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.

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.

11.4.1 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).

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.

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

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:

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).

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).

13.4.1 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).

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.

14.3.1 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.

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).

15.3.1 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.

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:

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;

  1. o Co-identifying information-sharing, technology transfer and regulatory and institutional capacity-building needs and available assistance; and

  2. 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