Potentials of quantitative and qualitative approaches to assessing ecosystem services

Ecological Indicators 21 (2012) 89–103

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Ecological Indicators journal homepage: www.elsevier.com/locate/ecolind

Potentials of quantitative and qualitative approaches to assessing ecosystem services Malte Busch a,∗ , Alessandra La Notte b , Valérie Laporte c , Markus Erhard c a

Helmholtz-Zentrum Geesthacht, Institute for Coastal Research, Human Dimensions of Coastal Areas, Max-Planck-Straße 1, 21502 Geesthacht, Germany European Commission Joint Research Centre, Institute for Environment and Sustainability, Via E. Fermi 2749, I-21027 Ispra (VA), Italy c European Environment Agency, Kongens Nytorv 6, 1050 Copenhagen K, Denmark b

a r t i c l e

i n f o

Keywords: Ecosystem services Qualitative assessment Quantitative valuation Offshore wind farming Forest and coastal ecosystem services Methodology integration

a b s t r a c t Quantitative and qualitative approaches to assessing and valuing ecosystem services have been compared using case studies from Italy and Germany. This paper describes and analyzes two different methods of applying an ecosystem service approach and discusses their relative strengths, shortcomings and peculiarities. This allows the conditions to be identified that best support the application of one or other of these methods. Suggestions are offered on how to integrate both methodologies in order to improve the implementation of an ecosystem service approach in decision-making processes. © 2011 Elsevier Ltd. All rights reserved.

1. Introduction The Millennium Ecosystem Assessment (MA) has highlighted the essential “non-marketed services” provided by ecosystems. Because their benefits (such as water purification, carbon storage, landscape beauty) often remain unrecognized, non-marketed services are regularly degraded as a result of actions taken to increase the supply of marketed ecosystem services such as food or timber production (Butchart et al., 2010; Harrison et al., 2010; CBD, 2010; MA, 2005). Although these ongoing degradations can have major socio-economic effects, their impacts are rarely acknowledged. Such degradations are, furthermore, predicted to worsen if no action is taken, resulting in potentially irreversible changes (MA, 2005; Rockstrom et al., 2009). The MA (2005) defines ecosystem services as the benefits that humans obtain from ecosystems, identifying four interacting categories: “provisioning” (such as food and timber), “regulating” (such as air and water purification), “cultural” (e.g., recreational opportunities), and “supporting” ecosystem services (i.e., services that underpin all of the above services, e.g., nutrient cycling). Ecosystem services are, however, strictly linked to the spatial dimension of the defined area in which those services are provided. The term “ecosystem service approach” used in this paper is based on this understanding and comprises the application of ecosystem services as units for assessing, describing and predicting the interrelationships between ecosystems and human activities.

∗ Corresponding author. Tel.: +49 4152 87 18 26; fax: +49 4152 87 2020. E-mail address: [email protected] (M. Busch). 1470-160X/$ – see front matter © 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.ecolind.2011.11.010

The assessment of ecosystem services is a first step towards documenting changes in their nature and availability. It includes the identification of pressures acting on these services, and the identification of human populations that are most vulnerable to the effects of such changes. In case there is no quantitative information to perform monetary valuation, qualitative assessments can be used to address state and trends. Scenario developments allow alternative future courses of action to be investigated and the likely effects of economic and political decisions on ecosystem services to be assessed. The various goods and services that humans derive from ecosystems are often subject to trade-offs, in that enhancing market valued one often comes at a cost to another non-valued service (see e.g., Carpenter et al., 2009; Raudsepp-Hearne et al., 2010). Careful investigations may, however, allow the identification of “win-win situations”. Alternatively, where negative trade-offs are identified optimization to maintain the multi-functionality of ecosystems may reduce negative impacts or service depletion might be compensated. In addition to the assessment of ecosystem services it is useful to be able to provide an economic quantification of these services. For this purpose, tools are being developed to ascribe monetary values to ecosystem services (e.g., willingness to pay). Monetary valuation of ecosystem services will allow comprehensive costbenefit analyses under different scenarios, which are essential if economic incentives are to be changed. Such changes may include, for instance, the removal of subsidies triggering ecosystem service degradation, and the development of payment schemes for ecosystem services that currently have no market value. A comprehensive socio-economic framework is required if the pressures acting on ecosystem services, the interactions between them, and the impacts that changes in ecosystem service provision

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have on both economics and human well-being, are to be investigated. The MA framework, which is often taken as a starting point for ecosystem assessments, is based on determinations of human well-being, ecosystem services, and direct and indirect drivers of change. It aims to provide a summary of key interactions between nature and humans, thereby simplifying the issues by focusing on those interactions most important (Ash et al., 2010). It departs from the more traditional linear DPSIR framework by considering a dynamic system in which changes have a feedback effect on the acting pressures. There have been various attempts at ecosystem services classification based on the MA framework, aiming to adapt the approach to case-specific requirements (e.g., different data availabilities), or focusing on different environmental units (e.g., marine or terrestrial units). Alternative frameworks and classifications have, for instance, been developed in accounting in order to avoid double counting (Balmford et al., 2008). A standardized classification scheme has been proposed and currently discussed in the context of the System of Environmental-Economic Accounts (SEEA) of the UN Statistical Division (Haines-Young and Potschin, 2010). Ecosystem services have become an important subject for research in recent years because of their ability to reflect and communicate human-environmental interactions. Initiatives on a global scale included a study on the economic significance of the global loss of biological diversity, namely “The Economics of Ecosystems and Biodiversity (TEEB)” study (2008). Moreover, SEEA is under revision to include measures of natural capital in order to capture the depletion or degradation of these resources, responding to the need to move beyond GDP indicators as measures of sustainability and human well-being (European Commission, 2009). Nevertheless, ecosystem services have mainly been investigated on local scales (for reviews of case studies see, e.g., TEEB, 2010). The objective of this paper is to test, examine and compare quantitative and qualitative approaches for ecosystem services assessments in order to highlight their relative applicability, their potential and shortcomings, as well as potential mutual interaction in terms of providing decision support. By providing case studies implementing both approaches, the paper expands the pool of practical experiences and contributes to the discussion on ecosystem service assessments. After recalling the conceptual background of the approach (Section 2), case studies illustrating qualitative and quantitative attempts to ecosystem service assessment are presented (Section 3). The case study section has two parts. For better comparability, first qualitative and quantitative methodologies applied are described for each case study (Sections 3.1 and 3.2), followed by a description of the respective results (Sections 3.3 and 3.4). The succeed discussion (Section 4) compares methodologies and results obtained initializing concluding comments applicable to future assessments (Section 5).

2. The ecosystem service approach: conceptual background The concept of ecosystem services addresses the multifunctionality of ecosystems, the overall list of direct and indirect products, functions and processes each ecosystem is able to provide. It allows the attribution of an – in theory – unlimited number of service layers, each with its own use and value. The MA (2005) provides the probably most common list of services. From a normative point of view the approach is utilitarian, which means the ecosystems are only analyzed in context of human use and wellbeing. This makes the approach especially applicable for economic valuation also addressing the impacts of (human-induced) environmental changes on service provision. As such it can demonstrate current gaps in economic accounting and requirements to include

currently non-valued services into the overall economic calculation the way forward from GDP to green GDP. Application of the ecosystem service approach implies three major issues: (i) classification, (ii) scale, and (iii) dealing with synergies, trade-offs and non-accountable services. The definition and classification of ecosystem services (e.g., Fisher et al., 2009) and methods of dealing with multiple ecosystem services in relation to multiple (human) beneficiaries remain subject of on-going discussion and research. The two items are often mixed which can lead to inconsistencies in the overall approach (Rounsevell et al., 2010). Dealing with multiple services and beneficiaries can also lead to double, or even multiple accounting which will bias the monetary valuation of natural capital in environmental and economic accounting schemes (Boyd and Banzhaf, 2007; De Groot et al., 2009). Classifications should be geographically and hierarchically consistent in order to allow comparisons to be made between different regions, and to allow the integration of more detailed local studies into a broader geographical context. If local and regional studies are to be extrapolated to other geographic areas, whether this involves being generalized for larger regions of interest or used as input for more detailed studies, scaling will need to be taken into account. Scaling applies rules and limitations to the transfer, aggregation/generalization, and representativeness of the values to be used in a broader and/or different context. This is an issue that has previously been addressed by Costanza et al. (1997) and is described, e.g., within the context of TEEB study in the EEA’s 2010 report. The proposal for a common international classification of ecosystem goods and services (CICES) for integrated environmental and economic accounting addresses both, classification and scale (Haines-Young and Potschin, 2010). It aims for inter-comparability across countries but should also be consistent across scales. Local or regional specific issues have to be addressed to provide the relevant information for decision making. At the same time service assessments should also be compatible to larger spatial scales if non-specific services such as carbon or water are addressed. These can then be used for assessments on global and national level or vice versa. Addressing trade-offs poses another challenge for ecosystem service based economic valuation and qualitative assessment. The multitude of services and beneficiaries, as well as the frequently contradictory and often non-linear interrelationships and interactions, form major obstacles to the implementation of monetary valuations of ecosystem services into economic accounting frameworks. Trade-offs occur when an increase in one ecosystem service results in a reduction in another, and they can affect spatial scale, temporal scale, and reversibility (Rodríguez et al., 2006). An increase in food production, for example, often results in reduced biodiversity and vice versa. In addition, many ecosystem services have either not yet been quantified or are non-quantifiable (e.g., specific cultural services). Making full use of the ecosystem service approach consequently requires multi-decision making and adequate decision support systems combined with defined priorities and targets. It can target the optimization of services provided by an ecosystem, but requires quantification, targets and valuation of currently non-valued services to identify sustainable solutions. Furthermore trade-offs can differ depending on the actors and economic assets involved. The appropriate approach may be more quantitative or more qualitative, depending on the type and number of services addressed and the scale analyzed. The setting of normative priorities needs to be case-specific and spatially explicit, since priorities may change from one area to another and across scales. The indicators used to describe the “good ecological status” of water bodies in the EU Water Framework Directive (European Commission, 2000) are an example of a one-dimensional normative setting. The legally binding framework allows identifying distances

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between current status (measurements) and target (indicators) and the quantification of measures and costs to close the gap between the both. In terms of more qualitative assessments, data are usually normalized using relative scaling. This is the appropriate method if qualitative information is available. Normalization has also been used for vulnerability assessments (Schröter et al., 2005) if quantitative information of ecosystem services has to be overlaid for spatial explicit multi-sectoral assessments. The conceptual problems of classification, scale and trade-offs are addressed by several authors (e.g., Maness, 2007; Rodríguez et al., 2006). There are examples for implementation as also outlined in the following chapters. The studies do not provide the complete picture but relevant additional information how to implement comprehensive ecosystem assessments on local scale. A full proof of concept for consistent assessment of multiple ecosystem services, their synergies and trade-offs is still object of further research and assessments.

3. Case studies This section presents examples of qualitative and quantitative methodologies used for ecosystem service assessment, and the results of case studies from Germany and Italy. While the German offshore wind farm (OWF) case study represents a semi-qualitative approach combining qualitative and quantitative information in its assessment, the case studies from Italy, which focus on ecosystem services provided by a forest and an island, constitute purely quantitative methodology. The different outputs allow the contextspecific suitability of the respective methodologies to be estimated.

3.1. Semi-qualitative methodology: OWF case study Qualitative approaches to investigate ecosystem services are not able to ascribe monetary values, nor do they aim to. Instead they offer a means of assessing potential changes induced within socio-ecological systems by specific drivers of change. Examples can be found in the work of the Swedish Environmental Protection Agency (2008) in the Baltic Sea, and several other research projects (e.g., Vihervaara et al., 2009). Qualitative analysis of ecosystem services constitutes the precondition to succeed rating of environmental and societal changes, relative to an actual state or reference condition. The expected response of ecosystem services can be analyzed based on models, estimates of known causal interrelations, or expected changes can be assessed through interviews or questionnaires. Qualitative rating can, however, be based on qualitative, quantitative, or a combination of both assessment processes (Carpenter et al., 2006). The German “Zukunft Küste – Coastal Futures” research project (2004–2010) offers a valuable example of an ecosystem services assessment combining qualitative and quantitative analysis, addressing the potential impacts of the introduction of an additional industrial use to marine areas with the aim to provide a qualitative rating of the identified impacts. The project focused on changes within the socio-ecological system relating to the construction and operation of OWFs within the German North Sea. The project aimed to answer the following questions:

1. What ecosystem services are provided by North Sea marine and coastal ecosystems? 2. Which of these ecosystem services will be impacted by the introduction of OWFs, and what process-reinforcing or processdiminishing impacts can be expected? 3. How will these impacts vary across different spatial scales?

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The condition of the North Sea at the beginning of the project in 2004, without any OWFs, was defined as the reference state. The project looked at the German Federal Government’s (2002) political target of OWF construction within the German North Sea Exclusive Economic Zone (EEZ) with a wind energy capacity of approximately 25,000 megawatts by 2030. The ecosystem services approach was used to assess both the potential ecological changes induced by the construction and operation of OWFs and the socio-economic implications of offshore wind energy. The relevant spatial scales for this assessment were identified and investigated, ranging from local to macro-regional (southern North Sea) scales, in order to make the potential impacts and their intensities traceable across a large geographic area. The appropriate scale varies depending on the specific ecosystem service category. While, for example, “local” was defined as the pile and its foundation of a single wind turbine when considering regulating or provisioning services, for cultural services it was taken as “islands and (coastal) municipalities”, since cultural services are exclusive to areas of human population. A transdisciplinary team (natural, environmental, and social scientists) used the MA (2005) in a modified version (Fig. 1) to identify and analyze relevant marine and coastal ecosystem services in relation to offshore wind energy (Fig. 5), and classified them into either ecological integrity, provisioning, regulating, or cultural, ecosystem services. Indicators of ecological integrity (energy cycling, nutrient cycling, storage capacity, minimization of nutrient loss, abiotic heterogeneity, biotic diversity and organization) were used to analyze the potential ecological impacts of OWFs (Müller, 2005). Ecological integrity addresses the support and preservation of those processes and structures that are essential to the self-organizing capacity of ecological systems (Barkmann et al., 2001). The concept addresses abstract matter flows and structures securing ecosystem functioning just like supporting services, and does not focus on individual species or parameters. These abstract processes are a precondition for the provision of regulating, provisioning, and cultural ecosystem services. Impacts on ecological integrity were quantitatively assessed by computer-based ecological modelling (ECOHAM, ECOPATH models) and the use of GIS-edited monitoring data on sea bird and fish distributions. Physical modelling (HAMSOM, MIKE 21 models) addressed potential impacts of OWFs on regulating services, indicated for example by changes in climate regulation capacity and sea bed control functions (Lenhart et al., 2010). Changes to provisioning services such as food (fish, aquaculture products) or electricity provision, were estimated in qualitative terms on the basis of causal linkages and literature reviews. The employment potential of OWF development for coastal municipalities within the case study region was calculated with an economic input–output model (Hohmeyer et al., 2010). Cultural services were addressed via questionnaires (n = 387) answered by local residents indicating their opinion on OWF-induced changes of the seascape, regional image, educational and recreational potential, etc. (see Gee and Burkhard, 2010). This combination of quantitative and qualitative methods generated a holistic picture of the potential impacts of OWFs on each ecosystem service category. A qualitative rating of the expected changes in the provision and distribution of ecosystem services was then achieved on the basis of these assessments. In order to allow the predicted impacts to be compared across different ecosystem service categories and spatial scales (local, regional, German EEZ, southern North Sea), the impacts were qualitatively rated on a relative scale from −2 to +2, according to expert estimations based on the project’s results. A rating with −2 was taken to represent a “strong diminishing impact” of OWFs on the relevant ecosystem service, while +2 indicated a “strong

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Requirements:

Identification of relevant Ecosystem Services (ES)

Modification of ES according to case-specific requirements

Identification of natural, socio-economic and cultural processes potentially influenced by OWFs

1.) OWP must have an impact 2.) Spatial scale must differs from mainly global MA approach 3.) Exclusive concentration on marine ES 4.) Modification of ES with terrestrial focus for marine conditions

Line and verify developed hypothesis

Rating of impacts to make impacts on ES traceable and comparable Fig. 1. Project-specific application of ecosystem services as a holistic evaluation tool by adjusting to the case-specific conditions and requirements (Busch et al., 2010).

process-reinforcing impact”. These categories were chosen to avoid the normative terms “positive” and “negative” for changes of ecological processes and structures responsible for the provision of ecosystem services. How changes in service provision are perceived can depend on personal preferences or the indicator applied. While there is, for example, a provable positive relationship between the establishment of wind turbine piles and increases in (underwater) biodiversity (DONG Energy, 2006), this would not necessarily be rated as a positive development. Only if the chosen indicators targeted increasing biodiversity would a positive rating be given. If the applied indicator was “ecosystem nativeness” the establishment of piles would be given a negative rating. In summary, results of the qualitative rating indicate whether ecosystem services are likely to respond to certain developments with process reinforcement or with diminishment. An improved understanding of the potential consequences of intended development allows better informed decision-making.

3.2. Quantitative methodology: “forest” and “island” case studies The quantitative approaches to assessing ecosystem services referred to in this paper are integrated environmental-economic accounting systems that consist of systematically and consistently framed data sets based on national accounting systems. Environmental accounting schemes originated from the well known weaknesses of the traditional System of National Accounts (SNA), a standardized international system aiming to provide integrated, complete system accounts that allow all significant economic activities to be compared. The SNA does not take into account quantitative depreciation of natural assets, qualitative degradation of environmental media, or most types of environmental expenditure and, by ignoring these aspects results in a misleading description of a socio-economic system’s performance. At present the System of Integrated Environmental and Economic Accounting (SEEA) (UNSD et al., 2003) proposed by the United Nations Statistics Division (UNSD) together with Eurostat, the IMF, the OECD, and the World Bank represents the accepted status quo from the conceptual and methodological points of view, and will soon (2012/2013) become an internationally agreed accounting standard.

Evolving from the SEEA, the notion of ecosystem accounts started to develop when an ecosystem capital account was proposed within the Land and Ecosystem Accounts (LEAC), supported by the European Environment Agency (EEA, 2006, 2010). The LEAC framework is based on two components: ecosystem capital and ecosystem services, assessed on the basis of land use and land cover data, and offers considerable potential for development (Weber, 2007, 2008). It unites different scale-specific tools (global, national and local), as reported in Fig. 2. Two case studies, addressing the monetary valuation of ecosystem services provided by a forest and an island, serve as examples for quantitative ecosystem service assessment at a local (action) scale. Both case studies are located in the Veneto Region of northern Italy. Asset accounts were set up for the selected case study areas, with monetary valuation being framed according to the Total Economic Value (TEV) notion. This approach takes into account use (direct and indirect) and non-use (option, bequest and existence) values (for which handbooks and guidelines have been published, e.g., by CBD, 2007; UNDP and GEF, 2006; OECD, 2002; The World Bank and IUCN, 2004; US-EPA, 2009; Defra, 2007; TEEB, 2010). Environmental accounts allow the TEV to be calculated in order to assess the potential losses induced by ecosystem change and to estimate trade-offs between competing users and possible uses, and thus to plan economy-wide strategies taking into account goods and services (Lange, 2003). The procedure, which integrates environmental accounts with the TEV approach, is summarized in Fig. 3. Once the services to be assessed and valued have been identified (Fig. 3) an asset balance can be established for each service, first in physical and then in monetary terms. The balance sheet consists of an opening stock, a record of all changes occurring during the period under consideration, and a closing stock. As illustrated in Fig. 4, asset balances considered should not be restricted to provisioning ecosystem services, but include all ecosystem services categories (regulating, habitat, cultural) identified in a consistent way. Since ecosystem services are related to their provision areas, Geographic Information Systems (GIS) enable the integration of this spatial dimension into monetary valuations. Geo-referenced environmental accounts can thus make monetary valuation spatially explicit (De Groot et al., 2009; Turner et al., 2010).

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Fig. 2. Overview of the context (scale, governance, accounts and payments) of Land and Ecosystem Accounts (LEAC). (Graphic from Jean-Louis Weber, 2008.) Available at: http://unstats.un.org/unsd/envaccounting/londongroup/meeting13.asp?sID=3 (accessed 02.01.11).

3.3. Application and findings: OWF case study Ecosystem services potentially impacted by OWF development were derived on the basis of an inventory of ecosystem services

provided by the North Sea (Fig. 5). The strongest impacts were found for components of ecological integrity (Fig. 6) and provisioning services (see Busch et al., 2010). A large number of cultural services are potentially impacted by OWF development (Fig. 5),

Choice of the economic valuation tool according to the goal to be achieved

Environmental accounts feasible for multiple uses constitute the conceptual framework adopted

Identification of the functions to be valued

Functions vary according to use destination and characteristics of the areas

Choice of valuation methods

Different valuation methods correspond to different functions; GIS can play different roles in relation to the respective methods

Valuation in physical terms

Processing of quantities, indicators and qualitative classes

Valuation in monetary terms

Calculate the monetary values to 'attach' to physical values

Aggregation of function values according to appropriate criteria

Productive and non-productive values are aggregated in order to obtain the TEV. Georeferenced data might originate from different values for each georeferenced unit

Fig. 3. Procedure for integrating environmental accounts, and Total Economic Value (TEV) (La Notte and Turvani, 2007).

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OPENING STOCKS due to anthropic action Δ INCREASE Δ DECREASE

due to natural cause due to other causes

CLOSING STOCKS

PROVISIONING SERVICES Food production Raw material provision Water provision … ...

REGULATING SERVICES Erosion prevention Air quality regulation Waste treatment … …

HABITAT SERVICES Maintenance of lifecycle migratory species Maintenance of genetic diversity

CULTURAL SERVICES Opportunities for recreation and tourism Spiritual experience … ...

Fig. 4. Procedure for establishing asset accounts for selected ecosystem services (own design; La Notte).

indicating the potential of OWFs to impact on the coastal societal setting in multiple ways (e.g., a different regional image, new knowledge added to the region, perceived land/seascape changes). Results for ecological integrity components indicate that a majority of the identified impacts of OWFs are expected to reinforce ecological processes like e.g. energy and nutrient cycling, storage capacity, abiotic heterogeneity and others on a local scale (e.g., pile and F scales, see Fig. 6), with the impacts diminishing on larger scales (e.g., EEZ and southern North Sea scales), indicating a typical dilution process. The expected local reinforcement of ecological integrity parameters relates to the assumption that artificial reef ecosystems

emerge, as observed at Danish OWF sites (DONG Energy, 2006). The establishment of hard substrates on the sea floor (e.g., piles, scour protection) provide additional habitat for epifaunal macrobenthos (e.g., sessile filtering organisms or suspension feeders). This higher biotic heterogeneity is assumed to increase underwater biodiversity and the accumulation of biomass within OWFs, reinforcing further ecosystem services such as nutrient cycling and storage capacity. Negative impacts for biodiversity are assumed for above the water surface: monitoring has indicated strong avoidance by certain sea bird species (especially gavia stellata), which are expected to be negatively impacted on both local and supra-regional scales due to the cumulative impacts

Fig. 5. Project-specific ecosystem services of relevance to offshore wind farm development and components of human well-being likely to be affected (Busch et al., 2010).

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Fig. 6. Impacts of offshore wind farms on ecological integrity portrayed for different spatial scales. Relevant integrity parameters are located on the edges of the diagrams. The single blue lines show the qualitative rating values (−2 to +2) connecting the impact ratings of each ecosystem service, while the hatched areas show the impact range, indicating if an ecosystem service is affected by several factors. To indicate these variations and to provide a sophisticated description, the individual rating values are marked by green spots. The reference state (actual condition in the North Sea, without offshore wind farms) is shown in orange and set to “0” (Busch et al., 2010). (For interpretation of the references to color in this figure legend, the reader is referred to the web version of the article.)

Food - fishery Food - mariculture

2

1

Impact

of proposed extensive OWF developments (Mendel and Garthe, 2010). The main benefit of OWFs is the technical transformation of wind (a regulating service) into electricity. Whether this makes wind energy a provisioning ecosystem services or an anthropogenic utilized environmental service is discussable. Further obvious provisioning services are influenced by OWF development (e.g., food – see Fig. 7). The food (fish) ecosystem service in particular shows interesting responses to OWF development. Safety regulations in German waters prohibit shipping within 500 m of wind farms. This results in the complete exclusion of fishing activities within these areas, allowing fish populations to recover. Wind farms can also provide additional nutrition and shelter as a result of the emergence of artificial reef ecosystems. Fish populations in the vicinity of OWFs are expected to benefit from these developments, but no changes are assumed on an EEZ scale. The food (mariculture) ecosystem service shows a positive response, which relates to the potential co-use of OWFs as mariculture sites. The establishment of turbine piles provides infrastructure suitable for mussel or algae cultivation (MichlerCieluch and Krause, 2008). This result in an increased potential to provide this ecosystem service that is exclusive to OWF areas, since water depth and harsh offshore conditions otherwise prevent this type of mariculture development.

0

-1

-2

OWF

close surroundings

EEZ

Spatial Scales Fig. 7. Impacts of offshore wind farming on two food-related provisioning ecosystem services across the investigated spatial scales (OWF: Offshore Wind Farm, close surroundings: proximity to offshore wind farms, EEZ: Exclusive Economic Zone) (Busch et al., 2010). (For interpretation of the references to color in this figure legend, the reader is referred to the web version of the article.)

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Table 1 Ecosystem services, valuation methodology and monetary values calculated for the “forest” case study. Ecosystem service

Measurement unit in physical terms

Valuation method

Monetary valuation data source

Timber provision

Tons of standing timber

Market price

Carbon stocked

Carbon tons

Emission trading scheme

Erosion prevention

Hectares classified in classes according to risk control and vegetation cover indices Hectare classified according to species varieties, sensitivity and pressure indices Hectare classification according to selected features

Replacement costs

Local timber auction/data provided by agents IETA Report (Lecocq and Capoor, 2005) Handbook on environmental engineering intervention costs

Gene pool protection

Recreation

Value transfer (contingent valuation)

Travel cost method

Total monetary value

These examples demonstrate that the ecosystem service approach can serve as a valuable instrument for assessing the expected impact of a human activity on the composition and provision of ecosystem services. Within each service category, changes and trends of relevance to decision-making (e.g., related to the siting of OWFs) were identified, analyzed and rated. Interesting co-use potentials and side effects that reinforced ecological processes were identified, as well as expected impacts that should be addressed in a precautionary manner within OWF approval procedures, before commencing large-scale installation. The difficulties and short-comings of this approach in dealing with a strongly delimited subject (a specific anthropogenic activity at a specific spatial scale) became obvious and identified the need for further structural debate. Another challenge relates to the use of average values when an ecosystem service is measured on the basis of multiple parameters. How to weight the emergence of additional species of filtering organisms colonizing the piles of wind turbines against the loss of certain sea bird species of increased conservation concern in terms of overall biodiversity. In the absence of a common unit to balance, qualitative research answers this challenge by indicating the range of impacts (see Fig. 6, hatched areas). The results also suggest that ecosystem service trade-offs may be difficult to avoid with OWFs, especially in consideration of the massive development intents. The ecosystem service approach proved to be a valid tool for identifying these trade-offs (e.g. offshore wind electricity vs. sea bird abundance or space for fisheries), thereby supporting better informed decision-making in terms of policy interventions. Ideally, monetary quantification of these trade-offs would be the next step. A more detailed analysis and description of the results of the “Zukunft Küste – Coastal Futures” project can be found in Lange et al. (2010). 3.4. Application and findings: “forest” and “island” case studies Geo-referenced accounts for selected ecosystem services were applied at a local level in two case studies, in order to estimate their monetary value. The two areas were chosen to test whether the whole procedure is feasible in different contexts, be it a single ecosystem (a forest) or a heterogeneous environment (an island and its surroundings), whether there is a default minimum reference unit (such as a forest management parcel) or whether one needs to be established (e.g. a grid), and whether all the relevant data are in place or whether they need to be adapted from other datasets and studies. In the “forest” case study, the monetary value of a forest (covering 3347 ha) was assessed to test the feasibility of a monetary valuation that includes more than just the market value of the timber and thus justifies conservation policies (La Notte, 2006). In the “island” case study, an island of 3.24 km2 within the Venetian Lagoon was studied to assess the monetary consequences of

Study on the diversity of biodiversity valuation (Christie et al., 2006) Valuation study undertaken for the area

1000 (D/ha)

% value

6

9.53

9

14.26

30

45.99

17

26.05

3

4.14

65.5

potential development projects impacting on the ecosystem services provided by the lagoonal system (La Notte and Turvani, 2007). Without presenting the details of the monetary valuation, that are briefly reported in Appendix A, the steps that were undertaken and how the assessment thus attained can assist in a trade-off analysis are briefly shown. In the “forest” case study the following services were considered and assessed: timber provision, climate regulation, erosion prevention, recreation, and gene pool protection. Table 1 presents the physical to assess each ecosystem service and the monetary valuation techniques applied. The calculated value per hectare is shown in Euros, together with its relevance shown as a percentage of the value calculated based on these five services. The spatial distribution of higher and lower monetary values within the forest parcels is illustrated in Fig. 8. However, only partial monetary values are presented for these forest parcels because additional ecosystem services (e.g., water purification, flood regulation, etc.) have either not been valued or cannot be valued in monetary terms (e.g. oxygen production, spiritual values, etc.). The calculation provides an opening stock value for the asset balance of each ecosystem service identified. The same services can then be assessed (using the same methodology) after 5, 10 or 15 years in order to detect any changes, in both physical and monetary terms. We can also hypothesize on how the monetary value will change if a particular policy is applied. The distribution of monetary valued services illustrated in Fig. 8 indicates spatial heterogeneity, with some parcels having higher monetary values than others. A high value for a particular parcel can be due to all services within that parcel having high values, or to low values for some services being compensated by very high values for others. It is proposed to investigate the spatial interrelationship between different parcels in order to understand how the monetary value of any one parcel responds to value changes in neighbouring parcels. The economic consequences of potential future developments can be explored on the basis of the monetary values determined for particular ecosystem services. What happens, for example, if a low market value results in timber no longer being harvested? The timber provision service will decline in value, while both the carbon stock and gene pool protection will increase. If the area is left unmanaged the recreational service value will decrease. The monetary value per parcel, as well as for the total area, will therefore change. The proportional influence of each ecosystem service on the total value will vary and, based on the “translation” into a commonly understood value per unit (e.g. D/unit), can be measured in a consistent way. The same approach was used for the “island” case study, but as well as the climate regulation and recreation ecosystem services of the “forest” case study, the food provision, flood prevention and

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Fig. 8. Mapping of the Total Economic Value calculated for the “forest” case study, and its location within the Veneto Region in northern Italy (La Notte, 2006). (For interpretation of the references to color in this figure legend, the reader is referred to the web version of the article.)

maintenance of life cycles and gene pool protection were assessed (Table 2). Although the approach is similar, the assessment techniques for the “island” case study are different. For recreation services, for example, the approach is the same in physical terms but the monetary technique is different. In both cases primary studies were specifically undertaken for the area under consideration. However, the main differences are the features taken into account to buffer the physical map (such as the type of tourism that takes place, the income that it brings, or the fact that the “forest” case almost exclusively cater for local tourism while the “island” case caters for mixed local-global tourism), and the monetarisation methods chosen. While the travel cost valuation method operates with “indirect revealed preferences”, the contingent

valuation method works with “direct stated preferences”, as defined in the valuation literature. Contingent valuation was preferred in this case because it meant that the results of a study undertaken specifically for this area could be utilized. The high average monetary values ascribed to ecosystem services in the “island” case study (Table 2) relate to its location within the Venetian Lagoon. The willingness to pay (used for valuing maintenance of life cycles and recreation) assessed via telephone interviews correlates well with the attractive location of the island. Moreover, the defensive expenditures (e.g., against flooding) have a higher value due to the island’s location. A comparison of the monetary values ascribed to specific ecosystem services within both Italian case studies reveals large

Table 2 Ecosystem services, valuation methodology and monetary values calculated for the “island” case study. Ecosystem service

Measurement unit in physical terms

Valuation method

Monetary valuation data source

Food provision

Produced quantities Cultivated hectares Carbon tons

Market price

Hectares within selected buffer areas Hectare classification according to selected environmental indicators Hectare classification according to selected features

Defensive expenditures

Interview with farmers and INEA data IETA Report (Lecocq and Capoor, 2005) Expenditure balance for St. Erasmo works Valuation study undertaken for the island

Carbon sequestration Flood prevention Maintenance of life cycles

Recreation Total monetary value

Emission Trading Scheme

Contingent valuation: existence value

Contingent valuation: use and option value

Valuation study undertaken for the island

1000 (D/ha) 33 0.5

% value 5.8 0.09

315

55.41

120

21.11

100

17.59

568.5

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Fig. 9. Mapping of the Total Economic Value calculated for the “island” case study; the black line marks the island within the Venetian Lagoon (La Notte and Turvani, 2007). (For interpretation of the references to color in this figure legend, the reader is referred to the web version of the article.)

disparities in terms of value placed on different ecosystem services. We need to keep in mind that different methods were used, and rather different ecosystem services have been quantified. Those valuations make sense in their local context for planning and decision making. To enable a comparison of different sites, ecosystem services assessed and valuation techniques applied need to be identical. Fig. 9 illustrates how services such as flood prevention, recreation, and maintenance of life cycle can increase the value of the island and its surroundings considerably. If only food production is considered (the only service that passes through a market) the highest values would be concentrated in the island. Recalculating the monetary value of the ecosystem services in a few years’ time will be interesting in order to investigate how the construction and operation of various artificial interventions, such as a long pier at the Lido inlet in front of the island, will affect the lagoon ecosystem. The internal lagoon environment is expected to change, and consequently the environmental indicators used to assess the physical mapping of the maintenance of life cycles and carbon sequestration ecosystem services will be affected. Moreover, the flood prevention service provided by the (barrier) island by reason of its location is also expected to change. The total monetary value obtained by summing up all the services will therefore be different and the weighting of each service will change. By ascribing monetary values to ecosystem services, quantitative assessments generate information that is easy to understand, compare and communicate. The drawback lies in the subjectivity involved in deciding how to assess each ecosystem service in physical and monetary terms. This can to some extent be overcome by

providing transparent metadata information: the more trackable the assessment process (with its assumptions, limitations, etc.) the more reliable the interpretation of the data will be. 4. Discussion Recalling the objective of this paper to test, examine and compare qualitative and quantitative approaches to ecosystem service assessment in order to highlight their relative applicability, their potential, shortcomings (Section 3) and mutual interaction (Section 5), the discussion will explicitly focus on the potential of both approaches to support practical decisions on policy interventions. The decision on which approach to choose will regularly depend on the following criteria: (1) the type of information available, (2) the spatial scale of the area a decision should be made on, being a proxy for the expected accuracy of produced evidence and (3) the type of (political) intervention it should serve. 4.1. Information/data availability A first criteria determining which approach to ecosystem service assessment to choose and what type of political intervention to inform, is information availability. Accounting for a complete set of ecosystem services requires a large amount of very detailed information which will not commonly be available. Consequently, ecosystem services can be overlooked, leading to the monetary value being underestimated as a result of not including the full set of benefits that ecosystems provide for human well-being. More flexible qualitative approaches can bridge the gap of limited data availability with well-grounded estimations, for example based on

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in-depth interviews, etc., and allow questions that are out of an accounting range to be addressed. The potential of quantitative assessments is to analyze restricted sets of ecosystem services going along with comprehensive information availability. It is highly relevant that data are consistent for all ecosystem services analyzed to allow a consistent application of valuation techniques. 4.2. Scale Another aspect that might influence the choice between a qualitative or quantitative approach is the spatial scale an ecosystem service assessment addresses. The OWF case study demonstrates the suitability of its approach to assess a broader set of ecosystem services and trace changes in service provision from local to macroregional scales. A qualitative approach, flexible to combine a range of different methods (see Section 3.1) and capable of including data being inconsistent for quantitative analysis, seem to be better applicable to large scale assessments. The central advantage might be that qualitative rating based on a relative scale (e.g. −2 to +2) allows merging different information into one common statement. Outputs can provide guidance for strategic political interventions and identify crucial aspects for further analysis. To assess ecosystem service change at local scale accurate data are required, as sensitivity to impreciseness increases due to a lower total amount of data. Under those circumstances quantitative methodology allows to address delimited subjects (e.g. a local forest ecosystem) in a more detailed way than qualitative assessments are capable of. An effect restricted to a specific area allowing addressing a narrower context appears to be beneficial in terms of conducting a precise monetary valuation. Nevertheless, quantitative assessments are not necessarily restricted to local scales. The methodologies are applicable at any scale if not limited by data availability. In cases like the OWFs case study where ecosystem service change is assessed across different spatial scales quantitative assessments need to be aware that in accordance to the adopted scale, the identification and measurement of the most relevant ecosystem services can change while valuation techniques require consistent application. 4.3. Type of political intervention to serve Depending on whether ecosystem service assessment is carried out to inform strategic or concrete management/financial decisions the chosen approach might vary as qualitative and quantitative ecosystem service assessments differ in terms of the outputs they are able to provide. The qualitative analysis of impacts of OWFs on related ecosystem services indicates trends like e.g. the reinforcement of biomass storage capacity within OWFs due to additional space for filtering organisms to colonize turbine piles and scour protection. Moreover, mutual interactions related to the provision of certain ecosystem services can be identified, e.g. OWFs compromise the availability of marine space for other marine uses like fishery in terms of access to fish resources (ecosystem service food). That information is valuable in terms of strategic decision making for example in context of marine spatial planning (MSP) dealing with the allocation of marine space between conflicting uses or support decision on whether OWFs could act as marine protected areas for certain genera (e.g. due to exclusion of fisheries). In contrast quantitative case studies provide concrete monetary values for individual ecosystem services, allowing to precisely trade-off different development scenarios in financial terms or investigate concrete costs of conservation efforts. Where e.g. to harvest timber without compromising the forest’s erosion prevention function and sparing forest parcels with high relevance for gene pool protection? Following a strategic decision like for

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example managing a forest in a sustainable way, quantitative ecosystem service assessments are able to provide the information required to plan and implement such management schemes. Consequently, quantitative approaches are especially adequate in situations where firm, clearly defined (economic) decisions of short-term significance are required. 4.4. General discussion After discussing the criteria for choosing the context specific appropriate approach to ecosystem service assessment to facilitate environmental decision-making it can be stated that qualitative assessments appear valuable to inform more abstract decisions at superior planning stages. They can provide wellgrounded estimations of large scale ecosystem response to certain human interventions, e.g. OWF development, and facilitate strategic planning and policy preparation. Qualitative rating of future developments can help address potential consequences in a precautionary way. This will assist in reducing the lopsidedness of trade-offs, for example, by identifying co-use potentials. The main advantage of qualitative ecosystem service assessment is therefore that it can indicate trends, including associated potentials and conflicts, while avoiding the risk of over-interpreting data that is insufficient for the delivery of quantitative information. Nevertheless, when considering a whole set of ecosystem services for multi-dimensional service assessments, normalization is regularly applied in order to present the variety of services and beneficiaries per spatial unit while ignoring their relative importance (including, e.g., scarcity, importance for biodiversity or recreation, etc.) in a spatial and environmental context. However, this limits the usefulness of such assessments for decision making (see Metzger et al., 2008). At the same time results derived by traditional accounting valuation techniques, will often better suit the information requirements to trade-off policy options at local scale. The monetary values ascribed to ecosystem services present an easy communicable currency that everyone is familiar with and allows smooth integration into existing economic schemes. Politicians addressing clearly located subjects will therefore regularly prefer to justify political actions based on a concrete cost-benefit analysis than on projected trends of a qualitative assessment. However, undertaking quantitative assessments can be very expensive and time consuming in the absence of physical and monetary assessments to build on, or if reclassification and harmonization of data is required. This conflicts with the needs of e.g. politicians who often think in terms of legislative periods and are in support of fast decisions. Consequently, the ultimate objective of a proposed study is an important factor when determining the methodology to be applied. The objective, shaped by the spatial scale under consideration and the decision maker to inform, determines the perspective on valuation and, depending on whether valuation is understood as a process of discussion or calculation, the methodology to be applied. 5. Conclusions Quantitative approaches that allow accounting and monetarisation of ecosystem services are characterized by the generation of results that can be applied to existing economic systems and are hence easy to integrate into decision-making processes. Monetary valued ecosystem services can be used to analyze the full costs and benefits of ecosystem change, and to enable non-marketed natural capital to be taken into consideration. If it is to be explicit and significant, quantitative information requires comprehensive data, the absence of which commonly limits the scope of quantitative ecosystem service assessments to a few specific services

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and a single defined scale. While this may be ideal in terms of clearly delimited decisions or trade-offs, it can cause problems when broader issues are addressed due to the exclusion of some ecosystem services and consequent underestimation of the overall monetary value. Qualitative approaches to ecosystem service assessment are characterized by their ability to provide a comprehensive analysis of ecosystem change and its consequences, taking into account a whole range of ecosystem services. This is due to the fact that qualitative methodology allows the integration of wellargued and informed interpretations of unquantifiable data and the consideration of causal linkages within assessment processes. Qualitative assessments are valuable in terms of providing an overview, indicating trends, and identifying trade-offs, which subsequently require an in depth analysis. Consequently, they are especially valuable for strategic decision-making processes and impact assessments on ecosystem services across various spatial scales. However, qualitative information can lack the explicitness and accountability of quantitative information and should be used as a proxy indicator. Both approaches are valuable in terms of exploring the impacts of alternative development paths on ecosystems and the human beneficiaries, or envisioning management options based on a stateof-art assessment. Depending on the firmness of development intents, the availability of data, and the scale of the investigation, decision-makers should select the appropriate assessment approach on a case-specific basis. Qualitative approaches could raise issues, ascribe weight to those issues, and suggest the most relevant scale for more precise quantitative studies aiming for detailed measurement and/or monetarisation of specific ecosystem services. Qualitative analysis would thus partially function as an initial analysis step, identifying starting points for quantitative measurements. A comparison of results obtained by both methodologies would allow a cross-check either reinforcing the results (if both assessments point in the same direction) or identify discrepancies. Any discrepancies would raise the question: “What did not work out, and why?” This could trigger methodological progress helping to address the big challenge for both qualitative and quantitative ecosystem service assessments: effectively capture the complexity of ecosystems. Acknowledgements The OWF case study was carried out within the research project ‘Zukunft Küste – Coastal Futures’ funded by the German Federal Ministry of Education and Research (BMBF FKZ 03F0476 A-C) during the period 2004–2010. Conceptually, the EU project “KnowSeas” funded by the European Community’s Seventh Framework Programme [FP7/2007-2013] under grant agreement number 226675, contributed to the work on this paper. Moreover the authors want to thank two anonymous reviewers for their helpful comments. The valuation of the Cansiglio forest was undertaken during a two year post-doc fellowship at the University of Padua, TESAF Department. The valuation of St. Erasmo island was funded by the University of Venice IUAV-Department of Planning. Appendix A. The appendix provides details about the data which have been used for calculating and mapping the values for ecosystem services in Cansiglio forest and St. Erasmo island (see Section 3.4). A detailed description of procedures and valuation techniques is given in La Notte (2006) for the first case study (Cansiglio forest) and La Notte and Turvani (2007) for the second case study

(St. Erasmo island). The following paragraphs briefly illustrate the main steps of calculation. In some cases (e.g. food and timber production) the value of resources/goods is used as a proxy for the ecosystem service and its market value (supply-demand sides) is the monetary value attributed. In other cases (e.g. erosion prevention, recreation) ecosystem services are valued through models and biophysical mapping procedures; supply-demand and demand side valuation techniques are used for the monetary valuation. Although we do agree that aggregating assets, services and different types of valuation techniques might be debatable, we would like to emphasize how the value of each service (or the resource used as a proxy to value the service) and its weight on the overall value calculated can change when different policies are applied. A.1. Cansiglio forest A.1.1. Timber provision We used the timber stock expressed in tons per hectare as a proxy. Timber services refer to standing timber from natural forests and plantations. The capacity of forests to produce timber was approximated using the geo-referenced inventory of the areas’ Forest Management Plan. The total cubic meters of timber are differentiated according to the species (beech, fir, larch, pine). For natural forest, the monetary value is calculated by multiplying the quantity of timber and the stumpage price per species. Ranges in timber value are due to timber quality and logging costs, which increase when territory conditions are imposing certain limits. In particular accessibility, slope and roughness are considered as factors making logging operations more expensive and thus decrease the value of woodland. For plantation and specifically those patches where even-aged beech wood is planted an additional analytical step is required before multiplying quantity by price: because clear cut occurs as part of a regular plantation cycle, it would not be correct to account for the full standing timber. Deferred discounted annuities are computed: a=

An × r

(1 + r)t − 1

where a is the annuity value; An is the steam wood value at the end of the cycle; r is the discount rate (1.6%); t is the patch age. A.1.2. Climate regulation We used tons of carbon stocked in soil and biomass as a proxy. Carbon stock valuation requires the aggregation of carbon accumulated in both above ground biomass and soil. The formula used to calculate standing timber carbon is: Qc = [Vj + (Vj × C1 ) + (Vj × C2 )] × C3 × C4 where Vj is the volume of stem wood in cubic meter; C1 is the coefficient to value branches and leaves according to the tree species. The adopted coefficients are usually higher for broadleaves and lower for conifers; C2 is the coefficient to value roots. C3 is the conversion factor from cubic meter to tons of dry matter; C4 is the conversion factor from tons of dry matter to tons of carbon. There is no precise formula to calculate soil carbon content. Even if only as rough estimate, we can refer to the content of organic carbon depending on type and depth of soils according to Batjes (1996). Monetary values are obtained firstly by transforming C (carbon) tons into CO2 tons, and secondly by multiplying the resulting tons by the price of the Emission Trading Scheme reported for Europe. A.1.3. Erosion prevention We used a model that values the erosion risk control. The model (Giau, 1996) implements a relationship between erosion risk and vegetation cover. The multiplicative model consists of an erosion

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risk index and a cover index, known as ‘parameter C’, addressing the protective function of the vegetation. Parameters taken into account for the erosion risk index are texture, soil depth, roughness, rainfall, temperature, slope and speed of pedogenesis. The monetary valuation is undertaken by using replacement costs defined in environmental engineering handbooks. For each hypothetical work to be applied at a specific location, as substitute for the forest cover, a maximum length of time is considered in order to calculate a yearly depreciation amount. The discount rate is 1.5%. CS1 = Csqm × 10, 000 CS2 = CS1 × #HA  r  CS3 = CS2 × qn − 1 where CS1 is the replacement cost per hectare; Csqm is the replacement cost per square meter; CS2 is the total replacement cost; #HA is the total number of hectares; CS3 is the yearly depreciation amount; r is the discount rate; qn = (1 + r); (1/qn − 1) = discount factor; n is the length of time of the intervention expressed in number of years. In order to assign correctly the hypothetical intervention to be undertaken across all patches, the territory was classified according to characteristics such as slope, soil depth, size, proximity to the road network, landslide and avalanche risk.

A.1.4. Recreation A model focusing on territorial characteristics valued by typical ‘users’ for recreational purposes was applied. A formerly conducted detailed study on the Cansiglio area allowed tracking where visitors come from, which kind of recreational activities they practice and what kind of landscape they prefer. Based on this information, spatial qualities critical for the recreational value were identified. Spatial queries were used to combine these qualities and (in some cases) buffer them to obtain a recreational-zoning map. Three studies specifically conducted on the Cansiglio forests were used to attribute monetary values to recreational features. The number of visitors was computed based on estimates according to Tosi (1989).

A.1.5. Gene pool protection We used a model based on the biophysical mapping of selected environmental indicators. These indicators were proposed by the project CartaNatura 1:50,000 for covering requirements of the EU Habitat Directive (92/43/EEC) and are based on the CORINE biotope classification of the European Union (EEA, 2005). CartaNatura values are based on an index processed from three macro-indicators referring to (i) ecological value of biodiversity, (ii) sensitivity of biodiversity and (iii) the impact of human pressure on biodiversity. According to the value of these indicators higher and lower values are attributed to each patch of the Cansiglio forest. Once this step was completed, a monetary value was attached. For the Cansiglio area no studies are available to value this kind of service. The Benefit Transfer technique was applied to value this function. The values are ‘stated preferences’ and was applied in UK, specifically in Cambridgeshire and Northumberland (Christie et al., 2006). Values are attributed to governmental hypothetical intervention such as protect rare and familiar species from further decline, that can be considered representative for the Ecological Value, slow down the rate of rare and unfamiliar species, that can be considered representative for the Ecological Sensitivity, and habitat recreation to restore functions lost because of human intervention, that can be considered representative for human pressure.

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A.2. St. Erasmo island A.2.1. Food provision We used the crop production of the island as a proxy. The island’s agricultural production is valued through INEA1 estimated prices considering the value (D) per hectare of agricultural area and the type of cultivation, specifically for St. Erasmo: horticulture, vineyards and permanent cultivation. Special attention was on the violet artichoke, the most valuable agricultural product on the island. In order to assess its price and the produced quantities, the organization of local producers was interviewed. A.2.2. Climate regulation We used the tons of carbon stock in soils and sandbanks as a proxy based on Batjes (1996) using the soil type and its depth for estimating organic carbon content. Once the carbon quantity was attributed, monetary values were calculated using the price of Emission Trading Permits per ton of CO2 in Europe. A.2.3. Flood prevention and pollution mitigation Given the position of the island within the lagoon and towards the Lido inlet, the protection against high tide and flooding can be considered as the major contribution to the safeguard of the island’s environment. Recently, several public projects were undertaken by the authority responsible for this objective (Magistrato alle Acque). The extent and the type of works relevant for our purposes were identified, located and buffered in physical terms in order to proceed with the valuation in monetary terms, through defensive expenditures. We use the costs of the public works undertaken. Specifically, sandbanks perform an important protective role in terms of pollutant sinks and therefore the reconstruction, existence and maintenance of sandbanks is important also for pollution mitigation. A.2.4. Recreation We use a model based on the characteristics of typical ‘users’ of this territory for recreational purposes. Recently, the island of St. Erasmo was object of a study by local research institutions (such as CORILA) and local Universities. The results were used for the assessment. We were able to specify the kind of recreational facilities, the typology of visitors and the monetary values to be attributed. The identified main attractions for recreational purposes were spatially zoned, distinguishing between those attractions where the relevance for tourists depends on distance and those where existence is more important. After production of a recreational-zoning map with a number of classes, we used the results of a contingent valuation study aimed at estimating the use and non-use value of the island (Alberini et al., 2004a, 2004b). Specifically, we applied the ‘use’ and ‘option’ values. The full monetary value was attributed to the highest class, while descending values were attributed to the lower classes. A.2.5. Maintenance of life cycles and gene pool protection We used a model based on the biophysical mapping of selected environmental indicators as described in the ‘Atlante della Laguna’ (Guerzoni and Tagliapietra, 2006). This publication includes a vast amount of environmental indicators which describe the physical and biological components of the Venetian Lagoon environment. We selected those indicators considered to be most important and then, identified and located the related areas to build qualitative classes. In order to assign a monetary value, we used the existence value elicited from the aforementioned contingent valuation

1 INEA is the National Institute of Agricultural Economics and all data was taken from its web site http://www.inea.it/.

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