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Ecosystem Services: concepts, methodologies and instruments for research and applied use

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Ecosystem Services: concepts, methodologies and instruments for research and applied use

Sergi Nuss-GironaMita Castañer

(eds.)

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Qualsevol forma de reproducció, distribució, comunicació pública o transformació d’aquesta obra només pot ser realitzada amb l’autorització dels seus titulars, llevat excepció prevista per la llei. Dirigiu-vos a CEDRO (Centro Español de Derechos Reprográficos) si necessiteu fotocopiar o escanejar algun fragment d’aquesta obra (www.conlicencia.com; +34 91 702 19 70 / +34 93 272 04 47)..

Direcció de la sèrie: Josep Vila SubirósCorrecció dels textos originals: Redactum

Coordinadora científica de la publicació: Emma Soy Massoni

Supported by a grant from lceland, Liechtenstein and Norway through the EEA Financial Mechanism. Operated by Universidad Complutense de Madrid

© del text: els autors© de l’edició: DOCUMENTA UNIVERSITARIA ® www.documentauniversitaria.com

Aquest llibre s’ha imprès amb paper que procedeix de boscos gestionats de manera sostenible (FSC, PEFC)

Dipòsit legal: GI-2.068-2015ISBN: 978-84-9984-308-7

Imprès a CatalunyaGirona, desembre de 2015

Dades CIP proporcionades per la biblioteca de la UdG

CIP 504.03 ECO

Ecosystem services : concepts, methodologies and instruments for research and applied use / Sergi Nuss-Girona, Mita Castañer (eds.). – Girona : Documenta Universitaria, 2015. -- p. ; cm. – (Quaderns de medi ambient ; 6) ISBN 978-84-9984-308-7

I. Nuss Girona, Sergi, ed. II. Castañer i Vivas, Margarida, ed.1. Serveis dels ecosistemes 2. Ecosistemes – Gestió 3. Zones Verdes 4. Economia ambiental

CIP 504.03 ECO

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contents

Presentation .................................................................................................................... 9Josep Vila i Subirós

Ecosystem Services: concepts, methodologies and instruments for research and applied use ....................................................................................... 11Sergi Nuss-Girona, Mita Castañer Vivas

State of the Art

Ecosystemic services: a general introduction ..........................................................23Jaume Terradas

Ecosystem services: application and conflicts ......................................................... 33Beatriz Rodríguez-Labajos

Assessment and application of ecosystem services

Socio-economic valuation of ecosystem services in Spain .................................... 51Marina García-Llorente, Cristina Quintas-Soriano, Pedro Zorrilla-Miras, María Loureiro, Carlos Montes, Javier Benayas, Fernando Santos-Martín

Monetary valuation of urban ecosystem services– operationalization or tragedy of well-intentioned valuation? An illustrated example ....................... 65David N. Barton

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A multi-scale assessment of regulating ecosystem services in Barcelona ........... 87Francesc Baró

Managing the side effects of bioenergy: a review of the impacts of biomass feedstock production on ecosystem services ..................................... 101Elena Gissi, Mattias Gaglio, Matelda Reho

Socio-cultural values of urban ecosystem services............................................... 113Johannes Langemeyer

Indicators of recreational value of urban green spaces ........................................127Emma Soy-Massoni, Graciela Rusch

Payment for ecosystem services: concept and examples ...................................... 137Rafael Sardá

Tools for regulating environmental services in Catalonia .................................. 151Joan Pons Solé

Testimonials and reflections from experiences

Price versus value of ecosystem services in the southern Alps ........................... 163Ann Brower

Enabling stakeholders to apply the ecosystem services concept in practice ........... 171Josep Lascurain

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Presentation

Josep Vila i SubirósDirector 2009-2015, Environment Institute, UdG1

This book represents the 6th volume of the collection “Quaderns de Medi Ambient” (Environment Notebooks) impulsed by UdG’s Environment Institute and Documenta Universitària publishers in 2009. A sixth volume devoted to the fifteenth edition of the International Summer School on the Environment (ISSE) focusing on ecosystem services, with a twofold theoretical-conceptual and practical-applied approach. A topic once again tied to the central questions of contemporary scientific discussions. With this framework, the goal of publishing the book returns to the idea of fostering dissemination and transfer of the knowledge delivered by the 2015 ISSE guest speakers, given their proven relevance and experience. Ultimately, the mission of the book is communicating results to the highest possible number of people interested in the topic of ecosystem services, from the public, private, university or civil society sectors, in order to maximize benefits from the efforts allocated by institutions supporting the celebration of this international seminar on the environment.

All together with a programme willing to open, one more year, the summer school to a non-strictly university audience, offering topics and discussion attractive for public officers and decision-makers, yet also for environmental consultancy professionals and environmentalist organization members. In practice, participants included people from all sectors, including a few unemployed, for whom we expect the ISSE to be a retraining opportunity, improving their professional record either for job seeking, or for new entrepreneurial activities.

We hope this modest contribution will help in addressing the importance of ecosystem services and value the benefits nature supplies us in a much more weighted and fair way. Services that should become a core factor in much more

1 www.udg.edu/ima

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decided and committed policies for the conservation of our natural heritage. The many and valuable environmental services associated to nature leave those who have marginalized resources for nature research and conservation without arguments, yet their misunderstanding and mismanagement has affected maintenance and improval of essential services that nature freely delivers to the society.

On the other hand it is to recall that all documentation of the ISSE is available at www.udg.edu/ISSE2015. A website including a link to video recordings of the presentations, for which the speakers’ necessary authorisations apply; also accessible from UdG’s Library website:

http://diobma.udg.edu/handle/10256.1/3954

I wouldn’t want to finish this brief presentation without sincerely thanking all the registered attendants and all the people that made it possible. For their fundamental task as coordinators of the XV ISSE, a special acknowledgement must go to PhD Mita Castañer and PhD Sergi Nuss-Girona, and to Emma Soy Massoni for her support in the release of this book. Likewise, thank you very much to UdG’s Library Services for facilitating the recording of the course, and to the Environment Institute Secretariat, from where an enormous managament work allowed to have everything timely in place. Last but no least, our sincere thanking to the NILS Science and Sustainability program, for financing this book and, together with UdG’s Social Council and UdG’s Geography and Territorial Thinking Chair, the XV ISSE. Without all of you, none of this would have been possible. Thank you very much.

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Ecosystem Services: concepts, methodologies

and instruments for research and applied use

Sergi Nuss-Girona*1 Mita Castañer Vivas

Environment Institute and Geography Department, University of Girona

Abstract

The ecosystem services (ES) framework opens the door to common ground between ecology science and economics in facing the challenge of planetary sustainability. Application of the available methodologies and instruments reveals both the high potential and the controversies related to this new socio-ecological management approach. Feeding on the research, reviews and testimonies presented at the XV International Summer School on Environment (ISSE) of the University of Girona (2015), and collected in form of a conference proceedings book, this paper carries out a brief introduction of the main concepts related the ES approach and the overarching international frameworks, policies and institutions. In order to facilitate an outlook of the book, a transversal presentation of the chapters follows, allowing readers to overview the variety of ES methods and topics that ISSE speakers work on.

Introduction

Payment for ecosystem services (PES) is not a novel practice, even if not under these instruments’ contemporary standardized schemes. As early as 1937 the USA approved the Federal Aid in Wildlife Restoration Act (or Pittman–Robertson Act [P-R Act]), which dedicates an 11% excise tax (earlier it was 10%) on hunting firearms and ammunitions exclussively to finance state wildlife protection areas and related management activities. Created after certain species at the border of extinction, it was understood that pressure over the resource need be managed and

*1 corresponding author [email protected]

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financed to continue enjoying outdoor recreation on the long-term (U.S. Fish and Wildlife Service, 2015). Outcomes of this ongoing regulation are twofold; in form of a come-back of notable species and more than USD 15 billion in Federal funds for the establishment, restoration and protection of wildlife habitats; research and land acquisition inclusive (U.S. Fish and Wildlife Service, 2015).

Nevertheless, PES is neither the only way, nor the only aim of applying the ecosystem services (ES) framework, in order to bridge economic development and ecological sustainability. Feeding on the research, reviews and testimonies presented at the XV International Summer School on Environment (ISSE) of the University of Girona (2015), and collected in form of a conference proceedings book, this paper carries out a brief introduction of the main concepts related the ES approach, and the overarching international frameworks, policies and institutions. In order to facilitate an outlook of the book, a transversal presentation of the chapters follows, allowing readers to overview the variety of ES methods and topics that ISSE speakers work on.

Common ground between ecology science and economics is increasingly crucial for planetary socio-ecological sustainability. Good example of this are the complex negotiations behind the 2015 COP21 climate change Agreement in Paris. Emerging economies and poor countries defended their right to support accelerated economic growth on carbon intensive technologies, in order to catch-up with industrialized states in terms of infrastructures, services and income for their societies. Yet, admitting the need to prevent global warming exceeding 2ºC by 2050 (and possibly 1.5ºC by 2100), all signing nations assumed shared responsibility in curving down global greenhouse gas emissions no later than 2030, provided funds for low carbon transitions of developing countries are stipulated and supplied by the wealthy ones (UNFCC, COP21, 2015). Hence, climate stability is an economic issue.

The ecosystem services approach

The essential idea of the ES approach is to understand how ecosystems contribute to humankind’s needs through products and services (De Groot, 1992), thereby to foster managment such that ecosystems-dependent wellbeing is granted on the long-term and included in development policies and economic accounts.Currently, ES is the dominant framework for addressing the ecology-economics challenge (Rodríguez-Labajos, in this book); with the Millenium Ecosystem Assessment ([MA] 2005) and the ‘The Economics of Ecosystems and Biodiversity’ (TEEB) project (Kumar and TEEB, 2010) as main referential documents. The MA reported evidence on the decline of ecosystem services under a scientifically agreed categorization and accounting system, evolved over a period of 40 years (Rodríguez-Labajos, in this book). TEEB1 delivered a review of monetary valuation methods for ecosystem products and services.

1 begun “at the meeting of the G-8 in Potsdam in 2007, with support from the German government, and the European Commission” (Labajos, in this book).

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Table 1. Ecosystem Services: definition, categories and examples

Ecosystem Services Definition: ‘the capacity of natural processes and components to provide goods and services that satisfy human needs, directly or indirectly’ (De Groot, 1992); “the benefits people obtain from ecosystems” (MA, 2005)

Ecosystem Service Categories

Regulating: “the capacity of natural and semi-natural ecosystems to regulate essential ecological processes and life support systems through bio-geochemical cycles and other biospheric processes” (De Groot et al., 2002)Supporting: “necessary for the production of all other ecosystem services” (MA, 2006); & “natural ecosystems provide refuge and reproduction habitat to wild plants and animals and thereby contribute to the (in situ) conservation of biological and genetic diversity and evolutionary processes” (De Groot et al., 2002)Provisioning: “Products obtained from ecosystems” (MA, 2005); “ecosystem goods for human consumption, ranging from food and raw materials to energy resources and genetic material” (De Groot et al., 2002).Cultural: “Nonmaterial benefits people obtain from ecosystems through spiritual enrichment, cognitive development, reflection, recreation, and aesthetic experiences” (MA, 2005)

Examples of goods and services provided by ecosystems (InforMEA, 2015)

• Food, fuel and fiber • Shelter and building materials • Purification of air and water • Detoxification and decomposition of wastes • Stabilization and moderation of the Earth’s climate • Moderation of f loods, droughts, temperature extremes and the forces of

wind • Generation and renewal of soil fertility, including nutrient cycling • Pollination of plants, including many crops • Control of pests and diseases • Maintenance of genetic resources • Cultural and aesthetic benefits • Ability to adapt to change

Mounting literature on how ecosystems contribute to the subsistence, resilience, economic progress, well-being, etc., of social systems (Martin-Ortega et al., 2015) run in parallel to growing interest by the international community on using the ES approach as a driver of economic greening. In this sense, flow of ES in economic terms is the “star” mechanism for raising, in front of economic agents (decision-makers, regulators, corporations, etc.), the discussion about nature’s free supply of environmental resources for economic activity and the costs associated to their degradation and missuse. According to UNEP’s InforMEA platform (2015), at least 40% of the global economy and 80% of the needs of people in developing countries depend on ecosystem products and services. So, even from a merely antropocentric and business focus ES conservation is convenient.

From a market-based approach green taxing and PES schemes would be among the preferred ways to internalize flows of ES in balance sheets, as these setups institutionalize direct funding for ES conservation (Sardà; Pons; in this book). However, risks from an ill-adoption of valuation methods mustn’t be shaded by

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apparent benefits (Kallis et al., 2013). Monetary tools should not be seen as surrogates for regulation and management, as the latter are often the only reliable ways to reach the desired quality in the ES to maintain. Many authors (Terradas; Rodríguez-Labajos; Barton; in this book) raise the risks of a simplistic and utilitarian look upon nature and ES, by which a certain amount of money will, through man-made technologies and restoration measures, offset depletion of, or compensate damage over essential and hyper-complex life-supporting ecological functions, such as the bio-geo-chemical processes involved in natural purification of air and water.

For instance, given groundwater is the major source of drinking water (FAO, 2013) a top priority is to ensure its quality on the long-term. Thus, effluent discharge on the natural environment relies on technological treatments, yet also on natural filtering and purification supplied by vegetation, soil and underground materials (hence, ecosystem services). Imbalance in this process often occurs, such as in nitrate groundwater pollution, deriving from the intensive application of fertilizers (mainly pig slurry) to crops and subsequent leaching to groundwater (Menció et al., 2016). Nowadays, this is a major threat for water resources, and it is the most frequent cause of poor groundwater chemical status in Catalonia (Agència Catalana de l’Aigua, 2009). Proper definition of an ecosystem service system —flow and limits of the ES; ownership; beneficiaries...— (Sardà, in this book) related to the fertilizers-groundwater pollution problem, might lead to develop a PES scheme requiring farmers to pay a rate for financing slurry/manure treatment to avoid excessive fertilization; aquifer decontamination technologies; and/or additional potabilization techniques for drinking waters in polluted areas. Even so, unless this kind of solutions are followed by a reduction of the pressures on the environment and water resources, such as the intensity of agricultural and livestock activities (i.e. limit the heads of livestock that a certain territory can support) and the amount of fertilizer applied on crops, an improvement of groundwater quality is not foreseen (Boy-Roura, 2013).

Operationalizing ecosystem services

In a context of intense academic debate around theory and practice of the ES approach, the international community is setting up related institutions and frameworks, under the notion that broadspread standardized study and evaluation of ES will lead to the proper development and adjusment of a variety of operational tools (managerial; regulatory; for decision-making; fiscal; market-based; to address awareness and governance...).

The first structural factor for the latter is a common methodological framework. On one hand, in 2003 the UN Statistical Commission (UNStats) initiatied the System of Environmental-Economic Accounting (SEEA); a “system for organizing statistical data for the derivation of coherent indicators and descriptive statistics to

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monitor the interactions between the economy and the environment and the state of the environment to better inform decision-making” (UNStats, 2015).

Upon growth of ecosystem valuation studies “the need to understand better how estimates at different places compare to each other has increased” (Haines-Young and Potschin , 2011). As well as the interest on generating an integrated framework, in which the complexity of ES (including products and services from subsystems and their interrelations) and economic activities (with their several nomenclatures, also in products and services) are commonly considered and addressed. There “needs to be consistency between countries in defining and naming elements of the accounts” (Haines-Young and Potschin , 2011). Hence, between 2009 and 2011 the EEA commissioned the development of the Common International Classification of Ecosystem Services (CICES). As of today (2015), CICES has reached version 4.3 and is widely accepted as the system to define, classify and therefore study ecosystem services.

International implementation and updating of SEEA and CICES are among the aims of the Intergovernmental Platform on Biodiversity and Ecosystem Services (IPBES), an independent body open to all UN members created in 2012 under sponsorship of UNEP, UNESCO, FAO and UNDP. IPBES’ missions include assessing, in dialogue with all stakeholders, ES at global and regional level; strengthening the capacity and knowledge foundations of the science-policy interface on biodiversity and ecosystem services, with regard to thematic and methodological issues; and conducting communication and reporting activities.

Currently, thanks to the SEEA, CICES and IPBES, the ES community enjoys a general set of institutions (organisational and instrumental) to work with.

Indeed, despite still a young field, ES has fastly gained relevance in sustainability policy, such as in UN’s 2015-2030 sustainable development goals (SDGs). Objectives 14 and 15 advocate for sustainable use of sea and terrestrial ecosystems and natural resources (UN, 2015). Specific targets request to “integrate ecosystem and biodiversity values into national and local planning, development processes, poverty reduction strategies and accounts” and “mobilize and significantly increase financial resources” for biodiversity, ecosystems and forests (this last in developing countries) by 2020 (UN, 2015). Growth and ES are also addressed in the SDGs. While sustained per capita GDP growth is expected at global scale, including a yearly +7% rate in the least developed countries, decoupling growth from environmental degradation is also targeted (UN, 2015). To this end, humankind must foster a low carbon circular economy, sustainable production and consumption of goods and services and resilient human settlements and infrastructures (UN, 2015). Alas, UN’s track record after the assessment of the Rio’92 objectives prepared for Rio+20 (UNEP, GEO 5, 2012; Nature, 2012) offers little confidence for the 2030 targets, but altogether, the SDGs suggest that ES analysis, accounting and regulation will further grow among the core issues of Multilateral Environmental Agreements.

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Following the work of MA and TEEB, and the 2011-2020 Nagoya Strategic Plan for Biodiversity (derived from the UN Convention on Biodiversity), the EU established its own specific framework for the deployment of the ES approach at continental scale, within the EU 2020 Biodiversity Strategy (also from 2011). Under Action 5 of the European strategy the Mapping and Assessment of Ecosystem Services initiative was launched, in response to calling “Member States to map and assess the state of ecosystems and their services in their national territory” (BISE, 2015). The EU expects to obtain a “critical evaluation of the best available information” on ecosystem services, in oder to guide decisions on biodiversity and ecosystems related policies (on water, climate, agriculture, forest, and regional planning). Against the challenges of facilitation, consistency and comparability, EU engagement has lead to producing the MAES Analytical Framework (Fig. 1).

Fig. 1. EU Mapping and Assessment of Ecosystem Services Analytical Framework

Source: Biodiversity Information Service for Europe (BISE), 2015.

The MAES Analytical Framework provides 12 ecosystem categories (drawn from the higher levels of the EUNIS Habitat Classification). A first version of a EU comprehensive 1 Ha spatially explicit map of ecosystems is available, based on Corine Land Cover (CLC) datasets, covering terrestrial and freshwater ecosystems, and 1 in 4 marine ecosystems (other crosswalk resources apply for the remaining marine ecosystems). Assessment of ES is based on the aforementioned CICES system, and a knowledge-exchange platform (the Ecosystem Services Partnership Visualisation tool; ESP-VT) “allows users to share information on ecosystem services maps, data and mapping methods” (BISE, 2015). Strong importance is allocated to mapping, as

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it is seen very useful in a twofold manner. First, for supporting “prioritization and problem identification, especially in relation to synergies and trade-offs among different ecosystem services, and between ecosystem services and biodiversity” (BISE, 2015). Additionally, “as a communication tool to initiate discussions with stakeholders” (BISE, 2015). Indeed, in a context of multiple interests over ecosystem services —such as in Gissi’s (in this book) review of the impacts of biomass feedstock production on ecosystem services; Baró’s (in this book) multi-scale assessment of regulating ecosystem services in Barcelona; or Brower’s (in this book) reflection on price versus value of ecosystem services in the Southern Alps— mapping emerges as a potent tool for assertive analysis and results interpretation.

Three key projects are currenlty implementing the MAES framework at different scales, in a variety of ecosystems and socio-ecological settings, with the ultimate goal of translating the ES approach “into operational frameworks that provide tested, practical and tailored solutions for integrating ES into land, water and urban management and decision-making” (OpenNESS Project, 2015). Two of them are represented in this book (and in UdG’s XV ISSE); OpenNESS and OPERAs, with a total of 5 papers (Baró; Bartón; Langemeyer; Lascurain; and Soy), dealing with urban and coastal ecosystem services and showcasing state-of-the-art European (applied) research in the field.

Presentation of book chapters and conclusion

The XV ISSE and this outcoming proceedings book is an attempt to display the variety of methods and instruments currently developing for the assessment, valuation and operationalitzation of ES.

Book chapters include an introductory reflection by Emeritus Professor in Ecology Jaume Terradas, in which the relevant concepts behind the ES framework are presented and challenged against planetary and local ecological sustainability. The strong antropocentric view behind greening the economy through ES tools is denounced, given the risks of missmanaging the biophysical matrix sustaining us may continue to develop. Rodríguez-Labajos, in chapter 2, travels along the evoution of ES science, first by presenting potential opportunities (in ecosystem management, in multi-stakeholder dialogue, etc.), to later criticize undesirable outcomes, such as cost-effective biodiversity protection and “green washing” practices, which can unfold from an ill-adoption of the ES approach.

A second group of papers deal with methods and instruments to assess ES of uneven nature, at different scales and in a variety of environments. García-Llorente et al. present an extensive socio-economic valuation of ecosystem services in Spain, with “a meta-analysis of the studies previously done in Spain, through the use of market-based instruments applied to the case of food from agriculture, and a social preference assessment combined with a choice experiment exercise”.

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Baró focuses on regulating ES (pollution control and climate regulation) from green infrastructure in Barcelona and its metropolitan regions, considering both the supply and demand of these services along the urban-rural gradient. In contrast to Baró, Gissi reviews the impacts and trade-offs between the provision of Biomass based energy sources (BBES) and other ecosystem services. Barton applies different econometric methods for trees, parks, and metropolitan forests in Oslo, to show and discuss methodological challenges (inaccuracy; double-counting; appropriate scale...) and the “tragedy of well-intentioned valuation” (Kallis et al., 2013), unless several ecological economic and political economic criteria are met. Langemeyer, and Soy-Massoni and Rusch study socio-cultural services of urban ecosystems, including agriculture patches and parks. Both authors underpinn how immaterial ES (recreational, experiential, spiritual, inspirational, etc.) represent the most valuable services for the citizenship; hence the direct link between ES and wellbeing in cities, counter-balancing, as shown by Baró, the limited contribution of green infrastructure for certain regulation services.So far, cited papers deal with ES assessment, for regulation, provision and socio-cultural services. Furthermore, the latter manuscripts address several biophysical accounting, monetary and non-monetary valuation, and mapping methods and techniques.The next two chapters, by Sardà and Pons, enter the field of operationalization through market, fiscal, regulatory and agreement based instruments and procedures. Building from literature review and research in integrated coastal management, Sardà unfolds the theoretical framework and successful PES cases. Pons explores the current legal and policy setting in Catalonia and Spain, to offer potential ways to adapt and deploy ES based instruments, including specific comments on land stewardship.Last but not least, the closing section of the book is dedicated to short reflective chronicles related to ES management cases. Brower presents the unexpected counter-productive effects of market arrangements in high ecological value land in New Zealand. Lascurain, in turn, addresses the conflicts and lessons learnt from regenerating natural ecosystem in highly urbanized areas, where social requirements from ecosystems clash with conservation goals. Thus, the critical role of governance and stakeholders’ engagement and enabling.The ES approach is an exciting new sustainable development field aiming to bridge ecology and economics. There is undoubtedly a wide fringe for the encounter of these two fundamental development disciplines. Yet, it still seems as if conservationists are trying to seduce and convince economists and decision-makers about the essential life-supporting (manufacturing processes included) role of nature, offering instruments to internalize it in business as usual. It would be a pitty to reduce such biospheric functions to taxes and/or voluntary payment scheemes, because the ultimate goal of this crosswalk language should be changing mentalities and raising awareness to the point of installing new socio-ecological values; leading society to economic models in which ES conservation is a premise and therefore an intrinsic component of profits.

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Bibliographical references

Agència Catalana de l’Aigua (2009). Pla de gestió de l’aigua de Catalunya. Available at: http://aca-web.gencat.cat/aca/documents/ca/planificacio/pla_gestio_Catalunya/pla_de_gestio_COMPLET.pdf [Accessed: 15/11/2015]

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Boy-Roura, M., 2013. Nitrate groundwater pollution and aquifer vulnerability: the case of the Osona region. PhD dissertation. Universitat de Girona.

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De Groot, R.S., Wilson, M.A., Boumans, R.M.J. (20o2). A typology for the classification, description and valuation of ecosystem functions, goods and services. Ecological Economics 41(3), 393–408

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Haines-Young, R., and Potschin, M. (2011). Common International Classification of Ecosystem Services (CICES): 2011 Update. Commissioned by the European Environment Agency (EEA). Available at: http://unstats.un.org/unsd/envaccounting/seeaLES/egm/Issue8a.pdf [Accessed: 15/11/2015]

Kallis, G., Gómez-Baggethun, E. and Zografos, C. (2013). To value or not to value? That is not the question. Ecol. Econ. 94, 97–105. doi:10.1016/j.ecolecon.2013.07.002

Kumar, P., TEEB (2010). The Economics of Ecosystems and Biodiversity: ecological and economic foundations. Earthscan, London and Washington D.C. doi:10.1017/s1355770x11000088

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Martin-Ortega, J., Jorda-Capdevila, D., Glenk, K. and Holstead, K. (2014). Defining ecosystem services-based approaches, in: Martin-Ortega, J., Ferrier, R., Gordon, I., Kahn, S. (Eds.), How Can Ecosystem Services-Based Approaches Help Addressing Global Water Challenges? Cambridge University Press, Cambridge, p. (in press).

Menció,A., Mas-Pla. J., Otero, N., Regàs, O., Boy-Roura, M., Puig, R., Bach, J., Domènech, C., Zamorano, M., Brusi, D., Folch, A. (2016). Nitrate pollution of groundwater; all right…, but nothing else? Science of the Total Environment 539: 241-251.

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OpenNESS Project (2015). Avalilable at: http://www.openness-project.eu/ [Accessed: 15/11/2015

UNEP, GEO5 (2012). Perspectivas del Medio Ambiente Mundial 5. Medio ambiente para el futuro que queremos. Novo Art, Panamá.

UN (2015). 2030 Sustainable Development Goals. Available at: https://sustainabledevelopment.un.org/ [Accessed: 15/11/2015]

UNFCC, COP21 (2015). Adoption of the Paris Agreement. Available at: http://unfccc.int/resource/docs/2015/cop21/eng/l09r01.pdf [Accessed: 12/12/2012]

UNStats (2003). System of Environmental-Economic Accounting (SEEA). Available at: http://unstats.un.org/unsd/envaccounting/seea.asp [Accessed: 15/11/2015]

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State of the Art

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23

Ecosystemic services: a general introduction

Jaume TerradasCentre for Ecology Research and Forestry Applications (CREAF)

Universitat Autònoma de Barcelona (UAB), Spain1

Abstract

This paper will consider some aspects of the concepts of ecosystem services and green infrastructure, and analyze some difficulties involved in their use that sometimes questioned their utility. We relate both concepts to the physical and social vulnerabilities of socio-economic systems and their complexity. We address the very important issue of sustainability, with special emphasis on that management can be approached by taking inspiration on nature’s functions and solutions. We will also insist in the importance of increasing ecological knowledge and ecological management. We think that, in many cases, the best management, when good ecosystems knowledge is available, can be developed through local populations empowerment. To reach this end, a special role for environmental educators will be essential.

Keywords

Ecosystem services, Green infrastructure, Conservation management, Communities empowerment, Environmental education

1 [email protected]

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24 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Introduction

From a human-biased point of view, ecosystems can be considered as something providing us with a large array of goods and services. These are from different types: products, environmental benefits and social benefits. Products include clean drinkable water, foods for humans or domestic animals (including meat, fish, herbs and grass, fruits, fungi, etc.), medicals, building materials (as wood, mud, clay, stones, cork, etc.), and many other materials with different uses. Many of these products, if exploited in a reasonable way, can be used indefinitely as renewable resources.

Environmental benefits from ecosystems are usually free: no effort or cost is required from humans to get them. Notorious examples are the biodiversity, the regulation of the atmospheric composition and climate, the regulation of biogeochemical cycling, the protection of soil from erosion, the hydrological regulation, the pollutants dispersion or dilution, the storage of carbon… More relevant social benefits from ecosystems include recreational, educational, cultural and aesthetic uses, research opportunities, psychological and health benefits, etc. These social benefits have, however, a number of economic consequences like those related to health, tourism, leisure and sport activities in the field, coast, sea, lakes or rivers, local gastronomy, etc.

Ecosystem services are called sometimes landscape services or green services. But in any case some aspects of the ecosystem services (ES) concept are controverted (Basnou et al., 2015). First of all, the human-centered view involved in the term “services” is dangerously shortsighted: ecosystem services (ES) are not a consequence of God providence to maintain humans in the world, but a result of ecosystem functions. These functions can effectively provide many services to humans, but other of their consequences can so well be considered as disservices (EDS). It can be pleasant for people living near natural or semi-natural areas to see trees, f lowers or a lake, but unpleasant, and even dangerous, to have mosquitoes and other hematophagous insects in the vicinity. A high biodiversity warrants healthy ecosystem function, even it seems to be related to health benefits for human population (Keesing and Ostfeld, 2015), but humans prefer to avoid big predators or poisonous snakes, and also they usually sacrifice more and more biodiversity in order to enhance production of food from crops or grasslands. Some human-managed ecosystems, like “dehesas” and other grasslands or some forests, still maintain a quite high biodiversity and are resilient in front of environmental f luctuations; others have a very low biodiversity and, usually, are more vulnerable to f luctuations, as greenhouse crops. It seems risky to evaluate Nature on the basis of the benefits that society obtains from it, because many functions of Nature do not seem to produce direct benefits to humans and they are, however, very relevant in life-on-Earth support. I will return to this point later.

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25Ecosystemic services: a general introduction

Cities are also ecosystems (Boyden et al., 1981, Barracó et al., 1999). Urban ecosystems depend on a large array of inputs of energy, foods and materials, in order to sustain their functions, and have to eliminate gaseous, liquid and solid wastes to avoid the accumulation of noxious substances inside them: this is the same that occurs in an ant colony or in the body of an animal, because urban ecosystems are essentially heterotrophic, they don’t produce the food they need from light, air, water and soil nutrients, like plants, and they have to take it from outside (farms, rangelands, fisheries, food industries, etc.). The role of primary producers in urban metabolism is modest (Baró et al., 2014). Budgets of energy and cycles of elements and compounds can be studied to understand ecosystem metabolism for any kind of ecosystem, including rain forests, small ponds or towns.

Ecosystems are systems that work, with more or less efficiency, and they can be disturbed by natural causes (climatological changes, extraordinary meteorological events, floods, wildfires, earthquakes, tsunamis, volcanic eruptions, etc.) and by multiple anthropogenic activities. If ecosystems functions are modified by any cause, their capacity to offer “services” to humans can be changed, even destroyed. This means that we cannot manage ecosystems on the basis of the services they offer us, we must do this on the basis of their ecological functions. This is a quite simple idea, but frequently the inverse option is applied in management, with awful results.

A maximizing immediate benefits approach is particularly frequent when international corporations exploit local products following a nomadic strategy or when Garret Hardin’s “commons tragedy” (Hardin, 1968) is at work, as occurs at a global scale with fishing. In those cases, the exploitation is frequently much over the capacity of the ecosystems to renew their production and, then, resources decrease or even vanish forever. Many examples of environmental mismanagement by ancient cultures have produced the collapse of ecosystem functions and, as a result, that of ecosystem services (for a description of cases see Diamond 2005). Easter Island is a usually mentioned example, but many other areas where relevant civilizations existed are now unproductive, scarcely populated regions. Plato already adverted about the catastrophic effects of erosion and many landscapes that once had fertile soils lost them by this process in large areas of the Mediterranean Europe, the Near and Middle East, Northern Africa, etc.

Vulnerability and resilience

Function collapse in ecosystems has as an obvious consequence; a great decrease of the environmental capacity to support human life. In any case, the vulnerability of ecosystems is highly variable. Vulnerability is an important concept that can be applied to entire ecosystems or to subsystems. The exposition to a disturbance is a component of vulnerability: more exposed systems are more vulnerable than less

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26 Ecosystem Services: concepts, methodologies and in truments for research and applied use

exposed. Delta planes, and human populations in them, are more vulnerable to an increase of sea level than rocky coasts.

Another component of vulnerability is the sensitivity of a system to the disturbance. A highly sensitive system is very vulnerable. A sensitive, exposed system has more probabilities to collapse as a consequence of a disturbance. In human systems, some (old persons, children, people depending on scarce resources, underdeveloped countries) are more sensitive, and then vulnerable, to most risks (natural disasters, social conflicts as war, epidemics, etc.), even if their exposition is similar to that of the rest of the population. Then, exposition combines with sensitivity to determine vulnerability.

The concept of vulnerability has to be complemented with that of resilience. After a catastrophic event, some systems can relatively fast rebuild, others are less resilient and can never recover or require much time for it. Vulnerability and resilience can be applied to many types of systems, but are useful concepts when we consider the effect of environmental changes (natural or human-caused) on ecosystem functions.

What is an ecosystem?

Ecosystem is a convenient term, but not a precise one. A small pond can be studied as an ecosystem, but also an entire ocean can be considered as a giant ecosystem. In some sense, the planet Earth is an ecosystem. Limits are decided by the study aims, because no ecosystem is completely isolated from neighbors. So, our reasoning about ecosystem functions can be applied to the great processes at a global scale, like carbon, nitrogen or phosphorus cycling, climate warming, biodiversity loss, the proliferation of biological invasions and so on. In the same way, we can apply it to a pond or, even, an aquarium.

When considering the Earth global system, a number of boundaries to not trespass have been proposed for some processes (Rockström et al., 2009), to avoid their collapse or the activation of uncontrolled positive feed-backs that will accelerate an undesired trend (i.e., methane emissions, due to permafrost fusion or the liberation of marine clathrates that add powerful greenhouse gases to the atmosphere, accelerating global warming). Unluckily, it seems quite plausible that some of these thresholds have already been crossed during the last decades.

A main lesson that humans have to learn is that ecosystems, of any dimensions, are always complex systems, meaning that their responses to an environmental change are usually not lineal and cannot be predicted with a great level of security. When a relevant variable passes some threshold (that usually is not known with great precision) the entire system can move to a different “attractor”: ecosystem

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27Ecosystemic services: a general introduction

functional processes, and their ES and EDS to humans, can then change in a sudden and perhaps disrupting way, as explained by catastrophe or chaos theories. This is the reason to recommend prudence in human actions potentially affecting some basic processes that support life and to claim for urgent increases of ecological knowledge and of management experience in order to take decisions aimed to enhance ES or benefits obtained from nature.

What are ecosystem services?

A number of definitions have been given for ES and green infrastructure (GI, we include here also water ecosystems that sometimes are considered as blue infrastructure). This variety of definitions creates confusion. For some authors, ES include ecosystem conditions, processes and life-support functions (see Daily, 1997), for others just the final services, that are always a product of natural and cultural elements (Carlisle et al., 2009), whereas they do not include intermediate services (processes and functions) in an economic analysis, because this would mean a double-accounting (intermediate necessary products and final services). Some authors consider that services are, more simply, benefits that people obtain from ecosystems (MEA 2005, Chan et al., 2012). Others (Boyd and Banzaf, 2007) consider that benefits are the product of f lows of ES, but not ES, and are somewhere in between ecosystems and human wellbeing and we can put economic values on them (Costanza et al., 1998; Chan et al., 2012, Reyers et al., 2013; Fisher et al., 2008)

The conceptual differences between authors when using the ES concept are a serious problem that adds to the anthropocentric bias. The very large amount of literature on the topic that has been produced in the last years includes some brave attempts to furnish a theoretical basis (Boyd and Banzhaf, 2007, Bastian et al., 2012; Logsdon and Chaubey, 2013) but a clear consensus has not been reached. Whereas some ecological economists have tried to evaluate ES in monetary values, many ecologists are unsatisfied with the real utility of these efforts, because methods of giving values can be sometimes easily objected, and are difficult to apply to processes that are simply essential to life-supporting systems; on the other side, most economists continue to ignore these quantifications. When environmental aspects are included in a project, most times an ecological assessment is required to guide some basic decisions, but the use of a cost-benefit analysis in monetary terms is rare or reserved to later phases of the project development, after essential decisions have been taken about what has to be preserved or not. So, after two decades of attempts to do monetary evaluations of ES, their use is still quite uncommon for developers, planners and decision-takers. It is obvious that some consequences from ecosystem functions can interact with social factors affecting well-being or vulnerability of human societies, or some part of them, and that we

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need to evaluate this. But, in many cases, we are not able to do these evaluations in an appropriate way because we are considering interactions between two systems (nature and society) that are very complex: if we use a monetary basis, we will leave aside a number of values that are too high to be accounted for (as clean air or drinkable water), or factors whose incidence is simply unknown. Other difficulties are linked to the fact that, even if monetary values can be calculated, for instance for the wood of a forest, values and benefits are different at different scales. From the perspective of global wood production (that of a multinational wood industry) a local forest can be irrelevant but, from a local point of view, the same forest can have a high value as a long-term capital resource.

Then, whereas for a modest number of ecological economists monetary evaluation of ES has the advantage that can be used to convince managers or planners about their benefits, for other people (it is my case) this anthropocentric approach is conceptually misleading and has been unable to reach its aims. Personally, I think that this approach must be substituted by another axed on ecosystem functions and their social connections. Some hope can be put on the European Commission decision to include in its R&I Agenda a contribution from a Nature-based Solutions Expert Group (EEA and ETC/ULS 2015). The idea consists in looking for the use and reinforcement of natural solutions to solve environmental problems, imitating the way organisms and communities face extreme environmental events. The aims would be to minimize catastrophic risks, to enhance human wellbeing and to make compatible the preservation of green areas with human settlements. Practical attempts have begun in coasts menaced by a rising sea (Temmerman et al., 2013; Giosan et al., 2014). These solutions must be efficient in energy and resources consumption and resilient in front of changes, but also adapted to local conditions. In general, this approach is based on Nature’s functions knowledge, not in the anthropocentric concept of ES, and can permit to find innovative ways of governance, new institutions, and new models for business and finances with private and public funds, involving any stake-holder and every citizen in the area.

Green infrastructure

The concept of green (or blue) infrastructure (GI) was introduced in order to make clear that systems supporting life are as essentials to social systems as artificial infrastructures (roads highways, and bridges, train ways, electric cables, sewage tubes, etc.) or human settlements are (see Rouse and Bunster-Ossa, 2013). Urban design must avoid considering natural or semi-natural areas as vacant land: they are life-supporting infrastructures. The concept of GI is not anthropocentric and it is not linked to an economic perspective. Difficulties in that case come from the complexity of the systems involved.

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29Ecosystemic services: a general introduction

First, GIs are formed by pieces of very variable size, from street trees or parterres to very large natural parks, from urban ponds or springs to oceans. They are multi-scalar systems forming networks. The idea of network has changed conservation approaches (Forman, 1995) tending to protect natural areas as islands inside a highly managed (urban settlements, crops, grasslands…) territory. Landscape ecology has emphasized the role of connections (for instance, natural corridors, like rivers) in order to maintain ecosystem functions and biodiversity. The essential change in conservation theory and practice is that GIs can only be preserved if they maintain their ecological functions, with efficient connectors between relatively undisturbed landscape patches. This means that, not only the free-access natural areas, but also areas that are private and closed, might be included in the protected network design. This supposes very serious difficulties for management.

So, GI is a less confusing concept than ES, but it includes an enormous variety of sizes, characteristics, legal and ownership situations, etc. Multi-criteria and multi-scalar approaches are required, as well as institutional and legal innovations and a strong implication of planners, administrations, experts, the totality of stakeholders and directly affected people at any level. Decisions have to be taken through agreements based in the best information available. For instance, after studies made on water management in a number of deltas around the world, sustainability can be increased and risks decreased at low costs by implementing nature-(GI) based solutions more than just gray infrastructures (Tessler et al., 2015).

People empowerment and environmental education

I have just mentioned the need of an implication of local populations. Some studies (Fraser et al., 2006) have demonstrated that well-informed local populations have been the best managers of local natural resources to preserve GI and ES, using them in a sustainable way, whereas top-down approaches have failed. The process of engaging people to select key indicators provides a valuable opportunity for community empowerment and education. Participation demands an effective feedback interaction between community inputs in selecting indicators and in planning and decision-making. This is only possible if local community, experts and policy-makers can use a common language on the process. A core role to achieve this has to be accorded to environmental education. Environmental education can find its major field of professional development in this very relevant area of bringing together top-down and bottom-up contributions in landscape management.

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30 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Bibliographical references

Baró, F., Chaparro, L., Gómez-Baggethun, E., Langemeyer, J., Nowak, D. J., and Terradas, J.. (2014). Contribution of ecosystem services to air quality and climate change mitigation policies: the case of urban forests in Barcelona, Spain. AMBIO, 43(4), 466-79. doi: 10.1007/s13280-014-0507-x.

Barracó, H., Parés, M., Pou, G., and Terradas, J. (1999). Ecologia d’una ciutat. (in Catalan, with texts in English and Spanish included). Ajuntament de Barcelona; 1999. 139 pp.

Basnou, C., Pino, J., and Terradas, J. (2015). Ecosystem services provided by green infrastructure in the urban environment. CAB Reviews, 10 (004). DOI: 10.1079/PAVSNNR201510004

Bastian, O., Haase, D., and Grunewald, K. (2012). Ecosystem properties, potentials and services—the EPPS conceptual framework and an urban application example. Ecological Indicators, 21,7–16. doi:10.1016/j.ecolind.2011.03.014

Boyd, J., and Banzhaf, S. (2007). What are ecosystem services? The need for standardized environmental accounting units. Ecological Economics, 2007,63,616–26.

Boyden, S., Millar, S., Newcombe, K., and O’Neill, B. (1981). The Ecology of a City and its People: The Case of Hong Kong. Australian University Press, Canberra.

Carlisle, S., Henderson, O., and Hanlon, P. (2009). Wellbeing”: a colateral casualty of modernity?. Social Science and Medicine, 69,1556–60.

Chan, K. M., Guerry, A. D., Balvanera, P., Klain, S., Satterfield, T., Basurto, X., ... and Woodside, U. (2012). Where are cultural and social in ecosystem services? A framework for constructive engagement. BioScience, 62(8), 744-756.

Costanza, R., d’Arge, R., De Groot, R., Farber, S., Grasso, M., Hannon, B., ... and Van Den Belt, M. (1998). The value of the world’s ecosystem services and natural capital. Ecological economics, 1(25), 3-15.

http://www.esd.ornl.gov/benefits_conference/nature_paper.pdfDaily, G. C. (1997). Nature’s Services: Societal Dependence on Natural Ecosystems.

Island Press, Washington, 392 pp.Diamond, J. (2005). Collapse: how societies choose to fall or succeed. Penguin

Group, N. York, 576 pp.Fisher, B., and Turner, R. K. (2008). Ecosystem services: classification for

valuation. Biological conservation, 141(5), 1167-1169.doi:10.1016/j.biocon.2008.02.019Forman, R. T.T. (1995). Land Mosaics. The Ecology of Landscapes and Regions.

Cambridge University Press, Cambridge, UK. ISBN: 9780521479806EEA, ETC/ULS (2015). Exploring nature-based solutions: The role of green

infrastructure in mitigating the impacts of weather–and climate change-related

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31Ecosystemic services: a general introduction

natural hazards. ESA Technical Report 12/2015, Publications Office of the European Union, Luxemburg, 61 pp.

Fraser, E. D., Dougill, A. J., Mabee, W. E., Reed, M., and McAlpine, P. (2006). Bottom up and top down: Analysis of participatory processes for sustainability indicator identification as a pathway to community empowerment and sustainable environmental management.  Journal of environmental management, 78(2), 114-127.

Giosan, L., Syvitski, J., Constantinescu, S., and Day, J. (2014). Climate change: Protect the world’s deltas. Nature 516, 31–33. doi:10.1038/516031ª.

Institute for European Environmental Policy (2011). Green Infrastructure Implementation and Efficiency, Final Report. London and Brussels, 206 pp. http://www.ieep.eu/assets/898/Green_Infrastructure_Implementation_and_Efficiency.pdf

Hardin, G. (1968). The tragedy of the commons. science, 162(3859), 1243-1248.Keesing, F., and Ostfeld, R. S. (2015). Is biodiversity good for your health?.

Science, 349(6245), 235-236.Logsdon, R. A., and Chaubey, I. (2013). A quantitative approach to evaluating

ecosystem services. Ecological Modelling, 257, 57-65.Millenium Ecosystem Assessment (2005). Ecosystems and Human Well-Being:

Biodiversity Synthesis. World Resources Institute, Washington, DC.Reyers, B., Biggs, R., Cumming, G. S., Elmqvist, T., Hejnowicz, A. P., and Polasky,

S. (2013). Getting the measure of ESs: a social-ecological approach. Frontiers in Ecology and the Environment, 11(5), 268–73. http://dx.doi.org/10.1890/120144.

Rockström, J., Steffen, W. L., Noone, K., Persson, Å., Chapin III, F. S., Lambin, E., ... and Foley, J. (2009). Planetary boundaries: exploring the safe operating space for humanity. http://www.ecologyandsociety.org/vol14/iss2/art32/

Rouse, D. C., Bunster-Ossa, I. F. (2013). Green infrastructure: a landscsape approach. American Planning Association, Planning Advisory Service, Report 571; Chicago, 159 pp.

https://www.planning.org/pas/reports/subscriber/archive/pdf/PAS_571.pdfTemmerman, S., Meire, P., Bouma, T. J., Herman, P. M. J., Ysebaert, T., and De

Vriend, H. J. (2013). Ecosystem-based coastal defence in the face of global change. Nature 504, 79–83. doi:10.1038/nature12859

Tessler, Z. D., Vörösmarty, C. J., Grossberg, M., Gladkova, I., Aizenman, H., Syvitski, J. P. M., and Foufoula-Georgiou, E. (2015). Profiling risk and sustainability in coastal deltas of the world. Science, 349(6248), 638-643.

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33

Ecosystem services: application and conflicts

Beatriz Rodríguez-LabajosInstitute of Environmental Science and Technology, Autonomous University of Barcelona, 08290 Cerdanyola del Vallès, Barcelona1

Abstract

The ecosystem service framework has become a dominant paradigm of environmental assessment, as a way to link the state of the ecosystems with different constituents of human wellbeing. The objectives of this chapter are to introduce the concept of ecosystem service and milestones in its use for both management and scientific assessment; to present some cases of instrumentation; and to outline ethic-ecological controversies derived from the ecosystem services approach in the prevalent economic model. To this end, well-established references from the literature are reviewed, connecting them with results of research undertaken by the author, in works that are cited along the text.

Keywords

ecosystem service; history; applications; conflicts

1 [email protected]

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34 Ecosystem Services: concepts, methodologies and in truments for research and applied use

1. Introduction

Those who have visited the Ifugao Rice Terraces in the Philippine Cordilleras quickly understand why this remarkable place is inscribed in the UNESCO World Heritage List. Impressive mountains and dramatic skies frame a landscape where the Ifugao indigenous people have shaped the hilly landscapes with agroecological zones that combine five key components, namely: forest (muyong or pinugo), swidden fields (habal), terraced paddies (payo), settlement districts (boble) and braided riverbeds (wangwang) (Butic and Ngidlo, 2003).

Well-known because of its ethnoecological richness (Conklin, 1967), this area has secularly been managed to ensure food provision (Acabado, 2012) despite evident environmental constraints, thanks to a culturally-mediated use of rainwater filtration systems, soil conservation and natural pest control. At the same time, an interesting landscape was shaped, showing patches of forest and mottled rice terraces of ever-changing array of colours. Nowadays, the place attracts thousands of visitors charmed by scenic aesthetics and cultural features of this region. Some could propose that aesthetic benefits provided by the area can be economically measured, e.g., estimating the travel costs incurred by the recreational users of the landscape. This outstanding place is not as unique as it would seem. Similar cultural landscapes exist, for instance, in the Guangxi and Yunnan Provinces (China), in Bali (Indonesia) and in Vietnam’s Northern Highlands.

Ecosystem functions like the processes of the water cycle or soil formation occur in nature beyond human agency although human decisions may affect them to a great extent (Rodríguez-Labajos and Martínez-Alier, 2012). In contrast, the example presented above shows that ecosystem services (ES) are socio-ecological processes, the maintenance of which is perceived as beneficial by the socio-economic system and that certainly entail human involvement. They have a focus on human interests, that is, societies attribute preferences for each ES, or for a pack of them, and have therefore normative views about their development. Different levels of human agency are then required for their existence, either cognitive, behavioural or through the application of different forms of human-made capital. While they are linked to material processes that occur in nature, their raison d’être is that they are useful to generate either goods (like food production) or services that allow further human achievements (like soil fertility), including the protection against undesired events (like most regulating services) (Rodríguez-Labajos and Martínez-Alier, 2013).

The objectives of this chapter are to introduce the concept of ecosystem services and milestones in its use for both management and scientific assessment; to present some cases of instrumentation; and to outline ethic-ecological controversies derived from the ecosystem service approach in the prevalent economic model. To this end, well-established references from the literature are reviewed, connecting them with results of research undertaken by the author, in works that are cited along the text.

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35Ecosystem services: application and conflicts

2. The brief history of the notion of ecosystem services and their assessment

Socioeconomic approaches to halt or reverse biodiversity loss are used in different ways. Rodríguez-Labajos and Martínez-Alier (2013) identify three of these approaches. Between the neoclassical tradition of economic studies analysing biodiversity under the lens of natural capital and externalities, and the prevalence of non-chrematistic cultural and livelihood values, there is an interpretation of biodiversity loss as a disruption of ecosystem functioning and ecosystem services provision. The type of assessments consistent with each approach, and the proposed conservation tools will therefore differ in each case.

As several authors review (Gómez-Baggethun et al., 2010; Mooney et al., 1997), the history of the notion ’ecosystem service’ is older than its well-known formalisation by G.C. Daily (1997). Prior to her work, the definition and categorisation of ecosystem functions (de Groot, 1992) warned about the need to defend the properties and processes of the ecosystems as eventually beneficial of human interest. A famous attempt to give money values to all environmental services from ecosystems at the global level (Costanza et al., 1997) opened the door to discussions about the use of Payment for ecosystem services (PES) as an instrument for conservation. This option has been blamed to be counterproductive, since money incentives may change the logic of conservation, and even promote further loss if the payments are not high enough to compensate for the opportunity costs. An improved understanding of the extent to which the use of economic incentives can undermine (“crowd out”) or reinforce (“crowd in”) people’s motivations to engage in biodiversity and ecosystem conservation, as in Rode et al. (2014), is therefore needed before engaging in PES.

Anyhow, Daily’s contribution, and many others’ before and since, allowed a scientific agreement on the need of a standardised account of the human dependence of ecosystems (Boyd and Banzhaf, 2007). The Millenium Ecosystem Assessment (MA) (2005) was the celebrated outcome of such agreement. The MA not only provided the evidence of declining ecosystem services but also demonstrated their links with the constituents and determinants of human wellbeing. Although humans cannot survive without such services, they are not made available through the market except in some very special cases (payment for pollination services, for instance, see Gallai et al., 2009).

Among the attempts to enhance its policy relevance, economic valuation was one preferred strategy. This inspired ‘The Economics of Ecosystems and Biodiversity’ (TEEB) project (Kumar and TEEB, 2010), that began at the meeting of the G-8 in Potsdam in 2007, with support from the German government, and the European Commission. Written between 2008 and 2011, the TEEB reports presented a synthesis of methods of valuation of ecosystem products and services, which is interesting in its wide scope (Sukhdev, 2008). In practice, the reports omitted the biophysical approaches [from ecology (resilience theory) and from thermodynamics principles], and also the methods of valuation from political sciences.

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36 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Figure 1. Evolution of ES theory and literature on the topic

“Bene�ts derived by humans from theproperties and processes of the ecosystems”

Source: Adapted from Matín-Ortega et al. , 2015: 5, with permission

The TEEB reports rather focussed on providing illustrative figures on the economic value of ES increase their visibility in policy making. Besides compiling ideas to promote ES protection, such as the ‘GDP of the poor’ o the notion of ‘net positive impact’ (explained below), the TEEB initiative developed an approach to ES assessment consisting in the following stages: a) identifying and evaluate the range of impaired ES; b) estimating and demonstrate the values of ecosystem services, with appropriate methods; and c) capturing such values looking for ‘economically informed’ policy tools (TEEB, 2010).

As ES became a consolidated field of studies, the number of related publications increased from several dozens a year by 2000 to nearly two thousand in the mid-2010s (Martin-Ortega et al., 2014). Along this dynamic discussion brought, some of the original standpoints were redefined, in relation to the definition of ES and their classification (Fisher and Kerry Turner, 2008; Fisher et al., 2009; Haines-Young and Potschin, 2013), and the distinction between those socio-ecological processes that are indeed perceived as benefits by the people, and those intermediate processes that should be rather considered in the side of the ecological functions. There is nowadays agreement in recognising the existence of a cascade linking the two ends of a ‘production chain’, between ecological structures and processes created or generated by living organisms and the benefits that people eventually derive (Haines-Young and Potschin, 2010: 115).

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37Ecosystem services: application and conflicts

3. Application of ecosystem-based assessments, the bright side

Today the use of the ecosystem service approach is widespread in environmental evaluation to inform spatial planning in terrestrial (Gret-Regamey et al., 2008) and marine areas (Lester et al., 2013). It is also used for assessing post-stress damages, such as oil spill episodes (Committee on the Effects of the Deepwater Horizon and National Research Council, 2012), mining impacts (Larondelle and Haase, 2012) or biological invasions (Pyšek and Richardson, 2010).

An example of application related to spatial planning is the work by Jordà-Capdevila et al., (2015) for the case of fifteen water f low-dependent ecosystem services (ES) in the Ter (Catalonia, Spain), a river with persisting intra–and inter-basin conf licts on water f lows. This modelling exercise allows analysing the ES response (and ensuing social reaction) to changes in water f low management, in a context of several tradeoffs and synergies driven by access to water use.

An example of assessment of post-stress damages, also in Catalonia, is related to the effects of alien species (AS) in aquatic ecosystems. While public attention often focusses on less than twenty species that generate economic costs, either damage or control costs, a fine-grained evaluation of the AS effects on ES confirms the multi-dimensionality of impacts (Rodríguez-Labajos, 2006). The higher number of impacts of 356 AS detected in Catalonia’s aquatic ecosystems takes place through direct impact on regulating services. This accounts for half the total number of impacts registered, in great part because of disrupted biocontrol services. Impact in supporting services is the next category affected (23 percent), followed by provisioning and cultural services (14 and 8 percent respectively). The species that cause costs only represent 7 percent of the total number of entries (Rodríguez-Labajos, 2014).

Besides the ample use of the ES approach for environmental evaluation, there is increasing mobilisation of the concept in decision contexts (Fisher et al., 2009), particularly governmental. A foremost example of this would be the Intergovernmental science-policy Platform on Biodiversity and Ecosystem Services (IPBES, www.ipbes.net), which work program aims to strengthen the knowledge-policy interface, same than the Intergovernmental Panel on Climate Change has done for the understanding of the human-induced climate change. Along these lines, the Common International Classification of Ecosystem Services (CICES, http://cices.eu) proposes a comprehensive framing of the concept of ecosystem services in support of the European Environment Agency, and linked with the United Nations Statistics Division initiative to revise the System of Economic and Environmental Accounts.

Regarding concrete policies, the Water Framework Directive (WFD 2000/60/EC) can be mentioned as an example. The WFD sees water not only as an economic

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38 Ecosystem Services: concepts, methodologies and in truments for research and applied use

resource but also as a basic element of ecosystems. The idea is that a better state of aquatic ecosystems will result in increased quality and quantity of available water. For this reason, the Directive has become a driver of ecological restoration, setting specific normative objectives of environmental improvement, applying cost-effective measures. This influential piece of legislation is relevant in the context of this article because it greatly increased the social visibility of the ES provided by rivers and other water bodies, moving away from the view of rivers as merely sources of money in the form of water abstraction or hydroelectricity (Rodríguez-Labajos and Martínez-Alier, 2013).

The ES framework has been praised for its ability to facilitate inter-disciplinary (Maynard et al., 2010) and inter-sectoral dialogues (Primmer and Furman, 2012). Currently, many interested individuals and organisations have created means of enhanced communication, coordination and cooperation, such as the Ecosystem Service Partnership (ESP, www.fsd.nl/esp), an international network of scientist and practitioners seeking a exchange between diverse approaches that reduces duplication of research effort. The ESP is associated with the journal ‘Ecosystem services’, which also presents itself as an interdisciplinary endeavour to improve the understanding of the dynamics, benefits and social and economic values of ecosystem services, and to create a scientific interface to policymakers in the field of ES assessment and practice, among other purposes (Braat, 2012).

The virtues of an ES-based approach involve, inter alia, enhanced awareness of socio-environmental linkages (Su et al., 2012), identification of distributional issues linked to ES loss (e.g., through the calculation of the so called GDP of the poor) (Christie et al., 2012) and practical evaluation of tradeoffs (Fisher et al., 2011; Lester et al., 2013).

The area of climate change, where monetary estimates of impacts dominate the discussion on the costs of inaction, provides an example of unveiled socio-environmental linkages through the ES approach. Looking at the literature, environmental pathways for the change in ecosystems due to climate change involve warming, frequency of extreme weather events, conditions of the water column and chemistry of sea water (including acidification, admittedly uncertain), among others. As a result, there would be biological impacts in the seasonal phenology, physiological performance of organisms, species population and distribution, productivity and community and ecosystem structure. In turn, decreased and/or spatially shifted availability of ES provision would manifest itself as losses in cropping, decreased productivity of fish stocks and valuable shell-forming invertebrates, declining forage provision, increase in the frequency of natural hazards, such as forest fires, biodeterioration of cultural assets, or decline of pollination services, among other examples compiled by Rodríguez-Labajos (2013). It is noteworthy that ecosystem change can also report increase in value of some ES, for instance due to the arrival of fish stocks to new areas, or certain cases of bioprotection of cultural assets.

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39Ecosystem services: application and conflicts

In order to illustrate the potential of the ES approach to identify distributional issues Rodríguez-Labajos and Martínez-Alier (2013) focus on one of the most innovative ideas from TEEB, the ‘GDP of the poor’. The contribution of forests and other ecosystems to the livelihoods of poor rural households is large in terms of their wellbeing, and therefore there is a significant potential for nature conservation efforts to contribute to poverty reduction. TEEB showed that ecosystem services and other nonmarketed natural goods account for 47 to 89% of the so-called ‘GDP of the poor’ (i.e., the total sources of livelihoods of rural and forest-dwelling poor households) in some large developing countries. One could argue that the GDP of the poor should not be measured in money but in kind, in terms of contributions to livelihood. However, TEEB made an attempt to translate livelihood values from ecosystem services into monetary values to emphasise their importance (ten Brink, 2011:118). The introduction of the notion of ‘the GDP of the poor’ provides an interesting link to a critique of uniform economic development. It also supports the movements of the ‘environmentalism of the poor’ (Martinez-Alier, 2014) in defence of biodiversity because this notion signals the importance of ecosystems as a resource base for livelihoods.

As an example of the evaluation of trade-offs, Rodríguez-Labajos (2013) reviews how ecosystem-based action against climate-change promotes provision of regulating ES, but also concurrent emergence of disservices, i.e. negative or unintended socio-environmental effects. An overview of trade-offs unavoidably encompasses the mitigation initiatives in the energy sector. Low-carbon generation sources, such as hydropower, wind power and nuclear power projects are being supported worldwide under the aegis of reduced CO2 emissions (a regulating service), but they also entail clashes with other types of ES provision. Leaving aside the effectiveness of biofuels as a mitigation strategy, there is evidence of their disruptive effect on food security, land tenure, labour rights and biodiversity in several parts of the world. Options like urban greening require copious amounts of water for irrigation, and often use alien plant species, becoming a precursor for the establishment of new damaging invaders.

A final example of trade-off helps us connect with topics elaborated in the next section. Environmental conflicts about water can be seen as conflict over who takes advantage and who loses access to environmental services when trade-offs appear. Rodríguez-Labajos and Martínez-Alier (2015) provide supporting examples of water-related ecological distribution conflicts around the world across the different ES categories. The complete appropriation of biophysical processes of river basins is often the foundation of large projects of economic development, as in the long-standing efforts by the different countries along the Mekong basin to build dams. However, trade-offs between ecosystem services are common, as in the case of irrigation and nature conservation or hydroelectric power production and supporting services. From there, there is an emergence of social conflicts that should be studied looking at the languages of valuation deployed and the power of those involved.

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40 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Figure 2. Effects and trade-offs between actions against Climate Change and ES provision

Source: Rodríguez-Labajos, 2013, DOI: 10.1002/wcc.247

4. Conflicts and tension to solve

All in all ES have become the foundation of a new conservation paradigm. Yet the literature recognises the need to solve critical questions for ES assessments with respect several aspects, in particular related to biophysical realism, the study of trade-offs, the consideration of off-site effects and stakeholder work (Seppelt et al., 2011).

The success of the ES approach among conservation scientists and practitioners is to great extent motivated by the desire of demonstrating the disruptive effects in ecosystem service provision from biodiversity loss. That is, the role of biodiversity in ecological processes that turn out to be beneficial to humans, admittedly complex (Elmqvist and Maltby, 2010), tends to be considered positive. This is confirmed by systematic reviews of literature that at the same time show that ES emerge from both positive and negative interactions occurring in complex systems that involve biotic and abiotic attributes (Harrison et al., 2014).

The need to better understand the complexities and many uncertainties involved in biophysical underpinning of ES, gets then even more intricate when human preferences on ecological processes need to be considered. The same biophysical process, e.g., a certain amount of water flows in the river, is perceived differently depending on the season, the location of this river and the type of water user. In this context, a new modelling approach is argued to integrate diverse value perspectives

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41Ecosystem services: application and conflicts

(engaging with stakeholders’ concerns and claims) and of environmental contour aspects (including droughts, wet years, and different options for managing flows). The differentiated perceptions can be modelled through the participatory design of service suitability curves in a way that spatial/temporal patterns and ES performance can be represented and analysed (Jordà-Capdevila et al., 2015).

Also illustrating the complexity of ES assessments, Tilliger et al. (2015) identify entangled interrelationships between agricultural landscapes and cultural ES, for the case of the Ifugao rice terraces described at the beginning of this paper. The authors discuss at a theoretical level how these relationships call into question the linearity of widely accepted assessment approaches, such as the cascade model mentioned above. This type of results underlines the need of dealing with a plurality of values hardly manageable in isolation when undertaking ES management and policy decisions.

The ES applications presented in the previous section are also limited by the issues of measurement (Boyd and Banzhaf, 2007), mapping (Burkhard et al., 2012; Maes et al., 2012) and valuation (Busch et al., 2012; Kumar and TEEB, 2010; Sijtsma et al., 2013), often intertwined (Gret-Regamey et al., 2008; Lester et al., 2013) in many of the latest conceptual developments.

Again the case of agricultural landscapes is used to illustrate the difficulties of using valuation approaches adequately. In the Philippine island of Luzón, two rice production areas were studied to understand the value of insect pollination. These areas belong to the provinces of Ifugao, which production system has been described above, and Laguna. Different than in Ifugao, rice production in Laguna is characterised by the implantation of modern agriculture. Landscape complexity, crop diversity and even cultural diversity are higher in Ifugao. However, when estimating the economic value of insect pollination, using standard methodologies, such value turns out to be much higher in Laguna than in Ifugao. This is due in part to the use of plant varieties that are on average more dependent on insect pollination. However, it is also due to higher volumes of crop production per capita associated to an intensive agricultural system. Arguably, the employed method is limited when recognising processes of self-consumption, or useful plants with medicinal uses or used in cultural practices. Nevertheless, the use of a monetary valuation approach in this case may provide misleading signals for management, that may give a boost to agricultural intensification based on the optimisation of ES provision (Puigdollers et al., 2015).

The issues with the valuation of ES have generated ample literature that would be difficult to summarise here. More recently, there is tendency to clarify that monetary valuation is not an issue per se, but there is a warning about the need of valuation methods to consider elements of additionality, equality, no complexity blinding and no dispossession (Kallis et al., 2013).

Coming back to TEEB, while this initiative is well informed about the importance of the environment for the livelihood of indigenous and poor rural people, some of its recommendations ‘greenwash’ large mining corporations. For instance, TEEB

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42 Ecosystem Services: concepts, methodologies and in truments for research and applied use

explicitly praised proposals (called “no net loss” or even “net positive impact” by Rio Tinto) to permit the destruction of a habitat if a certificate were presented confirming that an equivalent habitat was ‘created’ somewhere else. Making the certificates tradable would supposedly create a global market, supporting a flexible and cost-effective biodiversity protection system (Rodríguez-Labajos and Martínez-Alier, 2013). Biodiversity offsetting divides conservationists because instead of being used as last resort mechanisms to avoid further damage (that is, a sort of fine) may be used to enable fees to keep generating such damage and thus being counterproductive.

Moreover, the use of ES is linked to irreducible diverse value perspectives that may create divergent and even colliding views between stakeholders on what needs to be done. For instance, Jorda-Capdevila and Rodríguez-Labajos (2014) explain, for the issue of water flows, that flood irrigation of rice fields in their study area is seen as a water waste by some environmentalists and water managers. In contrast, other environmentalists and farmers argue that it contributes to replenish the groundwater and to preserve traditional landscapes. The authors detect that such divergences arise from traditional versus new environmentalist views, feelings toward wilderness versus what is exotic, private versus public use of the river, and monetary versus non-monetary values.

This makes unavoidable the emergence of trade-offs from management decisions and then the question is how these trade-offs are distributed. Local communities and other actors have disparate capacities to be heard and negotiate solutions that are favourable to them, which brings the distributions of costs and benefits in terms of ES to the arena of environmental justice and environmental governance.

From an environmental justice perspective, what activists demand is usually to halt the drivers that generate disruptive changes in the environment and in people’s wellbeing. That is the case of radical alternatives such the one of leaving fossil fuels unexploited (Espinosa, 2013). In this context, the use of ES terminology is seen with concern and even rejection (Yánez, 2015). However, some pragmatic recommendations for the use of ES evaluation in environmental justice struggles are also present in the recent developments of activist-led research (Zografos et al., 2014).

5. Conclusions

The ES framework has become a dominant paradigm of environmental assessment, as a way to link the state of the ecosystems with different constituents of human wellbeing. As shown through different examples, ES assessment is widespread and increasingly influential in environmental management and policies. Additionally, this framework has proven to facilitate the consideration of trade-offs and distributive issues.

Still, the use of the ES approach is not exempt from limitation. For instance, issues of biophysical realism and the inherent social complexity make it difficult to apply

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43Ecosystem services: application and conflicts

linear approaches to ES assessment. Also, ES valuation is itself legitimate, appropriate in some contexts. However, monetary valuation may turn to be counterproductive when it is part of a compensation of damages still to happen. Moreover, there is a risk to exclude the languages of valuation of less powerful actors.

All in all, ES management is a challenging task that need have to consider carefully equity and substitutability questions. Although criticism to the ES approach is frequent in the position of environmental justice organisations, ES assessment has been successfully used by environmental defenders in preventing projects that were likely to generate impacts on biodiversity.

Acknowledgements

The author acknowledges funding from the LEGATO and STACCATO projects and expresses her gratitude to Sara Blas, Dídac Jordá, Eloi Puigdollers, Clara Solé, Bianca Tilliger, Christos Zografos and the LEGATO partners for insightful discussions about the ES approach.

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Tilliger, B., Rodríguez-Labajos, B., Bustamante, J.V. and Settele, J. (2015). Disentangling Values in the Interrelations between Cultural Ecosystem Services and Landscape Conservation—A Case Study of the Ifugao Rice Terraces in the Philippines. Land 4, 888–913. doi:10.3390/land4030888

Yánez, I. (2015). ¡Y dale con los servicios ambientales! Marchamos porque no queremos un capitalismo “verde” en el Ecuador [WWW Document]. Acción Ecológica. URL http://www.accionecologica.org/editoriales/1844-iy-dale-con-los-servicios-ambientales-marchamos-porque-no-queremos-un-capitalismo-qverdeq-en-el-ecuador (accessed 10.29.15).

Zografos, C., Rodríguez-Labajos, B., Aydin, C.A., Cardoso, A., Matiku, P., Munguti, S., O’Connor, M., Ojo, G.U., Özkaynak, B., Slavov, T., Stoyanova, D. and Živčič, L. (2014). Economic tools for evaluating liabilities in environmental justice struggles. The EJOLT experience. EJOLT Report No. 16.

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Assessment and application of ecosystem services

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51

Socio-economic valuation of ecosystem services in Spain

Marina García-Llorente,1,2,* Cristina Quintas-Soriano,2,3 Pedro Zorrilla-Miras,4,5 María Loureiro,6 Carlos

Montes,2 Javier Benayas,2 Fernando Santos-Martín2

Abstract

The importance of ecosystems and their biodiversity in supporting human well-being through the supply of multiple ecosystem services is now widely recognized on the academy and from a conceptual point of view. However, the characterization and quantification of ecosystem services on social terms and its later translation on real implementation in environmental and land use policies is still very limited. The National Ecosystem Assessment of Spain addresses the process of economic valuation of priority ecosystems supported and complemented by a previous robust analysis of the biophysical dimension. It is the first nationwide economic assessment, with a mixed variety of methods which also capture services outside conventional markets and include social and cultural aspects, for both use and non-use values. Overall, twelve ecosystem services have been valued. Here we present the main finding obtained from: a meta-analysis of the studies previously done in Spain, through the use of market-based instruments applied to the case of food from agriculture, and a social preference assessment combined with a choice experiment exercise.

Keywords

Choice experiment, Market, Social preference, Spanish National Ecosystem Assessment, Systematic review, Valuation plurality.

1 Department of Applied Research and Agricultural Extension, Madrid Institute for Rural, Agricultural and Food Research and Development (IMIDRA), Madrid, Spain

2 Social-Ecological Systems Laboratory, Department of Ecology, Universidad Autónoma de Madrid, Madrid, Spain

3 Andalusian Center for the Assessment and Monitoring of Global Change (CAESCG), Biology and Geology Department, University of Almeria, Almería, Spain

4 School of Geosciences, University of Edimburgh, United Kingdom5 Terrativa Sociedad Cooperativa Madrileña, Spain6 Departamento de Fundamentos da Análise Económica, Facultade de Ciencias

Económicas e Empresariais, Universidade de Santiago de Compostela, Spain* Corresponding author email: [email protected]

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52 Ecosystem Services: concepts, methodologies and in truments for research and applied use

1. Introduction

Since the publication of the International Millennium Ecosystem Assessment (MA, 2005), interest in ecosystem service (ES) assessment has grown exponentially in environmental science and policy. Since the last years, the literature on ES has stressed the importance of integrating social, ecological, and economic dimensions of ES and biodiversity in environmental decision making (de Groot et al., 2010). Despite the academic progress, many important issues are still to be resolved in order to fully operationalize the ES framework in environmental policy targets as it has been acknowledge by the international ES initiatives like the Millennium Ecosystem Assessment (MA), The Economics of Ecosystem Services and Biodiversity (TEEB), and the Intergovernmental Platform for Biodiversity and Ecosystem Services (IPBES) and the European project Openness (Operationalisation of natural capital and ecosystems services: from concepts to real-world applications; http://www.openness-project.eu/; Gómez-Baggethun et al., 2014). In Europe, the EU Biodiversity Strategy (EC, 2011) has identified ‘the need to maintain and restore ecosystems and their services’, and commits that all Member States shall map and assess the state of ecosystems and their services in their national territories by 2014, assess the economic value of such services, and promote the integration of these values into accounting and reporting systems at EU and national level by 2020. Therefore a key challenge to be addressed in any evaluation of ES requires an integrated analysis, taking into account the multiple value dimensions, from the biophysical to the socio-cultural and economic assessment (Haines-Young and Postchin, 2010).

With this aim the National Ecosystem Assessment of Spain (Spanish NEA, www.ecomilenio.es) has addresses the economic valuation based on a robust analysis of the biophysical dimension (Spanish NEA, 2014) and with the implementation of mixed methodologies that include social and cultural aspects in the valuation process (Fig.1). Although in Spain it has been some experience before to value some ES in monetary units (VANE, 2008) this is the first nationwide economic assessment with a mixed variety of methods which also capture ES outside conventional markets and include both use and non-use values.

The fashion of ES valuation has also uncovered fundamental conceptual and methodological shortcomings, which should be taken into account (Kallis et al., 2013). The ecological complexity underlying the supply of ES cannot be completely translated into monetary value, and the hegemony of such value could be counterproductive if the final objective is understood in terms of the commodification of nature (Gómez-Baggethun and de Groot, 2010). At the same time, socio-cultural (also called non-monetary) techniques are required to bring to the table the multiple values of ES (i.e., cultural, educational, ethical, moral, historical, spiritual, inspirational, or therapeutic values), increasing the visibility of the intangible and incommensurable contributions provided by nature, which could be obscured under simplification into the metric of money (García-Llorente et al., 2016).

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53Socio-economic valuation of ecosystem services in Spain

Fig. 1. Multidimensional framework for assessing ES, including methods ranging from biophysical (supply-side) to socio-cultural and economic approaches (demand-side), and how this approach has been

incorporated into different steps during the Spanish NEA timeline.

Fig. 2. Conceptual framework of concentric relationship between the three dimensions of the evaluation. The biophysical dimension marks the ability of

ecosystems to provide services that contribute to human wellbeing and the socio-cultural context determines the economic dimension associated to the services.

The Spanish NEA understands human wellbeing as a good quality of life within the ecosystem’s biophysical limits following the postulates of Ecological Economics (Santos-Martin et al, 2013; Spanish NEA, 2014). An ecosystem’s capacity to supply services determines its range of potential uses by society, which influence its socio-cultural and monetary value (Martín-Lopez et al., 2014). These

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54 Ecosystem Services: concepts, methodologies and in truments for research and applied use

interdependencies (and the different information provided) explain why research on ES should combine the three value domains (biophysical, socio-cultural, and economic) to properly inform the environmental decision-making process (Fig. 2). The Spanish NEA attempts to build a common language between scientists and policy makers as well as a discussion forum around the idea of how to make the human dependence on ecosystems and their biodiversity explicit, working under a pluralistic and multidimensional framework.

Based on the results obtained in the biophysical assessment (Spanish NEA, 2014) we selected the target ES that were identified as priorities to value in economic terms at national level. Overall, twelve services were valued under various techniques (Table 1).

In the following sections we summarize the main findings using different methods, as well as some final conclusions and key remarks of this assessment.

Table 1. Ecosystems services evaluated economically with different valuation methods in the Spanish NEA.

Market based methods Meta-analysis Stated preference

Prov

isio

ning

Food X XWater X

Gene pool X

Regu

latio

n

Climate regulation XWater quality X X XErosion control X XNatural hazards X XBiological control X

Cul

tura

l

Recreation X XLocal ecological knowledge X

Spiritual X XAesthetic X

2. State of the art on the economic valuation of ecosystem services in Spain

Since the publication of Costanza et al. (1997), the number of scientific publications that considers the monetary valuation of ES has increased progressively and internationally. Numerous works have provided updated monetized estimates of the value of ES at global level, however no existing Spanish studies have been considered until now (De Groot et al., 2012; Costanza et al., 2014). To fill this knowledge gap and also as a first step to guide the assessment, this study aims

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55Socio-economic valuation of ecosystem services in Spain

to provide an overview of past and current economic valuation studies in Spain. To do that, we performed a systematic review of the published papers that were indexed in the ISI Web of Science (https://www.accesowok.fecyt.es/) until the end of 2012 focus on the economic valuation of ES in Spain. Overall, 150 papers were included in the review analysis.

Our results showed that the economic studies were mainly located in Andalusia (26.58% of studies located in this area), in Cataluña (17.72%) and in Galicia regions (5.06%) (Fig. 3). The majority of Spanish regions were associated with studies focused on different categories of ES, while two regions (the Canary Islands and the Balearic Islands) focused only on cultural ES.

Fig. 3. Number of economic valuation studies per regions.

Mainly the economic studies were located on coastal and marine systems (164 estimates, 25.38%), forests (135 estimates, 20.77%) and freshwater ecosystems (116 estimates, 17.85%) (Fig. 4). The ecosystem type with less number of economic value estimates were insular (12 estimates, 1.85%). Urban, arid systems, agroecosystems and mountains counted with more than 45 estimates each.

The greatest number of economic value estimates was primarily focused on cultural ES (41.76%), followed by regulating ES (34.98%) and provisioning ES (23.27%). Overall, nature tourism and recreation, freshwater and food were the ES with the highest number of economic value estimates. Although, significant associations were detected between the ecosystem type and services valued (Fig. 4).

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56 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Fig. 4. Number of economic value estimates per ecosystem types.

Finally, a meta-regression analysis was performed with the economic value estimates of ES extracted from the papers. More details of this research could be found in Quintas-Soriano et al. accepted.

3. Spatial valuation of ecosystem service using market-based methods

The biophysical assessment of the Spanish NEA detected which are the priority ES. From this list, those that could be assessed using market-based methods were: (1) agricultural production, (2) recreation (nature tourism), (3) natural hazards (wild fires), (4) water provision, and (5) water regulation. Here, we will present as an example the case of agricultural production valuation.

The valuation has been performed using a spatial explicit method, obtaining a mean value for crop production for the whole territory, but also a detailed map (pixel size + 5 km x 5 km) of crop production in biophysical units (tones) and monetary units (euro). Data came from the Ministry of Agriculture, Food and Environment of Spain to obtain the mean productivity of each crop per province, the mean price of crop and a mean of the area cultivated of each crop between 1996 and 2010 in each municipality. Finally we obtained the total annual agricultural production per municipality (Fig. 5 and 6).

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57Socio-economic valuation of ecosystem services in Spain

Fig. 5. Total agricultural production in Spain in biophysical units (tones/ha/year).

Fig. 6. Total agricultural production in Spain in monetary units (euro/ha/year).

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58 Ecosystem Services: concepts, methodologies and in truments for research and applied use

We also obtained the biophysical and monetary production of each type of crop in the different regions of Spain (Fig. 7 and 8). Comparing both maps we see that agricultural production is higher in Valencia region in tones, while in monetary units it is greater in Andalucia and La Rioja. This means that the production in Andalucia and La Rioja has a higher value per kilo than in Valencia. The production maps can be combined as well with other spatial information to analyze specific features. For example, we have combined it with the high nature value farmlands map produced by the European Topic Centre on Spatial Information and Analysis (EIONET).

Fig. 7 and 8. Production capacity of the agrarian systems of Spanish regions and proportion of the production generated by each type

of crop (tree crops, herbaceous crops and horticultures).

Finally we explored the tradeoffs between the agricultural production and the high nature value farmlands. We identified three different areas trends: areas with high economic value and low natural value; areas with low agricultural production and high natural value; and areas with high agricultural production and high natural value. Areas with high agricultural production and high natural value could be used as references of sustainable production systems, where a high agricultural production can be obtained while enhancing high natural value systems. We could detect that those areas correspond with traditional Mediterranean crops, like vineyards, olive trees, agroforestry systems (dehesas) and rain-fed cereals grown in areas with an appropriate climate and soil conditions to those crops. Also very productive systems that concentrate a high value birdlife (ricefields) are detected in the Ebro and Guadalquivir estuaries (Please visit www.ecomilenio.es to see more detailed information).

4. The socio-economic importance of ecosystem services following Spanish population preferences

The way in which society interacts with nature in terms of how we perceive, demand, use, enjoy, or value ES is usually polarised toward economic values (Seppelt et al., 2011); masking the wide range of social values attached to nature (Russell et al., 2013;

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59Socio-economic valuation of ecosystem services in Spain

Sandifer et al., 2015). To address this challenge, we had combined monetary and non-monetary techniques in a survey to analyse the Spanish population environmental concerns and preferences towards ES, including a willingness to pay exercise. In order to proceed, 800 valid online questionnaires were completed in 2014 involving a representative sample of the Spanish population; taking into account quotas for gender, age, and rural or urban inhabitants distributed across regions (Fig. 9).

Fig. 9. Sample effort per Spanish region having into account the actual population in each area. CL: Castilla y León, PV: País Vasco, CM: Castilla la Mancha.

The questionnaire had two differentiated parts, one including non-monetary aspects in terms of: social concern of environmental problems, perceptions of how environmental conservation affects human wellbeing, social importance of ES, respondent’s relationship with rural areas, and socio-demographic characteristics. The second section was based on a stated preference exercise through a choice experiment question that included final motivational reasons for the support obtained. This second section was done in order to better analyse social support for ES conservation in Spain. The choice experiment is an economic tool that evaluates the public’s preferences or support by asking respondents to choose from a series of alternatives of choice sets, each described in terms of different attributes and levels, related with ES conservation plans (Turner et al., 2010; García-Llorente et al., 2012). For this purpose, five ES were selected and valued: two regulating ES (soil protection and erosion control, water quality), two cultural ES (satisfaction for conserving species, local ecological knowledge) and, one provisioning ES (gene pool of native livestock breeds; in terms of agrobiodiversity) (Fig. 10).

These ES were selected following its importance in terms of provision but also its vulnerability in biophysical terms. Important regulating services and traditional cultural services associated with rural areas are declining (Spanish NEA, 2014) and so we used the choice experiment with the aim of exploring Spanish population prioritisation in regard to the design of an environmental policy to maintain and promote those five vulnerable ES.

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60 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Fig. 10. Example of a choice set containing in its Spanish original version. Each card contain two future options (1 and 2) a status quo option. Each profile is described in

terms of five ES and a payment. A questionnaire contained a sequence of nine choice sets.

Among the main results, three environmental concerns were prioritised: water scarcity, environmental pollution and climate change, with more than 70% of the population stating that these issues are very important or important and urgent measures should be taken (Fig. 11). Biodiversity loss, the lack of awareness of ecosystem values and the current lifestyles related with unsustainable production and consumption patterns were also highly mentioned.In addition, the population recognized that environmental degradation has a direct influence on human wellbeing, particularly on the dimensions related with basic materials, health and security. Ecosystems contribute to social connectedness in terms of social capital and cohesion and sense of belonging (Russell et al., 2013). In addition, the interaction with ecosystems have been related with psychological benefits (ex. reducing anxiety), cognitive (ex. improving learning capabilities), or even with the increase of social resilience (Sandifer et al., 2015). However, how ecosystems influence on our good social relationships and freedom for action and election was less evident (Fig. 12).

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61Socio-economic valuation of ecosystem services in Spain

Fig. 11. Social concern of environmental issues.

Fig. 12. Influence of environmental degradation in human wellbeing components.

Finally, in the choice experiment exercise, all ES were significant in the choice of a conservation and higher levels of any unique attribute increased the probability that a management option would be chosen. The most valued services (ranked from most to least important) were: water quality, biodiversity conservation, gene pool, local ecological knowledge, and erosion control. It is also remarkable that the main reasons to financially support conservation plans were related with: bequest, ecological and health values; followed by a precautionary investment (between others), understanding that it is better to invest now than investing higher in a later stage (Fig. 13).

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62 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Fig. 13. Motivations behind willingess to pay in relation with ecosystems values.

5. Conclusions

Two of the greatest challenges of any economic assessment process are: (1) risk of simplify ecological complexity and not following the biophysical and spatial boundaries, (2) reproduce the logic of market shaping the preferences, reducing them to mere metric or obscuring other reasons that promote conservation. In order to overcome these risks in this study we have: (i) made a special effort to represent the results in a spatially explicit way; (ii) used different complementary techniques: markets based method, stated preferences (choice experiments) and meta-analysis, and (iii) techniques have been contextualized by biophysical and social characteristics.

We believe that expressing the value of ES in economic and social importance is a powerful tool because: (1) the majority of planning decisions are based on economic information and thus better information of the importance of ecosystems in economic terms is crucial to achieve more accurate decisions, (2) visualize those ES without market value (i.e. regulating and cultural) is a necessary and, (3) it is a powerful tool of communication to society. Finally, we consider that any ES assessment, should in first place respect ecosystem biophysical limits; as ecosystem services are the proxy to visualize ecological functions relevance. Further knowledge is still required in terms of understanding the ecosystems integrity, their resilience, and thresholds under different management options, in order to properly contextualise any ES assessment and potential trade-offs.

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63Socio-economic valuation of ecosystem services in Spain

Acknowledgments

We really appreciate the positive insights and valuable comments made by B. Martín López on this research. Funding for this project was supported by the Biodiversity Foundation through the Spanish Millennium Ecosystem Assessment project (http://www.ecomilenio.es/), MGL was founded by a postdoctoral grant from the Spanish National Institute for Agriculture and Food Research and Technology (INIA), which is co-funded by the Social European Fund.

Bibliographical references

Costanza, R., de Groot, R., Sutton, P., van der Ploeg, S., Anderson, S. J., Kubiszewski, I., ... and Turner, R. K. (2014). Changes in the global value of ecosystem services. Global Environmental Change, 26, 152-158.

Costanza, R., d'Arge, R., De Groot, R., Farber, S., Grasso, M., Hannon, B., ... and Van Den Belt, M. (1998). The value of the world's ecosystem services and natural capital. Ecological economics, 1(25), 3-15.

De Groot, R., Brander, L., Van Der Ploeg, S., Costanza, R., Bernard, F., Braat, L., ... and Van Beukering, P. (2012). Global estimates of the value of ecosystems and their services in monetary units. Ecosystem services, 1(1), 50-61..

De Groot, R. S., Alkemade, R., Braat, L., Hein, L., and Willemen, L. (2010). Challenges in integrating the concept of ecosystem services and values in landscape planning, management and decision making.  Ecological Complexity,7(3), 260-272.

European Commission, 2011. Our life insurance, our natural capital: EU biodiversity strategy to 2020. Brussels.

García-Llorente, M., Castro, A. J., Quintas-Soriano, C., López, I., Castro, H., Montes, C., and Martín-López, B. (2016). The value of time in biological conservation and supplied ecosystem services: A willingness to give up time exercise. Journal of Arid Environments, 124, 13-21.

García-Llorente, M., Martín-López, B., Nunes, P. A. L. D., Castro, A. J., and Montes, C. (2012). A choice experiment study for land-use scenarios in semi-arid watershed environments. Journal of Arid Environments, 87, 219-230.

Gómez-Baggethun, E., and De Groot, R. (2010). Natural capital and ecosystem services: the ecological foundation of human society’, in Hester & Harrison (eds), Ecosystem Services: Issues in Environmental Science and Technology, Cambridge pp. 105–121.

Gómez-Baggethun, E., Martín-López, B., Barton, D., et al. (2014). EU FP7 OpenNESS Project Deliverable 4.1., State-of-the-art report on integrated valuation of ecosystem services. European Commission FP7.

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Haines-Young, R., and Potschin, M. (2010). The links between biodiversity, ecosystem services and human well-being. Ch 7. In: Raffaelli, D., Frid, C. (Eds.), Ecosystem Ecology: A New Synthesis. BES Ecological Reviews Series, CUP, Cambridge, in press.

Kallis, G., Gómez-Baggethun, E., and Zografos, C. (2013). To value or not to value? That is not the question. Ecological Economics, 94, 97–105.

Martín-López, B., Gómez-Baggethun, E., García-Llorente, M., and Montes, C. (2014). Trade-offs across value-domains in ecosystem services assessment.Ecological Indicators, 37, 220-228.

Millennium Ecosystem Assessment (2005) Ecosystems and Human Wellbeing. World Resources Institute, Washington, DC.

Quintas-Soriano, C., Martín-López, B., Santos-Martín, F., Loureiro, M., Montes, C., Benayas, J., and García-Llorente, M. (accepted). Ecosystem services values in Spain: a meta-analysis. Environmental Science & Policy.

Russell, R., Guerry, A. D., Balvanera, P., Gould, R. K., Basurto, X., Chan, K. M., ... and Tam, J. (2013). Humans and nature: how knowing and experiencing nature affect well-being. Annual Review of Environment and Resources, 38, 473-502.

Sandifer, P. A., Sutton-Grier, A. E., and Ward, B. P. (2015). Exploring connections among nature, biodiversity, ecosystem services, and human health and well-being: Opportunities to enhance health and biodiversity conservation. Ecosystem Services, 12, 1-15.

Santos-Martín, F., Martín-López, B., García-Llorente, M., Aguado, M., Benayas, J., and Montes, C. (2013). Unraveling the relationships between ecosystems and human wellbeing in Spain. PloS one, 8(9), e73249.

Seppelt, R., Dormann, C. F., Eppink, F. V., Lautenbach, S., and Schmidt, S. (2011). A quantitative review of ecosystem service studies: approaches, shortcomings and the road ahead. Journal of applied Ecology, 48(3), 630-636.

Spanish Millennium Ecosystem Assessment, SNEA (2014). In: Santos-Martin, F., Montes, C., Benayas, J. (Eds.). Ecosystems and biodiversity for human wellbeing. Synthesis of the key findings. Biodiversity Foundation of the Spanish Ministry of Agriculture, Food and Environment. Madrid, Spain 90 pp.

Turner, R.K., Morse-Jones, S., and Fisher, B. (2010). Ecosystem valuation. A sequential decision support system and quality assessment issues. Annals of the New York Academy of Sciences 1185, 79e101.

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65

Monetary valuation of urban ecosystem services-operationalization

or tragedy of well-intentioned valuation? An illustrated example

David N. BartonNorwegian Institute for Nature Research (NINA), Oslo, Norway1

1. Introduction

A recent review of published monetary valuation methods of ecosystem services found that only a small fraction of studies discussed how estimates were or could be used for different kinds of decision-support (Laurans et al., 2013). Possible explanations range from decision-makers not having sufficient training in economics, the cost of valuation, inaccuracies of monetary valuation, political strategies requiring opacity or ambiguity, regulatory frameworks not being conducive to ecosystem service valuation and finally, fundamental inadequacies of valuation. Kallis et al. (2013) summarise four fundamental reasons why monetary valuation of ecosystems is inadequate, including ecosystems’ high complexity and interconnectedness, multiple rationalities and values; dependence of values on distributional and institutional settings, and the fact that valuation is a social processes conditioned by value articulating institutions. Gomez-Baggethun and Barton (2013) review a number of additional technical challenges with valuing ecosystem services in urban contexts. Despite the large number of practical and theoretical limitations on monetary valuation, Kallis et al. (2013) suggest that monetary valuation of ES can still be policy relevant if it meets several ecological economic and political economic criteria. Failure to address these criteria can result in ‘well-intentioned valuation’, contributing to deterioration of environmental conditions, inequalities and redistribution of power, suppression of other languages of valuation and enclosure of the commons.

1 [email protected]

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In this chapter we test these criteria on four valuation studies conducted in Oslo by the EU FP7 project OpenNESS, which used a mix of original studies and value transfers (Barton et al., 2015a):

1. Meta-analysis of willingness-to-pay for green spaces in the built zone2. Hedonic pricing of green infrastructure in the built zone of Oslo3. Time use value of Marka peri-urban forest outside the built zone of Oslo4. Liability value of urban trees in the built zone

The two methods looking at recreation in green spaces (1) and the peri-urban forest (3) found annual values between one and several billion Norwegian kroner. The value of green spaces in property prices (2) and the liability value of city trees (4) revealed capital values in the range of tens of billions of Norwegian kroner. The study was quite widely reported in the Norwegian press and online. However, several of the valuation studies do not pass Kallis et al. criteria for when monetary valuation is desirable. Were the methods applied examples of ‘the tragedy of well-intentioned valuation?’

2. Fundamental inadequacies of monetary valuation — an illustrated guide

Kallis et al. (2013) detail what Laurens et al. (2013) also call ‘fundamental inadequacies’ of monetary valuation of ecosystem services:

1. Ecosystems are highly complex and interconnected (critical species and systems). Their values cannot be compressed into a single metric.

2. Multiple rationalities entail multiple values and other relevant valuation languages than those expressed in monetary terms

3. There is no unique value for environmental goods and services independent of the distributional and institutional settings within which such values are expressed.

4. Social processes of valuation, including monetary valuation, are value articulating institutions (VAIs). Different people exhibit different values depending on the socio-institutional environment in which they express them

Together these constitute a fundamental plea for plural values and plural value-articulating institutions. The key to Kallis et al.(2013) paper is that their plea does not exclude monetary valuation methods, but suggests that valuation languages are context dependent. The next sections provide an illustrated guide to some fundamental challenges for valuation of urban ecosystem services.

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67Monetary valuation of urban ecosystem services [...]

2.1. Ecosystems complexity and interconnectedness“Ecosystems complexity and interconnectedness cannot be compressed into a single metric” (Kallis et al., 2013). This is a theoretical position as much as it is practically evident when reviewing approaches to ecosystem mapping and modeling. A number of different metrics at different spatial scales and levels of species, population, and community organization are used to describe ecological importance or ‘ecological value’ (Figure 1).

Figure 1 Different levels of ecosystem function metrics

Source: own elaboration based on Duany Plater-Berk & Company landscape gradient illustration at reeassociationdesign.files.wordpress.com

• Metric type A — mapping of green structures and nature types indicate important green infrastructure features;

• Metric type B — structures and nature types have biophysical capacities as habitats in terms of food resources, shelter, range etc.;

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• Metric type C — individual organisms, including humans, have observable habitat choices/preferences;

• Metric type D — individual organisms’ habitat occupancy are mediated by functional traits and they are of importance for other organisms;

• Metric type E — groups of organisms have functional traits of importance for other communities;

• Metric type F — functional diversity define ecosystems and their capacity to provide ES. Functional traits have a spatial and temporal definition which define ecosystem functions’ spatial and temporal extent.

Mapping and modeling of metrics A-F provide (i) insight into multiple ‘ecological values’ and (ii) alternative spatial definitions of ‘ecosystem’, ‘ecosystem function’ and ‘ecosystem service’ along rural-urban gradients of green space fragmentation and built area. Monetary valuation of ecosystem services assumes that ecosystem services can be spatially and temporally demarcated. At the same time this makes ecological functions invisible across boundaries of commodification assumed implicitly or explicitly by monetary valuation method (Vatn, 2005b). Even a simplification such as in Figure 1 makes it evident that using only monetary valuation of ES for decision-support leads to a large loss of information. In the case study from Oslo we question whether there are decision-support contexts where monetary valuation may still be informative without the ‘ecological resolution’ outlined in Figure 1.

2.2. Other relevant valuation languages“There are other relevant valuation languages than those expressed in monetary terms” (Kallis et al., 2013). Figure 2 shows the hypothetical relationships between different kinds of values, human needs, and ecosystem services (Gómez-Baggethun et al., 2014). A given ecosystem services satisfies a set of needs at different levels —conceptualized here using Maslow’s (1943) needs hierarchy. In turn, economic, cultural and ecological values stem from different combination of needs at different levels. There is no a priori commensurability between the different types of needs in the hierarchy (Max-Neef, 1992). All the different needs must be met integrally for a human being to be a healthy and happy —they are so— called ‘functionalized elements’ of an individual (Vatn, 2005a). By extension different types of values and ecosystem services are non-commensurable. Recognizing multiple values is required to capture the diversity of needs and wants that nature can contribute to fulfill for society and individuals (Gómez-Baggethun et al., 2014).

The links in the model are hypotheses that need further evaluation. For example, in Figure 2 economic values are not relevant for the ‘higher needs’ of affections and sense of belonging, esteem and identity and self-actualization (self-realisation). These needs are on the other side associated with cultural ecosystem services. Interpreting the conceptual model literally suggests that economic values — and by extension monetary valuation — is not appropriate to address cultural ecosystem services. In the case study from Oslo we apply monetary valuation to several recreational

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(cultural) ecosystem services. Which types of needs are satisfied by recreation in urban green infrastructure? Can we argue that we are valuing cultural ecosystem services and if so are these monetary valuation examples fundamentally inadequate?

Figure 2. Different levels of human needs as a basis for different valuation languages.

Source: Gómez-Baggethun et al. (2014)

2.3. Context dependent values of ecosystem services“There is no unique value for environmental goods and services independent of the distributional and institutional settings within which such values are expressed” (Kallis et al., 2013). Valuation of ecosystem services is decision-context specific because values are an expression of preferences for alternative courses of action with alternative benefits and values. In this sense “total economic valuation” of ecosystems has limited usefulness for decision support (Brouwer et al., 2013). Figure 3 uses the ecosystem services cascade framework (Haines-Young and Potschin, 2010) to illustrate decision context dependency of values. The example from Barton et al. (2015a) is for catchment nutrient run-off, eutrophication and recreational lake use. The example illustrates the functional interaction of regulating and cultural services, but any cascade of effects would serve the same illustrative purpose.

The combination of spatial scale and resolution determines information about ecosystem structure. The number of locations and time steps in the catchment monitoring programme determine the information about ecological function of the catchment system. The number of locations and times we record perceptions determine the extent of our knowledge about the ecosystem services of nutrient mitigation provided by blue-green infrastructure in the catchment. The combined variation across ecosystem structures, function and service end-points describes biophysical

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heterogeneity. Ecosystem benefits are determined by how well we can identify potential lake user occasions and the population of lake users’ personal characteristics (child, man, woman, family, single etc.). Different catchment management measures determine the potential improvement in lake suitability according to a regulatory classification system for lake ecological status (red-yellow, red-green etc.). Individual willingness-to-pay for nutrient mitigation measures depends on how the management decisions are framed in terms of the number of management choice alternatives, and the time between investment in mitigation and improvement in lake ecological status. The extent to which researchers have identified the decision alternatives and horizons and the different individuals’ reactions to them, determine their knowledge of ecosystem values. In summary, values are plural because they are place, time, group and person specific. The combined variation from ecosystem service end-point, benefits and values is what we called socio-cultural heterogeneity in Figure 3.

Even for this relatively simple example of a rural catchment experiencing lake eutrophication we can envisage a large number of value contexts. For the Oslo urban example below we are dealing with a more complex picture, including fragmentation of green structures, high population density and high cultural diversity. Cities are ‘high context density environments’. They represent one of the most challenging contexts for ecosystem service valuation (Gómez-Baggethun and Barton, 2013).

Figure 3. Spatial and temporal context dependency of values

Source: adapted from Barton et al. (2015a)

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2.4. Social processes of valuation, including monetary valuation are value articulating institutions“Different people exhibit different values depending on the socio-institutional environment in which they express them” (Kallis et al., 2013). Social processes of valuation, including monetary valuation, are value articulating institutions (VAIs) (Vatn, 2005a). The issue here is the value elicitation context which must frame respondents choices and statements about choices in an institutional setting for them to be credible. Values plurality is the combined diversity of fragmented urban nature, citizens’ perspectives on fragments, and different institutional valuation frames (Figure 4). Sources of variation can be grouped into at least five types of value metric at different levels of organisation:

• Metric type 1 — Capacity of blue-green structures and spaces for different uses. Environmental quality. Amenity. Capacity is a metric of potential demand and value.

• Metric type 2 — Suitability for different uses, defined by minimum user requirements relative to capacity. Suitability is also a potential demand and value.

• Metric type 3 — Individual activity-location choices are conditional on blue-green structures’ capacity and suitability, and individuals capabilities. Individuals capabilities condition potential demand and value. Information on choices made by individuals defines what economists mean by demand. In this framing stated choices are indicators of potential demand, similar to value metric 1 and 2. Actual choices reflect actual demand, what economists call revealed preferences.

• Metric type 4 — Individual roles in social contexts determine norms which condition choices. Individuals can have different roles in different choice settings, enriching understanding of value plurality.

• Metric type 5 — Value articulating institutions (VAI) are different contexts where choices are constructed and importance stated, or frames (Kahneman and Tversky, 2000) where actual choices are revealed. Valuation methods are considered institutions in the theory of VAI. Conversely, institutions can be considered implicit valuation context, where for example opportunity costs follow from referendum choice, jurisprudence or regulation.

In the context of the Oslo case study below we compare monetary values generated using four different monetary valuation methods, which have their basis in four different institutional framings: willingness to pay municipal fees for park maintenance; value of recreation time in open access forests; hedonic pricing of green spaces in the property market; and compensation value of city trees as an economic liability on municipal property. We ask the question whether any of these value articulating institutions are favoured by Kallis et al. (2013) framework for monetary valuation.

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Figure 4. Different value articulating institutions’ frame preference formation and elicitation in different socio-cultural, individual, and physical contexts

Source: own elaboration based on Duany Plater-Berk & Company landscape gradient illustration at freeassociationdesign.files.wordpress.com

Table 1 Guiding framework criteria for desirable monetary valuation

Source: based on Kallis et al. (2013)

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3. Guiding Framework for monetary valuation

Kallis et al. (2013) suggest that the question is not “to value, or not to value” ecosystem services in monetary terms, but rather when and where to value monetarily. They propose a Guiding Framework of control questions for when monetary valuation of ecosystem services can be ‘desirable’. Table 1 summarises the questions which define the contexts in which monetary valuation is desirable from an ecological and political economic point of view.

A valuation method is desirable if it improves environmental conditions, reduces inequalities or redistributes power to the weak, does not supress other valuation language and value articulating institutions, and does not serve a process of enclosing commons. Valuation may be acceptable if some of the criteria are met only partially, depending on the context.

In the example from Oslo we ask whether the additionality and equality criteria can be evaluated ex ante or in the short term if environmental and distribution impacts of policies take time to observe. We have to interpret questions 1 and 2 in terms of subjective prior likelihoods. Question 3 must be assessed on the contents of the study at hand. Is a monetary valuation study that does not refer to other types of values interpreted as suppressing other languages, or can it be seen as one more type of advocacy in public for a with many voices? Kallis et al. (2013) hold monetary valuation to high standards based on the assumption that monetisation is already a dominant language, but this assumption has also been questioned (Laurans et al., 2013). Question 4 is also difficult to answer in the context of decision-support. Strictly speaking, it is possible to answer by looking at the rights allocation assumptions of the monetary valuation method. However, enclosure from privatization of rights ultimately depends on how policy-makers use information from valuation studies in the policy process. As a whole, Kallis et al. (2013) guiding framework could be seen as a form of ‘safe minimum standards’ for monetary valuation. We also note that Kallis and co-authors examples all concern implementation of economic instruments (damage compensation, water pricing, markets and public payments for ecosystem services). Their examples do not address ecosystem service assessment and valuation methods per se. In the case study below we therefore subject four monetary valuation methods to criteria in the Guiding Framework laid out in Table 1.

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Figure 7. Overlapping monetary values of urban cultural ecosystem services for the purpose of advocacy in Oslo

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4. Case study assessment of monetary valuation methods

Oslo Municipality covers an area of 454 km2 of which 287 km2 are forest, and 28 km2 of green space within the built zone (grey area in upper panel Figure 7). Norway’s capital has a population of 650 000 (2015), predicted to rise to 820 000 by 2030 (OsloKommune, 2015). It is Europe’s fastest growing capital city in percentage terms. Oslo’s peri-urban forest border, or ‘Markagrense’, is protected by law, obliging the municipality’s population growth to be accommodated through a densification strategy around transport nodes, in industrial transformation zones, or de facto through growth into neighbouring municipalities.

Several monetary valuation studies were carried out at municipal level (Barton et al., 2015a; Barton et al., 2015b) and neighbourhood/project level (Reinvang et al., 2014) in the same period as the public hearing process for the municipal plan for 2030 was under way (OsloKommune, 2015). Here we focus on the municipal level studies for the purposes of ‘awareness-raising’, or ‘advocacy’ sensu Laurens et al. (2013) (middle panel Figure 7). The methods used were

1. Willingness-to-pay for parks based on a meta-analysis benefits transfer2. Hedonic property pricing green space proximity3. Value of annual recreation time in Marka forest4. Compensation value for city trees using economic liability on municipal

land

The two methods looking at recreation in green spaces (1) and the peri-urban forest (3) found annual values between one and several billion Norwegian kroner. The value of green spaces in property prices (2) and the liability value of city trees

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(4) revealed capital values in the range of tens of billions of Norwegian kroner Cultural ecosystem service values are derived from assessments of trees and forest in different partially overlapping spatial contexts (illustration lower panel, Figure 7). Barton et al. (2015) acknowledge that monetary values overlap and are ‘double counted’ in some cases. Following Kallis et al. (2013) we next subject each monetary valuation method to the four questions of the guiding framework. Conclusions are summarised in Table 2.

Table 2. To do or not to do valuation for the purpose of advocacy?

4.1. Willingness-to-pay for parks based on a meta-analysis benefits transfer?

Barton et al. (2015) calculated total WTP for green spaces in Oslo using a meta-analysis function of willingness-to-pay to conserve green space (Brander and Koetse, 2011) adjusting for income and green space area. Will the environment be improved by using this valuation method? No. Probably not given the existing protected status that most green spaces have as parks in city landuse zoning. Unless a bespoke WTP study was targeted at the smallest unregulated green spaces in the city, it is difficult to argue that monetary estimates would make any contribution to existing zoning plans. For new neighbourhoods we might be able to answer ‘yes’ only if we can demonstrate that WTP was a decisive argument in establishing new parks. That does not seem likely. Is inequality reduced? No. Probably not, unless WTP provided the decisive argument for planning new parks in poor neighbourhoods with low access to green spaces. The value transfer does not have this spatial resolution. Does the valuation method supress value plurality? No. Given that the discussion on valuation of urban ecosystem services is only a few years old in Norway (Lindhjem and Sørheim, 2012) and there are few previous examples of monetary valuation of urban green space in Oslo (Strand and Wahl, 1997; Sælen and Ericson, 2013) I would argue that it has not been familiar in urban planning circles. If that

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is true monetary valuation actually adds a valuation language to the debate in Oslo on urban planning. Is a process of enclosure of the commons promoted by the valuation method? Maybe. The premise for WTP estimates transferred to Oslo from other studies is that households would pay a municipal tax that would go towards conserving green space. If the municipal taxes were not introduced, parks in Oslo would probably not decline, thanks to protection in the zoning plan. The services delivered by parks are not obviously conditional on a new tax. The language of the WTP contributes to a ‘discourse of enclosure’, but a municipal tax earmarked parks would not change the right of free access to parks. On the other hand, the reliability of the WTP estimates are open to attack on purely methodological grounds for being transferred to an inappropriate institutional context. In summary, the Guidance Framework would probably council ‘do not value’ city parks ecosystem services using value transfer of WTP. At best the value transfer method is an (admittedly low cost) undertaking with little impact on operational city planning. At worst it promotes a discourse of enclosure.

4.2. Hedonic property pricing of green space proximity?Barton et al. (2015) calculated the total incremental capital value of apartments’ proximity to parks, open spaces, water features in parks, Oslo’s coastline and peri-urban forest. Will the environment be improved by using this valuation method? Maybe. Not from applying the method itself. However, if estimates are found credible by planners, property developers and apartment owners, it advocates more use of green and blue structures in planning of new neighbourhoods, and conservation of green infrastructure in the face of urban densification. Is inequality reduced? No. Not initially. The hedonic value of green spaces is expressed through the property market and preferences of home-buyers who by definition have access to enough capital and have incomes to maintain a mortgage. If there is reason to believe their preferences for proximity to green space are different from those of households without access to mortgages, and hedonic pricing results are used to target particular types of green space, then hedonic pricing may increase inequality. The questions cannot be answered without knowing how value estimates are used for decision-support. Does the valuation method supress value plurality? No. The hedonic pricing method on its own does not recognise other types of value, but if reported in context with other valuation methods it promotes value plurality, as argued above. Is a process of enclosure of the commons promoted by the valuation method? No. The property market–used to reveal hedonic values of proximity to green space —is already ‘enclosed’ or commodified. On the contrary, the hedonic property method derives value from the fact that the proximal green spaces are public and open access to the private property owner. In summary, the Guidance Framework would council ‘maybe value’, depending on some further assumptions about the preferences of the poor for green space, and how the valuation results would be used for decision-support.

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4.3. Value of annual recreation time in Marka forest?Barton et al. (2015) calculated the total annual visits and hours spent in Oslo’s peri-urban ‘Marka’ forest using existing survey data for the adult population of Oslo (Synnovate, 2011). Total time spent in the forest is valued using different methods to test robustness: opportunity costs of labour after tax, choice experiment and travel cost from a previous studies (Sælen and Ericson, 2013). Will the environment be improved by using this valuation method? No. Not directly for the same reasons as under 4.1. The Marka forest is protected by a special law which regulates the management of the forest for recreational purposes.The aggregate valuation does not address trade-offs in how different recreational activities are zoned within the forest, nor forestry impact on recreation. Is inequality reduced? No. We cannot argue that aggregate estimates of value across a whole forest ecosystem helps groups with poor access. In fact, the method only accounts for time spent by adults responding to surveys. Time spent by children is not captured, and immigrants are often poorly represented in panels used for surveys of Oslo’s population. Does the valuation method supress value plurality? No. In this case non-monetary methods (time use) were used together with stated and revealed preference valuation methods. Is a process of enclosure of the commons promoted by the valuation method? No. The valuation methods make no assumptions about changes in property or use rights in order to elicit values. In summary, the Guidance Framework would probably council ‘maybe value’. The valuation methods are low cost, but also unlikely to make any operational impacts on zoning of forest landuses.

4.4. Compensation value for city trees using economic liability on municipal land?Barton et al. (2015) used Oslo municipality’s VAT03 assessment method for compensation value of damaged city trees (Randrup, 2005) to estimate the total economic liability value of 700 000–1,2 million city trees taller than 5 meters in Oslo’s built zone. Will the environment be improved by using this valuation method? Yes. The VAT03 assessment method is already applied to calculate compensation values for damage to individual trees on municipal land, arguably providing an additional incentive to protect Oslo’s trees. One could argue that using the method to calculate an aggregate potential economic liability contributes to making the responsibility for city trees more widely known in the population. However, we have no data on actual environmental impact of the VAT03 method, only a hypothesis awaiting further impact analysis. Is inequality reduced? Maybe. If we can argue that street trees are conserved to a greater extent, then we can also make a case that this favours poorer households in the inner city disproportionately, because the greatest concentration of trees on municipal land are found here. In the outer city, most individual trees are on private property, or in parks accessible to the whole population. Does the valuation method supress value plurality? No. The compensation value of municipal trees complements municipal regulation protecting trees. Compensation value comes into play in situations of negligent

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or wilful damage, motivated by the multiple ecosystem services values that are qualified in the VAT03 assessment methodology. Is a process of enclosure of the commons promoted by the valuation method? No. Quite the reverse. Applying the VAT03 method also to city trees on private land actually promotes the idea that there are public values of privately owned trees. In summary, the Guidance Framework would probably council ‘to value’.

5. Decision context dependent use of monetary valuation methods

By applying Kallis et al. (2013) Guiding Framework we concluded that we should probably not apply value transfer using WTP, maybe apply hedonic pricing and time use valuation, maybe apply recreational time valuation, and probably value the economic liability for city trees (Table 2). The framework is able to discriminate monetary valuation methods in our case study, as it did for economic instruments in the original paper. However, we saw that many of the answers depended on assumptions about the purpose of the valuation study. Barton et al. (2015) reported aggregate values from different types of urban green infrastructure framed as awareness-raising, with low requirements for accuracy and reliability. Valuation of urban ecosystem services in this case is advocacy for conservation of urban green infrastructure. Should Kallis et al. (2013) criteria be applied less rigorously in such an advocacy context? In the context of Oslo we would have to assess whether the ‘monetary narrative’ of willingness-to-pay for conserving urban parks is outweighed by the danger of promoting a ‘discourse of enclosure’. The answer depends on the extent to which different narratives currently play a role in the debate in Oslo.

Table 3. To do or not to do monetary valuation for decisive and technical purposes?

Source: based on Barton et al. (2015)

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Barton et al. (2015) also assess the possible decisive and technical purposes that these valuation methods could be applied to if they had higher spatial resolution. Table 3 summarises these recommendations showing what types of contexts we found could potentially be relevant for the different valuation methods when higher resolution and more data was available. Should Kallis et al. criteria apply equally across all these purposes?

From the examples used by Kallis et al. (2013) they seem to have been devised mainly for the context of instrument design and liability. Monetary valuation methods as used in priority-setting or accounting is not tested with their framework.

If we look more closely at the first criteria of their Framework —whether monetary valuation reduces environmental impact—we can show how the purpose of the valuation determines whether we answer ‘yes’, ‘maybe’ or ‘no’. Figure 3 illustrates the hypothesis that lies behind the horizontal axis in Figure 7 (middle panel); requirements for reliability are increasing as we pass from informative, through decisive, to technical purposes of valuation (in the sense defined by Laurens et al. (2013)).

Figure 8. Increasing demands for accuracy of monetary valuation in different decision-support contexts.

In figure 3 accuracy is illustrated with error bars in relation to the expected value of benefits and costs (i.e. a coefficient of variation). Reliability refers to whether decisions can be made repeatedly with the confidence level required by the decision-maker using the valuation information (Schroter et al., 2014).

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The advocacy in Barton et al. (2015a) uses monetary valuation methods to show that “nature in Oslo is worth billions”. Monetary valuation methods that only show “millions” for ecosystem services at municipal level do not have the same “big numbers” effect. What constitutes an ‘awareness-raising number’ depends on expectations and points of comparison. Billions are “big numbers” relative to site specific studies (Reinvang et al., 2014), and operate at the same magnitude as for example as municipal budgets. In the case of ecosystem accounting, the purpose is monitoring natural capital. Monetary valuation needs to have sufficient accuracy to identify trends in the value of ecosystem services. In priority-setting there are several distinctions to be made. A familiar distinction in benefit-cost analysis is between screening –taking action based on confidence about benefits being greater than costs —and ranking— identifying with confidence the decision alternative with the highest net benefits. Ranking requires greater accuracy than screening. For instrument design, such as water pricing or PES (Kallis et al., 2013), prices must either cover full economic costs, or in a role as incentives be higher than opportunity costs, but lower than willingness-to-pay (with some level of confidence). We see this as at least, or more, demanding than ranking. Finally, in a legal setting economic liability for interim damages (before ecosystem services recover) would seem to place the highest standards on monetary estimates. We assume here that reliability and accuracy of monetary valuation results must stand up to a high level of scrutiny in court by a jury. The extent to which punitive fines —negative incentives— are considered, complicates this picture.

Notwithstanding our simplifications, in each setting we can see that monetary valuation has insufficient reliability and accuracy if it leads to either no advocacy; no identification of trends; or even worse, to false positive errors in screening of policies; choosing a suboptimal design alternative; have no or perverse incentive effects; or entail lacking or excessive compensation and ‘unfairness’ to one of the parties. In each case we would argue that environmental impacts were not improved by the monetary valuation method because it was applied inappropriately for the requirements of the context.

Criteria for ‘desirable’ valuation in one setting may lead to the opposite conclusion in another setting. For example, cultural ecosystem service values in Barton et al. (2015) should not be aggregated because they partially assess trees and forest in different overlapping spatial contexts in Oslo. This would lead to double counting in the context of national accounts or priority-setting using benefit-cost analysis. However, for the purpose of advocacy, we argue that overlapping values provide mutual support (providing they are of the same magnitude). What carries greater weight in public advocacy, one big aggregate value, or several mutually supporting values? One argument or many supporting arguments?

Different decision-support contexts with different requirements for reliability, mean that the information value of monetary valuation methods —of any valuation method— varies with purpose. Information value —the benefits of improved decisions— should also be judged against the information costs of

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conducting different valuation methods (Figure 9). With the examples from Oslo we have implicitly been arguing that sometimes study costs may be justified if the information costs are low (value transfer), the purpose is advocacy, despite not meeting all the criteria for ‘desirable’ valuation.

As argued by Kallis et al. (2013) and illustrated in section 2, monetary values are highly context specific in a number of ways. Context specificity requires higher resolution of valuation methods. Urban ecosystem services challenge the capabilities of both socio-cultural and monetary valuation methods(Gómez-Baggethun and Barton, 2013). Further work is needed in exploring the boundaries of spatial and temporal resolution and decision-support contexts where monetary valuation methods complement oneanother. And where ecological and socio-cultural valuation methods start to have higher information value than monetary valuation methods.

Figure 9 To do or not to do monetary valuation is a question of a method’s information value relative to the accuracy and reliability

requirements in different decision-support contexts.

Source: based on Gomez-Baggethun and Barton (2013) and Schröter et al. (2014)

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Conclusions

In this chapter we reviewed different ‘fundamental inadequacies’ of monetary valuation methods. We concluded that the principle inadequacies relate in some way or another to lacking recognition of context. We presented Kallis et al. (2013) Guiding Framework for monetary valuation and applied it to four examples from Oslo found in Barton et al. (2015). Based on this test, we suggest that criteria for applying monetary valuation should include the decision-support context for which monetary valuation is intended (Gomez-Baggethun and Barton, 2013). We argue that Kallis et al. (2013) criteria were developed for decision contexts where valuation is being used to inform instrument design and economic liability, but lack a discussion of monetary valuation for use in priority-setting, accounting and advocacy. We also argue that paying more attention to the required accuracy and reliability of monetary valuation, conditional on decision-support context, spatial scale and resolution, will go some way to providing further guidance on when monetary valuation can meet the ‘environmental additionality’ criteria.

Whether monetary valuation methods are fit-for-purpose will depend in large part on how much context specificity is required by decision-makers. Whether we can afford to obtain the information or not, there is a lower limit to the information value of a valuation method when information costs exceed the net benefits of the decision under scrutiny. This lower limit must be real in both a temporal and spatial sense, although there is little research to support where it lies (it will depend on the context…).

An underlying premise of Kallis et al. (2013) framework is that monetary valuation is the dominant discourse in the place where the method is being applied, and perhaps also that its influence is on the increase. Barton et al.(2015) argue that this is not the case in Norwegian nature management policy, nor in municipal planning in Oslo. While there is considerable interest from authorities in whether monetary valuation of ES can make a difference in planning and operational decisions at municipal level, it is difficult to argue that monetary valuation has or will soon gain a hegemony in public discourse.

To do or not to do monetary valuation is a question of a method’s information value relative to the reliability requirements of different decision-support contexts. Just as monetary values are conditioned by institutional and socio-cultural context, so are the answers to when valuation is desirable.

Acknowledgements

The research was funded by the EU FP7 Project OpenNESS (contract no.308428).

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Max-Neef, M.A., 1992. ‘Development and human needs’, in P. Ekins and M.A. Max-Neef (eds), Real-life Economics: Understanding Wealth Creation, London, UK: Routledge, 197–213.

OsloKommune (2015). Oslo mot 2013. Kommuneplan 2015. Samfunnsdel og bytuviklingsstrategi. Del 1 15.04.15.

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Randrup, T.B. (2005). Development of a Danish model for plant appraisal. Journal of Arboriculture, 31(3), 114-123.

Reinvang, R., Barton, D.N., and Often, A. (2014). Verdi av urbane økosystemtjenester: Fire eksempler fra Oslo. Rapport nummer 2014/46. Vista Analyse A/S, NINA.

Schroter, M., Barton, D.N., Remme, R.P., and Hein, L. (2014). Accounting for capacity and flow of ecosystem services: A conceptual model and a case study for Telemark, Norway. Ecol Indic, 36, 539-551.

Schröter, M., Remme, R.P., Sumarga, E., Barton, D.N., and Hein, L. (2015). Lessons learned for spatial modelling of ecosystem services in support of ecosystem accounting. Ecosystem Serviecs, 13, 64–69.

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Synnovate (2011). Kommentarrapport. Undersøkelse om bruk av Oslomarka 10.-17. september 2011. Utarbeidet for Oslo Kommune, Bymiljøetaten av Erik Dalen, Synnovate 6. oktober 2011.

Sælen, H., and Ericson, T. (2013). The recreational value of different winter conditions in Oslo forests: A choice experiment. J Environ Manage, 131, 426-434.

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87

A multi-scale assessment of regulating ecosystem

services in Barcelona

Francesc BaróInstitute of Environmental Science and Technology (ICTA)

Universitat Autònoma de Barcelona (UAB)1

Abstract

Regulating ecosystem services (RES) such as climate regulation, air quality regulation or runoff mitigation can play a major role in enhancing human well-being and quality of life in urban areas. However, RES are often overlooked in urban decision making because their impact on human well-being can be difficult to measure, which causes a lack of reliable economic and biophysical RES assessments. In this paper, the results of two biophysical RES assessments (focusing on climate regulation and air quality regulation) carried out in the urban area of Barcelona are analyzed from a multi-scale perspective. The first is based on the application of the i-Tree Eco model in the municipality of Barcelona (local scale) whereas the second uses different spatial models to quantify and map the two RES in the Barcelona metropolitan region (regional scale). This multi-scale assessment considers both the supply of and demand for the RES along the urban-rural gradient and contributes to the integration of the ecosystem services framework into urban management and planning.

Keywords

Regulating ecosystem services; air quality regulation; carbon sequestration; Barcelona; Multi-scale assessment; urban management and planning.

1 [email protected]

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88 Ecosystem Services: concepts, methodologies and in truments for research and applied use

1. Introduction

Regulating ecosystem services (RES) can play a major role in enhancing human well-being and quality of life in urban areas. First, RES can contribute to climate change mitigation and adaptation strategies (e.g., climate regulation through carbon sequestration and storage, moderation of extreme events such as heat waves or flooding) hence increasing resilience and adaptive capacity in cities (Gómez-Baggethun and Barton, 2013). Second, RES can directly contribute to improve environmental quality in cities (e.g., air pollution removal, noise reduction) hence having a positive impact on citizens’ mental and physical health (Goméz-Baggethun et al., 2013; Baró et al., 2015). Third, RES often have a significant influence on the capacity to provide other ecosystem services (e.g., pollination is necessary to many edible vegetables). Most part of RES must be provided at the same location or close to beneficiaries because their delivery depends on proximity (an exception is climate regulation through carbon sequestration and storage) (Costanza et al., 2008; Syrbe and Walz, 2012). Therefore most RES need to be generated in cities or periurban areas so that urban population can benefit from their flow. However, RES are often overlooked in urban policy making and management because their impact on human well-being can be difficult to measure, which causes a lack of reliable economic and biophysical RES assessments (Villamagna et al., 2013).

The general aim of this paper is to assess both the provision of and the demand for RES in urban areas using a multi-scale approach (local and regional) in order to support urban policy and planning. Multi-scale assessments of ES offer a number of advantages compared to traditional single-scale evaluations. For example, they allow for the “validation of larger-scale conclusions by smaller scale assessments and create a context at larger scales for findings at smaller scales” (Scholes et al., 2013, Box 2). The assessment is applied to the urban area of Barcelona, Spain, covering two important RES in urban areas: air quality regulation (through air pollution removal) and climate regulation (through carbon sequestration and storage). It partially builds on two previous assessments carried out at the local scale (municipality of Barcelona; Baró et al., 2014) and regional scale (Barcelona metropolitan region; Baró et al., forthcoming). The research questions that the paper aims to answer include the following: Can the expected mismatch between RES provision and demand at the city scale be reconciled at the metropolitan (regional) level? Does the city scale assessment uncover RES provision potentially overlooked in the metropolitan assessment?

2. Case study and scales

The spatial scope of the assessment encompasses the Barcelona metropolitan region (BMR), North-East of Spain (Fig. 1). The BMR (5.04 million inhabitants in a total area of 3,244 km2, Statistical Institute of Catalonia2, year 2013) embeds 164 municipalities

2 www.idescat.cat

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89A multi-scale assessment of regulating ecosystem services in Barcelona

and seven counties, but its urban core is mainly constituted by the municipality of Barcelona (1.62 million inhabitants in an area of 101.2 km2, Barcelona City Council Statistical Yearbook 2013). Despite the BMR is one of the most densely populated urban regions in Europe (1,554 inhabitants per km2), it still contains a rich variety of natural habitats of high ecological value, including Mediterranean forests (1,184.6 km2; 36.5%) and shrub land (448.6 km2; 13.8%), extensive agro-systems (654.5 km2; 20.2%) including a substantial share of vineyard, and inland water bodies (24.1 km2; 0.7%) (see Fig. 1). The municipality of Barcelona is the second largest city in Spain and one of the most densely populated cities in Europe (16,008 inhabitants per km2). The total green space within the municipality of Barcelona (including urban parks, periurban forests and other green land covers) amounts to 28.9 km2 representing 28.6 % of the municipal area and a ratio of 17.9 m2 per inhabitant (Barcelona City Council Statistical Yearbook 2013, see Fig. 1). This ratio is very low in contrast to other European cities—especially in northern countries—where green space amounts to up to 300 m2 per inhabitant (Fuller and Gaston, 2009). Nonetheless, these low levels of green space are partly counterbalanced by the high number of single street trees, accounting for 162,816 specimens in 2013, a ratio of 100.4 street trees per 1000 inhabitants. This ratio is relatively high compared to other urban areas in Europe; mostly ranges between 50 and 80 street trees per 1000 inhabitants (Pauleit et al., 2002).

At both scales, there are strategic planning instruments intended to maintain and enhance ecosystems and biodiversity. At the regional level, the ‘Territorial Metropolitan Plan of Barcelona’ (PTMB, 2010) was approved in 2010 by the Government of Catalonia and protects about 74.1% of the BMR area from urbanization. At the local level, the City Council of Barcelona aims to boost the development of green infrastructure and to enhance urban biodiversity through the ‘Barcelona Green Infrastructure and Biodiversity Plan to 2020’ approved in 2013 (Barcelona City Council, 2013).

Abatement of air pollution is still a pressing challenge in most major urban areas worldwide, especially in regard to dioxide nitrogen (NO2) and particulate matter (WHO, 2014). The harmful impacts of air pollution on human health are consistently supported by scientific evidence (EEA, 2014). The city of Barcelona and other urban areas in the BMR have repeatedly exceeded the EU limit values for average annual concentrations of NO2 (40 µg/m3) and particles with diameter of 10 micrometers or less (PM10) in the last decade. Vegetation in urban landscapes can improve air quality by removing pollutants from the atmosphere, mainly through leaf stomata uptake and interception of airborne particles (Nowak et al., 2006).

Approximately 80% of worldwide energy consumption and greenhouse gas (GHG) emissions are associated with urban activities (Hoornweg et al., 2011). The European Commission launched the ‘Covenant of Mayors’ in 2008 involving local authorities, voluntarily committing themselves to implement more sustainable energy policies within their territories by reducing GHG emissions at the local level by at least 20 % until 2020. The City Council of Barcelona signed the ‘Covenant of Mayors’, committing to reduce by 23 % GHG emissions only derived from services and activities directly

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90 Ecosystem Services: concepts, methodologies and in truments for research and applied use

managed by the City Council. Other municipalities in the BMR have also set similar reduction targets. Urban and metropolitan forests can contribute to offset urban carbon emissions through carbon sequestration and storage (Escobedo et al., 2010).

Fig. 1. Main land covers in the Barcelona Metropolitan Region (BMR) and the municipality of Barcelona. Source: own elaboration based

on the spatial dataset “Habitats of Catalonia” (year 2013).

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91A multi-scale assessment of regulating ecosystem services in Barcelona

3. Data and methods

The multi-scale assessment was based on the definition and quantification of RES supply and demand indicators using different models and data sources (see an overview of RES indicators in Table 1 and Baró et al., 2014; forthcoming). The i-Tree Eco model was used to quantify RES provision within the municipality of Barcelona. The i-Tree Eco model has been used in more than 50 cities across the world, especially in the United States, to assess urban forest structure and ecosystem services (Nowak et al., 2008). i-Tree Eco protocols (Nowak et al., 2008; i-Tree User’s Manual, 2008) were followed to collect field data on urban forest structure within the municipality. See Baró et al. (2014) for a detailed description of the sample design and data collection process. The actual provision of air quality regulation was quantified on the basis of field data, air pollution concentration (NO2), and meteorological data. Fundamentally, i-Tree Eco model estimates dry deposition of air pollutants (i.e., pollution removal during non-precipitation periods), which takes place in urban trees and shrub masses. The (removed) pollutant f lux (F; in g/m2 s) is calculated as the product of deposition velocity (Vd; in m/s) and the pollutant concentration (C; in g/m3). Deposition velocity is a factor computed from various resistance components (for more details see Baldocchi et al., 1987; Nowak et al., 2008). The climate regulation provision was calculated based on the modeling results of net carbon sequestration. i-Tree Eco model calculates the biomass for each measured tree using allometric equations from the literature. Biomass estimates are combined with base growth rates, based on length of growing season, tree condition, and tree competition, to derive annual biophysical accounts for carbon sequestration. Several assumptions and adjustments are considered in the modeling process (for more details, see Nowak and Crane, 2002; Nowak et al., 2008). The demand for these two RES at the local scale was determined based on data of NO2 pollution levels (considering both municipal emissions and background pollution) and local GHG emissions. In both cases, data was extracted from the estimations of the Energy, Climate Change and Air Quality Plan of Barcelona corresponding to year 2008 (PECQ, 2011).

The quantification of RES supply and demand at the regional scale (BMR) followed a spatially explicit approach. For air quality regulation, the methodological framework provided by the Ecosystem Services Mapping tool (ESTIMAP) (Zulian et al., 2014) was used. ESTIMAP is a collection of spatial models for ES assessment originally developed to support policies at European scale such as the EU Biodiversity strategy (Maes et al., 2014). The NO2 removal indicator was mapped based on the spatial distribution of NO2 annual average concentrations. Concentrations of NO2 were estimated using Land Use Regression (LUR) models, a computation approach widely used for assessing air pollution at different scales (e.g., Briggs et al., 1997; Hoek et al.; 2008; Beelen et al. 2013). The LUR model was built using NO2 concentration

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92 Ecosystem Services: concepts, methodologies and in truments for research and applied use

measurements (year 2013) from the operational monitoring stations located in the BMR (n=40) as dependent variable, and a set of spatial predictor parameters (independent variables) related to land cover type, geomorphology, climate, and population (see Baró et al., forthcoming for more details), that were considered to be the most relevant for the distribution of NO2 concentrations. Annual NO2 removal was estimated as the total pollution removal f lux in the areas covered by vegetation, calculated as the product of NO2 concentration and deposition velocity maps (Nowak et al., 2006). Deposition velocity (Vd) was estimated as a linear function of wind speed at 10 m height and land cover type following Pistocchi et al. (2010). Air quality regulation demand was mapped based on NO2 concentration levels and housing population density. A cross-tabulation was carried out between both variables (see Baró et al., forthcoming for more details). The map of (un)satisfied demand (or balance between ES supply and demand) for air quality regulation was generated from the population living in areas where annual mean concentrations exceed the EU limit value (40 µg/m3 for NO2, see Baró et al., 2015). For climate regulation, the provision of carbon sequestration and storage was estimated based on above-ground tree biomass maps from Pino (2007). The author used data from two Spanish forest inventories (IFN2 & IFN3) and applied a LUR model considering various spatial predictors. Tree biomass was multiplied by 0.5 to estimate the stored carbon in trees and carbon sequestration was estimated from tree biomass net growth between the two inventories. Demand for climate regulation was based on carbon emissions estimated for each BMR municipality. Estimates were collected from municipal Sustainable Energy Action Plans (SEAPs) corresponding to the year 2012 by the Barcelona Regional Council. Unfortunately, total values do not include emissions from some relevant sectors such as industry or agriculture, so they should be considered a first order estimate. The balance between carbon emissions and carbon sequestration was carried out at the municipal level (sequestration rate minus emissions rate).

Rural-urban gradients have been used to analyze ecological patterns and processes in urban landscapes, including the consideration of ES indicators in some recent studies (Kroll et al., 2012; Larondelle and Haase, 2013). Following these approaches, rural-urban gradients of the supply, demand and (un)satisfied demand (balance) of the two RES considered here were computed using the resulting maps as described above. A 50-km concentric buffer with 1-km intervals was created around the city center of Barcelona (Catalunya square), covering almost all the BMR area. For each concentric ring, the average RES reclassified value (0-5 range) was calculated omitting null values.

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93A multi-scale assessment of regulating ecosystem services in Barcelona

Tabl

e 1.

Ove

rvie

w o

f RES

indi

cato

rs c

alcu

late

d fo

r the

mul

ti-sc

ale

asse

ssm

ent.

RES

Scal

eES

Indi

cato

rQ

uant

ific

atio

n un

itIn

put d

ata

Qua

ntif

icat

ion

met

hod

Mai

n re

fere

nces

Clim

ate

regu

latio

n

Regi

onal

(BM

R)C

arbo

n se

ques

trat

ion

(Act

ual s

uppl

y)kg

C / h

a ye

ar

Nat

iona

l for

est i

nven

tory

da

ta (I

FN2

& IF

N3)

Land

use

dat

a an

d ot

her

spat

ial p

redi

ctor

s.

Land

use

regr

essio

n m

odel

ing

Pino

(200

7)

Loca

l (B

arce

lona

m

unic

ipal

ity)

Car

bon

sequ

estr

atio

n (A

ctua

l sup

ply)

t C/ y

ear

Fiel

d da

taA

llom

etri

c equ

atio

ns fr

om

liter

atur

ei-T

ree

Eco

Mod

elBa

ró e

t al.

(201

4)

Regi

onal

(BM

R)C

arbo

n em

issio

ns

(Dem

and)

t C /

mun

icip

ality

ye

arC

arbo

n em

issio

ns p

er se

ctor

an

d m

unic

ipal

ity (y

ear 2

012)

Vari

ous (

see

refe

renc

es)

Barc

elon

a Re

gion

al C

ounc

il in

tern

al d

ata

Loca

l (B

arce

lona

m

unic

ipal

ity)

Car

bon

emis

sions

(D

eman

d)t C

/ ye

arC

arbo

n em

issio

ns p

er se

ctor

in

Bar

celo

na (y

ear 2

008)

Vari

ous (

see

refe

renc

es)

Baró

et a

l. (2

014)

PEC

Q (2

011)

Air

qu

ality

re

gula

tion

Regi

onal

(BM

R)N

O2 re

mov

al fl

ux(A

ctua

l sup

ply)

kg N

O2 / h

a ye

ar

Air

qua

lity

data

from

BM

R m

onito

ring

(yea

r 201

3)Va

riou

s spa

tial p

redi

ctor

s (s

ee m

ain

refe

renc

es)

Land

use

regr

essio

n m

odel

ing

(EST

IMA

P)Zu

lian

et a

l. (2

014)

Baró

et a

l. (fo

rthc

omin

g)

Loca

l (B

arce

lona

m

unic

ipal

ity)

NO

2 rem

oval

flux

(Act

ual s

uppl

y)t N

O2 /

year

Fiel

d da

taA

ir q

ualit

y da

ta fr

om

Barc

elon

a m

onito

ring

st

atio

ns (y

ear 2

008)

Met

eoro

logi

cal d

ata

(yea

r 20

08)

i-Tre

e Ec

o M

odel

Baró

et a

l. (2

014)

Regi

onal

(BM

R)Po

pula

tion

expo

sure

to

NO

2 con

cent

ratio

n(D

eman

d )

Dim

ensio

nles

s va

lue

and

inha

bita

nts

expo

sed/

ha

Air

qua

lity

data

from

BM

R m

onito

ring

(yea

r 201

3)Va

riou

s spa

tial p

redi

ctor

s (s

ee m

ain

refe

renc

es)

Popu

latio

n de

nsity

gri

d (y

ear

2011

)

Cro

ss-t

abul

atin

g an

d re

gres

sion

mod

elin

g

Beel

en e

t al.

(200

9)Zu

lian

et a

l. (2

014)

Baró

et a

l. (fo

rthc

omin

g)

Loca

l (B

arce

lona

m

unic

ipal

ity)

NO

2 em

issio

ns

and

conc

entr

atio

n (D

eman

d)t /

year

NO

2 em

issio

ns in

Bar

celo

na

and

back

grou

nd p

ollu

tion

impa

ct (y

ear 2

008)

Empi

rica

l dat

aBa

ró e

t al.

(201

4)PE

CQ

(201

1)

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94 Ecosystem Services: concepts, methodologies and in truments for research and applied use

4. Results

At the local level (Barcelona municipality) results show that the contribution of urban forests to climate change mitigation is very low, accounting for 0.47 % of the overall city-based GHG emissions. Considering only GHG emissions derived from the sectors that are directly managed by the City Council (reference emissions to meet ‘‘Covenant of Mayors’’ 23 % reduction target and representing 2.10 % of the total emissions) the contribution of urban forest is still modest but yet substantial, accounting for 22.55 % of the emissions (Table 2). Contribution of urban forests to air quality (NO2) is low considering only city-based emissions (0.52%) and also considering local emissions and the impact of background pollution (0.43%) (Table 2).

Table 2. Contribution of urban forests on air quality and climate change mitigation (municipality of Barcelona, year 2008). * indicates CO2eq emissions from services and activities directly managed by the City Council (‘‘Covenant of Mayors’’ policy target baseline emissions).

RES ES provision ES demand ES balance (%)

Biophysical value(t / year)

City-based emissions(t / year)

Background pollution impact (%)

City-based emissions

City-based emissions & background pollution

Climate regulation (CO2eq)

19,036 4,053,76684,403* N/A 0.47

22.55* N/A

Air quality regulation(NO2)

54.59 10,412.94 18.70 0.52 0.43

Supply, demand and (un)satisfied demand distribution maps at the regional scale (BMR) for the two RES are shown in Fig. 2 (climate regulation) and in Baró et al. (forthcoming) (air quality regulation). Both RES show similar spatial patterns. Supply values are especially relevant in periurban forest areas (such as the mountain range of Collserola) and other natural sites located in the hinterland. However, NO2 removal in natural areas such as in the Montseny mountain massif is relatively low in some zones because the driver of actual use is almost absent (pollutant concentration). The lowest supply values for both RES are located in urban and agricultural land. As expected, the municipality of Barcelona and adjacent middle-size cities show the highest demand values in the BMR for both ES analyzed. As observed in the local scale assessment, this urban agglomeration is characterized by a compact urban form, very high population density and a relative small share of inner green areas. The other middle-sized municipalities, located both along the coastline and hinterland, show mostly middle to low demand values. Smaller towns and sprawling urban areas mostly show very low to no relevant demand values in both cases. Finally, results clearly show that unsatisfied demand (when demand is higher than supply) is circumscribed to

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95A multi-scale assessment of regulating ecosystem services in Barcelona

the core city of Barcelona and several middle-size cities. Unsatisfied demand for climate regulation shows nearly the same spatial pattern as for demand (carbon emissions are higher than carbon offsets provided by the local vegetation in almost all municipalities) whereas unsatisfied demand for air quality regulation is principally limited to the urban areas surrounding the main roads and streets of Barcelona and adjacent cities where NO2 concentration is highest (above 40 µg/m3).

The urban-rural gradients of climate regulation (Fig. 2) and air quality regulation (Baró et al., forthcoming) for the BMR illustrate graphically the spatial patterns shown in the maps. Supply gradients show the impact of periurban forests such as Collserola (km 6 – 12 approximately) followed by a steady flat trend without any substantial increase (except for climate regulation at the outskirts of the BMR). Demand and unsatisfied demand gradients are also quite similar for both ES, showing highest values in the core city area (km 1-5) followed by a decreasing trend as the distance to the city increases.

5. Discussion and conclusion

This multi-scale assessment suggests that, in terms of urban planning and policy, strategies intended to reconcile RES supply and demand both at the local and regional level should mainly focus on drivers of demand (i.e., air pollution concentrations and CO2 emissions). Air quality improvements and climate change mitigation by urban vegetation are relatively low at the urban core, suggesting a limited effectiveness to address RES mismatches by increasing ES supply (e.g., implementing tree-planting programs or selecting trees with high air pollution removal or carbon sequestration capacity). Moreover, factors such as vegetation configuration and climate conditions can limit the ability of vegetation to remove air pollutants, especially at the patch scale such as in street canyons (Vos et al., 2013). Therefore, policy interventions should focus on reducing and limiting traffic in certain areas, increasing public transport, incentivizing the use of low-emitting vehicles (e.g., bicycles and electric vehicles), and enhancing planning towards shorter commuting needs. This could be done by a combination of prescriptive policy regulations (e.g., creating clean air areas through traffic restrictions) and economic incentives (e.g., implementing a tax on NO2 and carbon emissions from private transport and subsidizing public transport and public biking system). The assessment also highlights the importance of the periurban green areas (such as Collserola) in terms of RES supply. However, it is important to note that the provision of air quality regulation depends on proximity to beneficiaries, whereas climate regulation does not (Costanza, 2008). Thus, the enhancement of air quality regulation supply should be mainly dealt with at the local level (especially in cities such as Barcelona) whereas climate regulation could be dealt at the regional or higher level (except if specific municipal CO2 reduction targets are considered which can be assumed as local demands for this RES; see Baró et al., 2015).

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96 Ecosystem Services: concepts, methodologies and in truments for research and applied use

ES modelling entails some limitations and caveats that should be taken into account when analyzing its outcomes. First, an important limitation of i-Tree Eco and most dry deposition models is the level of uncertainty involved in the quantification of the air pollution removal rates due to the complexity of this process (Pataki et al., 2011). For instance, some sources of uncertainty include non-homogeneity in spatial distribution of air pollutants, particle re-suspension rates, transpiration rates or soil moisture status (Manning 2008). Estimation errors in climate regulation service values include the uncertainty from using biomass equations and conversion factors as well as measurement errors (Nowak et al., 2008). Estimates of carbon sequestration and storage also include uncertainties from factors such as urban forests maintenance

Fig. 2. Supply, demand and unsatisfied demand maps and urban rural gradient (50km) for the ES climate regulation (carbon

sequestration) in the BMR. See Table 1 for main data sources.

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97A multi-scale assessment of regulating ecosystem services in Barcelona

(e.g., intensity of pruning) or logging, wildfire effects, tree decay, or restricted rooting volumes, which are not accounted for in the models estimations (Nowak et al., 2008; Pataki et al., 2011; Pino 2007). Another issue not considered in the models relates to the ecological thresholds or tipping points (Andersen et al., 2009). An ecological threshold can be defined as a “point at which an (ecological) system experiences a qualitative change, mostly in an abrupt and discontinuous way” (Jax, 2014). It is often very difficult to determine when and under what conditions or pressures ecosystems experience thresholds which can affect their ability to provide ES (Gómez-Baggethun et al., 2011). In the case of air quality regulation, high pollutant concentrations can severely damage vegetation or lead to stomatal closure, reducing air pollution removal capacity and consequently f low (Robinson et al., 1998; Escobedo and Nowak, 2009).

In conclusion, this paper argues that both the regional and local scales should be considered in RES assessments in order to comprehensively support urban planning and policy (Scholes et al., 2013). Local assessments can take into account small RES providing areas such as street trees which are usually overlooked in regional assessments. These areas can have potentially a relevant impact in terms of RES supply at the local level such as in the case of Barcelona. The relevant scale of RES supply and demand also arises from this research (Geijzendorffer and Roche, 2014). Our results reflect that demand is highest in urban cores, but the actual use of the air quality regulation highly depends on the proximity between ES providing areas and benefiting areas, whereas it does not for climate regulation (Syrbe and Walz, 2012). This issue calls for a strong institutional coordination between local and regional authorities dealing with urban and environmental policy and for the harmonization of planning instruments at different scales.

Acknowledgements

This paper builds on various research projects focusing on urban ecosystem services. I thank the following people for contributing directly or indirectly to this work: Erik Gómez-Baggethun; Johannes Langemeyer; Lydia Chaparro; David J. Nowak; Jaume Terradas; Ignacio Palomo; Grazia Zulian; Pilar Vizcaino; Dagmar Haase; Coloma Rull; Margarita Parès; Montserrat Rivero; and Carles Castell.

This research was partially funded by the following organizations: ERA-Net BiodivERsA network through the Spanish Ministry of Economy and Competitiveness project ‘URBES’ (code PRI–PIMBDV-2011-1179); 7th Framework Program of the European Commission project ‘OpenNESS’ (code 308428); Fundación Iberdrola España (Ayudas a la Investigación en Energía y Medio Ambiene 2015); Barcelona City Council (Ajuntament de Barcelona); and Barcelona Regional Council (Diputació de Barcelona).

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101

Managing the side effects of bioenergy: a review of the impacts of biomass feedstock production

on ecosystem services

Elena Gissi,1,* Mattias Gaglio,1,2 Matelda Reho1

Abstract

Biomass based energy sources (BBES) are an Ecosystem Service (ES), which can contribute to achieve EU 2020 targets and to ensure energy security. Their effects on other ES can be different according to biomass typologies, as energy crops and residuals.

An ES approach, based on the cascade model, provides a framework in which trade-offs can be identified and assessed, taking into consideration biological interdependencies and management options.

The present review shows that energy crops affect several ecosystem functions, particularly related to soil and water, while trade-offs induced by residual biomass can be limited with proper management. However, efficiency constraints concerning their low density affect their exploitation potential.

This study highlights that the ES approach can provide a suitable tool for decision makers, with respect to the biomass feedstock chain, whose effects on ecosystems are often underestimated.

Keywords

Biomass Based Energy Sources, Ecosystem services, Trade-off, Crop and forest residues, Energy crops.

1 Department of Design and Planning in Complex Environment, University Iuav of Venice, Italy.2 Department of Life Sciences and Biotechnology, Section of Evolutionary Biology,

University of Ferrara, Via Borsari 46, 44121 Ferrara, Italy. email: [email protected]

* email of corresponding author: [email protected]

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1. Introduction

The theoretical framework of Ecosystem services (ES) is based on the understanding of ecological functioning of ecosystems, and their relationship with the socio-cultural context in which ES are estimated and benefited. According to the cascade model, proposed by Haines-Young and Postchin (2010), ES represent the link between ecological functions, depending in turn from several biophysical structures, and socio-cultural context (Haines-Young et al., 2012). ES depend on ecosystem functions; each ecosystem delivers several functions at once and each function can provide one or more ES. Consequently, different ES are biologically related with each other rather than mutually independent and their relationships can be complex (Bennet et al., 2009). Thus, some ES appear together repeatedly, producing bundles of ES (Raudsepp-Hearne et al., 2010).

Because of their complex interactions, the exploitation of an ES can result in change of status of other ES. That happens because two or more ES derive from the same ecological function, or from more ecological functions, which are related to a common driver (Bennet et al., 2009).

Moreover, economic interdependencies can strongly influence the delivery of related ES (Abler, 2004; Briner et al., 2013), as the results of drivers that orient alternative choices with respect to ES. In fact, ES can show relationships that change according to different management scenarios (Rodríguez et al., 2006; Briner et al., 2013), as well as to different planning options (Geneletti, 2013), giving place to alternative land use patterns.

Concerning ES interactions, it is possible to define 3 categories of different resulting situations: i) trade-off, i.e. the delivery of an ES results in a degradation of another (Bennett et al., 2009; Raudsepp-Hearne et al., 2010); ii) synergy, i.e. the delivery of an ES results in a increasing of another (Bennett et al., 2009; Briner et al., 2013); iii) neutrality, i.e. the delivery of an ES does not affect the level of others, they derive from different functions which are not related to each other, nor they are dependent from a same driver.

The aim of this study is to revise the trade-offs between Biomass based energy sources (BBES) feedstock and ES provided by agro-ecosystems, considering the relationship between several ES depending from related functions.

The knowledge and proper evaluations of the interrelations between ES provision and related impacts is necessary to scientifically inform decision making in energy policy.

The authors review the scientific literature, in order to investigate the relations between concurring ES and feedstock provision for BBES, rather than focusing on a single effect or impact per each ES.

Results of trade-offs have been described in order to explore management solutions that minimize the negative effects between ES, either mitigating or eliminating impacts that affect ecosystem functions.

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103Managing the side effects of bioenergy [...]

The study aims at understanding: i) which ES trade-offs are involved along BBES feedstock production; ii) how to maintain the supply of BBES without harming the capacity of ecosystems to provide other ES.

2. Materials and Method

BBES are inter-related with other ES because of biological inter-dependencies between ecological functions. Biomass is produced by agro–and forest ecosystems (the latter with respect to residues harvesting). Those ecosystems are exploited to obtain a surplus production of biomass, based on the ecosystem functioning of primary production.

Feedstock production implies a complex potential trade-off of other ES provision. Recent research has focused mostly on trade-off between BBES and food production (Johansson and Azar, 2007; Koh and Ghazoul, 2008; Ajanovic, 2011; Nonhebel and Kastner, 2011). Less attention has been paid to trade-off with non-market services, such as regulating, supporting and cultural services that are not often accounted for, when compared to the monetary benefits of provisioning services, as food, fiber or biomass (Meehan et al., 2013).

According to the cascade approach (HainesYoung and Postchin, 2010), the functions providing several ES have been related to study reciprocal behavior about the capacity to provide ES and BBES. Primary productivity has been analyzed with respect to i) soil functionality and water regulation, as regulating services; ii) biodiversity provision and iii) cultural services derived from landscape. They constitute the BBES-related bundle.

This study has been structured according to the different nature and origin of biomass (annual crops, perennial crops, agricultural residues, forest residues) and, consequently, different impacts and potentials in feedstock production. Finally, the trade-offs have been evaluated according to the management practices and their capacity to control the trade-off severity.

3. Results

3.1 Biomass primary production and soil functionalitySoil has several important ecological functions: substrate for vegetation, microclimate regulation, absorption of rainfall, water and nutrients availability for plants and animals, habitat for microorganisms with key roles in the ecosystems (Blanco-Canqui and Lal, 2009; Powlson et al., 2011).

Primary production is related to soil productive capacity, and therefore to its physical, chemical and biological characteristics. It can be reduced in consequence of a decrease of soil quality (Grigal 2000). Soil Organic Carbon (SOC) is related

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to primary production and, at the same time, it directly provides the ES ‘climate regulation’ through carbon storage in soil. Primary production is also responsible for other ES, such as soil protection and soil regeneration through maintaining soil quality (Blanco-Canqui, 2010; Janowiak and Webster, 2010). In this case, trade-off can emerge as several ES depend on the same ecological function, which can support the delivery of some ES against the decrease of others.

Among bioenergy crops, annual and perennial crops have different effects on soil functionality.

The cultivation of annual crops, as for instance corn, led to a decrease of SOC levels (Blanco-Canqui 2010), conversely, perennial crops can increase the organic matter of the soil (Brandao et al., 2010)

Moreover, perennial crops guarantee a continue land coverage, especially after the establishment period (EEA, 2006), providing an erosion prevention service. Contrarily, annual crops do not provide land coverage all year long, leading to erosion phenomena, which causes further depletion of SOC levels. Erosion processes are caused by water flow in slope areas, and by wind in flatted zones with intensive agriculture practices. They depend on slope and rainfall, with the latter expected to undergo increased peak levels in the future, according to general projections about climate change (IPCC, 2013).

Another driver that affects soil functionality concerns the management practices. Mechanization of agricultural activities, both for energy and food purposes, leads to an increase of soil erosion and degradation (EEA, 2005). This problem is particularly critical for annual crops, where intensive agriculture occurs.

With respect to residues, their removal from forests harms soil functionality, since it prevents the natural accumulation of SOC (Brandao et al., 2010) and it causes nutrient loss (Pimentel et al., 1981).

In addition, agricultural residual biomass plays an important role in soil regeneration and erosion prevention. Their removal affects soil stability and fertility, leading to a decrease of primary production, with magnitude those changes according to the different soil texture and mineralogy (Blanco-Canqui and Lal, 2009). Moreover, residual removal causes a loss of habitat and trophic sources for macro and micro-organisms, some of whom are key species due to their role in soil structural development nutrient cycle and microclimate regulation (Karlen et al., 1994).

3.2 Biomass primary production and water regulationThe impact of energy crops on water is due both to water consumption (Ballarin et al., 2008) and water quality decrease.

Water quality regulation and water provision rely on the self-depuration capacity of water bodies. This function is related to primary production through a common driver, agricultural intensive practices, due to the use of fertilizers and pesticides, which enhance biomass production maximizing yield and decreasing water quality.

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Furthermore, water quality is affected by high level of Dissolved Organic Carbon (DOC) due to organic matter runoff from agriculture fields, which increases water treatment costs and may lead to health issues associated with disinfection by products generated during the treatment process (Smith et al., 2013).

As described by Smith et al. (2013), the variables involved and the dynamics of water regulation are site-specific, according to regional and climatic context; besides, they vary in time. Collection activities adversely affect water quality immediately after harvest, and the capacity of the system to recover the precedent conditions needs a time range between 2 and 5 years (Aust and Blinn, 2004). Moreover, indirect impacts are due also to the construction of infrastructures for biomass transportation, which are recognized as the main cause of soil erosion and therefore water pollution (Grigal, 2000).

From the side of residues, their removal reduces provision on water services due to loss of sediments in runoff, inducing changes in the water cycle, because of the acceleration of evaporation and infiltration rates (Blanco-Canqui and Lal 2009).

3.3 Biomass primary production and biodiversity provisionPressures on biodiversity depend on different drivers as land use and agricultural practices performed in order to obtain biomass for energy purposes, affecting the availability and quality of habitats (Janowiak and Webster, 2010).

Converting forested areas into woody plantations (e.g. short rotation forestry of poplar (Populus spp.) or willows (Salix spp.)), or intensive crops cause strong pressures on habitats and species (Firbank et al., 2008). Contrarily, cultivation of energy crops in previously degraded lands, with absent or very low natural grade, can deliver a function of new habitat provision, leading to a positive “biodiversity balance” (Anderson and Fergusson, 2006).

In terms of biodiversity, corn cultivation is the most impacting bioenergy crop among the present alternatives, although this is the most advanced bioenergy feedstock for commercial production in rural areas (Groom et al., 2008).

Biodiversity concerns are site-specific. Lattimore et al. (2009) classify the effects on biodiversity at different levels: landscape, ecosystems, habitats, species and genes, specifying how risks and impacts vary according to the socio-geographic context. For instance, in developing countries the greatest risks arise from deforestation activities, uncontrolled timber harvesting and the replacement of tropical forest areas with intensive monoculture plantations.

At landscape level, bioenergy crops leads to modifications in habitat turnover (Firbank et al., 2008), as, for example, when new establishment of annual bioenergy crops causes new crops rotations, leading to a temporary loss of habitats for species.

Introducing genetically modified organisms, as most efficient energy crop species, can be a risk for ecosystems, and their effects are difficult to foresee, because of their ecological traits (Firbank et al., 2005).

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In forest ecosystems, presence of deadwood and slash provide habitat for rare species and increase habitat diversity (Vertek et al., 2011).

Plant species for energy scope are selected according to their yield, energy potential, tolerance to poor growing condition and with few resident pests; among them, many are non-native species, as described by Barney and Di Tommaso (2008). Alien species introduction can lead to a loss of several ecosystem functions by altering ecosystems balances outcompeting native species (Pejchar and Moony, 2009).

3.4 Biomass primary production and landscape cultural valuesAccording to the MEA (2005), cultural services are i) spiritual and religious values, ii) aesthetic values, and iii) recreation and ecotourism. Heterogeneity, as the function related to landscape spatial structure, has been related to landscape-based values, as cultural values and human appreciation (Dramstad et al., 2001; Tempesta, 2010), as well as to landscape preferences in agricultural landscapes (Dramstad et al., 2006).

Farmers’ preferences can be oriented to improve biomass primary production (and related monetary incomes) by adopting intensive monoculture suitable for energy purposes, leading to land use changes and landscape homogenization against diversity and heterogeneity, the latter recognized as main functions supporting landscape cultural values.

In surveys carried out in England and Wales, Upreti (2004) shows that local communities are particularly sensitive to landscape changes of rural areas as a consequence of the construction of biomass plants and the related supply chains. Indeed, efficient energy crops are likely to replace a more diversified landscape, composed from cultures of various species and variety, with negative effects on other ES such as those relating to biological control (Landis et al., 2008) and cultural values associated with the area. Residues harvesting avoids those land use change and the loss of landscape heterogeneity.

4. Discussion and conclusions

Trade-offs among BBES and ES depends on biomass feedstock type. No BBES is free of impacts on other ES (i.e. trade-offs), as well as other renewable energy sources exploitation, but different BBES can better fit in different site conditions.

Soil regeneration service is heavily impacted by annual crops, which affects soil quality, and is correlated with site conditions as slope and climate, as well as by management options related to intensive agriculture activity. A switch to perennial crops can play as synergy, supporting organic matter accumulation in soil and increasing carbon storage capacity. However, it can cause further water quality

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107Managing the side effects of bioenergy [...]

loss, due to massive use of fertilizers, resulting in a trade-off with the water regulation service.

Erosion prevention and soil regeneration are closely related services. Mitigation of erosion is possible by avoiding annual crops on steep land, or switching to perennial crops, which ensure constant land coverage. In order to support erosion prevention and soil regeneration services, as well as to maintain BBES supply, proper management options for residues can be adopted. Many forest management guidelines (Abbas et al., 2011) suggest leaving up to 30% of forest residues on soil surface and to distribute them more evenly, as well as to leave slash matter to promote natural regeneration from seed-bearing cones.

Furthermore, BBES produced from the vast use of energy crops result in trade-off with food production, as they are based on the same ecosystem function of primary productivity. This trade-off should be considered when supply chains from energy crops are planned, in order to avoid land use conflicts and livestock feed price increase (Koh and Ghazoul, 2008; Gissi et al., 2011). According to Hoogwijk et al. (2005), Fritsche et al. (2006) and Field et al. (2008), biomass crop systems should be allocated in degraded, abandoned or less productive lands, so to avoid land use conflicts.

Trade-offs between BBES and water quality regulation occur because of use of fertilizers to maximize the yield (Smith et al., 2013), which is necessary to perform good energy outputs in dedicated energy crops. Use of agricultural and forest residues can avoid severe trade-offs with water services, especially retaining a proper quantity of matter on site.

In this case, the trade-off between residues uptake and soil regulation depends on the management option negotiated between farmers and public authorities.

With respect to the water regulation service, buffer zones in agricultural land can significantly break down nutrients loads deriving from chemical fertilizers for energy crops, avoiding negative interaction between BBES and water quality regulation. Christen and Dalgaard (2012) show that a three-zone riparian buffer design can be suitable both to retain high percentages of nitrogen and phosphorous and to provide woody biomass. Furthermore, lands with high nutrients breakdown capacity should be more appropriate for energy crops establishment.

Also, trade-offs between BBES and biodiversity can be mitigated by using proper management options. Forest guidelines can be suitable to manage the impacts of biomass uptake on biodiversity in forest ecosystems. The latter should be avoided where endangered or rare species are present and during the bird nesting period, and it should be performed retaining leaves three foliage on site. Deadwood should be maintained on site in order to support forest biodiversity (Vertek et al., 2011).

Concerning BBES from agro-ecosystem, trade-offs with habitat and species diversity are restrained by the low natural degree of these environments. To maintain genetic diversity service, Pejchar and Mooney (2009) propose genotype-specific pre-

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introduction screening, consisting in risk analysis, climate-matching modeling, and ecological studies of fitness responses to various environmental scenarios. Finally, afforestation of degraded of rural areas for bioenergy purposes can have positive impacts on biodiversity, providing habitat for new species. Biodiversity is also important in delivering other ES (van Oudenhoven and de Groot, 2013). High levels of biodiversity support better performances in other ES provision, especially if compared with purely economic driven choices (Nelson et al., 2009).

Landscape issues significantly affect the social acceptability of BBES. Buchholz et al. (2009) fit the visual impacts and the diversity of cultures among the sustainability criteria for bioenergy systems, in agreement with findings from Vera and Langlois (2007) with particular reference to the need of long-term vision. In order to minimize the ecological footprint of energy crops, the land area needed to grow sufficient quantities of the feedstock should be minimized (Groom et al., 2008), controlling land use change. Mapping ecosystem services and their biological interdependencies and inter-variability can inform decision-making, anticipating possible dysfunctionalities in ES bundles, as joint products of a specific site (Burkhard et al., 2012). It aims at analyzing different benefits deriving from prioritizing some ES against others, considering that a scenario where all ES are maximized is unreal (Lester at al., 2013). The analysis of trade-offs based on ecosystem processes and functions is a precondition to support feedstock allocation and production for the delivery of BBES.

Acknowledgements

The present review have been performed within the research titled “Biomass supply chain and ecosystem services: operational criteria for the localization of power plants” which has been financed by Veneto Region (IT), under the Regional Operational Program of the European Social Fund (prog. code 2122/1/14/1686/2012), in collaboration with Guerrato S.r.l, Rovigo (IT).

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113

Socio-cultural values of urban ecosystem services

Johannes LangemeyerInstitute of Environmental Science and Technology (ICTA),

Universitat Autònoma de Barcelona (UAB), Spain Stockholm Resilience Centre,

Stockholm University, Sweden1

1. Introduction

We intuitively know that urban green spaces are important for human well-being in cities. However, the scientific evidence for the direct links between green spaces and human well-being remains poor (e.g., Pataki et al., 2011). The concept of urban green infrastructure provides a holistic and multi-functional understanding of interconnected green and blue spaces in cities (Ahern, 2011). Urban green infrastructure includes for example lakes, rivers, coastlines, wetlands, forests, single trees, parks and gardens. The ecosystem service approach has been introduced as a transdisciplinary framework to capture the contribution of ecosystems and urban green infrastructure to human well-being (TEEB, 2010; Pauleit et al., 2011). The emerging research field of urban ecology understands cities as coupled social-ecological systems (Niemelä et al., 2011; Pickett et al., 2013). From this perspective cities are seen as integrated parts of the earth’s larger ecosystems and green spaces nested within them (Langemeyer, 2015). Urban green spaces are often strongly shaped by urban planning and management, and ES from urban green spaces can thus be understood as co-produced by nature and humans (Andersson et al., 2014). Cities are usually described as location where the demand for ES exceeds their supply (e.g. Baró, in Nuss-Girona & Castañer, 2015). This puts special interest on maintaining, restoring and creating green spaces in cities, as green infrastructure, for the local supply with ecosystem services (Elmqvist et al., 2015).

Two important types of green spaces providing ecosystem services for the benefit of humans in cities are urban parks (Konijnendijk et al., 2013) and urban (horticulture) gardens (Langemeyer et al., 2015). The importance of urban parks, defined as “delineated open space areas, mostly dominated by vegetation and water, and generally reserved

1 [email protected]

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for public use” (Konijnendijk et al., 2013: 2), has been highlighted due to a number of benefits. Among those benefits are increased social cohesion, tourism, biodiversity, regulation of stormwater runoff, urban cooling as well as – to a limited extend – air pollution reduction, especially reduction of particulate matter (PM10) (Konijnendijk et al., 2013). Urban gardens have been critically important to cities throughout history for the production of food (e.g., Barthel and Isendahl, 2013). They further support pollinators (Jansson and Polasky, 2010), and provide refuge to a range of plants and animals (Pourias et al., 2015). Furthermore urban gardens have also been described for their importance as source of cultural ecosystem services, such as recreation and relaxation (e.g. Kaplan, 1973), environmental learning (e.g., Krasny and Tidball, 2009), the formation of sense of place and social cohesion (Tidball et al., 2014; Armstrong, 2000). For an overview of ecosystem services provided by urban gardens see Table 1.

This chapter provides a closer look at urban parks and urban horticulture gardens as sources of ecosystem services that provide multiple benefits to humans. More explicitly, I look at the values urban citizens attach to specific ecosystem services sustained by urban gardens and urban parks in Barcelona, Spain.

Table 1. Ecosystem services from urban gardens (Langemeyer et al., 2015a)Category Ecosystem Service Key References

Provisioning services Food provision Orsini et al. (2014); Pourias et al. (2015); Keatinge et al. (2011)

Provision of medicinal and seasoning plants

Airriess and Clawson (1994)

Ornamental plants Dunnet and Quasim (2000)Regulating services Improvement of soil quality Boen et al. (2013); Li et al. (2009)

Erosion prevention and water retention

Edmondson et al. (2014);Watts and Dexter (1997).

Local climate and air quality regulation

Edmondson et al. (2014).

Pollination and seed dispersal Andersson et al. (2007); Jansson and Polasky (2010)

Habitat or supporting services

Refuge for plants and animals Pourias et al. (2015)

Maintenance of genetic diversity

Barthel et al. (2010); Barthel et al. (2014)

Cultural services Recreation and relaxation Hawkins et al. (2011); Kaplan and Kaplan (1989); van den Berg et al. (2011)

Physical activity Park et al. (2011); Dunnet and Qasim (2000)

Nature experiences Wilson (1984)Environmental learning Beilin and Hunter (2011); Krasny and

Tidball (2009)Sense of place and social cohesion

Tidball et al. (2014);Okvat and Zautra (2014)

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115Socio-cultural values of urban ecosystem services

2. Valuation of urban ecosystem services

Figure 1 provides a conceptualization of the linkages between the human value system, green space governance and the ecosystem characterizing the urban green spaces. This chapter explicitly looks at the human value system, i.e. the perception and appreciation of ecosystem services from urban green spaces. Values concern appraisal or importance of ES as foundations of human societies (Atkinson et al., 2012, Dendoncker et al., 2013, Gomez-Baggethun and de Groot, 2010). Values of ES are often divided into three main categories: ecological, socio-cultural and economic values (Martín-López et al., 2014). The communication of ES values can serve different purposes, for example awareness raising, environmental accounting, priority setting, instrument design, and litigation in courts (Gómez-Baggethun and Barton 2013).

Figure 1.Urban green spaces as social-ecological systems (adapted from Langemeyer, 2015)

The conceptual framework builds on the ‘Ecosystem Service Cascade’ model (Potschin and Haines-Young, 2011). Governance institutions, including policy and planning shape urban ecosystems through management and practices. The ecosystem’s structure and processes

sustain ecosystem services, which are perceived and appreciated in the human value system.

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116 Ecosystem Services: concepts, methodologies and in truments for research and applied use

The assessment of ES values, i.e. the valuation of urban can be conducted through different methodological approaches, corresponding to the type of values (ecological, socio-cultural, and economic) that are to be assessed. Different types of values provide a different perspective on ecosystem services. Ecologic interests are best addressed through by examining ecological values defining the natural boundaries for a sustainable provision of ES and the potential to sustain ecosystem services over time (Pascual et al., 2010). Socio-cultural and economic valuation both address human preferences and express the social appraisal for ES (Scholte et al., 2015). Citizens’ value systems are well represented and informed through socio-cultural valuation approaches, while governance institutions are often interested in an economic quantification of values.

Ecological valuation approaches determine the capacity or potential of green infrastructure to provide ES, based on biophysical assessments, including for example energy or material f low accounting (Gómez-Baggethun and Martín-López, 2015). The examination and measurement of ecological values is what currently dominates in studies addressing urban ecosystem services (Haase et al., 2014). One example is given in Chapter 5 in which the biophysical capacity of urban green spaces for carbon sequestration and air pollution reduction is modelled.

Economic valuation, often synonymously referred to as monetary valuation, was established on the grounds of classical economic theory and the assessment of human wellbeing in terms of increases in individual utility (Gómez-Baggethun and Martín-López, 2015). Economic values are made explicit through methods such as contingent valuation methods, travel-cost-method and choice experiments (Atkinson et al., 2012). Economic valuation is often appealing and current decision-making is highly influenced by economic considerations. Following, ES assessments strongly rely on economic valuation techniques – urban assessments make no exception (Haase et al., 2014). However, decision-making exclusively based on economic values has been strongly criticized as reductionist (Munda, 2008:35; Spangenberg and Settele, 2010) and ethically questionable (Gómez-Baggethun and Ruiz-Pérez, 2011; Kosoy and Corbera 2010; Jax et al., 2013).

Socio-cultural valuation is often highlighted as an alternative or complementary approach to overcome the reductionism and ethical concerns embedded in economic valuation (Christie et al., 2012; Martín-López et al., 2012). Based on methods stemming from social sciences, such as psychology, ethnology and sociology, social valuation approaches are not linked to assumptions of rational choices and individual utility as economic valuation approaches (Parks and Cowdy 2013). Instead, socio-cultural valuation embraces methods to address either individual or group values, and to examine either utility values or “others-oriented” values (Scholte et al., 2015), through quantitative, qualitative and deliberative approaches (Kelemen et al., 2014), such as surveys, storytelling or participatory mapping (Kenter, 2014).

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117Socio-cultural values of urban ecosystem services

3. Values of urban green spaces in Barcelona, Spain

Cities across Europe increasingly engage with the concepts of urban green infrastructure and ecosystem services. The Barcelona City Council has recently launched Barcelona’s Green Infrastructure and Biodiversity Strategy (Barcelona City Council, 2013), as a strategic policy to enhance the citizens’ well-being through the implementation and improvement of green spaces. Barcelona, located in Northeast Spain, is home to about 1.62 million people (4 million in larger metropolitan area). Representative for major European Mediterranean cities, Barcelona is extreme densely populated (160 inhabitants/ha), and characterized by a very small amount of green spaces per capita (6.82 m2 greenery/inhabitant in the inner city) (IDESCAT, 2013). The small availability of green areas results in a high pressure on adjacent green spaces combined with an ongoing trend of urban sprawl (Barcelona City Council, 2013; Fuller and Gaston, 2009). In this context, a better understanding of the human values related to urban green spaces is critically demanded to steer the implementation and management of inner city green infrastructure with regard to citizens’ need and preferences.

Socio-cultural values of ecosystem services have recently been studied in urban horticulture gardens (Camps-Calvet, 2014; Camps-Calvet, 2015) and at the city’s largest urban park, Park Montjuïc (Langemeyer, 2012; Langemeyer et al., 2015). Over the second half of the 20th century, horticultural gardens and urban parks followed opposite developments in Barcelona. Horticultural gardens were increasingly replaced by built infrastructure and marginalized to the urban fringes (Domene and Saurí, 2007), while urban planners prioritized the creation of urban parks. Today they make 1,076 ha (about 30% of all green spaces), compared to an estimated amount of 30 ha of horticultural gardens (about 1 % of the city’s green spaces) (Barcelona City Council, 2013).

Related to their limited spatial extension, only few citizens directly benefit from the horticulture gardens in Barcelona. Camps-Calvet et al. (2015) estimate less than 1,000 users or direct beneficiaries from all urban gardens in Barcelona, either created by the municipality or through bottom-up squatting and citizens’ initiatives (27 gardens during the time of the assessment in 2013). Notwithstanding the limited number of beneficiaries, Camps-Calvet et al. (2015) found a wide spectrum of ES sustained by urban gardens. The study uses a Likert-scale ranking approach – a socio-cultural valuation approach based on numerically stated preferences (in this case on a scale from 0 to 5). Results from this study underline an outstanding importance of urban gardens for the production of cultural ES, providing both individual benefits through relaxation, stress reduction and nature experiences (e.g. biophilia), as well as group benefits through increased social cohesion, political fulfillment and place-making, i.e. increased quality of places through cooperation (cf. Healey, 2007). The study authors therefore suggest the promotion of urban gardens as green infrastructure to increase human well-being in cities. The study underlines the specific importance of urban gardens to buffer

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118 Ecosystem Services: concepts, methodologies and in truments for research and applied use

social exclusion especially in times of economic crisis, and the capacity to build up urban resilience through fostering urban gardening (cf. Camps-Calvet et al., 2015).

Park Montjuïc is Barcelona’s largest inner city park with over 200 ha and over 16 million annual visits (Barcelona City Council, 2010). Based on these figures and the conduction of a survey-based, economic travel-cost approach, Langemeyer et al. (2015) estimate an annual surplus value of 56,800,000€ (3.55 € per visit) only from the direct utility related to cultural ecosystem services. Socio-cultural values of ecosystem services were further examined by means of Likert-scale rankings (in this case on a scale from 1 to 10). Results, shown in figure 2, affirm the importance of cultural ecosystem services, including recreation, tourism and environmental education. Findings further underpin the importance of Park Montjuïc for so called supporting and regulating ecosystem services (TEEB, 2010; MA, 2005), among them habitat for species, local climate and air quality regulation.

4. Discussion and conclusions

The examples from Barcelona have shown that there are a couple of highly valued ecosystem services from urban green spaces. Studies from Barcelona, in line with a number of other urban ecosystem service assessments indicate the outstanding importance of cultural ecosystem services in urban environments, together with specific regulating ecosystem services and habitat for species (cf. Konijnendijk et al., 2013). Cultural ecosystem services, like recreation, aesthetic appreciation, spiritual experiences, sense of place and social cohesion, enrich human life with meanings and emotions and contribute to enhance the physical and mental health of city inhabitants (Gómez-Baggethun et al., 2013; Maas, 2006).

An important limitation – and potential aim for future research – in the assessment of socio-cultural values in Barcelona, is the lacking consideration of the number of beneficiaries in relation to the size of the available area. For example, urban gardens have been reported to provide multiple, highly important services to their direct beneficiaries, but the overall number of beneficiaries was estimated to be relative small (Camps-Calvet et al., 2015). Urban gardens as part of green infrastructure strategies might thus be reserved for providing benefits to social groups at risk of social inclusion, but do not seem to be appropriate to provide ecosystem service benefits to a larger urban society in a context where available space is extremely scarce. Urban parks may play a different role in the implementation of urban green infrastructure, assuming a stronger potential to provide benefits to a larger number of citizens.

A better understanding of values, expressing human demands for ecosystem services, may allow for adaptations of inner city green spaces and lower some of the pressure on adjacent ecosystems. However, value perceptions are not homogenous across different individuals or societal groups. For example, figure 2

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119Socio-cultural values of urban ecosystem services

Figure 2. Socio-cultural values attached to ecosystem services at Park Montjuïc, Barcelona

Results obtained through a survey using Likert rankings between 1 (very low importance) and 10 (very strong importance) among 198 park visitors (unpublished data from

Langemeyer 2012). Bars show the minimum, maximum, 1st and 3rd quartiles.

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shows strong deviations of stated values in relation to ecosystem services at Park Montjuïc. Other studies showed different appreciation of ecosystem services with regard to the sex, age and origin of the beneficiaries (Bieling et al., 2014; Martín-López et al., 2012) as well as different governance contexts (cf. Gómez-Baggethun and Kelemen 2008). Considering a pluralism of values attached to ecosystem services suggests a need for diversified green infrastructure strategies in cities, i.e. a mix of green space types addressing different needs of urban citizens. A diversified green infrastructure network may also better adapt to changes in the social value system, and changes in the values attached to ecosystem services, as seems to underlie the bottom-up creation of urban gardens in Barcelona (Camps-Calvet et al., 2015b).

Independent from the type of green infrastructure and the specific ecosystem services they provide, it becomes evident that parks and gardens provide multiple benefits simultaneously. This is a crucial difference between an implementation of green infrastructure and engineering-based approaches to urban challenges. Highlighting the multiple benefits ecosystem services from urban green spaces enable, is thus a promising strategy to convince policy-makers to stronger adapt green infrastructure strategies for the sake of human well-being in cities.

Acknowledgements

The research described in here has been conducted by a group of researchers at the Institute of Environmental Science and Technology (ICTA) at Universitat Autònoma de Barcelona, Spain, among them Erik Gómez-Baggethun, Francesc Baró, Laura Calvet-Mir and Marta Camps-Calvet and has been published in different journal articles. The research was funded by the Generalitat de Catalunya through an FI DGR scholarship (2012FI_B 00578).

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127

Indicators of recreational value of urban green spaces

Emma Soy-Massoni,1,* Graciela Rusch2

Abstract

This chapter aims to review the importance of green areas for recreational activities in urban contexts, where ongoing urban planning schemes may adversely affect ecosystem services, as well as human health and well-being. Europeans engage in outdoor and recreational activities very frequently. Therefore the study of green urban areas’ design should take into account qualities that influence usability.

Keywords

Recreational value, green areas usability, human well-being, structural elements, recreational features classification.

1. Ecosystem services of green urban areas

The presence of natural assets and components in urban contexts contributes to the quality of life in many ways (Chiesura, 2004) regarding a wide range of benefits and ecosystem services. Green urban areas provide water flow regulation and runoff mitigation, urban temperature regulation, air purification, etc. (Gómez-Baggethun and Barton, 2013), and some of them contribute directly to quality of life in the city, such as aesthetic and recreational services (Bolund and Hunhammar, 1999). Their benefits are primarily determined by the quantity and quality of green areas as well

1 Landscape Analyses and Management Laboratory. Geography Department. University of Girona, Spain

2 Norwegian Institute for Nature Research (NINA), Trondheim, Norway* Corresponding author email: [email protected]

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as their accessibility (Tyrväinen et al., 2005). In a global context where more than half the world’s population now lives in cities, compared with about 14% a century ago (United Nations, 2001), those services are crucial for population well-being (Kaplan and Kaplan, 1989). Some ecosystem services are locally produced in cities and can be available on the local scale; others need the possibility of transferring the service from where it is produced to the city where humans benefit from it (Bolund and Hunhammar, 1999). However, the quality of life for urban citizens is improved by locally generated services, e.g. recreational aspects that cannot be improved with the help of distant ecosystems.

Various studies have quantified the benefits of green areas in urban contexts, to inform planning processes accordingly. Many studies carried out an economic valuation of the urban green areas’ benefits (e.g. travel cost method, contingent valuation, hedonic price model, etc.). Distance from a green area is one of the most common variables influencing the price of dwelling (Morancho 2003, Barton et al., 2015). Land cover or habitat types are commonly scored according to their ecological potential to quantify flood control, air quality, noise buffering, urban cooling, biodiversity conservation, carbon sequestration, etc. (Farrugia et al., 2013, Radford and James, 2013, Tratalos et al., 2007). Mostly, the scoring system is based on expert judgments and scientific findings, and spatial analysis is used to represent ecosystem services values. However, the existing spatial data are generally more appropriate for use at regional levels than at local levels, given the resolution of regional data is too coarse to reflect the often heterogeneous nature of the local environment (Sheate et al., 2012), which is relevant to evaluate benefits at the local level. A good example is ESTIMAP, which stands for Ecosystem Service Mapping Tool, is a GIS model based approach to quantify and model ecosystem services (Zulian et al., 2013, Zulian et al., 2014). However, ESTIMAP covers the EU-28 using a 1 ha model resolution, a scale that is rarely useful at municipality level. In that sense, Farrugia et al. (2013) found that individual trees have a big influence on flood control and urban cooling ecosystem services and a small scale mapping was recommended. Tyrväinen et al. (2007) developed a social value mapping tool designed to capture the valuations of local green area characteristics.

2. Cultural ecosystem services provided by green areas

Cultural ecosystem services provided by green areas are highly important in the urban context, since green areas contribute to the quality of human habitat in cities (Martín-López et al., 2009), and increase the attractiveness of the city and promote it as tourist destination. Disclosing social and cultural values is especially important for decision-making processes in urban areas because of the very high cultural and social diversity in cities (Gómez-Baggethun and Barton, 2013). Three of the eleven highlighted ecosystem services in urban contexts are cultural: aesthetic, recreation and spiritual (Haines-Young and Potschin, 2008). Despite the wide importance reported,

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129Indicators of recreational value of urban green spaces

only a few scoring systems that assess ecosystem services of urban green areas include cultural values as indicators. A few examples are Radford and James (2013) who scored aesthetic, recreation and spiritual values at Great Manchester (UK) using a survey instrument designed to capture the quality of urban walking (Residential Environment Assessment Tool (REAT); Defra Project (Haines-Young and Potschin, 2008) assessed recreational value at Kent Thameside, London (UK), through network analyses and public consultation; Gundersen et al. (2015) assessed the recreation opportunity spectrum (ROS) in urban forests in Oslo using survey data and automatic counters; and Voigt et al. (2014) combined mapping of the biophysical characteristics of green areas with questionnaires to capture the users’ values of green area features.

3. Recreational value of green urban areas

The recreational aspects of all urban ecosystems are perhaps the highest valued ecosystem service in cities to make the city livable and pleasant for its citizens. Recreational activities based on enjoying the contact with nature are becoming increasingly widespread (Bell et al., 2007, Morancho, 2003): in Stockholm more than 90% of its residents visit parks at least once during the year, 45% do so every week, and 17% more than three times per week; Norwegians engage with outdoors activities 96 days per year (on average), a Danish study shows that nearly 82% of the Danish population visits a park at least once a week (Schipperijn et al., 2010); and almost 97% of the Helsinki residents surveyed in a finnish study (Neuvonen et al., 2007) participated in outdoor recreation during the year, and half of them embarked on a recreational outing daily or every other day.

Green spaces can increase the physical and psychological well-being of urban citizens, including relaxation and re-generative enjoyment (Ulrich et al., 1991). The results from a finnish study revealed a link between the need for restoration (worries and stress), the use of environmental self-regulation strategies (favorite places) and restorative outcomes (Korpela et al., 2010). A study from Sweden found that preferred urban green areas play specific roles and embrace qualities of clear importance to health and outdoor life (Grahn and Stigsdotter, 2010). The variables most predictive of the likelihood of restoration were the percentage of ground surface covered by grass, the amount of trees and bushes visible from the given viewing point, and apparent park size (Nordh et al., 2009).

Parks’ design and management, therefore, should take into account recreational requirements of all target groups where different age-groups and different cultural backgrounds have different motives to visit the park and different activities they are going to undertake (Schwab, 1993, Chiesura, 2004, Arnold and Shinew, 1998). Green spaces can play an important social integrative role if adequately designed (Chiari et al., 2000, Loukaitou-Sideris, 1995, Gobster, 1998). Moreover, people’s expectations and demands for recreational use are closely related to previous environmental experience (Tyrväinen et al., 2007).

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Overall, methods for monitoring and analyzing recreational experience include studies and indicators on preferences, use, and spatial composition of green spaces (e.g. Voigt et al., 2014, Caspersen and Olafsson, 2010, Soy-Massoni et al., in preparation). Those combined methods describe bio-physical characteristics and usability of urban green areas and contribute to the spatial urban planning design in order to achieve the optimal recreational value for its citizens.

4. Design of green urban areas and recreational value

The Greenwich Open Space Project (Burgess et al., 1990) shows that the most highly valued open spaces are those which enhance the positive qualities of urban life: variety of opportunities and physical settings; sociability and cultural diversity. Coeterier (1996) argues that people handle a combination of criteria when a green urban area is visited: land use, ground and water, historic character, naturalness and spaciousness. Coles and Bussey (2000) reported different characteristics of green space, such as size and the presence of facilities, to have an effect on its use. Voigt et al. (2014) proposed a classification of the facilities presents in green areas according three dimensions: natural features, abiotic side conditions and recreational infrastructures; covering a wide range of aspects that have relevance influence on the usability of the urban green areas, and therefore meeting the interest of a multidimensional assessment of green areas qualities. Regarding the urban green areas’ facilities, several authors reported trees, forest or wooded areas as important for the recreational value of green urban areas (Voigt et al., 2014, Kaczynski et al., 2008, Shores and West, 2008, Cohen et al., 2006, Nordh et al., 2011). At Clifton Backies in suburban York, a stretch of scrubby woodland with clearings which contains a diversity of flowers, birds and other wildlife is highly valued by the community (Shoard, 2003). From the Greenwich Open Space Project in south London (Burgess et al., 1990) emerged that open spaces are felt to provide a chance to experience nature, exploration and ‘adventure’; and they provide a variety of natural forms in contrast to the man-made environment (Harrison et al., 1987). Nordh and Ostby (2013) found structures that contributed the most to high ratings on psychological restoration in small urban parks were the “natural” categories: ‘a lot of grass’ followed by ‘a lot of flowers/plants’ and ‘water features’. Voigt et al. (2014), Nordh and Ostby (2013) and Dunnett et al. (2012) also found that proximity to water is highly valued. But despite the natural and water elements, other recreational infrastructures are important for public use of the green urban areas (Table 1): sport facilities and pathways, toilet facilities, provision for children/playgrounds, sitting features, lighting, dog facilities, drinking fountain and swimming areas, public transport access, and silence and tranquility areas. Presence of people can impact positively or negatively green areas usability depending on expectations of the visitors. Urban ecosystem services are effectively co-produced between biotic, abiotic and constructed structures.

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131Indicators of recreational value of urban green spaces

Because there is a distinction across types of green areas (more natural-more humanised) (Kopomaa, 1995) and across types of leisure activities (active/passive) (Voigt et al., 2014); some studies propose a distinction of green areas types depending on the experience class they offer. Gunderson et al. (2015) proposed a classification depending on intensity of use, to support a zoning plan based on a wilderness gradient in order to satisfy both different users’ requirements and conservation. In Copenhagen green areas were classified into seven ‘experience classes’ (wilderness; feeling of the forest; panoramic views, water, and scenery; biodiversity and land form; cultural history; activity and challenge; service and gathering) depending on their design and features present (Caspersen and Olafsson, 2010).

Table 1. Green areas’ structural elements cited as important for recreational value in the literature

Structural element Cited as important for recreational value of green urban areas

Forest/Trees/Wooded areasVoigt et al., 2014, Withford et al., 2011, Kaczynski et al., 2008, Shores and West, 2008, Cohen et al., 2006, Nordh et al., 2011, Gundersen et al., 2008

Grass Nordh and Ostby, 2013Flowerbed Voigt et al., 2014, Nordh and Ostby, 2013Wildlife f lora and fauna Shoard, 2003, Harrison et al., 1987, Gobster et al., 2004Water elements Voigt et al., 2014, Nordh and Ostby, 2013

Sport facilities/pathways Voigt et al., 2014, Özgüner and Kendle, 2006 , Kaczynski et al., 2008, Floyd et al., 2008, Cohen et al., 2006, Lloyd et al., 2008

Quiet and tranquility areas Carles et al., 1999Accessibility/Public transport links Özgüner and Kendle, 2006, Voigt et al. 2014Toilet facilities Özgüner and Kendle, 2006

Provision for children/Playgrounds Özgüner and Kendle, 2006, Ferré et al., 2006, Cohen et al., 2006, Dunnett et al., 2002

Sitting features Özgüner and Kendle, 2006Lighting Cohen et al., 2006

Dog facilities Cutt et al., 2008, Sanesi and Chiarello, 2006; Schipperijn et al., 2010

Drinking fountains Cohen et al., 2006Swimming areas Cohen et al., 2006

An important requirement for providing a high multitude of alternatives of activities and enjoyment is certainly the diversity of biotic and abiotic features, as well as man-made facilities (Voigt et al., 2014), but we also have to regard the size of the park. For instance, the size of the urban green areas is important for the preservation of fauna (Bolund and Hunhammar, 1999), and a significant climatic function can only be expected when park size exceeds one hectare (Tyrväinen

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et al., 2005). Urban forest size is also a determinant factor having influence on house prices (Kong et al., 2007). Studies in the UK have shown that urban parks have a minimum attractive size for visitors of about two hectares and that their attractiveness increases when linked with footpaths (Coles and Bussey, 2000; Tyrväinen et al., 2005). Urban parks, although on a smaller scale, also carry out the same environmental and recreational functions as forest and green areas (Morancho, 2003). Tyrväinen et al. (2007) found that values and qualities of green areas are to a certain extent independent of the area and land-use type, because some activities may be pursued and expectations fulfilled in various environments. Nordh (2009) and Nordh and Ostby (2013) demonstrated that small urban parks (pocket parks) could provide a potential for psychological restoration. However the relationship between green urban areas size and the diversity of facilities present is not well studied yet.

5. Final remarks

Undoubtedly green urban areas have a value for recreation in cities. Satisfying recreational experiences depends on the design and provisioning of natural and manmade features, facilities, and amenities meeting visitors’ interests and demands. Recent studies, dealing with the relationship between green urban areas characteristics and visitors’ activities and demands, propose an integrative method including a feature mapping exercise and questionnaire surveys to users, with the aim to understand the differences regarding demand and supply of recreational services. Thanks to these new approaches, regional and municipal planning policies are provided by a decision support system for use at municipal and regional levels, in order to facilitate future planning of the recreational landscape.

Acknowledgements

The present review have been performed within the OpenNESS Oslo case study under the supervision of David N. Barton (OpenNESS project: Operationalisation of Natural Capital and Ecosystem Services: From Concepts to Real-world Applications, funded from the European Union Seventh Framework Programme FP7-ENV.2012.6.2-1 under grant agreement n° 308428). Moreover the study was supported by a grant from Iceland, Liechtenstein and Norway through the EEA Financial Mechanism, operated by Universidad Complutense de Madrid.

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137

Payment for ecosystem services: concept and examples

Rafael SardáCentre d’Estudis Avançats de Blanes (CEAB-CSIC), Spain1

Abstract

As humans increase their footprint on the environment, the world Natural Capital degrades. This Natural Capital benefits humankind through the continuous offer of ecosystem goods and services and its degradation jeopardizes our ability to generate welfare in the future. Searching for solutions different economic arrangements have been created to reward the conservation of the Natural Capital and ensure the provision of its ecosystem goods and services. Amongst all of these mechanisms, the general concept of Payment for ecosystem services (PES) evolved. PES is a term used to describe a range of schemes through which the beneficiaries, or users, of ecosystem services provide payment to the stewards, or providers of those services. This paper reviews the basic concepts of these public, private or public-private schemes and provides general examples of these solutions. Finally, emphasis is given to enlighten its possible use for beach social-ecological systems in our environments.

Key words

Sustainable development. Ecosystem services. PES. Ecosystem Approach.

1 [email protected]

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138 Ecosystem Services: concepts, methodologies and in truments for research and applied use

1. Introduction

In recent decades, a global anthropogenic-driven transformational change has produced a large shift of the world’s natural environment (Crutzen, 2002; Rockström et al., 2009; Steffan et al., 2015). To balance present economic development and social prosperity with environmental protection has become a priority. From Rio 1992 to Johannesburg 2002, Rio 2012 and beyond, Sustainable Development and the recent social-ecological paradigm has made its way from the academia arena to the public, business and political agendas. The environment is pressuring us today to find adequate responses to what has been described as the largest human-caused environmental crisis.

To give visibility to this crisis, United Nations launched the Millennium Ecosystem Assessment (MEA) report (MEA, 2005). This report constituted a global audit on the world’s ecosystems and concludes that ecosystems have declined more rapidly and extensively over the past 50 years that at any other comparable time in human history, advising that this degradation jeopardizes not only ecosystems around the globe but also human activities because healthy ecosystems provide natural conditions that are indispensable for human well-being and societal welfare. The MEA report clearly stated that world ecosystems, if sustainably managed and protected, benefit current and future people and societies emphasizing the concept of ecosystem goods and services as the flow of benefits from nature to people. These concepts introduce a new jargon and framework that can be used in the management of public goods. On the biosphere, sunlight produces organic matter through the photosynthesis process. This organic matter travels around complex trophic chains giving us humans a living support. This produced biodiversity also helps us to obtain other types of goods but also to support other fundamental ecosystem services such as habitat, water depuration or carbon fixation, between others, that benefit humankind. During the last three decades, a large effort has been made to demonstrate how the Natural Capital (ecosystems’ structural units and their functions) is benefiting kind through the continued offer of ecosystem goods and services (Daily, 1997; de Groot et al., 2002; Gómez-Baggethum et al., 2010), Today we are learning how to value such goods and services in monetary terms through new markets and innovative schemes (www.ecosytemmarketplace.com).

Innovative economic arguments for linking public and private efforts to protect ecosystems and to ensure the provision of ecosystem goods and services are becoming more and more used today. Some of these arguments like the “payment for ecosystem services (PES)” could be a fundamental move in this way (Wendland et al., 2009; Farley and Costanza, 2010; Farley et al., 2010). PES schemes are voluntary transactions where a well-defined ecosystem service is being “bought” by at least one buyer from (a minimum of one) ecosystem service provider if, and only if, the ecosystem service provider secures its provision. PES schemes differ greatly from other conventional markets and transaction tools to address negative externalities of our activities following the “polluter pays principle”, by the introduction of the “steward earns principle”, or the “beneficiary principle,

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139Payment for Ecosystem Services: concept and examples

in which positive externalities are addressed from the very beginning (Gómez-Baggethun and Ruiz-Pérez, 2011).

The aim of this article is twofold; on one hand, to review fundamental ideas around the payment for ecosystem services concept and, on the other hand, toexplore through examples the raise of public, private and public-private innovative models of collaboration that help in the maintenance and/or restoration of vital ecosystem services. Establishing connections between ecosystem change and people’s benefits can lead to develop a much more proactive approach in the protection and conservation of vital ecosystem services.

2. Sustainability in the social-ecological paradigm

Our environmental crisis can be described by applying three main tendencies; human footprint is increasing, planet earth has a bounded carrying capacity for its natural resources and the natural capital is degrading. Altogether puts the old concept of Sustainable Development “meets the needs of the present without compromising the ability of future generations to meet their own needs” in a serious dilemma (Figure-1); we are running an unsustainable machine in an unsustainable way. Our Natural Capital, the extension of the economic notion of capital (manufactured means of production) to goods and services relating to the natural environment, should assure a valuable flow of these goods and services (natural resources) through various combinations of ecosystem functions which are in turn delivered by different components of its biodiversity.

In today’s world, natural resources cannot be treated as discrete entities that need to be analyzed separately; they are dependent of the social and economic systems with which they interact. The concept of social-ecological systems (Berkes and Folke, 1998) has been gaining acceptance to analyze this complexity. Social-ecological systems (Figure 1) are “complex adaptive systems in which humans are part of nature and the dynamics of both dimensions are strongly linked at equal weight”. When managed, these coupled co-evolving systems should focus on the ability to respond to feedbacks from the environment considering the tendencies of ecosystem goods and services that we obtain as benefits from the environment. A framework for the analysis of social-ecological systems is explained in Sardá (2013).

The social-ecological paradigm of our present times recognizes the mutual inter-associations between human societies and ecological processes that are necessary for the survival of both. This paradigm gives much more importance to the Natural Capital as resource provider and puts pressure in better defining the sustainable development path that we need to construct for the future of humanity “enhancing social capital and economic prosperity while maintaining the integrity of natural systems and its potential for the provision of ecological good and services”. By using this new definition two aspects become more relevant: a) a better care

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of ecosystem service provision, and b) a move in future regulations from impact-driven policies to state-driven policies. The PES schemes are solutions for the first demand, while the introduction of the Ecosystem Approach (EA) strategy (CBD, 1998) is a solution for the second one.

Figure 1.–Sustainability in the social-ecological paradigm.

3. The payment for ecosystem services (pes) concept

3.1. PES basic concepts.Market-based mechanisms have begun to be considered widely as important tools for the protection and restoration of ecosystem services. However these schemes are only one among other developed tools (regulatory schemes, direct provision by governmental directions, voluntary efforts played by stakeholders, and others).

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141Payment for Ecosystem Services: concept and examples

A large group of market-based mechanisms can be found in practice; taxes, fees and other related charges, tradable permits and offsetting schemes, certification tools, and PES schemes. From all of these market-based mechanisms, PES schemes evolved rapidly due to its special particularities; introducing economic considerations when talking about ecosystems services, win-win advantages, short to long-term financial considerations, social responsibility aspects for the private sector,… Nevertheless, these schemes also have difficulties because they are entirely dependent of a correct understanding of ecosystem services, a correct ecosystem mapping of the social-ecological system selected, a correct assignment of property rights, and a correct understanding of functional processes in ecosystems. A much better usage of these schemes could be obtained if they were considered inside of an environmental management tool that follows an ecosystem approach strategy.PES is a term used to describe a range of schemes through which the beneficiaries, or users, of ecosystem services provide payment to the stewards, or providers of those services. The beneficiaries may be individuals, communities, businesses or public bodies. Providers can be also individuals, communities, business or public bodies that are responsible and manage or simply live with these natural resources. Basically PES are mechanisms aimed to advance into the protection of the natural capital and the sustainable development of social-ecological systems.PES schemes are voluntary transactions. An agent (the beneficiary) realizes that an ecosystem service(s) upon which it relies is being degraded and/or depleted, then, through a PES mechanism, another agent can be paid to ensure the permanence of this ecosystem service(s) by doing different actions. Obviously, before signing these transactions, you need to demonstrate what is the present baseline of the service(s) under negotiations. The payment can vary in its form (money, fiscal deductions, intangible issues, …), yet the provider will always increase its personal value due to the transaction. These transactions are given by certain actions that are new, they work beyond legal obligations (additionality), and they are dependent on the delivery of the negotiated ecosystem service(s). The payment should cause the benefit to occur where it would not have otherwise (conditionality). The PES transaction must also take in consideration that what it is going to be done for the provision of the ecosystem service(s) does not harm other ecosystem services elsewhere.PES schemes have also detractors. There are people in complete opposition to the idea of pricing nature; however, many times, this is the only possibility to protect such services and no other alternative mechanisms are forecasted to avoid irreversible damages. PES mechanisms are not only about pricing but also about values. During the transaction, the provider of the services increase its returns (through the payment or even by increasing economic profits on the field) but normally, both the beneficiary and the provider increase their global value by the additional value introduced by the fact of improving its ecosystem services.

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There are different possibilities in PES transactions. Providers can gain economic returns from a PES agreement by “selling” actions ensuring an entire range of ecosystems services as a global package, or by “selling” different ecosystem services to different beneficiaries; the “buyers” that will benefit from these services. In addition, the “sellers” can obtain these returns from the transaction and, in addition, besides the beneficiaries, other people can benefit from the improved ecosystem services provided which, in turn, can also be an intangible gain as social responsibility actions.

3.2. PES typologies and examplesThe introduction of a PES scheme is largely dependent on the social-ecological system in which it is going to operate. This social-ecological system (Figure-1) “A spatially bounded region containing an ecosystem and a social system interacting with each other” can be local, regional, national or trans-national. In a simple typological way, PES schemes can be classified into three different types: public, private and combined public-private schemes (Table-1) (DEFRA, 2013). Different revisions on this topic can be found in Forest Trends et al., 2008; DEFRA, 2013; Welsh Government, 2014.

Table 1. Types of PES schemes (adapted from DEFRA, 2013).

PES Types Description

Public payment schemes Governments pay land or resource managers to enhance ecosystem services on behalf of the wider public.

Private payment schemesShelf-organised private deals in which beneficiaries of ecosystem services contract directly with service providers.

Public-Private payment schemes

Governments and private funds pay land or other resource managers for the delivery of ecosystems services.

If a private owner has an economic return for activities that may have been compromising ecosystem service(s) that are critical for others stakeholders, public managers or even these former stakeholders directly can work in co-operation with the owner to reduce such activities by adding some incentives and ensuring that the owner will take care of the provision of the ecosystem service(s). At the end, agreements will be considered in a win-win situation, the private owner increases the value of the returns and the beneficiaries can still enjoy their ecosystem service(s). The private owner sometimes can increase its economic returns and improve ecosystem service(s) at the same time; for instance, in the case of IKEA, better practices in the field for farmers, allow these people to improve incomes while reducing its negative footprint and improve ecosystem service(s) in the region. Table-2 is showing several examples of PES for the categories exposed above.

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143Payment for Ecosystem Services: concept and examples

Table 2. PES examples

Public payment schemes

1.–Catskill Mountains (New York, USA)

In order to preserve the quality of the city’s drinking water, the city government of New York realized that changing agricultural practices in the Catskill mountains and paying to preserve the rural nature of this area was more valuable that to invest in new water-filtration plants and its maintenance costs. The city government paid 250 M$ buying land to prevent development and is paying 100 M$ a year to farmers to minimize pollution. The complete deal (all land around reservoirs) was not closed but the lesson was learnt.

Appleton (2002)

2.–Tasmanian Forest (Australia)

The Australian and Tasmanian Governments conservation goals for forest were set to protect almost 50,000 ha of private landowners. They created a Forest Conservation Fund (FCF) that it was working using market-based mechanisms. Landowners could apply for the Fund in a competitive tender process that also implied valuing their properties following a Conservation Value Index.

(https://www.environment.gov.au/node/20482)

Private payment schemes

3. Perrier-Vittel (France)

Perrier-Vittel, a French bottler of mineral water, found it necessary to reforest part of heavily farmed watersheds and also to pay farmers to switch to modern facilities and organic farming in order to preserve the quality of some of its products as a consequence of underground contamination by nitrates.

Perrot-Maítre (2006).

4. IKEA and Better Cotton (India/Pakistan)

IKEA had problems with the cotton it was using for their products. IKEA developed a project aimed to inf luence farmers in Pakistan and India to change the way they cultivate cotton. By using fewer chemicals and pesticides, less water practices, and other simple tools, the life of these farmers improved in earnings and living standards benefited for better provisioning of ecosystem services such as water depuration and habitat condition. Altogether improve incomes for farmers but also have a lower impact on the environment.

Pogutz (2014)

5. Unilever and the

Sustainable Fisheries Initiative (International waters)

Many years ago, Unilever committed to source all its fish from sustainable stocks. As one of the largest fish buyers, the company decided to go for a more sustainable fishing and to secure its fish suppliers for the future. Although not completely successful, the company is still going in this way. The premium paid protects marine ecosystem services linked to not overexploited populations.

Pogutz (2014)

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144 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Public-Private payment schemes

6. Forest protection (Costa Rica)

The government-led Costa Rican PES Programme rewards forest owners for four bundled Environmental Services (ES) (watershed protection, carbon sequestration, landscape beauty and biodiversity protection) their forests provide. The system works also with the engagement of the private sector (mostly Hydroelectric Power (HEP) producers) who receive services such as streams-flow regulation, sediment retention an erosion control, and that is charged in part to private consumers of water used for multiple aspects. Finally public in general benefit with the supply of water and the maintenance of the area’s scenic beauty for recreation and ecotourism .

Porras et al. (2013)

7. The Panama Canal (Panamá)

The Panama Canal watershed is currently being reforested to facilitate the management of water flow into the channel and the avoidance of sediments and nutrients entering the Canal. In this case under the vigilance of the Government that cannot afford to pay this price, a Forestry Insurance Company is creating a 25 year bond to facilitate reforestation. Big companies using the Canal are paying for the bond. These companies actually are paying large insurances that get a reduced premium if they purchase the bond stimulating in this way the functioning of the PES mechanism, Money go to land owners for reforestation.

Brauman (2008)

8. Sierra de las Minas (Guatemala)

The Sierra de las Minas Biosphere Reserve (SMBR) Water Fund is a World Wildlife Fund (WWF) initiative to finance responsible water management. It promotes activities upstream that are expected to increase recharge and decrease erosion and support major users (industry) to increase their water use efficiency and reduce the impacts of their effluents. Savings are to be invested in the fund and channelled upstream to protect the buffer zone of the SMBR. The goal of the present PES initiative is to promote shade coffee and sustainable agriculture by channelling funds to the farmers living in the reserve’s buffer areas (large private farmers and forest owners, indigenous community forest owners, and cooperatives (coffee, sugar and vegetables). The Fund is managed by WWF and “Fundacion Defensores de la Naturaleza” (FDN) and paid by a total of 20 industrial users (large companies).

Mendez et al. (2004)

3.3. Further possibilities for PES in our regionIn Catalonia we have implemented some PES mechanisms; the “Xarxa de Custòdia del Territori (XCT)” is probably the best example, XCT is a land stewardship strategy intending to generate accountability among the owners and users for the conservation and proper use of natural, cultural and landscape resources and values. Stewardship materializes in voluntary agreements between the owners and managers of land, and land stewardship entities in order to maintain and recover the natural environment and landscape. However, the possibilities for application of PES schemes are large and still mainly unexplored.

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145Payment for Ecosystem Services: concept and examples

In Catalonia as well in other parts of Spain, beaches play a key role in the maintenance of the Tourism and Transport Industry, an essential sector for the economic welfare of the country (15.2% of the entire GDP in 2014, accounting for 160 billion euros). This importance is largely recognized, [for instance “Tourism main reason for visiting Andalucia is the use an enjoyment of beaches” (Consejeria de Turismo de Andalucia)]. However, besides these considerations, beaches are in constant erosion and they suffer important losses in the provision of their ecosystem services. Mostly this occurs because river flow (and its associated sedimentary balance) has been restricted by manmade interventions, coastal morphodynamics altered, and river mouths suffering intense occupation and being not able to play their ecological functions. The recreational service of the Lloret de Mar central beach was assessed using Travel Cost Methodology (TCM) (Ariza et al., 2012) in 73.8 M euros per year (beach investments in Lloret de Mar beaches this year barely rose 1 M euros). Using the same TCM methodology, three beaches of the city of Cádiz (Santa Maria del Mar, Victoria and Cortadura) totaled a recreational service of 141 M euros with an average investment budget in the city of 1.2 M euros per year (Alves, 2015). The difference between the recreational service provided by those beaches and the public investments for taking care of these resources is completely misbalanced. We deeply believe that natural resources as beaches are granted for us and they will always be with us, but tendencies today show that this may will not be the case (CIIRC, 2010). In order to guarantee future enjoyment of these recreational services, provision of watershed services such as water flow, flood protection, and functional habitat at the mouth of the rivers could be restored conveniently by using PES schemes as indicated in Figure.2. Still, we have plenty of information needed to ensure that this could work but PES mechanisms open up large possibilities in the management of natural public goods for the future.

Figure 2.–Schematic representation of a PES project structure. Downstream ecosystem service users as in the case of beach beneficiaries negotiate a deal with other stakeholders

that can improve the provision of watershed services. (Adapted from DEFRA, 2013).

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4. Discussion

Following the publication of the Millennium Ecosystem Assessment report (MEA, 2005), a new social-ecological paradigm emerged to change the way humans relate with natural systems. Natural systems are starting to be seen much more as capital assets that provide a flow of natural goods and ecosystem services to humanity , a flow that needs to be preserved. However, most people believe that public institutions should take care about these public goods and should be leading conservation planning for ecosystem services. In this way, mankind’s impact on the environment (from individuals or corporations) should be regulated by those institutions to take proper care of its use and abuse. The above comments sound logical but many times public management is not enough and other mechanisms need to be developed to take care of these systems. Market-based mechanisms such as PES schemes can play a major role at that time.

The use of payment schemes lies in highly technical issues (understanding and mapping ecosystem functions and services, proper assignment of property rights, economic considerations). Assessing its feasibility, providing implementable recommendations, diminishing data gaps, designing and implementing action programs, analyzing its possible use in agreements, are all needs that can explain the difficulties associated with implementation. However, a correct analysis can facilitate an integrated management of natural systems, something that it is included in all regulatory schemes these days. For example, clean drinking water, sediment transport, and coastal productivity are all benefits arising from a well-managed river ecosystem, but normally these services are managed independently in isolation one from another and tradeoffs over management options depend on partial negotiations and lobbying. In addition, sediment transport is an intermediate service for the final service of having a safe recreational space in a beach to lay-down in summer time, which is basic for the maintenance of the tourism industry. As we do not properly account for these relationships, many times we end up with decisions that are not the best ones for the entire society. In this case, beneficiaries of the latest service can be local actors such as hoteliers or global actors such as tour-operators, but, in any case, they are normally far from the room in which important decisions that affect their business are taken on the provision of ecosystem services that benefit them. The use of PES schemes will be better oriented if these mechanisms are used in conjunction with correct environmental management tools that follow an Ecosystem Approach (EA) strategy.

Recently, in order to widely adopt an EA strategy in management, a standardized stepwise process has been developed to ensure consistency in the development of management measures that address the aspirations of the stakeholders and meet legislative and regulatory requirements, the so-called Ecosystem-Based Management System (EBMS) (Sardá et al., 2014). The EBMS is a stepwise process that combines environmental quality and risk management system theory with the EA principles. The EBMS is an intended standard operational procedure for

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147Payment for Ecosystem Services: concept and examples

practicing the EA strategy. The EBMS departs from the prioritization of activities in policy cycles of continuous improvement based on adaptive management with the final objective to reach a desired vision. Once this prioritization is delivered, PES schemes can be one of the selected tools to deal with required objectives.

It is time to develop proper structured methodologies that help people to understand, valuate, and manage risks and opportunities arising from our dependence to ecosystems and the provision of its services. Included or not in larger management frameworks, PES schemes are one of these methodologies that can open solutions for the future.

5. Conclusions

As human beings increase their footprint on the environment, the world Natural Capital degrades. This Natural Capital benefits mankind through the continuous offer of ecosystem goods and services and its degradation jeopardizes our ability to generate welfare in the future. PES schemes open up the possibility to shift the way we understand Nature, its functionality and its provision of ecosystem goods and services. The introduction of PES schemes into our daily operational work should be based on a correct understanding and mapping of these ecosystem goods and services facilitating the introduction of the EA strategy, by reconciling the theory of environmental policy with the practice of environmental management. The new framework opened by the introduction of the ecosystem services concept into the management of public goods should help us to advance in the conservation of natural systems.

6 Acknowledgements

This work was carried out within the framework of the KNOWSEAS+ project (PIE-2014) and the PLAYA+ project (CGL2013-49061) of the National Research Plan of Spain in R+D+i. We thank the Secretariat of the XV International Summer School on the Environment to allow us to present the paper at the Conference.

7 References

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Alves, B. (2015). Beach user’s perception, coastal value and management implications. Doctoral Thesis University of Cádiz. 120 pp.

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Brauman. K. (2008). Emerging markets for ecosystem services: a case study of the Panama Canal Watershed. Ecology, 89 (9), 2667-2668.

Centre Internacional d’Investigació dels Recursos Costaners (CIIRC). (2010). Estat de la Zona Costanera a Catalunya. Departament Política Territorial i Obres Públiques. Barcelona.

Convention on Biological Diversity (CBD). (1998). Report of the workshop on Ecosystem Approach. Lilongwe, Malawi, 26–28 January 1998, UNEP/CBD/COP/4/Inf.9, p. 15

Crutzen, P.J. (2002). Geology of mankind: the Anthropocene. Nature, 415, 23.Daily, G.C. (1997). Nature’s Services. Island Press, Washington. 393 pp.DEFRA. (2013). Payments for Ecosystem Services: A Best Practice Guide. 82 pp

and annexes.de Groot, R., Wilson, M.A. and Boumans, R.M.J. (2002). A typology for the

classification, description and valuation of ecosystem functions, goods and services. Ecological Economics, 41, 393–408.

Farley, J. and Costanza, R. (2010). Payments for ecosystem services: from local to global. Ecological Economics, 69, 2060-2068.

Farley, J., Aquino, A. Daniels, A., Moulaert, A., Lee, D. and Krause, A. (2010). Global mechanisms for sustaining and enhancing PES schemes. Ecological Economics, 69, 2075-2084.

Forest Trends, Katoomba Group and UNEP. (2008). Payments for Ecosystem Services: getting started. A Primer. UNEP Publisihng Services. Nairobi. 65 pp.

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Gómez-Baggethun, E. and Ruiz-Pérez, M. (2011). Economic valuation and the commodification of ecosystem services. Progress in Physical Geography, 35, 613-628.

Méndez, J.C., Hernández, O., Cobos, C.R., and Ortiz, A. (2004). Economic valuation for the environmnetal service of wáter regulation. Souther Sierra de las Minas, Guatemala. In: Payment Schemes for environmental services in watersheds. FAO Regional Office for latin America and the Caribbean. 76 pp.

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Millennium Ecosystem Assessment (MEA). (2005). Ecosystems and Human Well-being Synthesis. Washington DC. Island Press.

Perrot-Maítre, D. (2006). The Vittel payments for ecosystem services, a “perfect” PES case?. International Institute for Environment and Development. London (UK). 24 pp.

Pogutz, S. (2014). IKEA and the Better Cotton Initiative. Oikos Case writing. 2012/2012/MU/nº0003. Bocconi University, Milano (Italy). 25 pp.

Pogutz, S. (2014). Ecosystem services and Business: Unilever and Palm Oil. Oikos Case writing. 2012/2013/MU/nº0003. Bocconi University, Milano (Italy). 23 pp.

Porras, I., Barton, D.N., Chacón-Cascante, A. and Miranda, M. (2013). Learning from 20 years of paument for Ecosystem Services in Costa Rica. International Institute for Environment and Development. London. 76 pp.

Rockström, J., Steffen, W., Noone, K., Persson, A. Stuart Chapin, F., Lambin, E.F., Lenton,T.M., Scheffer, M., Folke, C., Schellnhuber, H.J., Nykvist , B., de Wit, C.A., Hughes, T., van der Leeuw, S., Rodhe, H., Sörlin, S., Snyder, P.K., Costanza, R., Svedin, U., Falkenmark, M., Karlberg, L., Corell, R.W., Fabry, V.J., Hansen, J., Walker, B. Liverman, D., Richardson, K., Crutzen, P. and Foley, J.A.. (2009). A safe operating space for humanity. Nature, 461, 472-475.

Sardá, R. (2013). Ecosystem Services in the Mediterranean Sea: the need for an Economic and Business oriented approach. In: Mediterranean Sea. Ecosystems, Economic Importance and Environmental Threats. Novar Publ. New York. pp: 1-35.

Sardá, R., O’Higgins, T., Cormier, R., Diedrich, A. and Tintoré J. (2014). Proposal of a Marine Ecosystem-Based Management System (EBMS): linking the theory of environmental policy and practice of environmental management. Ecology and Society, 19 (4). http://www.ecologyandsociety.org/vol19/iss4/art51/

Steffan, W., Richardson, K., Rockström, J., Cornell, S.E., Fetzer, I., Bennett, E.M., Biggs, R., Carpenter, S.R., de Vries, W., de Wit, C.A., Gerten, D., Heinke, J., Mace, G.M., Persson, L.M., Ramanathan, V., Reyers, B. and Sorlin, S. (2015). Planetary boundaries: guiding human development on a changing planet. Ecology, 347 (6223).

Wendland, K.J., Honzák, M., Portela, R, .Vitale, B., Rubinoff, S. and Randrianarisoa, J. (2009). Targeting and implementing payments for ecosystem services: opportunities for bunding biodiversity conservation with carbon and ater services in Madagascar. Ecological Economics, 69, 2093-2107.

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151

Tools for regulating environmental services in Catalonia

Joan Pons SoléINSTA – Serveis jurídics ambientals, Tarragona, Spain1

Abstract

The regulation of environmental services through tools and legal mechanisms in Catalonia could be regulated from multiple ways. Firstly, land stewardship is a model for the relationship between the environment and the people has been greatly developed in our country, even with legal tools at their disposal. The benefits of land stewardship in regards to the conservation of ecosystem services are clear, so the door is open. But, on the other hand, knowledge of other mechanisms, such as payment for environmental services and tax incentives, might add to the available tools for improving the quality of ecosystem services and the welfare of the catalan society. are also, This article aims to address what could be a future regulatory system of environmental services in Catalonia. The author has taken references from experiences in different parts of the world, analyzing the existing legislative framework and the reality on the territory.

Keywords

Environmental services, land stewardship, environmental taxation, públic regulation, environmental law.

1 [email protected]

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152 Ecosystem Services: concepts, methodologies and in truments for research and applied use

1. Payment for environmental services in Catalonia

The possible implementation of a payment for environmental services (PES) system in Catalonia must have a strong analysis of what are the main environmental services and then determine ways to finance the conservation and the payment. Catalonia, like other countries, has already implemented some kind of PES.

According to Russi (2010), “the main elements of reflection derived from the analysis of ongoing Catalan experiences are:

In many areas, PES can be used as a tool for effective environmental policy.

PES promotes the mobilization and local and regional empowerment.

We need to advance in the direction to assess and promote the proper degree of conditionality and additionally.

PES programs can contribute significantly in terms of communication and environmental education.”

Legal framework applicable to the payment for environmental services

Spanish Law 42/2007, of natural heritage and biodiversity, specifically the article number 73, mentions the need to find mechanisms and policies to encourage positive externalities in conservation, restoration and improvement the natural, fixing carbon dioxide, soil conservation and hydrological regime and groundwater recharge. The aforementioned article makes an explicit reference to environmental services and the possible compensation, not speaking specifically of PES as such.

Catalan Law 8/2005, of protection, management and planning of the landscape, provides a Fund for the protection, management and planning of landscape, specifically in article number 16. Through this Fund, not yet developed, PES programs aiming to preserve landscape in Catalonia would be financed.

The existing legal framework provides regulatory pathways for the management of environmental services, yet often not specifically designed for this purpose. That is why it’s absolutely necessary to have a specific legislation that promotes PES in Catalonia. The preamble of the draft law of natural heritage and biodiversity of Catalonia mentions the need for regulation to protect environmental services of ecosystems, or at least it defines relevant threats to those services. After several references throughout the text, the latest version of the bill includes a specific article that opens the door to the implementation of PES in Catalonia.

Moreover, the Spanish law 43/2003 of Forestry amended on March 26 2013, incorporated as a subject of the law, in article number 4, the social function of the natural environment, particularly the environmental services provided.

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153Tools for regulating environmental services in Catalonia

The figure of the territorial contract derived from Spanish law 45/2007 of Rural Sustainable Development, is close to the land stewardship agreements, thus allowing to establish mechanisms that could be assimilated to PES.

Lozano and Ràbade (2013) believes that territorial contracts encourage the agricultural activity in favour of sustainable development of the rural economy, as a mechanism that acts on the conservation of agricultural environmental services. The author defines the territorial contracts as subsidies because it would be a transfer of funds that pursue public interest in the form of quantitative and qualitative environmental improvements. Payments offset positive externalities generated by farms; the improvement of agricultural ecosystems and the sustainable development of rural areas.

Finally, the recent Spanish law 21/2013 promotes the creation of conservation banks. We will see if the law implementation leads to the proliferation of payment mechanisms for environmental services and compensation of externalities.

Considerations for a PES in Catalonia

Within the reality of Catalonia PES should be specified in these areas:

• Private property: Catalonia has over 80% of the land in private hands. Therefore, it is absolutely necessary to take into consideration landowners in any PES scheme. Hence, mechanisms such as the “Selvans Program” promoted by the Province Council of Girona, which pays long-term income of felling mature / old forests to keep them standing, or Costa Rican experiences on payment for environmental services of forests, would be a reference because they fully involve the private sector in the management of environmental services.

• Tributary funding of PES: a way of financing PES could come from taxing the consumption of, for example, fossil fuels or water (as in Mexico for water services), among others. However, the strong structure and consolidated taxation in our country would be an obstacle to finding ways of funding from these instruments.

Different factors play against PES in Catalonia, among which the current economic situation that makes it unlikely to implement PES based strictly on public funds. It’s for this reason that we have to study the involvement of companies in the design and development of PES, as it has been done in Brazil and France, for the economic success of a PES project. Another important factor is the lack of tradition on PES in Catalonia. The conception of the environment as something that provides environmental services that must be preserved and, moreover, the notion of paying for a certain sustainable management of these services, are still far from the reality of the Catalan society. We will see if the theories of authors like Folch (2013) are well received by society.

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154 Ecosystem Services: concepts, methodologies and in truments for research and applied use

2. Land stewardship and environmental services

Under land stewardship2 we understand a set of voluntary agreements between landowners and stewardship entities, in the design of management strategies of environmental services from ecosystems. Pallarès (2010), frames voluntary agreements as “those legal transactions entered into by one or more administrations, and one or more natural or legal persons subject to conventional private law, a common legal status of the parties, which aims to improve environmental protection, scheduled or not, without forgetting the public purpose that may be pursued through the use of the mentioned instrument.”

Land stewardship agreements should be an incentive encouraging owners to visibilize the preservation of environmental services. According to Vázquez (2011) this can be done in two ways:

• By direct economic benefits (incentives and tax breaks, direct or indirect subsidies, etc.).

• Through the ability to sell in secondary markets with the economic value of the rights over resources and environmental services, depending on the state or status of the agreement or contract established.

The author highlights the fact that the real right holder (owner) will take advantage of the content of the sustainability values of the property, because:

• It increases the economic value of the property. • It gives access to grants and tax benefits. • It allows to operate directly on environmental values, within the parameters

of the proposed land stewardship. • It fosters marketing both sustainability and the values of the property.

Different types of environmental services of ecosystems and land stewardship activities on the environment can act on the full range of environmental services. Land stewardship can improve the supply of local produce, as well as providing regulatory services, such as erosion control, improved water quality, etc. Moreover, environmental services contribute to a higher cultural quality by improving aesthetics, landscape and promoting dissemination, research and environmental education. Finally, good land stewardship practices can clearly influence

2 Land stewardship is a strategy intending to generate accountability among the owners and users for the conservation and proper use of natural, cultural and landscape resources and values. Land stewardship materializes in voluntary agreements between the owners and managers of land, and land stewardship entities in order to maintain and recover the natural environment and landscape. Land trusts are public or private not-for-profit organizations that take an active part in preserving land and its values through mechanisms making land stewardship easier. (Definition of Xarxa de Custòdia del Territori)

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155Tools for regulating environmental services in Catalonia

sustaining ecosystem functions, on the basis of the joint system that gives meaning to all environmental services.

Examples such as Menorca’s agricultural contracts demonstrate that it is possible to preserve the environment and their environmental services, through the figure of land stewardship and the contracts between the steward and the property manager. In addition, this example also confirms that there is no need to display a complex legal framework for the establishment of these contracts.

Consolidated and internationally recognized experience on land stewardship in Catalonia, as well as the provision of a consistent enough legal framework, and ongoing steps to improve safekeeping tools, become one of the main routes for the direct and indirect management of environmental services. But these mechanism must not be isolated, they must be accompanied by fiscal measures, as discussed below, and possibly some PES strengthening programs.

3. Environmental taxation applied to environmental services

Environmental taxation can be defined as “any use of the tax system as a means of encouraging changes in behaviour that are considered positive from the environmental point of view” (Roca et al., 2005).

Authors like Buñuel (2002) mention that for a tax to be considered “green” or environmental, it needs to revert on the environmental vector for which it is issued, while discouraging its further degradation. Within this classification, the author also includes taxes that are not purely environmental, due a percentage of the proceeds going to environmental purposes.

According Boisán (2011), an agreement for a fee or remuneration of rights over environmental services must be reached. This, through administrative channels, takes shape by tax credits, such as a reduction in the regional income tax, property tax rebates, subsidies, etc. The author mentions that we are not talking about the creation of new special tax rates, but applying incentives and bonuses in existing taxes that reward sustainable management of environmental services.

Tax deductions

On the application of income tax deductions, article 34 of the Law 16/2008 of tax and financial measures provided in this aspect, but concentrated on donations to nonprofit organizations and environmental stewardship. Although the impacts of this tax deduction not exclusively has to do with preserving the environmental

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services of ecosystems, it contributes in part because most environmental and stewardship organizations address these services on their projects and actions.

Regarding inheritance tax, although it is difficult to make a connection between this concept and environmental services, the tenth section of article number 27 of Law 19/2010, regulating inheritance tax and donations in Catalonia, contemplates a reduction of 95% of this tax for the transfer of property declared part of the Catalan Plan for Natural Areas or the Natura 2000 Network. It highlights the fact that the tax credit will be maintained only if management encourages conservation and a model of environmental services. However, this bonus applies only in protected areas, reducing significantly the potential scope of the region where preservation is ensured.

Municipal environmental taxation

The implementation of these taxes is not caused by environmental aspects or created for this purpose, but the municipalities have taken advantage of existing taxes in which environmental elements can be included. Examples are taxes on motor vehicles, on economic activities, among others (Adame and Adame, 2003).

According to Puig (2009), from the environmental point of view, municipalites should encourage a funding model fostering resource efficiency and avoiding environmentally damaging practices.

Before changes in 2002 the Local Tributes Law only contemplated the possibility of rebates aimed at protecting the environment in the tax on vehicles. However, for some city governments such as Barcelona, this has not been a barrier for the introduction of environmentally aimed taxes on real estate, buildings and economic activities, installations and works. Other municipalities, such as Tarragona, introduced tax deductions on building taxes, with various bonuses on bioclimatic design, insulation, installation of solar panels, the use of reusable materials, etc.

After changes in the law in 2002, an attempt, yet incomplete, was made to solve this gap in green tax bonuses. Local councils were specifically entitled to establish incentives in the business tax and the construction, installations and works tax, in addition to those already existing for the motor vehicles tax. Nonetheless, municipal environmental taxation has not been designed to conserve environmental services.

Authors such as Puig (2011) recommend the possibility of new taxes on activities that may damage environmental services and apply raisings on the conservation of this elsewhere. Specific taxes that could be created:

• Taxes on activities affecting the territory as golf courses, quarrying, skiing, etc. and on activities that use rural and natural environments, such adventure sports, water activities, etc.

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157Tools for regulating environmental services in Catalonia

• Tax on land use changes. • Tax on harmful substances. • Tax on the use of raw materials for pharmaceutical purposes. • Tax on pesticides and fertilizers.

This situation exposed opens two ways to apply environmental taxation tools for regulating environmental services of ecosystems:

1. Creation of an environmental taxation system in Catalonia, with the creation of new types of tax levied from activities that might degrade environmental services and proceeds reverting directly to their conservation.

2. Application of incentives and tax rebates to existing taxes for those who apply conservation techniques for environmental services present on their property.

4. Conclusions for a final proposal for Catalonia

From the analysis of the experiences of management and regulation of environmental services in different realities we tried to respond to the main objective of the article: to determine what is the appropriate tool for the legal regulation of environmental services in Catalonia. At the beginning of the investigation we wondered if the Catalan legal framework allows for the regulation of environmental services, if payment for environmental services is applicable to Catalonia, if land stewardship would be an effective tool to ensure the protection of environmental services, and which legal experiences of other countries can apply here, or inspire the formulation of environmental services legislation in Catalonia.

In response to these issues we can say that Catalonia has a sufficient legal framework to promote mechanisms for regulating the environmental services of ecosystems. The regulatory framework regarding contracts and property rights allows for the establishment of mechanisms that facilitate the management and conservation of environmental services. The legislative framework allowing tax rebates and tax incentives types is available. Finally, PES experiences as might be the case of Selvans in Girona Province and Agricultural Agreements in Menorca, show no need to display a specific legal framework to implement these programs.

With the analysis of the three tools for regulating the environmental services of ecosystems in Catalonia: payment for environmental services, land stewardship and environmental taxation, the extracted conclusion is that if we focus only on one of these methodologies, the model will be incomplete. In Catalonia, the remarkable history of land stewardship initiatives should not be underestimated and revisited to address how land stewardship may revert in improving the

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environmental services of ecosystems. That is why the proposal is divided into two lines, which could be grouped in a Strategy for the Management of Environmental Services in Catalonia:

• Land stewardship as a mechanism for private conservation of ecosystem services, by complementing existing taxes with incentives for owners applying best environmental practices that contribute to these services.

• Payment program for environmental services from private lands (without land stewardship agreements) funded through a new framework for environmental taxation in Catalonia, supported by an environmental tax law itself.

The rapid environmental changes, the degree of degradation of environmental services and the effects of human activity on the environment, justify Catalonia’s short-term adoption of e a number of regulatory mechanisms in this field. That is why the implementation of a strategy for the management of environmental services in Catalonia would be a good instrument to articulate the proposals mentioned above. However, this must be accompanied by the Natural Heritage and Biodiversity Act of Catalonia, which requires urgent processing and approval.

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Boisán, I. J. (2011). Sobre la configuración de un derecho real de custódia territorial. Seminari de seguretat jurídica i fiscalitat de la custòdia del territori. Xarxa de Custòdia del Territori.

Buñuel, M. (2002).  El uso de instrumentos fiscales en la política del medio ambiente: teoria, pràctica y propuesta preliminar para España.  Madrid: Fundación Biodiversidad.

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Puig, I. (2011). Cap a la fiscalitat per a la custòdia del territori. Seminari de seguretat jurídica i fiscalitat de la custòdia del territori. Xarxa de Custòdia del Territori.

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Laws

• “Edicte del Consell de Menorca: Aprovació de les bases que han de regir la vuitena convocatòria d’ajuts per a promoure pràctiques sostenibles a les explotacions agràries de Menorca mitjançant la subscripció del “contracte agrari de la reserva de la biosfera” i la convocatòria d’ajuts corresponents a l’any 2011”. Sessió ordinària del Consell Insular de Menorca (28 de febrer de 2011).

• “Llei 1/2008 de contractes de conreu”. DOGC (3 de març de 2008). • “Llei 16/2008, del 23 de desembre, de mesures fiscals i financeres”. DOGC

(31 de desembre de 2008). • “Llei 16/2013, de 29 d’octubre, per la que s’estableixen determinades mesures

en matèria de fiscalitat ambiental i s’adopten altres mesures tributàries i financeres”. BOE (30 d’octubre de 2013).

• “Llei 19/2010, de 7 de juny, de regulació de l’impost de successions i donacions a Catalunya”. DOGC (11 de juny de 2010).

• “Llei 21/2013, de 9 de desembre, d’Avaluació Ambiental”. BOE (11 de desembre de 2013).

• “Llei 38/2003, de 17 de novembre, general de subvencions”. BOE (24 de novembre de 2003).

• “Llei 42/2007, de 13 de desembre, del patrimoni natural i de la biodiversitat”. BOE (14 de desembre de 2007).

• “Llei 45/2007, de 13 de desembre, per al desenvolupament sostenible del medi rural”. BOE (14 de desembre de 2007).

• “Llei 5/2006 del Llibre cinquè del Codi Civil de Catalunya relatiu als drets reals”. DOGC (24 de maig de 2006).

• “Llei 51/2002, de 27 de desembre, de reforma de la Llei 39/1988, de 28 de desembre, reguladora de les Hisendes Locals”. BOE (28 de desembre de 2002).

• “Llei 8/2005, de 8 de juny, de protecció, gestió i ordenació del paisatge”. DOGC (16 de juny de 2005).

• “Llei Federal de Drets”. DOF Mèxic (31 de desembre de 1981). • “Llei Forestal 7575 de 1996”. La Gaceta de Costa Rica (16 d’abril de 1996).

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• “Ordre AAM/133/2013, de 10 de juny, per la qual s’aproven els preus públics per a la prestació de serveis en els espais naturals de protecció especial”. DOGC (26 de juny de 2013).

• “Ordre AAM/254/2012, de 31 d’agost, d’aprovació del Pla pilot de recol·lecció de bolets dins l’àmbit del paratge natural d’interès nacional de Poblet i de la seva zona d’influència, i de creació del preu públic vinculat al Pla pilot”. DOGC (6 de setembre de 2012).

• “Projecte de llei d’avaluació ambiental”. Ministeri d’Agricultura, Alimentació i Medi Ambient (30 d’agost de 2013).

• “Proposició de Llei de fiscalitat ambiental”. Mesa del Congrés dels Diputats (21 de juliol de 2009).

• “Proposició de Llei de la biodiversitat i el patrimoni natural”. BOPC (18 d’abril de 2011), Tram. 202-00039/09.

• “Real Decret 1336/2011, de 3 d’octubre, pel que es regula el contracte territorial com instrument per a promoure el desenvolupament sostenible del medi rural”. BOE (4 d’octubre de 2011).

Web pages

• Asociación de Empresas de Servicios de los Ecosistemas (Accessed: May 30th, 2012) [comunicarseweb.com.ar]

• Avaluació dels Ecosistemes de Mil·lenni (Accessed: April 20th, 2012) [www.millenniumassessment.org]

• Comisión Nacional Forestal (Accessed: May 14th, 2012) [www.conafor.gob.mx] • Consell Insular de Menorca (Accessed: June 6th, 2012) [www.cime.es] • • Diputació de Girona (Accessed: June 9th, 2012) [www.ddgi.cat] • Ecosystem Services for Poverty Allevation – ESPA (Accessed: June 5th,

2012) [www.espa.ac.uk] • GOB Menorca (Accessed: June 6th, 2012) [www.gobmenorca.com] • Global Environmental Facility (Accessed: November 26th, 2013) [www.

thegef.org] • Iniciativa per Catalunya Verds – Esquerra Unida i Alternativa (Accessed:

June 18th, 2012) [www.iniciativa.cat] • Mercados de Medio Ambiente (Accessed: October 8th, 2013) [www.

mercadosdemedioambiente.com]. • Natural Environment Research Council (Accessed: June 5th, 2012) [www.

nerc.ac.uk] • Nestlé Waters (Accessed: June 5th, 2012) [www.nestle-waters.com] • Parlament de Catalunya (Accessed: May 3th, 2012) [www.parlament.cat] • Programa Sèlvans (Accessed: June 9th, 2012) [www.selvans.cat] • UICN – Comitè Espanyol (Accessed: April 20th, 2012) [www.uicn.es] • Xarxa de Custòdia del Territori (Accessed: April 19th, 2012) [www.xct.cat]

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Testimonials and reflections from experiences

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163

Price versus value of ecosystem services in the southern Alps

Ann BrowerDepartment of Environmental Management

Lincoln University, New Zealand1

Ideas about landscape services and their ownership can shape the landscapes themselves. In working landscapes, John Locke s ideas feature prominently. This appears to be the case in New Zealand South Island land reform

There is something different about the light in the high country of New Zealand’s South Island. The details of topography are so crisp as to appear fake, or somehow digitally enhanced. A clear day is like stepping out of Plato’s cave; a cloudy day is like stepping into Narnia.

But high country landscapes are changing: from a vast pastoral estate owned by the Crown, leased to 300 pastoral farmers, to private subdivisions; and from broad expanses of tussock to viticulture, golf courses, and pig farms. From 1856 to 1992, the Crown owned 2.4 million hectares of South Island high country land. This is 10% of New Zealand’s landmass, and 20% of the South Island. In comparison, Israel is 2 million hectares, Belgium is 1.7 million hectares. For 150 years, the Crown leased land to runholders “for pastoral purposes only” (Land Act 1948 s.51(1)(d)). Under these pastoral leases, runholders’ property rights were strong (they enjoyed 33-year perpetually renewable terms) but narrow —no subdivision, no golf courses, and no soil disturbance (Brower, 2008, chapter 3).

1 Author of Who owns the high country?, stimulating a national debate about the on-going South Island land reform that is transforming the landscapes of the Southern Alps. Email: [email protected]

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164 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Figure 1. Geographic setting of New Zealand’s Crown pastoral estate (2.4 M ha : 10 % of NZ)

Source: Land Information New Zealand

In 1992, Hon. Denis Marshall was a minister of the Crown in the centre-right National party government. Somewhat unusually, he held the ministerial portfolios of both Lands and Conservation. This is unusual because NZ’s Department of Conservation manages about one-third of the nation’s landmass. So Hon Mr Marshall was responsible for nearly half of the nation’s land. In that post, he directed officials to begin a process with the obscure name of “tenure review” to

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165Price versus value of ecosystem services in the southern Alps

divide the high country. Land with value for biodiversity, landscape, or recreation could become public conservation land as parks or reserves. And land “capable of economic use” would be privatised (Brower, 2008, chapter 2).

At the time, most people who knew anything about it saw something good in tenure review, and supported the idea. The runholders saw something good in getting freehold title to the most productive parts of the high country. Conservationists saw something good in removing sheep from fragile land at high altitude, allowing tussock to recover. And recreationists saw something good in opening access for trampers and hunters. Hence for interest groups at the time, high country tenure review seemed win-win. But these three groups represented their particular interests, not the interests of the public at large and not the interests of the New Zealand taxpayer (Brower, 2006). The Crown, after all, represents the New Zealand citizen and the New Zealand taxpayer.

On the ground, tenure review has often resulted in land above 1,000 metres becoming public conservation land administered by the Department of Conservation (DoC), while productive land along river valleys and on lake shores has been, and is being, privatised and opened to development. This privatisation has paved the way for new viticulture (where Denis Marshall now lives and owns a vineyard), subdivisions in the Queenstown Lakes District, and the ‘for sale signs’ that sprinkle the Central Otago hills.

In short, tenure review is a transfer of land and potential wealth from the New Zealand public to a few hundred farmers. Tenure review is also carving up landscape with a blunt knife. What began as 110 Crown-owned parcels with an average size of 5,938 hectares, on which tenant sheep farmers could only engage in extensive sheep grazing, has so far been subdivided into nearly 3,000 parcels, of which at least 550 are smaller than 50 hectares. Following tenure review, these 865 parcels are now in private ownership, ready and waiting for development. So far, only about one quarter of this new freehold land has been sold by about 40 of the new landowners, but these tenants turned owners have grossed over 300 million New Zealand dollars ($ from hereon).

At the end of each tenure review deal, the Crown buys the runholder’s leasehold rights to land that will shift into the public conservation estate; and the runholder purchases freehold rights from the Crown for land to be privatised. Judging by the freehold land prices above, it would appear that what the Crown sold — freehold title and the option to develop the shores of Lakes Wanaka, Wakatipu, Tekapo, Pukaki and more —is more valuable than what it bought— the option to graze the unforgiving, high altitude country deemed incapable of economic use. Yet somehow the Crown has lost money while selling the freehold.

From 1992-2015, the Crown has purchased runholders’ leasehold rights to 330,000 hectares for $117 million; and leaseholders have purchased freehold rights to 370,000 hectares for $62 million. The median price at which the Crown sold freehold rights is $278/hectare to date; and the median price at which the Crown purchased leasehold rights has been $77/hectare.

Hence, on net, the NZ government has paid the new owners of valuable land $55 million more than they have paid the Crown.

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166 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Figure 2. Lake Wanaka, New Zealand.

Source: A. Brower

Let us take the harshly beautiful landscape surrounding Lake Wanaka as an example. Between 2003 and 2005, the Crown privatised 18,957 hectares. In these deals, the Crown shifted 4,767 hectares of mostly higher elevation land into public conservation land. Thus, it privatised the most developable 80% of this land, and conserved the least productive 20%. At the conclusion of the deals, on net the Crown paid the former runholders (and new possessors of freehold title) $263,000.

In the case of Glendhu Station on the south shore of Lake Wanaka, the Crown privatised 91% (over 2000 hectares), while conserving fewer than 300 hectares. And the Crown paid $50,00 net to the new owners of 91% of Glendhu. These new owners have since applied for consent to build a $10 million golf course with villas and accommodation for 200. If that proposal fails, they have suggested proposing a pig farm on the shores of Lake Wanaka.

It is surprising for the Crown to pay more to purchase leasehold rights than to sell freehold rights if one subscribes to the following Common Law ideas about property and ownership:

• that property is a bundle of rights, not an absolute; • that renting and leasing are different from owning;

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167Price versus value of ecosystem services in the southern Alps

• that leasehold grazing rights, no matter how long-lived, are very different to freehold rights;

• that the option to develop land such as the shores of Lake Wanaka is more valuable than the right to graze the same shoreline.

According to this common law understanding of property rights, the Crown shouldn’t have lost money in the process of tenure review. I set out to find out why it had.

As I started to study the politics and economics of tenure review, it became clear that the entrenched power of the farming lobby had consistently and fundamentally altered the perceived truth about who owns the landscape values of the high country. I found that, rather than taking an active stance in land reform negotiations, mid-level bureaucrats directed the government contractors making deals to be neutral in negotiations.

The trouble is that the Crown and its taxpayers have a sizeable financial interest in the high country; and the Crown’s failure to advocate for that interest was a tacit agreement to lose. The Crown forfeited its rights, and the public’s, from the start. Farmers and some environmentalists could call land reform deals “win-win” only because the Crown was agreeing to lose millions (Brower, 2006).

The high country farmers were livid about my findings. Their fury suggested there might be another reason the Crown was giving away the high country and paying the lucky recipients. It suggested there are two very different understandings of property rights at play in the high country.

First, there is the modern, Common Law bundle-of-rights view which I had naively assumed everyone shared. But there is a much deeper, more cultural and romantic view, first propounded by the seventeenth century English philosopher, John Locke. In the Lockean view, when a person works and improves land he or she comes to own it. But in the modern legal view, the rights are shared; work and improvements do not change ownership unless the statute says so, and New Zealand’s Land Act 1948 does not say so. It seemed that the bureaucrats in charge of deals subscribed to the Lockean view —romantic, patriotic, and several centuries out of date.

Also during this controversy, the Minister of Land Information in the Labour government, David Parker, released some data I’d been denied three times by Land Information NZ bureaucrats. This was data about who paid whom how much money in each individual tenure review deal. Analysis revealed that:

1. the leaseholders who privatised the most land also got the best deal. That is, those who kept the least land got the worst price per hectare, and those who kept the most land got the best price per hectare;

2. while privatising high country land, the Crown was giving a bulk discount. The more land it privatised, the less it charged per hectare. Essentially, the Crown paid whatever it took to close tenure review deals (Brower et al., 2010).

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168 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Throughout, high country runholders stuck doggedly to the Lockean work-to-own view of ownership that made it seem clear that they owned the high country, and deserved to be paid as though a lease were the same as freehold. The bureaucrats followed the farmers’ Lockean lead.

This Lockean view is also intimately related to and bolstered by the cultural imagery of the high country. The “Southern Man” is not just this year’s Speight’s beer advertising campaign, but every year’s Speight’s campaign. These pop culture images encourage the perception that lessees’ labour created freehold ownership of the land. It seems that the Lockean ideas and cultural ideals had transformed renters into owners. The Lockean view of land ownership also justified the Crown paying public money while giving away public assets (Brower, 2008; chapters 9-10).

One would expect the Crown to stick up for the many, and for the public interest. The declassified payments data suggest that the Crown failed to try to stick up for the many, and succeeded admirably at failing. But neither corruption nor conspiracy was to blame. Rather, the resonant imagery of the do-it-yourself, work-to-own theory of property created a hegemony of farmer dominance. This dominance turned even the otherwise left-leaning Labour-led government under Rt Hon Helen Clark into the anti-Robin Hood —taking from the many to give to a few.

Figure 3. Richmond Station, Lake Tekapo, New Zealand.

Source: A. Brower

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169Price versus value of ecosystem services in the southern Alps

In June 2007, the Labour government changed the policy. This followed growing controversy, but was triggered by outrage over a deal to privatise nine kilometres of Lake Tekapo shoreline while paying the runholder $325,000. The Crown had always held the power to veto a deal at any time, but had never used it. Then, in one fell swoop, Cabinet all-but vetoed more than 50 deals on runs within five kilometres of a southern lake unless the deal met the Crown’s criteria. To be approved for funding, future deals near lakes would need to prevent shoreline subdivision or “significant alteration” and water pollution. In sum, the Crown announced it would no longer enter every tenure review proposed by a runholder, but would first consider whether tenure review could adequately protect conservation values. If the deal proceeded, it would be subject to a much more rigorous set of tests than before the 2007 changes.

It is striking that it took 16 years for a Cabinet to insist that each deal must be good for the Crown. These changes came none too soon for many. One DoC staffer commented, “If the minister hadn’t insisted on better outcomes, that land would be in pivot irrigators tomorrow. Guaranteed.”

But the changes to New Zealand’s land reform policy did not last long. On 26 August, 2009, the centre-right National-led cabinet repealed most of the Labour Government’s changes to tenure review. A 2009 Cabinet minute revokes the previous government’s attempts to protect lakefront land from intensive development, subdivision and other forms of alteration of the landscape. It suggests replacing the strong protection of lakefront landscapes with “time-limited covenants, for example preventing subdivision for a specified period to allow district plans to be amended”. One might also wonder why no one has made a fuss about renewed privatisation of hundreds of thousands of hectares of high country. Perhaps it is because the day after announcing changes to tenure review, the government revealed the possibility of mining on public conservation land.

Thus, 23 years later, we are back to where we started —with the quiet but steady privatisation of the New Zealand high country.

References

Brower, A. (2006). Interest groups, vested interests, and the myth of apolitical administration: The politics of Land Tenure Reform on the South Island of New Zealand. Report to Fulbright-New Zealand. at https://researcharchive.lincoln.ac.nz/bitstream/handle/10182/1196/brower_fulbright.pdf?sequence=1

Brower, A. (2008). Who owns the high country? Nelson, NZ: Craig Potton Publishers.

Brower, A., Meguire, P., and Monks, A. (2010). Closing the deal: Principals, agents, and sub-agents in New Zealand land reform. Land Economics, 86(3), 467-492.

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171

Enabling stakeholders to apply the ecosystem services concept in practice

Josep LascurainServeis Globals Mediambientals (SGM),

Barcelona, Spain1

A simple goal/question: is it possible to create a method of urban dune creation and management, similar to those employed on managing urban ray-grass lawns?

Over this argument there is a real opportunity to transform the fate of a lot of European Mediterranean dune systems which most are just urban, and so are landscapes which endure the high impact of human frequentation.

Beaches are the most visited landscape on the mediterranean coasts of Spain, France and Italy. Most of them are cleaned by mechanical sieving and the rate of human use is expanding from summer throughout the whole year. So, most of them are “urban” beaches and if there were dunes, they could be considered as urban dunes. Some characteristics of those dune ecosystems are the impaired aeolian sand transport processes, the disappearance of dune relief, and the disappearance of mobile and semi-fixed dunes along its biodiversity.

So, if we want to have new dunes, specifically if we want to create semi-fixed dunes, there is a need of engineering. The concept behind the Hybrid dunes project is about this: to learn how to construct urban dunes, and learn how to manage the huge demand of ecosystem services on such threatened landscapes.

The Hybrid dunes project is a “case study” (or exemplar) of the OPERAs project (http://www.operas-project.eu/). The development of the project has been done and supported by the Barcelona Metropolitan Administration. And it has a two-sided aim: learning to construct and manage urban dunes, and also learning how to manage the balance between a huge ecosystem service (ES) demand and a provision of those ES limited by our capacity to construct and restore dunes.

1 Coordinator of a case study in OPERAs, a European research project aiming to put cutting edge ecosystem science into practice. Email: [email protected]

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172 Ecosystem Services: concepts, methodologies and in truments for research and applied use

Figure 1: Social-ecological context and ecosystem services of urban beach dune systems

On this scenario, the different stakeholders play a key role. And enabling them to apply the ES can open the window to innovation on coastal management.

One of the most outstanding outcomes has been the discovery that the dune management can be an ecological and cheap alternative to the beach nourishment official strategy, with an annual budget around one million euros, and moving 100,000 cubic meters of sand from the south extreme of the metropolitan beach of Barcelona to the mouth of the Llobregat river each year. The creation of the so called “winter dunes” strategy, used in Emilia-Romagna coast to protect tourist beaches, can be adapted to reconstruct the lost profile of the dune-beach system and so, protecting the coast against the f lood and storm risk increased by the global climatic change. And what is most important, at a much lower cost and ecological impact. On this issue, the modification of the Spanish Shore Act (art. 9 RD 876/2014) opens the window to other stakeholders to propose and, by getting approval, execute works on coastal defence.

Governance can go beyond the administrative context of coastal management. The intensive use of the Barcelona’s metropolitan beaches (the huge demand of ES) is multi-faceted: the social use varies between the summer sun-bathing period (June-September) and the rest of the year; and the ways people use to arrive to the beaches are determinant on types of users, of their behaviours and needs.

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173Enabling stakeholders to apply the Ecosystem Services concept in practice

Figure 2: participative dune systems restoration

Source: J. Lascurain

The combination of location, means of transport and time of the year combine to shape different groups of dune visitors.The ecological impacts derived from social use are strongly dependent of the opportunity cost to arrive to the beaches. The time of the year and the way to arrive (big parking lots, railway stations, urban areas close to the beaches,..) can so create a certain taxonomy of the potential dune visitors.

The most relevant trade-offs of social use are: trampling over the dunes (flattening the dunes), unleashed dog walking (with impacts over marram grass, the dune constructor plant which needs strictly oligotrophic soils, and over fauna), and stray cat colonies feeding (with impacts over nearly extinct dune fauna).

There are also some surprising facts, local NGOs are strongly against the ES concept as a collateral of the feeling that nature must not be considered under the utilitarian point of view. There is a need to cope with narratives which collide with the ecosystem service (ES) concept. And the way to cope with that needs to go beyond the conventional wisdom. We have to identify the narratives that shape those feelings.

So, along the complex administrative environment, there is a complex social system, with different socials groups demanding different uses and also different moral narratives.

To cope with all this we are conducting different strategies:

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174 Ecosystem Services: concepts, methodologies and in truments for research and applied use

• Translate to geospatial information the different social groups: how they arrive there, where they come from, with which frequency, how much people, etc….

• Try to know their needs, priorities and try to uncover what shapes their narratives about “beach wilderness” by the use of interviews and online surveys.

• From this knowledge tailor a communication program (environmental education) in order to be most effective to all relevant groups.

And if we get all these aims, we could find that over the 25% of all dune ecosystems could be “hybrid dunes” in Catalonia.

Figure 3: Hybrid dunes – a green infrastructure approach to urban-bean dune system restoration

Source: J. Lascurain

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