Introduction

One important way in which the concept of ecosystem services has been put into practice is by providing incentives to the providers of these services to increase or maintain their provisioning levels. These incentives are typically monetary or in-kind compensations that are formalized in contracts. For example, a downstream community may pay upstream landowners to plant or protect forests and thereby maintain a continuous supply of water to the downstream community. This contract increases the ecosystem service “freshwater supply” in return for a payment. The general term that is used to refer to this type of contracts is “payments for ecosystem services” or PES. In this chapter we will provide detailed definitions of PES, assess important elements in the design of PES contracts, and look at potential and pitfalls in PES performance, based on case study evidence.

Our starting point is the argument made by Ferraro and Kiss (2002) that direct payment mechanisms like PES are in fact the most effective approach to nature conservation and ecosystem governance, as payments create a direct incentive for sustainable use. PES contracts are theoretically grounded in Nobel Prize-winning economist Ronald Coase’s theorem which states that “when trade in an externality is possible and there are no transaction costs, bargaining will lead to an efficient outcome regardless of the initial allocation of property rights” (Coase, 1960). Following this theorem, PES contracts would normally reflect some bargaining outcome between ecosystem service providers and beneficiaries. Since the potential beneficiaries have a certain willingness to pay (WTP) for the safeguarding of ecosystem service provisioning and the potential providers have a certain willingness to accept (WTA) compensation to cover their opportunity cost, we should expect PES contracts to occur as long as the WTP exceeds the WTA. As we will see further along in this chapter, the fact that transaction costs are not zero limits the potential of PES contracts, in addition to the (semi-) public good character of most ecosystem services, which complicates PES contract design.

Introduction

The popularity of local governance approaches in ecosystem management has increased over the last decades. There seem to be several reasons for why this has been the case. First, top-down approaches to ecosystem governance have not been very cost-effective as the costs of top-down enforcement of resource-use restrictions have been relatively high (Somanathan et al., 2009). This has been the case in water management, fisheries management, forest management, wetland management, nature conservation, etc. Also, especially in conservation-oriented approaches, the livelihood costs of prohibiting resource use, and sometimes even displacement of local communities, have been enormous (Cernea and Schmidt-Soltau, 2006). As a result, conservation agencies and governments have started exploring alternative approaches to ecosystem governance like integrated conservation–development approaches and community co-management in order to reduce adverse livelihood impacts, increase local participation, and reduce monitoring and enforcement costs. Second, an understanding has been emerging that ecosystems do not necessarily need to be managed by governments solely but that local communities can self-govern common-good resources effectively as well. Ostrom (1990) showed that communities can sustainably manage forests, pastures, wetlands, and other common-good resources, and that collective resource rights do not need to result in a “tragedy of the commons” if local governance mechanisms exist. This understanding, in combination with the finding that top-down governance of ecosystems has often not been very effective, has resulted in a decentralization of common-good resource management worldwide. Third, the human right to self-determination has become an important argument for local ecosystem governance, especially since people living in biodiversity-rich spots are often marginalized: not only are poverty and biodiversity strongly correlated, biodiversity hotspots coincide with indigenous peoples’ territories, indigenous peoples that are often not well represented in decision-making processes concerning their livelihoods (Colchester, 2004).

Introduction

Ecosystem services are hot, and they have been hot for a while. In 1998, Costanza et al. published their famous article about the societal value produced by ecosystems through ecosystem service delivery, an article which at the time this introduction was written had been cited more than 10 000 times. In 2005, the Millennium Ecosystem Assessment (MEA) framed the need to protect biodiversity and the world’s ecosystems in terms of ecosystem services (MEA, 2005). In 2009, “The economics of ecosystems and biodiversity” (TEEB) followed up by presenting an approach to help decision-makers recognize, demonstrate, and capture the values of ecosystem services and biodiversity. And by now, most (inter)national policies in the field of nature conservation refer to ecosystem services when explaining the need for nature conservation, biodiversity protection, and sustainable resource use.

Despite the popularity of the concept of ecosystem services, policy-makers and practitioners are struggling to implement the concept in practice. An important reason for this science–policy divide is the lack of an interdisciplinary framework that guides policy-makers through the definition and measurement of ecosystem services to their valuation and the translation of these values into effective policy instruments and governance arrangements (see also Daily et al., 2009).

Biodiversity-related ecosystem services

Biodiversity is being depleted at an alarming rate (UN 2010). This is a problem, not only because biodiversity is intrinsically valuable, but also because biodiversity is the underlying asset from which all ecosystem services are produced. Biodiversity is best conserved in protected areas and national parks (Bruner et al. 2001). Establishing protected areas often has significant consequences for local communities. On the one hand, negative impacts occur when local communities are denied access to natural resources or are relocated to more marginal parts (Cernea and Schmidt-Soltau 2006). But even when communities are allowed to continue living near or inside protected areas the impacts of protected area establishment can be great: increased populations of wildlife increase crop damages and sometimes even cause personal injuries (Woodroffe et al. 2005) and the resource-use restrictions inside protected areas limit resource harvests for subsistence use (Wilkie et al. 2006). On the other hand, recent studies indicate that protected area establishment can also have positive livelihood effects: protected areas attract ecotourism, which can generate significant revenues at regional scale (Andam et al. 2010, Sims 2010). If these revenues trickle down to the local level, then protected area establishment can also have positive livelihood impacts at local scale.