climate change mitigation
Many air pollutants and GHGs share common sources, not least energy production and use (Climate Watch, 2022[196]). As outlined in Chapters 1 and 3 with regards to air pollutants, ozone (O3) is not emitted directly, but results from reactions of heat and solar radiation with precursor gases (NOx and VOCs including CH4). Air quality is therefore affected by emissions of several pollutants and a by-product of the combustion of fossil fuels and biofuel, agriculture, manufacturing and extracting industries, energy production and use, forest fires and other processes that also generate GHGs (Bhattu et al., 2024[197]; Adams et al., 2015[198]; Mukherjee and Agrawal, 2017[199]). Analysis using the modelling toolbox developed for this Outlook suggests that population exposure to global PM2.5, which include BC, organic carbon and other particles with a diameter lower than 2.5 µm, are expected to decrease from around 23 µg/m3 in 2020 to around 19 µg/m3 in 2050. Equivalent regional estimations are discussed in more depth in Chapter 3.
Many air pollutants do not have a direct effect on climate but there are several important interactions (Lanzi and Dellink, 2019[200]), most of which are synergies where reductions in air pollutions benefit climate mitigation:
N2O is a GHG and its contribution to overall GHG emissions has been estimated at 2.6 Gt of CO2, which is about 4.5% of all GHG emissions (Gao and Cabrera Serrenho, 2023[201]).
Fluorinated gases (F gases) are also GHGs; they were introduced as a substitute for atmospheric ozone-depleting substances called chlorofluorocarbons (CFCs). Their share of all GHGs in CO2 equivalent (CO2e) is about 2% (IPCC, 2023[202]). They grew significantly when first introduced to replace CFCs. Emissions of F gases are now declining in high-income countries but not in all emerging and developing economies.
BC aerosols15, CH4, O3 and hydrofluorocarbons (HFCs) are also short-lived climate pollutants (SLCPs) that simultaneously contribute to climate change and compromise air quality. BC accelerates the melting of snow and ice and prevents clouds from forming. Estimating emissions of BC is subject to considerable uncertainties for example as BC emitted per fuel consumed by a given activity varies greatly under different condition of combustion and across countries, especially between developing and developed countries (Wang et al., 2014[203]). Wang et al (2014[203]) estimate a global increase of around 72% from 1960 to 2007 (Wang et al., 2014[203]), less than the increase in all GHGs over the same period (see also Chapter 3). Reductions in SLCPs undertaken to reduce environmental and health effects will also contribute to mitigating climate change.
Several aerosols have a cooling effect on climate. Sulphate (SO₄²⁻) aerosols (a by-product of SO2-emitting fossil fuel combustion as well as industrial processes) act to reduce the warming effect of GHG emissions by scattering sunlight and reflecting part of the sunlight away from Earth. Although about two-thirds of such emissions come from developing countries their impact on temperature is about the same as the lower share from developed countries, on account of the latter emissions being located at higher northern latitudes and being more evenly distributed zonally (Lin et al., 2022[204]). Nitrate aerosols (e.g. NH4NO3), formed from NH3 and NOx emitted from fossil fuel combustion, biomass burning, and agriculture, similarly also contribute to cooling, although this is more limited than for SO₄²⁻ aerosols (IPCC, 2021[205]). OC aerosols also have a cooling effect, although less than SO₄²⁻ aerosols and NH4NO3 (IPCC, 2021[205]). Overall, between 1750 to 2019, aerosols’ cooling effect is estimated to around -0.5 ºC – although there is uncertainty about the magnitude, with a range of ‑0.2 ºC to ‑0.9ºC (IPCC, 2023[206]).
Given the trade-off between sulphur emissions reductions and climate mitigation efforts, the use of sulphur aerosols in the higher atmosphere as part of as part of solar radiation management, called stratospheric aerosol injection, is a possible measure to limit warming. Stratospheric aerosol injection is a geoengineering method that refers to injecting different types of aerosols, typically SO₄²⁻, into the stratosphere to reflect sunlight, which should cause rapid global cooling. However, this method is controversial as the possible effects of large-scale deployment are not fully known yet. Some research suggests that it might cause, among others, changes in regional precipitation, acid rain, stratospheric warming and increases in stratospheric water vapour (UNEP, 2023[207]; CBD, 2016[208]). Sudden and sustained termination of stratospheric aerosol injection would also produce rapid temperature increases, known as termination shock (UNEP, 2023[207]; CBD, 2016[208]).
Synergies and trade-offs between air pollution control and biodiversity conservation
Air pollutants are associated with various adverse environmental impacts by acting as direct physiological stressors to plant species. For instance, the impact of O3 on plant physiology and interactions with pollinators is well-documented. High concentration of O3 can compromise plant growth, damages agricultural crops and forests, as well as intervene with the scents of flowering plants, altering the foraging behaviours of pollinators (Rollin et al., 2022[209]). Global relative yield losses are estimated to range between 7% and 12% for wheat, between 6% and 16% for soybean, between 3% and 4% for rice, and between 3% and 5% for maize (Van Dingenen et al., 2009[210]). Global losses from O3 on agriculture are estimated at USD 34 billion in 2010 and under business as usual are expected to be USD 36 billion in 2030 and USD 45 billion in 2050, with the People’s Republic of China (hereafter ‘China’) facing 23% of global damages and India 12% in the base year (Sampedro et al., 2020[211]).
Deposition of air pollutants, and especially SOx and NOx, through rainfall (commonly known as acid rain) can lead to acidification of soils and water bodies. The deposition of air pollutants can be in a different place from where the emission took place, as the pollutants are carried by wind over significant distances. The pollutants in the air react with water, oxygen and other chemicals to form sulfuric and nitric acids which fall to the ground, increasing the acidity of the soil and negatively affecting ecosystems, forests in particular.
Nitrogen deposition on land and in water bodies is mainly caused by the air pollutant NH3 from agricultural activities (see Annex 6.D) and NOX air pollutants from combustion processes. Excessive amounts of nitrogen introduced to an ecosystem lead to several negative impacts (United Nations Environment Programme/Food and Agriculture Organization of the United Nations, 2024[212]). In water bodies, it contributes to eutrophication, characterised by algal blooms and less available oxygen. In sensitive terrestrial ecosystems such as grasslands, if critical loads for nitrogen deposition are exceeded, sensitive species can be lost. In tandem, species that benefit from high nitrogen levels can flourish, which can change the structure and functioning of an ecosystem.
Although the impact of air pollution on animal species is less well-researched than on plant species, studies find a range of potential negative impacts. Air pollution can have lethal and sub-lethal effects on species through affecting respiratory organs (Sanderfoot and Holloway, 2017[213]). It can also negatively affect reproductive capacity (Carré et al., 2017[214]). These physical impacts on different species can cascade through the trophic chain, resulting in reduced abundance and altered community composition, threatening biodiversity and the critical ecosystem services it sustains.
Conversely, reducing air pollution through dedicated policies such as the assisted natural regeneration of forests (World Resources Institute, 2022[215]) can contribute towards biodiversity conservation and restoration. However, as outlined in the earlier sections, there are some potential trade-offs with biodiversity conservation when e.g. renewable energy policies are used as an instrument to reduce air pollution. Regulating air pollutant emissions, for example through mandated use of filters, does not have such trade-off effects.
Air pollution intensifies during wildfire events, resulting in the release of PM2.5, nitrogen dioxide (NO₂), O3, and the heavy metal lead. Implementing wildfire prevention measures can mitigate the risk of air contamination by these harmful pollutants. However, fuel-reduction treatments to create tree discontinuities may adversely affect local biodiversity by removing shrubs, trees, and species reliant on deadwood16 (Campbell and Ager, 2013[142]).17
6.4.2. Tools for synergistic air pollution control policies
Countries can conduct cost-benefit analysis of mitigation pathways for air pollutants to identify opportunities for synergistic policy. Including both data and input from authorities responsible for climate and biodiversity mitigation in cost-benefit analysis is key in identifying potential synergies. For example, in 2013, Mexico’s National Institute of Ecology and Climate Change conducted a cost-benefit analysis that integrated emission data from Mexico’s Special Programme on Climate Change and particulate data from the Climate and Clean Air Coalition to identify BC as a key priority for mitigation efforts (Molina Center for Strategic Studies in Energy and the Environment, 2013[216]). As a result, Mexico included BC for the first time in its 2015 Intended Nationally Determined Contribution under the Paris Agreement (Climate and Clean Air Coalition, 2019[217]). Building on this momentum, Mexico launched its 2019 Integrated SLCP Strategy to Improve Air Quality and Reduce the Impact of Climate Change, which will reduce BC emissions by around 53% by 2030 if nine identified mitigation measures in eight key BC source sectors are fully implemented (Climate and Clean Air Coalition, 2019[217]), exceeding Mexico’s NDC target.
Air quality Habitat Regulations Assessment (HRA) can help evaluate the impact of air pollutants on conservation areas and promote synergistic outcomes. For example, an HRA conducted on a proposed development plan in Greater Manchester, England, concluded that the development plan would generate air pollutants likely to negatively affect nearby conservation areas (Ricardo Energy and Environment, 2021[218]). In response, the developers submitted a revised plan 2023 that included mitigation measures such as a green belt, a boundary between the development and protected areas consisting primarily of hedgerows (Greater Manchester Combined Authority, 2023[219]). A subsequent HRA determined that the updated plan would exert no adverse air quality effects on the nearby conservation areas (The Green Manchester Ecology Unit, 2023[220]).
Countries can additionally conduct source apportionment18 studies for air pollutants to support the development of synergistic air quality management plans. In 2024, for example, Bangladesh’s Department of Energy and Ministry of Environment, Forest and Climate Change conducted a source appointment study using the Greenhouse Gas and Air Pollution Interactions and Synergies (GAINS) model to identify sources of emissions contributing to both air pollution and GHG emissions (Bangladesh Ministry of Environment, Forest and Climate Change, 2024[221]). Bangladesh synthesised these findings into their National Air Quality Plan 2024-2030, which outlines specific actions and responsible agencies for reducing air pollutants with GHG potential. One highlighted measure involves the increase of the environmental protection tax for highly polluting companies that fail to comply with emissions standards (Bangladesh Ministry of Environment, Forest and Climate Change, 2024[221]) – see also Annex 6.C for further information on pollution taxes.
To enhance the cross-border integration of synergies and trade-offs between air quality, climate change, and biodiversity, systematically using comprehensive environmental assessments and cost-effectiveness analysis is essential. For instance, adopting Natural Capital Accounting (NCA) would be highly beneficial. NCA aims to assess natural capital values, including e.g. the value of clean air by capturing the costs associated with air pollution, including healthcare expenses, productivity losses, and the degradation of ecosystems and the environment. Implementing annual NCA would systematically incorporate air quality into standardised accounting and reporting frameworks. Since 2019, for example, the United Kingdom’s Natural Capital team has published annual NCA reports that assess the economic value of various ecosystem services, including the regulation of air pollution, GHG, noise and urban heat. According to the 2024 NCA, the estimated annual stock value of the air pollution regulation service alone is GBP 133 billion (Natural Capital team, 2024[222]). Furthermore, NCA can facilitate international co-operation at the airshed level by enabling the exchange of comparable information, in alignment with the international standards set by the SEEA EA. The examples also show that vertical integration can help facilitate the effective implementation of different policies for air pollution.
Policies can promote horizontal and vertical integration to effectively manage air pollutants, with key tools outlined below. In many cases, improvements in air quality in an urban area require reductions in emissions outside that area. Research increasingly suggests the benefits of implementation of measures at the airshed level (Khan et al., 2024[223]; World Bank, 2025[224]; World Bank, 2023[225]), which requires adequate demarcation of the airshed area19 that can span different geographical scales (local, regional, national, international). Considering airsheds can help identify measures that are more cost effective than if considering only a narrow area within city lines, both by encouraging horizontal integration through sector-specific interventions in high-polluting sectors and through vertical integration by leveraging regional co-operation in joint air pollution strategies (World Bank, 2023[225]). In many cases, airsheds cross national boundaries, which in turn requires some form of international co-ordination and sharing of the burdens of reducing emissions in an acceptable way.20
Common tools of international co-ordination efforts include: (i) an overarching regulatory framework that sets emissions and air quality targets for participating jurisdictions, (ii) a well-funded central institution that ensures accountability and transparency, (iii) decentralised planning of abatement policies within the parameters set centrally, and (iv) economic incentives to reduce emissions, for example, through taxes and subsidies or by making access to funds conditional on abatement performance (World Bank, 2023[225]). An example regulatory framework for promoting vertical and horizontal integration is outlined in Box 6.5, for the case of the National Air Pollution Control Programme in the European Union.
There is no global framework for addressing air pollution, although there are key instruments, notably the regional Convention on Long-Range Transboundary Air pollution (1979). Their subsequent amendments have progressively imposed stricter and expansive requirements (e.g. BC and PM2.5 under the Gothenburg Protocol). In the EU, the National Air Pollution Control Programme provides the common format for discussing national plans for air pollution control, which contains an explicit reference to climate change. The reporting template ensures that “relevant climate change energy policy priorities” for key sectors including industry, transport and agriculture are considered in developing policies to control air pollution. The template also makes explicit references to the need for vertical integration (from international to local levels) for managing air pollution.
For instance, the Air Pollution Control Programme 2023-2030 in Spain highlights the importance of coherence and co-ordination between policies to tackle air pollution through 61 individual measured (across 12 packages) with other environmental policies such as climate change and energy transition plan, setting out the structure for informing local implementation on a range of sectors (e.g. transport policy of establishing low-emission zones in municipalities with over 50 000 inhabitants) through national-level planning. Reduction of burning pruning waste is also discussed in relation to synergies for alleviating the adverse impacts on soil biodiversity and improving soil carbon content as well as crop resilience against climate change.
In France, alignment to national objectives such as the national low carbon strategy and the national plan for reducing emissions of air pollutants including SO₂, NOₓ, NMVOCs, NH₃, and PM2.5 (MoET, 2022[226]) are ensured at the regional, local and municipal levels. At the local level, Climate Air Energy Territorial Plan helps operationalise policy measures to propel energy transition in sub-regional territories (MoET France, 2023[227]). Public inter-municipal co-operation bodies with over 20 000 inhabitants must establish a Climate Air Energy Territorial Plan, while smaller bodies may do so on a voluntary basis. As with the regional planning tools, these plans entail actions towards multiple objectives including air pollution control and climate mitigation. As of December 2023, 60% of the 750 bodies with over 20 000 inhabitants had adopted their Climate Air Energy Territorial Plan, and 97% had at least initiated the process (ADEME, 2024[228]). These regional and territorial plans also need to be aligned with Atmosphere Protection Plans at the local level. These plans are established under the authority of Prefects within urban areas of more than 250 000 inhabitants, and in areas where regulatory limits are exceeded or at risk of being exceeded. The plans set the objectives and measures to bring pollutant concentration below the regulatory thresholds. Together, these regional, local and municipal plans form the foundation for France’s approach to integrating climate, energy, and air quality targets. The vertical integration with national, regional and local frameworks ensures the implementation of strategic objectives while accommodating local specificities.
6.4.3. Integrating synergies of different types of air pollution policies with climate change mitigation
The synergies between air pollution and climate change identified in Section 6.4.1 act through a range of polices, including caps on emissions, taxes and sectoral control policies.
Caps on emissions of air pollutants
Caps on emissions of air pollutants can facilitate climate change mitigation when the effects of the pollutants that contribute to radiative forcing dominate. For example, Prinn et al. (2007[229]) considered the effects of caps set at 2005 levels on common air pollutants such as carbon monoxide, NOx and VOCs, and on CH4 – which is both an air pollutant and a major GHG – and SO₄²⁻ aerosols. First, as these air pollutants contribute to O3 creation, the caps can lower O3 concentrations causing less warming; second, they lower hydroxyl free radicals (OH·) that are considered the “detergents of the atmosphere” because they react with pollutants such as CH₄ – hence a reduction of OH· increases warming; third, as OH· are essential for converting SO₂ into SO₄²⁻ aerosols, they also lower SO₄²⁻ aerosols yielding more warming; and fourth, through O3 reduction caps allow more carbon uptake by ecosystems as O3 is toxic to plants (Franz and Zaehle, 2021[230]), leading to less warming. Overall, these effects roughly offset each other suggesting that an across-the-board air pollution strategy would have a relatively small net effect on the global mean surface temperature and sea level rise. The analysis does not, however, take account of overall demand for fossil fuels and on the choice of fuel, nor any caps on BC or organic carbon aerosols on climate. These effects, if included, could lead to more substantial synergies between capping air pollutant emissions and climate change mitigation.
In general, countries do not set caps on air pollution emissions uniformly. For example, Germany transposed the EU Medium Combustion Plant Directive into national law through default NOₓ caps for medium-sized combustion plants (Bundesamt für Justiz, 2019[231]). Parallelly, it includes provisions for its Länder (states) to apply stricter or accelerated limits regionally. The limits are determined to a significant extent by their effects on concentrations of pollutants, which in turn has impacts on human health and ecosystems. States like Baden-Württemberg or Bavaria enforce stricter NOₓ limits precisely because their air quality monitoring shows excessive NO₂ concentrations, exceeding EU limits. As the net climate effect of NOₓ emissions is generally considered to be warming, particularly through its role in increasing methane levels, such policies tend to reduce warming.
Ideally, any emission caps would take account of the benefits of the reductions in terms of their impacts relative to the costs and policies would be then selected to ensure the most cost-effective sources of emissions reductions were undertaken. The costs of reductions by source should take account not only of the direct costs but also any co-benefits, such as reductions in GHGs. In practice, standards are partly set based on benefits and costs using monetary estimates of these items but co-benefits such as lower GHG emissions are rarely accounted for. Consequently, the climate and other co-benefits of air pollution controls vary considerably. This can be seen by looking at policies to reduce PM. Here the co-benefits for climate can be quite large but vary by source of PM reduction. The World Bank led a study in South Asia that looked at measures at an airshed level across the sub-continent (World Bank, 2023[225]). Implementing key measures in South Asia to bring PM2.5 down to WHO interim target 1 (35 μg/m3) and WHO interim target 4 (10 μg/m3) by 2030 would reduce CO2 by 22 and 41 percent, BC by 81 and 89 percent, and CH4 by 21 and 28 percent, respectively. In the case of CO2, the largest contributions to CO2 reduction come from measures on power and heating plants and agriculture. In the case of BC, the biggest ones are from residential combustion and waste and in the case of CH4 they are from agriculture and waste (World Bank, 2023[225]).
A key aspect of setting limits on local pollutants is to integrate the implications for climate change as well. The issue has been discussed in the literature on co-benefits as well as on that of trade-offs. In the case of co-benefits, the focus has been on seeing air pollution reduction as a co-benefit of climate policy rather than the other way round. For the latter the most important trade-off is with respect to sulphate and other aerosol emissions, the reduction of which contributes to global warming. Reductions in aerosol emissions have been significant, not only in OECD countries but also in China, where they have declined 90% over the last two decades (The Economist, 2024[232]). In all, these are believed to have contributed to the observed global warming in recent years.
Air pollution tax policies
An analysis of the effects of air pollution tax policies on climate is included in a few selected studies. An example is (The Ex’tax Project, 2022[233]), which used the E3ME macroeconomic model to investigate the impacts of a range of hypothetical EU-wide taxes including those on all forms of transport, a price on carbon and taxes on air pollution (NOx, SO2, PM2.5 and NH3). Tax rates were based on the external costs although not implemented at the full external cost of each activity or pollutant.21 While the set of policies do reduce CO2 emissions (by a little more than 7%) – as well as positive employment and GDP effects – it is not possible to separate out the effects of air pollution charges from those of carbon prices.
While studies such as these show considerable benefits from environmental taxes, implementing them at the proposed rates faces considerable opposition at the local levels as well as difficulties in reaching agreement at the regional level (Levi, 2021[234]; Kallbekken and Sælen, 2011[235]).
Controls on emissions from transport
A number of large cities around the world have a congestion or vehicle user charge to drive in city centres that have significant environmental impacts (Chamberlain et al., 2023[236]). These can take form of area-based charges (e.g. London) and facility-based charges (e.g. tolled roads in Melbourne) (Veitch and Rhodes, 2024[237]). The rate of the charge can vary by the type of vehicle, with more polluting vehicles paying a higher charge and less polluting ones paying a lower charge or none at all. Significant reductions in air pollutants as well as carbon emissions were found in Milan (CO2 -35% and PM10 -18%), London (CO2 and PM10 -12%) and Stockholm (PM10 -18%) (EPOMM, 2015[238]).
Alternatively, some controls are set for Low Emissions Zones (LEZs), where polluting vehicles are either banned or only allowed to enter on payment of a higher charge. Several cities around the world have such controls. (Beedham, 2022[239]) evaluate the effectiveness of LEZs by using Tomtom’s traffic data. It estimates effectiveness of LEZs in Paris, Berlin and London for reducing CO2 (0.3-0.4% reduction), NOx (7-8% reduction) and PM (27-35% reduction). They emphasise that ‘most low emission zones focus on pollutants (NOx and PM), which heavily penalises diesel vehicles, and commercial vehicles such as trucks and vans in particular’. A fully integrative policy would weigh both the air pollutant and GHG emission characteristics of vehicles.
In all the studies reviewed there was no comparison of the effectiveness of direct regulations versus using charges as a means of managing LEZs. Given the limited evidence on this topic, further research is needed to evaluate the different instruments. Instruments that provide an incentive to drivers to switch to cleaner vehicles (potentially accompanied with a subsidy to phase out polluting vehicles) can generate stronger synergies with climate change mitigation. There are also some examples of policy approaches that integrate climate objectives in addressing air pollution from road transport (see Box 6.6).
Road transport is one of the largest contributors to air pollution in India, accounting for 27% of the country’s total outdoor air pollution (IEA, 2023[240]) and is responsible for 13% of the country’s CO2 emissions (IEA, 2021[241]). Launched in 2015, the Green Highways Policy in India set out a comprehensive framework for planting, transplanting and maintaining trees alongside and between highways to alleviate pollution. The main objective is to reduce air pollution and dust while addressing vegetation loss due to the development of highways by planting trees and shrubs. Acting as natural sinks for air pollutants, trees also help mitigate noise pollution and prevent soil erosion along embankment slopes (MoRTH, 2015[242]). Plantations alongside and between highways, as well as compensatory afforestation, are also carried out to offset carbon emissions resulting from tree felling and forest clearance during highway construction (CSIR-CRRI/CSIR-IIP/IORA/TERI, 2023[243]). The compensatory afforestation mechanism implemented by the Indian Ministry of Environment and Forests (MoEF) mandates planting a minimum of 50 plants or 10 times the number of trees felled for the construction of a highway, whichever is greater (MoEF, 2019[244]).
Over almost a decade, 46.5 million saplings have been planted along national highways under the Green Highways Policy (MoRTH, 2024[245]). These plantations and compensatory afforestation are estimated to sequester approximately 584 000 t of CO2 over a period of 20 years, while forest clearance and tree felling have emitted 653 000 tonnes. Furthermore, by allowing smoother traffic flow and more efficient fuel use, the improved highways are projected to avoid 25 Mt of CO₂ emissions over the same period (CSIR-CRRI/CSIR-IIP/IORA/TERI, 2023[243]).
Controls on emissions from domestic heating
The burning of solid fuels, and especially coal, for heating is a major source of air pollution in temperate climates. In Europe more than half of the particulate emissions come from the burning of solid fuels for heating (European Parliament, 2023[246]). Globally, data for contributions of all household air pollution (including cooking and heating) to particulate emissions is estimated at around 20% but with a high level of uncertainty (Chowdhury et al., 2023[247]). Encouraging fuel switching (away from coal and other solid fuels) and use of more efficient heating technologies (such as certified fireplaces or pellet stoves) can reduce the emissions from residential wood and coal heating devices. Filters and a switch to electric cooking stoves may further reduce health effects from indoor air pollution. Educational campaigns may also be useful tools to reduce emissions from residential solid fuel heaters (WHO, 2015[248]).
These interventions have demonstrable health benefits and also reduce GHGs such as CO2, although the amount of such reduction varies across the measures taken. A study in Ireland found that banning the marketing, sale and distribution of coal (specifically bituminous coal) improved air quality, carbon emissions and health, and reduced deaths from respiratory and cardiovascular causes (Clancy et al., 2002[249]). Connecting households using stoves to district heating systems that operate on clean fuels is also a very effective way of reducing air pollution as well as GHGs, as Werner (2017[250]) shows for Sweden.
Emissions controls from domestic heating can simultaneously contribute to GHG emissions reduction. The “Clean Air Programme” in Poland, for example, provides subsidies for replacing inefficient heating boilers and facilitating modernisation of buildings. With its dual environmental objectives to reduce PM emissions and GHG emissions, it has been progressively strengthened since its launch in 2018. For instance, the Programme has withdrawn co-financing support for coal-fired boilers in 2022 to improve consistency of the Programme with climate objectives.
The promotion of efficient stoves using pellets from processed biomass is seen as an effective way of reducing both air pollution and CO2 emissions. The European Commission proposals for updating the 2014 directive on air pollution control of domestic stoves imply that implementing eco-design standards involving such stoves, combined with energy labelling, are expected to reduce CO2, along with a fall in PM emissions, organic gaseous compound emissions and CO emissions. To achieve the adoption of these more efficient stoves countries are using a number of instruments, including partial grants for the new equipment, lower rates of value added tax, restrictions on the use of old stoves in some locations on particularly high pollution days and even a mandatory replacement of appliances (WHO, 2015[248]).
These measures can have considerable co-benefits for climate as examples given above indicate. Further evidence for this is provided by a World Bank study which found that replacing current wood stoves and residential boilers used for heating with pellet stoves and boilers and replacing chunk coal fuel with coal briquettes (mostly in eastern Europe and China) could provide significant climate benefits. It would also save about 230 000 lives annually across the world, with the majority of these health benefits occurring in OECD member states (Parson et al., 2013[251]).
Furthermore, action in this area is of particular importance to climate change in the Arctic (OECD, 2021[252]). This is because widespread dissemination of pellet stoves and coal briquettes have a disproportional benefit to mitigating warming from BC deposition in the Arctic (UNEP and WMO, 2011[253]). The World Bank found that replacement of wood logs with pellets in European stoves could lead to a 15% less warming in the Arctic (about 0.1 °C). For Arctic nations the modelling strongly indicates that the most effective BC reduction measures would target regional heating stoves for both climate and health benefits (Parson et al., 2013[251]).
The difficulties that governments have found in implementing programmes on wood stoves are significant. They arise in part from the costs involved to replace old equipment where subsidies are used as an instrument. Where regulations mandating replacement are used, they face a considerable backlash. Legislation had to be cancelled in Sweden on account of protests (Sahlberg et al., 2022[254]) in 2019 and more recently in Scotland (Johnson, 2024[255]). Finding the right combination of support to households and regulatory instruments remains a challenge.
Issues that arise include costs of alternatives for poor households, knowledge about alternative methods and conviction that they are effective and their compatibility with end-users (Boudewijns et al., 2022[256]). These challenges imply that economic support and information, in addition to adequate price signals, are key to achieving the adoption of clean fuel technology among mostly poor, rural households.
Controls on emissions from agriculture
Measures for reducing air pollution from agriculture include the following:
Reduction of emissions from burning of agricultural wastes. Policies that eliminate, or at least reduce, the burning of fields can be a cost-effective measure for the region. Compliance with regulations prohibiting burning has been a problem. Surveys show that farmers’ reasons for burning crop residues are mainly the high cost of incorporating, collecting, transporting, and processing crop residues in South Asia. Labour shortages, the marketability of the crop residue and the short time interval between harvest and next cropping seasons also influence farmers decision (Lin and Begho, 2022[257]). Recent evidence from India shows that cash transfers as payments for ecosystem services can reduce agricultural burning by up to 80 percent (Jack et al., 2022[258]), although Indonesia’s attempt to control agricultural burning through cash transfers has had mixed results (Falcon et al., 2022[259]). Other examples of payments for sustainable agriculture, have shown promising results in Latin America (Balseca et al., 2022[260])
Efficient use of fertiliser and management of livestock emissions. Agriculture also contributes to air pollution through the emissions of NH3 stemming from fertiliser use and manure management. Through chemical reactions NH3 is a major contributor to PM2.5. Adopting more efficient fertiliser management is important but challenging; see also Section 6.5. While reducing subsidies for fertiliser would be cost effective it faces strong opposition in countries where subsidies are used. Measures to better manage livestock emissions are available and involve combining controls for CH4 and NH3. However, implementing them at scale requires integrated management of both GHGs and local pollutants and considerable outlays (World Bank, 2023[225]; Sapkota et al., 2019[261]).
6.4.4. Integrating synergies of different types of air pollution policies with biodiversity
Air pollution policies that provide biodiversity benefits can involve determining threshold values for concentrations of air pollutants such that damage to ecosystems is limited, so-called critical loads or critical levels described for example in Annex 1 of the 2012 Gothenburg Protocol revision (UNECE, 2012[262]). For example, critical levels (as defined in article 1 of the Gothenburg Protocol) of NH3 are determined to protect plants in accordance with the Convention’s Manual on Methodologies and Criteria for Modelling and Mapping Critical Loads and Levels and Air Pollution Effects, Risks and Trends (UNECE, 2012[262]). Critical levels can then be translated into limits on emissions from sources that contribute to the concentrations. Threshold values are determined as levels above which damage to ecosystems is expected to be significant; the literature, however, does not demonstrate that costs of meeting threshold values relative to damages sustained have played a part in arriving at the values.22 The SEEA EA being set up by the UN under which spatially detailed accounts of ecosystem services are being drawn up will be a major factor in helping determine threshold values at the ecosystem level based on values of services and costs of abatement (United Nations et al., 2021[263]; United Nations, 2022[264]).
The impacts of air pollution on ecosystems and biodiversity are tracked through the effects of emissions of SOx and NOx on forests (acid rain) and through the effects of for example O3 on other ecosystems.
Acid rain
Measures to reduce transboundary SOx and NOx pollution, key drivers of acid rain, have had a significant spill-over effect on ecosystems and biodiversity. Given the large-scale effects on ecosystems and its transboundary nature, co-ordinated policy actions under the UNECE Convention on Long-range Transboundary Air Pollution (the Air Convention) on SOx and NOx emissions were undertaken in Europe and North America. These included for example the implementation of protocols that set national emission targets and identify specific measures for cutting emissions across a wide range of sectors (UNECE, 2015[265]). Air pollution emissions were substantially reduced through these actions (for the most important acidifying compound, SO2, emissions in Europe have decreased by 80% or more since the peaks around 1980–1990). Acid rain has thus been mitigated and ecosystem impacts including on European forests decreased (Grennfelt et al., 2019[266]).
In the US, similar actions took place. Congress passed several amendments to the Clean Air Act in 1990 that cut emissions of SO2 through a cap-and-trade scheme. The goal was a 50 percent reduction in such emissions from 1980 levels. That goal was achieved in 2008: SO2 emissions fell from 17.3 million tons in 1980 to 7.6 million tons in 2008. The effect has been remarkable. The Smithsonian notes that the rain falling in the Northeast today is about half as acidic as it was in the early 1980s. Consequently, surface waters have become less acidic and fragile ecosystems are beginning to recover (Willyard, 2010[267]).
Another region where acid rain is an issue is East Asia. In the two decades to 2000, countries in the region reduced emissions of SOx and NOx, thus lowering damages on terrestrial and aquatic ecosystems. Since 1975 China has decreased its emissions of SO2 by more than 75%. Acid rain remains, however, significant (Xuan et al., 2021[268]). Thus, potential solutions can address the problem from multiple scales and perspectives, including collecting reliable and relevant data, showing the synergies and trade-offs between emissions that cause acid rain and climate change, and recognising the need to take account of transboundary movement in agreeing on cost effective emissions reductions. The last makes use of dispersion models that are now well developed (Shah et al., 2000[269]). Lanzi et al (2022[270]) highlighted that integrated policy approaches across China, Japan and Korea can have very significant beneficial effects in comparison with domestic policies only.
Managing ground-level ozone concentrations
Policymakers have utilised a range of different mechanisms to address ground-level O3. These include emissions controls for O3 precursors, standards and guidelines which define air quality targets, alarm thresholds to advise sensitive sectors of the population to move indoors and O3 alarm plans to restrict emissions associated with land transport and industry during O3 peak episodes. A major part of regulations on O3 emissions and concentrations are set for health reasons (Box 6.7) but there are some that consider impacts on biodiversity, especially in areas where O3 is having a significant effect on ecosystems.
There is no global framework in place for the direct management of O3. The WHO has issued guidelines to policymakers for the most common air pollutants, including O3. These are intended to provide guidance for the development of policy measures aimed at reducing the impacts of air pollutants on human health (WHO, 2006[271]). Agreements are in place for some regions that set limits for O3 precursors – NOx and non-methane volatile organic compounds (NMVOCs). National limits for European countries are further based on scientific assessments of the thresholds of effects on natural ecosystems and human health, and the relative cost effectiveness of abatement options. The Protocol also defines critical levels, which together with integrated assessment modelling, are used to inform the negotiations on national obligations. For the European Union National targets for emissions are set in the National Emissions Ceilings (NEC) Directive. Also relevant for O3 control is the Integrated Pollution, Prevention and Control (IPPC) Directive. The Directive imposes a requirement for stationary sources of pollutants from new and existing industrial and agricultural sources with high polluting potential to be permitted only where certain environmental conditions are met. This includes sources associated with the energy industries, production and processing of metals, mineral industry, chemical industry, waste management and livestock farming (Fowler et al., 2008[272]).
Where regional agreements are in place to limit precursors, they have been quite effective. A review of the Gothenburg Protocol showed that the emissions of the O3 precursors NOX and NMVOCs have declined substantially as a result of emissions controls. In spite of this, however, as noted above there has not been a strong decline in O3 concentrations across North America and Europe. This partly reflects an increase in hemispheric background concentrations of precursors and indicates a need for further controls – e.g. on emissions from shipping and aviation, which are currently under consideration (Fowler et al., 2008[272]). In addition, episodic high O3 events caused by temperature variations prevent many areas from attaining O3 air quality standards. This is despite the large reductions in NOx emissions. A recent study finds that a time variant price for NOx emissions instead of a uniform price set by tradable allowances within a cap can be most cost effective in attaining O3 concentration standards (Holt and Linn, 2024[273]). For such a differentiated pricing policy, however the regulator must commit in advance to trading rules that depend on the forecast marginal effects of the regulated pollutant. The cost savings of a differentiated price depend on forecast accuracy and emissions abatement costs; if forecasts are sufficiently inaccurate, a uniform price could be less costly.
1. For the UNECE region the key agreement in place is the 1999 Gothenburg Protocol for, amongst others, VOC and NOx emissions. The Gothenburg Protocol was amended in 2012 to introduce national emission reduction commitments for 2020 and beyond, with a further revision process in progress (UNECE, 2023[274]).
Specific actions to address O3 concentrations’ impacts on ecosystems include selecting cultivars less sensitive to O3 and strong limits on emissions of CH4. Avnery et al (2013[275]) compare two O3 mitigation strategies applied worldwide: a gradual reduction in emissions of CH4, an important O3 precursor, and choosing crop varieties less sensitive to O3. The first increases global production of soybean, maize and wheat in 2030 by 2-8% of 2000 level, while the second could increase by around 12%. Benefits are dominated by improvements for wheat in South Asia, where O3-induced crop losses would otherwise be severe. Both measures have net benefits (i.e. costs of implementing them are less than the benefits). Combining the two strategies generates benefits that are less than fully additive, given the nature of O3 effects on crops.
Reduce burning of crop residues
As noted above, the burning of crop residues is a major source of air pollutants with health impacts as well as negative impacts on the ecosystem. It is reported to degrade the soil, increase the risk of erosion, and increase the soil temperature, consequently decimating soil microorganisms. This subsequently impacts the monetary cost involved in recovering soil fertility and the potential for further pollution through the increased use of fertiliser (Lin and Begho, 2022[257]). Actions to reduce such emissions are seen as a key part of a strategy to reduction PM2.5 in South Asia (World Bank, 2023[225]), with significant synergies for ecosystems and biodiversity conservation for the reasons given.
Nitrogen and phosphorus touch on all aspects of the triple planetary crisis (Annex 6.D). Given their centrality, maintaining a balanced nutrient cycle – which involves the movement of nutrients through soils, plants, animals, and the environment – is essential for water purification (Chapter 1) and long-term agricultural productivity, sustainability and resilience. Nutrients, particularly nitrogen and phosphorus, are essential for plant growth and are applied in agriculture through both inorganic fertilisers (made from minerals or synthetic chemicals) and organic fertilisers (especially manure).23 However, the management of these inputs varies widely across regions, often resulting in either nutrient deficiencies or surpluses24, both of which pose significant challenges. In some developing regions, notably Sub-Saharan Africa, the limited use of inorganic fertilisers and the diversion of organic inputs such as manure for non-agricultural uses such as energy for household cooking (Jones and Deuss, 2024[276]) lead to negative nutrient balances. This means that nutrient inputs are insufficient to replenish what is removed through plant uptake (Ludemann et al., 2024[277]). An insufficient nutrient cycle undermines long-term agricultural productivity and resilience. Over time, this may reduce soil fertility and reduce crop yields.
In contrast, in regions where fertiliser use is high, overapplication of nutrients can cause a range of pollution effects. Nitrogen reacts with other chemicals to form polluting compounds that have effects on the environment through air, soils, water, and ecosystems. Excess nutrients often leach into water bodies and act as fertilisers, triggering eutrophication and algal blooms (NOAA, 2024[278]). Excess algae can reduce or deplete dissolved oxygen available to aquatic life and can produce toxins that are harmful to people, animals, and aquatic ecosystems. Nitrogen compounds can also contaminate groundwater, posing health risks such as NO3- poisoning, even at low levels. Infants are especially vulnerable to NO3- in drinking water as their bodies do not break down NO3- as well as adults. Phosphorus fertilisation is also threatening water quality in regions such as Europe, East Asia, and South Asia (Ros et al., 2020[279]), although significant amounts of phosphorus are stored in the soils (and good soil management can prevent major losses to water bodies).
Striking the right balance in nutrient use is therefore essential. Integrated nutrient management that combines both organic and inorganic sources, tailored to local soil characteristics (such as pH, organic matter content, and nutrient holding capacity) and varying nutrient requirements of different crops25 (Al-Shammary et al., 2024[280]), is key to maintaining soil health, improving yields, and minimising negative environmental impacts. Strengthening farmers’ access to knowledge, inputs, and technologies that support efficient nutrient use can help close yield gaps while preserving the natural resource base critical for food security (Ignaciuk et al., 2021[281]).
6.5.1. The nitrogen cascade and phosphorus cycle touch on all pillars of the triple planetary crisis
Nitrogen easily changes chemical form and moves between air, soils, water and ecosystems, causing a cascade of damages across all pillars of the triple planetary crisis (Figure 6.4). For example, N2O is an important GHG and, in the stratosphere, also a powerful stratospheric ozone layer depleting substance that contributes to climate change. Similarly, NOx reduces air quality via the creation of (ground-level) ozone (O3). NH3 commonly originating from fertiliser and manure increases human health risks such as respiratory illnesses and cancer. Combined with NOx, NH3 contributes to the formation of nitrate aerosols such as ammonium nitrate (NH4NO3). Through the rain, these can lead to eutrophication in lakes, terrestrial and coastal areas, impacting fisheries and drinking water quality. Nitrogen can also damage ecosystems through leached NO3-, as well as acidification of soils and waterbodies. In the case of pollution through the soil, the application of fertiliser and manure increases the availability of soil mineral nitrogen (i.e. ammonium NH4+ and NO3−), which can result in the release of N2O that contribute to climate change.
Source: Sutton (2020[282]).
The global phosphorus cycle, which involves the movement of phosphorus between geological formations, soil, plants, and organisms, is not as well-defined as the global nitrogen cycle (Boyd, 2019[283]). Phosphorus is predominantly found in phosphate mineral-bearing ores, known as phosphate rock26. Phosphorus occurs naturally in most geological formations and soils in varying amounts and forms. Humans have enhanced the availability of phosphorus by mining phosphate rock. Rivers now transport about twice as much phosphorus as 300 years ago (Schlesinger and Bernhardt, 2020[284]). Only about 10% of the phosphorus delivered to the oceans is potentially available to marine organisms (Schlesinger and Bernhardt, 2020[284]). The remainder is bound to soil particles and sequestered by sediment through formation of iron, aluminum, and calcium phosphates. The phosphorus cycle has no significant gaseous component. The flow of phosphorus through the atmosphere in soil dust and seaspray is also much smaller than other transfers in the global phosphorus cycle (Schlesinger and Bernhardt, 2020[284]). Figure 6.4 visualises the intended and unintended flows of phosphorus. Phosphorus mainly accelerates the triple planetary crisis through the mining process itself, as well as through the subsequent agricultural applications of phosphorus as a fertiliser.
6.5.2. Synergies and trade-offs of nutrient pollution policies for biodiversity conservation and climate change mitigation
Policies to mitigate nutrient pollution consist of controls in the use of inputs that generate excess nutrients as well as controls on emissions of nutrients, such as for example the EU Nitrates Directive that for example sets limits on livestock manure application per hectare. In addition, there have been some fiscal incentives and in the form of taxes on inputs such as fertilisers (Andersen and Bonnis, 2021[285]; Andersen, 2022[286]), taxes on mineral phosphorus in commercial animal feed (Andersen, 2017[287]), and taxes on emissions of N2O (OECD, 2013[288]). These measures have synergies with climate change as they often result in a reduction on GHGs, and with biodiversity and ecosystems because the decline in nutrients causes less damage in these areas. Finally, lower levels of nutrients and the resulting lower emissions to the air and water also have health benefits.
Vice versa, policies undertaken with a climate change goal will reduce emissions of nutrients when agricultural sources of GHGs are covered and also reduce leakage of nutrients into the environment. Climate change-induced sea-level rise and shifts in precipitation accelerate the transport of dissolved and particulate phosphorus and nitrogen from soils, further exacerbated after the application of fertilisers and manures or drought (Lucas et al., 2023[289]). Climate change-related extreme rainfall events particularly increase phosphorus losses from agricultural lands to waterbodies (Our Phosphorus Future Network, 2022[290]). Hence, climate mitigation is essential to limit further leakage of nutrients into the environment.
Beyond taxes that target fertilisers or nutrient feed directly, a broad tax on GHGs that covers activities that also emit nutrients, such as agriculture or livestock, will reduce nutrient pollution as it scales down the economic activities that co-emit carbon and nutrients and incentivises a shift towards less-emitting sectors. Most carbon tax programmes, however, do not cover emissions from agriculture/livestock as these are hard to estimate and monitor. Denmark constitutes an exception as it is planning the world’s first carbon tax on emissions from livestock, to be introduced in 2030, as part of a vast agriculture plan called the Green Tripartite. From 2030, CH4 emissions caused by enteric fermentation, flatulence, and mature from cattle and pigs would be taxed at a marginal rate of DKK 300 (around EUR 40) per tonne of CO2e, a sum gradually rising to DKK 750 (around EUR 100) by 2035 (Danish Ministry of Taxation (Skatteministeriet), 2024[291]). However, a basic deduction (tax rebate) of 60 percent applied to the average emissions per type of animal to incentivise farms that have low emissions means the effective average tax rate will be DKK 120 (around EUR 16) per tonne emitted in 2030 and DK 300 (around EUR 40) in 2035 (Danish Ministry of Taxation (Skatteministeriet), 2024[291]).
Carbon taxes can also incentivise the use of technologies that reduce nutrient pollution. In the above Danish example of tax on emissions from livestock, calculations would be based for example on animal head count with different categories and a series of metrics that affect emissions per animal, e.g. feed composition and use of climate-friendly technologies (New Zealand Ministry of Foreign Affairs and Trade, 2024[292]). For example, in Denmark feed changes such as the use of 3NOP (Bovaer) as a methane-reducing feed additive is encouraged. Other carbon mitigation techniques considered in the tax include biogas production (Holmes, 2024[293]). Funds received from the new tax should also be returned to farmers to help reduce the burden over time through green technology innovation (New Zealand Ministry of Foreign Affairs and Trade, 2024[292]).
Equally, policies implemented to address biodiversity, such as protected areas, often limit agricultural activity which also lowers nutrient levels – especially when the threats to protected areas lead to controls of nutrient emissions in neighbouring areas. However, the recent Dutch Greenpeace litigation case against nitrogen pollution in protected areas illustrate that even in the presence of national targets, location specific policies are needed and demonstrates the challenge of balancing different economic activities when implementing a nitrogen target (Box 6.8).
The agricultural sector plays an important role in the Dutch economy: On average and in relative terms, Dutch agriculture occupies more land, generates more value and is more trade-oriented than in most OECD countries (OECD, 2023[294]). The Netherlands has built an agricultural sector that is a world leader in productivity and competitiveness.
In parallel, environmental challenges including nitrogen pollution have become increasingly urgent. Nitrogen pollution has been a problem in the Netherlands for many years, for a considerable part due to NH3 emissions from manure and fertilisers in the emissions intensive agricultural sector as well as due to NOx from the transport and industry sector (Government of Netherlands, 2025[295]).
Location-specific policies have been implemented for example through a court ruling on nitrogen deposition on sensitive landscapes, including protected areas. The Nitrogen Reduction and Nature Improvement Programme of 2021 sets binding targets for the percentage of the hectares of nitrogen-sensitive habitats in Natura 2000 protected areas on which the nitrogen deposition must be brought below critical deposition values. However, the negative environmental impacts of farms that are adjacent to Natura 2000 sites that are sensitive to nitrogen deposition have meant that nitrogen deposition on sensitive landscapes remains substantially above safe thresholds in most cases, which has impaired the quality and recovery capacity of natural habitats (OECD, 2023[294]). The government has tried to solve the nitrogen problems with an area-based approach that allocates a transition fund to reduce the impacts of adjacent farms, but further location-specific policies are needed to protect sensitive areas.
In an aim to reach the nitrogen targets, the Netherlands has considered trade-offs between different economic activities and imposed restrictions on for example the nitrogen-intensive construction sector, creating policy backlash. A 2019 court ruling put a temporary halt to all new development activity requiring permits to emit nitrogen, affecting agriculture and construction most strongly but touching many parts of the Dutch economy and placing many projects in limbo (OECD, 2023[294]).
Announcements of measures to reduce nitrogen emissions led to large farmer protests e.g. in 2019 (Tullis, 2023[296]) and have evolved into the Farmer-Citizen Movement party (called BoerBurgerBeweging, BBB). The BBB, as part of the new government, puts more emphasis on innovation and proposed to shift away from the area-based nitrogen approach.
Resolving policy backlash while sufficiently addressing the underlying nitrogen pollution problem (in the Netherlands known as the stikstofcrisis, the nitrogen crisis) is challenging, as a court ruling from January 2025 illustrates: the Court ordered the Dutch government to comply with the statutory nitrogen target for 2030, which means that the state must bring 50% of the area of nitrogen-sensitive nature below the critical deposition value by the end of 2030, with priority to be given to the most endangered nature types. If the Netherlands fails to achieve this goal, it must pay a penalty of EUR 10 million.
In response to the court's decision, a ministerial committee comprising four ministers was formed in 2024 to work on plans to tackle the nitrogen crisis. This committee was co-ordinated by the prime-minister’s office, highlighting the cross-cutting nature of the issue, although a final plan has not been delivered yet as the government collapsed in 2025 without a clear outcome of this committee.
Beyond helping to limit nutrient emissions and leakage, climate mitigation policies also help to alleviate the impacts of nutrient pollution, given that these are exacerbated by climate change. With rising temperatures, waterbodies are more prone to eutrophication due to increased biomass growth in waters that lower dissolved oxygen concentrations, reducing the self-cleansing capacity of waterbodies and increasing the risk of eutrophication (Forber et al., 2018[297]). This is in line with Birk et al. (2020[298])s’ findings that nutrients are not only the dominant stressors acting to degrade lakes but that parallelly the ecological effects of nutrients are exacerbated by stressors such as climate change.
A trade-off arises with respect to climate mitigation through reductions in NOx and NH3 emissions as a result of nutrient pollution control. NH3 reacts with the products of NOx to form secondary PM2.5. The resulting PM2.5 has a cooling effect on climate as it scatters light and promotes cloud formation (see Section 6.4). Conversely, reducing NH3 pollution will have a warming effect. However, this effect is small and relativised further by the fact that the formation NOx creates a GHG. Although these trade-offs should not be ignored, given the comparably strong synergies the mitigation of nutrient pollution is important to tackle the triple planetary crisis.
A further potential trade-off arises in the sense that increasing fertilisation leads to higher yields and thus lower land use requirements to produce the same amount of output, which could benefit both climate change mitigation and biodiversity depending on the alternative land use. OECD scenario research (Adenäuer, Laget and Cluff, 2024[299]) suggests that fertilisers are crucial for enhancing agricultural yield: a hypothetical elimination of fertiliser support in India prompts a rapid reduction in domestic fertiliser use and decrease in agricultural production. Globally, as the strong growth in agricultural output per unit of agricultural land is due in large part to a more intensive use of inputs, including fertilisers, reducing fertilisation could increase land use requirements for agriculture. Since 1961, global consumption of nitrogen fertilisers grew almost nine-fold, while consumption of phosphorus fertilisers nearly quadrupled (OECD, 2021[300]).
However, increasingly yield growth is driven by total factor productivity growth, such as through efficiency gains from better farm management practices (OECD, 2021[300]). Since the 1990s, total factor productivity growth has been the major factor driving the growth of global agricultural production (OECD, 2021[300]). Agricultural output increased also in periods where average nitrogen surpluses declined, such as from 1990 to 2009 (OECD, 2025[301]). The trade-off from decreased fertilisation might be particularly limited if nitrogen use efficiency (the ratio of nitrogen in crops compared to nitrogen applied as inputs through fertilisers or manure) is improved. Global nitrogen use efficiency lies only at around 50% globally, hence at better efficiency nitrogen pollution could be reduced without significant harm to yields (Lassaletta et al., 2014[302]).
6.5.3. Integrating synergies and trade-offs when assessing policies to reduce nutrient pollution
Policies to reduce nutrient pollution have strong synergies with protecting ecosystems and biodiversity. With regards to climate change, NH3 can have a cooling effect that is lost when emissions fall as a result of policies to reduce nutrient use and pollution. The effect, however, is small and moreover N2O creates a GHG that contributes to climate change. Overall, therefore, there are key synergies with climate change in reducing nutrient pollution.
Integrating synergies and trade-offs with nutrient pollution policies is most effective if the selection of measures takes account of the co-benefits of the measures with respect to climate change and biodiversity. A local case study highlighting Peruvian regulations on the collection of the natural fertiliser guano is shown in Box 6.9 below.
The Peruvian government aims to decrease soil pollution caused by artificial fertilisers by encouraging farmers and communities to use guano, a natural fertiliser obtained from the accumulation of bird excrement, while preserving collection areas’ biodiversity. Historically, the exploitation of guano, rich in nitrogen, phosphate and potassium, provided a natural source of nutrients for the agriculture before the invention of synthetic fertilisers, but its overexploitation damaged the biodiversity of the collection areas. During the 19th and 20th centuries, the overexploitation in the 22 Peruvian guano islands resulting from high demand on the international market caused the degradation of the lands and the decline of the fauna populations, starting with the guano birds themselves, whose nests were destroyed by the collection process (Doig-Alba et al., 2023[303]). The name “guano birds” refers to the three most important seabird species involved in the production of guano in Peru: the Guanay Cormorant, the Peruvian Booby and the Peruvian Brown Pelican. At the time, the labour-intensive extraction process using picks, shovels and brooms to loosen and collect the guano deposits which in some places were several metres thick, was often accompanied by the collection of bird eggs and the poaching of adult penguins (Chauvin, 2018[304]; Peru, 2024[305]). The population of guano birds declined from 53 million in the late 1880s to 4.2 million in 2011 (IUCN, 2013[306]).
While recognising the role of guano as a natural fertiliser and subsidising it, the Peruvian government has implemented measures to restore and protect biodiversity of the collection sites (Peru, 2024[305]; Peru, 2019[307]). To support local sustainable agriculture, guano is offered at subsidised prices to small farmers and communities when agricultural yields decline because of pest outbreaks or extreme weather conditions (Peru, 2024[305]; Peru, 2014[308]). In 2009, Peru designated the Guano Islands, Islets, and Capes National Reserve System as a protected natural area, part of Peru’s National System of Natural Areas Protected by the State (Sistema Nacional de Áreas Naturales Protegidas por el Estado - SINANPE), whose main objective is to maintain the conservation of biodiversity and ensure the sustainable use of wild flora and fauna resources (Peru, 2009[309]). On these islands, extraction is strictly regulated: harvesting is prohibited during the breeding season and rotational harvesting allows birds to recover and breed without interruption (Geo, 2024[310]; Peru, 2013[311]). The endangered Humboldt penguin population, historically negatively affected by guano collection, has increased since the creation of the reserve, with 8 025 penguins registered in 2019 (World Bank, 2020[312]).
Comprehensive EIAs and other integrative measurement tools play a key role in highlighting the effects of nutrients – and nutrient policies – on climate change, biodiversity and pollution. The interactions of phosphorus and nitrogen (and its derivatives across air, water and soils), (highlighted in Annex 6.D), also need to be taken into account. This can be complemented by cost-effectiveness analysis of different options to meet nutrient targets, or even undertaking a full cost-benefit analysis, again taking a broad scope to include all environmental impacts. Typically, cost-effectiveness assessments in this area have only looked at the direct costs of measures that meet a given target. Information is, however, available on the benefits of nutrient reduction in terms of health endpoints and climate change (reductions in GHGs). Estimates can also be made of reductions in environmental impacts, which will vary by location within the area for which the targets are set. Such benefits can be deducted from the direct costs to give the full costs of measures, and the selection can then be based on those that have the lowest net cost. For instance, in 2021, the European Commission conducted a cost-effectiveness assessment of countries’ proposed measures to meet targets set out in the Baltic Sea Action Plan 2021, including sub-basin nitrogen and phosphorus maximum allowable inputs (Baltic Marine Environment Protection Commission, 2021[313]). The assessment incorporated the climate adaptation and habitat creation benefits of spatial measures for nutrient pollution reduction, such as artificial or restored wetlands, stormwater ponds and mussel farms, and concluded that these measures are indeed cost-effective. The assessment further noted that targeted interventions in coastal areas are likely to yield additional biodiversity benefits (HELCOM ACTION, 2021[314]).
Full cost-benefit or even broad cost-effectiveness analysis of nutrient management policies are exceedingly rare. Countries do, however, take account of differences in biodiversity impacts of nutrient pollution when setting controls and limits on nutrient emissions in areas, particularly those sensitive to nutrient pollution. For example, as outlined in Section 6.3, the Netherlands applies stricter nutrient emission limits and manure regulations in nutrient-sensitive areas. It aims to reduce nitrogen deposition on at least 50% of the surface area of the nitrogen-sensitive nature areas within the Natura 2000 network to less than the critical deposition load by 2030. As noted earlier, the ecosystem costs of nutrient pollution vary greatly from one location to another. Furthermore, different measures will have different implications for linked watersheds. These need to be accounted for when selecting particular measures. For instance, New Zealand’s National Policy for Nutrient Management 2020 Amendment Act of 2024 provides specialised nutrient pollution targets for different water body types, including freshwater lakes and rivers, with accompanying species health targets for ecosystems that may be impacted by nutrient pollution, including fish and macroinvertebrates (New Zealand Government, 2024[315]). Also, the health and climate benefits of the measures are not generally part of the process that sets environmental targets. Integrating these into target setting for different locations would increase synergies between the different goals.
Making the implementation of measures for nutrient pollution management more effective will also increase synergies with other triple crisis objectives. While there are major gaps in tracking the environmental and other outcomes of different policies, some key findings emerge. One is that trust in new nutrient pollution management technologies and the way they work needs to be high (Li et al., 2023[316]). This requires their promotion to be accompanied with strong technical support. A second is that for regulations such as watershed management programmes, their adoption needs to be linked to accompanying measures such as information sharing, capacity building, technical assistance, training support for the local population and farmer-to-farmer communication networks that build trust and enhance understanding of the potential benefits of conservation practices (Piñeiro et al., 2020[317]). Agricultural extension services, both public and private, have been shown to have a positive impact on adoption rates. Third, subsidies, where they are offered to incentivise the adoption of particular measures need to take account of possible negative environmental and social consequences (Piñeiro et al., 2020[317]). For example, subsidies may increase the adoption of intercropping and residue mulching, but these practices may crowd out adoption of zero tillage. Finally, adoption of conservation methods depends a lot on them being profitable to the farmers in the short run (Piñeiro et al., 2020[317]). The evidence shows that when such technologies are offered in conjunction with measures that enhance the short-term profitability of agriculture (such as new crops, biological barriers and improved agricultural production), the adoption of conservation practices increases significantly. Considerations of resource efficiency and circular economy are also important (Box 6.10).
Phosphorus used as fertiliser is normally sourced from phosphate rock, a non-renewable resource. Therefore, the use of phosphorus through fertiliser, animal feed and other products leads to depletion of the fossil phosphate stock (Smil, 2000[318]; Van Vuuren, Bouwman and Beusen, 2010[319]). In contrast, nitrogen and its compound derivatives can be produced on a large scale as nitrogen is the most abundant gas in the Earth’s atmosphere and a renewable resource.
Reuse of phosphorus-containing waste and phosphorus recycling are increasingly supported at the policy level. In view of concerns over future phosphate rock availability, policy discourse has shifted from seeing phosphorus mainly as a pollutant in eutrophication control through downstream retention or removal of phosphorus in phosphorus landscape sinks towards increasingly considering reuse, recovery and recycling of phosphorus (Nanda and Kansal, 2021[320]; Roy, 2017[321]). This constitutes a move up in the phosphorus management hierarchy of Roy (2017[321]) that positions reuse and recovery through recycling provide above strategies where downstream retention is the primary goal (Figure 6.5). Bans of direct applications of sewage sludge containing phosphorus for agricultural application due to the concerns over contaminants (e.g. heavy metals) further underscore the importance of recovery through recycling.
In Germany, an importer of phosphorus, recycling from sewage sludge will be mandatory from 2029 for all wastewater treatment plants (WWTP) larger than 50 000 person equivalents (p.e.), equal to around 500 WWTP out of around 9300 WWTP (Bundesministerium der Justiz und für Verbraucherschutz, 2017[322]). These WWTPs will have to recover the phosphorus if the sludge contains more than 2% phosphorus per dry solid content or have to incinerate the sludge in mono-incinerators. Land application of sludge will only be allowed for WWTP below 50 000 p.e. (Bundesministerium der Justiz und für Verbraucherschutz, 2017[322]). By the end of 2023, municipalities in Germany were already obliged to report to the competent authority their planned measures for phosphorus recycling from January 2029.
Note: Preference in the phosphorus management hierarchy is a function of potential total benefits to the environment, society, and economy (Roy, 2017[321]). However, all three categories have a vital role to play in comprehensive phosphorus management and upstream source reduction remains crucial.
Source: Roy (2017[321]), inspired by EPA (2017[323]).
Phosphorus recycling and reuse are also supported at the regional level. Since 2017, phosphorus is listed in the European Union’s Critical Raw Materials list (EU, 2017[324]). The European Union 2020 Circular Economy Action Plan highlights the development of an Integrated Nutrient Management Action Plan (INMAP), with a view to stimulating the markets for recovered nutrients (European Commission, 2020[325]). The revised Urban Waste Water Treatment Directive also encourages EU member states to “prepare for reuse, recycling and other recovery of resources, in particular phosphorus and nitrogen” (European Parliament and Council of the European Union, 2024[326]). This directive also suggests that a minimum combined reuse and recycling rate should be set at EU level where members would be “able to choose whether to reuse or recycle, or both, urban wastewater or sludge, or both, in order to recover phosphorus” (European Parliament and Council of the European Union, 2024[326]).
However, one challenge in phosphorus recycling and reuse is obtaining phosphorus of consistently good quality, because of for example a variation in the quality of sewage sludge (Nanda and Kansal, 2021[320]). Concerns about energy efficiency of the recycling and reuse process have also been raised – although research progress has been made on enhancing phosphorus removal performance and its recovery potential and many technologies have been developed recently (Zheng et al., 2022[327]). Note: Nitrogen can also be recovered for reuse and recycling but is generally harder to recover at large scale in a useful, low-energy form because nitrogen readily changes oxidation states and is often lost as N₂/N₂O gas.
Agricultural development programmes for smallholder farmers in China demonstrate the potential effectiveness of combining technical support, information sharing, and subsidies to reduce nutrient pollution. In 2009, The China Agricultural University launched The Science and Technology Backyards (STB) programme, which places researchers and graduate students directly in farming communities to co-develop agricultural best practices that increase yields while minimising environmental impacts (FAO TECA, 2022[328]). Between 2019 and 2020, STB researchers collaborated with smallholder farmers and local government in Quzhou County to implement nitrogen pollution reduction initiatives. These included indoor technical training sessions, government subsidies for inhibitor-treated urea fertilisers and information on the expected yield benefits of deep fertiliser placement (Kang et al., 2023[329]). Compared with traditional practices, the co-developed methods increased nitrogen use efficiency for wheat and maize by 28% and 40%, respectively, while reducing NH3 emissions by 40% and 8% and improving profitability by 25% and 9% (Kang et al., 2023[329]). The STB programme continues to expand, reaching 127 programmes across 29 research institutions and 45 crop varieties as of 2018 (FAO TECA, 2022[328]).
Consultations with farmers can play a key role in developing synergistic nutrient pollution policies. In 2023, Canada held forums with farmers to discuss the voluntary national target of reducing fertiliser-related emissions to 30% of 2020 levels by 2030 (Agriculture and Agri-Food Canada, 2023[330]). The consultations revealed that concerns over potential agricultural losses could hinder adoption, and that incentives and best-practice sharing could help support implementation (Agriculture and Agri-Food Canada, 2023[330]). Canada’s On-Farm Climate Action Fund addresses these concerns over the potential agricultural losses by providing funding to farmers to support the adoption of beneficial management practices (BMPs) that decrease GHG emissions and/or increase soil organic carbon, helping farmers tackle climate change. For example, the fund allocated CAD 50 million to the Ontario Soil and Crop Improvement Association to support the adoption of farming practices such as nitrogen management and rotational grazing, which provide climate and nutrient benefits (Agriculture and Agri-Food Canada, 2025[331]). Although the fund was recently launched (2022), early synergies are already apparent. One farm in Ontario, for instance, used the grant to adopt a wide-drop sprayer, enabling a targeted 28% reduction in nitrogen use, allowing the farm to cut fertiliser costs while maintaining yield (Ontario Soil and Crop Improvement Association, 2025[332]).
Given that synergies of nutrient pollution policies with climate and biodiversity benefits are strong, the question that arises is where these synergies are strongest and which policies are best suited to take advantage of them. Several findings emerge:
Using cost-effective measures to ensure the lowest possible costs of nutrient reductions, in overall terms (i.e. taking into account synergies and trade-offs), can allow for higher targets to be set without increasing the burden on governments and society more widely. Here there is a role of market-based instruments such as a nitrogen/phosphorus tax or a tax on NH3 emissions. These have been studied and rates proposed but they have not been implemented. It would be worth looking more closely at the proposals and introduce such a tax – something the European Commission is currently examining (European Commission, 2024[333]). As above, the targets and tax rates should reflect the broad environmental impacts, and not only the immediate pollution benefits. Alongside market-based instruments, more detailed spatial planning could help create targeted local policies. Denmark has implemented a spatially targeted nutrient management system to reduce nitrogen runoff from agriculture into rivers, lakes, and coastal waters. Rather than applying uniform national limits, Danish authorities use detailed spatial maps that identify areas most vulnerable to nitrogen leaching and eutrophication (Hashemi et al., 2018[334]).
While taxes on fertilisers have had an effect in reducing nutrient application and thus nutrient emissions, they have fallen out of favour and are unlikely to be taken up. Where there is potential for making a change, however, is to reform subsidies to fertiliser use in countries where such subsidies are used. Repurposing such subsidies to other forms of support to farmers can help reduce nutrient pollution while not harming farmers’ livelihoods. There are moves in this direction. The World Bank, for example has given grants to pilot initiatives to realign fertiliser support and soil health programmes in Bangladesh, Ghana, Indonesia, Mozambique, Malawi and Tanzania (World Bank, 2024[335]).
Carbon taxes on fossil fuels can also reduce nutrient emissions by indirectly raising the price of nutrients through the higher production costs of fertilisers. The levels of such taxes and extent of coverage varies a lot across countries. Moving carbon tax levels closer to the estimated social cost of carbon (SCC) would make a substantial contribution to reducing nutrient pollution and its attendant biodiversity and health impacts. Most national governments review the evidence on SCC and from that recommend a range of values that could be used in such an exercise. As outlined in Section 6.5.1, potential trade-offs arise on agricultural income and food prices, although these might be limited with improved nitrogen use efficiency.
Environmental assessment tools for nutrient pollution
Nutrient flow assessments can support countries in developing strategies that promote synergistic outcomes. For example, in 2021, the EU Joint Research Committee collaborated with Member States to collect data regarding nitrogen and phosphorus flows, as well as national progress towards EU nutrient waste targets. This data was used to forecast the impacts of potential nutrient pollution interventions on biodiversity, water and air quality and GHG emissions (European Commission, 2025[336]). The European Commission will use this analysis to develop Integrated Nutrient Action Plans to help each Member Country meet the European Union’s goal to reduce nitrogen and phosphorus losses to the air, water and soil by 50% by 2030 (European Commission, 2025[336]). This objective is aligned with differentiated criteria outlined in the European Union’s Biodiversity Strategy to 2030, the Farm to Fork Strategy and the Zero Pollution Action Plan (European Commission, 2025[336]).
Countries can also conduct SEAs for policies targeting nutrient pollution to improve their practicability and effectiveness. Integrating climate and biodiversity considerations into Strategic Environmental Assessment criteria can identify additional opportunities to refine policies and advance synergistic outcomes. For example, in 2025, Ireland conducted a Strategic Environmental Assessment for its Fifth Nitrogen Action Programme (NAP) which found that, although the policy was likely to generate synergies across the UN Sustainable Development Goals, including those pertaining to climate and biodiversity, increased guidance on thresholds would support more effective implementation (RPS, 2025[337]). In response, Ireland committed to modelling the NAP’s impacts on water quality, a step that could inform the establishment of thresholds for future NAPs (RPS, 2025[337]).
This chapter highlights how some of the principles frameworks and mechanisms discussed in earlier chapters can be worked out more concretely when diving deeper into specific cases. The detailed case studies on renewable energy projects, protected area management, air pollution control policies and nutrient management clearly show how many environmental projects and proposals are inherently local, while contributing to significant global benefits when they are well designed and implemented. Each of the deep-dives presented above allows analysis of specific aspects of how the triple crisis interlinks:
the case of renewable energy expansion emphasises the need to identify the correct counterfactual, and compare effects of projects and policies to what the alternatives may be; this includes embedding spillover effects of the policy on the other aspects of the triple crisis not directly targeted, i.e. beyond climate change mitigation;
the case of protected areas highlights the risks that unmitigated environmental damage can do to an inherently synergistic policy – e.g. loss of pollution filtering functions due to climate change – thereby undermining not only co-benefits for the other aspects, but also the effective contribution to the core policy objective, in this case biodiversity protection;
the case of air pollution control measures and their co-benefits for climate mitigation and biodiversity conservation demonstrates the importance of a multi-effect, pollutant-specific approach to enhance synergies;
the case of nutrient pollution control highlights that environmental impacts are linked in various ways, from soil and water and air in the case of nitrogen and its compound derivatives, to pollution and resource management for phosphorus.
In all cases, the deep dives highlight some core aspects of integrated management of the triple crisis in both design and implementation. Some key considerations are presented in Figure 6.6.
Source: Authors’ own elaboration.
Five key policy recommendations can be formulated on the basis of the insights of these case studies – these key insights inform the roadmap for policy action developed in the next chapter.
First, synergistic outcomes are often possible but require active consideration and integrated management to materialise. A systematic approach can ensure that all relevant possible effects are considered; broad environmental impact assessments can help structure the assessment. This requires careful integration and alignment of policy instruments and tools for measurement and monitoring both vertically (across geographical scales – from local to regional to national and from airshed/watershed to global) and horizontally (i.e. by looking at spillover effects on other sectors, other regions and other environmental domains). Identifying the major trade-offs and synergies associated with a project or policy is essential to ensure they are taken into account.
Second, synergistic policy design does not necessarily mean replacing existing policies with new ones, but rather leveraging existing approaches, policies and modalities better in light of their synergistic effects and trade-offs. Using an appropriate policy instrument mix can prevent many trade-offs; this includes choosing an appropriate number of instruments to deal with different objectives, not least avoiding using one instrument to try and achieve many objectives, or having overlapping instruments that aim at the same objective. A clear hierarchy can perhaps not be established, but especially at the policy design phase, a careful consideration of the existing situation and relevant alternatives is merited. To generate biodiversity benefits, one can assign new areas to protect, but expansion of existing areas or revision of current areas to improve their connectivity are also relevant options that should be weighed. Countries can use environmental assessment tools both to help formulate and strengthen the effectiveness of policies and evaluate alternatives. Tools like environmental impact assessments and natural capital accounting can help countries determine the value of natural processes. Integrating considerations from all elements of the triple planetary crisis into these tools can help countries identify priority areas for policy consideration.
Third, (cost-effective) policies that provide incentives to address all aspects of the triple crisis are to be prioritised. Market-based policy instruments can provide such incentives, but may not always trigger sufficient responses or be administratively cumbersome, so an optimal toolkit contains a variety of policies. Hotspot areas where potential environmental impacts are concentrated are especially of concern and can be targeted with priority. Countries can for instance leverage spatial planning to create priority and non-priority zones to help maximise synergies while the avoiding potential downsides associated with burdensome legal requirements and limit environmental evaluations to those instances where trade-offs are potentially significant. Also important for cost-effectiveness is to weigh the cost and benefits of implementation.
Fourth, there is a need to measure a range of indicators for climate change, pollution and biodiversity impacts, to be able to identify synergies and trade-offs. Evaluation hinges on two key aspects (i) robust measurement of impacts and (ii) establishing the appropriate alternative. Renewable energy projects may lead to some unwanted impacts on pollution and biodiversity, but if the alternative is coal mining and coal power generation, then the relevant consideration is the difference in environmental impacts. Although it is not straightforward to measure the environmental impacts of renewable energy expansion and knowledge gaps remain (e.g. assessments of cumulative impacts), countries are adopting various approaches to reconcile the need to accelerate the expansion with more robust consideration of the risks. Examples include identifying priority areas for renewable expansion in which permitting processes are streamlined, as well as enhancing the circularity of end-of-life solar panels and wind turbine blades.
Fifth, policies need to be aligned across different scales to ensure consistency. Sound design and implementation will have to consider the relevant scope of the measure, both geographically (e.g. airsheds, cross-border rivers) and sectorally, to avoid undermining policy effectiveness and to exploit available synergies. Anticipating local impacts before crafting policy to proactively identify and address potential pitfalls in local implementation can be key to smooth implementation. Consultations that assess local authorities’ operational and technical capacity can help countries craft practicable policies while helping countries invest in required capacity building.
Many countries have established national targets and policies that integrate elements of the triple planetary crisis. Implementation at the local level, using environmental assessments and vertical integration, can be challenging but can also create important synergies of a truly integrated policy landscape, while avoiding trade-offs where possible.
EIAs are mandated for major development projects above a certain threshold (e.g. spatial requirements). It is defined in dual terms as: (i) a technical tool for assessing the environmental impact of proposed projects and (ii) an institutional procedure often linked to decision-making (International Association for Impact Assessment, 2009[338]). The use of EIA for renewables projects is already widespread across countries. At the level of the EU, the EIA Directive (Directive 2011/92/EU, as amended by the Directive 2014/52/EU) sets out the guidance for the procedures and high-level criteria for determining if projects require an EIA, including characteristics and location of projects (e.g. a proposed renewable energy project in a Natura 2000 site, which are designated protected areas for biodiversity conservation). Member states translate the guidance into more detailed criteria and thresholds such as the height of a mast and installed capacity. In Germany, for instance, wind power projects consisting of more than 20 turbines exceeding 50 metres (m) in height must go through an EIA process (Schumacher, 2019[339]). In France, projects that include at least one wind turbine generators with a height of more than 50 m or a mast between 12 and 50 m and installed capacity exceeding 20 MW are subject to an authorisation process (DGPR et al., 2019[340]).
In some countries, EIA requirements can be extensive in application of the mitigation hierarchy. For instance, the National Energy and Climate Plan (NECP) in Spain mandates every renewable energy project to undergo an EIA to obtain an Environmental Impact Declaration, which sets out a set of conditions and measures for implementation. These include preventative and corrective measures during construction and operation phases, as well as compensatory measures to offset residual impacts that could not be fully eliminated. Continuous monitoring throughout the projects is also mandated.
There can be a substantial variation in the scope of EIA within a country, including what triggers the EIA screening and the timelines (McMaster et al., 2021[92]). For example, in the province of Ontario, windfarm developers must obtain a Renewable Energy Approval (REA) from the Ontario’s Ministry of the Environment, Conservation and Parks for facilities with a nameplate capacity over 3 kW (Ontario, 2024[341]).
In the UK, large-scale onshore wind power projects exceeding a production output of 50MW or above are typically subject to an EIA. Wind power installations composed of more than two turbines or exceeding a production output of 5MW are subject to an EIA screening procedure.
In Japan, an EIA is mandatory for wind power plants of capacity greater than 50 MW and for solar projects of at least 40 MW while for wind plants between 37.5-50 MW and for solar projects of 30-40 MW the government it is decided on a case-by-case basis after a thorough screening process (Ministry of the Environment/Japan, 2022[342]). In New Zealand large-scale wind power farms that are considered to be of “National Significance” are subject to the New Zealand Environmental Protection Authority (NZEPA) evaluation that includes EIA called an “Assessment of Environmental Effects”.
The IUCN has developed six categories to classify protected areas based on their management objectives (Dudley, 2008[343]). These categories regulate human activities with varying levels of stringency, ranging from strict protection in Category Ia areas to sustainable use of natural resources in Category VI areas:
Other Effective Area-Based Conservation Measures (OECMs) are geographically defined areas distinct from protected areas that are governed and managed in ways that achieve positive and sustained long-term outcomes for the in-situ conservation of biodiversity, with associated ecosystem functions and services and, where applicable, cultural, spiritual, socio-economic, and other locally relevant values (Convention on Biological Diversity, 2018[344]). Unlike protected areas, OECMs may not have the primary objective of biodiversity conservation (such as sustainable resource management, cultural preservation, or human-wildlife conflict mitigation) but deliver effective conservation outcomes as a by-product of their management.
Key Biodiversity Areas are sites that contribute significantly to the global persistence of biodiversity. These areas are identified based on rigorous scientific criteria, including the presence of threatened species, geographically restricted species, ecological integrity, biological processes, and irreplaceability. KBAs are not legally designated but serve as a tool for conservation planning and prioritisation (IUCN, 2016[345]).
Natura 2000 is an ecological network of protected areas and the key instrument to protect biodiversity in the EU, designed to ensure the long-term survival of Europe’s most valuable and threatened species and habitats. Based on the Birds Directive (1979) and the Habitats Directive (1992), the network includes Special Protection Areas (for wild birds and their main habitats) and Special Areas of Conservation (for most in need habitats and species, excluding birds) and, collectively aiming to maintain or restore natural habitats and species at a favourable conservation status (European Commission - Environment, 2015[346]).
The IUCN Green List is a global standard for protected and conserved areas that demonstrates effective management and equitable governance. It aims to recognise and promote areas that deliver successful conservation outcomes through fair governance, sound design, and effective management practices (IUCN and World Commission on Protected Areas (WCPA), 2017[347]).
Complementing regulatory measures that are already relatively widespread, there are also emerging examples of economic and information policy instruments countries deploy to reduce risks for biodiversity and pollution associated with renewables. Notably, some of these instruments can be designed in a manner that simultaneously supports consideration of biodiversity and pollution risks.
Taxes with a rate proportional to impacts can incentivise actions to lower risks of biodiversity loss and pollution. While many countries have taxes on solar panels or wind turbines for which the rate is proportional to installed capacity, environmental taxes that are directly proportional to environmental impacts remain rare. Belgium presents an exception, where case law on the tax on wind turbine masts calls on local authorities to determine its value on the basis of “the extent of the environmental and landscape impact caused by the mast and blades of the wind turbine” (OFB, 2023[104]).
Conditional subsidies can incentivise renewable power and utilities companies to prevent pollution and alleviate impacts on biodiversity. Research and Development (R&D) subsidies can also be used to enhance the knowledge and evidence base of the impacts of renewables on biodiversity and pollution. Technological and data advances can play an important role in mitigating the environmental risks associated with renewables expansion. Subsidies can also promote research, development and demonstration of these technologies, such as AI bird-identification technology to support wind turbines’ shutdown-on-demand (OECD, 2024[2]).
In France, the national research agency has provided support to research projects such as the ePARADISE, which aims to develop a sensor for measuring the airflow near wind turbine blades, optimising their operations and minimising acoustic emissions (IFPEN, 2022[348]). In a similar vein, subsidies can support decision-support tools to prevent the adverse environmental impacts of renewables expansion. For instance, the UK Research and Innovation (UKRI) has awarded Plymouth Marine Laboratory with an approximately GBP 300 000 (~ USD 363 000) grant to fund the development of the Offshore Renewable Impacts on Ecosystem Services (ORIES) tool that helps stakeholders understand the impacts of planned offshore wind installations on marine biodiversity and ecosystem services.
Countries have also provided subsidies to enhance circularity of renewable technologies. Through its Circular Business Models for the Solar Power Industry (CIRCUSOL) programme, the European Commission has funded a range of demonstration projects to re-use, repair and refurbish end of life solar panels (CIRCUSOL, 2024[349]).
Similar developments are widespread for improving the circularity of wind turbine blades. In Spain, among the Strategic Projects for Economic Recovery and Transformation (Proyectos Estratégicos para la Recuperación y Transformación Económica - PERTE), there was a call in 2022 for projects to support innovative facilities for wind turbine blade recycling (Ministerio para la transición ecológica y el reto demográfico, 2021[350]). Furthermore, there will be an additional call with a budget of EUR 100M to support facilities for recycling not only wind turbine blades but also PV panels and batteries in 2025 within the PERTE framework. In France, the Zero wastE Blade ReseArch (ZEBRA) project seeks to develop a thermoplastic resin that, combined with a high-performance glass fibre, could be chemically recycled.27 Funded by the European Union, the EoLO-HUBs project, with 18 European partners from 7 countries, aims to propose and demonstrate novel solutions to recycle high value materials from the wind turbine blades, developing a set of innovative composite material recycling technologies. Similarly, in the United Kingdom, the project PRoGrESS – a three-year scheme, that seeks to deliver a circular model of wind turbine blades.
Green public procurement – is increasingly seen as a strategic lever for achieving environmental sustainability (OECD, 2024[351]). Call for tender for large renewable projects is a key policy tools used by countries to subsidise the deployment of large-scale renewables, providing an opportunity for integrating biodiversity and pollution consideration in procurement and tendering.28 For instance, in Spain, non-economic criteria will account for up to 30% of the overall rating in future calls for tender, allowing for the consideration of environmental impact, resilience, and local value chains.
Similarly, in the Netherlands, applications of the two wind farms in the Hollandse Kust West Wind Farm Zone were assessed based on the contribution of the project to the ecology of the North Sea, which accounted for 50% of the total points available (OECD, 2024[2]). There are also examples of assigning a score based on the budget allocation. For instance, in France, tenders for offshore wind projects are scored based on the budget they allocate to environmental measurement and a biodiversity fund.
Environmental labelling and information schemes (ELIS), defined as “policies and initiatives that aim to provide information about one or more aspects of the environmental performance of a product or service to external users” (Gruère, 2013[352]), can also help integrate the biodiversity and pollution impacts of renewables. Within ELIS, “ecolabels"29 can help identify environmentally preferable options. There are two main types of labels for the environmental quality of projects. The first type of labels is “producer-oriented” and includes those certifying the electricity produced at source. To be labelled, producers must meet certain eco-responsibility criteria, including the preservation of biodiversity and natural ecosystems, and public participation. For example, in the Netherlands the EcoCertified Solar label certifies that soil health and biodiversity are preserved (Wageningen University, 2025[353]). EcoCertified Solar excludes Natura 2000 and UNESCO-listed sites and labels projects that prioritise installation in cleared and degraded lands and spare natural habitats (OFB, 2023[104]). EcoCertified Solar label further requires an environmental management plan, including the complete restoration of the site after dismantling, prevention of soil compaction as well as the monitoring and evaluation of the effectiveness of the avoidance, mitigation and compensation measures.
The second type of labels is “consumer-oriented” and certifies the electricity distributed to consumers by renewable energy suppliers by providing greater transparency for buyers (OFB, 2023[104]). These labels can be used to enhance the effectiveness of Guarantees of Origin systems used in the jurisdictions including in the European Union and the United Kingdom. For example, the “EKOenergy” label developed in Finland is an internationally recognised label that certifies that electricity commercialised by labelled suppliers come from renewable sources that respects precise criteria such as the exclusion of production in protected areas. EKOenergy also excludes Natura 2000 sites, UNESCO-listed sites and Important Bird Areas according to BirdLife. In Switzerland, “NatureMade Star” labels both the electricity produced, and the electricity distributed to consumers, excluding as a matter of principle for instance wind farms located in Swiss regions listed in the Federal Inventory of Landscapes, Natural Sites and Monuments of National Importance. Another example is the EPEAT Registry, an ecolabel which provides information on photovoltaics to facilitate choice of solar panels that meet standards on chemical substance management, material circularity, water use and carbon footprint (Global Elecronics Council, 2025[354]).
Ecolabels can be used in conjunction to reinforce the incentives provided by economic instruments, as well as with other tools such as green power purchase agreements. However, ecolabels require a balance between ease of implementation and ecological requirements that are key to avoid any risk of greenwashing. In this context, “EcoCertified Solar”, “EKOenergy”, and “NatureMade Star” all conduct ex post evaluation and audits of production sites to determine whether they comply with the criteria set out in the specifications, which contributes to trust in the information the labels provide.
Green power purchase agreements can help take into account biodiversity and pollution in long-term electricity purchase agreements. In some countries, long-term contracts of about 15 years have been used to finance the development of renewable energy production facilities, making it possible to take a large part of the lifecycle of a renewable energy installation into consideration. These long-term contracts can help reconcile the challenges of preserving biodiversity with developing renewable energy in the region. For example, some renewable energy producers, such as the Danish solar energy company “Better Energy”, offer long-term contracts that contain biodiversity criteria including the choice of renewable energy sites that avoid sites with high ecological risks; preservation of groundwater, soil, flora and fauna; pesticide-free vegetation management at solar PV plants and measures to maintain wild flora and fauna in a good state of conservation (OFB, 2023[104]).
Countries can additionally leverage the auction process to incentivise developers to proactively integrate biodiversity and pollution considerations into renewable energy proposals. For instance, the European Union’s Net-Zero Industry Act of 2023 mandates that public authorities responsible for renewable energy auctions consider non-price criteria (NPC) in at least 30% of the auctioned volume or 6 gigawatts in each country (European Commission, 2024[355]). In 2024, Denmark’s NPCs foregrounded upstream and downstream pollution by including LCA, environmental monitoring, and nature inclusive design into their auction criteria (Renewables Grid Initiative, 2024[356]). Additionally, in 2024, the Netherlands released a permit auction with NPCs for biodiversity. The winning consortium, Noordzeker, partnered with the Dutch research institute for biodiversity to create a proposal that integrated biodiversity considerations by introducing measures to reduce marine mammal disturbance during construction and operation of the windfarm as well as integrating artificial reefs on 75% of the wind turbines (Netherlands Enterprise Agency, 2024[357]).
Landfill taxes can discourage the disposal of end-of-life wind turbine blades and stimulate more environmentally sustainable waste management, i.e. include biodiversity and pollution considerations in renewable energy projects. The cost of landfills for composite materials is around EUR 120 per tonne in Denmark and around EUR 130 per tonne (including a tax of 95 EUR per tonne) in the United Kingdom (Beauson et al., 2022[358]). However, in some cases, the cost of recycling composite waste is much higher than the landfill taxes for waste, suggesting that the current rate may only provide an inadequate incentive for recycling and sound waste management (Majewski et al., 2022[359]).
Extended producer responsibility (EPR) is “an approach that makes producers responsible for their products along the entire lifecycle, including at the post-consumer stage” (OECD, 2024[360]). Several countries including Germany and France have adopted the EU Waste regulations from Waste from Electrical and Electronic Equipment (WEEE) directives for management of end-of-life solar PV, making manufacturers responsible for the cost of collection and recycling of PVs (Ngagoum Ndalloka et al., 2024[361]). Unlike solar panels and batteries, wind turbine blades are very large structures and cannot be collected together with municipal waste streams, which has limited the emergence of dedicated EPR schemes thus far. For instance, in the EU WEEE Directive, wind turbines blades are excluded as they are considered “Large Scale Fixed Installations” although some electric and electronic components of wind turbines, such as generators and cables are within the scope of the Directive (Majewski et al., 2022[359]). Nonetheless, future EPR systems for wind turbine blades are conceivable, for instance, building on the existing EPR for end-of-life vehicles (Alexandre, Follenfant and Legait, 2019[362]).
Biodiversity offsets are measurable outcomes that result from actions to compensate for residual risks to biodiversity (OECD, 2016[363]). They can be categorised into (i) restoration offsets and (ii) averted loss offsets, which respectively refer to remediating damage and protection and maintenance of existing biodiversity (Bennun et al., 2021[47]). They can be used to mainstream biodiversity in renewable energy projects (OECD, 2024[2]). For example, restoration has helped the expansion of habitats for golden eagles at the Beinn an Tuirc Wind farm in Scotland (UK) (Wind Europe, 2017[364]).
However, offsetting is the last step in the mitigation hierarchy. Use of offsets should not be made at the expense of other instruments that are more aligned with the mitigation hierarchy. Importantly, offsets cannot be used to address the loss of irreplaceable or highly vulnerable biodiversity. Biodiversity offsets have in practice not often worked as intended and can therefore be considered with caution. The effectiveness of biodiversity offsets is difficult to ascertain; for instance, many offsets assume a certain background rate of biodiversity decline and implicitly counts aversion of this decline as gains (Maron et al., 2018[365]). Conservation banks (also known as biobanks) aim to deliver biodiversity gains in advance, before the impacts occur but are not widely adopted (Bennun et al., 2021[47]). Achieving equivalence of losses and gains is difficult in practice, as compensating for permanent losses, for instance, would require offsets to deliver in-perpetuity protection with ongoing funding if they are to achieve no net loss meaningfully (Maron et al., 2025[366]). In this context, there is a need for better understanding of what can be offset in practice. Offsets are meant to deliver a quantifiable benefits to compensate for the same amount of damage; lack of efficacy and shortfalls of offsets therefore constitute not only an opportunity cost, but also a realised loss of biodiversity that might not otherwise been permitted if it was not for the introduction of offsets (Maron et al., 2025[366]).
Target 3 of the Kunming-Montreal Global Biodiversity Framework aims to ensure the effective conservation of at least 30% of terrestrial and inland water areas by 2030. Achieving this target will require the designation of new protected areas to complement the current coverage of 17.6% of terrestrial and inland waters and 8.4% of marine and coastal areas (UNEP-WCMC and IUCN, 2024[180]).
Despite the designation, protected areas do not always effectively deter biodiversity loss caused by anthropogenic activities, and downgrading (i.e. reduction of legal restrictions on human activities), downsizing (i.e. decrease in its size) and ‘degazettement (i.e. loss of legal protection) of protected areas (PADDD) have affected both terrestrial and marine protected areas. Aska et al. (2023[367]) examined 1 743 tailing storage facilities from mining operations, which account for 36% of global mineral commodity production, and found that 9% are located within protected areas. Half of these facilities were established after the designation of the protected area. Of these, 40% were in IUCN Category VI areas, while a concerning 10% were in Category II areas (see definitions in Annex 6.B). Within marine protected areas, 43 PADDD events have been enacted until 2020, among which in Australia, Dominican Republic, Palau, and South Africa (Albrecht et al., 2021[368]).
Mainstreaming protected area management into national policies and decision-making frameworks is essential, alongside its incorporation into sectoral plans and strategies (UNDP, SCBD and UNEP-WCMC, 2021[369]). The integration of protected areas entails a two-fold process (Ervin, 2010[370]): (i) linking protected areas within a broader network of protected and managed lands and waters in order to maintain ecological processes, functions and services, such as ensuring their representativeness, consistency and connectivity; (ii) reducing habitat fragmentation and supporting the connectivity of networks of protected areas is key to ensure their effectiveness of conserving biodiversity, and their capacity to adapt to the effects of climate change.
Annex Figure 6.C.1 illustrates the concept of integration and mainstreaming of protected areas at the national level. The first level of the integration of the information on protected areas consists in measuring and collecting their extent and condition, and how much their coverage represents compared to the national terrestrial/marine area. The second level, so far rarely implemented, is to assess the ecosystem services of protected areas systematically, including at upper jurisdictional scales (local, national, supranational and international). The third level is to consider protected areas as part of a network, and to assess how representatively and consistently biodiversity is conserved and how many ecosystem services are provided when all networks are aggregated. The fourth level of integration is to account for migration corridors of species, to ensure that the network of protected areas is well-connected. The monitoring framework of the Kunming-Montreal Global Biodiversity Framework proposes the two following global indicators to assess the ecological representativity, and the well-connectedness of terrestrial protected areas (Hilty et al., 2020[371]): the Protected Connected land indicator (ProtConn) and the PARC-Connectedness indicator.
Incorporating protected areas into broader land-use plans, natural resource policies, and national strategies can maximise biodiversity benefits, enhance ecosystem services, and mitigate threats to biodiversity. Protected areas can be integrated into National Biodiversity Strategies and Action Plans (NBSAPs), National Adaptation Plans (NAPs), and Nationally Determined Contributions (NDCs). Embedding protected areas into these strategies facilitates vertical and horizontal integration, enabling the tracking of biodiversity goals alongside co-benefits for climate change and pollution. According to the Intergovernmental Panel on Biodiversity and Ecosystem Services (IPBES), mainstreaming biodiversity involves incorporating biodiversity-related actions or policies into broader development processes, such as poverty reduction or climate change mitigation. (Secretariat of the Convention on Biological Diversity, 2016[372]).
Protected areas monitoring can serve as a cross-sectoral indicator within national adaptation plans, relevant to agriculture, forestry, water management, infrastructure, and tourism (OECD, 2024[373]). For instance, the UK has included the protection and improvement of protected sites and other areas of important wildlife habitat as an adaptation objective, linked to the target of restoring 75% of terrestrial and freshwater sites to favourable condition by 2042 (Department for Environment, Food and Rural Affairs, 2023[374]).
Protected areas can create negative spillover effects on surrounding regions, called “leakage” or “displacement” effects. For example, restrictions on land or sea use within protected areas may displace harmful activities to unprotected areas, such as deforestation (Murray, McCarl and Lee, 2004[375]) or fishing (Sen, 2010[376]), and causing unintended degradation. Within a random selection of 120 protected areas established before 2001, in tropical forests of America, Africa and Asia, deforestation rates between 2001 and 2017 were higher in about half of the 10 km buffer zones of protected areas compared to inside them or in control areas (determined by propensity score matching) (Ford et al., 2020[377]). In more than three-quarters of these documented leakage cases, reduced deforestation in protected areas was not sufficient to offset the amount of deforestation in these buffer zones to a level that would be expected without protection.
Nevertheless, assessing the causal effects – and associated spillovers - of protected areas is challenging mostly due to data availability, selection bias and heterogeneity in treatment effects. Indeed, the causal estimation of the impacts of protected areas is complex due to the (i) the need of comparable conservation metrics and indicators across protected and unprotected areas, potentially covering a wide geographic area and hence different biodiversity and ecosystems ; (ii) site selection issues may arise since the status of protected areas is not randomly distributed and is often designed in higher biodiversity zones, with lower habitat conversion pressures and lower accessibility (Brodie et al., 2023[378]; Cooke et al., 2023[379]; Reynaert, Souza-Rodrigues and van Benthem, 2024[380]); (iii) the effects of protection may vary by type of ecosystem, economic restrictions, enforcement stringency, year of implementation and makes it difficult to build an average treatment effect (Vincent, 2015[381]; Reynaert, Souza-Rodrigues and van Benthem, 2024[380]).
Source: Authors’ own elaboration.
Policies to address air pollution can be classified as regulatory, legal, economic and social. They can act at the local, regional, national or international levels. Policies are implemented in packages, which include combinations of each kind, with action often taken at more than one level.
There is a very large range of policies that address emissions to air. Annex Figure 6.C.2 summarises the main pathways by which these are conducted. Primary among them are regulations that set emissions limits – e.g. to the air from vehicles, factories or power stations. Regulations also exist that set limits on concentrations of pollutants in different locations. Furthermore, governments impose restrictions on activities in different areas and at certain times, such as driving polluting vehicles in high congestion zones or burning waste when there is a fire risk. Lastly, there is a large set of fiscal instruments (taxes and subsidies) to encourage reduced emissions from several of these sources.
Source: Authors’ own elaboration.
Emissions standards for air pollution have generally been effective in lowering concentrations of key pollutants such as PM2.5 and NOx, especially in developed countries. Death rates from PM2.5 have been declining widely (India is a notable exception). In the EU, for example, in 2021, 253 000 premature deaths in the EU were attributable to fine PM, a 41% drop since 2005 (European Environment Agency, 2023[382]). Emissions of sulphur have also fallen but the pattern has been more mixed. From 1990 to 2015, Europe had the largest reductions in sulphur emissions in the first part of the period while the highest reduction came later in North America and East Asia. However, emissions from East Asia increased from 2000 to 2005 followed by a decrease, while in India a steady increase over the whole period has been observed (Aas et al., 2019[383]). The decreases in PM and NOx have been accompanied by a reduction in CO2, which represents a lowering of the warming effect. The fall in sulphur has the opposite effect but its magnitude is uncertain.
The use of fiscal instruments such a pollution taxes, permits and subsides has seen a considerable increase worldwide. The latest OECD PINE database identifies over 4 500 policy instruments relevant to environmental protection and natural resource management in 146 countries globally (OECD, 2024[176]), up from 3 900 policy instruments in 2023 (OECD, 2023[384]). Environmentally related taxes and fees are expanding in all countries. They represent over a third of the instruments in the PINE database. This now includes around 900 taxes and fees in 146 countries. Other economic instruments that address air pollution include tradable permits and offsets and voluntary approaches.
Although the number of instruments is large, information on their direct impacts is limited, and their spillover effects even more so. Studies on the direct effects have found that when taxes are high relative to abatement costs, they have reduced emissions considerably (e.g. taxes on SO2 and NOx in Sweden) (European Commission, 2024[385]). More commonly, however, where tax rates are low relative to abatement costs, the impacts are much smaller. This is the case for example in a number of European countries -- Italy, France and Spain, amongst others (Withana et al., 2014[386]). The use of waste as source of energy through incineration with energy recovery has increased with a reduction of GHGs.
Subsidies to encourage a switch to less polluting production and consumption have also been introduced, although less than taxes. The PINE OECD database identifies 215 environmentally beneficial subsidies and payments across 146 countries. Subsidies have a fiscal cost, which can be distorting to the economy. They also provide an incentive to increase production, which can raise pollution levels. In general, economists find such subsidies less effective than taxes in addressing pollution problems (Stavins, 2003[387]). An example of a subsidy is the scheme to get owners to scrap old more polluting cars for newer, less polluting ones which has been introduced (temporarily) in several countries. An ITF review of schemes finds that such schemes have a synergistic effect in reducing CO2 and air pollutants; the costs are relatively high, however, and do not make up for the value of the scrapped cars (ITF, 2011[388]). A car scrapping and recycling scheme was introduced in Cairo between 2013 and 2017 and has been successful in reducing CO2 emissions there by about 350 kilotons by 2018 (World Bank, 2018[389]) but it has not been evaluated in terms of its effects on air pollution.
Some jurisdictions set out frameworks for ensuring consistency of air pollution policies with climate and biodiversity objectives. For instance, the revised Ambient Air Quality Directive in the EU, setting higher ambition and stricter air quality standards for 2030 and the goal of zero pollution by 2050, highlights the need for consistency between air pollution policy and other policy domains, including climate and biodiversity (European Commission, 2024[390]).
Targets for nutrient inputs and emissions
Policies to mitigate nutrient pollution are often set in terms of targets for inputs in (agricultural) production (e.g. fertiliser use) or in terms of emission levels. The United Kingdom has a target of reducing nitrogen and phosphorus pollution from agriculture into the water environment in England by at least 40% from a 2018 baseline by 2038, with an interim target of 10% by 31 January 2028 (UK Department for Environment, Food and Rural Affairs, 2023[391]). It also aims to reduce phosphorus loadings from treated wastewater by 80% by 2038 against a 2020 baseline, with an interim target of 50% by 31 January 2028 (UK Department for Environment, Food and Rural Affairs, 2023[391]). The European Union includes several targets by 2030: a 50% reduction in nutrient losses (known as Farm to Fork strategy) without deterioration of soil fertility, a 20% reduction in fertiliser use, as well as a 25% reduction in the total ecosystem area where air pollution threatens biodiversity relative to 2016 (key Green Deal ambitions under EU law). In addition, targets are set for individual nitrogen-based air pollutants: a 13% reduction in NOx emissions and an 11% reduction in NH3 emissions by 2030 (European Parliament and the Council of the European Union, 2016[392]). Globally, there is the International Nitrogen Initiative (INI) to reduce nitrogen waste by 50% by 2030. This initiative is working with governments, businesses, and citizens to achieve this goal outlined in the Colombo Declaration on Sustainable Nitrogen Management (UNEP, 2019[393]).
While targets for NOx reductions have been met for most EU member states, this is not the case for NH3.30 More widely, there has been an increase in the production of NH3 but with a small reduction in NH3 emissions. Across the OECD region, NH3 emissions decreased in the period 2003-15 by 3.2%. This was a slower rate than during the period 1993-2005 when the decrease was 8.6% (OECD, 2019[394]). Since 2005, in many EU member states, NH3 emissions have decreased only slightly or in some cases even increased, with almost one third of EU member states having emission levels above the emission reduction commitments for 2020-2029 and requiring emission reductions to meet the commitment (EEA, 2024[395]).
Phosphorus surpluses decreased across almost all OECD countries since the 1990s, driven by reduced phosphorus fertiliser application rates, livestock population, crop-mix changes, and policy interventions, but phosphorus pollution remains at an alarming level (OECD, 2019[394]). Phosphorus fertiliser application rates fell for most OECD countries since the 1990s, possibly as a result of improved farm practices or usage of alternative fertilisers, but (as outlined in Chapter 2) were still substantial across all regions in 2020. Phosphorus losses from land to fresh waters have doubled in the last century and continue to increase in many regions (Our Phosphorus Future Network, 2022[290]).
Since the 1960s, the application of mineral nitrogen fertiliser has been increasing globally across all regions. Many OECD countries have used a system of high fertiliser use and high yields, and differences in fertiliser inputs with emerging and developing economies have persisted. Thus, OECD countries tend to release large amounts of reactive nitrogen to the environment but also recovering a greater portion of added nitrogen in crops. This recovery efficiency has not changed noticeably, implying additions of nitrogen to the environment have been going up (Conant, Berdanier and Grace, 2013[396]). Furthermore, differences in nitrogen absorption and utilisation efficiency between OECD and other countries has persisted. Improvements in efficiency especially in non-OECD countries offer scope for reducing nutrient runoff (Conant, Berdanier and Grace, 2013[396]). As outlined in Chapter 3, nitrogen and phosphorus deliveries in agriculture are expected to increase by 2050, particularly in lower- and middle-income regions (e.g. for lower-income regions +41% for nitrogen and +40% for phosphorus).
These figures highlight the need for more effective policies targeting the agriculture sector to apply well-established agricultural practices that bring down nutrient pollution. They include practices related to applying fertilisers and manure, and balancing feed for livestock. Nutrient use efficiency is the product of nutrient uptake efficiency (how much nutrient plants can take up) and nutrient utilisation efficiency (the yield per unit of nutrient that is taken). For nitrogen, nitrogen use efficiency currently varies from around 30% to 53% (Anas et al., 2020[397]). Nitrogen losses are up to 70% of total available nitrogen (Anas et al., 2020[397]). While nitrogen management policies introduced in the past decades by some OECD countries have succeeded in reducing excess nitrogen use by farmers, half of global mineral nitrogen fertiliser used is still lost for crops (Andersen and Bonnis, 2021[285]). For phosphorus, as much as 80% of the mineral is not (directly) taken up by plants (Our Phosphorus Future Network, 2022[290]). This could be reduced by adopting policies for improved agronomic approaches as discussed below, but good soil management is also essential to ensure the phosphorus does not leak to water bodies.
Other measures to reduce nutrient pollution
Countries have introduced a wide range of measures to reduce nutrient use, covering advice to farmers on more efficient application and management, subsidies for the adoption of better practices and, in a few cases, taxes on fertilisers and emissions of nitrogen.31 Common measures include: more informed fertiliser application, integrated nutrient management (i.e. optimal use of indigenous nutrient components), increased use of modified fertiliser (i.e. products that improve use efficiency by reducing various losses of nutrients associated with production system and by enhancing their beneficial use in plants), improved methods of nitrogen and phosphorus application, adoption of low nitrogen feed, coverage of manure when stored and low emission animal housing (Singh et al., 2018[398]; Giannakis et al., 2019[399]).
Piñeiro et al. (2020[317]) review nearly 18 000 papers on whether incentive-based programmes led to the adoption of sustainable practices (including nutrient management) and their effect on environmental, economic and productivity outcomes. The authors find a clear association between market and non-market-based incentives and environmental objectives in a subset of the papers and the review also shows an evidence base for environmental outcomes associated with cross-compliance incentives. The main cross-compliance incentives are Payments for Environmental Services (PES) and other agri-environment payments. These are incentives offered to farmers, or landowners, in exchange for managing their land to provide some type of ecological service, including water quality, forestry, soil erosion and air pollution. For example, in 2023, the European Commission approved the Netherlands’ EUR 1.47 billion in agri-environmental schemes aimed at reducing nitrogen deposition on conservation areas. This included a EUR 975 million scheme providing grants to compensate farmers for voluntarily ceasing agricultural practices, such as dairy cattle breeding, that contribute to nitrogen emissions near Natura 2000 sites. These schemes are part of the Netherlands’ broader strategy to meet EU environmental objectives under the European Green Deal (European Commission, 2023[400]).
Piñeiro et al. (2020[317]) find less consensus and documentation in the case of regulatory measures, where more measurement and reporting are needed. It is also important to note that most of the evidence on environmental outcomes is qualitative through assessment of farmers’ perceptions. Stronger identification strategies, and especially measurement and monitoring through environmental assessments, are needed to uncover the causal effect of the chain of incentives, adoption and outcomes.
In technical terms the abatement measures range from containment and capturing of manure nutrients, to employment of novel cleaner technologies and structural measures in the agricultural sector to reduce nutrient volumes. For the European Union, actions found to be most cost-effective in meeting the NH3 targets were, in order of lowest cost first where each would fully achieve the required emission reduction: adoption of low nitrogen feed, coverage of manure when stored, improving or substituting urea fertiliser application and use of low emission animal housing (Giannakis et al., 2019[399]). Information on adoption of such measures, however, is not available. Also, the cost- effective analysis was based only on direct costs and did not take account of climate or biodiversity co-benefits of the different actions.
Permit trading for nitrogen emissions can help reduce the cost of meeting an emissions reductions target. Hasan et al. (2022[401]) compare the use of a market trading of permits for nitrogen emission against a uniform reduction for all emitters in Denmark (Hasan et al., 2022[401]) and find a 56% lower cost of meeting the target of a 21.5% reduction in nitrogen emissions when using market trading of permits. A higher reduction target might be possible, with its corresponding greater GHG and other pollution reduction benefits, at the same overall cost as is experienced by the present system of uniform regulation.
Taxes on fertilisers
A tax on surplus nitrogen set in the range of the abatement costs (estimated at EUR 0.1 to 0.6 per kg of nitrogen) could generate significant environmental benefits (valued at EUR 0.3-10 per kg of nitrogen) as well as some reductions in GHGs and could be justified on these grounds. To advance the adoption of the measures, Andersen and Bonnis (2021[285]) argue for a tax on surplus nitrogen at a rate of EUR 1-2 per kg of nitrogen as a suitable starting point. No estimate is made, however, of the likely reduction in nitrogen or the decline in GHGs that could be achieved through this tax.
A tax can also be used as a complement to direct control measures, making the latter more effective. Liu et al (2023[402]) construct an integrated multi-scale framework for evaluating alternative nitrogen loss management policies for corn production in the US. This approach combines site- and practice-specific agro-ecosystem processes with a grid-resolving economic model to identify locations that can be prioritised to increase the economic efficiency of the policies. They find that regional measures, albeit effective in reducing local nitrogen loss, can displace corn production to the area where nitrogen fertiliser productivity is low and nutrient loss rate is high, thereby offsetting the overall effectiveness of the nutrient management strategy. This spatial spill-over effect can be suppressed by implementing such measures in tandem with nationwide tax policies. For example, a wetland restoration policy combined with split fertiliser application, along with a nitrogen loss tax could reduce nitrogen loss to the Mississippi River by 30% while only increasing corn prices by less than 2% (Liu et al., 2023[402]). The final tax is determined by the nationally uniform 5 nitrogen fertiliser price (EUR per kg of nitrogen applied), and the product of a nationwide tax rate of USD 1 per kg of nitrogen loss and the nitrogen loss intensity (kg of nitrogen loss per kg of nitrogen application) that varies by location. The nitrogen loss tax is, however, only a proposal.
Subsidies and other regulations on fertiliser
Regulations to limit mineral fertiliser use have been introduced for example in New Zealand, which imposes a cap on the nitrogen content of mineral fertilisers used on pastoral land (New Zealand MFE, 2023[403]). Accordingly, each hectare of grass pasture has a synthetic nitrogen fertiliser limit of 190 kg/ha/yr – anything above this requires a resource consent. On land planted with fodder crops, the limit of 190 kg/ha/yr may be exceeded to meet the needs of some types of plants, but the average across the pastoral system must not exceed 190. In the European Union, the Good Agricultural and Environmental Conditions standards require the establishment of buffer strips along watercourses to ‘protect river courses against pollution and run-off’. A ‘buffer strip’ is an area where fertilisers and plant protection products cannot be applied and must be three metres wide or more. Member States define this bottom width and may add other criteria. In Germany, farmers have not been permitted to apply urea without combining it with urease inhibitor or incorporating it into the soil within four hours (Jones and Deuss, 2024[276]).
At the same time, several developing countries still subsidise fertilisers to increase production but with possible negative impacts on the environment, health and climate. For example, between 2020-2022 the annual average value of fertiliser subsidies provided by the government of India totalled more than USD 22 billion (Jones and Deuss, 2024[276]). India is currently the second largest user, producer and importer of phosphorus-based fertilisers as well as the second largest user and producer, and first importer of nitrogen-based fertilisers (Adenäuer, Laget and Cluff, 2024[299]). For urea (the main form in which nitrogen is applied by Indian farmers), the government sets a Maximum Retail Price (MRP) and pays manufacturers the necessary subsidy to achieve this (Jones and Deuss, 2024[367]). Repurposing such subsidies to realign fertiliser support and soil health programmes can benefit the farmers livelihoods as well as climate and biodiversity (OECD, 2024[405]; Damania et al., 2023[406]; OECD, 2024[407]; Valin, Henderson and Lankoski, 2023[408]).
Bans on manure application on bare fields (after harvest) have been imposed for example in Denmark. From 2019/2020, no liquid manure can be applied to fields after harvest, except to grass or winter oilseed rape until October 1. Solid manure can be applied in autumn to fields before sowing winter crops and to bare soils from October 20 to November 15. In consequence, 92% of liquid manure was spring-applied in 2018, increasing from 55% in 1990 (Sommer and Knudsen, 2021[409]). Sufficient manure storage capacity is a prerequisite for complying with this regulation, and since 1993, the requirement has been a minimum 9-month storage capacity (or 6 months on farms where late application of manure was acceptable) (Sommer and Knudsen, 2021[409]).
Further regulatory restrictions encompass application techniques, as simple and cheap broadcasting spreads the manure unevenly and has given rise to concerns over NH3 volatilisation and nutrient losses in runoff resulting from broadcasting. In Denmark, this technology has been banned since 2002 and replaced by the trail hose application technology. The requirements have since been further tightened, and from 2019/20, slurry must be applied by injection to grassland and bare soils. Alternatively, acidified slurry may be applied with trail hoses to grassland, but acidified slurry must be injected to bare soil. Adding acid may take place in animal houses, during storage or while applying the liquid manure (Sommer and Knudsen, 2021[409]).
Taxes on nutrient feed
Beyond fertiliser-related policies, policies on nutrient management also encompass animal feed taxes, such as the animal feed mineral phosphorus tax that came into effect in Denmark in 2005. With a tax rate of DKK 4 (EUR 0.53) per kg of phosphorus, it targets commercial animal feed phosphate to encourage a switch to phytase (an enzyme that helps to enhance phosphorus uptake) to reduce the need for mineral phosphorus additives. The justification for the tax is the high loss rate of up to 90% associated with animal feed phosphorus supplied to large livestock installations, mainly pork and poultry. The tax is a measure to reduce the saturation of agricultural soil with phosphorus and to curb the leaching of phosphorus to waters (Andersen, 2016[410]).
Mineral phosphate use in animal feed has fallen by about 2 000 t (15%) since the introduction of the tax, and phytase use has increased (OECD, 2019[411]). The tax is thus believed to have improved overall efficiency in the use of animal feed. A new phosphorus regulation introduced in 2017 allows efficient feeding, for example using high doses of phytase, to meet requirements on phosphorus reductions. It has resulted in a stronger incentive to increase phytase and reduce mineral phosphorus use than the tax. To prevent double regulation, the tax was abolished in July 2019 (OECD, 2019[411]).
Taxes on NOx
Taxes on have also been added on NOx emissions. Sweden introduced such a tax in 1992 on nitrogen oxides from large combustion sources (e.g. power plants, industrial plants, waste incinerators) that was accompanied by a refund of the revenue raised (except administration costs) according to the amount of energy generated (OECD, 2013[288]). This ensures that facilities with low NOx emission intensities are net beneficiaries of the scheme. Originally at SEK 40 per kg, from 1 January 2008, the tax rate was increased to SEK 50 (around EUR 5.5) per kg, partly to maintain a strong abatement incentive (OECD, 2013[288]). Actors who continually measure and register emissions using measurement tools that meet specific requirements are permitted to calculate the charge based on their measurements, following extensive guidelines (Swedish Environmental Protection Agency, 2024[412]). NOx declarations are submitted via the e-service E-NOx32.
In the case of air emissions, NH3 from fertilisers and manure reacts with other chemicals to form particulates that are harmful to biodiversity. NH3 is one of the main sources of nitrogen pollution. A major effect of NH3 pollution on biodiversity is the impact of nitrogen deposition on plant species diversity and composition within affected habitats (Guthrie et al., 2018[413]). Common, fast-growing plants well adapted to high nutrient availability can thrive in a nitrogen-rich environment and tend to out-compete species which are more sensitive, smaller or rarer – for example, sensitive lichen and mosses can be damaged by even low concentrations of NH3 (Guthrie et al., 2018[413]). NH3 also impacts species composition through soil acidification (as the nitrogen compounds deposit on land through rainfall), direct toxic damage to leaves, and by altering the susceptibility of plants to frost, drought and pathogens. If changes in species composition and extinctions are large, some ecologically significant habitats may be lost (Guthrie et al., 2018[413]). These biodiversity impacts have not, however, been quantified in relation to changes in NH3.
The contributions of nitrogen pollution to climate change are complex. For example, NH3 emissions ultimately reach the ground (soil, vegetation, or water) and increase emissions of the GHG N2O (Figure 6.4). One estimate of the synthetic nitrogen fertiliser supply chain calculates it was responsible for global emissions of 1.13 Gt of CO2e in 2018, representing 10.6% of agricultural GHG emissions and 2.1% of global GHG emissions (Menegat, Ledo and Tirado, 2022[414]). The actual amounts of emissions, however, vary by type of fertiliser and methods of application.
Overall, reductions in applications of nitrogen and phosphorus as mineral fertiliser are considered to provide benefits to air, soil, and water. They also have significant benefits in terms of less damages to aquatic and terrestrial ecosystems, contribute to biodiversity conservation, pollution control, climate change mitigation, and to human health for the reasons given above. To tackle the environmental and public health challenges of the various nitrogen forms, OECD (2018[415]), in collaboration with the United Nations Economic Commission for Europe (UNECE) Task Force on Reactive Nitrogen, proposes a three-pronged policy framework consisting of (i) analysing nitrogen pathways to better manage environmental risks; (ii) addressing N2O emissions in climate change mitigation policies; and (iii) monitoring and managing residual nitrogen by measuring the effect of the above measures on the national nitrogen budget (Annex Box 6.D.1).
The three-pronged approach to respond to nitrogen pollution includes:
1. Analysing nitrogen pathways to better manage environmental risks
First, there is a need to manage local pollution risks by better understanding the pathways of nitrogen between sources and impacts (the so-called “spatially targeted risk approach”), for example with regards to the resilience of ecosystems to increased nitrogen loading.
The spatially targeted risk approach framework highlights the importance of Impact-Pathway Analysis (IPA)1 when applicable, and proposes to:
Identify the nitrogen sources of relevance to the impact at the local level and delineate the different nitrogen emission zones that converge towards the risk area.
Calculate the marginal abatement costs for reductions in release of the different nitrogen forms. For that, estimate the potential for new or additional emission reductions in each emission zone through cost-effective analysis.
Compare the cost-effectiveness of emission reductions for all sources of risk in the different emission zones to ensure coherence of interventions.
Estimate the marginal ancillary benefits of reducing nitrogen emissions in the different emission zones (i.e. the avoided damages along the pathways of nitrogen to the risk area.
The third step in IPA is yet not widespread in the implementation of nitrogen policies. For example, nitrogen sources other than agriculture – such as industry, wastewater, atmospheric sources and net nitrogen exchanges with marine waters – have not been included in assessing the cost-effectiveness of eutrophication risk management measures in Denmark.
2. Addressing N2O emissions in climate change mitigation policies
Second, account must be taken of the observed increase in global atmospheric concentrations of N2O that have consequences for both climate change and stratospheric ozone (the “global risk approach”).
Feasibility limitations of effective IPA might arise and highlight the need for an integrated policy approach. Nitrogen pathways do not follow the administrative boundaries and can extend beyond national borders. IPA has an important role to play in facilitating the adoption of new international provisions to manage transboundary nitrogen-related pollution.
IPA does not apply to N2O as the risk area (be it the risk of a greenhouse effect or the risk of depletion of the ozone layer) is global. Delineating emission zones, if at all possible, would be of no use for the purposes of cost-effective analysis (or cost-benefit analysis). The more sources are compared, the more likely it is to find the one for which emission reduction is the most cost-effective. In other words, as many sources of N2O as possible should be identified, wherever they are, so as to monitor them and compare the costs – and if possible the ancillary benefits – of reducing their emissions. This requires a global approach.
3. Monitoring and managing residual nitrogen by measuring the effect of the above measures on the national nitrogen budget
Third, there is a need to prevent “excessive” nitrogen entering the environment by developing strategies addressed to the different sources (on the basis of the most cost-effective means) to reduce them (the “precautionary approach”). Precautionary management copes with uncertainties in the nitrogen cascade in cases where conducting an IPA may not be scientifically possible yet.
The main policy responses are summarised in Annex Table 6.D.1.
Source: OECD (2018[415]).
1. Impact-Pathway Analysis (IPA) is an evaluation of the pathways that generate an impact (including through modelling) to estimate the expected benefits of possible emissions changes (OECD, 2018[415]).
2. It can be assumed that ecosystems respond differently to changes in nitrogen load, depending on the type of ecosystem and local conditions (OECD, 2018[415]).
The impacts of phosphorus mining on the triple planetary crisis
In terms of pollution impacts from phosphorus surface and underground mining activities, mining may transfer toxic metals and radioactive elements into the environment and contribute to air pollution. Land disturbance by phosphate mining activities increases the concentrations and loads of radioactive elements such as uranium or thorium, as well as many dissolved and toxic metals in the environment, such as arsenic, chromium, lead, mercury, nickel, vanadium and cadmium (Reta et al., 2018[416]). The beneficiation process (which separates the phosphate minerals from other minerals by removing much of the clay, sand and other impurities) increases the amount of such hazardous elements leaking into soil, water, air and the human food chain. The major airborne emissions from phosphorus mining occur in the form of fine rock dust from drying and grinding operations of phosphate rock (Reta et al., 2018[416]). Fluoride emissions and radon gas emissions constitute further challenges for air quality.
In terms of biodiversity impacts from surface and underground phosphorus mining activities, mines take up large amounts of habitat, with surface mining spanning particularly large areas and in many cases displacing plants and animals (Reta et al., 2018[416]). Surface mining entails the removal of the overburden to expose bare rock surfaces, not only causing destruction of pre-existing vegetation, but also occupying large areas of land for dumping the spoil (Yang et al., 2014[417]). Settling ponds or other waste disposal areas implies further changes to habitats and can impact the water quality of the surrounding water, hence also affecting aquatic ecosystems. In the case of dam failure, the tailings released into the environment can create widespread ecological damage, especially in aquatic environments (Islam and Murakami, 2021[418]).
Surface and underground phosphorus mining can also affect local carbon sinks and contribute to climate change through the energy use required for mining activities. New mining sites often involve clearing vegetation, which disturbs local carbon sinks. This can lead to carbon release from soils and loss of carbon sequestration capacity. The extraction, crushing, and transportation of rock requires significant energy, which contributes to GHG emissions. Although mining contributes a relatively small fraction of total emissions in the phosphorus supply chain from phosphate rock mining to phosphorus use in crop production, with Gong et al. (2022[419]) estimating around 8% in China for phosphorus mining versus 43% for phosphorus fertiliser manufacturing and 48% for crop production, its impact is still significant in absolute terms given the large emissions from crop production outlined Chapter 2.
At the moment, although most phosphate rock resources occur as sedimentary marine phosphate rocks, found especially in northern Africa, China, the Middle East, and the United States (Smit et al., 2009[420]), seabed mining is not being used. However, several companies have applied for licenses to exploit marine phosphate reserves in countries such as Namibia, New Zealand, Mexico and South Africa. Namibia had given clearance for seabed mining in 2016, but this was quickly suspended over environmental concerns, in particular potential impacts on fisheries that challenged the validity of the mining license. Potential impacts of seabed mining are explored in Annex Box 6.D.2.
Seabed mining has been proposed as an alternative to terrestrial mining but traditional types of seabed phosphorus mining such as dredging (using suction dredges), proposed for example by NMP (NMP, 2012[421]), could have severe and irreversible impacts on marine ecosystems and fishery resources. Mining proposals in South Africa have been found to overlap with marine protected areas and areas identified as critically endangered ecosystems (Safeguard our Seabed Coalition, 2016[422]). Potential impacts outlined in a report on sea-bed phosphorus include the direct destruction of seabed ecosystems such as corals, the direct destruction of spawning, breeding and feeding habitats for fish species, the burial and smothering of marine organisms both in the mining block and surrounding areas (Currie, 2013[423]). Assuming bulk mining plans would be similar to those of the NMP project in Namibia (NMP, 2012[421]), the dredging vessel would focus on a target mining area for several years and be at sea for the majority of that time, creating a semi-permanent source of acoustic disturbance. As large volumes of dredged sediment have to be transferred ashore frequently (every 37 hours in the Namibian example), a significant increase in shipping traffic would develop between the mining site and the selected port, creating persistent disturbance along the transfer route for species sensitive to sound.
Climate impacts from dredging could also arise for example due to reduced light penetration and damage to carbon-sequestering habitats. Sediment clouds potentially reduce the photosynthesis rates of marine flora that convert CO2 into OC while releasing oxygen, which can weaken the ocean’s role as a carbon sink (Currie, 2013[423]). If carbon-sequestering habitats degrade or die, stored carbon can be re-released as CO2 or CH4.
In terms of potential pollution impacts, offshore mining is particularly vulnerable to weather conditions that pose added risks of material spills and leakage into the environment (Reta et al., 2018[416]).
As an alternative to traditional sea-bed mining, bio-based mining from marine ecosystems through removing phosphorus in waterbodies could help reduce eutrophication and therefore eutrophication-related biodiversity losses. Bio-based approaches for marine phosphorus mining through microorganisms such as phosphate accumulating organisms (PAOs) could be exploited for removal and recovery purposes from marine ecosystems (Cakmak et al., 2022[424]). However, due to the experimental stage other environmental impacts are difficult to predict.
The impacts of agricultural applications of phosphorus on the triple planetary crisis
Phosphorus fertilisation is often used even when the soil has high levels of phosphorus as only a small fraction of it is available for uptake by cultivated plants – some authors suggest it may be less than one percent (Bünemann, 2015[425]). The main reason for phosphorus fertilisation is that farmers aim to avoid a deficit in phosphorus uptake as that would limit plant growth and thus diminish harvest. Phosphorus fertilisation leads to a variety of impacts.
In the case of phosphorus pollution through the soil from agricultural applications of phosphorus-based fertilisers, these can disrupt biodiversity, although the overall effects of such applications are species and location specific and depend for example on the underlying phosphorus limitations of the ecosystems, as well as the timeframe examined. Almeida et al. (2019[426]) find that the application of superphosphate, a phosphorus-based chemical fertiliser, decreased fungal biomass when combined with nitrogen. However, Korevaar and Geerts (2015[427]) found that species-richness initially declined but started to recover after about 25 years of fertilisation. Research in Australian grasslands indicated that superphosphate application affected invertebrate populations, with some groups declining in abundance (Oliver et al., 2005[428]).
Phosphorus applications through manure have also been linked to high microbial diversity and activity (White and Reddy, 2000[429]; Li et al., 2021[430]),. Given that manure contains both phosphorus and nitrogen, the effects are difficult to detangle specifically to phosphorus. A global meta-analysis by Nessel et al. (2021[431]) that synthesises data from 1 679 cases across 207 studies reveals that nitrogen and phosphorus additions decreased invertebrate abundance in both terrestrial and aquatic ecosystems, although the meta-analysis combines manure and chemical fertilisation. The study also notes that combined nitrogen and phosphorus enrichments had stronger negative impacts, particularly in tropical regions (Nessel et al., 2021[431]).
Research on the biodiversity effects of direct phosphate rock applications is still emerging. Being natural compounds, finely ground phosphate rock can be used in organic agriculture and has been used as an alternative to soluble chemical fertilisers – although most of the phosphate rock is treated to make the phosphorus more soluble. In agricultural settings phosphate rock is often applied in combination with a microbial phosphate solubilising inoculum that improves plant phosphorus acquisition and may affect the native bacterial and fungal communities. Trabelsi et al. (2017[432]) suggest that fertilisation with phosphate rock might foster bacteria with fast-growing rates, such as Bacteroidetes and Gammaproteobacteria, at high phosphorus levels.
In the case of phosphorus pollution in waters, primarily through agricultural runoff, sewage discharge, and industrial effluents (Jwaideh, Sutanudjaja and Dalin, 2022[433]), phosphorus pollution is the paramount nutrient for freshwater eutrophication and the main limiting factor for algal growth in most freshwater areas (Zhou et al., 2024[434]). When excess phosphorus enters freshwaters it hence rapidly stimulates algal blooms and eutrophication, overall leading to more than double the freshwater fish species loss than nitrogen pollution (Zhou et al., 2024[434]). Phosphorus-driven algal blooms also affect biodiversity by reducing for example plankton diversity (Amorim and Moura, 2021[435]). Globally, phosphorus losses from land to fresh waters have doubled in the last century and continue to increase (Our Phosphorus Future Network, 2022[290]).
Phosphorus pollution also has significant effects on climate change, in particular through the aforementioned eutrophication. Phosphorus-driven eutrophication and the decomposition of dead algae in these oxygen-deprived environments favours the production of the GHG CH₄. Eutrophic systems tend to emit more CH4 than oligotrophic systems, due to accumulation of organic matter on the sediment (Nijman et al., 2022[436]). The accumulated sediment also continues emitting CH4 even in cases where external nutrient inputs have been controlled. Empirical models demonstrate that an increase in eutrophication driven, in part, by increasing phosphorus loading could increase CH4 emissions globally by the equivalent of up to 33 percent of annual CO2 emissions from burning fossil fuels (Beaulieu, DelSontro and Downing, 2019[437]). Phosphorus has also been found to affect carbon cycling in mangrove sediment, resulting in changes in soil GHG emissions and carbon stocks of mangrove ecosystems (Qiu et al., 2024[438]). Similarly, in peatlands, Schillereff et al. (2021[439]) suggest that long-term elevated phosphorus deposition and accumulation strongly correlate with increased organic matter decomposition and lower carbon sequestration. Continued phosphorus pollution thus places the future carbon sink of peatlands at risk.
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← 1. More precisely pollution prevention and control; the shorter phrasing is used throughout in this chapter and is meant to include prevention.
← 2. The selection has been made in consultation with governments and topical experts but necessarily remains somewhat ad-hoc. Alternative deep-dives on cross-cutting issues such as water and chemicals are beyond the scope of this report but are integrated in the analysis of the other chapters.
← 3. This is equivalent to around 3.7 Gt of CO2e (IPCC, 2021[205]).
← 4. The same is true for fossil fuels, which are also heavily internationally traded.
← 5. See Chapter 3 of (OECD, 2024[2]) for a comprehensive and synthetic review of impacts of solar power, wind power and power lines on biodiversity.
← 6. Australia is also conducting further research about potential impacts to specific species and mitigation measures for effective and efficient regulation of both offshore and onshore windfarms.
← 7. This is in contrast with the time it currently takes. For instance, in some Member States of the EU, it can take for large renewable energy projects up to 9 years to obtain environmental permits (European Commission, 2022[441]).
← 8. Similarly, it can be challenging to consider the full upstream environmental impacts at the level of a project-specific EIA, depending on the scope of impacts that needs to be considered; specifically, assessment of indirect effects, e.g. from the use of energy in upstream production processes, can be complicated.
← 9. Unmanaged land refers to land cover types that are suitable for exploitation but not in active use for agriculture or forestry, nor used for human settlements.
← 10. These 31 parks include in total 5.6 million hectares of forests.
← 11. Articles 6(3) and 6(4) of Council Directive 92/43/EEC.
← 12. See definition in Annex 6.B.
← 13. Adopted in Decision 15/5.
← 14. Bolivia, Brazil, Colombia, Ecuador, Guyana, Peru, Suriname and Venezuela.
← 15. Aerosols refer to solid particles or liquid droplets dispersed in a gas (e.g. in air).
← 16. Deadwood refers to non-living wood material in forests, which can increase due to factors like storms and bark-beetle outbreaks (Milad et al., 2011[442]).
← 17. There are also interaction effects with the protected area management policies, as outlined in Section 6.3.
← 18. Source appointment refers to the practice of deriving information about pollution sources and the amount they contribute to ambient air pollution levels.
← 19. An airshed is a geographical area that shares a common flow of air due to topography or weather conditions and could extend across borders (World Bank, 2025[224]).
← 20. Such solutions require economic incentives so that the burden of the abatement costs is shared by everybody who benefits from the emissions reductions. This sharing can be accomplished by establishing regional funds or even by introducing a system of tradable emissions rights.
← 21. The imposition of these taxes is matched by a reduction in personal income taxes, employment taxes but a harmonisation of VAT at a standard rate of 22% and harmonised reduced rate of 12% for selected goods and services. In addition, income support to the lowest two quintiles is raised as are excise duties on tobacco products. These changes are assumed to be introduced gradually over the period 2021 to 2025.
← 22. Estimates of damages in monetary terms are not available on a systematic basis. To get these, more information on impacts is required at a highly granular level.
← 23. Especially in Latin America and Africa a large share of total nutrient inputs also comes from manure applied to soils (OECD, 2021[300]).
← 24. As outlined in Chapter 2, this is calculated as the difference between nutrient inputs (e.g. manure or synthetic fertiliser) and the amount of nutrients absorbed by plants.
← 25. For example, corn has a higher nitrogen demand during its growth stages while legume crops such as soybeans can naturally fix nitrogen from the atmosphere making it less dependent on external nitrogen inputs (Adenäuer, Laget and Cluff, 2024[299]). By contrast, legume crops have more demand for phosphorus for optimal nitrogen fixation compared to plants like cereals (Mitran et al., 2018[440]).
← 26. Phosphate rock includes both i) the phosphate mineral bearing rock (mostly apatite) that can be used directly as fertiliser and ii) a beneficiated concentrate of the phosphate mineral (Smit et al., 2009[420]). In the first case, phosphate rock is finely ground and applied directly as a fertiliser. In the latter case, beneficiation of the rock removes much of the clay, sand and other impurities, raising the phosphorus concentration by separating the phosphate minerals from other minerals (Smit et al., 2009[420]; FAO, 2004[444]).
← 27. However, the life-cycle carbon footprint of blades made with this new material has yet to be understood (Volard, 2023[443]).
← 28. This is based on compliance with technical guidelines designed to ensure compliance with the “do no harm” principle under the European Commission's Recovery and Resilience Facility Regulation (2021/C 58/01) (OFB, 2023[104]).
← 29. According to the ISO, environmental labelling and information schemes are classified into three main types: (i) Type I - ecolabels - multi criteria and third party verified (ISO 14024), (ii) Type II - self-declared environmental claims (ISO 14021), and (iii) Type 3 - environmental declarations (ISO 14025) (ISO, 2019[445]).
← 30. The EC expects the member states to use penalties that are “effective, proportionate and dissuasive” (EU Directive 2016/2284).
← 31. In addition, a few of studies have looked at the impacts of taxes that have yet to be implemented.
← 32. E-service available here: https://noxdeklaration.naturvardsverket.se/.