Introduction

Making decisions between alternative investments, projects, or policies that affect the provision of ecosystem services often involves weighing up and comparing multiple costs and benefits that are measured in different metrics and are incurred at different points in time. For example, the establishment of a new protected area might involve costs in terms of the purchase of land, compensation of local communities, and ongoing maintenance and enforcement costs; and benefits in terms of biodiversity conservation, recreational use, and improved watershed services. These costs and benefits are likely to be measured in different units, incurred by different groups and have different time profiles. Organizing, comparing, and aggregating information on such a complexity of impacts, and subsequently choosing between alternative options with different impact profiles require a structured approach. Methods for evaluation or appraisal of complex decision contexts provide systems for structuring the information and factors that are relevant to a decision.

Benjamin Franklin’s description of his own approach to making complex decisions sets out the intuition behind evaluation methods (Franklin, 1772):

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

In Chapter 3, the variation in ecosystem processes and functions was described using the plant functional trait approach. Due to variability in plant functional traits and other environmental conditions, as well as variation in human influence on ecosystems, the supply of ecosystem services is not homogeneously distributed across space. Besides spatial variation in the supply of ecosystem services, spatial variation in socioeconomic conditions makes the demand of ecosystem services dependent upon location as well. To understand the role of this spatial variation many ecosystem service assessments use observations, measurements, and models to create ecosystem service maps. Especially in the last couple of years an exponential increase has been observed in research, papers, and reports focused on mapping of ecosystem services. Ecosystem services can be mapped using various methods, of which applicability depends upon data availability, scope of the study, and time constraints. Regulating services are most commonly mapped, followed by provisioning services. The most frequently studied individual ecosystem services are climate regulation and food production (Martínez-Harms and Balvanera, 2012), and when multiple ecosystem services are mapped food production is almost always included (Crossman et al., 2013). Moreover, these studies vary in their scale from the global to the local level, vary in the type and number of ecosystem services incorporated, and map supply, demand, or a combination of both.

Introduction

The concept of ecosystem services is useful but also rather controversial. In part, this controversy relates to what is sometimes called the “commoditization” of ecosystem services (Muradian and Rival, 2012; see also the introduction of this book). Many people feel that the benefits that nature provides us with should not be conceptualized as “services” comparable to those supplied by traditional markets (e.g. the legal services of a lawyer or the financial services of a bank).

Nevertheless, many environmental economists have observed that market mechanisms can play an important role in balancing the “supply” and “demand” of ecosystem services. For example, farmers respond to subsidy schemes that offer them a fair reward for environment-friendly practices and consumers with “green” preferences tend to buy ecolabeled products.

Unlike actors in traditional markets, the suppliers and consumers of ecosystem services often do not directly engage in transactions with each other. This has to do with four typical and closely related features of many ecosystem services (also known as “market failures”):

• Their “public good” (or “collective good”) nature: Ecosystem services, such as those provided by pristine forests and clean rivers, are there for everyone to enjoy. No one can be excluded from the tangible and intangible benefits they provide and (to a certain extent) there is no rivalry in their use (that means that their enjoyment by any individual does not preclude the simultaneous enjoyment by others).

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).