Once the multifunctionality of landscapes and their services are identified, questions arise, like: How can we measure (value) the importance of these services, to get a basis for our decision making? How robust are the estimated values of ecosystem services? To answer these questions we have to address the terms “value” and “valuation”, which have different meanings in different disciplines:
Farber et al., 2002*), because ecosystems are seen to have an “intrinsic value”, which cannot be measured (Callicott, 1989). Nevertheless, some concepts of value are important in the natural sciences, and are commonly used to talk about causal relationships between different parts of a system: For example, referring to particular tree species and their value in controlling soil erosion in a high slope area, or to the value of fires in recycling nutrients in a forest (Farber et al., 2002*). Therefore, the ecological importance (value) of ecosystems is determined by ecological criteria such as integrity, resilience and resistance (health). Ecological measures of value encompass parameters such as complexity, diversity and rarity (de Groot et al., 2003). To integrate ecological values into landscape planning sustainable use-levels are often applied. Batabyal et al. (2003*), for instance, propose to use a scarcity value which is described by ecological thresholds, as a measure for sustainable managing. Their study presents a formal model that explicitly analyses the connections between thresholds and ecosystem management (Batabyal et al., 2003). The application of ecological modelling allows assessing the impact of environmental change and biodiversity loss on combined ecosystem services (Metzger et al., 2006*; Egoh et al., 2008*; Nelson et al., 2009*). Another approach to valuate the impact of land use change on ecosystem services is the application of reference systems, e.g. the potential natural vegetation (PNV) (Tüxen, 1956). Tüxen emphasized the big value of PNV-maps for different purposes in landscape planning and nature conservation, particularly for forestry, agriculture and landscape management. However, maps of the potential natural vegetation are less useful for purposes of detailed planning on larger scales in cultural landscapes, where the reconstruction of the PNV has only hypothetical character (Zerbe, 1998).
Pearce, 1991; Torras, 2000; TEEB, 2010*). To provide a common metric in which to express the benefits of diverse ecosystem services, the economic approach usually uses money as a general measurement unit. There exist many ways to translate the economic values into monetary terms. For details on valuation techniques see Dixon and Hufschmidt (1986); Peterson and Sorg (1987); Pearce and Turner (1990); Tietenberg (1992); Pearce and Moran (1994); Heal (2000*); Turner et al. (2003*) and the TEEB report (TEEB, 2010*). Chee (2004*), for instance, shows the principal methods for the monetary valuation and points out the pro and contra of these methods.
In general, there is a distinction between direct market valuation, indirect market valuation, contingent valuation and group valuation, each with its own associated measurement issues (de Groot et al., 2002*; MEA, 2003*). Whereas services, which are directly linked to the market, can be easily valued according to their market price, non-market services are often valued using the “willingness to pay” or “willingness to accept” compensation methods encompassing “avoiding cost”, “replacement cost”, “factor income”, “travel cost” and “hedonic pricing” (de Groot et al., 2002*). In the last years “contingent valuation” and “group valuation”, which are based on an open public deliberation, have also become appreciate techniques for estimating values (Jacobs, 1997; Sagoff, 1998).
All these different methods have gained increasing attention concerning ecosystem service valuation and have become an applicable tool for estimating service values. Following proponents of monetary valuation techniques, these economic methods are able to illustrate the distribution of benefits, improve understanding of problems and trade-offs and can thus facilitate decision making (e.g. Aylward and Barbier, 1992; Salzmann et al., 2001; de Groot, 2006*). However, economic valuation of ecosystem services has reached its limits (e.g. Heal, 2000; Farber et al., 2002*; Wilson and Howarth, 2002*; Chee, 2004*; Hein et al., 2006*). Although it may encourage management options, decisions makers have to take into account the overall objectives and limitations of economic valuation techniques (see Ludwig, 2000).
MEA, 2003*). However, although such cultural services play an essential part in the enhancement of human welfare, they are marginally present in the current research activities (Benayas et al., 2009*). This is considered as an increasing problem when the concept of ecosystem services is applied in cultural landscapes with typically long-lasting land use history, dynamic interactions of humans and nature, cultural patterns, and people’s identities and values. Therefore, the ecosystem service approach should be expanded by the “cultural landscape paradigm”, which includes humans as integral parts of landscapes, whereas other models in the present debate tend to see humans as impartial observers, as external drivers on ecosystems or as beneficiaries of environmental services (Matthews and Selman, 2006). Therefore, landscapes are seen as “social-ecological systems”, in which social, economic and environmental components are closely interwoven (Berkes et al., 2003). While conceptual and methodological developments in monetary valuation have aimed at covering a wide range of values, including intangible ones, it can be stated that socio-cultural values cannot be fully evaluated by economic valuation techniques. A psycho-cultural perspective of valuation would strongly suggest a transdisciplinary dialogue (Rist et al., 2004), aiming at cooperation between natural and social sciences research through debates on environmental ethics, tools and methods of social inquiry and socio-economic development as well as empowerment (Kumar and Kumar, 2008*).
Since the last two decades many publications have dealt with different interpretations and implementations of the term “value” in the context of ecosystem services (e.g. Costanza et al., 1997*; Bishop, J. T., ed., 1999; Odum and Odum, 2000; Howarth and Farber, 2002; Chee, 2004*; Farber et al., 2006*; Kumar and Kumar, 2008), which shows the big interest and importance of this topic. Following Costanza (2000*) valuation is a basic need of human beings. Any choice and trade-offs between competing alternatives imply valuations, which are simply the relative weights given to the various aspects of decisions. Therefore, valuation ultimately depends on the specific goal or objective of an item (Costanza, 2000). For a long time the main focus has been on the utilitarian approach. However, individual utility maximization has become constrained when sustainability and social equity were also included as goals into the valuation concept (Costanza and Folke, 1997*). According to the MEA and also to the TEEB approach the “total value” of an ecosystem and its services has to include three types of value domains, namely the ecological (environmental), economic and socio-cultural value (Toman, 1998; de Groot, 2006*). For example, hunting a game gives us food (health) and income but also cultural identity (as a hunter).
A special issue on valuation of ecosystem services, published in the journal Ecological Economics, discusses in detail the background, pro and contra of these three value approaches (de Groot et al., 2002; Farber et al., 2002*; Limburg et al., 2002; Wilson and Howarth, 2002*). One common problem in the valuation process is that information is often only available for some value domains and often in incompatible units.
Valuation can be conducted in many different ways (Pagiola et al., 2004). The MEA (2005) and TEEB (2010*) for instance focus on assessing the value of changes in ecosystem services resulting from management decisions or other human actions. This type of valuation is most likely to be directly policy relevant. The change in value can be assessed by either explicitly estimating the change value or by comparing the current value with the future value resulted by the alternative management regime. At landscape scale the (land use) change value approach proved also very useful to present all the different stakeholder positions, and their linkages, in a rather objective and clear manner to support management discussions. Depending on the goal of the valuation and on data availability, monetary as well as non-monetary valuation approaches are applicable (Gómez-Baggethun et al., 2010). In the further section we introduce some examples of valuation methods, which demonstrate important steps within the ecosystem service approach. As economic valuation has been implemented in many research studies and is also the main focus of the TEEB project, we provide also some important examples based on monetary valuation methods, although we do not place great emphasis on economic valuation within this review.
Adger et al., 1995; Pimentel et al., 1995; Costanza and Folke, 1997; Pimentel et al., 1997; Hein et al., 2006*). This total economic value can be seen as an economic indicator, which provide as measure of gross national product or genuine savings policy-relevant information on the state of the economy (MEA, 2003*). Costanza et al. (1997*), for instance, whose publication presented an important milestone in the valuation process, attempted in their study to find the total economic value for a range of different ecosystem services at the global (biospheric) level. The current economic value of 17 ecosystem services for 16 biomes was estimated, based on published studies and a few original calculations. In general, they estimated unit area values for ecosystem services (in $ ha–1 yr–1) and multiplied them by the total area of each biome. This approach has stimulated considerable debate and had not only to accept very sharp criticism from ecologists but also from economists (e.g. Opschoor, 1998; Turner et al., 1998; Bockstael et al., 2000*; Xiaoli and Wie, 2009*). Some of the core objections to Costanza’s model can be summarized as follows (Xiaoli and Wie, 2009*): the model did not adequately incorporate several factors which impact on ecosystem services, such as regional differences, spatial heterogeneity and social development. Neither can values estimated at one scale be expanded by a convenient physical index of area, such as hectares, to another scale, nor can two separate value estimates, derived under different contexts, simply be added together (Bockstael et al., 2000). However it has to be stated, that the objective of this world wide study was not to present accurate values, but to show how valuable the natural world is (Pearce, 1998). Since 1997, many studies were conducted to identify and quantify the value of ecosystem services. Whereas some of them based their results on Costanza et al. (1997) estimated values, others tried to modify Costanza’s model by including new approaches (e.g. Sutton and Costanza, 2002; Williams et al., 2003; Xiaoli and Wie, 2009).
To visualise that ecosystem services are spatially variable, and to identify key areas to be protected for the purpose of sustainable development, the “spatially explicit measure” represents a welcome method. It provides a mechanism for incorporating spatial context into ecosystem services evaluation (Chen et al., 2009*). Explicit value transfer becomes a useful method assessing ecosystems or landscapes, if valuation data is absent or limited (Bateman et al., 2002; Troy and Wilson, 2006*; Brenner et al., 2010*). Values and other data from the original study site are transferred to the designated policy site (Loomis, 1992). Troy and Wilson (2006*), for example, presented in their paper a decision support system framework, which was built upon the value transfer methodology. In each case study a unique typology of land cover, to which ecosystem service estimates were available from the literature, was developed. Standardized ecosystem service value coefficients were broken down by land cover class and service type for each case study. Therefore, scenario and historic change analyses according to ecosystem services could have been conducted. However, this approach also suffers from limitations, such as availability of data, strength of the data and comparability between the source data and policy context (Troy and Wilson, 2006). Whereas some ecosystem services are easily transferable because they are provided at large scales (e.g. the avoided greenhouse gas costs of carbon sequestration), other local scale services may have limited transferability (e.g. flood control values) (Farber et al., 2006*).
Recognizing the limitations of value transfer, advanced research has focused more on spatially-explicit ecological and economic models, to explain the effect of human policies on ecosystem services and subsequently on human welfare (Naidoo and Ricketts, 2006; Barbier et al., 2008; Polasky et al., 2008; Nelson et al., 2009*). Such models show the spatial heterogeneity of service provision and supply a framework for regulatory analysis in the context of, for example risk assessment, non-point source pollution control, wetlands restoration and avalanche protection (Bockstael et al., 1995*). The application of integrated modelling supported by GIS to simulate environmental change scenarios, especially climate change, has become a useful tool to help decision-makers in selecting sustainable and economically feasible development strategies (see Bockstael et al., 1995*; Higgins et al., 1997; Boumans et al., 2002; Gret-Regamey et al., 2008*; Chen et al., 2009). For example, in the Alpine region a study integrated into a single GIS platform several ecosystem process models simulating the provision of ecosystem services simultaneously with economic valuation procedures, in order to visualize climate change effects (Gret-Regamey et al., 2008). However, modelling is costly of data and measurability requirements and therefore, studies often address relatively small spatial scales, at which it is achievable to develop ecological-economic models. In addition, most models usually focus only on a few ecosystem services and neglect the impact of biodiversity loss on combined ecosystem services. Only some authors tried to integrate the interactions between biodiversity and multiple ecosystem services in their studies (e.g. Metzger et al., 2006; Egoh et al., 2008*; Nelson et al., 2009).
The recent TEEB project, mainly based on economic valuation, concentrates on assessing the consequences of changes resulting from alternative management options, rather than for attempting to estimate the total value of ecosystems (TEEB, 2010). Within this project best practice examples from around the world are presented. However, the review of case studies undertaken by TEEB shows that, in many instances, more efficient but less precise methods have been used, hence the results must be interpreted with appropriate care. Especially, in more complex situations involving multiple ecosystems and services, and/or different ethical or cultural convictions, monetary valuations seems to be less reliable or unsuitable. Nevertheless, monetary assessments are important for internalizing so-called externalities in economic accounting procedures and in policies that affect ecosystems, especially where the alternative assumption is that nature has zero (or infinite) value (de Groot, 2006*).
Paetzold et al., 2009*). A specific Norwegian quality assessment, for example, evaluates current provision of services relative to their provision 100 years ago (Pereira et al., 2005*). Paetzold et al. (2009) propose to evaluate the status of an ecosystem in terms of its sustainable provision of ecosystem services in relation to the societal expectations. Thereby for each ecosystem service the quality is defined by the ratio of its sustainable provision to the expected level of service delivery. Thus, systems that provide services in a satisfactory and sustainable way can therefore be regarded as being of better quality than those that do not. One major challenge is to select, or develop appropriate indicators that, for example, assess the sustainability aspect of a service or societal expectations (McMichael et al., 2005). In addition, it is difficult to obtain context-specific data on the provision and demand for many services (Chan et al., 2006). According to Martin and Blossey (2009) an ecosystem service cannot have a discrete value, because it depends on stakeholder preference and changes with quality and time frame. They suggest the following framework considering the quality of ecosystem services, the weighting, and the issue of time scale: TV = ∫tx 1S1 + y2S2… + znSn, where TV is the total value of a system; S1, S2, and Sn are service functions; 1, 2, and n include measures of quality; x, y, and z are the respective weights of the service functions 1, 2, and n; and t is the time frame considered. Habitat quality encompasses, for example, taxonomic diversity, suitability for rare species and historic composition of the site. The weighting of services depends mainly on the background and preferences of decision makers.
In the UK the merits of a “habitat, service and place based perspective” to the assessment of ecosystem services are emphasized (Haines-Young and Potschin, 2008). The habitat perspective is based on the use of a matrix of habitats and their related services. Pressures, respectively impacts on the services are additionally identified to assess state and trends of each service associated with England’s ecosystems. Since there is no commonly agreed terminology of pressures it is difficult to make such an assessment consistent. A clear advantage of using habitats as framework for representing the output of ecosystem services is that as distinct ecological units they could be seen in terms of “bundles” of services that they can deliver. It is generally known that most ecosystems are multifunctional, as structures and processes within them are capable of generating a wide range of different services (de Groot, 2006). The quality assessment of each habitat depends on the condition of their services and on the weighting of the service related indicators and their pressures. Although the habitat approach sounds very promising it also has its shortcomings, especially considering the multifunctionality of ecosystems. In most cases the links and interlinks between services might be overlooked. For policy relevance often costs-benefit analyses are conducted, because the exploitation of services usually has both costs and benefits for the society.
A wide range of studies illustrate that multifunctional landscapes are not only ecologically more sustainable and socio-culturally preferable but frequently also economically more beneficial than landscapes that only provide few ecosystem services (Balmford et al., 2002; Turner et al., 2003; Naidoo and Adamowicz, 2005). Therefore, Willemen et al. (2010*) propose to assess landscape values by referring to the total potential provision of goods and services at multifunctional locations. For each landscape the capacities of all landscape functions are normalized and summed up (see Gómez-Sal et al., 2003; Gimona and Van der Horst, 2007). Finally, a weighted value can be assigned to each landscape.
In the context of environmental assessment land use management decisions are often guided by some kind of transdisciplinary process, such as suggested by the concept ‘integrated planning assessment’ or more specifically the ‘quality of life capital’ approach (Potschin and Haines-Young, 2003; Haines-Young and Potschin, 2007). Thereby a “Leitbild” is used to describe what is viable in future, with regard to ecological sustainability and to the service preferences of society. Thus, the “Leitbild” concept can be applied as a reference system for service assessment in a given landscape.
To integrate in landscape planning not only environmental but also socio cultural values, great emphasis has to be placed on the expectations of inhabitants, tourists and the general public (Hunziker et al., 2008). By integrating different social groups into the valuation process both conflicting and compatible views about landscape change may arise. However, these insights are important for steering landscape development in a stakeholder-related sense and for recognising and reducing conflicts of interest (see Backhaus et al., 2007; Soliva et al., 2008). The underlying idea is that an integrated and multi-dimensional approach will be more likely to capture the full range of values, including those which may be context specific (local, regional, national, and global). Schama (1995), for instance, show how landscape perception is over-formed by cultural and national identity.
In general, case studies of socio-cultural assessment methods are lacking (Benayas et al., 2009). Christie et al. (2008) give an overview of non-economic techniques for assessing the importance of biodiversity to people in developing countries. Also, Pereira et al. (2005) provide some interesting non-monetary assessment methods.