3.5 Biodiversity and protected areas 40
3.5.3 Protected area governance/management approaches 46
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While the IUCN categories of protected areas have been criticised on various points, they currently provide the best succinct overview of the multiplicity of protected area types (Stolton, 2010b). It is encouraging to note that efforts for their improvement and refinement are already underway as has been shown above.
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and preservation, and second those which distinguish between centralised state control and devolved local control (Borgerhoff Mulder and Coppolillo, 2005; Brockington et al, 2008).
They then map many conservation strategies or approaches on this matrix as shown in Figure 3.2 below.
Centralised
Preservation Use
Decentralised
Figure 3.2: A typology of conservation practice
Source: Borgerhoff Mulder and Coppolillo (2005: 300)
Salafsky et al (2002) have also developed yet another taxonomy which attempts to map the wide diversity of conservation activity, strategy and intervention. They divide conservation activities into protection and management, law and policy, education and awareness, and incentives. This complex category system, a modified version of which is outlined in Table 3.4 by Brockington et al (2008), shows that the incentives will influence the relationships between people and nature and also those among people. Moreover, in the laws, policies, educational ideas and mechanisms of protection, there will be countless interactions with the economy and markets (Brockington et al, 2008). Table 3.4 categorises the types of tools available to conservation practitioners. Columns contain broad categories of tools, while each cell contains a broad approach and two examples of more specific strategies (italic font) under this approach.
Timber Certification
Buffer Zones
National Park Extractive
Reserves
Community Wildlife Management
Areas Village Forest
Reserves
Indigenous Protected Areas Private
Reserves
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Table 3.4: A taxonomy of biodiversity conservation approaches and strategies Protection
and
Management
Law and Policy Education and Awareness Changing Incentives
Strictly Protected Areas:
Reserves and private parks
Legislation and Treaties:
developing international treaties and lobbying governments
Formal Education:
developing school curricula and teaching graduate students
Conservation Enterprises:
linked (for example, ecotourism) and unlinked (for example, jobs for poachers)
Managed Landscapes:
conservation easements and community- based
management
Compliance and Watchdog:
developing legal standards and monitoring compliance with standards
Non-Formal Education:
media training for scientists and public outreach via museums
Using Market Pressure:
certification (positive incentives) and boycotts (negative incentives)
Protected and Managed Species:
bans on
killing specific species and management of habitat for species
Litigation:
criminal
prosecution and civil suits
Informal Education:
media campaigns and community awareness raising
Economic Alternatives:
sustainable agriculture/
aquaculture and promoting alternative products
Species and Habitat
Restoration:
reintroducing predators and recreating wetlands
Policy
Development and Reform:
research on policy options and devolution of control
Moral Confrontation:
civil disobedience and moneywrenching/
ecotourism
Conservation Payments:
quid-pro-quo
performance payments and debt-for-nature swaps
Ex-Situ Protection:
captive
breeding and gene banking
Non-Monetary Values:
spiritual/cultural/
existence values and links to human health
Source: Brockington et al (2008: 11)
Conservation has been described as “an incredibly broad church…riven with conflict”
(Brockington et al, 2008: 6). Indeed, such a ‘broad church’ image has been reflected in the many and diverse protected area categories and governance types discussed above. This is because there are sharp disagreements about ethics, morals, practices and compromises
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among conservationists (Barbault, 2011; Brockington et al, 2008; Doak et al, 2014; Minteer and Miller, 2011). For example, there are conservationists who prefer wild places without human interference, while others prefer peopled landscapes (Brockington et al, 2008; Minteer and Miller, 2011). There are also some for whom landscape is irrelevant and only species matter, while on the other hand there are conservationists whose concern is strictly their love of particular places and for whom global considerations are irrelevant (Brockington et al, 2008).
There are also some particularly deep divisions among conservationists about some specific issues, for example the debate about trade in live animals or their products (Brockington et al, 2008). A particular case to illustrate this is that of parrot trade. The World Parrot Trust, alongside some conservationists, insist that the trade in these wild birds is resulting in many deaths leading to species loss (Brockington et al, 2008). On the other hand are those who feel that this can be used to raise funds for conservation, and banning will worsen the situation by driving illegal traders underground (Brockington et al, 2008; Cooney and Jepson, 2006; Roe, 2006). Another contentious issue surrounds the trade in ivory. An increasing elephant population in Southern Africa has resulted in overcrowding leading to calls for the legalisation of ivory trade (Brockington et al, 2008). Certainly, under these situations, a multiplicity of conservation approach typologies is not only inevitable, but also justified and helpful (Brockington et al, 2008).
As shown earlier, during the late 19th century and much of the 20th century, the dominant conservation strategy in the world was the establishment of protected areas through state action in the form of national parks (Carter et al, 2008; Kreuter et al. 2010; Massey et al, 2014; Romero et al, 2012; Sloan et al, 2014). Indeed, national parks and other government- protected areas have long served as the conventional tool for biodiversity conservation since the establishment of Yellowstone National Park in the USA in 1872 (Langholz, 2009). Since then, governments worldwide have set aside more than 108 000 protected areas covering 30 million km2 of land, with many countries having reached the international standard of formally protecting 10% of their territorial surface area (Langholz, 2009).
It, however, became apparent in the second half of the 20th century that national parks were not adequate as a tool for effectively protecting the world’s flora and fauna from extinction (Kreuter et al, 2010; Langholz, 2003; Langholz, 2009; Nelson, 2010; Romero et al, 2012). A diverse array of factors increasingly challenges the prevailing state-centric natural resource
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policies and management practices across much of the world (Nelson, 2010). The struggle for solutions has led to new approaches, with most of this new conservation effort being directed on communal and private land outside the formally protected national parks (Bond, 2004;
Langholz, 2003; Romero et al, 2012).
A general trend in protected area governance approach, since the second half of the 20th century, has thus been a shift towards decentralisation and sustainable use. Such an approach has most probably been hatched out of the sustainable development thinking that emerged roughly around the same time period. The focus of this study is on community-conserved areas and private protected areas, and these are discussed in greater detail as stand-alone sections below.
3.6 Community-based natural resource management (CBNRM)
The involvement of local people in conservation has become a major feature of conservation policy across the world (Adams and Hulme, 2001; Balint, 2006; Cronkleton et al, 2012;
Rawlins and Westby, 2013). Disappointment with existing state-led conservation strategies to protect wildlife prompted many analysts, starting in the 1980s, to advocate for a greater involvement of communities and local populations in protection strategies (Agrawal and Redford, 2006; Horwich and Lyon, 2007). It is increasingly being noted that a great deal of hope can be placed in the ability of rural communities to conserve nature, and conservation by local communities is often claimed to be a more equitable and/ or effective alternative to many types of protected areas particularly fortress conservation (Brockington et al, 2008; Liu et al, 2010). Local residents have evolved with their surrounding environment over several centuries and retained traditional ecological knowledge and activities facilitating biodiversity conservation (Liu et al, 2010). Such traditional ecological knowledge, especially as it relates to resource use, can complement modern conservation systems and aid biological research, while supporting a more equitable and culturally sensitive method of management (Cox et al, 2014; Liu et al, 2010). Community conservation areas are also widely perceived to be a means of expanding the conservation estate, ensuring land is managed for conservation purposes beyond the boundaries of formal protected areas (Brockington et al, 2008; Cox et al, 2014; Rawlins and Westby, 2013). Advocates of community-based conservation further insist that it will result in increased support for conservation values and more prosperous and/ or empowered people (Brockington et al, 2008; Rawlins and Westby, 2013). This section
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explores the concept of community-based conservation or community-based natural resource management.
3.6.1 The origin and growth of community-based conservation
The predecessors of community-based conservation include the concept of buffer zones introduced by the UN Educational, Scientific and Cultural Organisation’s (UNESCO) Man and the Biosphere programme in 1979 and ICDPs popularised in the late 1980s and early 1990s (Bauch et al, 2014; Blom et al, 2010; Campbell and Vainio-Mattila, 2003; Fisher et al, 2005; García-Amado et al, 2013; Niedziałkowski et al, 2014; Pfund, 2010; Scherl et al, 2004). Since the 1980s, conservation organisations have been implementing approaches that aim to build support among local communities by sharing social and economic benefits from protected areas (Bauch et al, 2014; Horwich and Lyon, 2007; Scherl et al, 2004). Such initiatives involved compensating local people for lack of access to protected areas and providing alternative income sources that would allow people to benefit economically from conservation while refraining from environmentally destructive practices (Cox et al, 2014;
Fisher et al, 2005; Scherl et al, 2004). A lot of international development agencies provided funding in support of biodiversity conservation through ICDPs in the 1990s (Scherl et al, 2004). However, buffer zones and ICDPs have been criticised for their failure to adequately involve local populations in planning and decision-making (Agrawal and Redford, 2006;
Bauch et al, 2014; Campbell and Vainio-Mattila, 2003; García-Amado et al, 2013; Horwich and Lyon, 2007; Wells and Brandon, 1993). Many ICDPs have failed to produce ‘win-win’
conservation and development outcomes (Blom et al, 2010; Larson et al, 1997; Pfund, 2010;
Wells and Brandon, 1993), as evidenced by a general failure to limit unsustainable resource use and a lack of demonstrable improvements in peoples’ livelihoods (Agrawal and Redford, 2006; Bauch et al, 2014; Scherl et al, 2004). Some of the major shortcomings of the first generation ICDPs have been summarised by McShane and Wells (2004 cited in Scherl et al, 2004: 30):
The flawed assumption that planning and money alone were sufficient to achieve
‘win-win’ scenarios;
Attempting to implement ICDPs within the framework of a time-bound ‘project cycle’ and failure to adapt to the pace of local communities by trying to meet externally imposed deadlines;
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Failure to identify, negotiate and implement trade-offs between the interests and claims of multiple stakeholders;
Lack of adaptive management and flexibility to respond to evolving scenarios;
Failure to cede significant decision-making to local stakeholders so that ICDPs remained outside local systems, thereby reducing the likelihood that any gains they may have achieved would persist beyond the project life;
Perceived or actual bias towards the interests of either the protected area management agency or an environmental NGO;
A focus on activities (social programmes and income creation through alternative livelihoods) rather than impacts (on biodiversity);
Addressing local symptoms while ignoring underlying policy constraints or conversely dealing with macro-level issues while ignoring local realities; and
Regarding ‘local communities’ as a homogenous entity when the reality was a wide range of different stakeholders with different needs and aspirations.
Another attempt to involve local communities in protected areas, besides through buffer zones and first generation ICDPs, has been through inclusive or collaborative management approaches (Ameha et al, 2014; Fischer et al, 2014; Fisher et al, 2005; Ming’atea et al, 2014;
Scherl et al, 2004; Cronkleton et al, 2012). Indeed the formation of partnerships for active participation in the day-to-day management of protected areas is becoming more and more widespread (Scherl et al, 2004). Co-management arrangements have come out of the realisation that local communities have roles to play in resource management, conservation and development, and also from the reality that forest-dependent communities have demanded a recognition of their rights and have thus increasingly been difficult to exclude (Cronkleton et al, 2012; Ming’atea et al, 2014). Collaborative management or co- management involves a partnership between stakeholders, especially protected area authorities and local communities (Cronkleton et al, 2012; Fisher et al, 2005; Trimblea et al, 2014). Systems of co-management between local communities and technical advisors such as government protected area authorities, NGOs or private contractors can ensure that local communities have a stake in decision-making and receive a significant share of the benefits from protected areas (Brockington et al, 2008; Scherl et al, 2004; Wells and Brandon, 1992).
However, collaborative management arrangements as a way of involving local communities in conservation, just as with buffer zones and first generation ICDPs, has not been well
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received. Brockington et al (2008) have noted that the label ‘co-management’ is problematic as it implies equality between the participants. In reality, protected area authorities have often been the dominant partners, wielding most of the decision-making powers over conservation and the benefits derived from it, at the expense of their powerless and long-marginalised local-community partners lacking the capacity and experience to flourish in such institutional environments (Brockington et al, 2008; Cronkleton et al, 2012; Lopes et al, 2013; Scherl et al, 2004; Zhu et al, 2014). The sharing of decision-making under collaborative management arrangements often does not eliminate the power imbalances inherent in top-down management approaches (Chen et al, 2014; Cronkleton et al, 2012). Collaborative management also conceals a considerable diversity of practices, and a variety of specific historical and political circumstances that would have given rise to the arrangements (Brockington et al, 2008; Rodwell et al, 2014).
Community conservation has been hailed as a more realistic move for ensuring that local communities are more directly involved in decision-making and benefit sharing from conservation. Compared to the above attempts (buffer zones, first generation ICDPs and collaborative management) to involve local communities in conservation, community-based conservation is different in that it places the community’s involvement at the centre of conservation, rather than the mechanisms for achieving it such as parks, projects or land use zoning (Campbell and Vainio-Mattila, 2003; Cox et al, 2014; Rawlins and Westby, 2013).
Community conservation projects see local rural people as the solution to habitat degradation, whereas ICDPs see them as the problem (Horwich and Lyon, 2007). According to Western and Wright (1994: 9), the central tenet of community-based conservation is “the co-existence of people and nature, as distinct from protectionism and the segregation of people and nature”. Adams and Hulme (2001: 13) have defined community-based conservation as “those principles and practices that argue that conservation goals should be pursued by strategies that emphasise the role of local residents in decision-making about natural resources”.
Community-based conservation represents a more inclusive and people-oriented approach to conservation and is a reaction to the failure of exclusionary conservation in a world in which social and economic factors are increasingly seen as key to conservation success (Berkes, 2004). The counter-narrative of community-based natural resource management, especially through community conservation areas, effectively displaced the narrative of ‘fortress conservation’ by emphasising development through community conservation and sustainable
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use (Balint, 2006; Berkes, 2004; Horwich and Lyon, 2007; Murphree, 2002). While, as shown earlier, buffer zones, first generation ICDPs and collaborative management are viewed as early attempts at community-based conservation, the level of locals’ involvement in decision-making under such arrangements was too low compared to community conservation areas largely owned, managed and controlled by local communities. For this reason, community-based conservation initiatives through community conservation areas have often been referred to as a new generation of ICDPs, more truly representative of the community conservation counter-narrative.
The community-based conservation counter-narrative was quickly accepted within conservation circles, and several factors acted as catalysts towards this more people-centred conservation approach (Adams and Hulme, 2001; Cranford and Mourato, 2011; Fabricius, 2004; Gustavssona et al, 2014). First, community conservation equates conservation with sustainable development and hence captures the huge upwelling of policy commitment from the UNCED in 1992 which argued that conservation goals should contribute to, and not conflict with, basic human needs (Adams and Hulme, 2001; Cranford and Mourato, 2011).
This argument led some commentators, in turn, to argue that fortress conservation must be abandoned because of its adverse impacts on the living conditions of the rural poor (Adams and Hulme, 2001; Reed, 2008; Reed et al, 2009). Community conservation therefore provides a conceptual framework within which the conservation of biodiversity and the challenge of meeting human needs must be integrated (Cottrell et al, 2013; Holden et al, 2014).
A second reason for the success of the community conservation narrative was that it developed at a time of significant shifts in the dominant discourses of development (Adams and Hulme, 2001; Romero et al, 2012). During the 1970s ‘top-down’ technocratic or blueprint approaches to development came under increasing scrutiny as they failed to deliver the economic growth and social benefits that had been promised (Adams and Hulme, 2001;
Romero et al, 2012; Turner and Hulme, 1997). A new perspective emerged arguing that
“development goals could only be achieved by bottom-up planning, decentralisation, process approaches, participation, and community organisation” (Adams and Hulme 2001: 17). The impact of such alternative development thinking was witnessed through commitments by aid donors and development planners towards adapting participatory development approaches in the early 1990s (Adams and Hulme, 2001). The link that had earlier been forged between conservation and development gave conservationists strong motivation to do the same,
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especially as this enabled them to tap into new sources of funds by making their activities fit development aid budgets (Adams and Hulme, 2001).
A third reason that catalysed the success of the community conservation paradigm lies in the renewed interest in the 1980s in the market as a means of delivering development (Adams and Hulme, 2001; Romero et al, 2012; Toye, 1993). This renewed market-oriented development thinking argued that, to achieve public policy goals (including conservation, development or sustainable development), economic incentives must be set correctly for all of the main actors through market mechanisms (Adams and Hulme, 2001). This therefore required non-market actors such as state agencies, which distorted markets and thus made public policy goals unachievable, to be abolished (Adams and Hulme, 2001). The implication of this to conservation goals was less regulation and more entrepreneurial action by local communities, individual businessmen and private companies thereby unlocking the economic values of conservation resources and making them part of the local and global economy (Adams and Hulme, 2001; Cranford and Mourato, 2011; Gustavssona et al, 2014).
Adams and Hulme (2001) further note that community conservation also fitted well into the
‘New Policy Agenda’ for foreign assistance that had developed in Washington in the early 1990s, driven by benefits of neo-classical economics and liberal democratic theory.
Community conservation seemed to be in tandem with neo-classical economics as it recognised the importance of economic incentives and markets; meant a reduced role for the state; and created spaces for communities, including villages, private individuals, companies and groups of companies, to be more involved in conservation (Adams and Hulme, 2001;
Cranford and Mourato, 2011). Community conservation was also in line with liberal democratic theory as, by helping communities to organise themselves to manage natural resources, it deepened the democratisation process (Adams and Hulme, 2001).
Another reason for the success of community conservation came out of the realisation that conservation goals can often not be achieved within the boundaries of formal protected areas, even if they were quite large (Adams and Hulme, 2001). It has increasingly been realised within conservation biology that mobile wildlife species cannot be sustained on small preservation islands such as national parks and buffer zones but needed large dispersal areas to ensure healthy breeding stock and to respond to local extinctions and environmental changes (Adams and Hulme, 2001; Goswami et al, 2014). This led to the realisation that conservation needed to reach out of protected areas into the wider landscape, with