11.1 Introduction
This chapter analyzes the potential for transformative change for biodiversity conservation in the governance of protected areas and other conserved areas (which incorporates other effective area-based conservation measures or OECMs). This is achieved by analyzing efforts to achieve Aichi Target 11 under the UN Convention on Biological Diversity (CBD) strategic plan to 2020, and discussing the need for a new outcome-based approach under the CBD’s Post-2020 Global Biodiversity Framework (GBF), which is under discussion at the time of writing but expected to be adopted during 2022. Under Aichi Target 11,Footnote 1 governments collectively agreed to designate 17 percent of terrestrial and inland waters and 10 percent of coastal and marine areas as protected areas and OECMs that are effectively and equitably managed, ecologically representative, well connected and integrated into the wider landscape and seascape. It is widely considered to be the Aichi Target that governments have made most progress on delivering (UNEP-WCMC et al., 2018).
The issue is discussed here through the conceptual lens of transformative governance, which is defined in Chapter 1 of this book and is understood to address the underlying causes of biodiversity loss through governance mechanisms that are integrative, inclusive, transdisciplinary, anticipatory and adaptive (Reference Chaffin, Garmestani and GundersonChaffin et al., 2016; Reference GustonGuston, 2014). How to follow up Aichi Target 11 with a new area-based target has formed a key part of the CBD discussions in advance of the adoption of a new Post-2020 GBF. A range of perspectives emerged during these discussions. These can be summarized as: full implementation of Aichi Target 11; more ambitious area-based targets (such as a 30 percent area-based target or “half-earth” approach); “new conservation,” which intends to integrate conservation with neoliberal economic approaches; and a “whole earth” approach, which aims to find a balance between human and nonhuman needs (Reference Bhola, Klimmek and KingstonBhola et al., 2020, see also Chapter 12). This chapter does not fit neatly within any of these categories but contributes to the discussion by recognizing the valuable role played by protected and conserved areas, and the need for their continued prioritization at the policy level, and provides recommendations for their implementation under the Post-2020 GBF.
We understand transformative change in the context of protected and conserved area governance as referring to their contribution to the effective conservation of existing biodiversity, as well as its restoration, where possible. This chapter thus begins with an introduction to protected and conserved area governance, before examining the extent to which Aichi Target 11 has stimulated action to achieve effective biodiversity conservation outcomes. Outcomes are understood as being the consequences of project interventions and provide reliable indicators of long-term conservation impacts, either success or failure (Reference Howe and Milner-GullandHowe and Milner-Gulland, 2012; Reference Kapos, Balmford and AvelingKapos et al., 2009). Biodiversity outcomes as used in this chapter refer to the status of biodiversity elements such as species and ecosystems. Equity outcomes refer to the fair sharing of power, responsibility and benefits in natural resource management, as well as strengthening governance arrangements including legal entitlements and making decisions more transparent, inclusive and equitable.
We therefore first review the academic literature on protected and conserved areas through the lens of transformative governance (in Section 11.2), including recent literature analyzing the strengths and weaknesses of policy efforts to reach Aichi Target 11 (in Section 11.3). In Section 11.4, we use three case studies through which to explore the transformative change needed, before drawing conclusions related to the potential for a new outcome-based approach to protected and conserved area governance. The three case studies were selected to encompass different continents (Africa, Asia and Latin America), different ecosystems (terrestrial and marine) and different governance approaches.
11.2 Governance of Protected and Conserved Areas
Protected areas have been viewed as a mainstay of actions to conserve biodiversity and have long been at the fore of conservation and research efforts (Reference Andam, Ferraro, Pfaff, Sanchez-Azofeifa and RobalinoAndam et al., 2008; Reference Rands, Adams and BennunRands et al., 2010). It has been shown that well-managed protected areas are effective in conserving biodiversity and can reduce habitat loss and maintain species populations (Reference Bruner, Gullison, Rice and FonsecaBruner et al., 2001; Reference Leverington, Lemos, Pavese, Lisle and HockingsLeverington et al., 2010; Reference Watson, Dudley, Segan and HockingsWatson et al., 2014), as well as provide a range of societal benefits (Reference Stolton and DudleyStolton and Dudley, 2010). The widely accepted International Union for the Conservation of Nature (IUCN) definition of protected areas is “a clearly defined geographical space, recognized, dedicated and managed, through legal or other effective means, to achieve the long-term conservation of nature with associated ecosystem services and cultural values” (Reference DudleyDudley, 2008). This places nature conservation objectives firmly at the center. They have a prominent position within global environmental governance fora, such as the CBD, which has a dedicated program of work on protected areas, and have long been promoted as an important conservation tool by IUCN and its member organizations through the World Commission on Protected Areas (WCPA). Other international environmental conventions have also placed a high importance on designating and managing protected areas, such as the Ramsar Convention on Wetlands of International Importance and the World Heritage Convention concerning the Protection of the World’s Cultural and Natural Heritage (agreed in 1971 and 1972, respectively).
The concept of other effective area-based conservation measures (or OECMs) was introduced for the first time in the international policy arena in 2010, as an additional way by which national governments could meet Aichi Target 11. At the time, there was no accepted definition, and it was not until 2018 that the CBD adopted a decision that defined OECMs as “a geographically defined area other than a Protected Area, which is 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” (CBD, 2018). OECMs are expected to involve a wider array of stakeholders in governance arrangements, particularly IPLCs, and spiritual and religious groups (Reference Laffoley, Dudley and JonasLaffoley et al., 2017), and provide an opportunity to engage rights-holders and promote equitable and diverse partnerships in conservation efforts (IUCN-WCPA, 2019). Their governance arrangements are therefore expected to be more complex than those of traditional protected areas and are likely to require strengthening or the gaining of official recognition of informal arrangements (Reference Dudley, Jonas and NelsonDudley et al., 2018). OECMs are now commonly referred to as “conserved areas,” which is the term we adopt throughout the rest of this chapter.
Protected area governance is well-documented in the academic literature and has been influenced by integrative governance (IG) concepts (Reference Visseren-HamakersVisseren-Hamakers, 2015; Reference Visseren-Hamakers2018). This includes polycentric governance, under which there are several centers of decision-making (Reference Carlisle and GrubyCarlisle and Gruby, 2019), and multilevel governance, in which decision-making takes place at different scales in support of common goals (Reference Bennett and SatterfieldBennett and Satterfield, 2018). Terrestrial protected areas, for example, have been heavily influenced by multilevel collaborative governance with growing interest in scaling-up to the landscape level with an increased focus on transboundary and connectivity issues (Reference LockwoodLockwood, 2010). Forests, in particular, have been valued for assets other than biodiversity, notably for timber, and more recently for carbon sequestration and, as a consequence, governance arrangements and stakeholder engagement for forest areas are often complex (Reference Nagendra and OstromNagendra and Ostrom, 2012; Reference Reinecke, Pistorius and PregernigReinecke et al., 2014). Forest governance also reflects IG approaches and has been heavily influenced by the concept of networked governance, which recognizes diverse configurations of stakeholders interacting at multiple levels, with a diffusion of authority (Reference Jedd and BixlerJedd and Bixler, 2015). Marine protected areas (MPAs) have similarly been highly influenced by shared governance approaches between governments and local communities (Reference Bown, Gray and SteadBown et al., 2013). Locally managed marine areas (LMMAs), in which nearshore waters are actively managed by communities, have been widely adopted across the tropical western Pacific, for example, as a way of achieving biodiversity conservation and fisheries management objectives simultaneously (Reference Jupiter, Cohen, Weeks, Tawake and GovenJupiter et al., 2014).
Protected area governance has also been influenced by particular concepts under the transformative governance agenda. In addition to adopting increasingly integrative approaches, protected and conserved area governance has also become more inclusive and adaptive. Recent efforts to scale-up both terrestrial and marine protected areas to the landscape/seascape scale, for example, including through the newer emphasis now being given to conserved areas, have moved toward more inclusive forms of governance, including by the increased and more formal involvement of IPLCs in governance mechanisms (Reference Premauer and BerkesPremauer and Berkes, 2015). Studies indicate that adaptive management approaches that have integrated local communities in co-governance arrangements have been the most successful for terrestrial protected areas (Reference Dawson, Martin and DanielsenDawson et al., 2018; Reference Premauer and BerkesPremauer and Berkes, 2015). Similarly, the most successful approaches to MPA governance have found a balance between top-down and bottom-up approaches with a diversity of institutions involved (Reference McCay and JonesMcCay and Jones, 2011) that take an adaptive approach with room for experimentation in management strategies (Reference Bown, Gray and SteadBown et al., 2013). A global analysis of both marine and terrestrial protected areas by Reference Oldekop, Holmes, Harris and EvansOldekop et al. (2015) suggested that conservation benefits for biodiversity were highest when protected areas also delivered positive socioeconomic outcomes for local people, and that a co-management approach between local communities and conservation organizations delivered the greatest benefits to both local people and biodiversity.
Although the study by Reference Oldekop, Holmes, Harris and EvansOldekop et al. (2015) supports the joint achievement of biodiversity and socioeconomic outcomes, it is important to recognize that in practice protected area governance is impacted by debates concerning the ownership of, access to and governance of natural resources (Reference OstromOstrom, 1990). Tensions often exist over competing objectives to be achieved (Reference Anthony, Szabo and Lopez-PujolAnthony and Szabo, 2011) and have increased in complexity due to expectations that protected areas will achieve a wide range of objectives (Reference Watson, Dudley, Segan and HockingsWatson et al., 2014). In high-income countries this tends to reflect an increasing move toward achieving multiuse areas for a wide range of social and economic goals and ensuring the continuation of a range of ecosystem services (Reference Hammer, Mose, Scheurer, Siegrist and WeixlbaumerHammer et al., 2012). In low-income countries, such tensions are often more concerned with how to enable local, and often marginalized, communities to achieve social and economic justice and livelihood goals, alongside nature conservation goals (Reference McShane, Hirsch and TrungMcShane et al., 2011; Reference Shahabuddin and RaoShahabuddin and Rao, 2010). This situation is likely to increase in complexity with the more recent inclusion of conserved areas, as by their very definition biodiversity conservation is not necessarily the main objective but is rather one of a number of objectives or an outcome resulting from management that is primarily for another purpose other than conservation. The management of trade-offs is, therefore, potentially even more complex in the case of conserved areas and there are still no commonly agreed minimum criteria of accepted biodiversity outcomes for potential OECMs that can be managed for purposes other than biodiversity conservation. Discussion is ongoing, for example, regarding the appropriate balance between biodiversity conservation and fisheries management objectives for conserved areas (Reference Diz, Johnson and RiddellDiz et al., 2018).
Resolving such issues and achieving conservation and socioeconomic outcomes is related to both the quality of protected and conserved area governance as well as management effectiveness, and there is often blurring and confusion between these two issues (Reference LockwoodLockwood, 2010). While protected area management is concerned with the means and actions to achieve given objectives, protected area governance is concerned with decisions on what the objectives are, how decisions are taken and who has power, authority and responsibility and should be held accountable (Reference Borrini-Feyerabend, Hill, Worboys, Lockwood, Kothari, Feart and PulsfordBorrini-Feyerabend and Hill, 2015). In the last few decades there has been a shift from mostly state-driven, top-down governance approaches to a range of approaches to protected area governance, summarized in a typology adopted by the IUCN (Reference LockwoodLockwood, 2010). It encompasses four main types: state governance; shared governance, which is more collaborative in nature between state and nonstate actors; private governance (i.e. governance by nonstate actors) and governance by IPLCs (Reference Borrini-feyerabend, Dudley and JaegerBorrini-Feyerabend et al., 2013). This IUCN categorization by governance-type is distinct from but complementary to the more widely cited IUCN protected area categories system, which classifies protected areas according to their management objectives (Reference DudleyDudley, 2008).
Increasing attention has been given to assessing the management effectiveness of protected areas (for example, Reference Bruner, Gullison, Rice and FonsecaBruner et al., 2001 and Reference Leverington, Lemos, Pavese, Lisle and HockingsLeverington et al., 2010), but this has not been matched by efforts to examine whether protected area management and governance are leading to positive outcomes for biodiversity. This is despite the recent boom in satellite remote sensing tools that can provide relatively cheap and rapid assessments of terrestrial biodiversity (Reference Luque, Pettorelli, Vihervaara and WegmannLuque et al., 2018), including for tropical forest ecosystems (Reference Mulatu, Mora, Kooistra and HeroldMulatu et al., 2017). In particular, little attention has been paid to the role of protected area governance in achieving effective conservation (and, where relevant, restoration) of biodiversity, and the role of conserved areas in this regard needs further examination. It is particularly important to examine biodiversity outcomes in the context of increasing pressures from a range of underlying drivers linked to unsustainable, global patterns of consumption and trade (Reference Folke, Österblom and JouffrayFolke et al., 2019). There has been limited examination, for example, of how effective the approach set by the overarching international biodiversity governance agenda has been in stimulating action toward achieving biodiversity outcomes and addressing underlying drivers of loss. In the next section, we thus turn to considering the overall strengths and weaknesses of protected area governance approaches agreed at the international policy level, by examining the implementation of the Aichi Targets (2011–2020) through a transformative governance lens and further analyzing the new and growing role of conserved areas.
11.3 Strengths and Weaknesses of International Policy Approaches to Protected and Conserved Area Governance
11.3.1 CBD Aichi Target 11
Although the prominence of protected areas within the CBD and other international environmental conventions has ensured that high-level goals exist to stimulate government action, there has been a clear gap between such aspirational targets and the realization of actual outcomes on the ground. Lack of national level implementation is commonly cited as a problem and has been referred to as “perhaps the most significant factor in the failure of international biodiversity law” (Reference Jóhannsdóttir, Cresswell and BridgewaterJóhannsdóttir et al., 2010:146), including the implementation of provisions under the CBD that are soft and open-ended in nature. There has been a lack of implementation of national biodiversity strategies and action plans (NBSAPs) in many countries, with limited progress made toward achieving many of the Aichi Biodiversity Targets set in 2010 (Reference Buchanan, Butchart, Chandler and GregoryBuchanan et al., 2020; Reference Maxwell, Cazalis and DudleyMaxwell et al., 2020; also see Chapter 3). CBD Aichi Target 11, however, provided a quantified target for the percentage of terrestrial (17 percent) and marine (10 percent) areas to be conserved and has been more successful than the other Aichi targets in stimulating government action. The coverage of terrestrial protected areas, for example, increased from 10.9 percent in 2011Footnote 2 to 14.9 percent in 2018 (UNEP-WCMC et al., 2018). The growth of MPAs has been particularly dramatic in recent years, with a fifteen-fold increase since 1993 when the CBD came into force, to a total of 16.8 percent of national waters having been designated by 2018 (UNEP-WCMC et al., 2018). According to the World Database on Protected Areas, the global coverage of terrestrial and marine protected areas as of January 2021 is 16 percent and 8 percent respectively.Footnote 3 Although Aichi Target 11 brought the designation of protected areas to the fore, much less attention has been paid to implementing the second half of the target, which is concerned with ensuring that such protected areas are effectively and equitably managed, and ecologically representative (Reference Maxwell, Cazalis and DudleyMaxwell et al., 2020; Reference Watson, Darling and VenterWatson et al., 2016a), as we next discuss.
Effective Management
Efforts have been made to assess the management effectiveness of protected areas in the academic literature (for example analyses by Reference Bruner, Gullison, Rice and FonsecaBruner et al., 2001 and Reference Leverington, Lemos, Pavese, Lisle and HockingsLeverington et al., 2010), by IUCN and through publicly funded partnership initiatives. One such example is the European Union (EU) funded BIOPAMA program, managed by IUCN, which aims to support data management and analysis in Africa, Caribbean and Pacific (ACP) countries to facilitate better decision-making for protected areas.Footnote 4 A range of new tools have been developed, such as the Protected Areas database on Protected Area Management Effectiveness (PAME),Footnote 5 the Integrated Management Effectiveness Tool (IMET), the Management Effectiveness Tracking Tool (METT) (Reference Hockings, Stolton, Dudley and DeguignetHockings et al., 2018), and the IUCN green list of protected and conserved areas (IUCN and WCPA, 2017), a certification program and standard for the effective management and fair governance of protected and conserved areas. One weakness of some of these approaches has been the focus on management structures and procedures, with less attention given to assessing whether protected areas are effectively conserving and restoring biodiversity, which cannot necessarily be inferred from PAME assessments (Reference Coad, Leverington and KnightsCoad et al., 2015; Reference Maxwell, Cazalis and DudleyMaxwell et al., 2020).
Furthermore, efforts made to assess management effectiveness have not necessarily been matched by efforts by national governments to support work on the ground to protect biodiversity within sites (Reference Geldmann, Deguignet and BalmfordGeldmann et al., 2021). Studies that intersect these Earth observation data with networks of protected areas (so as to assess their effectiveness) show that many protected areas slow, but fail to halt, human pressures and biodiversity loss within their borders (Reference Verma, Jones and RheindtVerma et al., 2019). Protected area management is unable to address many of the underlying drivers of biodiversity loss linked to unsustainable production and consumption patterns and global trade, which is placing increasing pressure on biodiversity around the world (Reference Folke, Österblom and JouffrayFolke et al., 2019). At least a third of protected areas are reported to be facing intense human pressure (Reference Jones, Venter and FullerJones et al., 2018b), and around 40 percent of protected areas globally are estimated to face major deficiencies in management (Reference Leverington, Lemos, Pavese, Lisle and HockingsLeverington et al., 2010), making it difficult to resist and adapt to such pressures. The study by Reference Jones, Venter and FullerJones and colleagues (2018b) also found that human pressure had increased in 55 percent of protected areas between 1993 and 2009. Deficiencies in management effectiveness are partly due to a lack of allocation of the necessary finance and resources, including of well-trained staff responsible for site management (Reference Geldmann, Coad and BarnesGeldmann et al., 2018). The general lack of finance for protected areas is well-documented (for example, Reference Mccarthy, Donald and ScharlemannMcCarthy et al., 2012; Reference Waldron, Mooers and MillerWaldron et al., 2013), with the Global South facing the greatest shortfalls in budgets and staffing (Reference Coad, Watson and GeldmannCoad et al., 2019).
Equitable Management
Governments have also given insufficient attention to achieving the equitable management component of Aichi Target 11 (Reference Hagerman and PelaiHagerman and Pelai, 2016). This is despite an increase in co-governance arrangements with an increasing trend of participation in biodiversity governance from nonstate actors and the emergence of multistakeholder partnerships, such as the Congo Basin Forest Partnership (CBFP) (Reference Pattberg, Kristensen and WiderbergPattberg et al., 2017), which could represent a shift toward more inclusive forms of governance. The EU, for example, has been supporting public–private partnerships (PPPs) as a key tool for protected area governance, through its development aid programs (European Commission, 2014). In the case of MPAs, a key challenge is how to ensure that local communities remain meaningfully involved in governance with a greater focus on increasing the coverage of areas under designation and on scaling-up (Reference Gruby and BasurtoGruby and Basurto, 2014; Reference McCay and JonesMcCay and Jones, 2011). Progress has also been limited toward achieving CBD Aichi Target 18, which was dedicated to the full and effective participation of IPLCs, but which did not provide a measurable target (Reference Fajardo, Beauchesne and Carbajal-lópezFajardo et al., 2021).
The importance of equity considerations is not to be underestimated, particularly as the scale and global importance of involvement by Indigenous Peoples in protected area governance is now becoming apparent. A recent study showed such communities impact the governance of approximately 40 percent of sites worldwide (Reference Garnett, Burgess and FaGarnett et al., 2018), and the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES) Global Assessment recognized their involvement as critical to achieving transformative governance (Reference Díaz, Settele and BrondízioIPBES, 2019). The valuable contribution made by Indigenous People to conservation is particularly significant in the case of forests (Reference Fa, Watson and LeiperFa et al., 2020), which harbor around 75 percent of global terrestrial biodiversity (FAO, 2016). The need to halt tropical deforestation is recognized as being one of the most pressing and urgent global environmental challenges (Reference Franklin and PindyckFranklin and Pindyck, 2018), with CBD Aichi Target 5 aiming to halve, and where feasible bring to zero, the loss of natural habitats including forests. Primary tropical forest is of disproportionate value for biodiversity and ecosystem services (Reference Morales-Hidalgo, Oswalt and SomanathanMorales-Hidalgo et al., 2015), but forest loss continues unabated in low-income tropical countries (Reference Keenan, Reams and AchardKeenan et al., 2015). Although the conservation outcomes achieved by community conserved areas (CCAs) vary widely depending on the context (Reference Rao, Nagendra, Shahabuddin, Carrasco, Joppa, Baillie and RobinsonRao et al., 2016), there is increasing evidence that securing land rights for Indigenous Peoples over forest land is an effective and important conservation management strategy (Reference Fa, Watson and LeiperFa et al., 2020; Reference Oliveira, Asner and KnappOliveira et al., 2007; Reference Watson, Evans and VenterWatson et al., 2018).
Furthermore, governance by Indigenous Peoples is having a positive conservation impact outside of protected areas, with studies showing that areas under their governance harbor as much biodiversity as protected areas (Reference Schuster, Germain, Bennett, Reo and ArceseSchuster et al., 2019; Reference Sheil, Boissiere and BeaudoinSheil et al., 2015). The high focus on site designation contrasts strongly with the very limited attention given to delivering Aichi Targets concerned with conserving biodiversity outside of protected areas, including lands governed by Indigenous People (Reference Hagerman and PelaiHagerman and Pelai, 2016). Although many of the Aichi Targets are concerned with areas outside of protected areas, very little attention has been given to the 83 percent and 90 percent of undesignated terrestrial and marine areas, respectively. This is despite the significant biodiversity and ecosystem services they harbor and provide, which if lost would be an unmitigated disaster for both nature and people (Reference Jones, Klein and HalpernJones et al., 2018a; Reference Maron, Simmonds and WatsonMaron et al., 2018).
Ecological Representation
Limited attention has also been given to ensuring the ecological representativeness component of Aichi Target 11. There is evidence that protected area designation has not necessarily targeted areas with high levels of threatened species but has instead been established in areas that minimize conflicts with agriculturally suitable land (Reference Venter, Magrach and OutramVenter et al., 2017). The same is true of MPAs, which have failed to include all ecoregions, with area selection being influenced by socioeconomic factors (Reference Jantke, Jones and AllanJantke et al., 2018). Existing tools such as the global standard for identifying key biodiversity areas (KBAs) (IUCN, 2016), for example, could help governments identify the most valuable areas for biodiversity, but no such standard has been formally adopted by governments globally to help guide designation of protected and conserved areas (Reference Visconti, Stuart and BrooksVisconti et al., 2019). Although the CBD Aichi Targets are global, they are often interpreted at a national level, with an assumption that all governments will try to achieve the 17 percent and 10 percent targets within their countries. Within the EU, an approach was adopted under the Birds and Habitats Directives that enabled a network of protected areas (known as Natura 2000) to be selected based on ecological representativeness at the continental level (Reference Maiorano, Amori and MontemaggioriMaiorano et al., 2015). No equivalent framework exists at the global level by which to incentivize countries with disproportionately high levels of biodiversity, such as megadiverse countries (Reference Yang, Cao and HouYang et al., 2020), to designate a larger percentage area of their territories. This lack of incentive is compounded by the overlap of key areas for biodiversity with areas facing high levels of poverty (Reference Fisher and ChristopherFisher and Christopher, 2007), with low-income countries facing the greatest relative shortfalls in site designation that would ensure ecological representativeness (Reference Butchart, Clarke and SmithButchart et al., 2015).
11.3.2 Influence of the International Climate and Forest Governance Agendas
Protected area governance has been strongly influenced during the past decade by the climate regime, not least due to the introduction of the REDD+ initiative under the UN Framework Convention on Climate Change (UNFCCC), which aims to reduce carbon emissions from deforestation and forest degradation and encourage the conservation and enhancement of forest carbon stocks (UNFCCC, 2007). Although REDD+ was intended to target areas where deforestation is highest, in many cases this has included protected areas with large expanses of forest habitat (Reference Scharlemann, Kapos and CampbellScharlemann et al., 2010). This new focus on carbon as the main value to be conserved represented a departure from previous forest governance approaches, which had tended to focus on finding a balance between biodiversity conservation and economic use, notably timber extraction (Reference McDermottMcDermott, 2014). The involvement of new stakeholders from a climate perspective has brought an extra level of complexity and stimulated networked forest governance arrangements (Reference Reinecke, Pistorius and PregernigReinecke et al., 2014; Reference Visseren-Hamakers, Mcdermott, Vijge and CashoreVisseren-Hamakers et al., 2012).
In its early days, REDD+ was mainly implemented through projects that targeted specific forest areas, and in many cases such projects were spearheaded by conservation organizations (see, for example, Reference FergusonFerguson, 2009). As REDD+ has evolved, however, there has been a shift toward implementation through integrative and transdisciplinary governance approaches at the landscape level that engage stakeholders from different land-use sectors. Agribusiness companies seeking to reduce their impacts on deforestation, for example, have engaged in initiatives such as deforestation-free supply chains, sustainable commodity roundtables and certification schemes (Reference Boucher and EliasBoucher and Elias, 2013). This is often supported by provincial governments working to deliver emission reductions as part of nested efforts to deliver nationally appropriate mitigation actions (NAMAs), for example, through low emission rural development (LED-R) activities (Reference Nepstad, Irawan and BezerraNepstad et al., 2013). Such initiatives, which have developed subsequent to the adoption of the Aichi Biodiversity Targets, may overlap with efforts to scale-up terrestrial protected areas to the landscape scale and to designate conserved areas. This is expected to lead to more complexity and confusion in governance mandates due to the diversity of stakeholders involved who represent different interests, and bring different perspectives of “landscape” as either ecosystems and habitats, commodity production areas, administrative areas or territories with land rights (Reference Weatherley-Singh and GuptaWeatherley-Singh and Gupta, 2017).
In sum, under the framework of the Aichi Targets, the considerable progress made by national governments in designating sites has not been matched by efforts to ensure effective and equitable management, nor ecological representativeness, combined with limited consideration as to how to achieve conservation outcomes for biodiversity. Efforts made to achieve Target 11 have been undermined by the lack of progress in achieving other Aichi Targets, which are complementary and necessary to fully address the drivers of biodiversity loss but more difficult to measure and achieve. This includes targets that are concerned with conserving biodiversity outside of protected areas, for example Target 5, which aims to “at least halve, and where feasible bring close to zero, the rate of loss of all natural habitats”; Target 10, which is concerned with decreasing pressures on coral reefs; Target 7, on the sustainable management of areas under agriculture, forestry or aquaculture; Target 15, on enhancing ecosystem resilience; and Target 18, on the full and effective participation of IPLCs (Reference Fajardo, Beauchesne and Carbajal-lópezFajardo et al., 2021; Reference Hagerman and PelaiHagerman and Pelai, 2016; Reference Watson, Jones and FullerWatson et al., 2016b). There has also been little effort by high-income and importing countries to achieve Aichi Targets that focus on underlying drivers of biodiversity loss, such as Target 3 on the phasing out of harmful subsidies, and Target 4 on sustainable consumption and production.
There is, therefore, a need to ensure that protected and conserved area governance approaches (including international targets) achieve a better balance between site designation, and equitable and effective management and ecological representativeness. Furthermore, the impact on governance of newer landscape approaches, both due to the relatively recent inclusion of conserved areas as well as the engagement of new stakeholders from the climate sector, is still unclear. To inform this discussion, in the next section, we analyze the links between governance and biodiversity outcomes at the field level, from which lessons can be learned to inform recommendations for a transformative GBF.
11.4 A Transformative Policy Agenda: An Outcome-Based Approach to Protected and Conserved Area Governance
The last decade under the policy framework of the Aichi Targets has not provided a transformative governance agenda with clear outcomes for biodiversity conservation. We therefore discuss some of the policy and governance changes needed to redress this issue through an approach based on achieving biodiversity outcomes, in the light of the Post-2020 GBF and the growing importance of conserved areas. We draw on three case study examples that highlight how incorporating different transformative governance approaches can work in practice.
11.4.1 Case Studies
In this section, we present three case studies of protected areas and/or conserved areas with different forms of governance, in which the Wildlife Conservation Society (WCS) has been working, in order to draw some common lessons to inform a transformative policy agenda. The three selected case studies provide examples from different continental regions, varying ecosystem types and a range of governance scales and approaches. The first is Makira National Park and REDD+ project in Madagascar, the second is Kyeintali marine fisheries OECM in Myanmar, and the third is the Madidi-Tombopata Landscape in Bolivia and Peru (an area which encompasses both protected areas and OECMs).
Makira National Park and REDD+ Project
Makira National Park in the MaMaBay landscape in northeast Madagascar makes up the largest remaining intact humid rainforest in Madagascar, a country known for its unique endemic biodiversity. Containing half of Madagascar’s remaining coastal forest, a quarter of its lowland forest, 50 percent of all its flowering plant species, as well as coral reefs, mangroves and wetlands, the MaMaBay forest landscape receives some of the highest rainfall rates in the country. Despite its size and importance, the forests of Makira remain under threat from deforestation and unsustainable natural resource extraction. As the human population grows, traditional hillside rice cultivation (known as “tavy”) has become a major driver of forest loss.
The Makira National Park is managed collaboratively by the WCS as a “delegated manager” with “local community managers” of natural renewable resources,Footnote 6 thereby providing an example of inclusive governance. This institutional arrangement is based on Madagascar’s 1996 Secured Local Managed Forests (GELOSE) Law and the 2001 Contracted Management of Forests Decree, which delegate the management of some natural resources to Community Based Groups (COBAs). These rules and regulations underline local communities as the main actors in forest management and restore the legitimacy of local management of common resources (Reference SarrasinSarrasin, 2009). The Makira project is one of the world’s first forest carbon mitigation projects (thereby demonstrating an integrative approach), and at times REDD+ has provided a financial mechanism through which to fund the activities of COBAs and provide benefits to communities, although it has been necessary to secure supplementary income from official development assistance (ODA).
Combined efforts by WCS and the COBAs has strengthened the overall management structure of the Makira National Park. Park staff now work with communities to promote the sustainable use of natural resources through awareness-raising of COBA rules and regulations and environmental education activities. This is resulting in a decrease of anthropic pressure on the forests, demonstrated by the reduction of slash and burn, illegal settlements and clearing. For example, there was a decrease in the areas cleared from 834 to 605 hectares between 2016 and 2018. Improving the situation in future will require the reinforcement of joint patrols and enhancing access to justice for local communities, including by tackling some of the underlying drivers of biodiversity loss, such as corruption.
Kyeintali, Marine Fisheries Co-management Area
Kyeintali is located in the southern Rakhine state of Myanmar, one of the country’s poorest regions. Seventy-eight percent of the population live in poverty and over 80 percent are largely dependent on small-scale fishing for their livelihoods and subsistence. Traditional fishers (primarily men) and fish-workers who process the fish (primarily women) are rarely involved in decision-making or planning processes (Reference Matthews, Mizrahi and BoydMatthews et al., 2020). These coastal households are highly susceptible to impacts from the evident fisheries depletion. Recent interviews suggest that catches have more than halved in the past few years and provide evidence of bycatch of threatened species, though information is guarded and poorly documented (WCS, 2018).
The Kyeintali Inshore Fisheries Co-Management Area is now governed by the Kyeintali Inshore Fisheries Co-Management Association (KIFCA), which includes local community members (one man and one woman from each nearby village). Advisory and working committees composed of representatives from the government, police and Rakhine Fisheries Partnership support KIFCA. These groups were formed after a lengthy participatory process facilitated by WCS, which included the collection of scientific data on fishing activities, biodiversity and socioeconomic needs; detailed consultations with the fishing dependent communities and management planning in which the communities proposed their own no-take zones, seasonally closed areas, gear-restricted zones and protected turtle nesting beaches (Reference Exeter, Htut and KerryExeter et al., 2021; WCS, 2018). Following this process, the area was officially accepted by the government in 2018 as Myanmar’s first marine fisheries co-management area (Reference LattLatt, 2019).
Factors critical to success include inclusive and adaptive governance approaches. The engagement and recognition of the needs of all local stakeholders, combined with coordination of activities among the fisheries department, local coastal conservation association and local communities, for example, has been very important. Support from the Rakhine Fisheries Partnership helped secure strong relationships with Kyeintali fishers. This engagement was further supported by fair and open elections to select members of the management association, with efforts made to deliberately include women from the communities. Management decisions by the participants and stakeholders are also being supported by scientific evidence (primarily through GPS-based tracking of fishing activity, and household and market surveys), with the potential for adaptive management based on the outcomes of scientific surveys.
The process for developing the co-management area was slower than anticipated because this decentralized form of management is very new in this national context. However, other communities are now interested in developing similar management schemes. As the zoned areas were proposed by the communities themselves, it is expected that levels of compliance will be high. Such compliance will be key, as one of the greatest limitations to achieving sustainable fishing in coastal Myanmar is a lack of enforcement of marine-related regulations. In areas where enforcement is low, compliance must be won through local support, therefore a co-management area in which communities have a strong voice can be an appropriate strategy to recover local fish stocks, while also achieving biodiversity outcomes as a complementary goal.
Greater Madidi-Tombopata Landscape
The Madidi-Tambopata landscape is found in northwestern Bolivia and neighboring Peru and stretches from the High Andes to the tropical lowlands. It covers 14 million hectares, five national protected areas, three subnational protected areas and eight indigenous territories, as well as communities of ten indigenous groups, providing an example of inclusive governance. Connectivity and overlap between protected areas and indigenous lands across the Amazon is critical to maintaining intact forests, which are necessary for wide ranging species, such as the jaguar, as well as for maintaining globally important ecosystem services such as climate mitigation and freshwater provision (Reference Painter, Montoya, Varese, Rodríguez and AndersonPainter et al., 2017). The WCS has been working in the Greater Madidi-Tambopata landscape in Bolivia for two decades to support efforts by Indigenous People to secure legal recognition of their ancestral territories and increase their capacity to manage their lands and waters.
This is partly being achieved by the development of Indigenous Life Plans (or territorial management plans) for 1.8 million hectares of titled and claimed Indigenous territory. These plans enable Indigenous People to protect their lands, as well as using and managing natural resources in line with environmental, social and economic sustainability criteria, reflecting an inclusive and integrative approach. Such plans also contribute to the preservation of indigenous cultural identity and the revalorization of ancestral knowledge. They identify areas for achieving integrated conservation and development objectives, as well as connectivity corridors that link protected areas and Indigenous territories, to enhance the conservation of intact forest and healthy wildlife populations.
Management capacity-building processes have resulted in increased awareness among IPLCs of the environmental, economic and sociocultural value of their territories and have helped to secure local land rights. Local Indigenous People have worked together in the ordering and titling of their territories and benefit from increased security in access to and use of natural resources and the development of productive enterprises. The lives of Amazonian Indigenous Peoples depend on maintaining a harmonious relationship with nature for their spiritual, social, cultural and economic development. The Indigenous territorial management model has been developed by Indigenous People from their perspective and cultural identity, and also strengthens their commitment to conservation.
These three case studies demonstrate the critical importance of incorporating elements of transformative governance (particularly inclusive, integrative and adaptive approaches) to long-term success at the local level. In the next section, we draw some recommendations and discuss how these can benefit the development and implementation of policies at the international level under the Post-2020 GBF.
11.4.2 Moving toward a Transformative, Outcome-Based Approach to Conservation
As described in Section 11.3.1, implementation of Aichi Target 11 over the past decade has mainly focused on site designation, with limited attention given to achieving the second half of the target, relating to effective and equitable management and ecological representativeness. Instead of a quantified target that reflects the size of the area designated combined with measures of effective management, a transformative governance agenda for protected areas under the Post-2020 GBF needs to be based on achieving measurable outcomes for biodiversity. Discussions on the Post-2020 GBF, which are still ongoing at the time of writing, have still tended to focus on extending the coverage of protected and conserved areas (Reference Bhola, Klimmek and KingstonBhola et al., 2020; Reference Woodley, Locke and LaffoleyWoodley et al., 2019), but parallel discussions on measurable biodiversity targets (Reference Díaz, Zafra-Calvo and PurvisDíaz et al., 2020; Reference Geldmann, Deguignet and BalmfordGeldmann et al., 2021), combined with increasing recognition of the need to integrate IPLCs, means there is scope for the implementation of an outcome-based goal.
As shown by the three preceding case studies, achieving equitable management goals by involving IPLCs in the governance and management of protected and conserved areas and landscapes is a slow and time-consuming process but absolutely vital to ensuring conservation goals are achieved in the long term. The elements of transformative governance that have been incorporated within the case studies have not only been inclusive of IPLCs, but have worked toward ensuring the recognition and enforcement of their rights. The case studies demonstrate that a range of actions are needed, depending on the context, to ensure ownership and buy-in of IPLCs as well as achieving conservation outcomes. These actions may involve environmental education, strengthening of community access to and rights over land, enforcement measures, spatial planning and tools that enable scientific findings to be combined with local knowledge. As highlighted in the example from Myanmar, an inclusive approach is associated with increased compliance with regulations in areas with limited capacity for enforcement, thereby tackling a driver of biodiversity loss.
Ensuring new governance arrangements are equitable and inclusive is, therefore, of paramount importance. Although we advocate here a new governance approach based on the achievement of biodiversity outcomes, this must be accompanied by the achievement of equitable outcomes for IPLCs. The inclusion of conserved areas can assist in this regard, as such areas do not have to have biodiversity conservation as their primary purpose and can instead be managed for socioeconomic, cultural or other purposes. As mentioned, the role played by IPLCs has been undervalued and under-recognized until recently (see, for example, Reference Díaz, Settele and BrondízioIPBES, 2019). IPLC-led governance will be crucial in enabling conserved areas to contribute to biodiversity conservation as demonstrated in the three case studies described here. A greater role for IPLCs in decision-making and policy-setting at regional, national and international levels is only likely to facilitate the achievement of biodiversity outcomes and the management of potential economic and social trade-offs. The importance of properly including IPLCs has gained traction within discussions on the Post-2020 GBF, but much work remains to ensure this is fully embedded and implemented (Reference Fajardo, Beauchesne and Carbajal-lópezFajardo et al., 2021).
There is also a need for greater equity at the global scale, particularly in terms of the distribution of financial resources for biodiversity conservation. Although the Global South harbors most of the world’s important biodiversity, and the Global Environment Facility (GEF) was in part established to facilitate finance to these areas, the amount of financing under this mechanism is still inadequate. Discussions on CBD resource mobilization have moved toward increasing consideration of private sector sources of finance, to complement development aid and public sector support (OECD, 2020). New governance arrangements for conserved areas and landscape-level approaches under the climate regime may facilitate the resource mobilization agenda by involving the private sector and ensuring that climate finance simultaneously achieves biodiversity outcomes, although, as shown by the Makira National Park case, it often needs to be complemented by other types of finance, including ODA. An approach based around biodiversity and equity outcomes could be accompanied by a financing framework, under which the areas or countries with the highest biodiversity values are identified and prioritized due to the increased focus on achieving ecological representation and biodiversity outcomes. This could facilitate transfers of finance from high-income, importing countries to low-income countries, and the establishment of new, innovative forms of financing mechanisms, which could even include performance-related payments, based on the achievement of biodiversity outcomes.
At a global level, considerable work is underway to better assess management effectiveness (and to a limited extent, governance effectiveness) of protected areas (Reference Geldmann, Deguignet and BalmfordGeldmann et al., 2021). Conservation areas or OECMs are, by their nature, considered to be “effective” conservation measures and, according to the accepted definition, they must result in biodiversity outcomes, regardless of the primary objective for which there are managed. Rather than attempting to assess the effectiveness of protected area management with a relatively small amount of monitoring effort dedicated to monitoring outcomes, a monitoring approach can be adopted for both protected and conserved areas (i.e. OECMs) that focuses more strongly on biodiversity outcomes. This will enable more responsive to adaptive and anticipatory governance responses. Post-2020, implementing an outcome-based approach will be more practical and cost-effective than it has been in the preceding decade, at least for terrestrial sites, due to the rapid advances being made in the area of remote sensing tools and the availability of proxy data on which to base an estimate of biodiversity outcomes (Reference Watson and VenterWatson and Venter, 2019).
Rapid advances in remote sensing that can monitor biodiversity outcomes can also assist in the future designation of terrestrial protected and conserved areas by ensuring that areas are selected for their biodiversity value. More attention needs to be given, however, to the development of equivalent tools for marine areas. This would ensure that an outcome-based approach is taken to site designation as well as to management and governance. In the case of conserved area designation, criteria need to be developed that provide a common global understanding of what constitutes an accepted biodiversity outcome. This would prevent a situation in which, for example, a marine area managed primarily for fisheries is designated because it results in positive outcomes for one or two specific species, despite little discernible benefit for a wider assemblage of species, and even potential harm due to bycatch. Recent discussions to designate much larger areas for biodiversity, such as the “half-earth” approach to set aside half of the earth for nature, as proposed by Reference WilsonWilson (2016), would indeed help to achieve greater ecological representation in some areas. The concept has received some criticism, however, from those who view this as a land-grab by the conservation community (Reference Dudley, Jonas and NelsonDudley et al., 2018), and for some ecoregions not enough natural habitat remains to meet this goal without substantial restoration efforts (Reference Dinerstein, Olson and JoshiDinerstein et al., 2017; Reference Mappin, Chauvenet and AdamsMappin et al., 2019). A transformative agenda needs to go beyond target-setting and aim for ecological representation and achievement of biodiversity outcomes at the global level, accompanied by equity outcomes. As noted previously, this should be complemented with a mechanism for increasing financial resources to low-income countries that are high in biodiversity. It would also need to provide scope for restoration of ecoregions that have been most depleted.
As shown by the case study from Bolivia, new outcome-based approaches to transformative governance are necessary where protected areas with the primary goal of biodiversity conservation are located in large landscape and seascape areas where biodiversity conservation and equity are being achieved even in the areas that are not managed for biodiversity per se. It also gives space for adaptive, integrative and anticipatory governance responses by decreasing some of the pressures on biodiversity. Given the increasing pressure on land from many sectors, conserved areas with the primary purpose of food production and carbon storage will necessarily play an increasingly important role. Such an approach will, however, require the development of an accepted global standard, with associated criteria to determine what constitutes positive biodiversity outcomes. This will be facilitated by new cost-effective monitoring based on satellite remote sensing.
The move toward landscape approaches under REDD+ and other policy initiatives by provincial governments under the climate regime, such as LED-R initiatives, present a potential opportunity for integration with the emerging landscape-based approach to protected and conserved area governance. New points of intersection between the climate, forest, agriculture and biodiversity policy agendas could facilitate innovative, integrative forms of governance and financing for large areas. There may be potential for conserved areas to be designated and managed primarily for their carbon values, with biodiversity benefits as a major outcome. The involvement of new stakeholders from different interest groups in governance arrangements, particularly from the agricultural and climate sectors, and the potential complications that can arise from competing objectives, however, should not be underestimated. The establishment of an outcome-based approach with agreed minimum standards and criteria for achieving biodiversity outcomes could assist in managing and agreeing such trade-offs by ensuring that a certain threshold of biodiversity value must be met.
To be truly transformative, the implementation of this approach also needs to be combined with efforts by high-income or consuming countries to address some of the underlying drivers of biodiversity loss linked to unsustainable consumption and global trade that negatively impacts biodiversity in low-income countries (Reference Lenzen, Moran and KanemotoLenzen et al., 2012). This type of integrative governance approach is now gaining traction within forest policy discussions, as governments in consuming countries consider how they ensure that their imports of agricultural commodities are deforestation-free (Reference Weatherley-Singh and GuptaWeatherley-Singh and Gupta, 2018). Such policies need to be advanced to ensure that supply chains do not cause biodiversity loss in producer countries. This would also help the Post-2020 GBF to address the pressures on biodiversity, and not just the responses to those pressures (OECD, 2019).
A review and revision of the IUCN management and governance categories would also be necessary to enable a new approach based around biodiversity and equity outcomes. Notwithstanding the crucial need to retain a number of protected areas that are managed primarily for biodiversity conservation, these could be incorporated within a set of IUCN protected and conserved area categories that is based around the achievement of biodiversity and equity outcomes. The current IUCN categorization of governance types could also be reviewed and expanded to encapsulate and include the full range of conserved areas and their outcomes for biodiversity, to better reflect their governance and management structures.
11.5 Conclusion
Although the international policy framework, and particularly CBD Aichi Target 11, has stimulated further progress in protected and conserved area governance, especially in site designation, this falls short of meeting the criteria for transformative biodiversity governance. Efforts have been made in recent years to ensure such areas become more inclusive (through, for example, co-governance arrangements that engage IPLCs). The valuable role played by IPLCs is starting to gain recognition, including in new discussions around conserved areas, but governance needs to go beyond including them as beneficiaries, to recognizing and strengthening their rights as active stakeholders. This will require a considerable investment of time and resources at local and landscape levels to conduct inclusive consultations, build capacity where needed, especially in terms of access to technology, and to find solutions that meet the needs of IPLCs and that reflect their own visions for their territories, which will ultimately be more sustainable.
There has also been a focus on achieving adaptive management that allows for some experimentation in management approaches. This principle can be used to quickly integrate new scientific findings, which are now providing more timely, up to date information on species and habitats and the human pressures they are facing. This increase in information is occurring from the local to global level, enabling decision-making to be better informed and potentially also anticipatory, for example, by modeling the impacts of human pressures and facilitating future site designations.
Greater information on biodiversity outcomes will also enable finance to be directed to the areas of greatest biodiversity value, thereby helping to achieve greater ecological representation. The potential scaling-up of protected areas to become part of decision-making governance structures at landscape and seascape scales, including conserved areas, is expected to open the door to greater integrative approaches and new forms of financing, although such arrangements bring a new level of complexity.
In conclusion, a new approach based around the delivery of biodiversity outcomes could help drive forward a transformative governance agenda. Its success will also depend on the long-term engagement of IPLCs and the achievement of equity outcomes. Reviewing the IUCN governance and management categories would be an additional small first step toward building a more supportive policy framework at the international level that facilitates transformative change. If such efforts are combined with actions taken by high-income, importing countries to increase the sustainability of their consumption and trade patterns, and thereby tackle some of the underlying drivers of biodiversity loss, this would be even more transformative.
12.1 Introduction
News on the state of the environment does not seem to be improving. Despite some holding on to “conservation optimism,”Footnote 1 the general conclusion in the academic and policy literature is that global biodiversity, the global climate and the state of other environmental indicators are bad, and getting worse (CBD, 2020; European Environment Agency, 2019; IPBES, 2019; Reference Lenton, Rockström and GaffneyLenton et al., 2019; Reference Newbold, Hudson and ArnellNewbold et al., 2016; Reference Tucker, Böhning-Gaese and FaganTucker et al., 2018; Reference Watson, Shanahan and Di MarcoWatson et al., 2016; WWF, 2018). This has resulted in growing calls for transformative change in the way we govern biodiversity, and the environment more broadly (Reference Bennett, Blythe, Cisneros-Montemayor, Singh and SumailaBennett et al., 2019; Reference Scoones, Stirling and AbrolScoones et al., 2020). Making incremental, adaptive changes to the current system and structures is no longer considered sufficient to move us to a sustainable future; rather, deeper, more fundamental transformation is needed (Reference Blythe, Silver and EvansBlythe et al., 2018). In relation to biodiversity conservation, an important example of this new emphasis is the 2019 Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES) report, which argues that “nature can be conserved, restored and used sustainably while simultaneously meeting other global societal goals through urgent and concerted efforts fostering transformative change” (IPBES, 2019: 7). The report realizes this is not easy, but insists:
Since current structures often inhibit sustainable development and actually represent the indirect drivers of biodiversity loss, such fundamental, structural change is called for. By its very nature, transformative change can expect opposition from those with interests vested in the status quo, but such opposition can be overcome for the broader public good.
Clearly, “transformative change” is an extremely complex proposition, and precisely what it means is widely debated and contested (Reference Brown, O’Neill and FabriciusBrown et al., 2013; Reference Scoones, Stirling and AbrolScoones et al., 2020). The IPBES (2019: 9) report, however, provides many suggestions, including a particularly important one: “A key constituent of sustainable pathways is the evolution of global financial and economic systems to build a global sustainable economy, steering away from the current limited paradigm of economic growth.” The European Environment Agency, likewise, states that economic growth should no longer be pursued at the expense of the environment and urges governments to “deliver transformative change in the coming decade” (European Environment Agency, 2019: 10).
Coming from major international reports, these are not just “regular” transformative suggestions; they are radically transformative suggestions that go to the roots (“radix”) of the problem of contemporary unsustainability.Footnote 2 Demand for such change is echoed by a growing number of civil society groups, networks and social movements battling the myriad environmental and social conflicts caused by unfettered economic growth and consumption.Footnote 3 And while the global COVID-19 pandemic had many governments and institutions scrambling to get back to “normal,” it also amplified the demands for transformative change. The key questions, then, become: How do we act on these demands and suggestions? What do they imply for environmental governance and biodiversity conservation?
In this chapter, we support and advance arguments for a fundamental structural transformation that envisions radically different institutional and societal structures. This view is in line with the current volume and increasingly shared by many calling for transformative change (e.g. Reference Chaffin, Garmestani and GundersonChaffin et al., 2016; Reference Martin, Teresa Armijos and CoolsaetMartin et al., 2020; Reference Massarella, Nygren and FletcherMassarella et al., 2021). At the same time, many actors still believe that transformative change can happen without directly and explicitly challenging the capitalist underpinning of contemporary institutional and societal structures (Reference FeolaFeola, 2020). We argue that many of the solutions put forward for transforming biodiversity conservation follow this belief. More specifically, we argue that even seemingly “radical” new approaches, such as “neoprotectionism” (focused on creating space for protected areas) and “new conservation” or natural capital approaches (championing the use of market-based mechanisms to integrate people and nature) are not actually transformative. Although they call for radical shifts – both in symbolism and how we govern biodiversity on a global scale – they do not sufficiently address or challenge the main driver of biodiversity loss: the neoliberal capitalist model that dominates our global political economy. In fact, by not responding holistically and critically to the global challenges we are facing, including currently disturbing authoritarian trends in global governance systems and an increasing concentration of corporate governance, these proposals for transformative change may even set us back.
We therefore argue that the only way to properly conceptualize transformative change is to combine radical reformism in the short term with an intermediate to long-term vision for fundamental structural transformation that directly challenges our contemporary capitalist political economic model and its newfound turn to authoritarianism. In doing so, we emphasize, following Reference Scoones, Stirling and AbrolScoones et al. (2020), that our structural approach can and should be seen in conjunction with – not necessarily against – what they call systemic and enabling approaches that focus more on complex system change and values and actions of different actors. The latter, however, can only gain (appropriate) direction through a critique of the dominant political economy and hence why we emphasize structural transformation. In what follows, we contribute to the current volume by presenting a vision for structural transformation under the banner of “convivial conservation.” Convivial conservation is a vision, a politics and a set of transformative governing principles that moves biodiversity governance beyond market-based mechanisms and a central focus on protected areas (PAs). We outline and analyze these three elements and propose the idea of “biodiversity impact chains” (BICs) to operationalize some of the transformative governance aspects of convivial conservation in practice.
BICs, in essence, aim to politicize transformative environmental governance by drawing more concrete connections between differentiated actors, and their variegated impacts on biodiversity, in a highly uneven conservation field. This allows us not only to understand that those with the largest footprints must change their lives the most in order to redress biodiversity loss, but also that spatial proximity to conservation areas should be of less concern to conservation action than is often the case (see also Chapter 14 of this volume). BICs, therefore, help us to gain a clearer view of the structural pressures on biodiversity, and how these need to be mediated or challenged in order to achieve structural transformation. In the penultimate section of the paper, we develop this perspective in more detail in order to explain, in the conclusion, how a convivial transformation may be our most realistic chance to respond positively to the global biodiversity crisis. First, however, we summarize the arguments for fundamental structural transformation and what we believe this should entail.
12.2 Authoritarian Currents and the State of Biodiversity (Conservation)
As is apparent from the preceding discussion, the necessity for fundamental transformation is becoming increasingly obvious in global environmental governance circles and, indeed, within global governance more generally. As shown by many other chapters in this volume, most environmental indicators around climate, oceans, biodiversity, forests and more are so alarming that even most mainstream commentators now call for forms of change beyond mere nudges within the general parameters of the current system. Much evidence from the current transformations literature could be presented here, but for an overview we refer to Chapters 1 and 4 in this volume and Reference Massarella, Nygren and FletcherMassarella et al. (2021). What we want to add is a more sociological analytic, namely that the mainstream system in which global environmental governance approaches have been operating is increasingly leading to forms of authoritarian populism and right-wing extremism. Prominent examples include the recent Trump regime in the United States, the Bolsonaro regime in Brazil and the Modi regime in India, among others (Reference KielyKiely, 2021; Reference Saad-Filho and BoffoSaad-Filho and Boffo, 2021). All of these regimes articulate narrow versions of both nation and nature, to the extent that Indigenous and other minority groups are frequently cast as the enemies of national economic progress, often violently so. Indeed, one key constant across these regimes is that they have come to power with the support of major extractive industries and have, in turn, unapologetically exercised their power in support of these industries to directly attack and dismantle forms and institutions of environmental protection that stand in their way (Reference KielyKiely, 2021; Reference McCarthyMcCarthy, 2019; Reference Saad-Filho and BoffoSaad-Filho and Boffo, 2021).Footnote 4
We argue that these worrying trends need to be acknowledged and challenged directly for transformative conditions to arise (Reference MasonMason, 2019). After all, as Reference PolanyiPolanyi (1957) argued as far back as the 1950s, the rise of authoritarianism is the ultimate response to the threat that social and environmental protection poses to the continued advancement of neoliberal capitalism. As crises of capitalism are increasingly accompanied by crises of legitimation (of the continuation of “business-as-usual”), authoritarianism offers a solution to both. This “authoritarian fix” allows for capital accumulation to continue by removing barriers to the exploitation of natural resources and labor, while simultaneously removing the need for legitimation (Reference BruffBruff, 2014: 125; Reference PoulantzasPoulantzas, 1978). Thus, it is hardly coincidental that many of the new authoritarians have sought to undermine or withdraw from global institutions focused on climate change mitigation, at the precise moment of a growing political tension between environmental protection and economic business-as-usual. The dissolution of restrictions on agriculture and mining have gone hand-in-hand with a denial of the scientific truth of environmental degradation, and widespread attacks on agencies producing spatial data on deforestation and defaunation (Reference Neimark, Childs and NightingaleNeimark et al., 2019). By undermining protections at all levels, new authoritarian regimes thus act to sustain a capitalist economy that demands continuous growth in order to remain stable (Reference Büscher and FletcherBüscher and Fletcher, 2020). From this perspective, the fight against environmental catastrophe is also a fight against authoritarianism, given how the latter is directly implicated in the defense of the current capitalist political economy (Reference KielyKiely, 2021; Reference McCarthyMcCarthy, 2019; Reference Saad-Filho and BoffoSaad-Filho and Boffo, 2021; Reference Scoones, Edelman and BorrasScoones et al., 2018).Footnote 5
Given this context and these threats, it is little surprise that many in the conservation community feel great anxiety and pressure. And while they do often agree that transformation is needed, it seems very difficult to break out of the neoliberal consensus-mold many organizations embraced in the 1980s and 1990s. As documented in the literature (Reference AdamsAdams, 2017; Reference FletcherFletcher, 2014; Reference MacDonaldMacDonald, 2010; Reference MacDonald and CorsonMacDonald and Corson, 2012), since the 1980s conservation organizations have increasingly conformed to the general, consensus-oriented “sustainable development” models that have thoroughly neoliberalized biodiversity conservation (Reference Fletcher, Dressler, Anderson and BüscherFletcher et al., 2019). Indeed, Reference BüscherBüscher (2013) identifies consensus and anti-politics as two of three foundational elements of a general neoliberal conservation politics that pervaded the 1990s and early 2000s (with “marketing” being the third). Since the late 2000s, and especially triggered by the 2007/2008 financial crisis, the international political context has changed rapidly, leading – inter alia – to the abovementioned authoritarian developments. One would expect that, from the imperative to oppose these forces, a more political and less consensus-oriented approach to environmental governance would ensue. Yet, this has only marginally proven to be the case thus far.
For example, the WWF flagship Living Planet report, released two days after Bolsonaro was elected in November 2018, calls for a “new global deal for nature and people” and urges “decision-makers at every level” to “make the right political, financial and consumer choices to achieve the vision that humanity and nature thrive in harmony on our only planet.” To operationalize this “ambitious pathway,” WWF, together with other organizations, will launch a new research initiative based around “systems modelling” to help “us determine the best integrated and collective solutions and to help understand the ‘trade-offs’ we may need to accept to find the best path ahead” (WWF, 2018: 8). Similarly, the European Environment Outlook 2020 paints a grim picture of prospects for European biodiversity and argues that its “message of urgency cannot be overstated.” At the same time, it states that “transformative change will require that all areas and levels of government work together and harness the ambition, creativity and power of citizens, businesses and communities” (European Environment Agency, 2019: 7; 17). On a superficial level, this may be correct, but it leaves out which businesses, types of activities and communities (such as the oil, coal, infrastructure, large-scale agriculture and other communities) will inevitably have to “lose” (that is, to degrow, and rapidly so) in order to reach a more sustainable overall state.
To a degree, we can understand that conservation and government organizations want to be careful politically. But a big problem with this conciliatory, mainstream approach is that it is easy to ignore for alt-right and authoritarian (-leaning) politicians and movements, and their corporate backers. Another problem is that it often does not lead to more political or politicized action to demand structural change, and may – unintentionally – lead other actors to take politicized action into dubious terrains of increasingly militarized, even ecofascist, forms of environmental protection that often further marginalize local communities (Reference Duffy, Massé and SmidtDuffy et al., 2019). As a result, we have seen more direct-action movements such as Extinction Rebellion, Fridays for Future and others rapidly take center stage in environmental politics, while, from within the conservation community, we have also seen more radical alternatives emerge to challenge mainstream approaches.
Two of the more prominent conservation communities espousing discontent at the status quo are “neoprotectionists” and “new conservationists.” New conservationists have been quite radical in a sense, as they have started criticizing the key elements on which the global conservation movement has been built since the nineteenth century: protected areas and the ideas of “pristine” nature and wilderness. Instead, they suggest a full integration of conservation into dominant, capitalist political economic systems for conservation to stand a chance in the future and maintain or retain legitimacy (Reference Kareiva, Marvier and LalaszKareiva et al., 2012). In this way, they build on a growing trend within mainstream dominant approaches to conservation, represented, among others, by the Capitals Coalition, which aims to turn nature and natural resources into a form of “capital” that can be traded on markets and used to offset more regular forms of development (Reference FletcherFletcher, 2014; Reference Fletcher, Dressler, Anderson and BüscherFletcher et al., 2019).
Yet another community of conservationists – “neoprotectionists” – strongly contest the new conservationists. Neoprotectionists believe that the new conservation strategy would not only be the death of conservation, but of the entire planet (Reference Wuerthner, Crist and ButlerWuerthner et al., 2014; Reference Wuerthner, Crist and Butler2015; Reference WilsonWilson, 2016). Basing their conservation objectives and strategies on conservation biology science, neoprotectionists believe that to ensure long-term viability of an ecosystem, nature must be set aside from the influence of people (Reference Locke, Wuerthner, Crist and ButlerLocke, 2015). Such ideas have important lineages to colonial conservation strategies, in which fences, fines and ideas about “pristine wilderness” were crucial tools to evict people from protected areas and to keep them out. According to neoprotectionists, we need to go back to protected areas and wilderness protection, but on a scale hitherto unseen. Some even argue that only if at least half the planet becomes a system of nature reserves can the ecological processes critical to human and planetary survival persist (Reference WilsonWilson, 2016).
Since earlier versions of these movements were suggested, they have also morphed, nuanced and developed. Neoprotectionist approaches, for instance, have reduced their emphasis on “protected areas only” somewhat to focus also on other conservation measures. They have also given more attention to social goals related to conservation, seemingly embracing a “social turn” that aims to bridge nonhuman nature and people (Reference EllisEllis, 2019; Reference Locke, Ellis and VenterLocke et al., 2019). While inclusion of Indigenous knowledge in conservation is now discussed by neoprotectionists (Reference LockeLocke, 2018), it still remains quite vague, with sparse and somewhat superficial references to land rights and integration of Indigenous knowledge in policymaking, while separating humans from nature via protected areas is maintained (Reference LockeLocke, 2018). It is unclear how this emergent “social turn” will manifest and be integrated into neoprotectionist visions on protected areas, a concern further highlighted by recent research finding that protecting half of the Earth might negatively affect over one billion people and result in widespread social and environmental injustices (Reference Schleicher, Zaehringer and FastréSchleicher et al., 2019). Of particular importance, climate mitigation and adaptation are now widely discussed and tentatively integrated into protected area targets in order to accommodate broader regimes, such as around the sustainable development goals (SDGs). In this regard, Reference Dinerstein, Vynne and SalaDinerstein et al. (2019) proposed a “Global Deal for Nature” including half Earth approaches that they believe should be paired with the 2015 Paris Climate Agreement.
Clearly, the debate on biodiversity governance is dynamic, diverse and rapidly changing in response to ongoing socioecological dynamics. Within these diverse dynamics, however, two core issues remain central: how to relate people to the rest of nature and how to situate conservation vis-à-vis the political economy of neoliberal capitalism. And despite more recent iterations that nuance earlier and more radical proposals to mix people and nonhuman nature through “natural capital” valuation, or separate people and nature on an unprecedented global scale, it is doubtful whether the dominant options currently on the table can provide a productive way forward. As argued in Reference Büscher and FletcherBüscher and Fletcher (2020), none of the current approaches will provide the fundamental structural transformations needed, as they do not directly confront the drive for continual accumulation of capital at the heart of the neoliberal capitalist economy. Neither do they sufficiently engage with the social injustices that have historically plagued both protectionist and market-based approaches to environmental governance (Reference Martin, McGuire and SullivanMartin et al., 2013). Nor do they take into account the vast differences in ways of knowing nature, environmental values and perspectives on what makes “good governance” (Reference Sikor, Fischer, Few, Martin, Zeitoun and SikorSikor et al., 2013). We therefore need a different approach to transformation that can bring about the “substantial, profound and fundamental change” required (Reference Massarella, Nygren and FletcherMassarella et al., 2021). We outline one pathway to transformative change through the alternative approach of convivial conservation.
12.3 Convivial Conservation: Vision, Politics, Governance
Convivial conservation emphasizes the vision, politics and governance mechanisms needed for a realistic, structural transformation of biodiversity protection. This is because convivial conservation is founded on a political ecology approach that is critical of contemporary capitalism, the global and unsustainable political economy it has spawned over the last centuries and the recent increase in global authoritarianism (Reference Büscher and FletcherBüscher and Fletcher, 2020). This makes convivial conservation itself a political economic approach to environmental governance, characterized by questions such as: How can we understand political economy and international development from the perspective of integrated socioecological dynamics around biodiversity? Or, how can a concern for biodiversity become central to the ways we (need to) rethink the relationship between political economy and development generally? And how does this lead to the implementation of concrete policies and measures at all levels that are sustainable, equitable and just? In short, convivial conservation is a critical-constructive approach that, contrary to practice-oriented, consensus and neoliberal approaches, bases its strategy on a critique of the structural context within which actors and organizations maneuver.
While a fuller elaboration of the convivial conservation vision has been published elsewhere (Reference Büscher and FletcherBüscher and Fletcher, 2020), it can be summarized as a postcapitalist, political economic approach to conservation that aims to integrate and reconnect people and nature in landscapes across different scales, spaces and times. The convivial conservation vision functions within the broader transformative vision of degrowth: an overall quantitative downsizing of economic throughput to ecologically sustainable levels coupled with widespread wealth redistribution to make this reduction “socially sustainable” (Reference D’Alisa, Demaria and KallisD’Alisa et al., 2015; Reference Hicks, Levine and AgrawalHicks et al., 2016; Reference Holland, Peterson and GonzalezHolland et al., 2009; Reference KallisKallis, 2011; Reference RaworthRaworth, 2017; Reference Wilkinson and PickettWilkinson and Pickett, 2010). Within these overarching contexts, convivial conservation defines specific parameters for a fundamentally different form of conservation that does not separate people and nature. This means that protected areas and urban centers, as the two quintessential “end-points” of traditional human–nature dichotomies, have to be connected more, with the ultimate aim of achieving a better balance between human and nonhuman lives and needs across urban and rural spaces.
Convivial conservation envisions five fundamental shifts for conservation: moving from protected to promoted areas; from a framing of saving nature to one of celebrating human and nonhuman nature; from touristic voyeurism to engaged visitation; from a focus on spectacle to a focus on everyday environmentalisms and from privatized expert technocracy to common democratic engagement (Reference Büscher and FletcherBüscher and Fletcher, 2020: 163–174). In line with the themes of the book, this chapter focuses on how this vision is also a politics and form of governance. Central to convivial conservation is the fact that it politicizes conservation – meaning that it explicates the interests of different actors and how they may or may not be compatible, and always function within broader frameworks of power. Convivial conservation, therefore, is not focused on achieving consensus and does not believe that all actors with widely differential interests can or want to come together to promote biodiversity conservation. Rather, it conceptualizes biodiversity conservation as a political struggle caught up in histories and contexts of power that provide structural and agentic challenges and barriers. In this struggle, commonalities need to be sought and created, but not at the expense of the overall political direction of the convivial conservation vision, which, as mentioned, entails (moving toward and encouraging) degrowth, wealth redistribution and, ultimately, postcapitalism. In this sense, convivial conservation also aligns with environmental justice movements that conceptualize political struggle as an imperative part of radical transformation (Reference PellowPellow, 2017; Reference Temper, Walter, Rodriguez, Kothari and TurhanTemper et al., 2018).
This brings us, finally, to governance, or the way that actors steer, direct and influence affairs in particular directions. Governance mechanisms include, among others, legal regimes, state and other forms of organization, (formal and informal) institution-building or breaking, and more, in both material and discursive forms. What constitutes biodiversity governance is very broad and encapsulates a wide range of actors, activities and approaches. However, the concept of “transformative governance” established in Chapter 1 of this book is more specific, and it is this concept that is at the center of convivial conservation. Transformative biodiversity governance is understood as a product of deliberate and political acts that directly challenge embedded power structures, dominant agendas and framings, and mainstream approaches to conservation (see Chapter 1). In order to disrupt embedded hierarchies and power structures and bring about this transformative governance, we must first critically interrogate the (historical and contemporary) framings, responsibilities and roles of different actors within biodiversity conservation.
Table 12.1 provides a heuristic basis for such an analysis, depicting our conceptualization of four broad categories of conservation actors and organizations. We regard rural lower classes (category 4) as those actors who often live in or with biodiversity and who (still) depend on the land for subsistence, especially in tropical countries. They are often (seen as) poor and the ones who have least contributed to global problems of biodiversity loss (historically and contemporarily). Yet they are most often targeted in conservation interventions and forced or “incentivized” to change their livelihoods to meet biodiversity targets. Category 3 actors comprise urban, semiurban or semirural middle and lower classes throughout the world, who are not directly land-dependent for subsistence and who participate and rely on local and global labor and consumer markets. Through their consumption and place in global markets, they do heavily influence biodiversity in many places, but are often not part of or specifically targeted by conservation interventions, except as potential donors.
Table 12.1 Generic categorization of classes important for conservation
| 1. Upper classes |
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| 2. Land-owning capitalist classes |
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| 3. Middle and lower classes |
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| 4. Lower rural classes |
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Category 2 actors are land-owning capitalist classes such as major capitalist farmers and/or landholders for large agro-industry. They are often targeted by conservation, not as part of community-based interventions, but as partners in the conservation effort or as targets of (so-called) activist interventions or forms of resistance. In many places (Indonesia, Brazil, Central Africa and so forth), these classes are also part of violent frontiers of land conversion, and hence difficult to target and engage with (Reference CampbellCampbell, 2015). Finally, category 1 actors comprise the global upper classes that are, politically, economically or otherwise, at the helm of the global capitalist system (often referred to as the “transnational capitalist class”; see Reference SklairSklair, 2001). These elite actors are often both urban and rural – owning multiple properties, including in rich residential areas in cities to be close to elite political-economic circles, but also with second, third or even more properties in rural, semirural and biodiversity-rich spaces, including large estates and private reserves (Reference HolmesHolmes, 2012). Upper-class elites are often recruited as funders or included on boards of conservation organizations, but rarely targeted as part of conservation initiatives aiming at behavioral or livelihood change, as they are often either seen as unreachable or as doing good for the environment through their philanthrocapitalism or other forms of conservation-related charity (including through the privatization of nature/parks, etc.). Hence the upper classes have a strange double role, as they are at the helm of the system that keeps the pressure on biodiversity intense and high, while also considered either untouchable or even to be championing conservation through their large donations to conservation causes, NGOs and more (Reference EdwardsEdwards, 2008; Reference Ramutsindela, Spierenburg and WelsRamutsindela et al., 2011).
While empirical reality is much more complex than this table can depict, its point is that currently dominant conservation paradigms focus mostly on category 4 actors in terms of whose lives need to change. Convivial conservation would change this and target actors according to their differential responsibilities and accountabilities in relation to both the direct and indirect impacts of their actions on biodiversity, as well as the relative power these actors possess within broader structures of capital accumulation. Paraphrasing Reference Moore and MooreMoore (2016), it is about identifying, targeting and “shutting down the relations” that produce biodiversity loss, not just about geographical proximity.
In this way, we might reverse the model of “polycentric” governance proposed by Ostrom and others (e.g. Reference Ostrom and CoxOstrom and Cox, 2011). In this standard model, governance is seen to start with local people and then must consider their embeddedness within overarching structures of governance with which they must contend to assert their space for self-governance. In our vision, by contrast, effective conservation governance would start by addressing actors in these superordinate levels in order to first target their actions, then work down toward the local people in direct contact with the biodiversity in question. In this way, the pressures exerted on local conservation initiatives can be proactively addressed at their source rather than merely retrospectively in relation to their impacts.
We should clarify that this governance model pertains only to the ways that conservationists frame and confront threats to conservation, not to how decision-making regarding effective conservation should proceed. The latter must embody deeply democratic forms of engagement in which local actors, those generally affected heaviest by conservation measures, are placed at center stage (see also Chapter 8). A convivial conservation politics, therefore, must simultaneously center local people as key decision-makers in conservation planning and decenter them as the central targets of interventions aimed at fostering behavioral change. This analysis gives rise to a number of questions, and in a short chapter it is not possible to work out all the details of the convivial conservation vision and the politics it necessitates. Our analysis does, however, point to the need for transformative governance mechanisms that disrupt this conservation class structure, “trigger regime shifts” and ultimately alter the “structures and processes that define the system” (Reference Chaffin, Garmestani and GundersonChaffin et al., 2016: 400).
We have previously put forward some suggestions for transformative governance mechanisms, including a program of historical reparations directed at category four actors, developing “integrated conservation landscapes” that prioritize human and nonhuman coexistence (Reference Büscher and FletcherBüscher and Fletcher, 2020), and alternative finance mechanisms such as “conservation basic income” for those living close to areas of high biodiversity (Reference Fletcher and BüscherFletcher and Büscher, 2020). In the remainder of this chapter, we discuss the rationale for BICs as both a political methodology and a transformational governance mechanism. The basic idea behind BICs is simple: to better understand and politicize the relationships between different actors and the impacts that their livelihoods and consumption choices have on the conservation of particular forms of biodiversity. BICs challenge many of the embedded assumptions that we have previously outlined in this section by refocusing attention onto those with the largest footprints – likely to be in class 1 and 2 – while challenging the problematic focus on class 4 actors. In doing so we open up the potential for transformative change in biodiversity governance, as the focus of conservation discourses, actions and interventions shifts onto those with the biggest footprints.
12.4 Biodiversity Impact Chains
The idea of BICs is partly inspired by the value chain literature, which studies value supply chains to see how commodities are produced, distributed and consumed, and to study social, political and environmental issues along the way (Reference Bair and BairBair, 2009). The value chain literature has developed in numerous directions, including how value chains relate to forms of more sustainable production or the tracing of knowledge as a valuable commodity in its own right (Reference BüscherBüscher, 2014; Reference Guthman and BairGuthman, 2008; Reference PontePonte, 2019). A classic example comes from Reference HartwickHartwick’s (1998: 426) focus on gold, where she shows how production, processing and consumption dimensions are connected through “vertical” long-distance relationships but also consist of “horizontal” dimensions of local interrelationships along various points on the chain. She contends that the production of one commodity can imply multiple chains, while along points on a singular chain “halo-effects” can occur. In this way, wider social and environmental effects are brought about by particular activities along the chain.
A major critique in much of the literature on value chains is that they have quite a linear understanding of the chains they describe and a very simplistic or instrumental idea of the “value” they envision. According to Reference StarostaStarosta (2010: 435):
[W]hat commodity chain studies do is simply to offer, through an essentially inductive-empirical methodology, a typological description of the immediate outer manifestations of the determinations at stake. This failure firmly to explain the nature of GCCs [Global Commodity Chains] is expressed, for instance, in the disjuncture between the portrayal of the particular dynamics internal to each industry and the general dynamics of the “system as a whole.”
Like others (Reference PontePonte, 2019), Reference StarostaStarosta (2010: 455) argues that we should pay more attention to irregular circulatory dynamics of value, rather than “captive governance structures” that work according to linear models of how value is produced. The same lessons apply for how we should study the idea of “impact chains.” Like value or commodity chains, the last decades have seen a major literature develop around the idea of impact, including in relation to “cross-sectoral cumulative impacts” that we draw upon and are inspired by (Reference Baird and BarneyBaird and Barney, 2017).
Building on these important considerations, we imagine BICs as a political methodology and a governance mechanism to (further) study, map and steer political economic activities in particular bioregions (both urban and rural, and everything in between) and how they relate to specific ecosystems and biodiversity that provide the (raw) materials for these activities. In many cases, this is impossible to establish given the complex considerations above. Hence, we consider starting with specific ecosystems wherein this dependency can be most directly established. These could include (fresh) water, as the distances between water and their use – although they can be large – are often local or regional. As the important case of the drought in Cape Town, South Africa, in 2018 shows – a recent example of a major global city facing an acute water crisisFootnote 6 – the conservation of water sources is critically important, and depends on complex political-ecological factors, some of which can be directly controlled and some not (such as climate change). But once the availability and sustainable supply of water are more-or-less known, needs and interests can be renegotiated accordingly, which is precisely what happened in Cape Town, where more pressure was put on major water users in particular to conserve.Footnote 7
Other examples could relate to locally specific biodiversity and their needs vis-à-vis inhabited (urban or rural) landscapes. But all of these are still, in many ways, local or regional. Given the thoroughly global nature of today’s value and impact chains, it is critical to also map and study global connections so as to more directly highlight the political implications and biodiversity impacts of richer lifestyles. There are two ways to do this, both of which are already being explored in practice: first, to start from a specific and important ecosystem or species and “work up” toward the main actors or economic sectors that impact it; or, second, to “work down” from particular actors and economic sectors to show their cumulative impacts on different biodiversity and ecosystems. In what follows, we provide some first tentative examples of both, after which we wrap up the section by suggesting how we can take this concept forward as part of a broader move to operationalize the transformative governance of convivial conservation.
12.4.1 Working down the Biodiversity Impact Chain
Conservation areas and biodiversity are often – and rather self-evidently – said to be impacted mostly by “local people” aiming to fulfill their livelihood needs by utilizing surrounding natural resources. This is, among other factors, the basis of much of the “community-based conservation” literature (Reference Dressler, Büscher and SchoonDressler et al., 2010), as well as an explicit assumption of many elite actors involved in conservation. One example concerns famous Virgin billionaire-entrepreneur Richard Branson. In a video supporting conservation in Africa, he asks the question, “what is Africa?” and answers bluntly that “Africa is its animals. That is the beauty of Africa, that’s what makes it different from the rest of the world. And to lose those animals would be catastrophic.” Branson blames “dwindling wildlife numbers” on “Africa’s increasing (human) populations” and argues that Africa should “increase the amount of land for the animals and by increasing the amount of land for the animals, that will help human beings.”Footnote 8
Unfortunately, this neocolonial discourse is not uncommon when it comes to conservation in Africa (Reference Mbaria and OgadaMbaria and Ogada, 2017). Convivial conservation challenges colonizing discourses and practices by more clearly identifying the impacts of extra-local actors, and especially global elites who have the largest footprints. In the case of Branson, his environmental impacts are quite well-documented and provide a pertinent example. Branson, after all, owns several luxury game reserves around the world and has voiced some of the largest climate commitments of any elite actor. Together, these could constitute quite an environmental legacy were it not for the fact that scholars have thoroughly debunked these commitments. Naomi Reference KleinKlein (2015: 251–252), for example, argues that “Branson set out to harness the profit motive to solve the climate crisis – but the temptation to profit from practices worsening the crisis proved too great to resist. Again and again, the demands of building a successful empire trumped the climate imperative.” Scott Reference PrudhamPrudham (2009), similarly argued that Branson’s environmentalism did nothing to limit further capitalist expansion, including the resource extraction and use this entails. However, while these authors may show that Branson is far from an environmental hero, his precise impact on biodiversity is unclear and needs more research.
At the same time, this research also needs to be extended to aggregate sectors instead of (only) individuals. Our own research on the high-end tourism sector in South Africa provides a short example of how a BIC analysis could work by analyzing the impact of all four conservation classes (Table 12.1) on biodiversity. Adjacent to the world-famous Kruger National Park, philanthrocapitalists such as Richard Branson have their own residences on private protected lands (“upper class,” category 1), while lodge operators and large tourism companies own enormous tracts of private lands (“land-owning capitalist class,” category 2) for relatively wealthy tourists to enjoy (“upper class,” category 1 and “land-owning capitalist class,” category 2). Furthermore, some wealthy South Africans, Europeans and others own properties on so-called “wildlife estates,” sometimes as a permanent residence but often also as “second homes” (again categories 1 and 2, but also 3) (Reference Koot, Hitchcock and GressierKoot et al., 2019).
Meanwhile, the inequality between these classes and the “middle and lower classes” (category 3) and “lower rural classes” (category 4) remains enormous, and people from the latter two categories are often associated with causing most of the problems of conservation, including poaching (Reference Duffy, Massé and SmidtDuffy et al., 2019). However, these people also provide substantial “conservation labor” (needed for the first two class categories to enjoy nature) and, through the tourism industry, are increasing the value of private land, thereby reducing the chances of the middle and lower and lower rural classes to claim land for other purposes (Reference RamutsindelaRamutsindela, 2015; Reference SodikoffSodikoff, 2009), perpetuating and fortifying socioeconomic inequality. Despite a variety of such negative social and environmental consequences, the tourism industry often champions itself for its sustainable contribution to conservation (including much support for militarized anti-poaching conservation initiatives) and community development. However, initial research from several of this chapter’s authors suggests that tourism’s contributions are actually quite meager. More research is needed to accurately evaluate the impacts that all of the classes outlined here have on the national park and its biodiversity, and we posit that BICs as political methodology would enable such an analysis (see also Reference Mugo, Visseren-Hamakers and Van der DuimMugo et al., 2020).
12.4.2 Working up the Biodiversity Impact Chain
The other way to operationalize impact chains is to work “up” from specific biodiverse spaces, and document the direct and indirect pressures on these areas. Unlike the aforementioned top-down impact-chain mapping, this is an area where a lot of work is already being done. NGOs like Greenpeace, Friends of the Earth, the Rainforest Action Network and many others are well known not just for their (direct) actions but also for their research linking environmental impacts on specific areas to specific actors. The Rainforest Action Network, for example, published a report in 2017 tracking the impact chains on Southeast Asian rainforest, especially those in the Leuser Ecosystem in Sumatra, Indonesia (RAN, 2017). According to the report, it
profiles key environmental, social and governance (ESG) performance issues of 8 companies operating in Southeast Asia’s tropical forest-risk commodity sectors. The 8 companies profiled – Felda Global Ventures Holdings, Indofood Sukses Makmur, IOI Corporation, Wilmar International, Asia Pulp and Paper Group, Oji Holdings Corporation, Marubeni Corporation, and Itochu Corporation – were found to have had a range of serious ESG violations in their own operations or direct supply chains. These violations include: use of child and forced labour; conflicts with local communities over violations of their tenure rights; tropical deforestation and destruction of carbon-rich peatlands; threats to biodiversity; corruption; and illegality.
But the report doesn’t just highlight the responsibility of the companies directly involved in the destruction of biodiversity and other misdemeanors; it goes all the way up to specific institutional investors, which they argue are equally responsible for the impacts on biodiversity:
The forest-risk commodity sector operations of the 8 companies profiled in this report have been enabled by at least 6.38 billion USD in bond- and shareholdings at the most recent filing date in May 2017 by institutional investors (asset managers, insurance companies, pension funds) and have received more than 32.67 billion USD in loans and underwriting facilities between 2010 and 2016.
They then list the investors and bank and highlight that these “have both a moral and corporate responsibility, and a fiduciary duty to understand and address the harmful ESG impacts … which they are connected to” (RAN, 2017: 3).
This type of work is critical and puts the spotlight where it belongs: on the wealthy, often extra-local actors that have disproportionate (negative) impact on biodiversity. A similar “working up” approach was also recently applied by Amazon Watch to destruction of the Amazon and Cerrado biomes in Brazil, in their report entitled “Complicity in Destruction” (Amazon Watch, 2019). Home to 10 percent of the world’s biodiversity and 20 percent of its flowing freshwater, it is hard to imagine a convivial conservation transition without a concerted international effort to curb rapid deforestation and land conversion that has increased by more than 50 percent since 2016 (Amazon Watch, 2019). The report echoes research implicating soy and beef production for over 80 percent of forest land conversion in Brazilian Amazonia, and while noting the difficulty in following the exact trail to consumption destinations, it outlines clearly the global financial sources underwriting local and multinational companies implicated in the commodity chain. Among the largest creditors and equity investors in companies active in the Amazon and Cerrado, including those fined for illegal practices, were Barclays, Capital Group, BlackRock, Bank of America, Citigroup, JPMorgan Chase, BNP Paribas, Santander, HSBC, Credit Suisse, Vanguard, Morgan Stanley and Fidelity Investments (Amazon Watch, 2019: 19–24). Illegal timber supply chain links were also found with major importers in France, Belgium, the Netherlands, Denmark, the UK and the USA. Ultimately, Amazon Watch calls for a no-deforestation policy by global financiers, which are effectively underwriting the rapid decline of the world’s most biodiverse region, and sees scope for targeting EU and North American governments, given their accounting for 18.3 percent and 11 percent of Brazilian agricultural exports, respectively.
The importance of viewing the soy and beef industries together in this conservation impact chain is not incidental. Research has shown that despite the primary driver of Amazon deforestation by far being cattle production, this has occurred partly as a result of displacement of medium and smaller cattle ranchers from land now occupied by soy (Reference Barona, Ramankutty, Hyman and CoomesBarona et al., 2010). Perhaps even more salient has been Brazil’s efforts to “flex” its soy crop for animal feed processing and biofuel production in order to maintain a degree of domestic control – and significant revenues – as China monopolized Brazilian whole bean exports after 2008 (Reference Oliveira and SchneiderOliveira and Schneider, 2016). Maintaining a Brazilian soy-crushing and animal feed production capability effectively depends on constantly expanding domestic cattle production, or else losing out to global competition.
With China now crushing the bulk of Brazilian soy to make chicken, pig, salmon and cattle feed for markets worldwide, the “working up” of Amazonian biodiversity destruction simultaneously results in a “working down” to numerous examples of agro- and aqua-industrial pollution and ecosystem decline across worldwide cases from Norwegian salmon to Vietnamese shrimp, and beef industrial expansion across much of Asia. In Brazil itself, then, the conversion of some 200,000 square kilometers of highly biodiverse Cerrado forest and savanna for monocrop GM soy, with associated intensive pesticide use and seed consolidation by a tiny list of corporate players, has meant a wholesale collapse of pre-existing nature and agrarian livelihoods, while also enabling biodiversity destruction associated with agribusiness around the globe (Reference Oliveira and HechtOliveira and Hecht, 2016). Arguments that we need to continually expand food production to feed a growing population are quickly countered by deeply uneven global access, distribution and profiteering from corporate-led food systems that themselves increasingly depend on ecological catastrophe and the undermining of local food production in favor of export markets (Reference McMichaelMcMichael, 2014).
Finally, with regard to Amazonian biodiversity decline, “scaling up” also highlights the complicity of the global financial and market connections already identified in the rise of authoritarian government. The close association of extractivism with the new Latin American far right is well covered in the literature (Reference Arsel, Hogenboom and PellegriniArsel et al., 2016; Reference McCarthyMcCarthy, 2019; Reference Saad-Filho and BoffoSaad-Filho and Boffo, 2021), yet often understated are the simultaneous attacks on protected areas in the Amazon and elsewhere – especially those managed by Indigenous Peoples – that the expansion of mining and the cattle–soy nexus necessitates. The dismantling of Brazilian government ministries for Indigenous Peoples and the environment is effectively now preventing any regulation of Amazonian conservation. The same list of global financiers noted above thus profit from the authoritarian enforcement of biodiversity decline, a fact further highlighting the urgent need for institutional control of global finance.
More examples can be mentioned, but what is clear is that the transformation to convivial conservation would rely on a dramatic extension and normalization of such research and exposure endeavors. In doing so, the precise details of the impact-mapping in the above examples should be as important as the sociocultural and political-economic process that accompanies it. Again: we see this methodology and governance mechanism as a politicization tool that connects different actors from Table 1 in relation to how biodiversity is conserved or not. This political process can then further map the needs and interests of stakeholders in the short term, and also how these needs might change as the overall economy shifts toward degrowth, sharing the wealth and convivial conservation. In addition, the planning process could start to create awareness of how people in bioregions can contribute to degrowth and sharing of wealth. This is how an active process of shifting needs and interests (and hence, ultimately, human nature itself), and challenging the vested interests associated with the creation of capitalist needs and interests, might start or be further encouraged. Moreover, “impact chains” can never do justice to all the different types of impacts generated through activities, especially the complicated climate-related impacts. The point is therefore not to get one-on-one impacts “measured” precisely but rather to complicate, and politicize, the capitalist governance of biodiversity by incorporating direct and indirect pressures and by targeting and challenging these from two sides (bottom-up and top-down).
12.5 Conclusion
Along with climate change, inequality and, more recently, a global pandemic, biodiversity loss is considered to be one of the world’s most pressing challenges. As such, calls for transformative change in the ways biodiversity is governed and conserved are growing. However, major differences on how to approach transformative change exist, and some prominent responses to the biodiversity crisis that consider themselves transformative do not actually address underlying structural drivers of destruction. We therefore argue that these responses, including neoprotectionism and new conservation, should not be considered transformative in the way we have defined the term. Instead, and in line with a growing number of academics, social movements and civil society groups, we contend that fundamental structural transformation is needed to achieve the biodiversity and wider environmental governance capable of adequately addressing the growing biodiversity crisis. In this chapter we have built on the vision of convivial conservation, put forward as a necessary and realistic alternative – one that has fundamental structural transformation at its core.
We have also outlined a practice tool – biodiversity impact chains – as an example of a transformative governance mechanism that reframes perspectives on biodiversity conservation by politicizing the uneven relationships and impacts that different actors have with and on biodiversity. BICs can be seen as a tool for governance of transformations (Chapter 1) as they aim to steer the transformative change outlined in this chapter as part of the convivial conservation vision. Two characteristics of transformative governance highlighted in Chapter 1 are reiterated here as particularly important in relation to BICs. First, BICs are inclusive as they emphasize the interests of different actors and how such interests impact biodiversity. Second, BICs are integrative as they connect actions and solutions across scales. BICs also demonstrate the need for transformative governance to expand yet further and provide a mechanism through which the very framing of biodiversity and its conservation is politicized, challenged and disrupted. Local communities are still typically conceptualized as the recipients, or targets, of biodiversity governance interventions – even in cases where this governance is thought to be transformative. BICs support an alternative approach – one that could support policymakers in better targeting interventions in a more impactful and transformative way.
BICs are just one tool in the convivial conservation toolbox that we and other diverse actors are developing, and in line with other transformative movements such as degrowth. The convivial conservation vision, however, goes beyond the use of individual tools, and the focus, we argue, must be on broader “whole earth” transformation (Reference Büscher, Fletcher and BrockingtonBüscher et al., 2017). This requires what Reference WarkWark (2015) calls “alternative realism,” in contrast to “capitalist realism,” asserting that there is no viable alternative to the existing order – and a questioning of many of the assumptions that underpin conservation as we know it. This may seem impossible, but if, as Reference Olsson, Bodin, Folke, Armitage and PlummerOlsson et al. (2010: 280) argue, “transformational change is most likely to occur at times of crisis, when enough stakeholders agree that the current system is dysfunctional,” then this moment could be the opportunity to make the fundamental, structural changes that are needed.
13.1 Introduction
Agricultural land systems, covering about 40 percent of the world’s ice-free terrestrial surface, are the single largest contributor to biodiversity loss worldwide (Reference Chapin, Zavaleta and EvinerChapin et al., 2000; Reference Montanarella, Scholes and BrainichIPBES, 2018a; Reference Díaz, Settele, Brondízio and Ngo2019). Agricultural practices have been linked to staggering losses in critical ecosystems such as tropical forests and ecologically functional species such as pollinators, raising concerns of losing biodiversity as both an intrinsic global value and as a central pillar of food security and ecosystem functions (Reference Potts, Imperatriz-Fonseca and NgoIPBES, 2016; Reference Laurance, Sayer and CassmanLaurance et al. 2014; Reference Ramankutty, Mehrabi and WahaRamankutty et al., 2018). Conserving biodiversity in this sector is crucial beyond this intrinsic value (see Chapter 2), since biodiversity in agricultural landscapes supports ecosystem services that sustain human well-being through provisioning services such as food production, regulating services including flood and climate control or stabilization, and supporting services such as pollination and soil fertility (Reference Potts, Imperatriz-Fonseca and NgoIPBES, 2016; Reference Rounsevell, Fischer, Torre-Marin Rando and Mader2018b; Reference Díaz, Settele, Brondízio and Ngo2019; Reference Scherr and McNeelyScherr and McNeely, 2008; Reference Tscharntke, Clough and WangerTscharntke et al., 2012). There are a wide range of approaches proven to enhance synergies and reduce conflicts between biodiversity, food production and livelihood objectives, such as agroecology, permaculture, organic agriculture, agroforestry and “nature-inclusive” agriculture (Reference Bouwma, Zinngrebe, Runhaar, Dries, Heijman, Jongeneel, Purnhagen and WesselerBouwma et al., 2019; Reference Chapin, Zavaleta and EvinerChapin et al., 2000; Reference Chappell and LaValleChappell and LaValle, 2011; Reference RunhaarRunhaar, 2017; Reference Scherr and McNeelyScherr and McNeely, 2008). Climate change, the projected rise in global food demand and changing diets are projected to further increase pressures on food systems and land use (FAO, 2017a). The challenge for transformational policies is to disincentivize unsustainable practices while incentivizing biodiversity-friendly food production approaches. While healthy diets (Chapter 5) and animal welfare (Chapter 9) are also fundamental components of future food systems, this chapter focuses on governance of agricultural land use.
Conserving and enhancing biodiversity in agriculture is central to some of the most prominent international environmental agreements and conventions. The Convention on Biological Diversity (CBD) aims to ensure sustainable management and biodiversity conservation (Aichi Target 7 of the 2011–2020 Strategic Plan) and keep resource extraction within sustainable limits (Aichi Target 4). The impending Post-2020 CBD Global Biodiversity Framework (GBF), which is expected to be approved in 2022, is also expected to reflect the importance of sustainable agriculture. The importance of agricultural biodiversity has been reconfirmed by the 2015 United Nations Sustainable Development Goals (SDGs), particularly SDG15 (Life on Land), SDG2 (Zero Hunger) and SDG8 (Sustainable Production and Consumption). In 2017, the UN Framework Convention on Climate Change also initiated a work stream aiming to promote sustainable agricultural systems (UNFCCC, 2017).
Within these international conventions, as well as in national-level governance frameworks, an increasingly important way to promote biodiversity conservation in agricultural landscapes is through the mainstreaming of biodiversityFootnote 1 into public and private governance of the agricultural sector, a strategy that was specifically advocated in the CBD’s 2011–2020 Strategic Plan. This chapter analyzes the progress in mainstreaming biodiversity into public and private sector agricultural policies worldwide by employing the concept of biodiversity policy integration (BPI). BPI analyzes the consideration of biodiversity in all sectors and levels of policymaking and implementation, providing a conceptual approach to identify leverage points for transformative change. In this chapter, we analyze BPI in agricultural landscapes, which adds to the toolbox of the transformative biodiversity governance framework. We review available literature on BPI in agricultural policies in developed countries (with a focus on the European Union [EU]) and developing countries (with a focus on tropical countries). Recognizing the important role of nonstate actors in biodiversity governance, we also include private sector governance in our analysis, defined here as rules and standards developed and monitored by firms or nongovernmental organizations (Reference Grabs, Auld and CashoreGrabs et al., 2020).
This chapter proceeds as follows. We first provide an overview of trends and threats to biodiversity, highlighting the necessity to integrate biodiversity in the governance and management of agricultural landscapes (Section 13.2). We then introduce our analytical approach (BPI) and how it relates to the broader literature on environmental policy integration and mainstreaming (Section 13.3), before analyzing to what extent and how biodiversity is integrated into agricultural governance in developed and developing countries (Section 13.4). Based on these analyses, we discuss four central leverage points for transformative biodiversity governance in agricultural landscapes and reflect them with the analytical dimensions of this book (Section 13.5), before concluding with key lessons (Section 13.6).
13.2 Current Trends and Key Threats to Biodiversity
This section focuses on two principal mechanisms through which agriculture impacts biodiversity: land use change for agricultural expansion and management choices on agricultural land – that is, intensification, specialization and enlargement of farms (Reference Ramankutty, Mehrabi and WahaRamankutty et al., 2018). After introducing these issues within the broader contemporary debate, we discuss central arguments for segregated (“land-sparing”) versus integrated (“land-sharing”) approaches.
13.2.1 Land Use Change
Land use change for the production of feed, fuel, biofuels and livestock is one of the major drivers of biodiversity loss (Reference Díaz, Settele, Brondízio and NgoIPBES, 2019; MEA, 2005). Between 2000 and 2010, 80 percent of deforestation worldwide was directly attributable to the agricultural sector (Reference Hosonuma, Herold and de SyHosonuma et al., 2012). Agriculture currently occupies 38 percent of the world’s terrestrial land surface, with about 12 percent devoted to crops and about 25 percent to livestock rearing and grazing (Reference Foley, Ramankutty and BraumanFoley et al., 2011). Of the area used for cereal production, 31 percent is devoted to animal feed (Reference Mottet, de Haan and FalcucciMottet et al., 2017). Although land clearing has slowed since the 1950s relative to the previous century in temperate latitudes, it has shifted to tropical highly biodiverse forests in Latin America, Southeast Asia and Africa (Reference Díaz, Settele, Brondízio and NgoIPBES, 2019; Reference Ramankutty, Mehrabi and WahaRamankutty et al., 2018). In addition to loss of ecosystems and their intrinsic value, deforestation of biodiverse, tropical forests reduces carbon sinks, which are important for mitigating climate change (Reference Bunker, DeClerck and BradfordBunker et al., 2005; Reference Pachauri and MeyerIPCC, 2014).
The causes of agricultural expansion into intact ecosystems differ by region. In Africa, subsistence and small-scale farming drives the majority of expansion and deforestation (Reference Díaz, Settele, Brondízio and NgoIPBES, 2019; Reference Seymour and HarrisSeymour and Harris, 2019). In contrast, deforestation in South America (particularly in the Amazon) and Southeast Asia is primarily driven by commercial agriculture supplying international markets, most notably since the 1990s (Reference Hosonuma, Herold and de SyHosonuma et al., 2012; Reference Díaz, Settele, Brondízio and NgoIPBES, 2019; Reference Seymour and HarrisSeymour and Harris, 2019). Though the majority of agricultural commodities are consumed domestically, global trade of a select few agricultural commodities – notably soybeans (of which the majority is used for animal feed globally), beef and palm oil – is a major external driver of ecosystem loss (Reference DeFries, Herold, Verchot, Macedo and ShimabukuroDeFries et al., 2013; Reference Green, Croft and DuránGreen et al., 2019; Reference Henders, Persson and KastnerHenders et al., 2015; Reference Meyfroidt, Lambin, Erb and HertelMeyfroidt et al., 2013). As a prominent example, oil palm plantations supplying global markets have been responsible for over 80 percent of agricultural land expansion in South Asia since the 1990s (Reference Gibbs, Ruesch and AchardGibbs et al., 2010). Countries that consume these commodities are thus contributing to ecosystem and biodiversity loss, as recognized in recent attempts to reduce “imported deforestation” (Reference Bager, Persson and dos ReisBager et al., 2021). The long-term effects of land use change are often underestimated as – particularly in biodiversity-rich regions – species continue to be lost even if the agricultural land has been abandoned (Reference Gibson, Lee and KohGibson et al., 2011).
13.2.2 Management Choices
Agriculture has undergone significant structural changes since the Second World War. New farming practices falling under the paradigm of “industrial agriculture” were strongly subsidized by governments, particularly in developed countries and in some developing countries, as part of the “Green Revolution.” This “agricultural modernization” relied heavily on mechanization, genetic alterations of crops (e.g. hybridization, genetically modified organisms) and the use of chemical inputs to increase productivity (Reference Bosc and BélièresBosc and Belières, 2015; Reference Duru, Therond and FaresDuru et al., 2015). Three overarching and interrelated trends can be distinguished: intensification, specialization and scale enlargement (Reference Aubert, Schwoob and PouxAubert et al., 2019; Reference Poux and AubertPoux and Aubert, 2018).
Intensification refers to increasing productivity on a given parcel of land through the heavy use of inputs (such as pesticides and fertilizers). Though this may increase profits, and in some cases also food security, it generally drives biodiversity loss as it is currently practiced (Reference Batáry, Gallé and RieschBatáry et al., 2017; Reference Hendershot, Smith and AndersonHendershot et al., 2020; Reference Rasmussen, Coolsaet and MartinRasmussen et al., 2018). Studies point to the detrimental impacts on biodiversity in general, and on soil biodiversity and insects in particular, especially through mechanization and pesticide use (see, for example, Reference Orgiazzi, Panagos and YiginiOrgiazzi et al., 2016; Reference Sanchez‐Azofeifa, Pfaff, Robalino and BoomhowerSanchez-Bayo and Wyckhuys, 2019; Reference Seibold, Gossner and SimonsSeibold et al., 2019; Reference Tsiafouli, Thébault and SgardelisTsiafouli et al., 2015). Globally, pesticide sales and use continue to increase, with hundreds of older generation pesticides that are highly toxic to vertebrates and invertebrates still being used in developing countries, although banned in many developed countries (Reference Schreinemachers and TipraqsaSchreinemachers and Tipraqsa, 2012). Through run-off, pesticides and fertilizers also have biodiversity impacts reaching far beyond the farm (Reference Beketov, Kefford, Schäfer and LiessBeketov et al., 2013; Reference Van Dijk, Van Staalduinen and Van der SluijsVan Dijk et al., 2013; Reference Yamamuro, Komuro and KamiyaYamamuro et al., 2019). Solutions related to increasing efficiency, such as precision agriculture, can contribute to sustainability and food security through the reduction of inputs (Reference Shukla, Skea and Calvo BuendiaIPCC, 2019). However, recent work shows that implementation remains a problem (Reference Lindblom, Lundström, Ljung and JonssonLindblom et al., 2017). Moreover, such solutions do not address many of the underlying problems of conventional intensification, including the need for energy-intensive inputs (Reference KremenKremen, 2015).
Secondly, specialization describes a shift away from diversified crop production to monocultures and a separation of crops and livestock systems. At the macro level, specialization is driven by the logic of economies of scale and the creation of regional or national comparative advantages in trade (Reference Abson, Lemaire, Kronberg, Recous and de Faccio CarvalhoAbson, 2019). As a prominent example, Brazil has developed a significant comparative advantage in soybean production by using soybeans as a “flex crop” with multiple processing pathways that differentiate the product into a food grain, livestock feed or fuel (Reference OliveiraOliveira, 2016). However, these regional advantages come at a cost – extreme specialization of food and agriculture is a major driver of the decline in biodiversity at genetic, species and ecosystem levels (Reference Bélanger and PillingFAO, 2019; Reference Díaz, Settele, Brondízio and NgoIPBES, 2019). While agronomic research and technical expertise have focused on the production of a few key staple crops (wheat, corn and rice initially, now followed by oilseeds, e.g. soybeans and rapeseed), technical knowledge on other crops remains low (Reference Bélanger and PillingFAO, 2019; Reference Magrini, Anton and CholezMagrini et al., 2016). Furthermore, specialization conflicts with the idea of multifunctional production and its potential for contributing to food security (Reference Bommarco, Vico and HallinBommarco et al., 2018; Reference Misselhorn, Aggarwal and EricksenMisselhorn et al., 2012), climate-smart landscapes (Reference Scherr, Shames and FriedmanScherr et al., 2012) and viable farming income, despite potential trade-offs in efficiency (Reference Lakner, Kirchweger, Hoop, Brümmer and KantelhardtLakner et al., 2018).
Lastly, scale enlargement entails a trend toward fewer but larger farms. Although there is still a wide variety of farm types and sizes around the world, a productivist ideology has led farms to increase in size overall in order to benefit from economies of scale, which enables cost reductions and helps farmers remain competitive (Reference DuffyDuffy, 2009). This strategy is capital- and input-intensive, requiring high investments in machinery and chemical inputs that are only considered worthwhile if farm output is high, lowering costs per unit of production (Reference McIntyre, Herren, Wakhungu and WatsonMcIntyre et al., 2009). Concentration across the agri-food industry, and the resulting control exerted by a small number of companies on farmers, has further encouraged a consolidation and enlargement trend (Reference Folke, Österblom and JouffrayFolke et al., 2019; IPES-Food, 2017). Scale enlargement contributes to biodiversity loss principally through the destruction of seminatural landscape features, such as hedges, field margins and permanent prairies, which maintain heterogeneity and connectivity of habitats at the landscape level (Reference Poux and AubertPoux and Aubert, 2018; Reference Tscharntke, Clough and WangerTscharntke et al., 2012).
13.2.3 Land-Sharing and Land-Sparing in a Telecoupled World
For many decades, the dominant global discourse on food security has resulted in the notion that there is direct competition for land between biodiversity conservation and agricultural production and that the two are incompatible (Reference Butler, Vickery and NorrisButler et al., 2007; Reference Henle, Alard and ClitherowHenle et al., 2008; Reference Steffan-Dewenter, Kessler and BarkmannSteffan-Dewenter et al., 2007; Reference Tscharntke, Clough and WangerTscharntke et al., 2012). This has led to a simplified framing in which “land-sparing” (segregating intensive agriculture from conservation lands) and “land-sharing” (more extensive agriculture that contributes to conservation) are viewed as a dichotomy, though neither of them singularly has the full potential to address the challenge of sustainable agriculture (Reference KremenKremen, 2015). Instead, we argue that a combined approach of both large, protected regions and wildlife-friendly farming areas is critical to conserving biodiversity (Reference KremenKremen, 2015; Reference Kremen and MerenlenderKremen and Merenlender, 2018).
The land-sparing logic argues that effective biodiversity conservation on nonagricultural land (see Chapter 11) depends on the separation of agricultural land from protected areas, necessitating the intensification of production on agricultural land to “free up” land for conservation. However, since the effectiveness of protected areas correlates with the pressures from its surroundings (Reference Kremen and MerenlenderKremen and Merenlender, 2018; Reference Watson, Dudley, Segan and HockingsWatson et al., 2014), conservation in these designated areas will still depend on the management of external or internal pressures. Therefore, the idea of completely separating the interactions between biodiversity conservation and agricultural production areas is conceptually flawed, as landscape structures are shaped by cultural dynamics and human–nature interactions, as well as geographical and climate conditions, making ecological and productive systems mutually interdependent (Reference Fischer, Batáry and BawaFischer et al., 2011; Reference Fischer, Abson and Butsic2014). In addition to localized detrimental impacts of intensive farming, the land-sparing approach can also have far-reaching impacts on biodiversity: Land-sparing in one area can have spill-over effects that drive relocation and expansion of production in other regions, rather than leading to an overall reduction of biodiversity threats (Reference MeyfroidtMeyfroidt, 2018; Reference Meyfroidt, Lambin, Erb and HertelMeyfroidt et al., 2013; Reference Rudel, Schneider and UriarteRudel et al., 2009). Even in regions where the extension of agricultural land use remains relatively constant (such as within the EU), the “imported land” needed to satisfy consumer demand continues to grow (Reference Asici and AcarAsici and Acar, 2016, Reference Teixidó-Figueras and DuroTeixidó-Figueras and Duro, 2014; Reference Yu, Feng and HubacekYu et al., 2013). This shows that consumption decisions and agricultural management in a globalizing world are “telecoupled” (Reference Friis, Nielsen and OteroFriies et al. 2016; Reference Sun, Tong and LiuSun et al., 2017). Therefore, while protected areas remain crucial to maintaining biodiversity, the land-sparing approach requires policy integration.
In contrast, land-sharing recognizes agriculture as “both the greatest cause of biodiversity loss and the greatest opportunity for conservation” (Reference Hendershot, Smith and AndersonHendershot et al, 2020: 393, emphasis added). Land-sharing approaches recognize the need and potential for agricultural land to help protect biodiversity through a range of practices, as agricultural expansion and its (inadequate) management drive biodiversity loss. While this is a good idea in theory, the above-described trajectories show that land conversion and management choices continue to invade important ecosystems and fail to produce sound ecological structures. At the same time, the separation of sufficiently large areas seems necessary for the conservation of certain ecosystem values and habitats (Reference Kremen and MerenlenderKremen and Merenlender, 2018; Reference Watson, Dudley, Segan and HockingsWatson et al., 2014).
Hence, while a conceptual separation of land-sparing and land-sharing can help to identify socio-ecological trade-offs, it has largely failed in identifying solutions for addressing them (Reference Fischer, Abson and ButsicFischer et al., 2014). We argue that in transformative biodiversity governance, area-based (land-sparing) and integrated (land-sharing) approaches offer a complementary toolkit to address direct and indirect drivers of biodiversity loss in agricultural landscapes, and that biodiversity policy integration is crucial in both of these approaches.
13.3 Conceptual Framework for Biodiversity Policy Integration
Biodiversity policy integration (BPI) is an analytical tool derived from the broader literature of environmental policy integration (EPI) (Reference ZinngrebeZinngrebe, 2018). EPI can be defined as “the incorporation of environmental objectives in non-environmental policy sectors such as agriculture, energy and transport” and can be considered transformative because of its “aim to target the underlying driving forces, rather than merely symptoms, of environmental degradation” (Reference Persson, Runhaar and Karlsson-VinkhuyzenPersson et al., 2018: 113). Governance elements and processes that support EPI have been widely studied, particularly in European and OECD countries (see e.g. Reference Jordan and LenschowJordan and Lenschow, 2010; OECD, 2018; Reference Persson, Runhaar and Karlsson-VinkhuyzenPersson et al., 2018; Reference RunhaarRunhaar, 2016; Reference Runhaar, Driessen and UittenbroekRunhaar et al., 2014; Reference Runhaar, Wilk, Persson, Uittenbroek and Wamsler2018; Reference Runhaar, Wilk, Driessen, Biermann and Kim2020, Reference Visseren-HamakersVisseren-Hamakers, 2015). This literature shows that no single instrument can realize policy integration, but rather, EPI needs a suite of complementary instruments and mechanisms (Reference Persson and RunhaarPersson and Runhaar, 2018; Reference Runhaar, Wilk, Driessen, Biermann and KimRunhaar et al., 2020).
In this chapter, we use BPI as an analytical tool deriving from EPI literature, with a focus on biodiversity (Reference ZinngrebeZinngrebe, 2018). To date, empirical analyses of policy integration between agriculture and biodiversity are scarce. A Web of Science search for the terms “agriculture” AND “policy integration” AND “biodiversity” resulted in six articles, all of which are included in the analysis in this chapter (Reference Karlsson-Vinkhuyzen, Kok, Visseren-Hamakers and TermeerKarlsson-Vinkhuyzen et al., 2017; Reference Karlsson-Vinkhuyzen, Boelee and Cools2018; Reference Söderberg and EckerbergSöderberg and Eckerberg, 2013; Reference Somorin, Visseren-Hamakers, Arts, Tiani and SonwaSomorin et al., 2016; Reference ZinngrebeZinngrebe, 2018, Reference Zinngrebe, Pe’er and SchuelerZinngrebe et al., 2017). Other combinations of search terms were also explored: “biodiversity” OR “mainstreaming biodiversity” AND “production landscapes,” “agricultural policy,” “coherence,” “inclusion,” “social capital” and “capacity.” These also returned few hits of direct relevance that included concrete examples. Reference Redford, Huntley and RoeRedford et al. (2015) note that publications by practitioners involved in public and private biodiversity mainstreaming programs and projects are severely deficient in the peer-reviewed literature, particularly those focused on developing countries. Therefore, to capture relevant gray literature, we also applied the following Google searches. “mainstreaming biodiversity” AND “production landscapes” (yields sixty-seven results) and “mainstreaming biodiversity” AND “agricultural policy” (yields ninety results). Titles and abstracts were screened to select relevant publications.
In order to analyze the extent to which biodiversity considerations have been incorporated in agricultural policies, we distinguish five dimensions of BPI (see Figure 13.1) (Zinngrebe et al., 2018; for similar approaches see Reference Kivimaa and MickwitzKivimaa and Mickwitz, 2006 and Reference Uittenbroek, Janssen-Jansen and RunhaarUittenbroek et al., 2013):
1. Inclusion: the extent to which the objective of biodiversity conservation is included in political sectors. This is measured by the extent to which a sector has reframed a biodiversity objective into sector-specific targets and specific biodiversity indicators.
2. Operationalization: the extent to which a sector has adopted or adjusted policy instruments and monitoring and enforcement mechanisms to implement biodiversity objectives (see also Reference RunhaarRunhaar, 2016), and the uptake of biodiversity values in internal evaluation processes.
3. Coherence: the extent to which objectives and policy instruments within a sector complement rather than contradict each other. This is measured by the extent to which policies within a sector are internally consistent and direct sector activities toward biodiversity objectives.
4. Capacity: the level of institutional development, available resources and political mechanisms that ensure the implementation of instruments identified in the “operationalization” dimension, as well as the extent to which other actors are supported by their organization (“social capital”) (Reference Zinngrebe, Borasino and ChiputwaZinngrebe et al., 2020).
5. Weighting: the importance given to biodiversity objectives in relation to other political objectives. Weighting further analyzes whether biodiversity, as natural capital, is regarded as substitutable by other forms of capital and whether ecological limits are recognized.
In the next section, we use this analytical framework to analyze the current state of BPI in agricultural governance along the five dimensions. However, we note that while the BPI framework assesses the level of integration at a specific point in time, transformative governance is adaptive, requiring dynamic policy design and institutional reconfigurations to iteratively improve BPI performance. In Section 5, we draw on our BPI analysis to reflect on enabling factors and barriers and discuss them in relation to the transformative governance analytical framework of this book.
Figure 13.1 Five dimensions of biodiversity policy integration.
13.4 Taking Stock: Assessing the Level of Biodiversity Policy Integration in Agricultural Governance
13.4.1 Inclusion
In many developing countries with available studies, biodiversity is not an explicit target in agricultural policies (Reference ZinngrebeZinngrebe, 2018; Reference Zinngrebe, Borasino and ChiputwaZinngrebe et al., 2020). While most Parties to the CBD identify the need for both ex-situ and in-situ biodiversity conservation, only 3 percent have mainstreamed biodiversity in their agricultural policies, plans and programs (Reference Lapena, Halewood and HunterLapena et al., 2016). Among the exceptions is Kenya, where the Ministry of Agriculture in Busia County has set a performance target for establishing a biodiversity policy (Reference Hunter, Borelli, Olsen Lauridsen, Gee and NodarHunter et al., 2018). Similarly, Costa Rica has a biodiversity law setting general standards (although in rather generic terms) to also be considered in agricultural landscapes, which has been regarded as “one of the most comprehensive efforts to implement … the Convention on Biological Diversity” (Reference MillerMiller, 2006: 359). Despite few government-led policy initiatives to advance BPI in developing countries, international organizations have been active in pushing for integrated instruments and planning procedures, which we include in the following sections.
In the EU, various policies have aimed to integrate biodiversity objectives into the agricultural sector to differing degrees. Most recently, the European Green Deal includes a “Farm to Fork” strategy that explicitly aims to reverse biodiversity loss by aiming for a “neutral or positive impact” within agri-food systems (EC, 2019; 2020a). As an additional element, the EU Biodiversity Strategy for 2030 includes area-based targets aimed at protecting 30 percent of its terrestrial area, with “at least 10 percent of utilized agricultural area under high diversity landscapes,” and a life-cycle assessment assuming responsibility for outsourced environmental impacts as well as a reduction of the overall EU’s global footprint (EC, 2020b, section 2.2.2). The key legal instruments underpinning the EU’s conservation policies date back several decades: the Birds and Habitats Directives established the Natura 2000 network, which covers almost 18 percent of the EU’s terrestrial surface area (Reference Bouwma, Zinngrebe, Runhaar, Dries, Heijman, Jongeneel, Purnhagen and WesselerBouwma et al., 2019). Almost 90 percent of all Natura 2000 sites are subject to agriculture or forestry activities, making BPI highly relevant (Reference Tsiafouli, Apostolopoulou and MazarisTsiafouli et al., 2013). The Habitats and Birds Directives do not, however, include targets or indicators related to land use systems or ecosystem services. Instead, they have the objective of maintaining healthy habitats for selected species (Reference Bouwma, Zinngrebe, Runhaar, Dries, Heijman, Jongeneel, Purnhagen and WesselerBouwma et al., 2019). Similarly, the European Common Agricultural Policy (CAP) speaks more generally of “sustainable management of natural resources and climate action” in the 2013–2020 period and uses a farmland bird index and High Nature Value farmland index as proxies for biodiversity (EC, 2013). Since 2018, a proposal by the European Commission that includes a strategic objective on the protection of biodiversity, enhancement of ecosystem services and preservation of habitats and landscapes (Target F, EC, 2018) has been negotiated by EU institutions. While this proposal takes a comprehensive approach to envisioning sustainability in agriculture, the proposed indicators target farm management and land use in general and have been assessed as insufficient for monitoring biodiversity (Reference Pe’er, Bonn and BruelheidePe’er et al., 2020).
Overall, countries face challenges in translating international biodiversity targets into nationally determined targets (Reference Chandra and IdrisovaChandra and Idrisova, 2011; Reference Velázquez GomarVelázquez Gomar, 2014). In an analysis of 144 national biodiversity strategies and action plans (NBSAPs) developed by countries that signed the CBD, 72 percent of developing countries and 58 percent of developed countries acknowledge agriculture explicitly as a threat to biodiversity conservation (Reference Whitehorn, Navarro and SchröterWhitehorn et al., 2019). Despite this, only 23 percent of the developing and 33 percent of the developed countries address the question of trade-offs between agriculture and conservation (Reference Whitehorn, Navarro and SchröterWhitehorn et al., 2019). More tellingly, almost no national agricultural plan cross-references the countries’ NBSAPs (Reference Pe’er, Zinngrebe and MoreiraPe’er et al., 2019; Reference ZinngrebeZinngrebe, 2018). This means that although these NBSAPs may be well developed by environmental ministries and include agriculture-related targets, these goals do not reach the actors they need to engage, such as agricultural ministries and the network of actors in the agricultural sector. In some agricultural policies, the need for considering “sustainability,” the “environment” or certain land use practices are mentioned, but without linking it to specific ecological criteria or policy instruments (Reference ZinngrebeZinngrebe, 2018).
13.4.2 Operationalization
The operationalization of biodiversity-related objectives into policies differs strongly between developing and developed countries. In many developing countries, operationalization of policy instruments is poorly executed (e.g. Reference Carew-ReidCarew-Reid, 2002; Reference HuntleyHuntley, 2014); regulatory frameworks are weak, poorly implemented or nonexistent (Reference HuntleyHuntley, 2014) and some countries have started to develop their environmental governance framework only in the past decade (e.g. Reference VijgeVijge, 2018). Nevertheless, some advancement in operationalization is visible, particularly in Latin America, including Costa Rica, Mexico, South Africa, Australia and Brazil (Reference Harvey, Komar and ChazdonHarvey et al., 2008; Reference HuntleyHuntley, 2014; Reference Somarriba, Beer, Alegre-Orihuela, Nair and GarritySomarriba et al., 2012).
Costa Rica made significant advancements in the institutionalization of payment for ecosystem services schemes, aimed at enhancing forest biodiversity on agricultural land (Reference Sanchez‐Azofeifa, Pfaff, Robalino and BoomhowerSanchez‐Azofeifa et al., 2007). However, these payment schemes are regarded as insufficiently funded in the long-term and to complement but not substitute regulatory interventions by governments (Reference Schomers and MatzdorfSchomers and Matzdorf, 2013; Reference Wunder, Engel and PagiolaWunder et al., 2008). In South Africa, the national Biodiversity Act sets bioregional plans, biodiversity assessments and biodiversity action plans as legal instruments for BPI operationalization at the regional spatial scale (Reference Botts, Skowno and DriverBotts et al., 2020). Additionally, “conservation farming” is supported by stringent regulation, involvement of nongovernmental organizations and farmer communities, effective communication with farmers and scientific and technical support for farmers (Reference Donaldson, Pierce, Cowlings, Sandwith and MacKinnonDonaldson, 2012). In Brazil, operationalization focuses on specific tools such as national plans promoting agroecology and organic production (Biodiversity International 2016), an “agrobiodiversity index” assessing private sector performance (Reference Tutwiler, Bailey, Attwood, Remans and RamirezTutwiler et al., 2017) and a national school food program mandating 30 percent of federal funds toward procurement from family farms using agroecological production approaches (Reference Johns, Powell, Maundu and EyzaguirreJohns et al., 2013).
In the private sector, producers and companies have started responding to the demand for deforestation-free commodities. Initiatives such as the Consumer Goods Forum, Tropical Forest Alliance, the New York Declaration on Forests, the Amsterdam Declaration Partnership, various beef and soy moratoriums and voluntary commitments under the Business for Nature coalition are, however, nonbinding and coexist with nonsustainable policies (Reference Stabile, Guimarães and SilvaStabile et al., 2020).
In Europe, the main biodiversity-related instruments of the 2014–2020 CAP are direct subsidies to farmers conditioned on fulfilling “greening” obligations (Ecological Focus Areas) and cross compliance, as well as voluntary agri-environmental and climate measures (AECMs). These specific “deep green measures” have been found to produce strong local impacts (Reference Batáry, Dicks, Kleijn and SutherlandBatáry et al., 2015; Reference Pe’er, Lakner and MüllerPe’er et al., 2017). However, the weak performance of “greening” (Reference Pe’er, Zinngrebe and HauckPe’er et al., 2016) and the low allocation of funding to AECMs are central arguments for identifying the CAP’s toolbox as weak “green architecture” (Reference Pe’er, Zinngrebe and MoreiraPe’er et al., 2019). The new Post-2020 CAP proposal will continue to link direct payments to weak, unspecific targets (similar to cross compliance), while allowing for EU member states to use voluntary “eco-schemes” to support specific landscape features (Reference Pe’er, Bonn and BruelheidePe’er et al., 2020). Simultaneously, area-based instruments linked to the EU Birds and Habitats Directives are being used. However, evaluations of Natura 2000 indicate that only about a third of the sites have developed specific management plans for biodiversity conservation and only 4 percent show an improvement of habitats (Reference Bouwma, Zinngrebe, Runhaar, Dries, Heijman, Jongeneel, Purnhagen and WesselerBouwma et al., 2019; EEA, 2015). Literature suggests that effective implementation of Natura 2000 sites depends on a joint implementation with policies such as agri-environmental measures (Reference Bouwma, Zinngrebe, Runhaar, Dries, Heijman, Jongeneel, Purnhagen and WesselerBouwma et al., 2019; Reference Lakner, Zinngrebe and KoemleLakner et al., 2020).
13.4.3 Coherence
Even in cases where conservation is included as one of the targets in agricultural policies, and when policies have been appropriately reconfigured to achieve those targets, they may still run counter to specific biodiversity conservation policies in the environmental sector. Often, decisions about trade-offs between productivity and conservation are avoided or not explicitly addressed, and a patchwork of incoherent policies result in a lack of incentives for biodiversity-friendly farming.
One barrier to coherent agri-environmental policies is a lack of horizontal coherence, notably, a lack of coordination between ministries and agencies at the national level. Insights from Indonesia, Uganda, Peru and Honduras show that while different regulatory processes for agricultural landscapes exist for the governmental sphere and for sustainability markets in the private sector, they are incoherent and generally favor conventional practices, rather than biodiversity-sound management systems such as agroforestry (Reference Zinngrebe, Borasino and ChiputwaZinngrebe et al., 2020). Even in Costa Rica, which has relatively strong environmental laws and regulations, incoherent policies have been reported (Reference Brockett and GottfriedBrockett and Gottfried, 2002; Reference LansingLansing, 2014). One general issue is that ministries of finance and planning – which generally hold decision-making power on large-scale investment allocations – are often not in regular consultation with the ministries responsible for biodiversity governance (Reference SwiderskaSwiderska, 2002).
Besides a lack of horizontal coherence (i.e. between sectoral policies at one level of governance) there is also often a lack of vertical coherence (i.e. between national and subnational biodiversity strategies). Vertical coherence is especially pertinent in developing countries, since many are in the process of decentralizing their governance systems (Reference Carew-ReidCarew-Reid, 2002; Reference Hunter, Özkan and Moura de Oliveira BeltrameHunter et al., 2016; Reference SwiderskaSwiderska, 2002). The few existing studies indicate that vertical integration across political levels for the implementation, enforcement and monitoring of biodiversity conservation in agricultural landscapes is generally low (e.g. Reference ZinngrebeZinngrebe, 2018). Nevertheless, the example of local stakeholder networks in Ethiopia illustrated that despite low coherence at the national level, local collaboration can lead to coherent management approaches (Reference Jiren, Bergsten and DorresteijnJiren et al., 2018). In Rwanda, the successes of watershed management plans in enabling dialogue and policy coordination across ministries of agriculture, fisheries and rural and social development at both local and national levels are another promising exception (FAO, 2017b). Based on selected case studies from countries within Africa and Latin America, the FAO (2017b) highlights that management models that take an ecosystem-based approach can serve as a lever for coordination, integration and synergies, though this has not been sufficiently applied to improve coherence. In South Africa for instance, bioregional plans enhance both coherence in local land use planning and across core sectoral strategies at the national level (Reference Botts, Skowno and DriverBotts et al., 2020). Deliberations in trade-off options between conservation and other goals is part of the planning process for this purpose (Reference Redford, Huntley and RoeRedford et al., 2015). The international Biodiversity for Food and Nutrition Project, funded by the Global Environment Facility, shows how, in Brazil, Kenya, Turkey and Sri Lanka, a sound evidence-base on how biodiversity supports nutritional outcomes, and the establishment of multistakeholder and multisectoral steering committees, improves coherence across agriculture and food policies (Reference Beltrame, Oliveira and BorelliBeltrame et al., 2016; Reference Beltrame, Eliot and Güner2019).
The EU is a strong advocate of policy coherence across sectors, as acknowledged in a large number of official EU documents. However, while most EU policies are coherent at the level of objectives, they provide incoherent incentives at the implementation stage, and therefore have not managed to effectively or efficiently reverse declining biodiversity trends (Reference Pe’er, Lakner and MüllerPe’er et al., 2017). For example, while the EU Birds and Habitats Directives aim to conserve biodiversity, the CAP’s fundamental targets, defined by the Treaty of Rome in 1957, direct agricultural policy toward increased productivity, low food prices and supporting farmers’ incomes. Another example of incoherence in the CAP is the aforementioned Ecological Focus Areas, which obligates each farm of more than fifteen hectares to dedicate 5 percent of its land to conservation activities. In reality, this instrument primarily results in measures with a low contribution to biodiversity, such as catch crops and nitrogen-fixing crops (Reference Cole, Kleijn and DicksCole et al., 2020; Reference Pe’er, Lakner and MüllerPe’er et al., 2017). Watering down ecological standards in federal implementation processes, as well as misconceptions about farmers’ motivations to engage in biodiversity conservation, reduce the CAP’s potential to contribute to conservation (Reference Brown, Kovács and HerzonBrown et al., 2020). In the EU proposal for a post-2020 CAP (EC, 2018), direct payments will continue to dominate and low ecological targets continue to persist (Reference Pe’er, Bonn and BruelheidePe’er et al., 2020). Overall, studies show that despite the EU’s rhetoric for policy coherence, large inconsistencies in the instruments and implementation of EU policies remain (Reference De Schutter, Jacobs and ClémentDe Schutter et al., 2020; Reference Nilsson, Zamparutti and PetersenNilsson et al., 2012).
Within the EU, there are also strong calls for enhancing coherence of EU policies with non-aid policies that impact developing countries. These calls have grown since the 1990s, when Europe’s need for agricultural biodiversity and production land substantially increased and was therefore transferred to other parts of the world. This policy blind-spot results in the EU’s contribution to tropical deforestation and biodiversity loss in developing countries (Reference Fuchs, Brown and RounsevellFuchs et al., 2020). However, while the EU and member states such as Denmark, the Netherlands, Sweden and the UK (which was an EU member at the time of analysis) have tested approaches for policy coherence for development, implementation performance has been weak (Reference CarboneCarbone, 2008; see also Reference Pendrill, Persson and GodarPendrill et al., 2019). Civil society actors have created a proposal to streamline EU policies into a “Common Food Policy” for Europe (Reference De Schutter, Jacobs and ClémentDe Schutter et al., 2020; IPES-Food, 2019). Blueprints describe an integrated food policy framework that promotes healthy diets and sustainable food systems through coherence across policy areas and governance levels, including by aiming to relocalize food production and to reduce dependence on global food imports (Reference De Schutter, Jacobs and ClémentDe Schutter et al., 2020; IPES-Food, 2019). It remains to be seen to what extent the integrated approach of the European Green Deal, and its “Farm to Fork” strategy, can translate such suggestions into practice.
13.4.4 Capacity
While there is generally higher institutional capacity in developed countries relative to developing countries, the aforementioned division between the institutional processes of the environmental and agricultural sectors undermines social capital for BPI in most countries.
In developing countries, the capacities to develop biodiversity (and other environmental) policies are limited to environmental ministries or departments. In Indonesia, Uganda, Honduras and Peru, social capital and capacities for training, financial support and regulation exist, but are not targeted at ecologically sound forms of production (Reference Zinngrebe, Borasino and ChiputwaZinngrebe et al., 2020). The availability of institutional capacities is further undermined by unclear mandates between government agencies, high turnover among government officials resulting in discontinuous policy formulation and execution, and a lack of experienced biodiversity research institutions or centers of excellence (Reference ZinngrebeZinngrebe, 2018; Reference Zinngrebe, Borasino and ChiputwaZinngrebe et al. 2020). In the public policy arena, there is a lack of knowledge on and awareness of the linkages between biodiversity and agriculture or food security (Reference Beltrame, Oliveira and BorelliBeltrame et al., 2016; Reference Chandra and IdrisovaChandra and Idrisova, 2011). This is largely due to lack of training, funding, incentives for experts to work in the environmental field (Reference Chandra and IdrisovaChandra and Idrisova, 2011), biodiversity-focused science–policy interfaces, and institutionalized mechanisms for the participation of Indigenous Peoples and local communities (which hold critical local ecological knowledge) in monitoring, reporting and verification initiatives (Reference Vanhove, Rouchette and Janssens de BisthovenVanhove et al., 2017). Mexico tackles these issues via multistakeholder roundtables, consisting of agricultural, rural development and research agencies, Secretaries of States, academia, NGOs and private actors, which coordinate sector activities, financing and science-policy mechanisms at the national and state level (Reference Tutwiler, Bailey, Attwood, Remans and RamirezTutwiler et al., 2017). In Uganda, the agricultural ministry, under the direction of the Ministry of Finance, Planning and Economic Development, has to allocate a portion of their budget to conservation activities (IIED, UNEP-WCMC, 2015). Their staff receive training and a dedicated conservation expert from the environmental ministry to help prepare plans, while policy actors use learning lessons from the ground to inform the national macroeconomic framework (IIED, UNEP-WCMC, 2015). In South Africa, implementation of the Biodiversity Act is supported by pilot projects, regular monitoring and a national science-policy institute and multiagency committees, which align partnerships and cofinancing (Reference Botts, Skowno and DriverBotts et al., 2020).
Within the EU, implementation of agricultural and biodiversity policies is supported by institutions at the European, national and subnational levels. However, lack and variance of capacity among different members states has also been identified as a barrier to implementation of agricultural policy proposals that contribute to environmental protection (Reference Erjavec, Lovec, Juvančič, Šumrada and RacErjavec et al., 2018). Political decision-making and implementation processes of theoretically synergistic policies are designed and implemented by separated policy regimes (Reference Pe’er, Bonn and BruelheidePe’er et al., 2020), undermining social capital and potential synergies. Capacity problems are further enhanced by budgetary imbalances between agricultural and environmental instruments. Although the CAP is the EU policy with the highest budget (€58.4 billion in 2020), the majority of this is dedicated to direct income support. As a result, most of the budget in the 2015–2020 CAP (approximately €40 billion in 2017) was spent on direct payments that support land-intensive and biodiversity-threatening forms of farming, such as intensive animal breeding and monocultures (Reference Pe’er, Zinngrebe and MoreiraPe’er et al., 2019). Furthermore, though Natura 2000 has demonstrated improvements in biodiversity within agricultural areas, funding per hectare is considerably lower than for greening or agri-environment climate measures (Reference Pe’er, Lakner and MüllerPe’er et al. 2017), hardly compensating farmers for resulting costs from forgone incomes due to management restrictions and lower rents, and thus not providing sufficient incentive for adoption by farmers (Reference Bouwma, Zinngrebe, Runhaar, Dries, Heijman, Jongeneel, Purnhagen and WesselerBouwma et al., 2019). Additionally, contradictory technical advice by agricultural extension services and administrative hurdles have hampered effective implementation of biodiversity measures (Reference Zinngrebe, Pe’er and SchuelerZinngrebe et al., 2017).
13.4.5 Weighting
Even where biodiversity policy objectives are present and have been operationalized through concrete instruments with allocated capacity, political discourses are dominated by productivist narratives. The political framing in which food production must increase above all else provides little incentive to phase out agricultural subsidies that support the dominant model but are harmful to biodiversity (Reference Bouwma, Zinngrebe, Runhaar, Dries, Heijman, Jongeneel, Purnhagen and WesselerBouwma et al., 2019; Reference Fouilleux, Bricas and AlphaFouilleux et al., 2017; Reference Roche and ArgentRoche and Argent, 2015). In 2015, OECD countries provided $100 billion in direct and indirect subsidies that stimulated intensive agricultural production (OECD, 2019: 73). Although certification and other schemes are partly driving growth in organic and sustainable practices, the overwhelming policy bias and dominance of conventional agricultural methods gives these practices limited scope for truly scaling-up (Reference Aubert, Hege and KinniburghAubert et al., 2018).
In developing countries, both policies and politics also prioritize agricultural intensification and expansion (Reference Wilson and RiggWilson and Rigg, 2003; Reference Zinngrebe, Borasino and ChiputwaZinngrebe et al., 2020). Biodiversity narratives in Peru show that even conservationists do not dare to talk about limits to production carrying-capacity. Adverse impacts on ecological functionality and related pollution and water-management issues remain untargeted key drivers for biodiversity loss (Reference ZinngrebeZinngrebe, 2016a; Reference Zinngrebe2016b). Another example is China, where, though the Law of Agriculture provides for wetlands conservation, the priority is placed on the draining and cultivation of wetlands for food security, resulting in lower priority and trade-offs for biodiversity (Reference Ongley, Rong and HaohanOngley et al., 2010). Despite successful instruments for supporting agrobiodiversity and integrated natural resource management, agricultural expansion and intensification dominates decision-making considerations (Reference Laurance, Sayer and CassmanLaurance et al., 2014).
Similarly, in the EU, the political discourse and resulting policies are oriented toward increasing productivity for human nutrition (Reference Erjavec, Erjavec and JuvančičErjavec et al., 2009; Reference Freibauer, Mathijs and BrunoriFreibauer et al., 2011; IPES-Food, 2019). Despite the emergence of new discourse elements targeting multi-functionality and liberal markets, central policy elements support productivity (Reference Alons and ZwaanAlons and Zwaan, 2016; Reference Erjavec and ErjavecErjavec and Erjavec, 2015). Following this policy design, even the implementation of conservation mechanisms, such as Ecological Focus Areas, is biased toward measures supporting increased productivity of agricultural lands (e.g. cash crops and nitrogen-fixing crops) (Reference Pe’er, Zinngrebe and HauckPe’er et al., 2016). This is one of the stated reasons for why the CAP has not managed to reverse biodiversity loss (Reference Pe’er, Lakner and MüllerPe’er et al., 2017). Some argue that the CAP is also not likely to do so in the near future, considering the content of current proposals for a post-2020 CAP (Reference Pe’er, Zinngrebe and MoreiraPe’er et al., 2019). This strongly conflicts with the European Green Deal, which explicitly aims to halt biodiversity loss due to agriculture (EC, 2019).
13.5 Looking Forward: Toward Transformative Biodiversity Governance in Agricultural Landscapes
The previous section highlighted the overall very modest advances of BPI in agricultural landscapes. Given that the majority of global and national biodiversity targets are vague and the agricultural sector is not held accountable for its biodiversity performance, there is little guidance for investments in operationalization and capacity-building. Likewise, biodiversity policies are mostly “added on” to regulations of agricultural landscapes, receiving a low share of support compared to that for conventional farming systems focused on productivity. Given the significant agri-food system lock-ins and incumbent power dynamics, more effective BPI will not be implemented spontaneously – rather, the required shifts will need leadership at various levels (Reference Oliver, Boyd and BalcombeOliver et al., 2018; Reference Runhaar, Wilk, Driessen, Biermann and KimRunhaar et al., 2020). We argue that political will is required as a key driving force to overcome lock-ins and improve BPI performance (see Figure 13.2). In the following paragraphs, we present four central leverage points specifying the dimensions for the transformation of biodiversity governance for agricultural landscapes.
Figure 13.2 Improving the BPI level through transformative governance in adaptive learning circles.
A first transformative factor is the creation of a coherent sustainability vision based on inclusive biodiversity governance, which will guide implementation and induce accountability among implementing agents. As we showed in the previous section, the BPI dimensions of inclusion and coherence suffer from a lack of clear orientation, and the weighting is geared toward specific production-oriented interests. Decisions on agricultural policy are often dominated by small but well-organized interest groups that marginalize values of biodiversity conservation and downplay societal mandates such as the biodiversity targets under the CBD (Reference Brown, Kovács and HerzonBrown et al., 2020, Reference Pe’er, Zinngrebe and MoreiraPe’er et al., 2019). Stakeholder groups differ in the way they envision appropriate use of land and nature, leading to different, often disconnected, discourses that are not equally reflected in policy design and implementation processes (Reference Velázquez GomarVelázquez Gomar, 2014; Reference ZinngrebeZinngrebe, 2016a). Questions of accountability and legitimacy of planning will depend on the extent to which potentially conflicting values are acknowledged and diverse value systems and perceptions are reflected in democratic planning and participatory implementation processes (Reference Díaz, Pascual and StensekeDíaz et al., 2018; Reference Runhaar, Wilk, Driessen, Biermann and KimRunhaar et al., 2020; Reference Termeer, Stuiver, Gerritsen and HuntjensTermeer et al., 2013; Reference ZinngrebeZinngrebe, 2016b). Likewise, a positive perspective of what “sustainable agricultural landscapes” entail in a given context helps to orient the decisions and activities of political and nonpolitical actors. There are various alternatives to the dominant productivist model, including agroecology, sustainable intensification, agroforestry, and “nature-inclusive” agriculture (Reference Brouder, Karlsson and LundmarkBrouder et al., 2015; Reference Shukla, Skea and Calvo BuendiaIPCC, 2019; Reference Loos, Abson and ChappellLoos et al., 2014; Reference Perfecto and VandermeerPerfecto and Vandermeer, 2010; Reference Plieninger, Muñoz-Rojas, Buck and ScherrPlieninger et al., 2020; Reference Tscharntke, Clough and WangerTscharntke et al., 2012; Reference van Noordwijkvan Noordwijk, 2019; Reference Zinngrebe, Borasino and ChiputwaZinngrebe et al., 2020). Agroforestry, as a specific example of an agroecological approach, has the potential to support ecosystem functions and biodiversity in both developed (Reference Torralba, Fagerholm, Burgess, Moreno and PlieningerTorralba et al., 2016) and developing countries (Reference van Noordwijkvan Noordwijk, 2019). More concretely, objectives can be formulated around agroecological infrastructure such as hedges, trees and other seminatural habitats that protect multiple taxonomic groups and ecosystem services (Reference Barrios, Valencia and JonssonBarrios et al., 2018; Reference Fagerholm, Torralba, Burgess and PlieningerFagerholm et al., 2016; Reference Gonthier, Ennis and FarinasGonthier et al., 2014; Reference Plieninger, Torralba, Hartel and FagerholmPlieninger et al., 2019; Reference Plieninger, Muñoz-Rojas, Buck and Scherr2020; Reference Poux and AubertPoux and Aubert, 2018; Reference Torralba, Fagerholm, Hartel, Moreno and PlieningerTorralba et al., 2018). Scenarios form an effective method for a participatory visioning process involving policymakers and other actors to deliberate options for land use and assess their implications for food security within a land-constrained world facing climate change (e.g. Reference Aubert, Schwoob and PouxAubert et al., 2019).
A second transformative factor that gives more weight to biodiversity in decision-making on trade-offs is social capital for integrative governance. Especially in developing countries, institutional capacities for implementing policies are severely lacking and often result in institutional gaps between policy integration “on paper” and the implementation of concrete policy instruments (Reference Runhaar, Wilk, Driessen, Biermann and KimRunhaar et al., 2020). Overlapping and unclear competences also create “responsibility gaps” in which no actor actually takes leadership in regulation or wider governance (Reference Sarkki, Niemelä and TinchSarkki et al., 2016). Efforts to improve mainstreaming and fill these gaps have not resulted in institutional reconfigurations favoring effective implementation (Reference HerkenrathHerkenrath, 2002; Reference Prip and PisupatiPrip and Pisupati, 2018). However, environmental impact assessments of large agricultural projects, or approval and monitoring of agroforestry concessions, can improve the operationalization of conservation objectives (Reference Slootweg and KolhoffSlootweg and Kolhoff, 2003; Reference ZinngrebeZinngrebe, 2018). In Europe, both agricultural and environmental policies are well developed, but not institutionally connected in decision-making and implementation structures (Reference Pe’er, Zinngrebe and MoreiraPe’er et al., 2019). Involving farmers in local implementation processes and partnerships with conservationists is an important strategy for improving biodiversity conservation leadership and outcomes in both developing (Reference Harvey, Komar and ChazdonHarvey et al., 2008) and developed countries (Reference Buizer, Arts and WesterinkBuizer et al., 2016; Reference Pe’er, Zinngrebe and MoreiraPe’er et al., 2019; Reference Persson, Eckerberg and NilssonPersson et al., 2016). A collaborative process of aligning policy packages of information, regulation and finance can help overcome fragmentation between political actors and produce coherent incentive systems for conservation practices (Reference Zinngrebe, Borasino and ChiputwaZinngrebe et al., 2020). Such a collaborative process should not only advance top-down implementation of (inter)national regulatory frameworks, but also cover a diverse range of locally based agricultural management practices. The Reference Díaz, Settele, Brondízio and NgoIPBES Global Assessment (2019), for example, highlights a wide number of studies documenting the importance of small agricultural landholdingsFootnote 2 in contributing to biodiversity conservation in different ecosystems (Reference Batáry, Gallé and RieschBatáry et al., 2017; Reference Belfrage, Björklund and SalomonssonBelfrage et al., 2015; Reference Fischer, Brosi and DailyFischer et al., 2008).
A third point of leverage is harnessing private initiatives for integrative governance. Private sector and market-based mechanisms can help with operationalization, provide new sources for institutional capacity, and increase coherence with farming interests (see Chapter 5). Engaging private actors is critical, particularly due to the rise and extent of private governance in the agricultural sector globally. Private actors can help incentivize biodiversity-friendly agriculture through various market opportunities, finance mechanisms, and public–private partnerships and other cooperative mechanisms. For example, numerous cases of the landscape approach have shown cooperation between governmental and private actors, such as co-funding from corporate actors in the maintenance of ecosystem services (Reference Van OostenVan Oosten, 2013). Private agricultural standards (including voluntary programs, such as various organic certifications) have become an integral part of agri-food chain governance (Reference Henson and ReardonHenson and Reardon, 2005; Reference Verbruggen, Havinga, Verbruggen and HavingaVerbruggen and Havinga, 2017). Sustainability certifications (potentially) open new markets (FAO, 2017b) and provide opportunities for the scaling-up of environmental sustainability criteria, including for biodiversity (Reference Runhaar, Melman and BoonstraRunhaar et al., 2017). Particularly in countries that import large quantities of agricultural goods with high biodiversity impacts, government procurement of certified agricultural products can support and incentivize private sector actors in achieving biodiversity goals (Reference FransenFransen, 2018). The use of economic instruments by firms, such as payment for ecosystem services, can also help provide financial incentives for other actors to engage in biodiversity-friendly farming and production processes (Reference Donaldson, Pierce, Cowlings, Sandwith and MacKinnonDonaldson, 2012; Reference Harvey, Komar and ChazdonHarvey et al., 2008; Reference Sanchez‐Azofeifa, Pfaff, Robalino and BoomhowerSanchez‐Azofeifa et al., 2007).
However, to improve biodiversity outcomes, private initiatives need to be accompanied by political regulation and cooperation between private and public actors (Reference Folke, Österblom and JouffrayFolke et al., 2019, Reference Lambin, Gibbs and HeilmayrLambin et al., 2018; Reference Runhaar, Melman and BoonstraRunhaar et al., 2017; Reference Runhaar, Wilk, Driessen, Biermann and Kim2020). So far, land use change and management choices exercised by powerful transnational corporations have had a range of detrimental consequences for biodiversity (Reference Folke, Österblom and JouffrayFolke et al., 2019). In the agri-food sector, consolidation is extremely high among corporations controlling fertilizers, agrochemicals and seeds, as well in the production of specific commodities such as coffee, bananas, soy, palm oil and cocoa (Reference Folke, Österblom and JouffrayFolke et al., 2019). Private initiatives and certification schemes connecting consumer support for sustainable production systems have not yet proven effective in reversing detrimental environmental impacts (Reference Dietz, Estrella Chong, Grabs and KilianDietz et al., 2019; Reference Lambin, Gibbs and HeilmayrLambin et al., 2018; Reference Pendrill, Persson and GodarPendrill et al., 2019). Experiences with green certification show that private standards need to be complemented with adequate regulatory frameworks to avoid deforestation and other detrimental effects to biodiversity, while simultaneously providing sufficient economic incentives for farmers (Reference Dietz, Estrella Chong, Grabs and KilianDietz et al., 2019; Reference Lambin, Gibbs and HeilmayrLambin et al., 2018).
Knowledge integration and learning for informed and adaptive governance is necessary to develop context-specific policy solutions for complex societal challenges. This can help in identifying suitable strategies for operationalization and (targeted) capacity-building. Experiences in participatory land use planning have shown how different knowledge systems can be integrated at the community level to build adaptive capacity and adopt more sustainable land use practices (Reference Rodríguez, Cisneros, Pequeño, Fuentes and ZinngrebeRodríguez et al., 2018). While the EU has a wide range of instruments for conservation in agricultural landscapes, it does not yet use all available knowledge to inform the improvement of these instruments from one funding period to the next (Reference Pe’er, Bonn and BruelheidePe’er et al., 2020). Social capital can facilitate the input and reflection of available knowledge (Reference Zinngrebe, Borasino and ChiputwaZinngrebe et al., 2020). Policy learning based on available experiences has the potential for overcoming complete policy failure and fragmentation (Reference FeindtFeindt, 2010; Reference ZinngrebeZinngrebe, 2018). Reference FeindtFeindt (2010) argues that stronger institutionalized support for policy integration, balanced representation and wider societal engagement is needed to hold back powerful actors from dominating the policy arena to defend the status quo. Certain levels of flexibility and a complementary structure of CAP support and Natura 2000 instruments have shown synergistic effects in increasing the willingness of farmers to adopt conservation measures (Reference Lakner, Zinngrebe and KoemleLakner et al., 2020). In addition, the integration of local knowledge has been shown to improve both farmers’ engagement in reflexive learning processes and policy performance, in the EU context on the CAP’s agri-environmental measures (Reference Goldman, Thompson and DailyGoldman et al., 2007; Reference Prager, Reed and Scott.Prager et al., 2012) and in developing countries, for example in the context of conservation farming in South Africa (Reference Donaldson, Pierce, Cowlings, Sandwith and MacKinnonDonaldson, 2012) or in Mesoamerican landscapes (Reference Harvey, Komar and ChazdonHarvey et al., 2008).
13.6 Conclusion
Low levels of biodiversity policy integration in agricultural policy in both developing and developed countries is a determining factor in the continued biodiversity loss within agricultural landscapes and beyond. While land-sparing approaches have proven to be indispensable for the conservation of certain components of biodiversity (Reference Le Saout, Hoffmann and ShiLe Saout et al., 2013; Reference Watson, Dudley, Segan and HockingsWatson et al., 2014), a more integrated land-sharing approach is necessary to enable a transformation of current trajectories toward sustainable farming, in order to bend the curve of biodiversity loss while also ensuring food security, climate resilience, enhanced animal welfare and improved rural livelihoods.
With the exception of EU policies, in most countries, specific biodiversity-related objectives are missing in agricultural policies. Worldwide, the underlying drivers of biodiversity loss from agriculture are not sufficiently addressed. In particular, the objective of phasing out policies supporting threats to biodiversity and a strongly productivist-oriented agricultural sector overpowers the idea of sustainable agriculture. Instead of coherent targets and complementary institutional structures, conservation has generally been treated as an add-on to business-as-usual agricultural policy. Trade-offs considering biodiversity and ecological limits are seldom explicitly recognized in agricultural policies, and no country expresses a long-term vision for the development of sustainable agricultural landscapes. Political discourses remain centered on prioritizing intensive food production, thereby marginalizing the potential functions of agricultural landscapes for biodiversity conservation. Based on our BPI analysis, we extract the following recommendations for transformative biodiversity governance:
1. Inclusive governance needs to genuinely incorporate multiple stakeholder views and perceptions, and negotiate and develop clear, coherent visions and definitions of sustainable agriculture to legitimate policies and decision-making.
2. Integrative governance can be improved by building social capital as a means to creating favorable actor constellations and institutional structures incentivizing and prioritizing biodiversity-sound practices.
3. Integrative governance can benefit from complementing public and private initiatives in coherent governance structures.
4. Informed and adaptive governance requires a continuous and participatory reflection of governance systems to guide institutional learning processes toward sustainable agricultural landscapes.
We argue that the Post-2020 Global Biodiversity Framework should focus on the transformation of agricultural governance systems by concretely addressing key leverage points and providing specific guidance for member states to address country-specific drivers and potentials for sustainable innovation through biodiversity policy integration. Eventually, however, the dynamic of this transformative process will be conditioned by political will and active leadership at all levels.
14.1 Introduction
The governing of nature has been an essential part of the story of urbanization. Whether through the conversion of rivers for transportation, the creation of urban drainage systems for wastewater removal or the installation of parks for their recreational and aesthetic value (Reference GandyGandy, 2004; Reference Gleeson, Low, Low, Gleeson, Elander and LidskogGleeson and Low, 2000; Reference RydinRydin, 1998), nature has played a critical role in urban development. Yet, conservationist thinking, which has dominated environmental governance and policy, has tended to equate the environment as belonging to either “rural” or “wilderness” places that needed to be protected from the encroachment of (urban) society (Reference Owens and BrehenyOwens, 1992). As a result, much of the governance of biodiversity at the urban scale during the twentieth century was focused on the designation and enforcement of protected areas (Reference Vaccaro, Beltran and PaquetVaccaro et al., 2013). Yet such dualistic thinking has ignored the ways in which nature inhabits the city, whether intended or otherwise, from domestic gardens to public parks, urban sewers to derelict corners of the city, as well as the potential benefits that such forms of biodiversity can bring to the city.
It has only been since the late 1980s that how cities might contribute toward local, national and global sustainability has begun to be recognized. While climate change has tended to dominate this agenda, cities also have a range of different yet substantial roles in addressing the loss of nature: as habitats for biodiversity and threatened species (Reference Aronson, La Sorte and NilonAronson et al., 2014; Reference Hall, Camilo and ToniettoHall et al., 2017; Reference Ives, Lentini and ThrelfallIves et al., 2016; Reference Soanes and LentiniSoanes and Lentini, 2019); as locations for people to connect with nature (Reference Soanes and LentiniSoanes and Lentini, 2019); as key jurisdictions in global and multilevel governance (Reference Pattberg, Widerberg and KokPattberg et al., 2019) and as important consumption arenas driving biodiversity loss globally (Reference Díaz, Settele and BrondízioDíaz et al., 2019). Nonetheless, it was not until 2008 that the first Global Biodiversity Summits of Local and Subnational Governments was held in parallel to the Conference of the Parties to the Convention on Biological Diversity. These summits have since taken place biannually and are intended as a means through which to reinforce the recognition and involvement of local and subnational governments in contributing to CBD objectives and targets. While the initial version of both the Strategic Plan for Biodiversity 2011–2020 and the 20 Aichi Biodiversity Targets makes no direct references to cities or urban areas, a subsequent assessment of the Aichi targets and the 2030 Agenda for Sustainable Development found that Sustainable Development Goal 11 on Sustainable Cities and Communities corresponded to six (2, 4, 8, 11, 14 and 15) of the Aichi targets (CBD, 2016). At the same time, the 2010 Decision X/22 of the Convention on Biodiversity laid out explicit terms on which the Parties to the convention were to be encouraged to recognize and facilitate the work of subnational and local authorities through the development and implementation of local biodiversity strategies and action plans (LBSAPs). Over the past decade, the urban dimension of biodiversity issues has come to be increasingly recognized.
Yet despite this, in practice biodiversity governance has yet to gain widespread traction at the local level, and local biodiversity planning has been critiqued for an overly narrow approach, the exclusion of diverse values for nature and limited effectiveness (Reference Bomans, Steenberghen, Dewaelheyns, Leinfelder and GulinckBomans et al., 2010; Reference Elander, Alm, Malbert and SandstrÖmElander et al., 2005; Reference EvansEvans, 2004; Reference Wilkinson, Sendstad, Parnell, Schewenius, Elmqvist, Fragkias and GoodnessWilkinson et al., 2013). In this chapter, we explore how the governance of urban nature is evolving in response to the increasing urgency of this agenda. In so doing, we follow the distinction put forward by Reference Patterson, Schulz and VervoortPatterson et al. (2017) and highlighted in Chapter 1 between governance for transformation, where governance creates the conditions by which transformative change can emerge; governance of transformations, where governance is deliberately intended to advance transformative change in terms of either processes or outcomes that involve systemic or structural shifts in current socioecological orders; and transformations in governance, where governance regimes – their architectures, agency, power and so forth – are themselves transformed. We find that, internationally, urban biodiversity governance is being transformed both in terms of its intentions (governance for transformation) – moving from a concern only with reducing the threat of cities to biodiversity to also realizing their benefits (Section 14.2) – and in terms of the forms that governance is taking (transformation in governance) – through the growth of governance experimentation in cities and the growth in transnational governance networks (Section 14.3). These shifts are changing the outcomes of what biodiversity governance in the city is seeking to realize – from a focus on specific places and parts of nature to a broader engagement with multiple socio-natures and the ways in which working with nature can generate sustainability benefits for a diverse range of communities. At the same time, within urban policymaking and practice on the ground, there has yet to be a significant effort to address the ways in which cities contribute to the underlying drivers of biodiversity loss through explicitly linking their roles and responsibilities in reducing waste, combating climate change and shaping production and consumption with biodiversity agendas. We return to these points in conclusion (Section 14.4) and reflect on their implications for the ways in which cities can contribute to transformative biodiversity governance.
14.2 Transforming Biodiversity in the City: from Threat to Opportunity?
If, for the most part of modern urban development, cities were regarded as separated from nature, the global environmental challenges facing society in the twenty-first century have abruptly erased any such boundaries. As the IPBES Global Assessment makes clear, cities are a primary driver of biodiversity loss through urban expansion and pollution, as well as affecting the loss of nature globally through the consumption practices of urban residents and the global value chains of urban economies. The detailed analysis presented in the Nature in the Urban Century report (Reference McDonald, Colbert and HamannMcDonald et al., 2018) asserts that urban growth was responsible for the loss of 190,000 km2 of natural habitat between 1992 and 2000 and could threaten 290,000 km2 of global natural habitat by 2030. Cities located in globally important biodiversity hotspots bear special significance in this context. Biodiversity hotspots are areas of exceptional concentrations of endemic species that are simultaneously undergoing a high rate of loss of habitat. It was estimated that in 1995, 20 percent of the world’s population was living in global biodiversity hotspots, which accounts for about 12 percent of the earth’s surface. Population growth in these hotspots was estimated to be 1.8 percent per annum (Reference Cincotta, Wisnewski and EngelmanCincotta et al., 2000).Footnote 1 Such impacts are not only felt in areas with particular biodiversity value: urbanization and increased impervious surfaces are also having severe impacts on urban wetlands and waterways (Reference Booth, Roy, Smith and CappsBooth et al., 2016). In short, even though the impact of individual cities will be highly varied, the weight of evidence suggests that urbanization processes are “catastrophic for native species, and … a well-known threat to biodiversity worldwide” (Reference Garrard, Williams, Mata, Thomas and BekessyGarrard et al., 2017: 1).
For the most part, it has been this discourse of the in-situ impacts of urbanization on biodiversity either within the city boundary or at its expanding edge that has shaped how the potential role of cities in governing biodiversity has been framed (Reference Bulkeley, Kok and XieBulkeley et al., 2021). Over the past decade, the Convention on Biodiversity has primarily focused on the spatial planning capacities of cities as essential to managing urban encroachment on biodiversity and on the importance of protected areas for biodiversity conservation. Yet this underplays two other important ways in which cities are connected to the biodiversity challenge. First, as the IPBES Global Assessment makes clear, cities have a significant role in shaping the drivers of biodiversity loss – from climate change to consumption. Second, as urbanists have long recognized, cities are intricately connected to and dependent on nature – from water resources to urban parks (Reference GandyGandy, 2002; Reference Swyngedouw, Kaika, Bridge and WatsonSwyngedouw and Kaika, 2000). There is now a growing interest in the ways in which cities can benefit from both the ecosystem services that nature provides and also how urban nature and biodiversity contribute to less readily quantified values, such as heritage, well-being, stewardship and reverence, and provide an essential form of connection between nature and people in the urban milieu. Urban nature is increasingly recognized for its capacity to not only support biodiversity conservation, but also to generate additional environmental, economic and social benefits – or what are termed “nature’s contribution to people” (Reference Kabisch, Frantzeskaki and PauleitKabisch et al., 2016). This is reflected in a growing interest in urban nature-based solutions, an umbrella term used to encompass ecosystem-based adaptation (Reference Geneletti and ZardoGeneletti and Zardo, 2016; Reference Munang, Thiaw and AlversonMunang et al., 2013), green infrastructure and ecosystem services (Reference Cohen-Shacham, Walters, Janzen and MaginnisCohen-Shacham et al., 2016; Reference Dorst, Raven, van der Jagt and RunhaarDorst et al., 2019; Reference Nesshöver, Assmuth and IrvineNesshöver et al., 2017; Reference Pauleit, Zölch, Hansen, Randrup, Konijnendijk van den Bosch, Kabisch, Korn, Stadler and BonnPauleit et al., 2017). Nature-based solutions provide a means through which cities not only have the potential to benefit directly from nature, but also contribute to addressing the global challenge of the loss of biodiversity. In the rest of this section, we examine how cities are currently undertaking action that can contribute to three key elements of biodiversity governance – protecting or conserving nature, restoring nature and fostering the value of nature’s contributions to people through thriving with nature. We suggest that there is significant evidence that cities can no longer be viewed simply as a threat to biodiversity, but are transforming their role to one of significant opportunity. In doing so, they are adopting new means of governing nature in the city, which in turn are leading to the transformation of biodiversity governance within and beyond its boundaries.
14.2.1 Urban Biodiversity Conservation
The main goal of conservation is to prevent further loss and degradation of natural ecosystems and resources (Reference YoungYoung. 2000), although in practice this can include the preservation, maintenance, sustainable use and enhancement of the components of biological diversity as well as exploring how society lives in harmony with nature. Although cities have been seen to hold little conservation value, there is increasing recognition of the role that urban green spaces, waterways and wetlands play in conservation, and its wider contributions to human health and well-being (Reference Aronson, La Sorte and NilonAronson et al., 2014; Reference Endreny, Santagata and PernaEndreny et al., 2017; Reference Parris, Amati and BekessyParris et al., 2018). Cities also provide habitat for threatened species, and some threatened species are found exclusively in urban areas (Reference Soanes and LentiniSoanes and Lentini, 2019). Reference Ives, Lentini and ThrelfallIves et al. (2016: 117) analyzed the distribution of Australia’s listed threatened species and found that 30 percent are found in cities and that “Australian cities support substantially more nationally threatened animal and plant species than all other non-urban areas on a unit-area basis.” Globally, while a large number of species have been disadvantaged or made locally extinct by urbanization, urban areas have also provided range expansions for other species, including fruit-eating bats (Reference Williams, McDonnell, Phelan, Keim and Van Der ReeWilliams et al., 2006) and nectar-feeding birds that feast on the well-watered and productive plants found in urban gardens.
It is therefore a misconception that cities cannot contribute significantly to biodiversity conservation (Reference Soanes, Sievers and CheeSoanes et al., 2019), yet even where this is recognized, there is significant debate concerning how this contribution can be realized. As the main preference is “given to conserving large, highly connected areas,” “relative ambivalence [is] shown toward protecting small, isolated habitat patches” even though they are “inordinately important for biodiversity conservation” (Reference Wintle, Kujala and WhiteheadWintle et al., 2019: 909). Far from being delivered through systematic forms of urban (biodiversity) planning, urban reserve or park systems are often small, fragmented and disconnected, located on leftover, undevelopable land or squeezed in size due to urban development pressures and economic imperatives, but nonetheless have been shown to make important contributions to conservation (Reference Kendal, Zeeman and IkinKendal et al., 2017). While “effective conservation planning requires an understanding of species-habitat relationships” that goes beyond the simplified single species focus (Reference Threlfall, Law and BanksThrelfall et al., 2012: 41), in urban areas the same species that may be valued as threatened species may also be labeled as pests. The grey-headed flying fox is listed as a vulnerable species in Australia, but in Melbourne, a colony of the animals was evicted from the botanic gardens for causing roosting damage to trees. Black-legged kittiwakes, a threatened gull species that nest and breed in areas along the quayside in Newcastle, UK, are blamed for mess and noise, such that while birdwatchers and wildlife enthusiasts celebrate their presence, local businesses are less enthusiastic and have used various means (spikes, nets, electric shocks) to attempt to prevent nesting. At the same time, as cities experience the impacts of climate change, what it is appropriate to conserve is also coming into question as much-loved and threatened urban species may not be able to flourish under changing conditions (Reference LennonLennon, 2015; Reference Prober, Doerr, Broadhurst, Williams and DicksonProber et al., 2019). These complexities point to the challenges of governing biodiversity in human-dominated landscapes, suggesting that the forms of nature that are or are not valued cannot be established through scientific knowledge but, as the Introduction to this volume suggests, require the bringing together of diverse forms of knowledge often in a transdisciplinary manner.
Governing for urban conservation is therefore no straightforward matter, but rather shot through with contention over which kinds of nature should be conserved, for whom and under which conditions. The importance of fragmented urban nature and small, disconnected spaces in cities for biodiversity conservation also suggests that addressing biodiversity goals involves multiple sites and actors that are not directly engaged in the formal land-use planning or regulatory systems of local authorities. Indeed, as we discuss further below, it appears that urban conservation governance is being transformed – rather than being led by urban planning, it is now taking place through a whole host of initiatives and programs, including those conceived as nature-based solutions to diverse sustainability challenges, that are undertaken by a range of urban actors, including private and civil society organizations. In this context, rather than requiring more integrated governance, as Chapter 1 suggests, fragmented and diverse forms of governance are potentially a more viable means through which to transform the capacity of cities to address the loss of nature.
14.2.2 Urban Biodiversity Restoration
While conservation mainly focuses on preventing ongoing and future losses, restoration seeks to actively reverse such degradation (Reference Garson, Garson, Plutynski and SarkarGarson, 2016). Similar to conservation activities, restoration activities differ greatly in their spatial scale and in terms of the sheer magnitude of intervention they entail (Reference Garson, Garson, Plutynski and SarkarGarson, 2016). With cities’ roles in biodiversity conservation being increasingly recognized, more attention is also being directed toward the restoration of urban green spaces for biodiversity habitat (Reference Butt, Shanahan and ShumwayButt et al., 2018). Restoration activities have focused on habitat improvement and planting; creating artificial structures for nesting, shelter or to facilitate faunal movement and connectivity between sites; control of pest or invasive species; and community engagement and education programs, including citizen science and site or species monitoring programs (Reference Threlfall, Soanes and RamalhoThrelfall et al., 2019). Green spaces that include understory cover and increased structural complexity of vegetation have been shown to improve biodiversity outcomes, and therefore restoration efforts that “redress the dominance of simplified and exotic vegetation … with an increase in understorey vegetation volume and percentage of native vegetation will benefit a broad array of biodiversity” (Reference Threlfall, Mata and MackieThrelfall et al., 2017: 1874). Furthermore, with studies showing the “inordinately important” contribution of small, isolated habitat patches for biodiversity conservation, the restoration of these small patches of urban green space, wetlands and waterways should be “urgently prioritised” (Reference Wintle, Kujala and WhiteheadWintle et al., 2019: 909). This in turn implies a model of governance that extends beyond the capacities of local governments to include a host of actors who own and manage urban land and water systems.
Cities allow for a diversity of restored habitats that serve to improve conditions for biodiversity in public and private lands (Reference Aronson, Lepczyk and EvansAronson et al., 2017). For example, practices such as the restoration of native prairie vegetation along roadsides has been shown to increase bee species richness (Reference HopwoodHopwood, 2008). Moreover, urban green and blue spaces are being increasingly recognized for their capacity to not only support biodiversity conservation (Reference Dunn, Gavin, Sanchez and SolomonDunn et al., 2006; Reference Goddard, Dougill and BentonGoddard et al., 2010; Reference Miller and HobbsMiller and Hobbs, 2002; Reference NiemelaNiemela, 1999), but also to generate additional environmental, social-cultural, and economic benefits, including managing water quality, fostering community inclusion and generating new opportunities for business (Reference Haase, Kabisch and HaaseHaase et al., 2013; Reference Kabisch, Qureshi and HaaseKabisch et al., 2015), as well as fostering the functioning of ecosystems for climate change mitigation and adaptation (European Commission, 2015). The restoration of Merri Creek, a waterway in Melbourne, Australia, has seen the return of a range of species, including birds such as the sacred kingfisher, and pollution-sensitive insects to restored wetlands beside the creek (Reference Bush, Miles and BainbridgeBush et al., 2003; Reference McGregor and McGregorMcGregor and McGregor, 2020). The restoration has provided many opportunities for community involvement in replanting, rubbish collection and so on, underpinning a remarkable community reconnection with the creek and a renewed sense of shared ownership (Reference Bush, Miles and BainbridgeBush et al., 2003). In the Netherlands, many citizen grassroots initiatives around urban nature exist, but their contribution to restoration in a classical sense (i.e. conserving rare/Red list species) is limited, not only because of their spatial scale but also because their objectives in terms of social, economic and environmental outcomes are not always aligned with such outcomes (Reference Mattijssen, Buijs, Elands and ArtsMattijsen et al., 2018). While restoration efforts have focused on public land, there is an increasing recognition of the potential contribution of greening the private realm. The City of Melbourne has recently joined a number of other cities, including Seattle, Helsinki and Malmo, in establishing a “Green Factor Tool” to encourage the integration of greening in new buildings and developments by private developers (Reference Bush, Ashley, Foster and HallBush et al., 2021; City of Melbourne, 2017).
However, restoring urban habitat brings to the fore the potential for increasing conflicts between humans and nonhumans in these urban “shared habitats.” For example, in an Australian urban creek restoration project, neighboring residents viewed the return of native birds and lizards either neutrally or favorably, but there were fears about the return of snakes, which created conflict (Reference Maller and FarahaniMaller and Farahani, 2018). In another case, we can see that while water-sensitive urban design treatments to address urban issues of flooding and stormwater management can enhance biodiversity (Reference Parris, Amati and BekessyParris et al., 2018), they may contain high levels of contaminants, including pesticides and heavy metals from stormwater runoff, which can potentially endanger water quality for human use (Reference Sievers, Hale, Swearer and ParrisSievers et al., 2019). At the same time, what counts as “restoration” is also continually a matter of negotiation and contestation. Global environmental change poses challenges to traditional practices of “restoring ‘degraded’ ecosystems to a ‘natural state’ of acceptable historic variability” (Reference LennonLennon, 2015) such that the end goals of restoration are far from clearly determined by science alone. Further, “novel ecosystems,” which are composed of “non-historical species configurations” and dominate many urban landscapes, are rarely considered as worthy of either conservation or restoration, despite providing rich species assemblages and biodiversity habitat (Reference Planchuelo, von Der Lippe and KowarikPlanchuelo et al., 2019).
As with the governing of urban biodiversity conservation, interventions and practices aimed at enabling the restoration of urban nature for biodiversity is fraught with conflicts, indeterminacy and the potential for exclusionary processes that revere some forms of nature at the expense of others (Reference Tozer, Hörschelmann, Anguelovski, Bulkeley and LazovaTozer et al., 2020). While questions of the design and implementation of such schemes have been debated, there has been less consideration of “new principles that can help guide goal-setting for nature conservation and ecological restoration in dynamic environments” (Reference Prober, Williams, Broadhurst and DoerrProber et al., 2017: 477), particularly in the face of climate change (Reference Prober, Doerr, Broadhurst, Williams and DicksonProber et al., 2019). Indigenous people’s perspectives and knowledge, which have critical contributions for connecting past, present and future natural and cultural heritage, must be embedded in these debates for new principles as well as broader planning and implementation of conservation and restoration activities. Indigenous knowledge and perspectives are “crucial for long-term, sustained biodiversity conservation, land and water management” (Reference Threlfall, Soanes and RamalhoThrelfall et al., 2019: 3). As has been found with conservation initiatives, recognizing the key role that restoration in cities can play toward realizing global biodiversity goals also suggests that multiple actors and modes of governing beyond traditional forms of land-use planning and regulation will need to be harnessed if its potential is to be realized. While this may take place through the development of more inclusive forms of governance, we suggest it will also involve forms of protest, contestation and conflict over whose nature should be conserved or restored.
14.2.3 Thriving with Urban Biodiversity
As the IPBES Global Assessment makes clear, in addition to seeking to conserve and restore nature, a central concern for biodiversity governance in the coming decade will be to ensure that nature’s contribution to people is preserved and enhanced (Reference Díaz, Settele and BrondízioDíaz et al., 2019). In short, to ensure that cities can thrive with nature. How, why and with what consequence it is possible to consider nature as generating a contribution to individuals and to society has been subject to intense debate, as scholars, activists and policy-makers take issue with the extent to which such contributions are framed as instrumental – a means to a human end – or as ensuing from a sense of connection, spirituality or well-being derived from knowing and being in nature (Reference Gavin, McCarter and BerkesGavin et al., 2018; Reference Pascual, Balvanera, Díaz, Pataki and O’FarrellPascual et al., 2017). Attempts to identify so-called ecosystem services that contribute to societal needs and to calculate their monetary value have in particular been subject to a strong critique that doing so reduces the actual contribution that nature makes to society to a narrow range of attributes and functions that can be captured in this way (Reference Schröter, van der Zanden and van OudenhovenSchröter et al., 2014). Recent years have witnessed something of a move away from this position to a recognition of the multiple ways in which nature contributes toward society, as well as the continued importance of recognizing the intrinsic value of nature itself (Reference Díaz, Pascual and StensekeDíaz et al., 2018).
This shift in conceiving of nature as providing singular and functional benefits for society to a position in which the multiplicity of nature’s contributions is recognized can also be witnessed at the urban level. The growth and increasing prominence of the discourse of nature-based solutions, particularly in the European Union, draws explicitly on the idea that nature can contribute to addressing the challenges facing cities, for example in terms of air or water quality, while at the same time generating a wide range of benefits, such as flourishing biodiversity and enhanced well-being, that are not so readily captured in functional or economic terms. Despite the novelty of the term, it is clear that historically urban nature has played these multiple roles, offering a means through which cities could function more effectively but also creating more or less formalized spaces of connection, solidarity and spirituality for diverse communities. In Victorian Britain, for example, formalized parks were seen to provide havens from city life for reflection and recreation. In cities that experienced colonization, Indigenous communities continued to maintain and fight for rights in order to continue to access both food resources and their cultural and spiritual connections to land and water.
As cities now seek to realize diverse goals for urban sustainability, working with and for nature has come to play a vital role. In Tianjin, China, for example, the Ecological Wetland Park is a constructed, artificial wetland with an approximate size of 630,000 m2 located in one of the largest industrial, logistics and free-trade centers of the country. Its aims are not only to enhance the environmental quality of the industrial park, but also to generate space for biodiversity, a thriving economy and enhanced social well-being. In Winnipeg, Canada, a grassroots-run neighborhood group – the Spence Neighbourhood Association – is working with Indigenous communities and local stakeholders to transform more than fifty vacant lots into edible community gardens and parks. Besides their conservation value, these urban green spaces provide important social, economic and environmental benefits. For example, in the Ogimaa Gichi Makwa Gitigaan garden, which opened in 2012, the inclusion of indigenous plants not only contributes to the conservation of local species, but also allows community members to utilize traditional knowledge while learning about horticultural practices. These examples show that cities are transforming their development approach by seeking to thrive with nature in multiple ways. Yet the multifunctionality of nature-based solutions provides both opportunities and challenges. While they are frequently asserted, the benefits, synergies and trade-offs of interventions designed to generate a contribution to society need to be better investigated (Reference Raymond, Frantzeskaki and KabischRaymond et al., 2017). Multifunctionality is also problematic in view of the organization of local governments and the private sector in specialized “silos” (Reference Dorst, Raven, van der Jagt and RunhaarDorst et al., 2019; Reference Kabisch, Frantzeskaki and PauleitKabisch et al., 2016), meaning that while in principle the idea of generating diverse contributions to society is regarded as a benefit, such interventions can lack the political champions or consistent backing required to ensure that they are taken up as part of urban development.
Understanding who benefits and how from urban nature’s contribution to people is not only important from the perspective of their uptake, but also in relation to their consequences. Research has documented a persistent phenomenon of green gentrification emerging in relation to efforts to develop and enhance nature’s contributions to people within cities, leading not only to forms of demographic change and displacement, but also exclusion from the very benefits that nature is supposed to generate (Reference Anguelovski, Connolly, Masip and PearsallAnguelovski et al., 2018; Reference Wolch, Byrneb and NewellWolch et al., 2014). Such processes not only serve to reproduce and deepen urban inequalities, but also to sustain particular dominant views about which forms of nature can best contribute to society, generating elitist and often exclusionary views of what “counts” as the kinds of urban nature and biodiversity that should be conserved, restored and generated (Reference Mattijssen, Buijs, Elands and ArtsMattijssen et al., 2018; Reference Tozer, Hörschelmann, Anguelovski, Bulkeley and LazovaTozer et al., 2020). Rather than taking for granted how nature-based solutions should intervene to contribute to biodiversity, if they are to ensure that diverse communities are to thrive with nature in the city, it is vital that the kinds of nature and biodiversity that are being generated and the auxiliary benefits they carry are subject to scrutiny by those who may need the benefits of nature most. Rather than assuming that nature and biodiversity are automatically of benefit to urban residents, it is critical that the ways in which urban nature has historically been used to repress and exclude different communities is considered in efforts to govern urban biodiversity and its wider benefits, or there is a significant risk that such interventions will contribute to, rather than transform, urban inequalities (Reference Kuras, Warren and ZindaKuras et al., 2020). While measures to support inclusive governance, as suggested in Chapter 1, can seek to make alternative voices heard, without more fundamental changes to the structures of power within which decisions about urban futures are made, and an acknowledgment that contestation and conflict may be a necessary part of generating alternatives, inclusive governance is unlikely to be sufficient.
14.3 Transformative Urban Governance for Biodiversity?
Our analysis suggests that cities are now engaged in a vast array of efforts toward conservation, restoration and thriving with nature, both through their efforts to maintain existing forms of urban nature and through the increasing focus on nature-based solutions as interventions by which to accomplish multiple sustainability goals. Urban biodiversity governance is not confined to the actions of municipal authorities, but undertaken through a wide range of interventions. In this section, we examine how urban biodiversity governance is being transformed as a result, and with what consequences for the capacity of cities to engage in the transformative governance of biodiversity. We first examine the multiple modes of governing through which cities are mobilizing their actions on biodiversity. We then turn to examine how the urban governance of biodiversity is being transformed by the growth of transnational initiatives, generating a growing “urban biodiversity complex.” We suggest that these transformations in the ways in which governing biodiversity in the city are taking place each generate new forms of transformative capacity, but that this is yet to be recognized within the global biodiversity governance landscape.
14.3.1 Transforming the Modes of Governing Biodiversity in the City
If the governance architectures envisaged by international organizations a decade ago assumed that municipal authorities might be involved in contributing to the global governance of biodiversity through the development of LBSAPs that contribute to national biodiversity strategies and action plans (NBSAPs) and global goals (Reference Puppim de Oliveira, Doll, Moreno-Peñaranda, Balaban and FreedmanPuppim de Oliveira et al., 2014), this form of vertical alignment or integration is relatively rare, with only a fraction of national plans containing urban goals, and the majority of strategies and plans developed at the urban scale operating relatively independently of national biodiversity planning (Reference Xie and BulkeleyXie and Bulkeley, 2020). In part this is due to issues of capacity and competing demands within municipalities. Planning tools and mechanisms are often limited in their coverage of, or ability to address, biodiversity. Further, making the case to invest municipal funds into natural assets is also challenging in the face of pressing city needs such as housing and poverty alleviation. Nonetheless, using the planning system to assign protected areas within and on the borders of cities has remained popular as a model to govern urban biodiversity (Reference Vaccaro, Beltran and PaquetVaccaro et al., 2013). However, these governance approaches have drawn criticism for their top-down character, exclusionary stipulations and the associations of this form of governance with the control of nature (Reference Vaccaro, Beltran and PaquetVaccaro et al. 2013). Cities located in biodiversity hotspots face different challenges, as it appears that many of them lack planning approaches that are specifically geared toward harmonizing the need to simultaneously secure globally important biodiversity and the need to accommodate growing cities (Reference Weller, Drozdz and KjaersgaardWeller et al., 2019).
As well as being shaped by the challenges of implementing biodiversity planning on the ground, the lack of alignment or integration between global, national and local policy and planning for biodiversity is a result of the increasingly complex, fragmented and multiple forms through which urban biodiversity governing takes place. Analysis of fifty-four examples of urban nature-based solutions in eighteen cities found that no fewer than twelve different modes of governing were being deployed in order to govern urban nature, ranging from those that were wholly without the involvement of municipal or other government actors, such as those undertaken by philanthropic donors or community organizations, through to those that were wholly enacted by municipalities through their capacities to finance, build and implement infrastructure projects (Reference BulkeleyBulkeley, 2019). Across these modes of governing, the forms of regulation and land-use planning associated with traditional forms of biodiversity strategies and action plans were relatively muted, in comparison to a diverse range of governing mechanisms related to incentives, persuasion, provision, enabling and so forth. This reflects a broader phenomenon now extensively documented in the literature on urban sustainability governance, which suggests that experimentation has come to be a critical means of governing the city toward sustainability (Reference BulkeleyBulkeley, 2019; Reference Bulkeley, Broto and EdwardsBulkeley et al., 2014; Reference Evans, Karvonen and RavenEvans et al., 2016). As Reference Karvonen, Turnheim, Kivimaa and BurkhoutKarvonen (2018: 202) explains, “experiments might not simply serve as one-off trials to provide evidence and justification for new … policies, regulations, and service provision through existing circuits of policymaking and regulation. Instead, these activities are emerging as a new mode of governance in themselves.” In this view, governance by experimentation is increasingly operating alongside and indeed replacing traditional “plan-led” forms of urban governance in the face of growing fragmentation of authority and the growth of the number of actors with a stake in urban futures (Reference BulkeleyBulkeley, 2019).
Alongside the trend in the growth of biodiversity governance experimentation, analysis suggests that a specific form of intervention – nature-based solutions – is also gaining momentum (Reference Almassy, Pinter and RochaAlmassy et al., 2018). The governance of nature-based solutions shows strong parallels to other forms of urban experimentation (Reference Dorst, Raven, van der Jagt and RunhaarDorst et al., 2019), which are often characterized by participation, collaboration and learning (catalyzing local and tacit knowledge), which can contribute to inclusive, transdisciplinary and adaptive governance (Reference FrantzeskakiFrantzeskaki, 2019; Reference Munaretto, Siciliano and TurvaniMunaretto et al., 2014; Reference Plummer, Armitage and de LoëPlummer et al., 2013; Reference ReidReid, 2016; Reference Triyanti and ChuTriyanti and Chu, 2018). Indeed, collaborative forms of governance dominate the design and implementation of nature-based solutions in European cities and beyond (Reference Almassy, Pinter and RochaAlmassy et al., 2018; Reference BulkeleyBulkeley, 2019). While significant barriers to mainstreaming nature-based solutions remain – not least with respect to knowledge about their value and operation, the disruption they pose to existing ways of undertaking urban development, and access to finance – it is apparent that at least some forms of nature-based solutions are becoming systematically deployed. For example, as a response to changing predictions of the nature and extent of urban flooding, “sponge city” and “sustainable urban drainage” approaches are now routinely used, often creating and restoring habitat within cities and contributing to conservation goals as well as generating contributions to social well-being, health and economic development. Overall, we can suggest that the growth of urban governance experimentation is fueling the uptake of nature-based solutions, which provide forms of intervention that work across a landscape of fragmented authority and a plethora of agendas around which nature-based interventions can gather, while the multiple benefits that nature-based solutions promise serves to attract more, and more varied, actors toward governing biodiversity in the city through experimentation.
Yet despite the evident ways in which urban biodiversity governance is being transformed as a result, there is less clear evidence that urban nature-based solutions are effectively addressing issues of urban inequalities, and indeed a growing literature suggests that they could have precisely the opposite effect, casting doubt on their transformative reach. Research on the phenomenon of “green gentrification” points to the ways in which urban (re)development projects that bring nature into the city can have a significant effect on widening inequalities, displacing residents as land values and house prices rise and failing to secure access to new forms of urban nature for communities who may already suffer from multiple forms of social exclusion (Reference Anguelovski, Connolly, Masip and PearsallAnguelovski et al., 2018; Reference Wolch, Byrneb and NewellWolch et al., 2014). For example, the now famed High-Line project in New York, while often celebrated as an economic regeneration development in the city, has also been critiqued as effectively serving the interests of business, tourists and higher income groups at the expense of the (former) residents of the neighborhood (Reference Anguelovski, Connolly, Masip and PearsallAnguelovski et al., 2018). Equally important, efforts to bring nature into cities can serve to reproduce particular ideas about what constitutes valuable or appropriate forms of nature, failing to take account of the manifold and often contested values for nature held by diverse communities. For example, the views and values of Indigenous communities concerning the kinds of nature that should be included in urban plans are often overlooked. This suggests that the governing of urban nature can be far from transformative, serving to reproduce existing social inequalities and the systems of capitalist urban development that are in many senses responsible for driving the loss of nature globally. On the other hand, where issues of social inclusion, the multiple values of nature and justice are taken into account, there is gathering evidence that efforts at governing urban nature can be transformative. In Winnipeg, for example, an initiative has been developed to harness Indigenous knowledge to develop community gardens in vacant lots in the city to provide space for alternative nature in the city and address issues of isolation and poor mental health among these social groups. How, by and for whom urban nature is governed is therefore critical in shaping its potential to be transformative of urban inequalities. Advocating for grassroots actions, the notion of urban nature stewardship offers opportunities for scientific and policy partnerships with local communities (Reference Connolly, Svendsen, Fisher and CampbellConnolly et al., 2014; Reference Krasny and TidballKrasny and Tidball, 2012), highlighting the importance of the openness and inclusiveness of urban nature governance that allows the participation of different stakeholders. Yet while such approaches can be transformative for those involved, many of the issues regarding exclusion and inequality at large remain challenging to address through such interventions. This in turn suggests that alongside any efforts at more inclusive governance, there needs to be space for dissent and contestation so that the nature of ongoing inequalities and their consequences can be made visible.
14.3.2 Transnational Transformations?
In parallel to the shift from an urban planning approach to biodiversity governance at the local level toward urban experimentation and nature-based solutions, we can see that the governance of urban biodiversity is also evolving in the international arena. First, within the Convention on Biological Diversity itself there has been a renewed commitment to the importance of urban action, notably through the development of the Edinburgh Process, through which local and subnational governments have been mandated by the Secretariat of the CBD to put forward their proposals for how the post-2020 Global Biodiversity Framework should advance and support their potential contributions. To date, this constituency has focused primarily on the need to ensure that the post-2020 Framework contains an explicit mandate for local and subnational action on the goals and targets agreed internationally, to replace the previous policy architecture agreed a decade ago. Second, and often in parallel, governance arrangements and initiatives concerned with the global governance of urban development have begun to recognize the potential value of urban nature. For example, The New Urban Agenda,Footnote 2 adopted in 2017, refers to the value of cities and human settlements that protect ecosystems and biodiversity as well as to the importance of encouraging nature-based solutions and innovations as part of urban development processes. Cross-cutting both arenas, initiatives and arrangements that are primarily concerned with the governing of climate change have increasingly signaled the potential of urban nature-based solutions as a means through which to address climate challenges as well as the biodiversity and urban sustainability agendas, for example in the report of the Global Commission on Adaptation and other initiatives highlighted at the 2019 UN Climate Action Summit that took place in New York (GCA and WRI, 2019; UNCC, 2019; UNDP, 2019). Across the UN environmental governance landscape, it is evident that the potential for urban responses to play an important role in transforming biodiversity governance is increasingly recognized.
At the same time, it is critical to recognize that the global architecture for urban biodiversity governance is not confined to the workings of international conventions, but also encompasses a range of actors and networks that operate transnationally. Of these, the first to be established (in 2006) was the Cities Biodiversity Centre, part of ICLEI Local Governments for Sustainability, who were appointed in 2019 as the representative of local and subnational governments within the CBD Secretariat. Over the past two years, ICLEI’s Cities Biodiversity Centre has partnered with the newly formed IUCN Urban Alliance and The Nature Conservancy to form the CitiesWithNature platform, intended to provide a focal point for urban action toward the post-2020 biodiversity agenda. To date, 174 cities from 58 countries have committed to action under the CitiesWithNature umbrella. The involvement of the IUCN and The Nature Conservancy in such initiatives is particularly significant, marking a growing interest in urban biodiversity from organizations that have traditionally been concerned with conservation and restoration in a relatively conventional sense and for whom cities have been marginal to their interests. A similar urban biodiversity initiative was launched by the Secretariat of the Ramsar Convention in 2017 – the Wetland City Accreditation scheme.Footnote 3 In Europe, a number of urban projects designed to develop and implement urban nature-based solutions are being supported under the Horizon 2020 Sustainable Cities and Communities program, with a total budget of approximately 200 million Euros. These transnational initiatives primarily seek to enhance the ways in which cities are governing biodiversity within their own territories, creating a means through which both urban biodiversity planning and the increasingly diverse forms of experimentation that cities are deploying to govern biodiversity are recognized, aggregated and shared, and learning between cities is fostered. In this way, they both benefit from the fragmentation of authority to govern urban biodiversity and, through fostering new and more varied initiatives, serve to contribute toward it.
A further, if currently embryonic, trend is the emergence of transnational initiatives that are seeking to engage cities in addressing their contribution to the underlying drivers of biodiversity loss and in so doing contributing to governance for transformation – primarily through taking measures either to support ecosystems on which cities depend or to improve the sustainability of production and consumption. The World Resources Institute has developed the Cities4Forests initiative, aimed both at improving the quality and quantity of urban forest biodiversity and enhancing the role that cities play in protecting “nearby” and “faraway” forests. One of the sixty cities who have now joined this initiative, Raleigh in North Carolina, USA, has developed a water levy to pay for partnership work with upstream landowners to protect water quality in the catchment from which it draws its own water supply. As well as taking measures to protect its surrounding forest area, Kigali, Rwanda, another member of the initiative, is partnering with the Rwandan Ministry of Environment to fulfill the aim of planting trees across 43,000 hectares of land nationwide. In October 2020, twenty-six cities came together under the European Circular Cities Declaration, founded by The Collaborating Centre on Sustainable Consumption and Production (CSCP) together with ICLEI, the Ellen MacArthur Foundation, Eurocities, UNEP and other partners to accelerate the transition to a circular economy at the city level in order to reduce their impact on climate change and biodiversity. The long-established C40 Cities Climate Leadership Group is currently promoting the use of nature-based solutions to enhance building efficiency and the adaptive capacity of cities in the face of climate change, while its Food System Networks promotes regenerative urban agriculture to decrease production emissions, close yield gaps, increase food security, support local producers, decrease food miles, mitigate urban heat island effect and reduce building energy demand (through roof and wall gardens). What is notable in these initiatives is that biodiversity is often not positioned centrally to urban actions, but rather that potentially transformative forms of governing biodiversity through urban action are emerging as a “co-benefit” of urban efforts to reconfigure their economies and address climate change. Such outcomes are therefore being generated through the fragmentation, rather than integration, of governance.
There is therefore a growing density and diversity in the multilevel governance arrangements, networks, initiatives and projects through which the urban governing of biodiversity is taking shape. Taken together, the growing governance complex through which the governing of urban biodiversity is taking place, as well as the diversification of modes of governing through which it is being implemented, suggest that this is an arena of biodiversity governance that has been substantially transformed over the past decade. The transformation of the architectures, arrangements, networks and substance of urban nature governing, away from a specific form of urban planning concerned primarily with nature conservation and largely isolated from wider urban sustainable development and climate change goals and toward a much more fragmented, multiple and encompassing approach, not only represents a transformation in the governance dynamics at play, but has arguably also served to shift the governing of urban biodiversity on to a more transformative footing. By bringing a whole host of new actors into the realm of urban biodiversity governance and transforming both the capacities and purpose of governing biodiversity in the city, the transformation of urban biodiversity governance is arguably paving the way for a more transformative approach to biodiversity on the ground.
14.4 Conclusions
Cities hold considerable potential for conserving and restoring biodiversity, and will be critical to ensuring that society can thrive by preserving and enhancing nature’s contribution to people. As we discussed in this chapter, there is now a growing realization of the importance of urban governance for nature. Of the themes of transformative governance raised in the Introduction to this book, we find most evidence of transformations in governance when it comes to the role of cities in biodiversity governance. First, we have argued that biodiversity governance is being transformed within cities themselves. Rather than being confined to urban planning, we find a growth of urban experimentation as various initiatives and nature-based solutions are now being undertaken by municipal authorities and their partners, as well as a range of private and community actors, to protect, restore and thrive with nature. Second, the growing recognition of cities as key agents of change and as presenting opportunities for governing biodiversity represents a transformation in biodiversity governance internationally, which has traditionally focused on cities as a threat to biodiversity and has tended to be dominated by a focus on the nation-state. This in turn is leading to a transformation in the global architecture for biodiversity governance, such that cities are now given more prominence within the global Convention on Biological Diversity. In parallel, we witness a growth of transnational networks seeking to both advocate for cities within international fora and to foster urban responses, a phenomenon both generated by and contributing to the fragmentation of authority to govern urban nature. In short, the rise of cities on the biodiversity agenda is leading to transformations in how and by whom biodiversity is governed both within the urban arena and beyond.
However, some fundamental issues persist and form the key challenges that will need to be addressed if we are to realize a transformation in how urban biodiversity governance is pursued and to what ends – in short, if we are to generate governance for transformations. The first issue concerns how matters of biodiversity can become mainstream within urban development and how cities come to be positioned within biodiversity governance (and vice versa). Despite a growing recognition of its importance, biodiversity is relatively marginalized in policymaking and planning in cities. Among most of those transnational networks/initiatives that incorporate biodiversity goals and targets, biodiversity is usually regarded as a “co-benefit” of urban efforts to reconfigure their economies and address climate change. This not only limits the attention given to biodiversity per se, but also means that the underlying drivers of biodiversity loss beyond the city limits receive limited attention – for example in terms of cities addressing the impacts of their consumption or of waste in terms of their effects on the loss of biodiversity or in terms of how they compromise the capacity of other communities to realize the benefits of nature. On the one hand, a continued emphasis on the win–win potential of initiatives for addressing biodiversity while also attending to other critical urban priorities will be necessary to maintain its position on the urban agenda, yet at the same time it will be crucial that cities come to see themselves as having a fundamental role in governing nature within and beyond their own boundaries through further embedding this issue in key policy arenas and through the actions of critical stakeholders in urban development. We suggest that it is unlikely that governance for transformative action that addresses the underlying urban drivers of biodiversity loss will be found through existing institutions, but will rather require new coalitions and partnerships that bring urban actors together with those in the business and finance sectors as well as through place-to-place partnerships. Rather than expecting this to be a fully joined-up or integrated process, as with the climate agenda, we might witness a growing fragmentation and complexity of governance in order to address the critical issue of transformative change.
Second, and related, a transformative approach to biodiversity governance would necessarily need to challenge which forms of urban nature come to count in the pursuit of urban sustainability. As nature-based solutions are gaining traction, the delicate relationship between nature and society that coexists within cities becomes particularly salient, even if such forms of “hybrid nature” are not afforded much value in terms of conservation or do not represent the restoration of previously lost ecologies. Cities are spaces for new kinds of mundane nature that bring significant worth to everyday life and also provide the space for novel ecologies that consist of what might be termed invasive or non-native species, around which forms of human and nonhuman association and community are often developed. Questioning which forms of nature are seen to belong or are to be excluded from the city, by whom and to what purpose, in turn might lead to a transformation in how urban biodiversity should be understood, conserved, restored and prioritized in order that diverse communities can thrive with nature. Such an effort will require more inclusive forms of governance, as suggested in the Introduction to this book, but it also suggests that we will need to leave space open for dissent, contestation and protest in order to realize transformative governance for biodiversity.
Last but not least, how issues of social exclusion and injustice can be addressed (rather than exacerbated) is a significant problem, but one that must be solved if biodiversity governance is to become truly transformative. While a focus on inclusive governance points to the importance of ensuring equitable processes, governance for transformation also requires that we focus on the outcomes that are generated through interventions for biodiversity governance and how such forms of governance either serve to reproduce or challenge existing socioeconomic and power inequalities. Given that some nature-based solutions projects risk excluding minority or Indigenous communities in the project design and implementation process, displace residents who cannot afford the resulting rising house prices and can serve urban elites at the expense of others, there is a growing concern that the governing of urban nature will entrench forms of neoliberal economic development and social exclusion. Transformative biodiversity governance will necessarily involve a fundamental reordering of structures of power and knowledge that can enable social and environmental justice to be secured and enhanced, and as such is likely to be highly contested and often contradictory and fragmented. Focusing on the underlying drivers of the loss of biodiversity and the diminished and unequal contributions that nature makes to people will, as other contributions to this volume make clear, be necessary if governance is to be transformative. This in turn suggests that it will not be sufficient for global institutions and transnational networks to promote urban action on nature, but that they will need to play a critical part in building the capacity and vision needed for cities to ensure that they take action for nature within and beyond urban boundaries that not only contributes to global biodiversity goals but also ensures social justice.
15.1 Introduction
The ocean’s enormity and depth are illustrated by the limited ability of humankind to comprehend it. The current science and policy seascape remains largely fragmented, and as a result the integrity of marine life and the well-being of those (human and nonhuman) dependent on a healthy ocean is being negatively impacted. Fragmented governance is an indirect driver of ocean biodiversity loss due to its inability to provide synergistic solutions to address simultaneously multiple direct drivers for such loss (overfishing, land-based and marine pollution, and climate change). This governance problem is well known (Reference Kelly, Ellis and FlanneryKelly et al., 2019; Reference Watson-Wright, Valdés, Werle, Boudreau and BrookesWatson-Wright and Valdés, 2018), and to some extent it is being addressed in ongoing international negotiations on an international instrument on marine biodiversity of areas beyond national jurisdiction (A/RES/72/249, 2017).
This chapter will shed new light on these well-known problems by applying the lens of “transformative governance,” understood as “formal and informal (public and private) rules, rule-making systems and actor-networks at all levels of human society (from local to global) that enable transformative change … towards biodiversity conservation and sustainable development more broadly,” with a view to “respond[ing] to, manag[ing], and trigger[ing] regime shifts in coupled socio-ecological systems at multiple scales” (Reference Visseren-Hamakers, Razzaque and McElweeVisseren-Hamakers et al., 2021: 21; see also Reference Chaffin, Garmestani and GundersonChaffin et al., 2016 and Chapter 1 of this volume). We share the editors’ views that there is a need to shift away “from the technocratic and regulatory fix of environmental problems to more fundamental and transformative changes in social-political processes and economic relations” (Reference OtsukiOtsuki (2015: 1; see also Chapter 1 of this volume). This can also help us to better understand how ocean biodiversity can contribute to “other environmental and social justice issues”Footnote 1 that are interwoven with the ocean in less visible ways than terrestrial biodiversity, such as poverty (Reference Singh, Cisneros-Montemayor and SwartzSingh et al., 2018) and resource-grabbing (Reference Virdin, Vegh and JouffrayVirdin et al., 2021).Footnote 2
In particular, the chapter will illustrate the broad recognition of the vital need for integrative and inclusive governance of ocean biodiversity, to ensure that solutions also have sustainable impacts at other scales and in other sectors, and to empower those whose interests are currently not being met and represent transformative sustainability values.Footnote 3 The complementary roles of adaptive governance (enabling learning, experimentation, reflexivity, monitoring and feedback) and anticipatory (precautionary) governance will also be touched upon. The latter has been extensively debated in international legal scholarship (Reference GustonGuston, 2014; Reference Birnie, Boyle and RedgwellBirnie et al., 2009), so we will reflect on how the former can contribute to the latter. Fundamentally, however, the chapter will focus on the role of transdisciplinary governance (the recognition of different knowledge systems and the inclusion of underrepresented types of knowledge) in supporting integration, inclusion and learning in ocean affairs for transformative change.
Accordingly, this chapter will first engage in a brief analysis of the major underlying causes of marine biodiversity loss, by drawing on global synthesis reports. Second, considering the extensive literature assessing existing regulatory mechanisms and their effects on the status and uses of marine biodiversity, this chapter proposes to focus specifically on the lessons learned for transformative ocean governance in the context of area-based management and spatial planning from the international to the local level. Finally, an alternative governance approach will be proposed as a possible way forward, building on the factual and legal interdependencies between human rights and marine biodiversity. The chapter will suggest taking a broader approach to fair and equitable benefit-sharing to shift toward transformative governance for the ocean at different scales.
15.2 Marine Biodiversity Loss: Causes and Consequences
The ocean is an integrated physical and biological system that provides a multitude of planetary services. These include the provision of half of the oxygen we breathe, absorption of 26 percent of anthropogenic CO2 emissions from the atmosphere, and rich and diverse life (UNGA, 2016: A/70/112). The full extent of the ocean’s biodiversity is not fully known or understood, but there is sufficient knowledge indicating that marine life is declining dramatically, albeit not yet irreversibly (Reference Serrao-Neumann, Davidson and BaldwinSerrao-Neumann et al., 2016). Additionally, we have limited understanding of the intrinsic, as well as the social and cultural, values of marine biodiversity, and its multiple contributions to human identity and well-being (IPCC, 2019).
The causes of marine biodiversity loss are numerous, pervasive and interconnected. Globally, the major direct drivers include overexploitation, climate change and pollution. The increasing number of zoonotic pathogens associated with biodiversity loss is also affecting marine life, as well as humans (Reference Morand, Lajaunie, Morand and LajaunieMorand and Lajaunie, 2017). Examples include outbreaks of influenza in seabird populations, and distemper morbillivirus in seal colonies (Reference Bogomolni, Gast and EllisBogomolni et al., 2008; Reference Morand, Lajaunie, Morand and LajaunieMorand and Lajaunie, 2017; Reference Waltzek, Cortés-Hinojosa, Wellehan Jr. and GrayWaltzek et al., 2012). This led to calls for a more comprehensive global approach in 2020 as the COVID-19 pandemic raged (Reference CorlettCorlett, 2020; Reference OstfeldOstfeld, 2009), and serves as a reminder of the links between human well-being and healthy, resilient ecosystems. The following subsections will explore threats to marine biodiversity on the basis of seminal global scientific assessments (UNGA, 2016: A/70/112; FAO, 2020; IBPES, 2019; IPCC, 2019).
15.2.1 Exploitation of Living and Nonliving Marine Resources
The exploitation of marine resources has brought about the largest relative impact on biodiversity since 1970 (IPBES, 2019). Illustrative examples may be drawn from fisheries and aquaculture, as well as the projected impacts of commercial mining activities in the deep seabed, all of which can contribute to habitat and biodiversity loss in the ocean.
Fishing has had the most impact on marine biodiversity in the past fifty years, including impacts across scales on target and nontarget species, habitats and ecosystems (IPBES, 2019). Combined with the effects of climate change, fishing is expected to remain a leading driver in worsening the state of marine biodiversity (IPBES, 2019). Funded by harmful government subsidies, commercial fishing fleets have expanded geographically and into deeper waters that were previously not financially viable to exploit (IPBES, 2019; Reference Sumaila, Ebrahim and SchuhbauerSumaila et al., 2019), directly contributing to a global decline in fish stocks (FAO, 2020). Fishing above sustainable levels causes negative impacts on marine biodiversity and reduces fish productivity and ecosystem functioning (FAO, 2020). Bycatch caused by nonselective fishing methods impacts marine biodiversity, and some fishing gear, such as bottom trawls and pelagic drift nets, also cause damage to habitats and biodiversity. The United Nations has recognized that the threat of illegal, unreported and unregulated (IUU) fishing goes beyond the depletion of fish populations, and there is a close nexus between the illegal activities in fisheries and transnational organized criminal activity, known as fisheries crime (A/63/111, 2008).Footnote 4 Fisheries crime threatens fish stocks and undermines the international goal to conserve and use the ocean for sustainable development (A/RES/70/1, 2015; A/RES/60/31/2006). Finally, the impacts of fisheries crime are being exacerbated by climate change (Reference CheungCheung, 2016; IPBES, 2019; NIC, 2016).
Aquaculture, whether it is coastal farming or offshore aquaculture (Reference HolmerHolmer, 2010), has been promoted as a means to address both overfishing and food security, but may have a negative impact on the environment and biodiversity, mainly arising from excess feed, pesticides and medicines leaching into the marine environment (Reference Tovar, Moreno, Mánuel-Vez and Garcı́a-VargasTovar et al., 2000). Aquaculture may affect ecosystems and biodiversity with the loss of critical habitats like mangrove or wetlands, with consequences for coastal protection (Reference Páez-OsunaPáez-Osuna, 2001), or the alteration of hydrologic regimes by the use of structures such as fish cages (Reference Eng, Paw and GuarinEng et al., 1989). The intensification of aquaculture has a dramatic effect on seabed fauna and their abundance (Reference DianaDiana, 2009; Reference Tsutsumi, Kikuchi and TanakaTsutsumi et al., 1991). In turn, coastal pollution (agriculture, hydrocarbon, heavy metals) and marine pollution affect the success of aquaculture (Reference Eng, Paw and GuarinEng et al., 1989).
15.2.2 Pollution
Pollution is the direct or indirect introduction by humans of substances that result or are likely to result in deleterious effects to the environment (UNCLOS, Art. 1(4)). Marine and coastal areas are highly vulnerable to pollution from activities on land or at sea, which have a direct impact on marine biodiversity. Land-based pollution comes in many forms, including nutrient run-off (untreated sewage), agricultural and industry run-off such as pesticides, heavy metals or oils entering river systems and then the open ocean (UNEP/EA.4/Res.11, 2019). Marine pollution can come from a variety of activities at sea, including plastics from discarded fishing gear, dumping from vessels and underwater noise (UNEP/EA.3/L.19, 2018).Footnote 5 Marine environmental pollution has gathered international attention, as captured in Sustainable Development Goal (SDG) 14.1: “By 2025, prevent and significantly reduce marine pollution of all kinds, in particular from land-based activities, including marine debris and nutrient pollution.”
Plastic pollution is pervasive in the marine environment, and the widespread impacts of macro- and microplastics on marine biodiversity at all levels are sobering. Addressing plastic pollution presents a complex governance challenge and is subject to intensified international attention. For example, the UN has highlighted the pervasive nature of plastic pollution, highlighting that between 4.8–12.7 million tons of plastic enters the ocean annually (UNEP/EA.3/L.19, 2017). The vast majority of this (~80 percent) is from land-based sources,Footnote 6 while the rest comes from maritime activities, including fishing (Reference Isensee and ValdesIsensee and Valdes, 2015), which requires stronger monitoring and control by states to prevent plastic entering ocean systems (Reference HawardHaward, 2018) and potentially new measures at the international level (Reference Borrelle, Rochman and LiboironBorrelle, et al., 2017).
Deep-seabed mining for minerals and rare-earth metals at a commercial scale occurs in areas within national jurisdiction and may soon be a reality in the Area (which is the seabed beyond the jurisdiction of any state; one of the two areas outside national jurisdiction, together with the high seas) (Reference Casson, Alexander and MillerCasson et al., 2020).Footnote 7 Noise and light pollution, as well as sediment plumes, may have a harmful effect on marine species, while the mining itself may permanently destroy deep-sea habitats and may impact communities relying on fish stocks, with potential human rights implications (Reference Miller, Thompson, Johnston and SantilloMiller et al., 2018). Deep-sea sediments act as long-term stores of atmospheric carbon, meaning mining activities may pose an additional climate risk by releasing carbon through sediment disturbance (Reference Sala, Mayorga and BradleySala et al., 2021).Footnote 8 Climate change is also predicted to alter deep-ocean environments and to be exacerbated by other deep-sea extractive activities such as oil and gas extraction and bottom fishing (Reference Levin, Wei and DunnLevin et al., 2020).
15.2.3 Climate Change
There is scientific consensus that human-induced climate change is altering the physical and chemical makeup of the ocean (Reference Stocker, Qin and PlattnerStocker et al., 2013). The main impacts of climate change on the ocean are warming (IPCC, 2019), acidification and deoxygenation, which simultaneously occur due to increasing carbon dioxide (CO2) and other greenhouse gas emissions (Reference Beaugrand, Edwards and RaybaudBeaugrand et al., 2015; Reference Molinos, Halpern and SchoemanMolinos et al., 2016). These changes are expected to persist throughout this century, as levels of CO2 increase to those unseen in human times (Reference Gattuso, Magnan and BilléGattuso et al., 2015). Transformative governance has thus been recommended by the Intergovernmental Panel on Climate Change to address and adapt to these issues (IPCC, 2018).
The consequences of climate change on marine biodiversity include species extinction, local changes in species richness, proliferation of invasive species, ecosystem collapse, and disruption of ecosystem functioning and services (Reference Beaugrand, Edwards and RaybaudBeaugrand et al., 2015; Reference Cheung, Lam and SarmientoCheung et al., 2009; FAO, 2018; IPCC, 2019; Reference Molinos, Halpern and SchoemanMolinos et al., 2016). In addition, climate change is projected to decrease net ocean primary production and fish biomass (IPBES, 2019). Changes in the distribution of fish populations from historical locations can affect livelihoods, income and food security (IPCC, 2019), and increase conflicts between fishers, communities, authorities and states, highlighting a need for adaptive governance in the conservation and management of marine species (Reference Spijkers, Singh and BlasiakSpijkers et al., 2019; SROCC, 2019).
Roughly half of the CO2 emitted by anthropogenic activities between 1800 and 1994 is stored in the deep ocean as organic matter from absorption by planktonic organisms (Reference Sabine, Feely and GruberSabine et al., 2004). Since 1980, this uptake has been between 20 percent and 30 percent of total anthropogenic CO2 emissions, causing an increase in ocean acidification (IPCC, 2019). Acidification of the ocean decreases its ability to uptake and store carbon (IPBES, 2019), and leads to habitat destruction, with coral reef ecosystems particularly under threat (IPCC, 2019), alteration of marine food webs (Reference Feely, Sabine and LeeFeely et al., 2004; Reference Kleypas, Buddemeier and ArcherKleypas et al., 1999) and sensory perception changes in marine species (Reference Dixson, Munday and JonesDixson et al., 2010; Reference Munday, Dixson and DonelsonMunday et al., 2009; Reference Munday, Dixson and McCormick2010).
As a result of both climate change and pollution, ocean deoxygenation has become a pervasive yet overlooked issue. Deoxygenation is caused by the warming of ocean waters, from agricultural run-off into rivers and from the atmosphere from the burning of fossil fuels (Reference Laffoley and BaxterLaffoley and Baxter, 2019). This causes species loss, resulting in changes in ecosystem structure and function (Reference Laffoley and BaxterLaffoley and Baxter, 2019). There has been a marked loss in ocean oxygen levels from the surface to 1000 m depth since 1970, leading to the prevalence of oxygen minimum zones, which are uninhabitable for many marine species (IPCC, 2019).
15.2.4 Lessons Learned
While our global understanding of the multiple threats to marine biodiversity is growing, ocean science is “still weak in most countries” due to limited holistic approaches for understanding cumulative impacts of various threats, and lack of capacity to conduct science (A/71/733, 2017). Low- and middle-income countries face the greatest challenges in this regard: to prevent and mitigate negative development impacts connected to the ocean, participate in traditional and emerging ocean activities (Reference Blasiak, Jouffray and WabnitzBlasiak, 2018), and predict and harness the socioeconomic benefits of ocean conservation (Reference Blasiak, Jouffray and WabnitzBlasiak, 2018). As a result, scientific understanding of the effectiveness of conservation and management responses is poor, meaning it is more difficult to predict the productivity limits and recovery time of marine ecosystems in these countries. Meanwhile, the negative social, economic and cultural impacts of degraded mangroves and corals on local communities are increasingly noted (CBD, Decision XII/23, 2014), as are the negative impacts of declining fisheries on the human rights to food and culture (A/67/268, 2012). The urgency of advancing ocean science, in and to the benefit of all countries, is expected to take centerstage globally, with the UN declaring 2021–2030 as the Decade of Ocean Science for Sustainable Development (UNESCO, 2020).
This situation is compounded by limited efforts to bridge different knowledge systems (notably Indigenous and local knowledge), which contributes to marginalizing these knowledge holders from relevant decision-making, even if these groups are disproportionally affected by the negative consequences. Furthermore, limited understanding of the benefits that derive from a healthy ocean for society and the economy fuels a “disconnect” between some communities and the ocean (Reference Jamieson, Singleman, Linley and CaseyJamieson et al., 2021). In effect, only recently have global scientific reviews highlighted the multiple dependencies of people’s right to health on the marine environment (WHO/CBD, 2015; A/HRC/34/49, 2017; A/75/161, 2020).
From a transdisciplinary governance perspective, all the facts observed and anticipated scenarios in the global reports analyzed above are not equally known, and even less equally predictable. For instance, if the recent rate of fishing capture is maintained, the collapse of some fisheries is almost certain, while others, especially close to the shores of the more important fishing nations, have already collapsed, leading these states to travel greater distances, thereby replicating the process elsewhere. It is also projected that the warming and acidification of the ocean will exacerbate this. In contrast, the severity and the intensity of the impacts that will result from deep-sea mining is very difficult to evaluate, as are as the effects of all the occurring changes that are cascading through unpredictable interactions. Here, the limited predictability of changes in the state of the ocean and marine resources is not a matter of observation, monitoring techniques or models (Reference MazzegaMazzega, 2018). Rather, unpredictability is intrinsic to the complex dynamics of the ocean system, emphasizing the need for ocean governance to be anticipatory and adaptive.Footnote 9
Furthermore, while the main trends summarized above represent scientific consensus, these global syntheses of current knowledge are based on a small fraction of the volume of articles annually published on these themes.Footnote 10 The limitation of these systematic reviews is of particular concern because the impacts of human activities and environmental changes on biodiversity are for the vast majority manifesting at relatively local scales, in specific ecosystems or biomes. They require careful observations and analysis in context (Reference Allan, Weisser and FischerAllan et al., 2013).
15.3 An Assessment of Existing Mechanisms for Ocean Governance
The international legal framework for the ocean is considered “critical” to make progress in all target areas of SDG 14: “life below water” (A/71/733, 2015). The international framework, though, is notoriously so complex and fragmented (sectorally and geographically) that it presents colossal challenges to effective, let alone transformative, ocean governance. To an extent, fragmentation is the result of historical processes of international lawmaking. The earliest marine treaties focused on clarifying the rights and obligations of states over portions of the ocean,Footnote 11 establishing safeguards,Footnote 12 regulating discharge of wastes and pollution from shipping,Footnote 13 and managing fishing resources. The next wave of treaties prioritized specific objectives, including the protection of (marine) species.Footnote 14 However, the narrow scope and diverse approaches encapsulated within these instruments often failed to consider the impacts on ecosystems in a holistic and integrated manner (Reference KimballKimball, 2001; Reference MossopMossop, 2007). As these treaties resulted in a patchwork approach to marine management, early attempts at integrated ocean governance began with the negotiations of the 1982 United Nations Convention of the Law of the Sea (UNCLOS).Footnote 15
UNCLOS, commonly referred to as the “constitution of the oceans,” firmly embodies elements of customary international law, as well as several innovative features for a more comprehensive approach to the regulation of ocean activities, including on the basis of a general obligation to protect and preserve the marine environment. UNCLOS, however, heavily relies on other international instruments and mechanisms, thereby confirming the continued relevance of sectoral and regional governance approaches.
Table 15.1 Main biodiversity-related changes
Direct drivers (climate change: CC; fisheries: F; exploitation of nonliving resources: E), spatial scales (local, regional, global), concerned conventions and organizations analyzed in the chapter. An x indicates that the authors understand the conventions concerned, or the decisions adopted under them, or the instruments deployed by the organizations have sought to address these changes and drivers. A question mark indicates the conventions or their decisions may be applicable to these changes and drivers, but need further study. The table is meant as a basis for discussion with other legal and nonlegal experts, as the understanding of governance landscape may be subject to differing interpretations.
For instance, the UN Fish Stocks Agreement (UNFSA) implements UNCLOS Articles 63–68, and 116–120 on straddling and highly migratory fish, and sets out obligations to ensure sustainable fishing activities and mitigate the impacts of fishing on the marine environment and biodiversity, applying the precautionary principle when scientific information is inadequate or absent (Art. 6). UNFSA, in turn, is significantly underpinned by regional, collaborative approaches (Arts. 9 and 15). Arguably, therefore, UNFSA both requires, and sets the conditions for, an integrative, anticipatory and inclusive approach at the regional level, which, with the correct synergies, may be scaled up to the global level. Examples of such approaches will be discussed in Section 15.2.3.
While UNCLOS reflects to some extent the evolution of natural sciences and ecosystem management by referring to the interrelatedness of the problems of ocean spaces and the need to consider them as a whole, a parallel legal development under international environmental law has also contributed to a more integrative and inclusive approach to ocean governance. This is the case of the Convention on Biological Diversity (CBD)Footnote 16 and its objectives of conservation, sustainable use, and fair and equitable benefit-sharing (Reference Morgera and RazzaqueMorgera and Razzaque, 2017). Over the years, the CBD has provided integrative tools to complement earlier biodiversity-related treaties, including the Convention on International Trade in Endangered Species of Wild Fauna and Flora (CITES) and the Convention on the Conservation of Migratory Species of Wild Animals (CMS) (UNEP-WCMC, 2012), and contributed to addressing the nexus between the ocean, climate change and biodiversity (Reference MorgeraMorgera, 2011; Reference Diz, Morgera and RazzaqueDiz, 2017). It has also addressed an increasing number of new and emerging human activities that pose challenges to biodiversity conservation and sustainable use, such as renewables development, which can increase demands for ocean space (UNCTAD/DITC/TED/2014/5). In doing so, the CBD has also addressed the specific concerns of Indigenous peoples and local communities (IPLC), and highlighted the importance of their knowledge (Reference MorgeraMorgera, 2020), thereby contributing to defining inclusive and transdisciplinary ocean governance.
These developments have occurred under the CBD ecosystem approach (CBD Decisions V/6, 2000; VII/11, 2004), which aims at integrating the management of land, water and living resources, and balancing the three objectives of the Convention, as well as integrating different legal and management strategies, depending on local, national, regional or global conditions (CBD Decision V/6, 2000, Annex, para. 5), through adaptive management and precaution (thereby contributing to adaptive and anticipatory governance) (Reference MorgeraMorgera, 2011). The ecosystem approach also aims to integrate modern science and Indigenous and local knowledge (CBD Decision V/6, 2000, Principle 11), as well as equity concerns, recognizing that human beings and their cultural diversity are an integral component of many ecosystems (CBD Decision V/6, 2000, para. 2). Under this umbrella, one of the key obligations under the CBD is to establish a system of protected areas (CBD, Art 8[a]). This was complemented with a target of a 10 percent increase in marine protected areas (MPA) coverage by 2020 among the Aichi Biodiversity TargetsFootnote 17 by implementing effective and equitable protection of marine and coastal areas, particularly those important for biodiversity and ecosystem services (Aichi target 11).Footnote 18 Scientific guidance for the development of representative MPA networks had been previously adopted by CBD Parties in 2008 (CBD Decision X/2, 2010, target 11),Footnote 19 and “ecologically or biologically significant marine areas” (EBSAs) have been described by states as meeting the scientific requirements to benefit from enhanced conservation and management measures, protected status and impact assessments.Footnote 20 That said, commentators (Reference Diz, Johnson and RiddelDiz et al., 2018) have underscored that while progress has been made toward the 10 percent target in quantitative terms, the qualitative elements of the MPA target (effectively and equitably managed, ecologically representative and well connected systems), which would contribute to inclusive and integrative governance, have received far less attention (Reference Rees, Foster, Langmead, Pittman and JohnsonRees et al., 2018).
Also linked to the ecosystem approach, the guidance elaborated under the CBD in relation to marine spatial planning places a focus on the need to identify stakeholder roles and interests, promoting a deeper understanding of their dependence on ecosystem services, enhancing collaboration across different cultures, and demonstrating fairness, transparency and inclusiveness, including by employing a long-term historical perspective on how current conditions and issues evolved in a given area (CBD Decision XIII/9, 2016). This approach can address one of the main sources of opposition to the creation of MPAs: rather than pitting conservation against fisheries as competing interests, it could support the co-development of MPAs as integral components of ecosystem-based fisheries management (Reference Rees, Sheehan and StewartRees et al., 2020). This approach can also support the fair and equitable sharing of benefits arising from the establishment of MPA networks (discussed in Section 15.3.1) with ecosystem stewards and traditional knowledge holders, thereby contributing to integrative, inclusive and transdisciplinary governance (Reference MorgeraNtona and Morgera, 2018).
15.3.1 A Common but Differentiated Strategy: The Use of Area-Based Management Tools in Achieving Integrative Governance of the Ocean
UNCLOS,Footnote 21 as well as treaties aimed at improving safety at sea,Footnote 22 support area-based management tools (ABMTs)Footnote 23 such as MPAs (Reference Baxter, Laffoley and SimardBaxter et al., 2016; Reference De SantoDe Santo, 2018; Reference WarnerWarner, 2019), and previous experiences led by regional organizations serve to illuminate key opportunities and challenges (Reference De SantoDe Santo, 2018). ABMTs have in effect been promoted from early on in the regional context, most notably through the Regional Seas Programme, which was birthed from early attempts by UNEP to catalyze a more specialized and integrated methodology at the regional level (Reference Akiwumi and MelvasaloAkiwumi and Melvasalo, 1998).Footnote 24 Described as one of UNEP’s most significant achievements in the past thirty-five years,Footnote 25 the concept’s linchpin is to engage neighboring countries in comprehensive and specific actions for the sustainable management and use of the marine and coastal environment (A/9625, 1974). An additional advantage of the framework is the opportunity that the Regional Seas Programme provides stakeholders to share experiences and support more integrative ocean governance. For instance, relevant states participating in the regional seas Abidjan Convention in West Africa have cooperated with the Benguela Current CommissionFootnote 26 for the management of the Benguela Large Marine Ecosystem (Reference Cochrane, Augustyn and FairweatherCochrane et al., 2009), and the OSPAR Convention for the Protection of the Marine Environment of the North-East AtlanticFootnote 27 provides almost complete coverage of the Eastern Atlantic.Footnote 28 This has led to exchanging knowledge and capacity, as well as ensuring coherent implementation of the ecosystem approach, beyond the scope of the respective conventions. That said, there is widespread understanding that UNCLOS provides limited guidance on MPA networks, and progress has been too limited in areas beyond national jurisdiction. For these reasons, ABMTs are currently being addressed in international negotiations on a new international instrument on marine biodiversity of areas beyond national jurisdiction (Reference De SantoDe Santo, 2018).
Regional fisheries management organizations (RFMOs) have also established ABMTs. The advantage of RFMOs is that they can adopt targeted management measures that are adapted to the political and ecological characteristics of a given region. The key difference with regional seas organizations is that RFMOs can adopt measures that are binding on their member states. Many RFMOs now include an ecosystem and precautionary approach to fisheries.
While such provisions do not confer upon RFMOs the mandate to regulate activities other than fisheries, they generally allow them to conduct cumulative impact assessments to evaluate the aggregate effects of human activities on the ecosystems in their regulatory area.
Nevertheless, RFMOs are still not cooperating with other organizations to the extent necessary to ensure cross-sectoral cooperation for MPAs, other area-based management and risk assessments “in adopting integrated and coherent conservation and management measures within ecologically meaningful boundaries (or ecosystem-based units/ functional units)” (Reference Diz and NtonaDiz and Ntona, 2018: 19; Reference Kenny, Campbell and Koen-AlonsoKenny et al., 2018). Thus, their sector-focused approach to management still poses an obstacle to the integrated management of fisheries (Reference Leroy and MorinLeroy and Morin, 2018; Reference Pentz, Klenk, Ogle and FisherPentz et al., 2018).
For that reason, synergies between the Regional Seas Programme and RFMOs have been pursued. One approach has been to focus on large marine ecosystems (LMEs),Footnote 29 wide areas of ocean space along the planet’s continental margins, spanning 200,000 km2 or more. LMEs are another type of ABMT that include both ocean space and connected coastal land areas, such as river basins and estuaries (Reference Sherman and AlexanderSherman and Alexander, 1986), to maintain and restore ecosystem functions. As discussed in Section 15.3.2, the establishment of the Benguela Current Commission between Angola, Namibia and South Africa, as the three states that border the LME, is an example of transformative ocean governance. The connection between the Regional Seas Programme, RFMOs and LMEs is being deepened by the Sustainable Ocean Initiative, led by the CBD (CBD, 2016).
Against this background, a case study will serve to illustrate progress and continued challenges in creating MPAs as a leading ABMT methodology that is integral to marine spatial planning for balancing ocean uses to support sustainable development and enhance ocean governance. (Reference Finke, Gee and GxabaFinke et al., 2020a; Reference Kirkman, Holness and HarrisKirkman et al., 2019). The next subsection will thus identify lessons learned in ensuring integrative and inclusive ocean governance, understood as inclusivity of diverse representative species and biodiversity hotspots, as well as of varied human dependences on marine ecosystems through stakeholder engagement, securing of resource rights, and the recognition of Indigenous and local knowledge systems that can contribute to biodiversity conservation goals (Reference MacKinnon, Lemieux and BeazleyMacKinnon et al., 2015).
15.3.2 Experiments in Integrated and Inclusive Approaches: The Benguela Current Commission and South Africa’s MSP Process
The Benguela Current Commission is a notable example of integrating and upscaling efforts between the Regional Seas Programme, RFMO and a large marine ecosystem (CBD, 2016).Footnote 30 The establishment of the Commission resulted from the cooperation over two decades in ocean governance between Angola, Namibia and South Africa toward a multisectoral ocean governance approach.Footnote 31 Cooperation culminated in several international instruments, including the 1999 Strategic Action Programme for the Ecosystem, which was given effect through a voluntary 2007 Interim Agreement on the Establishment of the Benguela Current Commission.Footnote 32 This was to ensure effective longstanding transboundary cooperation and the sustainable management and protection of the LME (Reference O’Toole, Shannon, Hemper and ShermanO’Toole and Shannon, 2003). In 2013, the Interim Agreement was replaced by the Benguela Current Convention (BCC), cementing the legal status of the Benguela Current Commission.Footnote 33
Several remarkable features of the BCC make it a good basis for more inclusive and integrative ocean governance. First, the BCC addresses the complex legacy of fragmented governance left by colonial and political histories (Reference Cochrane, Augustyn and FairweatherCochrane et al., 2009), including Angola’s independence and forty years of debilitating war (Reference Cochrane, Augustyn and FairweatherCochrane et al., 2009), Namibia’s independence from South Africa,Footnote 34 and the end of apartheid in South Africa (Reference Finke, Gee and GxabaFinke et al., 2020a), with the social impacts spilling over into the establishment and effectiveness of South Africa’s MPA system (Reference Sowman and SundeSowman and Sunde, 2018).
Secondly, the Commission links the Benguela Current Large Marine Ecosystem with the neighboring Agulhas and Somali LMEs, which is vital, as these boundaries are highly dynamic and the neighboring warmer waters directly influence the Benguela ecosystem and its living marine resources (Reference Heileman and O’TooleHeileman and O’Toole, 2001).
Thirdly, the arrangement reinforces the framework under the Abidjan Regional Seas Convention, as well as relevant regional fisheries arrangements.Footnote 35 Finally, there is an established linkage between the Benguela Current Commission and the Orange-Senqu Commission that comprises the four riparian statesFootnote 36 fed by the largest river discharging into the Benguela LME (Reference Finke, Gee, Kreiner, Amunyela and BrabyFinke et al., 2020b). This in turn allows a link between ocean management and a wetland of international importance under the Ramsar Convention.Footnote 37
The BCC allows its members to manage transboundary resources holistically while balancing different ocean users’ needs with conservation imperatives. Its objective is “to promote a coordinated regional approach to the long-term conservation, protection, rehabilitation, enhancement and sustainable use of the [LME], to provide economic, environmental and social benefits” (BCC, Art. 2). According to the BCC, member states must be guided by principles on sustainable use and management, precautionary and prevention (BCC, Art. 4; Reference VranckenVrancken, 2011), thereby providing the legal basis for integrative and anticipatory governance.
Member states and the Commission are guided by a five-year Strategic Action Programme (Reference Hamukuaya, Attwood and WillemseHamukuaya et al., 2016), which addresses the following eight themes: living marine resources; nonliving marine resources; productivity and environmental variability; pollution; ecosystem health and biodiversity; human dimensions; enhance the economic development potential; and governance (Reference HamukuayaHamukuaya, 2020). The Strategic Action Programme is based on a transboundary diagnostic analysis, consisting of a scientific and technical assessment to identify important transboundary issues related to the marine environment and their impacts on the environment and socioeconomy of the region (Reference Hamukuaya, Attwood and WillemseHamukuaya et al., 2016). Both instruments are reviewed and updated every five years.Footnote 38 The Commission included marine spatial planning into its 2015–2019 Strategic Action Programme (Reference Finke, Gee and GxabaFinke et al., 2020a) to support a variety of ecosystems and sectors, make contributions to the existing economies of member states and tackle increasing demands on the region’s marine space (Reference Finke, Gee, Kreiner, Amunyela and BrabyFinke et al., 2020b). This is in line with the progress already made under the Benguela Ecologically or Biologically Significant Areas Project (Reference Kirkman, Holness and HarrisKirkman et al., 2019), the Second National Biodiversity Strategy and Action Plan (to implement the CBD) of NamibiaFootnote 39 and Angola,Footnote 40 and the three countries’ commitment to implementing an ecosystem approach to fisheries (Reference Kirkman, Blamey and LamontKirkman et al., 2016).
Through the Benguela Current Commission, a regional working group for MSP was established to foster cooperation between different stakeholders (Reference Finke, Gee and GxabaFinke et al., 2020a), including government officials, technical experts and representatives of civil society, supporting the implementation of MSP within the three states and enabling information exchange, mutual learning and capacity-building in the form of expertise (Reference Finke, Gee and GxabaFinke et al., 2020a). These are not limited to the region. The regional working group has engaged with the European Commission, the Baltic Marine Environment Protection Commission and the Baltic Sea Spatial Planning Organization (Reference Finke, Gee and GxabaFinke et al., 2020a). A valuable output from the regional working group is enabling a uniform approach to MSP in the region (Reference Finke, Gee and GxabaFinke et al., 2020a). For the successful implementation of MSP within the region, however, extensive data is required on the state of the marine area, the impact of human activities and the effect of external pressures such as climate change.
To date, the Benguela Current Commission has undertaken projects to inform the regional MSP process, such as the spatial biodiversity assessment of marine and coastal biodiversity in the ecosystem, focusing on the ecosystem threat status, ecosystem protection levels and priority areas for protection (Reference Holness, Wolf and LombardHolness et al., 2012). In addition, through the Marine Spatial Management and Governance Project (MARISMA), member states have been supported in describing the region’s EBSAs, in line with the CBD, as part of MSP.
The main challenge facing the Benguela Current Commission, in addition to lack of long-term funding for the MSP process, is how to engage with stakeholders across different sectors as part of its efforts to strategically organize the use of the marine space, to avoid conflicts and limit threats while ensuring the long-term sustainable development of the blue economy in the region.Footnote 41 The challenge facing the Commission is, therefore, encompassing inclusive, transdisciplinary and adaptive governance.
Regarding national efforts, there are currently no MPAs legislated in Angola.Footnote 42 In Namibia, the Namibian Islands are currently the sole MPA, but will be one of seven marine areas that have been described as an EBSA under the CBD (Reference Finke, Gee, Kreiner, Amunyela and BrabyFinke et al., 2020b). South Africa has legislated forty-four MPAs in line with Operation Phakisa MPA Network.Footnote 43 Of the three states, only South Africa promulgated legislation specifically on marine spatial planning (Marine Spatial Planning Act of 2018). Nevertheless, Namibia and Angola have established similar institutional structures to South Africa, enabling different government agencies to work together to implement MSP through the National Working Groups by using experts of the MARISMA project (Reference Finke, Gee, Kreiner, Amunyela and BrabyFinke et al., 2020b). The three states are thus developing plans sequentially to focus on one marine area at a time to integrate learning from one planning process into the next (Reference Finke, Gee, Kreiner, Amunyela and BrabyFinke et al., 2020b).
In South Africa, researchers and government partners have identified Algoa Bay in the Eastern Cape as a case-study area for developing the first marine spatial plan, with a view to using lessons learned for the development of marine area plans as set out in the Marine Spatial Planning Act (Reference Dorrington, Lombard and BornmanDorrington et al., 2018). Algoa Bay has been extensively researched and is home to government-funded research platforms, therefore providing a substantial body of data, allowing an understanding and management of the complexity of legal and socioeconomic requirements, on one hand, and environmental (physical, chemical and biological) considerations, on the other (Reference Dorrington, Lombard and BornmanDorrington et al., 2018). The development of the Algoa Bay marine spatial plan is following the Intergovernmental Oceanographic Commission of UNESCO (IOC-UNESCO) ten-step approach, underpinned by the CBD ecosystem approach principles, which include recognition of Indigenous knowledge systems (CBD Decision V/6, 2000, Principle 11). This case study can, therefore, become an entry point for recognizing human rights as part of the governance of the ocean and its resources, integrating different systems of knowledge. In addition, the case study is viewed through a systems approach lens and the development of system dynamic tools/models that provide opportunities for scenario-planning and determining possible inter-sectorial impacts and environmental impacts (Reference Lombard, Ban and SmithLombard et al., 2019). Algoa Bay, therefore, entails a research-stakeholder-led “enabling approach” to developing capacities for the “governance of transformations” (i.e. governance to actively trigger and steer a transformation process).Footnote 44 It aims to bring together natural science findings and methods across fisheries, marine ecology and oceanography, with social sciences, law and art to support transdisciplinary, integrative, adaptive and inclusive ocean governance. Algoa Bay provides an example that could be scaled up not only to the national but also the regional level, including with a view to supporting the Benguela Current Commission and the Western Indian Ocean in constructively engaging with stakeholders over trade-offs, by expanding their current integrative and anticipatory governance approaches to include inclusive, adaptive and transdisciplinary approaches. Lessons learned are providing guidance for the development of the Western Indian Marine Spatial Planning Strategy (Reference Lombard, Clifford-Holmes and SnowLombard et al., 2021). This is for marine planning at a regional scale, rather than at local levels, which is considered key for the development of a sustainable blue economy (Reference Friess and Grémaud-ColombierFriess and Grémaud-Colombier, 2021).
15.3.3 Ways Forward
Among the possible ways forward for transformative ocean governance in all its dimensions at different scales, this section will investigate the potential of the interdependence between human rights and marine biodiversity to address indirect drivers of biodiversity loss, including power dynamics.
From an international law perspective, even if the CBD and its guidelines do not use explicit human rights language, they have made significant conceptual and normative contributions to the relationship with human rights, specifically with regard to Indigenous peoples’ rights to natural resources (Reference MorgeraMorgera, 2018a). As a result, the CBD and its instruments have been increasingly relied upon by international human rights bodies (A/HRC/37/59, 2018). This recognition has implications both for national-level action, as well as for international cooperation, at the global and regional levels (A/HRC/34/49, 2017, paras. 36–48), and can have a bearing on the inclusiveness and integration of ocean governance. Notably, human rights can help address, from a legal perspective, the “politics of transformative change,”Footnote 45 preventing a shifting of the burden of response onto the vulnerable; paying attention to social differentiation, through the lens of nondiscrimination; and addressing issues of power and legitimacy. In other words, human rights can serve to address questions of justiceFootnote 46 in ocean governance. The integration of international human rights law into the interpretation and application of the law of the sea, however, is not very advanced (Reference Barnes, Bogojević and RayfuseBarnes, 2018).
One way in which human rights considerations can be put into practice in the context of ocean governance, with a view to making it more integrated and inclusive, is reliance on the international legal concept of fair and equitable benefit-sharing, which is already included in the law of the sea and international human rights law, and has been elaborated upon under the CBD (Reference MorgeraMorgera, 2018b). As will be argued below, fair and equitable benefit-sharing can support transformative governance in terms of framing and agenda-setting, leadership, financial investment, capacity for learning and increasing institutionalization.Footnote 47
Fair and equitable benefit-sharing norms in the law of the sea are conceived narrowly in relation to deep-seabed mining and marine scientific cooperation (UNCLOS, Arts. 82(1) and (4), 242–244 and Part XI; Reference NoyesNoyes, 2011; Reference Salpin, Morgera, Buck and TsioumaniSalpin, 2013), and they are currently being developed with regard to bioprospecting in areas beyond national jurisdiction as part of the negotiations of a new legally binding instrument on marine biodiversity of these areas (Reference MorgeraMorgera, 2018–19). Benefit-sharing has, however, become a broader obligation in international biodiversity law (Reference MorgeraMorgera, 2016) arising from the conservation and sustainable use of natural resources (both within and outside national jurisdiction, beyond access to genetic resources) to address equity and sustainability issues as part of the ecosystem approach (Contra Reference BaslarBaslar, 1998).Footnote 48 Along parallel lines, under international human rights law, benefit-sharing has been identified as a safeguard to protect the human rights of Indigenous peoples (A/HRC/27/59, 2018, Principle 15; Reference MorgeraMorgera, 2019), small-scale fishing communities (A/RES/73/165, 2019; Reference Morgera and NakamuraMorgera and Nakamura, forthcoming) and rural women (CEDAW/C/GC/34, 2016), including in connection with their effective participation in the creation and management of protected areas. In addition, benefit-sharing is part and parcel of the human right to science (the right of everyone to benefit from scientific advancements), which reveals the human rights dimensions of interstate obligations related to scientific cooperation, capacity-building and technology transfer (International Covenant on Economic, Social and Cultural Rights, Art. 15(3); Reference MorgeraMorgera, 2015).
That said, benefit-sharing implementation is often dominated by a transactional logic to obtain a “green light” for conservation or development projects, rather than redress power asymmetries that threaten biodiversity conservation and sustainable use (Reference Martin, Akol, Phillips and SikorMartin et al., 2014). A different interpretation, however, emerges from CBD guidance that is more aligned with human rights standards. This interpretation focuses on the active participation of beneficiaries in the identification of benefits, which relies on an iterative, concerted and good-faith dialogue to develop a common understanding as part of mutual learning and an adaptive approach. Based on a combined reading of interpretative materials, “sharing” principally conveys the idea of agency, as opposed to the passive enjoyment of benefits (Reference MancisidorMancisidor, 2015), and therefore a shift away from unidirectional (likely, top-down) or one-off flows of benefits. In addition, benefit-sharing usually relies on a menu of benefits, the nature of which can be economic and noneconomic. This arguably allows taking into account, through the concerted, dialogic process of sharing, the beneficiaries’ needs, values and priorities through a contextual selection of the combination of benefits that may best serve to lay the foundation for partnership (Reference MorgeraMorgera, 2016). The expressions “fair and equitable,” which is generally left to subsequent negotiations, can be interpreted to express the rationale of balancing competing rights and interests (Reference BurkeBurke, 2014), with a view to integrating both procedural and substantive dimensions of justice (Reference KlägerKläger, 2011) into a relationship regulated by international law that is characterized by power imbalances (Reference KlägerKläger, 2011).
Applied at the multilateral level, this interpretation of benefit-sharing can support the voice of developing countries in co-identifying the benefits and needs for transformative ocean governance through the integrated implementation of capacity-building, technology transfer, scientific cooperation and information-sharing obligations (Reference MorgeraMorgera, 2016). In particular, this can be applied to the creation and management of MPA networks, with a focus on equity and power imbalances in ocean science production and area-based management and impacts at local levels. It could also support the co-development of MPAs as integral components of ecosystem-based fisheries management based on better understanding of the dependence on ecosystem services for different actors and sectors. As the Post-2020 Global Biodiversity Framework indicates, this would be aligned with the broader goal of valuing and maintaining nature’s contributions to people through conservation and sustainable use “for the benefit of all” and would take into account the importance of spatial approaches to this end:
The number of people who can benefit from nature’s contributions to people depends not only on nature’s ability to provide the benefit, but also on societies’ ability to manage their distribution, fairly and equitably, within and between generations.
This approach is aligned with the innovative theory of change in the Global Biodiversity Framework, which emphasizes “a whole-of-government and society approach” for transformative change and the role of a rights-based approach and cross-scale partnerships for ensuring that “biodiversity is used sustainably in order to meet people’s needs,” notably gender equality, youth inclusion, and the full and effective participation of Indigenous peoples and local communities in the implementation of this framework (CBD/POST2020/PREP/2/1, 2020).
This co-identification and delivery of benefits can be supported by a process of institutionalization:Footnote 49 multilateral facilitative and brokering arrangements can serve to operationalize relevant duties of cooperation with a view to ensuring equitable distribution across different regions, monitoring of effectiveness, and learning from experience. The need for such an approach has already been demonstrated in other international processes, such as the International Seabed Authority (ISA) and the International Maritime Organization (IMO) (Reference Morgera and NtonaMorgera and Ntona, 2018). In addition, benefit-sharing is a key element to recognizing Indigenous peoples and local communities for their global contributions to the conservation and sustainable use of biodiversity, and to respectfully integrate their knowledge systemsFootnote 50 in relation to MPA creation and management at different levels. This could allow for the co-identification of benefits and needs for transformative ocean governance beyond the current state-centric model, with a view to enhancing both transdisciplinary and inclusive ocean governance.
The key elements of a benefit-sharing inspired multilateral approach to transformative ocean governance would then be the following:
Joined-up thinking on the implementation of various international obligations on scientific cooperation and information-sharing, financial and technological solidarity, capacity-building and their human rights dimensions (integrative and transdisciplinary governance);
Dialogue to enhance collaboration across sectors, among duty-bearers and among human rights-holders, to contribute to the achievement of international biodiversity, ocean, climate change and human rights objectives (integrative governance);
Deliberation and mutual learning with a view to setting priorities to the benefit of the most vulnerable (inclusive governance);
The provision of international institutional support for facilitating and brokering scientific cooperation opportunities;
o Co-identifying information-sharing, technology transfer and regulatory and institutional capacity-building needs and available assistance; and
o Building, and assessing the effects of partnerships, including public–private partnerships (adaptive governance);
Multistakeholder identification and assessment of obstacles, co-development of proposals for enhancement, joint monitoring and reflection on lessons learned on emerging transformative approaches (inclusive and adaptive governance); and
Transparency about, and assessment of, the distribution of benefits across regions, as well as good practices and lessons learned at the local, national and regional levels, with a view to ensuring fairness and equity in benefit-sharing (arising from the dialogue and incrementally shaping funding and governance across scales – adaptive governance).Footnote 51
15.4 Conclusions
These elements could be applied in the context of area-based management and spatial approaches under the ongoing negotiations of an international instrument on marine biodiversity of areas beyond national jurisdiction (Reference Morgera, De Lucia, Nguyen and Oude ElferinkMorgera, 2022), and under the Sustainable Ocean Initiative. This chapter focuses on the latter, as an already institutionalized opportunity for transformative governance. The Initiative has become a regular process to facilitate the exchange of experiences, to identify options and opportunities to enhance cross-sectoral collaboration toward internationally agreed goals and to discuss the need for specific tools, guidelines or other initiatives to strengthen collaboration among not only regional seas conventions and RFMOs, but also sectoral international organizations like the Food and Agriculture Organization of the United Nations, the IMO and the ISA (Reference Diz and NtonaDiz and Ntona, 2018). The Initiative could take the approach outlined above to understand the reasons why “many protected areas are not effectively or equitably managed,” as well as “the importance of focusing on biodiversity outcomes rather than spatial area” included within MPAs, and the “provision of ecosystem services and to maintain integrity of planetary ecological processes” (CBD/SBSTTA.24/3/Add.2, 2021, paras. 54–56). Equally, the Initiative could provide a forum to reflect on equity issues across scales in interregional scientific cooperation, notably in relation to carrying out fisheries assessments in data-poor environments (Reference Kenny, Campbell and Koen-AlonsoKenny et al., 2018), implementation of the precautionary approach to fisheries (UNFSA, Art. 6 and Annex II; A/Conf.210/2016/5, 2016, para. 36), habitat protection in the context of conflicts of use (i.e. fishing or fishing survey activities vs seismic activities) (NAFO, 2016), and the effects of climate change and ocean acidification on marine ecosystems (A/RES/72/73, 2018, para. 196). Furthermore, scientific and participatory methodologies for assessing coastal communities’ and coastal and marine ecosystems’ vulnerabilities to climate change and ocean acidification are a crucial area of scientific cooperation and capacity-building to identify adaptation measures in most vulnerable regions (Reference Cochrane, Rakotondrazafy and AswaniCochrane et al., 2017).
A reflection has already been started on the role of the Regional Seas Programme for contributing to the Post-2020 Biodiversity Framework (CBD/SBSTTA/24/INF/24, 2021). Based on the key challenge facing the Benguela Current Commission and the findings from the Algoa Bay case study in South Africa, the SOI could share learning across scales on integrating social and natural sciences insights, as well as different knowledge systems. This could support regional seas organizations to engage in complex stakeholder engagements and deliberations on trade-offs in a constructive manner, to maximize the potential for transformation, by expanding their current integrative and anticipatory governance approaches to inclusive, adaptive and transdisciplinary approaches. The Initiative could also provide a forum to engage with the increasing concentration of businesses in the blue economy and explore how to build fair partnerships with the private sector in the context of MPA networks at different scales (Virdin, 2021). These efforts could contribute to strengthening the adaptive and transdisciplinary governance dimensions of efforts on EBSAs and ABMTs across scales, contributing to implementing CBD obligations to monitor biodiversity components that require urgent conservation measures and those that offer the best potential for sustainable use through international technical and scientific cooperation on conservation and the sustainable use of biodiversity (CBD, Arts. 7 and 17–18). It could also support CBD Parties in providing the evidence base to identify processes with (likely) significant adverse impacts on biodiversity conservation and sustainable use (CBD, Art 7 (c)), as well as to assess and minimize adverse impacts (CBD, Art. 14), while building capacity by sharing cross-regional learning on transboundary MSP approaches (CBD, Art. 12; CBD/EBSA/EM/2017/1/INF/1, 2017).
At the national level, this rights-based interpretation of benefit-sharing could be explored as part of marine spatial planning processes. It could support bottom-up forms of deliberations (Reference Cotula and WebsterCotula and Webster, 2020), characterized by the agency of beneficiaries, the respect of human rights, and mutual understanding of different benefits and priorities in MPA creation and other area-based management tools, as well as in the sustainable use of marine resources and the advancement of ocean science. Such dialogues could be informed by interdisciplinary and transdisciplinary research (Reference Morgera, Parks, Schroeder, Heyvaert and Duvic-PaoliMorgera et al., 2021) to assist different actors in the respectful and constructive engagement with beneficiaries’ choice and capabilities, knowledge systems, and different worldviews of nature and development, and an understanding of different benefits and risks across scales (Reference MorgeraNtona and Morgera, 2018). The partnership that is being built among researchers from different disciplines, different sectors of government and different knowledge holders could also contribute to the contextual application of the precautionary principle and new technologies (anticipatory governance), through learning, experimentation and reflexivity (adaptive governance). Research is equally needed to document good practices in integrating the evidence base across marine sciences and social sciences through inclusive approaches, with a view to understanding barriers and opportunities to scaling up to the national, regional and international levels.
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